Faculteit Bio-ingenieurswetenschappen

Academiejaar 2013 – 2014

Millipede communities in small forest patches in contrasting agricultural landscapes

Willem Proesmans

Promotor: Prof. dr. ir. Kris Verheyen

Copromotor: Prof. dr. Dries Bonte

Tutor: Pallieter De Smedt

Masterproef voorgedragen tot het behalen van de graad van

Master in de bio-ingenieurswetenschappen:Bos & Natuur

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Table of contents

Abstract…………………………………………………………...…………………………………...... 5 Samenvatting…………………………………………………………………………………………….7 Foreword……………………………………………………………………………………….…………9 Introduction……………………………………………………………………………………………...10 1. Habitat fragmentation ...... 10 1.1. Situation in Western Europe ...... 10 1.2. Consequences of habitat fragmentation ...... 11 a. Processes……………………………………………………………………………….11 b. Fragmentation and Arthropod community //…………………………………………..13

1.3. Importance of forest age ...... 14 2. Millipedes ...... 15 2.1. Systematics ...... 15 a. Classification of Diplopoda within Arthropoda…………...... 15 b. Internal phylogeny and classification………………………………………………..16 2.2. Morphology ...... 17 2.3. Ecology ...... 18 a. General……………………………………………………………………………….18 b. Community composition and habitat requirements………………………………….19 c. Ecosystem services: millipedes as litter decomposers……………………………….21 d. Millipedes and ecological restoration………………………………………………..21 e. Detrimental effects: pest species and invasions……………………………………...22 Goals…………………………………………….……………………………………………….24 Material & methods…………………………………………………………………………….25 1. Sampling sites ...... 25 2. Sampling ...... 25 3. Data analysis ...... 29 3.1. General effects on diversity, species richness and density ...... 29 3.2. Unconstrained ordination: DCA ...... 30 3.3. Constrained ordination: RDA and CCA ...... 31 3.4. Indicator species analysis ...... 31 Results……………………………………………………………………………………………33

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1. Difference between regions ...... 33 2. Diversity and density differences between forests ...... 39 3. Detrended Component Analysis (DCA) ...... 43 4. Indicator species analysis ...... 47 Discussion………………………………………………………………………………………..51 1. Differences in species composition between regions ...... 51 2. Factors influencing community composition and abundance at the regional level ...... 52 2.1. Effect of land use intensity ...... 52 2.2. Effect of forest age ...... 54 2.3. Effect of patch size ...... 57 2.4. Effect of location in the forest ...... 58 2.5. Effect of tree species ...... 59 2.6. Effect of tree, shrub, herb and moss cover ...... 59 2.7. Effect of dead wood ...... 60 2.8. Other factors ...... 60 3. Impact on ecosystem function ...... 60 4. Implications for landscape management ...... 61 Conclusion……………………………………………………………………………………….62

Future research...……………………………………………………………………………….63

Acknowledgements……………………………………………………………………………..64

References………………………………………………………………………………………65

Appendix 1: tree species…………………………………………………………………….....76

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Abstract

In the intensively used Western European landscape, only small forest fragments remain in many places. These fragments may, despite their small size still provide important ecosystem services, such as biological control, pollination, nutrient recycling,… However, these services are possibly underappreciated and the forest fragments are therefore often not protected.

Nutrient recycling is an important ecosystem service, in which litter decomposers play an important role. Millipedes are, together with woodlice, harvestmen and earthworms among the most important decomposers in forest ecosystems. They increase the decomposition mostly by physical fragmentation and by facilitating establishment of soil micro-organisms.

However, millipedes are known as very slow dispersers, and therefore the community may be adversely influenced in fragmented landscapes with intensive land use. Also, forest size and age might influence species composition. To assess the influence of these factors, this study investigated forests differing in age and size located in agricultural landscapes in five regions in Western Europe. In each region, sixteen forest fragments in an intensively used (‘open field’) and a less intensively used (‘bocage’) agricultural landscape were investigated.

The study shows that the species composition differed markedly between the open field and the bocage landscape. Forest fragments in the intensively used open field landscape housed a more species rich community than in the bocage landscape (on average 7.0 species per patch vs 6.2). Furthermore, typical indicator species could be found for the open field landscape, where especially Polydesmus inconstans and Cylindroiulus caerulescens were very abundant. These species are known as typical inhabitants of more open habitat and have also been reported as pest species on crops.

Furthermore, some species are typical for old forests, typically having a low dispersal capacity and the need for habitat with long term stability. In three of the five regions, clear differences in community composition between the old and recent forest fragments were very clear, with especially Glomeris marginata as an indicator species of old forest.

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In contrast to age and surrounding landscape, the size of the forest fragments was only of limited importance. Millipede relative abundance was slightly higher in small forest fragments, but species composition was similar. This could be explained by the fact that millipedes only need small surfaces to support a viable population. Therefore, even the protection of very small forest fragments may be meaningful for this taxonomic group.

The conclusion of this investigation is that the surrounding landscape has a major influence on species composition. Forest fragments in intensely used landscapes house a species community with several species typical for more open habitat. These species are often known pest species that might need forests as refuges or as breeding habitat. Furthermore, certain species are confined to old forest remnants. To safeguard the continued existence of these species, especially the old forest fragments deserve attention and should be better protected in the future.

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Samenvatting

Het landschap in West-Europa wordt zeer intensief door de mens gebruikt. Daardoor blijven op veel plaatsen slechts kleine fragmenten bos over. Ondanks hun beperkte grootte kunnen deze fragmenten toch belangrijke ecosysteemdiensten voorzien, zoals biologische controle, bestuiving, nutriënt cycling,... Het belang van deze diensten wordt echter mogelijk onderschat en de bosfragmenten zijn hierdoor vaak onbeschermd.

Nutrient recycling is een belangrijke ecosysteemdienst, waarin strooiselverteerders een essentiële rol spelen. Miljoenpoten behoren, samen met pissebedden, hooiwagens en regenwormen tot de belangrijkste verteerders in bosecosystemen. Ze versnellen strooiselvertering vooral door fysische fragmentatie en door het bevorderen van de vestiging van micro-organismen in de bodem.

Miljoenpoten staan echter gekend als zeer trage verspreiders. Hierdoor kan de gemeenschap worden beïnvloed in landschappen met intensief landgebruik. Leeftijd en grootte van bosfragmenten kan ook soortsamenstelling mee bepalen. Om het effect hiervan te evalueren werden in dit onderzoek bossen van verschillende leeftijd en grootte behandeld in vijf regio’s in West-Europa. In iedere regio werden 16 bosfragmenten in een intensief gebruikt (‘open field’) en een minder intensief gebruikt (‘bocage’) landschap onderzocht.

Dit onderzoek toont aan dat soortensamenstelling duidelijk verschilt tussen het ‘open field’ en het ‘bocage’ landschap. Bosframgenten in het intensief gebruikte landschap zijn over het algemeen soortenrijker (gemiddeld 7,0 soorten vs 6,2). Verder zijn er enkele zeer typische indicatorsoorten voor het ‘open field’ landschap, waar in het bijzonder Cylindroiulus caeruleocinctus en Polydesmus inconstans zeer talrijk waren. Deze soorten zijn typisch voor open habitat en zijn gekend als plaagsoorten op landbouwgewassen.

Verder zijn bepaalde soorten typisch voor oud bos, vooral sooretn met een beperkte dispersiecapaciteit en de noodzaak voor habitat dat stabiel blijft op lange termijn. In drie van de vijf regio’s bestonden duidelijke verschillen in soortensamenstelling tussen oude en recente bosfragmenten, waarbij vooral Glomeris marginata naar voren kwam als een soort die typisch is voor oud bos.

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In tegenstelling tot leeftijd en omringend landschap was de grootte van de bosfragmenten van zeer beperkt belang. De dichtheid aan miljoenpoten was iets hoger in kleine bosfragmenten, maar de soortensamenstelling was gelijkaardig. Dit kan verklaard worden doordat miljoenpoten waarschijnlijk slechts een kleine oppervlakte nodig hebben om een leefbare populatie te handhaven. Het heeft dus ook nut om de allerkleinste bosfragmenten te beschermen.

De conclusie van dit onderzoek is dat het omringende landschap een grote invloed heeft op soortsamenstelling. Bosfragmenten in intensief gebruikt landschap bevatten een gemeenschap met verschillende soorten typisch voor open habitat. Deze soorten zijn vaak gekende plaagsoorten die bossen mogelijk als refugia of voortplantingshabitat gebruiken. Andere soorten zijn dan weer beperkt tot oude bosfragmenten. Om het voortbestaan van deze soorten te garanderen verdient de bescherming van deze fragmenten extra aandacht.

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Foreword

This MSc-thesis is part of the European ‘smallFOREST’-project (http://www.u- picardie.fr/smallforest/uk/), which attempts to assess the importance of small patches of deciduous forests in intensively used agricultural landscapes in terms of biodiversity and ecosystem services. These forests often consist of patches of different quality, age, size… which might have an impact on their biodiversity and the ecosystem services they provide.

In most parts of Western Europe, centuries of intensive land use have reduced the forest cover, and often only small forest fragments remain. The smallFOREST-project attempts to quantify biodiversity and ecosystem function of small forest fragments in Western Europe. This study focuses on millipedes, which are important decomposers. However, the project as a whole is much broader and attempts to assess the importance of these forests as carbon stocks, habitat for biological control agents and source of biodiversity, but also looks at their importance for social and recreational uses.

In the end this project will hopefully provide us with a broad, multidisciplinary view on the importance of small forest fragments and provide society with guidelines for management of these forests and of agricultural landscapes in general, an issue that will become even more important in the future with the increasing anthropogenic pressure on the natural environment.

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Introduction

1. Habitat fragmentation

1.1. Situation in Western Europe In Western Europe human population pressure has caused fragmentation of the original forest into small areas, often surrounded by a matrix of intensively used human-dominated land. These forests might, despite their small size, still play an important role in conserving biodiversity by acting as refuges or stepping stones (e.g. HONNAY et al. 1999; COUSINS & ERIKSSON 2008). However, they are often not protected and are prone to degradation. In the light of the current biodiversity crisis, protection of these habitat patches might be important. In addition, the presence of forest fragments in agricultural landscapes can serve many functions, such as biological control (TSCHARNTKE et al., 2007), pollination (RICKETTS, 2004), wind speed regulation, … .

Of the 22,932,740 km² area of Europe, 10,199,400 km² (44.5%) is covered with forests (FOREST

EUROPE, 2011). However, the forested area is not spread equally over Europe (table 1) and the forest landscape is highly fragmented. 40% of all forest land lies within 100m distance of another land use, thus suffering from edge effects and being prone to invasions by diseases and pest species (ESTREGUIL et al., 2012). Despite an average annual increase of 0.8% in forest area, forest connectivity does not necessarily increase and even minor forest loss can severely increase the isolation of forest fragments (ESTREGUIL et al., 2012). As shown in figure 1, the degree of forest fragmentation is highly variable. Eastern Europe and Scandinavia typically contain a large forested area that is well connected. In Northwestern Europe, the few forest that remains is highly fragmented due to anthropogenic factors, such as agriculture. In Southern Europe, forests also seem to be fragmented, but this is nowadays, at least partly, due to natural factors, such as climate and soil properties – often caused by historic land use – that do not allow the growth of forest on these locations. These forests are also of a distinct type, largely consisting of coniferous and sclerophyllous trees. The forests in this study therefore suffer more from anthropogenic fragmentation than the other European forest types. Most coniferous forest is situated in mountainous regions and in the northern, less populated countries and thus suffers less from

10 habitat fragmentation. Therefore, studying the effects of forest fragmentation in this forest type is very relevant in the context of West European landscapes.

Table 1: Total forest area and forest index of the countries in this study. Except for Sweden all these countries have a forest index below the European average (Forest Europe, 2011).

Total forest area (km²) Total area (km²) Forest index (%)* Belgium 7,060 30,280 23 France 175,720 550,100 32 Germany 110,760 348,770 32 Sweden 306,250 410,310 75 * The forest index for the country as a whole does not necessarily correlate with the forest index in the study areas as locations with a highly fragmented forest landscape were chosen (e.g. in Sweden).

Figure 1: Forest connectivity in Europe. The degree of connectivity was calculated per 25 x 25 km squares. The index varies from 0% (forest patches completely unconnected) to 100% (all forests connected), taking into account the inter- patch distances and patch sizes and calculated for species that can disperse up to 1km (map FOREST EUROPE 2011). 1.2. Consequences of habitat fragmentation a. Processes

In the first place, habitat fragmentation leads to habitat loss, which is the major threat to terrestrial ecosystems (SALA et al., 2000). Additionally, habitat fragmentation may increase the effects of habitat destruction, thus acting in a synergistic way (FAHRIG, 2002). Due to

11 fragmentation, the extinction threshold increases, making species more prone to extinction

(FAHRIG, 2003). However, the effects of habitat fragmentation seem to be weaker than the effects of habitat loss and the effect of fragmentation is not always straightforward (EWERS & DIDHAM,

2006; FAHRIG, 1997). Furthermore, species might not respond immediately to habitat loss and fragmentation, but show an extinction debt, and go extinct years after the habitat was degraded

(KUUSSAARI et al., 2009).

Because of fragmentation, the average size of forest fragments decreases. Species that need a minimum patch size might go extinct when the area of forest fragments decreases beneath a certain threshold (FAHRIG, 2001).

In addition, by fragmenting forests, the relative share of forest edges is increased. This makes that a larger portion of the forest area is influenced by edge effects. MURCIA (1995) distinguishes three different edge effects: abiotic effects, which change the environmental conditions because of the proximity to a different type of habitat; direct biological effects, which are changes in the community caused by the physical condition at the edge (e.g. higher temperature) and indirect biological effects caused by species interactions such as parasitism, competition and predation.

Known effects are changes in species composition (GIBB & HOCHULI, 2002; RAND et al., 2006) including larger susceptibility to invasion by exotic species (COLLINGE, 1996), higher maximum temperature (FETCHER et al., 1985), more temperature fluctuations (CHEN et al., 1995; FETCHER et al., 1985), higher wind speed (CHEN et al., 1995) lower moisture content (CHEN et al., 1995;

REDDING et al., 2003), faster mineralization and higher nutrient availability due to higher mineralization rates and atmospheric deposition (WEATHERS et al., 2001).

Smaller forest fragments are more sensitive to these effects, as a larger share of these fragments can be considered as ‘edge’ (BARBOSA & MARQUET, 2002). A synergistic interaction between area effects and edge effects, which increases the effects even more can also be present (EWERS et al. , 2007). The impact of edge effects differs, however, which is probably caused by several confounding factors, such as the matrix quality. The matrix forms a habitat for certain species that can influence the data by increasing the diversity at the edges, thus acting as ‘noise’ in species data sets that investigate the effect on diversity of forest species (COOK et al. , 2002;

KUPFER et al., 2006; LÖVEI et al., 2006). Furthermore, forests are not necessarily the only

12 relevant habitats for many species, as they can compensate for habitat loss by using resources in the matrix (EWERS & DIDHAM, 2006).

In addition to the known effects of habitat fragmentation on biodiversity, the effects of global climate change may act in a synergistic way by impeding migration and gene flow (OPDAM &

WASCHER, 2004). While models indicate that many specialists with poor dispersion abilities seem to cope with climate change, the combination with habitat loss and fragmentation may sharply increase the chance of extinction (MCLAUGHLIN et al., 2002; TRAVIS, 2003). This indicates that forest ecosystems in the highly fragmented landscape of West Europe may be at an even higher risk due to climate change.

b. Fragmentation and arthropod communities

Habitat fragmentation is known to have a severe impact on arthropod community composition. Generalist species become more abundant in disturbed, fragmented landscape, while specialists decline (DEVICTOR et al., 2008; MARVIER et al., 2004). In this way, fragmentation can make ecosystems prone to invasions of generalist species (MARVIER et al., 2004). Species that disperse slower, such as most millipede species, also suffer from habitat fragmentation.

Decomposers might go extinct in forest fragments because of a lack of suitable resources, changes in microclimate and stochastic events (DIDHAM et al., 1996). In carrion beetles, which are, like millipedes, important decomposers with low mobility, both the diversity and the abundance are known to decrease. Additionally, the community structure is known to shift to small-bodied generalists and more mobile groups, such as muscid flies will make up a larger portion of the decomposer guild (GIBBS & STANTON, 2001). In termites, the community is known to shift from soft bodied, soil-feeding species to more hard-bodied litter-feeders, which are less sensitive to changes in microclimate (DE SOUZA & BROWN, 1994). Forest fragmentation has proven to influence the community composition of millipedes in the Tropics (GALANES &

THOMLINSON, 2011), showing a decrease in species richness in more isolated forest patches.

Habitat fragmentation might not only change the community composition, but also the ecosystem function. Decomposition by carrion beetles is known to slow down in fragmented forests (KLEIN, 1989). In small forest fragments, litter decomposition seems to slow down and decomposition rates become more unpredictable and variable (DIDHAM, 1998). However, VASCONCELOS &

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LAURANCE (2005) did not find an effect of invertebrate abundance, species richness or species composition on decomposition rates, despite finding a clear difference in decomposition rate between primary and second-growth forest.

At the within species level, genetic variability in fragmented arthropod populations may be reduced, which may have an adverse impact on the resilience of the population to changes in habitat (KELLER & LARGIADÈR, 2003). Due to fragmentation, species also tend to become less abundant and widespread (GONZALEZ et al., 1998; HADDAD & BAUM, 1999), with rare species being more prone to extinction (GONZALEZ & CHANETON, 2002).

Forest fragmentation also influences biotic interactions between species. By decreasing pathogens and parasitoid loads, forest fragmentation might, for example, increase outbreaks of pest species (KRUESS & TSCHARNTKE, 2000; ROLAND, 1993).

1.3. Importance of forest age Stability through time is an important feature of forest ecosystems. Old forests house communities that differ distinctly from recently planted ones. The definition of ‘old’ forests differs per country, depending on the data available to determine the forest’s minimum age. Attempts to define old forest also make use of structural features, such as the amount of large trees, canopy layering, amount of dead biomass and number of dead trees (FRANKLIN & VAN

PELT, 2004).

Arthropod communities show marked differences between old forest and recent forest (SIPPOLA et al., 2002). This can influence ecosystem function. For example, the number of aphids in the canopy is an order of magnitude larger in recent, regenerating forests than in old forests, thereby causing significant damage to the trees (SCHOWALTER, 1989). Old forests seem to house a larger predator diversity, which decreases herbivory (SCHOWALTER, 1989). In general species diversity seems to increase with site productivity, amount of coarse woody debris and tree species diversity (SIPPOLA et al., 2002).

Because of human land use, the fast turnover in land use has caused old forests to become rare and fragmented in Western Europe. Despite this, they may still provide important ecosystem services and house a distinct community of soil arthropods.

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2. Millipedes Millipedes form a very diverse and species-rich taxon of terrestrial organisms. Despite their abundance and ecological importance they are relatively poorly studied and little is known about their ecology, morphology and systematics. Here, a general introduction is given.

2.1. Systematics a. Classification of Diplopoda within Arthropoda

Millipedes (Diplopoda) belong to the arthropods (Arthropoda), which also contains insects, crustaceans and arachnids. Arthropoda forms a vast taxon, with about a million known species, and tens to hundreds of millions of species that still await description. It is estimated that at least 80% of all known animal species fall within this taxon. Arthropods are characterized by their segmented body, their exoskeleton of α-chitin, ecdysis (molting) and jointed segmental appendages. Due to their adaptive diversity, they survive in virtually every environment and often dominate ecosystems.

Figure 2: (a.) Phylogeny of Arthropoda, based on Giribet et al. (2001), (b.) Internal phylogeny of Myriapoda, based on Regier et al. (2005)

Within Arthropoda, Myriapoda forms a smaller, monophyletic subtaxon of about 16,000 known species. They are characterized by their segmented body with an elongate trunk consisting of many leg-bearing segments, which is not divided into a thorax and an abdomen. They form the sister taxon of Pancrustacea (Crustacea + Hexapoda, figure 2a) (GIRIBET et al. 2001). Myriapoda 15 consists of four subtaxa (figure 2b). In addition to the well-known millipedes (Diplopoda, 12,000 known spp., SIERWALD & BOND 2007; BREWER et al. 2012) and centipedes (Chilopoda, 3,300 known spp., ANDERSSON et al. 2005), two small taxa: Pauropoda (750 known spp., ANDERSSON et al. 2005) and Symphyla (200 known spp., ANDERSSON et al. 2005), consisting of minute, centipede-like animals exist. An internal phylogeny of Myriapoda was carried out by REGIER et al. (2005) and shows that Pauropoda forms the sister taxon of Diplopoda.

All members of Myriapoda are terrestrial, but because of their relatively permeable cuticle they are prone to dessication, which often limits them to humid environments, such as forest soils.

Millipedes form the largest taxon within Myriapoda, with about 12,000 known extant species

(SIERWALD & BOND, 2007). It is, however, estimated that at least 80,000 species exist (SIERWALD

& BOND, 2007), though empirical support for this claim seems to be lacking (BREWER et al., 2012).

b. Internal phylogeny and classification

The tremendous morphological diversity is reflected in the fact that of the 2,947 described genera of millipedes, 68% is monotypic. They are classified in 145 taxa of the ‘family’ level (SIERWALD

& BOND, 2007). The internal phylogeny of Diplopoda is poorly studied, and many questions remain (BREWER et al., 2012). For a thorough revision on millipede ecology, we refer to

SIERWALD & BOND (2007), which contains a phylogenetic analysis of Diplopoda using both morphological and molecular characteristics. Here, only the most important West-European taxa are discussed.

Glomeridae: a large, holarctic taxon with often unclear taxonomy (KIME & ENGHOFF, 2011), represented in this study by two species: Glomeris marginata and G. intermedia. The species are relatively short and capable of rolling up as a mean of protection. Both species in this study are known as typical inhabitants of forests (BERG et al., 2008). They may look like woodlice, but they have more than seven pairs of legs. Another special feature of this taxon is that the last leg pairs of the male are modified to gonopodes.

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Polyzoniidae: in this study only represented by Polyzonium germanicum, which is typical for moist forests (ANDERSSON et al., 2005). Typical for the family is the pointy head with reduced mouthparts. The gonopodes are situated at the seventh body ring of the male.

Craspedosomatidae: all species in this area have 29 body segments. They are characterized by compound eyes in the shape of an equilateral triangle. Another conspicuous characteristic is the presence of lateral protuberances at each of their body segments. In this study, the taxon is represented by Craspedosoma rawlinsi.

Chordeumatidae: grouped together with Craspedosomatidae in Chordeumatida. They also have 29 body segments. Leg pairs 7-11 are modified in the male and serve as gonopodes and associated organs. Mostly small, white species, represented in this study by the genera Melogona, Mycogona and Orthochordeumella.

Polydesmidae: species in the study area have 19 or 20 segments. The most striking characteristic of this taxon is the dorsoventrally flattened body. The seventh leg pair in males is modified to gonopods, the morphology of which is essential for certain identification in males. For females, the shape of the epigyne is the most important characteristic for species identification. In this study, the taxon is represented by the genera Polydesmus, Propolydesmus and Brachydesmus.

Blaniulidae: grouped together with Julidae in the taxon Julida. Long, very slender, snake-like animals with the gonopods on the seventh body segment. Here, the taxon is represented by the species Choneiulus palmatus, Blaniulus guttulatus and Proteroiulus fuscus.

Julidae: the largest group in this study. Most species, including the genera Cylindroiulus, Julus, Tachypodoiulus, Leptoiulus, Allajulus and Ommatoiulus belong to this taxon. They are snake- like, mostly dark-coloured species that often reach relatively large sizes. The shape of the protuberance on the telson is often very important for a correct identification, together with the shape of the gonopods, which are situated at the seventh body segment.

2.2. Morphology Here, the general external morphology of millipedes is described. For a more thorough review on millipede morphology and physiology we refer to Hopkins & Read (1992). Millipedes form a morphologically diverse group, from the isopod-like Glomeridae, to the small, hairy Polyxenidae

17 or the long, worm-like Julidae. They vary in size from 2mm to nearly 30cm. The most important diagnostic feature for millipedes is the occurrence of diplosegments (SIERWALD & BOND, 2007): the body rings are in reality composed of two segments, each of which carries two pairs of legs. The first trunk segment behind the head is the collum, which lacks legs. The next three segments all carry one pair of legs.

The gonads open on or behind the second leg pair (third body ring). The antennae consist of seven to eight segments and four sensory cones. Another feature typical for millipedes, but not occurring in all species is the Tömösvary organ, which is situated at the base of the antennae. The function of this organ is unknown and most research dates back to the beginning of the 20th century, where it was suggested that it may be a humidity receptor (HENNINGS, 1906),a pressure receptor (PFLUGFELDER, 1933) or a sound receptor (MESKE, 1960). The mouth parts consist of a pair of mandibles and a gnathochilarium.

In Helminthomorpha, the largest taxon of millipedes, males possess gonopods, which are modified legs on body ring 7. These help in sperm transfer and their morphology is a useful feature in identifying species. In Pentazonia, the other major clade, the males possess modified legs, called telopodes, at the caudal body end, with the same function.

2.3. Ecology a. General

The ecology of millipedes was reviewed in detail by HOPKINS & READ (1992). Millipedes occur on all continents, with the exception of Antarctica. They form an ecologically and morphologically diverse taxon that is very abundant in forest soils (DAVID, 2009; GOLOVATCH &

KIME, 2009). Most species live on the soil or in litter, but some species occur in the vegetation or under tree bark (BERG et al., 2008). Some species are adapted to more extreme habitats, such as the sea shore, deserts, tundra and floodplains (GOLOVATCH & KIME, 2009) Furthermore, some species are burrowers that live almost exclusively underground. HOPKINS & READ (1992) consider five life forms, based on their morphology and behavior:

 ‘Bulldozers’ or ‘rammers’, in our study area mainly represented by the snake-like Julida  ‘Wedge types’ or ‘litter-splitters’, short, stout species, such as Polydesmida  ‘Borers’, here represented by Chordeumatida and Polyzoniida 18

 ‘Rollers’, capable of rolling themselves up when threatened, such as Glomerida  ‘Bark dwellers’ with a tiny, soft body, e.g. Polyxenida.

These morphotypes can be classified in five functional groups (GOLOVATCH & KIME, 2009): stratobionts that live in litter and the uppermost soil, which are the most diverse and abundant group among millipedes and on which this study focuses; pedobionts, usually small and slender species that live in the mineral soil; troglobionts that live in caves; under-bark xylobionts, living under the bark of trees and often possessing a flattened or miniature body and epiphytobionts, which are often small and live on trees and other plants.

Being small, slow animals, millipedes are prone to predation and infestation by parasites. Mites are often considered as parasites, but this relationship might be commensal (SIERWALD & BOND,

2007). Furthermore, millipedes are often infested with parasitic nematodes (BLOWER, 1985). In addition, dipterans belonging to Phaeomyiidae and Sciomyzidae are parasitoids on millipedes

(SIERWALD & BOND, 2007). Predation by beetles (BAKER, 1985; HERBERT, 2000; SNIDER, 1984) , slugs (HERBERT, 2000), spiders (BAKER, 1985) and small mammals (BAKER, 1985) is common.

Because of this predation pressure, millipedes have evolved chemical defence mechanisms. Most species have repugnatorial glands and produce toxic repellents (HOPKINS & READ, 1992). The basal Polyxenida lack these glands, but have detachable bristles that can entangle predators

(EISNER et al., 2006).

Most millipedes are detritivores and they play an important role in fragmentation of dead plant material, thereby stimulating microbial activity. As this has proved to be an important ecosystem function, the role of millipedes as decomposers will be thoroughly discussed in the next paragraph. In addition to dead organic matter, some species feed on living plants and may be considered as pest species. Because of the intensive agriculture in the study area, the role of millipedes as pest species may be of importance. Therefore, it will be discussed in the following paragraphs.

b. Community composition and habitat requirements

Despite the fact that millipedes form a very diverse taxon that is often dominant in forest ecosystems, local communities usually consist of no more than two dozen species, even in the

19 tropics (GOLOVATCH, 1997). This is in accordance with the fact that most species have very small distribution ranges (GOLOVATCH & KIME, 2009), with many species being restricted to a single cave, forest patch or mountain. This high degree of endemism is especially striking in the tropics, while in the boreal region, some species, such as Polydesmus inconstans and Polyzonium germanicum have a large, sometimes even pan-European distribution (KIME & ENGHOFF, 2011).

In temperate Europe, most millipede species are associated with deciduous forests (KIME &

GOLOVATCH, 2000; KIME, 1992). Several studies are known to link millipede community composition to abiotic and biotic factors, such as tree species (GAVA, 2004; MEYER & SINGER,

1997; STAŠIOV et al., 2012; WYTWER et al., 2009), chemical soil properties (KIME, 1992;

STASIOV, 2005; STAŠIOV et al., 2012; TAJOVSKY & WYTWER, 2009; TOPP et al., 2006), soil humidity (TAJOVSKY & WYTWER, 2009; WYTWER et al., 2009), humus type (BRANQUART et al.,

1995; DAVID et al., 1993; WYTWER et al., 2009), humus content (STASIOV, 2002), slope gradient and direction (TOPP et al., 2006) climate (WYTWER et al., 2009), forest age (WYTWER et al.,

2009) and large scale zoogeographical patterns (ENGHOFF, 1993; GOLOVATCH, 1992; WYTWER et al., 2009). Most of these studies were carried out in the Eastern European plain and in primeval forest in this region, usually on a small scale. In Western Europe, where human impact on the landscape is a lot more significant, no extensive large-scale studies on millipede communities have been published. Additionally, the influence of the surrounding landscape has generally been ignored in these studies.

At the local scale, forests existing of trees that produce quickly decomposing litter, such as Ulmus spp., Carpinus betulus and Alnus glutinosa support richer communities than forests with species such as Betula pubescens and Larix decidua (STAŠIOV et al., 2012). Chemical characteristics of the soil also prove to be important, with the forest age (STASIOV, 2009), soil pH and Ca-content being the most significant factors in determining the community structure (SMITH et al., 2006; STAŠIOV et al., 2012; TAJOVSKY & WYTWER, 2009) and N-content in litter the most important factor for species richness, with a negative relationship between nitrogen content and species richness (STASIOV, 2009).

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c. Ecosystem function: millipedes as litter decomposers

Millipedes are very abundant and can make up more than 30% of the soil fauna biomass

(BATTIGELLI et al., 1994) and more than half of the invertebrate decomposer biomass in mature forest (MORÓN-RIOS & HUERTA-LWANGA, 2006). Generally they have a greater biomass in temperate ecosystems than in the tropics and are often abundant in degraded environments, such as cropped surfaces or cleared forests (CRAWFORD, 1992) They play an important role in the soil food web as decomposers of plant organic matter (e.g. ANDERSON 1987; CRAWFORD 1992; RUAN et al. 2005) and in the maintenance of soil structure (ALTIERI, 1999).

Millipedes are important decomposers. Research on the American species Harpaphe haydeniana showed that they can consume up to 20% of their own biomass, or 36% of the total litter production in a conifer forest (CÁRCAMO et al., 2000). They aid in litter decomposition by fragmenting leaf litter (KHEIRALLAH, 1990), thus increasing the surface area and by facilitating establishment of soil bacteria on litter after having passed through the gut by reducing nutrient deficiency of the microflora (MARAUN & SCHEU, 1996; TAJOVSKY et al., 1991). The chemical transformations of ingested litter are largely due to several microbial symbionts (RAMANATHAN

& ALAGESAN, 2012). However, because of overgrazing on microbial fauna, high densities of macrofauna such as millipedes can decrease mineralization rates (HANLON & ANDERSON, 1980). Additionally, the decomposition rate may differ depending on litter quality, which, in its turn depends on the season (DAVID & GILLON, 2002; MARAUN & SCHEU, 1996; VAN WENSEM et al., 1993).

In addition, millipedes are significant producers of methane gas, in contrast to other soil invertebrates, such as slugs, earthworms and woodlice, thereby significantly influencing meso- and microenvironments (ŠUSTR & ŠIMEK, 2009).

d. Millipedes and ecological restoration

Millipedes are often mentioned together with isopods and earthworms in the context of ecosystem restoration. Their impact on soil structure and chemistry can help to accelerate restoration of degraded habitats, given that the right species are used and the minimum habitat requirements are met (SNYDER & HENDRIX, 2008). Despite this, the ecology of most species is not known well enough to put them effectively to use. Earthworms, on the other hand, already

21 play an important role in ecosystem restoration by inoculating them on degraded terrains (BUTT,

1999). However, the fact that millipedes are often very tolerant to heavy metals (GRELLE et al.,

2000; READ et al., 1998) might be interesting for ecological restoration of polluted sites.

Additionally, millipedes can, together with isopods and earthworms, be good indicator species to determine the degradation of habitats, in the same way as aquatic macroinvertebrates are often used as indicator species for water quality (SNYDER & HENDRIX, 2008). DUNGER et al. (2001) mention a succession in arthropods, including millipedes from pioneer species to forest species on an afforested abandoned mine site. Similar studies using arthropods, including millipedes, have been performed in tropical rainforest (NAKAMURA et al., 2003), colliery spoil heaps

(TAJOVSKÝ, 2001) and coastal dunes (REDI et al. , 2005; VAN AARDE et al., 1996). These studies indicate a change in community composition from pioneers to forest species and an increase in density, diversity and species richness with time. Certain species could be used as indicator for habitat connectivity, other habitat requirements, or ecosystem processes (e.g. presence of millipede species indicates that sufficient leaf litter is produced). The problem remains, however, that the ecology of most species is insufficiently known.

e. Detrimental effects: pest species and invasions

In addition to these beneficial effects, millipedes have long been known as economically important pest species on several agricultural crops (BLOWER, 1985). Several species, such as Blaniulus guttulatus, Cylindroiulus caeruleocinctus and Brachydesmus superus are known to affect crops such as potatoes, maize and carrots (ALLEN & FILOTAS, 2008; BRUNKE et al., 2012). C. caeruleocinctus is a species that is extremely abundant on agricultural fields, and often makes up more than 95% of all the individuals. The species can infest 15-24% of all potatoes on a field

(BRUNKE et al., 2012). In greenhouses, millipedes have been observed to feed on vegetables

(MESSELINK & BLOEMHARD, 2007). Despite their reputation as pest species, in many cases presence of millipedes is often far less detrimental than other co-occurring taxa such as wireworms (Elateridae) and probably they only eat already infested crops (BRUNKE et al., 2012).

Despite being not very mobile and not being able to disperse over large distances on their own, several species are known as invasive, mostly in human settlements, but also in disturbed and in natural habitat (SIERWALD & BOND, 2007). More than half of all the British species are also

22 occurring in North-America due to human transport (KIME, 1990). Some of these, such as the previously mentioned C. caeruleocintus can inflict damage to agricultural crops in areas where they are introduced (BLOWER, 1985; BRUNKE et al., 2012). In the study area, most invasive species occur in greenhouses and human settlements (ANDERSSON et al., 2005; BLOWER, 1985) and –usually rare– exotic species from other parts of Europe (ENGHOFF, 2010; KIME, 2004), but invasions may occur without being noticed because of the lack of monitoring and difficulties with identifying species (SNYDER et al., 2006).

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Goals

The goal of this study is to assess the factors that influence millipede community composition in small deciduous forest fragments. Not only the intrinsic properties of the forests, such as occurring tree species, humus type and soil pH, but also factors acting on a larger scale, such as climate and latitudinal position will be taken into account. In addition, the intensity of land use will be assessed at the landscape-level and will be used as an explanatory factor. Not only the species composition, but also the diversity and abundance of millipedes will be examined. This research is unique in that it combines these factors in the context of small forest fragments in intensively used agricultural landscapes. This makes that the results may have implications for the management of these small forest fragments because of the important role of millipedes in litter decomposition, but also because of their potential detrimental effect as pest species.

It is hypothesized that the species composition will differ between the landscapes. As a limited number of species are very abundant on agricultural fields, they may be present in larger numbers in more intensively used landscape. In general, the diversity and abundance of milliepdes is expected to be lower in intensively managed agricultural landscapes (DAUBER et al. 2005;

CALLAHAM et al. 2006; ATTWOOD et al. 2008; RAHMAN et al. 2011). Furthermore, the community might differ between old and recent forests, with forest species with a lower mobility as typical species for old forests.

The results of this study, combined with concurring investigations on the communities of other detritivores, such as woodlice and harvestmen and predators such as carabid beetles, centipedes and spiders should provide much information on the effect of the mentioned factors on biodiversity and ecosystem function in small forest fragments. This might have implications for forest management in densely populated regions with highly fragmented forest remnants.

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Material & methods

1. Sampling sites This study makes use of the smallFOREST network, which studies small deciduous forest fragments in eight regions in Western Europe along a latitudinal gradient. Here, samples from five regions were analyzed: Central and Southern Sweden, Western Germany, Belgium and Northern France. (figure 3) All regions contain typical Western European landscapes with more or less intensive agriculture and small forest remnants between the fields.

In each region, two 5x5 km² windows were selected: one intensively used landscape (Open field, O) and one more or less extensively used landscape (Bocage, B). In each landscape, 16 deciduous forest patches

Figure 3: Sampling localities analyzed in this were selected according to age and size: four young, study small patches, four old, small patches, four young, large patches and four old, large patches (figure 5). Only in the Bocage-window of South- Sweden, twelve forest patches were old growth forest and four were recent forest.

2. Sampling In each region, two samplings were carried out: one in the spring and one in the summer of 2013 (table 2). To synchronize the trapping dates in the different regions along the latitudinal gradient, the amount of growing degree hours since January 1 was used. These Figure 4: Schematic view of the arrangement of the pitfall traps in were calculated from regional weather station a forest patch data, at less than 30km from the sampling localities (DE FRENNE et al., 2011).

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Figure 5: Landscapes where the samplings took place. Each window measures 5x5 km. Forest fragments that were sampled are encircled in red (white when on red background). Two forests in Central Sweden where the data was omitted are encircled in light grey.

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Figure 5 (continued) The samplings were carried out with an arrangement of two pitfall traps (ø 10 cm, depth 15 cm), separated by a cardboard fence (100x30 cm) with one trap oriented at the centre of the forest and the other at the edge, to measure potential fluxes of arthropods from the forests to the surrounding fields and vice versa. This setup is demonstrated in figure 4. The design is primarily used to measure a flux in spiders, which seems to correlate with predation rates (MENALLED et al., 1999) but it might also uncover a flux in millipedes, some of which are known pest species on agricultural crops. Four of these arrangements were placed in each forest patch: two at the centre of the forest and two at the southern edge, totaling eight pitfall traps per forest per trapping period, with two replications per forest (figure 4). Sampling design at the forest level is shown in figure 3. Due to temporal restrictions, only the first replication of the first sampling period of Southern Sweden was analyzed.

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The traps were set out for two weeks and filled with ethylene glycol as conservation fluid and a drop of detergent to prevent trapped animals from floating on the surface. A small aluminum roof was put on the traps to shelter them from rain (figure 6). After collecting, the animals were stored in 70% ethanol, sorted out and identified up to the species level. As females of Figure 6: setup of the pitfall traps. Picture: Edwin Brosens. Julus scandinavius, Ophyiulus pilosus and Julus terrestris cannot be separated with certainty, females of these species were identified as ‘Julidae sp.’ and they were omitted from the statistical analyses on species richness and the ordinations.

Table 2: Exact sampling dates per region. Some sampling campaigns took more than one day, so multiple dates are given.

Region Window Start period 1 End period 1 Start period 2 End period 2 C-Sweden B 12/06/2013 26/06/2013 31/07/2013 14/08/2013 13/06/2013 27/06/2013 01/08/2013 15/08/2013 O 10/06/2013 24/06/2013 29/07/2013 12/08/2013 11/06/2013 25/06/2013 30/07/2013 13/08/2013 S -Sweden B 05/06/2013 18/06/2013 / / 06/06/2013 19/06/2013 O 07/06/2013 20/06/2013 / / 08/06/2013 22/06/2013 W -Germany B 23/05/2013 06/06/2013 08/07/2013 22/07/2013 24/05/2013 07/06/2013 09/07/2013 23/07/2013 27/05/2013 10/06/2013 O 28/05/2013 11/05/2013 10/07/2013 24/07/2013 29/05/2013 12/05/2013 12/07/2013 26/07/2013 Belgium B 13/05/2013 27/05/2013 01/07/2013 15/07/2013 14/05/2013 28/05/2013 O 15/05/2013 29/05/2013 02/07/2013 16/07/2013 16/05/2013 30/05/2013

In addition to the sampling of millipedes, relevant descriptors were collected. Here, percentage cover of the moss, herb, shrub and tree layer, dominant herb, shrub and tree species, percentage of cover by coarse woody debris and surrounding land use were identified.

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3. Data analysis

3.1. General effects on diversity, species richness and density To assess whether the number of species differed significantly between the regions, a generalized linear model in R 2.15.2 was fitted only including the five regions in this study as explanatory variables. As the model indicated that the region was significant, pairwise t-tests between the regions were conducted to assess the differences. As the difference in millipede diversity between the bocage and the open field landscape was highly variable between the regions, the same analysis was also carried out for each window individually to provide extra insights in the patterns determining species composition.

The effect of patch size, forest age and surrounding land use on diversity, species richness and abundance were investigated. In the analysis, the data were merged per forest, with the data from the two periods, the two replications and the centre and edge combined to one sample by adding the specimens from all traps in the forest. The Shannon diversity indices were calculated for every forest as a measure of alpha-diversity.

Forest size, age and landscape intensity were included in a generalized linear mixed model with region as a random factor. For every model, first a full model, containing the mentioned explanatory variables and their interaction factors were fitted. The least significant variable was dropped every time until a model with only significant variables (p<0.05) was left. No significant explanatory variables were found to influence the Shannon diversity, which is therefore omitted from the results of this study.

The only factor that proved significant in determining the number of species was the intensity of surrounding land use. The general millipede abundance significantly depended on forest age, size and intensity with a significant interaction term between forest age and intensity. The final models used are displayed in table 3.

Table 3: Statistical models for calculating species number, abundance and diversity

Response variable Explanatory variable(s) Species number Intensity Density Intensity+Age+Size+Intensity:Age Shannon-diversity -

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The model described above used the forest patches individually as distinct data points, instead of the centre and the edge of each forest to prevent pseudoreplications within the forests. However, to test a difference in abundance between the centre and the edge of the forests, the same model was used with location (centre vs. edge) and the interaction between location and size as explanatory variables.

3.2. Unconstrained ordination: DCA Ordination techniques were used as explorative explanatory methods. CANOCO 4.5 was used to perform the analysis (TER BRAAK & ŠMILAUER, 2003). A detrended correspondence analysis

(DCA) was performed (HILL & GAUCH, 1980). This technique extracts the largest gradient in species abundance data without regard to environmental variables (unconstrained ordination)

(LEPŠ & ŠMILAUER, 2003). The environmental variables were added to the plot as passive variables to score their importance in determining the community composition.

This analysis gives extra information in addition to constrained ordination techniques. Additionally, this method helps to recognize the abundance pattern of species as linear or unimodal and thus aids in choosing the type of constrained ordination technique (RDA or CCA)

(LEPŠ & ŠMILAUER, 2003; TER BRAAK & ŠMILAUER, 2003). First, a DCA was carried out for all regions together, followed by a DCA for each region individually.

In every forest, the centre and the edge were considered as two distinct sampling points, but replications 1 and 2 and sampling period 1 and 2 were grouped together as one sample. Species were log transformed (y=log(N+1)) to prevent bias. Detrending by segments was used as detrending method. Furthermore, downweighting of rare species was enabled, to prevent bias caused by accidental occurrences of rare species.

First, a DCA with samples from all regions was performed. The dataset used for the DCA combining the forests from all regions underwent an y=log(N+1) transformation. The location (centre-edge), size (small-large), age (old-recent), intensity (bocage-open field) and region (C- Sweden – S-Sweden – W-Germany – Belgium – N-France) were used as passive nominal variables in this analysis together with tree, shrub, herb and moss cover, amount of dead wood and dominant tree species (table 4).

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Then, a DCA was performed for each region individually with the same variables. In Central- Sweden, the samples from two forests in the open field landscape were omitted from the data as the pitfall traps were filled with ants and the millipedes were not identifiable anymore. The effect of location inside the forest (centre vs. edge) and the effect of patch size are omitted from the plots as these showed no clear signal. The percentage cover of the tree, shrub, herb and moss layer and the percentage of coarse woody debris in relation to the total standing biomass were used as continuous variables and forest age (old vs. recent) and land use intensity (bocage vs. open field) were categorical variables. In the dataset from North France, no data about tree, shrub, herb, moss and dead wood cover or about dominant tree species were available. For these explanatory variables, the correlation with the ordination axes and the significance of this correlation was calculated.

Table 4: Environmental variables used in the DCA

Variable Meaning Type Categories Tree% Percentage cover of tree layer Continuous Shrub% Percentage cover of shrub layer Continuous Herb% Percentage cover of herb layer Continuous Moss% Percentage cover of moss layer Continuous Dead wood% Percentage dead wood i.r.t. total standing biomass Continuous Location Location in the forest Categorical Centre, Edge Tree species Dominant tree species Categorical See appendix 1 Size1 Size of the forest Categorical Small, Large Intensity1 Intensity of surrounding land use Categorical Bocage, Open field Age1 Forest age Categorical Old, Recent Region1,2 Region where the sampling took place Categorical C-Sweden, S-Sweden, W-Germany, Belgium, N-France 1Variables used in the combined analysis 2Variables used only in the combined analysis 3.3. Constrained ordination: RDA and CCA As much of the data, especially the data on physical and chemical characteristics of the soil was not available, and therefore a large amount of the variation remains unexplained, a constrained ordination was not performed.

3.4. Indicator species analysis The data was subjected to an indicator species analysis to determine which species are typical for certain conditions at the forest level. This analysis was performed with the indicspecies-package

31 in R 2.15.2. The indicator value of each species was determined. Two variables were used for this: A, being the positive predictive value that gives the probability that the site where an individual of species S was found belongs to target site group G, or A=P(G|S) and B, being the sensitivity, or proportion of target sites where the species was found, B=P(S|G) (DE CÁCERES et al., 2012; DUFRENE & LEGENDRE, 1997) .The indicator value is then calculated as:

(De Cáceres, Legendre, & Moretti, 2010). To test the statistical significance, 999 permutations were run. Here the data was randomly reordened 999 times and tested against a test statistic.

The analysis was carried out for bocage and open field landscape (table 15) and for recent and old growth forest (table 16). Because the species composition is very different between the regions due to zoogeographical factors, the analysis was always carried out per region.

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Results

1. Difference between regions A large difference in abundance was found between the different regions and the windows within each region. In total, 10,667 individuals were identified. The total number of individuals per region and per window is displayed in table 5. In total, 35 species were found. These species, the abbreviations used for them and occurrence in each region is shown in table 12.

Table 5: Number of individuals and species per region and per window Bocage Open field Total individuals Total species C-Sweden 421 3173 3594 18 S-Sweden* 622 670 1292 16 W-Germany 508 602 1110 15 Belgium 1123 1447 2570 22 N-France 602 1341 2101 21 *In southern-Sweden, only the first replication of the first sampling period was carried out Despite the number of species found in all regions being more or less similar, ranging from 15 in West Germany to 22 in Belgium, the number of individuals was highly variable between the regions. Additionally, within the regions, the number of individuals often differs greatly between the windows.

The average number of species per forest was also calculated per region. In Belgium, on average 8.5±1.06 species were found per forest patch, which is significantly more than all the other regions, where the average number varied from 5.8 species in West Germany to 6.7 in South Sweden (table 6).

Table 6: Differences in species number between regions. A generalized linear mixed model with log-link function was used.

Estimate Std. Error z value p-value Avg species Intercept (Belgium) 2.140 0.061 35.295 <0.001 8.50 ± 1.06 Central Sweden -0.303 0.095 -3.214 0.001 6.27 ± 1.09 South Sweden -0.233 0.093 -2.509 0.012 6.73 ± 1.10 West Germany -0.380 0.095 -3.994 <0.001 5.81 ± 1.09 North France -0.366 0.098 -3.747 <0.001 5.90 ± 1.10

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Table 7: Pairwise t-tests between regions to test significant differences in species number between the forest patches of different regions. The p-values are given in the table.

Central Sweden South Sweden West Germany Belgium South Sweden 0.430 West Germany 0.436 0.115 Belgium <0.001 0.003 <0.001 North France 0.535 0.162 0.886 <0.001

As these differences might be caused by one of the two windows in each region (bocage or open field), the same analysis was carried out for open field and bocage landscapes separately.

Table 8: differences in species numbers in forest patches in the bocage landscape of each region. A generalized linear mixed model with log-link function was used here.

Estimate Std. Error z value p-value Avg species±StDev Belgium (Intercept) 2.147 0.085 25.135 <0.001 8.56 ± 1.09 C-Sweden -0.550 0.141 -3.897 <0.001 4.94 ± 1.15 S-Sweden -0.285 0.130 -2.187 0.029 6.44 ± 1.14 W-Germany -0.409 0.135 -3.025 0.002 5.69 ± 1.14 N-France -0.454 0.137 -3.312 <0.001 5.44 ± 1.15

Table 9: Pairwise t-tests between the bocage landscapes of the different regions to test significance of differences in species number between them in individual forests. The p-values are given in the table.

Cental Sweden South Sweden West Germany Belgium South Sweden 0.056 West Germany 0.335 0.335 Belgium <0.001 0.007 <0.001 North France 0.520 0.200 0.747 <0.001 The Belgian bocage landscape is by far the most species rich, with on average 8.56 species per forest (table 8). The average species number is significantly greater than in any other region. Especially the bocage landscape in Central Sweden is very species-poor, with on average 4.94 species per forest, though this is not significantly less than any other region except from Belgium (table 9).

Table 10: Differences in species numbers in forest patches of the open field landscape of each region

Estimate Std Error z value p value Avg species±StDev Belgium (Intercept) 2.132 0.086 24.780 <0.001 8.43 ± 1.09 C-Sweden -0.080 0.129 -0.624 0.532 7.78 ± 1.14 S-Sweden -0.177 0.132 -1.335 0.182 7.06 ± 1.14 W-Germany -0.351 0.134 -2.624 0.009 5.94 ± 1.14 N-France -0.267 0.139 -1.920 0.055 6.46 ± 1.15

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In the open field landscape the Belgian forests are still the most species-rich (table 10). However, the species-richness is only significantly higher than North France and West Germany (table 11). The differences in the open field landscape are therefore less pronounced than in the bocage landscape.

Table 11: Pairwise t-tests between the open field landscapes of the different regions to test significance of differences in species number between them. The p-values are given in the table.

Central Sweden South Sweden West Germany Belgium South Sweden 0.403 West Germany 0.028 0.173 Belgium 0.431 0.101 0.003 North France 0.131 0.483 0.534 0.021 Table 12 shows some interesting patterns. Some species, such as Julus scandinavius are very common in every region, while other species, such as Melogona voigtii were only found in one region in limited numbers. Allajulus nitudus, Tachypodoiulus niger and Cylindroiulus punctatus are more abundant in the southern regions, while Polydesmus denticulatus and P. inconstans were trapped in higher numbers in the northern study sites.

In Central-Sweden a big difference between the bocage and the open field exists. In the open field landscape, far more individuals were found. More than half of the 3173 specimens belong to Cylindroiulus caeruleocinctus. Furthermore, P. denticulatus, P. inconstans and Ommatoiulus sabulosus were very abundant. It is important to notice that most individuals (2900) were collected during the first sampling period, while the second yielded far less specimens (273).

The bocage-landscape gives an entirely different result. The abundance is almost ten times lower in this window. P. denticulatus is also very common here. Furthermore, Julus scandinavius was relatively abundant in these samples. The total number of individuals in this window was the lowest of all windows in this study. As in the open field-window, the number of individuals was far higher during the first sampling period (356 vs. 65).

In South Sweden, the species composition between the open field and the bocage window was clearly different. In the bocage landscape, Unciger foetidus and Polyzonium germanicum were very abundant, while these were a lot scarcer in the open field landscape, where Cylindroiulus caeruleocinctus was the most common species, followed by Polydesmus inconstans. Ophyiulus pilosus and Julus scandinavius were quite common in both windows. The number of animals found in the open field was slightly higher than in the bocage landscape: 670 in open field and

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622 in bocage. Because of time restrictions, only the first replication of the first sampling period was analyzed in this region.

Table 12: Species encountered in this study. Green: Abundant, species occurs in more than 40% of all forests, more than 60 individuals; Yellow: Common, species occurs in more than 25% of all forests, more than 30 individuals; Orange: Rare, species occurs in more than one forest, more than 15 individuals; Red: Very rare, species occurs in only one forest, or less than 15 individuals; White: species does not occur.

Species C. Sweden S.Sweden* W. Germany Belgium N. France Abbreviation Allajulus nitidus Alla niti Blaniulus guttatus Blan gutt Brachydesmus superus Brac supe Brachyiulus pusillus Brac pusi Choneiulus palmatus Chon palm Chordeuma sylvestre Chor sylv Craspedosoma rawlinsi Cras rawl Cylindroiulus caeruleocintus Cyli caer Cylindroiulus latestriatus Cyli late Cylindroiulus punctatus Cyli punc Enantiulus nanus Enan nanu Glomeris intermedia Glom inte Glomeris marginata Glom marg Julus scandinavius Julu scan Julus terrestris Julu terr Leptoiulus belgicus Lept belg Leptoiulus kervillei Lept kerv Melogona gallica Melo gall Melogona voigtii Melo voig Mycogona sp Myco sp Nemasoma varicorne Nema vari Ommatoiulus sabulosus Omma sabu Ophiodesmus albonanus Ophi albo Ophyiulus pilosus Ophy pilo Orthochordeumella pallida Orth pall Polydesmus angustus Poly angu Polydesmus complanatus Poly comp Polydesmus coriaceus Poly cori Polydesmus denticulatus Poly dent Polydesmus inconstans Poly inco Polyzonium germanicum Poly germ Propolydesmus testaceus Prop test Proteroiulus fuscus Prot fusc Tachypodoiulus niger Tach nige Unciger foetidus Unci foet Total number of species 18 16 15 22 21 *Only the first replicate of the first sampling period has been analyzed in South Sweden

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Figure 7: Pie chart with the abundance of all species in per window and per region. Species that represented less than 5% of the total number of specimens are grouped under 'other'.

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Figure 7 (continued): Pie chart with the abundance of all species per window and per region. Species that represented less than 5% of the total number of specimens are grouped under 'other'.

The German samples contained 508 specimens in the bocage window and 602 in the open field window. Glomeris marginata, Julus scandinavius and Polydesmus angustus were very abundant in both landscapes. Ommatoiulus sabulosus was especially abundant in the open field window, but also occurred in bocage landscape. In this region the first sampling yielded far more specimens than the second one. This was especially clear in the open field landscape (first period: 469; second period: 133), but also in the bocage window (first period: 313; second period: 195).

In Belgium, the most common species was by far Tachyopodoiulus niger. Leptoiulus kervillei was also very abundant in both landscapes. In the open field landscape, Polydesmus coriaceus and Cylindroiulus caeruleocinctus were much more abundant than in the bocage landscape. In the bocage window, Glomeris intermedia and G. marginata were very abundant. In the open

38 field, more individuals were found than in the bocage landscape, with respectively 1447 and 1223 specimens. The first period yielded more specimens than the second (open field: 801 vs. 646; bocage: 602 vs. 521).

In North-France, the most common species is Tachypodoiulus niger. In the bocage window, more than two thirds of all specimens belong to this species. Leptoiulus kervillei is also common in both landscapes. Except for these species, the bocage landscape does not contain many abundant species. In the open field window, Cylindroiulus caeruleocinctus is very abundant, along with Polydesmus angustus, Allajulus nitidus, Melogona gallica and Propolydesmus testaceus. Glomeris intermedia, occurs in both windows in low abundance.

2. Diversity and density differences between forests The number of species and the Shannon diversity index were calculated for each forest. Forest size and age were no significant predictors for both variables (results not shown). The intensity of the land use in the surrounding landscapes, however, was significant for the number of species (p=0.0297) with on average 6.2 species per forest in bocage landscape and 7.1 species in open field landscape. In table 13, bocage landcapes are considered as the baseline (intercept). The factor intensity accounts for the difference with open field landscapes.

Table 13: Statistical model with variables explaining number of millipede species per forest. Only the significant variables are reported.

Estimate Std. Error z-value p-value Avg species±StDev Intercept (bocage) 1.819 0.071 25.539 <0.0001 6.17 ± 1.07 Intensity (open field) 0.136 0.063 2.174 0.0297 7.06 ± 1.06

The number of individuals caught in each forest was correlated with forest age, size, intensity of land use and an interaction between the intensity of land use and the forest age. The results are given in table 14. The average number of individuals in large, recent forest situated in bocage landscape is 31.82. On average, 1.39 times more individuals were found in old than in recent forest, 1.19 times more in small than in large forest and 1.49 times more individuals in an open field than in a bocage landscape. Furthermore, a significant interaction term between land use intensity and forest age was observed.

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Table 14: Generalized linear mixed model with log-link function for variables explaining millipede density at the forest level..

Estimate Std. Error z-value p-value Intercept 3.788 0.204 18.609 <0.001 Intensity (Open field) 0.401 0.028 14.140 <0.001 Age (Recent) -0.332 0.034 -9.769 <0.001 Surface (Small) 0.177 0.020 8.684 <0.001 Intensity:Age 0.478 0.042 11.283 <0.001 Despite on average having more individuals on old forest, when the interaction with land use intensity is considered, other patterns are visible. In the bocage landscape, millipede density is higher in old forests, but in open field landscape, the opposite is true: recent forests have more millipedes than old forests. This interaction between land use intensity and forest age is visualized in figure 8.

Figure 8: Boxplots of the interaction term between intensity and forest age. In open field landscape, recent forests have a higher density of millipedes, while in bocage landscape, old forests have a higher millipede density In addition to these models, another model with the centre and the edge of each forest as distinct data points, instead of each forest as a single data point was used (table 15). This approach was performed to test the difference in millipede activity density between the centre and edge of the forests. To prevent statistical bias due to pseudoreplicates, only the effects of location and location:size should be considered. Other effects are only left there to show their relative importance compared to location.

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Table 15: Millipede activity density in forest fragments: differences between centre and edge. As edge and centre are pseudoreplicates, only the absolute values of the variables location and size:location should be considered. The others are only there to compare their influence with location in the forest.

Estimate Std. Error z value p-value Intercept 2.866 0.201 14.232 <0.001 Intensity (Open field) 0.669 0.037 17.907 <0.001 Age (Recent) -0.340 0.043 -7.933 <0.001 Size (Small) 0.087 0.030 2.926 0.003 Location (Edge) 0.282 0.040 4.07 <0.001 Size:Location 0.162 0.040 4.069 <0.001 Age:Location -0.136 0.040 -3.328 <0.001 Intensity:Location -0.096 0.043 -2.223 0.026 Intensity:Age 0.490 0.043 11.287 <0.001 Samples from the forest edge contain significantly more individuals (factor x1.33). Furthermore, a synergistic interaction between size and location was significant. In large forests, the difference between edge and centre is more pronounced, however, figure 9 shows that this interaction is only of limited importance. The effect of location in the forest is less than the effect of age or intensity, but was larger than the effect of patch size.

Figure 9: Interaction between location in the forest and forest size in relation to the number of individuals per sample. In old forests, the difference between millipede density in the edge and the centre of the forest is more pronounced than in recent forests. The interaction between forest age and location is displayed in figure 10.

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Figure 10: Interaction between location in the forest and forest size in relation to the number of individuals per sample. A third interaction was visible between land use intensity and location in the forest. In open field landscapes, the number of individuals in the forest edge in relation to the number of individuals in the centre is relatively lower than in the bocage landscapes (figure 11). This interaction was, however, only slightly significant (p=0.026).

Figure 11: Interaction between land use intensity and location in the forest in relation to the number of individuals per sample.

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3. Detrended Component Analysis (DCA)

Figure 12: DCA with all sampled forests in Central sweden (green), South Sweden (blue), West Germany (yellow), Belgium (red) and North France (brown). Samples from bocage landscapes are indicated by triangles, samples from open field landscapes are circles. First, a DCA of all regions was performed (figure 12). This ordination showed a clear separation between the regions. The first ordination axis, which has a gradient length of 5.636 and explains 14.9% of all variance, seems to coincide with a north-south gradient. Belgium and North France have very similar faunas, as do Central and South Sweden. The second axis does not give a clear separation between the regions, but for each region individually, except for Belgium and West Germany separates the forests from bocage and open field landscape.

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Figure 13: DCA's of the different regions: a. Central-Sweden, b. South Sweden, c. West Germany, d. North France, e. Belgium 44

Table 16: Length of gradient of the first two ordination axes of the different regions together with the cumulative variance and the variance explained by the studied environmental variables

Region Axis 1 Axis 2 Central Sweden Length of gradient 2.57 3.056 Cumul. % variance explained 20.1 37.5 Cumul % species-environment relation 30.4 46.3 South Sweden Length of gradient 2.83 4.34 Cumul. % variance explained 18.4 33.4 Cumul % species-environment relation 29.4 50.2 West Germany Length of gradient 3.11 2.25 Cumul. % variance explained 19.4 35.5 Cumul % species-environment relation 22.4 41.6 Belgium Length of gradient 3.08 2.34 Cumul. % variance explained 15.8 26.1 Cumul % species-environment relation 23.4 35.0 North France Length of gradient 2.74 2.32 Cumul. % variance explained 19.0 30.0 Cumul % species-environment relation 21.9 72.8

Results of the ordination are given in figure 13 and tables 16-17. The gradient lengths of the first ordination axis in the different regions varies between 2.57 for Central Sweden and 3.11 for West Germany, which indicates a linear response of the species to changing environmental variables. The total variance explained in the ordination plots ranged from 30% in North France to 37.5% in Central Sweden, while the cumulative percentage variance caused by the environmental variables explained by the first two ordination axes ranged from 35% in Belgium to 50.2% in South Sweden. In France, this was even higher; being 72.8%, but this is caused by the fact that no information was available for most variables in this region.

The most striking contrast was between open field and bocage landscapes. The land use intensity was a significant factor in all regionas. In Central Sweden, samples from the open field landscape contained far more individuals than samples from the bocage landscape. Cylindroiulus caeruleocinctus, Polydesmus inconstans and Ommatoiulus sabulosus were the most typical species from the open field landscape, while Julus scandinavius was relatively more abundant in the bocage landscape. The first ordination axis seems to separate both windows very clearly In South Sweden the difference was also striking, with Polyzonium germanicum and Craspedosoma rawlinsi being very abundant in bocage landscape and C. caeruleocintus and P. inconstans being

45 dominant in the open field landscape. In West Germany, the difference between both landscapes was less clear, but Glomeris marginata and Tachypodoiulus niger occurred more in the bocage landscape, and O. sabulosus in the open field landscape. In Belgium, the separation between both landscapes was also not very clear, but still some species were typical of one landscape: C. caeruleocinctus and P. inconstans were relatively abundant in the open field landscape, while Glomeris intermedia and Polydesmus angustus preferred the bocage landscape. The second ordination axis separates both landscapes. In North France, the fauna of the bocage landscape was relatively poor, both in density and species number, while in the open field landscape, some typical species, such as P. inconstans, C. caeruleocinctus and Melogona gallica reached high densities. Most information on the distinction between both landscapes seems to be contained on the second ordination axis in this region.

Table 17: Correlation between variables and first two ordination axes. Significant correlations (p<0.05) are indicated in bold.

C Sweden S Sweden W Germany Belgium North France Axis 1 Axis 2 Axis 1 Axis 2 Axis 1 Axis 2 Axis 1 Axis 2 Axis 1 Axis 2

Int. (Open field) 0.809 0.080 0.603 0.579 0.509 0.231 0.323 0.451 0.296 0.753 Age (Old) 0.066 0.320 -0.574 0.046 0.073 -0.534 -0.578 0.036 0.320 -0.309 Size (Large) 0.053 -0.206 -0.277 -0.27 0.037 -0.032 -0.128 -0.261 -0.050 -0.070 Loc. (Centre) 0.029 -0.137 0.034 -0.160 -0.089 0.062 -0.091 -0.035 0.064 0.003 Tree% -0.255 0.120 0.101 0.087 -0.025 -0.467 -0.360 -0.048 - - Shrub% 0.039 0.172 -0.287 -0.240 0.032 0.179 -0.228 0.099 - - Herb% -0.004 -0.212 0.564 0.012 0.144 0.177 0.471 -0.118 - - Moss% -0.284 0.071 0.202* -0.019* 0.056 0.046 0.228 -0.057 - - Dead wood% 0.021 0.309 -0.230 0.058 0.175 -0.092 0.249 0.351 - - Popucana ------0.487 -0.318 - - Querrobu 0.291 0.162 0.198 0.034 -0.306 -0.210 -0.347 0.412 - - Fagusylv - - -0.474 -0.150 -0.161 -0.411 -0.313 -0.113 - - * Only two forests in South Sweden did have a moss layer. Therefore this data is not displayed in the ordination plots and is not discussed any further The difference between old and recent forests is less clear in Central Sweden, where age was not a significant factor. In South Sweden, the difference is very clear in the bocage landscape, but is less pronounced in the open field landscape. The first ordination axis seems to separate forests by age. In West Germany, the most information about forest age is contained on the second ordination axis. Glomeris marginata is a very clear indicator of old forest in this region. In Belgium, Glomeris intermedia, G. marginata and Polydesmus angustus are typical for old forest, while C. caeruleocinctus is found more often in recent forests. The separation is very clear in this 46 region, with the most information about forest age on the first ordination axis. In North France, the difference between old and recent forest fragments was only clear in the open field landscape, as the bocage landscape was relatively species-poor.

The effects of size of the forest fragments (small vs. large) and location of the traps within the forest (centre vs. edge) were also tested, but no clear contrast between these values was observed, and therefore these are omitted from the plots. Size had a slightly significant effect in South Sweden and location in the forest was significant in North France, but the effect was minimal. Effect of dominant tree species was also not visible on most plots, except for South Sweden and Belgium. In South Sweden, the species assemblages in forests where beech (Fagus sylvatica) was dominant differed clearly from forests dominated by sessile oak (Quercus robur). In Belgium, forests with Canada poplar (Populus x canadensis) also had a distinct community composition.

The effects of dead wood and cover by tree, shrub, herb and moss layer were assessed for every region, except for North France. Brachydesmus superus seems to be associated with dead wood in Central Sweden, while Cylindroiulus punctatus shows some affinity with dead wood in West Germany and Belgium and Julus scandinavius in Belgium and South Sweden. The effect was, however, never very consistent between the regions. G. marginata was associated with forests with a dense tree cover in the regions where it occurred, being West Germany and Belgium, while C. caeruleocinctus and P. inconstans usually preferred more open habitat, which is especially clear in Central Sweden and Belgium. In West Germany and Central Sweden, O. sabulosus also prefers more open forest fragments with lower tree cover. The effects of shrub, herb and moss cover were far less consistent between the regions and were therefore difficult to interpret.

4. Indicator species analysis An indicator species is displayed for bocage and open field landscapes and for old and recent forest. In this table, the A-value indicates the proportion of sites where the indicator species was found, that effectively belong to the habitat for which the species is an indicator (e.g. if species X is an indicator for open field and forty specimens were found in open field landscape and ten in bocage landscape, the A-value is 0.80). The B-value indicates the proportion of forests belonging to the selected habitat, where the species occurs (e.g. if species X occurs in four of the 16 forest

47 fragments in open field landscape, the B-value is 0.25). The indicator value is then calculated as the product of the square roots of these values. Species that were found to be significant indicators (p<0.05) are indicated in bold in the table.

Table 18: Indicator species of bocage and open field windows per region. A: proportion of forests where the indicator species was found, that belong to the respective window. B: proportion of sites of the respective window that the species was found in. Ind. val.: indicator value of the species. Significant (p<0.05) indicator species are in bold. Species are sorted by decreasing indicator value.

Region Intensity Species A B Ind. val. p-value C-Sweden Bocage Cras rawl 1.00 0.13 0.35 0.467 Open field Cyli caer 1.00 1.00 1.00 0.001 Poly inco 0.99 0.93 0.96 0.001 Brac supe 0.95 0.86 0.90 0.001 Omma sabu 0.85 0.93 0.89 0.001 Ophy pilo 0.80 0.64 0.72 0.030 Prot fusc 0.76 0.43 0.57 0.137 Blan gutt 1.00 0.14 0.38 0.204 Nema vari 1.00 0.14 0.38 0.198 Cyli punc 0.63 0.21 0.37 0.510 Chon palm 0.77 0.14 0.33 0.459 Alla niti 1.00 0.07 0.27 0.482 Enan nanu 1.00 0.07 0.27 0.482 Julu terr 1.00 0.07 0.27 0.477 Unci foet 1.00 0.07 0.27 0.461 S-Sweden Bocage Poly germ 0.95 0.69 0.81 0.003 Open field Cyli caer 0.99 0.86 0.92 0.001 Alla niti 1.00 0.14 0.38 0.194 Brac pusi 1.00 0.07 0.27 0.448 Melo voig 1.00 0.07 0.27 0.488 W-Germany Bocage Tach nige 1.00 0.25 0.50 0.088 Open field Omma sabu 0.85 0.63 0.73 0.028 Cras rawl 1.00 0.25 0.50 0.100 Belgium Bocage Glom inte 1.00 0.31 0.56 0.046 Nema vari 1.00 0.19 0.43 0.220 Open field Poly inco 0.95 0.44 0.65 0.020 Cyli caer 1.00 0.25 0.50 0.108 N-France Bocage Poly inco 1.00 0.24 0.49 0.097 Ophi albo 1.00 0.12 0.34 0.515 Open field Cyli caer 0.96 0.76 0.86 0.002 Poly angu 0.93 0.71 0.81 0.011 Melo gall 0.93 0.65 0.78 0.002 Prop test 0.98 0.47 0.68 0.013 Poly dent 0.95 0.47 0.67 0.011 Ala niti 0.90 0.47 0.65 0.126 Brac pusi 1.00 0.29 0.54 0.042

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Blan gutt 1.00 0.12 0.34 0.456

The indicator species analysis for open field and bocage landscapes is displayed in table 18. In general, the open field landscape has several typical indicator species, in contrast to the bocage landscape, which was especially in Central Sweden and North France very species-poor. Cylindroiulus caeruleocinctus and Polydesmus inconstans are clear indicator species for open field landscape in most regions where they occur. These species are almost exclusively found in the open field landscape, where they are often very abundant. Only in North France, P. inconstans was found more often in the bocage landscape, but the species was generally rare in this region and there was no significant relation. Ommatoiulus sabulosus is also a very clear indicator of open field landscapes. Despite also occurring in bocage landscapes, the species is far more widespread in the open field landscape. Only two species were significant indicator species for bocage landscape: Polyzonium germanicum, which was only found in South Sweden and almost only occurred in the bocage landscape, where it was found in almost 70% of all forests and Glomeris intermedia in Belgium, which only occurred in the bocage landscape, but only in one third of all forests.

The results of the indicator species analysis for old and recent forest are given in table 19. In general, less indicator species were found than for the contrast between open field and bocage landscape. No clear indicator species were found for Central Sweden or North France. In South Sweden, Unciger foetidus and C. caeruleocinctus were indicator species for recent forest, with U. foetidus being found in all recent forests. In other regions, no indicator species for recent forests were found. Craspedosoma rawlinsi and P. germanicum were almost exclusively found in old forests in South Sweden, but P. germanicum was not very abundant here, and therefore is just not significant (p=0.064).

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Table 19: Indicator species of recent and old forest per region. A: proportion of sites where the indicator species was found, that belong to the respective age cateogry. B: proportion of sites of the respective forest age class that the species was found in. Ind. val.: indicator value of the species. Significant (p<0.05) indicator species are in bold

Region Age Species A B Ind. val. p-value Central Sweden Recent Ophy pilo 0.96 0.44 0.65 0.117 Blan gutt 1.00 0.13 0.35 0.477 Nema vari 1.00 0.13 0.35 0.467 Old Cras rawl 1.00 0.14 0.38 0.187 South Sweden Recent Unci foet 0.85 1.00 0.92 0.007 Cili caer 0.72 0.70 0.71 0.041 Brac supe 0.78 0.50 0.62 0.112 Brac pusi 1.00 0.10 0.32 0.339 Melo voig 1.00 0.10 0.32 0.315 Alla niti 0.86 0.10 0.30 0.759 Old Cras rawl 0.95 0.60 0.76 0.016 Poly germ 0.92 0.55 0.71 0.064 Prot fusc 0.82 0.30 0.50 0.293 West Germany Recent Prot fusc 0.91 0.25 0.48 0.287 Cras rawl 0.86 0.19 0.40 0.479 Old Glom marg 0.91 0.63 0.75 0.013 Belgium Recent Brac pusi 0.91 0.44 0.63 0.150 Nema vari 1.00 0.19 0.43 0.199 Cyli caer 0.91 0.19 0.43 0.200 Old Glom marg 0.99 0.50 0.70 0.016 Poly angu 0.81 0.50 0.64 0.048 Glom inte 0.99 0.25 0.50 0.182 North France Recent Brac pusi 1.0000 0.25 0.50 0.097 Old Poly dent 0.89 0.38 0.58 0.057 Blan gutt 1.00 0.13 0.35 0.503

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Discussion

The results indicate that land use intensity and forest age have an important effect on community composition. Here, first the differences between the regions will be discussed. Then, the other characteristics, which influence the species composition and abundance at a regional level, will be examined, beginning with the land use intensity and forest age, as these are the major factors influencing species composition. Other factors such as size, dominant tree species and cover of tree layer will also be discussed briefly. With these results, general guidelines for landscape management will be given.

1. Differences in species composition between regions The gradient length of the first axis of the DCA where Central Sweden, West Germany, Belgium and North France were analyzed together is 5.636 (figure 12). This indicates that the species have a unimodal response curve at this scale. This is to be expected as the environmental gradient is much larger here, and overlaps with most part of the realized niche of several species.

The lengths of the axes in DCA are expressed in standard deviation units of species turnover

(HILL & GAUCH, 1980). The gradient length of the axes can be considered as a measure of beta- diversity (LEPŠ & ŠMILAUER, 2003). For each region individually, the length of the longest gradient varied from 2.57 in Central Sweden to 3.11 in West Germany. These values indicate that the response curves of most species within these regions could best be approximated by a linear response model. The differences between the forests seem to be the largest in West Germany. In Central-Sweden, the contrast between the open field and the bocage landscape is very clear.

The community composition differs clearly between the regions. The plot in figure 12 shows a very clear separation between the four analyzed regions, where the first axis of ordination seems to correlate with the latitudinal position of the sampling region. This might in the first place be caused by differences in climatologic factors. Additionally, as most millipede species have very small ranges of occurrence (GOLOVATCH & KIME, 2009), the difference may be caused by zoogeographical factors, such as dispersal limitation. This makes the comparison of millipede communities in different regions very difficult (MEYER & SINGER, 1997).

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The number of species found in each forest varies between 5.8 in West Germany and 8.5 in Belgium (table 6). Only in Belgium a significantly higher number of species was found than in the other regions. The higher species richness in Belgian forests is especially obvious at the bocage landscape. This may be in line with the fact that more intensively used landscapes were generally more species-rich: of all regions in this study, the difference in intensity between the open field and the bocage landscape was the smallest in Belgium, which was the only region where the average species number per forest patch was higher in the bocage than in the open field landscape (tables 8-11). The high overall intensity of land use in Belgium probably explains the high species richness. Soil and humus characteristics may also influence species richness. Data on these characteristics might provide new insights in the near future, but are not available for now.

2. Factors influencing community composition and abundance at the regional level

2.1. Effect of land use intensity The influence of landscape variables on soil biota is poorly known (GOLOVATCH & KIME, 2009). However, in this study the intensity of land use in the surrounding landscape was found to have a great influence on the millipede community. This seems to contradict the vision of WITH & CRIST (1995) that taxa with lower dispersal ability are less influenced by landscape composition.

First of all, the activity density of millipedes was significantly higher in forests in the open field landscapes. However, the differences between the regions are very pronounced: in West Germany and South Sweden, almost no difference in density was observed, while in South Sweden and North France, far more individuals were caught in the open field landscape. The high density is caused by a few indicator species.

In Central Sweden, Cylindroiulus caeruleocinctus made up more than half of all individuals trapped in the open field landscape, while not occurring in the bocage landscape. In addition to this species, several others, most notably Polydesmus inconstans and Ommatoiulus sabulosus occur in high numbers and are mainly restricted to the open field habitat. These few species probably explain the difference in activity density between both windows. While P. inconstans and C. caeruleocinctus are widely known as highly synanthropic species (BERG et al., 2008), O.

52 sabulosus is a species with a very high ecological adaptability, both occurring in natural and in highly disturbed habitats (KANIA & TRACZ, 2005).

A striking characteristic is that these three species are among the few known millipedes that are considered as pest species (BLOWER, 1985; BRUNKE et al., 2012; KANIA & TRACZ, 2005). Therefore, these species also use the matrix surrounding the forests as habitat, probably for foraging, but they may be dependent on forests and other stable semi-natural habitats for reproduction. Additionally, arthropods from agricultural landscape are known to use semi-natural habitat adjacent to crop fields as overwintering habitat (PFIFFNER & LUKA, 2000).

The mass occurrence of these species is a phenomenon that has already been observed and described dozens of times, especially for O. sabulosus, but also for C. caeruleocinctus and Tachypodoiulus niger, another species that was very abundant in this study, though not limited to open field landscapes (KANIA & TRACZ, 2005; VOIGTLÄNDER, 2005). In general, areas influenced by man are more affected by these mass occurrences (VOIGTLÄNDER, 2005). A further explanation for the high density of these synanthropic species is that these are probably dependent both on the semi-natural habitat and on the surrounding agricultural land, and are therefore have a higher mobility and therefore a higher chance of getting trapped in the pitfall traps.

In addition to abundance, the species richness was also higher in the open field landscape. Almost no species were typical indicators for bocage landscape, exept for Glomeris intermedia in Belgium, which only occurred in the bocage landscape and Polyzonium germanicum, which was found almost exclusively in the bocage landscape in South Sweden. It seems therefore that the typical stenotopic species such as Glomeris marginata and the more eurytopic species occur in forests in both landscapes, while true synanthropic species, such as C. caeruleocinctus are truly dependent on intensively managed agricultural landscapes. Therefore, in addition to the ‘usual’ species pool that is found in both landscapes, a number of synanthropic species are added to the open field landscape, thus increasing the total diversity.

When looking at the ordination plots (figure 13), the land use intensity is the major factor influencing species composition. In each region the intensity is the variable with the highest correlation with at least one of the two species axes (table 17). Only in Belgium the correlation

53 was less clear, as the bocage landscape was also relatively intensively managed here, but even then the effect was very clear. The main differences are, as discussed before, the presence of synanthropic species, such as C. caeruleocinctus and P. inconstans in the open field landscape.

An important remark that should be made is the fact that both windows in each region were separated a few dozens of kilometres from each other, and therefore some spatial autocorrelation might exist between forest patches in the same window. However, this effect is unlikely to give consistent differences between both windows in all five regions. Therefore this spatial factor is probably only of minor importance compared to intensity of land use.

Our findings seem to contradict earlier research that concluded that landscapes with higher agricultural intensity have lower biodiversity and abundance of millipedes (ATTWOOD et al.,

2008; CALLAHAM et al. , 2006; DAUBER et al., 2005; RAHMAN et al., 2011). However, these studies mostly investigated the agricultural matrix itself and not the forest patches. Therefore, the diversity may be lower at a landscape scale in the open field landscape than in the bocage landscape, but higher at the forest patch level due to spillover from species from agricultural habitat. DAUBER et al. (2005) also found a high dependence on surrounding landscape by millipedes.

2.2. Effect of forest age Many millipede species are slow dispersers (DUNGER & VOIGTLÄNDER, 1990). Therefore it can be expected that at least some species will be typical indicators for old forests and that the species communities will differ between old and recent forests.

Millipede abundance was higher in old forest. One explanation is that certain typical species for old forest, such as Glomeris marginata often occur in very large densities. However, when the interaction with land use intensity is considered, it seems that in the open field landscape the opposite is true. The reason for this might be that typical indicator species for bocage landscape, such as Glomeris intermedia and Polyzonium germanicum, though not significant indicators for old forest, are still far more common in old forest than in recent forest, and often occur in large numbers. Therefore, these species may cause the higher millipede density in old forests in the bocage landscape. In the open field landscape, the most abundant species are typical for more open habitats. Species such as Cylindroiulus caeruleocinctus are present in very high numbers

54 and show in several regions a higher, though statistically not always significant, affinity for recent forests (e.g. in South Sweden and Belgium). This interaction effect is therefore probably caused by the autecology of individual species.

Despite the total number of species being not significantly different between old and recent forest patches, the community composition is clearly different in most regions. In Central Sweden, no difference between old and recent forests was observed. In North France the differences in community composition were less pronounced than in regions, where the correlation with at least one of the two first ordination axes was at least 0.5 (table 17). For Central Sweden, this might be because the most typical species for old forest do not have a range that extends this far northward and that because of this northern position the fauna is impoverished, compared to other regions.

Several studies have investigated succession in millipede communities, usually after afforestation on degraded terrains, such as abandoned mines in forest (DUNGER et al., 2001; SCHREINER et al.,

2012; TAJOVSKÝ, 2001) and dune ecosystems (REDI et al., 2005; VAN AARDE et al., 1996), fallow ground (SCHREINER et al. 2012) and cleared rainforest (NAKAMURA et al., 2003). SCHREINER et al. (2012) found C. caeruleocinctus as typical for open, fallow ground, but observed that this species disappeared when the forest began to grow. In our study, C. caeruleocinctus was only in South Sweden a significant indicator for recent forest. However, in other regions the presence of this species, which nearly exclusively was found in open field landscape, might be due to spillover from agricultural fields. In Belgium, almost 99% of all individuals of this species were also found in recent forest. The same conclusion can be drawn about Polydesmus inconstans, which is known as a species of open landscapes and an early successional pioneer (BERG et al.,

2008; DUNGER & VOIGTLÄNDER, 1990; TAJOVSKÝ, 2001). Another species that is typically considered as an early pioneer is Craspedosoma rawlinsi (DUNGER & VOIGTLÄNDER 1990;

TAJOVSKÝ 2001). However, in our study this species was a significant indicator for old forest in South Sweden. The reason for this might be that the landscapes in this region only contained only nine recent forests and that therefore the setup may be unbalanced with too few recent forests for drawing reliable conclusions. In other regions, this species was too rare to draw firm conclusions.

SCHREINER et al. (2012) considered Julus scandinavius as an indicator for ageing forests. However, J. scandinavius was in our investigation never seen as an indicator of old forest in any region studied. The forests investigated by SCHREINER et al. (2012) were of a relatively recent

55 age, with the oldest forests being about 150 years old and most forests younger than 100 years.

This seems to coincide with the results from DUNGER & VOIGTLANDER (1990) and TAJOVSKÝ (2001), who considered J. scandinavius as an intermediate successional species that appeared in higher numbers a few decades after afforestation but that is present in lower numbers in forests of any age. Unciger foetidus is also considered as an intermediate successional species (DUNGER &

VOIGTLÄNDER 1990; TAJOVSKÝ 2001), but this species was in our study only present in significant numbers in South Sweden, where it seemed to be a species typical for recent forest. However, because of the low number of recent forests in this region, it is difficult to assess the reliability of this result.

The clearest indicator species for old forest in our study was G. marginata, which was a clear indicator of old forest in both regions where it was found, being West Germany and Belgium. G. intermedia was also nearly exclusively found in old forests in Belgium, but due to its low abundance, it was not considered as a significant indicator species in the analysis (table 19).

BERG et al. (2008) mentions these species to be typical for old forests. Investigations on afforested mine sites did not find these species after several decades (DUNGER & VOIGTLÄNDER 1990), which is consistent with the status of these species as indicators for old forest.

P.germanicum, which was found in this study only in South Sweden, is known as a late successional species from afforested former mine sites (Dunger & Voigtländer, 1990, 2009), and colonization by this species is known to take several decades. However, in this study it was just not significant (p=0.064), despite 92% of all specimens being found in old forest. This might partially be caused by the uneven sampling design between old and recent forest in this region.

Polydesmus denticulatus is mentioned as an old forest species by DUNGER & VOIGTLÄNDER (1990, 2009), but this association was not found in this study, except in North France, where it was slightly insignificant (p=0.057) with almost 90% of all individuals of the species trapped in old forest. Polydesmus angustus is considered as an old forest species in Belgium, but was barely significant in the indicator species analysis (p=0.048, table 19), while in North France and West Germany, where this species was far more common, it was not associated with old forests. Therefore, this result is probably biased and should be considered with care.

As most species are slow dispersers, habitat stability through time is an important characteristic to maintain viable populations. Recent forests are populated with faster dispersing species and

56 generalist species that can also survive in more open habitat and do not need forest for their continued existence. In old forests, slow dispersing forest specialists, such as Glomeris marginata and G. intermedia are a typical component of the species assemblage, but species that occur in recent forests usually also occur in old forests. This explains why typical indicator species were found for old forests, but not for recent forests. Another reason might be because the recent forests still differed slightly in age: some were very young poplar stands with very open canopy, while other ‘recent’ forests were several decades old, which could have an important influence on the results, as the early successional stages seem to happen in a few years, or at most a few decades (DUNGER & VOIGTLÄNDER 1990). To assess what species are typical for recent forest, the exact age of forests, or distinct categories for very recent forests should be used. As old forest communities are clearly different from recent species assemblages, it is important to maintain these old forest fragments to conserve the typical species for old forests.

2.3. Effect of patch size Small forests seem to have a slightly higher density in millipedes by a factor 1.19. Previous studies on several species groups, that tried to find a relation between population density and patch area, found different results, going from a positive relation to a negative, or even no relation at all (BOWMAN et al., 2002; HAMBÄCK & ENGLUND, 2005). In this study a significant difference in abundance was found. Small forest fragments seem to contain higher densities of millipedes, which may be caused by edge effects. Millipedes living in the fields surrounding the forest may reach the centre of the small forest fragments and thereby will be trapped more often. As these species (e.g. Cylindroiulus caeruleocinctus, Polydesmus inconstans,…) often occur in higher densities than typical forest species, this may explain the high number of individuals caught in small forest fragments.

No significant differences in species richness between small and large fragments were found in this study. Only in South Sweden, a slightly significant relation with the ordination axes was observed. Therefore, the influence of patch size is omitted from the DCA-plots as the effect on community composition was negligible. It may be possible that most millipede species only need a very small minimum surface to support a viable population and that forest size has no influence on the species composition. Additionally, several species live in more open habitats and are not solely dependent on forests. Therefore, presence of these species is independent of forest size.

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2.4. Effect of location in the forest The activity density of millipedes was higher at the forest edges by a factor 1.33. This is probably caused by the fact that the species of more open habitat were caught in higher densities. As these species are very abundant in open habitat, they will often reach the edge of the forest. However, in large forests, less individuals will get to the centre, which is demonstrated by the interaction effect (figure 9). Furthermore, the difference between edge and centre is greater in old forests (figure 10). This might be because some recent forests were still very open, which enabled the abundant species from open habitats to go deeper into the forest. In intensively used landscapes the difference between centre and edge is also relatively smaller than in bocage landscape, but this interaction is only small and barely significant (figure 11). This interaction may be because the absolute number of individuals is much higher in the open field landscape than in the bocage landscape, but as it has relatively low significance, it may even be caused by random variation.

The effect of location on the DCA was negligible and the correlation with the first two ordination axes was very small (table 17). No differences in diversity and species composition were noticed between samples from the centre and from the edge of a forest. This is in contradiction with the findings of WEIERMANS & VAN AARDE (2003), who investigated millipede communities in coastal dunes and found different communities in the centre and in the edges. However, as the study focused on a completely different ecosystem, comparisons should be made with care.

Possibly the resemblance between the communities in the centre and the edge of the forests is because the forests were all relatively small and all experienced edge effects over their complete surface (Barbosa & Marquet, 2002). Therefore, the differences in biotic and abiotic circumstances between both locations were not sufficient to cause a difference in millipede species composition. However, as the millipede density was higher in forest edges, especially in larger forest patches, certain differences still exist. Even then, species with more affinity to meadows and other open habitats, such as C. caeruleocinctus were often found in the centre of large forests, which may indicate a relatively high mobility of these species, thus enabling spillover from adjacent fields, or a generalist way of life of these species, where they can also live in forests, though in lesser numbers.

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2.5. Effect of tree species In most regions, the effect of tree species was limited, and the dominant species are not displayed on the ordination plot. However, in Belgium, forests that were dominated by oak, beech and poplar showed clear differences. These differences are, however, at least in part caused by the correlation between tree species composition and forest age: most forests with poplar are only recently planted and house communities typical for recent forest. Oak and beech forests that were sampled are often old growth forests, which influences their community composition. Therefore, the impact of dominant tree species should be considered with care.

Earlier studies showed a clear correlation between dominant tree species and millipede communities (GAVA, 2004; MEYER & SINGER, 1997; STAŠIOV et al., 2012; WYTWER et al., 2009). Probably the main effects of tree species on millipede community composition are by changing the amount of light that reaches the surface and by influencing the quality and quantity of litter production, thus also changing humus quality and soil pH. However, in this study, the main focus was not on tree species, and therefore our discussion is rather limited. In the mentioned studies, the number of tree species investigated was limited, which made data easier to interpret. In our study, many different dominant tree species occurred, some only in one or a few forests, which made the influence of tree species very hard to interpret.

2.6. Effect of tree, shrub, herb and moss cover The effect of moss and shrub cover showed no consistent trend over the regions. Additionally, the correlation with the ordination axes was relatively low (table 17). Moss cover was only a significant explanatory variable in the DCA of Central Sweden, which is probably caused by the fact that forests in bocage landscape had a higher moss cover. The effect of shrub cover was negligible in each region.

Herb and tree cover showed a much higher correlation with the ordination axes and had a more or less consistent effect over the regions. Except for South-Sweden, where the effect of tree cover was rather limited and Central Sweden, where the relation with the ordination axis was not significant, both had opposite effects. The most probable explanation is that a dense tree cover not only influences the species composition but also the herb layer. Therefore the correlation between herb cover and millipede community may not imply any causal effect.

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The correlation values of tree cover with the ordination axes were slightly higher, but still much lower than those of intensity and age.

2.7. Effect of dead wood The only species that is explicitly mentioned in literature to be associated with dead wood is

Cylindroiulus punctatus (BERG et al., 2008). However, the species is very common throughout its range, even in regions like Belgium, where forests have only very little dead wood. Therefore, C. punctatus does probably need only small amounts of dead wood. Still, the ordination plots show a positive relation between presence of C. punctatus and presence of coarse woody debris. Glomeris marginata also seems to show an affinity with dead wood, but it is unclear whether there is a causal link. G. marginata is nearly uniquely found in old forests. As the amount of coarse woody debris might be higher in these forests, the correlation might not be caused by a causal effect. In the ordination plots, no significant relation between amount of dead wood and the ordination axes was observed.

2.8. Other factors Soil and humus chemistry are often considered as the most important factors in influencing the community composition of millipedes. The data on these factors is still being processed, but will certainly provide more insights in the community composition. In general mull humus is known to harbor a richer fauna, not only for Diplopoda, but also for Gastropoda, Isopoda and

Lumbricidae (DAVID et al., 1993).

3. Impact on ecosystem function Soil fauna diversity, number of trophic levels and presence of keystone species have a strong impact on decomposition, but the effect of diversity within functional groups is not very clear at the moment (HÄTTENSCHWILER et al., 2005). HEEMSBERGEN et al. (2004) indicate that positive effects on litter decomposition are observed when functionally dissimilar species are present, while functionally similar species such as P. denticulatus and the isopod Oniscus asellus might inhibit each other by competing on the same resources. However, to really assess the impact of community composition on nutrient cycling, the autecology of each species should be investigated and the effects of complementarity or inhibition by other species groups, such as isopods, should be evaluated.

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4. Implications for landscape management The surrounding land use has a clear impact on the species composition in forests. Therefore, forest fragments in different landscapes may have different importance for biodiversity and ecosystem function. Even in very intensively managed landscapes, the typical forest species generally seem to survive in the forest patches, while the total species pool is supplemented with typical species of open landscape.

Additionally, old forests seem to house different communities with typical species. To conserve these species, it is essential to conserve the old forest fragments. As size of the forest fragments had a negligible effect on species composition, even small forest fragments in very intensively used agricultural landscapes can probably play an important role in conserving millipede biodiversity and function, often housing populations of rare species, such as Glomeris intermedia

(BERG et al. 2008), and deserve protection.

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Conclusion

Millipede communities show large differences between forest patches. Within each region, the effect of surrounding land use intensity seems to have the greatest influence on the community, especially by supplementing the population with species from more open habitat, which can often become very abundant, or even dominant in the forest patch. These species are probably not dependent on forests, and mainly occur due to spillover from the adjacent agricultural fields.

Another very important characteristic of forests is age. Many millipede species are slow dispersers and therefore are not able to colonize recent forests. These species are mostly restricted to old forest patches. To guarantee the continuing existence of these species it is especially important to conserve these old forest patches.

Other characteristics play a less important role in the millipede community composition. Some species show a slight affinity with dead wood, which might partly be caused by a correlation between forest age and presence of dead wood. Tree cover also seems to be positively correlated with forest age and negatively with herb cover. Therefore, the individual contribution of these characteristics is very hard to separate. Furthermore, other factors such as humus and soil chemistry, of which no data was available, might provide an important contribution to species composition.

Size of the forest patches only seems to influence millipede density. More millipedes were trapped in small forest fragments, mainly because of spillover of very abundant species from the agricultural landscape to the forest. This effect was also seen when comparing the edges of each forest with their centres.

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Future research

This research provided extra knowledge on the factors that influence biodiversity in semi-natural habitat in intensively used landscape. However, several factors, such as soil physics and chemistry, humus characteristics and interactions with other species groups have not been evaluated. As these factors can provide us with extra information and can explain more variation between the communities, it might be interesting to take them into account.

Especially the interactions with other organisms are important and interesting factors that have not been treated in this thesis. Other detritivores such as woodlice and earthworms might have complementary roles, or they might compete with each other. Information on these interactions will provide us with more insights on ecosystem function and ecological importance of millipedes.

Additionally, within the millipedes, the morphological diversity is enormous. It is therefore a reasonable question whether this is also refelected in their functional diversity: rollers that live most of their lives in the litter layer, such as Glomeris marginata probably play a completely different role in nutrient recycling than Cylindroiulus punctatus, which is most often associated with dead wood, or Cylindroiulus caeruleocinctus, which occurs more often in open habitats and is known to feed on living plant tissue. Therefore, a good knowledge on the ecology of all relevant species, which is now only scarcely available, is indispensable to evaluate the functional diversity of these species.

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Acknowledgements

First of all I would like to thank my tutor, ir. Pallieter De Smedt for helping me throughout the development of this thesis. His corrections and constructive advice have been a tremendous help in the development of this thesis.

Furthermore I am very grateful to my supervisor, prof. dr. ir. Kris Verheyen for offering me the opportunity to work on this subject, and for thoroughly reading and correcting this manuscript, which made major improvements to this paper.

I also want to thank prof. dr. Dries Bonte, who acted as my co-supervisor during this thesis for offering workspace at his lab. Furthermore I am very grateful for him contacting prof. dr. Matty Berg.

Prof. dr. Matty Berg revised the reference collection that I composed at the start of this thesis. This has been of great help for me, and it certainly aided in identifying the correct species.

Several researchers helped with the sampling campaign by putting up pitfall traps in Sweden, Germany and France and sorting out the animals in these traps. It is thanks to them that this research was possible on such large geographical scale and with such great amount of specimens.

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Appendix: dominant tree species

Dominant tree species in the sampled forests

Species Abbreviation Acer pseudoplatanus Acerpseu Alnus glutinosa Alnuglut Betula pendula Betupend Betula pubescens Betupube Carpinus betulus Carpbetu Castanea sativa Castsati Corylus avellana Coryavel Fagus sylvatica Fagusylv Fraxinus excelsior Fraxexce Malus sylvestris Malusylv Picea abies Piceabie Pinus sylvestris Pinusylv Populus x canadensis Popucana Populus tremula Poputrem Prunus avium Prunaviu Prunus padus Prunpadu Prunus serotina Prunsero Quercus robur Querrobu Quercus rubra Querrubr Salix alba Salialba Salix caprea Salicapr Sambucus nigra Sambnigr Sorbus aucuparia Sorbaucu Tilia spp. Tilia Ulmus spp. Ulmus

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