Inthe last two decades many studies have documented the local extirpation of plant and animal species from fragmented tropical rainforests [20,24,37], particularly from forest edges [34,38]. This spatially nonrandom pattern of species impoverishment has pervasive effects on the subsequent community dynamics and ecosystem function [21,34], but its impact on tree evolutionary diversity has never been examined despite the implications to conservation [5,10]. One ofthe major conclusions of this study is that the local extirpation oftree species from forest edges results in a significant lossoftreephylogeneticdiversity. Such a loss is observed at the plot scale as a decrease by 11% inthephylogenetic distance between any two randomly chosen individuals and an increase by 17% inthe distance between a given individual and its closest non-conspecific relative, indicating that edge effects inthe study area are much more profound than previously envisioned and documented. Given that almost half ofthe remaining hyper- fragmentedBrazilianAtlanticforest is less than 100 m from the nearest edge , it is likely that theedge-relatedlossoftreephylogeneticdiversity is even more relevant at the regional scale. Theforest edges and forest fragments we studied have been embedded in a stable landscape for as long as 200 years. The fact that forest fragments showed intermediate MPD and MNTD between degraded forest edges and conserved old-growth interior areas provides new insights into the manifestation of late edge effects infragmented tropical rainforests; a phenomenon that is currently poorly understood owing to the scarcity of long-term data or studies in old fragmented landscapes. On one hand, our results reinforce the notion that edge effects are a continual phenomenon intheBrazilianAtlanticforest that drives small forest fragments toward early- to mid-successional systems . On the other hand, the intermediate condition faced by small forest fragments suggests that even two centuries of fragmentation may not be long enough to allow the full spectrum ofedge effects to be seen in their core areas, which already exhibit evidence of many types ofedge effects [22,23,25,29]. In fact, the role of greater time lags inthe manifestation of fragmentation effects on tree assemblages has received little attention inthe habitat fragmentation literature . This oversight arises not only from the misinterpretation of habitat fragmentation as a static phenomenon rather than a dynamic process , but also from not considering the exceptionally long lifespan of many old-growth tropical trees (several centuries in some cases ).
ABSTRACT – To mitigate the impact of greenhouse gases, the United Nations created a mechanism to advance forest carbon and biodiversity saving initiatives through Reducing Emissions from Deforestation and Forest Degradation (REDD+). However some methods to recognize the co-benefits among carbon and biodiversity have failed in highly fragmented landscapes. Therefore our study tested the potential for the existence of carbon-biodiversity co- benefits in a Tableland Atlantic Rain Forestin Brazil. Inside 240 10x10m plots , we measured three rsources of carbon present inforest fragments: trees, lianas and standing dead trees. We then related this carbon sources to species richness, community structure, endemic species richness and IUCN Red listed tree species. To evaluate these relationships we used generalized linear mixed models, selecting the best performing model on the basis of their corrected Akaike Information Criterion value, ideal for small sample sizes (AICc). We measured a total of 4,140 trees, 8,236 lianas and 277 standing dead tree individuals. We estimated that theforest fragments we sampled contain 424.39Mg ha -1 of carbon and 443 species of trees, of which 188 are AtlanticForest endemic species and 36 are considered as threatened species by the IUCN. Our results showed there is a significant spatial congruence between biodiversity and carbon stocks infragmented landscapes of tropical forest. This relationship, however, is strongerin larger fragments, where carbon stocks are significantly larger and the number of species with high conservation value is greater. In conclusion, the REDD+ co-benefits scheme could be use in a fragmented landscape, even one subjected to high fragmentation levels. This suggests that additional REDD+ funds could be used to enhance the carbon and biological value through the management offorest fragments
Unfortunately, large trees appear to be highly vulnerable to habitat fragmentation and the establishment of artiﬁcial forest edges. Long-term tree inventories inthe Amazon region have documented a chronic, increased mortality of large trees near forest edges surrounded by open habitats (Laurance et al., 2000; Nascimento and Laurance, 2004). Speciﬁcally, large trees may face a 40% increase in mortality near edges (particularly inthe aftermath of habitat fragmentation) and die three times faster than those inhabiting spots located over 300 m into theforest (Laurance et al., 2000) due to elevated rates of uprooting and breakage caused by increased wind turbulence near forest edges: the ‘‘wind damage hypothesis’’ (sensu D’Angelo et al., 2004). In addition to this direct, edge-induced mortality, surface forest ﬁres and logging may constitute aggravating sources of large tree mortality in human- impacted, fragmented landscapes (Barlow et al., 2003). However, increased mortality has been documented primarily in recently created fragmented landscapes, severely limiting accurate insights about the long-term persistence of large trees in consolidated, human-dominated landscapes. Assuming that theedge-induced mortality observed inthe Amazon region represents a more persistent fragmentation-related effect, one would expect a profound impoverishment or even the structural collapse ofthe large-tree stand in aging, hyper-fragmented landscapes, regardless of whether these landscapes have experienced expressive forest ﬁres and logging. This prediction has not been examined thoroughly despite its far-reaching consequences for biodiversity conservation in landscapes vastly dominated by edge-affected habitats.
Furthermore, we employed other four well-known phylobeta- diversity measures to compare theforest types within the Southern BrazilianAtlanticForest (see Table 1). COMDIST is a phylobe- tadiversity measure that computes the mean phylogenetic distance among species occurring in two different sites . For this reason, this phylobetadiversity measure captures variation associated with the more basal nodes linking species . Computing COMDIST values without considering the variation in species abundances is equivalent to compute thephylogenetic distinctness (Rao’s D) proposed by Hardy & Senterre . Thus, we opted for using only the former in this study. On the other hand, by standardizing Rao’s D values by the mean within-site phylogeneticdiversity it is possible to obtain another phylobetadiversity measure (Rao’s H, ), which captures phylobetadiversity patterns related to more terminal nodes inthetree . COMDISTNT  measures the mean phylogenetic distance between every species in a plot and the nearest phylogenetic neighbor in another site (Table 1). It is, therefore, a ‘‘terminal node’’ metric . The last phylobetadi- versity method used in this study was UniFrac , which measures, for each pair of sites, the fraction ofthe total branch length ofphylogenetictree that is exclusive to each site (Table 1). Since more basal nodes are likely to be shared by most species, UniFrac captures phylobetadiversity patterns related to more terminal nodes . This method is mathematically equivalent to the Jaccard index when a star phylogeny is considered . UniFrac gives very similar (but not exactly similar) results when compared to PhyloSor , which is another well-known phylobetadiversity measure . For this reason, we opted for using only the former. COMDIST, COMDISTNT and Rao’s H were computed inthe R environment (available at http://www.r- project.org), using the package picante 1.6–2 (, available at http://cran.at.r-project.org/web/packages/picante/). UniFrac was computed using the R package GUniFrac 1.0 (available at http://cran.r-project.org/web/packages/GUniFrac/index.html).
sportive lemur (Lepilemur edwarsi) populations in a very fragmented region on North-western Madagascar. They implemented two methods of analysis, one based on mitochondrial DNA (mtDNA) and the other on nuclear DNA (microsatellites). Their results showed an overall medium to low genetic diversity within populations. They also found evidence for a population collapse inthe last hundred years, even inthe largest forest analysed, having this species completely disappeared from the most isolated forest fragments (Craul et al., 2009). Likewise, the effects offorest fragmentation have also been demonstrated for other species, like the red ruffed lemur (Varecia rubra) (Razakamaharavo et al., 2009). On the other hand, a study was implemented for the golden-crowned sifaka (Propithecus tattersalli) from the Loky-Manambato region (North-eastern Madagascar), the same region of this study. The results showed that despite the high level oftheforest fragmentation observed inthe area, the populations still presented a high level of genetic diversity (Queméré et al., 2009). However the authors noted that the high expected heterozygosity values in their samples were associated with a small number of alleles which has been suggested to happen in populations which were previously large and subjected to a demographic bottleneck (Nei et al., 1975) and it has been linked to habitat fragmentation in other primates as well (e.g. Olivieri et al., 2008).
Despite that epiphyte biomass rarely accounting for more than 2% ofthe dry mass of a forest, its photosynthetic biomass, photosynthesis rate and ion capture rate can equal, or even surpass, that ofthe trees (Nadkarni 1984; Coxson & Nadkarni 1995, Benzing 1990, Hofstede et al. 1993). In certain forests, epiphytes comprise up to 63% ofthe photosynthetic biomass (Walker & Ataroff 2002) and 45% ofthe mineral content (Nadkarni, 1984). Additionally, epiphytes can inter- cept and accumulate substantial quantities of dead organic material, a rich source of nutrients for fauna and vegetation existing above the ground, influence water dynamics and the microclimate (Pócs 1980, Hofstede et al. 1993; Bohlman et al. 1995, Freiberg & Freiberg 2000). Epiphytes play an important role inthe primary productivity and nutrient cycling within ecosystems, contributing significant quantities of biomass. By absorbing nutrients from rain, mist or suspended particulate matter (Nadkarni 1984), epiphytes are capable of rapidly reintegrating energy and ions into an ecosystem (Matelson et al. 1993). Thus, they are not only influenced by the envi- ronmental changes caused by theedge effect but also play an important and decisive role in effecting those changes.
Ciids (Coleoptera: Ciidae) are minute tree-fungus beetles occurring in almost all terrestrial ecosystems. They feed and live inside persistent basidiomes of macrofungi, and are thus considered being micetobiont organisms and the most diverse and abundant taxon of this guild in tropical and subtropical lands [1,2,3,4]. There are few registers of ciids utilizing other resources or living outside fungi, the most common are micropterous and apterous species sometimes found in leaf litter or associated to dead wood [5 see for a summary]. However, there is no register about their association to another resource instead of fungi [6,7]. Such dependence makes them strictly vulnerable to variations of quantity and quality of fungi resources. When a species is more specialized, more seldomly it explores other resources [8,9,10]. Based on this, many studies have proposed that there are physic-chemical characteristics of fungi that stimulate ciids to establish a population, or prevent when they are in a resource to which they are not adapted for [1,11,12,13].
animal-dispersed and late successional species (Herlin & Fry 2000). However, some studies have shown species-specific distributions in corridors with different widths (Cervinka et al. 2013). Our results show that there were more pioneer and early successional species in narrow hedgerows, pos- sibly due to lateral light penetration, and higher zoochoric abundance, reflecting an increase in some animal-dispersed pioneer trees, such as Tabernaemontana catharinensis. Nonetheless, colonization by pioneer species is an important mechanism creating microhabitats favorable to late species and attracting more seed dispersing animals (Wunderle 1997; Florentine & Westbrooke 2004). Boughey et al. (2011) found that some bat species use hedgerows as narrow as 2 m, reinforcing the value of linear structures in providing food resources and shelter for these mammals.
TheAtlanticForest is home bristle 1-8% of all the flora and fauna ofthe world. In particular bats this biome represent about 64% of all species of bats in Brazil. Data on the composition and abundance of species of bat communities have never been synthesized and analyzed, as well as data about the spatial distribution and structure of bat communities in this biome, especially as their dissimilarity in a meta community perspective that can be broadly defined as a set of ecological communities in different locations (potentially, but not necessarily connected by dispersion), while a community is a group of species in a given location. In view of this I sought to evaluate the spatial pattern of collections made in this biome and the structuring of bat communities along theBrazilianAtlanticForest, as its beta diversity. Therefore gather a database of 57 articles totaling 342 locations through literature review. Realized statistics focused on spatial distribution of sampling, the sampling effort (median: 19140h.m²) and for the analysis of beta diversity and metacomunidades. The results show that there are collections across theAtlanticForest with higher density in PR outbreaks, south of Rio de Janeiro, PB-PE, south of Bahia. Sampling gaps are reported in ES, AL, CS and RS. The bat communities along theBrazilianAtlanticForest has a high heterogeneity (~ 0.9) mainly due to the turnover component and a low component of nesting. This result is valid for analysis in four spatial scales: locations, grid 2.5°, 5 and between 4 biogeographic regions oftheAtlantic. The structure of arrays of occurrence of species per sample unit (as defined above) presentam a random structure (non-coherent) without latitudinal gradient. Four inventories were conducted in four biogeographical regions oftheAtlanticForest biome within the project Network Inventory: diversity patterns, biogeography and endemic species of mammals, birds, amphibians, fruit flies and parasites intheAtlanticForest (CNPq / PPBIO) inthe rainy season totaling a 99.600 h.m² effort. 935 bats were collected, and tissue samples for DNA extraction purpose and parasitological diagnosis of trypanosomes (T. cruzi et Leishmania spp.), Hantavirus and helminths. The place with the highest species richness was the APA Pratigi BA, followed by Rebio Guaribas, PB; Serra dos Orãos, RJ State Park and the Sierra Board, SC. Recorded the first occurrence ofthe bat Ametrida centurio Grey, 1847 for theAtlanticForest, extending its distribution in more than 1000 km. The specimen was collected using canopy networks to 9m in REBIO Guaribas.
space (Mouchet et al. 2010). The high values of functional evenness (0.67 to 0.80) may be due to both filtering and limiting similarity processes, which have been reported to increase functional evenness (van der Plas et al. 2015). The similarities of distribution in reproductive traits also explain the high functional evenness since it will be maximized when species and abundance distributions are more even inthe functional space (Mouchet et al. 2010). Among these traits, the predominance of bee pollination system was expected, as were the traits related to bee pollination system, pollen-flowers, dish-shaped corollas as well as the predominance of androgynous flowers. Diverse Small Insects (DSIs) pollination system was also predominant. The presence of DSIs may be due to the presence of many flowers with less restrictive morphology, such as dish-shaped corollas. Flowers with dish- shaped corollas are visited by many different insects, such as bees, butterflies, moths, flies and wasps (Bawa 1990; Martins & Batalha 2007). For dispersal traits, the predominance of biotic dispersal system was also expected, as were traits related to biotic dispersal systems, such as black fruits. Dry fruits were also well represented inthe studied sites due to the large abundance of Alchornea and Tibouchina species.
Twelve sampling units were set up at each protected area and separated into six transects in each differently disturbed site. These transects were at least 120 m apart (median = 150 m). Each sampling unit comprised five pitfall traps arranged in a line and placed approximately two meters apart from each other. The pitfall traps were made of 500 ml plastic containers that were 8.5 cm in diameter, buried so the opening would be flush with the ground and protected from rain and falling leaves by a styrofoam cover. A mixture of 69.9% water, 30.0% propy- lene glycol, 0.1% formaldehyde and some drops of detergent (to break the surface tension and facilitate arthropod collec- tion) was used as preserving liquid. The traps were left open inthe field for six days per month. Sampling at the Santa Virgínia nucleus was done from November 2004 to May 2005. Like- wise, the Boracéia Biological Station and the Parque das Neblinas were sampled from November 2005 to May 2006, and collections at the Paranapiacaba Biological Reserve took place from November 2006 to May 2007.
SANTOS, T.R.M., 2008 [viewed 17 June 2013]. A influência de espécies introduzidas de peixes nas interações tróficas da ictiofauna da lagoa Carioca, Parque Estadual do Rio Doce, MG [online]. Belo Horizonte: Universidade Federal de Minas Gerais, 135 p. Masters Dissertation in Ecology, Conservation and Management of Wildlife. Available from: http://www.icb.ufmg.br/pgecologia/ dissertacoes/D205_Thiago_Russell_Miguel_Santos.pdf SPEZIALE, K.L., LAMBERTUCCI, S.A., CARRETE, M. and TELLA, J.L., 2012. Dealing with non-native species: what makes the difference in South America? Biological Invasions, vol. 14, no. 8, pp. 1609-1621. http://dx.doi.org/10.1007/s10530-011-0162-0. SUNAGA, T. and VERANI, J.R., 1991. The fish communities ofthe lakes in Rio Doce Valley, Northeast Brazil. Verhandlungen - Internationale Vereinigung für Theoretische und Angewandte Limnologie, vol. 24, no. 4, pp. 2563-2566.
296. Bierregaard, R.O.Jr.; Laurance, W.F.; Gascon, C.; Benitez- Malvido, J.; Fearnside, P.M.; Fonseca, C.R.S.; Ganade, G., Malcolm, J.R., Martins, M.B., Mori, S., Oliveira, M., Rankin- de Mérona, J., Scariot, A., Spironello, W., and Williamson, B. - 2001. Principles ofForest Fragmentation and Conservation inthe Amazon. In Bierregaard, R. O., Jr., Gascon, C., Lovejoy, T. E., and Mesquita, R. (eds.). 2001. Lessons From Amazonia: The Ecology and Conservation of a FragmentedForest - Yale University Press, New Haven, Connecticut, USA. Chapter 29. Pp. 371-385.
system, Oliveira-Filho (2009) proposed a new classification system for vegetation types in extra-Andean tropical and subtropical South America, aimed primarily at describing vegetation types at finer scales than those allowed by the IBGE system (< 1:10 vs. < 1:100,000). Because of its finer scales, the new system also allows the incorporation of ecological features that cannot be evaluated with the IBGE scale, particularly those related to soils, topography and drainage. As a starting point, the new system defines five main vegetation formations—forest, scrubland, savanna, grassland and man-made—which are subdivided into 16 first order vegetation types. That is the first level of classifica- tion, which deals with topological aspects ofthe plant mass, such as height, openness and leaf texture (e.g., broadleaved dwarf forest, stiff-leaf scrub, and parkland savanna). From there, up to five hierarchical attributes may be appended to make up the vegetation types, though the author stresses that some levels should be omitted as necessary, because the system is intended to allow flexibility inthe interest of user-friendliness. The five levels include climatic regime (rain, cloud, seasonal, semi-arid and maritime), leaf-flush regime (evergreen, semideciduous, deciduous, alternate and ephemeral), thermal regime (tropical/subtropical), elevational belt (coastal, lower plains, upper plains, lower highlands, and upper highlands) and substrate (sandy, rocky, eutrophic, slope, etc.). While the IBGE lists 28 formations, the new system—hereafter referred to as the Oliveira-Filho system (Oliveira-Filho 2009)—allows up to 202 combina- tions, if the substrate is disregarded. Nevertheless, because ofthe above-mentioned flexibility, the user rarely needs to apply all levels.
deciduous forest (SSF), mixed ombrophilous forest (MOF), and dense ombrophilous forest (DOF). Twelve study sites were sampled: four in SSF, five in MOF and three in DOF (Figs 1 and 2). The sites were selected according their size (larger than 500 ha) and history of disturbance (low disturbance); this procedure permitted a selection of representative sites of each forest type. We used the point count method with limited distance (100 m, H UTTO et al. 1986, B IBBY et al. 1992). Twelve point counts were placed in each site; each point was 200 from the other and at least 300 m from theforestedge. Point counts were performed inthe morning, beginning with increased bird activity, and fin- ishing around 3 hours later; the time for sampling in each point was 15 minutes, and another 15 minutes was the time necessary for the displacement ofthe observer and his assistant from one point to another. Six points were performed in each morning; the other six points were sampled the following day. We avoided performing point counts in days with rain and stronger wind.
The municipal district of Saquarema has outstanding differences in terms ofthe conservation of its restingas between the western and eastern portions. The western portion, locally denominated Nova Barra, has been thor- oughly destroyed after the removal ofthe restinga veg- etation for urbanization purposes that, visibly without planning, allowed the occupation ofthe whole area by properties, which extend close up to the beach. Even the beach vegetation (halophilous-psammophilous ground vegetation) has been removed for the construction of roads, which may reduce the capacity of fixation ofthe dunes. As a consequence, this coastal road now suffers serious problems related to the constant accumulation of sand on the highway and to the collapse of highway pas- sages due to the mechanical action ofthe sea waves, dur- ing periods of high tide. Inthe few remaining patches of sand vegetation still existent in Nova Barra and Jaconé, individuals ofthe sand lizard Liolaemus lutzae are still found. Inthe few remaining small patches of restinga vegetation near the beach in Jaconé, the threatened bird Formicivora littoralis can still be found. However, these patches are surrounded by properties and some of them are already for sale (Vecchi and Alves, no prelo). This is the only bird species endemic both to restinga habitats and to the State of Rio de Janeiro (Gonzaga and Pacheco, 1990).
The proportion of polymorphic loci detected for populations of M. coriacea was similar to the values detected for Gaultheria fragrantissima (86 %), which is also a dioecious shrub, usually found on forest edges (Apte et al. 2006). Interestingly, even herbaceous species of Primulaceae have a high proportion of polymorphic loci, ranging from 76 to 83 % in Androsace tapete (Geng et al. 2009), and from 58 to 64 % in Primula interjacens (Xue et al. 2004). Like M. Table 4. Within-population genetic diversityin ten populations of Myrsine coriacea. N = number of individuals from each locality; H’ = Nei’s genetic diversity (Nei 1973); I = Shannon’s genetic diversity index (Lewontin 1972); P - percentage of polymorphism. 1 – Alegre; 2 – Mimoso do Sul; 3 – Castelo; 4 – Dores do Rio Preto (Macieira); 5 – Domingos Martins; 6 – Iúna; 7 – Vargem Alta; 8 – Venda Nova do Imigrante; 9 – Muqui; 10 – Dores do Rio Preto (Casa Queimada).
In resource-limited environments, slow-growing species invest more resources in producing immobile defenses, such as lignin or total phenols, displaying lower herbivory rates (Coley 1983, Coley et al. 1985, Endara & Coley 2011). Tree species adapted to these environments exhibit lower capacity to absorb nutrients and to perform photosynthesis (Chapin 1980). Thus, for these plants, it is probably more costly to replace lost parts than to defend them (Coley et al. 1985, Fine et al. 2004, 2006, Agrawal 2006, Fine & Mesones, 2011). Although, many studies dealing with herbivory intheAtlantic rainforest are available, most are related to the community level or seed/fruit predation (Souza et al. 2013, Rossetti et al. 2014; Galetti et al. 2015, Corrêa et al. 2016). Leaf herbivory studies considering functional characteristics ofAtlanticforest trees in contrasting chemical and physical soils have not yet been performed. Therefore, the seasonally dry forestinthe Northern area ofthe state of Espírito Santo, Brazil, can be considered an ideal forest formation to assess the relationship between herbivory, nutritional aspects and chemical defenses, because it displays two distinct and peculiar phytophysiognomies under the same climatic condition: the Mussununga, with sandy soils and lower diversity, and the Mata Alta, with clayey soils and higher floristic diversity (Simonelli et al. 2008). Mata Alta soils and humus are more nutrient rich (with a lower C/N ratio) than Mussununga soils (Garay et al. 2003, 2016). Myrtaceae is recognised as one ofthe most important families in both types of forests, both in terms ofthe number of species and individuals ( Jesus & Rolim 2005, Simonelli et al. 2008, Giaretta et al. 2016). Furthermore, certain species of this family produce secondary compounds, such as terpenoids and phenols (Cooper 2001, Chaieb et al. 2007, Keszei et al. 2010) and display physical defenses, like leaf hardness, which provides them not only with chemical resistance, but also mechanical resistance to leaf-chewing insects (Sanson et al. 2001). In this context, we expected to find that Mussununga vegetation plants, with lower water content and higher fibre content, should therefore be less attacked by herbivorous insects compared to Mata Alta plants, which have higher water content and lower fibres. Only plants ofthe Myrtaceae family were investigated, due to their representativeness and the abundance of individuals inthe studied areas. The use of plants from the same family also allows for control regarding phylogenetical aspects.
Since no fruits developed after spontaneous and hand self-pollination inthe studied species, they probably are self-incompatible. Self-incompatibility has been suggested as prevalent in tropical trees (Bawa 1974; Bawa et al. 1985b). In fact, several tropical Leguminosae tree species, such as Swartzia pickelli (Lopes & Machado 1996) and Swartzia apetala (Moço & Pinheiro 1999) inthe north-eastern Atlantic rainforest, Pseudopiptadenia contorta (Prata-de- Assis-Pires & Freitas 2008) and S. multijuga (Wolowski & Freitas 2010) in a montane Atlantic rainforest, Senna silvestris (Carvalho & Oliveira 2003) and Copaifera langsdorffi (Freitas & Oliveira 2002), both intheBrazilian savanna, and Caesalpinia calycina (Lewis & Gibbs 1999) inthe caatinga, have been demonstrated as self-incompatible. However, since the number of treated flowers and individuals in this study was low, more experiments are necessary to detect the mating-system of these species. Furthermore, mixed mating was also reported for S. multijuga (Ribeiro & Lovato 2004). The low fruit set in natural conditions ofthe studied species may be because of self-deposition due to high floral constancy and behavior of pollinators, which remain for long periods foraging on the same individual plant promoting self-pollination followed by stigma clogging and pollen discount (Wilcock & Neiland 2002). Low fruit set under natural conditions and/or following hand cross-pollinations has been reported in many trees ofthe legume and other families (Bawa & Webb 1984; Bawa & Bullock 1989; Lopes & Machado 1996; Jausoro & Galetto 2001; Freitas & Oliveira 2002; Carvalho & Oliveira 2003; Prata-de-Assis-Pires & Freitas 2008).
A Shannon diversity (H’) index of 2.96 and a Pielou’s evenness (J) index of 0.71 were obtained. Studies in non-alluvial MOF of this region showed diversity and evenness values ranging from 2.79 and 0.70 in upper montane MOF (Higuchi et al., 2013) and 3.74 and 0.86 in montane MOF (Higuchi et al., 2012a), respectively. The low values observed inthe upper montane forest are the result of a more restricted environment due to the intense cold found inthe highest altitudes of this region. The limiting environment is also the reason why thediversity and evenness values are usually reduced in alluvial forests; in this case, the limitation is due to water excess. According to Junk (1993), flood stress and increased erosion and sedimentation rates reduce tree species diversity. Other studies in alluvial forests have also revealed low diversity values, such as Vilela et al. (2000), Budke et al. (2004) and Silva et al. (2009), who found H’ values of 0.93, 2.73, and 2.36, respectively. A higher diversity value was probably observed because there are also slopes in this area (some plots from theforest/non-forest matrix edge), where non-common species have been found in flood areas, thereby increasing richness. In addition, a range ofdiversity values may be observed in response to the environmental heterogeneity of alluvial forests, which are found in different climates, soil conditions, vegetation matrices and disturbance histories, as well as factors related to flooding, such as the degree and time of flooding (Silva et al., 2012a).