Changes in plant community of Seasonally Semideciduous Forest after invasion by Schizolobium parahyba at southeastern Brazil

Acta Oecologica 54 (2014) 57e64

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Original article

Changes in plant community of Seasonally Semideciduous Forest after invasion by Schizolobium parahyba at southeastern Brazil Rodolfo Cesar Real de Abreu a, Francisco Ferreira de Miranda Santos b, Giselda Durigan c, * a Research Center on Water Resources and Applied Ecology, Engineering School of São Carlos, University of São Paulo, Avenida Trabalhador São Carlense, 400, CP 292, CEP 13560-970 São Carlos, SP, Brazil1 b Graduate Program in Botany, Department of Botany, University of Brasília, Campus Universitário Darcy Ribeiro, Asa Norte, CEP 70910-970 Brasília, DF, Brazil2 c Forestry Institute of São Paulo State, Assis State Forest, CP 104, CEP 19802-970 Assis, SP, Brazil3

a r t i c l e i n f o

a b s t r a c t

Article history: Received 6 March 2013 Accepted 28 March 2013 Available online 25 April 2013

The recognition of a species as invasive is generally accepted when it comes from another continent or even from another country, but requires strong evidences of negative impacts to support control actions when the invasive species comes from another region in the same country. Schyzolobium parahyba e the ‘guapuruvu’, is a Brazilian tree native from the evergreen type of the Atlantic Forest, which has been recorded as invader in a number of remnants of the Seasonally Semideciduous Forest e SSF. We hypothesized that this giant and fast growing invasive tree changes the structure and composition of the understory, thus impairing the forest dynamics. We assessed the invasive population in the whole fragment, and, within the portion invaded, we sampled the regenerating plant community 1) under the largest alien trees, 2) under a native species with similar ecology (Peltophorum dubium), and 3) randomly in the forest. Density, basal area and richness under S. parahyba were remarkably lower than under the equivalent native species or in the understory as a whole. Floristic composition of the plant community was also distinct under S. parahyba, possibly due to increased competition for soil water. Even though the alien species has occupied, as yet, a small proportion of the forest fragment, it dominates the overstory and threatens the regeneration processes under its canopy. In view of our findings, we recommend extirpation of the species from SSF, as well as avoiding cultivation of the species away from its native range. Ó 2013 Elsevier Masson SAS. All rights reserved.

Keywords: Atlantic Forest Biological invasion Guapuruvu Nascent foci Community structure Plant invasion Understory

1. Introduction A clear distinction between natural range expansion and invasion is not easy (Vermeij, 2005; Valéry et al., 2008). Nevertheless, new arrivals to a community are managed typically depending on whether is an extra range natural dispersal or a human-mediated range expansion (Wilson et al., 2009). An invasive species has been defined as an alien species that sustains self-replacing populations over several life cycles, forms reproductive offspring in very large numbers at considerable distances from the parent and/or site of introduction, and has the potential to spread over

* Corresponding author. Fax: þ55 18 33238330. E-mail addresses: [email protected] (R.C.R. Abreu), [email protected] (F.F.M. Santos), [email protected] (G. Durigan). 1 http://buscatextual.cnpq.br/buscatextual/visualizacv.jsp?id¼K4718652D4. 2 http://buscatextual.cnpq.br/buscatextual/visualizacv.jsp?id¼K4231178H4. 3 http://lattes.cnpq.br/6907261164017372. 1146-609X/$ e see front matter Ó 2013 Elsevier Masson SAS. All rights reserved. http://dx.doi.org/10.1016/j.actao.2013.03.013

long distances (Richardson et al., 2011). The recognition of a species as invasive is generally accepted when it comes from another continent or even from another country and leads to ecological changes or economic losses (Pimentel et al., 2005). Although invasive species recorded outside their natural range within the same country can cause severe impacts on natural ecosystems, control of these species may still encounter resistance from the public and policy makers (Simberloff, 2003a). Resistance to control these invaders is greater when the species is noted for some human value (e.g. large size, fast growth, beauty). Such is the case for the guapuruvu e Schizolobium parahyba (Fabaceae: Caesalpinioideae) in the Brazilian Atlantic Forest. The Atlantic Forest in the state of São Paulo has two main subdivisions, related to the climate conditions (Oliveira-Filho and Fontes, 2000): the Evergreen Forest e along the coast, where the climate is always wet (average annual precipitation about 2500 mm yr
1, Sentelhas et al., 2003), and the Seasonally Semideciduous Forest (hereafter SSF) e located farther inland and

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characterized by a dry winter average annual precipitation about 1500 mm yr
1 (Sentelhas et al., 2003). Historically, guapuruvu is restricted to the evergreen type (Andrade and Vecchi, 1916; Bentham, 1876; Correa, 1926), where it was reported by Correa (1926) as “the tallest tree in the Atlantic coastal forest” (Correa, 1926). Recent evidence of the guapuruvu as invader at SSF was obtained from 232 records of herbarium specimens from the Species Link data base (http://www.splink.org.br/, accessed 08/12/2012). S. parahyba has been recorded in the Atlantic Forest region since the nineteenth century. All exsiccates and records of guapuruvu at SSF before 1985 corresponded to cultivated specimens. However, 23 exsiccates were collected from fragments of SSF after 1985, and several new occurrences were mentioned in the literature (Matthes et al., 1988; Nicolini, 1990; Souza et al., 2009; Nogueira, 2010). The recent range expansion of guapuruvu into the SSF likely occurred as a result of human-mediated invasion. Guapuruvu is a fast growing tree (Backes and Irgang, 2002; Lorenzi, 2002), recommended and widely cultivated as ornamental or for restoration purposes, being included in almost 50% of forest restoration projects away from its native range, in the western São Paulo state (Barbosa et al., 2003). From the trees planted as ornamentals or for forest restoration, the species has easily invaded patches of SSF, and the consequences of this process have not yet been evaluated. The functional performance of S. parahyba in restoration plantings and possible unintended negative consequences on the restored communities were never assessed. In spite of the evidence of recent invasion of the SSF by the guapuruvu, it has not been recognized as invasive species and a threat to the biodiversity. The potential for a catastrophic loss in biodiversity is great. The Atlantic Forest as a whole supports one of the highest degrees of species richness and rates of endemism on the planet and is among the most threatened biomes in the world (Ribeiro et al., 2009); it has been elected as one of the global biodiversity hotspots for conservation (Myers, N., et al., 2000). Of the various sub-types of the Atlantic Forest, the SSF is one of the most threatened by habitat fragmentation, since it has the most suitable soils and relief for agriculture and was the first to be deforested by the European settlers. While 11.8% of the Atlantic Forest (sensu lato) remains, the remaining proportion of SSF is only 7.1%, scattered among thousands of small and isolated fragments (Ribeiro et al., 2009) and susceptible to a wide range of edge effects, including biological invasions. Given the potential consequences of unchecked expansion of guapuruvu on SSF, it is crucial that we monitor effects on the plant community at early stages of its invasion. We assessed the recent spread of the giant trees of Schizolobium parahyba into a SSF fragment and its impact on the invaded ecosystem by assessing the plant community regenerating under the largest alien trees, under an ecologically similar species (Peltophorum dubium), and at random locations in the invaded portion of the forest. We hypothesized that the fast growing invader, adapted to the evergreen rain forest, can change the native community of the semideciduous forest in its composition and structure, constraining the forest dynamics and threatening the native species populations. 2. Material and methods 2.1. Study species e the alien Schizolobium parahyba and the native Peltophorum dubium From the species of the SSF, we choose P. dubium (Spreng.) Taub. (another species from the family Leguminosae e Caesalpinioideae), as the “native equivalent” to S. parahyba. S. parahyba and P. dubium are functionally very similar species (Carvalho, 2003). Both are fast

growing trees, classified as pioneers or early secondary, with yellow flowers pollinated by bees. Their seeds are dispersed by wind or gravity, and persist in the soil seed bank for years, since they have tegumentary dormancy. Both are emergent trees, surpassing 20 m in height, and 100 cm in d.b.h., with broadly umbelliform crowns. Reproductive processes start at 6e8 years for S. parahyba and 7e12 years for P. dubium, but young individuals are very rare in the understory, since both species are heliophylous. In spite of being legumes, neither is N-fixing. The main differences between S. parahyba and P. dubium are their native range within the Atlantic Forest (the evergreen forest and the SSF, respectively), as well as their growth rhythm (up to 45 and 19.6 m3 ha
1 yr
1, respectively), wood density (0.22e0.40 and 0.75e0.90 g cm
3, respectively), and longevity. While S. parahyba very rarely lives over 40 years (but can reach 70 years; Callado and Guimarães, 2010), P. dubium can often surpass a hundred years. P. dubium is reported as deciduous and S. parahyba as semideciduous in their respective native ranges (Carvalho, 2003). In the SSF, however, the guapuruvus behave as deciduous, loosing all their leaves in the dry season e an indication of soil water deficiency e and recovering them after the first rains in the spring. Another relevant distinction is that P. dubium is resistant to wind while S. parahyba easily breaks with the wind and opens gaps. 2.2. Study area The study site is a SSF fragment of 73 ha, located at 22 470 2000 S and 50 470 2000 W with altitudes ranging from 426 to 455 m (a.s.l.) e at Tarumã municipality, southwestern São Paulo State, Brazil, about 500 km west from the native range of guapuruvu (Fig. 1). Local climate is tropical, with average annual rainfall around 1400 mm and a dry season in the winter, which can last up to five months (Brando and Durigan, 2004). High fertility and clayish texture characterize the local soil, classified as Eutrudox. Soil and topography are both homogeneous, and there is no record of recent disturbances such as fires or tree cutting in any portion of the forest. The forest fragment is far from water bodies, and isolated from other forest remnants at least during the last 50 years. It has been surrounded by sugar cane plantations, with high proliferation of lianas along the edges and in natural gaps. Well-preserved remnants of the same forest type e the SSF have been studied in the region, where basal area reaches 30 m2 ha
1 (Durigan et al., 2000). The spread of the invasive trees over the native forest at the study site started from a stand of S. parahyba planted in 1965 for ornamental purposes, in a 30  140 m area (6 m  6 m spaced), between the edge of the native forest and the main entrance of the farmhouse. 2.3. Extent of invasion and population structure We assessed the alien population of S. parahyba by a census of all individuals from 1 m in height, established within the native forest and also under the planted stand. We searched for individuals of S. parahyba through the whole fragment, at first from the edges, since the emergent trees are easily visible, and later by transects from the planted stand towards the core area of the fragment. In the area where the adult trees were recorded and at least 50 m around them, the whole terrain was explored in search for young individuals. Each individual was labelled, mapped (coordinates taken with a GPS) and measured (d.b.h. e diameter at breast height e and canopy diameter, as its projection on the ground). 2.4. Invasion impacts To evaluate the impact of the alien species on the native forest, we compared the structure and composition of the native plant

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Fig. 1. Study area location (black square) and natural range of Schizolobium parahyba (in grey) in the state of São Paulo, Brazil. The new range is 500 km west from the limits of its natural range “Serra do Mar” e a mountain chain that reaches about 1300 m a.s.l.

community regenerating under the canopies of S. parahyba and under the “native equivalent” e P. dubium. From each species, ten adults with d.b.h.  30 cm were randomly selected as sample units. The understory was surveyed in a circular plot of 50 m2 around each stem. In addition, we sampled ten plots randomly distributed in the invaded area (hereafter referred to as ‘forest plots’), at least 10 m far from the projection of the canopies of S. parahyba or P. dubium. The 30 plots, representing the three ‘treatments’, were randomly distributed within the invaded portion of the forest fragment. All individuals of tree species with height  50 cm and d.b.h. < 5 cm regenerating in each plot were measured and identified. The minimum height  50 cm has been the inclusion criteria for most studies of the understory of the Atlantic Forest (e.g. Souza and Batista, 2004; Tabarelli and Mantovani, 1997). Below this height, young trees are difficult to identify and mortality is very high, particularly during the dry season. By excluding individuals

with d.b.h.  5 cm, our analyses were restricted to young plants (maximum age estimated in 15 years on the basis of average tree growth in the region) established under the canopies of the invaders or the native trees sampled, and thus after gap closure, since both heliophylous species (the invader and the native) certainly occupied forest gaps in the past. 2.5. Data analyses We first assessed the extension of the forest fragment that was colonized by the alien species. The invaded area was quantified as the polygon comprising the canopies of all individuals of the alien species growing in the native forest fragment, mapped by ArcGis (ESRI, 1999). Considering this polygon, we estimated the density (individuals ha
1), canopy cover (%), and basal area (m2 ha
1) of the alien trees in the invaded portion of the forest. We also assessed inequality in population size structure using Gini coefficient and

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Lorenz curve (Weiner and Solbrig, 1984), performed in R (R Development Core Team, 2009), with ineq package (Zeileis, 2012) for Gini calculations and boot package (Canty and Ripley, 2010; Davison and Hinkley, 1997) for 5000 bootstrap calculations. The invasion impact on the native plant community was evaluated by comparing density, basal area, and floristic composition of the assemblages regenerating under S. parahyba (the alien tree), P. dubium (the native equivalent) and the forest plots. Species were classified according to their growth rates (fast, moderate or slow), shade tolerance (tolerant or intolerant) and adult size (large, median, small) (Durigan et al., 2004). Density and basal area of the regenerating community among the three sets of plots were compared conducting a single, one-way ANOVA with three treatments (S. parahyba, P. dubium, and forest plots) followed by post hoc Tukey test for multiple comparisons. We performed an Indicator Species Analysis (ISA e Dufrêne and Legendre, 1997) to the floristic data, to test the null hypothesis that the species from the understory establish at random in the habitat types. The ISA indicates which species better represent a specific habitat or treatment. To be conservative, after performing the analyses with the whole set of species recorded, we included only the species with frequency higher than 40% of plots in a specific habitat (occurring at least in four plots). For the post hoc Tukey test we used the lme4 package (Bates et al., 2011) and for ISA the labdsv package (Robert, 2012). Multivariate analyses of community composition were conducted using nonmetric multidimensional scaling (NMDS). We used species abundance data from each 50 m2 plot under S. parahyba, under P. dubium, and in the forest plots. The NMDS was calculated with the vegan package (Oksanen et al., 2010) in R, using a BrayeCurtis dissimilarity matrix. 3. Results The invasive population occupied an area of 5400 m2 contiguous to the planted stand, which corresponds to a small proportion (0.7%) of the forest fragment. The individual established deepest within the interior of the forest (6 cm d.b.h. and 6 m high) was at a 32 m distance from the planted stand. Although the invaded area is not large, the invaded portion of the forest was clearly dominated by the guapuruvus. The projections of the large canopies of

S. parahyba, taller than the native trees, covered about 70% of the invaded forest. A total of 33 invading trees were recorded in the forest fragment, corresponding to a density of 65 individuals ha
1 considering just the invaded area and not the planted trees. The invasive population was composed mainly by large mature trees (Fig. 2A) with only 11% of individuals lower than 20 cm d.b.h. On the other hand, under the planted stand the population size structure fits to the reverse-J curve, with the young individuals being the majority (Fig. 2B). The basal area of guapuruvus in the invaded community totalled 10 m2 ha
1. The Gini coefficient for the invasive population was 0.307 (95% CI ¼ [0.195e0.367]) and in the planted stand it was 0.589 (95% CI ¼ [0.263e0.609]). The Lorenz curves were different between both environments (Fig. 2). The density of regenerating native tree species (Fig. 3A) differed among habitat types (F ¼ 84.56; P < 0.001). Tukey test results indicated that density of native regenerating trees under S. parahyba (2040 individuals ha
1) was lower than under P. dubium (11,480 individuals ha
1; P < 0.001), and forest plots (8260 individuals ha
1; P < 0.001). No difference was found between the forest plots and the P. dubium understory (P ¼ 0.27). The basal area of regenerating plants also differed among habitat types (F ¼ 12.99; P < 0.001; Fig. 3B). Basal area under the guapuruvus (0.66 m2 ha
1) was lower than under P. dubium (1.46 m2 ha
1; Tukey HSD P < 0.001) and forest plots (1.73 m2 ha
1; P < 0.001). No difference was found for basal area between P. dubium and the forest plots (P ¼ 0.68). Richness was lower under the invasive S. parahyba (40 species recorded under the ten trees) than under the native P. dubium (51 species recorded) or in the forest plots (46 species). In addition, the assemblage under S. parahyba is remarkably different from that under P. dubium or in the forest plots (Table 1). ISA showed that most species surveyed under P. dubium or in the forest plots were late successional (shade-tolerant and slow-growing) usually recorded in undisturbed forests (Table 2), while a series of fast growing pioneer or drought adapted species (e.g. Aloysia virgata, Gallesia integrifolia, Guarea macrophylla, Maclura tinctoria) were exclusive to or more abundant in the understory of S. parahyba. NMDS community ordination (Kruskal stress ¼ 0.201) indicated that the assemblage under S. parahyba trees differed from the

Fig. 2. Size structure (diameter classes) of the invasive population of Schizolobium parahyba in: A) Native forest fragment; and B) in the planted stand. Lorenz curves representing the inequality measures from both stands. Fragment in light grey and planted stand in dark grey. The Gini coefficient corresponds to the area between each Lorenz curve and the black diagonal; G0 ¼ 0.307 in the invaded area and G0 ¼ 0.590 in the planted stand.

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Fig. 3. Comparisons among the understory of Schizolobium parahyba adult trees (invasive species; n ¼ 10), Peltophorum dubium (native species; n ¼ 10), and forest plots (n ¼ 10) in Seasonally Semideciduous Forest, Tarumã, SP e Brazil. Boxplot (median  quartiles): A) Density (individuals ha
1); and B) basal area of native plants regenerating (height  50 cm, d.b.h. < 5 cm).

assemblage under P. dubium trees, which was similar to the forest plots. Differences in community composition under non-native trees and the other plots were largely described by scores from NMDS axis 1 (Fig. 4). 4. Discussion 4.1. Invasiveness of Schizolobium parahyba The extra-range dispersal pathway of S. parahyba over the SSF followed two steps: first it was human-mediated and classified as cultivation type (Wilson et al., 2009) e when the species was planted as an ornamental or in ecological restoration programs away from its native range (Barbosa et al., 2003). In the past few decades, the species has spread from plantations to the native forest remnants. In this second step, the leading-edge dispersal type has occurred, when the species expands from the edge of the new range (Wilson et al., 2009). The area invaded by the guapuruvus after 42 years of plantation is not large, being restricted to the native forest edge, an area corresponding to less than 1% of the fragment. What is probably limiting the spread of guapuruvus to the core areas of the fragment in this first cycle of invasion is the short-distance wind dispersal, limited by the heavy seeds (Carvalho, 2003). In the invaded portion of the forest, however, the basal area of guapuruvus corresponds to one third of the average basal area of mature stands of seasonally forest in the same region (Durigan et al., 2000), and their tall canopies cover about 70% of the area, shading the native trees. The invasive species is virtually dominating the forest structure in the invaded area. The population of guapuruvus invading the forest does not follow the reverse-J pattern (negative exponential) a pattern for established and persistent species (Crawley, 1997). As expected for pioneer or early secondary species (Carvalho, 2003), at the shaded environment inside the forest, the guapuruvus probably occupied forest gaps in the past, but are not successfully regenerating at present. The Gini’s coefficient and the Lorenz curve did not show size asymmetries in the invasive population. The size structure of the invasive population, however, can potentially change in the near future if some natural or artificial disturbances occur close to the parent trees, where a persistent seed bank of the species exists (Carvalho, 2003). A lag phase is common between establishment and spread, when the invasive species adapt to the new community (Sakai et al., 2001; Barney, 2006; Abreu and Durigan, 2011). Longevity of Schizolobium parahyba is unknown in the new range, but after 42 years less than 30% of the planted trees are still alive. In the native range they usually do not live for more than 70

years (Callado and Guimarães, 2010), easily falling or breaking branches (Lorenzi, 2002; Backes and Irgang, 2002). The predicted falling of the elder giant trees will certainly open new large gaps. The fast growing invasive species may overtake native ones in the gaps closure, starting a new invasion cycle, from the seed bank. The fast growth and the short life cycle of the guapuruvus give the species a high potential to considerably reduce the gap turnover, accelerating the invasion process. Often, natural or artificial disturbances trigger biological invasion processes (Williamson, 1996). In the fragmented SSF, tree fall gaps are the most prominent natural disturbance. They can increase light penetration, expose mineral soil and change the environmental conditions in the understory, promoting the germination of species including the aliens. Gap related biological invasions were noticed in a deciduous forest in USA where the invasive shrub Rubus phoenicolasius Maxim. first established into tree fall gaps before spreading through the forest understory (Gorchov et al., 2011). In tropical forests of Tanzania, Maesopsis eminii Engl., an early successional zoochorous tree introduced for wood production, escaped from plantations and outcompeted native species first establishing into tree fall gaps but then reaching even patches of undisturbed forests (Binggeli and Hamilton, 1993). The forest edges adjacent to these plantations were all colonized by this invasive species, and the greatest distance between colonized forest patches and source sites over which dispersal and recruitment succeeded was 3.97 km (Cordeiro et al., 2004). Although the invasion by Schizolobium parahyba is restricted as yet to gaps close to the parent trees, theoretically, the species can slowly invade all the gaps of the invaded forest. 4.2. Invasion impact on the native community The fast growing invasive species caused remarkable changes in the forest regeneration processes and, naturally, could threaten the community assembly and dynamics of the forest studied here. Apparently, the invasive species modified the microhabitat, acting as a biotic filter, hindering the establishment of seedlings of many tree species and favouring others, including itself e a common pattern in invasion ecology (Williamson, 1996). This process was previously observed in the Caatinga biome (Pegado et al., 2006), after the invasion by Prosopis juliflora in northeastern Brazil. Adult trees acting as filters and determining the composition of the understory have often been detected in tropical forests. In the semideciduous Atlantic Forest, which is the vegetation type of this study, some functional attributes of the adult trees related to competition for light and soil water have been considered as possible explanations for the understory composition (Gandolfi

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Table 1 Tree species (height  50 cm) regenerating under adult trees of Schizolobium parahyba (invasive species), Peltophorum dubium (native species), and the forest plots in Seasonally Semideciduous Forest (Tarumã, SP, Brazil). Growth ¼ growth rhythm (slow, medium, or fast); size ¼ adult tree size (small, medium, or large); ST ¼ shade tolerance (positive or negative). * Late successional species. Species

Growth Size ST S. parahyba P. dubium Forest

Aegiphila sellowiana Alchornea glandulosa Allophyllus edulis* Aloysia virgata Aspidosperma polyneuron* Astronium graveolens* Balfourodendron riedelianum* Bauhinia longifolia Bougainvillea glabra Bunchosia pallescens* Calliandra foliolosa* Campomanesia guazumaefolia* Campomanesia xanthocarpa* Casearia gossypiosperma Casearia sylvestris Cedrela fissilis Centrolobium tomentosum Chomelia pohliana Chorisia speciosa Chrysophyllum gonocarpum* Cordia ecalyculata Cordia trichotoma Cordia superba Croton floribundus Cupania vernalis* Diatenopteryx sorbifolia* Endlicheria paniculata* Esenbeckia febrifuga* Eugenia blastantha* Eugenia rhamboi* Eugenia uniflora* Gallesia integrifolia Guapira hirsuta* Guapira graciliflora* Guapira opposita Guarea guidonea Guarea kunthiana* Guarea macrophylla Handroanthus heptaphyllus Holocalyx balansae* Inga uruguensis Lonchocarpus guilleminianus Lonchocarpus muehlbergianus Machaerium stipitatum Machaerium nyctitans Machaerium paraguariense Maclura tinctoria Matayba elaeagnoides Maytenus aquifolium* Metrodorea nigra* Myroxylon peruiferum* Ocotea indecora* Ocotea velutina Parapiptadenia rigida Patagonula americana Peltophorum dubium Picramnia ramiflora* Picramnia sp. Pilocarpus pennatifolius* Piptadenia gonoacantha Pterogyne nitens Plinia trunciflora Randia calycinna* Rhamnidium elaeocarpum* Rheedia gardneriana Rollinia sp. Rollinia sylvatica* Schizolobium parahyba Seguieria floribunda Senagalia polyphylla

F F S F S M S M M M M S S M M M M M M S M M M F M S M F S S S F M M M M M M M S F F F M S M F M S S S M M F S F M M S F M S M M M M M F M F

S M S M L L L M M M S M M M M L L S L M M M L M M M M S S S M L M M M M M M L L M M M M M L M M M M L M M M L L M M M M M M S M M M M L M M



þ
þ þ þ þ þ þ þ þ þ þ þ þ þ þ
þ þ


þ þ þ
þ þ þ
þ þ þ þ þ þ
þ


þ
þ
þ þ þ þ þ þ
þ
þ þ þ

þ þ þ þ þ þ
þ


22 2 0 6 7 5 0 1 11 0 0 0 4 14 0 1 5 0 0 1 0 2 0 1 2 0 0 0 2 0 0 7 0 0 0 0 3 6 1 1 0 1 0 3 0 0 5 0 0 5 1 3 0 5 0 0 0 1 0 0 1 0 0 1 0 4 1 10 0 0

0 0 5 0 35 19 22 0 6 2 22 1 18 11 1 0 5 1 1 22 1 0 0 0 7 9 2 1 0 2 3 0 23 15 0 3 3 0 0 6 1 0 2 22 0 0 0 2 7 118 0 7 0 0 1 1 2 0 9 4 0 0 3 4 1 0 8 0 12 10

0 0 2 0 11 9 13 0 8 0 8 1 17 5 1 0 1 1 0 5 0 0 1 1 18 0 0 7 3 2 2 0 6 0 10 0 0 0 0 5 0 0 1 2 2 1 0 2 9 134 0 2 2 0 0 0 0 0 6 2 0 1 5 7 0 0 3 0 14 4

Table 1 (continued ) Species

Growth Size ST S. parahyba P. dubium Forest

Syagrus romanzoffiana Tabernaemontana catharinensis Trichilia catigua* Trichilia claussenii* Trichilia elegans* Trichilia pallida* Urera baccifera Zanthoxylum rhoifolium Total

M M M S S M F M

M M S M S M S M

þ þ þ þ þ þ þ þ

0 1 1 4 4 2 1 1 159

0 3 20 20 99 3 0 0 605

1 4 19 13 39 3 0 0 413

et al., 2007; Souza, 2007). The differences in species composition under the invasive trees in this study reflect a habitat that is no longer suitable for most native species that characterize the SSF. The lower density of plants regenerating under the canopy of S. parahyba compared to the non-invaded plots was likely caused by regeneration failure of many of the native species. Light conditions for the regenerating young plants were very similar under the alien and the native adult trees, since both are deciduous species that establish in gaps. Therefore, we believe that competition for soil water triggered by the guapuruvus may be the main constraint for native species establishment and growth. In some cases, invasive plant species are better competitors than the native ones (Vilà and Weiner, 2004). A review of pair-wise comparative studies on water consumption by invasive and native species, however, suggest that the invaders do not necessarily consume relatively more water than natives (Cavaleri and Sack, 2010). Even if it is not a global pattern, it could nonetheless be the case for guapuruvus. A tree-ring study of S. parahyba in the Atlantic coastal forest e where the species is native, demonstrated that annual girth growth is correlated with precipitation (Callado and Guimarães, 2010). Some physiological adaptations and growth characteristics of the species, such as low wood density, height, and fast growth are closely related to high water use (Goldstein et al., 1998; Kagawa et al., 2009; Asbjornsen

Table 2 Habitat preference of the species according to Indicator Species Analysis. Only species recorded in at least 40% of plots of a habitat type were considered. Indval ¼ the indicator value for each species; n ¼ number of plots where the species was recorded. P ¼ p values. *Late successional species. P. dubium and forest

Indval

n

P

Balfourodendron riedelianum* Calliandra foliolosa* Campomanesia guazumaefolia* Chrysophyllum gonocarpum* Esenbeckia febrifuga* Eugenia rhamboi* Eugenia uniflora* Guapira hirsuta* Guapira opposita Holocalyx balansae* Maytenus aquifolium* Metrodorea nigra* Pilocarpus pennatifolius* Piptadenia gonoacantha Seguieria floribunda Senegalia polyphylla Tabernaemontana catharinensis Trichilia catigua* Trichilia claussenii* Trichilia elegans*

0.40 0.44 0.42 0.33 0.38 0.55 0.40 0.41 0.42 0.34 0.43 0.32 0.40 0.43 0.37 0.40 0.39 0.35 0.34 0.36

14 10 6 12 6 4 5 10 11 7 10 24 12 5 12 8 4 17 17 21

0.001 0.001 0.001 0.001 0.001 0.001 0.004 0.001 0.001 0.004 0.001 0.001 0.001 0.001 0.001 0.001 0.012 0.001 0.001 0.001

S. parahyba

Indval

n

P

Aloysia virgata Gallesia integrifolia Guarea macrophylla Maclura tinctoria

0.35 0.42 0.35 0.42

4 4 5 4

0.05 0.005 0.02 0.005

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63

different techniques, mixed or alone, brought beneficial results at least four times greater than the costs (De Wit et al., 2001). The decision for eradication depends on both the evaluation of the costs and benefits, and the potential to be successful (Myers J.H., et al., 2000). Eradication of guapuruvus from fragments of SSF would be beneficial to the environment and economically feasible in this nascent foci phase. The elimination of the propagule sources (invading and planted individuals) is strongly recommended to increase success of extirpation. Extirpation is feasible once provided actions begin before reproductive maturity of recent colonists (wsix years), when the invasive species still occurs at low densities. Control efforts must be periodically repeated until the depletion of the long-lived seed bank. 5. Conclusions

Fig. 4. Nonmetric multidimensional scaling (NMDS) ordination graph of native species abundance under adult native trees of Peltophorum dubium (open squares, n ¼ 10), invasive Schizolobium parahyba (filled triangles, n ¼ 10), and the forest plots (filled circles, n ¼ 10). Kruskal Stress ¼ 0.201. Each symbol represents plant species regenerating (height  50 cm, d.b.h.  5 cm) in a single 50 m2 circular plot.

et al., 2011). Such evidence of high water consumption, in addition to the wet climate of its native ecological range, suggests that the key limiting resource under the invasive species is soil water. The assumption that water uptake is more rapid or complete underneath the invasive species is consistent with the observation that most native species that managed to survive at the understory of S. parahyba are adapted to environments with a greater water stress, such as the fast growing pioneers Aegiphylla sellowiana and Aloysia virgata (Lorenzi, 2002), among others. 4.3. Management recommendation Invasion by guapuruvus in the SSF fragments is recent (beginning around four decades ago) and not yet recognized by the general public or plant ecologists in Brazil. The management of exotic species involves dealing with humans’ values that can change at each generation, especially when people are brought up in the presence of exotic species and get used to them (Schlaepfer et al., 2011). In these cases, a non-native species can become part of native ecosystems, being no longer seen as a species that causes problems to the natural ecosystem, and thus eradication may encounter public resistance (Schlaepfer et al., 2011). In addition, the impressive size and beauty of the mature guapuruvus lead people to find it hard to accept that they must be extirpated. These can be major obstacles to developing control actions, especially if amongst the most important recommendations to obtain success in controlling invasions is to act in the initial phases of the invasion or on nascent foci is amongst the most important recommendations to obtain success in controlling invasions (Moody and Mack, 1988; Puth and Post, 2005). Once a species is recognized as being invasive, population control with management becomes necessary (Simberloff, 2003b). As verified by De Wit et al. (2001), analyzing different possibilities for managing the invasive species Acacia mearnsii in South Africa, the decision not to manage the species was shown to be an unsustainable means of protecting the natural ecosystem in the medium and long-term. On the other hand, management options with

When moved from its native range, the evergreen forest, to the vicinity of SSF fragments as an ornamental tree or for restoration purposes, Schizolobium parahyba became a threat to native biodiversity. It is now able to colonize the native forests in the new range and displace late successional native species. Therefore, invasion of the semideciduous forest by the guapuruvus and subsequent replacement of the native community justifies efforts to control the spread of this species, especially in protected areas of this forest type. We consider that control actions must be immediate on the basis of the precautionary principle. In view of our findings we recommend extirpating the species from the remnants of SSF and avoiding its plantation in areas away from its native ecological original range. Acknowledgements Our gratitude to Mariana Regina Durigan, Juliano Fernandes Corrêa, Beatriz Zidioti Ferreira and Natashi A.L. Pilon for helping in the field work, to Dr. Eduardo Pinheiro for drawing Fig. 1. We thank the National Council for Scientific and Technological Development (CNPq) for research grant to GD and PhD grant for RCRA and CAPES/ FULBRIGHT for a research grant to RCRA. We also thank J. Stephen Brewer for his contributions to the final version of the manuscript, and the valuable suggestions from anonymous reviewers. References Abreu, R.C.R., Durigan, G., 2011. Changes in the plant community of the Brazilian grassland savanna after 22 years of invasion by Pinus elliottii Engelm. Plant Ecology and Diversity 4, 269e278. Andrade, E.N., Vecchi, O., 1916. Les bois indigènes de São Paulo. Serviço Florestal, São Paulo, Brasil. Asbjornsen, H., Goldsmith, G.R., Alvarado-Barrientos, M.S., Rebel, K., Van Osch, F.P., Rietkerk, M., Chen, J., Gotsch, S., Tobon, C., Geissert, D.R., Gomez-Tagle, A., Vache, K., Dawson, T.E., 2011. Ecohydrological advances and applications in plant-water relations research: a review. Journal of Plant Ecology 4, 3e22. Backes, P., Irgang, B., 2002. Árvores do Sul. Instituto Souza Cruz, Porto Alegre, Brasil. Barbosa, L.M., Barbosa, K.C., Potomati, A., Martins, S.E., Asperti, L.M., Melo, A.C.G., Carrasco, P.G., Castanheira, S.A., Piliackas, J.M., Contieri, W.A., Mattioli, D.S., Guedes, D.C., Santos-Júnior, N.A., Silva, P.M.S., Plaza, A.P., 2003. Recuperação florestal com espécies nativas no Estado de São Paulo: pesquisas apontam mudanças necessárias. Florestar Estatístico 14, 28e34. Barney, J.N., 2006. North American history of two invasive plant species: phytogeographical distribution, dispersal vectors, and multiple introductions. Biological Invasions 8, 703e713. Bates, D., Maechler, M., Bolker, B., 2011. LME4: Linear Mixed-effects Models Using S4 Classes. R package version 0.999375-42. Bentham, G., 1876. Leguminosae II et III (Caesalpiniae, Mimosae). In: Martius, K.F.P., Eichler, A.G., Urban, I. (Eds.). Flora Brasiliensis, vol. 15, pp. 1e528. Rio de Janeiro. Binggeli, P., Hamilton, A.C., 1993. Biological invasion by Maesopsis eminii in the East Usambara forests, Tanzania. Opera Botanica 121, 229e235. Brando, P.M., Durigan, G., 2004. Changes in Cerrado vegetation after disturbance by frost (São Paulo State, Brazil). Plant Ecology 175, 205e215. Callado, C.H., Guimarães, R.C., 2010. Estudo dos anéis de crescimento de Schizolobium parahyba (Leguminosae: Caesalpinioideae) após episódio de mortalidade em Ilha Grande, Rio de Janeiro. Revista Brasileira de Botânica 33, 85e91.

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