Ecological Engineering 10 (1998) 275 – 286
Restoration of species-rich limestone grassland
communities from overgrown land: the importance
of propagule availability
Meelis Pärtel a,*, Rein Kalamees a, Martin Zobel a, Ejvind Rosén b
b
a
Institute of Botany and Ecology, Uni6ersity of Tartu, Lai 40, EE2400 Tartu, Estonia
Department of Ecological Botany, Uppsala Uni6ersity, Villa6ägen 14, S-75236 Uppsala, Sweden
Received 16 September 1997; received in revised form 22 January 1998; accepted 3 February 1998
Abstract
A field experiment was established in the eastern coast of the Baltic Sea to test the relative
roles of the availability of propagules and light competition in the restoration dynamics of
a former calcareous grassland overgrown by pine (Pinus syl6estris L.). The treatments were:
clear-cutting of trees with additional grazing and transplantation of sods from an open
grassland with all possible sources of propagules. Species richness and composition were
studied on a small-scale during 7 years in the transplanted patches of 20 × 20 cm and in their
surroundings of 50 ×50 cm. In the cut and grazed site the species richness increased.
Transplantation of sods from an open species-rich grassland did not result in higher richness
even in their closest surroundings. In the forest, transplanted patches lost their high species
richness by the second year. In the cut and grazed site, transplanted patches remained
species-rich, but after 3 years, control patches reached the same level of species richness. In
landscapes where former species-rich limestone grasslands are overgrown, but the local
species pool has not yet changed, restoration of semi-natural grassland communities does not
require the additional input of diaspores of grassland species. Transplantation of sods is
potentially important method of community restoration in case of impoverished local species
pools. © 1998 Elsevier Science B.V. All rights reserved.
Keywords: Alvar; Community management; Dispersal; Grassland restoration; Species pool;
Succession; Transplantation
* Corresponding author. Tel.: +372 7 441325; fax: + 372 7 441272; e-mail: [email protected]
0925-8574/98/$19.00 © 1998 Elsevier Science B.V. All rights reserved.
PII S0925-8574(98)00014-7
276
M. Pärtel et al. / Ecological Engineering 10 (1998) 275–286
1. Introduction
The area of species-rich semi-natural grasslands has decreased considerably
during the last decades throughout Europe (Rosén, 1982; Gibson et al., 1987;
Bakker, 1989; van Dijk, 1991; Ejrnæs and Bruun, 1995; Kiefer and Poschlod, 1996).
Traditional cultural landscapes have been considered to be especially valuable,
representing the history of environmentally friendly land-use in Europe over several
thousands of years. In most cases species-rich grasslands need continuous management — grazing of suitable intensity or mowing (Rosén, 1982; Gibson et al., 1987).
Otherwise, overgrowing by taller species like graminoids, shrubs and trees will
eliminate, through light competition, many typical grassland species resulting in
considerably lower species richness (Rosén, 1982; Kull and Zobel, 1991; Bobbink
and Willems, 1993).
A scientific nature protection study should be based on community ecology
experiments (van der Maarel and Klötzli, 1996). Experiments to study the restoration of species rich grasslands have been carried out in several countries. Most of
them deal with restoration processes on abandoned cultivated land (Gibson and
Brown, 1991a,b; Olff and Bakker, 1991; Berendse et al., 1992; Bobbink and
Willems, 1993; Hutchings and Booth, 1996; Willems and van Nieuwstadt, 1996).
Less work has been done on former grasslands which have been overgrown by
shrubs and trees forming a species-poor secondary brushwood. The latter problem
has been more actual in the calcareous grasslands of northern Europe, where most
of the clearing experiments have been carried out (Laasimer, 1981; Hæggström,
1983; Rosén, 1988; Zobel et al., 1996), but also elsewhere (Kiefer and Poschlod,
1996). Though the importance of long-term studies has been stressed (Rosén, 1995;
Bakker et al., 1996b), there is a deficit of long observation series in restored
grasslands, which is largely a result of problems with long term funding (Tilman,
1989). Despite the fact that the existing restoration experiments have shown some
success, the resulting ‘restored communities’ are mostly far from being real ‘ancient’
grasslands, the development of which will probably take centuries not only in
abandoned fields (Gibson and Brown, 1991b) but also in overgrown areas (Kiefer
and Poschlod, 1996; Zobel et al., 1996). Aerts et al. (1995) concluded that heathland
restoration requires the active introduction of Calluna 6ulgaris (L.) Hull propagules.
Restoration of a species rich grassland requires the immigration or introduction of
propagules of a large number of species. The importance of different diaspore
sources during restoration succession is mostly unknown. The studies of the seed
banks of grazed and overgrown alvar grassland communities have shown that
although the seed bank may be important for very early successional stages (Kiefer
and Poschlod, 1996) or for populations of certain species (Willems, 1988), its
overall importance as a seed source for species rich community restoration is
negligible (Bakker et al., 1996a; Bekker et al., 1997; Kalamees and Zobel, 1997).
There are very few studies of the seed rain in calcareous grasslands. Poschlod
(1993), Jackel and Poschlod (1994), Kiefer and Poschlod (1996) were able to trap
mainly the seeds of ruderal species, bushes and trees. Recently, the importance of
seed dispersal by sheep has been described (Fischer et al., 1996). Though there is no
M. Pärtel et al. / Ecological Engineering 10 (1998) 275–286
277
doubt that the results of community restoration depend a great deal on the local
species pool (Pärtel et al., 1996), the presence of a diaspore does not inevitably
mean that it will germinate and establish in particular conditions.
In Estonia, a small country on the eastern coast of the Baltic Sea, alvars—calcareous grasslands on thin soils —were quite widespread until the middle of this
century (Laasimer, 1965, 1981). After that the changes in land use were drastic—
grazing of animals ceased in many places. Today only a fraction of the grasslands
remain and only a few are still grazed. In 1990, a field experiment was established
to study the possibilities of restoring a species-rich alvar grassland from land
overgrown by Pinus syl6estris L. Clear-cutting of trees and grazing were used as
restoration treatments. The results showed that small-scale species richness increased after the improvement of light conditions (Pärtel and Zobel, 1995).
In the present study, we were interested in: (a) the changes in species richness and
species composition during the 7 years of restoration management, and (b) whether
the availability of diaspores was limiting the increase of richness and shoot density
and change of species composition. The transplantation of sods from an open
grassland to a community under restoration management was used as an experimental treatment. This means that the contribution of all possible sources of
propagules of grassland species (seed bank, rhizomes, seed rain and vegetative
dispersal from transplanted plants) were artificially increased. Also, we tried to
estimate the role of different seed sources in restoration succession from overgrown
land to open grassland.
2. Methods
The study site was located in the coastal zone of western Estonia (58°37%N,
23°35%E) in a former alvar grassland (Pärtel and Zobel, 1995). The initial grassland
community belonged to the Filipendula 6ulgaris— Trifolium montanum association
with common species such as Anthyllis 6ulneraria, Antennaria dioica, Briza media,
Festuca o6ina, Galium 6erum, Helictotrichon pratense, Hieracium pilosella, Thymus
serpyllum etc. (Laasimer, 1965). Nomenclature of species follows that of the
Estonian Flora (Eesti NSV floora, 1959–1984). The grassland had become overgrown with Pinus syl6estris since grazing ceased approximately 40 years ago. The
field layer under the trees still contained many species found in grasslands, but it
had a significantly lower shoot density and a higher herb canopy. The thickness of
the soil layer was 10 – 12 cm, below that lies a mixture of coarse limestone pebbles.
Soil pH was 6.9 – 7.2 (Kalamees, 1995). The mean standing crop of the field layer in
the overgrown area was 116 g and in the open grassland 168 g dry weight per 1 m2
(Erikson, 1993). The soil seed bank of the overgrown community contains 22
species (Kalamees and Zobel, 1997).
In 1990 a homogeneous young forest expanded in the area. Two similar adjacent
sites of approximately 20×20 m were chosen. One was left untouched while in the
other site, pines were cut and grazing by sheep was introduced (one ewe and one
lamb for 1 week twice a year). From 1994 onwards, grazing was simulated by
cutting all above ground vegetation and by disturbing the moss layer.
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M. Pärtel et al. / Ecological Engineering 10 (1998) 275–286
In 1990 four patches of sod (20 ×20 cm) from an open grassland approximately
100 m away, were transplanted into both sites. Transplanted patches and an area of
50× 50 cm around each patch were marked with sticks. A similar system of
vegetation sampling was used in the study community without transplantation: six
control patches of 20× 20 cm with surrounding areas of 50× 50 cm. All rooted
vascular plant species were recorded from each transplanted and control patch and
from their surroundings. Recordings were made in the beginning of July during 7
years, 1990 – 1996.
Differences from the initial situation (1990) were tested for each following year.
The differences between patches or their surrounds with and without transplantation during one year were checked. T-test was used for the statistics.
The percentage of species in the surrounds of the control patches which belonged
to persistent seed bank (Kalamees and Zobel, 1997), and the percentage of
wind-dispersed and animal-dispersed species (Hodgson et al., 1995; Poschlod et al.,
1996) were calculated for each year. Comparisons between manipulated and
non-manipulated sites, as well as between the initial situation and all the following
years were carried out. The study plots were compared with the local species pool
which was defined as all species in the local landscape unit (approximately 1 km2),
consisting of open alvar grasslands and overgrown grasslands of different successional stages.
3. Results
Species richness dynamics are presented in Fig. 1. In the forest, species richness
in transplanted patches was higher than in the control only in the first year,
afterwards the levels of species richness equalized. There were no differences in
species richness in the surrounds of transplanted and control patches. In the cut and
grazed area, species richness in the transplanted patches remained constant while in
the control patches a continuous increase of the richness was observed. Similar
increases in species richness were also observed in the surrounds of both transplanted and control patches.
Species which were not in the initial community in 1990 but which arrived into
the surrounds of transplanted or control patches are indicated in Table 1. Each
species, which was either present in some of the transplanted patches in 1990, or
which was present in the forest soil seed bank is identified. Similarly, the ability to
disperse by wind or by animals is also indicated for each species. Nine to ten species
arrived in the surrounds of both experimental and control patches in the forest,
regardless of whether there was a patch in the centre or not. In the cut and grazed
area, 20 and 21 species appeared, respectively.
There were more invading species in the cut and grazed site per year, than in the
forest, but the final cumulative number of species per whole series did not differ
much.
The percentage of wind-dispersed species in the local species pool was 27 and the
percentage of animal-dispersed species were 33. In Fig. 2, the mean percentage of
Transplanted
Alchemilla 6ulgaris A,
Artemisia campestris,
Campanula glomerata,
Potentilla tabernaemontani,
Vicia cracca
(Cerastium holosteoides)W
Polygala amarella A,
Hieracium praealta W
(Campanula rotundifolia)SB W,
Carlina 6ulgaris W, (Medicago
lupulina)SB, Helictotrichon
pubescens A
Carex caryophyllea SB, (Linum (Polygala comosa)A,
catharticum)A, Listera o6ata W (Trifolium repens)A
(Ranunculus polyanthemos)A
Control
Transplanted
Astragalus danicus TR SB
Thymus serpyllum TR SB
Carlina 6ulgaris W, (Medicago
lupulina)SB
(Campanula rotundifolia)SB W,
Carex caryophyllea SB A
Helictotrichon pubescens A
Hieracium praealta W, Euphrasia
officinalis W
1991
1992
1993
1994
1995
1996
Arenaria serpyllifolia SB W,
Leucanthemum 6ulgare, Linum
catharticum SB A
Hieracium pilosella SB W,
Anthyllis 6ulneraria W,
Cerastium holosteoides W,
Ranunculus polyanthemos A
Control
Hieracium praealta W, Astragalus
danicus SB, Dactylis glomerata SB,
Trifolium pratense A
Campanula rotundifolia)SB W,
Carex caryophyllea SB, (Knautia
ar6ensis)A, Ranunculus
bulbosus SB A
(Campanula rotundifolia)SB W, Trifolium repens A, Euphrasia
Astragalus danicus SB,
officinalis W, Fragaria 6iridis A,
(Trifolium pratense)TR A
Poa compressa, Polygala amarella A, Listera o6ata W
Anten naria dioica TR W,
Carex caryophyllea SB A,
Hypericum perforatum SB W,
(Leontodon autumnalis)W,
Leucanthemum 6ulgare TR,
Polygala amarella A
Hieracium pilosella TR SB W,
Anthyllis 6ulneraria TR W,
Linum catharticum TR SB A
Cut+grazed
Forest
Year of arriving
Table 1
Newly arrived species in the surrounds of transplanted and control patches of 20×20 cm in two sites under different management regimes
M. Pärtel et al. / Ecological Engineering 10 (1998) 275–286
279
47
40
—
30
50
40
10
9
22
56
44
22
37
31
30
30
30
40
42
20
22
Transplanted
Transplanted
Control
Cut+grazed
Forest
—
38
33
43
52
21
31
Control
Abbreviations: TR –species was transplanted into the particular site in 1990, SB — species was found in the local forest seed bank, W—species is
wind-dispersed, A— species is animal dispersed. Species in brackets were observed in the particular series only during 1 year. The cumulative numbers of
arrived species per site are given. Note that transplanted series had 4 and the control series 6 replicates. Percentages of different types of species dispersal
are given —since one species may use different methods to reach a site simultaneously, the sum of percentages is higher than 100%.
%TR
%SB
%W
%A
Inital number of
species per series
Number of invaded
spp.
Final number of
species per series
Year of arriving
Table 1 (Continued)
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M. Pärtel et al. / Ecological Engineering 10 (1998) 275–286
M. Pärtel et al. / Ecological Engineering 10 (1998) 275–286
281
wind and animal-dispersed species in the surrounds of the control patches is
given. In the forest, the percentage of both dispersal types was lower than the
mean for the local species pool and it did not change during the 7 years. In the
cut and grazed site, the initial percentages of wind- or animal-dispersed species
were similar to that in the forest. Following the start of restoration management,
both percentages reached the levels characteristic for the local species pool. In
the case of wind-dispersed species, this level was already achieved after the first
year while in the case of animal-dispersed species, the level characteristic of the
local species pool was reached after 3 years.
The percentage of species which had viable seeds in the local soil seed bank
was slightly lower than or equal to the percentage in the local species pool (Fig.
2). However, there were no evident differences between the forest and the cut
and grazed area. Also, there were no statistically significant differences between
years.
Fig. 1. Left side: dynamics of species richness in transplanted grassland patches of 20 ×20 cm (white
columns, n =4) and similar areas (control patches) of the original community (dark columns, n = 6)
during 7 years. Right side: dynamics of species richness in the surrounds of the transplanted patches
(50 × 50 cm, white columns) and the control patches (dark columns). Means9S.E. are given. An
asterisk above a column indicates a significant difference from the initial year (1990, t-test, P B 0.05).
Underlined year numbers indicate significant difference between transplanted and control series within
the same study site (t-test, PB 0.05).
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M. Pärtel et al. / Ecological Engineering 10 (1998) 275–286
Fig. 2. The mean percentage9 S.E. of wind- and animal-dispersed species, and the percentage of species
found in the local forest soil seed bank in the surrounds of the control patches in the forest (squares) and
in the cut and grazed site (circles) during 7 years. Significant differences between sites are indicated by
the underlining of the respective year, significant differences from the initial value (1990) are indicated
by filled symbols (PB 0.05). The dashed line indicates the corresponding percentage in the local species
pool.
When the species, which newly arrived at the study plots, are considered (Table
1), one can see a few species (Astragalus danicus, Carex caryophyllea, Dactylis
glomerata, Medicago lupulina) which are present in the seed bank but which lack
the ability to disperse effectively by wind or animals. In these cases, one would
expect that they originate from the local seed bank.
M. Pärtel et al. / Ecological Engineering 10 (1998) 275–286
283
4. Discussion
In an overgrown alvar grassland, species richness per unit area started to increase
when restoration management (cutting of trees and grazing by sheep after that)
were initiated. Rapid changes took place during the first 4 years, after that
succession slowed. The lack of diaspores was not inhibiting succession from young
pine forest to open grassland. Patches of sod from grassland, transplanted into the
forest, already lost their high species richness by the following year. Patches,
transplanted into a cut and grazed site, kept their high richness, but did not
influence even their close surroundings—species richness increased there just as
quickly as it did in control plots. The influence of a transplanted sod is evidently
stronger if the local species pool has impoverished.
The presence of diaspores may have a significant effect on the restoration of
populations of particular species and whole communities (Poschlod and Jordan,
1992; Bobbink and Willems, 1993; Jackel and Poschlod, 1994; Fischer et al., 1996;
Willems and van Nieuwstadt, 1996). The different result of the present study may
be explained by the nature of the landscape around the study sites. In the strongly
fragmented landscapes of Central Europe the increase of species richness under
restoration management may be directly dependent on the arrival of diaspores of
certain species (Jackel and Poschlod, 1996). In the Northern European countries,
where natural and seminatural landscapes spread more or less continuously over
large areas, the effect of isolation on species richness is not usually evident
(Eriksson et al., 1995). As we noticed earlier (Pärtel and Zobel, 1995), the
overgrowing of the former alvar grassland in the current study area by Pinus
syl6estris significantly reduced species richness per area, but none of the species
disappeared totally from the local species pool. Consequently, rapid reimmigration
is possible and restoration management does not require additional introduction of
diaspores.
The proportion of wind- and animal-dispersed species in the species list was
relatively lower in the overgrown area. After the cutting of trees, these proportions
increased and reached the levels in the local species pool. One may conclude that
dispersal by wind is the most rapid way for species to arrive in sites under
restoration management; dispersal by animals becomes important some years after
the clear-cutting. Most diaspores travel only short distances (Harper, 1977; Verkaar
et al., 1983; McEvoy and Cox, 1987; Poschlod et al., 1996), but the structure of the
canopy certainly has an effect on the distance of dispersal. Wind-dispersal is
expected to be more efficient in the case of an open canopy. Seed rain may also
include ruderals which do not belong to a particular (semi)natural community
(Peart, 1989; Skoglund, 1990; Poschlod and Jordan, 1992; Holzapfel et al., 1993;
Kiefer and Poschlod, 1996), but their establishment is more probable when there
are ‘free’ gaps in the field layer. In our study sites, we did not notice the
establishment of any ruderal species, the majority of the invading species were
typical for calcareous grassland communities.
Kiefer and Poschlod (1996) found that in a cleared area on an overgrown former
calcareous grassland, colonization took place primarily as a result of the germina-
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M. Pärtel et al. / Ecological Engineering 10 (1998) 275–286
tion of diaspores from the diaspore bank. We expected more species from the bank
in the grazed site, since small disturbances by animals may expose the soil surface
and activate the seed bank (Graham and Hutchings, 1988). However, our results
did not provide any evidence that the seed bank significantly contributes to the
increase in the number of species following restoration management since the
proportion of species having persistent seed bank did not differ between sites and
between years. It is probable that newly appeared individuals of certain species, like
some legumes, Carex caryophyllea and Dactylis glomerata originate from the seed
bank, but these changes have negligible influence on the level of species richness.
We have to agree with previous authors (Thompson and Grime, 1979; Bakker et
al., 1996a; Bekker et al., 1997) that limestone grassland species were rather poorly
represented in the seed bank of the overgrown grassland and therefore the seed
bank cannot be an effective source for the restoration of such grasslands.
As conclusion, the restoration of species-rich calcareous grassland community is
successful in landscapes where the local pool of appropriate species is still present.
No artificial input of diaspores will be needed. In isolated sites where the species are
not any more available, much more efforts are needed to retrieve former grassland
and the result is still doubtful. Transplantation of sods from species-rich sites can
have a positive effect in these cases.
Acknowledgements
J. Püttsepp, H. Zingel, R. Mändla, M. Suurkask and E. Möls assisted during
field work. This research was financed by the Estonian Science Foundation (Grants
2391 and 3280) and the Royal Swedish Academy of Science (for E. Rosén). We
would like to thank E. Roosaluste for critical comments on the manuscript.
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