ARROYO 2013 PlANT B DIVERSITY IN FRAGMENTED RAIN FORESTS TESTING

Journal of Ecology 2013, 101, 1449–1458
doi: 10.1111/1365-2745.12153
Plant b-diversity in fragmented rain forests: testing
floristic homogenization and differentiation hypotheses
€ s2, Federico Escobar2, Felipe P. L. Melo3,
Víctor Arroyo-Rodríguez1*, Matthias Ro
ulio A. Santos4, Marcelo Tabarelli5 and Robin Chazdon6
Bra
1
noma de Me
xico, Morelia, Michoaca
n 58190,
Centro de Investigaciones en Ecosistemas, Universidad Nacional Auto
Mexico; 2Red de Ecoetología Instituto de Ecología A.C., Xalapa, Veracruz 91000, Mexico; 3Departamento de Ecologia
tica e
Aplicada, Universidade Federal de Pernambuco, Recife, PE 50670-901, Brazil; 4Departamento de Sistema
~o Pessoa, PB 58051-900, Brazil; 5Departamento de Bota
^nica,
Ecologia, Universidade Federal da Paraíba, Joa
Universidade Federal de Pernambuco, Recife, PE 50670-901, Brazil; and 6Department of Ecology and Evolutionary
Biology, University of Connecticut, Storrs, CT 06269-3043, USA
Summary
1. Land-use change is the main driver of global biodiversity loss, but its relative impact on species
turnover (b-diversity) across multiple spatial scales remains unclear. Plant communities in fragmented rain forests can undergo declines (floristic homogenization) or increases (floristic differentiation) in b-diversity.
2. We tested these alternative hypotheses analysing a large vegetation data base from a hierarchically nested sampling design (450 plots in 45 forest patches in 3 landscapes with different deforestation levels) at Los Tuxtlas rain forest, Mexico. Differences in b-diversity across spatial scales (i.e.
among plots, among patches, and among landscapes) were analysed using multiplicative diversity
decompositions of Hill numbers.
3. Plant b-diversity among plots within forest patches decreased in landscapes with higher deforestation levels, leading to floristic homogenization within patches. This homogenization process can be
explained by the loss of rare and shade-tolerant plant species, and the recruitment and dominance of
disturbance-adapted species, and can limit the accumulation of species (c-diversity) in landscapes
with higher deforestation.
4. Nevertheless, the landscape with the highest deforestation level showed the highest floristic differentiation among patches. This landscape showed the greatest isolation distances between patches;
a landscape spatial pattern that can limit the interchange of seeds (and species) between patches.
Because the study patches are undergoing secondary succession following disturbances (e.g. logging,
edge effects), different disturbance regimes and increased distance among patches could lead to
higher b-diversity.
5. Synthesis. These findings indicate that patterns of floristic homogenization and differentiation
depend on the landscape configuration and on the spatial scale of analysis. At the landscape scale,
our results suggest that, in accordance with non-equilibrium dynamics and the landscape-divergence
hypothesis, patches located in landscapes with different forest cover and different connectivity can
experience contrasting successional pathways due to increasing levels of compositional differentiation between patches. These novel findings add further uncertainties to the maintenance of biodiversity in severely deforested tropical landscapes and have key ecological implications for biodiversity
conservation planning.
Key-words: community assembly, compositional dissimilarity, diversity partitioning, environmental filtering, habitat fragmentation, multiscale analysis, secondary succession, wet forest
*Correspondence author. E-mail: [email protected]
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society
1450 V. Arroyo-Rodríguez et al.
Introduction
Tropical forests are undergoing accelerated and extensive
land-use changes world-wide, and fragmented forests have
replaced continuous tracks of forest in most regions (FAO
2011). Because tropical forests maintain the most biologically
diverse communities on Earth, an increasing fraction of global
biodiversity is living in forest patches embedded in matrices
dominated by agricultural lands (Perfecto & Vandermeer
2008; Gibson et al. 2011). In these emerging environments, a
myriad of threatening factors can alter the composition and
structure of remaining assemblages (Gardner et al. 2009);
however, the relative impact of such a new habitat configuration on the spatial dissimilarities in species composition
(b-diversity) remains poorly understood, particularly in the
tropics (Karp et al. 2012).
Beta diversity is a key component of species diversity, which
is essential for understanding how species diversity is organized
and maintained (Condit et al. 2002; Flohre et al. 2011). Patterns of b-diversity among localities can be related to processes
operating at both local and regional scales (Lawton 1999). At
local scales, niche structure, biological interactions and environmental characteristics play a more important role (Ricklefs
1987), and all these attributes are known to be altered in fragmented forests (Fischer & Lindenmayer 2007; Gardner et al.
2009). Yet evolutionary and biogeographical features such as
dispersion, extinction and colonization processes are more
important factors limiting species turnover at larger spatiotemporal scales (Ricklefs 1987; Myers et al. 2013), particularly in
fragmented tropical landscapes, where metapopulation dynamics are expected to determine patterns of patch occupancy and
long-term population persistence (Perfecto & Vandermeer
2008). Thus, a multiscale approach is required to accurately
assess the impact of land-use change on b-diversity (Karp et al.
2012) and to have a better understanding of the mechanisms
that contribute to the maintenance of species diversity in fragmented landscapes (Wiens 1989; Harrison & Cornell 2008).
This understanding is particularly needed for plant communities, as they represent key structural and functional elements
of forest ecosystems. Tropical rain forests not only maintain
very high local plant species richness (a-diversity), but also
outstanding rates of species turnover, conferring a very high
regional (c) species diversity to these forests (Condit et al.
2002; Macía et al. 2007). This turnover has been related to
environmental variability within and among sites (e.g. Vormisto et al. 2000; Condit et al. 2002; Tuomisto, Ruokolainen
& Yli-Halla 2003; Macía et al. 2007), as well as to geographical distances that limit plant dispersion (Hubbell 2001; Condit et al. 2002; Myers et al. 2013). However, the spatial
distribution and determinants of plant b-diversity have been
rarely studied in fragmented landscapes, and hence, the
response of this key component of species diversity to
changes in the landscape configuration is unclear, but crucial
for understanding the biological dynamics in fragmented landscapes (but see Galanes & Thomlinson 2009).
Evidence from fragmented rain forests indicates that many
tree species can disappear from small forest patches,
particularly those with large seeds, emergent and/or shade tolerant (Silva & Tabarelli 2000; Laurance et al. 2006; Santos
et al. 2008; Tabarelli, Peres & Melo 2012). This local plant
extirpation has been related to both biotic and abiotic conditions that constrain key ecological processes such as seed dispersal, seedling recruitment, growth and survival in smaller
forest patches and along forest edges (e.g. Laurance et al.
2002, 2006; Santos et al. 2008). A number of disturbanceadapted native species (e.g. short- and long-lived lightdemanding species), however, can thrive in fragmented forests
with pronounced edge effects (Laurance et al. 2006; Santos
et al. 2012). In some cases, such species can even dominate
plant communities, contributing to increased levels of floristic
homogenization (i.e. reduced b-diversity) across multiple spatial scales (L^
obo et al. 2011; Tabarelli, Peres & Melo 2012).
Furthermore, remaining forest patches in fragmented landscapes are usually restricted to steep slopes, particularly in
sharply undulating terrains, such as the Los Tuxtlas mountain
range in Mexico (Guevara, Laborde & Sanchez-Ríos 2004),
potentially reducing habitat heterogeneity and b-diversity
among forest patches and across landscapes.
Alternatively, increases in b-diversity can also emerge in
fragmented rain forest landscapes, as fragmentation limits interpatch dispersal patterns, and can therefore promote the
compositional differentiation between forest patches and landscapes (Hubbell 2001). This is particularly expected, as the
defaunation process that typically occurs in fragmented rain
forests usually reduces animal-mediated seed dispersal and
seedling recruitment in forest patches (Melo, Dirzo & Tabarelli 2006). Thus, because old-growth forest patches tend to
undergo regeneration related to different post-fragmentation
disturbances (e.g. logging, edge effects), different disturbance
regimes and increased distance among patches can lead to
higher levels of b-diversity. The differences in altitude (Guevara, Laborde & Sanchez-Ríos 2004; García-Aguirre et al.
2010) and/or in other patch and landscape features (e.g.
microclimatic conditions) can also contribute to higher levels
of community divergence among patches and landscapes
(Laurance et al. 2007). This floristic differentiation can drive
contrasting successional trajectories in fragmented landscapes,
depending on the biotic and abiotic characteristics of the sites,
the life-history traits of the species established, and the temporal and spatial contexts of the landscapes (Laurance et al.
2007; Chazdon 2008), potentially affecting the maintenance
of biodiversity in these human-modified landscapes (Chazdon
et al. 2009; Melo et al. 2013).
We tested these two alternative hypotheses (i.e. floristic
homogenization vs. differentiation) evaluating differences in
plant b-diversity in the Los Tuxtlas region of Mexico; a
strongly deforested but highly diverse mountainous tropical
forest (Arroyo-Rodríguez et al. 2009). We adopted a multiscale approach, in which three landscapes with different landscape spatial patterns were compared. We used a large
vegetation data set from a hierarchically nested sampling
design (i.e. 450 plots located in 45 forest patches from three
fragmented landscapes), to evaluate differences in b-diversity
across three spatial scales: among plots, among forest patches,
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1449–1458
Plant b-diversity in fragmented rain forests 1451
and among landscapes. This approach is needed in fragmentation studies because the response of species to changes in the
landscape spatial pattern can strongly vary among landscapes
(Laurance et al. 2007; Arroyo-Rodríguez et al. 2009; Pardini
et al. 2010). Through this approach, we can test whether the
ecological mechanisms that govern the structuring of tree species composition in fragmented landscapes operate at small
and/or at larger spatiotemporal scales.
Variations in b-diversity were assessed with multiplicative
diversity decompositions of Hill numbers, one of the most
appropriate analytical approaches for diversity evaluations
(Jost 2006; Tuomisto 2010; Chao, Chiu & Hsieh 2012). With
this method, we assessed the relative importance of rare and
dominant species on b-diversity (Tuomisto 2010; Chao, Chiu
& Hsieh 2012). We thus identified the spatial scale (from
plots to landscapes) and abundance level (from rare to
dominant species) at which we observed higher differences in
b-diversity. If deforestation and forest fragmentation lead to a
process of floristic homogenization, b-diversity within and
among patches will decrease in landscapes with higher deforestation; however, if plant communities are undergoing a process of floristic differentiation, b-diversity will increase in
landscapes with higher deforestation. Because a large number
of plant species in the tropics have small populations, we
also predicted that b-diversity will be higher when considering rare species than when considering dominant species, particularly in the landscape with a lower extent of
deforestation, in which we expected to find a higher number
of rare species. Finally, to have a better understanding of the
determinants of b-diversity among patches, we tested whether
this component of species diversity was related to patch location (i.e. geographical distances between sampled patches)
and/or to interpatch differences in patch size, altitude and/or
terrain’s slope, as all these patch-scale metrics differed significantly among landscapes, and they can determine the environmental conditions (e.g. soil, weather characteristics) within
the patches. In this case, we expected higher b-diversity values between those patches more isolated between each other
and between those with higher differences in size, altitude
and/or slope.
Materials and methods
STUDY SITE
The Los Tuxtlas region is located in the south-east of Veracruz State,
Mexico (18°8′–18°45′ N, 94°37′– 95°22′ W; Fig. 1). The climate is
warm and humid, with a mean annual temperature of 25 °C and
annual rainfall between 3000 and 4600 mm. The elevation ranges
from 0 to 1738 m at sea level (a.s.l.). The region is volcanic in origin, and dates from the Tertiary up to the Plio–Pleistocene. The last
eruption (San Martín Volcano, 1738 m a.s.l.) was in 1793, and lava
flows cover most of the region, mainly as basalt. Thus, soils are
mainly developed from volcanic ash of basaltic composition (GarcíaAguirre et al. 2010). Los Tuxtlas was decreed a Biosphere Reserve in
1998, because of its high and unique biodiversity. The Reserve covers
an area of 155122 ha and represents the northern limit of tropical rain
forest distribution in the Neotropics. The original dominant vegetation
type below 700 m a.s.l. was tropical rain forest, but it was heavily
deforested and fragmented since 1970s, and the remaining rain forest
(c. 5%) is mainly surrounded by a matrix of pastures and croplands
(Guevara, Laborde & Sanchez-Ríos 2004).
EXPERIMENTAL DESIGN AND VEGETATION SAMPLING
The sampling was carried out between June 2004 and July 2005
using a hierarchically nested sampling design (Arroyo-Rodríguez
et al. 2009, 2012). We selected four hierarchical spatial scales, from
larger to smaller: region (n = 1), landscapes (n = 3), forest patches
(n = 15 per landscape) and plots (n = 10 per forest patch). Based on
the remaining forest cover, within the region, we delimited three landscapes (Fig. 1), considering that they (i) represent a gradient of rain
forest deforestation, (ii) are situated between 0 and 400 m a.s.l. (to
avoid larger changes in vegetation related to soil and altitude; see
García-Aguirre et al. 2010), (iii) all have similar areas (c. 5000 ha)
and ages of fragmentation (c. 50 years), and (iv) are naturally delimited by the coast and large rivers.
The landscapes were digitized through ARCVIEW 3.2 software (Environmental Systems Research Institute, Redlands, CA, USA) using
recent (2001) aerial photographs (1 : 20000), orthophotographs and
field information. Also, a 15-m-resolution digital elevation model from
the National Institute of Geography and Information (INEGI, freely
available from http://www.inegi.gob.mx) was processed to obtain the
altitude and terrain’s slope of each forest patch, that is, considering all
the 15-m pixels that constituted each patch. Although the three landscapes have been highly deforested, we found notable differences in
deforestation levels (Fig. 1; Table S1 in Supporting Information). The
landscape with the lowest deforestation level (LDL) covers 5356 ha,
24% of which corresponded to rain forest distributed among 75
patches ranging from 0.5 to 700 ha; the landscape with intermediate
deforestation level (IDL) covers 4965 ha, 11% of which correspond to
rain forest distributed among 88 patches ranging from 0.5 to 76 ha;
and the landscape with the highest deforestation level (HDL) covers
5046 ha, with only a 4% of remaining rain forest in 46 patches ranging from 0.5 to 68 ha. Mean patch size and shape complexity was significantly greater in LDL than in the other two landscapes; while mean
distance to the nearest patch was significantly greater in HDL, patches
within HDL had significantly lower (flatter) slopes and were located at
lower altitude than patches in LDL and IDL (Fig. 1; Table S1).
Within each landscape, we sampled vegetation in 15 randomly
selected forest patches using Gentry’s (1982) protocol. We chose this
method because it is logistically simple, economical and appropriate
for the analysis of species diversity in tropical forests (Gentry 1982).
Within each patch, we randomly located 10 50 9 2 m plots (0.1 ha),
excluding a 10-m-wide edge. Edge effect exerts its greatest influence
within 100 m of the forest edge (Laurance et al. 2002), but we could
not exclude a 100-m-wide edge because most of the patches sampled
(38 vs. 7 patches) did not have interior habitat after eliminating
100 m of edge from each patch (see Arroyo-Rodríguez et al. 2009).
Thus, edge effects are expected to be important in most patches.
Within each plot, we recorded all species of trees, shrubs, lianas,
palms and herbs (i.e. Heliconiaceae) with diameter at breast height
(DBH) ≥ 2.5 cm. Species not identified in the field were collected
for subsequent identification in the MEXU (Institute of Biology,
UNAM, Mexico City) and XAL (Institute of Ecology A.C., Xalapa,
Veracruz) herbaria (see further details in Arroyo-Rodríguez et al.
2009, 2012). In general, the tree inventory was reasonably accurate
with our sampling effort, with 84–90% of the known species
recorded in each landscape (see Arroyo-Rodríguez et al. 2009, 2012).
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1449–1458
1452 V. Arroyo-Rodríguez et al.
Fig. 1. Location of the three landscapes
studied in the Los Tuxtlas Biosphere
Reserve, south-eastern Veracruz, Mexico.
Black polygons represent studied forest
patches (LDL, lowest level of deforestation,
24% of remaining forest cover; IDL,
intermediate deforestation level, 11%; HDL,
highest deforestation level, 4%).
DATA ANALYSES
q
To analyse patterns of tree species turnover at multiple spatial scales,
we used multiplicative diversity decompositions of effective numbers
of species (so-called Hill numbers) in its unweighted form (Jost 2007;
Jost, Chao & Chazdon 2011). Hill numbers (qD) represent true diversities, as they obey the replication principle (Jost 2007; Tuomisto
2010). They are in units of ‘species’, and hence, they can be plotted
on the same graph to construct diversity profiles that can be useful to
characterize the species abundance distribution of a community and to
provide complete information about its diversity (Chao, Chiu & Hsieh
2012). For S species, gamma (c) diversity of order q is defined as:
q
Dc ¼
X
S
q
p
i¼1 i
1=ð1qÞ
eqn 1
i denotes the mean relative abundance of the ith species in
where p
the N communities (Jost 2007; Tuomisto 2010), and q is a parameter
that determines the sensitivity of the measure to the relative abundances. Because this measure is undefined for q = 1, c-diversity of
order 1 can be estimated as:
X
S
1
p
Dc ¼ exp log
p
eqn 2
i
i¼1 i
When q = 0, diversity represents the species richness, which is not
sensitive to abundances and so gives disproportionate weight to rare
species (Jost 2006; Tuomisto 2010). When q = 1, diversity is equivalent to the exponential of Shannon’s entropy index, and weights each
species according to its abundance in the community, without favouring rare or abundant species (Jost 2007). The Hill number of order 1
can be therefore interpreted as the number of ‘typical species’ in the
community (Chao, Chiu & Hsieh 2012). Finally, if q = 2 (equivalent
to the inverse Simpson concentration), abundant species are favoured
and rare species are discounted, and hence, this diversity can be interpreted as the number of ‘very abundant’ or ‘dominant’ species in the
community (Jost 2010; Chao, Chiu & Hsieh 2012).
When considering several communities, alpha (a) component of
diversity can be analysed with the following formula (Jost 2007):
Da ¼
X
1=ð1qÞ
1
S
1 XS q
q
p
þ
p
þ.
.
.
i1
i2
i¼1
i¼1
N
N
eqn 3
where pi denotes the relative abundance of the ith species in each of
the N communities. Again, for the particular case of q = 1, a-diversity can be estimated as:
XS
1 XS
1
Da ¼ exp ðpi1 ln pi1 Þ þ
ðpi2 ln pi2 Þ þ . . .
i¼1
i¼1
N
eqn 4
Then, using the multiplicative diversity partition, b-diversity is calculated as: qDb = qDc/qDa. This beta can be interpreted as ‘effective
number of completely distinct communities’ (Jost 2007), as it ranges
between 1 (when all communities are identical) and N (when all communities are completely different from each other).
To evaluate changes in b-diversity at multiple spatial scales, we
considered the following decompositions, from larger to smaller: (i)
b-diversity between landscapes: qbland = qcregion/qaland; (ii) b-diversity
between patches: qbpatch = qcland/qapatch ; and (iii) b-diversity between
plots: qbplot = qcpatch/qaplot . To assess whether the magnitude in bdiversity differed across spatial scales, we compared the relative compositional dissimilarity between communities using the transformation
of beta (qDb) proposed by Jost (2006) for communities with different
numbers of samples (i.e. landscapes: n = 3; patches: n = 15; plots:
n = 10): qDS = 1 - [(1/qDb – 1/N)/(1 – 1/N)] , where N is the number
of samples. qDS = 1, when all the samples are completely distinct,
and qDS = 0, when all are identical.
To test whether patterns of b-diversity within patches differed
among landscapes, we used generalized linear models. We selected a
normal distribution and an ‘identity’ link function after verifying that
q
bplot’s errors fit the normal distribution (Shapiro–Wilk test) (Crawley
2002). When we found differences among landscapes, we used contrast tests to identify which landscapes differed from each other (Crawley 2002). Finally, we used Mantel tests to assess whether variations
in plant b-diversity between patches were related to patch location
(i.e. log-transformed geographical distances between sampled patches)
and/or to differences between patches in size, altitude and slope. The
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1449–1458
Plant b-diversity in fragmented rain forests 1453
P-values were calculated using the distribution of the R coefficients
obtained from 10 000 permutations. All analyses were performed
using the R statistical software (v. 2.15.2) R Development Core Team,
Vienna, Austria.
this landscapes, while IDL and HDL landscapes were dominated by 3 and 2 species, respectively (Fig. 3). The highest
proportion of singletons and doubletons was found in LDL
(24.4% of species) and HDL (23.8%); only 9.3% of species
in IDL were singletons and/or doubletons (Figs 2 and 3).
Results
OVERVIEW ON PLANT SPECIES DISTRIBUTION
We recorded 9451 plants belonging to 75 families, 207 genera and 364 species in 4.5 ha (450 plots). The number of species varied among landscapes, being higher in the landscape
with the lowest deforestation level (LDL, 242 species), followed by the landscape with the highest deforestation level
(HDL, 181 species) and the landscape with intermediate
deforestation level (IDL, 161 species; Fig. 2a). Despite the
fact that we registered more individuals in IDL, the mean
individual-based accumulation curves and the 95% confidence
intervals showed that the number of species was significantly
higher in LDL than in IDL and HDL (Fig. S1). In general,
most species (c. 60%) were restricted to one landscape, and
only 20.3% were distributed in all three landscapes (Fig. 2b).
The species restricted to one landscape were more frequent in
LDL, as almost half (i.e. 114 of 242 species) of the species
located in this landscape had a restricted distribution. LDL
was by far dominated by a single palm species (Astrocaryum
mexicanum), composing 11% of all individuals sampled in
PATTERNS OF b-DIVERSITY ACROSS SPATIAL SCALES
As predicted, b-diversity among landscapes (i.e. within the
entire region) was relatively higher when considering rare
species (0bland = 1.9), than when considering typical
(1bland = 1.6) and abundant (2bland = 1.5) species (Fig. 4a). In
LDL and HDL, b-diversity among patches was higher than
within the entire region (Fig. 4a). In contrast, IDL showed
the lowest level of b-diversity among patches for any order q
(Fig. 4b), as in this landscape we observed the highest
percentage of species with shared distribution both at the
landscape (Fig. 2) and patch scales (Fig. S2). As found in the
entire region (Fig. 4a), b-diversity among patches in LDL
was 1.6 times greater when considering rare species
(a)
(b)
Fig. 2. Plant species distribution in three landscapes with different
deforestation levels in Los Tuxtlas, Mexico. The number (and percentage) of species is indicated in each case. We show differences
among landscapes in (a) species with both restricted and shared (i.e.
species present in more than one landscape) distribution; and (b) species distributed in one, two or three landscapes. From the species
restricted to one landscape, we indicate the number of singletons (i.e.
species represented by only one individual). Landscapes: LDL, lowest
deforestation level; IDL, intermediate deforestation level; HDL, highest deforestation level.
Fig. 3. Relative abundance of plant species (log10 pi) sampled in
three landscapes with different extents of deforestation located in Los
Tuxtlas, Mexico (LDL, lowest deforestation level; IDL, intermediate
deforestation level; HDL, highest deforestation level). The x axis
ranks each species in order from most to least abundant. Species representing c. 30% of all individuals in each landscape are indicated, as
well as the species with one and two individuals (singletons and doubletons, respectively). Underlined species dominate tree communities
in the three landscapes. Superscript letters refer to their light requirements: ST, shade-tolerant (old-growth forest species); P, pioneers
(early colonizers); and LD, light-demanding (late colonizers) species.
Tree species: A.mex., Astrocaryum mexicanum; C.chi., Croton schiedeanus; D.arb., Dendropanax arboreus; H.amp., Hampea nutricia;
L. mex., Lunania mexicana; M.lon, Myriocarpa longipes; O. obl.,
Orthion oblanceolatum; P. chi., Psychotria chiapensis; R. gua., Rinorea guatemalensis; R. muc., Rollinia mucosa; S. and., Siparuna andina; S. don., Stemmadenia donnell-smithii; T. mex., Tapirira
mexicana; V.gua., Vochysia guatemalensis.
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1449–1458
1454 V. Arroyo-Rodríguez et al.
(a) 0.9
Region
LDL patch
Compositional dissimilarity
0.8
IDL patch
HDL patch
0.7
LDL plot
IDL plot
0.6
HDL plot
0.5
0.4
(b) 4.5
LDL
4
IDL
HDL
β patch
3.5
3
2.5
2
1.5
(c)
6
LDL
IDL
β plot
5
HDL
4
3
2
q=0
q=1
q=2
(i.e. 0bpatch) than when considering abundant species (2bpatch).
In contrast, HDL showed the highest b-diversity of dominant
species among patches, even greater than among landscapes
and among plots (Fig. 4a). The turnover of rare species
among patches in this landscape was also very high and very
similar to that observed in LDL (Fig. 4b).
Considering q = 0 and q = 1, b-diversity was higher
among plots (within patches) than among patches (within
landscapes) and among landscapes (Fig. 4a). bplot was significantly higher in LDL than in HDL and IDL when considering
the orders 0 (GLM, v2 = 31.8, d.f. = 2, P < 0.001) and 1
(v2 = 10.5, d.f. = 2, P = 0.005) (Fig. 4c). For q = 2, bplot
tended to be higher in LDL (Fig. 4c), but we did not find significant differences (v2 = 3.9, d.f. = 2, P = 0.12). For all
Fig. 4. Plant b-diversity in three landscapes
with different deforestation levels in Los
Tuxtlas, Mexico. To assess whether the
magnitude in species turnover differed across
spatial scales, in panel (a), we compared the
relative compositional dissimilarity among
landscapes (i.e. for the entire region), among
patches and plots using the transformation of
beta proposed by Jost (2007) (see Materials
and methods). b-diversity among patches and
plots are also indicated in panels (b) and (c),
respectively. In the latter case, we had 15 data
points (i.e. 15 patches) per landscape, and
hence, means ( SE) per landscape are
indicated. In all cases, we evaluated three
orders of q (0, 1 and 2). Landscapes: LDL,
lowest deforestation level; IDL, intermediate
deforestation level; HDL, highest deforestation
level.
landscapes, bplot was between 1.7 and 1.8 times higher when
considering rare species than when considering abundant
species (Fig. 4c).
EFFECTS OF PATCH LOCATION, ISOLATION, SIZE,
ALTITUDE AND SLOPE ON b-DIVERSITY AMONG
PATCHES
In general, the results indicated that b-diversity is more
related to the geographical distance between sampled patches
than to the differences between patches in patch size, altitude
and slope (Table 1). Higher b-diversity values were found
among patches more isolated from each other, particularly
in LDL and HDL landscapes (Table 1). b-diversity among
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1449–1458
Plant b-diversity in fragmented rain forests 1455
patches also increased with increasing interpatch differences
in altitude, but only in LDL did we find significant correlations after applying the Bonferroni correction (P < 0.001).
Differences in patch size and slope between patches also
increased bpatch, but the correlation coefficients were very
small (generally R < 0.2) and not significant for any order q
(Table 1).
Discussion
Our multiscale evaluation of b-diversity in severely deforested
tropical landscapes indicates that plant species turnover is
associated with the landscape configuration and varies across
spatial scales. Three general patterns emerge: (i) the landscapes
with higher deforestation levels (IDL and HDL) showed a
lower b-diversity (floristic homogenization) within patches; (ii)
the landscape with highest deforestation (with only 4% of
remaining forest cover) showed the highest b-diversity (floristic differentiation) among patches, particularly considering
abundant species; and (iii) b-diversity among patches was
more related to the geographical distance between sampled
patches than to variations in patch size, altitude and/or terrain’s
slope. These novel findings show that plant communities in
fragmented landscapes can undergo either declines (floristic
homogenization) or increases (floristic differentiation) in bdiversity depending on the landscape spatial configuration and
on the spatial scale plant communities are examined.
CAUSES OF FLORISTIC HOMOGENIZATION AND
DIFFERENTIATION
The few existing studies showing biotic homogenization in
human-modified tropical landscapes suggest that this process
results from ecological filters related to intensive land-use
changes (Karp et al. 2012), and to the alteration of seed dispersal, seedling recruitment, growth and survival in strongly
deforested landscapes (Michalski, Nishi & Peres 2007; L^
obo
et al. 2011; Tabarelli, Peres & Melo 2012). Our results sup-
port this idea and demonstrate that floristic homogenization
within forest patches emerges in landscapes with < 11% of
remaining forest cover (i.e. IDL and HDL).
The decline in b-diversity within patches can be related to
the loss of rare (i.e. with smaller populations) and shade-tolerant species, as well as to the recruitment and dominance of
disturbance-adapted species in landscapes with higher deforestation. Rare species were less frequent in IDL and HDL than
in LDL. For example, the number of species with < 5 individuals was 19% and 60% lower in HDL and IDL, respectively, compared to LDL (Fig. 3). Rare species are more
vulnerable to local disappearance in forest patches (Fischer &
Lindenmayer 2007). Furthermore, as approximately 80% of
rare species in LDL were shade-tolerant species and these are
the most extinction prone in small forest patches and near forest edges (Laurance et al. 1998, 2002; Santos et al. 2008),
rare and shade-tolerant species declined in IDL and HDL
(Fig. S3). This pattern is probably related to higher edge
effects in IDL and HDL, as these landscapes presented smaller patches (Table S1), without relatively unaltered interior
areas (see Arroyo-Rodríguez et al. 2009). Under this situation, short- and long-lived light-demanding plant species can
be favoured, recruiting in large numbers along forest edges
and small patches (Laurance et al. 2006; Santos et al. 2012);
an idea consistent with our previous analyses, in which we
recorded higher numbers (and percentage) of pioneer and
light-demanding plant species in IDL and HDL than in LDL
(Arroyo-Rodríguez et al. 2009).
Floristic homogenization at the plot and patch scales was
particularly evident in the landscape with intermediate deforestation, probably because the extent of landscape connectivity (i.e. distance to nearest forest patch) in this landscape was
similar to that in the landscape with lowest deforestation
(Table S1). This landscape spatial pattern can favour interpatch seed dispersal processes, facilitating the recruitment of
disturbance-adapted species in different plots and patches.
Supporting this idea, the number of individuals was more
evenly distributed among dominant species, which were
Table 1. Correlations between tree b-diversity among patches and forest patch location (geographical distance between sampled patches) and
interpatch differences (D) in patch size, altitude and slope in three landscapes with different deforestation levels studied at Los Tuxtlas, Mexico†
Landscapes‡
b-diversity component§
Patch location
D Patch size
D Altitude
D Slope
LDL
0
bpatch
bpatch
2
bpatch
0
bpatch
1
bpatch
2
bpatch
0
bpatch
1
bpatch
2
bpatch
0.188
0.337**
0.393**
0.335**
0.181
0.060
0.522**
0.448**
0.323*
0.054
0.104
0.112
0.197
0.216*
0.182
0.240*
0.175
0.065
0.412**
0.346**
0.241*
0.12
0.042
0.024
0.209*
0.179
0.122
0.158
0.197
0.295*
0.070
0.003
0.106
0.110
0.111
0.039
1
IDL
HDL
†We indicate the correlation coefficients calculated with Mantel tests. Significant coefficients are indicated with asterisks (*P < 0.05;
** P < 0.001, after applying a stringent Bonferroni correction to reduce the likelihood of type I statistical errors).
‡Landscapes: LDL, lowest deforestation level; IDL, intermediate deforestation level; HDL, highest deforestation level.
§bpatch represents the species turnover between patches. We considered three orders q (0, 1 and 2), which determine the sensitivity of each
b-diversity component to the relative abundances.
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1449–1458
1456 V. Arroyo-Rodríguez et al.
mainly pioneer and light-demanding species (Fig. 3). Also,
species in IDL were more widely distributed at the patch
(Fig. S2) and landscape scales (Fig. 2). For example, in this
landscape, 33 species (21%) were distributed in at least 11
patches, and 6 species (4%) were recorded in the 15 patches.
From these six species, four (Croton schiedeanus, Tapirira
mexicana, Rinorea guatemalensis and Casearia sylvestris)
were light-demanding species and two were shade-tolerant
(Pseudolmedia oxyphyllaria, Psychotria clivorum), suggesting
that biotic and abiotic conditions in this landscape promoted
the dispersal and establishment of these species.
In contrast, the landscape with highest deforestation level
showed the greatest isolation distances between patches
(Table S1). This spatial pattern can significantly limit the
interchange of seeds (and species) among patches (Hubbell
2001; Condit et al. 2002; Myers et al. 2013), increasing
b-diversity among patches. In fact, in this landscape, we
found the largest correlation coefficients between the distance
between sampled patches and bpatch for any order q (Table 1),
indicating that the increment of the geographical distance
among patches in landscapes with only 4% of remaining forest cover can result in a sharp increment of tree species turnover among patches. In addition, the Los Tuxtlas region has
experienced extensive defaunation (i.e. most mid- and largesized seed dispersers have disappeared; Dirzo & Miranda
1991), but it has been more severe in the landscape with
highest deforestation due to its lower amount of forest cover.
As an example, howler monkeys (Alouatta palliata) – an
important seed disperser (Arroyo-Rodríguez et al. 2013) – are
almost absent in HDL (with only 8.7% of patches occupied
by this species), whereas in IDL and LDL, monkeys occupy
between 21 and 27% of the patches (Arroyo-Rodríguez,
Mandujano & Benítez-Malvido 2008). Mid- and large-sized
herbivorous and/or seed predators are also virtually absent
(e.g. Cuniculus paca, Dasyprocta punctata, Odocoileus virginianus, Tayassu pecari) or extinct (e.g. Tapirus bairdii,
Mazama americana, Pecari tajacu) in the region, and as consequence, monospecific and extensive carpets of tree seedlings are commonly found below parental trees in the region
(Dirzo & Miranda 1991). Overall, these processes can reduce
the spatial distribution of plant species (e.g. 78% species in
this landscape were distributed in less than 5 patches, Fig.
S3), increasing floristic differentiation among patches.
CONSEQUENCES FOR RECOVERY DYNAMICS AND THE
MAINTENANCE OF SPECIES DIVERSITY
An increasing proportion of species in the tropics inhabits
severely deforested and hyperfragmented rain forest landscapes. As others (Tabarelli, Peres & Melo 2012; Melo et al.
2013), our study indicates that in these emerging human-modified landscapes, b-diversity within forest patches will decline,
leading to increased levels of floristic homogenization at the
patch scale. This pattern can greatly limit the total number of
species (c-diversity) that landscapes and regions can accumulate (Flohre et al. 2011). Correspondingly, c-diversity in IDL
and HDL was significantly lower than in LDL (Fig. S1), and
mean ( SD) a-diversity within patches was significantly
higher in LDL (61 9 species) and IDL (65 9 species)
than in HDL (49 14 species) (Arroyo-Rodríguez et al.
2009). Thus, because a-diversity within patches did not
decline in IDL, the lower c-diversity in this landscape compared to LDL was mainly due to the loss of b-diversity
among patches, whereas the lower c-diversity in HDL was
due to both reduced a- and reduced b-diversity.
Interestingly, our findings indicate that the increased
distances among patches in the landscape with highest deforestation could limit the interchange of species, promoting floristic differentiation among patches. The study patches are
undergoing continuous regeneration following post-fragmentation disturbances such as edge effects and extraction of forest
products (e.g. logging). Because such disturbances are known
to have important deleterious effects on plant communities
(Laurance et al. 2002; Brown & Gurevitch 2004; Tabarelli,
Lopes & Peres 2008), differences among patches in disturbance regimes can also increase plant b-diversity among
patches, particularly in landscapes dominated by edge-affected
patches (i.e. HDL). This suggests that, in accordance with
non-equilibrium dynamics (Hubbell 2001; Myers et al. 2013),
and the ‘landscape-divergence hypothesis’ (Laurance et al.
2007), patches located in landscapes with different forest
cover and different connectivity can experience contrasting
successional dynamics. Based on the biotic and abiotic characteristics of each forest patch, the landscape spatial context
surrounding each patch, the life-history traits of the species
found in each patch, the land-use history, along with many
stochastic events (Chazdon 2008; Tabarelli, Lopes & Peres
2008), patches located in severely deforested rain forests will
most probably exhibit increasing differentiation in ecosystem
functioning, as well as in species composition. As proposed
by Laurance et al. (2007), ‘differences (among patches) in
characteristics such as forest dynamics, carbon storage, functional guild composition and species invasions could gradually accumulate over time, leaving an increasingly pervasive
signature of divergence on community composition and functioning’. However, if there is a functional convergence among
early successional species or edge-colonizing species, the
functional divergence among patches will be probably lower
than the compositional differentiation. This topic constitutes a
very important avenue for future research.
Overall, these findings add further uncertainties to the
future of biodiversity in severely deforested tropical landscapes (Chazdon et al. 2009; Melo et al. 2013). Evidence
indicates that the dominance of some tree species within the
forest can impede/retard forest succession, whereas others can
act as facilitators (Mesquita et al. 2001). In fact, if seed dispersal processes are disrupted, and the biotic and abiotic conditions within the forest limit the recruitment of old-growth
forest species, forest patches can move towards an early successional system dominated by pioneer species (Tabarelli,
Lopes & Peres 2008). Therefore, the successional dynamics
within each patch, as well as its ability to maintain original
levels of diversity, will vary extensively among patches
depending on the landscape spatial pattern (Pardini et al.
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1449–1458
Plant b-diversity in fragmented rain forests 1457
2010). This phenomenon should be therefore considered in
future management and conservation plans, as it could imply,
for example, that ‘protected areas in different landscapes
could preserve biologically and functionally different components of ecosystems’ (Laurance et al. 2007). Nevertheless, it
is important to note that tree communities can be highly
dynamic, particularly in fragmented landscapes (Laurance
2002), and hence, further long-term studies are required to
assess the changes in tree communities over time and to thus
improve our understanding of the homogenization vs. differentiation processes in human-modified landscapes.
Acknowledgements
We thank B. G
omez-Chagala and R. Mateo-Gutierrez and family for their
invaluable help. M. Peredo-Nava (XAL herbarium) and G. Castillo-Campos
helped with taxonomy and the identification of specimens. This study was
funded by the Departamento de Biodiversidad y Ecología Animal (Instituto de
Ecología A.C.). The Centro de Investigaciones en Ecosistemas of the Universidad Nacional Aut
onoma de Mexico provided logistical support.
References
Arroyo-Rodríguez, V., Mandujano, S. & Benítez-Malvido, J. (2008) Landscape
attributes affecting patch occupancy by howler monkeys (Alouatta palliata
mexicana) in fragmented landscapes at Los Tuxtlas, Mexico. American Journal of Primatology, 70, 69–77.
Arroyo-Rodríguez, V., Pineda, E., Escobar, F. & Benítez-Malvido, J. (2009)
Value of small patches in the conservation of plant-species diversity in
highly fragmented rainforest. Conservation Biology, 23, 729–739.
Arroyo-Rodríguez, V., Cavender-Bares, J., Escobar, F., Melo, F.P.L., Tabarelli,
M. & Santos, B.A. (2012) Maintenance of tree phylogenetic diversity in a
highly fragmented rainforest. Journal of Ecology, 100, 702–711.
Arroyo-Rodríguez, V., Andresen, E., Bravo, S.P. & Stevenson, P.R. (2013)
Howler Monkeys: Examining the Biology, Adaptive Radiation, and Behavioral Ecology of the Most Widely Distributed Genus of Neotropical Primate
(eds M. Kowalewski, P.A. Garber, L. Cortes-Ortiz, B. Urbani & D. Youlatos), Developments in Primatology: Progress and Prospects book series,
Springer, New York (in press).
Brown, K.A. & Gurevitch, J. (2004) Long-term impacts of logging on forest
diversity in Madagascar. Proceedings of the National Academy of Sciences
USA, 101, 6045–6049.
Chao, A., Chiu, C.H. & Hsieh, T. (2012) Proposing a resolution to debates on
diversity partitioning. Ecology, 93, 2037–2051.
Chazdon, R.L. (2008) Chance and determinism in tropical forest succession.
Tropical Forest Community Ecology (eds W.P. Carson & S.A. Schnitzer),
pp. 384–408. Blackwell, Malden.
Chazdon, R.L., Harvey, C.A., Komar, O., Griffith, D.M., Ferguson, B.G., Martínez-Ramos, M., et al. (2009) Beyond reserves: a research agenda for conserving biodiversity in human-modified tropical landscapes. Biotropica, 41,
142–153.
Condit, R., Pitman, N., Leigh, E.G., Chave, J., Terborgh, J., Foster, R.B., et al.
(2002) Beta-diversity in tropical forest trees. Science, 295, 666–669.
Crawley, M.J. (2002) Statistical Computing. An Introduction to Data Analysis
Using S-Plus. John Wiley & Sons, London.
Dirzo, R. & Miranda, A. (1991) Altered patterns of herbivory and diversity
in the forest understory: a case study of the possible consequences of contemporary defaunation. Plant–Animal Interactions: Evolutionary Ecology in
Tropical and Temperate Regions (eds P.W. Price, T.M. Lewinsohn, G.W.
Fernandes & W.W. Benson), pp. 273–287. John Wiley & Sons, New
York.
FAO (2011) State of the World’s Forests 2011. Food and Agriculture Organization of the United Nations, Roma.
Fischer, J. & Lindenmayer, D.B. (2007) Landscape modification and habitat
fragmentation: a synthesis. Global Ecology and Biogeography, 16, 265–
280.
Flohre, A., Fischer, C., Aavik, T., Bengtsson, J., Berendse, F., Bommarco, R.,
et al. (2011) Agricultural intensification and biodiversity partitioning in European landscapes comparing plants, carabids, and birds. Ecological Applications, 21, 1772–1781.
Galanes, I.T. & Thomlinson, J.R. (2009) Relationships between spatial configuration of tropical forest patches and woody plant diversity in northeastern
Puerto Rico. Plant Ecology, 201, 101–113.
García-Aguirre, M.C., Alvarez,
R., Dirzo, R., Ortiz, M.A. & Eng, M.M. (2010)
Delineation of biogeomorphic land units across a tropical natural and humanized terrain in Los Tuxtlas, Veracruz, Mexico. Geomorphology, 121, 245–256.
Gardner, T.A., Barlow, J., Chazdon, R., Ewers, R.M., Harvey, C.A., Peres,
C.A., et al. (2009) Prospects for tropical forest biodiversity in a human-modified world. Ecology Letters, 12, 561–582.
Gentry, A.H. (1982) Patterns of Neotropical plant species diversity. Evolutionary Biology, 15, 1–85.
Gibson, L., Lee, T.M., Koh, L.P., Brook, B.W., Gardner, T.A., Barlow, J.,
et al. (2011) Primary forests are irreplaceable for sustaining tropical biodiversity. Nature, 478, 378–381.
Guevara, S., Laborde, J. & Sanchez-Ríos, G. (2004) Los Tuxtlas. El paisaje de
la Sierra. Instituto de Ecología, A. C. & European Union, Xalapa.
Harrison, S. & Cornell, H. (2008) Toward a better understanding of the regional causes of local community richness. Ecology Letters, 11, 969–979.
Hubbell, S.P. (2001) The Unified Neutral Theory of Biodiversity and Biogeography. Princeton University Press, Princeton.
Jost, L. (2006) Entropy and diversity. Oikos, 113, 363–375.
Jost, L. (2007) Partitioning diversity into independent alpha and beta components. Ecology, 88, 2427–2439.
Jost, L. (2010) The relation between evenness and diversity. Diversity, 2, 207–
232.
Jost, L., Chao, A. & Chazdon, R. (2011) Compositional similarity and b (beta)
diversity. Biological Diversity: Frontiers in Measurement and Assessment
(eds A. Magurran & B. McGill), pp. 66–84. Oxford University Press, New
York.
Karp, D.S., Rominger, A.J., Zook, J., Ranganathan, J., Ehrlich, P.R. & Daily,
G.C. (2012) Intensive agriculture erodes b-diversity at large scales. Ecology
Letters, 15, 963–970.
Laurance, W.F. (2002) Hyperdynamism in fragmented habitats. Journal of Vegetation Science, 13, 595–602.
Laurance, W.F., Ferreira, L.V., Rankin-de Merona, J.M. & Laurance, S.G.
(1998) Rain forest fragmentation and the dynamics of Amazonian tree communities. Ecology, 79, 2032–2040.
Laurance, W.F., Lovejoy, T.E., Vasconcelos, H.L., Bruna, E.M., Didham, R.K.,
Stouffer, P.C., et al. (2002) Ecosystem decay of Amazonian forest fragments:
a 22-year investigation. Conservation Biology, 16, 605–618.
Laurance, W.F., Nascimento, H.E.M., Laurance, S.G., Andrade, A., Fearnside,
P.M., Ribeiro, J.E.L., et al. (2006) Rain forest fragmentation and the proliferation of successional trees. Ecology, 87, 469–482.
Laurance, W.F., Nascimento, H.E.M., Laurance, S.G., Andrade, A., Ewers,
R.M., Harms, K.E., et al. (2007) Habitat fragmentation, variable edge effects,
and the landscape-divergence hypothesis. PLoS ONE, 2, e1017.
Lawton, J.H. (1999) Are there general laws in ecology? Oikos, 84, 177–192.
L^obo, D., Leao, T., Melo, F.P.L., Santos, A.M.M. & Tabarelli, M. (2011) Forest fragmentation drives Atlantic forest of northeastern Brazil to biotic
homogenization. Diversity and Distributions, 17, 287–296.
Macía, M.J., Ruokolainen, K., Tuomisto, H., Quisbert, J. & Cala, V. (2007)
Congruence between floristic patterns of trees and lianas in a southwest
Amazonian rain forest. Ecography, 30, 561–577.
Melo, F.P.L., Dirzo, R. & Tabarelli, M. (2006) Biased seed rain in forest edges:
evidences from the Brazilian Atlantic forest. Biological Conservation, 132,
50–60.
Melo, F.P.L., Arroyo-Rodríguez, V., Fahrig, L., Martínez-Ramos, M. & Tabarelli, M. (2013) On the hope for biodiversity-friendly tropical landscapes.
Trends in Ecology and Evolution, 28, 461–468.
Mesquita, R.C.G., Ickes, K., Ganade, G. & Williamson, G.B. (2001) Alternative
successional pathways in the Amazon Basin. Journal of Ecology, 89, 528–537.
Michalski, F., Nishi, I. & Peres, C.A. (2007) Disturbance-mediated drift in tree
functional groups in Amazonian forest fragments. Biotropica, 39, 691–701.
Myers, J.A., Chase, J.M., Jimenez, I., Jørgensen, P.M., Araujo-Murakami, A.,
Paniagua-Zambrana, N., et al. (2013) Beta-diversity in temperate and tropical
forests reflects dissimilar mechanisms of community assembly. Ecology
Letters, 16, 151–157.
Pardini, R., Bueno, A.A., Gardner, T.A., Prado, P.I. & Metzger, J.P. (2010)
Beyond the fragmentation threshold hypothesis: regime shifts in biodiversity
across fragmented landscapes. PLoS ONE, 5, e13666.
Perfecto, I. & Vandermeer, J. (2008) Biodiversity conservation in tropical agroecosystems – a new conservation paradigm. Annals of the New York Academy of Sciences, 1134, 173–200.
Ricklefs, R.E. (1987) Community diversity: relative roles of local and regional
processes. Science, 235, 167–171.
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1449–1458
1458 V. Arroyo-Rodríguez et al.
Santos, B.A., Peres, C.A., Oliveira, M.A., Grillo, A., Alves-Costa, C.P. & Tabarelli, M. (2008) Drastic erosion in functional attributes of tree assemblages
in Atlantic forest fragments of northeastern Brazil. Biological Conservation,
141, 249–260.
Santos, G.G.A., Santos, B.A., Nascimento, H.E.M. & Tabarelli, M. (2012)
Contrasting demographic structure of short- and long-lived pioneer tree species on Amazonian forest edges. Biotropica, 44, 771–778.
Silva, J.M.C. & Tabarelli, M. (2000) Tree species impoverishment and the
future flora of the Atlantic forest of northeast Brazil. Nature, 404, 72–74.
Tabarelli, M., Lopes, A.V. & Peres, C.A. (2008) Edge-effects drive tropical forest fragments towards an early-successional system. Biotropica, 40, 657–661.
Tabarelli, M., Peres, C.A. & Melo, F.P.L. (2012) The ‘few winners and many
losers’ paradigm revisited: emerging prospects for tropical forest biodiversity.
Biological Conservation, 155, 136–140.
Tuomisto, H. (2010) A diversity of beta diversities: straightening up a concept
gone awry. Part 1. Defining beta diversity as a function of alpha and gamma
diversity. Ecography, 33, 2–22.
Tuomisto, H., Ruokolainen, K. & Yli-Halla, M. (2003) Dispersal, environment,
and floristic variation of western Amazonian forests. Science, 299, 241–244.
Vormisto, J., Phillips, O.L., Ruokolainen, K., Tuomisto, H. & Vasquez, R.
(2000) A comparison of fine-scale distribution patterns of four plant groups
in an Amazonian rainforest. Ecography, 23, 349–359.
Wiens, J.A. (1989) Spatial scaling in ecology. Functional Ecology, 3, 385–397.
Received 4 March 2013; accepted 6 August 2013
Handling Editor: Thomas Kitzberger
Supporting Information
Additional Supporting Information may be found in the online version of this article:
Table S1. Spatial characteristics of three landscapes with different
deforestation levels studied at Los Tuxtlas, Mexico.
Figure S1. Differences among landscapes in the number of plant species considering individual-based species accumulation curves.
Figure S2. Distribution of plant species according to the number of
patches occupied in each landscape.
Figure S3. Distribution of plant species according to their total abundances and ecological group.
© 2013 The Authors. Journal of Ecology © 2013 British Ecological Society, Journal of Ecology, 101, 1449–1458