This article appeared in a journal published by Elsevier. The attached
copy is furnished to the author for internal non-commercial research
and education use, including for instruction at the authors institution
and sharing with colleagues.
Other uses, including reproduction and distribution, or selling or
licensing copies, or posting to personal, institutional or third party
websites are prohibited.
In most cases authors are permitted to post their version of the
article (e.g. in Word or Tex form) to their personal website or
institutional repository. Authors requiring further information
regarding Elsevier’s archiving and manuscript policies are
encouraged to visit:
Author's personal copy
Landscape and Urban Planning 97 (2010) 92–97
Contents lists available at ScienceDirect
Landscape and Urban Planning
journal homepage: www.elsevier.com/locate/landurbplan
Evaluating a vegetation-recovery plan in Mediterranean alpine ski slopes:
A chronosequence-based study in Sierra Nevada (SE Spain)
, Mercedes Molina-Morales
, Eva María Ca
, Julio Pe
Departamento de Botánica, Facultad de Ciencias, Universidad de Granada, Avda. Fuentenueva s/n, E-18071, Granada, Spain
Departmento de Biología Animal, Facultad de Ciencias, Universidad de Granada, Avda. Fuentenueva s/n, E-18071, Granada, Spain
Received 17 July 2009
Received in revised form 27 April 2010
Accepted 30 April 2010
In this paper, we assess the results found in the restoration of vegetation on ski runs in the Mediterranean
high mountain, contrasting different issues widely used for evaluating recovery plans, such as cover, rich-
ness, diversity, growth, and qualitative species composition, with the aim of establishing their relative
validity as well as ﬁnding a straightforward model to assess the success of the restoration of degraded
areas. Ski runs were selected in Sierra Nevada ski station (SE Spain) in which hydroseeding was performed
from 2002 to 2005. The sampling design was based on a chronosequence approach, using natural areas
established as ‘models’ (i.e. target for long-term restoration) to evaluate the restoration success based
on the similarity to the model. Although parameters such as growth, cover, and even richness or diver-
sity reached similar values to the ones in the model areas after 4 years (i.e. natural perennial mountain
pastures), other indicators such as composition, measured in a qualitative way as the ratio of colonizing
species to total species, showed different occurrence values for the most abundant species. Moreover,
when the whole pool of species was taken into account using discriminant analysis, the results differed,
showing that although the process performed well, the recovery (sensu stricto) requires longer periods
than the duration assessed to be fully successful. The results showed that common parameters, such as
growth, cover, richness, or diversity, when used solely may lead to misinterpretation, and therefore addi-
tional methods to compare composition, such as the discriminant analysis, are strongly recommended.
© 2010 Elsevier B.V. All rights reserved.
Since the mid-twentieth century, Mediterranean mountains
have undergone a vigorous increase in tourist activities, especially
those related to alpine skiing (Lasanta et al., 2007). Ski resorts
yield large economic returns (Elsasser and Messerli, 2001) and pro-
vide services and infrastructure improvements for mountain areas
(Snowdon et al., 2000) but on the other hand inﬂict heavy impact
on the environment and landscape (e.g. Benthem, 1973; Pignatti,
1993; Needham and Rollins, 2005). Speciﬁcally, the construction
of ski runs is one of the major causes of pastureland loss in the
Alps (Watson, 1985; Urbanska, 1997). To build the ski runs, all
the vegetation is eliminated, the upper soil layers are disturbed,
stones are removed and the topography is modiﬁed, resulting in
the complete elimination of the vegetation and the general distur-
bance of the ecosystem (Gros et al., 2004; Barni et al., 2007). The
Corresponding author. Tel.: +34 958 241000x20223; fax: +34 958 243254.
E-mail addresses: firstname.lastname@example.org (J. Lorite), email@example.com (M. Molina-Morales),
firstname.lastname@example.org (E.M. Ca
nadas), email@example.com (M. Ballesteros).
processes of primary succession of the vegetation are extremely
slowed down (Körner, 2003) due to the adverse high-mountain
conditions. Moreover, maintenance activity speciﬁc to ski resorts,
such as the machine-grading of ski slopes and artiﬁcial snowing,
can strongly limit the recovery of the vegetation. The machine-
grading (i.e. the complete removal of topsoil and vegetation, as
well as the leveled of the surface) causes severe and lasting impact
on alpine vegetation, which proves difﬁcult to mitigate even by
recovery measures, particularly at higher altitudes (Wipf et al.,
2005). The impact of artiﬁcial snowing depends on the current
state of the vegetation, the environmental objectives of a spe-
ciﬁc ski resort, and the properties of the snow, being the sum of
these usually negative, but not always (Rixen et al., 2008; Wipf
et al., 2005). As a result of the management, extensive areas of soil
remain bare in these environments (Tsuyuzaki, 1995; Urbanska and
Fattorini, 2000). Thus, restoration is vital to prevent erosive pro-
cesses as well as to restore ecosystem structure and functionality
(Muller et al., 1998). Restoration should minimize resource deple-
tion and ensure long-term ecosystem recovery, which in terms
of ecological restoration is called rehabilitation (e.g. Bradshaw,
0169-2046/$ – see front matter © 2010 Elsevier B.V. All rights reserved.
Author's personal copy
J. Lorite et al. / Landscape and Urban Planning 97 (2010) 92–97 93
Details of the hydroseeding seed mixture.
Species Seed mass % purity gg(m
) Seeds (m
Agrostis nevadensis 0.320 0.200 30.0 0.42 8 1.6
Arenaria tetraquetra subsp. amabilis 0.193 0.610 90.0 0.02 5 1
Dactylis glomerata subsp. juncinella 0.280 0.470 47.5 1.11 88 17.8
Deschampsia ﬂexuosa 0.120 0.180 25.0 3.18 119 24.1
Festuca indigesta 0.750 0.910 63.0 1.69 129 26.1
Helianthemum apenninum subsp. apenninum 0.670 0.950 75.0 0.25 26 5.3
Hormathophylla spinosa 0.762 0.490 86.0 0.2 11 2.2
Reseda complicata 0.041 0.398 80.0 0.12 93 18.9
Sideritis glacialis 1.218 0.420 57.0 0.12 2 0.4
Thymus serpylloides 0.700 0.210 75.0 0.47 11 2.2
Seed mass, weight per 100 seeds; % purity, percentage of purity; g, % of seed germination; g (m
), weight of seeds applied per square meter; seeds (m
), number of viable
seed applied per square meter; %, percentage of seed belonging to one species in the mixture (data provided by technical environmental staff of CETURSA Sierra Nevada S.A.).
Some studies have pointed out the changes in plant-species
composition, dynamics, biomass, etc. caused by the building of
ski runs and later restoration using the different techniques (e.g.
Watson, 1985; Urbanska, 1995, 1997; Tsuyuzaki, 1990, 1995, 2002;
Muller et al., 1998; Urbanska and Fattorini, 2000), but all of them
have focused on alpine ecosystems sensu stricto. Restoration of
Mediterranean high-mountain ecosystems have not yet been stud-
ied despite that there are major differences with respect to Alpine
mountains (Grabherr et al., 2003) due to climatic irregularity and
the existence of two stress periods for plants: winter, owing to low
temperatures; and summer, due to the typical Mediterranean sum-
mer drought (Giménez-Benavides et al., 2007). The present study
was conducted in Sierra Nevada (SE Spain), a good example of a
Mediterranean mountain with a ski station (the southernmost in
This study assesses the results of plant restoration works on
ski runs under Mediterranean conditions, comparing results from
a chronosequence and evaluating different issues frequently used
as indicators, such as cover, richness, diversity, growth, and species
composition. The aim was to establish a methodology to assess the
success of the vegetation recovery in ski runs and other disturbed
2. Materials and methods
2.1. Study area
The study area is located at the ski station of Sierra Nevada (SE
W), in the Sierra Nevada mountain. This mountain
area covers about 2100 km
included in the Baetic range. Sierra
Nevada is a major Mediterranean diversity hotspot (Médail and
Quézel, 1997, 1999); in fact, above 2000 m asl occur nearly 100
endemic and rare taxa (Lorite et al., 2007a) as well as rare and/or
endemic plant communities (Lorite et al., 2007b). Geologically,
the study site is constituted of siliceous rocks, mainly micaschists
(Jabaloy et al., 2008). The soils belong to the suborders Orthents
and Ochrepts, the soil units being Anthropic Regosols and Humic
Regosols, with sandy loam or loamy sand containing a high pro-
portion of gravel texture (Delgado et al., 2007). In general, the
climate is typically Mediterranean although with high-mountain
features. The average annual rainfall of the area is 925 mm with
mainly snow-based precipitations and an extended period of sum-
mer drought, while the mean annual temperature is 3
C, hottest month: 14
C; lowest: −14.6
C) (for further information see Gómez, 2002). The ski station,
built in 1964, is provided with 115 ski runs and 102.88 skiable kilo-
meters, ranging between 2100 and 3300 m asl. In years with high
rainfall, the average snow depth exceeds 2 m at most ski slopes dur-
ing the ski season, although in some areas more than 5 m can be
reached, in contrast to dry years, when the snow is almost absent.
Therefore, the ski resort produces artiﬁcial snow when necessary.
The ﬁrst snow machines were installed in 1989, and at present day
produce artiﬁcial snow for 32 km of snow runs (source: Cetursa
Sierra Nevada S.A.; www.cetursa.es).
2.2. Ski-run restoration
The ski-run construction at the site dates back to the early 1960s,
for which the topsoil (10–40 cm of depth, according Delgado et al.,
2007) was removed and discarded, exposing the underlying min-
eral soil and bedrock. The ﬁrst restoration works date from 2000
(Gálvez, pers. com.), although not until 2002 did the restoration
program begin in strict terms. Since then the ﬁrm that manages
the ski runs has undertaken sowing by hydroseeding (see further
details on the technique in Merlin et al., 1999), in all cases dur-
ing the autumn (throughout October) using the same mixture of
autochthonous seeds every sowing period (see Table 1), harvested
in the surroundings of the ski runs (see Table 1). The assessed period
in this study corresponds to the hydroseedings performed over 4
years: 2002 (ca. 1.5 ha), 2003 (ca. 2.8 ha), 2004 (ca. 3.1 ha), and 2005
(ca. 1.9 ha).
2.3. Sampling design
Field samples were carried out during the summer of 2006. A
chronosequence was established for this purpose (Kent and Coker,
1992; Foster and Tilman, 2000), including the areas seeded from
2002 to 2005 (H02, H03, H04, H05, hereafter) (Fig. 1). The plant
community we sought to restore was a native Mediterranean-
alpine perennial pasture of Festuca indigesta. It is also composed
of others species such as Deschampsia ﬂexuosa, Koeleria crassipes,
Fig. 1. Conceptual scheme established for the monitoring of the chronosequence.
On the x-axis, time (in years) from the recovery action (note that 0 corresponds to
the evaluating time in 2006). On the y-axis, evaluation issues settled to evaluate the
recovery success (see Section 2 for further information).
Author's personal copy
94 J. Lorite et al. / Landscape and Urban Planning 97 (2010) 92–97
Results of growth estimation as in Dactylis glomerata (A), Festuca indigesta (B), and Reseda complicata (C).
Species Chronosequence Model
H05 H04 H03 H02 Mo
H 2.34 ± 0.14 c 3.38 ± 0.18 c 2.82 ± 0.16 c 5.40 ± 0.32 b 10.07 ± 0.69 a
D 3.41 ± 0.25 d 8.15 ± 0.41 c 8.72 ± 0.51 c 15.79 ± 0.93 b 23.58 ± 1.48 a
B 39.20 ± 11.70 b 266.97 ± 37.50 b 248.42 ± 37.60 b 1729.17 ± 379.40 b 7606.90 ± 1390.80 a
H 2.11 ± 0.12 d 3.65 ± 0.19 c 4.03 ± 0.17 c 5.44 ± 0.25 b 9.94 ± 0.48 a
D 3.13 ± 0.19 d 8.53 ± 0.44 c 9.84 ± 0.51 c 12.57 ± 0.42 b 20.27 ± 1.19 a
B 25.33 ± 4.12 b 328.81 ± 64.73 b 461.53 ± 67.85 b 928.49 ± 90.70 b 5167.18 ± 919.20 a
H 2.18 ± 0.12 d 13.33 ± 0.99 c 17.96 ± 1.57 c 36.43 ± 2.06 b 56.62 ± 2.70 a
D 1.74 ± 0.13 d 22.45 ± 1.76 c 23.99 ± 1.82 c 43.84 ± 1.99 b 69.06 ± 2.80 a
B 9.00 ± 1.86 c 3214.66 ± 614.00 c 5196.50
± 1593.70 c 37067.14 ± 5922.57 b 132726.69 ± 16192.70 a
H, height (cm); D, diameter (cm); B, biovolume (cm
). Different letters in the same line indicate signiﬁcant differences in post hoc Tukey–Cramer test at p < 0.05. n =30inall
S. glacialis, Avenula laevis, Thymus serpylloides, Arenaria tetraquetra
subsp. amabilis, and Jasione amethystina. This community lives in
a mosaic with juniper-genista patches being dominant in higher-
slope areas under the same ecological conditions of hydroseeded
areas (see Lorite, 2002 for further information on this community
type). This community is widely distributed in the study area and
fulﬁlls the objectives and structure (i.e. plant architecture of the
community) required for ski-run maintenance and skiing. Thus, this
natural plant community was established as a model (Mo, here-
after) to evaluate the restoration success, comparing this model to
the chronosequence of the restored areas using hydroseeding. With
this aim, a number of evaluation issues were used, such as cover,
growth of key plant species, richness, diversity, and composition
(see Fig. 1). A stratiﬁed sampling was made to guarantee the repre-
sentation. We deﬁned the following non-overlapping strata: all the
chronosequence stages (H02–H05) and the model areas (Mo), three
slope ranges: I: 0–30%, II: 30–60%, III: >60%, and two exposures: I:
and II: 180–360
. The samples were randomly placed for
We performed 147 linear transects of 25 m with three con-
tact points per meter, 75 contacts per transect in total, for which
the number of occurrences of perennial species was recorded
(chamaephytes, hemicryptophytes, and geophytes). Afterwards
every species was classiﬁed by its behavior in colonizing and non-
colonizing species, following Molero and Pérez Raya (1987) and
Lorite et al. (2007c). To estimate the size of three key species (F.
indigesta, Dactylis glomerata subsp. juncinella, and Reseda complicata
(see Table 1)), we randomly collected 150 individuals (30 individu-
als × 5 chronosequence stages). For each individual, we measured
height and diameter, after calculating the volume of the semi-
spheroid form (volume of semi-spheroid = ((4/3)r
h)/2); where r
is radius and h is height), given the cushion-like or low tussock habit
of the species. The nomenclature of the species followed Blanca et
2.4. Data analysis
The raw data matrix was used to compute the cover, size of key
species, richness, diversity (Shannon–Wiener index; see Magurran,
1988) and frequency of plant colonizing species (calculated as the
ratio colonizing species to total number of species encountered
per transect). The data for the different variables were compared
throughout the chronosequence using a one-way ANOVA (normal-
ity was checked by the Shapiro–Wilk test, and homoscedasticity
by the Bartlett test). To test differences in frequency of colonizing
species between the model areas (perennial Mediterranean-alpine
pastures of F. indigesta) and the hydroseeded areas, the Wilcoxon
non-parametric test was used.
In addition, to compare the ﬂoristic composition of the different
samples, a discriminant analysis was used (see Hair et al., 1998).
This is a powerful technique when the variable to discriminate is
nominal with multiple levels (chronosequence in our case: H02,
H03, H04, H05, and Mo) and the independent variables (species) are
metric (number of individuals). All the statistical analyses were per-
formed using JMP 6.0 (SAS Institute). Throughout the text, means
are followed by ±SE.
3.1. Key species growth, cover, richness, and diversity
Data concerning the size of key species in terms of biovolume
(Table 2) showed that, for the three key species studied, there were
signiﬁcant differences between the 4 years of hydroseeding and
the model areas, in height (H), diameter (D), and biovolume (B);
only in the older seedings did the data approximate the model
areas, although there were signiﬁcant differences in all cases (see
Table 2). In Reseda complicata, size was faster than in Dactylis glom-
erata subsp. juncinella or F. indigesta, and therefore we could expect
individuals, in less time, to reach a size similar to that found in the
With regard to the total cover (Table 3), chronosequence H02
reached similar values to those of the areas Mo (39.45 ± 1.89 vs.
41.71 ± 3.07), even for the case H04 (34.00 ± 2.66), which led us to
deduce that the cover was a parameter with easy recovery.
For richness, H03 and H02 showed no signiﬁcant differences
compared with Mo. Diversity showed even fewer signiﬁcant dif-
ferences than did richness, and only H05 had signiﬁcantly lower
diversity (1.17 ± 0.14) compared to the rest of the chronosequence
Results of the one-way ANOVA comparing: cover, richness, diversity and ratio of colonizing species to total species encountered in the chronosequence.
Chronosequence n Cover Richness Diversity Ratio col.:tot.
H05 13 18.87 ± 2.27 b 5.14 ± 0.61 c 1.17 ± 0.14 b 0.641304 c
H04 26 34.00 ± 2.66 ab 7.23 ± 0.35 bc 1.54 ± 0.07 ab 0.292415 b
H03 14 24.57 ± 2.73 b 7.89 ± 0.57 abc 1.70 ± 0.12 a 0.286822 b
H02 41 39.45 ± 1.89 a 8.17 ± 0.33 ab 1.64 ± 0.054 a 0.325763 b
Mo 53 41.71 ± 3.07 a 8.80 ± 0.38 a 1.71 ± 0.046 a 0.120024 a
Different letters in the same line indicate signiﬁcant differences in post hoc Tukey–Cramer test at p < 0.05.
Author's personal copy
J. Lorite et al. / Landscape and Urban Planning 97 (2010) 92–97 95
Data on life form, distribution, and abundance of the most abundant species sampled.
(n = 53) H
(n = 94) p-Value
Dactylis glomerata subsp. juncinella (Bory) Stebbins & Zohary 0 Hc N s.l. 779 1.58 ± 0.54 6.83 ± 0.62 <.0001
Festuca indigesta Boiss. 0 Hc N-A 608 4.64 ± 1.13 3.74 ± 0.39 0.0057
Spergularia rubra (L.) J. Presl & C. Presl 1 Th/Hc W 295 0.90 ± 0.23 2.45 ± 0.51 0.2597
Reseda complicata Bory 0 Ch N s.l. 202 0.77 ± 0.29 1.60 ± 0.24 0.0068
Thymus serpylloides Bory 0 Ch N s.l. 183 2.85 ± 0.71 0.45 ± 0.10 0.0042
Agrostis nevadensis Boiss. 0 Hc N-A 142 1.44 ± 0.47 0.71 ± 0.10 0.2608
Jasione amethystina Lag. & Rodr. 0 Ch N s.l. 124 1.98 ± 0.36 0.28 ± 0.11 <.0001
Artemisia absinthium L. 1 Ch W 121 0.00 ± 0.0 1.17 ± 0.27 0.0002
Polygonum aviculare L. 1 Th/Hc W 106 0.083 ± 0.05 0.99 ± 0.22 0.0008
Carex nigra (L.) Reichard 0 Hc W 88 1.82 ± 0.93 0.00 ± 0.00 0.0031
Arenaria tetraquetra subsp. amabilis (Bory) H. Lindb. ﬁl. 0 Ch N s.l. 83 1.52 ± 0.32 0.10 ± 0.04 <.0001
Pilosella pseudopilosella (Ten.) Soják 1 Hc W 78 0.40 ± 0.16 0.57 ± 0.14 0.5031
Carduus carlinoides subsp. hispanicus (Kazmi) Franco 1 Hc N s.l. 76 0.33 ± 0.12 0.58 ± 0.16 0.3470
Herniaria boissieri Gay 1 Ch N s.l. 64 0.44 ± 0.13 0.42 ± 0.10 0.4517
Genista versicolor Boiss. 0 Ch N s.l. 61 0.83 ± 0.42 0.20 ± 0.06 0.3895
Paronychia polygonifolia (Vill.) DC. 1 Hc WM 48 0.42 ± 0.23 0.27 ±
Linaria aeruginea subsp. nevadensis (Boiss.) Malag. 0 Hc N s.l. 46 0.19 ± 0.09 0.36 ± 0.07 0.0385
Festuca pseudoeskia Boiss. 0 Hc N s.l. 45 0.87 ± 0.3 0.03 ± 0.02 <.0001
Rumex angiocarpus Murb. 1 Hc WM 44 0.10 ± 0.09 0.38 ± 0.1 0.0149
Lotus corniculatus subsp. glacialis (Boiss.) Valdés 0 Hc B 44 0.81 ± 0.26 0.05 ± 0.03 <.0001
Festuca clementei Boiss. 0 Hc SN 42 0.08 ± 0.07 0.37 ± 0.12 0.1018
Eryngium glaciale Boiss. 1 Hc N-A 40 0.50 ± 0.18 0.16 ± 0.05 0.0776
Plantago holosteum Scop. 1 Ch B 32 0.27 ± 0.12 0.18 ± 0.06 0.9281
Sideritis glacialis Boiss. 0 Ch B 29 0.52 ± 0.22 0.04 ± 0.02 0.0149
Arenaria armerina Bory 0 Ch WM 29 0.54 ± 0.19 0.03 ± 0.02 0.0002
Dianthus brachyanthus Boiss. 0 Ch WM 28 0.52 ± 0.23 0.03 ± 0.02 0.0007
Anthyllis vulneraria subsp. pseudoarundana H. Lindb. 0 Hc N s.l. 26 0.42 ± 0.14 0.06 ± 0.02 0.0005
Trisetum glaciale (Bory) Boiss. 0 Hc SN 25 0.52 ± 0.17 0.00 ± 0.00 <.0001
Species with more than 25 contacts in all transects.
Behavior of the species (1, colonizing; 0, non-colonizing).
Biotype (Ph, Phanerophyte; Ch, Chamaephyte; Hc, Hemicryptophyte; Th, Therophyte).
Distribution of the species (SN, Sierra Nevada; N s.l., Sierra Nevada, Sierra de Baza and Sierra de los Filabres; B, Baetic Mountains (Sierra Nevada and other nearby
mountains); N-A, Nevadense and North Africa; WM, Western Mediterranean; W, wide).
Number of contacts per species in 147 transects.
Average number of contacts per transects of the species in model areas.
Average number of contacts per transects of the species in hydroseeded areas.
p-Value from non-parametric Wilcoxon tests (values of p < 0.05 in bold).
3.2. Composition-based parameters
The ratio of colonizing species to total species signiﬁcantly dif-
fered among all the hydroseedings performed (H05, H04, H03, and
H02) compared with Mo (Table 3). Furthermore, we found that H04,
H03, and H02 did not signiﬁcantly differ, while the ratio decreased
the ﬁrst year (H05) after the hydroseeding and later remained con-
stant (H04, H03, and H02).
When the areas Mo (n = 53) and the hydroseeding areas (n = 94)
were analyzed as two groups, with respect to the most abun-
dant species (n total of contacts >25), signiﬁcant differences were
detected (see Table 4). Three groups of species were established,
according to their performance: Group 1: signiﬁcantly more abun-
dant species in the model area, such as F. indigesta, Thymus
serpylloides, Jasione amethystina, Carex nigra, Arenaria tetraquetra
subsp. amabilis, F. pseudoeskia, Lotus corniculatus subsp. glacialis,
Sideritis glacialis, Arenaria armerina subsp. armerina, Dianthus
brachyanthus, Anthyllis vulneraria subsp. pseudoarundana, and Trise-
tum glaciale. Group 2: the most abundant species on the ski runs.
The cases of Artemisia absithium or Polygonum aviculare were
notable for their abundance on the ski-run area (1.17 ± 0.27 and
0.99 ± 0.22 contacts per transect, respectively) but absent or scarce
in the model area. Group 3: species not showing signiﬁcant differ-
ences between the hydroseeding areas and the models. This group
included species that frequently appear in pastures but seem to
adapt successfully in order to colonize open and altered areas, such
as Eryngium glaciale, Pilosella pseudopilosella or Spergularia rubra.
Lastly, results from discriminant analysis, show signiﬁcant dif-
ferences between the three groups of samples formed (see Fig. 2):
(1) H05, (2) Mo, and (3) H02, H03, and H04, this last group con-
Fig. 2. Biplot showing the results from discriminant analysis, with ellipsoids of
probability of 95% superimposed. Mo, H02, H03, H04 and H05 are the stages of
the chronosequence (see Section 2 for further information).
Author's personal copy
96 J. Lorite et al. / Landscape and Urban Planning 97 (2010) 92–97
sists of three quite overlapping choronosequences (see probability
ellipsoids of 95% in Fig. 2). These results show that frequency and
abundance of plant species signiﬁcantly differed between Group 3
(H02, H03 and H04) and both Mo areas as well as for those more
recently seeded (H05).
4. Discussion and conclusions
Results from the hydroseeding were quite good, showing the
importance of sowing-adapted species, as pointed out by other
authors (e.g. Quezel, 1977; Urbanska, 1997; Muller et al., 1998). The
cover of the pastures as well as the one of the restored area was rela-
tively low compared with other alpine pastures of temperate areas
(i.e. 50–60% in Japan; Tsuyuzaki, 1995, 2002). This illustrates the
comparatively harsher conditions of the Mediterranean mountain
due to the additional effect of the summer drought (Giménez-
Benavides et al., 2007).
According to most authors (e.g. Gómez-Campo and Malato-
Beliz, 1985; Sainz Ollero and Hernández Bermejo, 1985; Blanca
and Molero, 1990; Heywood, 1996; Blanca et al., 1998), the main
objective for the restoration of areas with high levels of diversity,
such as the one studied, is to encourage the recovery of the most
natural and original-like ecosystem. In most of the studies deal-
ing with ski-run restoration (e.g. Tsuyuzaki, 1990, 1995; Urbanska,
1997; Grismer et al., 2008), the most frequently used parameters
to assess success are: cover, richness, diversity, and size, while
only a few papers include other methods to evaluate species com-
position (e.g. Sluis, 2002). In our study, the cover, richness, and
diversity results reﬂect that 3–4 years after the restoration, or, as
in the size case in a 5–6 years term, the recovery of the area can
be achieved. The results found by other authors in studies compar-
ing ski runs and native communities indicate the opposite, with
degraded areas showing low values of cover, richness, and diver-
sity (Wipf et al., 2005), although the areas surveyed, in most cases,
have not been subjected to a recovery plan. According to our results,
if we examine other parameters that appraise composition, using
either a straightforward approach such as the ratio of colonizing
species to total species or comparisons between the most abun-
dant species, or else using a more complex method of counting all
the species present and using a discriminant analysis, we found
a completely different outcome. Although the recovery appeared
to be performing well, the process was far from being concluded,
with a larger richness and abundance of colonizing species than in
natural pasturelands. This ﬁnding matches the common idea that
colonization and succession processes in the high mountain are
very slow (see Körner, 2003 for a revision), especially where the
uppermost soil layer has been removed (Urbanska, 1995; Muller
et al., 1998; Gros et al., 2004). Consequently, our study site should
be considered rehabilitated rather than restored (sensu Bradshaw,
Therefore, we conclude that the parameters for evaluating the
success of this kind of restoration, such as cover, richness, diver-
sity, etc., are useful only for assessing the ability of a recovery plan
to slow down erosive processes and improve soil structure and
composition (e.g. Grismer et al., 2008). Although these parameters,
particularly species richness and diversity, have been suggested as
an indicator of a perceived community or system quality (Bowles
and Jones, 1999; Woodward et al., 1999), the assemblage of species
composition may be equally or more important in judging restora-
tion and management success (Henderson, 1999), taking into
account that the main target of recovery plans should be to restore
the native-plant-community structure and composition. As some
authors have pointed out (e.g. Sluis, 2002), the maintenance of
species richness and mechanisms causing species patterns are not
sufﬁcient to allow the re-creation of patterns of species found
in remnant grassland communities. Our study aims in the same
direction, and thus we conclude that the above-mentioned param-
eters should never be used solely to assess the recovery stage of
the vegetation but jointly with others that take into consideration
the assemblage of species in comparison with a selected recovery
model. In this sense, the discriminant-analysis technique, which,
as far as we know has not previously been used in similar studies,
provides a useful tool for fairly straightforward and powerful anal-
ysis while enabling the user to check whether a group signiﬁcantly
differs from another/others. This is deﬁned as different states of a
label-type variable, the difference being based on the composition
(species appearing) and frequency of each taxon.
Although some authors have pointed out the originality of the
different ecological processes in the Mediterranean mountains
(e.g. Grabherr et al., 2003; Giménez-Benavides et al., 2005, 2007),
studies remain scarce, and the work presented here represents
the ﬁrst analysis of the restoration of degraded areas in Mediter-
ranean mountains. For this reason, together with the absence of
nearby areas with similar features, our results cannot be com-
pared with others, highlighting the need for further studies in other
Mediterranean mountains, in order to establish speciﬁc patterns
and methodologies for these speciﬁc environments.
We conclude that under current management-planning frame-
works, which include concepts such as Limits of Acceptable Change
(LAC), Visitor Impact Management (VIM) or Visitor Experience, and
Resource Protection (VERP) (see Needham and Rollins, 2005), it is
important to develop indicators or tools to evaluate the success
of restoration measures. In this context the methods used here;
chronosequence, restoration model, and a multivariate approach,
such as discriminant analysis to evaluate the composition, all
together may constitute a useful and reliable way to evaluate the
relative success of vegetation-recovery plans, not only on ski slopes
but also in other degraded areas.
This research has been ﬁnanced by Cetursa Sierra Nevada S.A.
The authors appreciate the contribution of Lea Gálvez, from the
technical staff of Cetursa, who promoted the research and provided
us the information regarding the recovery plan and the manage-
ment of the ski resorts. We also would like to thank David Nesbitt for
linguistic advice. Three anonymous referees improved signiﬁcantly
the style, as well as the content of the manuscript.
Barni, E., Freppaz, M., Siniscalco, C., 2007. Interactions between vegetation, roots,
and soil stability in restored high-altitude ski runs in the Alps. Art. Antarct. Alp.
Res. 7, 25–33.
Benthem, R., 1973. Recreational and environmental planning. Biol. Conserv. 5, 1–5.
Blanca, G., Molero, J., 1990. Peligro de extinción en Sierra Nevada (Granada, Espa
In: Hernández Bermejo, J.E., et al. (Eds.), Conservation Techniques in Botanic
Gardens. Koeltz, Koenigstein, pp. 97–102.
Blanca, G., Cueto, M., Martínez-Lirola, M.J., Molero, J., 1998. Threatened vascular ﬂora
of Sierra Nevada (Southern Spain). Biol. Conserv. 85, 269–285.
Blanca, G., Cabezudo, B., Cueto, M., Fernández-López, C., Morales, C., 2009. Flora
Vascular de Andalucía Oriental, 4 vols. Consejería de Medio Ambiente. Junta de
Bowles, M., Jones, M., 1999. Vegetation Proﬁles and Species Richness Indices for
Chicago Region Graminoid Plant Communities Described and Sampled by the
Illinois Natural Areas Inventory. The Morton Arboretum, Lisle, IL.
Bradshaw, A.D., 1997. What do we mean by restoration? In: Urbanska, K.M., Webb,
N.R., Edwards, P.J. (Eds.), Restoration Ecology and Sustainable Development.
Cambridge University Press, Cambridge, pp. 9–14.
Delgado, R., Sánchez-Mara
nón, M., Martín-García, J.M., Arnada, V., Serrano-
Bernardo, F., Rosúa, J.L., 2007. Impact of ski pistes on soil properties: a case
study from a mountainous area in the Mediterranea region. Soil Use Manage.
Elsasser, H., Messerli, P., 2001. The vulnerability of the snow industry in the Swiss
Alps. Mt. Res. Dev. 21, 335–339.
Foster, B.L., Tilman, D., 2000. Dynamic and static views of succession testing the
descriptive power of the chronosequence approach. Plant Ecol. 146, 1–10.
Author's personal copy
J. Lorite et al. / Landscape and Urban Planning 97 (2010) 92–97 97
Giménez-Benavides, L., Escudero, A., Pérez-García, F., 2005. Seed germination of high
mountain Mediterranean species, altitudinal, interpopulation and interannual
variability. Ecol. Res. 20, 433–444.
Giménez-Benavides, L., Escudero, A., Iriondo, J.M., 2007. Local adaptation enhances
seedling recruitment along an altitudinal gradient in a high Mountain Mediter-
ranean plant. Ann. Bot. 99, 723–734.
Gómez, A., 2002. Geomorphological map of Sierra Nevada; glacial a periglacial geo-
morphology. Consejería de Medio Ambiente. Junta de Andalucía, Sevilla.
Gómez-Campo, C., Malato-Beliz, J., 1985. The Iberian Peninsula. In: Gómez Campo,
C. (Ed.), Plant Conservation in the Mediterranean Area. Dr. W. Junk Publishers,
Dordrecht, pp. 47–70.
Grabherr, G., Nagy, L., Thompson, D.B.A., 2003. An outline of Europe’s alpine areas.
In: Nagy, L., Grabherr, G., Körner, C., Thompson, D.B.A. (Eds.), Alpine Biodiversity
in Europe. Springer-Verlag, Berlin, Heidelberg, pp. 3–12.
Grismer, M.E., Schnurrenberger, C., Arst, R., Hogan, M.P., 2008. Integrated monitoring
and assessment of soil restoration treatments in the Lake Tahoe Basin. Environ.
Monit. Assess. 150, 365–383.
Gros, R., Monrozier, L.J., Bartoli, F., Chotte, J.L., Faivre, P., 2004. Relationship between
soil physico-chemical properties and microbial along a restoration chronose-
quence of alpine grassland following ski run construction. Appl. Soil Ecol. 27,
Hair, J.F., Anderson, R.E., Tatham, R.L., Black, W.C., 1998. Multivariate Analysis, 5th
edition. Prentice Hall International, Inc., New Jersey.
Henderson, R.A., 1999. Response to Henry Howe. Ecol. Restor. 17, 189–192.
Heywood, V.H., 1996. Endemism and biodiversity of the ﬂora and vegetation of Sierra
Nevada: Environmental consequences. In:ChacónMontero, J., Rosúa Campos, J.L.
(Eds.), Sierra Nevada. Conservación y Desarrollo Sostenible, vol. V. Universidad
de Granada, Madrid, pp. 191–201.
Jabaloy, A., Galindo, J., Sanz, C., 2008. Guía geológica. Granada; guías de Natrualeza.
Diputación de Granada, Granada.
Kent, M., Coker, P., 1992. Vegetation Description and Analysis. A Practical Approach.
CRC Press, Boca Raton.
Körner, C., 2003. Alpine Plant Life. Functional Plant Ecology of High Mountain Ecosys-
tems. Springer-Verlag, Berlin, Heidelberg.
Lasanta, T., Laguna, M., Vicente-Serrano, S.M., 2007. Do tourism-based ski resorts
contribute to the homogeneous development in the Mediterranean mountains?
A case study in the Central Spanish Pyrenees. Tourism Manage. 28, 1326–1339.
Lorite, J., 2002. La vegetación de Sierra Nevada. In: Blanca, G. (Ed.), Flora amenazada
y endémica de Sierra Nevada. Editorial Universidad de Granada, Granada, pp.
Lorite, J., Navarro, F.B., Valle, F., 2007a. Estimation of threatened orophytic ﬂora and
priority of its conservation in the Baetic range (S. Spain). Plant Biosyst. 141, 1–14.
Lorite, J., Gómez, F., Mota, J.F., Valle, F., 2007b. Orophilous plant communities of
Baetic range in Andalusia (south-easter Spain): priority altitudinal-island for
conservation. Phytocoenologia 37, 625–644.
Lorite, J., Salazar, C., Valle, F., 2007c. Floristic analysis of almeriensian Sierra Nevada
(Almería, SE Spain). Flora Mediterranea 17, 9–23.
Magurran, E., 1988. Ecological Diversity and its Measurements. Princeton University
Médail, F., Quézel, P., 1997. Hot-spots analysis for conservation of plant biodiversity
in the Mediterranean Basin. Ann. Mo. Bot. Gard. 84, 112–127.
Médail, F., Quézel, P., 1999. Biodiversity hotspots in the Mediterranean Basin: setting
global conservation priorities. Conserv. Biol. 13, 1510–1513.
Merlin, G., Di-Gioia, L., Goddon, C., 1999. Comparative study of the capacity of ger-
mination and of adhesion of various hydrocolloids used for revegetalization by
hydroseeding. Land Degrad. Dev. 10, 21–34.
Molero, J., Pérez Raya, F., 1987. La ﬂora de Sierra Nevada. Avance sobre el catál-
ogo ﬂorístico nevadense. Servicio de Publicaciones Universidad de Granada,
Muller, S., Dutoit, T., Alard, D., Grevilliot, F., 1998. Restoration and rehabilitation
of species-rich grassland ecosystems in France: a review. Restor. Ecol. 6, 94–
Needham, M.D., Rollins, R.B., 2005. Interest group standards for recreation and
tourism impacts at ski areas in the summer. Tourism Manage. 26, 1–13.
Pignatti, S., 1993. Impact of tourism on the mountain landscape of central Italy.
Landsc. Urban Plan. 24, 49–53.
Quezel, P., 1977. Forests of the Mediterranean basin. In: Quezel, P., Tomaselli, R.,
Morandini, R. (Eds.), Mediterranean Forests and Maquis: Ecology, Conservation
and Management, MAB Tech. Notes 2. UNESCO, Paris, pp. 9–32.
Rixen, C., Freppaz, M., Stoeckli, V., Houvinen, C., Huovinen, K., Wipf, S., 2008. Altered
snow density and chemistry change soil nitrogen mineralization and plant
growth. Arct. Antarct. Alp. Res. 40, 568–575.
Sainz Ollero, H., Hernández Bermejo, J.E., 1985. Sectorización ﬁtogeográﬁca de la
Peninsula Ibérica e Islas Baleares: La contribución de su endemoﬂora como
criterio de semejanza. Candollea 40, 485–508.
Sluis, W.J., 2002. Patterns of species richness and composition in re-created grass-
land. Restor. Ecol. 10, 677–684.
Snowdon, P., Slee, B., Farr, H., Godde, P.M., 2000. The economic impacts of different
types of tourism in upland and mountain areas of Europe. In: Godde, P.M., Price,
M.P., Zimmermann, F.M. (Eds.), Tourism and Development in Mountain Regions.
CAB International, Wallingford, pp. 137–145.
Tsuyuzaki, S., 1990. Species composition and soil erosion on a ski area in Hokkaido,
Northern Japan. Environ. Manage. 14, 203–207.
Tsuyuzaki, S., 1995. Ski slope vegetation in Central Honshu, Japan. Environ. Manage.
Tsuyuzaki, S., 2002. Vegetation development patterns on skislopes in lowland
Hokkaido, Northern Japan. Biol. Conserv. 108, 239–246.
Urbanska, K.M., 1995. Biodiversity assessment in ecological restoration above the
timberline. Biodivers. Conserv. 4, 679–695.
Urbanska, K.M., 1997. Restoration ecology research above the timberline: coloniza-
tion of safety islands on a machine-graded alpine ski run. Biodivers. Conserv. 6,
Urbanska, K.M., Fattorini, M., 2000. Seed rain in high-altitude restoration areas in
Switzerland. Restor. Ecol. 8, 74–79.
Watson, A., 1985. Soil erosion and vegetation damage near ski lifts at Cairngorm.
Scotland Biol. Conserv. 33, 363–381.
Wipf, S., Rixen, C., Fischer, M., Schmid, B., Stoeckli, V., 2005. Effect of ski piste prepa-
ration on alpine vegetation. J. Appl. Ecol. 42, 306–316.
Woodward, A., Jenkins, K.J., Schreiner, E.G., 1999. The role of ecological theory
in long-term ecological monitoring: report on a workshop. Nat. Areas J. 19,
[Show abstract] [Hide abstract] ABSTRACT: An intense debate exists on the effects of post-fire salvage logging on plant community regeneration, but scant data are available derived from experimental studies. We analyzed the effects of salvage logging on plant community regeneration in terms of species richness, diversity, cover, and composition by experimentally managing a burnt forest on a Mediterranean mountain (Sierra Nevada, S Spain). In each of three plots located at different elevations, three replicates of three treatments were implemented seven months after the fire, differing in the degree of intervention: "Non-Intervention" (all trees left standing), "Partial Cut plus Lopping" (felling 90% of the trees, cutting the main branches, and leaving all the biomass in situ), and "Salvage Logging" (felling and piling the logs, and masticating the woody debris). Plant composition in each treatment was monitored two years after the fire in linear point transects. Post-fire salvage logging was associated with reduced species richness, Shannon diversity, and total plant cover. Moreover, salvaged sites hosted different species assemblages and 25% lower cover of seeder species (but equal cover of resprouters) compared to the other treatments. Cover of trees and shrubs was also lowest in Salvage Logging, which could suggest a potential slow-down of forest regeneration. Most of these results were consistent among the three plots despite plots hosting different plant communities. Concluding, our study suggests that salvage logging may reduce species richness and diversity, as well as the recruitment of woody species, which could delay the natural regeneration of the ecosystem.
- "phanerophytes, chamaephytes, hemicryptophytes, or geophytes, according Raunkjaer's system; Raunkjaer, 1934), while annual species were noted and used for the calculation of vegetation cover but not classified into species. Non-annual species (hereafter referred to as perennials but also including biennials) were classified according to their post-fire regeneration strategy (seeders, resprouters, or both), in accordance with the literature (Lorite et al., 2007, 2010; Blanca et al., 2009) and expert knowledge (Appendix S1). Nomenclature of species throughout the manuscript follows Flora de Andalucía Oriental (Blanca et al., 2009) or Flora iberica (Castroviejo et al., 1986e2012). "
[Show abstract] [Hide abstract] ABSTRACT: Background/Question/Methods Bog exploitation for horticultural purposes leaves large surfaces of residual peat that remain devoid of vegetation for decades. Restoration of those bogs is necessary to mitigate the loss of local biodiversity. However, tools to assess the success of restoration works have not been rigorously defined yet. We used vacuum-milled peat extracted bogs restored by the moss transfer technique in Eastern Canada as a model system to test an approach for assessing restoration success, based on plant composition. A total of 188 plots in 12 restored bogs that had been restored from 4 to 11 years ago and continuously monitored were clustered in three success categories, according to their characteristic vegetation composition. Then, vegetation composition in the plots was analyzed retrospectively at the third year since restoration to obtain the combination of indicator species that best discriminated between the success categories using linear discriminant analysis (LDA). Results/Conclusions LDA classified correctly 86% of the cases into three success categories: a first one representing Successful restoration, with dominance of Sphagnum, a typical bog genus that is able to initiate self-regulatory mechanisms leading back to bog ecosystems (restoration goal); a second one representing Failure, with dominance of bare peat; and a third category, interpreted as a dead-ended successional pathway, dominated by Polytrichum strictum, a pioneer moss that usually facilitates Sphagnum colonization. Recently restored bogs were finally used to illustrate the use of our predictive tool and suggest different management strategies.
- "Multivariate analyses can be used effectively to develop integrative tools for evaluating success since they make it possible to synthesize environmental information, thereby explaining most system variability on fewer dimensions. Among the panoply of existing multivariate techniques, linear discriminant analysis (LDA, Fisher, 1936; Rao, 1948 Rao, , 1952) is one of the few that can be used specifically for prediction purposes, although it has seldom been applied for this aim in ecology (Legendre and Legendre, 2012), especially in the evaluation of restoration projects (but see Syvaranta et al., 2008 and Lorite et al., 2010). We combined several indicator species, as well as key environmental and management variables, through LDA modeling to predict success in attaining desired trajectories shortly (3 years) after restoration work (i.e., application of the restoration technique ). "
[Show abstract] [Hide abstract] ABSTRACT: In recent decades, the use of some subalpine mountain grasslands in the Central Spanish Pyrenees has changed. Ski resorts have been developed and cattle herd management has shifted from the traditional "rotational-type" system in which grazing cattle are overseen by a herder to a "continuous-type" system that does not involve a herder. In 2005, the locations of 30 floristic inventories performed in 1972 were revisited and inventories were repeated in two adjacent similar areas, although one had been used for the development of ski runs and the other had not. The objective was to assess the effects of those changes on plant diversity and other characteristics of the grasslands. In both areas, plant diversity was significantly higher in 2005 than it was in 1972. Both areas had been grazed by cattle to a similar extent; thus, the results suggest that diversity was affected primarily by the change in the livestock grazing system. Livestock grazing within the skiing area appears to have counterbalanced any reduction in plant diversity that would have occurred because of the construction and use of ski runs. In the skiing area, legume cover and pastoral value decreased, the Ellenberg Nitrogen Index reflected lower soil nutrients available to plants, and the cover of plant species that regenerate by seeds increased between 1972 and 2005; such changes did not occur in the non-skiing area. The abundance of ruderal species increased more in the skiing area than in the non-skiing area. Between 1981 and 2000, the amount of bare ground increased only in the skiing area.
- "In mountain ecosystems that have extreme climates, the bare soil created in the process of building and maintaining ski runs is susceptible to erosion (Ries, 1996; Rixen et al., 2003; Gros et al., 2004). Those patches become desiccated and overheated, which leads to further erosion in the summer (Watson, 1985 ). Harsh highmountain conditions and the mechanical damage caused to plants by the construction and maintenance of ski runs greatly retards primary succession and the recovery of the vegetation cover (Rixen et al., 2003; Wipf et al., 2005; Roux-Fouillet et al., 2011 ); consequently , extensive areas can remain bare in these environments (Tsuyuzaki, 1994; Urbanska, 1997; Rixen et al., 2008b; Lorite et al., 2010). Ski runs can have negative effects on plant diversity and species richness (Forbes, 1992; Goñi and Guzmán, 2001; Wipf et al., 2005) and lead to changes in floristic composition (Puntieri, 1991; Fahey and Wardle, 1998; Tsuyuzaki, 2002; Titus and Tsuyuzaki, 1998; Ruth-Balaganskaya and Myllynen-Malinen, 2000; Rixen et al., 2003; Wipf et al., 2005; Roux-Fouillet et al., 2011). "