Do forest-dwelling plant species disperse along landscapecorridors?

Jaan Liira • Taavi Paal

Received: 23 April 2012 / Accepted: 4 February 2013 / Published online: 22 February 2013

� Springer Science+Business Media Dordrecht 2013

Abstract Woody corridors in fragmented land-

scapes have been proposed as alternative habitats for

forest plants, but the great variation in species-specific

responses blurs the overall assessment. The aim of this

study was to estimate the dispersal success of forest-

dwelling plants from a stand into and along an attached

woody corridor, and to explain the observed patterns

from the point of view of species’ dispersal traits and

corridor properties. We sampled 47 forest–corridor

transects in the agricultural landscapes of southeastern

Estonia. Regionally common forest-dwelling species

(observed in at least 10 % of seed-source forests) were

classified on the basis of their ecological response

profile—forest-restricted species (F-type) and forest-

dwelling generalists (G-type). Species richness and

the proportion of F-type species decreased sharply

from the seed-source forest core to the forest edge and

to the first 10–15 m of the corridor, while G-type

species richness remained constant throughout the

transect. Corridor structure had a species-specific

effect—F species were promoted by old (C50 years)

and wide (C10 m) corridors, while G species were

supported by young and narrow corridors with ditch-

related soil disturbances. Moderate shade (canopy

cover \75 %) was optimal for all forest-dwelling

species. Large dispersule weight, and not seed weight,

dispersal vector or Ellenberg’s indicator values, was

the trait that differentiated F species from G species.

We conclude that most woody corridors are only

dispersal stepping-stone habitats for habitat generalist

species, and not for specialists. Only century old

corridors can relieve the dispersal limitation of forest-

restricted species.

Keywords Corridor habitat � Dispersal mode �Dispersal limitation � Forest plant dispersal �Niche space � Plant functional type

Introduction

The repeating clear-cut cycles, the intensified silvi-

cultural management, and particularly the conversion

of forest into arable land, have caused a decline in

forest biodiversity (Bengtsson et al. 2000; Riitters

et al. 2012). Small residual stands within agricultural

landscape matrix are prone to edge effects and

continuing habitat fragmentation (Belanger and

Grenier 2002; Harper et al. 2005; Liira et al. 2012).

The fragmentation effect is amplified by the low

dispersal ability of forest plants (Jacquemyn et al.

2003; Flinn and Vellend 2005). The dispersal limita-

tion of forest plants is explained by relatively large

J. Liira � T. Paal (&)

Institute of Ecology and Earth Sciences, University

of Tartu, Lai 40, 51005 Tartu, Estonia

e-mail: taavi.paal@ut.ee

J. Liira

e-mail: jaan.liira@ut.ee

123

Plant Ecol (2013) 214:455–470

DOI 10.1007/s11258-013-0182-1

seeds, the dominance of short-distance seed dispersal

vectors, the short period of seed dormancy, low

seedling recruitment rate, and the long pre-reproduc-

tive period (Bierzychudek 1982; Verheyen et al. 2003;

Whigham 2004). All of this contributes to a high rate

of extinction debt in fragmented forest landscapes

(Tilman et al. 1994; Vellend et al. 2006).

The creation or maintenance of woody corridors

has been proposed as a means to improve connectivity

between fragments, with the ultimate aim of preserv-

ing forest biodiversity in fragmented landscapes

(Baudry et al. 2000; Bailey 2007; Gilbert-Norton

et al. 2010). Woody corridors are defined as woody

linear features in a landscape, i.e., hedgerows, lines of

trees or alleys. Corridors’ ability to harbor forest plant

diversity has been shown to be dependent on corridor

structure and plant traits (de Blois et al. 2002; Deckers

et al. 2004; Wehling and Diekmann 2008). Less

attention has been devoted to the quantified assess-

ment of the dispersal success of plants from the seed

source forest along corridors (Corbit et al. 1999; Sitzia

2007; Wehling and Diekmann 2009). Knowledge of

corridor use by forest species is also geographically

limited, because the majority of research on plants in

woody corridors has been done in Western European

hedgerow (‘bocage’ in France) landscapes of the

temperate nemoral forest region, where hedgerows

have existed for centuries (Baudry et al. 2000;

Wehling and Diekmann 2008; Wehling and Diekmann

2009; Jamoneau et al. 2011). Classical hedgerows are

specific landscape elements with heaped-up soil under

the shrubs. Few studies have been performed con-

cerning regions in which corridors are not uplifted

landforms but are instead sometimes structured with

ditches, such as in the boreonemoral region, and to our

knowledge the relationship between drainage ditches

and the presence of forest plants in corridors has not

been explicitly studied (some aspects are discussed by

Niemela 2001; Aavik et al. 2009). Therefore, the

overall evaluation of the importance of woody corri-

dors in supporting forest biodiversity remains inde-

terminate (McCollin et al. 2000; Davies and Pullin

2007).

We contribute to the current state of knowledge of

forest plant dispersal by quantifying the efficiency of

small-scale plant dispersal into and along woody

corridors and by adding a new biogeographic dimen-

sion—the boreonemoral region. For that purpose we

surveyed a set of transects consisting of a single seed-

source forest and a target corridor surrounded by a

contrasting non-seed-source matrix habitat (arable

land). This sampling design resembles studies of

colonization fronts (Bossuyt et al. 1999; Roy and de

Blois 2006; Brunet et al. 2012), and its approach differs

from that of many other dispersal studies about forest

plants, which use landscape windows with multiple

seed-source forests per target stand (Jacquemyn

et al. 2003; Jamoneau et al. 2011). We expected to

reveal species-specific small-scale dispersal patterns

by studying short-distance dispersal (e.g., myrmec-

ochory and ballochory) and long-distance dispersal

(e.g., zoochory or anemochory). We addressed the

following questions: (i) how successful are forest-

dwelling species in their dispersal into corridors, (ii)

what is the set of critical conditions in the corridor

(habitat age and structural characteristics) required to

support the immigration of forest species, and (iii) do

plants with various emergent dispersal patterns differ

in their dispersal properties and environmental

requirements?

Methods

Study sites

Estonia is located in the hemiboreal vegetation zone

of Northern Europe, between the boreal and nemoral/

temperate forest zones (Ahti et al. 1968). Average

annual precipitation varies between 600 and

700 mm year-1, and average temperatures range

from 16.5 to 17 �C in July and from -5 to -7.5 �C

in February (Aunap 2011). The main soil types in

agricultural areas are podsols, luvisols, and various

gleysols.

There is great variability in the origin and structural

properties of woody corridors in Estonia. Corridors

structured with trees usually originate from historic

alleys planted around rural manors at the end of the

19th century, or from windbreaks planted at field

boundaries in the mid-20th century. Corridors domi-

nated by shrubs and small trees situated along drainage

ditches or stream banks are of stochastic origin. In

western and northern Estonia, hedgerows have formed

along stone fences. Many spruce hedgerows are

planted parallel to main roads as shelterbelts to

provide protection from snow. Woody corridors are

456 Plant Ecol (2013) 214:455–470

123

usually not managed, except for the trimming of

planted hedgerows along roads, occasional mowing in

alleys and cutting during the maintenance of ditches.

The study was carried out in the summer of

2009/2010 in an area of 120 9 120 km (coordinates

of central point: 58�270100, 26�2905000) in locations with

flat terrain 30–100 m a.s.l. in central and south-eastern

Estonia. The pre-selection of suitable corridors was

conducted via the Estonian Land Board’s web map

server (xgis.maaamet.ee). Study sites were selected

according to predefined criteria for transects

(Table 1). Both habitat types could contain drainage

ditches created to remove excess water in wet seasons

and to improve forest productivity in areas with flat

relief. Ditches were more frequent on forest margins

and in corridors, but were also constructed to affect the

forest core area. As the criteria were very strict, only

47 transects were found and sampled (Fig. 1).

Data collection

The survey was carried out from late May to early

August over two summers, a period when both spring

ephemerals and summer plants were visible. The

transect consisted of eight sampling locations along

each transect—in the forest core (20 m from the forest

edge), at the forest edge (at the base of the corridor),

and at six locations along the corridor (Fig. 1). In the

corridor, sampling locations were positioned with an

exponential distance step—5, 10, 15, 25, 50, and

100 m from the forest edge (on two occasions the

maximum length of the corridor was 75 m, and that of

another corridor was 50 m). An increasing step size

was selected for corridor sampling because the

probability of plant occurrence has been shown to

decrease exponentially in proportion to absolute

distance (Honnay et al. 2002; Wehling and Diekmann

2009; Brunet et al. 2012), and in analyses the

exponential distance is converted into an equivalent

step after logarithmic-transformation, as this is more

adequate for statistical analyses.

Forest vegetation (i.e., the species pool in the seed-

source forest) was sampled within a 2 9 2 m quadrat

in the forest core, and the list was completed by

recording additional species in a 10 m radius around

Table 1 Criteria for forest–corridor transect selection

Criterion Data source

1. Corridor orientation—more or less perpendicular to the forest margin Topographical map

2. Surrounding land-use type—historically open landscape, i.e. non-forest land; presently arable

field or rotational grassland

Topographical

map ? historical maps

3. Soil type—similar among forest and corridor, and suitable for boreonemoral plant species Soil maps

4. Corridor structure—shrubs and/or trees at least 2 m high and dominated by deciduous species Field survey

5. Connection point between forest and corridor—continuous closed canopy Field survey

6. Forest properties—historically continuous, minimum 1 ha; the overstory canopy (first layer) trees

have a maximum age of at least 50 years

Historical maps ? field

survey

7. Sampling of neighboring corridors—at least 150 m apart and structurally or historically

contrasting, to avoid the effects of spatial auto-correlation

Aerial photographs

Fig. 1 A map showing the locations of the sampling transects

in Estonia and the sampling design

Plant Ecol (2013) 214:455–470 457

123

the quadrat. Forest edge vegetation was sampled in a

2 9 2 m quadrat located at the attachment point of the

corridor, and the list was extended with additional

species recorded along the forest boundary in a 20 m

long and 1–2 m wide strip. In all corridor sampling

locations a 2 9 2 m sample quadrat was positioned at

the central line of the corridor, and additional species

were recorded in the direct vicinity around that quadrat

under the woody canopy (ca 1–2 m where possible).

The outer limit for sampling in the forest boundary

strip and corridor was the tree canopy’s projection on

the ground. In the presence of a drainage ditch, the

quadrat was positioned in the upper half of the ditch,

making it possible to avoid sampling on wet soil and to

evaluate the effect of soil disturbances caused by ditch

digging. In preliminary analyses we realized that the

scarcity of forest plants in 2 9 2 m quadrats did not

provide adequate information about their presence/

absence in each section of a transect, and therefore in

later analyses we use only the extended list of species

(quadrat ? additional species). The plant nomencla-

ture follows Leht (2010).

The structural characteristics recorded along each

transect were: canopy cover, corridor width, signs of

disturbance (i.e., cutting of trees, small-scale soil

disturbances) and the presence of a ditch (large-scale

soil disturbance). The canopy cover of woody plants

was estimated visually for forest, edge and as an

average per corridor. The historic continuity of shade

conditions (i.e., the age of woody structures), was

estimated using historical maps and aerial photos, and

was re-evaluated in the field. Map data for the periods

1888–1913, 1907–1939, 1945–1952, 1978–1989,

1990–2000, and 2003-present day were provided by

the Estonian Land Board. The GIS-based estimate of

corridor age was adjusted during the field work using

tree diameter or rings on the cut trunks of the largest

trees. The correction was only made upward if the

tree-based estimate suggested the habitat to be older

than the GIS-estimated age. In the analyses we used a

threshold of 50 years to distinguish between young

and old corridors, because it has been shown that ca

50 years is a critical age for forest species to be able to

populate new forest patches (e.g., Brunet et al. 2011).

Species classification

Species were classified hierarchically in two steps

(Appendix A). First, species were categorized as

common forest-dwelling species (the main group of

interest) and common corridor/open-land species (as a

control group for scaling plant trait values). Species

observed in at least 10 % of the sampled forests were

defined as common forest species in the region

(cf. Aavik et al. 2009; Liira et al. 2012). Second,

within the common forest-dwelling species, grouping

as forest-restricted and generalist species was done

according to species distribution profile along tran-

sects. The distribution profiles were estimated by

fitting regression models to data on the species’

occurrence frequency (percentage) versus log-trans-

formed distance. In these models, the linear and

second-order polynomial trend-line was tested. Spe-

cies were classified accordingly: ‘‘forest-restricted

species’’ or ‘‘species of F response type’’—maximum

occurrence frequency in forests (statistically signifi-

cant negative linear or polynomial regression trend),

‘‘corridor species’’ or ‘‘C species’’—species with

increasing frequency in corridors (the linear slope

estimate is positive or the polynomial profile shows an

increase), and ‘‘U species’’—plants having a high

frequency both in the forest and also in the distant part

of the transect (the polynomial term has a significant

positive parameter estimate, and the profile increases

in both directions, assuming bird-dispersal or other

types of zoochory). The remaining common forest

species that were frequent along the transect but

lacked an apparent distribution pattern (i.e., the

statistical regression model was non-significant) were

defined as ‘‘generalist species s.s.’’ However, in a

broad sense, C and U species can also be considered to

be forest-dwelling generalists, and can be pooled with

the last group into a joint type called ‘‘generalist

species s.l.’’ or ‘‘species of G response type’’. The

dispersal ecology of G species can be generalized as

shade tolerant species with good dispersal ability and/

or species with broad ecological niches, but they

definitely stand in contrast to F species in their ecology

due to their use of corridor habitats. Also, in the case of

G species, dispersal direction, whether from forest to

corridor or vice versa, cannot be uniquely defined.

Among residual species that were not defined as

common forest species, we extracted species observed

in at least 10 % of corridors and classified them as

common open-land species (O species). Later we used

them as the control group.

Information on plant traits was gathered via

Internet databases and different published studies

458 Plant Ecol (2013) 214:455–470

123

(Appendix B). Species environmental requirements

(average of realized niche) were described in terms of

Ellenberg’s ecological indicator values (the score on a

relative scale from 1 to 9 describing the realized niche

centroid for a species estimated for Central Europe;

Ellenberg et al. 1991). The weight of seeds and

dispersules (seeds with special soft covers and addi-

tional structures) and dispersal vectors were used to

describe dispersal properties.

Data analysis

The difference in traits between F, G, and O response

types was compared using generalized linear models

(GLZ) and general linear models (GLM). The weights

of seeds and dispersules were compared between

response types using the GLZ with settings of the log

link function and gamma distribution of the model’s

residuals. The GLZ with binomial error-distribution

and logit link function were used in the pair-wise

analysis of dispersal vectors within growth form.

Differences in Ellenberg’s ecological indicator values

were analyzed using general linear models (GLM; i.e.,

the normal error-distribution). In all these analyses,

plant growth form (graminoid or other forb; excluding

horsetails and ferns) was considered to be a con-

founding factor. Ferns and horsetails were excluded,

because both are wind-dispersed (anemochorous)

with spores. Both analyses were performed in Statis-

tica ver. 9.

Differences in species richness between transect

sections (forest, edge, 5 m, 10 m, etc.) were analyzed

using the mixed model (SAS ver. 9.2), considering the

autocorrelation within transect and response groups.

For that purpose we included two random factors in the

model: the transect as a random subject within the

response type, and second within the transect section

with the type AR(1) setting (spatial autocorrelation

intensity with autoregressive change) for the covari-

ance matrix (Littell et al. 1996). As the response-type-

specific trends were of special interest to us, interaction

terms with response type were included in the model.

Homogeneity groups between factor levels were tested

using the Tukey HSD test in Statistica ver. 9.

An additional mixed model was built to estimate

the effect of corridor structural characteristics on the

species richness of the response groups, using the data

from only the corridor (distance sections 5–100 m).

Structural data, which were entered in the model as

independent factors, were carefully selected to avoid

collinearity. To simplify the interpretation, all struc-

tural characteristics were analyzed as two-level cate-

gorical predictors. Corridor width was classified as

narrow (\10 m) or wide (C10 m), canopy cover as

open (closure\75 %) or closed (closure C75 %), and

corridor age as young (\50 years) or old (C50 years).

Results

The dominant tree species in the surveyed forest–

corridor transects were Alnus incana, Betula pendula,

Salix spp, and Tilia cordata, and in the understorey

A. incana, Salix spp, Prunus padus, Picea abies, and

Sorbus aucuparia. Drainage ditches were present in 32

and roads in 14 corridors. The historical age of the

corridors ranged from 15 to more than 110 years

(corridors with woody features present on the earliest

detailed maps available), and 16 corridors had been

constantly present for more than 50 years. The width

of the corridors varied between 3.5 and 20 m, and

by thresholding at 10 m we obtained 33 narrow and

14 wide corridors.

We recorded 272 herbaceous species. The average

number of species per transect was 61, ranging from

19 to 103 species. The most frequent plant species per

transect were Rubus idaeus (100 % of transects),

Taraxacum officinale agg. (98 %), Deschampsia

cespitosa (96 %), Fragaria vesca (96 %), Paris

quadrifolia (96 %), Dryopteris carthusiana (94 %),

Urtica dioica (94 %), and Dactylis glomerata (91 %).

Of the 272 species, 89 were classified as common forest-

dwelling species (Appendix C) and 40 as common

open–land (corridor) species (O species, Appendix D).

The distance-based frequency count profile con-

structed for each of the 89 common forest-dwelling

species resulted in 70 statistically significant and 19

non-significant regression models. According to the

trends revealed in the models, we classified these 89

species into four dispersal response types (Appendix

C, Fig. 2): (i) 60 common forest species as the F

response type (species with maximum occurrence

frequency in forests, e.g., Oxalis acetosella, D.

carthusiana and P. quadrifolia), (ii) five corridor

species that were also common in forests as the C

response type (increasing frequency in corridors, e.g.,

T. officinale agg., D. cespitosa and D. glomerata), (iii)

Plant Ecol (2013) 214:455–470 459

123

five species with a U-shaped pattern (high frequency

in the forest and in the distant part of the transect, e.g.,

Equisetum pratense, Convallaria majalis and Scirpus

sylvaticus), and (iv) 19 species as habitat generalists

s.s. (frequent along the transect without an apparent

distribution pattern, e.g., Anthriscus sylvestris, Veron-

ica chamaedrys, and Filipendula ulmaria). The low

count of C and U species was insufficient for statistical

analysis, and therefore in later analyses we used the

pooled G response type (C, U, and generalist s.s.

together), because their ecology is comparable within

the limits of our study question.

The generalized linear model (GLZ) test results for

seed and dispersule weight revealed a significant pattern

between dispersule weight and species dispersal response

types (Fig. 3; Table 2). F species had a much greater

average dispersule weight than G and O species. Unex-

pectedly, seed weight did not statistically differ between

response types. However, average estimates of both traits

were greater in forbs than in graminoids (Table 2).

Graminoids and forbs possessed a variety of

dispersal modes (Table 3), but almost none of these

had a uniform differentiation between the F and G

species when growth form was considered. The only

significant differences were revealed as a higher

proportion of anemochorous and hydrochorous forbs

within G species than in F species (respective contrast

tests: 20 % of F species vs. 50 % G species; p = 0.026;

and 11 % F species vs. 38 % G species; p = 0.03).

Our comparison of Ellenberg’s indicator values

showed statistically significant differences only for the

indicator value of light requirements. Increased

requirements were observed for O species, positioning

F species and G species in the same homogeneity

group (p \ 0.0001; Table 2; Fig. 4). The indicator

value for soil moisture did not differ between response

types, but only showed a larger requirement for

moisture among graminoids than forbs (p = 0.006;

Table 2; Fig. 4). The non-significant differences

between species’ environmental requirements for soil

(a) (b) (c)

Fig. 2 Species profiles for the three species groups: a common

forest species for which occurrence frequency declines from the

forest along the corridor—the F species (n = 60), b common

forest species with an increasing or U-shaped trend of

occurrence frequency along the transect, or frequent species in

all steps along the corridor without a significant response

pattern—the G species (n = 29), c species that have a low

occurrence frequency in forests, but are frequent in corridors—

the O species (n = 58)

Fig. 3 The seed and dispersule weights of different species

groups (mean ± 95 % confidence intervals). F species—forest-

restricted species, G species—generalist species, O species—

open-habitat species common in corridors. The letters indicate

statistical homogeneity groups. The letter range is unique to

each model

460 Plant Ecol (2013) 214:455–470

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conditions provide supporting proof of adequate

transect selection, avoiding confounding trends.

The species richness of common forest-dwelling

species (F ? G species) decreased from forest to

boundary to corridor along the transect. The average

number of species in seed-source forests was 33 (29

F ? G species), at forest boundaries 27 (22) and in

corridors 17 (13). More than half of F species (57 %;

n = 34) were recorded in at least 10 % of corridors.

From forest to corridor, we detected a significant

response-type-specific pattern of species richness

along the distance (the main effect of distance classes

and the interaction term with dispersal response

groups in the mixed model; p \ 0.0001). The signif-

icant interaction term indicated that the species

richness of F-type species declined sharply from the

forest (21 species) to the boundary (13) and to 5 m,

and then remained almost constant throughout the

corridor (6), while the species richness of the G-type

did not differ between distance classes (in average 7

species; Fig. 5). In contrast, the average number of O

species per transect was 3.

In proportional units, the percent of F species

within common forest-dwelling species (F ? G-type)

decreased linearly along the transect, from 74 % in the

forest to 40 % at 50 and 100 m (Fig. 5; p = 0.0001).

The results of the mixed model analysis regarding

only species richness data from the corridor part

Table 2 The p values for analyses used to compare dispersal

traits (dispersule weight and seed weight with GLZ analysis:

the log link and gamma distribution of residuals) and

Ellenberg’s ecological indicator values (GLM analysis: the

normal distribution) between F, G, and O response types with

growth form as confounding factor (levels: graminoid vs. other

forb; horsetails and ferns excluded)

Growth form (contrast) Response type

df = 1 df = 2

Log dispersule weight 0.007 (graminoid \ forb) 0.001

Log seed weight 0.029 (graminoid \ forb) n.s.

Ellenberg’s light value n.s. \0.0001

Ellenberg’s soil moisture value 0.006 (graminoid [ forb) n.s.

Ellenberg’s soil fertility value n.s. n.s.

The average estimates for the response types are shown in Fig. 3 and 4

df degrees of freedom

Fig. 4 Average Ellenberg’s indicator values for F, G, and

O-type species (mean ± 95 % confidence intervals). The lettersindicate statistical homogeneity groups. The letter range is

unique to each model

Fig. 5 Distance profiles for the number of species of F and G

response types (mean ± 95 % confidence intervals) analyzed

within the mixed model (Table 4), and the proportion of F

species within common forest-dwelling species. The lettersdenote statistical homogeneity groups between distance classes

based on the results of post hoc comparisons using the Tukey

HSD multiple comparison test, and the letter range is unique for

each model

Plant Ecol (2013) 214:455–470 461

123

(5–100 m) showed only a few significant effects of

corridor structure (Table 4). We found that the

richness of both response types was uniformly reduced

in corridors with an intense tree and/or shrub canopy

(canopy closure C75 %; Fig. 6). The presence of a

ditch had a significant effect on species richness, but in

pair-wise comparison the supportive effect was only

evident for G species, while F species were unaffected

(Fig. 6, although the interaction term was non-signif-

icant; Table 4). Corridor width had a significant

response-type-specific effect on richness (the

interaction term was significant). G species richness

was enhanced in narrow (\10 m) corridors, while the

species richness of the F response type was marginally

independent of corridor width, even if it showed

higher values in wider corridors. Corridor age had an

effect on both response types, but the effect was

reciprocal—the species richness of the G response

type was greater in younger corridors (\50 years), and

the richness of F-type species was greater in older

corridors (C50 years). The interactions between dis-

tance along the corridor and corridor structural

characteristics were non-significant in the model,

indicating that none of the structural characteristics

enhanced species dispersal along corridors (Table 4).

Discussion

Different studies have addressed the question of the

suitability of corridor habitats for forest-dwelling

plants (McCollin et al. 2000; Deckers et al. 2004;

Davies and Pullin 2007). We recorded more than half

of the forest-restricted species (F-type) in at least 10 %

of corridors, which points to their potential for

dispersal from seed to source forests. The rate of

dispersal pressure, expressed in terms of the propor-

tion of forest-dwelling species found over corridors,

seems particularly high in the pooling of F and G

species. Contrary to the dispersal potential from

source forests, the establishment rate of F species

was quite low, i.e., six species per corridor from the

Table 3 The observed frequency of dispersal modes (%) for the dispersal response groups within growth forms

Growth form Ferns and horsetails (%) Graminoids (%) Forbs (%)

Response type F-type G-type F-type G-type F-type G-type

Anemochory 100 100 40 20 20* 50*

Myrmecochory – – 40 – 25 11

Endozoochory – – 20 40 36 39

Epizoochory – – 20 20 16 11

Ballochory – – – – 16 6

Unspecified – – 60 60 30 39

Hydrochory – – 40 20 11* 33*

No. of species 6 1 10 10 44 18

The percentage shows the frequency of a certain dispersal vector type in a given species group. Please note that the same species can

have several dispersal vectors suggested in the literature, and therefore percentages add up to [100 within the column

* Statistical differences between response types within a growth form

Fig. 6 Structural characteristic classes of corridors affecting

the number of species of the F and G response types based on the

results of the mixed model (Table 4). The letters denote

statistical homogeneity groups and structure classes based on

the results of post hoc comparisons using the Tukey HSD

multiple comparison test, and the letter range is unique for each

factor. Error bars denote 95 % confidence intervals

462 Plant Ecol (2013) 214:455–470

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average number of species available in the source

forest (21 species).

The observed patterns fit within the observed range

of other regions (Corbit et al. 1999; Roy and de Blois

2006; Roy and de Blois 2008; Wehling and Diekmann

2009). The broad overlap in species richness between

our study and various other published studies shows

that the overall conclusions are independent of the

study-specific definition of forest specialist species,

whether the classification is made by experts (Corbit

et al. 1999; Roy and de Blois 2008), using Ellenberg’s

ecological indicator values for light (McCollin et al.

2000; Sitzia 2007) or an occurrence-based two-step-

classification-algorithm is applied, as in our case. Our

method was expected to extract the regional forest

species pool, which has a sufficient frequency of

species for statistical analyses, and is largely inde-

pendent of the subjectivity of experts or the choice of

traits (Aavik et al. 2009; Liira et al. 2012). The

resulting list of forest-dwelling species was in general

agreement with earlier studies, with some variations—

e.g., F and G species did not differ in their Ellenberg’s

indicator values, and some ancient forest species

occurred to be habitat generalists.

The classification inconsistency between studies

hints at the incomplete knowledge of the true ecolog-

ical profile of forest species, i.e., the underestimation

or biased estimation of niche breadth (Hill et al. 2000;

Lohmus and Kull 2011; Paal et al. 2012). The

underestimation of habitat niche breadth can easily

Table 4 The result of the

mixed model analysis test

regarding the effect of the

structural characteristics of

corridors on the species

richness of the F and G

response types

Only data from within the

corridor (5–100 m) were

used. Transect-related

autocorrelation among

response types and along

the transect were defined

with the covariation matrix

within the random factor

df degrees of freedom in the

model

Fixed factor df factor;

df error

F

statistic

p value

Distance 5; 209 1.03 n.s.

Canopy cover 1; 209 6.46 0.012

Ditch 1; 209 6.18 0.012

Width 1; 209 0.72 n.s.

Age 1; 209 3.00 n.s.

Distance 9 canopy cover 5; 209 1.35 n.s.

Distance 9 ditch 5; 209 0.15 n.s.

Distance 9 width 5; 209 0.34 n.s.

Distance 9 age 5; 209 0.59 n.s.

Response type 1; 84 0.11 n.s.

Response type 9 distance 5; 209 6.51 \0.0001

Response type 9 canopy cover 1; 209 0.07 n.s.

Response type 9 ditch 1; 209 0.76 n.s.

Response type 9 width 1; 209 5.03 0.026

Response type 9 age 1; 209 5.31 0.022

Response type 9 distance 9

canopy cover

5; 209 0.26 n.s.

Response type 9 distance 9 ditch 5; 209 0.65 n.s.

Response type 9 distance 9 width 5; 209 1.12 n.s.

Response type 9 distance 9 age 5; 209 1.67 n.s.

Covariance tests Z p

Covariation between response

types

3.72 0.0001

Transect-specific covariation 3.37 0.0007

AR(1)-type covariation between

distance classes

9.85 \0.0001

Plant Ecol (2013) 214:455–470 463

123

occur in regions where specific habitat types, e.g.,

semi-open woodlands, have became lost. As a result,

when one applies the biased classification in hetero-

geneous landscapes, niche underestimation will affect

the conclusions (Paal et al. 2012; Vojta and Drhovska

2012). We therefore suggest that the quantification of

species’ niche breadth should be undertaken in

conditions where species have had equal probabilities

to disperse into all potential conditions, whether along

a single environmental gradient or over habitat types.

In agreement with some studies, we revealed the

extreme dispersal limitation of forest-restricted spe-

cies (F species) along corridors (Corbit et al. 1999;

Honnay et al. 2002; Sitzia 2007; Wehling and

Diekmann 2009), and the fact that forest generalists

already became dominant among common forest-

dwelling species at a short distance from the forest

edge (ca 10–15 m). However, the dispersal direction

of G species between forest and corridor is indeter-

minate, because they can be common in forests simply

because of the immigration into disturbance- or gap-

dynamic-related microhabitats (Brunet et al. 2011;

Jamoneau et al. 2011; Liira et al. 2012). The dispersal

limitation of forest species has been explained via

specific dispersal-related traits, predominantly prop-

erties associated with seed size, such as short seed

bank survival, short dormancy period, and dispersal

vector type (Hermy et al. 1999; Graae 2000; Roy and

de Blois 2006). Dispersal trait specificity was partially

confirmed by our study, but only in terms of dispersule

weight, contrary to widely tested dispersal vector or

seed weight. The latter ones were mostly constrained

to plant growth form, i.e., evolutionarily restricted.

Larger dispersule weight distinguished F species from

G or O species. In ecological terms, dispersule weight

combines information about seeds’ supplementary

structures (i.e., dispersal aiding pappi, wings, awns,

fruity coats), and is related to dispersal type. In that

regard, the usefulness of dispersule weight is largely

underestimated. This might also explain the observed

discrepancy in the role of short and long-distance

dispersal vectors in species dispersal success on the

landscape scale (Jacquemyn et al. 2003; Brunet 2007;

Liira et al. 2012).

The success of seedling establishment has also been

shown to be limiting in corridors (Schmucki and de

Blois 2009), but we found that micro-habitats with soil

disturbance at ditch banks were only exploited by G

species. This indicates that the establishment of forest-

restricted species is not limited by the availability of

regeneration gaps with lower competition intensity.

Therefore, the use of forest-dwelling species as one

large group containing both F and G-type species

would leave a falsely optimistic message about forest

species’ efficacy in using corridors.

We found that over-intensive shade from trees and

shrubs causes uniform resource limitation for all

forest-dwelling species, independent of their response

type (Liira and Sepp 2009; Van Couwenberghe et al.

2011; Liira et al. 2012). The effect of corridor width is

frequently associated with light availability because of

the edge effect (Sitzia 2007; Roy and de Blois 2008;

Wehling and Diekmann 2008), but this is not always

the case (Corbit et al. 1999) and also our corridors

were usually closed at the margins. The observed

negative effect of shade intensity also conflicts with

the statement about the resemblance of closed-canopy

corridors to forest interiors (Roy and de Blois 2008;

Wehling and Diekmann 2008). We would instead

explain variation between studies through the defini-

tion of forest-dwellers, i.e., the pooling of G and F

species into a single group can have a unpredictable

effect on the overall relationship.

Habitat age is commonly seen as a factor supporting

forest species richness in newly-formed forest frag-

ments and corridors (Jacquemyn et al. 2003; Flinn and

Vellend 2005; Roy and de Blois 2008), even if its

effect has not always been evident (Vellend et al.

2007; Wehling and Diekmann 2009; Baeten et al.

2010). We did, however, find that the long-term

continuity of shady corridors had an effect related to

response type—only the richness of F species

increased with habitat age, while G species were

successful only in younger corridors. The observed

contrasting responses of F and G species on corridor

age suggests that corridors with long historical con-

tinuity will obtain specific qualities which is difficult

to define, but somehow these support forest-restricted

species but do not support the accumulation of

generalist species.

The results of our survey imply that woody

corridors cannot be considered to be primary alterna-

tive habitats for forest-dwelling plant species. Forest-

dwelling plants are able to disperse out from forests,

but dispersal along corridors is stochastic and has a

low success rate. The highest richness of forest

specialists can be expected in old and wide corridors

with intermediate canopy closure, but in practice these

464 Plant Ecol (2013) 214:455–470

123

conditions are easier to find in forest patches (Brunet

2007; Liira et al. 2012; Vojta and Drhovska 2012).

Therefore, small and large forest patches should be

preferred in landscape planning for forest biodiversity

and the dispersal limitation of habitat-restricted spe-

cies is mostly relieved by the sufficiently long duration

of stable habitat conditions. We would like to point out

that the ecology of forest-dwelling species in woody

corridors should be assessed with care, considering

their fundamental niche breadth, availability in the

local and regional species pool, and there must also be

awareness of the potential reciprocal relationships

among response types.

Acknowledgments This project was supported by the

University of Tartu (SF0180012s09), the Estonian Science

Foundation as national grant ETF7878 and through the

BiodivERsA project smallFOREST, and the European Union

through the European Regional Development Fund (the FIBIR

Centre of Excellence). Alexander Harding proofread the text.

Appendix A

The hierarchical classification tree of species. Capital

letters denote the response types: F-type—species

preferring forest habitat, U-type—species with max-

imum occurrence in forests and in the end of the

transect, C-type—species preferring a corridor habitat,

G-types.s—generalist species sensu stricto, G-types.l—

generalist species sensu lato, and O-type—open-land

species.

Appendix B

List of databases and studies used for gathering

information about seed and dispersule weight and

dispersal types. aStudies/databases for seed and/or

dispersule weight and bstudies/databases for dispersal

type.

aAustrheim G, Evju M, Mysterud A (2005) Herb

abundance and life-history traits in two contrasting

alpine habitats in southern Norway. Plant Ecol

179:217–229.bCampbell JE, Gibson DJ (2001) The effect of seeds

of exotic species transported via horse dung on

vegetation along trail corridors. Plant Ecol

157:23–35.aCosyns E, Claerbout S, Lamoot I, Hoffmann M

(2005) Endozoochorous seed dispersal by cattle and

horse in a spatially heterogeneous landscape. Plant

Ecol 178:149–162.a,bCouvreur M, Vandenberghe B, Verheyen K,

Hermy M (2004) An experimental assessment of

seed adhesivity on animal furs. Seed Sci Res

14:147–159.bDavy AAJ (1980) Deschampsia caespitosa (L.)

Beauv. J Ecol 68:1075–1096.bDupre C, Ehrlen J (2002) Habitat configuration,

species traits and plant distributions. J Ecol

90:796–805.bDzwonko Z, Loster S (1992) Species richness and

seed dispersal to secondary woods in southern

Poland. J Biogeogr 19:195–204.aEriksson O, Ehrlen J (1991) Phenological variation

in fruit characteristics in vertebrate-dispersed

plants. Oecologia 86:463–470.bFischer SF, Poschlod P, Beinlich B (1996) Exper-

imental studies on the dispersal of plants and

animals on sheep in calcareous grasslands. J Appl

Ecol 33:1206–1222.aFitter AH, Peat HJ (1994) The ecological flora

database. J Ecol 82:415–425.aFroborg H (2001) Seed size and seedling emer-

gence in 16 temperate forest herbs and one dwarf-

shrub. Nord J Bot 21:373–384.aGorb SN, Gorb EV (1995) Removal rates of seeds

of five myrmecochorous plants by the ant Formica

polyctena (Hymenoptera: Formicidae). Oikos

73:367–374.a,bGrime JP, Hodgson JG, Hunt R (1988) Compar-

ative plant ecology: a functional approach to

common British species. Allen and Unwin, London.aGrime JP, Mason G, Curtis AV, Rodman J, Band

SR, Mowforth MAG, Neal AM, Shaw S (1981) A

comparative study of germination characteristics in

a local flora. J Ecol 69:1017–1059.bHerault B, Honnay O (2005) The relative impor-

tance of local, regional and historical factors

Plant Ecol (2013) 214:455–470 465

123

determining the distribution of plants in fragmented

riverine forests: an emergent group approach.

J Biogeogr 32:2069–2081.bJongejans E, Telenius A (2001) Field experiments

on seed dispersal by wind in ten umbelliferous

species (Apiaceae). Plant Ecol 152:67–78.aLindborg R (2007) Evaluating the distribution of

plant life-history traits in relation to current and

historical landscape configurations. J Ecol 95:555–564.bPakeman RJ, Digneffe G, Small JL (2002) Eco-

logical correlates of endozoochory by herbivores.

Funct Ecol 16:296–304.aPoschlod P, Kleyer M, Jackel A-K, Dannemann A,

Tackenberg O (2003) Biopop—a database of plant

traits and internet application for nature conserva-

tion. Folia Geobot 38:263–271.bSoukupova L (1992) Calamagrostis canescens:

population biology of a clonal grass invading

wetlands. Oikos 63:395–401.aSera B (2005) Diaspores—potential or real power

of wild plants? Life cycle. Ekologia (Bratislava)

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Trientalis europaea L. J Ecol 90:404–418.

Appendix C

Species profiles for the common forest species

Common forest species n_Site n_Forest Beta Beta p R2 Resp. type

logDist logDist2

Actaea spicata 19 16 -2.72 2.15 0.007 0.93 F

Aegopodium podagraria 41 30 -0.86 0.014 0.73 F

Anemone nemorosa 27 23 -2.37 1.61 0.004 0.97 F

Angelica sylvestris 34 17 n.s. G

Anthriscus sylvestris 41 18 n.s. G

Asarum europaeum 5 5 -2.36 1.69 0.041 0.87 F

Athyrium filix-femina 33 33 -2.52 1.85 0.010 0.93 F

Calamagrostis arundinacea 12 10 -2.56 2.03 0.038 0.82 F

Calamagrostis canescens 10 5 -2.79 2.74 0.019 0.79 U

Campanula persicifolia 16 7 n.s. G

Carex cespitosa 16 6 n.s. G

Carex digitata 11 11 -2.46 1.82 0.032 0.87 F

Carex elongata 10 10 -2.55 1.89 0.009 0.94 F

Carex pallescens 8 7 -0.81 0.029 0.65 F

Carex sylvatica 9 8 -2.27 1.48 0.006 0.97 F

Carex vaginata 7 5 -0.80 0.029 0.65 F

Chrysosplenium alternifolium 8 7 -0.93 0.002 0.86 F

Circaea alpina 6 6 -2.55 1.89 0.009 0.94 F

Cirsium oleraceum 25 24 -2.07 1.25 0.033 0.94 F

Convallaria majalis 16 12 -3.03 2.70 0.002 0.95 U

Crepis paludosa 31 27 -2.44 1.74 0.013 0.93 F

Dactylis glomerata 42 13 0.91 0.004 0.83 C

Deschampsia cespitosa 45 28 -2.42 2.82 0.008 0.87 C

Dryopteris carthusiana 44 43 -2.41 1.66 0.005 0.97 F

Dryopteris expansa 5 5 -2.55 1.89 0.009 0.94 F

Dryopteris filix-mas 36 35 -2.67 2.03 0.002 0.97 F

Elymus caninus 14 11 -2.59 2.15 0.045 0.78 F

Epilobium angustifolium 15 11 -2.61 1.98 0.009 0.93 F

Epilobium montanum agg. 28 18 n.s. G

466 Plant Ecol (2013) 214:455–470

123

Table a continued

Common forest species n_Site n_Forest Beta Beta p R2 Resp. type

logDist logDist2

Equisetum pratense 36 23 -2.73 2.86 0.012 0.82 U

Equisetum sylvaticum 11 7 -2.75 2.24 0.013 0.89 F

Festuca gigantea 15 10 n.s. G

Festuca rubra 28 5 n.s. G

Filipendula ulmaria 37 30 n.s. G

Fragaria vesca 44 39 -2.76 2.27 0.014 0.88 F

Galeobdolon luteum 10 9 -0.78 0.038 0.61 F

Galeopsis tetrahit 15 7 -2.96 2.91 0.005 0.89 U

Galium album 30 11 n.s. G

Galium palustre 8 7 -2.68 2.07 0.005 0.94 F

Galium uliginosum 8 7 -0.82 0.025 0.67 F

Geranium palustre 21 5 0.81 0.027 0.66 C

Geranium sylvaticum 7 6 -2.66 2.24 0.036 0.79 F

Geum rivale 39 34 -2.24 1.47 0.025 0.93 F

Geum urbanum 41 31 -2.56 1.94 0.018 0.90 F

Gymnocarpium dryopteris 15 15 -2.55 1.89 0.009 0.94 F

Hepatica nobilis 6 5 -0.80 0.032 0.63 F

Impatiens noli-tangere 7 6 -2.66 2.09 0.013 0.90 F

Impatiens parviflora 31 29 -0.88 0.010 0.77 F

Juncus effusus 13 9 -2.42 1.70 0.007 0.95 F

Luzula pilosa 23 20 -2.58 1.95 0.011 0.93 F

Lysimachia vulgaris 37 25 n.s. G

Maianthemum bifolium 28 24 -2.57 1.95 0.015 0.91 F

Melampyrum nemorosum 24 8 -2.79 2.52 0.024 0.80 F

Melica nutans 13 12 -2.23 1.46 0.021 0.94 F

Mercurialis perennis 7 6 -2.84 2.30 0.002 0.96 F

Moehringia trinervia 31 15 n.s. G

Mycelis muralis 29 28 -2.52 1.85 0.011 0.93 F

Myosotis scorpioides 12 10 n.s. G

Oxalis acetosella 41 40 -2.36 1.59 0.001 0.99 F

Paris quadrifolia 44 42 -2.48 1.85 0.028 0.88 F

Poa nemoralis 18 7 n.s. G

Poa palustris 15 5 -2.05 2.53 0.031 0.76 C

Poa trivialis 21 6 n.s. G

Prunella vulgaris 11 6 n.s. G

Pyrola rotundifolia 11 7 n.s. G

Ranunculus acris 24 5 n.s. G

Ranunculus auricomus 16 6 n.s. G

Ranunculus cassubicus 15 11 -0.80 0.032 0.63 F

Ranunculus repens 35 21 -2.77 2.23 0.006 0.93 F

Rubus idaeus 47 42 -0.85 0.015 0.72 F

Rubus saxatilis 29 23 -2.60 1.97 0.010 0.92 F

Scirpus sylvaticus 16 7 -2.62 2.73 0.026 0.75 U

Scrophularia nodosa 21 9 -2.56 2.08 0.048 0.78 F

Plant Ecol (2013) 214:455–470 467

123

Appendix D

Open-habitat (O species) species used in the study

Open-habitat species Number of sites

Achillea millefolium 15

Agrostis capillaris 14

Alchemilla sp 17

Alopecurus pratensis 14

Anthoxanthum odoratum 7

Arctium tomentosum 14

Artemisia vulgaris 20

Calamagrostis epigeios 16

Calamagrostis purpurea 5

Caltha palustris 7

Campanula glomerata 10

Campanula latifolia 5

Campanula patula 13

Carex acutiformis 5

Carex flava 6

Carex hirta 22

Carex nigra 5

Carex panicea 9

Appendix continued

Open-habitat species Number of sites

Carex spicata 6

Cirsium arvense 23

Cirsium palustre 5

Elymus repens 32

Epilobium hirsutum 7

Epipactis helleborine 6

Equisetum arvense 8

Equisetum palustre 6

Festuca ovina 5

Galium aparine 9

Galium boreale 16

Galium mollugo 5

Glechoma hederacea 11

Helictotrichon pubescens 5

Hieracium umbellatum 14

Hypericum maculatum 17

Knautia arvensis 9

Lapsana communis 8

Lathyrus pratensis 17

Leucanthemum vulgare 12

Table a continued

Common forest species n_Site n_Forest Beta Beta p R2 Resp. type

logDist logDist2

Scutellaria galericulata 19 10 -2.57 2.03 0.033 0.84 F

Solidago virgaurea 21 17 -2.77 2.24 0.007 0.92 F

Stachys sylvatica 4 4 -2.55 1.89 0.009 0.94 F

Stellaria holostea 7 6 -2.77 2.20 0.003 0.95 F

Stellaria media 34 15 -0.88 0.008 0.78 F

Stellaria nemorum 6 6 -0.87 0.012 0.75 F

Taraxacum officinale agg. 46 20 -0.85 1.75 0.006 0.96 C

Trientalis europaea 15 15 -2.54 1.85 0.004 0.96 F

Urtica dioica 43 30 -0.87 0.012 0.75 F

Vaccinium myrtillus 13 13 -2.50 1.82 0.012 0.93 F

Valeriana officinalis 31 18 -0.78 0.037 0.62 F

Veronica chamaedrys 41 28 n.s. G

Veronica officinalis 10 8 -2.46 1.81 0.028 0.88 F

Viola epipsila 5 5 -2.57 1.98 0.024 0.87 F

Viola mirabilis 15 14 -2.39 1.63 0.004 0.97 F

Viola riviniana 16 15 -2.49 1.85 0.025 0.89 F

n_Site number of species occurrences in the whole transect (forest, boundary, corridor), n_Forest number of species occurrences in

the forest part of transects, Beta standardized slope estimate, logDist log-transformed distance, p p value of the regression model, R2

coefficient of determination for a model, Resp. type resulting definition of species (F species preferring forest habitat, G generalist

species s.s., U species with maximum occurrence in forests and in the end of the transect, C species preferring a corridor habitat)

468 Plant Ecol (2013) 214:455–470

123

Appendix continued

Open-habitat species Number of sites

Luzula multiflora 5

Lychnis flos-cuculi 8

Lycopus europaeus 5

Mentha arvensis 5

Phalaris arundinacea 13

Phegopteris connectilis 5

Phleum pratense 16

Phragmites australis 8

Pilosella officinarum 13

Poa angustifolia 5

Poa pratensis 24

Potentilla erecta 9

Ranunculus fallax 11

Rumex acetosa 7

Trifolium medium 10

Trollius europaeus 6

Tussilago farfara 7

Vicia cracca 13

Vicia sepium 20

Viola canina 19

Number of sites shows the number of transects, where species

was observed

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