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The ongoing decline of many plant species in Northwest Europe indicates that traditional conservation measures to improve the habitat quality, although useful, are not enough to halt diversity losses. Using recent databases, we show for the first time that differences between species in adaptations to various dispersal vectors, in combination with changes in the availability of these vectors, contribute significantly to explaining losses in plant diversity in Northwest Europe in the 20th century. Species with water- or fur-assisted dispersal are over-represented among declining species, while others (wind- or bird-assisted dispersal) are under-represented. Our analysis indicates that the 'colonization deficit' due to a degraded dispersal infrastructure is no less important in explaining plant diversity losses than the more commonly accepted effect of eutrophication and associated niche-based processes. Our findings call for measures that aim to restore the dispersal infrastructure across entire regions and that go beyond current conservation practices.
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LETTER
Dispersal failure contributes to plant losses
in NW Europe
Wim A. Ozinga,
1,2
* Christine
Ro
¨mermann,
3
Rene
´e M. Bekker,
4
Andreas Prinzing,
5
Wil L. M.
Tamis,
6
Joop H. J. Schamine
´e,
1,2
Stephan M. Hennekens,
2
Ken
Thompson,
7
Peter Poschlod,
8
Michael Kleyer,
9
Jan P. Bakker
4
and Jan M. van Groenendael
1
Abstract
The ongoing decline of many plant species in Northwest Europe indicates that
traditional conservation measures to improve the habitat quality, although useful, are not
enough to halt diversity losses. Using recent databases, we show for the first time that
differences between species in adaptations to various dispersal vectors, in combination
with changes in the availability of these vectors, contribute significantly to explaining
losses in plant diversity in Northwest Europe in the 20th century. Species with water- or
fur-assisted dispersal are over-represented among declining species, while others (wind-
or bird-assisted dispersal) are under-represented. Our analysis indicates that the
Ôcolonization deficitÕdue to a degraded dispersal infrastructure is no less important in
explaining plant diversity losses than the more commonly accepted effect of
eutrophication and associated niche-based processes. Our findings call for measures
that aim to restore the dispersal infrastructure across entire regions and that go beyond
current conservation practices.
Keywords
Colonization deficit, dispersal infrastructure, dispersal vectors, diversity loss, eutrophi-
cation, functional traits, land-use changes.
Ecology Letters (2009) 12: 66–74
INTRODUCTION
Growing concern about the ongoing loss of biodiversity has
resulted in increased efforts throughout the world to protect
endangered species and to conserve and restore endangered
ecosystems (Balmford et al. 2005). In view of the large input
of financial and human resources into nature conservation
and restoration, improving our insights into the mechanisms
behind observed losses of biological diversity is therefore
one of the great scientific challenges for the 21st century.
Changes in plant species composition of vegetations in man-
made temperate landscapes are often explained in terms of
habitat degradation, especially by eutrophication (Vitousek
et al. 1997; Bobbink et al. 1998; Grime 2001; Tilman et al.
2002; Stevens et al. 2004; Suding et al. 2005). Restoration of
habitat quality, however, often fails to deliver the expected
plant diversity (Dobson et al. 1997; Bakker & Berendse
1999). It has therefore been questioned during the past
1
Aquatic Ecology and Environmental Biology Research Group,
Department of Ecology, Radboud University Nijmegen,
Toernooiveld 1, NL-6525 ED Nijmegen, The Netherlands
2
Centre for Ecosystem Studies, Wageningen University and
Research, Alterra, PO Box 47, NL-6700 AA Wageningen,
The Netherlands
3
Department of Physical Geography, University of Frankfurt am
Main, PO Box 111932, D-60054 Frankfurt am Main, Germany
4
Community and Conservation Ecology Group, University of
Groningen, PO Box 14, NL-9750 AA Haren, The Netherlands
5
Universite
´de Rennes 1, Unite
´Mixte de Recherche CNRS 6553
ÔEcobioÕ: Ecosyste
`mes–Biodiversite
´–Evolution, Campus de
Beaulieu, Ba
ˆtiment 14A, 263 Avenue du Ge
´ne
´ral Leclerc, 35042
Rennes Cedex, France
6
National Herbarium of the Netherlands, Leiden Branch Insti-
tute of Environmental Sciences Leiden, PO Box 9518, NL-2300
RA Leiden, The Netherlands
7
Department of Animal and Plant Sciences, University of
Sheffield, Sheffield S10 2TN, UK
8
Institute of Botany, University of Regensburg, D-93040
Regensburg, Germany
9
Landscape Ecology Group, Carl-von-Ossietzky University of
Oldenburg, D-26111 Oldenburg, Germany
*Correspondence: E-mail: wim.ozinga@wur.nl
Ecology Letters, (2009) 12: 66–74 doi: 10.1111/j.1461-0248.2008.01261.x
2008 Blackwell Publishing Ltd/CNRS
decade to what extent community composition is also
constrained by limited rates of seed dispersal on a regional
scale (Tilman 1997; Poschlod & Bonn 1998; Poschlod et al.
1998; Ro
¨mermann et al. 2008; Turnbull et al. 2000; Mouquet
et al. 2004; Ozinga et al. 2005; Nathan 2006). Metapopula-
tion theory asserts that regional survival of species requires
that local populations are connected by sufficient rates of
dispersal (Hanski 1998). The newly emerging concept of
ÔmetacommunitiesÕextends metapopulation theory towards
the community level and tries to integrate the processes that
drive the dynamics of species across local and regional scales
(Leibold et al. 2004).
In contrast to mobile animal species, plants depend for the
transport of their seeds among sites on external vectors,
including water, wind, birds and large mammals, each with
their own characteristics. At the landscape level, these
dispersal vectors act like a complex Ôdispersal infrastructureÕ
(Poschlod & Bonn 1998; Ozinga et al. 2004). Although
changes in the relative availability of dispersal vectors during
the 20th century have been documented in many parts of the
industrialized world (see Appendix S2 in Supporting infor-
mation), their impact on plant diversity has never been
examined in large-scale studies. We might expect that such
impacts will be strongly vector-specific. For example,
dispersal by water has been restricted by the regulation of
the natural flood regimes of rivers and brooks for the purpose
of flood control, while dispersal by large mammals has
declined due to the change from livestock grazing on common
grounds to grazing in fenced fields or livestock housing. In
Northwest Europe, these changes in dispersal infrastructure
mostly took place some 50–150 years ago (see Appendix S2).
The potential effect of changes in the availability of
dispersal vectors on species losses is founded upon two
premises. First, vascular plant species show interspecific
differences in the kind of dispersal vectors that can effectively
transport their seeds (known as Ôdispersal syndromesÕ
(Poschlod & Bonn 1998; Ozinga et al. 2004). Second, in
any terrestrial habitat type, a variety of dispersal syndromes is
represented in the habitat species pool, and a variety of
dispersal vectors may be available in the landscape (Ozinga
et al. 2004). A decline in the availability of specific dispersal
vectors is then expected to result in a decline of those species
that, given their traits, depend on these vectors. The effect of
limited availability of dispersal vectors on plant diversity has
never been tested on large spatial and temporal scales, due to
a lack of suitable data. By using two large databases that only
recently became available, we test the hypothesis that
differences between species in population trends over the
20th century in Northwest Europe can be explained by
interspecific differences in their dispersal traits.
The first database contains information on long-term
changes in the frequency of occurrence of flowering plant
species (Angiospermophyta), and is based on repeated
floristic inventories of over 200 000 grid cells in three
countries (the Netherlands, Great Britain, and Germany)
recorded over the 20th century. The second is a recently
completed database containing quantitative information for
more than 20 key plant characteristics for over 3000 vascular
plant species in NW Europe. This combination allowed us
to compare characteristics of declining and non-declining
species. We selected the plant characteristics best able to
discriminate between two competing explanations for plant
diversity losses:
(1) Nitrogen requirement. Loss of low-productivity habitats and
eutrophication of remaining habitat patches (with
associated niche-based processes) are currently regarded
as one of the major drivers of species losses in large parts
of the world (Vitousek et al. 1997; Bobbink et al. 1998;
Grime 2001; Tilman et al. 2002; Stevens et al. 2004). For
the possible effects of changes in water and light
availability see Appendix S1 in Supporting information.
(2) Dispersal capacity. The ability to track the changes in
habitat configuration, through seed dispersal in space
(long-distance dispersal by various dispersal vectors)
and or time (formation of a persistent soil seed bank)
can be a major determinant of regional species dynamics
(Tilman 1997; Turnbull et al. 2000; Leibold et al. 2004;
Ozinga et al. 2005; Nathan 2006).
MATERIAL AND METHODS
Quantification of trends in frequency of occurrence
during the 20th century
Trends in frequency of occurrence were assessed using
published national surveys of the occurrence of vascular
plant species in grid cells (quadrats). As trend data are
sensitive to various sources of bias and to differences in
spatial and temporal scale (Telfer et al. 2002; Hartley &
Kunin 2003; Tamis 2005), we used a binary classification for
species trend: declining vs. not declining. Stochastic effects
of rarity were included in the analysis by assigning to each
species a rarity index related to its frequency of occurrence
at the beginning of the period over which the trend analysis
was performed. Rarity in itself may increase the risk of local
extinction due to random processes such as demographic
and environmental stochasticity or genetic drift (Gilpin &
Soule´ 1986; Nee & May 1997; Hubbell 2001; Tilman 2004).
The species lists from various data sources were linked using
the SynBioSys species checklist (Schamine´e et al. 2007;
http://www.synbiosys.alterra.nl/eu/). Technical details of
the survey methods of national lists of declining species
differ between countries and therefore the three countries
were analysed separately. This diversity of approaches
provides a valuable check of the generality of our results
across regions.
Letter Dispersal as key to plant diversity losses 67
2008 Blackwell Publishing Ltd/CNRS
The Netherlands
Trends during the 20th century were based on the
occurrences of plant species in the Netherlands in 1-km
2
grid cells during two periods: 1902–1949 and 1975–1998
(over 7 million records; Van der Meijden et al. 2000). The
analysis was based on a selection of 7374 grid cells with
multiple observations within the grid cell across both
periods (nearly 25% of the land surface of the Netherlands)
and corrected for temporal differences in sampling intensity
(Van der Meijden et al. 2000; Tamis 2005). Species were
labelled as declining if the number of grid cell occurrences
had declined by at least 25% over the 20th century,
representing local extinction at the 1-km
2
scale. The
historical frequency of occurrence was defined as the log
3
transformed number of grid cells occupied in the 1902–1949
period, ranging from 1 (very rare) to 9 (very common)
(Tamis 2005).
Great Britain
The list of declining species for Great Britain was based on the
change index published in New Atlas of the British and Irish
flora (Preston et al. 2002). The change index is based on the
comparison of the results of two nationwide surveys of British
plant distribution at a 10 ·10-km scale (1930–1969 and
1987–1999) and takes into account differences in recording
intensity (Telfer et al. 2002). In contrast to the change index
for the Netherlands, this index cannot be interpreted as a
percentage of change in the number of occupied grid cells, but
refers to the change in the frequency of occurrence compared
to that of an Ôaverage speciesÕ. The change indices of all species
sum to zero. In this study, species were regarded as declining if
they had a change index of )0.30 or less. The historical
frequency in the species pool was derived from the log
2
transformed number of 10 ·10-km quadrats for the 1930–
1969 period. This transformation resulted in data that better
approached a normal distribution and that are better
comparable to the Dutch data (i.e. ranging from 1 for very
rare species to 9 for very common species).
Germany
The list of declining species for Germany is based on the
trend index (tendency to decline or increase) given by
Ellenberg et al. (2001), ranging from 1 (almost disappeared)
to 9 (strongly expanding). This trend index is based on a
combination of floristic data over the 20th century and
expert judgments by several experienced botanists, and
refers to changes in species frequency in 110-km
2
quadrats
and their dominance within these quadrats. Declining
species were defined as those with a trend index £3. As
floristic inventories in the first half of the 20th century in
Germany focused on rare species only, it was not possible to
find reliable information on species frequency in the
historical species pool for all species. Therefore, we have
analysed the German data without this variable (unlike the
two other datasets).
Classification of nitrogen requirement
To compare the effects of changes in the frequency of
occurrence due to dispersal limitation with those due to
habitat change (e.g. eutrophication), we used Ellenberg
indicator values for nitrogen requirement (Ellenberg et al.
2001). These indicator values are species-specific scores,
ranging from 1 to 9, for the optimal occurrence of species
along environmental gradients. For Great Britain, we used
adjusted indicator values (Hill et al. 1999). Evidence for the
accuracy of Ellenberg indicator values has been provided by
several studies reporting a close correlation between average
indicator values and corresponding measurements of
environmental variables (see Diekmann 2003 for a review).
For the possible effects of changes in water and light
availability, see Appendix S1.
Classification of dispersal traits
Data on dispersal ability by various vectors were extracted
from the LEDA database with life-history traits of the
Northwest European flora (Knevel et al. 2003, 2005; Kleyer
et al. 2008) and adapted to a binary classification (Table 1
and 2). We considered the following dispersal vectors, all
capable of providing highly effective long-distance dispersal
(> 100 m): water, wind, the fur of large mammals, the
digestive tract of large mammals and the digestive tract of
frugivorous birds. Humans as a complex dispersal vector
were not taken into account, as this would involve various
trait syndromes, and comparative data for large sets of
species are lacking. We aggregated the available data into a
binary classification, assigning each species to one of two
classes for each dispersal vector: Ô1Õif the species has
attributes for long-distance dispersal by a given vector and
Ô0Õif the species has no such attributes (Tables 2). Although
the binary classification of the continuum is less precise for
individual species, it allows generalizations at the level of
large species pools. It is important to note that many species
have a high dispersal potential (i.e. a Ô1Õin the database) for
more than one long-distance dispersal vector (Ozinga et al.
2004). As regards dispersal through time, species were
classified as being capable of accumulating a persistent seed
bank if their seeds can remain viable in the soil for 1 year, as
indicated by a seed longevity index 0.3 (Knevel et al. 2005).
Analysis
The relative importance of nitrogen requirements and
dispersal traits for the probability of a negative trend in
frequency of occurrence was quantified for each country by
68 W. A. Ozinga et al. Letter
2008 Blackwell Publishing Ltd/CNRS
means of multiple logistic regression, which is considered an
effective method to analyse binary ecological data (McCul-
lagh & Nelder 1989; Trexler & Travis 1993). The statistical
analyses were performed using
SPSS
16.0 (SPSS Inc. 1989–
2007). Variables were tested for inclusion in the model using
the likelihood ratio test, which assesses the improvement of
the fit between the predicted and observed values of the
response variable caused by adding the predictor variable.
Only variables for which the likelihood ratio v
2
had a
P-value < 0.05 were included in the model. The relative
effect of individual variables was assessed by means of Wald
v
2
. Wald statistics and the corresponding probability are
based on the squared ratio of the unstandardized logit
coefficient to its standard error.
Excluded from the analyses were species restricted to
aquatic or alpine habitats, species that are often planted
(such as many trees), apomictic microspecies (which can be
regarded as pseudoreplicates) and a few species groups
presenting taxonomic problems (see Van der Meijden et al.
2000 and Preston et al. 2002 for details for the Netherlands
and Great Britain respectively). Actual species numbers are
listed in Table 3.
As model parameters for individual variables are
expressed as differences in logistic values, which are difficult
to interpret in an ecologically meaningful way, the effects of
dispersal traits have been illustrated graphically for the
Netherlands, the country for which the most detailed
information was available.
Table 1 Variables used in the multiple logistic regressions
Plant characteristic Classification
Frequency in historical
species pool
Number of occupied grid cells in the first recording period (log-3 transformed for the Netherlands, log-2
transformed for Great Britain): 1 (very rare) – 9 (very common)
Dispersal potential – water Potential for Long Distance Dispersal by water (0 = low, 1 = high)
Dispersal potential – wind Potential for Long Distance Dispersal by wind (0 = low, 1 = high)
Dispersal potential – fur Potential for long distance dispersal by mammalian fur (0 = low, 1 = high)
Dispersal potential – dung Potential for long distance dispersal by mammalian dung (0 = low, 1 = high)
Dispersal potential – birds Potential for long distance dispersal by bird droppings (0 = low, 1 = high)
No LDD Species with no attributes for long distance dispersal by any of the five vectors considered (0 = no, 1 = yes)
Seed longevity Persistence in the soil seed bank (0 = seeds persist in the soil < 1 year, 1 = seeds persist in the soil 1 year)
Nitrogen requirements Ellenberg indicator value for nitrogen requirements (1 = low, 9 = high)
For further details on the classification of dispersal traits see Table 2.
LDD, long-distance dispersal.
Table 2 Classification criteria for the capacity for long-distance dispersal by individual dispersal vectors
Dispersal vector Criterion
Dispersal potential –
water
Propagules float on water surface for at least 7 days
Dispersal potential –
wind
Falling velocity of propagules after a phase of acceleration (terminal velocity in m s):
< 0.5 (for species with a release height < 0.2 m); < 0.6 (release height 1 m); < 0.75 (release height 2 m)
Dispersal potential –
dung
High survival of seeds after passing through the digestive tract (at least three germinating seeds and relative
abundance in dung higher than 5% of relative abundance in the diet) and seeds frequently eaten
Dispersal potential –
fur
Propagules with awns, spiny teeth, burrs, pappus with barbs, style with barbs, hooked hairs or excreting
viscid substances
Dispersal potential –
birds
High survival of seeds after passing through the digestive tract (at least three germinating seeds and relative
abundance in dung higher than 5% of relative abundance in the diet) and morphological adaptations
to attract birds (fleshy fruit)
Table 3 Number of species included in the analysis for the three
countries (C; total number of species 1274), relative to the total
number of terrestrial, non-alpine species (A)
The
Netherlands
Great
Britain Germany
(A) Total number of terrestrial,
non-alpine species
1351 1583 2226
(B) Subset of A with
trend data
1268 (94) 1252 (79) 1851 (83)
(C) Subset of B with data on
plant characteristics
1017 (80) 841 (67) 1085 (59)
(D) Subset of C labelled
as declining
322 (32) 355 (42) 558 (51)
Values in parenthesis represent percentages.
Letter Dispersal as key to plant diversity losses 69
2008 Blackwell Publishing Ltd/CNRS
The robustness of the results was tested for the Dutch
dataset (see Appendix S1), as this dataset has been recorded
with the highest resolution and has the smallest proportion of
missing values (Table 3). A potential problem in evaluating
the importance of individual variables is that they might be
interrelated (multicollinearity). We checked for this poten-
tially confounding effect by calculating Pearson correlations
between the explanatory variables, and by performing a
combination of conditional and marginal tests for all
explanatory variables. In addition we checked for confound-
ing effects due to relationships between dispersal traits and
other environmental variables by comparing the performance
of variables in an Ôenvironmental modelÕ(including prefer-
ences for nitrogen, moisture and light), a Ôdispersal modelÕ
(excluding the environmental variables) and a full model (in
which individual variables were entered in the full model).
To check for possible confounding effects of phylo-
genetic non-independence, we performed a post hoc test of
bivariate relationships between each of the independent
variables and the species trend, using phylogenetically
independent contrasts. Technical details on these additional
tests are available in Appendix S1.
RESULTS
Overall, dispersal traits make a large and significant
contribution to explaining interspecific patterns of species
losses, of the same order of magnitude as the effect of
eutrophication (Table 4, Appendix S1). Interspecific differ-
ences in dispersal traits are thus good predictors of the
extinction risk for plant species. The results are consistent
across all three countries (Table 4).
Interaction effects were insignificant and did not change
the effect of the dispersal vectors on the risk of species
decline. The results proved to be unbiased by possible
confounding effects such as multicollinearity among vari-
ables, correlation of dispersal traits with other environ-
mental conditions and phylogenetic non-independence of
species as data points (see additional analyses in Appendix
S1). Moreover, we found that the dispersal model
performed better than the environmental model (Table S1).
The direction of the relationship between dispersal traits
and extinction risk differs between dispersal vectors. Species
with a high potential for dispersal in the fur of large
mammals or by running water are significantly more likely to
decline than those using other dispersal vectors (Fig. 1). On
the other hand, species with a high potential for dispersal by
wind or birds are less likely to decline. Remarkably, even
those species with no adaptations for long-distance dispersal
are doing better than those adapted to dispersal by water or
animal fur. The results also demonstrate that species with
the ability to accumulate a persistent soil seed bank
(Ôdispersal through timeÕ) perform relatively well (Fig. 1).
Independent from the effect of dispersal vectors is the
effect of eutrophication. Species that are adapted to nutrient-
poor conditions are over-represented among the declining
species (Table 4). Interspecific differences in the risk of a
negative population trend can thus be predicted from the
combination of nitrogen requirement of a species (indicating
risks on local extinction due to eutrophication) and adapta-
tions to various dispersal vectors. On the other hand, there is
no consistent effect of historical abundance. We expected rare
species to be more likely to decline (cf. Hubbell 2001). This is
true for the Netherlands, but the opposite is true in the UK.
Table 4 Effects of plant characteristics on the probability of decline
Variable
The Netherlands Great Britain Germany
B SE Wald v
2
Sign. B S.E. Wald v
2
Sign. B S.E. Wald v
2
Sign.
Frequency in historical species pool )0.28 0.04 57.5 < 0.001 0.42 0.06 46.3 < 0.001
Nitrogen requirements )0.36 0.04 75.6 < 0.001 )0.19 0.04 22.1 < 0.001 )0.33 0.03 108.1 < 0.001
Dispersal potential – fur 1.34 0.19 47.8 < 0.001 0.65 0.18 13.5 < 0.001 0.83 0.16 25.6 < 0.001
Dispersal potential – water 1.21 0.18 46.6 < 0.001 n.s. 0.65 0.15 18.8 < 0.001
Seed longevity )1.01 0.17 36.2 < 0.001 n.s. n.s.
Dispersal potential – birds )1.33 0.42 10.0 0.002 n.s. )0.91 0.31 8.8 0.003
Dispersal potential – wind )0.99 0.32 9.7 0.002 )0.83 0.29 8.3 0.004 )0.46 0.23 4.0 0.047
Dispersal potential – dung n.s. )0.33 0.15 4.5 0.035 n.s.
No LDD n.s. n.s. n.s.
Constant 2.36 0.30 64.0 < 0.001 )2.24 0.46 24.208 < 0.001 1.29 0.17 60.1 < 0.001
Results are given for multiple logistic regressions for three countries, with decline during the 20th century as the dependent variable and plant
characteristics as the independent variables. Wald v
2
gives an indication of the strength of the effect for individual variables. Positive values of
B indicate that the decline is increased by the given variable, negative values indicate less than average decline. Model performance as
indicated by NagelkerkeÕsR
2
: the Netherlands, 0.40; Germany, 0.20; Great Britain, 0.16. The highest performance in the Netherlands
probably reflects the more detailed information on species trends available for that country.
Sign., significance; LDD, long-distance dispersal.
70 W. A. Ozinga et al. Letter
2008 Blackwell Publishing Ltd/CNRS
Widespread nature conservation activities in the UK have
generally succeeded in slowing the decline of very rare species,
while scarce species (often not targeted by conservation
activities) are doing worse. In the Netherlands, very rare
species have generally performed worse than scarce species.
There was no correlation between historical abundance and
other explanatory variables (Appendix S1).
DISCUSSION
Our results imply that differences between species in
adaptations to various dispersal vectors are an important
but largely overlooked factor in explaining losses in plant
diversity in Northwest Europe in the 20th century, with
water- or fur-assisted dispersal being over-represented
among declining species. This is what we had expected, as
free roaming furred mammals and freely running water
almost disappeared from the Northwest European land-
scape (see Appendix S2 for historical overview). Our
analysis indicates that dispersal limitation due to a degraded
dispersal infrastructure is no less important in explaining
declines in regional plant diversity than the effects of
eutrophication and associated niche-based processes. A
possible clue for the relative independence of niche- and
dispersal-based processes is provided by the fact that, in any
terrestrial habitat type, a variety of dispersal syndromes is
represented in the habitat species pool (Ozinga et al. 2004).
The crucial process of actual seed dispersal depends on the
availability of dispersal vectors, and thus impoverishment of
the dispersal infrastructure limits the effectiveness of the
regional species pool as a seed source for local colonization.
In the emerging field of metacommunity ecology (Leibold
et al. 2004), the degree of dispersal between local commu-
nities is regarded a key parameter in the assembly and
disassembly of local communities. Metacommunity studies
traditionally focus on trait-neutral processes (cf. Hubbell
2001), but our results underscore the importance of taking
into account differences between species in their dispersal
traits. The present study thus indicates that the degree of
connectivity of local communities through seed dispersal is
not a general characteristic of local plant communities but
instead should be differentiated for individual species and
linked to the availability of the various dispersal vectors
across the landscape.
Our result that species with the ability to accumulate a
persistent soil seed bank perform relatively well (Fig. 1)
indicates that possession of a seed bank may buffer species
from local extinction. The delayed response of long-lived
plant species (above or belowground) to habitat fragmen-
tation and degradation is well known as Ôextinction debtÕand
represents a future ecological cost of present day and past
land-use changes (Tilman et al. 1994; Lindborg & Eriksson
2004; Helm et al. 2006; Ozinga et al. 2007). It can be
expected that there are also delays, following degradation of
the services provided by dispersal vectors, before species
reach a new equilibrium corresponding to the dispersal
services currently provided by the landscape. Restoration
measures in most NW European landscapes are therefore
likely to take many years to show the desired effects in terms
of (re)establishment of less-mobile species. To keep up the
metaphor of extinction debt, this time delay in (re)coloni-
zation after restoration measures can be termed Ôcoloniza-
tion deficitÕ. Our results suggest that the colonization deficit
differs between species and landscapes. On larger spatial
and temporal scales, an impoverished dispersal infrastruc-
ture may hamper the ability of many species to track
changing habitat configurations due to climate change.
For endangered species, which have become rare in the
regional species pool and have low dispersal abilities,
management at larger spatial and temporal scales at the
level of the landscape will be necessary to prevent regional
extinction.
0
10
20
30
40
50
60
70
Birds Wind Dung of
mammals
No LDD
adaptations
Water Fur of
mammals
Percentage decreasing species
Transient seed bank
Persistent seed bank
n = 44 n = 13 n = 58 n = 46 n = 127 n = 275 n = 131 n = 135 n = 125 n = 96
n = 110 n = 135
Avg. for species
with transient
seed bank
Avg. for species
with persistent
seed bank
Figure 1 Percentage of plant species declin-
ing by more than 25% over the 20th century,
for five long-distance dispersal (LDD) vec-
tors. Species are divided into two groups for
each dispersal vector, according to their
ability to accumulate a persistent soil seed
bank. Horizontal lines indicate the average
percentages (geometric means) for species
with a persistent soil seed bank (black) and
those with a transient soil seed bank (grey)
(data from the Netherlands, the country with
the most detailed information on species
decline).
Letter Dispersal as key to plant diversity losses 71
2008 Blackwell Publishing Ltd/CNRS
Options here are rehabilitation of river and brook systems
leading to more natural flooding dynamics in terms of
magnitude, frequency, duration and timing (flood-pulse
concept, cf. Tockner & Stanford 2002), which in turn may
lead to enhanced seed dispersal (see Appendix S2), especially
for species with floating seeds. Floating seeds are not
restricted to wetland plants but also occur among plant
species from dryer habitats (see Appendix S1). Although
inundations of high-elevation areas (e.g. river dunes) might
be sporadic and short, our results indicate that their impact
on colonization probabilities and hence on local species
composition is probably high.
Another option is the creation of robust ecological
networks for large mammals such as the Pan-European
Ecological Network (Council of Europe 2000, Opdam et al.
2003; Lindenmayer et al. 2008). However, for mammals with
a large home range such as Red deer (efficient vectors
for long-distance seed dispersal, see Appendix S2),
Groot Bruinderink et al. (2003) showed that even with
robust ecological networks the spatial cohesion of large
parts of Central-Europe will remain too low. From a plant
perspective, this implies that the creation of ecological
networks, although very valuable, will on its own not be
sufficient for the conservation and restoration of plant
diversity and stresses the importance of complementary
approaches.
Governments in the European Union spend roughly 35
billion a year on agri-environmental schemes that cover a
quarter of farmland in the EU (Whitfield 2006). The
challenge will be to develop new varieties of traditional
farming systems that meet modern criteria with regard to
biodiversity and socioeconomic sustainability, as todayÕs
rapidly vanishing low-intensity farming systems with herded
and or free-ranging livestock on common grounds
(cf. Bignal & McCracken 1996) will increase the rate of
seed dispersal in the landscape (see Appendix S2).
In the end, for the most endangered species, some form
of direct management at the species level may be required,
which may include conscientious but deliberate re-intro-
duction schemes.
In conclusion, our research has shown that in fragmenting
landscapes, dispersal is an underrated key process in
explaining plant diversity losses, and there is an urgent need
to face the consequences of this conclusion by designing a
different, efficient and cost-effective form of nature conser-
vation for the 21st century. Traditional habitat restoration
measures that are directed at improving local habitat quality,
although very useful, may be insufficient to halt losses in plant
diversity. Our findings clearly show that survival of sessile
plant species in fragmented landscapes requires Ômoving
corridorsÕsuch as free flowing waters, dispersing birds and
free ranging or herded large mammals. Hence, the effects on
the regional persistence of endangered vascular plant species
provided by ecological networks such as the EUÕs prestigious
and costly ÔNature 2000Õframework will critically depend on
the parallel conservation or restoration of an appropriate
infrastructure of dispersal vectors.
ACKNOWLEDGEMENTS
We thank the contributors to the LEDA database and to the
floristic surveys. Research was supported by the European
Commission (LEDA-project, Marie Curie Individual Fellow-
ship and a CNRS-ATIP funding), by the European Science
Foundation (ASSEMBLE) and by the Dutch Science Foun-
dation (NWO Stimulation Programme Biodiversity).
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SUPPORTING INFORMATION
Additional Supporting Information may be found in the
online version of this article:
Appendix S1 Check for possible confounding effects.
Appendix S2 Overview of changes in dispersal infrastruc-
ture in NW Europe.
Letter Dispersal as key to plant diversity losses 73
2008 Blackwell Publishing Ltd/CNRS
Please note: Wiley-Blackwell are not responsible for the
content or functionality of any supporting materials sup-
plied by the authors. Any queries (other than missing
material) should be directed to the corresponding author for
the article.
Editor, Marcel Rejmanek
Manuscript received 24 July 2008
First decision made 19 August 2008
Second decision made 29 September 2008
Manuscript accepted 13 October 2008
74 W. A. Ozinga et al. Letter
2008 Blackwell Publishing Ltd/CNRS
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In this publication five classes are treated: Lycopsida, Equisetopsida, Pteropsida, Pinopsida and Magnoliopsida. The present proposal replaces the 'FLORON-Rode Lijst 1990' (Gorteria 16, 1990, 2-26). It meets the requirements of the IUCN criteria. Species are classified according to a combination of rareness and decrease into eight categories: 'Not Evaluated' (NB), 'Data Deficient' (OG), 'Actually Not Threatened' (TNB; IUCN: 'Low Risk/Least Concern'), 'Near Threatened' (GE), 'Vulnerable' (KW), 'Endangered' (BE), 'Critically Endangered' (EB), and 'Disappeared' (VN; IUCN: 'Regionally Exctinct'), the latter five forming together the Red List (RL2000). For this purpose 1490 species have been considered. The new List (Rode Lijst 2000) counts 499 species (38% of all relevant species): GE: 114 (23%), KW: 136 (27%), BE: 102 (20%), EB: 90 (19%), VN: 50 (10%). RL2000 is the result of treatments of the two national flora databases FLORIVON and FLORBASE, giving a picture of the decreasing and rare part of the Dutch flora between ca. 1935 and ca. 1999. The numbers of records presented are based on the 1x1 kilometer grid of the topographical map, which internationally is comparable with a single-population size. RL2000 could largely be based on sound calculations. In comparison with the former List (RL90) 191 species changed in status: 74 are new, 117 species do not return on RL2000. After correction for differences in criteria, RL2000 is 10% longer than its predecessor. In a comment on RL2000 the species are treated according to habitat type. The most threatened habitat types are (number of species between brackets): all habitats on calcareous soil [arable fields (91%), wetlands (89%), grasslands (74%), forest edges (69%), woodlands (51%)], fen meadows (88%), heath land [wet (80%), dry (76%)], nutrient-poor waters (70%) and salt marshes (52%).
Article
During the evolution of the Central European landscape and especially since the settlement of man there has been a permanent change of processes affecting dispersability of plants. In a traditional man-made landscape there was the highest diversity of dispersal processes combined with a high diversity of land use practices. In the actual man-=made landscape most of these processes became lost or changed. Due to the rules of seed prescription many weeds became extinct, which were spread in former times with uncleaned seed. Traditional manure contained huge amounts of diaspores whereas today animal slurry with low contents of diaspores or mineral fertilizer are used. Changing harvest methods have selected the dominance of weeds which ripen later and have light diaspores. Herded and transhumant domestic livestock decreased or became locally extinct, which was probably the most important dispersal vector in the Central European man-made landscape. Artificial flooding practices favoured the migration of species in meadows of mountain and flooding practices favoured the migration of species in meadows of mountain and floodplain regions. Whereas in the traditional man-made landscape all habitats were more or less connected due to alternating management or grazing, today most habitats are isolated. With respect to restoration efforts in habitats dispersal processes or vectors should be included before planning. If there is no possibility of restoring traditional or similar dispersal processes, artificial reintroduction of species is the only option.
Article
1. The historical role of agriculture in creating semi-natural vegetation is still not fully appreciated by many ecologists, conservationists, policy-makers or the general public. Nor is the fact that for many European landscapes and biotopes of high nature conservation value, the only practicable, socially acceptable and sustainable management involves the continuation of low-intensity farming. Consequently, too much emphasis is placed on attempting to ameliorate damaging effects of agricultural management rather than supporting ecologically sustainable low-intensity farming practices. 2. More than 50% of Europe's most highly valued biotopes occur on low-intensity farmland. However, most of this farmland has no environmental policy directly affecting it; most management decisions are taken by farm businesses and determined primarily by European and national agricultural officials. As a result, there continues to be intensification or abandonment of traditional practices, changes which are equally damaging to the nature conservation value. 3. However, the nature conservation importance of low-intensity farming systems is gradually being recognized. Reforms and reviews of agriculture policy are providing a variety of potential opportunities for maintaining such systems. Unfortunately, initiating change through policy is a slow process. There is therefore also a pressing need to look for other opportunities to maintain surviving systems and, where possible, to reinstate those recently lost. 4. Although these systems may be considered low-intensity in terms of chemical inputs and productivity, they are usually high-intensity in terms of human labour. Therefore, the processes that make the low-intensity farmed countryside biologically rich and diverse must be understood, but at the same time mechanisms to make life easier and more rewarding for the people who work such farmland must be found. 5. Ecologists and conservationists should think less of 'remnants of habitat being left amongst farmland' and more of a farmland biotope for which optimum management practices need to be developed. At the same time the current emphasis on site-based conservation should be complemented by strategic initiatives that promote wise management of the wider countryside.
Article
Questions: 1. Which plant traits and habitat characteristics best explain local above-ground persistence of vascular plant species and 2. Is there a trade-off between local above-ground persistence and the ability for seed dispersal and below-ground persistence in the soil seed bank? Locations: 845 long-term permanent plots in terrestrial habitats across the Netherlands. Methods: We analysed the local above-ground persistence of vascular plants in permanent plots (monitored once a year for ca. 16 year) with respect to functional traits and habitat preferences using survival statistics (Kaplan-Meier analysis and Cox’ regression). These methods account for censored data and are rarely used in vegetation ecology. Results: Local above-ground persistence is determined by both functional traits (especially the ability to form long-lived clonal connections) and habitat preferences (especially nutrient requirements). Above-ground persistence is negatively related to the ability for dispersal by wind and to the ability to accumulate a long-term persistent soil seed bank (‘dispersal through time’) and is positively related to the ability for dispersal by water. Conclusions: Most species have a half-life expectation over 15 years, which may contribute to time lags after changes in habitat quality or -configuration (‘extinction debt’). There is evidence for a trade-off relationship between local above-ground persistence and below-ground seed persistence, while the relationship with dispersal in space is vector specific. The rate of species turnover increases with productivity.