PLANT EFFECTS ON SOILS IN DRYLANDS:
IMPLICATIONS FOR COMMUNITY DYNAMICS
AND ECOSYSTEM RESTORATION
Jordi Cortina1 and Fernando T. Maestre2
1 Departament d’Ecologia, Universitat d’Alacant Ap. 99 03080 Alacant, Spain; 2 Department
of Biology, Duke University, Phytotron Building, Box 90348, Durham, NC 27708 USA.
Almost 50% of the emerged land (6.15 106 ha) are considered drylands
(Reynolds and Stafford Smith, 2002). They are distributed in four continents,
covering from 31% (South America) to 75% (Australia) of the continental
land area. Drylands encompass areas with a wide range of conditions in
relation to the ratio Precipitation/Evapotranspiration (P/E), from hyperarid
(P/E<0.03) to Dry Subhumid (0.5<P/E<0.75) (UNESCO, 1977). Throughout
the text, we will use the term drylands in a broad sense, meaning areas where
precipitation is so scarce that it becomes the main factor controlling biological
processes. This definition corresponds to what Noy Meir (1973) refers to as
In areas like the Mediterranean, drylands have been subjected to a long
history of land use (Hillel, 1992; Grove and Rackham, 2001). But this is not
always the case, as different dryland areas have supported extremely
contrasted human population densities. The boundary between subhumid to
semiarid conditions may be particularly sensitive to disturbances, and
population buildup in subhumid and semiarid zones have probably favoured
the expansion of areas characterized by discontinuous vegetation cover. This
process is not recent, as population density was relatively high in areas such
as the Mediterranean basin 2000 years ago (Grove and Rackham, 2001). For
example, Butzer (1990) suggests that population density in the Iberian
Peninsula was close to the carrying capacity for agricultural systems several
times since 2,000 BP. Phases of economic welfare and social integration
correspond to higher demands for resources, including land, leading to defor-
From: Binkley, D., and O. Menyailo (eds). 2005. Tree Species Effects on Soils: Implications
for Global Change. NATO Science Series, Kluwer Academic Publishers, Dordrecht.
estation and, eventually, land degradation. The 20th century has shown an
unprecedented increase in population density in many drylands (Le Houérou,
1992). But, unlike what happened in the past, population density has mainly
concentrated in coastal areas. As a result, recent population increase has been
paralleled by a reduction in the intensity of inland land use (and an increased
demand for resources overseas).
Changes in land use are relevant in the context of species effect on soil
properties for two reasons. First, land use, together with natural factors, have
favoured land degradation in vast areas. For example, alpha grass (Stipa
tenacissima) steppes, once covering more than 8 million ha in the Maghrib,
are being destroyed at a rate of 1% per year (Aïdoud, 1989; Le Houérou,
2001). Past disturbances may interact with plant species to generate
unexpected effects. For example, litterfall inputs may not be the same if
individuals are established in a slope that had either been terraced or left
aside. Second, with few exceptions, most individual plants may have
established in their present locations not long ago. Examples of changes in
land use dating less than 50 years are common (Grove and Rackham, 2001;
Bonet, 2004). In contrast, the effect of land uses such as agriculture or grazing
on ecosystem structure and functioning may last for centuries (Bruun et al.,
2001; Maestre 2004). Changes in soil phosphorus content in the surroundings
of Roman farms have been detected up to 2,000 years after abandonment
(Dupouey et al., 2002). These changes were large enough to affect P
concentration in oak leaves. Thus, information accumulated in soils for
centuries adds noise to the interpretation of species effect on soil properties
There are many ways in which plants may affect soils. They can modify
soil properties directly, e.g. through inputs of organic matter and nutrients,
and indirectly, by affecting abiotic and biotic conditions that influence soil
properties. For example, symbiosis with specific strains of mycorrhizal fungi
can be relevant for the establishment of their own progeny or for the
colonization by other species (Palenzuela et al., 2002; Azcón-Aguilar et al.,
2003). Under dry conditions, shade may affect litter decomposition directly
(Duguay and Klironomos, 2000; Verhoef et al., 2000), and as a result of
changes in microclimatic conditions (Jackson and Caldwell, 1992; Cortina
and Vallejo, 1994). Sometimes, changes in soil properties affect plant
composition and growth, generating positive feedbacks (Northrup et al., 1995;
Morehead et al., 1998). On the other hand, plants may affect soil properties in
ways that may appear subtle to us, but may be rather evident to other
organisms (Bais et al., 2003, 2004). This complex network of interactions
hampers the interpretation of plant-soil relations. Unfortunately, our
knowledge is still quite broad, and surprises arise too often.
Despite the intrinsic variability in soil properties and the complexity of
soil-plant interactions, plant effects on soils play a major role in population,
community and ecosystem dynamics in drylands. Several features of drylands
may contribute to this. By definition, climatic conditions are harsh and soil
conditions are unfavourable for plant growth due to salinity, low organic
matter, P immobilisation, etc. Thus, any slight modification of microclimate
or soil properties may have disproportionate effects on other organisms (e.g.,
Pugnaire et al., 1996a; Moro et al., 1997; Maestre, 2003). Furthermore,
xerophytes are commonly small and isolated, as compared to plants from
temperate and tropical areas. This may favour litterfall, throughfall, stemflow
and fine root inputs accumulating underneath the canopy, and thus
intensifying plant effects. Finally, isolated plants act as obstacles for runoff
carrying organic matter, nutrients and sediments (Greene et al., 2001). Thus, it
is not surprising that certain plants and vegetation patches have higher
infiltration rates, improved soil structure and nutrient content, and higher
biological activity, creating “hotspots” of favorable soil conditions that have
been referred to as “fertile islands” or “resource islands” (Whitford, 2002).
As we just mentioned, plant alteration of soil and microclimate conditions
may affect other organisms to the point of controlling the composition and
function of the whole ecosystem (West, 1989; Schlesinger and Pilmanis,
1998; Montaña et al., 2001). Identifying these interactions is crucial to
understand and to manage dryland ecosystems. In this review, we will
describe the way plants affect soil properties in drylands, and how these
feedbacks, together with changes in microclimatic conditions, affect plant-
plant interactions and community dynamics. In the last section, we will
discuss the use of this knowledge for the restoration of degraded dryland
PLANT EFFECTS ON WATER DYNAMICS AND USE
In drylands, water scarcity controls plant effects on soils in several ways.
Biological processes, including plant productivity, are strongly limited by
water. Water limitation results in discontinuous plant cover, which is often
arranged as a two-phase mosaic of vegetated and bare ground patches
(Valentin et al., 1999). Spatial patterns of vegetation, together with
morphological and physiological attributes of plants greatly affect water
fluxes and availability (Bromley et al., 1997, Cerdà, 1997). Indeed, the spatial
pattern of plant patches is essential to maintain ecosystem composition and
function in drylands (Noy Meir, 1973; Tongway et al., 2001), and can be used
as an indicator of the degradation status of arid and semiarid lands (Wu et al.,
2000; Bastin et al, 2002; Maestre and Cortina, 2004a).
At the patch level, canopies differ in their capacity to intercept water. For
small rains, which normaly account for a large number of rain events,
interception can be high (Martínez-Meza and Whitford, 1996; González-
Hidalgo and Bellot, 1997; Bellot et al., 2001), and may promote the formation
of dry shadows (Valladares and Pearcy, 2002). Stemflow and throughfall are
strongly dependent on plant architecture (Martínez-Meza and Whitford, 1996,
Domingo et al., 1998). Funnel-like structures, as found in many shrubs,
favour moisture concentration around the base of the stems (West, 1989;
Martínez-Meza and Whitford, 1996), whereas tussock grasses may
concentrate throughfall inputs around the canopy edge (Puigdefábregas et al.,
1999). Rainfall redistribution may affect deep soil moisture content as well, as
water fluxes at the base of the stem may circulate down the soil profile
following the roots and root pores (West, 1989; Martínez-Meza and Whitford,
1996; Ryel et al., 2003). These fast tracks for water fluxes may explain why
increases in moisture content following a rainfall event are not necessarily
sequential, from top to bottom, but water may saturate deeper soil profiles
earlier than shallow ones (González-Hidalgo et al.,2003). Species with the
highest stemflow, and with deep roots may favour the recharge of deep soil
horizons (Martínez-Meza and Whitford, 1996). These traits are common in
some desert and Mediterranean plants (Cannon, 1911; Kummerov, 1981). On
the other hand, hydraulic lift −water translocation to upper or lower parts of
the soil profile through the rooting system − has been described in drylands
worldwide (Caldwell et al., 1998; Burguess et al., 1998; Filella and Peñuelas,
2003; Gutiérrez and Squeo, 2004). This movement of water may affect N
mineralization and microbial activity as well (Caldwell et al., 1998). The
Figure 1. Moisture content of the 0-20 cm soil underneath Stipa tenacissima tussocks that were
either left undisturbed (black dots) or bended to suppress shadow (white dots) in a steppe
located in SE Spain. Data represent means and standard errors (n = 6). Redrawn from Appendix
A in Maestre et al. (2003a).
Volumetric soil moisture (%)
J F M A M J J A S O N D J
F M A M
shade provided by plant canopies, together with litter accumulation, influence
the soil water balance in the surroundings of plant patches (Fig. 1). It is
important to note that shade may not only affect soil moisture content, but
may also influence individual plant performance and plant community
composition by modifying evaporative demand, and light quantity and quality
Stormy rains are common in some drylands, such as the Mediterranean
and West Peru (De Luís et al., 1997; Arntz and Fahrbach, 1996). For example,
in Spain, more than 100 rain events with intensity higher than 100 mm in 24
h, including seven rain events higher than 500 mm in 24 h, were recorded
between 1901 and 1989 (Olcina, 1994). There is increasing evidence that the
amount and frequency of large vs. small rainfall events may control
ecosystem function and composition in drylands (Ehleringer et al., 1999). One
of the consequences of rainfall concentration is that part of the water does not
accumulate in soils, but is either drained or exported as runoff (Bellot et al.,
2001; De Luís et al., 2001). Surface runoff can have strong effects on plant-
soil interactions because it may transport sediments and organic matter within
or beyond the slope, and because it favours the increase in soil fertility and
water availability under plant canopies (Cerdà, 1997; Puigdefábregas et al.,
1999). Structures in contact with the surface soil (lower branches, multiple
stems, litter, tussock plants, etc.) favour runoff retention (Rostagno, 1989;
Bromley et al., 1997), and attributes of plant patches like size, width parallel
to the slope, and spatial pattern are critical to define the ability of an
ecosystem to retain and use those resources transported by runoff (Ludwig et
Most of the water inputs in drylands are lost by evapotranspiration. Thus,
in Mediterranean watersheds, annual evapotranspiration, and not drainage, is
well correlated with precipitation (Piñol et al., 1991). Vegetation
management, involving shifts from woody to herbaceous vegetation, partial
clearing and formation of stone pavements, etc. has been used to collect more
drainage water (Hillel, 1992; Lavee et al., 1997; Burch et al., 1987). In
general, evapotranspiration is proportional to LAI, and suppression of woody
vegetation commonly results in increasing soil moisture availability, seepage
drainage or watershed runoff (Burch et al., 1987; Bellot et al., 2001), although
exceptions exist (Dodd et al., 1998). On the contrary, increases in woody
cover (e.g. such as shrub encroatchment) may lead to a reduction in surface
water availability or groundwater levels (Puigdefábregas and Mendizábal
1998; Ohte et al., 2003; Bellot et al., 2004) (Fig. 2). The increase in the cover
of woody invasive species can have important economic consequences (Le
Maitre et al., 2002). Species identity may also be relevant for water losses.
Obvious examples are summer-deciduous or semi-deciduous species (e.g.
Cistus salviifolius, Cistus albidus, Euphorbia dendroides, etc.), that reduce
evapotranspiration losses by leaf shedding (Ne’eman and Goubitz, 2000). The
difference between drought escapers, drought avoiders and drought tolerant
species (Levitt, 1980), could be helpful in this context. Comparisons in water
use between pairs of species are frequent. For example, pines commonly close
stomata at relatively high water potential, and thus may reduce water losses
during drought as compared to other species. Pistacia lentiscus and Quercus
coccifera, two coexisting shrubs common in the western Mediterranean basin,
Volumetric soil moisture (%)
Apr AprMay May JunJun JulJul AugAug SepSep
Volumetric soil moisture (%)
Figure 2. Changes in moisture content of the surface soil (0-10 cm depth) of adjacent
Mediterranean semiarid grasslands with (white circles) and without (black circles) a dense
cover of Pinus halepensis (3470 trees ha-1). Bars correspond to rainfall events recorded
during the period of study (April-September 1996). Soil moisture data represent means and
standard errors (n = 5). Redrawn from Bellot et al. (2004).
Julian Date (days)
0 100 200300 400500600
NET EFFECT (rel. units)
Tussock vs Open
Figure 3. Net effect of the tussock grass Stipa tenacissima on soil moisture content (0-20 cm
depth) as compared to the arithmetic addition of the net effects of transpiration, runoff and
shadow in a Mediterranean semiarid steppe of SE Spain. Net effects are calculated from the
standardized differences in moisture content between tussock and open microsites (net
tussock effect), tussock and herbicided tussock (net transpiration effect), tussock with and
without runoff (net runoff effect) and standing and bended tussock (net shadow effect). From
original data in Maestre et al. (2003a).
show contrasted water use strategies (Vilagrosa et al., 2003). Other factors
being equal, the water spender P. lentiscus is likely to deplete soil moisture
sooner after a rain that Q. coccifera. But, to our knowledge, no attempt to
systematically relate plant strategy to withstand drought, and water dynamics
at a plot and catchment scale has been made so far.
Finally, plants may affect water availability by favouring the formation of
calcite and laterites, and thus modifying soil volume (Viles, 1990). Water
uptake from the vadose zone may lead rhizoconcretion precipitation around
particular roots, whereas phreatophytes may favour different forms of calcite
precipitation in the zone of capillary rise (Thomas, 1988). These processes
result in the reduction of root absorptive capacity, root death, and finally the
decrease in available soil.
Given the diversity of effects of plants on water fluxes, it is not surprising
that the effect of plants on water availability, a critical soil property in
drylands, may substantially vary according to plant community composition
and structure. Furthermore, it is difficult to explain variations in moisture
content as the additive effect of changes in single fluxes (Fig. 3). On the other
hand, plant effects on soil moisture content may be short-lived, being
restricted to the lapse between a rainfall event (and, eventually, homogeneous
high moisture content), and soil dessication, or to particular periods of the
year (Belsky et al., 1993) (Fig. 4). This is particularly true for shallow soils
and surface soil horizons.
ORGANIC MATTER AND NUTRIENTS: ISLANDS OF
Litterfall inputs are relatively low in drylands due to constrains in plant
productivity (Berg et al., 1999; Breckle, 2002), but they may be substantially
higher immediately underneath the canopy. Variability in litter decomposition
rates is very high, ranging from some of the lowest rates recorded (0.07 yr-1;
Hart et al., 1992), to relatively high values (Gallardo and Merino, 1992). For
Mediterranean ecosystems, Aerts (1997) found that short term decay rates
were largely variable, averaging 0.35 years-1. These values were very close to
those of temperate ecosystems (0.36 years-1), and substantially lower than
decay rates in tropical areas (2.33 years-1). Within a given site, variability in
litter decomposition rates can be very high (Gallardo and Merino, 1992),
suggesting that species identity may control soil organic matter (SOM)
dynamics and nutrient availability. In Mediterranean ecosystems litter decay
rates are related to actual evapotranspiration (AET) and litter quality (Aerts,
1997), but the amount of variability explained by single climatic or chemistry
parameters is very low. In desert ecosystems, AET underestimates
decomposition rates, probably due the activity of soil fauna (Whitford et al.,
1981). The relationship between litter decay rates and a widely used index of
Shrubs ShrubsShrubs P. halepensisP. halepensis P. halepensisQ. ilexQ. ilex Q. ilex
Litter (Oi) accumulation (Mg/ha)
20 20 20
Forest floor accumulation (Mg/ha)
Litter (Oi) accumulation (Mg/ha) Litter (Oi) accumulation (Mg/ha)
Forest floor accumulation (Mg/ha)
Forest floor accumulation (Mg/ha)
Figure 5. Litter (upper graph) and forest floor (lower graph) accumulation in Mediterranean
forests and shrublands dominated by different woody species. Data from Fons (1995), Ferran
(1996), Huesca et al. (1998), Serrasolsas and Vallejo (1999), Cortina (1992), Berg et al. (1993),
Sevink et al. (1989), and Van Wesemael and Veer (1992). Only maximum values were taken
from studies reporting measurements at several locations.
recalcitrance, the lignin to nitrogen ratio, is often poor in drylands (Gallardo
and Merino, 1992; Aerts, 1997). The resistance of the external layers of leaves
and needles may be more important drivers of decomposition rates in these
environments than the quality of the whole tissue (Gallardo and Merino,
1992; Cortina and Vallejo, 1994).
Forest floor accumulation in drylands is commonly lower than in more
mesic environments because of low litterfall, and relatively high
decomposition rates and vertical transfers (Vallejo et al., 1998; Fig. 5).
However, particularly dry conditions and surface accumulation of rock
fragments may favour the formation of xeromoder type duff layers (Sevink et
al., 1989; Fons, 1995). This is probably due to reduced transfer to lower soil
horizons, as CO2 efflux is promoted by the presence of a stone layer (Casals et
al., 2000). As previously mentioned, small height and isolation may increase
the potential of dryland vegetation to concentrate litterfall, and locally
promote forest floor build-up. Plant species differ in forest floor
accumulation, and morphology (Peltier et al., 2001), although other factors
Days after irrigation
02468 1012 1416
Volumetric soil moisture (%)
Figure 4. Changes in volumetric soil moisture content in the surface soil (0-20 cm)
underneath Stipa tenacissima tussocks (black circles) and in open areas (white circles) after
irrigation in a semiarid steppe in SE Spain. Soil moisture data represent means and standard
errors (n = 10). Redrawn from Maestre et al. (2001).
such as lithology may be more important in determining forest floor
properties (Sevink et al., 1989; Fons, 1995). Aphyllous species (crassulacean,
many legumes, etc.), and species with no spontaneous leaf shedding (e.g.
tussock grasses, palms) accumulate small amounts of litter; whereas species
loosing a substantial amount of their foliage every year can create a relatively
thick forest floor (Bochet et al., 1999; Peltier et al. 2001). Under
Mediterranean conditions, conifers may accumulate thicker forest floor layers
than hardwoods (Peltier et al., 2001).
Litter incorporation into the mineral soil and SOM stabilization can be
greatly affected by soil fauna (Whitford, 2002). Soil fauna may comminute
partly decomposed litter and favour SOM transport down the profile
(Anderson, 1988; Bertrand and Lumaret, 1992; Romanyà et al., 2000a). It is
also affected by plant species and plant cover type. For example, in New
Mexico deserts, shrub cover may negatively affect the presence and activity
of some mammals and promote the abundance of rabbits, but may not affect
ants and termites (Krogh et al., 2002; Jackson et al., 2003). In Mediterranean
environments, millipedes are more frequently found under Quercus coccifera
than under Q. ilex, Q. pubescens or Brachypodium ramosum (Bertrand and
Lumaret, 1992). However, Maestre and Cortina (2002) did not find any
relationships between the spatial pattern of earthworm casts and the spatial
pattern of grass species in a semi-arid steppe in SE Spain. Given the important
effects of soil fauna on SOM dynamics, infiltration and nutrient turnover in
drylands (Whitford, 2002), further studies on the interactions between plants,
fauna and soils are needed.
SOM decomposition, as plant productivity, is negatively affected by
water limitations (Meentemeyer, 1978), and SOM content in dryland soils is
commonly low (Vallejo et al., 1998). SOM is usually higher underneath the
canopy of isolated plants and vegetation patches than in open areas (see
below). Organic matter accretion as shrubs increase in size has been described
(Pugnaire et al., 1996a, Tirado, 2003). Fine root inputs and historical changes
in the spatial pattern of vegetation distribution may, however, attenuate these
differences (Puigdefábregas et al., 1999), although plant effects on SOM may
last less than two decades (Burke et al., 1999; Romanyà et al., 2000b;
Martínez-Mena et al., 2002). It is often difficult to evaluate species effect on
SOM because the age of individuals under comparison may be very different.
This is particularly true for resprouting species. For example, in a review on
the effects of 20th century afforestations with Pinus halepensis (one of the
most common tree species in the Mediterranean) on semiarid ecosystem
properties, Maestre and Cortina (2004b) found that most studies on SOM
dynamics reported lower SOM contents underneath pine canopies than
underneath patches of relatively undisturbed shrubland. We don’t know if
differences arise from contrasted species properties or from time, and thus we
don’t know if pines will ever reach the SOM levels found under established
shrubs. Contrasts in soil CO2 efflux can be substantial under different types of
vegetated patches within a given ecosystem (Maestre and Cortina, 2003; Fig.
SUSD BR ECBCBG
CO2 efflux (mg CO2-C·m-2·h-1)
Figure 6. Soil respiration from different microsites in a Stipa tenacissima steppe in SE Spain.
Measurements were taken on May 2001. Bars represent means and standard errors of ten
replicates per microsite. SU: Upslope of S. tenacissima tussocks, SD: Downslope of S.
tenacissima tussocks, BR: under the canopy of the sprouting perennial grass Brachypodium
retusum EC: bare ground areas covered with earthworm casts, BC: bare ground areas covered
with biological crusts, BG: bare ground areas covered with physical crusts. From original data
in Maestre and Cortina (2003).
6), an indication that species may greatly differ in their overall effects on
SOM content and quality. Plants can greatly modify nutrient accumulation
and availability through mechanisms such as weathering, nitrogen fixation,
runoff capture, concentration of animal feces, etc. (West, 1989; Kelly et al.,
1998; Eviner and Chapin, 2003). Nitrogen fixing plants, legumes and
actinorhizal plants, are widespread in drylands. This is remarkable, as soil
properties that are common in drylands (such as low P availability, salinity
and low water availability) may negatively affect N fixation rates (Reddell et
al., 1991; Azcón and Altrash, 1997). Sprent (1987) suggested that deep
rooting and adaptations to minimize water loss, such as the presence of
phyllodes, may have contributed to their success. Experimental measures of N
fixation rates range from less than 1 kg N ha-1 year-1 to 390 kg N ha-1 year-1
We have reviewed published literature on changes in soil properties under
vegetation patches in drylands. The data set includes 31 references, and more
than 40 species and vegetation types covering a rainfall gradient from less
than 100 mm to more than 1000 mm, with most studies in the range 200-250
Table 1. Some examples of N inputs through biological fixation in drylands.
1Rundel et al. (1982),
5Youngberg and Wollum (1976), 6Kummerov et al. (1978), 7Johnson and Mayeux (1990),
8López-Villagra and Felker (1997), 9Moro (1992), 10Aronson et al. (2002), 11Sharifi et al.
(1982), 12Unkovich et al. (2000).
mm. All communities show discontinous cover of woody vegetation and
herbaceous tussocks, with most studies reporting cover values ranging from
20% to 50%. Lythology and soil types are diverse, from acidic (pH 4.5) to
alkaline (pH 8.5). We used the index RII (Armas et al., 2004) to compare soil
properties underneath the canopy and in intercanopy areas. RII is calculated
where VARc and VARi are the values of a given soil property for the canopy
and intercanopy areas, respectively. RII ranges from –1 to +1, with positive
values indicating increases in the variable under study in canopy microsites.
For the 41 comparisons of soil organic carbon content (TOC) between soils
underneath vegetated patches and soils in open areas (Fig. 7), average soil
depth is 10 cm (range 1-40 cm). TOC ranged from 0.1 to 10%. Vegetated
patches showed an average increase in TOC up to 450%, with most values
falling between 50% and 100%. In no case TOC was higher in open areas
than in vegetated areas.
Results from the 31 studies comparing total N (TKN) content are similar to
those of TOC (Fig 8). TKN ranged from 0.02% to 1.7%. The relative
N fixation rate (kg ha-1 year-1)
4Gauthier et al. (1985),
TKN in surface soil (%)TKN in surface soil (%)
00 0.20.2 0.40.4 0.60.6 0.8 0.811
Net Canopy Effect - RII (rel. units) Net Canopy Effect - RII (rel. units)
-1.0-1.0-0.5-0.5 0.0 0.00.5 0.5 1.01.01.5 1.5
Frequency (number of cases)
Frequency (number of cases)
Figure 7. Left: Total organic carbon concentration in surface soils under the canopy of
vegetation patches (CANOPY, solid line), and in open areas (OPEN, broken line) in several
comparative studies conducted in drylands. Right: Net canopy effect (RII) on total soil carbon
concentrations in surface soils. This effect was calculated as (TOCc-TOCi)/(TOCc+TOCi),
where TOCc and TOCi are total carbon concentration in canopy and intercanopy areas,
TOC in surface soil (%)TOC in surface soil (%)
0022446688 101012 12
Frequency (number of cases)
Net Canopy Effect - RII (rel. units)Net Canopy Effect - RII (rel. units)
-1.0-1.0 -0.5-0.5 0.00.00.5 0.51.01.0 1.51.5
Frequency (number of cases)
Figure 8. Left: Total organic nitrogen concentration in surface soils under the canopy of
vegetation patches (CANOPY, solid line), and in open areas (OPEN, broken line) in several
comparative studies in drylands. Right: Net canopy effect (RII) on total soil carbon
concentration in surface soils. This effect was calculated as (TKNc-TKNi)/(TKNc+TKNi),
where TKNc and TKNi are total nitrogen concentrations in canopy and intercanopy areas,
Frequency (number of cases)
0 20 40 60
Olsen P in surface soil (mg/kg) Olsen P in surface soil (mg/kg)
0 2040 6080 80
Net Canopy Effect - RII (rel. units)Net Canopy Effect - RII (rel. units)
-1.0 -1.0-0.5 -0.50.0 0.00.50.5 1.0 1.01.5 1.5
Frequency (number of cases)
Figure 9. Left: Concentration of available phosphorus (bicarbonate extraction) in surface soils
under the canopy of vegetation patches (CANOPY, solid line), and in open areas (OPEN,
broken line) in several comparative studies in drylands. Right: Net canopy effect (RII) on total
soil carbon concentration in surface soils. This effect was calculated as (PAVc-
PAVi)/(PAVc+PAVi), where PAVc and PAVi are available phosphorus concentrations in
canopy and intercanopy areas, respectively.
increses in TKN underneath vegetated patches were in the same range as for
TOC. The number of studies showing higher TKN in open areas as compared
to vegetated areas was very low, and of those, none showed a substantial
decrease in TKN. Phosphorus availability (Olsen’s bicarbonate extraction; n
= 12; Fig. 9) ranged from 1-56 ppm in open areas to 2-65 ppm underneath
vegetated patches. Most studies found a ca. 50% increase in P availability
underneath vegetated patches.
BARE AREAS THAT ARE NOT
As mentioned before, precipitation in drylands is not high enough to
maintain a continuous cover of vascular plants (Specht, 1988), leaving what
are commonly, and wrongly, referred to as bare ground areas. Soil surface
conditions in these areas is relevant for ecosystem functioning (Tongway and
Ludwig, 1997), and its degradation may impair water fluxes to plant patches
(Eldridge et al., 2000), and modify ecosystem-level processes like soil
respiration (Maestre and Cortina, 2003). Bare areas are frequently covered by
communities of cryptogams (mosses, lichens, cyanobacteria, liverworts and
green algae), commonly referred to as biological crusts (West, 1990). Crusts
are an important source of soil organic carbon (Beymer and Klopatek, 1991),
fix atmospheric nitrogen (Rychert and Skujiņš, 1974), reduce wind and water
erosion (Belnap, 1995), increase soil stability (Belnap and Gardner, 1993),
and have an important effect on soil-water interactions (Eldridge et al., 2000).
Biological crusts are preferentially established on fine textured slightly
alkaline soils, with low content of surface rock fragments (Vitousek et al.,
2002), In drylands, smooth and rugose crusts dominate (Belnap, 2001).
Depending on crust composition, net annual carbon input has been
estimated between 0.4 and 37 g C m-2 year-1 (Evans and Lange, 2003). In arid
areas this may represent a significant input of organic matter. These figures
correspond to relatively high instantaneous rates of C fixation, up to ca. 5
µmol CO2 m-2 s-1 (García-Pichel and Belnap, 1996; Lange et al. 1997; Lange
et al., 1998), as crusts remain inactive for long periods.
When biological crusts incorporate cyanobacteria and cyanolichens, they
may fix substantial amounts of atmospheric nitrogen. Several studies have
measured N fixation rates higher than 10 kg N ha-1 year-1 (Rychert and
Skujiņš, 1974; Belnap, 2002; Evans and Lange, 2003), although much lower
rates are reported (Jeffries et al., 1992; Aranibar et al., 2003). Part of this N
may be lost by denitrification and ammonia volatilization (Vitousek et al.,
2002). These estimations are, however, subjected to great uncertainty, as
substantial spatial and temporal variability adds to methodological limitations.
Nevertheless, N inputs in arid ecosystems coming from biological crusts can
be relevant at an ecosystem scale (Evans and Ehleringer, 1993; Billings et al.,
Biological crusts affect water fluxes in various ways. Increased surface
roughness may favour runoff reduction and water infiltration (Warren, 2001).
However, the relatively flat morphology of biological crusts in arid and
semiarid areas not subjected to freezing may reduce the magnitude of this
effect. The net effect of biological crusts on infiltration rate depends on soil
texture, and on the identity of the organisms dominating the crusts. In sandy
soils, biological crusts may increase microporosity and reduce infiltration. In
soils with higher fine particle content, biological crusts improve aggregation
and create macropores, thus increasing infiltration rate. Moss-dominated
crusts favor infiltration, and lichen- and cyanobacteria-dominated crusts
reduce it (Maestre et al., 2002a). Crusts may also affect erosion and sediment
transport by altering runoff, attaching soil particles together, and physically
protecting the surface soil. Accordingly, disturbance of the biological crust
may favour increased sediment yield (Belnap and Gillette, 1997).
Biological crusts and vascular plants interact in a number of ways.
Vascular plants usually outcompete biological crusts, but they can also take
advantage of the microenvironment created by the former. Environmental
modifications promoted by plants, such as lower surface soil temperatures,
reduced radiation, and decreased wind speed on the soil surface, promote
changes in the composition, dominance and spatial pattern of the organisms
forming these crusts (Eldridge, 1999; Maestre et al. 2002a, Maestre 2003). On
the other hand, biological crusts can directly affect the establishment (Prasse
and Bornkamm, 2000), survival (Eckert et al. 1986), nutrient status (Harper
and Belnap, 2001), and water relations (DeFalco et al., 2001) of vascular
plants by altering soil surface topography, modifying water and nutrient
fluxes, chelating metals, secreting growth promoting compounds, favouring
mycorrhizal abundance or increasing pH (Belnap and Harper, 1995; Belnap et
al., 2001; Li et al., 2002; Pendleton et al., 2003).
PLANT RESPONSE TO DISTURBANCES MODULATE
PLANT EFFECTS ON SOILS
In environments that are prone to disturbances such as wildfire or grazing,
resistance and resilience against disturbance are relevant traits affecting plant
species effect on soils. As previously described, species that maintain a
relatively humid microclimate under the canopy will affect soil processes
directly. But they will also affect the probability of ignition and fire severity
(Elvira and Lara, 1989; Wheelan, 1995), and thus attenuate the effects of
wildfire, including changes in plant community composition. Persistance after
disturbances that remove aboveground parts, such as wildfire, is increased by
resprouting. This is a common trait in Mediterranean vegetation (Kummerov,
1981), that can be crucial to ensure fast soil protection after disturbance
(Vallejo and Alloza, 1998). Resprouters are likely to affect soil properties in a
given location more intensely than species that depend on the seed bank to
reestablish, despite that some obligate seeders may show some degree of
persistance at small spatial scales (Moreno and Oechel, 1994).
Plant architecture may also affect stability against disturbances. It is well
known that trees may increase erosion by raindrop splash as compared to
shrubs, due to the formation of big throughfall drops falling from more than
8-9 m above the surface soil (Viles, 1990). On the other hand, vertical
continuity of the canopy may favour the combustion of all vegetation strata.
Accordingly, self-pruning may influence fire effects on soils. Vertical
structure can also be relevant in shrubs. Ulex parviflorus is a spiny leafless
leguminous shrub that colonizes abandoned agricultural fields in the western
Mediterranean. Senescent stems form early and, without abscission, remain as
standing necromass, increasing the fuel load. When the canopy is not closed,
the surface layer next to this species, devoid of forest floor, may be colonized
by grasses such as Brachypodium retusum. Cistus albidus is a coexisting
shrub in these areas; its leaves twist during drought and may fall as summer
progresses. They usually form a relatively thick forest floor devoid of
herbaceous layer, with little foliar biomass concentrating in the top of the
branches. Vertical continuity of the fuel load is higher in Ulex parviflorus,
and this may be the cause for higher combusting power underneath this
species (Fig. 10).
Forest floor accumulation affects ecosystem resilience in several ways.
Forest floor protects the surface soil from rainfall splash, limiting the
formation of physical crusts (Thomas, 1988), and reducing the risk of erosion.
This is particularly relevant in areas where plant cover has been previously
removed by wildfire or clearing. For example, low severity fires may not
completely destroy the forest floor (Gillon et al., 1999). Furthermore, patches
of relatively thick forest floor may withstand further rains, even high intensity
rains, and thus protect the underlying soil from erosion (García-Cano, 1998).
Considering all factors involved in the comparison between Ulex parviflorus
and Cistus albidus, soil protection after wildfire is lower in the former
because it accumulates less forest floor, it shows higher combustion of surface
litter, and because ashes are readily washed away after fire.
COMMUNITY DYNAMICS DRIVEN BY CHANGES IN
Plant-plant interactions, resulting from the net output of positive and
negative interactions, are crucial for ecosystem composition, structure and
dynamics in drylands (Whitford, 2002). Most studies on negative plant-plant
relations in these areas have focused on trophic interactions, resource
depletion and competition, and allelopathy (Scholes and Archer, 1997; Kröpfl
et al., 2002; Whitford, 2002). Other negative interactions have comparatively
PRE-BURN SPECIES - litter type
Weight loss (%)
Figure 10. Weight loss through combustion of Cistus albidus (ca) and Ulex parviflorus (up)
litter located underneath the canopy of individuals of Cistus albidus (CA) and Ulex parviflorus
(UP) prior to an experimental fire. Bars represent means and standard errors of 30 individuals
per species and litter type. Location was the only significant factor (F=88, p=0.01).
Unpublished data from M.F. García-Cano.
received much less attention, probably due to the inherent experimental
difficulty in differentiating trophic and non-trophic factors. Negative
interactions can be the result of plant effects on soils, such as increased
salinity, caliche formation, etc. (Cortina and Vallejo, 2004). On the other
hand, positive interactions may result from increases in resource availability,
amelioration of microclimate and soil conditions, increases in pollination and
propagule dispersal rates, and defense against pathogens and herbivores
Positive and negative interactions are highly dynamic in drylands, and
their balance may depend on site properties, climatic conditions, species
identity and development stage (Pugnaire et al., 1996b; Maestre and Cortina,
2004c; Gómez-Aparicio et al., 2004). Bertness and Callaway (1995), and
Callaway and Walker (1997) suggested that positive interactions should be
more intense under high environmental stress or consumer pressure, and
should depend on the size of the facilitator. Within individual shrub canopies,
soil resources and microclimate show complex spatial patterns (Halvorson et
al., 1994; Moro et al., 1997), and their interaction promote the emergence of
different niches that increase in number and availability as shrubs increase in
size (Pugnaire et al., 1996a). Increased habitat availability, together with the
amelioration of the harsh climatic conditions, promote an increase in the
strength of facilitative interactions with the increase in the size of the
facilitator canopy (Callaway and Walker, 1997). Other plant traits –e.g.
related to the capacity to concentrate resources or to deter predators- can be
considered to estimate the ‘nursing’ power of a given species. For example,
the capacity to build up nebkhas or mounds of sediments underneath plant
canopies has been associated with canopy compactness (density) and openess
(a trait related to the presence of branches at the soil level), as well as to the
degree of mycorrhization (Carrillo-García et al., 1999). According to these
authors, dense canopies, whether open or closed, should be prone to nebkha
formation, whereas only open ones would allow the presence of understorey
plants. These observations are likely to depend on slope. Evidences on the
superior capacity of shrubs to alter soil properties are, however, not
conclusive (Mazzarino et al., 1996; Schlesinger et al., 1996). The capacity to
accumulate litter may also be relevant to establish nurse-protegé interactions,
although evidence on the net effect of litter on seedling emergence and
establishment is contrasting (Fowler, 1988; Owens et al., 1995; Milton, 1995).
Functional matrices, as those proposed by Ervine and Chapin (2003) to
characterize plant capacity to alter soils, could incorporate other relevant traits
to describe nursing capacity.
It has been suggested that nurse-protegé interactions are more common in
arid and semiarid communities than in other environments (Flores and Jurado,
2003). According to these authors, the Fabaceae and Mimosaceae (which may
be capable of fixing atmospheric N) are among the most common nurse
families (20% and 7% of the reported species, respectively), suggesting that N
inputs may be a major driver of positive interactions in drylands. The capacity
to fix N of a potential nurse plant is not, however, solid evidence of
nutritional facilitation (Barnes and Archer, 1999). Defense against grazing
and trampling may also be important, as many nurse species have thorns
(Acacia spp., Prosopis spp., Cactaceae, with 11%, 4% and 5% of the nurse
species, respectively), and unpalatable leaves.
As previously mentioned, nurse-protegé relations are commonly the
combination of independent factors, both positive and negative. However,
very few studies have performed manipulative field experiments to dissect the
net effects of a given plant-plant interaction into their underlying positive and
negative effects (Holzapfel and Mahall, 1999; Maestre et al., 2003a). This
knowledge is necessary to understand community dynamics and to develop
sound management programs. We evaluated the effects of the perennial
tussock grass Stipa tenacissima on the native late-successional shrub Pistacia
lentiscus in semiarid Mediterranean steppes. Stipa tenacissima (alpha grass) is
a tussock grass distributed in the western Mediterranean basin from arid to
dry subhumid conditions (100-500 mm; White, 1983). It is one of the
dominant species in steppes, that have been strongly affected by human
activities carried out during centuries, such as wood harvesting and fiber
cropping (Barber et al., 1997). After cessation of human activities, shrub
patches that were once part of these steppes hardly recover because of past
management practices and inherent restrictions to plant growth. Shrub patches
are, however, crucial for the composition, stability and function of semiarid
steppes, despite their low contribution to the total plant cover (Maestre and
Cortina, 2004a). In addition, the area covered by late-successional sprouting
shrubs is the most influencing individual variable on perennial plant species
richness in these steppes (Maestre, 2004).
Stipa tenacissima tussocks can retain runoff and sediments from upslope,
a process that affects their own performance (Puigdefábregas et al., 1999).
These tussocks also contribute to resource concentration, for example by
favouring infiltration underneath the canopy (Cerdà, 1997). As a consequence
of the changes in soil, microclimate and biological conditions, woody
vegetation gets preferentially established close to alpha canopies (Maestre et
al., 2001, 2003a), where they can withstand higher degrees of climatic stress
(García-Fayos and Gasque, 2002; Maestre et al., 2003a). Previous studies had
shown that S. tenacissima improved soil fertility (Puigdefábregas and
Sánchez, 1996; Bochet et al. 1999; Aïdoud et al., 1999), but not nitrogen
content (Bessah et al., 1999), reduced irradiation and soil temperature
(Maestre et al., 2001), and received runoff water (Puigdefábregas et al., 1999)
as compared to adjacent bare ground areas. We wanted to estimate the weight
of each of these factors in increasing the survival and growth of introduced
woody seedlings. We established a manipulative experiment in which we
planted seedlings of a common shrub species, Pistacia lentiscus, upslope of
alpha tussocks that were either undisturbed, herbicided or bended, or upslope
of alpha tussocks where runoff had been excluded (Maestre et al., 2003a). We
also planted P. lentiscus seedlings in undisturbed open areas. Finally, we
carried out a laboratory experiment to test the effect of soil properties on P.
lentiscus seedlings. In contrast to our expectations, runoff did not affect
seedling survival. Seedlings planted in alpha soil showed a trend towards
better growth and nutritional status, but the overall effect of soil type was not
statistically significant. Finally, shadow was the most important factor
affecting seedling establishment. So, in quantitative terms, the three factors
ranked: shade>>soil fertility>runoff. No grazing on Pistacia lentiscus was
We performed a similar experiment with Pinus halepensis in semiarid
plantations of SE Spain. It is one of the most common tree species in the
Mediterranean basin, indeed one of the few tree species that can thrive under
semiarid conditions. Pinus halepensis forests have extended in the last
decades due mainly to afforestation (Pausas et al., 2004). Under semiarid
conditions, P. halepensis plantations show poor growth and cover, and
spontaneous colonisation by sprouting shrubs is scarce (Maestre and Cortina,
2004b). Community composition (e.g. bird richness) has been related to the
abundance and structure of shrubs in these forests (López and Moro, 1997).
Pine canopies frequently show a relatively dense herbaceous understorey
dominated by the perennial grass Brachypodium retusum (Bautista and
Vallejo, 2002), whereas the establishment of woody seedlings in these
microsites seems to be impeded (Maestre et al., 2003b). We established a
series of field and glasshouse experiments to evaluate the relative importance
of direct interactions (soil fertility, allelochemicals, shadow and water
availability, and competition) and indirect interactions (through pine effects
on the herbaceous layer) (Maestre et al., 2004). We found that the direct
effects of pine on introduced seedlings, including competition, were rather
small. Thus, pine death by girdling, and the resulting decrease in belowground
competition, while maintaining a protective shadow, did not affect seedling
performance. In contrast, suppression of the herbaceous layer greatly
increased seedling survival and growth.
In a review of 31 studies on facilitation we found that improved soil
properties, including increased nutrient availability, was the most common
mechanism of facilitation that was mentioned (25 out of 31 studies). Of these,
8 studies attributed the positive interactions to higher nitrogen availability.
Half of the studies mentioned shade as an important driver of positive
interactions. Finally, only three studies attributed facilitation to the presence
of litter. However, most of the studies reviewed were observational. On the
other hand, for logistic reasons manipulative experiments commonly use
simplified designs, focusing on integrated factors (such as soil fertility or
shading) rather than specific processes. So our knowledge on the relative
importance of the various drivers of facilitation in drylands is still very poor,
and does not allow for sound generalizations.
IMPLICATIONS FOR ECOSYSTEM MANAGEMENT
Being a priority in land management in a wide variety of biomes, the
restoration of degraded ecosystems is especially important in drylands, as they
are being degraded and desertified at a fast rate throughout the globe
(Reynolds, 2001, Abahussain et al., 2002; Reynolds and Stafford-Smith,
2002). Degraded ecosystems in drylands are usually characterised by a
reduced plant cover and impoverished plant species diversity (Jauffret and
Lavorel, 2003; Maestre, 2004; Cortina et al., in press). Despite the specific
objectives of their restoration that may differ depending on the degree of
degradation, and on climatic, biotic and socio-economic constraints,
restoration programs often aim to increase plant cover by directly introducing
plant individuals, primarily woody species (Whisenant, 1999; Young, 2000;
Vallejo et al., 2000, Cortina et al., in press). This management action is
crucial to stop further degradation, to combat desertification and to foster the
recovery of the structure, composition and function of degraded ecosystems in
these areas (Castillo et al., 1997; Reynolds, 2001; Cortina and Vallejo, 2004).
However, if the target area is extensively degraded, restoration efforts could
be initiated with actions focusing on the recovery of ecosystem structure by
increasing the number of patches and reducing the downslope distance
between them. This can be done by inserting brush piles parallel to land
contours. Experiments conducted in Australia have shown the effectiveness of
this technique in creating fertile patches and ultimately rehabilitating
degraded landscapes (Ludwig and Tongway, 1996; Tongway and Ludwig,
1996). Such brush piles would reduce soil and nutrient losses, and would act
as filters rather than barriers. They would also provide suitable microsites for
enhancing the establishment, growth and survival of perennial plants in the
short term. Once this intervention has reduced degradation, the next step to
restore these systems should be the introduction of seedlings of native woody
If the target area holds some plant cover, restoration efforts can take
advantage of facilitative interactions among plants. Positive interactions, and
among them, those mediated by changes in soil properties, can be crucial to
maintain the integrity of dryland ecosystems. Accordingly, they can be very
helpful if not essential for reassembling pieces of degraded ecosystems. In the
previous section we have seen that there are numerous evidences of
facilitative interactions in drylands. These interactions can be used to promote
the establishment of species of interest, and this has been increasingly
recommended as a restoration technique (Maestre et al., 2001; Castro et al.,
2002; Gómez-Aparicio et al., 2004). On the other hand, a thorough
understanding of the mechanistic basis of positive interactions may allow the
identification of the main drivers of facilitation, and thus their use in
restoration. For example, shadow may be crucial in the first stages of seedling
establishment. Nurse plants may not be always available, or their use may be
restricted for logistic or economic reasons. In this case, the use of
ecotechnological tools to reduce incoming radiation such as treeshelters, can
be a suitable alternative to improve seedling performance (Cortina et al., in
press). This technique is also convenient when consumer pressure is high.
Amelioration of soil fertility can be easily achieved by using various types of
soil amendments. Residues with high SOM and nutrient content, such as
composted domestic refuses and sewage sludge are becoming increasingly
available (Valdecantos et al., 2002; Fuentes et al., 2002a, 2002b; Valdecantos
et al., in press). As the quality of these products improves, they will be
increasingly used for the restoration of areas where soil fertility hampers
succession. Nowadays, there is a vast array of techniques used in ecosystem
restoration, deriving from mechanisms like those previously described (Table
2). Further knowledge on this type of interactions will help to develop new
ecotechnology in this area, and to improve the success of restoration
Table 2. Some techniques used for the restoration of degraded semiarid ecosystems and the
ecosystem component and main processes that are affected. From Tongway et al. (2004).
1Ludwig and Tongway (1996), 2Tongway and Ludwig (1996), 3Bonet (2004), 4Wunderlee
(1997), 5Cortina et al. (2001), 6Valdecantos et al. (2002), 7E. De Simón, pers. com., 8Hillel
(1991), 9Lavee et al. (1997), 10Vallejo et al. (2003), 11Bellot et al. (2002), 12Azcón and Barea
(1997), 13Caravaca et al. (2003), 14Maestre et al. (2002b), 15Whisenant et al. (1995), 16Boeken
and Shachak (1994), 17Buttars et al. (1998), 18Belnap (1993), 19Vilagrosa et al. (1997).
Throughout the 20th century drylands have been the focus of an
impressive amount of research. In some cases, findings have supported
traditional knowledge and practices. But very often research has provided
entirely new perspectives on dryland functioning. Still, it is somewhat
discouraging that some of the questions that we are trying to answer, such as
those on the rooting depth of some species, or on the effects of site
preparation and the use of organic amendments, were already posed more
than 2,000 years ago (e.g. Theophrastus, Peri Phyton Historia). This delay
may partly result from the vast diversity of interactions that are involved. In
this review, we have seen that interactions are strongly dependent on plant
identity, or as a first approach, to plant functional types (Eviner and Chapin,
2003). We have also seen that plant-plant interactions may change depending
Branches, mulch, etc. Sinks
Sediment, runoff and seed
eventually water and
Perches Birds rest 3,4
Organic amendments and
Stones around introduced
Islands of fertility Local soil improvement 5,6
Patches with low
Treeshelters Nurse plants
incomming radiation and
Increase in resource
against pathogens and stress
Resource capture, mainly
water. Increase in available
Soil protection, runoff
Processes associated with
Field and nursery
Cyanobacteria inoculation Biological crusts 17,18
Nurse species plantation Nurse plants 19
on site properties, climatic conditions, or plant age. However, we know very
little on the importance of such interactions on community dynamics and
large-scale processes. In most cases, we just guess they are relevant. For
example, we still don’t know how important for succession is disturbance that
may be followed by a short-term unfavourable climatic period. Must we
asume that the interaction observed in a particular short-term study will
prevail in the long-term, and will control community organization? When will
facilitative interactions turn into competitive – or positive net balances turn
into negative- as plants grow? It is clear that more research is needed in this
area; particularly manipulative experiments, long-term studies and modelling.
This review has also shown that we can successfully estimate plant effects
on background soil fertility. But as our knowledge on some areas of soil
science such as soil biochemistry, soil microbiology and soil fauna increases,
we realize that some plant-plant interactions that were previously attributed to
other factors are actually occurring in the soil arena. This information is
needed to evaluate the relative importance of soil processes on plant-plant
interactions and community dynamics.
Drylands are threatened by a combination of natural factors and human
activities (Reynolds and Stafford Smith, 2002). Further knowledge on the
effects of plants on soils, and on the importance of such changes on long-term
ecosystem dynamics will help to improve dryland management and
We thank Dan Binkley and Oleg Menyailo for organizing the NATO
Advanced Research Workshop Trees and Soil Interactions. Implications to
Global Climate Change. This minireview has been possible thanks to funding
from NATO, EC DGII (project CREOAK, QLRT-2001-01594), and CICYT-
FEDER (projects FANCB, REN2001-0424-C02-02 / GLO, and BIOMON,
REN2000-0181P4-03). Fernando T. Maestre is a Fulbright fellow, and greatly
acknowledges support from the Dirección General de Universidades and
Fondo Social Europeo. We thank our colleagues from Fundación CEAM and
the Department of Ecology, Universitat d´Alacant for fruitful discussions on
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