Ecography 34: 8593, 2011
doi: 10.1111/j.1600-0587.2010.06250.x
# 2011 The Authors. Journal compilation # 2011 Ecography
Subject Editor: Francisco Pungnaire. Accepted 18 January 2010
Demographic processes of upward range contraction in a long-lived
Mediterranean high mountain plant
Luis Giménez-Benavides, Marı́a José Albert, José Marı́a Iriondo and Adrián Escudero
L. Giménez-Benavides (luis.gimenez@urjc.es), M. J. Albert, J. M. Iriondo and A. Escudero, Área de Biodiversidad y Conservación, Univ. Rey
Juan Carlos-ESCET, Tulipán s/n. ES-28933 Móstoles, Madrid, Spain.
We analyzed demographic data of a long-lived high mountain Mediterranean plant, Silene ciliata Poirret, over a 4-yr
period. Selected populations were located at contrasting altitudes at the southernmost margin of the species (Sierra de
Guadarrama, central Spain), representing a local altitudinal range at the rear edge of its overall distribution. Previous
studies have suggested that differences in the reproduction and performance of individuals at upper and lower
populations may have implications for population dynamics. We used matrix analysis to assess their demographic
behaviour. Life Table Response Experiments were used to identify the life history stages most relevant to observed
differences in population growth rates between populations.
Transition matrices revealed great spatio-temporal variability in demographic traits. Seedling recruitment was very low
each year in all populations. Maximum longevity of S. ciliata individuals in the lower peripheral population was much
lower compared to the central population, probably due to higher adult mortality. Population growth rate (l) showed a
declining trend at the lowest altitude and a relatively stable trend at the central population. Long-term simulations also
indicated a great risk of quasi-extinction at the lowest population. Our results suggest that rear edge populations of
S. ciliata at Sierra de Guadarrama are suffering demographic processes that may be leading to the latitudinal displacement
of the species’ range.
Peripheral populations of plant species are important
reservoirs of intraspecific genetic diversity and evolutionary
potential (Lesica and Allendorf 1995, Hampe and Petit
2005, Jump and Peñuelas 2005), as well as functional
drivers of ecosystem stability (Eriksson 2000). Studies of
individual plant performance across a species’ range
frequently find lower survival and/or reduced fecundity at
range margins compared to the range center (Jump and
Woodward 2003, Giménez-Benavides et al. 2007a, b,
Marcora et al. 2008). However, a major concern is whether
reductions in fitness components really affect population
growth and persistence. In fact, many times the persistence
of plant species is not crucially dependent on reproductive
success and seedling establishment (Pico and Riba 2002,
Garcı́a 2008, Iriondo et al. 2008).
Differences in life-history traits, such as life-span, will
largely determine the species dependence on sexual regeneration. In long-lived perennial species, the impact of
limited seed output on population maintenance is difficult
to determine due to the complexity of recruitment, but a
tradeoff between sexual regeneration and persistence of
already established individuals has been suggested (Garcı́a
and Zamora 2003). Typical examples of persistence due to
longevity and/or vegetative reproduction are more frequent
in stressful and unstable environments such as arid, alpine
and rocky habitats (Grime 2001, Garcı́a and Zamora 2003)
where geographical limits of reproduction do not necessarily coincide with actual range limits (Gaston 2003).
Therefore, to get a complete view of the factors shaping
geographical range limits, the components of individual
performance must be integrated into population dynamic
models across species’ distributions (Angert 2006, Foden
et al. 2007). This task is especially relevant today, when ongoing climate warming and other anthropogenic impacts,
such as habitat fragmentation and changes in land use, are
currently threatening peripheral populations. Despite this,
there are few detailed comparisons of population dynamics
of central vs marginal populations (Nantel and Gagnon
1999, Stokes et al. 2004, Angert 2006, Samis and Eckert
2007). Moreover, while demographic and evolutionary
traits underlying the expansion of species at their leading
edge have been more extensively studied during the last few
decades (Petit et al. 2004), population dynamics responsible
for range contractions at the rear margins of species’
distributions have not received sufficient attention (Hampe
and Petit 2005). This general lack of mechanistic studies is
even greater in high mountain environments, even though
they provide an excellent opportunity for the study of range
margins (Angert 2006, Körner 2007).
85
In mountain plants, conditions for regeneration and
survival are hierarchically arranged within their distribution
range. Firstly, they are more suitable in the latitudinal
centre of their distribution area than in the periphery.
Secondly, they also appear structured within each mountain
island. Similar to latitudinal range displacements, recent
altitudinal shifts in the abundance and distribution of
species inhabiting mountain environments have been
documented during the last few decades. Several studies
have revealed contemporary alterations of species richness
in high summits (Grabherr et al. 1994, Gottfried et al.
1999, Virtanen et al. 2003, Walther et al. 2005, Pauli et al.
2007, Erschbamer et al. 2009). The migration of lowland
plant species to higher elevations forces subsequent displacements of alpine species (Theurillat and Guisan 2001).
Altitudinal shifts in vegetation belts and distribution ranges
of species have already been documented (Walther et al.
2002, Klanderud and Birks 2003, Peñuelas and Boada
2003, Lesica and McCune 2004). Therefore, there is an
urgent need for accurate forecasting of the consequences of
this process. Important progress on species’ distribution
modelling has recently been made, but most of these models
do not explicitly take into account the essential mechanisms
operating at individual and population scales (Thuiller et al.
2008, Morin and Thuiller 2009). These factors may cause
important bias and inaccuracy in current projection models,
and are a probable cause of the divergence found among
coarse and fine resolution models when compared (Trivedi
et al. 2008). At least at the population level it seems basic to
monitor and model demographic trends at rear populations.
Despite this necessity, demographic studies of high mountain plants are still scarce compared to those of lowland
species, and few studies have documented the populationlevel dynamics driving altitudinal displacements (Doak and
Morris 1999, Diemer 2002, Angert 2006).
In the present work, we analyzed demographic data of a
long-lived Mediterranean high mountain plant, Silene ciliata
(Caryophyllaceae), at different altitudes within its southernmost margin of distribution (central Spain). During the last
45 yr, mean air temperature has increased by 1.88C in this
area, and days of snowcover per year have decreased by
19.7 d (Giménez-Benavides et al. 2007a). In addition to
direct impacts of climate warming on the species’ performance, the area has suffered a substantial bottom-up shrub
encroachment (Sanz-Elorza et al. 2003) with potential
consequences for the persistence of S. ciliata rear populations. The final objective of the present work is to assess
whether the reproductive and recruitment failure observed at
the rear edge of the species (Giménez-Benavides et al. 2007a,
2008) results in regressive population dynamic in the lower
population compared to higher altitude populations. We
argue that the breakdown of sexual regeneration at lower
limits could only be balanced out by a long lifespan and
reduced adult mortality. Otherwise, the species could suffer a
high risk of peripheral extinction and altitudinal range
contraction under the present global warming context. We
analyzed demographic performance of the species over a
4-yr period using transition matrix models and long-term
simulations. Specifically, the questions addressed were: 1) are
S. ciliata populations at the rear altitudinal edge experiencing
a declining population trend? 2) Are the upper and lower
populations along an altitudinal gradient driven by the
86
same demographic processes? And, if they differ, 3) what are
the vital rates responsible for the observed differences in
population growth rates at different altitudes? 4) Do
populations at different altitudes differ in their probability
of quasi-extinction in the long term?
Methods
Plant species and study site
Silene ciliata (Caryophyllaceae) is a long-lived perennial
plant that grows in main mountain ranges of the Balkan
Peninsula, the Appenines, the Massif Central in France and
the northern half of the Iberian Peninsula, covering a
latitudinal range from 408N to 468N (Tutin et al. 1995). In
mountain ranges of central Spain, where the species reaches
its southernmost margin, it grows from 1900 m (treeline
zone) up to the highest summits (ca 2600 m). The species
typically grows in a compact cushion-shape. Its flowering
period extends from late June to early-mid September.
Flowering stems (133 per adult plant) are 15 cm in
height and bear 15 flowers. Hand-crossing experiments
indicate that S. ciliata is a self-compatible species. However,
passive autogamy is restricted by a pronounced protandry
so it requires pollinators (Giménez-Benavides et al. 2007a).
Although many alpine species are highly clonal (Forbis
2003), no evidence of vegetative propagation was observed
in this species when several individuals were dug up
(Giménez-Benavides unpubl.).
The study area was in the Sierra de Guadarrama (Peñalara
Natural Park), a mountain range located in central Spain,
50 km north of Madrid city (408N, 38W). Mean annual
precipitation at Navacerrada Pass weather station (1800 m,
8 km southwest of the study site) is 1350 mm, and is
concentrated from late autumn to early winter. A marked
drought season occurs from late May to October (Fig. 1).
Snowfall generally begins in October and the snow-free
season begins in MayJune (Palacios et al. 2003).
This work is part of a broader study of factors
controlling the distribution and performance of S. ciliata
along the local altitudinal gradient. Three populations were
selected for this demographic approach. The first population was located in the vicinity of Laguna Chica (hereafter
Laguna), a small glacial lake situated in a moraine deposit
in the treeline zone (1970 m). Vegetation is dominated by a
dense shrub cover of Cytisus oromediterraneus and Juniperus
communis subsp. alpina intermingled with a low-dense
stand of Pinus sylvestris. Here, S. ciliata is displaced by the
shrub species and only grows in small, isolated pasture
patches dominated by Festuca curvifolia. The second
population was on the Dos Hermanas peak (hereafter
Dos Hermanas), a summit flat area situated at 2250 m,
dominated by a Cytisus-Juniperus shrub formation and
patchy xerophytic fellfields of Festuca curvifolia. This
fellfield community bears extreme winds and a relatively
short snowcover period, and is characterized by the
abundance of cushion plants (Escudero et al. 2004). The
third population was located at the summit of the highest
peak of the mountain range, the Peñalara peak (hereafter
Peñalara), at 2440 m. This area is dominated by the Festuca
fellfield and shrub species are scarce. As a consequence of
Figure 1. Climatic data at the Navacerrada Pass weather station (40846?N, 4819?W; 1860 m, located 8 km south-west of study sites).
Columns and lines represent monthly mean precipitation and temperature, respectively, during the period 19462006 and in the study
years, 2003 to 2006.
reported temperature increase and reduction of snow cover,
probably combined with a moderate reduction in livestock
grazing, the Cytisus-Juniperus shrub belt is encroaching
and replacing the Festuca cryophilic pastures colonized by
S. ciliata (Sanz-Elorza et al. 2003).
Census scheme
Data were collected from 2003 to 2006. One permanent
plot was established in each population for demographic
monitoring. Plot size varied between populations due to
differences in microhabitat characteristics and plant density
but populations were larger enough and similar in slope and
orientation to be considered representative at each altitude
considered. At Laguna (1970 m), we initially monitored
all plants available within a small population (128 individuals in a 9 m2 plot), while at Dos Hermanas (2250 m)
and Peñalara (2440 m) we tagged 266 (7 m2 plot) and 168
individuals (10 m2 plot), respectively, within a larger,
continuous population. All plants found within each plot
were mapped to allow subsequent location. Plants were
monitored every year at the end of the reproductive season
(SeptemberOctober). Plant size was estimated as maximum
cushion diameter. Total number of inflorescences per plant
was counted in a single visit as an estimate of reproductive
output. Previous studies suggest that inflorescence number
is a good surrogate of fruit production (Pearson’s r790,
549 and 580, pB0.0001, for Laguna, Dos Hermanas and
Peñalara respectively, Giménez-Benavides et al. 2007a).
Silene ciliata seedlings emerge at the beginning of the
growing season, suffering extremely high mortality during summer (Giménez-Benavides et al. 2007b). Thus, all
seedlings found within the plots at the end of the growing
season were registered as a basis for estimating annual
recruitment rates. Unfortunately, the plot at Peñalara was
vandalized during the second year of study, so parameter
estimations of only one transition could be obtained.
The possible uncertainty in plant population structure
and demographic parameters derived from the establishment of a single plot per population was assessed by
comparing them with five extra plots of 5 m2 randomly
placed at each population. Plant size of every plant in these
extra plots was measured in 2005, and size structures of
the permanent plots were compared with those of the
respective extra plots from the same altitude by cross-tabs
and chi-square test.
Stage classification
We established five stage classes, one seedling class and four
reproductive classes. The first corresponded to seedlings of
about 1 cm in diameter that germinated in spring. Seedlings
that survived into the next growth period grew into a
reproductive class. Reproductive classes were obtained by
classification of individuals by k-means clustering (Hartigan
1975), except seedlings, using pooled data from all
populations: small (plants of 1.52.5 cm diameter), medium (34.5 cm), large (58 cm) and extra-large (]8.5 cm).
87
Matrix construction
We constructed Lefkovitch matrices for each population
and time interval using estimates of inflorescence production, seedling recruitment and transition probabilities
between reproductive stages. Transitions were built from
the underlying vital rates (survival, growth and fertilitity),
following Morris and Doak (2002). When sample size of
some stages was small (n B13) mainly seedlings and small
reproductive plants at Laguna survival rates were obtained
from average transition frequencies across all years for each
population (Menges and Dolan 1998, Angert 2006).
The reproduction terms in the matrix were estimated as
follows. Mean number of inflorescences per class was used
to calculate the proportional contribution of each adult class
to total reproductive effort. Thus, the reproduction term for
each reproductive class in each transition was estimated
following the equation:
Fi;t ;t 1 sdlt 1 Ri;t
;
4
X
(Ri;t ni;t )
i1
where Fi,t,t1 is the reproduction matrix element of class
i (small, medium, large or extra-large) for the period t to
t 1, sdlt1 is the total number of seedlings censused in
the population in time t 1, Ri,t is the proportional
contribution of class i to total reproductive effort, and ni,t
is the number of individuals in class i surviving at time
t. Thus, seedlings censused in the following year were
allocated among the four reproductive classes according
to their proportional reproductive effort in the previous
year.
These estimations assumed that seedlings in time t 1
germinated from seeds produced in time t, as occurs in the
absence of a permanent soil seed bank. Field and lab assays
showed that germination capacity of S. ciliata seeds can
reach 100% over a one year period (Giménez-Benavides
et al. 2005, 2007b). Thus, we assumed that soil seed
bank does not play an important role in the dynamics of
S. ciliata populations. A recent seedbank study conducted
on this fellfield community also supports the absence
of a permanent seedbank in this species (Garcı́a-Camacho
2009).
Transition matrix models project population size according to the equation:
x(t 1)Ax(t );
where x is a vector of the number of individuals in different
plant stages (stage distribution at time t) and A is a matrix
of probabilities and fertilities that defines the survival and
reproduction of individuals in each stage between time t
and time t1 (i.e. matrix elements). The transition matrix
A is derived from a life cycle graph that shows the possible
transitions between stages (Caswell 2001). The life cycle
graph for our model system is shown in Fig. 2.
The dominant eigenvalue for each transition matrix was
used to calculate the finite rate of increase, l. Bootstrapped
matrices were generated by randomly sampling individuals
with replacement within stage classes, using a Matlab routine
(Matlab 7.0, MathWorks). Two-thousand bootstrapped
transition matrices were used to obtain bias-corrected 95%
confidence intervals for l.
Maximum plant longevity was estimated for each population separately following Forbis and Doak (2004). A
starting vector of one seedling and zero reproductive
individuals was multiplied by the mean matrix with all
fecundities set to zero (Caswell 2001). Year by year, the
resulting vector was multiplied by the mean matrix until
the summed probability of survival for all stage classes
reached 0.01.
Figure 2. Life cycle of Silene ciliata. Each arrow represents a one-year transition. S denotes survival (stasis in the same class, and growth or
regression to a different class) and F fertility.
88
Life Table Response Experiments
X
(l )
(dh) 1l
(ai;j
ai;j
)
l(l ) :l(dh) 1ai;j A m
i;j
j
where l(dh) is the population growth rate in Dos Hermanas,
and Am is the mean matrix from both populations. The first
factor of the sum (in parenthesis) denotes the differences,
and the result of multiplying them by the appropriate
sensitivity are the contributions (Caswell 2001). We carried
out an LTRE analysis for each time interval. All matrix
analyses were performed with Matlab 7.0.
Probability of quasi-extinction
We calculated the probability of reaching a quasi-extinction
threshold for each population by computer simulations. We
used the Matlab routine developed by Morris and Doak
(2002), which estimates the quasi-extinction time cumulative distribution function for a structured population
in a stochastic environment. Proportions of individuals
were equal to the initial stage structure of each population.
Environmental stochasticity was implemented by using the
three available transition matrices for each population. It
was based on the climatic variability among the periods
of time considered. The years 2004, 2005 and 2006
had milder temperatures and summer rainfall (data not
shown, Navacerrada Pass weather station, 40846?N,
4819?W, 1860 m), whereas in 2003 the most severe
European heatwave took place (Schär et al. 2004) (Fig.
1). Maximum temperatures of the complete climate series
were reached in summer, no rainy days were recorded in
July, and only one in August (Giménez-Benavides et al.
2007a). Therefore, two contrasting scenarios were considered: in the ‘‘realistic’’ scenario the probability of occurrence of an extremely warm and dry year (first transition
matrix) was fixed to 0.1 whereas in the ‘‘warming’’ scenario
this probability was substantially increased by making the
three transition matrices equally probable. We set the quasiextinction threshold at 50 individuals and time horizon at
250 yr. Peñalara was not included in the simulations
because only one transition was available. The starting
population density was the number of individuals of the
initial plots, i.e. 128 and 266 individuals for Laguna and
Dos Hermanas, respectively.
40
Individuals (%)
We used Life Table Response Experiments (LTREs) to
identify the matrix elements most relevant to the observed
differences in population growth rate between populations
(Caswell 2001). We focused on a one-way fixed design
where the two populations (Laguna and Dos Hermanas)
were of interest in themselves. We chose the Dos Hermanas
matrix as the reference matrix, i.e. a baseline for measuring
population effects. Thus, population growth rate in Laguna
was defined as:
50
Laguna (Low)
Dos Hermanas (Middle)
Peñalara (High)
30
20
10
0
Seedling
Small
Medium
Large
Extra-large
Stage classes
Figure 3. Size structure of S. ciliata populations according to stage
classes established in transition matrices. Data from the three
permanent plots set in 2003. n128 individuals in Laguna, 266
individuals in Dos Hermanas and 168 individuals in Peñalara.
large and extra-large classes. On the contrary, Dos
Hermanas and Peñalara populations were consisted of
smaller plants, with a majority of medium and large plants
and B15% of extra-large individuals. The seedling class
was scarcely represented in all populations. The results
obtained in our permanent plots did not differ in terms of
size structure from the surrounding extra plots (low: x2 13.4, p0.640, medium: x2 22.2, p 0.332, high: 19.4,
p0.248), thus the permanent plots could be considered
representative of their corresponding population.
Matrix analyses
Transition matrices also revealed great differences in
demographic traits between populations (Supplementary
material Table S1). Adult survival was higher at Dos
Hermanas for all stage classes and all time intervals. Mean
survival per class increased at Dos Hermanas from small to
extra-large plants, while mean survival was higher for large
plants at Laguna. Seedling survival showed high variability
between years and populations. The climatically extreme
year (20032004 transition) showed higher mortality only
for seedling and extra-large plants in Laguna (Supplementary material Table S1).
Population growth rates (l) showed a declining trend
at the lowest altitude and relatively stable population
dynamics at the central and highest population. Bootstrapping confidence intervals of l were much wider at
Laguna than at Dos Hermanas (Fig. 4).
Estimated maximum plant longevity inferred from
average transition matrices was 147 yr at Dos Hermanas
and only 23 yr at Laguna.
Life Table Response Experiments
Results
Plant size structure showed great differences between
populations (Fig. 3). Laguna population was dominated
by large plants, with 70% of individuals belonging to
LTRE analyses revealed greater differences in matrix elements between populations for the seedling stage, in addition to greater variability in differences (Fig. 5a). However,
the higher seedling survival at Laguna slightly contributed
to differences in l between populations (Fig. 5b). Lower
89
1.10
Laguna (Low)
Dos Hermanas (Middle)
Peñalara (High)
0.8
0.6
2003–04
2004–05
2005–06
0.4
LTRE differences
Median population growth rate
1.05
(a)
1
0.95
0.2
0
–0.2
–0.4
–0.6
0.9
–0.8
(b)
0.85
Seedling
Small
Medium
Large
Extra-large
0.03
Fecundity
Survival
0.02
0.75
2003–04
2004–05
2005–06
Time intervals
LTRE contributions
0.01
0.8
0
–0.01
–0.02
–0.03
–0.04
–0.05
Figure 4. Lambda values and 95% confidence intervals (from
2000 bootstrapped transition matrices) for each population and
time interval. Dashed line highlights a stable population (l 1).
–0.06
Seedling
Small
Medium
Large
Extra-large
–0.07
Stage classes
Probability of quasi-extinction
The probability of reaching a threshold density of 50
individuals was very different between Laguna and Dos
Hermanas under both scenarios. Such probability reached
100% after only ca 16 yr at Laguna, irrespective of the
scenario considered. However, at Dos Hermanas 100%
probability of quasi-extinction was reached in about 240 yr
under the ‘‘realistic’’ scenario, and was reached 50 yr earlier
under the ‘‘warming’’ scenario (Fig. 6).
Discussion
Demographic processes at contrasting altitudes
Our study clearly revealed that the S. ciliata population at
the lowest altitude presented a pronounced decline and
showed a demographic behaviour highly different from
those at the central and highest altitudes. These differences
were great enough to suggest that the studied peripheral
population of S. ciliata at its lower margin is not able to
withstand present conditions, which may lead to its local
extinction. This scenario was supported by the long-term
simulations showing great quasi-extinction hazard at the
lowest altitude, irrespective of the climatic scenario projected (Fig. 6). In the central population, finite rate of
increase was relatively stable, and time to quasi-extinction
was much longer than at the rear edge, indicating that
90
Figure 5. LTRE differences (a) and contributions (b) to differences in l between populations, per each stage class. Contributions
were grouped per vital rate (survival and fecundity). Bars represent
the mean value9SD for all time intervals. The Dos Hermanas
matrix was used as the reference matrix.
population declining is not likely to be occurring at this
altitude. However, quasi-extinction probability under a
warming scenario (with a higher chance of extreme years)
showed that this population is also vulnerable to changing
conditions. The larger confidence intervals of l at Laguna
1
Cumulative probability of quasi-extinction
survival of all reproductive stages at Laguna most greatly
contributed to differences in l. Fecundity contributions
increased with stage class but were always substantially lower
than survival contributions (Fig. 5b).
0.9
Laguna (Low)
0.8
Dos Hermanas (Middle)
0.7
0.6
0.5
0.4
0.3
0.2
0.1
0
0
50
100
150
200
250
Years into the future
Figure 6. Average of ten simulated cumulative distribution
functions of 5000 simulations, for the time to reach a quasiextinction threshold of 50 individuals for both populations of
S. ciliata. The plot only shows the ‘‘warming’’ scenario, in which
all time intervals are equally probable.
compared to those at Dos Hermanas reflected higher
uncertainty in population growth estimates, as a consequence of higher variability in plant fates. These results were
expected because marginal populations are usually near
the limit of their physiological tolerance and, consequently,
are more vulnerable to environmental stochasticity (Gaston
2003). Although we could only provide one transition, the
most favourable trend is expected at the highest population
(Peñalara peak), based on the positive finite rate of increase
(l 1.059) and the significantly higher values of adult
survival (100% for all reproductive stages).
Seedling recruitment observed in the field was highly
variable and mean values were very low for all the studied
years. This is a common feature in long-lived, high mountain
species (Forbis 2003, Forbis and Doak 2004). Furthermore,
recruitment in Mediterranean mountains is seriously limited
by high seedling mortality during summer drought (Castro
et al. 2004, 2005, Cavieres et al. 2005). The climatic diagram
presented in Fig. 1 precisely showed that mean monthly
temperature during the growing season of all the studied
years was 1.528C over the mean of the last 60 yr. Moreover,
mean monthly precipitation was severely reduced just in
July, the harsher month of the summer period. The studied
period is therefore representative of the warming trend
occurred in the last half of the century in this area (Wilson
et al. 2005, Giménez-Benavides et al. 2007a). Previous
results in S. ciliata showed that recruitment was mainly
limited by low seed production and low seedling emergence
and survival, probably due to environmental harshness in
summer, in contrast to biotic factors such as flower and fruit
predation, which had a minor effect on the probability
of plant recruitment (Giménez-Benavides et al. 2008).
Although a reciprocal sowing experiment detected some
evidence of local adaptation in seedling establishment along
this gradient (Giménez-Benavides et al. 2007b), the probability of recruitment (estimated as the probability of an
ovule becoming a 2-yr-old plant) was 20 to 40-fold higher
in medium and higher populations compared to the
lower population (Giménez-Benavides et al. 2008). Reduced
fecundity at species’ distribution limits has been widely
observed, and may be the main factor responsible for lower
population densities and aged population structures (Garcı́a
et al. 2000, Dorken and Eckert 2001, Jump and Woodward
2003, Marcora et al. 2008). As predicted by life history
theory (Grime 2001), extended longevity is expected to allow
long-term persistence of remnant populations in harsh
environments with high interannual climate variation, while
waiting for eventual recruitment episodes (Morris and Doak
1998, Garcı́a and Zamora 2003, Garcı́a 2008). In fact, our
LTRE analyses suggested that observed differences in l
between populations, much more pronounced at the seedling stage, were only slightly explained by the fecundity term
(Fig. 5b). Thus, fecundity didn’t seem an important factor
explaining differences in population growth rate between
altitudes. For this reason, demographic trends of long-lived
plants at their rear edge populations cannot simply be
inferred from their current recruitment rates, as they are
more determined by adult mortalities (Hampe and Petit
2005). Moreover, long-lasting longevity have been proved to
buffer the variability in vital rates associated with the
variability in climate, hence reducing the vulnerability of
long-lived species to climate change (Morris et al. 2008).
Interestingly, the effects of plant size on survival did not
follow the same trend at both altitudes. A decrease in plant
survival of extra-large plants occurred at the lower population, while this stage reached 100% probability of survival
at the central population. Size-dependence of demographic
fates defined as transition probabilities has been commonly
observed in a great variety of plants and environments
(Hortvitz and Schemske 1995). Greater adult survival has
also been observed in other alpine cushion-form Caryophyllaceae such as Silene acaulis (Morris and Doak 1998),
Minuartia obtusiloba and Paronychia pulvinata (Forbis and
Doak 2004). Smaller size stages are expected to be more
vulnerable to losses of above-ground tissues during adverse
environmental conditions (e.g. seasonal drought), and
consequently are expected to show higher rates of mortality.
However, our results suggest that environmental conditions
at the lower limit may also be affecting survival of large
plants by seriously reducing their life span. Indeed,
maximum longevity of S. ciliata individuals, inferred from
transition matrices, was very low at Laguna population
(23 yr) compared to Dos Hermanas (147 yr). In general,
maximum longevity of S. ciliata is relatively low when
compared to S. acaulis, another cushion plant from arcticalpine habitats (Morris and Doak 1998). The lifespan of
S. acaulis may extend over 300 yr and demographic studies
carried out in this species did not detect mortality among
large plants (Morris and Doak 1998). The results found at
Peñalara peak the highest population of the Guadarrama
mountain range were more in accordance with this
pattern (100% adult survival, resulting in a maximum
longevity of over 350 yr). Evidence of changes in plant
life-history along altitudinal gradients has been reported
previously (Körner 2007). Von Arx et al. (2006) detected
significantly older plants and lower growth rates at higher
altitudes by means of herb-chronology in three forb species,
corresponding to a more conservative life-history.
Population dynamics of Silene ciliata: towards an
altitudinal range contraction
Our results highlight the relevance of survival and longevity
for dissecting the processes that may be driving such distinct
population dynamics. Together with reproductive limitations (Giménez-Benavides et al. 2007a, 2008), rear edge
populations of S. ciliata at Sierra de Guadarrama are
suffering other demographic processes, resulting in low
adult survival, which may force them to an altitudinal range
shift. Populations inhabiting the rear edge will become
completely extirpated if current demographic processes
prevail. As upward shift is unviable, since the species
actually colonizes the highest summits of the major
mountain ranges in this southern margin, range shifts will
therefore be irremediably associated with a decrease in
habitat area. Biotic causes, apart from direct climatic effects
on survival and reproduction, are undoubtely involved in
the expected habitat contraction. The encroachment of
the high mountain xerophytic pastures (the main niche of
S. ciliata) by montane shrub species already detected in the
area (Sanz-Elorza et al. 2003), is probably one of the major
sources of risk. Remnant patches of suitable habitat are
currently colonized by small-sized populations dominated
91
by adult and senescent plants with extremely low proportions of seedlings and juveniles and a high degree of
isolation. Reduced population sizes and isolation are
common factors limiting individual reproductive performance (Leimu et al. 2006), leading to a feedback process
towards local extinction. Under this scenario, long-term
persistence would only be possible by the longevity of
established individuals, but our results highlight that longevity is also seriously eroded at this range margin. These
findings contrast with the assumption that populations of
many alpine species are not likely to be affected substantially by climate warming due to their long lifespan
(Steinger et al. 1995, Diemer 2002).
Evolutionary implications in response to climate
change
As noted above, extreme longevity coupled with occasional
recruitment episodes may support the demographic stasis
and even growth of perennial plant populations. However,
a more important consequence of population dynamics
governed by extreme longevity arises in the context of global
change. The evolutionary adaptation of populations to
changes in environmental conditions varies over both space
and time as a consequence of natural selection operating on
fitness components, and eventually fixed by sexual regeneration. Rapid climate change may act as a potent agent of
natural selection within populations and, in this context,
the adaptive potential of a given population will be partially
ruled by the frequency of sexual regeneration, being
annual plants the quickest to adapt because of their short
generation time (Jump and Peñuelas 2005). By contrast, in
long-lived perennials with delayed regeneration time, the lag
of adaptation will be significantly longer. Moreover, in
S. ciliata an extremely low recruitment rate is coupled to
size-dependent reproduction in stressful years, especially in
its lowland rearing edge (Giménez-Benavides et al. 2007a,
2008). In years of extreme summer drought, small-sized
individuals have a much lower flowering probability,
seriously limiting opportunities for adaptive selection.
In conclusion, although high longevity is the last strategy
to assure the long-term persistence of remnant populations
of S. ciliata, it is also critically reduced at its lowland range
limit. This situation is affecting the population growth rate,
eventually forcing the upward shift and the contraction of
its regional distributional range. Further demographic
studies are required to gather the necessary data to move
from simple bioclimatic niche models to process-based
models that take into account both climate change and
population dynamics.
Acknowledgements The authors especially thank Pedro QuintanaAscencio for his help with Matlab programming and the staff of
Parque Natural de las Cumbres, Circo y Lagunas de Peñalara who
gave them permission to work in the area. They also thank Nuria
Ortega, Vera Ortega and Raúl Garcı́a-Camacho who helped with
the field work and Lori De Hond for her linguistic assistance. This
work was supported by projects ISLAS (CGL2009-13190-C0301), SIL-HAD (CGL2009-08755) and LIMITES (CGL2009-
92
07229) funded by the Ministerio de Ciencia e Innovación (Spain)
and REMEDINAL2 funded by Comunidad de Madrid.
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