Agroforest Syst (2009) 76:389–400
DOI 10.1007/s10457-008-9165-y
Are drought and wildfires turning Mediterranean cork
oak forests into persistent shrublands?
Vanda Acácio Æ Milena Holmgren Æ Francisco Rego Æ
Francisco Moreira Æ Godefridus M. J. Mohren
Received: 18 January 2008 / Accepted: 24 July 2008 / Published online: 11 September 2008
Ó The Author(s) 2008. This article is published with open access at Springerlink.com
Abstract In the Iberian Peninsula Mediterranean
oak forests have been transformed into a mosaic
landscape of four main patch-types: forests, savannas,
shrublands and grasslands. We used aerial photographs over a period of 45 years (1958–2002) to
quantify the persistence and rates of transitions
between vegetation patch-types in southern Portugal,
where cork oak is the dominant tree species. We used
logistic regression to relate vegetation changes with
topographical features and wildfire history. Over the
45 years, shrublands have been the most persistent
patch-type (59%), and have been expanding; forests
are also persistent (55%) but have been decreasing
since 1985; savannas and grasslands were less
persistent (33% and 15%, respectively). Shrublands
persistence was significantly correlated with wildfire
occurrence, particularly on southern exposures after
1995. In contrast, forest persistence decreased with
V. Acácio (&) G. M. J. Mohren
Forest Ecology and Forest Management Group,
Wageningen University, P.O. Box 47,
6700 AA Wageningen, The Netherlands
e-mail: vanda.acacio@wur.nl
V. Acácio F. Rego F. Moreira
Centro de Ecologia Aplicada ‘‘Prof. Baeta Neves’’,
Instituto Superior de Agronomia, Universidade Técnica
de Lisboa, Tapada da Ajuda, 1349-017 Lisbon, Portugal
M. Holmgren
Resource Ecology Group, Wageningen University,
P.O. Box 47, 6700 AA Wageningen, The Netherlands
wildfire occurrence, and forests were more likely to
change into shrublands where wildfire had occurred
after 1995.
Keywords Alternative stable states
Cistus Semiarid Shrub encroachment
Vegetation shifts Quercus suber
Introduction
Mediterranean forests are one of the biodiversity
‘‘hotspots’’ on earth (Myers et al. 2000; Olson and
Dinerstein 2002). Due to a long history of human
disturbance, approximately 70% of the original
Mediterranean forests and shrublands were already
destroyed by 1990 (Millennium Ecosystem Assessment 2005). The remaining cover is considered to be
in critical or endangered conditions (Olson and
Dinerstein 2002; Mooney et al. 2005). Despite this
long history of degradation, forest cover has
expanded in some areas of Mediterranean Europe
since the 1960s after agricultural abandonment and
rural exodus (Mazzoleni et al. 2004). In contrast, the
opposite trend has apparently taken place in the drier
evergreen oak forests of the southern Iberian Peninsula. Here, canopy trees have decayed reducing tree
density (Gonçalves 1991; Ferreira 2000; Vicente and
Alés 2006), tree seedling mortality is high (Anon
1990; Montero et al. 1994) and xerophytic Cistus
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390
shrublands have expanded in areas formerly covered
by oak forests and savannas (Calvão and Palmeirim
2004).
Since in semi-arid Mediterranean ecosystems, tree
seedlings often establish under the shade of adult
nurse shrubs and trees (Holmgren et al. 2000; Pulido
and Dı́az 2003; Castro et al. 2004; Gómez-Aparicio
et al. 2004), removal of the original vegetation makes
conditions for seedling establishment difficult, resulting in a positive feedback of increasing vegetation
loss that can be difficult to reverse (Mulligan et al.
2004). Persistent loss of original vegetation has been
described for the five Mediterranean-type regions
around the world (e.g. Friedel 1987 for South Africa;
Westoby et al. 1989 for Australia; Laycock 1991 for
California; Puigdefábregas and Mendizabal 1998 for
the Mediterranean Basin; and Holmgren 2002 for
Chile).
Succession has traditionally been conceived as a
relatively linear process, but evidence from a variety
of ecosystems indicates that persistent alternative
vegetation or ecosystem states may occur, particularly after disturbances (Friedel 1991; Laycock 1991;
Scheffer et al. 2001). In these cases, succession is
strongly delayed or practically stopped in a condition
called arrested succession (Putz and Canham 1992;
Sarmiento 1997).
In the southern Iberian Peninsula, original forests
dominated by cork oaks (Quercus suber) and Holm
oaks (Quercus ilex) have been transformed into an
agroforestry system. Long-term human management
has included combinations of clearing, livestock
grazing, and ploughing for agriculture cultivation
(Joffre et al. 1999; Pulido et al. 2001). During the
1960s socio-economic changes (industrialization,
immigration to cities, emigration, and opening of
international markets) led to rural exodus and a
gradual abandonment of crops and pastures (Joffre
et al. 1991; Pinto-Correia 2000).
At present, oak systems in southern Portugal are a
mosaic of four types of patches: forests, oak savannas
(traditionally known as montados in Portugal and
dehesas in Spain), shrublands and grasslands. Oak
savannas and grasslands are maintained through
grazing and ploughing in the traditional agroforestry
system (Marañón 1988; Huntsinger and Bartolome
1992; Dı́az et al. 2003). In the absence of human
management, the usual pathway of natural succession
is through the gradual colonization by different shrub
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Agroforest Syst (2009) 76:389–400
species (pioneer Cistus shrubs, followed by other
shrub species like Arbutus and Erica), followed by
oak natural regeneration leading eventually to forest
recovery (Natividade 1950; Gonçalves 1991). However, under a combination of dry conditions (either
drought periods or south facing slopes) and wildfires,
forest recovery may be impeded and a transition to
Cistus shrublands more likely. The last decades have
been characterized by drier conditions (Cabrinha and
Santo 2000; Esteban-Parra et al. 2003; Pausas 2004)
and a higher wildfire incidence (Pausas 2004; Anon
2006) whereas ploughing and livestock grazing have
tended to decrease in Portugal (Alves et al. 2003;
Pinto-Correia and Vos 2004).
In this paper, we quantify the rate of change
between vegetation patch-types (cork oak forests,
cork oak savannas, shrublands and grasslands) in the
Mediterranean oak forests of southern Portugal
during 45 years (1958–2002) and assess the role of
drought and wildfire to explain these transitions. We
use aerial photographs and a Geographic Information
System (GIS) to quantify the persistence of vegetation patch-types and the rates of transitions between
them. We project our results for the coming decades
and discuss the forest changes in southern Portugal in
relation to the trends described for other European
Mediterranean regions.
Materials and methods
Study area
The study area is located in Serra do Caldeirão, a
mountain ridge in the northeastern part of the
Algarve province, southern Portugal, with an approximate area of 11,000 ha. Climate is Mediterranean,
with an average annual temperature of 16.6°C and
average annual precipitation of 900 mm (45-year
period: 1958–2002, Barranco do Velho station).
Altitude ranges from 146 to 588 m. Soil type is
homogeneous, consisting mainly of schist lithosols
(soils with hard rock at less than 10 cm depth). These
soils are shallow, with low fertility and prone to
erosion. Cork oak is the dominant tree species and
cork extraction represents the main economic activity
for local people. Our study area is one of the most
continuous remnants of cork oak forests of southern
Portugal.
Agroforest Syst (2009) 76:389–400
391
Image processing
We used hardcopy aerial photographs and digital
orthophotos (rectified copy of an original aerial
photograph) from five different years covering a
45-year period: 1958 (scale 1:26,000), 1972 (scale
1:8,000), 1985 (scale 1:5,000), 1995 (scale 1:40,000)
and 2002 (scale 1:5,000). Aerial photographs (1958,
1972 and 1985) were previously scanned and then
orthorectified and georeferenced with ENVI 3.4 (Anon
2001a) in order to produce geometrically correct
images and project them to the same coordinates
reference as the digital orthophotos (1995 and 2002)
(Transverse Mercator projection, datum WGS84).
Digital elevation models with a spatial resolution of
8 m were used for altitude coordinates (orthorectification process), and ground control points (between 10
and 15 per photo) were taken from the 2002orthophoto for map coordinates (georeferentiation
process). We obtained a RMS Error of about 10 m.
Photo interpretation and land cover classification
We used a regular 0.5 9 0.5 km grid of 441 points
covering the entire study area, and photo interpreted on
screen a 50 m-radius circle (sampling unit) around
each photographic point. Photo interpretation was
performed on a photo-by-photo basis. The grid was
created with Arcview GIS 3.2 (Anon 1999) and laid
over the aerial photographs and orthophotos (imported
into the GIS as scanned images). Our grid has a
resolution higher than the usual 4 9 4 km grids used
for monitoring forest conditions in Europe and allows
detecting changes in vegetation types (Köhl et al.
1994). Systematic point sampling is commonly used in
forest inventory and gives better estimates than simple
random sampling for large areas. Each of the 441
sampling units was classified into one of the following
five vegetation patch-types for each image year:
1.
2.
cork oak forests with high cork oak density
(more than 100–150 trees/ha) and a diverse
shrubby layer (e.g. Arbutus unedo, Viburnum
tinus, Erica arborea, Pistacia lentiscus, Rhamnus
alaternus, Phillyrea latifolia); this class corresponds to a tree cover higher than 30%;
cork oak savannas where cork oak density is
lower than in forest patches (usually less than
100 trees/ha) with sparse shrubs in the
3.
4.
5.
understory; this class corresponds to a tree cover
between 10% and 30%;
shrublands dominated by Cistus ladanifer shrubs,
pure or mixed with Ulex argenteus, Genista
hirsuta and Lavandula stoechas shrubs; in this
class, less than 10% of the sampling unit is
covered by trees;
grassland patches dominated by cultivated croplands, semi-natural pastures, or fallow land; in
this class, less than 10% of the sampling unit is
covered by trees.
others (e.g. plantations, pine forests, eucalyptus
forests, riparian forests, and urban).
When more than one vegetation patch-type
occurred in a sampling unit, it was classified based
on the predominant type. Sampling units with only
one type of patch-type were predominant (more than
50% of the total units) and when not, we could
usually clearly assign one dominant class to each
unit. We used tree cover percent rather than tree
density since in many cases it was not possible to
count the number of trees on the images. Boundaries
for forest density (10% and 30% of ground cover)
follow the National Forest Inventory norms (Anon
2001b).
Before photo classification (2003), we visited the
study area several times to recognize the different
vegetation types and compare them with the patterns
on the aerial photos of 2002. Photo interpretation was
carried out by a group of four persons and results were
cross-validated within the group. Approximately 10%
of the sampling units were validated in the field.
Accuracy assessment of land cover classification
Accuracy assessment was evaluated with the calculation of the error matrix (Walsh and Burk 1993). The
error matrix allows us to quantify the overall
accuracy of the land cover classification procedure,
dividing the number of units in the diagonal cells of
the matrix by the total number of field-validated
units. It allows us also to estimate the probabilities of
classification of a unit of class i in class j, as nij/nj
(where i represents the observed class, j the predicted
class, and nj the sum of all training observations
predicted as class j). Overall accuracy was about
82%. The highest omission error occurred in savannas (0.4) due to confusion with cork oak forests and
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392
pines, followed by forests (0.24) and shrublands
(0.14), due to confusion with pine forests. On the
other hand, pine forests and shrublands showed the
highest commission errors (0.36 and 0.33, respectively), confused mainly with cork oak forests.
Quantification of vegetation changes
The information for each sampling unit was stored as a
five-layer Arcview GIS database, where each layer
represented the landscape at a single point in time. Five
transition matrices (Usher 1981; Rego et al. 1993)
were built for the five-year set (1958–1972, 1972–
1985, 1985–1995, 1995–2002), and for 1958–2002
(overall period). We estimated the (1) percentage of
land covered by each type of vegetation patch
(expressed as the % of total number of points); (2)
transitions between vegetation patch-types by counting the number of sampling units of any vegetation
patch-type that changed into any other between two
discrete time periods (expressed as percentages); (3)
rates of change (percentage of change/year) between
vegetation patch-types, dividing the percentage of
transitions from one patch-type to another by the total
number of years during which the transitions occurred;
(4) projection of changes for the coming decades,
multiplying the observed rates of transitions for a
given period by the number of hectares of each
vegetation patch-type registered in 2002, and estimating its area for 2003; multiplying again the estimated
area by the transition rates and estimating a new area
for the following year; calculations were started in
2002 and were successively repeated until 2050. We
projected two scenarios: a conservative one based on
the average of the four rates of transition observed for
the overall period (1958–1972; 1972–1985; 1985–
1995; 1995–2002), and a second one based on the most
recent trends (1995–2002). To project such changes
we assumed that the probabilities of transitions
between vegetation patch-types were constant
throughout the whole period of projection, following
a Markovian process (Rego et al. 1993).
Correlates of vegetation changes
Topographical variables (slope and exposure) were
derived from digital elevation models (Portuguese
Military Geographic Institute) and stored as two
Arcview GIS layers. We assigned one class of slope
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Agroforest Syst (2009) 76:389–400
and exposure to each sampling unit based on the most
common type. The slope categories were: low to
medium (0–20%); steep (20–30%); very steep
([30%). The exposure categories were: north facing
slopes (including NW, N and NE), south facing slopes
(including SW, S and SE) and others (E and W).
Wildfire occurrence was available as spatially
referenced and digitized data for every year between
1984 and 2002 (General Directorate of Forests). The
limits of the burned area per year were stored as an
Arcview GIS layer and overlaid with the sampling
units in order to identify the sampling units that
burned between 1984 and 2002. The maximum
number of times a unit was burned in the period
1984–2002 was two.
Statistical analysis
We overlayed the Arcview GIS layers containing the
information on slope, exposure and wildfire occurrence with the layer containing the vegetation
transition types (e.g. forest to shrubland) for each
time period (1958–1972, 1972–1985, 1985–1995,
1995–2002). We used logistic regression to explore
the relative importance of slope steepness, exposure
and wildfire occurrence on the observed vegetation
changes. We analysed (1) transitions from each
vegetation patch-type (except ‘‘others’’) to shrublands
and (2) vegetation type persistence (except ‘‘others’’)
for each of the 4 periods, yielding a total of 28
models (4 persistences plus 3 transitions, times 4 time
periods). For each model, the predictor variable took
the value 1 if a transition (or persistence) occurred for
each sampling unit during the specified period, or 0 if
not. Slope categories were ranked from 1 to 3,
respectively for 0–20%, 20–30%, and [30%. Exposure and wildfire (presence/absence) were assumed as
categorical variables. Forward stepwise selection was
used, and variables entering the model selected based
on the likelihood-ratio test (Hosmer and Lemeshow
1989). To check whether the models could be
improved, some variables were square-transformed
and interactions between all variables explored, using
the procedures suggested by Hosmer and Lemeshow
(1989). Consequently, we created the simplified
dummy variables Exposure_South and Exposure_North, which took the value 1 for the specified
exposure and 0 for the others categories. Model
goodness of fit was assessed through the likelihood
Agroforest Syst (2009) 76:389–400
393
ratio statistic and the v2 test. Due to a low number of
‘‘occurrences’’ in some transition types, statistical
analyses could not be undertaken.
Results
Land cover changes
During the 45-year study period there were significant changes in land cover. Between 1958 and 1972,
shrublands expanded rapidly (54% increase) while
grassland cover registered a decline of 481%
(Table 1). In those 15 years, shrublands became the
patch-type with the highest cover (33.8%). By 1972,
only 4.8% of the total area was covered by grassland
patches which remained as such onwards (Table 1).
During the following two decades, shrublands
slightly decreased (around 28% of the total area) to
became again the predominant patch-type in 2002
(32.4%). Cork oak forests, the most abundant patchtype in 1958 (about 30%), decreased during the next
45 years to cover around 24.5% of the area in 2002
(Table 1). Savannas initially expanded but have
decreased since 1972 onwards. During the whole
period, only shrublands and the category ‘‘others’’
registered net increases (51.9% and 62.4%, respectively, Table 1). The high net increase of ‘‘others’’ is
mainly due to plantations and pine regeneration and
expansion (both sum up about 80% of the category
‘‘others’’ in 1958 and 90% in 2002).
Patch-dynamics: persistence and transition rates
between vegetation patch-types
The dynamism of this landscape can be better
appreciated in Fig. 1 showing the percentage of
Table 1 Temporal changes of vegetation patch-types (% of
total number of sampling units) and net changes from 1958 to
2002
1958 1972 1985 1995 2002 Net changes
1958–2002
Forests
29.9
26.8
31.1
29.3
24.5
Savannas
19.7
24.9
21.5
21.1
18.6
-5.9
Shrublands 15.6
33.8
28.1
27.9
32.4
?51.9
Grasslands
Others
-22
27.9
4.8
3.9
3.6
6.3
-342.9
6.8
9.8
15.4
18.1
18.1
?62.4
transitions between vegetation patch-types from 1958
to 2002 in a conceptual state-and-transition model.
Forests and shrublands were the most persistent
patch-types, contrasting with savannas and grasslands. The largest changes occurred in grasslands,
switching mostly to shrublands (40%), but also to
savannas (25%) and others (20%). Other important
transitions occurred in savannas, as they changed to
shrublands (28%) or forests (23%). Despite the high
persistence of cork oak forests (55%), an important
fraction became shrublands (20%).
Forests changed faster into savannas than into
shrublands (Fig. 2), but more forest patches changed
into shrublands (20%) than into savannas (13%)
during the overall period (Fig. 1), which indicates
that most of these changes are not direct transitions,
implying a gradual loss of the tree canopy: forest
patches change first into savannas and then these
savannas change into shrublands. In any case, forests
are changing at an increasing rate since 1985.
Oak savannas have become shrublands at increasing rates, while less savannas have returned to forests
(Fig. 2). Between 1972 and 1984, forests recovered
when savannas switched into forests at the highest
rate, but since then there has been a clear decreasing
trend. At present, savannas change six times faster
into shrublands than into forests.
Shrublands have been the most persistent patchtype (59%) changing very little into something else
(around 7% into forests, grasslands or savannas)
(Fig. 1). The vegetation transition rates indicate that
cork oak recruitment seems to be particularly difficult
in shrublands. Only 7% of shrublands changed into
savannas and forests compared to 25% of grasslands
changing into savannas, and 23% of savannas
becoming forests (Fig. 1). Grasslands have been
transformed mainly into shrublands and in a smaller
extent, into savannas (Fig. 1), both at a decreasing
rate since 1958 (Fig. 2).
Projection of changes for the coming decades
Projection of our results for the coming decades
shows an increase of shrublands and a decrease of
forests and savannas (Fig. 3). Obviously, the conservative approach based on the transition rates for the
45 years (1958–2002) suggests smoother changes
than the one projected using the most recent rates
(1995–2002). It is interesting to notice that present
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Agroforest Syst (2009) 76:389–400
33%
29 units
Savannas
25%
31 units
13%
17 units
28%
24 units
7%
5 units
5%
4 units
7% 5 units
Grasslands
20% 26 units
Shrublands
15%
18 units
55%
72 units
59%
41 units
Fig. 1 State-and-transition model with observed transitions
(%) between vegetation patch-types in southern Portugal
(1958–2002). Changes (% and number of units) for each
patch-type are based on the number of sampling units classified
as that patch-type in 1958 that changed (or remained) in 2002.
Rate of transitions from forests to:
transitions
per year
Forests
9%
6 units
40%
49 units
Dotted lines indicate frequency of transitions \10%. The % of
transitions from each patch-type to ‘‘others’’ are not shown in
the Figure; such % is the difference between 100% and the
total % of transitions shown from each patch-type
Rate of transitions from savannas to:
5%
5%
4%
4%
3%
3%
2%
2%
1%
1%
58-71
72-84
85-94
95-02
0%
0%
Savannas
Shrublands
Forests
Grasslands
Rate of transitions from shrublands to:
5%
transitions
per year
23%
20 units
4%
3%
2%
1%
Shrublands
Grasslands
Rate of transitions from grasslands to:
5%
58-71
4%
3%
72-84
2%
1%
95-02
85-94
0%
0%
Forests
Savannas
Grasslands
Forests
Savannas
Shrublands
Fig. 2 Rate of transitions (% of transitions per year) from each vegetation patch-type to the others (1958–2002)
rates of transition suggest a future increase in
grassland patches, which have increased in the last
years (see Fig. 2).
The role of topographic variables and wildfire
history on vegetation changes
Forest persistence increased with slope steepness
during each time period and decreased with wildfire
123
occurrence since 1985 (Table 2). In contrast, shrubland persistence was significantly correlated with
wildfire, particularly on southern exposures after
1995 (Table 2). Grassland persistence was higher on
lower slopes and particularly when they were located
on southern exposures. Savanna persistence was not
influenced by any studied variable except during the
period 1958–1972, when they were more likely to
persist on steeper and no north-facing slopes
Agroforest Syst (2009) 76:389–400
395
Projection based on the rates of change between 1958-2002
160
140
120
100
80
Area of occupation (ha)
60
40
20
0
2002
2008
2014
2020
2026
2032
2038
2044
2050
Projection based on the rate of change between 1995-2002
160
140
120
100
80
60
40
20
0
2002
2008
2014
2020
2026
2032
2038
2044
2050
Years
Grasslands
Shrublands
Savannas
Forests
Fig. 3 Projection of the area (ha) of occupation of vegetation
patch-types for 2002–2050, based on the rates of change
observed in 1958–2002 (average) and 1995–2002
(Table 2). Of the three potential patch transitions to
shrublands, wildfire increased the probability of
transition from forests to shrublands in the period
1995–2002, and grasslands were more likely to
change into shrublands if located on lower slopes
(until 1985). None of the studied variables influenced
the transitions from savannas to shrublands (Table 2).
Discussion
Overall vegetation changes
Shrubland encroachment has been the most conspicuous landscape change in this part of the southern
Iberian Peninsula since 1958 when forest understory
use for agriculture and pastures was abandoned in
many areas (Alves et al. 2003; Pinto-Correia and Vos
2004). Shrublands increased during the last 45 years at
the expense of the other vegetation patch-types.
Persistent shrub encroachment has also been found
in other Mediterranean ecosystems like the Italian
‘macchia’ (Pignatti 1978; Scarascia-Mugnozza et al.
2000), the cork oak forests of southern France
(Trabaud and Galtié 1996), as well as in the Spanish
dehesas, an oak savanna system comparable to the
Portuguese montados (Huntsinger and Bartolome
1992). In contrast, forest expansion has been described
as the dominant landscape change throughout Mediterranean Europe since the 60s (Mazzoleni et al. 2004).
There, shrublands are usually reported as intermediate
phases prior to forest recovery (Mazzoleni et al. 2004).
Shrublands and forests are more persistent vegetation types (59 and 55% respectively) than savannas
(33%) and grasslands (15%). The main vegetation
changes along the study period (above 20% in Fig. 1)
were the (a) replacement of grasslands by shrublands,
savannas and others, (b) replacement of savannas by
shrublands and forests, and (c) replacement of forests
by shrublands. Grasslands decreased in more than
300% and since 1972 only 4.8% of the total area has
remained covered by grassland patches, probably
maintained for family subsistence close to human
settlements.
Patch-type dynamics
Shrubland persistence was positively correlated with
wildfire occurrence, particularly on southern exposures since 1995. Shrublands usually grow in very dry
and degraded soil conditions and are commonly
dominated by Cistus spp., especially on south-facing
slopes (Nuñez et al. 1986; Diniz 1994; Seng and Deil
1999). In addition, Cistus shrublands are active
pyrophites, thus their persistence can be supported
by a positive feedback mechanism triggered by
frequent wildfires. Particularly, Cistus ladanifer is a
highly flammable species due to its external resins
(Trabaud 1981). Our results point towards the same
pattern with a tendency of Cistus ladanifer shrublands to be maintained by wildfire occurrence and on
south-facing slopes, where conditions are drier and
more limiting for seedling recruitment and cork oak
survival. Cistus shrubs will survive wildfires and
maintain themselves as long as the interval between
successive fires is higher than the time that is needed
for seed bank restoration. Cistus seed banks seem to
be restored every 3 years for Cistus ladanifer in
Spain (Ferrandis et al. 1999) or only 2 years for
Cistus sp. in California (Montgomery and Strid
1976), which is much less than the time interval
between successive fires in the region.
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Table 2 Results of multivariate logistic regression to predict persistence and transitions between vegetation patch-types
Fo ? Sr
Sa ? Sr
Gr ? Sr
Sr ? Sr
Fo ? Fo
Sa ? Sa
1958–1972
1972–1985
1985–1995
1995–2002
n = 3 (*)
n = 0 (*)
–
Fire (?) P = 0.057
n=5
n=5
–
–
–
–
n=8
n=6
n = 13
n = 26
Slo (-) P = 0.049
Slo (-) P = 0.003
n = 3 (*)
n = 1 (*)
n = 77
n=8
–
–
Fire (?) P = 0.007
Fire 9 Asp_S (?) P = 0.004
n = 57
n = 108
n = 97
n = 105
Slo (?) P = 0.002
Slo (?) P \ 0.001
Slo (?) P = 0.002
Slo (?) P \ 0.001
n = 106
n = 105
Fire (-) P = 0.026
Fire (-) P = 0.027
n = 115
n = 103
Slo 9 Asp P = 0.026
–
–
–
Slo 9 Asp_O (?)
n = 66
n = 66
n = 58
Slo 9 Asp_S (?)
n = 67
Gr ? Gr
Slo (-) P \ 0.001
Slo (-) P = 0.001
Slo (-) P = 0.002
Slo (-) P = 0.003
n = 19
Slo 9 Asp_S (-) P = 0.002
Slo 9 Asp_S (-) P = 0.001
Slo 9 Asp_S (-) P = 0.002
n = 11
n = 12
n = 14
For each combination of transition/persistence type and time period, variables entering the model are shown, as well as their direction
of association (?/-) with the response variable, and significance (Likelihood ratio test). Vegetation types include shrublands (Sr),
forests (Fo), savannas (Sa) and grasslands (Gr). Variables include slope (Slo), exposure (Asp), exposure south (Asp_S), exposure
‘‘others’’—east or west (Asp_O) and Wildfire (Fire). n = number of occurrences. The asterisk (*) signals low frequency-transitions
that were not considered for analysis. For all analyses, sample size was 441 sample units
Shrublands expanded initially at grassland patches.
The rate of this transition has been decreasing across
time, and was particularly evident on flatter areas
until 1985, perhaps because deeper soils contribute to
faster shrub encroachment, after the abandonment of
agriculture and pastures. In contrast, cork oak forests
have been decreasing since 1985. Although cork oak
forests can be highly persistent, a significant proportion (20%) has turned into shrublands. Our results
showed that forests persisted longer on steeper slopes,
likely because they were less accessible to human
interventions. After 1995, we observe a clear effect of
wildfire occurrence on tree density: when no wildfire
occurs forests persist at steeper slopes, whereas
occurrence of wildfire causes tree loss and conversion
of forests into shrublands. Although the cork oak is
partially protected against fires by bark insulation and
is able to resprout after fire, frequent or intense
wildfires may kill adult trees, especially if wildfires
123
occur immediately after cork extraction (Cabezudo
et al. 1995; Pausas 1997; Moreira et al. 2007), which
takes place in the summer, when the wildfire season
begins in Mediterranean Europe. In addition, the rate
of tree recovery might decrease with increasing
wildfire frequency because of lower resprouting
success and higher oak seedling mortality (Dı́azDelgado et al. 2002).
After an initial recovery following agriculture
abandonment (1958–1984), savannas have also been
changing into shrublands since 1985. Initial cork oak
recruitment in old fields at the beginning of abandonment was also found in Eastern Spain, followed
by shrub encroachment (Pons and Pausas 2006).
Shrub encroachment could be limiting tree recruitment onwards, especially on drier conditions. Our
results suggest that savannas were more likely to
persist in steeper slopes and drier exposures that did
not face north.
Agroforest Syst (2009) 76:389–400
Forests and savannas have been converted into
shrublands at an increasing rate. Forest replacement
by shrublands has been positively associated with
wildfire incidence since 1995, but the transition of
savannas to shrublands was not associated with any
studied variable. Drought has probably played an
important role on the replacement of savannas by
shrublands: savannas have lower tree density and
therefore seedlings and saplings are exposed to higher
irradiance and water stress than in forests. Patches
converted into shrublands have been highly persistent
and rarely switched into a different vegetation type.
Experimental evidence indicates that shrubland persistence can be explained by oak seedling recruitment
limitation in multiple phases (Acacio et al. 2007).
Grassland persistence has been higher on lower
slopes, probably because these are the areas where
farming activities are maintained. Grasslands tend to
persist on southern slopes likely due to soil and
climatic conditions slowing successional changes.
Mechanisms limiting forest recovery
Limited cork oak recruitment in shrublands might be
more related to seedling recruitment than to seed
availability. Indeed our transition rates from grasslands to savannas and from savannas to cork oak
forests suggest similar probabilities of cork oak
recruitment in grasslands and savannas patches despite
the likely difference in seed availability between these
two types of patches (Pulido and Dı́az 2003). Dense
Cistus stands may preclude cork oak regeneration in
different ways. It is known that Cistus shrubs have
allelopathic effects, inhibiting seedling germination
and survival (Chaves and Escudero 2001; Lobón et al.
2002). There could be also limitations in acorn
dispersal, increased acorn predation (Herrera 1995;
Leiva and Fernández-Alés 2003), or limitations in
seedling germination and or survival under the dense
and dry shrub cover (Retana et al. 1999).
Experimental work at the study area showed that
different mechanisms of oak recruitment limitation
(namely, seed source limitation, dispersal limitation,
germination limitation, and establishment limitation)
were significantly more severe in Cistus shrublands
than in oak forests and savannas, hence cork oak
seedling recruitment in shrubland was impeded in
multiple ways (Acacio et al. 2007). Cork oak recruitment is also rare in the shrublands of eastern Spain
397
(Pons and Pausas 2006) where very low oak seedling
survival rates have been found under Cistus shrubs
(Gómez-Aparicio et al. 2004; Pulido and Dı́az 2005).
Cork oak forests under climate change
Mean annual temperatures and rainfall intensity with
erosive potential (number of days with rainfall
[13 mm) have clearly increased in southern Portugal
since 1972 (Water Institute, National Meteorological
Institute and Regional Coordination Commission of
Algarve).
The increase of mean annual temperature and
rainfall extremes during the last decades follows the
trends predicted by present climate change models
projecting a higher frequency of droughts and intense
rain events for Mediterranean-climate regions (Cubash
et al. 1996; McCarthy et al. 2001). Under this scenario,
wildfire frequency and soil erosion are expected to
intensify. This will undoubtedly reduce forest cover
and limit regeneration, while facilitating shrubland
persistence and expansion. Comparable results have
been found in Spanish and French cork oak forests
where higher wildfire frequency was also related to a
decrease in forest resilience and their switch into
shrublands (Dı́az-Delgado et al. 2002; Trabaud and
Galtié 1996, respectively). Also modelling results
show that increasing wildfire frequency promotes
shrublands dominated by Cistus and limits Quercus
growth (Pausas and Vallejo 1999). Moreover, Cistus
species are well adapted to drought through physiological responses (Werner et al. 1998). In addition to
drought and wildfires, cork oak mortality has been
correlated to pathogens whose incidence might interact
with climatic conditions (Moreira and Martins 2005).
Recent climate simulation models also predict a
very severe impact of climate change on cork oak
forests and savannas in Portugal due to difficulties in
regeneration and increase of tree mortality, especially
in the south of the country, while shrublands are
expected to expand at the expense of cork oak areas
(Pereira et al. 2002). Such future trends are shown in
our projections, especially the one based on the most
recent rates. Predictions based on the most recent
trends are more reliable than the ones based on the
average rate (1958–2002), since transitions are more
likely to change over time. Nevertheless, we probably
underestimate the changes by assuming that the
probabilities of transitions follow a Markovian
123
398
process, which is clearly unlikely given the varying
transition rates from 1958 to 2002.
Our findings indicate a serious threat to the cork oak
production system in southern Portugal, which is the
basis of the local economy and has high conservation
value, protected by European Union (Habitats Directive 92/43/EEC). The combination of increasing
temperatures and increasing wildfire frequency has
contributed to the expansion of shrublands in previous
cork oak dominated areas. Shrublands may represent
an alternative persistent state of lower vegetation
biomass on the most degraded soils and dry areas,
where conditions for forest recovery become extremely difficult. If observed trends continue, the
sustainability of this unique landscape of the Western
Mediterranean area will be in jeopardy.
Acknowledgements This research was funded by the
Portuguese Foundation for Science and Technology
(Fundação para a Ciência e a Tecnologia), fellowship
number SFRH/BD/5008/2001. We also would like to thank
to: the Portuguese Military Geographic Institute, Portuguese
Geographic Institute and National Pulp Industry Association
for kindly providing the photographic material used in this
project; Rute Palmeiro, Susana Pereira, Tiago Dias and Miguel
Porto for the help with the photo interpretation work; João
Carreiras and Filipa Marques for the help with ENVI and
Arcview software; Patrick Jansen for insightful comments; and
Maria José Vasconcelos for kindly providing information on
burned areas. M. Holmgren thanks the Dutch NWO Meervoud
Programme (836.05.021).
Open Access This article is distributed under the terms of the
Creative Commons Attribution Noncommercial License which
permits any noncommercial use, distribution, and reproduction
in any medium, provided the original author(s) and source are
credited.
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