Download Heathlands confronting global change: drivers of biodiversity loss

Annals of Botany 111: 151 –172, 2013
doi:10.1093/aob/mcs257, available online at www.aob.oxfordjournals.org
INVITED REVIEW
Heathlands confronting global change: drivers of biodiversity
loss from past to future scenarios
Jaime Fagúndez*
Departamento de Bioloxı́a Animal, Bioloxı́a Vexetal e Ecoloxı́a, Facultade de Ciencias, Universidade da Coruña,
15071 – A Coruña, Spain
* E-mail [email protected]
Received: 3 July 2012 Returned for revision: 23 August 2012 Accepted: 18 October 2012 Published electronically: 6 December 2012
† Background Heathlands are dynamic plant communities characterized by a high cover of sclerophyllous, ericoid
shrubs that develop over nutrient-poor soils. Interest in the preservation of these habitats in Europe has increased
over the last decades, but over this time there has been a general decline in habitat quality, affecting community
structure, ecosystem functions and biodiversity. Negative drivers that trigger these changes include land-use
changes (i.e. habitat destruction and fragmentation), pollution, climate change, natural succession and human
management, as well as the presence of invasive exotic species.
† Scope Based on recent scientific literature, the effect of each of these potential drivers on a wide set of factors,
including physiological traits, species richness and diversity, community structure, ecosystem functions and soil
conditions, is reviewed. The effects of these drivers are generally understood, but the direction and magnitude of
factor interactions, whenever studied, have shown high variability.
† Conclusions Habitat loss and fragmentation affect sensitive species and ecosystem functions. The nature of the
surrounding area will condition the quality of the heathland remnants by, for example, propagule pressure from
invasive species. The dominant ericoid shrubs can be out-competed by vigorous perennial grasses with increased
atmospheric nitrogen deposition, although interactions with climate and management practices may either counteract or enhance this process. Grazing or periodic burning promotes heath loss but site-specific combined treatments maintain species diversity and community structure. Climate change alone moderately affects plant
diversity, community structure and ecosystem functions. Combined with other factors, climatic changes will condition heath development, mainly with regard to key aspects such as seed set and seedling establishment, rare
species occurrence and nutrient cycling in the soil. It is essential to address the effects of not only individual
factors, but their interactions, together with land-use history, on heathland development and conservation in
order to predict habitat response to future scenarios.
Key words: Heathlands, Europe, pollution, climate change, land-use change, natural succession, biological
invasions, grazing, burning, biodiversity, fynbos, environmental interactions.
IN T RO DU C T IO N
Heathlands are human-shaped habitats that play a crucial role
in the traditional European landscape, with a widely recognized strong cultural and natural value. Heathlands add substantially to many ecosystem services, such as food and
water supply, carbon sequestration, recreation, hunting, landscape and biodiversity conservation (Webb, 1998; Alonso
et al., 2001; Wessel et al., 2004). The importance of heathland
conservation has been highlighted in the last decades as a
general process of habitat loss and degradation is taking
place. These habitats, with their narrow specialist organisms
and complex interactions, are now threatened (De Graaf
et al., 2009). Protection and restoration plans for European
heathlands have been promoted by the European Union and individual countries to counteract the negative effects of different drivers of biodiversity loss over habitat quality (Pywell
et al., 2011). This interest has also been reflected in scientific
literature with many papers, especially in the 1990s and 2000s,
dealing with heathland biodiversity, conservation and restoration (Fig. 1).
The term heathland has been applied to a wide range of
habitats with some characteristics in common. From an ecological and physiognomic view, these are here taken as communities in which perennial sclerophyllous shrubs are
dominant and which develop over nutrient-poor soils
(Specht, 1979). These communities are usually dominated by
species included in the Ericeae tribe of the Ericaceae family
(Calluna, Daboecia and Erica genera; Oliver, 2000; Stevens
et al., 2004). Here, the focus is on the European temperate
heathlands, although some studies are discussed from other
similar habitats (i.e. other heathland types) such as the
fynbos biome from South Africa. The different heathland communities within temperate Europe are associated with different
climatic conditions (Mediterranean, oceanic or continental
heaths), edaphic conditions (acid or basic soils), waterlogging
(wet or dry heaths), altitudinal range (lowland, upland or
alpine heaths) and other ecological variables that condition
the community structure and composition. Heathlands are
mainly successional communities that replace woodlands,
except in the latitudinal and altitudinal tree limits, or hyperoceanic temperate areas in Scotland and in the west of
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152
Fagúndez — Heathlands confronting global change
Ireland, when heathlands may become the climax vegetation
(Specht, 1979; Cross, 2006; Fenton, 2008). The oceanic
climate of Atlantic regions depresses the potential tree-line
but, rather than temperature extremes, wind is a key factor
60
Number of papers
50
heath* biodiversity
heath* conservation
heath* restoration
40
30
20
0
< 1990
1990
1991
1992
1993
1994
1995
1996
1997
1998
1999
2000
2001
2002
2003
2004
2005
2006
2007
2008
2009
2010
2011
10
Year
F I G . 1. Scientific publications in the period 1900– 2011 that include the
topics heath* conservation, heath* restoration, and heath* biodiversity from
ISI Web of Knowledge search performed in May 2012. R 2 values of tendency
lines are: heath* biodiversity, 0.95; heath* conservation, 0.84; heath* restoration, 0.79. The asterisks (*) indicate ‘words starting with’.
Britton et al. (2008); Britton
and Fisher (2007, 2008)
controlling the vegetation type (Grace, 1997; Crawford,
2000). Sclerophyllous ericoids predominate except in the
most exposed mountain areas (Hodd and Sheehy
Skeffington, 2011). In extreme montane oceanic environments,
such as in western Ireland and Scotland and the Faroes, the
ericoid vegetation gives way to bryophyte heath (Averis
et al., 2004; Hodd and Sheehy Skeffington, 2011) A description of the different European heathland communities is
beyond this review and may be obtained from other sources
(Gimingham et al., 1979; Thompson et al., 1995; Dı́az,
1998; Loidi et al., 2007, 2010). The location of the study
sites from a selection of the papers reviewed here, covering
most of the geographic area studied, is shown in Fig. 2.
A large number of studies deal with different aspects of
heathland conservation, considering different threats, most of
which are included in the main drivers of biodiversity loss
as described by the Convention of Biological Diversity
(Secretariat of the Convention on Biological Diversity,
2006): (a) land-use changes (habitat loss and fragmentation);
(b) biological invasions; (c) climate change; (d ) pollution
and eutrophication; (e) overexploitation of natural resources.
Habitat loss has been extensively described throughout
Europe (e.g. Rose et al., 2000; Piessens et al., 2004) leading
to a strong fragmentation and isolation of the remaining
heath patches. The consequences of this process on the community have been studied and different responses for ecological groups recorded (e.g. Piessens et al., 2005). Pollution,
mainly from nitrogen (N) deposition, is a major driver of
heathland conversion to grasslands and has received much attention for this reason (Heil and Diemont, 1983; Moore, 1995;
Måren and
Vandvik (2009)
Roem et al.
(2002)
Alonso et al. (2001)
Werkman and
Callaghan (2002)
Piessens et al.
(2004, 2005, 2006),
Piessens and Hermy
(2006)
Beier et al.
(2004)
Beier et al.
(2004)
Mohamed et al. (2007), Härdtle
et al. (2009), von Oheimb (2010)
Power et al. (1998, 2001),
Manning et al. (2005)
Mitchell et al. (1997, 1999, 2000),
Rose et al. (2000), Manning et al.
(2005), Carvalheiro et al. (2010),
Hooftman and Bullock (2012)
Mikkelsen et al. (2008),
Larsen et al. (2010),
Albert et al. (2011a,b)
Beier et al.
(2004)
Cristofoli et al.
(2010a,b)
Mobaied et
al. (2011)
Tsaliki and
Diekmann
(2010, 2011)
Lonati et al. (2009),
Ascoli and Bovio (2010)
Calvo et al.
(2005, 2007)
Carrillo-Gavilán
et al. (2011)
Canals and
Sebastià (2002)
400
0
Beier et al. (2004),
Sardans et al. (2008a,b,c),
Prieto et al. (2009)
km
F I G . 2. Location of study sites of selected papers performing field surveys or experiments discussed in this review. For clarity, some references by similar teams
of researchers from the same localities have been omitted. Beier et al. (2004), for example, stands also for Schmidt et al. (2004), Emmett et al. (2004), Wessel
et al. (2004), Peñuelas et al. (2004), Gorissen et al. (2004) and Llorens et al. (2004a).
Fagúndez — Heathlands confronting global change
Britton and Fisher, 2007), while increasing CO2 concentrations
and climate change have shown a high variability whenever
measured across Europe, with several experimental studies
performed during the last decade (e.g. Peñuelas et al., 2004;
Albert et al., 2011a, b). As heathlands are not particularly
prone to species invasions, they are rarely referred to in that
context (e.g. Carvalheiro et al., 2010), but species such as
Rhododendron ponticum are recognized as important disruptors of heathland diversity and structure. Management
(i.e. resources exploitation) is in turn a very important issue
for heathland maintenance, and is related to all other drivers
and key processes such as fire occurrence (e.g. Mitchell
et al., 2009).
The effects of these drivers have been measured using a
wide range of traits, the most common of which being cover
of the key shrub species in relation to trees or grasses (e.g.
Alonso et al., 2001; Britton et al., 2003). Other studies
measure plant diversity on heathlands (e.g. Britton and
Fisher, 2007; Calvo et al., 2007) and/or focus on target
species considered as to be restricted to heathland habitats
(e.g. Kleijn et al., 2008; Tsaliki and Diekmann, 2011).
Different physiological variables have also been measured as
direct indicators of plant response to drivers, such as photosynthesis and respiration rates (e.g. Albert et al., 2011a, b), or development during key steps of the plant’s life cycle, such as the
soil seed bank and seedling survival (e.g. del Cacho et al.,
2012). Changes in plant contents of different nutrients have
been related to soil composition, which is an indirect
measure of habitat quality (e.g. De Graaf et al., 2009), considering parameters such as acidity and nutrient availability
(Kleijn et al., 2008). Moreover, soil biota is also an important
indicator of habitat quality (Payne et al., 2011), including the
crucial presence and diversity of ericoid mycorrhiza (Read
et al., 2004). Changes in ecosystem services can also be
used as a measure of heathland condition (e.g. Wessel et al.,
2004) as well as biotic interactions such as pollination activities (Forup et al., 2008) or food webs (Carvalheiro et al.,
2010).
The key threats to heathland ecosystems are now relatively
well known, but less information is available on how these
effects will interact and whether synergistic, antagonistic or
additive interactions will result from its confluence in future
scenarios. Information on the evolution of these drivers from
past times is also scarce but necessary for understanding
on-going processes and their possible future effects. Here,
the main factors influencing heathland conservation and biodiversity are reviewed, and a first approach to a holistic
view, considering past processes, recent drivers and projections for the future of heathland ecosystems is outlined.
O R IG IN A ND D E VE LO P M E N T O F P R E S E N T -D AY
H E AT H L A ND S IN E UROP E
Heathland flora evolved during the Mesozoic, a particular
evergreen sclerophyllous vegetation adapted to dry environments and poor soils, especially with limited phosphorous
(P) availability (Specht, 1979). Sclerophyllous characteristic
elements from fossil records in Europe have been dated from
up to 40 Mya BP (million years before present), a period of subtropical climatic conditions in the northern hemisphere
153
(Dallman, 1998). During the cold and warm alternating
phases of the Pleistocene, heathland vegetation dominated
the intermediate periods between woodland expansions at the
warmest moments and the cryoxeric grasslands of the colder
phases (Daniau et al., 2010). Traits such as the ericoid
leaves developed as an efficient adaptation to the physiological
water stress derived from low temperatures. Evergreen sclerophyllous shrubland communities also became adapted to
the Mediterranean climate, which appeared on earth only
approx. 3 Mya BP, characterized by the arid conditions of the
warm season (Mariotti and Dell’Aguila, 2012). Typical heath
components such as the Ericeae (Calluna, Daboecia and
Erica) have undergone different colonizing and retreat episodes across the northern European region concurrently with
the climatic shifts that took place during the late Tertiary. As
an example, Erica scoparia shows narrow genetic variation
among the continental Mediterranean populations compared
with Macaronesian populations (Désamoré et al., 2012). This
suggests that the present-day distribution is the result of different episodes of expansion and retreat, and a recent recolonization which probably occurred at the end of the last glacial
period. This process of rapid recolonization also occurred
with the tree heath (Erica arborea), which originated in the
East African mountain region approx. 5 – 2 Mya BP and colonized the Mediterranean and adjacent areas in Europe
several times, with the Iberian Peninsula serving as
late-Tertiary refuges (Désamoré et al., 2011). However, a
rapid recolonization during the Holocene did not come from
these refuges but from eastern African/Arabian populations.
Southern populations of Calluna vulgaris (hereafter Calluna)
also show a higher allozyme variation than northern ones, suggesting the existence of refuges and rapid south – north recolonization after the ice retreat (Mahy et al., 1999).
The last glacial period ended approx. 10 000 years BP and a
general expansion of broadleaf deciduous forests took place,
replacing open habitats such as heathlands. Heathlands have
been present in Europe during the whole Holocene period, although changes in their extension and dominance occurred in
association with climatic shifts and human occupancy.
Holocene pollen records show a general presence of Calluna
along its entire modern distribution, only limited by ice-cover
at the northern-most region of Europe (Fig. 3). From approx.
5000 years BP there is a general increase in the Calluna –
Erica pollen records throughout Europe (Prøsch-Danielsen
and Simønsen, 2000; Karg, 2008; Carrión et al., 2010),
which coincides with human settlements and the beginning
of a generalized deforestation process. The heathland extent
reached its peak during the Bronze age and the early Iron
age, approx. 4000 – 2000 years BP in western Europe (Webb,
1998; Odgaard and Rasmussen, 2000; Karg, 2008; Nielsen
and Vad Odgaard, 2010), although its expansion was highly
asynchronous: some areas were not transformed into heaths
until the Middle Ages, approx. 1000 years BP, mainly
uplands (Muñoz-Sobrino et al., 1997; Kooistra, 2008; Olsson
and Lemdahl, 2009; Sköld et al., 2010), a period at which lowlands were losing heath extent by transformation into agrosystems (Fyfe et al., 2003). The strong leaching process after
the removal of large wooded areas resulted in the acidification
and impoverishment of the soils, favouring heathlands. The
heathland landscape became dominant in many parts of
154
Fagúndez — Heathlands confronting global change
4000–0 years BP
0
8000–4000 years BP
400
km
12000–8000 years BP
14 000–12 000 years BP
Actual distribution
F I G . 3. Holocene pollen records and actual distribution of heather (Calluna vulgaris). Except for the northern-most localities during the late Würm glacial
period (14 000 years BP) which were covered by ice, a similar distribution for the species during the Holocene after ice retreat is shown. Pollen data were extracted
from the European Pollen Database (http://www.europeanpollendatabase.net). Present-day localities were extracted from GBIF (http://data.gbif.org).
western and southern Europe until a generalized decline
started around 400 years BP, but most evidently since the
18th century as a result of land-use changes related to the industrial revolution, urban development and intensification of
forestry and agriculture practices (Stevenson and Thompson,
1993; Prøsch-Danielsen and Simønsen, 2000; Berglund
et al., 2008; Nielsen and Vad Odgaard, 2010).
There is evidence of human use of heathlands since the
Bronze Age in western Europe (Karg, 2008). In the northern
oceanic stands such as the Faroe Islands, pre-human settlement
heathland expansion has been related to climatic shifts
(Hannon et al., 2005), but is mostly explained by
human-induced clearances of woodlands (Kooistra, 2008).
The traditional management of heathlands was similar for
Fagúndez — Heathlands confronting global change
approx. 2000 years, and included different grazing systems,
peat and turf cutting and removing, mowing to provide cattle
with winter feeding, periodical burning to promote young regrowth and wood extraction for different uses such as the
manufacture of smoking pipes or as a source of charcoal
(Webb, 1998; Izco et al., 2006; La Mantia et al., 2007). In
the Mediterranean, fluctuating pollen records show complex
vegetation pattern changes during the Holocene (Carrión
et al, 2010; Kouli, 2011). A higher presence of the sclerophyllous shrub vegetation extension may be explained by an increasingly arid period during the mid-Holocene at approx.
5000 years BP (Sadori and Narcisi, 2001; Fletcher, 2007;
Carrión et al., 2010; Kouli, 2011). However, the
human-induced increase in fire frequency was probably a
major cause of evergreen Quercus woodland depletion and
the promotion of shrublands in Mediterranean areas
(Fletcher, 2007; Colombaroli et al., 2008) and, therefore, a
major force of landscape shaping. Fire is also strongly
related to the human management of northern and western
heathlands, such as those at the Baltic coasts (Savukynienė
et al., 2003; Olsson and Lemdahl, 2009), but differences
with the Mediterranean region rely on the frequency of fires
and the extent of the burned areas.
An understanding of the role of past human uses and climate
on present-day heathlands may provide a basis for predicting
future changes in relation to land-use changes, global
climate change and other drivers. The implementation of this
information on modelling the future of dynamic vegetation
such as that of heathlands can be a powerful tool that is yet
to be fully explored (Marinova et al., 2012).
D R I V E R 1 : H A B I TAT DE S T RU C T I O N
A ND F R AG M E N TAT I ON
Heathlands were once a widely extensive habitat in a humanmade cultural landscape, but a general substitution of traditional by other land-use management types has taken place
in recent times. A general decrease in heathlands throughout
temperate Europe during the last two centuries has been extensively reported. In Belgium, only 1 % of the heathland area
from 1775 remains (Piessens et al., 2004) and in England
only 16 % of the lowland heathland extensions from the
early 19th century are left (Alonso, 2004), as most of the
area has been transformed by afforestation, urban development
or agricultural intensification. The heathland area has declined
by 60 –70 % in Sweden and Denmark from 1900 to 1960, and
by 95 % in The Netherlands during the same period (Newton
et al., 2009). Upland heathlands (moorlands) have also
declined significantly, by as much as 90 % in Ireland and
England during the last 200 years (Thompson et al., 1995).
On the other hand, in certain areas like montane
Mediterranean regions, heathlands have increased their extension due to land abandonment (Mouillot et al., 2005; Azevedo
et al., 2011). Besides the obvious negative effect of a general
reduction of heathland’s extension, habitat fragmentation and
isolation are a key aspect of heathland conservation
(Cristofoli et al., 2010a). The remaining areas of heathlands
throughout Europe are strongly fragmented, with remnants alternating with other natural or mainly managed habitats
155
(Veitch et al., 1995; Piessens et al., 2006). In Dorset,
South-West England, a general decline of lowland heathlands
has been documented since the 19th century (Webb, 1990;
Veitch et al., 1995; Rose et al., 2000; Hooftman and
Bullock, 2012). An accurate analysis of land-use change performed by Hooftman and Bullock (2012; Fig. 4) showed an
overall habitat loss of 56 % for heathlands, but also a decrease
in the mean patch size of 88 %, an increase in the distance
between patches, and a lower heathland occupancy on the surrounding area of each heathland patch. Heaths have mainly
been transformed to woodlands (broadleaf and coniferous)
and to more anthropogenic land uses (arable land, improved
grasslands or built-up areas). Overall, semi-natural habitats
decreased by 74 % and were substituted by anthropogenically
made habitats in the region. Therefore, heathlands have
decreased in extension, fragment size, connectivity and
species’ suitability with regard to the surrounding habitats.
Habitat fragmentation affects diversity and
community structure
One of the negative effects of fragmentation generally
described in the literature is regarding species richness.
Piessens et al. (2004) stated that plant species richness of
heathland areas is affected by fragmentation, but that this
negative effect relies mainly on isolation and not on patch
size. However, species richness is not a clear indicator of
habitat quality as a response to habitat fragmentation
(Kemper et al., 1999; Piessens et al., 2005) because it may
be strongly influenced by an increasing edge effect, since
species richness may be augmented by generalist or invasive
species from external habitats, which strongly depends on
the perimeter/area ratio, i.e. fragment size and shape. In contrast, community composition, measured as a shift in biological traits such as dispersal mode or pollination syndrome, may
better explain habitat changes. According to Piessens et al.
(2005), a given biological trait may determine the survival
of a certain species for an isolated heath area. Plants with a
persistent soil seed bank showed the highest survival rates
for isolated patches. Other traits, such as long-distance dispersal abilities, seem to play a weak role in patch colonization.
The soil seed bank is a key factor in the maintenance or recovery of European heathlands (Willems, 1988; Aerts et al., 1995;
Valbuena and Trabaud, 2001; Måren and Vandvik, 2009). A
persistent soil seed bank is characteristic for most of the dominant species in heathlands such as Calluna or Erica
(Thompson and Band, 1997). Whenever the seed bank has
been removed or depleted, heathland recovery will be severely
retarded (Aerts et al., 1995; Pakeman and Marshall, 1997;
Mitchell et al., 1998; Degn, 2001; del Cacho et al., 2012). If
fragmentation takes place, the seed bank will play a key role
in the long-term survival of the species (Bossuyt and Hermy,
2003; Cristofoli et al., 2010b). The rarest species are,
however, under-represented in the seed bank compared with
their occurrence in the community (Bakker and Berendse,
1999) and may therefore undergo local extinction even if recovery from the seed bank is achieved. Again, rare species
are more sensitive to processes that may only marginally
affect the community structure.
156
Fagúndez — Heathlands confronting global change
1930s
Mesotrophic grasslands, rough
Mesotrophic grasslands, managed
Calcareous grasslands, managed
Calcareours grasslands, rough
Woodlands
Coniferous woodlands
Arable
Improved grasslands
50
Acid grasslands, managed
Acid grasslands, rough
Heathlands
Fens, marsh and swamps
100
Heathlands 1930s (n = 242)
Heathlands 2000 (n = 952)
Connectivity (%)
30
20
10
Heathlands 1930s
Heathlands 2000
60
40
>100
75–100
50–75
25–50
10–25
7·5–10
5–7·5
2·5–5
1–2·5
0·75–1
0·5–0·75
20
0·25–0·75
0
Inland bare ground
Built-up areas and roads
Littoral sediments, incl. saltmarshes
Water
80
40
0·06–0·25
% of all connected areas
2000
0
100
200 300 400 500 600
Distance from source (m)
700
Field site size classes (ha)
F I G . 4. Land-use changes in Dorset, UK, and fragmentation effects on heathland habitats. Several types of semi-natural and anthropogenic habitats in the area of
Dorset (251 422 ha) were mapped from 1930s and 2000 (top of the figure, with key). Fifty-six per cent of total heathland area ( purple) was lost in the interval and
transformed to woodlands, anthropogenic land-uses or grasslands. At the bottom of the figure, two graphics show the effects of habitat fragmentation. On the left,
the graph shows a loss of large patches (.10 ha) and an increase in small patches (,2.5 ha). Note that the total number of patches increases significantly. On the
right, the graph shows that the probability of seed dispersal from a given patch to other patches (i.e. connectivity, expressed as a percentage) at different distances
decreases significantly because of habitat loss and isolation. Modified from Hooftman and Bullock (2012).
Population genetics and fragmentation
In a strongly fragmented landscape, the potential geneticdiversity loss associated with species population isolation
will also negatively affect biodiversity. Genetic drift, inbreeding depression and allelic impoverishment will eventually
cause the extinction of a given species, and an overall geneticdiversity loss. Rare species, such as Gentiana pneumonanthe,
did not show a direct relationship between genetic variation
and population size in Belgium, probably because isolation
has taken place in recent times, and the species is a long-living
perennial (Raijmann et al., 1994). However, it is clear that
there is a low gene flow between populations, meaning that
inbreeding depression from isolation will eventually take
place.
Three recent papers deal with population genetics related to
habitat fragmentation of three different heathland specialist
species: the ground beetle (Drees et al., 2011), a heath bee
(Exeler et al., 2010) and the smooth snake (Pernetta et al.,
2011). In a study of the genetic-diversity pattern among
smooth snake (Coronella austriaca) populations in a southern
England landscape dominated by pine plantations with heathland remnants, Pernetta et al. (2011) found a significant correlation between allozymic differentiation and geographical
distance, the so-called isolation-by-distance model, which corresponds to a continuous genetic population structure. This
suggests that the smooth snake is capable of using corridors
such as roadside verges to range between heathland patches
and therefore minimizing the genetic diversity loss for the
species. They also found a sex-biased population structure
explained by the sedentary behaviour of females in contrast
to a higher capacity of dispersal by males, promoting gene
flow between populations.
Fagúndez — Heathlands confronting global change
A strong gene flow between populations was also found for
a heathland specialist bee (Andrena fuscipes) in north-western
Germany (Exeler et al., 2010). Again, the isolationby-distance model better explains the genetic differentiation
among patches in a common structure for a long-distance
flying species with no effective barriers to gene flow. The relation to geographic distance was, however, weak and it is
understood that other non-measured ecological variables may
play a role in the population structure of the species. A study
of the ground beetle (Poecilus lepidus) populations established
that the main factor to explain genetic variability was the
heathland patch size where the species was present (Drees
et al., 2011). Patch size was positively correlated to allelic
richness and heterozygosity. In contrast, Exeter et al. (2010)
found no correlation with fragment size, although a recent
bottleneck event was deduced at the smallest recorded heathland patch of only 300 m2. A minimum patch size is thus critical for habitat quality, although this effect greatly varies
among ecological groups such as the flying bees or the sedentary beetles. Drees et al. (2011) concluded that heath
fragments of ,50 ha are ineffective for the long-term conservation of genetic variability of the ground beetle. Heathlands
are dynamic ecosystems with recurrent episodes of disturbance
such as fire or management changes. Thus, a minimum effective population size must be calculated over predictions of fluctuations with strong population declines to avoid the genetic
erosion of specialist species. On the other hand, the population
size of specialist species is not always in accordance with
patch size, as this correlation was not observed in two heathland Genista species studied in Germany by Tsaliki and
Diekmann (2010, 2011).
Matrix and corridors are determinant in the
fragmentation effect
In addition to fragment size and isolation, the existence of
corridors and the nature of the matrix strongly affect the
habitat quality and species’ populations of heathland remnants.
Piessens et al. (2006) studied a set of heathland patches with
different land uses in the adjacent areas. A strong edge
effect was recorded in heath patches adjacent to the different
land uses, reflected in plant community composition changes
and soil properties. The abundance of heather (Calluna)
decreased towards the edge in all cases, and it was mostly
replaced by grasses. Available P, nitrates and soil pH, as eutrophication indicators, increase towards the edge. The
authors determined the same average distance from the
margin as the edge effect limit for all bordering land uses.
However, the negative effect within this belt was higher for
heath patches adjacent to cropland, with higher eutrophication
in the soil. The nature of the matrix (i.e. surrounding landscape) in a heathland mosaic was also important for a wide
range of flower-visiting insects (Dupont and Nielsen, 2006).
Patch size was not correlated with species richness but community composition and, therefore, plant – insect assemblages
are affected by fragmentation. Small remnants have a higher
diversity, although this is explained by the occurrence of generalist species against heath-specialists. External habitats are
therefore enriching patches with a higher edge percentage,
which highlights the importance of the surrounding habitats.
157
Conversely, the need for patch connection through corridors
to maintain species in heath patches has been stated for different ecological and taxonomic groups (Noordijk et al., 2011;
Pernetta et al., 2011). The permeability of the surrounding
areas and the existence of suitable corridors, will determine
the effective connectivity between patches.
The classical approach to island biogeography, which
involves considering isolation-by-distance alone, is not suitable for a complex landscape such as heathlands. The matrix
composition is therefore critical (Ricketts, 2001). An approach
to the fragmentation effect over heathland landscapes should
consider the surrounding habitat types in terms of effective
corridors for key species such as pollinators (Webb and
Vermaat, 1990). For example, a complex interaction exists
between populations of the Dartford warbler (Sylvia undata)
in Dorset and the different surrounding habitats (van der
Berg et al., 2001).
The heathland plant community is highly resilient to fragmentation and isolation, with a resultant low diversity loss
(Piessens et al., 2005). It also means that a big effort is
needed to restore the heath plant community after habitat destruction (Cristofoli et al., 2010b). It may be the case that
local extinctions are taking place over a long time period
and, therefore, there is an actual extinction debt in heathland
ecosystems (Piessens and Hermy, 2006; Cristofoli et al.,
2010a), but this theoretical effect needs testing.
Fragmentation affects other taxonomic groups that may act
as dispersers or pollinators (Webb and Vermaat, 1990; van
der Berg, 2001; Noordijk et al., 2011), but plant reproduction
success is weakly influenced by population size (Tsaliki and
Diekmann, 2011). However, other studies point to an interaction effect of eutrophication and fragmentation that may
affect the reproductive success of rare heath species (Vergeer
et al., 2003).
Soil conditions are related to habitat quality independent of
fragment or population size (Piessens et al., 2004; De Graaf
et al., 2009; Tsaliki and Diekmann, 2010). Thus, species diversity and habitat quality of heathland patches are, as a general
rule, little or not dependent on patch size as long as the key
environmental factors (mainly soil properties) are maintained.
Land-use changes may strongly affect the habitat quality of the
surrounding landscape (i.e. matrix and corridors) resulting in a
higher environmental pressure to heathland remnants.
D R I V E R 2 : PO L L U T I O N A N D E U T ROP HI CAT I O N
One of the main characteristics of heathlands is that they
develop over poor soils, with low values for most nutrients
and pH (Mitchell et al., 1999; De Graaf et al., 2009).
Although some heathlands occupy basic soils, especially
around the Mediterranean basin (Canals and Sebastià, 2002;
Sardans et al., 2008a – c) most temperate heaths live on acid
soils (De Graaf et al., 2009). The main limitations to plant
growth in these conditions are exchangeable P, potassium
(K) and N (Bobbink et al., 2010). Other edaphic chemical variables that limit plant growth in heathlands are aluminium toxicity, low contents of minor nutrient concentrations such as
calcium (Ca+) or magnesium (Mg+), or the ratios between
these nutrients (Kleijn et al., 2008). Moreover, different hydrological conditions such as slow-draining soils of wet
158
Fagúndez — Heathlands confronting global change
heathlands diminish the acidification process and, whenever
drained, result in a high acidification, higher soil respiration
rates, and lower N availability (Jansen et al., 2004). Soil conditions are particularly important in maintaining heathland
habitat quality, as stated for habitat fragmentation.
Several pollutant sources globally affect the temperate
biomes, mainly through the atmospheric deposition of
sulphur (S) or N. Atmospheric N deposition is today considered one of the main drivers of biodiversity loss in northern
temperate biomes (De Schrijver et al., 2011; McClean et al.,
2011). Reduced and oxidized forms of organic nitrogen have
increased the deposition rates of N and will most likely continue for the following decades (Tilman et al., 2002). High atmospheric N depositions result in a higher N concentration in
the leaves of ericaceous species and a subsequent increase in N
: P and N : K ratios in leaf tissues (Britton et al., 2008) which
may result in a higher herbivorous pressure over the species.
Litter deposition then adds to organic soil N inputs where
higher decomposition rates are recorded, which in turn
decreases the C : N ratio (Power et al., 2001). Nitrogen leaching increases, eventually affecting water quality, especially in
upland landscapes (Britton et al., 2008), although this has not
always been observed (Power et al., 2004). The complete cycle
is thus affected on the system, with a strong negative effect
over the habitat quality.
Overall species richness is affected, but mainly specialist and rare
species
Species richness decreases with increasing N deposition for
all vegetation types, but in heathlands this negative effect is
among the highest (Maskell et al., 2010). Species richness,
as an indicator of habitat quality, decreases when eutrophication occurs (De Graaf et al., 2009), especially with a loss of
heath specialists (De Graaf et al., 2009) and rare and endangered species (Kleijn et al., 2008). Overall, acidity and Al :
Ca ratio determine community composition and species richness, but NH4 concentration and the NH4 : NO3 ratio are the
key factors affecting rare species that show a much narrower
tolerance for changes in these parameters (Kleijn et al.,
2008). Species composition may also be altered by high N
availability, by shifting from an N-limited to a P-limited or
an NP-co-limited environment (Power et al., 2001; Härdtle
et al., 2009; von Oheimb et al., 2010), although this effect
needs further experimental work (von Oheimb et al., 2010).
Eutrophication also promotes invasion by generalist species
with a negative effect on heath specialists (Mitchell et al.,
2000). However, markedly different results may be observed
in different heath communities, since N addition resulted in
higher species richness in an alpine Cantabrian heathland
(Calvo et al., 2007). Perennial herbs were enhanced, but no
negative effect over ericoid shrubs was recorded, suggesting
that the species loss associated with N addition needs to be
ascertained locally.
Atmospheric N deposition alters heathland
community structure
The increase in atmospheric N depositions affects community structure and composition in heathlands by enhancing
grass species over shrubs (Heil and Diemont, 1983; Roelofs,
1986; Heil and Bobbink, 1993; Moore, 1995; Bakker and
Berendse, 1999; Britton et al., 2001; Terry et al., 2004).
Nitrogen addition increases shoot growth, biomass and flowering of ericaceous species such as Calluna (Power et al., 1995;
Gidman et al., 2005; Britton and Fisher, 2008; von Oheimb
et al., 2010), but there is a resultant overall negative effect
due to competition with grass species such as the purple
moor grass (Molinia caerulea) against heather (Calluna), especially at the seedling stage when Molinia more efficiently uses
N inputs (Friedrich et al., 2011). This is of particular importance since germination and seedling establishment is crucial
for heath development (Power et al., 2001; Roem et al.,
2002; Britton et al., 2003; Friedrich et al., 2011). Seedling establishment of Deschampsia flexuosa also increased significantly with higher N deposition values, especially when combined
with uncontrolled fire (Power et al., 2001). The switch from
heather dominance to heath grasses is therefore conditioned to
a certain level of disturbance in the community (Power et al.,
1998; Britton and Fisher, 2007) from intensified management
(grazing or cutting), accidental fire, or by heather beetle
(Lochmaea suturalis) attacks. Any of these factors can break
up the canopy cover of the shrub, promoting the germination
of grass seeds (Terry et al., 2004). In addition, the plantcommunity changes observed with higher N availability may
serve as indicators for negative effects on soil biota (Payne
et al., 2011).
D R I V E R 3 : I N C R E A S I N G CO 2 A N D CL I M AT E
C HA NG E
Increasing atmospheric CO2 as a result of human intervention
is affecting the climate by global warming, changes in precipitation amounts and annual distribution, as well as extreme climatic events. Over the last century, an increase in the mean
annual temperature of 0.9 8C has been recorded in continental
Europe, and future scenarios predict an increase of the mean
annual temperature of 0.1 to 0.4 8C per decade for the 21st
century (Harrison et al., 2006; Alcamo et al., 2007). The
mean annual precipitation is predicted to increase in northern
Europe and decrease in southern Europe, but the summer
drought will be drier and last longer throughout the region
(Alcamo et al., 2007). Global change is affecting biological
systems throughout the world according to a wide set of indicators, including plant communities (Rosenzweig et al., 2007).
Direct consequences are changes in phenology, plant productivity or metabolic rates, and these effects determine the species
distribution, competition and community structure (Peñuelas
and Filella, 2001; Walther, 2003; Feehan et al., 2009). These
changes will probably disrupt ecological functions, species
interactions and ecosystem services (Fitter and Fitter, 2002;
Memmott et al., 2007; Feehan et al., 2009). The projected biodiversity losses in alpine and boreal biomes are among the
most directly affected by global warming, whereas others
such as tropical forests will be mainly affected by land-use
changes (Sala et al., 2005). In contrast, biodiversity loss in
temperate heathlands seems to be balanced between these
two main drivers, and is highly dependent on their interaction
Fagúndez — Heathlands confronting global change
effects (Sala et al., 2005; de Chazal and Rounsevell, 2009).
The diversity and structure of heathlands will be affected by
climatic changes, e.g. arctic-montane vascular plant species
in oceanic heaths are likely to decrease in extent due to predicted temperature rises, which will also favour the expansion
of more generalist competitive species (Walther et al., 2005).
These will affect bryophyte-rich heaths, which will also
reduce in extent due to a greater seasonality of rainfall
(Hodd and Sheehy Skeffington, 2011).
Climate change effects have been measured in heathlands
from historical records (Berry et al., 2002; Peñuelas and
Boada, 2003), manipulating experiments (Beier et al., 2004;
Jentsch et al., 2007; Mikkelsen et al., 2008; Albert et al.,
2011a) and the implementation of predictive models
[Coquillard et al., 2000; Berry et al., 2002; also Midgley
et al. (2002) in fynbos]. This large set of published data provides the basis for understanding recent changes in heathlands
associated with climate change, and to predict, to a certain
extent, the future of heathlands under global change scenarios.
Climate-change effects over phenology of heath species are
heterogeneous
The distribution of plant-cycle events throughout the year
(i.e. phenology) is markedly dependent on temperature and
precipitation (Peñuelas and Filella, 2001; Walther, 2003;
Feehan et al., 2009). Phenology is a key ecological process
that involves plant growth and ecosystem functions such as
plant – pollinator interactions (Memmott et al., 2007).
Moreover, shifts in plant phenophases are one of the most accurate biological indicators of climate change (Walther, 2003;
Jentsch et al., 2009). In heathlands, drought alters the flowering time span and the vegetative growth period of common
shrub species (Peñuelas et al., 2004), but these are less
affected by warming. The growth period may last longer as
early spring temperatures rise, but this does not necessarily
result in higher productivity because overall growth is
limited by lower water availability in drier regions (Bernal
et al., 2011). In a Mediterranean heathland experiment,
warming and drought benefitted Erica multiflora over
Globularia alypum (Llorens et al., 2004b). Both are sclerophyllous shrub species common in basic poor soils of the
western Mediterranean basin, but E. multiflora is a resprouter
species with a thick lignotuber to store reserves (Sardans
et al., 2006) and therefore less dependent on short-term
resources, such as soil-water availability. This apparent disruption of species competition due to climatic changes must be
considered in the specific conditions of the Mediterranean
climate, which is characterized by a strong inter-annual variation and unpredictability (Mariotti and Dell’Aguila, 2012).
Globularia alypum, a species with both seeder and resprouter
strategies (Saura-Mas and Lloret, 2007) may compensate for
changes with a higher ecological plasticity.
Flowering time is affected in a species-specific manner,
either delayed or advanced by warming or drought whenever
extreme climate events take place (Jentsch et al., 2009).
However, there is a strong interaction between the plant community and phenological shifts. A drought event significantly
delayed mean Calluna flowering peak and extended flowering
time, but only in complex stands with high functional diversity
159
(Jentsch et al., 2009). Delays and increasing lengths of flowering time occur in heath species under drought treatments
(Peñuelas et al., 2004), while general patterns show an
overall earlier bloom in a warmer climate (Fitter and Fitter,
2002; Peñuelas et al., 2002). The delay on flowering of the ericaceous species is stronger under a Mediterranean climate
(Peñuelas et al., 2004; Prieto et al., 2008) but it is either unaffected or slightly brought forward in temperate heaths.
Water availability during the flowering season is the main limitation for flowering in the Mediterranean stands, whereas
northern heaths are less affected by drought and are favoured
by mild spring temperatures.
Climate changes nutrients cycles and limitants
in heathland soils
Global warming and rising CO2 concentration increases
soil biological processes, enhancing litter decomposition and
nutrient availability. On the other hand, water stress lowers
biotic activity in the soil (Peñuelas et al., 2004; Sowerby
et al., 2008). Complex site-specific interactions are found for
warming and drought in soil respiration rates (Sowerby
et al., 2008). In oligotrophic plant formations such as heathlands, higher rates of mineralization may increase the N : P
ratio in the soil and eventually drive a change from
N-limited to N-P limitation (von Oheimb et al., 2010). This
process is accelerated by increasing air temperature whenever
water stress is kept under critical values, mainly in
non-Mediterranean climates (Schmidt et al., 2004). The negative effects of high N deposition on habitat quality is thus conditioned by climate (Britton et al., 2001; Schmidt et al., 2004;
Sardans et al., 2008a – c).
Plant physiological response to CO2, warming and drought
differ in magnitude and direction
Increasing CO2 atmospheric concentrations result in higher
plant productivity and increasing rates of photosynthesis and
respiration (Albert et al., 2011a, b), which is enhanced by
warming but reduced by drought in experimental treatments
on a Calluna Atlantic temperate heath. A non-additive response to interactions between these treatments was recorded
for photosynthetic performance (Albert et al., 2011b), suggesting complex interactions between warming, drought and rising
CO2 levels.
Warming treatment resulted in a higher plant productivity in
another controlled experiment on different European heathlands (Peñuelas et al., 2004; Wessel et al., 2004), while
drought reduced plant productivity, respiration rates and photosynthesis, mainly in Mediterranean heathlands. However, the
interaction between warming and drought factors was not
studied in this experiment.
Sclerophyllous shrubs are less affected than grasses with
regard to photosynthesis performance (Niinemets et al.,
2011), but the production of secondary compounds such as terpenoids lowers under drought treatments, which is an effect
that is also observed under increasing N addition (HoflandZijlstra and Berendse, 2009). Warming and drought increased
herbivore attack by the heather-beetle in temperate Calluna
stands (Peñuelas et al., 2004), probably due to the depletion
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Fagúndez — Heathlands confronting global change
of phenolic compounds which constitute a defence against herbivores. Moreover, a decline in the allelopathic compounds of
heather litter increases the mineralization rate and promotes
germination of other species. However, positive competition
of heather against heath graminoid species relies more on its
higher efficiency under nutrient-poor soil conditions than on
the allelophatic properties of its litter (Berendse et al, 1994;
Hofland-Zijlstra and Berendse, 2010).
Plant diversity is altered mainly due to effects at the
seedling stage
Plant responses to climate change are asymmetric, with differences across regions and ecosystems (Walther, 2003).
Moreover, different species co-occurring in a single habitat
may be affected differently by climatic shifts (Llorens et al.,
2004a, b). Species richness, as the general sum of species occurring at a site, is therefore a complex measure to indicate
habitat changes and must be contrasted with species composition. Plant diversity changes were measured under experimental manipulation for warming and drought over different
European heathlands (Peñuelas et al., 2007). No significant
differences due to warming or drought were recorded in the
7-year experiment across sites. However, recovery after fire
in a Mediterranean site of the north-east Iberian Peninsula
decreased under both treatments, but mainly under the
drought conditions. Contrary to biomass, plant cover or net
primary productivity, which responded significantly to treatments and time, plant richness was unaffected by climatic
changes. Disturbances such as wildfires will delay regeneration
processes and decrease plant richness at least during the early
stages in Mediterranean stands subjected to increasing aridity
(i.e. summer drought). Heathland regeneration relies both on
resprouting and seed germination (Ojeda et al., 2001; Lloret
et al., 2004). Seedling diversity lowers under drought and
warming treatments (Lloret et al., 2004; Prieto et al., 2009;
del Cacho et al., 2012). Long-term changes on plant richness
and diversity will occur due to climatic shifts, not over
mature stages but during early phases of vegetation recovery
from disturbance such as wildfires.
DR IV E R 4 : V E GE TAT IO N DYN AM I C S AN D
MA NA GEME NT
Most heathlands are commonly described as successional
vegetation types, intermediate steps within dynamic succession series that end in the climax vegetation (Mallik, 1995;
Pakeman and Nolan, 2009). Potential drivers of change over
heathland vegetation are therefore the natural forces that
affect community structure towards a highly mature vegetation
type such as broadleaf or coniferous woodlands. Changes may
also point towards a regression in natural succession, by means
of a general loss in shrub cover and substitution by pastures or
bare soils. In other cases, succession takes place favouring
heathlands. Mountain meadows, where summer grazing was
the traditional management, are now switching towards a
dwarf shrub heath community due to the abandonment of
these practices (Pasche et al., 2004), although this represents
a transition phase that will eventually continue succeeding to
woodlands.
However, natural succession is a complex transition between
community types conditioned by climate, soil properties,
species seed bank or dispersal, and local management, and
has therefore been described as site-specific (Britton et al.,
2000; Manning et al., 2005; Britton and Fisher, 2007). This
is in accordance with other views that discuss the generally
accepted phytosociological approach of a straight succession
towards a single potential vegetation type (Carrión and
Fernández, 2009). Mitchell et al. (1997, 1999, 2000) described
more than one successional trajectory (i.e. transitions between
communities) in lowland heaths (Fig. 5). Soil properties
explained most of the community variation on a canonical
multivariate analysis. Higher pH values and available Ca and
Mg is associated with birch woodlands, while higher N concentrations (NH4 and NO3) are found in sites invaded by
bracken. Low pH and high N values were found on plots
invaded by Pinus sylvestris and Rhododendron ponticum. In
a different study, a transition towards a Betula-dominated
scrub was limited by both seed availability of the invasive
species and the heath age and structure (Manning et al., 2005).
Time span is not the main factor for transition between communities, as several biotic and abiotic factors may act as limitants or promoters of succession. Management plays a critical
role in heathland existence, preservation and/or restoration as
it has been considered in the scientific literature (Webb,
1998). European heathlands have been periodically controlled
by grazing, burning or cutting, all of which are actions that
remove plant biomass and prevent natural succession.
Grazing maintains community structure when other factors are
controlled
The use of grazing, generally associated with other management practices such as periodic burning, is one of the main
factors that have shaped the European heathland landscape
for millennia (Thompson et al., 1995; Webb, 1998; Evans
et al., 2006). During the last two centuries, a general transformation process towards a higher intensification has taken
place, leading to higher stocking rates and the change of
mixed grazing by wild and domestic herbivores to a singlespecies grazing use (Gordon et al., 2004). On the other
hand, abandonment of traditional pastoral practices in heathlands implies a transition towards scrub and woods (Marrs
et al., 1986; Webb, 1998; Bokdam and Gleichman, 2000).
Evidence shows that intense grazing results in a loss of
cover of shrub species and an increase in graminoids
(Bullock and Pakeman, 1997; Bokdam and Gleichman,
2000; Newton et al., 2009). These changes are conditioned
by grazing species, stocking rates, management (herding, seasonal free ranging) or in combination with other practices such
as cutting or burning (Bullock and Pakeman, 1997; Newton
et al., 2009; Celaya et al., 2010). The use of different
grazing animals affects heathlands in different ways, promoting transitions towards other vegetation types at different
time rates (Putman et al., 1987; Jáuregui et al., 2009; Celaya
et al., 2010; DeGabriel et al., 2011; Wehn et al., 2011).
There is also a strong interaction between grazing species
and season (Putman et al., 1987; Hester and Baillie, 1998).
For example, cattle use heathers evenly throughout the year,
while other herbivores such as ponies have foraging peaks in
Fagúndez — Heathlands confronting global change
161
Axis 2
4
Na
3
NH4
Heath
B
PS
PA
R
ENV
K
2
NO3
1
LOI
Axis 1
0
–3
–2
–1
Ca
1
0
2
3
4
5
6
Mg
–1
P
pH
–2
–3
–4
F I G . 5. Two-dimensional canonical correspondence analysis biplot showing the relationship between the different successional stages and the soil nutrients on
the Dorset Heaths, UK. The soil nutrient vectors are shown by arrows. Coding for stands are: Heath, open heathland; B, Betula spp. is the major invader; PS,
Pinus sylvestris is the major invader; PA, Pteridium aquilinum is the major invader; R, Rhododendron ponticum is the major invader. Coding for environmental
values (ENV) are: Ca, exchangeable calcium; K, exchangeable potassium; LOI, percentage loss on ignition; Mg, exchangeable magnesium; Na, exchangeable
sodium; NH4, extractable ammonium-N; NO3, extractable nitrate/nitrite-N; P, extractable phosphorus. From Mitchell et al. (2000).
the spring, when the growth of grasses gets higher (Pratt et al.,
1986; Putman et al., 1987).
DeGabriel et al. (2011) compared the plant community
structure and species diversity in Scottish upland heathlands
with the presence or absence of sheep grazing on different
scales. Sheep removal was correlated with a higher grazing
pressure from deer, but an overall lower heather utilization.
Alpha diversity was correlated to the sheep livestock, mostly
due to a switch towards grasslands when the number of
sheep increased. Deer and moderate sheep grazing is the
optimal scenario to maintain community structure and alpha
and beta diversity. Higher diversity values were recorded in
intermediate stages during the transition towards grasslands,
a richer community, whereas absence of sheep may cause a
turn towards the degenerate phase of the Calluna cycle.
Newton et al. (2009) described great variation in the response of communities subjected to different experimental
grazing studies and concluded that it would be difficult to
give an exact general prescription for maintaining heath by
grazing. Grazing is an effective management practice for
heathlands but, in general, it requires to be combined with
other practices such as burning (Barker et al., 2004; Härdtle
et al., 2009). Again, this depends on local conditions and the
particular heathland type. Traditional grazing practices may
be sufficient for heathland maintenance in the Erica scoparia
community in central France (Gachet et al., 2009) but not, for
example, in Calluna stands in north-east Iberian mountain
heathlands (Bartolomé et al., 2000).
Fire may promote both regressive and progressive successional
stages
Fire is a key factor in the existence of most heathland vegetation types (Schwilk et al., 1997; Harris et al., 2011). The
sclerophyllous vegetation under Mediterranean climates, such
as fynbos in South Africa, chaparral in California or
Mediterranean heathlands, are generally fire-prone habitats
with high fire frequencies (Delitti et al., 2005; Mouillot
et al., 2005; Azevedo et al., 2011). Fire regimes range from
large, intense wild-fires to prescribed winter burning performed to maintain heath structure. Successional dynamics
on heathlands are directly correlated to fire regimes
(Sedláková and Chytrý, 1999; Mohamed et al., 2007; Måren
and Vandvik, 2009; Mitchell et al., 2009) and the use of controlled fire is an effective management tool for heathland preservation (Sedláková and Chytrý, 1999; Mohamed et al., 2007;
Davies et al., 2008). On the other hand, a general increase in
fire frequency associated with global warming may benefit
162
Fagúndez — Heathlands confronting global change
certain heath communities against other habitats such as woodlands, particularly under a Mediterranean climate (Mouillot
et al., 2002).
The increasing frequency of fire regimes generally causes a
regression on succession, promoting grasses against shrubs
(Power et al., 2001; Brys et al., 2005; Jacquemyn et al.,
2005). However, recent evidence supports the positive effect
of fire on tree encroachment on heathlands, promoting progressive succession (Borghesio, 2009; Ascoli and Bovio, 2010).
Ascoli and Bovio (2010) studied birch (Betula pendula)
and aspen (Populus tremula) encroachment on Callunadominated heathlands on the Alps piedmont. Fire reduced
the continuous vegetation canopy, removed allelophatic compounds and promoted tree encroachment via the facilitation
of seed germination and seedling establishment. Later, strong
suckering regeneration of aspen resulted in a general loss of
Calluna cover and the presence of characteristic heath
species such as Genista tinctoria or Gentiana pneumonanthe.
It is notable that the effect over Molinia shows the opposite
sign, with higher cover values underneath the tree stands.
Fire therefore causes a regressive succession from heathlands
to pastures, but also promotes positive succession favouring
different tree species. Below critical values, there is a risk
that fire may transform the heathland landscape into a
mosaic of woodlands and pastures. However, this particular response may be associated with this particular heathland type,
and thus new contrast studies are needed in other regions.
A need for combined treatments on heathland management
Several studies have revealed that incorrect policies have
been developed regarding heathland conservation. In
Montseny Natural Park, an isolated heathland region in a
Mediterranean area in eastern Spain, prohibition of fire has
resulted in a general transition of the Calluna heath towards
a Mediterranean forest or bracken invasion (Bartolomé et al.,
2005). Grazing alone has not resulted in the maintenance of
heath but a generalized transition towards scrub and forest.
Bokdam and Gleichman (2000) stated that grazing by freeranging cattle is not enough to prevent heath transition, both
towards grassland or woodland, in Dutch lowland heathlands.
They predict a future landscape formed by a wood – pasture
mosaic unless additional management is implemented, such
as limiting grazing to seasonal and day-time herding.
Sedláková and Chytrý (1999) state that the use of cutting as
an alternative to fire for biomass out-take retards Calluna recovery and enhances grass encroachment. A combined management of prescribed burning and grazing is proposed to
maintain heath structure and diversity. However, cutting may
be an efficient substitute to grazing for biodiversity conservation according to Calvo et al. (2007) who performed a manipulated experiment in an alpine Calluna stand in northern Spain.
A combined fertilization and cutting treatment showed the
highest biodiversity. Graminoid species and forbs were
enhanced, but no competition effect was observed with
Calluna, although the field experiment lasted for 5 years and
projections on long-term process should show cautiousness.
Heathland management is, without doubt, the key instrument for habitat maintenance, but the appropriate practices
to be used are strongly correlated to climate and regional
environmental characteristics. Lessons from traditional practices are especially useful regarding fire frequency, grazing
levels or cutting intervals, and may provide practical information on the conservation plans to be implemented.
D R I VE R 5 : I NVAS IV E E XOT I C S P E C IE S
The presence and spread of non-native species disrupting ecological processes throughout biomes worldwide is considered
one of the main causes of biodiversity loss (Hulme et al.,
2009; Butchart et al., 2010), even though these invasions
are, to some extent, not always a driver but a consequence
of a habitat quality loss triggered by other factors (Didham
et al., 2005). Heathlands are among the habitats with a lower
frequency of exotic invaders and are less affected with
regard to its ecosystem functions (Lambdon et al., 2008;
Chýtry et al., 2009). Mediterranean heathlands have a low
number of invasive species when compared with other
Mediterranean areas, and some heath types have no invasive
species ever recorded (Vilà et al. 2007). In contrast, the high
impact invasive exotic vascular plants are having on the
South African fynbos habitat threatens endemic native
species (Roura-Pascual et al., 2009).
In spite of its low invasiveness, European temperate heathlands occur on highly managed landscapes, and some important biological invasions have been recorded with a high
environmental and economic cost (Tyler et al., 2006).
However, this driver has not been given as much attention
by heathland ecologists as other issues, and the term ‘invasion’
is frequently used without distinction for the appearance of
native species as a result of natural succession and the
spread of exotic species that alter the community (e.g.
Mitchell et al., 2000).
Trees such as Prunus serotina or Robinia pseudoacacia,
shrubs like Gaultheria shallon, Rosa rugosa or
Rhododendron ponticum and grasses such as Panicum acuminatum are among the alien species naturalized in temperate
European heathlands. Some invasions are from coniferous
species such as Pinus strobus, Picea sitchensis or Douglas
fir (Pseudotsuga menziesii) (Broncano et al., 2005;
Carrillo-Gavilán et al., 2011), which compete with other
native gymnosperms involved in the natural succession
process. A favourable competition between Douglas fir and
the native silver fir (Abies alba) was described by
Carrillo-Gavilán et al. (2011) for a montane heathland landscape in north-east Spain. Higher germination rates were
found for Douglas fir, while the two species showed high mortality at early seedling stages. Success of the Douglas fir is
probably limited to the amount of available seeds. Propagule
pressure from nearby areas is thus the key process for invasiveness of the species, in accordance with the general view that
considers propagule pressure as a major factor on plant invasions (Simberloff, 2009).
Invasions in heathlands are driven mostly by habitat suitability directly influenced by soil conditions, as observed for
Rhododendron ponticum (Mitchell et al., 2000; Erfmeier and
Bruelheide, 2010). A negative feedback is promoted because
Rhododendron-invaded areas showed the lowest values of
plant diversity, either for total species or for heath specialist
numbers (Mitchell et al., 2000). In a different study, cover
Fagúndez — Heathlands confronting global change
of dominant species, either shrub or perennial grasses, was directly inverse to the presence of the alien Panicum acuminatum
in Calluna heathlands of northern Italy (Lonati et al., 2009).
Management should therefore limit tree encroachment
(Ascoli and Bovio, 2010) but also extend the loss of dominant
species’ cover to prevent spreading of the invasive species.
Whenever large amounts of biomass are removed by combined
factors such as fire and grazing in short intervals,
P. acuminatum will succeed in that particular region.
The negative effects of invasive species, however, go
beyond competition with native species, as complex interactions at trophic levels (i.e. food webs) are altered in invaded
heathlands. A significant negative effect of the invasive
species Gaultheria shallon over certain specialist parasitoid
and herbivorous species was recorded in a lowland heath in
Southern England (Carvalheiro et al., 2010). In contrast, and
to a certain level of invasiveness, generalist groups were positively affected by G. shallon. A lack of resources for specialist
herbivores, which do not feed on the exotic species, and disruption of the food chain by parasitoids – herbivores – plants,
causes a major threat to the habitat quality, and probably
affects other organisms and ecosystem functions. Disruption
of species assemblages in the community, including parasite
and host interactions, is thus a major cause of habitat quality
loss (Henson et al., 2009). An indirect effect of the presence
of an invasive species is the spread of Phytophthora spp.,
which infects Rhododendron ponticum and also other heathland species in the British Isles (Beales et al 2009). The
direct impact of the invasive parasite may be enhanced by
the presence of the invasive plant species.
I N T E R ACT I O N B E T W E E N D R I V E R S
The understanding of ecological processes involving different
drivers over ecosystems is a major task for ecologists in the
future (Darling and Côté, 2008). When dealing with biodiversity loss, very different responses have been recorded when
more than one factor is analysed, either additive effects, synergies or antagonisms. Non-additive interactions can be considered as ecological surprises, but recent knowledge points to a
high prevalence of these surprises in ecological research
(Lindenmayer et al., 2010). In the topics covered, many
studies analyse one or few drivers with the limitation of excluding other shift sources from the study. As an example,
several models have been proposed on the heathland future development under increasing N-deposition scenarios, but no
effect derived from global warming, or even climate, was considered (Power et al., 2004; Terry et al., 2004; Evans et al.,
2006; but see Britton et al., 2001). However, shifts occurring
under high N deposition values are similar to the rising temperature effects predicted for the future, such as changes in
plant growth or flowering (Peñuelas et al., 2004). Strong interactions between the described drivers are found whenever they
are analysed as factors in a single model, but the magnitude
and direction of interactions is generally unpredictable, as in
other ecosystems (Brook et al., 2008; Darling and Côté,
2008; de Chazal and Rounsevell, 2009). Moreover, effects of
different factors are frequently species specific and presumably
different communities and ecosystems will respond differently
163
to interactions (Werkman and Callaghan, 2002; Britton et al.,
2003; Kreyling et al., 2008).
Among the most studied interactions, eutrophication is probably the best-known since it is directly related to management
and heath dynamics and is critical for ecological restoration
and conservation (Piessens et al., 2005; De Graaf et al.,
2009). The increasing N concentrations in heath systems
come from atmospheric depositions or other factors such as
intense grazing (van der Wal et al., 2003) or burning
(Mohamed et al., 2007). In particular, higher values of NH4
are found after fire and there is a generalized loss of elements
through leaching, including N. The C : N ratio in plants is not
affected by fire, but C : P and N : P ratios increased in the grass
Deschampsia flexuosa. This suggests a negative effect on
D. flexuosa in competition with Calluna due to the latter’s advantage in uptake of the limitant P. The N content in Calluna
shoots increases with N addition (Pitcairn et al., 1995; Hicks
et al., 2000; Britton et al., 2008) and burning (Britton et al.,
2008), but there is no interaction effect according to Britton
et al. (2008). In contrast, the opposite effect of atmospheric
N deposition has widely been accepted with perennial
grasses favoured against heather (Heil and Bobbink, 1993;
Terry et al., 2004; Bobbink et al., 2010). Plant diversity
lowers with higher N availability and the soil N : P ratio
(Roem et al., 2002; De Graaf et al., 2009). A higher NH4 concentration mostly affects rare and endangered species (Kleijn
et al., 2008; De Graaf et al., 2009), and may constitute a
trigger factor for the spread of exotic invasive species
(Mitchell et al., 2000; Perry et al., 2010). Biomass removal
through other management practices, such as cutting or
grazing, buffers the ecosystem from the long-term effects of
N atmospheric deposition (Power et al., 2001) but, conversely,
removing the shrub canopy enhances grass competition by the
facilitation of seed germination, a particular stage in which
grasses are favoured by higher N deposition rates (Britton
et al., 2001; Friedrich et al., 2011).
Climate changes may also condition the effects of N atmospheric deposition (Whitehead et al., 1997; Gordon et al., 1999;
Britton et al., 2001; Werkman and Callaghan, 2002). Drought
is a key factor in the Calluna – grass and Calluna – bracken
competition (Gordon et al., 1999; Britton et al., 2001). The
dramatic substitution of heathlands by grasslands described
in The Netherlands (Heil and Diemont, 1983) does not occur
in the UK, at least to the same extent, and this may be
explained by a higher annual precipitation, a factor that
lowers the negative effect of N deposition (Gidman et al.,
2005; Helliwell et al., 2010), together with frequent cutting
(Britton et al., 2001; Fig. 6). Soil moisture is thus the key
factor affecting N mineralization in the soil, although higher
rates of respiration and decomposition are also recorded at
higher air temperatures (Emmett et al., 2004). Carbon sequestration will increase with higher N inputs (de Vries et al.,
2009), as has been seen in forest systems; however, the C : N
ratio response remains unclear, both in soil and biomass, to
recent changes, since increasing N deposition and a higher
CO2 concentration may act synergistically over plant productivity (Larsen et al., 2011).
Whitehead et al. (1997) measured the growth of Calluna
under high CO2 concentrations and N addition. Photosynthesis
and growth was enhanced both by increasing CO2 and as a
164
Fagúndez — Heathlands confronting global change
lluna cover
900 mm
100
60
40
Nd
epo
ha –1 sition
yea (kg
r –1 N
)
Years of stable Ca
80
10
20
20
0
5
35
10
15
20
Cutting
25
intervals
(years)
70
30
lluna cover
700 mm
100
60
P E R S P E C T I V E S FOR T H E F U T U R E
10
20
20
0
epo
ha –1 sition
yea (kg
r –1 N
)
40
5
10
15
20
Cutting
25
intervals
(years)
35
70
30
Nd
Years of stable Ca
80
lluna cover
550 mm
100
80
60
10
20
20
0
epo
ha –1 sition
yea (kg
r –1 N
)
40
5
10
15
20
Cutting
25
intervals
(years)
35
70
30
Nd
Years of stable Ca
species alters the specific response to each factor. Heather
growth was neither affected by increasing temperature or N addition, nor by its interaction, in pure stands. In mixed plots with
bracken, growth was negatively affected by both factors due to
the competition effect, since bracken reacted positively to rising
temperature and comparatively to N addition. Flowering showed
a different response, as it increased with higher temperature in
pure stands, but showed no significant response on mixed
plots. Bracken’s frond height increased and shaded heather,
which was the major cause for the negative effect over Calluna.
More long-term field experiments are needed to provide a
scientific-based approach to interactions between drivers of
biodiversity changes in heathlands. The system response to
higher nutrient availability together with different management types or climatic conditions are best known, but restricted
to competition, plant and soil chemical composition, and
effects over single species (Alonso et al., 2001; Werkman
and Callaghan, 2002; Barker et al., 2004; Härdtle et al.,
2009). Other factor interactions, such as the presence of invasive species combined with climate change, are desirable for
future predictions on heathland responses to these drivers.
F I G . 6. Interactive effect of cutting-time intervals, N deposition and precipitation in the maintenance of Calluna cover in a modelled lowland heathland.
Three scenarios with mean anual rainfall values of 550 mm, 700 mm and
900 mm where modelled. The frequency of cutting needed to maintain
Calluna cover decreases with higher precipitation values and higher N deposition. The 700-mm scenario showed a high sensitivity to N deposition, with a
management change needed at low levels of N deposition (20 kg N ha21
year21). Modified from Britton et al. (2001).
result of N and P addition (no response from N alone), but the
interaction was antagonistic. In contrast, bracken was enhanced
by both factors and synergistically by its interaction. This study
was performed under laboratory conditions, but a field study by
Werkman and Callaghan (2002), which also covered bracken
and heather response, suggests that competition between the
Heathlands are highly dynamic habitats, with strong biotic and
abiotic interactions affected by external drivers, despite its appearance as a homogeneous, stable ecosystem. All of the
factors studied influence diversity patterns, community structure and the general quality of the habitat. However, most of
these factors are species specific, site specific or strongly dependent on other factors such as local climate. Moreover,
shrubland types of markedly different climatic and geographic
areas with different species composition, as covered here, may
be affected differently by these drivers. As an example, a different response to experimental warming and drought in a
north – south gradient of European heathlands was recorded
for factors such as primary productivity, plant richness or
soil microbial activity (Peñuelas et al., 2004, 2007; Sowerby
et al., 2005). Even in stands from similar latitudes, differences
are observed due to abiotic local conditions or stochastic
forces such as heather-beetle attacks (Peñuelas et al., 2004).
Moreover, studies on interactions among factors are still insufficient but point to unpredictable and locally specific
responses. A summary of the general tendencies and future
projections that can be cautiously extracted from the available
information follows.
No clear effects on species richness and diversity but rare and
endangered species will decline
Overall species richness and diversity is a key indicator of
habitat quality, and the benefits of diversification in plant communities are many (Isbell et al., 2011). However, the temperate
European heathlands are relatively species poor, moderately
homogeneous and with a low species turnover (Webb and
Vermaat, 1990; De Graaf et al., 2009; Loidi et al., 2007,
2010). The overall number of species will probably increase
under intensely managed areas where a mosaic of grass and
shrubs is created (Calvo et al., 2005, 2007; Celaya et al.,
2010; Muñoz et al., 2012), but this does not necessarily
Fagúndez — Heathlands confronting global change
mean a higher habitat quality. For example, a higher number of
species was recorded in Betula-dominated and Pteridiuminvaded plots, although the heath-specialist species decreased
at both stages compared with the pure heathland stands
(Mitchell et al., 1999). A general species loss will take place
under higher N-deposition scenarios (Maskell et al., 2010).
Soil conditions, particularly higher NH4 concentrations, are
the key factors that affect plant richness. Rare and specialized
species will decline and most probably undergo local extinction processes under future conditions of soil eutrophication
(Roem and Berendse, 2000; van der Berg et al., 2005;
Kleijn et al., 2008; De Graaf et al., 2009). However, this
effect shows a high variability between regions and may
depend strongly on local climate and historical land-use
(Britton et al., 2001; Calvo et al., 2007).
Habitat fragmentation will affect heath species richness but
these changes have been retarded and thus heathland flora in
western European remnants show a case of extinction debt
(Piessens and Hermy, 2006; Cristofoli et al., 2010a). A
longer time span is needed for an accurate quantification of
biodiversity loss derived from an absence of connectivity
(Lindborg and Eriksson, 2004). On the other hand, restored
heathland patches will need time to recover diversity values
similar to undisturbed areas (Cristofoli et al., 2010b).
However, although fragmented and continuous patches are
similar in species richness even for heath specialists (Dupont
and Nielsen, 2006), other trophic or plant-associated groups
such as pollinators will decline, affecting all trophic levels
and plant – animal interactions (Piessens et al., 2005). The surrounding habitats at different scales, which buffers remnants
but may also provide propagule pressure of invasive species
and facilitate succession, will be critical for habitat conservation (Mitchell et al., 2000; Dupont and Nielsen, 2006;
Carrillo-Gavilán et al., 2011). Dynamics on heathland vegetation implies the alternation of different species and ecological
groups and, therefore, mosaics of heaths at different mature
stages together with remnants of natural and semi-natural
vegetation in extensive agriculture landscapes, i.e. higher landscape heterogenity, enhanced plant diversity (Firbank et al.,
2008; Muñoz et al., 2012). Global biodiversity threats
include a general switch towards intensified farming methods
(Tscharntke et al., 2005), indirectly affecting plant richness
and diversity in heathlands (DeGabriel et al., 2011).
Future climatic changes will not necessarily result in a decrease of species richness, but recovery after disturbances
such as wildfires will be retarded (Peñuelas et al., 2007),
since seedling diversity and survival will significantly decrease
under warmer, but mostly drier environments (Lloret et al.,
2004; Prieto et al., 2009; del Cacho et al., 2012). These
effects will be particularly significant in the Mediterranean
region, where wildfire frequency is predicted to rise (Delitti
et al., 2005). Moreover, extreme climatic events such as the
2003 summer heat-wave will probably increase in frequency
(Schär and Jendritzky, 2004) and provide the conditions for
wildfires to occur (Garcı́a-Herrera et al., 2010). Extreme climatic events are a key factor for the future modelling of
European heathlands (Peñuelas et al., 2007; Kreyling et al.,
2008) but predictions on event frequency and ecosystem response will be extremely difficult.
165
Community structure is affected by all drivers
Nitrogen deposition, global warming, intensive grazing, recurrent fires or rising CO2 concentrations promote a transition
in which perennial graminoids benefit in their competition
with shrub species (Berendse et al., 1994; Britton et al.,
2001; Power et al., 2004; Brys et al., 2005; Jaquemyn et al.,
2005; Hofland-Zijlstra and Berendse, 2009, 2010; Newton
et al., 2009; DeGabriel et al., 2011), while nutrient-limited
soils, prescribed burning or mixed grazing will benefit
heather (Mohamed et al., 2007; Harris et al., 2011). In addition, lack of management, wildfires in certain areas, high N
deposition and the transformation of surrounding areas promotes a transition towards scrub and woodland (Mitchell
et al., 2000; Manning et al., 2005; Mobaied et al., 2011).
Some authors have suggested that both transition forces may
occur simultaneously, resulting in the substitution of continuous heathland vegetation areas by a mosaic of grasslands and
woodlands (Bokdam and Gleichman, 2000; Ascoli and
Bovio, 2010). These changes also imply abiotic alterations
such as nutrient cycles that promote a certain species or
group above others, i.e. grasses versus shrubs.
Soil nutrient cycles will perform differently under a changing climate (Sardans et al., 2008a – c). For example,
drought conditions will reduce P and K capture by
Mediterranean heaths, reduce the concentration in leaves and
stems, and add to a diminution in the rate of photosynthesis,
while warming increases its concentration and promotes translocation from stems to leaves (Sardans et al., 2008b).
Moreover, rates of N atmospheric deposition are predicted to
increase, as is CO2 concentration, and both of these parameters
increase metabolic rates and biomass (Gorissen et al., 2004;
Albert et al., 2011a, b). Different responses have been
recorded between regions (Peñuelas et al., 2004) and species
(Llorens et al., 2004a, b), which means that heathland
species’ competition may depend on several environmental
factors. Competition and community structure is a major
factor that conditions the effects of different drivers. For
example, Calluna is negatively affected by warming and N
addition only when in competition with bracken (Werkman
and Callaghan, 2002), and flowering time is retarded by
drought only in complex-vegetation stands (Jentsch et al.,
2009).
General extent of heathlands will continue to decline
The abandonment of traditional practices of cultivated or
grazed areas is occurring in different areas, mainly in upland
heaths. Heathlands may be favoured by it, and re-colonization
of grasslands by different shrub communities including heathlands, will take place as a result of natural dynamics (Mouillot
et al., 2002, 2005; Azevedo et al., 2011). Mouillot et al.
(2002) predicted a higher heathland extent derived from a
higher frequency in fire regimes in the Central Massif of
France. Certain heathland communities in areas with a
Mediterranean climate have also extended due to traditional
land-use abandonment and increasing fire frequency (Moreira
et al., 2001; Mouillot et al., 2005; Azevedo et al., 2011). In
the Atlantic region, efforts have also been made to restore
166
Fagúndez — Heathlands confronting global change
heathland habitats, e.g. the programme to re-establish a substantial extension of lowland heathland in the UK (Pywell
et al., 2011). However, the general decline in lowland heathland landscapes observed for the last 200 years continues
2008
2000
(Veitch et al., 1995; Rose et al., 2000; Arianoutsou, 2001;
Mobaied et al., 2011; Hooftman and Bullock, 2012), and projections for the future predict a loss of heathland occupancy.
Mobaied et al. (2011) estimate a general loss of heathland
2016
Betula pendula + Molinia caerulea
2024
2032
Pinus sylvestris heath woodland
Homogenous stands of Calluna (building and
mature phases)
Open, discontinuous stands of Calluna (pioneer
and degenerate phases)
Betula pendula heath woodland
Betula pendula heath woodland
+ Molinia caerulea
Path
Erica tetralix + Calluna vulgaris
Erica tetralix
Swamp
Wet facies Erica tetralix + Molinia caerulea
Temporary swamp Molinia caerulea
Wet facies Calluna vulgaris + Molinia caerulea
Juncus + Molinia caerulea
Dry formation of Molinia caerulea
Campylopus and lichens
Sandstone
Rumex acetosella
Groves Betula pendula + Populus tremula
4·0
Measured
Model
3·5
Heathland
M. caerulea communities
Woodlands
Area (ha)
3·0
2·5
2·0
1·5
1·0
0·5
0
2000
2008
2016
Year
2024
2032
F I G . 7. Vegetation changes in a heathland landscape measured from direct observations in 2000 and 2008, and modelled predicted changes for 2016, 2024 and
2032 in the Fontainebleau massif (France). At the top of the figure, the 4.4-ha plot for two study periods (2000 and 2008) and three modelled scenarios (2016,
2024 and 2032) are shown (with key). Below, the area of heathland, woodland and Molinia caerulea grassland is shown for the same intervals. Heathland communities decrease while M. caerulea increases its cover, while woodlands increase slightly mainly due to the expansion of Pinus sylvestris. Modified from
Mobaied et al. (2011).
Fagúndez — Heathlands confronting global change
habitats by substitution with either grasslands or woodlands,
and predict a continuous heath loss measured from habitat
transitions (Fig. 7). In Mediterranean areas this tendency is
also observed due to profound changes in human land-use
and fire regimes (Arianoutsou, 2001). As an example,
machis vegetation declined by 30 % in favour of Pinus
brutia and Cupressus sempervirens forests in western Crete
over 50 years (Arianoutsou, 2001). In fynbos, future predictions of occupancy area and species loss by Midgley et al.
(2002) are even more catastrophic, with a total loss of .50
% and the extinction of a large proportion of endemic
species over the next 50 years considering the predicted
climate change. In western Europe, where the study sites of
most of the publications discussed here are located, heathland
occupancy will depend on management practices, land-use and
conservation policies, and social interests. A general concern
regarding the importance of the heathland landscape as a
natural and cultural heritage is now widely accepted, and
several heathland types are among the priority habitats of the
EU Habitats Directive (Ostermann, 1998). However, there is
also a need to maintain habitat quality, in terms of abiotic
factors and biotic interactions, in order to avoid the degeneration of existing heathlands despite efforts to preserve them.
Management policies must consider past, present and future
scenarios taking all possible shift drivers into account, and
specific conservation plans must be designed for each area.
CON CLU DI NG REMA RK S
The structure and biodiversity of heathlands are affected by
several drivers, but the effects are remarkably different
among sites, communities and species. Underlying biotic and
abiotic factors and driver interactions, together with landscape
history, explain the different responses recorded in the literature. As an example, eutrophication from N deposition has
been widely recognized as a major cause of heathland loss
by shift to grassland in The Netherlands (Heil and Bobbink,
1983, and references in the pollution paragraph), but it has
not been clearly demonstrated in other areas (Stevenson and
Thompson, 1993). Higher precipitation counteracts this negative effect to a certain extent and it has been proposed as a possible explanation for the diminishment of this process in the
British Isles (Britton et al., 2001; Fig. 6). In the Cantabrian
Mountains, a positive effect of N addition over diversity of
heath-specialist species with no competition against heather
(Calvo et al., 2007) suggests that local climate conditions or
historical management are conditioning the habitat response
to this driver. Whereas lowlands of western Europe had a
major transformation enhancing heathlands at approx. 3000
years BP, mountain areas in the north of the Iberian
Peninsula did not undergo an intense afforestation process
and heathland expansion until the Middle Ages, approx.
1000 years BP or even afterwards (Carrión et al., 2010, and
references therein). Climate differences and historical uses
add to the regional factors conditioning vegetation response
to drivers.
Strong interactions between the studied drivers suggest a
complex model in which additive interactions, synergies or
antagonisms are taking place. Nutrient cycling, occurrence
of rare species and early stages of development (seed bank
167
and seedlings) are key factors that are negatively affected by
the main drivers and their interactions. Competition between
heathers and grasses or progressive succession towards woodlands relies on a weak equilibrium affected by all drivers that
must be addressed locally to provide the appropriate management practices for the future conservation of European
heathlands.
ACK NOW LED GE MENTS
M. Sheehy Skeffington (National University of Ireland,
Galway), and an anonymous reviewer provided useful comments that improved this review. J. Izco (University of
Santiago de Compostela) made comments on an early draft.
I thank the authors and publishers who gave me permission
to use their figures.
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