Land use history drives differences in functional composition and losses in functional diversity and stability of Neotropical urban forests

Journal Pre-proof Land use history drives differences in functional composition and losses in functional diversity and stability of Neotropical urban forests Marcela V. Pyles, Luiz F.S. Magnago, Erica Rievrs Borges, Eduardo van den Berg, Fabr´ıcio Alvim Carvalho

PII:

S1618-8667(19)30717-4

DOI:

https://doi.org/10.1016/j.ufug.2020.126608

Reference:

UFUG 126608

To appear in:

Urban Forestry & Urban Greening

Received Date:

17 September 2019

Revised Date:

13 January 2020

Accepted Date:

22 January 2020

Please cite this article as: Pyles MV, Magnago LFS, Borges ER, van den Berg E, Carvalho FA, Land use history drives differences in functional composition and losses in functional diversity and stability of Neotropical urban forests, Urban Forestry and Urban Greening (2020), doi: https://doi.org/10.1016/j.ufug.2020.126608

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Title: Land use history drives differences in functional composition and losses in functional diversity and stability of Neotropical urban forests

Running headline: Functional changes and stability of urban forests

Authors: Marcela V. Pyles1,2*, Luiz F.S. Magnago³, Erica Rievrs Borges¹, Eduardo van den Berg 2, Fabrício Alvim Carvalho1

Instituto de Ciências Biológicas, Universidade Federal de Juiz de Fora, Campus Universitário, Zip Code

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36036-900, Juiz de Fora, Brazil

Instituto de Biologia , Departamento de Ecologia e Conservação, Universidade Federal de Lavras (UFLA),

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Zip Code 37200-00, Lavras, Brazil

Centro de Formação em Ciências Agroflorestais, Universidade Federal do Sul da Bahia (UFSB), Campus

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Jorge Amado, Zip Code 45604-811, Ilhéus, Brazil

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*Correspondence author. Email: [email protected] Telephone: (+55 32) 988509396

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Marcela V. Pyles ORCID: 0000-0002-4794-125X

ABSTRACT

With urbanisation rapidly transforming our world, a better understanding of how urban ecosystems can contribute to biodiversity conservation is urgently needed. Here, we investigated the functional composition and diversity of plant assemblages in different types of urban forests with the intent to predict the stability of the functions provided by these forests through the functional redundancy and response diversity of species.

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For this, we used data from non-urban mature forests, remnant urban forests, urban forests regenerated from croplands and urban forests regenerated from soil denudation, all located in the Brazilian Atlantic Forest. Our results showed that species functional composition shifted among urban and non-urban forests. However, functional diversity was lost only in urban forests with some previous use, indicating that the natural regeneration of urban forests after previous land use results in functionally poor forests. Urban forests regenerated from soil denudation (more intense disturbance) had a lower functional diversity. Conversely, urban forests regenerated from cropland (less intense disturbance) showed only a reduction in the number of functions

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provided. Our analysis also revealed the reduced functional redundancy and response diversity of urban forests with previous land use, which may indicate a greater vulnerability of the functions provided by these forests.

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Thus, we emphasise in this study the importance of land use history for decision making in urban forest conservation policies and highlight the crucial role of natural remnant urban forests as reservoirs of functional

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diversity and stability in highly degraded and fragmented landscapes.

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Keywords: urbanisation, land use change, functional redundancy, response diversity, tropical forests

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INTRODUCTION

Urbanisation is rapidly transforming our world. The establishment and expansion of urban areas causes deep,

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and likely permanent, transformations at the landscape level and shifts in habitats, biophysical processes and biogeochemical cycles (Grimm et al., 2000). These changes, in turn, affect biodiversity (Seto et al., 2012) and threaten the delivery of essential human health and well-being ecosystem services, including climate regulation, erosion prevention and production of food and fiber (Brack, 2002; Escobedo et al., 2011; Li et al., 2013). As urban areas are expanding on average twice as fast as their populations (Angel et al., 2011; Seto et al., 2012),

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understanding how urbanisation can affect ecological functions is critical to global biodiversity conservation (Aronson et al., 2017) and urgent to the development of tools that support such expansion along sustainable paths (Pickett et al., 2014).

In urban environments, biodiversity is affected by myriad drivers: landscape configuration (i.e., size, shape, patch connectivity), biotic interactions, land use history, human population density of the surrounding urban matrix, economic inputs and management activities (Aronson et al., 2017). All these factors differentially 2

alter habitat quality and, consequently, the composition, diversity and functionality of species (Chapin et al., 1997). Furthermore, their interaction challenges the widespread effects of urbanisation on biodiversity (Williams et al., 2015). For instance, urban ecosystems are, for the most part, either natural remnant forests or forests that regenerate naturally after the abandonment of some human activity, such as wood exploitation, fire suppression, agricultural cultivation or earthmoving (Kowarik and Lippe, 2018). The different environmental conditions left by urban expansion (e.g., high light incidence and edge density) or by land abandonment (e.g., initial floristic composition and soil degradation) lead the urban forests towards alternative successional

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trajectories in which the assembly of the plant communities tends to be divergent (e.g., Swan et al., 2016)

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From a broad perspective, the biodiversity of urban forests is marked by the extinction of specialised native species (Bierwagen, 2007; Schleicher et al., 2011) and a large influx of locally exotic species (Essl et al.,

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2011; Kowarik, 2011). These phenomena occur because not all species respond equally to urbanisation-driven changes. Subsequently, the anthropogenic environmental conditions filter out species that are not adapted while

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creating new habitat opportunities for new species and plant communities (Godefroid and Koedam, 2007; Corlett, 2015). Thus, the species similarity tends to increase through space over time, inducing a biotic

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homogenisation of urban biodiversity (see La Sorte et al. 2014; Yang et al. 2015; Zeeman et al. 2017), which may have important consequences for the provision and stability of ecosystem functioning (Mangels et al.,

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2017).

With ecosystem functions being restricted to the ones provided by functionally similar species, there is an imminent risk of species succumbing at the same time, eliminating specific ecosystem functions and services altogether (Mori et al., 2013). Functionally similar species tend to also be similar in their responses and

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susceptibility to disturbances (Mori et al., 2013). However, if they respond differently to environmental changes, it seems that the presence of functionally similar species may provide a certain safety net for the maintenance of the ecosystem processes in case of local species extinction (Elmqvist et al., 2003). Thus, ecosystem functioning stability or ecosystem functioning continuity is conditioned to the presence of two key mechanisms: functional redundancy (species numbers that contribute similarly to an ecosystem function, Laliberté et al., 2010) and response diversity (different responses of redundant species to disturbances and environmental changes, Elmqvist et al., 2003). As in urban areas, the pressure for land use change is constant

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(Colding, 2007), and the risk of species extinction is high (Hansen et al., 2005). Thus, knowledge about the forest responses to future disturbances is especially important (Pickett et al. 2016) because it can prevent loss of important traits and functional diversity and guarantee the permanence of ecosystem functions and services in these ecosystems (Oliver et al. 2015).

Here, we investigated how functional composition and diversity of forest fragments are affected by urbanisation and by urbanisation plus previous land use (previous cropland and previous soil denudation). The

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anthropogenic disturbances present in urban areas are considered to be ‘press’ (or chronic) disturbances and they tend to persist or increase in magnitude and frequency over time (Nimmo et al., 2015). Thus, through

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functional redundancy and response diversity, we also aimed to suggest about the continuity (stability) of ecosystem functions provided by these forest fragments. While functional redundancy is related to functional

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stability because the presence of species that play the same role in the ecosystem functioning ensures its continuity in the case of species loss (Yachi and Loreau, 1999), the response diversity is related to functional

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stability because the greater the variation of responses, the greater the security of renewal and reorganisation of the ecosystem (Elmqvist et al., 2003). We addressed two questions: (i) to what extent do urban forests with and

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without previous anthropic use differ from non-urban forests in functional composition and diversity, and (ii) are the functions provided by urban forests less stable than those provided by non-urban forests? We

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hypothesised that urban forests should present lower functional diversity indices and shifts in functional composition towards a prevalence of species with a high colonisation ability (abiotic dispersion and small seed size). Additionally, due to the extensive edge areas affected by edge effects (urban expansion) and previous land use, we expected to find an increase of species with typical early-successional traits (shorter stature and wood density and compound leaves) in urban forests. These shifts should result in urban forests that are more

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functionally homogeneous. Moreover, due to this greater homogeneity, we predicted that, although the urban forests should present more species that play the same role in the ecosystem (greater functional redundancy), the species should also be more similar to each other with regards to their response traits (less response diversity), and this characteristic would reduce the functional stability of these forests.

MATERIAL AND METHODS

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Study sites and data sampling

This study was conducted in 12 Atlantic Forest fragments located in state of Minas Gerais, Southeastern Brazil, in the municipalities of Juiz de Fora, Lima Duarte, Rio Preto and Santos Dumont (21°24' to 22°1'S and 43°18' to 43°55'W; Table S1). The region experiences a mesothermic climate, characterised by dry winters and temperate summers (Cwb – Köppen Classification; Alvares et al., 2013). The annual mean precipitation ranges from 1,497 to 1,585 mm, and the mean annual temperature ranges from 17.6 to 18.9°C (Alvares et al., 2013).

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The original vegetation of the region is classified as Montane Semideciduous Seasonal Forests (IBGE, 2012) and belongs to the Brazilian Atlantic Forest domain. The predominant soil in the region is latosol (Oxisoils;

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CETEC et al., 2010). Despite different previous use (see below), all fragments experience the same

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environmental conditions.

In this study, we sought to encompass the main types of urban ecosystems: natural remnant forests and

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forests that regenerated after previous use (Table S1). Thus, we chose forests with different land use histories and classified them as follows: (i) natural remnant urban forests (REM), forests that had partial suppression (>

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80 years) promoted by urbanisation expansion, but without any known previous land use (N = 3); (ii) urban forests regenerated from croplands (CROP), forests that regenerated naturally (~70 to 80 years) after complete abandonment of crop use (coffee and/or pasture; N = 3); (iii) urban forests regenerated from soil denudation

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(DEN), forests that regenerated naturally after complete vegetation and soil removal (~60 years; N = 3). For purposes of comparison, we also chose three mature forests outside the urban matrix that have no known history of previous use or disturbance (Non-urban, > 80 years, N = 3). The forest classification was determined through satellite images and interviews with residents. In all forests, the plots were established in the core areas (i.e., ≥

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30 m from the forest edge).

The data were collected from 2013 to 2016 by several authors. In each forest, 10 permanent plots (20

m × 20 m) were randomly assigned, yielding a total of 120 sample plots (4.8 ha in total). During the sampling (which was performed at different time periods), all live trees with a diameter at breast height (DBH) ≥ 4.8 cm were tagged, identified to the species level and had their DBH measured and height estimated. Our database comprised 7,202 trees that belong to 383 species and 68 botanical families.

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Functional traits

The following functional traits were selected due to their association with the processes of natural successional trajectory and recovery from past disturbance (Van Der Sande, et al. 2016): maximum height (Hmax, m), which is an indicator of the adult stature of a species and is related to the species longevity and life-history strategy (King et al. 2006); wood density (WD, g.cm-3), which is an indicator of stem construction costs, stability and hydraulic conductivity (Poorter et al. 2010); leaf compoundness (LC, 0 = simple and 1 = compound), which

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reflects the species evaporation and heat balance; seed size (SS, 0 = small and 1 = large), which, although usually related to the competitive vigour of the seedlings (Kitagima, 2007), is also an important life history trait

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for trees, correlated to a suite of morphological and physiological traits of pioneer species (small seeds) and shade-tolerant species (large seeds; Poorter and Rose, 2005; Osuri and Sankaran, 2016); and dispersal mode

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(DM, 0 = abiotic and 1 = biotic), which is an indicator of the ability of plants to colonise habitats and is especially important in fragmented urban landscapes because they can improve predictions of dispersal

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probability and seed bank composition (Kraft et al., 2015).

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The continuous traits (Hmax and WD) were measured following standardised protocols (Cornelissen et al., 2003; Pérez-Harguindeguy et al., 2013), and the binary traits were classified from field observations, local informants, herbaria and the literature. The species maximum height was calculated as the 95th-percentile

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height of all trees of the species. Species WD was obtained from the Global Wood Density database (filtered by Tropical South America; Zanne et al., 2009). For the species with no available WD, we used the genera or family WD mean values. The species SS was classified as small when the length was ≤ 1.5 cm and large when the length was ≥ 1.6 cm (Santos et al. 2008). Details about the trait values for each species are presented in

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supplementary material (Table S2).

Response and effects traits

The functional traits were classified as response traits when shifts in environmental conditions promote changes in these traits (or in the value) and as effects traits when the changes in these traits (or in the value) cause effects on environmental conditions or ecosystem functioning (Lavorel and Garnier, 2002; Cornelissen et al., 2003). This division is challenging but important because different types of environmental change (e.g., land use 6

change) can cause shifts in biodiversity (as response) that in turn affect ecosystem functioning (as effect; Suding et al., 2008).

According to Lavorel and Garnier (2002), traits that influence biogeochemical processes are effect traits and traits that influence regenerative processes are response traits. Thus, Hmax, WD and LC are effect traits, while SS and DM are response traits (e.g., Laliberté et al., 2010). However, several studies demonstrated that plant height and WD can be related to competition and they generally increase during succession (e.g.,

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Prach et al., 1997; Kahmen and Poschold, 2004). Yet, response to succession has also been reported for leaf traits associated with resource economy (Prach et al., 1997; Díaz and Cabido 2001; Kahmen and Poschold,

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2004; Louault et al., 2005). Thus, we decided to classify Hmax, WD and LC as both response and effect traits

Functional composition and functional diversity

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and SS and DM as response traits.

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Functional composition was assessed and discussed through functional groups (FGs), which are sets of species with similar functional trait values (Díaz and Cabido, 2001). Here, we utilised all traits (response and effect) to

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shape the FGs because our aim was to explore the community assembly mechanisms and ecosystem functioning of urban forests (e.g., Mumme et al., 2015). First, to calculate the dissimilarity matrix between species, we used

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a generalisation of Gower’s distance that allows mixed trait types (e.g., continuous, ordinal and categorical; Pavoine et al., 2009). Using a dendrogram of Ward’s clustering method from the ‘FD’ package in R (Laliberté et al., 2015), we determined the number of groups by visual inspection. The consistency of each species’ membership to its assigned FG was validated by permutational multivariate analysis of variance (PERMANOVA; Anderson, 2001). Nonmetric multidimensional scaling ordination (NMDS) was also used to

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visualise the separation of FGs.

To analyse functional diversity we used two indices: functional richness (FRic) and functional

dispersion (FDis). FRic is an indicator of the species volume that occupies the niche space of a community (Villéger et al., 2008), and FDis is an indicator of species distribution in the niche space weighted by abundance (Laliberte et al., 2010). High FRic values are associated with communities that have large numbers of functions,

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while high FDis values are associated with a high degree of differentiation between species that perform such functions.

Functional stability

To assess functional stability we used two measures: functional redundancy (FR) and response diversity (RD; Mumme et al., 2015). FR is an indicator of species numbers that contribute similarly to an ecosystem function (Laliberté et al., 2010) and is related to functional stability because the presence of species that play the same

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role in the ecosystem functioning ensures its continuity in the case of species loss (Yachi and Loreau, 1999).

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RD is an indicator of how redundant species present different responses to disturbances and environmental changes (Elmqvist et al., 2003). It is related to functional stability because the greater the variation of responses,

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the greater the security of renewal and reorganisation of the ecosystem (Elmqvist et al., 2003).

First, we identified groups of functionally redundant tree species using a set of functional effects traits.

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Then, we quantified RD within these FGs by measuring FDis, based on response traits. Redundant groups were identified through a Ward´s clustering method based on the effect-trait dissimilarity matrix estimated by the

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Gower distance. With the consistency of each species’ membership to its assigned functional effect group validated by PERMANOVA (Anderson, 2001), the FR indices were calculated as the ratio between the species

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richness (S) and the amount of functional effect groups (Laliberté et al., 2010). FDis represents RD by reflecting the functional differences among the species in a community (Craven et al., 2016). Thus, through dispersion variation among species that belong to the same FG, we assessed how different these species were in terms of

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functional response (traits), even if they play the same role in the ecosystem.

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Statistical analysis

We used generalised linear mixed models (GLMM) to assess the functional differences among the non-urban and urban forests (REM, CROP and DEN). The significance threshold was set at < 0.05. The site was considered to be a random factor, to account for the possible lack of independence between plots within the sites (Bates et al., 2014). Beta-binomial models corrected by the observation-level method were performed for modelling overdispersion of FG relative abundances (i.e., proportion data; Harrison, 2014). A Gaussian error distribution

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was used for functional diversity and RD indices analyses (normality of the residuals was confirmed by the

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Shapiro–Wilk test).

All analyses, figures and graphs were performed and generated using the R platform (R-Core-Team,

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2017) and the following packages: Vegan (Oksanen et al., 2016), multcomp (Bretz et al., 2015), lme4 (Bates et

RESULTS

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Functional composition and diversity

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al., 2014), lmerTest (Kuznetsova et al., 2016), MuMIn (Barton, 2016) and ggplot2 (Wickham and Chang, 2016).

A total of 383 tree species in 7,202 individuals were recorded and then grouped into 11 functional groups (Figs.1

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and S1) with a significant separation (PERMANOVA r² = 0.30-0.94, P = 0.002-0.001; Figs. S2 and S3). In urban forests regenerated from soil denudation, seven functional groups showed smaller relative abundances (FG1, FG2, FG4, FG5, FG8, FG9 and FG11) and two had a higher relative abundance in comparison with the other forest categories (FG6 and FG10). In urban forests regenerated from croplands, five FGs had a lower

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relative abundance (FG2, FG4, FG5, FG9 and FG10) and two showed higher relative abundance (FG3 and FG6). Natural remnant urban forests had one FG with a decreased relative abundance (FG8) and one FG with an increased relative abundance (FG6). One group (FG7) showed no significant changes in the relative abundance in any urban forests (Fig. 2 and Table S4).

No specific trait was shared by species of the functional groups that decreased its relative abundance in urban forests (Fig. 2 and Table 1). Among the three groups that showed proportion increases in some of the urban forests, two traits were shared by the species: taller stature than the average of all sampled individuals 9

and small seeds (Fig. 2 and Table 1). In FG3, the species also had lower-than-average wood density, simple leaves and biotic seed dispersal. Approximately 40% of all individuals in urban forests regenerated from croplands were contained in this group. FG6 was the only group that had higher a proportion of individuals in all urban forests when compared to the non-urban forests. In addition to tall stature and small seeds, this group was formed by species with higher-than-average wood density, compound leaves and abiotic dispersion. Curiously, 89% of the species and 95% of the individuals of this group belong to Fabaceae botanical family (Table S5). Finally, FG10, with 31 species (~7% of the total sampled), accounted for almost 72% of all

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individuals found in urban forests regenerated from soil denudation. This finding suggests that these forests are dominated by species with similar characteristics: higher-than-average height and wood density, simple leaves,

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abiotic dispersion and small seeds. In this group, two species were dominant: the exotic Pinus elliottii Engelm

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(~37% of individuals) and Eremanthus erythropappus (DC.) MacLeish (~32% of individuals; Table S5).

The GLMMs showed that FRic was negatively affected by the previous land use of urban forests (Fig.

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3A and Table S7). Urban forests regenerated from soil denudation differed significantly from non-urban forests and showed extremely low FRic values (t = -6.30, p < 0.001). Urban forests regenerated from croplands also

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presented significantly negative effects on forest FRic (t = -2.24, p = 0.044), but less than the urban forests regenerated from soil denudation. Natural remnant urban forests did not present significant differences in FRic

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indices when compared with non-urban forests (t = -0.49, p = 0.629). In relation to the FDis index, only urban forests regenerated from soil denudation showed values that were statistically different from those found in non-urban forests (Fig. 3B). In these forests, the FDis index was statistically lower (t = -2.30, p = 0.040). CROP (t = -0.82, p = 0.423) and REM (t = 0.125, p = 0.902) urban forests were not significantly different from non-

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urban forests.

Functional stability

The clustering method shaped six different functional effect groups (Figs. S3 and S4) with significantly intergroup dissimilarity (PERMANOVA r² = 0.30-0.92, p = 0.001; Fig. S5). The FR indices showed significantly lower values in urban forests with previous land use when compared to non-urban forests (Fig. 4A). Urban forests regenerated from soil denudation showed the lowest values (t = -8.18, p < 0.001), followed

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by urban forests regenerated from croplands (t = -4.41, p < 0.001). Natural remnant urban forests did not present statistical difference in relation to non-urban forests (t = -1.67, p = 0.119). RD indices also showed statistically significant differences, but only between the non-urban forests and the urban forests with some previous use (cropland and soil denudation; Fig. 4B). The lowest values were found in urban forests regenerated from soil denudation (t = -6.802, p < 0.001), followed by urban forests regenerated from croplands (t = -3.705, p = 0.003). Natural remnant urban forests did not present a statistical difference in the RD index when compared to non-

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urban forests (t = -0.939, p = 0.366).

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DISCUSSION

Functional shifts in plant assemblages in urban forests

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The urban forests showed different tree functional trait compositions when compared to non-urban forests. However, these shifts did not follow a single path as hypothesised (greater abundance of species with early-

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successional traits or high colonisation ability). The only traits shared by functional groups that were increased in urban forests were high maximum height and small seed size. Small seeds guarantee a colonisation advantage

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because they are produced in large quantities and can persist for longer in the soil seed bank to germinate under favourable conditions (Poorter and Rose, 2005). However, the higher maximum height could be taken as an

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unexpected tree investment in urban forests, where lower light is expected due to both the "hard edges", typical of urban environments (see Forman, 2014), and the heterogeneity in light availability found in forests at secondary succession stages (Lebrija-Trejos et al. 2011). Nevertheless, other studies report an increase in the proportion of high-stature trees in urban forests (Williams et al., 2015; Guerra et al., 2017), and it is likely that the greater probability of local extinction of short plants (Preston, 2000) and the longer life span of taller trees

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(Duncan and Young, 2000) may explain the success of tall-statured plant species in urban forests (Williams et al., 2015).

The proportion of each functional group varied substantially among urban forests, results that demonstrate the high functional dissimilarity among them. Even under the same environmental conditions (i.e rainfall and temperature regime and surrounding landscape matrix), differences previous land use can lead to different forest successional pathways (Arroyo-Rodríguez et al., 2013), and this may explain this dissimilarity 11

in functional composition. The group formed by species with traits that allow a greater ability to colonise and survive in disturbed areas was the only group that achieved ecological success (that is, increased relative abundance) in all urban forests. Seed dispersion by abiotic vectors is a recognised advantage in urban forests given the survival limitations imposed by urban structures and conditions on some groups of animals, including birds (Rottenborn, 1999) and mammals (Villaseñor et al., 2014). Compound leaves can prevent excessive water loss and regulate temperature through the movement of leaves and leaflets, a particularly important strategy for dealing with the high incidence of solar radiation (Wright et al., 2017). Finally, the higher wood density found

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in urban forests guarantees trunk resistance to physical damage by herbivores, pathogens and woody debris (Van Gelder et al., 2006; McCarthy-Neumann and Kobe, 2008) and can be indicative of hydraulic safety (Sterck

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et al., 2006). Furthermore, FG6 is formed predominantly by species of the family Fabaceae, which may fix nitrogen from the atmosphere. The local disturbance history and prior land use determines (and often limits)

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nitrogen availability (Powers and Mar, 2017; Winbourne et al., 2018).Considering the differences in the potential of nitrogen fixing genera by Fabaceae subfamilies (62% Papilionoideae, 54% Mimosoideae and 5%

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Shifts in functional diversity in urban forests

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Caesalpinioideae; Hedin et al., 2009), approximately 45% of FG6 can fix N 2.

The dominance of certain functional groups is reflected in the reduced functional diversity of urban forests

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regenerated from soil denudation and is related to the strong environmental filtering caused by land use conversion (Cornwell et al., 2006; Flynn et al., 2009; Lohbeck et al., 2014). The high abiotic filters (i.e., high light incidence, exposed soil) associated with the abandonment of post-use areas, along with the urban matrix that adds dispersion filters (i.e., great distance between forest fragments), establishes adverse environmental

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conditions that drive the community assembly through non-random patterns of species co-occurrence (Götzenberger et al., 2012). The strong environmental filters in these forests can selectively remove species that do not share adaptations (Naeem and Wright, 2003), a phenomenon that restricts forest recolonisation, survival or regeneration of species with specific and similar traits (Lebrija-trejos et al., 2010). This factor explains the decreases in functional richness and functional dispersion, respectively.

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The urban forests regenerated from croplands also presented loss of functional richness, an expected finding due to the previous land-use conversion (Tscharntke et al., 2005). However, in these forests, the functional dispersion indices were not reduced. These data indicate that although they provide fewer functions, there is some functional dissimilarity between the co-existing species. Several studies demonstrated that even near abandoned fields with the same fallow age, soil type and climatic conditions do not necessarily follow a single path of succession (Chazdon et al., 2007; Lebrija-trejos et al., 2010; Norden et al., 2011). Here, the

abandonment conditions probably prevented the similarity of traits in these habitats.

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distinct initial composition of these forests (e.g., Coffea arabica and Brachiaria spp.) and different

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Contrary to our hypothesis, the natural remnant urban forests did not show a lower functional dispersion when compared with non-urban forests. The 80 years of urbanisation around the forests did not

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apparently alter the functional volume occupied by species or the distribution of species within the community. In fact, the recognised direct impacts of urbanisation on species diversity, such as habitat loss, habitat

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fragmentation and the introduction of new species (Kowarik, 2011), did not visibly occur in these forests. This finding confirms the relative importance of natural remnants in comparison to other types of urban ecosystems

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found in previous studies (Schwartz et al., 2013; Aronson et al., 2014; Kowarik and von der Lippe, 2018) and highlights the role of these forests as the main reservoirs of functional diversity and potential ecosystem

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services.

We are aware that there are other potential drivers of community functional dispersion and composition that were not tested in this particularly study, especially concerning size and shape of the fragments. A recent study (Borges et al. 2020) measured the landscape configuration of these forests and found

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that only total area and perimeter-to-area ratio are significantly different between the denuded forests and other forest categories. Therefore, the results found for the forests with soil denudation might also be related to their significantly smaller sizes, which impose additional filters related to a higher exposition to edge effects and leads to a community dominated by functionally similar species unable to sustain ecological processes (Santos et al., 2008).

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Functional stability

Urban forests with previous land-use exhibited significant reductions in both functional redundancy (the number of species within functional effect groups) and response diversity (the dispersion of traits among effect group members). The relatively low functional redundancy and richness of urban forests with cropland and denudation histories is typical of species-poor assemblages (Díaz and Cabido, 2001; Petchey et al., 2007) and may have dramatic consequences on ecosystem function, especially when the range of species reactions to

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environmental change is low (i.e., low response diversity; Elmqvist et al., 2003). When species with similar functional effect traits respond similarly to environmental conditions and changes, even small disturbances can

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lead to a major loss and complete extinguishment of certain ecosystem functions (Elmqvist et al., 2003).

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Conversely, the natural remnant urban forests maintained the same levels of functional redundancy and response diversity of non-urban forests. Thus, in natural remnant urban forests, the possible events of

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random loss of one species may not affect ecosystem functioning, as its function can be compensated by others within the same functional effect group (see Fonseca and Ganade, 2001; Bruno et al., 2016). Moreover, response

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diversity should represent another mechanism that avoids the elimination of most species of a given functional group by disturbances (Mori et al., 2013). The conservation of ecosystem functions or their stability is most effective when functional redundancy and response diversity operate together (Stavert et al., 2017). Thus, the

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functions provided by remnant urban forests appear to be safe.

CONCLUDING REMARKS ON THE FUNCTIONALITY OF URBAN FORESTS

This study provides a substantial understanding on the influence of urbanisation and land use history on the

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functional composition and diversity of tropical tree communities. Moreover, it highlights the need for including information on land use history to fully understand the ecology of urban forests. Here, we assumed that ecosystem services and functions are positively correlated with functional diversity. Based on our findings, we have some considerations to ensure continuity in the provision of ecosystem functions and services by urban forests: (i) urban remnants should be a priority in city plans, because these forests support many ecosystem functions of urban environments; (ii) as the natural regeneration of urban forests after previous land use has resulted in functionally poor forests, species enrichment of some functions may increase the variety of 14

ecosystem processes, and, over time, it is possible that these forests follow its development toward more natural ecosystems. Still, we revealed that a specific functional group of Fabaceae family are well adapted to urban environments under different environmental conditions and history. Thus, it is likely that restoration strategies that include these species may be safer with regards to the species survival in urban environments.

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Conflict Of Interest

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The authors declare no conflict of interest in manuscript "Land use history drives differences in functional composition and losses in functional diversity and stability of Neotropical

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urban forests " (Ref. UFUG_2019_636).

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APPENDICES

Table 1: Functional group description based on distinctive functional traits S

Hmax

WD

LC

DM

SS

FG1

10

14.84

0.59

Compound

Abiotic

Small

FG2

30

11.42

0.62

Simple

Biotic

Small

FG3

58

16.74

0.52

Simple

Biotic

Small

FG4

44

14.36

0.57

Compound

Biotic

Small

FG5

66

8.27

0.63

Simple

Biotic

Small

FG6

29

14.08

0.71

Compound

Abiotic

Small

FG7

20

13.29

0.68

Compound

Abiotic

Large

FG8

30

13.70

0.72

Simple

Biotic

Large

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FG

22

FG9

14

21.16

0.72

Simple

Abiotic

Large

FG10

31

13.85

0.64

Simple

Abiotic

Small

FG11

51

15.85

0.81

Simple

Biotic

Small

FG: Functional groups, S: species richness, Hmax: Maximum height, WD: Wood density, LC: Leaf compoundness, DM: Dispersal mode and SS: Seed size. Continuous traits shown according to the difference between the mean values found within species of each group and the mean value found for all species. Species

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Hmax mean = 13.70 (m) and Species WD mean = 0.64 (gcm-3).

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Fig. 1: NMDS ordination plot showing the compositional dissimilarity of tree species based on their traits

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combinations. Each point represents one species, with 95% confidence ellipses for each of functional group.

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Fig. 2: Histogram of functional groups abundance relative by forest category. "*" shows a significant difference in relation to non-urban forests (p-value <0.05). Non-urban: Non-urban forests; REM: Remnant urban forests;

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CROP: Urban forests regenerated from cropland use; DEN: Urban forests regenerated from soil denudation.

Fig. 3: Histograms show the functional richness (A) and functional dispersion (B) across forest categories. Error bars represent the 95% of confidence intervals with n=30. "*" P< 0.05, "**"P<0.01 and "***"P < 0.001.

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Non-urban: Non-urban forests; REM: Remnant urban forests; CROP: Urban forests regenerated from cropland

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use; DEN: Urban forests regenerated from soil denudation.

Fig. 4: Histograms show the functional redundancy (A) and response diversity (B) across forest categories.

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Error bars represent the 95% of confidence intervals with n = 30."*"P< 0.05,"**"P<0.01 and "***"P < 0.001. Non-urban: Non-urban forests; REM: Remnant urban forests; CROP: Urban forests regenerated from cropland

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use; DEN: Urban forests regenerated from soil denudation.

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