Composition and structure of a diverse tree ...

Plant Ecol (2014) 215:1469–1481 DOI 10.1007/s11258-014-0407-y

Composition and structure of a diverse tree community at the edges of a Brazilian Amazon rainforest island surrounded by marshes and mangroves Luciana Oliveira dos Santos • Leiliane Oliveira dos Santos Moirah Paula Machado de Menezes • Colin Robert Beasley • Ulf Mehlig

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Received: 27 September 2013 / Accepted: 21 August 2014 / Published online: 6 September 2014 Ó Springer Science+Business Media Dordrecht 2014

Abstract Man-made forest edges affect tree community composition and structure, but there is little information on the effects of natural edges. To detect tree community changes related to natural edges, we investigated a terrestrial forest island, bordered by grassland and mangrove landscapes, on the Brazilian Amazon coast. Forest structure and tree species composition were examined in 10-m-wide and 100-m-long plots from the forest edge to the interior in relation to both landscapes (8 grassland and 4 mangrove plots). Elevation and soil quality did not reveal strong spatial variability; tidal inundation established a boundary for forest expansion without influencing the forest interior. The tree community consisted of 82 species. No distinct changes in vegetation structure and composition with distance from the border were detected. Ordination procedures

gave weak indications for shifts in community composition with distance from the edge and in respect to edge type. Single species occurred more frequently either close to or more distant from the grasslandforest interface. No such tendencies were detected at the mangrove-terrestrial forest interface, probably because the lack of structural differences between terrestrial forest and mangrove canopies did not permit the establishment of edge-related microclimatic gradients. Due to the isolation of the forest island and the harsh coastal environment, the tree community was dominated by generalist species which are well adapted to the conditions at the grassland–forest edge. Furthermore, the patchy distribution of frequent species and the high number of rare species made it difficult to detect spatial patterns related to the forest edge. Keywords Edge effect  Core area  Spatial smoothing  Split-window analysis  Para´  Braganc¸a

Communicated by K. Harper.

Electronic supplementary material The online version of this article (doi:10.1007/s11258-014-0407-y) contains supplementary material, which is available to authorized users.

Introduction

L. O. dos Santos  L. O. dos Santos  M. P. M. de Menezes  C. R. Beasley  U. Mehlig (&) Laborato´rio de Biologia Vegetal, Instituto de Estudos Costeiros (IECOS), Federal University of Para´ (UFPA), Campus Braganc¸a. Alameda Leandro Ribeiro s/n, Braganc¸a, PA 68600-000, Brazil e-mail: [email protected]

Forest edges have been the focus of scientific interest for a long time (Harper et al. 2005). Besides the shifts in species composition and richness generally expected for ecotones (Lloyd et al. 2000; Ewers and Didham 2006), the effects of edge creation by fragmentation of large forest areas have also been

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under scrutiny in species-rich tropical regions (e.g. Laurance et al. 2002, 2011). The distance up to which an edge influences the forest’s flora and fauna, and the identification of an unaffected ‘‘core area’’ (Laurance and Yensen 1991) are of special importance for nature conservation, as are the possibly indirect and nonlinear responses of the forest ecosystem to edge creation within a given landscape matrix (e.g. Murcia 1995; Laurance 2008; Dodonov et al. 2013). The influence of edge creation on the vegetation has been extensively studied in the central Amazon, focusing on the impact of forest fragmentation due to clear-cutting (for review, see Laurance et al. 2002, 2011). Within this context, microclimatic gradients have been detected between the forest border and interior (Kapos 1989; Malcom 1994; Camargo and Kapos 1995; Didham and Lawton 1999), but the extent and quality of edge-induced changes differed substantially. Generally, edges were exposed to a more variable temperature regime (e.g. Didham and Lawton 1999), whereas parameters related to humidity showed a less consistent pattern (Camargo and Kapos 1995; Didham and Lawton 1999). When margins were ‘‘closed’’ by secondary vegetation, microclimatic edge effects were partly mitigated (Camargo and Kapos 1995), but different successional pathways induced different edge effect scenarios (Didham and Lawton 1999). The biological consequences of edge creation in the Amazon include tree mortality (Ferreira and Laurance 1997; Mesquita et al. 1999; Laurance et al. 1998, 2000; Laurance and Curran 2008) and changes in vegetation cover and composition (Malcom 1994; Camargo and Kapos 1995; Didham and Lawton 1999). A dramatic and long-lasting increase in forest dynamics with rapid substitution of many species of the original forest community by successional species near forest fragment edges (Laurance et al. 2006a, b; Santos et al. 2012) was one of the most striking effects of fragmentation. In contrast to man-made edges, the vegetation of natural forest borders has received less attention (e.g. Lloyd et al. 2000; Harper and Macdonald 2001; Erd}os et al. 2013). For tropical regions, available information is incomplete (e.g. Ramos et al. 2008, 2010). Recently, Hennenberg et al. (2008) showed that microclimatic gradients exist across the border of natural forest islands in a western African savanna. Within the same area, Hennenberg et al. (2005a) detected vegetation changes across the grassland-forest ecotone and used

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these to define the forest core area. To the best of our knowledge, natural forest borders in the Amazon have not been investigated yet. We present here a study of a forest island in a rather peculiar geographical setting. The coast of the Amazon region features one of the world’s largest mangrove belts (Spalding et al. 2010) with many bays and estuaries, within which the formation of barrier islands is common (Souza-Filho et al. 2009). The nonmangrove, scrubby or herbaceous vegetation of the outer beaches and dune landscapes of these islands has received the attention of local botanists (see Silva et al. 2010, for review) but the existence of islands covered by terrestrial evergreen broadleaf forest, completely separated from the mainland by mangroves and marshes, has been neglected (but see Behling et al. 2001; Abreu et al. 2006). In comparison with the species-poor mangroves of the Atlantic coast and the beach vegetation, these forests have a much higher plant diversity (Abreu et al. 2006, and own, unpublished data). Our forest island’s overall limits are associated with the influence of the surrounding marine-tidal environment. However, the forest borders are adjacent to two different kinds of landscape matrix, tall mangrove forest on one side and a grassland-marsh environment characterised by low vegetation on the other. Therefore, our study area provides an interesting test case for examining possible edge- and matrix-related differences in species composition and forest structure. Moreover, a thorough understanding of the mechanisms governing composition and structure of plant communities at the interface between land and sea becomes more and more important in the context of scenarios of sea level rise (e. g, Salinas et al. 1986; Shirley and Battaglia 2006; Lara et al. 2010). Our study focuses on the following questions: (1) are tree community changes and changes in forest structure detectable along transects from the border to the forest interior? and (2) are there landscape matrix-related differences with respect to such changes?

Material and methods Study area The study area is located on the Ajuruteua Peninsula, Braganc¸a district, Para´, Brazilian Amazon coast

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A

B Fig. 1 Position of the study area in Brazil (small map) and on the coast of Para´ (A based on Wessel and Smith (1996)) and location of transects within the study area (B). Wells were used to monitor groundwater table and salinity (L. O. Santos, unpublished data)

(Fig. 1) and belongs to the federally protected Reserva Extrativista Caete´-Taperac¸u. Ajuruteua Peninsula vegetation is dominated by mangrove forests on extensive mud flats surrounding inactive sandy barrier islands (Cohen et al. 2005; Souza-Filho et al. 2009; Mehlig et al. 2010). The peninsula encompasses the Salinas do Roque marshland-mangrove complex with patches of evergreen terrestrial forest on scattered paleodune ridges. This study focuses on the largest of these forest patches (08550 25.400 S, 468400 11.900 W), covering an area of approximately 33 ha. Before construction of a road in the early 1980s, the Salinas do Roque area was completely isolated by the surrounding mangroves; the distance to non-mangrove terrestrial vegetation on the mainland was at least 10 km. A preliminary description of this particular forest patch was given by Abreu et al. (2006). The patch abuts well-developed mangrove forest along its western border. Elsewhere, it is surrounded by grassland and open marshes. The mangroves bordering

these marshes in an easterly direction exhibit bushy growth, with heights\6 m. Areas in the vicinity of the forest patch have been used for cattle farming and small-scale family horticulture some decades ago. There is no detailed information about the prior usage of the forest itself. According to locals, the forest has never been clear-felled in their memory; the current limits of the forest are almost identical to those seen on the oldest available satellite image (Corona KH-4, late 1960s) and subsequent Landsat imagery (1985 onwards). The presence of small patches with anthropogenic Amazonian dark earth (e.g. Ka¨mpf et al. 2003) containing pre-Columbian ceramic fragments documents the presence of man on this paleodune for [2,000 years (Silveira et al. 2011, and U. Salzmann and U. Mehlig, unpublished data). The regional climate is warm and humid, with peak rainfall between January and May and a dry season between September and November. The long-term mean air temperature at Tracuateua, 30 km SW of the

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study site, is 26.0 8C; the average total annual precipitation is about 2,470 mm (INMET 1992, 2014). Topography, tidal inundation and soil characteristics Field work was conducted between April 2008 and October 2009. Starting at 12 points distributed along the margin of the forest patch, we established 100-mlong transects. Each transect was perpendicular to a linear stretch of the forest border (Fig. 1). Distances between transect starting points varied from 200 to 300 m (\200 m between transects 9 and 10). A segment along the eastern border of the forest was excluded because of damage caused by fire spreading from the neighbouring marshland. The landscape matrix surrounding the forest patch next to starting points of transects 1–8 consisted of low vegetation. At the starting points of the remaining transects, the patch bordered mangrove forest equalling the terrestrial forest in height (Fig. 1). Along each transect, relative topographic elevation was recorded by theodolite. To check whether transects were within reach of tidal inundation, small plastic vessels were fixed at 5-cm intervals along PVC tubes installed upright at the starting point of each transect before equinox high tides (greatest tidal amplitude of the year) in March and September 2009. During subsequent ebb tides, inundation levels were determined by reading the height of the top-most water-filled vessel. Beginning at a distance of 5 m from the transect starting point, the soil was probed to a depth of 1 m with a Pu¨rckhauer auger at 10 m intervals. Soil layers were immediately characterised by visual and tactile examination. Soil layer colour was recorded with the help of colour charts (Munsell Book of Soil Color Charts, 1994 revised edition, Kollmorgen Instruments, Macbeth); the soil was slightly moist throughout soil sampling. No samples were taken along transects 9 and 10 to avoid disturbance of an archaeological Amazonian dark earth site (Silveira et al. 2011) crossed by these transects. Floristic inventory and forest structure Twelve 100-m-long and 10-m-wide plots were established, each centred lengthwise on a transect line. All treelets and trees with stem diameter at breast height (dbh; measured at 1.3 m) C2.5 cm within the plots

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were surveyed. Species, tree height estimates and dbh measurements were recorded. Multi-stemmed trees with ramifications below 1.3 m were considered as single individuals for tree density calculation. Basal area coverage was calculated summing up the crosssectional areas of all stems equal to or greater than the established minimum diameter. The location of each plant within the respective plot was determined by recording its position along the transect and the distance to the transect centre line by means of measurement tapes and a laser distance metre. Positions were later transformed to geographic coordinates using start/endpoint coordinates of the transect from handheld GPS measurements and the resulting bearing. Calculation of forest structure parameters (frequency: percentage of 10 9 10 m sub-plots where species occurred; relative density: percentage of total number of trees contributed by the respective species; relative dominance: percentage of total basal area contributed by the respective species; relative frequency: percentage of the sum of all the frequencies contributed by the respective species; importance value: sum of relative density, relative dominance and relative frequency) followed Mu¨ller-Dombois and Ellenberg (1974). Vouchers of all tree species found in the study area were deposited in the herbarium of the Institute for Coastal Research (IECOS), Federal University of Para´, Campus Braganc¸a (herbarium code: HBRA; Thiers continuously updated). Citation of plant genera and families follows Stevens (2001) and APG III (2009). Data analysis The data were analysed with GNU R 2.13 and 3.02 (R Core Team 2011, 2013). Point patterns obtained from tree positions were analysed by spatial smoothing of tree density and covariates with corresponding standard error estimates (R-package spatstat; Baddeley and Turner 2005, 2006; Baddeley et al. 2012). To analyse relationships between the tree community and distance from the margin and border type, we conducted an ordination of tree basal area coverage in the 10 9 10 m sub-plots by non-metric multidimensional scaling (NMDS) with a distance matrix based on the quantitative Jaccard (Ruzˇicˇka) index. We then fitted the environmental factors distance from margin and border type (grassland/mangrove) on the

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ordination (envfit in R-package vegan, Oksanen et al. 2013). We further used constrained correspondence analysis (CCA; ter Braak 1986; Oksanen et al. 2013) on the basal area coverage data of the sub-plots with subsequent permutation tests of significance to analyse variation due to the factors distance and border type, and their interaction. Changes in species richness and diversity along the transects were examined separately for transect plots starting at the grassland- and at the mangrove-forest interface. Sliding windows of 10 9 10 m were shifted along the transects of the respective border type in steps of 1 m. At each step, the average richness at this distance from the margin, the corresponding standard deviation and diversity indices (Fisher’s a, Shannon) were calculated. Calculations were done in R, with diversity indices from the package vegan (Oksanen et al. 2013). Transect-wise split-window analysis for detection of discontinuities in tree community gradients (Cornelius and Reynolds 1991; Hennenberg et al. 2005a) was performed by means of an ad-hoc Python script. Split-window analysis moves an observation window step-wise along a gradient and compares pairs of halfwindows by a distance measure; here, euclidean distance based on tree abundance was used. Mean and standard deviation from 1,000 random permutations of half-window positions at each step were used to obtain Z-scores and corresponding confidence envelopes (Cornelius and Reynolds 1991); significant peaks of Z-scores along the transect indicate transition zones in the tree community.

Results Topography, tidal inundation and soil characteristics

1473 Table 1 Frequency (F), relative density (rD), relative dominance (rDom), relative frequency (rF) and importance value (IV) of the 15 species with the highest IV in the forest structure inventory of trees with diameter at breast height C2.5 cm in the studied forest patch Species

F (%)

rD (%)

rDom (%)

rF (%)

IV

Protium heptaphyllum

60

12.2

6.0

8.7

26.9

Eschweilera ovata

56

8.6

10.1

8.2

26.8

Aniba citrifolia

38

4.7

10.1

5.5

20.4

Simarouba amara

33

5.1

7.3

4.8

17.2

Garcinia madruno

42

7.9

2.9

6.1

16.9

Attalea maripa

27

3.4

6.9

3.9

14.2

Himatanthus articulatus

20

5.0

4.3

2.9

12.2

Pouteria ramiflora

27

3.3

4.9

3.9

12.1

Astrocaryum vulgare

29

4.4

2.1

4.2

10.8

Attalea speciosa

16

1.7

6.7

2.3

10.8

Tapirira guianensis

19

2.5

2.3

2.8

7.6

Byrsonima spicata

15

1.7

3.4

2.2

7.3

Licania sp. 1 Vitex orinocensis

18 5

2.3 0.7

1.9 3.5

2.6 0.7

6.9 5.0

Talisia cerasina

11

2.9

0.1

1.6

4.6

layer of organic material. The thickness of the dark brown to very dark grey upper layer in auger samples was variable (7–93 cm) within and among transects. This A horizon was frequently followed by a light grey layer, otherwise by a layer with brown or yellowish colours with iron concretions. At transect 11, a small patch of Amazonian dark earth was found about 65 m from the forest border. Transects 9 and 10 crossed several patches of Amazonian dark earth (Silveira et al. 2011). Floristic inventory and forest structure

Within-transect differences in topographic elevation were small (0.4–1.8 m). We observed a drop-off towards the mangrove within the first 5 to 10 m of transects 9–12; the transects then sloped gradually upwards towards the internal part of the forest. With the exception of a shallow ridge (0.7 m above average elevation) crossed by transect 12, the profiles were relatively even. Tidal inundation at equinox high tides did not reach the forest. Soil samples along all transects consisted of predominantly fine-grained sand, covered by a thin

We recorded 82 different species from 61 genera/41 families (61 species when only trees with dbh C 10 cm were considered). Protium heptaphyllum (Aubl.) Marchand (Burseraceae) and Eschweilera ovata (Cambess.) Miers (Lecythidaceae) had the highest importance values (Table 1). Both occurred in more than 50 % of the 10 9 10 m sub-plots and were the only species found along all transects. Only 19 of the 82 species had a relative density[1 %, 21 a relative dominance [1 %. Shannon’s diversity index

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Table 2 Minimum and maximum distance from margin of tree species, obtained from the individual trees’ coordinates in plots along transects starting at the grassland and mangrove edges. Only tree species with a relative density of at least 2.5 % in the respective area are shown. The inventory included trees with diameter at breast height C2.5 cm Species

Distance to grassland (m)

Distance to mangrove (m)

Min.

Min.

Max.

Tapirira guianensis

1.9

94.6

Himatanthus articulatus

5.0

85.8

Astrocaryum vulgare

1.3

94.9

Attalea maripa

47.4

Protium heptaphyllum Licania sp. 1

1.2 22.4

Garcinia madruno Aniba citrifolia Eschweilera ovata

1.3

99.1

Max.

6.4

99.3

99.5

0.8

98.2

99.4 98.0

3.5

99.1

9.5

99.1

7.4

96.5

11.1

91.8

7.2

95.7

3.2

97.9

Agonandra brasiliensis

21.5

59.6

Casearia guianensis

2.1

61.1

Talisia cerasina

1.1

76.8

Pouteria ramiflora

7.0

92.6

Simarouba amara

6.0

98.2

Cecropia sp.

4.3

99.3

6.3

98.8

for the whole area was 3.6 (3.4 for trees C10 cm dbh), and corresponding values for Fisher’s a were 20.9 and 18.0, respectively. A complete analysis of general floristic and structural properties of the forest patch will be communicated elsewhere. Most of the more common species occurred within a broad range of distances from the margin (Table 2). Species showing an affinity to the grassland–forest interface (Andira surinamensis (Bondt) Splitg. ex Pulle, Fabaceae-Faboideae; Coccoloba latifolia Lam., Polygonaceae; up to 12.4 m from the edge within plots) occurred only occasionally in the plots but were observed regularly (and exclusively) at the forest margin between transects. Species restricted to the surroundings of the mangrove–terrestrial forest interface were too rare to allow further conclusions about their spatial distribution. For species with high importance values, it was possible to estimate the probability of occurrence along the transects by smoothed histograms (Fig. 2, Online resources S1). Most of these species had higher numbers of individuals at some distance from the

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grassland–forest interface. Similar tendencies at the mangrove–terrestrial forest edge, if at all observable, were weak. Himatanthus articulatus (Vahl) Woodson (Apocynaceae) occurred more frequently near the border in transects at the grassland edge; however, a cluster of small individuals (dbh: \6.5 cm) of this species occurred within a single plot at greater distances from the margin (Fig. 2/Mangrove). Smoothed estimates of tree density (on average: 863 individualsha-1) as a function of distance from the margin did not show any consistent pattern, neither for transects originating at the forest–grassland interface nor for those in contact with the mangrove forest (Online resource S2). Regions with high stem density were patchy (Fig. 3). Likewise, smoothed estimates of breast height diameter did not vary consistently with distance from the margin but were inversely related with tree density (not shown). NMDS ordination (Online resource S3) did not show distinctly separated groups. Fitting of factors distance from margin and border type yielded significant results (p \ 0.001), but with low R2 (0.23 and 0.05 for distance and border type, respectively). The final stress value of 0.24 indicates that results of this ordination may not be sufficiently clear to support further interpretation. CCA permutation tests yielded significant outcomes for both distance (p \ 0.01) and border type (p \ 0.02) but not for the respective interaction (Online resource S3). However, constrained inertia was low (5 %) compared to unconstrained inertia. Both mean richness and diversity indices varied along the transects without showing a clear tendency related to distance from the margin or border type (Fig. 4). Mean richness per 10 9 10 m sub-plot increased slightly from the grassland-forest interface towards the interior but variation among transect plots, especially near the border, was considerable. Split-window analysis did not indicate any significant discontinuity in tree community composition along transects (Fig. 5).

Discussion The Salinas do Roque forest island featured a moderate number of tree species compared to what is expected for natural forest in the eastern Amazon. Most of the eastern Amazon plots within a 500 km

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A

B

Fig. 2 Smoothed empirical probability density functions for occurrence of four selected species along the transects starting at the grassland–forest interface (A) and at the mangrove–forest interface (B), respectively

radius of our study site cited in the compilation of ter Steege et al. (2000) scored much higher values for Fisher’s a (on average 49.6 ± 18.2, n = 8). However, terrestrial woody vegetation from the Amazon coastal zone typically has a lower number of species. Silva et al. (2010) report a total of 86 tree species for dune forests along the entire coast of the Brazilian state of Para´, and the inventories of individual sites reviewed by Silva et al. (2010) rarely exceeded 25 woody species (Silva et al., unpublished data). The relatively small size of the forest island and its isolated position within the marine environment of the northern Brazilian mangrove belt probably do not allow the establishment of a more diverse tree community, favouring species with well-developed long-distance dispersal strategies (cf. Matallana et al. 2005). Furthermore, the rainfall regime is seasonal, causing abrupt changes in the level of the groundwater table (L. O. Santos, unpublished data), which are typical of sandy soil environments in the Amazon (Anderson 1981). These environmental factors may prevent the establishment of less hardy species. Several genera and species of our inventory are common components of natural and secondary non-inundated forests in the region (e.g. Pires et al. 1953; Alvino et al. 2005; Arau´jo et al. 2005; Prata et al. 2010).

Abiotic conditions such as slope and the characteristics of the soil layers along our transects were relatively homogeneous, as can be expected from the geological setting (Cohen et al. 2005; Souza-Filho et al. 2009). We have no reason to suspect that these variables confound our analysis of changes in the tree community along the transects. Possible exceptions are the patches of archaeological Amazonian dark earth within three transect plots starting at the mangrove–terrestrial forest interface, which, in part, spatially overlapped with tree density maxima. However, the patches were too small to derive general conclusions about their influence on the local tree species composition. The mangrove–terrestrial forest interface was characterised by a small but abrupt transition in elevation, preventing tidal inundation from affecting the terrestrial vegetation. In terms of elevation, the transition between grassland and terrestrial forest was less pronounced. Parts of the herbaceous vegetation close to the grassland-forest interface suffered occasional tidal inundation; however, inundation did not reach the forest border itself. Salinity of the groundwater down to a depth of 2 m was zero or, exceptionally, extremely low in wells at the grassland-forest interface (L. O. Santos, unpublished data; Fig. 1). Groundwater

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Fig. 3 Smoothed density maps (r = 3) for the 12 transect plots (highlighted outline) within the forest island (light grey area). Density is indicated by a colour gradient; transects originating at the mangrove–terrestrial forest interface are marked with a star Fig. 4 Diversity indices (Shannon, Fisher’s a) and mean species richness along transects starting at the grassland-forest interface (left) and the mangroveforest interface (right), respectively. Values were calculated for moving windows of 10 9 10 m width. Grey underlay: span of standard deviation

salinity imposes stress on terrestrial trees and is therefore of interest especially in agriculture and environmental dryland management research (e.g.

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Cramer et al. 2002). However, data on woody coastal dune vegetation and groundwater salinity are scarce (e.g. Nishijima et al. 2003; Geßler et al. 2005; Armas

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Fig. 5 Z-scores of splitwindow discontinuity analysis of tree communities along transects. Window width: 10 m. Dotted lines indicate 95 % confidence envelopes based on 1,000 random permutations of half-window positions. Transects originating at the mangrove–terrestrial forest interface are marked with a star

and Pugnaire 2009). From the available information and our own observations, we conclude that a very low concentration of salt in the groundwater may be tolerated at least temporarily but that soil infiltrated by sea water creates a barrier to the establishment of terrestrial tree species in our study area. Similar to Nishijima et al. (2003), we assume that, at least near the margins of the forest patch, freshwater floats on top of a wedge of saline groundwater of tidal origin. Furthermore, a below-ground freshwater flow from the more elevated parts of the paleodune is probably sustaining a non-saline groundwater layer thick enough to support the terrestrial trees at the forest edge (cf. Nishijima et al. 2003). Where the slope between paleodune and surrounding terrain is less steep, the thickness of this freshwater layer probably declines gradually and grassland will develop where it is too shallow to permit the establishment of deeprooting trees. Edge influence on microclimatic parameters has been well documented for forest fragments in the neotropics (e.g. Kapos 1989; Malcom 1994; Camargo and Kapos 1995; Didham and Lawton 1999; Dodonov et al. 2013; Ewers and Banks-Leite 2013). Less information is available on the influence of the surrounding landscape matrix (Camargo and Kapos

1995; Didham and Lawton 1999; Mesquita et al. 1999; Nascimento et al. 2006). Recently, Dodonov et al. (2013) tested a matrix height-based measure of edge contrast as a predictor for microclimatic and forest structure variables, but only a small part of the variability found at fragment edges could be explained by edge contrast, possibly due to a rather narrow range of edge contrasts. We are not aware of any studies concentrating on natural forest borders in the Amazon, however, microclimatic edge effects at forest borders ‘‘closed’’ by secondary vegetation in the central Amazon (Camargo and Kapos 1995; Didham and Lawton 1999; Ferreira and Laurance 1997) and elsewhere (e.g. Denyer et al. 2006) are probably qualitatively comparable to those at natural forest edges. Hennenberg et al. (2008) detected seasonally varying depths of edge influence of different microclimatic parameters in West African natural forest islands within a savanna matrix. In the same area, it was possible to single out ‘‘forest interior’’ species with occurrence in the core area of forest islands (Hennenberg et al. 2005a) as well as species typically occurring at the forest-grassland interface (Hennenberg et al. 2005b). However, the respective study area was dominated by seasonally deciduous trees, and

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microclimatic edge effects depending on vegetation cover were probably more pronounced than can be expected for our study area. We assume that microclimatic edge effects in our study area affected principally the forest–grassland interface, as the matrix of tall, closed-canopy mangrove forest and the terrestrial forest at the mangrove– terrestrial forest interface were structurally quite homogeneous (cf. Abreu et al. 2006) and therefore not exposed to the rapid changes in temperature and wind conditions typically encountered at edges between low vegetation and forest (Laurance and Curran 2008). In parts of the study area, the stretch of grassland next to the forest border was relatively narrow, and mangrove shrubs were found at short distances opposite to the terrestrial forest (Fig. 1). Due to the low height of these mangroves, any mitigating influence on microclimatic edge effects would probably be weak. It was not possible to reveal any consistent trend in forest structure parameters with increasing distance from the edge, in spite of the application of smoothing techniques. The latter allowed us to identify trends more easily by suppressing ‘‘noisy’’ components in the data; in combination with permutation tests, trends and random variation could be distinguished. Likewise, we did not find any evidence for a transition zone from an edge-tolerant tree community to an edge-avoiding, ‘‘core area’’ species group by means of split-window analysis. A few species seemed to grow preferentially near the grassland–forest interface, but numbers found within plots were too low to provide statistically robust evidence for this observation at the species level. The occasional occurrence of species in the forest interior that otherwise showed a preference for growing close to the edge (e.g. Himatanthus articulatus, a species with pioneer characteristics; Thompson et al. 1998) probably indicates gap succession. Succession on randomly encountered previous tree fall gaps may contribute to the observed variation in density as well as in diversity and species richness, confounding possible gradient-related changes in forest structure and composition. The commonest species were able to occupy the whole range between margin and forest interior, with slightly elevated densities more distant from the grassland–forest interface. Tendencies to grow next to or more distant from the margin were, if at all, more distinct at the grassland–forest interface than at the mangrove–

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terrestrial forest boundary. This hints at an influence of microclimatic parameters which are more likely to be of importance at an edge bordering open areas with low vegetation (see above). The weak tendencies for community change detected by the applied ordination techniques indicate that distance from the margin and border type were both accompanied by shifts in species composition. However, only a small fraction of the variability among releve´s could be attributed to the assumed gradients. Species richness and diversity varied widely along the transects, and there was no clear tendency linking changes in these parameters to the border type or to the distance from the margin. Edge-related shifts in richness (Lloyd et al. 2000; Ewers and Didham 2006; Erd}os et al. 2013) due to mixing of plant communities would not, however, be expected in our case, as the tidal inundation inhibits mixing at the mangrove-forest interface and herbaceous plants from the grassland-forest interface were excluded from our analyses. In conclusion, shifts in community species composition with increasing distance from the border were obscured by high overall variability. Likewise, border type/matrix-related differences were rather blurred. How do we explain this weak response of the tree community to an edge-influenced environment? First of all, ambient conditions on a seasonally dry sandy paleodune ridge isolated by mangroves are probably harsh enough to exclude more sensitive species, even in the forest interior. Consequently, there is no distinct ‘‘core area’’ community. Conditions at the forest island’s ‘‘closed’’, natural border are probably not limiting for the surviving, generalist species. Their spatial distribution will depend rather, e.g., on dispersal strategies and only to a lesser degree on edgerelated microclimatic conditions; species with preference for growing near the edge are probably those that are the most light-demanding. Second, even a reasonably well-sampled, only moderately species-rich tropical tree community like the one examined here has a high percentage of rare species, for which spatial distribution patterns are hard to reveal. It should, however, not be concluded from our results that the failure to detect a specific core area tree community implies necessarily resilience against any form of disturbance affecting the forest edge. While many of the tree species present might be able to establish themselves near the border and adapt their growth

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form accordingly, it is likely that larger specimens of the same species growing inside the forest are susceptible, e.g. to wind damage (Ferreira and Laurance 1997; Laurance et al. 1998, 2000; Santos et al. 2012) after removal of the sheltering canopy at the edge. Therefore, the natural forest borders should be maintained intact, avoiding any severe disturbance, e.g. fire in the neighbouring grassland. Acknowledgments The first author received a scholarship provided by the Support Program for Restructuring and Expansion of the Federal Universities (Restruturac¸a˜o e Expansa˜o das Universidades Federais, REUNI) of the Brazilian Ministry of Education (MEC). The Chico Mendes Biodiversity Conservation Institute (ICMBIO) and the Council of the Caete´-Taperac¸u Extractive Reserve provided the necessary permits for our work. The authors wish to thank editor Karen Harper and two anonymous reviewers for their helpful comments. We are grateful to Vitor A. N. Braganc¸a for his support during field work.

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