1 INTRODUCTION

Across the tropics, agricultural development and industrialization have resulted in the clearance of primary forests while urbanization has led to the abandonment of marginal agricultural lands (Guariguata & Ostertag, 2001; Melo, Arroyo‐Rodriguez, Fahrig, Martinez‐Ramos, & Tabarelli, 2013; Wright & Muller‐Landau, 2006a). As a consequence, forests regenerating in previously deforested areas—commonly called secondary forests (SFs)—have become an increasingly prominent feature of tropical landscapes and now account for a majority of remaining forest cover in many regions; for example, all forests in Puerto Rico and Costa Rica are secondary regrowth (Chazdon, 2003; Lugo & Helmeri, 2004), while SFs account for 63% of remaining forest cover in Southeast Asia (Mukul, Herbohn, & Finn, 2016). Given that the socioeconomic forces driving the expansion of SFs are unlikely to abate in the near future (Barlow et al., 2018), SFs are projected to increase in relative and absolute extent (Aide et al., 2012; Chazdon & Guariguata, 2016; Chazdon et al., 2009). Previously labeled “forests of the future” (Orihuela‐Belmonte et al., 2013; Sánchez‐Azofeifa et al., 2005; Wright, 2005), SFs have become important forests of the present.

The widespread replacement of relatively undisturbed primary forest (UPF) by SF has profound implications for global climate change and biodiversity conservation. Tropical forests store—principally in the form of plant biomass (Aguiar et al., 2016)—37% of the planet’s terrestrial carbon (U.S. D.O.E., 2010), and deforestation and forest disturbance release more carbon into the atmosphere than all other sources except fossil fuel combustion (Basham et al., 2016; Bonan, 2008; van der Werf et al., 2009). Tropical forests are also host to two‐thirds of all terrestrial species (Dirzo & Raven, 2003). Humanity’s ability to mitigate catastrophic climate change and avert mass species extinctions therefore depends, in part, on the capacity of SFs to recover the biomass and biota of UPFs. In addition, given that funding for both carbon and biodiversity conservation is far less than needed to meet globally agreed conservation targets (Basham et al., 2016), an understanding of potential synergies and trade‐offs between these differing dimensions of SF regrowth is needed to support the design of successful restoration strategies.

While the structural features of a forest, including plant biomass, often approach values typical of UPF in under a century of secondary succession (Feldpausch, Riha, Fernandes, & Wandelli, 2005; Fountain‐Jones et al., 2015; Guariguata & Ostertag, 2001), biotic recovery is subject to much greater uncertainty and debate (Gardner, Barlow, Parry, & Peres, 2006; Whitworth, Downie, May, Villacampa, & MacLeod, 2016; Wright & Muller‐Landau, 2006b). Estimates of the time required for SFs to regain the species richness of UPFs range from decades to centuries (Dunn, 2004; Martin, Newton, & Bullock, 2013; Whitworth et al., 2016). The rate at which the species composition of SF converges to that of UPF is even less certain, with estimates of recovery timescales ranging anywhere from decades to millennia (Chazdon, 2008; Letcher & Chazdon, 2009). Indeed, some findings suggest that SFs will inevitably contain a severely impoverished subset of primary forest specialists and thus lack the capacity to return to a pre‐disturbance state (Jakovac, Bongers, Kuyper, Mesquita, & Peña‐Claros, 2016; Kettle, 2012; Moura et al., 2014). And, although large‐scale studies have revealed a high degree of congruence between carbon and biodiversity in the tropics (Cavanaugh et al., 2014; Strassburg et al., 2010), evidence regarding the nature of the biomass–biodiversity recovery relationship in regenerating secondary forests is both conflicting and extremely limited (Gilroy et al., 2014; Martin et al., 2013).

Patterns of biotic recovery during secondary succession are not expected to be consistent, and will vary along broad geographical, temperature, and soil fertility gradients (Nichols et al., 2007; Wright & Fridley, 2010) and as a consequence of stochastic events, such as chance dispersal (Chazdon, 2008; Norden et al., 2015). Nonetheless, the starkly differing findings that emerge from previous research may also be strongly influenced by a variety of methodological limitations. These limitations can be grouped into four main categories: (i) Site selection bias: Most SF analyses focus on sites that have only recently begun regenerating (Dent & Wright, 2009). Extrapolations of recovery prospects from the earliest stages of succession may miss important later‐stage shifts in regeneration pathways (Whitworth et al., 2016); (ii) Taxonomic Sampling bias: Many studies infer the biodiversity value of SFs by sampling a single taxonomic group (Dunn, 2004). In fact, almost three‐quarters of all studies focus solely on woody vegetation (Quesada et al., 2009). Whether the successional dynamics of large‐stemmed plants can serve as a dependable proxy for all SF biodiversity remains relatively unexplored (Hilje & Aide, 2012); (iii) Insufficient sampling effort: To accurately determine SF recovery rates requires: (a) sufficient replication at all stages of regeneration and (b) a sufficient sample of co‐located UPFs to provide a meaningful recovery baseline. Few studies apply such sampling effort (Barlow et al., 2007); and (iv) Bifurcation of scale: The scale at which SF research is conducted is largely split between (a) macroscale meta‐analyses spanning thousands of kilometers, which rely on coarse‐grained data and, therefore, likely fail to capture important inter‐ and intra‐regional variability (Gardner et al., 2013) and (b) microscale intensive studies of plots covering only a few tens of kilometers (Barlow et al., 2007; Peres et al., 2010).

We seek to address these limitations by undertaking a detailed mesoscale assessment of the recovery of SFs spanning 800 km of the Brazilian Amazon (Figure 1). Brazil contains the largest remaining expanse of tropical forests, with over 60% of the Amazon rainforest lying within its borders (FAO, 2010). As in other tropical regions, agricultural abandonment on deforested land has led to a proliferation of SFs across Brazil: In the last three decades, the area of the Brazilian Amazon occupied by regenerating secondary forests increased fivefold, from <3 million ha in 1980 to over 15 million ha in 2012 (Aguiar et al., 2016; Jakovac et al., 2016). Moreover, as part of the Bonn Challenge, Initiative 20 × 20, and its Forest Code law, Brazil is committed to the restoration of an additional 12 million ha of forest by 2030 (Chazdon et al., 2016; Crouzeilles et al., 2016; Mukul et al., 2016). Despite these bold commitments, there is significant uncertainty regarding restoration priorities and the extent to which SFs are able to meet legally mandated minimum ecological standards (de Souza, Vidal, Chagas, Elgar, & Brancalion, 2016).

Figure 1 Open in figure viewer PowerPoint Study area and design. Municipality‐level forest loss (a) and fragmentation (b) across the Brazilian Amazon. The locations of the state of Pará and the municipalities of Paragominas and Santarém–Belterra (in the east and west of Pará, respectively) are shown in white. Municipalities in light grey had no native forest cover. The distribution of primary forest, secondary forest, and water bodies in Pará (c), Santarém (d), and Paragominas (e). White represents non‐forest land. Also shown in these latter two panels are the distribution and age of the secondary forest study plots and the distribution of the undisturbed primary forest plots (dark green diamonds). (d) Floral and faunal sampling within the study plots [Colour figure can be viewed at wileyonlinelibrary.com

We surveyed 59 forests undergoing natural regeneration following agricultural abandonment, along with 30 UPF reference sites, in two deforestation frontier regions of the eastern Amazon. Alongside large‐stemmed plants, the most commonly sampled group in SF studies, we sampled scarabaeine dung beetles, birds, and small‐stemmed plants. Each of these groups plays a key functional role in secondary succession—through, for example, primary or secondary seed dispersal, control of herbivorous insects, and nutrient cycling—and together provide a powerful and complementary set of bioindicators of ecosystem‐wide change (Audino, Louzada, & Comita, 2014; Gardner, Hernández, Barlow, & Peres, 2008; Guariguata, Chazdon, Denslow, Dupuy, & Anderson, 1997; Moura et al., 2013; Reid, Harris, & Zahawi, 2012). To investigate the drivers of succession in regenerating forests, we also measured a suite of forest structure, current and historical landscape context, and topoedaphic environmental variables at the plot and landscape scales. These data provide one of the most comprehensive assessments of tropical SFs to date, and we used them to address five questions: (i) Have our SFs regained the biomass and biodiversity typical of UPFs? (ii) How do species richness and composition recover in regenerating forests relative to biomass? (iii) At what rates do biomass and biodiversity recover toward a UPF state? (iv) Will managing SFs for carbon necessarily protect biodiversity, or are patterns of species occurrence driven by factors other than biomass? and (v) Are there thresholds in species’ responses to biomass that can help guide forest management decisions?