Local and landscape drivers of tropical forest regeneration in agricultural landscapes of the Atlantic Forest of Brazil

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(1)1. University of São Paulo Luiz de Queiroz College of agriculture. Local and landscape drivers of tropical forest regeneration in agricultural landscapes of the Atlantic Forest of Brazil. Ricardo Gomes César. Thesis presented to obtain the degree of Doctor in Sciences. Area: Forest Resources. Option in: Conservation of Forest Ecosystems. Piracicaba 2018.

(2) Ricardo Gomes César Forester. Local and landscape drivers of tropical forest regeneration in agricultural landscapes of the Atlantic Forest of Brazil versão revisada de acordo com a resolução CoPGr 6018 de 2011. Advisor: Prof. Dr. PEDRO HENRIQUE SANTIN BRANCALION. Thesis presented to obtain the degree of Doctor in Sciences. Area: Forest Resources. Option in: Conservation of Forest Ecosystems. Piracicaba 2018.

(3) 2. Dados Internacionais de Catalogação na Publicação DIVISÃO DE BIBLIOTECA – DIBD/ESALQ/USP. César, Ricardo Gomes Chronossequence and landscape effect in tropical forest succession / Ricardo Gomes César. - - versão revisada de acordo com a resolução CoPGr 6018 de 2011. - - Piracicaba, 2018. 165 p. Tese (Doutorado) - - USP / Escola Superior de Agricultura “Luiz de Queiroz”. 1. Florestas secundárias 2. Restauração florestal 3. Ecologia de paisagens 4. Regeneração natural I. Título.

(4) 3. ACKNOWLEDGEMENTS. Conmigo vienen, vienen los de atras! - Calle 13 There are many, many people behind the dozens of pages and four years of this work. The name of a few of them should be in the cover, along with me and my tutors. However, this would violate scientific writing conventions and would lead to a long and awkward discussion about the restrictions of scientific publications and the relevant people for the development of science. In other words, a discussion that I would certainly lose. Besides, I already violated these conventions several times before getting to this final version so, to avoid further hassle for the library staff, mentioning these people that are very dear to me just in this section will have to do. Regarding gratitude, the first time I felt it was for my family, so it would be fair to start from them. They were also the first to hear me talking about nature, science, the universe and everything else. For caring, hearing, inspiring, laughing and arguing (and not sleeping a single time while I was explaining my thesis), I thank my parents Francisco Ignácio Giocondo César and Marli Valverde Gomes, my sister Sofia Gomes César and my beloved Glaucia Zaina Gonsalves. I am also thankful for my grandparents Ricardo Gomes Filho and Antonieta Valverde Gomes for the weekends in the country house and for the first books about nature. I was also inspired by my grandfather and forester Geraldo de Barros Cesar, which I would like to have known better. Finally, I am thankful for my aunts Marisa and Magali Valverde Gomes, for being always there supporting me! Glaucia, in a 14 billion-old universe that extends to who-knows-where, with millions of species in this 210 million km² planet, where seven billion people live in thousands of cities, you make the best place to be at your side, and the best time now. I hope that the reader notices that this work was masterfully guided mainly by two people that pruned hedges, opened doors, pointed paths and donated themselves. For better or worse, all graduate students have a tutors. I was lucky to have tutors that became friends along this work. Every good tutor questions, research, write and apply. But you do all of this smiling and looking in the eyes of people. I have so much to say to both of you, that I prefer to keep it in simple words that encompass a lot: thank you Pedro and Robin! Your love move forests! (To Pedro this acknowledgement extends to the end of my undergraduate studies, when we first worked together). These pages look clean, but they hide a lot of sweat, dust and cuts from many volunteers that faced the remaining and forgotten secondary forests, where we collected most of our data. I’m very thankful to GEPEM for their help since the early stages of this work. Thank you Bianca Torres, Hellen Pecchi, Kerolin Amarante, Julia Martins, Ana Carolina Yamaguchi and Karen Beneton. I also thank GADE and mainly Leandro Degrandi. I thank Paula Meli, Adriano Adinolfi Ito (best tapioca in the field!), Marina Peluci, Monica Borda Niño, Saulo E. X. F. de Souza, Felipe Brancalion, Carol Giudice Badari, Flávia Garcia Flórido (that endured several wasp stings), Renan Afoacy (that rescued me out of the forest when I injured my eye) and Henrique Sverzut Freire de Andrade. Interested and motivated people like you make this kind of project possible..

(5) 4. I’m super thankful to Vanessa Moreno, a strong ally during data gathering in the field. Thank you for your attention to details, for reviewing our data, helping in almost every field expedition of this work and having the energy to sample even more forests. Thank you for your help in the ups and downs, stresses and laughs, ticks and mud of the field work. And for the brainstorm to guess if it will rain or not. I am very thankful for the help of Alex Mendes: the one that never spoiled Game of Thrones, the watcher of canopies, guardian of epiphytes, the one that went North and back, the wise of endless patience. Thank you Alex! Research is always bigger than any single person, therefore I acknowledge for my colleagues that are also investigating the different aspects of second-growth forests in the Corumbataí watershed. Thank you Vanessa Moreno, Alex Mendes, Vanessa Oliveira, Daniella Schweizer, Monica Borda Niño, Danilo Almeida (thanks for the help with statistical analyses in the first chapter!), Juliano van Melis (thank you very very much for all the help in the second chapter and in many other analyses from other projects!), Marina Peluci, Eduardo Alexandrino, Alessandro Palmeira and Paulo Guilherme Molin. I just noticed that, even with so many people collecting data and asking questions about these forests, there is still so much more to collect and ask. These pages also travelled a lot! The foreign collaborators that I had the privilege of working with were Prof. Jos Barlow, Prof. Fernando Espírito-Santo, Alessandro Palmeira and Leighton Reid. Many “obrigados” to all of you. These pages are also stained with laughter, conversations and philosophic discussions. For enchanting the everyday routine, I am thankful to LASTROP: Andréia Moreno, Daniella Schweizer (and Carlos and Lucia), Paula Meli, Andreia Alves Erdmann, Carina Camargo Silva, Juliano van Melis, Daniel Palma Perez Braga (the main source of philosophical discussions), Vanessa Erler Sontag (who will build a time machine someday), Danilo Almeida, Fabrício Hernani Tinto (for the food supply), Frederico Domene, Luciana Maria Papp (who also gets extra points for the food supply), Marina Melo Duarte (for helping eating all the food supply with me), Monica Borda Niño, Nino Tavarez Amazonas, Vanessa Souza Moreno, Luis Eduardo Bernardini, Carol Giudice Badri, Prof. Edson José da Silva Vida and Pedro Henrique Santing Brancalion. For making (much, much) more than their professional requirements in all material logistics, equipment, bureaucracy, keys, resources and conversations; I am very thankful to Andréia Moreno, Giovana Oliveira and Jeferson Polizel. You certainly help moving the wheels of graduate (and undergraduate) activities work better. For contributing to my mental sanity and insanity during all this process, I thank my friends from “Raízes e Asas”, mainly Dante Moretti, Raquel Galvani, Fabio Camolesi, Amarílis Ibanez, Ivo Racca, Bárbara Contarini, Paulo Santini, Jéssica Telhada and Beatriz Abud. Finally, I acknowledge the FAPESP for the funding granted by Processes #2014/14503-7 and 2017/05662-2..

(6) 5. “A minha surpresa é só feita de fatos De sangue nos olhos e lama nos sapatos”. Chico Buarque de Hollanda.

(7) 6. SUMMARY. RESUMO………………………………………………………………………………………………………….7 ABSTRACT……………………………………………………………………………………….........................8 1. INTRODUCTION .............................................................................................................................. 9 2.. EARLY. ECOLOGICAL. OUTCOMES. OF. NATURAL. REGENERATION. AND. TREE. PLANTATIONS FOR RESTORING AGRICULTURAL LANDSCAPES……………………...……..17 2.1.. INTRODUCTION ................................................................................................. 17. 2.2.. METHODS ............................................................................................................ 19. 2.3.. RESULTS ...............................................................................................................24. 2.4.. DISCUSSION .........................................................................................................29. 3. SURROUNDING LAND USE AND FOREST COVER AS MAJOR DRIVERS OF BIOMASS AND TREE. DIVERSITY. RECOVERY. BY. SECOND-GROWTH. TROPICAL. FORESTS. AGRICULTURAL LANDSCAPES………………………………………………………………………..39 3.1.. INTRODUCTION .................................................................................................39. 3.2.. METHODS ............................................................................................................ 41. 3.3.. RESULTS ...............................................................................................................48. 3.4.. DISCUSSION .........................................................................................................52. 4. FINAL CONSIDERATIONS………………………………………………………………………........61 APPENDIXES .......................................................................................................................................65. IN.

(8) 7. RESUMO. Fatores locais e de paisagem sobre a regeneração natural em paisagens agrícolas da Mata Atlântica brasileira. Florestas estabelecidas pelo plantio de mudas de espécies nativas (PL) e por meio do estabelecimento de florestas secundárias pela regeneração natural (FS) são as principais comunidades geradas durante a restauração florestal em larga escala. A escolha dessas estratégias está condicionada potencial de regeneração natural do local, mas tão importante quanto a decisão sobre métodos de restauração, são as diferenças das comunidades que essas escolhas podem gerar. As FS são heterogêneas e, enquanto existe uma literatura crescente dos fatores que afetam a chance do estabelecimento das FS, poucos trabalhos abordam os fatores que influenciam os atributos dessas florestas. Nesse contexto, nosso trabalho busca identificar as diferenças entre PL e FS e as variáveis locais e de paisagem que afetam os atributos das FS. Para tal, amostramos a comunidade arbórea de florestas estacionais semideciduais de Mata Atlântica estabelecidas naturalmente (FS) e por PL em paisagens agrícolas na bacia do Rio Corumbataí, no estado de São Paulo. Observamos que os PL apresentam biomassa semelhante às SF e maior riqueza de espécies. No entanto, as PL também apresentam menor abundância de indivíduos jovens, indivíduos zoocóricos e lianas. A composição de espécies entre essas florestas também difere. As FS estabelecidas em plantios abandonados de eucalipto apresentaram riqueza de espécies e biomassa de espécies nativas semelhantes a outras florestas secundárias. No entanto, os atributos das SF variam consideravelmente. Nesse contexto, as FS apresentam elevado potencial de provimento de alimento para a fauna e estocagem de carbono de maneira custo-eficiente, enquanto que as PL podem ter sua permanência em longo prazo comprometida pela falta de indivíduos jovens. Em seguida, investigamos as variáveis que direcionam a heterogeneidade observada nas FS utilizando modelos mistos lineares generalizados para estimar a influência de variáveis locais e de paisagem na biomassa, densidade de espécies, área basal de árvores zoocóricas e estrutura filogenética das FS amostradas. Plantios de cana-de-açúcar próximos as FS reduzem a biomassa e área basal de indivíduos zoocóricos, enquanto que a cobertura florestal da paisagem aumentou a densidade de espécies e a diversidade filogenética. A idade da floresta apresentou importância secundária ou nula para os atributos estudados. Nossos resultados ressaltam a importância de práticas agrícolas que minimizem os danos em florestas próximas e de mecanismos que favoreçam a cobertura florestal nativa em paisagens agrícolas, a fim de fomentar o potencial dessas florestas em prover serviços ecossistêmicos e conservar a biodiversidade. A escolha entre facilitação do estabelecimento de FS ou PL visando a restauração florestal está condicionada ao contexto local e de paisagem onde serão realizadas as ações de restauração. Apesar de ambas as abordagens apresentarem potencial para cumprir os objetivos dos projetos de restauração, atenção especial deve ser dada ao recrutamento de novos indivíduos para manter a sustentabilidade de PL, enquanto que práticas agrícolas menos impactantes e paisagens agrícolas com maior cobertura florestal nativa podem aumentar o potencial de SF em prover serviços e conservar a biodiversidade. Palavras-chave: Florestas secundárias; Restauração florestal; Ecologia de paisagens; Regeneração natural.

(9) 8. ABSTRACT. Local and landscape drivers of tropical forest regeneration in agricultural landscapes of the Atlantic Forest of Brazil. Forests established through native seedling planting (PL) and the establishment of secondary forests through natural regeneration (SF) are the main outcomes of large scale forest restoration. The decision making process of these approaches is conditioned by resilience. But the different outcomes of these approaches are as important as the decision making. SF are heterogeneous and - although there is a growing literature of the drivers of forest establishment – few works analyzed drivers of attributes of these recently established forests. In this context, our work aims to identify the differences between PL and SF and the local and landscape variables that affect SF attributes. To do so, we sampled the tree community in seasonal semideciduous forests of the Atlantic Forest established naturally (SF) and PL in agricultural landscapes in the Corumbataí Watershed, São Paulo State, Brazil. We observed that PL has similar biomass to SF and higher species richness. However, PL also showed lower abundance of young trees, animal-dispersed trees and lianas. Species composition between PL and SF also differs. SF established in abandoned eucalypt plantings showed species richness and biomass of native species similar to other SF forests. However, SF attributes vary greatly. In this context, SF show a large potential for providing food for fauna and storing carbon in a cost-efficient way. While PL can also provide these benefits, it may have its long-term sustainability compromised by the lack of regenerating trees. We then proceeded to investigate drivers of the heterogeneity observed in SF using generalized linear mixed models to estimate the effect of local and landscape variables on the biomass, species density and basal area of animal-dispersed trees of the SF sampled. SF surrounded by sugarcane plantations had lower biomass and basal area of animaldispersed trees, while native forest cover in the landscape increased species density of SF. Forest age showed little or no importance in predicting SF attributes. These results highlight the importance of low impact agricultural practices and of strategies that increase native forest cover in agricultural landscapes, in order to increase the potential of SF to provide ecosystem services and conserve taxonomic diversity. The choice between establishing PL or fomenting SF for forest restoration is conditioned to the local and landscape context where restoration actions will be carried out. Although both approaches can potentially fulfill the objectives of restoration projects, special attention must be given to the recruitment of new individuals to maintain PL sustainability, while less impacting agricultural practices and more forested agricultural landscapes may increase the SF potential to provide ecosystem services and conserve biodiversity. Keywords: Second-growth forests; Forest restoration; Landscape ecology; Natural regeneration.

(10) 9. 1. INTRODUCTION Scaling up forest restoration is increasingly necessary in order to mitigate the current environmental crisis and reach ambitious international environmental goals (MELO et al., 2013a). Strategies to reach these objectives must evaluate the potential for natural regeneration in areas to be restored and prioritize approaches that establish the desired native communities at the best cost-efficiency possible (HOLL; AIDE, 2011). Interventions for the restoration of ecossystems can be classified in a gradient of human intervention, from abandoning cultivated areas for natural regeneration (low intervention) to high diversity seedling plantings (high intervention). Both approaches are complementary and valuable to fulfill the demands of forest restoration, and their contexts and costs are constantly discussed in the literature (DE GROOT et al., 2013; ZAHAWI; REID; HOLL, 2014). Nevertheless, the communities established by these approaches are the true legacy of restoration interventions and, ultimately, will determine the value of restoration actions for conservation of biodiversity and provision of ecosystem services. Thus, identifiying the differences in the communities established when both approaches were properly employed is key for long-term adaptative management of areas under restoration and to achieve their associated goals. Restoration ecology slowly consolidated as a science in the early works of Aldo Leopold (1949). Currently, the International Society for Ecological Restoration (SER) defines restoration as the “process of assisting the recoveryof an ecosystem that has been degraded, damaged or destroyed”. This definition emphasizes ecological restoration as an intentional human activity, differing it from ecosystem recovery through natural processes without human intervention. The investiment on ecological restoration only justifies itself if it the area restored is able to sustain itself indefinitely through time, an aspect that is just now being considered and analyzed (REID et al., 2017). In Brazil, restoration ecology and ecological restoration have historically developed through several dynamic phases which culminated in our current (but not consolidated) practices of active forest restoration (BELLOTTO et al., 2009). Brazil has the privilege of relying on relatively robust envinromental legislation, such as the Native Vegetation Protection Law (Law 12.651/2012). Such legislation and market pressures have historically acted as leverage mechanisms to foment restoration practices in pre-determined sites, mostly through large-scale plantations (RODRIGUES et al., 2009, 2011). In this context, desired restoration outcomes must be somewhat predictable and acquired as soon as possible. Thus, tree seedling plantings were broadly used near watercourses and on other environmental protection areas to establish forests. There are also many case studies of other restoration techniques, such as nucleation (BECHARA et al., 2016) and direct seedling (COLE et al., 2011) being successfully employed for forest restoration. Forest succession is a subject that has been extensively studied in ecology (CLEMENTS, 1916; PICKETT; CADENASSO; MEINERS, 2008), and was part of the foundation of the scientific investigations in this field, a long time before restoration ecology established as a science (CLEMENTS, 1916). The early works of forest succession focused on the alterations in the composition, structure and function of ecosystems on the long term, with the main objective to describe temporal changes in the plant community and to evaluate if and how these ecosystems attributes are recovered in second-growth forests (CHAZDON, 2014; MARTIN; NEWTON; BULLOCK, 2013; NORDEN et al., 2009). These early stages of successional studies focused only on natural disturbances (i.e. non-human mediated, such as gap openings by falling trees, landslides, underground soil exposure by falling trees and burrowing animals, river floods and dune movement) to explain sucessional processes. At this phase, we theorized that successional changes would ultimately lead to a somewhat predictable and stable climax community given enough time (CLEMENTS, 1916; PHILLIPS, 1934)..

(11) 10. As successional research developed, the predictable and climax-based sucessional paradigm was increasingly questioned as several authors criticized the inflexibility of the stable climax concept and argued in favor of the different possible successional trajectories (CONNELL; SLATYER, 1977; EGLER, 1954; WHITTAKER, 1953). One notable essay that greatly fomented the change in ecology paradigm were the sucessional models based on facilitation, tolerance and inhibition proposed by (CONNELL; SLATYER, 1977). From the 1980s onward the paradigm of forest sucession gradually changed from a convergent, climax-based, stable perspective that did not encompass human disturbance, to a more fluid sucessional perspective that encompassed continuous change that could lead to several trajetories caused by abiotic conditions, species conditions, landscape context and human and natural disturbances. Forest succession research is frequently based on chronossequences, which is the analysis of similar forest formations in the same region, but with different ages (LETCHER; CHAZDON, 2009; WALKER et al., 2010). This method is more commonly used, since the alternative method (direct observation) is more difficult to monitor given the long duration of vegetation change, mainly in old-growth forests. The chronossequences method alone may not be ideal to study ecological succession (WALKER et al., 2010), given that remnants may have undergone distinct past disturbances and are currently subjected to distinct environmental and landscape context. Therefore, the successional pattern observed in one remnant may not occur in other remnants in the same region (ARROYORODRIGUEZ et al., 2015; WALKER et al., 2010). Such restriction highlights the role of several factors, such as land use before forest regeneration, landscape matrix, fragmentation and human disturbances in successional trajectories. Although these factors are considered as an unwanted “noise” for chronossequence studies, they also represent an opportunity to identify factors other than age that affect forest succession. Only more recently factors such as remnant size and connectivity, matrix, dispersal and establishment limitation were considered to determine successional trajectories and the potential of second-growth forests to recover biodiversity and ecosystem services (ARROYO-RODRIGUEZ et al., 2015; CHAZDON, 2014; LÔBO et al., 2011). The conceptual approaches of forest succession considered succession as a process that is more stochastic than deterministic (PICKETT; CADENASSO; MEINERS, 2008). However, recent approaches highlight the importance of considering several factors, including human and landscape factors, to study successional trajectories, in order to differentiate accurately between stochasticity and the factors that are indeed influencing successional pathways (ARROYO-RODRIGUEZ et al., 2015; MESQUITA et al., 2015). Landscape ecology has been showing that most ecological process are scale- and context-dependent, and are not influenced only at the local scale in a hermetic and non-interactive way (CRK et al., 2009). This field of research has been increasingly incorporated in restoration and sucessional studies (CROUZEILLES et al., 2016; SLOAN; GOOSEM; LAURANCE, 2015). Several methods and projects in the field of ecological restoration and restoration ecology act on the landscape level, by establishing, for example, corridors and stepping-stones in order to foment the recovery of biodiversity and ecosystem services of the native vegetation (METZGER; BRANCALION, 2016) Therefore, currently there is a knowledge gap about actions that aim to favor forest recuperation by manipulating the landscape (PEREIRA; DE OLIVEIRA; TOREZAN, 2013). The field of landscape ecology is increasingly used for conservation and restoration purposes due to the effects of fragmentation dynamics in natural habitat patches. Fragmentation is defined as the loss of continuity of natural habitat patches caused by habitat loss. The creation of patches of once continuous natural habitat implies not only forest loss, but also a series of alterations in populations connectivity, alterations in biotic and abiotic conditions and exposure to natural and human disturbances from the edge (know as “edge effect”). The effects of.

(12) 11. fragmentation are vast and have been studied since the coinage of the term in 1980s with still much ground to cover (HAILA, 2002). Some examples of ecosystem alterations caused by fragmentation include alteration in temperature, moisture and wind regimes (ARROYO-RODR??GUEZ et al., 2016; MAGNAGO et al., 2016), tree and liana biomass (MAGNAGO et al., 2016), species composition and distribution, by favoring species more adapted to the new disturbance and environmental regime (SFAIR et al., 2016; TABARELLI; PERES; MELO, 2012) and reduction in species density (ARROYO-RODRÍGUEZ et al., 2012) among others. Additionally, the effect of fragmentation in a given patch are species-specific (DA SILVA; ROSSA-FERES, 2016) and depends on the biome studied and patch surroundings. Nevertheless, less intensive agricultural practices and high native habitat cover can mitigate most of fragmentation effects (ARROYO-RODR??GUEZ et al., 2016; MELO et al., 2013b; VILLARD; METZGER; SAURA, 2014). Despite the negative effects of habitat fragmentantion, the conservation potential of forest patches in agricultural landscapes in Brazil cannot be ignored, as these forests house a great number of native species, many of them not found in conservation units, and are a valuable source of propagules for forest restoration and natural regeneration (CHAZDON et al., 2009a; FARAH et al., 2017). However, many of the remnant species in these forest patches are very rare and their offspring is unable to establish under the new biotic and abiotic regimes of the fragmented habitat, therefore they may be extinct in the future without intervention (ARROYO-RODRÍGUEZ et al., 2013; LÔBO et al., 2011). A global meta-analysis by Meli et al. (2017) showed that, when natural regeneration occurs, it develops in forests similar to actively planted forests for restoration purposes. Chazdon and Guarigata (2016) argues that, overall, forests established without human intervention show less structural development, but similar species density when compared to planted native forests. However, site specific results vary: SHOO et al. (2016) observed that native tree plantings had greater and higher forest cover, more species richness and more wind-dispersed and large-seeded species than naturally established forests in tropical Australia. Working on 10-years-old cloud forests in Equador, Wilson and Rhemtulla (2016) concluded that seedling planting favored natural regneration while forests established without human intervention seemed to be in arrested sucession. Understanding the local and landscape drivers of forest succession could colaborate to seize its beneftis to ecological restoration projects (WALKER; WALKER; HOBBS, 2007). Understanding site-specific potential for natural regeneration could prevent over-interviening in a given restoration site, which would not only increase restoration costs but also may hinder spontaneous vegetation (SAMPAIO; HOLL; SCARIOT, 2007). Tropical forest restoration can be a expensive activity, but these costs are highly variable (from ten to ten thousand dollars per hectare), how much restoration will cost at a given site will depend not only on the environmental and social conditions of the restoration site and project objectives, but also on the selection of cost-effective restoration interventions (BENINI; ADEODATO, 2017; DE GROOT et al., 2013). We highlight that forest establishment without human intervention depends on several biotic and abiotic, local and landscape factors, and should not be used indiscrimately due to its low costs (ARROYO-RODRIGUEZ et al., 2015). Downsides of relying only in forest succession over plantings to increase native forest cover in restoration projects include its impredictability, monitoring for longer time frames, disagreement with legal and market/certification requirements and social perceptions of “bushy” areas as project failures (CHAZDON; GUARIGUATA, 2016; ZAHAWI; REID; HOLL, 2014). As all restoration actions, there is not a “one shoe fits it all” solution and careful diagnostics are needed to assess the better intervention needed at each site (HOLL; AIDE, 2011)..

(13) 12. Sometimes the lines that divide production, spontaneous regeneration and active restoration can be blurred. Mono-specific forest plantings can be used as a first step to restore degraded areas by alleviating environmental conditions for spontaneously regenerating plants, attracting seed dispersers and generating income for landowners in the early stages of restoration (BRANCALION et al., 2012a; JOHNSTONE et al., 2016). If these plantings are managed with reduced impact techniques, they may favor the gradual recover of ecological memory by the gradual accumulation of a seed and seedling bank over time, which may develop quickly once (or even before) commercial plants are removed from plots (JOHNSTONE et al., 2016). Mono-specific plantings accumulate biomass faster than naturally regenerating forests (BONNER; SCHMIDT; SHOO, 2013)but house less (BARLOW et al., 2007)or similar (FONSECA et al., 2009) native woody plant species than naturally regenerating forests. Additionally, commercial forestry plantings may have a buffer effect around native forest remnants, reducind edge effect and disturbance when compared to other land uses, such as pastures (SOUZA et al., 2010). The role of silvicultural plantings to foment regeneration of native species depends on the biome where the planting is carried out, landscape context, species planted, previous land use and the silvicultural management techniques implemented (BROCKERHOFF et al., 2013; LAMB, 1998). Given the wide array of local and landscape, biotic and abiotic and human-mediated and natural drivers that influence the outcomes of restoration initiatives by either active seedling planting or natural regeneration without human intervention, identifying the the effect of these variables on the outcomes of these different approaches is a promising initiative to support biodiversity and restoration projects in agricultural landscapes. In this context, our study aims to compare high-diversity tree plantings with naturally established forests without human facilitation and to identify the drivers of the tree community attributes of the latter in agricultural landscapes of the Atlantic Forest in São Paulo, Brazil. We expect that forests established through tree seedling planting to have higher species richness, more animal-dispersed trees and biomass than forests spontaneously established through natural regeneration. We expect the main drivers of naturally established forests attributes to be surrounding and previous human land use and forest age. The results from this study are particularly relevant in the Brazilian context, in order to monitor restoration projects under the environmental law 12.651/2012 (“Lei de Proteção da Vegetação Nativa”, in Portuguese. They also seize the momentum created by national large-scale restoration programs, such as the PLANAVEG and the planning and priorization for large scale forest landscape restoration in initiaves developed by the private and third sectors such as the Atlantic Forest Restoration Pact. In the global context, we hope to contribute to ambitious international restoration goals such as the Bonn Challenge and the 20x20 Initiative, which will only be achieavable if we make informed decisions on the potential for natural regeneration in restoration sites and increase restoration cost-efficiency. This study is divided in four parts: 1) this overall introduction to the subjects approached; 2) the first scientific manuscript comparing naturally established second-growth forests after different previous land uses and mixed tree plantations; 3) the second manuscript analyzing the factors that drive naturally established second-growth forest attributes in agricultural landscapes; 4) final conclusion..

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(18) 17. 2. Early ecological outcomes of natural regeneration and tree plantations for restoring agricultural landscapes. Copyright by the Ecological Society of America. Expanded version of the manuscript accepted for publication in the journal Ecological Applications at 10/19/2017. Copyright aggreements require that this manuscript must be cited as: César, R. G.; Moreno, V. S.; Coletta, G. D.; Chazdon, R. L.; Ferraz, S. F. B.; Almeida, D. R. A.; Brancalion, P. H. S. (2018) Early ecological outcomes of natural reeneration and tree plantations for restoring agricultural landscapes. Ecological Applications, v 20, n 2, p 373-384.. ABSTRACT Mixed tree plantings and natural regeneration are the main restoration approaches for recovering tropical forests worldwide. Despite substantial differences in implementation costs between these methods, little is known regarding how they differ in terms of ecological outcomes, which is key information for guiding decision-making and costeffective restoration planning. Here, we compared the early ecological outcomes of natural regeneration and tree plantations for restoring the Brazilian Atlantic Forest in agricultural landscapes. We assessed and compared vegetation structure and composition in young (7-20 years old) mixed tree plantings (PL), second-growth tropical forests established on former pastures (SGp), on former Eucalyptus spp. plantations (SGe), and in old-growth reference forests (Ref). We sampled trees DBH 1-5 cm (saplings) and trees DBH>5 cm (trees) in a total of 32 20 x 45 m plots established in these landscapes. Overall, the ecological outcomes of natural regeneration and restoration plantations were markedly different. SGe forests showed higher abundance of large (DBH >20 cm) non-native species – of which 98% were re-sprouting Eucalyptus trees – than SGp and PL, and higher total aboveground biomass; however, aboveground biomass of native species was higher in PL than in SGe. PL forests had lower abundance of native saplings and lianas than both naturally established second-growth forests, and lower proportion of animal dispersed saplings than SGe, probably due to higher isolation from native forest remnants. Rarefied species richness curve was lower in SGp, intermediate in SGe and Ref and higher in PL, whereas rarefied species richness curves of saplings was higher in SG than in Ref. Species composition differed considerably among forest types. Although these forests are inevitably bound to specific landscape contexts and may present varying outcomes as they develop through longer time frames, the ecological particularities of forests established through different restoration approaches indicate that naturally established forests may not show similar outcomes to mixed tree plantings. The results of this study underscore the importance that restoration decisions need to be based on more robust expectations of outcomes that allow for a better analysis of the cost-effectiveness of different restoration approaches before scaling-up forest restoration in the tropics. Keywords: Active restoration; Ecological monitoring; Forest biomass; Passive restoration; Restoration ecology, Second-growth forests; Forest succession; Tropical forest restoration 2.1.. INTRODUCTION Selection of the best restoration approaches is largely defined by the potential for autogenic regeneration. of the target site and by management objectives, both of which determine the level of intervention required to foster ecosystem recovery (MCDONALD; JONSON; DIZON, 2016). Because of the spatial variation of these two driving factors, a wide gradient from passive to active restoration is observed, ranging from land abandonment and site protection to highly costly interventions to reconstruct ecological communities in degraded sites (HOLL; AIDE, 2011). The scale of restoration initiatives also affects the degree of intervention. Lower levels of intervention are adopted when ecosystem recovery is planned for larger spatial and temporal scales in landscapes that offer appropriate conditions for natural regeneration (CHAZDON; URIARTE, 2016), while more intensive and costly interventions have been employed in restoration projects requiring faster results but at smaller spatial scales, as in the case of mandatory projects to comply with environmental laws (BRANCALION et al., 2016). For tropical forests, a number of science-based guidelines have been developed to guide the selection of the most appropriate restoration.

(19) 18. approach by practitioners, with the goal of increasing restoration effectiveness and cost reduction (HOLL; AIDE, 2011; STANTURF; PALIK; DUMROESE, 2014). Natural regeneration without human assistance is the most widely used restoration approach for restoring tropical forests (CHAZDON; GUARIGUATA, 2016), while mixed species tree planting is preferred for active restoration (HOLL; AIDE, 2011; RODRIGUES et al., 2011). These contrasting approaches can potentially affect the structure and composition of restored forests, with implications for expected outcomes and rates of recovery. During natural regeneration, for instance, the species pool is limited by the interaction of natural local dispersal and colonization processes, vegetation cover in the surrounding landscape, and micro-site conditions for tree species establishment (HOLL et al., 2000; HOOPER; LEGENDRE; CONDIT, 2005). Human activities, such as seed collection and seedling production in nurseries, overcome dispersal limitations of species that will be introduced in mixed tree plantings in restoration sites, allowing the species composition of the restored forest to vary independently of the natural potential of tree species to recolonize the site (BRANCALION et al., 2012b). The contribution of landscape dynamics also differs between these restoration approaches. Whereas natural regeneration is favored in closer proximity and connectivity to forest remnants (ARROYO-RODRIGUEZ et al., 2015; BARNES; CHAPMAN, 2014; SLOAN; GOOSEM; LAURANCE, 2015), mixed tree plantings are commonly established in sites further from remnant forests or in landscapes with reduced forest cover with low levels of seed rain (HOLL; AIDE, 2011; RODRIGUES et al., 2009). Economics and ecology are integrated. into decision making frameworks to selection restoration. approaches in different socio-ecological contexts (HOLL, 2017). Thus, understanding the ecological outcomes of natural regeneration and mixed tree planting in contexts in which both are needed and viable is a first step towards the selection of restoration approaches with higher cost-effectiveness, an emerging challenge for up-scaling forest restoration efforts globally (BIRCH et al., 2010). Although natural regeneration has considerably lower costs (but see Zahawi, Reid, & Holl (2014)), few studies have compared the ecological outcomes between this approach and tree planting while controlling for other factors (such as forest age, soil type and prior land use), thus preventing more accurate cost-effectiveness analyses (but see Gilman et al., 2016). The few comparative publications to date have shown marked differences between natural regeneration and mixed-species restoration plantations. Overall, young (5-10 years old) mixed tree plantations show higher tree species richness and lower canopy openness than naturally established forests in Australian wet tropics uplands (SHOO et al., 2016), Andean Ecuador (WILSON; RHEMTULLA, 2016), humid forests of Costa Rica (HOLL et al., 2013; HOLL; ZAHAWI, 2014) and seasonal tropical forests in Brazil (BRANCALION et al., 2016). In contrast, Gilman et al. (2016) observed that species richness and composition of naturally recruiting individuals in the understory of mixed tree plantings vs spontaneously regenerating forests were similar in wet forests in Costa Rica, but mixed tree plantings had higher aboveground biomass due to the planted trees. These differences may emerge from the comparison of passive and active restoration in conditions where the potential for autogenic restoration was naturally low, thus limiting the full potential of natural regeneration to recover forest structure and diversity, which is known to take several decades or more (CHAZDON, 2014; MARTIN; NEWTON; BULLOCK, 2013). The ecological outcomes in single species tree plantings are influenced by the composition, management treatments, harvesting cycles and context (LAMB, 1998). As an alternative to mixed tree plantings, single species plantations consistently house less native plant biodiversity than second-growth forests in their understory, but such plantations accumulate biomass at faster rates (BARLOW et al., 2007; BONNER; SCHMIDT; SHOO, 2013). Evidently, the results of these comparisons are context dependent..

(20) 19. This study compares early outcomes of natural regeneration and mixed tree planting in scenarios where both restoration approaches succeeded in reestablishing a forest community. We compared structure, species richness and composition of saplings and trees in forests undergoing restoration established through mixed-species tree plantings with natural regeneration in pastures and Eucalyptus spp. plantations. We also compare these secondgrowth forests with old-growth "reference" forests near the study region. We expected that second-growth forests naturally established on pastures and abandoned Eucalyptus plantations will show, when compared to mixed tree plantings: i). higher abundance of tree saplings and climbers, because tree plantings are usually carried out in areas more isolated from seed sources (HOLL; AIDE, 2011); and because mixed tree plantings may require more time to support understory colonization by seeds produced by planted trees;. ii). lower abundance of bigger trees and reduced aboveground biomass (AGB) of native species. Since management interventions (fertilization, weeding, control of leaf-cutter ants, isolation from fires and cattle grazing, and regular spacing) will favor tree growth in mixed tree plantings, in contrast with selforganizing, second-growth forests;. iii). lower species richness. Due to the reduced species pool in agricultural landscapes and the high diversity of species introduced in mixed tree plantings in the region;. iv). higher proportion of abiotic-dispersed trees and more pioneer species, due to restoration guideline in Brazil and because practitioners favor biotic-dispersed species (Pedro H. S. Brancalion, personal communication) and a balanced proportion of fast- and slow-growing species in mixed tree plantings in our study region (RODRIGUES et al., 2009);. v). a more similar composition to old-growth forests, since composition of second-growth forests will be essentially determined by the same regional species pool, while plantations introduce species that are outside the regional species pool or are locally rare.. 2.2.. METHODS. 2.2.1.. Study site We studied second-growth forests established without human assistance and mixed tree species plantings. in the Corumbataí river basin of São Paulo State, southeast Brazil, in seasonal semi-deciduous tropical forests of the Atlantic Forest biome. Old-growth, reference forests were sampled as close as possible from the Corumbataí river basin, because no single conserved forest was found within it (23.8 ± 4.9 km from the basin, min: 17.2 km, max 28.5 km). The climate of the region is classified as Cwa according to the Köppen classification (ALVARES et al., 2013), with dry winters and wet summers, mean annual precipitation of 1,367 mm and mean temperatures of 20.5 oC (minimum and maximum monthly averages of 15.6 oC and 29.5 oC, respectively). Altitude varies from 470 to 1060 meters asl. The basin occupies an area of 1,700 km² and the main soil types are Acrisols (44%) and Ferralsols (22%). The basin is an ecotone between the Atlantic Forest (42% of the basin) and Cerrado (tropical savanna) (58%). The seasonal semi-deciduous forests characterize the “interior” biogeographical zone of the Atlantic Forest, the second most threatened of this biome, with only 7% native cover remaining (RIBEIRO et al., 2009). The recent increase of native forest cover in some regions in this biome (BAPTISTA; RUDEL, 2006; REZENDE et al., 2015) provide a unique opportunity to study the dynamics of tropical forest regeneration in highly deforested agricultural landscapes. Most of the deforestation occurred in the 19 th century for coffee plantation, which subsided in the early 20th century when coffee was gradually replaced by pastures and intensive agriculture in flatter terrains, mostly.

(21) 20. sugarcane plantations. During the 1970s, the region developed industrially, resulting in the migration of the rural population to urban centers and land abandonment followed by forest regeneration in some areas unsuitable for mechanized agriculture (DEAN, 1977). This contributed to a doubling of native forest cover in the six landscapes analyzed in this study (described in the next section) from 8 to 16% between 1962 and 2008 (FERRAZ et al., 2014). Currently, the main land uses in the Corumbataí watershed are pastures and sugarcane fields, occupying 43.7% and 29.4% of the basin, respectively. Native forest remnants, active Eucalyptus plantations for firewood and timber, and other land uses (buildings, roads, water bodies, other agricultural plantations, etc.) occupy 12.4%, 7.3% and 7.2% of the entire basin, respectively. Native forest cover in this region is represented by a mosaic of disturbed remnants of different sizes and levels of human-mediated disturbances, mostly resulting from cattle grazing and recurrent fires coming from sugarcane fields, which were burned historically before manual harvesting. More recently, burning of sugarcane fields was legally prohibited and mechanized harvesting has predominated, which favors forest regeneration in slopes that can no longer be used for mechanized production (MOLIN et al., 2017)..

(22) 21. Figure 1: Top: location of the selected landscapes in the Corumbataí basin, São Paulo State, Brazil. Bottom: highresolution satellite images showing overall landscape context of mixed tree plantings for forest restoration (A) and naturally established second-growth forests (B) sampled in this study. Each forest type is highlighted in red in the bottom figure. Note that second-growth forests are next to existing forest remnants while tree plantings are located along riparian zones within sugarcane plantations, more isolated from forest fragments.. 2.2.2.. Landscape selection for sampling forests undergoing restoration To locate second-growth forest in the 1,700 km² Corumbataí River watershed, we selected six landscapes. following the diversity variability analysis proposed by Pasher et al. (2013). First, we divided the basin into 1×1, 2×2, 3×3, 4×4 and 5×5 km square grid cells. For each grid size, we calculated the Shannon landscape diversity index.

(23) 22. (MCGARIGAL; MARKS, 1994) based on a 30m-resolution land-use map from 2002. Finally, we plotted the mean landscape diversity index of each grid size against the cell size. This method identified the 4×4 km cell (16 km²) as the smallest cell grid that shows no variation when compared to landscape diversity index of larger sizes, thus representing the landscape diversity of the study area. Using the 4 km square grid cell, we submitted the study region to a moving window analysis and calculated, for each pixel, the proportion of sugarcane, pasture and native forest cover of a sampling window centered on it. Finally, we selected six of the 4 x 4 km landscapes that had, in 2008 (latest image available), at least 70% of agricultural matrix and at least 10% native forest cover. To distribute the six selected landscapes more evenly across the study region, three landscapes were chosen randomly in the southern part of the basin and three in the northern part of the basin. For more details, see Ferraz et al. (2014).. 2.2.3.. Experimental design In the six chosen landscapes described previously, we located second-growth forests that established. without human assistance on planted cattle pastures (hereafter “SGp”) and in Eucalyptus spp. plantations abandoned after harvesting (hereafter “SGe”), the two most common prior land uses of second-growth forests in this region. When we refer to these second growth forests at the same time we use the abbreviation “SG”. We identified SG forests that were present in the 2008 satellite imagery but not in the 1995 imagery; since data gathering occurred in 2015, we estimated that sampled SG forests were 7-20 years old. Before the establishment of SGp, pastures had low productivity (i.e. <2 animals per hectare, without regular fertilization or other inputs), harbored very few isolated native trees and were covered by non-native fodder grasses, mainly Urochloa spp. In SGe, Eucalyptus spp., a nonnative species, was hand-planted and harvested using a chainsaw (which probably occurred 6-8 years after planting). Some of the harvested eucalyptus resprouted, resulting in a mixed community of resprouting Eucalyptus spp. and naturally regenerating native trees. Except for the resprouting Eucalyptus spp. in SGe, no seedlings were planted in SG forests and all individuals sampled originated from natural regeneration. High-diversity mixed species tree plantations (hereafter “PL”) of the same age range (7-20 years) were identified in the same study region (basin) but not in the same 16 km² landscapes where we sampled SG forests, due to the lack of mixed tree plantings for restoration plantings in these small landscapes. These tree plantings were typically established by sugarcane mills to comply with Brazilian environmental law (see details in Rodrigues et al. (2011)). Mixed tree plantings were established in sites used for decades for intensive sugarcane cultivation, where low potential for autogenic recovery limited the use of passive restoration approaches. Over 50 species of native tree seedlings were planted in a 3 × 2 m spacing (1,666 seedlings/ha) (RODRIGUES et al., 2011). Mixed tree plantings were protected from fire, fenced to prevent cattle invasion, and planted seedlings were fertilized and weeded for 2–3 years after planting. Given that natural regeneration is more likely to occur near existing forest remnants (RODRIGUES et al., 2011; SLOAN; GOOSEM; LAURANCE, 2015) and on slopes (MOLIN et al., 2017), mixed tree plantings for forest restoration usually occur in more isolated areas (RODRIGUES et al., 2011); accordingly, the landscape context of SG and PL forests differed (Figure 1). Among the 18 SG forests sampled, only one was not adjacent to a native forest remnant. On the other hand, the average distance of the seven PL forests sampled from the closest native forest remnant was 1,275 m (min: 10 m, max: 3300 m); all PL forests sampled were <100 m from rivers or water.

(24) 23. reservoirs and all, except one, was surrounded by large sugarcane plantations. Forest stand area where each plot was installed was: SGp: 21 ± 4 ha, SGe 50 ± 10 ha, PL 48 ± 36 ha. Finally, we identified old-growth conserved forests (Ref) of the same vegetation type in other agricultural landscapes near the study region. We selected four forests with no history of large disturbances in the last 100 years, protected from human and cattle encroachment and belonging to relatively large forest areas for the regional context (50 to 250 hectares), where small remnants predominate. Examples of the forest types sampled are in Supplementary File S3). We installed a 20 × 45 m (900 m²) plot to gather vegetation data in seven SGp forests, 11 SGe, and ten PL (hereafter collectively referred as “forest types”), and in four Ref sites, totaling 32 plots. We installed a single, long and continuous plot because both SG and PL forests were distributed in narrow patches at the limits of expanding forest remnants or in buffers around water streams, respectively. To avoid pseudo-replication, we did not place plots in forests with the same previous land use located in the same continuous forest remnant - or in the same continuous planting area for PL. In each 900 m² plot, we measured the diameter at breast height (DBH) and identified to the highest taxonomic level possible all living rooted trees and shrubs DBH≥5 cm (hereafter “trees”). We carried out superficial excavations to confirm if stems of the same species next to one another belonged to the same individual tree and we measured all stems DBH>5 cm of the same tree. Additionally, we installed a 4 × 30 (120 m²) subplot at the center of each plot to count and identify all trees and shrubs with DBH 1-5 cm (hereafter “saplings”). All sampled individuals were classified according to successional group (pioneer or non-pioneer), dispersal syndrome (animal-dispersed, abiotic dispersal or not classified) and species origin (native or non-native to the study region) (CREES; TURVEY, 2015; SWAINE; WHITMORE, 1988). We evaluated whether individuals sampled in PL were introduced during planting or established naturally by observing the distribution of individuals in the planting lines. Additionally, we estimated liana abundance by walking two 45-m transects in each plot and counting the number of touches of lianas diameter > 1 cm in a 2-m high pole, a method adapted from Vidal et al. (1997). In two PL plots we were not able to sample lianas and saplings.. 2.2.4.. Data analyses To analyze forest structure, we divided all sampled trees into three diameter classes based on their largest. stem: 1-5 cm, 5-20 cm, and >20 cm; the first class was sampled in the 120 m² subplot, while the others were sampled in the larger 900 m² plot. We analyzed the abundance of trees in each diameter class per plot separately for native and non-native species. We estimated AGB of each stem based on the equation 7 developed in the work of Chave et al. (2014). This equation requires only tree DBH and wood density. Data on wood density of the sampled trees were obtained from several references, but mainly from Chave et al. (2009) and Zanne et al. (2009). When wood density data was not available for a given species, we would use the following wood density values, in this order: i) average of the species of the same genus sampled in this study, or ii) average of species of the same genus in Zanne et al. (2009), or iii) average of the species of the same family sampled in this study. For species identified only to the genus or family level, we followed the steps mentioned previously. For unidentified morphospecies, we considered wood density as the average density of all species sampled in the study site. We used a specific equation developed locally to estimate AGB of Eucalyptus spp. trees (CAMPOS; SILVA; VITAL, 1992). Aboveground biomass of trees, abundance of native and non-native species for each diameter class, species density of native species, liana abundance, and proportion of pioneers and animal-dispersed native trees and.

(25) 24. saplings were compared among forest types using ANOVA and means were compared by the Tukey multiple comparison procedure (α = 0.05) when data were normally distributed. We used a general mixed model considering the logarithmical distribution when data did not have a Gaussian (normal) distribution. In both cases, means between forest types were compared by the Tukey multiple comparison procedure (α = 0.05). Rarefied species richness curves, indicator species and species composition analyses were carried out using the R package “vegan” (OKSANEN et al., 2016). We included Ref in these analyses. Rarefied species richness accumulation curves were developed for trees and spontaneously regenerating saplings of each forest type using the function “rarefy”. Given the low abundance of spontaneously regenerating saplings in PL, we were unable to develop a rarefaction curve for this forest type. We used the function “indval” to identify indicator tree species in each forest type with indicator value >0.5 and p<0.05. To compare species composition of saplings and trees among forest types, we calculated the Chao-Jaccard dissimilarity index (CHAO et al., 2004) between each plot and created a graph using non-metrical multidimensional analysis to visualize similarity among plots of different forest types using the “mds” function. We considered only native species to calculate rarefied species richness curves, and both native and non-native species for indicator species and non-metrical multidimensional analyses. Trees that could not be identified due to the absence or difficulty to sample vegetative or reproductive material were not considered for calculating species density, diversity, rarefied richness curves and composition, while unidentified morphospecies for which vegetative material was collected were included in these analyses. Given that non-native tree species and pioneer native species are well described and easily identifiable in the field, we classified morphospecies that were not identified and trees that we could not collect vegetative material – due to the absence of vegetative material or difficulty to reach the canopy - as native non-pioneer species in order to analyze abundance, AGB and proportion of pioneers. Given the low number of repetitions and the distance from other forest types, Ref forests were not included in the statistical comparison of structure, biomass, species density, liana abundance and proportion of biotic-dispersed trees and pioneers among forest types, but the data obtained in these forests are included in the results for reference and discussion purposes. 2.3.. RESULTS. 2.3.1.. Forest structure We sampled a total of 1025 saplings and 3167 trees, of which 68 (6.6%) and 410 (12.9%) belonged to. non-native species, respectively. We could not identify to the species level 143 (13.9%) and 198 (6.0%) of all saplings and trees, respectively. Overall, abundance of saplings of native species differed among forest types, being higher in SG forests (Supplementary File S1). Most of the saplings (87.9%) in PL were planted, with low abundance of spontaneously regenerating individuals. Forests of different types had similar abundance of trees DBH 5-20 cm, but SG showed lower abundance of larger native trees (DBH >20 cm) when compared to PL (Figure 2). Liana abundance differed among forest types, being lower in PL, whereas SGp and SGe did not differ (Supplementary File S1). The abundance of saplings of non-native species was similar among forest types; no Eucalyptus saplings were found in any of the sites (Supplementary File S1, Figure 2). However, SGe had the highest abundance of nonnative tree species among the largest trees (DBH > 20 cm), mostly large Eucalyptus spp. individuals. In SGe, resprouted Eucalyptus spp. represented 0%, 4.1 ± 2.2% and 64.0 ± 6.8% of trees with DBH 1-5 cm (saplings), 5-20 cm, and >20 cm, respectively..

(26) 25. Figure 2: Tree community structure in 7-20 years old second-growth tropical forests established over pastures (SGp) and Eucalyptus spp. plantations (SGe), mixed-species tree plantings (PL) and reference forests (Ref) in agricultural landscapes of southeastern Brazil. Abundance of native and non-native species per hectare is shown in the left and right graphs, respectively. Community was divided in the three diameter at breast height (DBH) classes listed in the vertical axis of each graph at the left. Most non-native trees in SGe are resprouting Eucalyptus spp. Abundance of native species DBH 1-5 cm was compared using ANOVA (α = 0.05), while the abundance of native species of other DBH classes and all non-native classes were non-parametrical and a generalized mixed linear model considering lognormal distribution of data was used (α = 0.05). In both cases, means were compared by the Tukey multiple comparison procedure (α = 0.05). Ref forests were not included in statistical analyses. Boxplots followed by the same letter do not differ statistically in each DBH class. N=4 for Ref forests and N=7, 11 and 10 for SGp, SGe and PL forests, respectively. Note that graphs have different scales. Aboveground biomass (AGB) differed among forest types (Supplementary File S1). SGe showed similar biomass of native species to SGp, but less than PL, while total AGB (native + non-native species) was higher in SGe in the other forests (Supplementary File S1 and Figure 3). Virtually all AGB stocked in PL (97.3%) was from planted trees..