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
1
2
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.
2.
Importance of forest age .................................................................................................. 14
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
3
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.
5
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.
6
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.
8
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.
9
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).
Belgium
France
Germany
Sweden
Total forest area (km²)
7,060
175,720
110,760
306,250
Total area (km²)
30,280
550,100
348,770
410,310
Forest index (%)*
23
32
32
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 interpatch 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 &
13
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.
14
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.
16
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 snakelike, 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).
20
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 mesoand 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).
23
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.
24
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
study
were selected according to age and size: four young,
small patches, four old, small patches, four young,
large patches and four old, large patches (figure 5). Only in the Bocage-window of SouthSweden, 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
were calculated from regional weather station
Figure 4: Schematic view of the arrangement of the pitfall traps in
a forest patch
data, at less than 30km from the sampling localities (DE FRENNE et al., 2011).
25
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.
26
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.
27
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
C-Sweden
S-Sweden
W-Germany
Belgium
Window Start period 1
B
12/06/2013
13/06/2013
O
10/06/2013
11/06/2013
B
05/06/2013
06/06/2013
O
07/06/2013
08/06/2013
B
23/05/2013
24/05/2013
27/05/2013
O
28/05/2013
29/05/2013
B
13/05/2013
14/05/2013
O
15/05/2013
16/05/2013
End period 1
26/06/2013
27/06/2013
24/06/2013
25/06/2013
18/06/2013
19/06/2013
20/06/2013
22/06/2013
06/06/2013
07/06/2013
10/06/2013
11/05/2013
12/05/2013
27/05/2013
28/05/2013
29/05/2013
30/05/2013
Start period 2
31/07/2013
01/08/2013
29/07/2013
30/07/2013
/
End period 2
14/08/2013
15/08/2013
12/08/2013
13/08/2013
/
/
/
08/07/2013
09/07/2013
22/07/2013
23/07/2013
10/07/2013
24/07/2013
12/07/2013
26/07/2013
01/07/2013
15/07/2013
02/07/2013
16/07/2013
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.
28
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
Species number
Density
Shannon-diversity
Explanatory variable(s)
Intensity
Intensity+Age+Size+Intensity:Age
-
29
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 (CSweden – 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).
30
Then, a DCA was performed for each region individually with the same variables. In CentralSweden, 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
Tree%
Shrub%
Herb%
Moss%
Dead wood%
Location
Tree species
Size1
Intensity1
Age1
Region1,2
Meaning
Percentage cover of tree layer
Percentage cover of shrub layer
Percentage cover of herb layer
Percentage cover of moss layer
Percentage dead wood i.r.t. total standing biomass
Location in the forest
Dominant tree species
Size of the forest
Intensity of surrounding land use
Forest age
Region where the sampling took place
Type
Continuous
Continuous
Continuous
Continuous
Continuous
Categorical
Categorical
Categorical
Categorical
Categorical
Categorical
Categories
Centre, Edge
See appendix 1
Small, Large
Bocage, Open field
Old, Recent
C-Sweden, S-Sweden,
W-Germany, Belgium,
N-France
1
Variables used in the combined analysis
Variables used only in the combined analysis
2
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.
32
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
C-Sweden
S-Sweden*
W-Germany
Belgium
N-France
Bocage
421
622
508
1123
602
Open field
3173
670
602
1447
1341
Total individuals
3594
1292
1110
2570
2101
Total species
18
16
15
22
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.
Intercept (Belgium)
Central Sweden
South Sweden
West Germany
North France
Estimate
2.140
-0.303
-0.233
-0.380
-0.366
Std. Error
0.061
0.095
0.093
0.095
0.098
33
z value
35.295
-3.214
-2.509
-3.994
-3.747
p-value
<0.001
0.001
0.012
<0.001
<0.001
Avg species
8.50 ± 1.06
6.27 ± 1.09
6.73 ± 1.10
5.81 ± 1.09
5.90 ± 1.10
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.
South Sweden
West Germany
Belgium
North France
Central Sweden
0.430
0.436
<0.001
0.535
South Sweden
West Germany
Belgium
0.115
0.003
0.162
<0.001
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.
Belgium (Intercept)
C-Sweden
S-Sweden
W-Germany
N-France
Estimate
2.147
-0.550
-0.285
-0.409
-0.454
Std. Error
0.085
0.141
0.130
0.135
0.137
z value
25.135
-3.897
-2.187
-3.025
-3.312
p-value
<0.001
<0.001
0.029
0.002
<0.001
Avg species±StDev
8.56 ± 1.09
4.94 ± 1.15
6.44 ± 1.14
5.69 ± 1.14
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.
South Sweden
West Germany
Belgium
North France
Cental Sweden
0.056
0.335
<0.001
0.520
South Sweden
West Germany
Belgium
0.335
0.007
0.200
<0.001
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
Belgium (Intercept)
C-Sweden
S-Sweden
W-Germany
N-France
Estimate
2.132
-0.080
-0.177
-0.351
-0.267
Std Error
0.086
0.129
0.132
0.134
0.139
z value
24.780
-0.624
-1.335
-2.624
-1.920
34
p value
<0.001
0.532
0.182
0.009
0.055
Avg species±StDev
8.43 ± 1.09
7.78 ± 1.14
7.06 ± 1.14
5.94 ± 1.14
6.46 ± 1.15
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.
South Sweden
West Germany
Belgium
North France
Central Sweden
0.403
0.028
0.431
0.131
South Sweden
West Germany
Belgium
0.173
0.101
0.483
0.003
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
35
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
Allajulus nitidus
Blaniulus guttatus
Brachydesmus superus
Brachyiulus pusillus
Choneiulus palmatus
Chordeuma sylvestre
Craspedosoma rawlinsi
Cylindroiulus caeruleocintus
Cylindroiulus latestriatus
Cylindroiulus punctatus
Enantiulus nanus
Glomeris intermedia
Glomeris marginata
Julus scandinavius
Julus terrestris
Leptoiulus belgicus
Leptoiulus kervillei
Melogona gallica
Melogona voigtii
Mycogona sp
Nemasoma varicorne
Ommatoiulus sabulosus
Ophiodesmus albonanus
Ophyiulus pilosus
Orthochordeumella pallida
Polydesmus angustus
Polydesmus complanatus
Polydesmus coriaceus
Polydesmus denticulatus
Polydesmus inconstans
Polyzonium germanicum
Propolydesmus testaceus
Proteroiulus fuscus
Tachypodoiulus niger
Unciger foetidus
Total number of species
C. Sweden
S.Sweden*
W. Germany
Belgium
N. France
Abbreviatio
Alla niti
Blan gutt
Brac supe
Brac pusi
Chon palm
Chor sylv
Cras rawl
Cyli caer
Cyli late
Cyli punc
Enan nanu
Glom inte
Glom marg
Julu scan
Julu terr
Lept belg
Lept kerv
Melo gall
Melo voig
Myco sp
Nema vari
Omma sabu
Ophi albo
Ophy pilo
Orth pall
Poly angu
Poly comp
Poly cori
Poly dent
Poly inco
Poly germ
Prop test
Prot fusc
Tach nige
Unci foet
18
16
15
*Only the first replicate of the first sampling period has been analyzed in South Sweden
36
22
21
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'.
37
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.
Intercept (bocage)
Intensity (open field)
Estimate
1.819
0.136
Std. Error
0.071
0.063
z-value
25.539
2.174
p-value
<0.0001
0.0297
Avg species±StDev
6.17 ± 1.07
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.
39
Table 14: Generalized linear mixed model with log-link function for variables explaining millipede density at the forest
level..
Intercept
Intensity (Open field)
Age (Recent)
Surface (Small)
Intensity:Age
Estimate
3.788
0.401
-0.332
0.177
0.478
Std. Error
0.204
0.028
0.034
0.020
0.042
z-value
18.609
14.140
-9.769
8.684
11.283
p-value
<0.001
<0.001
<0.001
<0.001
<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.
40
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
Intensity (Open field)
Age (Recent)
Size (Small)
2.866
0.669
-0.340
0.087
0.201
0.037
0.043
0.030
14.232
17.907
-7.933
2.926
<0.001
<0.001
<0.001
0.003
Location (Edge)
Size:Location
Age:Location
Intensity:Location
0.282
0.162
-0.136
-0.096
0.040
0.040
0.040
0.043
4.07
4.069
-3.328
-2.223
<0.001
<0.001
<0.001
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.
41
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.
42
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.
43
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
Central Sweden
Length of gradient
Cumul. % variance explained
Cumul % species-environment relation
South Sweden
Length of gradient
Cumul. % variance explained
Cumul % species-environment relation
West Germany
Length of gradient
Cumul. % variance explained
Cumul % species-environment relation
Belgium
Length of gradient
Cumul. % variance explained
Cumul % species-environment relation
North France
Length of gradient
Cumul. % variance explained
Cumul % species-environment relation
Axis 1
Axis 2
2.57
20.1
30.4
2.83
18.4
29.4
3.11
19.4
22.4
3.08
15.8
23.4
2.74
19.0
21.9
3.056
37.5
46.3
4.34
33.4
50.2
2.25
35.5
41.6
2.34
26.1
35.0
2.32
30.0
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
W Germany
Belgium
Axis 1
Axis 2
S Sweden
Axis 1
Axis 2
Axis 1
Axis 2
Axis 1
Axis 2
North France
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
C-Sweden
S-Sweden
W-Germany
Belgium
N-France
Intensity
Bocage
Open field
Bocage
Open field
Bocage
Open field
Bocage
Species
A
B
Ind. val.
p-value
Cras rawl
1.00
0.13
0.35
0.467
Cyli caer
Poly inco
Brac supe
Omma sabu
Ophy pilo
1.00
0.99
0.95
0.85
0.80
1.00
0.93
0.86
0.93
0.64
1.00
0.96
0.90
0.89
0.72
0.001
0.001
0.001
0.001
0.030
Prot fusc
Blan gutt
Nema vari
Cyli punc
Chon palm
Alla niti
Enan nanu
Julu terr
Unci foet
0.76
1.00
1.00
0.63
0.77
1.00
1.00
1.00
1.00
0.43
0.14
0.14
0.21
0.14
0.07
0.07
0.07
0.07
0.57
0.38
0.38
0.37
0.33
0.27
0.27
0.27
0.27
0.137
0.204
0.198
0.510
0.459
0.482
0.482
0.477
0.461
Poly germ
Cyli caer
0.95
0.99
0.69
0.86
0.81
0.92
0.003
0.001
Alla niti
Brac pusi
Melo voig
Tach nige
1.00
1.00
1.00
1.00
0.14
0.07
0.07
0.25
0.38
0.27
0.27
0.50
0.194
0.448
0.488
0.088
Omma sabu
0.85
0.63
0.73
0.028
Cras rawl
1.00
0.25
0.50
0.100
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
Bocage
Cyli caer
Poly inco
Ophi albo
1.00
1.00
1.00
0.25
0.24
0.12
0.50
0.49
0.34
0.108
0.097
0.515
Cyli caer
Poly angu
Melo gall
Prop test
Poly dent
0.96
0.93
0.93
0.98
0.95
0.76
0.71
0.65
0.47
0.47
0.86
0.81
0.78
0.68
0.67
0.002
0.011
0.002
0.013
0.011
Ala niti
0.90
0.47
0.65
0.126
Brac pusi
1.00
0.29
0.54
0.042
Open field
48
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).
49
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
Central Sweden
South Sweden
Age
Recent
Old
Recent
Old
West Germany
Recent
Belgium
Old
Recent
Old
North France
Recent
Old
Species
A
B
Ind. val.
p-value
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
Cras rawl
1.00
0.14
0.38
0.187
Unci foet
Cili caer
0.85
0.72
1.00
0.70
0.92
0.71
0.007
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
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
Prot fusc
0.91
0.25
0.48
0.287
Cras rawl
0.86
0.19
0.40
0.479
Glom marg
0.91
0.63
0.75
0.013
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
Glom marg
Poly angu
0.99
0.81
0.50
0.50
0.70
0.64
0.016
0.048
Glom inte
0.99
0.25
0.50
0.182
Brac pusi
1.0000
0.25
0.50
0.097
Poly dent
0.89
0.38
0.58
0.057
Blan gutt
1.00
0.13
0.35
0.503
50
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 betadiversity (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).
51
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.
57
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.
58
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.
59
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.
60
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.
61
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.
62
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.
63
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.
64
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Appendix: dominant tree species
Dominant tree species in the sampled forests
Species
Acer pseudoplatanus
Alnus glutinosa
Betula pendula
Betula pubescens
Carpinus betulus
Castanea sativa
Corylus avellana
Fagus sylvatica
Fraxinus excelsior
Malus sylvestris
Picea abies
Pinus sylvestris
Populus x canadensis
Populus tremula
Prunus avium
Prunus padus
Prunus serotina
Quercus robur
Quercus rubra
Salix alba
Salix caprea
Sambucus nigra
Sorbus aucuparia
Tilia spp.
Ulmus spp.
Abbreviation
Acerpseu
Alnuglut
Betupend
Betupube
Carpbetu
Castsati
Coryavel
Fagusylv
Fraxexce
Malusylv
Piceabie
Pinusylv
Popucana
Poputrem
Prunaviu
Prunpadu
Prunsero
Querrobu
Querrubr
Salialba
Salicapr
Sambnigr
Sorbaucu
Tilia
Ulmus
76