Effects of Defaunation on Plant Functional Diversity in a Neotropical Forest

FUNCTIONAL TRAIT MEDIATION OF PLANT-ANIMAL INTERACTIONS: EFFECTS OF DEFAUNATION ON PLANT FUNCTIONAL DIVERSITY IN A NEOTROPICAL FOREST

A DISSERTATION SUBMITTED TO THE DEPARTMENT OF BIOLOGY AND THE COMMITTEE ON GRADUATE STUDIES OF STANFORD UNIVERSITY IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY

Erin Leigh Kurten August 2010

This work is licensed under a Creative Commons Attribution- Noncommercial 3.0 United States License. http://creativecommons.org/licenses/by-nc/3.0/us/

This dissertation is online at: http://purl.stanford.edu/bb408gp7470

ii I certify that I have read this dissertation and that, in my opinion, it is fully adequate in scope and quality as a dissertation for the degree of Doctor of Philosophy.

Rodolfo Dirzo, Primary Adviser

I certify that I have read this dissertation and that, in my opinion, it is fully adequate in scope and quality as a dissertation for the degree of Doctor of Philosophy.

Peter Vitousek

I certify that I have read this dissertation and that, in my opinion, it is fully adequate in scope and quality as a dissertation for the degree of Doctor of Philosophy.

David Ackerly

Approved for the Stanford University Committee on Graduate Studies. Patricia J. Gumport, Vice Provost Graduate Education

This signature page was generated electronically upon submission of this dissertation in electronic format. An original signed hard copy of the signature page is on file in University Archives.

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Preface

Dissertation Abstract This dissertation examines how terrestrial vertebrates, as seed dispersers, seed predators and herbivores, influence plant functional trait composition in tropical forests and thereby diversity. I conducted this work in the Barro Colorado National Monument (BCNM) in Central Panama, where a long term mammal exclosure experiment has been ongoing, and in neighboring Parque Nacional Soberanía (PNS), which together with the BCNM forms a defaunation gradient driven by hunting. I first comprehensively review what is known about how the loss of vertebrates in tropical forests alters plant-animal interactions, plant demography, and plant diversity. Defaunation consistently lowers primary dispersal and creates a seed shadow that is more dense around the parent tree and less dense at sites farther away. However, it also often lowers seed predation by rodents, and as a consequence, species with rodents as seed predators and dispersers often benefit from defaunation. While demographic and diversity responses tend to be more mixed, a few consistent trends emerge. Community dominance tends to increase in response to defaunation. Often, plants with particular functional traits or abiotic or unhunted dispersal agents are favored by defaunation. I next examined how community-level functional trait composition shifts in seedling communities (Chapter 2) and sapling communities (Chapter 3) which have experienced exclosure from terrestrial mammals. Seedling communities in exclosures had higher median seed mass than paired plots open to the mammal community, but treatments did not differ in their leaf traits (leaf mass per area and laminar toughness) or wood density. In contrast to the seedling community, the sapling community did show significant shifts toward higher specific leaf area and lower leaf toughness in response to herbivore exclosure, primarily due to an increased dominance of species with those traits, and secondarily due to differences in the

species present in each treatment type. These data, combined with data from PNS, also suggest that hunting results in community mean wood density in seedling communities, due to a disproportionate number of high wood density species relying on hunted animals for their seed dispersal. Finally, I investigated the seed size response to changes in mammal abundance by measuring vertebrate seed predation rates in a protected and hunted forest (Chapter 4). I found that in central Panama, seed mass does not correlate well with either body size of the seed predator, or vertebrate seed predation rates. I suggest that rather than formulate seed predation rates as a linear function of seed predator abundance, these interactions may be better modeled as threshold-dependent processes.

This work suggests that terrestrial vertebrates play an underappreciated role in maintaining plant diversity and that pan-tropical levels of unsustainable hunting may indirectly lead to losses of plant biodiversity.

Acknowledgments

My deep gratitude goes to my two advisors, Rodolfo Dirzo and David Ackerly, for giving me the flexibility and support to pursue this work. Rodolfo’s generosity as a teacher and advisor, and his tireless contributions to teaching, research and conservation, both at Stanford and internationally, are a great inspiration to me. I am thankful to David for graciously and supportively allowing me to build a network of intellectual support that allowed me to successfully complete this journey. He has also been an excellent intellectual role model for me, challenging me to grapple with complexity and nuance and to confront problems from new perspectives. Joe Wright has been a generous and supportive mentor and collaborator throughout my time at the Smithsonian Tropical Research Institute, and some of this work would not have been possible without his contributions. Peter Vitousek contributed many insightful suggestions throughout the course of this work as a committee member, and his service activities, from facilitating collaboration among scientists in Hawaii, to the First Nations’ Futures Program, have also inspired me. I also thank Fio Micheli for her enthusiasm as a committee member and for helping me frame my work in a broader context. Walter Carson very generously permitted me to work on the long-term mammal exclosure experiment he established in the Barro Colorado National Monument in Panama and shared with me his long-term datasets. The communities at Stanford, Berkeley and STRI were essential in shaping my graduate school experience. At Stanford, Will Cornwell, Nathan Kraft, Virginia Maztek, Steve Allison, Stephen Porder, Katie Amatangelo, Jen Funk, Camila Donatti, and Mauro Galetti provided helpful feedback and support throughout the various stages of my work. Doug Turner was an essential help in analyzing leaf nutrients, both for chapter three, and work not included in this thesis. Alex Royo, Allen Herre, Scott Magnan, Liza Comita, Mike Tobin,

Patrick Jansen, Noelle Beckman, Roland Kays, Jackie Willis, Stephan Schnitzer generously shared their advice and knowledge of the field site with me, improving the feasibility and execution of this work. Many thanks go to the botanists at BCI. Without their assistance, this work would not have been possible. Andrés Hernández, Oldemar Valdes, David Brassfield, and Osvaldo Calderón were always happy to help me identify whatever leaf, fruit, seed, or flower I brought to their office. Andrés and Oldemar in particular taught me most of what I know about the BCI flora. Many paid and volunteer field and lab assistants helped to make this work possible. Lissie Jiménez helped immensely with the collection and processing of the thousands of leaves collected for Chapter three. Ana Patricia Calderón and Rousmery Bethancourt contributed many early mornings conducting mammal transect surveys. Clare Sherman was such a dedicated help in the lab and fieldwork for chapter 4, that I sometimes had to remind her to take a break and have some fun. Susan Rebellon helped with the pilot studies for chapter four, and also with leaf sample processing at Stanford. Gaspar Bruner, Karen Kapheim, Adam Roddy, and David Bethancourt, helped me recover (most of) my leaf samples after they were destroyed in a freezer accident. I would also like to thank staff at Stanford and STRI who helped make the logistical aspects of this work easier. The competence of Pam Hung, Oris Acevedo, Belkys Jiménez, Orelis Arosemana, and Marcela Paz made them a pleasure to work with. The Falconer library and copy staff, as well as Allen Smith, did their best to get me the literature I needed, despite STRI’s rigorous firewall. Valerie Kiszka and Jennifer Mason helped with countless administrative tasks and advice during my time as a student at Stanford.

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Contents

Preface iv Acknowledgements vi List of tables ix List of figures x

Introduction………………………………………………………………………... 1 1 Contemporaneous defaunation and cascading effects on tropical forests…...… 3 Introduction…………………………………………………………………. 3 Scope of Review……………………………………………………………. 4 Methodology………………………………………………………………... 5 Plant-animal interactions………………………………………………….... 7 Seed dispersal……………………………………………………….... 7 Seed Predation………………………………………………………… 14 Herbivory & Trampling…………………………………………….… 17 Plant Demography……………………………………………………….…. 17 Recruitment…………………………………………………………… 17 Seedling survival……………………………………………………… 19 Standing abundance……………………………………………….….. 20 Linking Dispersal and Seedling Recruitment………………………… 22 Community Diversity….…………………………………………………… 24 Seedling density………………………………………………………. 24 Diversity……………………………………………………………… 24 Plant Functional Groups……………………………………………… 25 What is defaunation? ………………………………………………………. 27 Heterogeneity Among Studies……………………………………………… 28 Conclusions……………………………………………………………….… 29 Appendix 1.1……………………………………………………………..… 30

2 Reduced seed dispersal as a consequence of hunting lowers community-level wood density in a Neotropical forest ……………………………………… 35 Abstract…………………………………………………………………… 35 Introduction……………………………………………………………..…. 36

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Methods…………………………………………………………………… 38 Results…………………………………………………………………….. 41 Discussion………………………………………………………………… 45 Acknowledgments………………………………………………………….. 49

3 Terrestrial mammalian herbivores influence the distribution of defense and nutrient traits in a Neotropical forest………………………………………….. 51 Abstract………………………..…………………………………………… 51 Introduction…………………………………………………………………. 52 Methods……………………..……………………………………………… 54 Results……………………..……………………………………………….. 56 Discussion……………..…………………………………………………… 61 Acknowledgments………………………………………………………….. 65

4 Hunting does not alter seed predation rates as a function of seed size in a Neotropical forest……………………………………………………………… 67 Abstract…………………………………………………………………..… 67 Introduction………………………………………………………………... 68 Methods…………………………………………………………………..… 72 Results………………………………………………………………………. 76 Discussion………………………………………………………………..… 81 Conclusion…………………………………………………………………... 86 Acknowledgments………………………………………………………….. 86

Bibliography………………………………………………………………………. 87

List of Tables

3.1 The number of species and stems for which traits were measured ………….. 57 3.2 R2 values for pair -wise correlations between species mean trait values ……... 57 3.3 Effect of species and exclosure on variation in leaf toughness……….. 61 4.1 Mean fresh seed masses of study species……………………………... 75

List of Figures

1.1 Primary seed dispersal in defaunated sites relative to non-defaunated sites…………………………………………………………………… 9 1.2 Differences in seed caching in defaunated sites relative to non- defaunated sites……………………………………………………… 10 1.3 Changes in seedling distribution as a consequence of defaunation…… 13 1.4 Differences in vertebrate seed predation rates as a consequence of defaunation……………………………………………………………. 16 1.5 Differences in invertebrate seed predation rates as a consequence of defaunation……………………………………………………………. 17 1.6 Plant recruitment responses to defaunation for seedlings and saplings.. 19 1.7 Seedling survival in defaunated sites relative to non-defaunated sites.. 20 1.8 Differences in seedling densities in defaunated sites relative to non- defaunated sites……………………………………………………….. 22 1.9 Community-level herb densities in defaunation forest comparisons and mammal exclosure experiments………………………………….. 24 1.10 Differences in species richness, dominance, and diversity for defaunation forest comparisons and mammal exclosure experiments... 26 2.1 Vertebrate activity in open and exclosure plots……………………….. 41 2.2 Proportion of seedlings by dispersal mode class in open and exclosure 42 treatments…………………………………………………………….. 2.3 Proportion of seedlings by life form class in open and exclosure 42 treatments…………………………………………………………….. 2.4 Median seed mass is significantly higher in exclosures……………… 43 2.5 Effects of exclosure and hunting on community median wood density. 44 2.6 Associations between dispersal agents and species wood specific gravity………………………………………………………………… 45 3.1 Changes in plot-level trait means, unweighted by species abundance... 58 3.2 Abundance-weighted, plot-level trait means in open and exclosure plots…………………………………………………………………… 59 3.3 Intraspecific differences in species’ mean leaf toughness……………. 60 4.1 Model of how seed predation should vary with seed mass as a function of defaunation intensity…………………………………….. 71 4.2 Map of Lake Gatun study area……………………………………….. 73 4.3 Animal abundances in BCI and PNS in 2008………………………… 77

4.4 Seed predation rates as a function of seed size……………………….. 78 4.5 Number of palm seed cached in BCI and PNS………………………... 79 4.6 Identity of species removing seeds…………………………………… 80 4.7 Published studies examining seed predation of large-seeded palms….. 85

Introduction

Tropical forests are among the most biodiverse ecosystems on the planet. Many mechanisms have been proposed to explain the maintenance of that diversity, including negative distance- or density- dependent mortality, niche partitioning, and neutral processes. However, little attention has been paid to the role vertebrate consumers play in maintaining tropical diversity. Evidence from defaunated tropical forests suggests that these animals play a critical role in diversity maintenance (Chapter 1). This dissertation examines how terrestrial vertebrates, as seed dispersers, seed predators and herbivores, influence plant community composition in tropical forests and thereby levels of diversity. Specifically, I used plant functional traits as a lens through which to observe changes in seedling and sapling communities, identify which guilds of consumers were responsible for the changes, and elucidate trait- mediated mechanisms for the observed change. This work suggests that terrestrial vertebrates play an underappreciated role in maintaining plant diversity and that pan- tropical levels of unsustainable hunting may indirectly lead to losses of plant biodiversity. I first examined how community-level functional trait composition shifts in seedling communities which have been protected from terrestrial mammals (Chapter 2). I conducted this work in the Barro Colorado National Monument in Central Panama, where a long term mammal exclosure experiment was established in 1993- 94, and where terrestrial mammals are relatively abundant. I found that seedling communities in exclosures did not differ in their dispersal mode or in the relative abundance of free standing and climbing growth forms, as may be expected in an experiment that did not manipulate primary dispersal agents. Seedling communities in exclosures had higher median seed mass than paired plots open to the mammal community, but treatments did not differ in their leaf traits (leaf mass per area and 2 INTRODUCTION laminar toughness) or wood density. These results were validated with similar data from a defaunation gradient in the same region of Central Panama. One key contrast to the exclosure, however, was that seedling communities in defaunated sites had a higher representation of species with abiotic dispersal modes, lianas, and species with lower wood densities, which is consistent with the fact that primary dispersers are impacted by hunting. I next examined the sapling community in the exclosure experiment to evaluate the effects of herbivores specifically, and to identify the relative contributions of altered species present, abundance, and trait expression to the differences in functional trait composition observed (Chapter 3). In contrast to the seedling community, the sapling community did show significant shifts toward higher leaf nitrogen and lower leaf toughness in response to herbivore exclosure, primarily due to an increased dominance of species with those traits, and secondarily due to differences in the species present in each treatment type. Interestingly, I also found evidence that intraspecific differences in leaf traits were also contributing minorly to changes in community mean leaf toughness, though whether this is the result of differential mortality among genotypes or microhabitats, or a plastic response to decreased mammalian herbivory is unknown. Finally, I investigated the seed size response to changes in mammal abundance by measuring vertebrate seed predation rates in a protected and hunted forest (Chapter 4). I aimed to both test a model of how seed predation rates should vary with seed size and defaunation intensity, and potentially clarify discrepancies in community level seed-size responses to hunting at different sites. I found that in central Panama, seed mass does not correlate well with either body size of the seed predator, or vertebrate seed predation rates. In fact, I found little difference in seed predation rates between the protected and hunted forests, despite large differences in key seed predators such as peccaries and agoutis. I suggest that rather than formulate seed predation rates as a linear function of seed predator abundance, these interactions may be better modeled as threshold-dependent processes. 3

Chapter 1

Contemporaneous defaunation and cascading effects on tropical forests

Erin. L. Kurten, Mauro Galetti

INTRODUCTION

Large bodied vertebrates have been subject to human hunting for millennia. At the end of the Pleistocene, a diversity of large mammals became extinct worldwide, with human overhunting likely being one of the major drivers (Barnosky et al. 2004). The extinction of mammoth, giant sloths, giant kangaroos, giant deer, and many more megafauna in such a short time likely changed the structure and composition of their associated plant communities (Zimov et al. 1995, Guimares et al. 2008, Johnson 2009). The extinction of large vertebrates is not a phenomenon restricted to the past, but rather continues in the present day. While scientists still debate what caused the Pleistocene megafaunal extinction (Alroy 2001; de Vivo & Carmignotto 2004; Koch & Barnosky 2006; Webb 2008), there is little doubt that human activities are resposible for threatening the persistence of approximately 22 percent of all mammal and 12 percent of all bird species in the world today (Pimm et al. 2006). Particularly in tropical forests, animal populations are currently in decline due to both 4 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS unsustainable hunting and habitat fragmentation throughout Asia (Corlett 2007), Africa (Fa & Brown 2009) and Latin America (Peres & Palacios 2007). Whether changes in vegetation structure and composition have occurred as a result of the extinction of the megafauna or climate change in the past, the effects of comtemporaneous defaunation on vegetation is measurable. In many parts of the tropics, plants have lost their major seeds dispersers, seed predators, and herbivores, likely altering plant demography, spatial distribution, genetic diversity and selection on seed and plant defense traits within species, with cascading effects on community composition and diversity.

SCOPE OF REVIEW

Because the animals most vulnerable to defaunation, as well as their less vulnerable competitors, interact with plants as seed dispersers, seed predators, seedling predators, and herbivores, contemporaneous defaunation is likely to disrupt plant-vertebrate interactions. Several papers have outlined in detail how species interactions and plant communities are likely to change as a consequence of these disruptions (Wright 2003, Dirzo et al. 2007, Muller-Landau 2007). Disruptions of seed dispersal are likely to have negative effects on plant recruitment by preventing individuals from escaping distance-dependent and density-dependent mortality ( sensu Janzen 1970, Connell 1971). Reduced seed dispersal may also prevent light-demanding species, or species with other specific environmental requirements, from reaching sites favorable for recruitment (Muller-Landau 2007, Brodie et al. 2009). Changes in seed predation, seedling predation, and herbivory may have positive or negative effects on species recruitment by altering seed and seedling survival (Wright 2003, Dirzo et al. 2007, Muller-Landau 2007). Changes in seed predation and seedling predation may have further indirect benefits for invertebrate seed predators and herbivores. Overall, the net effect of defaunation on plant diversity has been hypothesized to be negative (Wright 2003, Muller-Landau 2007) This is both because it is thought that species experiencing reduced seed dispersal will not persist if they cannot escape 5 distance-dependent or density-dependent mortality, and because seed predators and herbivores will not suppress populations of competitively dominant species. Here we synthesize what is currently known about the indirect effects of defaunation on tropical plants, in the context of hunting, forest fragmentation, and animal exclosure. We divide our analysis into three sections, addressing effects on plant-animal interactions, population demography, and community diversity. Our first goal was to evaluate the extent to which the hypothesized changes mentioned above are supported by empirical evidence. Our second goal was to focus on plant species or community responses that show mixed responses to defaunation, and try to clarify why such variability may exist. Our third goal was to identify areas of study which have received little attention and warrant future attention.

METHODOLOGY

IDENTIFICATION OF STUDIES Here we summarize literature in the field as of 2009. We identified studies primarily by searching literature databases for publications on aspects of plant ecology with a “defaunation” or “hunting” component. We supplemented this collection with other studies cited in that literature, as well as unpublished theses of which we had knowledge. In total, thirty-seven studies comprise the quantitative portion of this review (Appendix 1).

STANDARDIZATION OF RESPONSE VARIABLES FOR COMPARISON Together, the variation in study design, defaunation intensities compared, species studied, and response variables measured introduce too much heterogeneity to conduct a formal meta-analysis of responses (Hedges & Olkin 1985). Yet, we have attempted to present a review which is quantitative rather than qualitative. We have done so by calculating effect sizes for each response variable. The effect size estimator we use here is the percent difference in the response variable reported between defaunated and non- defaunated sites as follows: Effect size = ∗ 100, 6 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS

where RD and RND are the magnitudes of the response variables in the defaunated and non-defaunated sites, respectively. In cases where the denominator, RND , was zero, we used the smallest possible non-zero value for RND (e.g. one seed, in the case of absolute seed dispersal). When the minimum possible response could not be inferred (e.g., a percent abundance), we conservatively estimated the effect size as 100 percent. For example, where size class data were reported, sapling to seedling ratios were calculated to allow for comparison of recruitment rates across studies. When studies reported response variables from multiple defaunated or non- defaunated sites, the data were averaged to derive a single value for defaunated and non-defaunated states. In studies of defaunation gradients, data on the abundance or presence-absence of vertebrate species relevant to the study, as well as the authors’ categorical characterizations of the relative levels of defaunation were used to define sites as “defaunated” or “non-defaunated” for purposes of comparison. In these cases, sites experiencing the lowest levels of defaunation were often grouped with “no defaunation” sites, while sites showing medium to high levels of defaunation comprised the “defaunated” sites.

QUANTIFICATION OF DEFAUNATION INTENSITY We attempted to compare effects as a function of defaunation intensity. We restricted these analyses to Neotropical sites (70% of studies) because more extensive and comparative data on vertebrate communities was reported for this region. We selected twelve genera of mammalian frugivores, granivores, and browsers that differ in body mass and sensitivity to defaunation pressures: Tapir, Tayassu, Pecari, Odocoileus, Mazama, Ateles, Allouata, Cebus, Agouti, Dasyprocta, Sciurus and Proechimys . These genera are wide-ranging in the Neotropics, though the particular species may vary. Because some studies did not report animal abundances or densities, but rather species presence or absence, we calculated the degree of defaunation for each site as the percent of focal genera that appeared to be locally extirpated in that site, though present historically.

SEED SIZE DATA In several cases, the perturbations in plant-animal interactions as a consequence of defaunation are hypothesized to vary as a function of seed size. We therefore attempted to rank species by seed size in order to present the data in a way that would address these hypotheses. In most cases, either seed mass or seed length was reported. In the case of Celtis durandii , one Astrocaryum species and one unspecified Dipteryx species for which seed sizes were not reported, an approximate seed size was estimated by assigning that species the value of a congeneric species from another site. Species for which both length and mass were reported were used to approximate a relative size rank for the species for which only mass or length were reported. Alternatively, we could have used published seed mass-seed length correlations to estimate missing size parameters and used one measure of seed size to assign size ranks. However, this would generally change the positions of only a few species with unusual seed mass-to-seed length proportions (e.g. Gustavia superba ), and we felt that those species were better assigned a rank by considering the size parameters in the context of the plant-animal interaction and the natural history of that species.

PLANT -ANIMAL INTERACTIONS

Seed dispersal Seed dispersal is thought to promote plant recruitment in tropical forests by facilitating escape from natural enemies (Janzen 1970, Connell 1971) and by helping species with particular environmental requirements for germination and survival (e.g. light- demanding species) reach favorable microhabitats (Muller-Landau 2007). Seed dispersal establishes the spatial distribution and diversity of species in the seed and seedling banks. A single seed can be dispersed multiple times by different agents. Here we refer to primary seed dispersal as physical removal from the parent tree and deposition on the ground. Primary seed dispersal can be performed by biotic agents such as primates, bats, and birds, or by abiotic agents such as wind. Once a seed has dispersed from the parent tree, it may still disperse further. This we refer to here as secondary 8 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS dispersal. We focus on secondary dispersal by biotic agents such as rodents, though in some circumstances, water and gravity may also facilitate dispersal on the ground.

PRIMARY SEED DISPERSAL Seed dispersal by biotic agents should decrease as defaunation intensity increases and the abundance of seed dispersers is reduced. This should affect large-seeded species to a disproportionate degree for two reasons. First, their predominate dispersal agents tend to be larger-bodied, and therefore more vulnerable to defaunation (Peres & van Roosmalen 2002, Holbrook & Loiselle 2009). There also appears to be less redundancy of seed dispersal agents for large-seeded species, relative to smaller-seeded species (Peres & van Roosmalen 2002, Nuñez-Iturri et al . 2008, Donatti et al. 2009). When seed dispersal has been directly measured, either as the quantity of seeds removed by primary dispersers such as birds or primates, or as the proportion of seed crop removed from a parent tree, seed dispersal is lower in defaunated forests in almost all cases examined (Fig. 1.1). The magnitude of the decrease appears to be moderately correlated with seed size. It is important to keep in mind that tropical seed mass distributions span more than seven orders of magnitude (I.J. Wright et al. 2007). The species reported here represent only the very upper range of that distribution and were generally selected for study because they were most likely to show dispersal declines. Almost nothing is known about how hunting alters biotic dispersal across the broader range of seed sizes (Muller-Landau 2007). Two studies examining how perturbations of bird communities affect the dispersal of smaller-seeded species (< 1 cm fruit diameter) have shown contrasting results. Dispersal of Bocageopsis multiflora in defaunated sites was actually more than 2-fold higher than in non-defaunated sites (Fig. 1.1). The cause of this increase is unknown, but has been suggested to result from an increase in the relative abundance of generalist, frugivorous birds in highly fragmented sites (Cramer et al . 2007). In the case of Celtis durandii , a decrease in avian forest specialists that was not compensated for by forests generalists appears to be responsible for in an overall decrease in seed dispersal (Kirika et al . 2008). 9

Genus & Source Large (Community), Tonga (26) Leptonychia, Tanzania (9) Dipteryx , Costa Rica (19) Duckeodendron, Brazil (10) Carapa, Costa Rica (19) Attalea , Panama (36) Gustavia, Panama (3) Astrocaryum, Panama (36) Antrocaryon, Cameroon (35) Virola, Panama (3) Choerospondias, Thailand (7) Pouteria, Brazil (2) Pourouma, Brazil (2) Virola, Ecuador (20) Euterpe, Brazil (14) Bocageopsis, Brazil (10) Small Celtis, Uganda (22) -100 -50 0 50 100 150 % Difference in Seeds Dispersed

FIGURE 1.1 Seed dispersal is lower in defaunated sites relative to non-defaunated sites in almost all cases. Species are ranked by seed size, with Leptonychia being largest (11.15 cm long) and Bocageopsis and Celtis being smallest (< 1 cm long). Numbers in parentheses denote the study number in Appendix 1.

SEED CACHING A special case of seed dispersal is the caching of seeds in the ground by scatter-hoarding rodents. While individual cached seeds may be predated at a later time, seeds that are not subsequently recovered are afforded protection from invertebrate seed predators. For some species, such as large seeded palms with specialist bruchid beetle seed predators, seed survival is heavily dependent upon seed caching by rodents. 10 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS

In most cases, seed caching was higher in defaunated sites relative to non- defaunated sites (Fig. 1.2). This is likely due to the fact that the rodents primarily responsible for seed caching in these forests, agoutis ( Dasyprocta spp.) and squirrels (Sciurus spp.) may actually benefit from a reduction in the abundance of predators and larger-bodied competitors at low to medium levels of defaunation (Wright 2003, Dirzo et al. 2007, Peres & Palacios 2007). Therefore, while defaunation may have a negative effect on primary dispersal, this negative effect often does not extend to secondary seed dispersal by terrestrial rodents.

Genus & Sou rce La rge Carapa, Costa Rica (18) Attalea , Panama (24) Astrocaryum, Panama (24) Lechthis, Costa Rica (18) Pentaclethra, Costa Rica (18) Astrocaryum, Brazil (17) Astrocaryum, Brazil (13) Minquartia, Costa Rica (21 ) Hymenaea, Venezuela (4) Welfia, Costa Rica(18) Otoba, Costa Rica (18) Virola, Costa Rica (18) Clarisia, Costa Rica (11) Small Virola, Costa Rica (11) -200 -100 0 100 200 300 400 500 % Difference in Seeds Cached

FIGURE 1.2 Differences in seed caching rates in defaunated sites relative to non- defaunated sites. Species are ranked by size, with Carapa being the largest (20 g) and Virola the smallest (2 g). Absence of bar indicates no difference. Numbers in parentheses denote the study number in Appendix 1.

Defaunation gradients in Venezuela and Brazil were exceptions to the trend of higher seed caching rates in defaunated forests (Asquith et al . 1999, Galetti et al . 2006). In both of these cases, the defaunated sites were small fragments in which agoutis were virtually absent. Consistent with the loss of this important terrestrial seed disperser, seed caching in defaunated sites was lower than in non-defaunated sites in these studies (Fig. 1.2). Squirrels do not compensate for the loss of agoutis (Donatti et al. 2009). In contrast to primary dispersal, no relationship between seed size and differences in caching rates were apparent between defaunated and non-defaunated sites. This may again be due to the fact that agoutis are capable of handling large seeds and fruits much larger than their relatively small gape size might suggest. Therefore, in the Neotropics secondary dispersal of larger-seeded species may occur despite some degree of defaunation, provided that agoutis persist in the site. Smaller-bodied scatter-hoarding rodents, such as squirrels, exist in Africa and Asia, however to our knowledge, there is no functional equivalent to the agouti in Paleotropical forests. The larger Paleotropical rodents that do exist, such as Cricetomys species, are larder-hoarding rodents that tend to take seeds to burrows where they will not survive (Guedje et al . 2003). Therefore, the seed dispersal “buffer” that agoutis provide in the Neotropics is less likely to be present in other systems.

SPATIAL DISTRIBUTION OF SEEDS AND SEEDLINGS The spatial distribution of seeds and seedlings on the forest floor can also be an indicator of changes in seed dispersal. In particular, if seed dispersal is lower in defaunated forests, one would expect to find a greater proportion of seeds or seedlings undispersed under parent trees relative to what is observed in non-defaunated forests. Likewise, one would expect to see fewer seeds or seedlings at distances away from parent trees relative to non-defaunated forests. Indeed for many species, seed or seedling numbers under parent trees were 2- to 12-fold higher in defaunated sites relative to non-defaunated sites, and were lower 12 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS away from parent trees (Fig. 1.3). These patterns are consistent with the loss of biotic dispersal agents. One notable exception is Hymenaea on land-bridge islands in Venezuela. Hymenaea shows the highest increases in seed pods remaining under parent trees in sites without agoutis relative to sites with agoutis. However, it is one of the few cases in which seedling densities are actually lower under parent trees in defaunated forests (Fig.1.3). This is not due to distance or density dependent mortality effects as mediated by specialist invertebrates or pathogens (Janzen 1970, Connell 1971). Rather, Hymenaea relies on agoutis to open its seed pods so that seeds may germinate. In sites lacking agoutis, the seeds cannot escape the seed pod, and seedling densities are correspondingly low (Asquith et al . 1999). One study estimated changes in dispersal distances with defaunation. Cramer et al. (2007) reported a 60 percent and 80 percent decrease in mean and maximum dispersal distances respectively for the large-seeded Duckeodendron cestroides in defaunated fragments, relative to continuous forest. In the same study, the smaller- seeded Bocageopsis multiflora mean and maximum dispersal distances were not significantly different. This suggests that large-seeded species may become more spatially aggregated in defaunated forests. Those plant species requiring special microhabitats for germination and recruitment (e.g. light gaps) may have greater difficultly reaching suitable sites. Meanwhile, smaller seeded species may be less affected.

Genus & Sou rce Large Choerospondias, Thailand (7) Leptonychia, Tanzania (9) Under Parent Balanites, Gabon (5) Attalea , Panama (37) Attalea , Panama (37) Attalea , Panama (36) Dysoxylum, India (31) Chisocheton , India (31) Astrocaryum, Panama (36) Hymenaea, Venezuela (5) Polyalthia, India (31) Hymenaea, Venezuela (4) Syagrus, Brazil (1) Small Antrocaryon, Cameroon (35) -250 0 250 500 750 1000 1250 1500 2350 2600 Large Attalea , Panama (37) Attalea , Panama (37) Away From Attalea , Panama (36) Parent Astrocaryum, Panama (36) Duckeodendron, Brazil (10) Dysoxylum, India (31) Chisocheton , India (31) Polyalthia, India (31) Small Bocageopsis, Brazil (10) -100 -50 0 50 100 150 250 300 350 Seeds Seedlings % Difference in Seed or Seedling Abundance

FIGURE 1.3 The number or proportion of seeds or seedlings is generally higher immediate vicinity of parent trees and generally lower away from parent trees, in defaunated sites relative to non-defaunated sites. Species are ordered by seed size. Numbers in parentheses denote the study number in Appendix 1.

COMPENSATION AMONG DISPERSAL AGENTS An important question regarding seed dispersal in tropical forests is whether a decrease in dispersal by the principle dispersal agent may be compensated for by other dispersal agents. Comparisons of primate diets suggest that smaller species of primates only disperse a nested subset of the plant species that are dispersed by larger primates. Even when co-occurring dispersers 14 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS appear to have significant dietary overlap on the basis of species consumed, a quantitative diet analysis reveals that dispersers differ in the fruit species that make up the major portion of their diets (Poulsen et al . 2002). This suggests that even dispersers with diet overlap may not compensate for one another. Similarly, among frugivores that feed on Virola , alternative dispersers do not shift their diet preferences enough to compensate for the loss of the principle dispersers (Holbrook & Loiselle 2009). Decreases in primary seed dispersal summarized here (Fig. 1.1) suggest that this lack of compensation among primary dispersal agents may be fairly general, at least for larger-seeded species. There are no studies that directly investigate the degree to which an increase in secondary dispersal by terrestrial vertebrates (e.g. Fig 1.2) may compensate for reduced dispersal by birds and primates. However, in cases where species have lost their primary dispersal agent, the number of seeds and seedlings remaining under adult trees in defaunated sites is often higher than in non- defaunated sites (Fig. 1.3). These data provide indirect evidence that, at least in the cases studied, terrestrial seed dispersers do not fully compensate for loss of principle avian and primate dispersers. It is possible, however, that terrestrial dispersers are partially compensating for the loss of other dispersers, and that the observed changes in seed and seedling densities under and away from parents would be even more extreme in the absence of terrestrial dispersal.

Seed Predation

PRE -DISPERSAL SEED PREDATION Beckman and Muller-Landau (2007) investigated effects of defaunation on pre-dispersal seed predation for two species of contrasting seed size in central Panama. They found that pre-dispersal seed predation by mammals was significantly lower in the defaunated sites for the larger-seeded Oneocarpus mapora . Authors did not observe mammalian pre-dispersal seed predation of small-seeded Cordia bicolor . Pre-dispersal seed predation by insects between sites was not significantly different for either species. 15

POST -DISPERSAL SEED PREDATION After primary dispersal, seeds may be predated upon by terrestrial mammals such as rodents, peccaries, pigs and deer. The majority of studies comparing seed predation rates between defaunated and non-defaunated sites have performed manipulative experiments, setting out arrays of seeds and monitoring the fates of those seeds. However a few studies of species with slowly decomposing endocarps have looked at “standing” rates of seed predation in the field (Wright et al. 2000, Wright & Duber 2001, Galetti et al. 2006). In 22 of the 31 cases, seed predation rates were lower in defaunated sites, while in nine cases, seed predation rates were higher (Fig. 1.4). In a conceptual model, larger-seeded species were hypothesized to experience higher seed predation rates in moderately defaunated forests relative to non- defaunated forests (Dirzo et al . 2007). At severe intensities of defaunation, however, large-seeded species were hypothesized to experience lower seed predation rates, relative to sites with moderate to no defaunation. These changes in seed predation rates were posited to be in response to changes in the abundance of seed predators of medium body size. Using our estimates of defaunation intensity for Neotropical sites, we examined how differences in seed predation rates for different plant species varied with defaunation intensity. These empirical data do not generally support the hypothesis that seed predation of larger seeds in moderately defaunated sites is higher than non-defauanted sites, while being lower in highly defaunated sites. (Fig. 1.4). Among the largest-seeded species studied, seed predation increased and decreased in about the same number of cases across all levels of defaunation examined. The conceptual model also hypothesized that seed predation rates of smaller seeded species would increase monotonically with increasing defaunation intensity, as small rodents experienced release from competition and predation. However empirical data suggest that seed predation rates are generally lower for smaller-seeded species in defaunated sites, regardless of defaunation intensity (Fig. 1.4).

16 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS

250

200

150

100

-50 % Difference in Seed Predation Seed in Difference % -100 0.0 20.0 40.0 60.0 80.0 100.0 120.0 Defaunation Intensity in Defaunated Site

FIGURE 1.4 Differences in seed predation rates as a function of the intensity of defaunation in the defaunated sites. Circle sizes indicate the relative seed size of the plant species.

INVERTEBRATE SEED PREDATION It has been suggested that a reduction in vertebrate seed predation may benefit invertebrate seed predators in two ways. First, invertebrates experience a release from competition for seed resources (Muller-Landau 2007). Secondly, larvae developing in seeds may experience decreased mortality due to the reduction of seed predation by vertebrates (Silvius 2002). Through these mechanisms, defaunation may lead to higher abundances of invertebrates, and thereby higher rates of invertebrate seed predation. Few studies have investigated how defaunation indirectly alters invertebrate seed predation rates. Consistent with expectation, bruchid beetle seed predation of palms was higher in defaunated sites for every species investigated (Fig. 1.5). However, in most cases, the absolute increase in invertebrate seed predation rates only partially compensated for the decreases in vertebrate seed predation (Wright et al . 2000, Wright & Duber 2001, but see Galetti et al. 2006). 17

Genus & Source Att alea , Panam a (37) Att alea , Panam a (36) Astroca ryum, Br az il (17) Astrocaryum , Panama (36) Syagrus, Br az il (1) Syagrus, Brazil (1) 0 500 1000 1500 % Difference in Invertebrate Seed Predation FIGURE 1.5 Seed predation rates by invertebrates for four species of palms are higher in defaunated sites relative to non-defaunated sites. Species are ranked by size, with Attalea being the largest and Syagrus the smallest. Numbers in parentheses denote the study number in Appendix 1.

Herbivory & Trampling

Few observational studies have evaluated differences in herbivory or trampling as a consequence of defaunation. Dirzo & Miranda (1991) an absence of vertebrate herbivory in Los Tuxtlas, a defaunated site. Alves-Costa (2004) categorized herbivory of the palm Syagrus romanzoffiana into classes by percent damage and also found that the frequency individuals experiencing little to no herbivory was 66% higher in defaunated sites. Mortality by trampling has not been evaluated for live seedlings in the context of defaunation, however trampling has been evaluated with seedling models. In Bolivia, “trampling” of seedling models by vertebrates was 75 percent lower in defaunated forests (Roldán and Simonetti 2001).

POPULATION DEMOGRAPHY

RECRUITMENT Reduced seed dispersal in defaunated forests has been hypothesized to have negative impacts on plant recruitment by preventing individuals from escaping distance-dependent and density dependent mortality (Muller-Landau 2007). Empirically, plant recruitment has been reported either in terms of new seedlings 18 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS entering the population in a determined length of time, or as the ratio of juvenile plants to seedlings. Authors could define seedlings variously as first year germinants or by size class. In about half of cases, recruitment was about 2-fold higher in defaunated forests than in non-defaunated forests, both for seedlings and saplings (Fig. 1.6). In the other half of cases, recruitment is lower in defaunated forests. There have been virtually no investigations into why recruitment rates change. The variation in changes in recruitment do not seem to be consistently related to the degree of defaunation across studies, loss of particular dispersers or loss of particular seed predators. Changes in vertebrate seedling predation and herbivory, as well as distance and density dependent mortality as mediated by invertebrate herbivores and pathogens (Janzen 1970, Connell 1971), can be responsible for altered seedling recruitment. However, no studies have recorded herbivory rates or causes of plant mortality, so no definitive data exists on how each of these mechanisms contribute to the changes in recruitment observed. One study of Syagrus found that sapling recruitment near the parent tree was slightly lower in defaunated forests, while recruitment far from the parent tree was more than 2-fold higher in defaunated forests (Alves-Costa 2004). This may be because non-vertebrate natural enemies are more than compensating for the loss of vertebrate herbivores near the parent trees, but not at far distances, for this species. However, no data exist with which to evaluate this explanation. It is also possible, in some cases, that differences in recruitment are not actually due to defaunation. In some studies of seedling recruitment, only one defaunated site and one non-defaunated site were compared. This opens the possibility that some differences in recruitment were due to other environmental factors that vary across the landscape, such as precipitation or edaphic conditions. The driver of defaunation is another potential source of variation. In some cases, hunting was the cause of defaunation, altering primarily the animal community. However in other studies fragmentation or a combination of fragmentation and hunting was the cause of defaunation. As fragmentation alters many aspects of the microclimate, it is possible that changes in seedling recruitment are due to the effects of fragmentation on 19 microclimate, rather than or in addition to the effects of a reduction in herbivore abundance.

Genus & Sou rce Attalea , Panama (36) Seedlings Dipteryx , Peru/Panama (34) Astrocaryum, Brazil (17) Astrocaryum, Panama (36) Syagrus, Brazil (1) Saplings (Community), Malaysia (31) Leptonychia, Tanzania (9) Balanites, Gabon (5) Astrocaryum, Brazil (17) Syagrus (near), Brazil (1) Syagrus (far), Brazil (1) -100 0 100 200 300 400 500 % Difference in Recruitment

FIGURE 1.6 Plant recruitment does not respond consistently to defaunation across species for either seedlings or saplings. Numbers in parentheses denote the study number in Appendix 1.

SEEDLING SURVIVAL As with recruitment, responses of seedling survival to defaunation are quite variable (Fig. 1.7). Asquith et al . (1997) found that differences in seedling survival could vary among species within the same study system. Survival was lowest on highly defaunated islands in Lake Gatun, Panama for two of the three species tested. Exclosure experiments verified that spiny rats, free of competition from larger mammals, reach population sizes 10-fold greater on these islands relative to the non-that sites (Adler 1996), and are responsible for the increased mortality. The third species, which did not show a difference in survival between sites, Virola surinamensis , experienced 100% mortality at both defaunated and non-defaunated sites (Asquith et al . 1997). Differences in survival rates can also be variable within a species in the same region, likely due to differences in the defaunation intensities being compared. Two 20 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS studies have compared the survival of Gustavia superba in defaunated sites in the Lake Gatun region of central Panama. In one case, the defaunated site was hunted relative to the non-defaunated site, but the animal community was not so impoverished as to lead to a competitive release of small rodents. In that study, G. superba had higher survival rates in the defaunated sites. However, when the defaunated sites were islands with only high populations of spiny rats present, survival of G. superba was lower in the defaunated site. This difference in the degree of defaunation intensity evaluated in the two studies, and compensatory changes in abundance of non-hunted species, are likely to account for the contrasting responses in seedling survival, and illustrate that changes in seedling survival as a consequence of defaunation may not change monotonically.

Genus & Sou rce Dipteryx , Panama (3) Gustavia, Panama (3) Dipteryx , Peru/Panama (33) Syagrus, Brazil (1) Virola, Panama (3) Gustavia, Panama (32) Pentaclethra, Costa Rica (18) -75 -50 -25 0 25 50 75 100 125 % Difference in Seedling Survival

FIGURE 1.7 Seedling survival does not show consistent changes as a consequence of defaunation, even within sites (e.g. study 4) or species (e.g. Gustavia superba in Panama). Numbers in parentheses denote the study number in Appendix 1.

STANDING ABUNDANCE Seedling densities do not respond uniformly to defaunation. Species that are dispersed and predated upon by scatterhoarding rodents tend to increase in abundance in defaunated sites (Fig. 1.8). Likewise the increase in seedling abundance of the elephant dispersed Balanites wilsoniana , despite reduced seed dispersal, was attributed to low seed and seedling predation in the defaunated site 21

(Babweteera et al. 2007). Meanwhile, species that are bird dispersed have lower abundances in non-defaunated sites. These data suggest, for species which are rodent dispersed and predated, any increases in distance-dependent or density-dependent mortality by invertebrates or pathogens, as a consequence of reduced dispersal, do not compensate for reductions in vertebrate seed and seedling predation, resulting in a net increase in survival for these species. The same does not appear to be true for bird dispersed species, although there is no theoretical reason why these species should not experience high reductions in vertebrate seed predation. All studies involving rodent dispersed species were conducted on palm species, most of which are preferred food species for agoutis, a dominant Neotropical seed predator. More studies in Neotropical forests are required to determine if these species are unique in experiencing a net benefit from defaunation, or if other species preyed upon by agoutis also increase in abundance in defaunated sites. These studies should include species that have birds or primates as their primary dispersal agents. In contrast to the rodent-dispersed palms, little is known about the seed predators of the bird-dispersed species studied. However, at least two of the three studies of bird-dispersed species were conducted in Paleotropical forests with different assemblages of seed predators. Studies of vertebrate seed and seedling predation in the Paleotropics may determine whether differences observed between rodent- dispersed species and bird-dispersed species may be due to differences in the seed and seedling predator community, rather than differences between species with different dispersal agents per se .

22 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS

Genus & Sou rce Astrocaryum, Brazil (17) Rodent Dispersed Dipteryx , Peru (33) Syagrus, Br az il (1) Astrocaryum, Bolivia (29) Syagrus, Br az il (1) Attalea , Panama (36) Att alea , Panam a (37) Astrocaryum, Panama (36)

Bird Dispersed Chisocheton , India (31) Polyalt hia, India (31) Dysoxylum, India (31) Euterpe, Brazil (14) Leptonychia, Tanzania (9) -100 0 100 200 300 Elephant Dispersed Balanites, Gabon (5) -400 0 400 800 1200 % Difference in Seedling Densities FIGURE 1.8 Differences in seedling densities in defaunated sites relative to non- defaunated sites. Numbers in parentheses denote the study number in Appendix 1.

LINKING DISPERSAL AND SEEDLING RECRUITMENT It is clear, by several measures, that seed dispersal decreases as a consequence of defaunation (Figs. 1.1 & 1.3), though effects on seed caching are variable (Fig. 1.2). What is less clear is whether decreased dispersal translates to decreases in plant recruitment. Investigators focusing on decreased seed dispersal as a consequence of defaunation often suggest that decreased dispersal will result in decreases in seed and seedling survival and plant recruitment. This is because undispersed seeds are assumed to be unable to escape distance- dependent and/or density-dependent mortality (Janzen 1970, Connell 1971). However, there are several problems with this assumption. First, there are many species for which distance- and/or density-dependent mortality have not been detected (Hyatt et al . 2003, Uriarte et al . 2005). Secondly, spatial data on seedling densities also show that reduced seed dispersal does not generally lead to great decreases in seedling abundances under parent trees in defaunated forests (Fig. 1.3). This suggests that though seed and seedling mortality may be higher under parent 23 trees, the higher number of seeds that fall below the parent tree in defaunated forests compensate for this higher mortality and seedling densities under parent trees remain relatively similar to those observed in non-defaunated forests. Finally, as seen above, vertebrate seed and seedling predation rates may also vary with defaunation (Figs. 1.4 & 1.5). Seedling density data suggest that reduced rates of vertebrate seed predation result in increased survival rates that more than compensate for any increases in distance-dependent or density-dependent mortality by non-vertebrates, at least for rodent-dispersed species and species like B. wilsoniana (Fig. 1.6). Taken together, these data suggest that the assumed negative effects of reduced seed dispersal on plant recruitment as a consequence of hunting may not be general, but rather manifest themselves only in species which (1) experience strong natural enemy-mediated, negative density-dependent mortality (2) rely exclusively upon a few, large-bodied, hunted vertebrates for seed dispersal, and (3) do not experience a demographic benefit of reduced seed predation by rodents with increasing defaunation intensity. Interestingly, there are few studies that actually investigate changes in dispersal and seedling recruitment simultaneously in a single species. In the case of the large seeded palm Attalea butyracea in Central Panama, seed dispersal is reduced in defaunated forests (Wright & Duber 2001). In addition, the intensity of bruchid beetle seed predation does increase with proximity to the parent tree, and the effect intensifies with defaunation. However, increased bruchid beetle seed predation did not fully compensate for the decrease in vertebrate seed predation associated with the defaunation, and seedling densities of the palm were higher in defaunated sites, despite the decrease in seed dispersal. Brodie et al. (2009) modeled the effects of decreased dispersal on seedling recruitment for Choerospondias axillaris in Thailand and did find evidence that decreased dispersal would translate to decreased recruitment in this species. However this was because defaunation deprived C. axillaris of favorable regeneration sites, since C. axillaris seems to depend on light gaps for regeneration and could not disperse to gaps without its vertebrate disperser (Brodie et al . 2009). It was not a consequence of distance-dependent or density-dependent mortality . 24 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS

COMMUNITY DIVERSITY

SEEDLING DENSITY Seedling or herb density at the community level is consistently higher where exclosure experiments (Ickes et al . 2001, Royo & Carson 2005) or defaunation due to hunting and fragmentation (Dirzo & Miranda 1991) completely eliminate vertebrate herbivory (Fig. 1.9). Interestingly, at many moderately defaunated sites, community seedling densities are lower, relative to non-defaunated sites. It may be that terrestrial herbivores initially increase in abundance as a consequence of a release from predation, and perhaps competition. However, decreased seedling densities at the community level are in contrast to higher seedling densities of many individual species in defaunated sites (Fig. 1.8). This discrepancy further highlights the need to understand how defaunation alters plant recruitment across a broader range of species, and to reconcile changes in recruitment at the species level with changes at the community-level. Site & Source Uganda (8) Forest comparison Uganda (8) Exclosure study Panama (38) Bolivia (29) Panama (30) Peru (27) Malaysia (21) Panama (30) Malaysia (21) Mexico (12) -100 -50 0 50 100 150 % Difference in Seedling Density

FIGURE 1.9 Community level seedling and herb densities generally increase in cases of high defaunation but decrease in others. Numbers in parentheses denote the study number in Appendix 1.

DIVERSITY Several short-term exclosure studies have shown that species richness initially increases with herbivore removal (Fig. 1.10a). However, this appears to be an effect of increased stem densities increasing the likelihood that rare species will be present inside the exclosures. In an exclosure experiment in Panama, rare herbs 25 increased 2-fold (Royo & Carson 2005) (Fig. 1.10b). However, defaunation also appears to increase species dominance as well (Fig. 1.10b). Diversity indices appear to mask the changes in community composition that occur in exclosure experiments, because the effects of increased species richness and increased dominance tend to cancel each other out, such that exclosures show only small decreases in indices of diversity. In forest comparisons, species richness was generally lower. Unlike short-term exclosure experiments, defaunated forests have experienced decades of defaunation, likely allowing competitively dominant plant species to outcompete rarer and less dominant species, reducing diversity. Consistent with this hypothesis, species dominance was higher in the cases that reported it (Fig. 1.10b). Responses in diversity indices were mixed (Fig. 1.10c), again suggesting that these indices may be less informative than directly evaluating changes in species richness and dominance.

PLANT FUNCTIONAL GROUPS Nuñez-Iturri et al. (2007, 2008) and Wright et al. (2007) both found little difference in total species richness between defaunated and non-defaunated sites. However, the lack of differences in species richness belied important changes in plant community composition. Species with abiotic or unhunted, biotic dispersal agents consistently show higher abundance in hunted sites (Wright et al . 2007, Nuñez-Iturri et al . 2008, Terborgh et al . 2008). In contrast, species dispersed by game animals show decreased abundance in defaunated sites. Consistent with this finding, lianas, which tend to be abiotically dispersed (Hardesty & Muller-Landau 2005, Wright et al. 2007), were higher in relative abundance in defaunated sites in central Panama (Wright et al . 2007). 26 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS

Site & Source AA Site & Source B Mexico (12) Panama (30) (game) Peru (28) Brazil (1) (game) Peru (27) Panama (30) Uganda (8) Panama (30) Peru (28) Mexico (12) Uganda (8) 0 100 200 500 600 700 Bolivia (29) % Difference Species Dominance Bolivia (29) C Panama (38) Site & Source Mexico (12) Peru (27) Malaysia (21) Malaysia (21) Colombia (25) Panama (30) Panama (38) (abiotic) Peru (28) -80-40 0 40 (non-game) Peru (27) % Difference in Diversity Colombia (25) (abiotic) Peru (27)

-75 -50 -25 0 25 50 75 Forest comparison % Difference Species Richness Exclosure study FIGURE 1.10 Diversity responses to defaunation. (a) species richness, (b) species dominance, and (c) species diversity (Shannon-Weiner, Simpson, or Fisher’s α indices). Peruvian richness differences are generally for groups with different dispersal agents; dominance for the Panama study is similarly for subsets of common and rare herbs. Numbers in parentheses denote the study number in Appendix 1.

The studies found important differences in community seed mass responses. In Peru, the relative abundance of large seeded species was lower in defaunated sites, while in Panama, the community mean seed mass was higher in defaunated sites (Wright et al . 2007, Nuñez-Iturri et al . 2008). The discrepancy appears to be due to differences in faunal communities between Peru and Panama. In the Peruvian study, defaunation greatly reduces populations of large primate dispersers, while not greatly affecting the relatively low populations of seed predators such as agoutis. In contrast, defaunation in Panama, defaunation greatly reduces agouti populations, but does not greatly affect seed dispersal by large primates, as these animals are rare or absent even in the non-defaunated sites. Studies of the effects of defaunation effects on other functional traits are rare. Both a forest comparison study in Panama and an exclosure study in Panama found 27 that community median leaf mass per area and leaf laminar toughness in the seedling community did not change in response to defaunation (Ch. 2). However, saplings (40- 100 cm) in the exclosure experiment did have lower community mean leaf toughness, and higher community mean leaf nitrogen relative to controls (Ch. 3). Wood density was the only functional trait that had different responses to defaunation in the forest comparison and exclosure studies. This is likely due to the fact that primary dispersal is altered in the forest comparison, but not in the exclosure study (Ch. 2).

WHAT IS DEFAUNATION ? The heterogeneity in response to defaunation, particularly at the level of plant population dynamics and community composition, suggest that more precision is needed in characterizing exactly what is meant by defaunation, including its drivers and the effects on the particular mammal community being studied. Among the studies reported there, there was variation in the levels of defaunation in both “non- defaunated” and “defaunated” forests across studies. In some cases, sites with similar species compositions would be considered “non-defaunated” in one study, and “defaunated” in another. Unfortunately, terms such as “hunted”, “fragmented,” or “disturbed” are usually used to classify sites into treatments, but do not correspond to a specific change in a vertebrate community. This problem could be addressed by reporting a robust characterization of the vertebrate community. Often, studies did not actually report differences in either abundances of key species (e.g. a particular seed disperser) or the vertebrate community as a whole. In other cases, the presence or absence of species were reported, but not animal abundances. Reporting presence/absence data will not capture situations in which animals such as rodents or smaller primates actually increase in population size as defaunation reduces the predation and competition they experience. In other cases, a species may be present, but in such rare numbers so as to be functionally extinct. Though estimating animal abundances is labor-intensive, we suggest that further clarity and progress in understanding the effects of defaunation on plant 28 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS communities will rely on exactly this level of precision in characterizing sites used in comparative studies.

HETEROGENEITY AMONG STUDIES

Understanding the heterogeneity and sources of potential bias in the studies examined here is important both for understanding the limitations and generality of the data, as well as identifying areas for further research. Research on effects of defaunation on plants has predominantly focused on Neotropical mammals. Over 70 percent of studies were conducted in the Neotropics, while only 15 percent and 10 percent of studies were conducted in Africa and Asia respectively, with the remaining 5 percent conducted in Australia and Polynesia. As this review has demonstrated, responses to defaunation can be quite variable even within the Neotropics, where faunal communities are relatively similar. More studies in the Paleotropics are needed to understand how plant responses to defaunation in these forests may differ from those observed in Neotropical forests. Only 15 percent of studies reported effects of avian defaunation, with the remainder examining effects of mammalian defaunation. Many of those studies focused on rodents and/or primates. Despite the fact that many of the larger vertebrates such as peccaries, pigs, deer, tapir and elephants are most vulnerable to defaunation, only a few studies investigated the effects of losing these animals from the community. Future studies should seek to understand the impacts of losing relatively understudied groups of vertebrates, as well as understanding how losses of dispersers, seed predators, and herbivores may synergize. It is also important to note that forest comparisons generally had low replication. The majority of studies only had one non-defaunated site, and sometimes only one defaunated site. We suggest that a better understanding of patterns and mechanisms could come from increasing the number of study sites examined, and including a range of defaunation intensities within a single study, with better quantification of the differences in vertebrate communities. We suspect that many of 29 the interactions and patterns discussed in this review do not vary linearly with animal abundance, but rather, change in a step-wise or threshold dependent manner as key interaction partners are removed from a system. Studying gradients of defaunation intensity would allow scientists to better elucidate the levels of defaunation at which changes in key interactions occur, and perhaps allow for the reconciliation of discrepancies in responses among studies.

CONCLUSIONS

It is now well documented that larger-seeded species experience reduced dispersal, and increased aggregation around parent trees as a result of defaunation. However the overall consequences for species recruitment appear variable. More work is required to link reduced dispersal to subsequent stages of the seedling recruitment process to determine whether reduced dispersal poses a general threat to the persistence of large- seeded species. Responses to defaunation at the community level are also variable, and future studies should work to reconcile the changes observed that the species interaction or population levels with changes in composition at the community level.

30 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS

APPENDIX 1.1 Reference information and raw data for response variables for studies referenced in figures. Ndef = value for non-defaunated site(s); Def = Value for defaunated site(s); % ∆ = Percent change from non-defaunated site to defaunated site.

No Reference Site Response Variable Ndef Def % ∆ 1 Alves-Costa (2004) Brazil Diaspores predated (%) 31.2 12.6 -60 1 Alves-Costa (2004) Brazil Insect seed predation (near) 10.7 23.4 119 1 Alves-Costa (2004) Brazil Insect seed predation (far) 1.6 24.2 1376 1 Alves-Costa (2004) Brazil Seed density 0.16 0.18 8 1 Alves-Costa (2004) Brazil Seedlings under parent 0.06 0.72 1187 1 Alves-Costa (2004) Brazil Juvenile:Seedling (near) 2.33 1.16 -50 1 Alves-Costa (2004) Brazil Juvenile:Seedling (far) 2.65 9.51 259 1 Alves-Costa (2004) Brazil Seedling density 26.8 46.8 75 1 Alves-Costa (2004) Brazil No. species/No. plants 0.23 0.14 -37 1 Alves-Costa (2004) Brazil Frequency of low herbivory 24.1 40.0 66 1 Alves-Costa (2004) Brazil No. new plants 41.2 118.4 187 1 Alves-Costa (2004) Brazil Seedling mortality (%) 40.1 42.7 6 2 Andresen (2003) Brazil % Pouteria seeds buried 40 32.5 -19 2 Andresen (2003) Brazil % Pourouma seeds buried 36 19 -47 3 Asquith et al. (1997) Panama Seed removal 0.01 0.05 900 3 Asquith et al. (1997) Panama Seedling survival 0.21 0.07 -69 3 Asquith et al. (1997) Panama Seed removal 0.01 0.08 700 3 Asquith et al. (1997) Panama Seed dispersal 0.56 0.24 -58 3 Asquith et al. (1997) Panama Seedling survival 0.49 0.22 -56 3 Asquith et al. (1997) Panama Seed removal 0.00 0.00 0 3 Asquith et al. (1997) Panama Seed dispersal 0.54 0.19 -65 3 Asquith et al. (1997) Panama Seedling survival 0.00 0.00 0 4 Asquith et al. (1999) Venezuela Pod removal 1.00 0.14 -86 4 Asquith et al. (1999) Venezuela Seed removal 0.92 0.60 -35 4 Asquith et al. (1999) Venezuela Seed caching 0.17 0.00 -100 4 Asquith et al. (1999) Venezuela Seed density under parent 4.0 100.0 2400 4 Asquith et al. (1999) Venezuela Seedling density under parent 22.0 1.8 -92 5 Babweteera et al. (2007) Gabon Juvenile density 85 1169 1275 5 Babweteera et al. (2007) Gabon Sapling:Seedling 0.21 0.00 -98 5 Babweteera et al. (2007) Gabon Pole:Seedling 0.08 0.01 -92 5 Babweteera et al. (2007) Gabon Prob. Juvenile's under conspecific 58.0 89.0 53 6 Beckman & Muller-Landau (2007) Panama Primary seed removal 42.0 24.1 -43 6 Beckman & Muller-Landau (2007) Panama Predispersal seed predation 55.0 25.0 -55 6 Beckman & Muller-Landau (2007) Panama Secondary seed removal 0.0 10.6 100 6 Beckman & Muller-Landau (2007) Panama Primary seed removal 21.6 14.6 -32 6 Beckman & Muller-Landau (2007) Panama Predispersal seed predation 7.0 7.0 0 6 Beckman & Muller-Landau (2007) Panama Predispersal seed predation 20.0 0.0 -100 6 Beckman & Muller-Landau (2007) Panama Secondary seed removal 80.0 10.8 -87 7 Brodie et al. (2009) Thailand Proportion fruits under adults 0.19 0.68 252 7 Brodie et al. (2009) Thailand Seeds dispersed 0.15 0.01 -91 8 Chapman & Onderdonk (1998) Uganda Seedling density, non-valley sites 12.5 5.5 -56 8 Chapman & Onderdonk (1998) Uganda Seedling richness, non-valley sites 5.0 3.1 -39 8 Chapman & Onderdonk (1998) Uganda Seedling density, valley bottom 2.5 1.3 -50 31

8 Chapman & Onderdonk (1998) Uganda Seedling richness, valley bottom 1.3 1.0 -23 9 Cordeiro and Howe (2003) Tanzania Seed dispersal 12.8 3.0 -76 Proportion indiv under adults 9 Cordeiro and Howe (2003) Tanzania (<10 m) 57.3 88.5 54 9 Cordeiro and Howe (2003) Tanzania Juvenile:Seedling 0.4 0.3 -27 9 Cordeiro and Howe (2003) Tanzania Seedling abundance 163.7 96.7 -41 10 Cramer et al. (2007) Brazil Seeds dispersed 48.0 16.0 -67 10 Cramer et al. (2007) Brazil Seeds dispersed 5.0 12.0 140 10 Cramer et al. (2007) Brazil Ave. dispersal distance 7.5 3.0 -60 10 Cramer et al. (2007) Brazil Ave. dispersal distance 23.0 18.0 -22 10 Cramer et al. (2007) Brazil Max. dipersal distance 20.0 4.0 -80 10 Cramer et al. (2007) Brazil Max. dipersal distance 17.0 18.0 6 10 Cramer et al. (2007) Brazil Seeds away from parent 27.0 1.0 -96 10 Cramer et al. (2007) Brazil Seeds away from parent 8.0 6.0 -25 11 DeMattia et al. (2004) Costa Rica Seeds destroyed 100.0 93.6 -6 11 DeMattia et al. (2004) Costa Rica Seeds destroyed 68.8 65.6 -5 11 DeMattia et al. (2004) Costa Rica Seeds destroyed 100.0 84.1 -16 11 DeMattia et al. (2004) Costa Rica Seeds destroyed 31.2 15.6 -50 11 DeMattia et al. (2004) Costa Rica Seeds destroyed 93.8 75.0 -20 11 DeMattia et al. (2004) Costa Rica Seeds destroyed 28.0 12.2 -57 11 DeMattia et al. (2004) Costa Rica Seeds destroyed 18.5 24.5 33 11 DeMattia et al. (2004) Costa Rica Seeds cached 25.0 3.1 -88 11 DeMattia et al. (2004) Costa Rica Seeds cached 3.07 9.22 200 12 Dirzo & Miranda (1991) Mexico Herbivory 0.27 0.00 -100 12 Dirzo & Miranda (1991) Mexico Seedling density 22.6 52.8 134 12 Dirzo & Miranda (1991) Mexico Seedling species richness 6.65 2.30 -65 12 Dirzo & Miranda (1991) Mexico Seedling diversity 3.71 1.07 -71 12 Dirzo & Miranda (1991) Mexico Seedling species dominance 14.0 2.0 -86 13 Donatti et al. (2009) Brazil Seeds predated 0.18 0.03 -83 13 Donatti et al. (2009) Brazil Seeds cached 0.33 0.13 -61 14 Fadini et al. (2009) Brazil Seed dispersal 1.6 0.9 -42 14 Fadini et al. (2009) Brazil Seed predation 99.7 7.4 -93 14 Fadini et al. (2009) Brazil Seedling density 1000 80 -92 15 Farwig et al. (2006) Kenya seed removal 0.97 1.57 62 16 Fleury & Galetti (2006) Brazil Seed predation 15.8 48.0 203 16 Fleury & Galetti (2006) Brazil Seed predation 15.8 0.0 -100 17 Galetti et al. (2006) Brazil Seed predation (%) 6.9 0.3 -96 17 Galetti et al. (2006) Brazil Seed caching (%) 12.2 3.4 -72 17 Galetti et al. (2006) Brazil Seed predation (%) 38.4 2.0 -95 17 Galetti et al. (2006) Brazil Seed predation (%) 34.4 69.6 102 17 Galetti et al. (2006) Brazil Seedling density 148.2 34.5 -77 17 Galetti et al. (2006) Brazil Seedling:Adult 6.96 0.35 -95 17 Galetti et al. (2006) Brazil Sapling:Seedling 1.98 10.07 409 18 Guariguata et al. (2000) Costa Rica Secondary seed removal 0.41 0.13 -67 18 Guariguata et al. (2000) Costa Rica Secondary seed removal 0.31 0.37 20 18 Guariguata et al. (2000) Costa Rica Secondary seed removal 0.88 0.94 7 18 Guariguata et al. (2000) Costa Rica Secondary seed removal 0.59 0.91 56 18 Guariguata et al. (2000) Costa Rica Secondary seed removal 0.66 0.88 33 18 Guariguata et al. (2000) Costa Rica Secondary seed removal 0.87 0.38 -56 18 Guariguata et al. (2000) Costa Rica Secondary seed removal 0.98 0.40 -59 32 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS

18 Guariguata et al. (2000) Costa Rica Secondary seed predation 0.41 0.13 -67 18 Guariguata et al. (2000) Costa Rica Secondary seed predation 0.31 0.37 20 18 Guariguata et al. (2000) Costa Rica Secondary seed predation 0.77 0.91 18 18 Guariguata et al. (2000) Costa Rica Secondary seed predation 0.55 0.91 64 18 Guariguata et al. (2000) Costa Rica Secondary seed predation 0.57 0.78 38 18 Guariguata et al. (2000) Costa Rica Secondary seed predation 0.73 0.38 -48 18 Guariguata et al. (2000) Costa Rica Secondary seed predation 0.95 0.40 -58 18 Guariguata et al. (2000) Costa Rica Seed caching 0.00 0.00 0 18 Guariguata et al. (2000) Costa Rica Seed caching 0.00 0.00 0 18 Guariguata et al. (2000) Costa Rica Seed caching 0.03 0.11 300 18 Guariguata et al. (2000) Costa Rica Seed caching 0.01 0.03 400 18 Guariguata et al. (2000) Costa Rica Seed caching 0.10 0.09 -7 18 Guariguata et al. (2000) Costa Rica Seed caching 0.00 0.13 100 18 Guariguata et al. (2000) Costa Rica Seed caching 0.00 0.03 100 18 Guariguata et al. (2000) Costa Rica Tertiary seed removal/predation 0.50 0.81 63 18 Guariguata et al. (2000) Costa Rica Tertiary seed removal/predation 1.00 0.80 -20 18 Guariguata et al. (2000) Costa Rica Tertiary seed removal/predation 1.00 1.00 0 18 Guariguata et al. (2000) Costa Rica Seedling survival 0.00 0.07 100 19 Guariguata et al. (2002) Costa Rica Seeds dispersed 0.11 0.03 17 19 Guariguata et al. (2002) Costa Rica Seeds dispersed 0.00 0.02 -67 20 Holbrook & Loiselle (2009) Ecuador Seeds removed (primary) 89.4 66.8 -25 21 Ickes, K., et al. 2001. Malaysia Seedlings 75.5 117.5 56 21 Ickes, K., et al. 2001. Malaysia Sapling density 107.1 142.3 33 21 Ickes, K., et al. 2001. Malaysia Mortality 10.0 8.7 -13 21 Ickes, K., et al. 2001. Malaysia Recruits 14.9 48.8 228 21 Ickes, K., et al. 2001. Malaysia Species richness 50.6 55.7 10 21 Ickes, K., et al. 2001. Malaysia Diversity 45.8 39.7 -13 21 Ickes, K., et al. 2001. Malaysia Liana proportion of stems 25.9 30.5 18 22 Kirika et al. (2008) Uganda Seed dispersal 14124 3714 -74 23 Kurten (2010a) Panama seed mass 0.21 0.32 58 23 Kurten (2010a) Panama SLA 20.0 22.4 12 23 Kurten (2010a) Panama Leaf toughness 1.49 1.49 0 23 Kurten (2010a) Panama Wood density 0.65 0.64 -2 24 Kurten (2010b) Panama Secondary seed predation 0.50 0.81 63 24 Kurten (2010b) Panama Secondary seed predation 0.90 0.85 -5 24 Kurten (2010b) Panama Secondary seed predation 0.53 0.11 -79 24 Kurten (2010b) Panama Secondary seed predation 0.00 0.00 0 24 Kurten (2010b) Panama Seed caching 1.16 1.33 15 24 Kurten (2010b) Panama Seed caching 0.83 1.16 40 25 Lizcano (2006) Colombia Richness 8.02 11.24 40 25 Lizcano (2006) Colombia Shannon Index 0.82 0.75 -9 26 McConkey & Drake (2006) Tonga Seed dispersal 0.34 0.03 -90 27 Nuñez-Iturri et al. (2007) Peru Sapling richness - total 47.4 50.4 6 27 Nuñez-Iturri et al. (2007) Peru Sapling richness -large mammal 16.6 7.8 -53 27 Nuñez-Iturri et al. (2007) Peru Sapling richness - nongame 27.1 36.6 35 27 Nuñez-Iturri et al. (2007) Peru Sapling richness - abiotic 3.70 6.00 62 27 Nuñez-Iturri et al. (2007) Peru Sapling density - total 78.3 90.0 15 27 Nuñez-Iturri et al. (2007) Peru Sapling density -large mammal 26.8 11.3 -58 27 Nuñez-Iturri et al. (2007) Peru Sapling density - nongame 45.5 67.8 49 33

27 Nuñez-Iturri et al. (2007) Peru Sapling density - abiotic 5.91 10.98 86 28 Nuñez-Iturri et al. (2008) Peru Species richness (primate disp) 1.60 0.60 -63 28 Nuñez-Iturri et al. (2008) Peru Species richness (abiotic disp) 1.40 1.80 29 28 Nuñez-Iturri et al. (2008) Peru Species richness 145 100 -31 28 Nuñez-Iturri et al. (2008) Peru Density (primate disp) 1.90 0.90 -53 28 Nuñez-Iturri et al. (2008) Peru Density (abiotic) 3.60 11.30 214 29 Roldan & Simonetti (2001) Bolivia Seed predation 0.43 0.64 49 29 Roldan & Simonetti (2001) Bolivia Trampling survival 0.78 0.90 15 29 Roldan & Simonetti (2001) Bolivia Seedling density 4.21 3.92 -7 29 Roldan & Simonetti (2001) Bolivia Species richness 35.0 34.0 -3 29 Roldan & Simonetti (2001) Bolivia Seedling spp:Adult spp 0.61 0.59 -5 29 Roldan & Simonetti (2001) Bolivia Seedling composition 4.21 3.92 -7 30 Royo & Carson (2005) Panama Species richness 16.8 13.1 -22 30 Royo & Carson (2005) Panama Herb Density 1.21 1.19 -2 30 Royo & Carson (2005) Panama Density dominants 0.42 0.35 -17 30 Royo & Carson (2005) Panama Density rare 0.34 0.16 -53 30 Royo & Carson (2005) Panama Cover 7.90 5.90 -25 30 Royo & Carson (2005) Panama Cover of dominants 2.65 1.50 -43 31 Sethi & Howe (2009) India Total seedlings 2.55 0.76 -70 31 Sethi & Howe (2009) India Seedlings under parent 0.37 0.60 64 31 Sethi & Howe (2009) India Seedlings away from parent 2.19 0.16 -93 31 Sethi & Howe (2009) India Total seedlings 2.17 0.81 -62 31 Sethi & Howe (2009) India Seedlings under parent 0.99 0.35 -65 31 Sethi & Howe (2009) India Seedlings away from parent 1.18 0.46 -61 31 Sethi & Howe (2009) India Total seedlings 4.16 0.81 -80 31 Sethi & Howe (2009) India Seedlings under parent 2.96 0.58 -81 31 Sethi & Howe (2009) India Seedlings away from parent 1.20 0.24 -80 32 Sork (1987) Panama Secondary seed predation 0.94 0.41 -56 32 Sork (1987) Panama seedling survival 0.69 0.97 42 Peru/ 33 Terborgh & Wright (1994) Panama Seed predation 0.35 0.82 136 Peru/ 33 Terborgh & Wright (1994) Panama Seedling density 38.3 29.3 -23 Peru/ 33 Terborgh & Wright (1994) Panama Seedling recruitment 0.25 0.96 276 Peru/ 33 Terborgh & Wright (1994) Panama Seedling mortality 0.20 0.29 43 34 Terborgh et al. (2008) Peru Recruit. Bird dispersed 0.23 0.24 4 34 Terborgh et al. (2008) Peru Recruit. Bat dispersed 0.10 0.11 16 34 Terborgh et al. (2008) Peru Recruit. Lg. primate dispersed 0.25 0.19 -24 Recruit. Sm. arboreal mammal 34 Terborgh et al. (2008) Peru dispersed 0.33 0.24 -26 Recruit. Terrestrial mammal 34 Terborgh et al. (2008) Peru dispersed 0.04 0.02 -50 34 Terborgh et al. (2008) Peru Recruit. Wind disp. 0.05 0.17 278 35 Wang et al. (2007) Cameroon Diaspore under adult 10.0 50.0 400 35 Wang et al. (2007) Cameroon Seed dispersal 0.98 0.58 -41 36 Wright et al. (2000) Panama Seed predation 90.8 47.7 -47 36 Wright et al. (2000) Panama Seed predation 0.03 0.34 1033 36 Wright et al. (2000) Panama Seed density under parent 180 2749 1427 36 Wright et al. (2000) Panama Seed density away from parent 2.65 1.79 -32 36 Wright et al. (2000) Panama Prop. seeds dispersed 0.93 0.39 -58 34 CHAPTER 1: CASCADING EFFECTS OF DEFAUNATION ON TROPICAL PLANTS

36 Wright et al. (2000) Panama Seedling density 0.0 0.0 300 36 Wright et al. (2000) Panama Seedlings:Adult 11.0 28.9 163 36 Wright et al. (2000) Panama Seed predation 91.7 21.7 -76 36 Wright et al. (2000) Panama Seed predation 0.1 0.3 460 36 Wright et al. (2000) Panama Seed density under parent 910 1290 42 36 Wright et al. (2000) Panama Seed density away from parent 4.4 2.8 -37 36 Wright et al. (2000) Panama Prop. seeds dispersed 0.9 0.1 -93 36 Wright et al. (2000) Panama Seedling density 0.1 0.2 129 36 Wright et al. (2000) Panama Seedlings:Adult 113.8 92.5 -19 37 Wright et al. (2001) Panama Seed density under parent tree 24.6 35.7 46 37 Wright et al. (2001) Panama Seed density away from parent 0.9 0.2 -81 tree 37 Wright et al. (2001) Panama Seedling density under parent 0.0 0.1 282 Seedling density away from 37 Wright et al. (2001) Panama parent 0.0 0.1 282 37 Wright et al. (2001) Panama Seedling density (total) 0.0 0.1 282 37 Wright et al. (2001) Panama Seed predation-rodents 0.8 0.4 -46 37 Wright et al. (2001) Panama Seed predation-bruchid 0.1 0.4 266 38 Wright et al. (2007) Panama Community mean seed mass 90 100 11 38 Wright et al. (2007) Panama Prop. seedlings (hunted disperser) 0.7 0.4 -40 38 Wright et al. (2007) Panama Prop. seedlings (unhunted 0.2 0.4 75 disperser) 38 Wright et al. (2007) Panama Proportion lianas 0.1 0.4 150 38 Wright et al. (2007) Panama Seedling density 15.2 9.4 -38 38 Wright et al. (2007) Panama Seedling Richness 70.7 69.8 -1 38 Wright et al. (2007) Panama Seedling S-W Diversity 2.6 3.2 23

Chapter 2

Reduced seed dispersal as a consequence of hunting lowers community-level wood density in a Neotropical forest

Authors: Erin L. Kurten, S. Joseph Wright, Andrés Hernandéz , Walter P. Carson

Abstract

Defaunation alters interactions between plants and their seed dispersers and predators, and is sometimes associated with decreased diversity at the community level. A more cryptic plant response to defaunation is an altered distribution of life histories and functional traits in the community. In this study, we censused 12,769 seedlings in a long-term terrestrial mammal exclosure experiment (LTEE) in Central Panama to examine how vertebrate seed dispersers and seed predators shape relative abundances of species classified by dispersal mode, growth forms, seed mass, leaf mass per area (LMA), leaf toughness, and wood density. We then compared exclosure results with data from a natural comparison of 38,250 seedlings in nearby hunted and unhunted sites. The exclosure experiment excludes terrestrial seed predators and herbivores but not seed dispersers such as primates, birds, and bats. In contrast, the hunting 36 CHAPTER 2: CONSEQUENCES OF HUNTING FOR SEED DISPERSAL AND WOOD DENSITY comparison reduces the abundance of all three groups. We tested the hypotheses that (1) where seed dispersal is disrupted, species with unhunted dispersal modes, and lianas, which are abiotically dispersed, are favored, (2) where seed predation is reduced, larger-seed plant species are favored, and (3) where herbivory is reduced, species with lower LMA, leaf toughness, and wood density are favored. Consistent with our expectations, dispersal-driven changes in the relative abundance of growth forms or dispersal mode classes were not observed in the terrestrial mammal exclosure experiment, while higher median seed mass as a consequence of reduced seed predation was observed. In contrast to our hypothesis, leaf functional traits did not differ as a consequence of defaunation either in the forest comparison or in the LTEE. Interestingly, wood density responses differed between our two experiments. Species dispersed by mammals and large birds had significantly higher mean wood density than wind dispersed species, suggesting that seed dispersal, rather than seedling herbivory, may mediate decreases in community median wood density in hunted forests.

Introduction

Over-hunting has caused severe declines, and even local extirpations, of hunted wildlife populations in tropical forests throughout Latin America (Peres & Palacios 2007), Africa (Fa & Brown 2009), and Asia (Corlett 2007). Defaunation perturbs plant interactions with the frugivores, granivores, and herbivores that play critical roles in establishing the composition and spatial distribution of seedlings in tropical forest seedling banks. Reduced seed dispersal in hunted forests results in higher proportions of undispersed seeds under parent trees of species reliant on hunted biotic dispersal agents (Wright & Duber 2001, Cordiero & Howe 2003, Babweteera et al 2007, Sethi & Howe 2009). Vertebrate seed predation rates may increase or decrease, depending on the severity of defaunation and the size of the seeds and seed predators affected (Guariguata et al . 2000, Galetti et al . 2006, Dirzo et al. 2007). Vertebrate herbivory rates can also be drastically reduced (Dirzo and Miranda 1991). The net effect of these perturbations can be the loss of plant diversity (Dirzo and Miranda 37

1991, Chapman & Onderdonk 1998), with up to 66% of species failing to recruit in the most extreme case documented. Other important changes may also be occurring in the plant community even if species richness is unaffected. Species with hunted dispersal agents, such as primates or large birds, consistently decrease in abundance in hunted forests, while species with abiotic or unhunted dispersal agents increase (Wright et al . 2007b, Nuñez-Iturri et al . 2008, Terborgh et al . 2008). Lianas, the majority of which are abiotically dispersed in this forest (Hardesty & Muller-Landau 2005), also increase in abundance in hunted sites (Wright et al . 2007b). Where dispersal of large, primate-dispersed seeds is disrupted by local extirpation of ateline monkeys, the relative abundance of those plant species is lower than in sites where ateline monkeys are present (Nuñez-Iturri et al . 2008). In sites where hunting decreases rodent seed predator populations, the median seed mass of the seedling community is higher than in unhunted sites (Wright et al. 2007b). In this study, our first goal was to move beyond investigations of dispersal mode classes and seed mass to examine the effect of defaunation on other functional traits. Lower abundances of browsers and other seedling predators in hunted sites should lead to reduced vertebrate herbivory and seedling predation. Leaf mass per unit area (LMA), low leaf laminar toughness (hereafter referred to as “toughness”), and wood density are correlated with palatability to vertebrates or seedling survival (Wardle et al. 2002, Alvarez-Clare & Kitajima 2007). We hypothesized that species with low leaf area per unit mass, low leaf laminar toughness, and lower wood density should be favored in hunted sites. A weakness of many forest comparison studies in defaunation work is that they lack replication of hunted and unhunted forests. Even when there is replication, differences in climate, soils, forest age, fragmentation or other anthropogenic factors may covary with hunting intensity in comparative studies, making it difficult to attribute the patterns observed to hunting alone. Therefore, a second goal of this study was to validate results from forest comparisons with data from a long-term exclosure experiment (LTEE). We conducted this study in Central Panama, where a hunting 38 CHAPTER 2: CONSEQUENCES OF HUNTING FOR SEED DISPERSAL AND WOOD DENSITY defaunation gradient has been previously documented (Wright et al. 2007b). Within the unhunted, protected sites, a terrestrial mammal exclosure experiment has been ongoing for the past fifteen years (Royo & Carson 2005). Because hunting typically disrupts multiple plant-animal interactions simultaneously, it can be difficult to determine the relative influence of lost seed dispersers, seed predators, or herbivores on resulting seedling communities in comparative studies. A third goal of this study was to improve the extent to which changes in the seedling community could be attributed to changes in particular interactions. Unlike the hunted forest comparison, our LTEE excludes terrestrial seed predators and herbivores from exclosure plots, but does not alter the abundance of primary dispersal agents such as primates and most large birds. By comparing results between the forest comparison and exclosure approaches, we can identify which changes in the seedling community are due to changes in seed dispersal and which are due to changes in seed or seedling predation. . We predicted that seedling communities in exclosures would have higher median seed massas a consequence of lower seed predation, consistent with the forest comparison (Wright et al. 2007). In contrast to the forest comparison experiment, we predicted the LTEE would not show differences in the relative abundances of dispersal mode classes or growth forms. We expected that median LMA, toughness and wood density would be lower in exclosure plots relative to paired plots open to terrestrial mammals, and in hunted forests relative to unhunted forests, as a consequence of reduced herbivory.

Methods

Study sites. The Barro Colorado National Monument (BCNM) is a group of islands and mainland peninsulas located in Lake Gatun in the Republic of Panama. Because of forest guard patrol and researcher presence, activity of poachers is virtually nonexistent on Barro Colorado Island (BCI) itself and is highly reduced on Gigante (GIG) and other peninsulas. The Parque Nacional Soberanía (PNS) is a protected area 39 contiguous with the BCNM, which is not actively patrolled. Intensity of poaching within the PNS varies with accessibility, but is much greater than within the BCNM (Wright et al. 2000).

Exclosure experimental design. Between late-1993 and mid-1994, eight pairs of fenced, exclosure plots and open, control plots were established within the BCNM (Royo & Carson 2005). Paired fenced and unfenced plots are 30 m x 45 m and are approximately 5 m apart. Exclosure fences are constructed of 12.7 x 12.7 cm galvanized steel fencing 2.2 m tall and buried 0.25 m deep. A secondary 1.3 x 1.3 cm mesh surrounds the lower 70 cm and also extends 0.25 m below ground. Twenty- eight 0.5 x 0.5 m subplots were established in each of the 16 plots in a stratified random fashion to census seedlings. To avoid fence effects, no subplots were established within a 5-7 m buffer zone of a plot edge. Within each subplot, each plant less than 50 cm in height was tagged, measured, and identified. Seedling census data were collected in 2006.

Exclosure effectiveness. To assess differences in the animal communities between open and exclosure treatments, Reconyx RC-55 infrared cameras (Reconyx, Inc., Holmen, WI) were used to monitor mammal activity in the 16 plots between August and October 2008. Each plot was monitored at 6-7 locations, for a total of 41 camera trap days per plot. A camera trigger was counted as a new “visit” if (1) it was a different species than the prior trigger, or (2) if 60 minutes had elapsed between triggers (Di Bitetti et al. 2008). We chose this method of quantification because unmarked individuals of most mammals in this community cannot be distinguished by photograph, and because this measure of activity likely describes animal impact on each plot more accurately than abundance. Because it was difficult to accurately count individuals for social animals such as peccaries (Tayassu tajacu ) and coatis (Nasua narica ), activity by these species was analyzed as the number of visits by social groups rather than visits by individuals.

40 CHAPTER 2: CONSEQUENCES OF HUNTING FOR SEED DISPERSAL AND WOOD DENSITY

Hunting intensity comparison. At 9 protected sites (BCNM) and 11 intensely hunted sites (PNS), all adult trees larger than 20 cm diameter at breast height (dbh) within a 1 ha plot were mapped and identified by Condit et al. (2002) and Wright et al. (2007b). Between June and December 2004, all woody plants less than 50 cm tall in an 8 x 8 m subplot were censused by Wright et al. (2007b). Seedling plots were at the center of each tree plot, or as close as possible in three plots in which tree falls occurred between adult and seedling censuses. A total of 38250 individuals comprising 312 identified species were recorded (Wright et al. 2007b).

Plant traits. Wright et al. (2010) determined life history traits, dry seed mass, leaf mass per area (LMA), laminar toughness, and wood density for species in central Panama. In this study, dry seed mass was measured as endosperm plus embryo mass after oven drying to constant mass at 60ºC. Leaf mass per unit area was measured on shade leaves as the mass of a standard-sized leaf disc after oven drying to constant mass at 60ºC divided by the area of the disc. Laminar toughness was measured as the force necessary to cut across a leaf, perpendicular to the midvein (Lucas et al . 2000). Wood specific gravity, or mass after drying at 100ºC per unit volume, was used as the unit of wood density. Data on dispersal agents have been assembled from published studies and long-term personal observations.

Analyses . We used a MANOVA to test whether mammal activity differed between open and control plots for each vertebrate species observed. Paired t-tests were used to test for differences in the relative abundance of growth forms, dispersal mode classes, and functional traits. For the hunting comparison, an ANCOVA was used to determine whether the correspondence of functional traits between seedling and adult communities differed with hunting. Statistical analyses were performed with SYSTAT © 11.0 (Richmond, CA, U.S.A.) and R version 2.7.2 (R Development Core Team, 2008).

Results

EXCLOSURE EFFECTIVENESS . Fenced plots excluded non-climbing, terrestrial granivores and herbivores from exclosures, including agoutis ( Dasyprocta punctata ), paca ( Agouti paca ), deer ( Mazama Americana and Odocoileus virginianus ), and peccaries ( T. tajacu ) (Fig. 1). Most climbing animals appeared to be undeterred by the exclosure treatment. Spiny rats ( Proechimys steerei ) showed higher activity in exclosures than controls, perhaps responding to a lack of competition for food, greater cover, or a reduction in ocelet abundance (Felis pardalis ) in the exclosures. Most birds were not affected by the exclosures, a notable exception being tinamous, which are seed predators (Erand et al . 1991).

0 Agouti ***

0 20 40 60 80 100

Anteater 0 Armadillo 0 Capuchin Coati† 0 Deer 0 Mouse 0 Ocelot * Opposum * 0 Paca ** 0 Peccary† ** 0 Rabbit Spiny Rat Squirrel

0 Tinamou Birds(other)

0 2 4 6 8 10 12 14

Visits/plot

FIGURE 2.1. Total visits per plot by species for control plots (dark grey) and exclosure plots (light grey). Zeros denote no individuals observed in exclosures. *** p < 0.001, ** p < 0.01, * p < 0.05, p < 0.1; † groups, not individuals 42 CHAPTER 2: CONSEQUENCES OF HUNTING FOR SEED DISPERSAL AND WOOD DENSITY

DISPERSAL AGENTS & LIFEFORMS . A total of 12,769 seedlings of 322 species were recorded in the 2006 LTEE seedling census. Of those, dispersal agents were known for 247 species (76.7%) and 12575 stems (98.5%), and growth form was known for 241 species (74.8%) and 12468 stems (97.6%). Consistent with our expectations, there was no difference in the proportions of seedlings with different dispersal agents (hunted: t = 1.161, P = 0.2651; unhunted t = 1.861, P = 0.089) (Fig. 2) or growth forms (free-standing: t = 1.7131, P = 0.1096) (Fig. 3) in the LTEE.

Hunted Mixed Unhunted Proportion of seedlings of Proportion 0.0 0.2 0.4 0.6 0.8 1.0

Open Excl Open Excl Open Excl

Treatment by Disperser

FIGURE 2.2. Proportion of seedlings by dispersal mode class is not significantly different between treatments in exclosure experiment.

Climbing Free Standing Proportion of seedlings of Proportion 0.0 0.2 0.4 0.6 0.8 1.0

Open Excl Open Excl

Treatment by Growth form

FIGURE 2.3. Proportion of seedlings by life form is not significantly different between treatments in exclosure experiment. 43

SEED MASS . In the LTEE, seed dry mass was known for 166 (51.6 %) of species and 10866 (85.1%) of stems. Exclosures removed all but the smallest seed predators. The reduction in agouti activity was particularly dramatic (Fig. 4). Consistent with these findings, as well as results from the forest comparison (Wright et al . 2007b), median seed mass was 44% higher in exclosure plots, relative to controls ( t = 2.315, P = 0.02261). Diaspore Dry mass (mg) mass Dry Diaspore 100 150 200 250

Open Excl

Treatment

FIGURE 2.4. (A) Median seed mass is significantly higher in exclosures ( p < 0.02261).

LEAF TRAITS . LMA and laminar toughness were known for 11,173 stems (87.5%) and 10,904 stems (85.4%) respectively in the LTEE, and for 33,839 stems (87.5 %) and 29,962 stems (77.5 %) respectively in the hunting experiment. Exclosure and hunting treatments were expected to reduce the number of terrestrial vertebrate seedling predators in a way analogous to seed predators. A reduction in vertebrate herbivores was expected to favor species with low LMA and small laminar toughness. Contrary to expectation, neither the exclosure nor the hunting experiment showed significant differences in median LMA or laminar toughness, nor was there a significant difference between protected and hunted sites when the correspondence 44 CHAPTER 2: CONSEQUENCES OF HUNTING FOR SEED DISPERSAL AND WOOD DENSITY between seedling and adult communities was evaluated by ANCOVA (data not shown).

WOOD DENSITY . Adult wood specific gravity was known for 208 species (64.6% %) and 11,620 stems (91.0 %) in the LTEE, and for 239 species (63.2 %) and 28,488 stems (73.3 %) in the hunting experiment. Stem tissue density is strongly positively correlated with seedling survival in this forest (Alvarez-Clare & Kitajima 2007). If greater stem tissue densities confer a protective effect against herbivores, we would expect higher wood density species to be favored in open plots relative to exclosures, and protected sites relative to hunted sites. Median wood densities were greatly reduced in hunted site seedling communities, relative to both protected site seedling communities (pooled variance t=3.86, P = 0.0011, one-tailed test) (Fig. 5a), and their surrounding adult communities (F 1,18 = 16.2, p < 0.001) (Fig. 5b). However, the LTEE showed no differenced in median wood density ( t = 0.024, P = 0.49) (Fig. 5a).

) 0.65 AA B Hunting Exclosure 0.60

0.55

0.50 0.55 0.60 0.65 Wood SpecificGravity (g/cm3) 0.45 0.65 0.50 0.45 0.50 0.55 0.60 Wood density of tree seedlings (g/cm3 Open Excl Protect Hunt Wood density of canopy trees (g/cm3) Treatment

FIGURE 2.5. (A) Median wood density is not significantly different between treatments in the exclosure experiment (p = 0.0238), but is significantly lower in hunted sites relative to protected sites (p = 0.0011). White and grey boxes indicate analogous treatments (B) Seedling median wood density is lower than adult median wood density for tree species in hunted sites (dark circles) but not in protected sites (light triangles) (F 1,18 = 16.2, p < 0.001). The dashed line represents is the 1:1 line at which adult and seedling communities have the same median values.

We tested whether differences in primary dispersal may be responsible for the discrepancy in wood density responses, as it was for differences in dispersal mode and growth form. We used an ANOVA to compare mean wood densities across dispersal 45 mode classes. Interestingly, species with unhunted dispersal agents had significantly lower mean wood density than either species with hunted dispersal modes, or species dispersed by agents in both categories (ANOVA, F 2, 277 , p = 0.0017) (Fig. 6a). Specifically, wind dispersed species had significantly lower wood densities compared to both mammal and large bird dispersed species (ANOVA, F 6, 461 , p = 0.0057) (Fig. 6b).

0.7 0 . a B ab A a a a b 0.6 ab ab ab b 0.5

0.4

0.3

0.2 Mean Wood SpecificGravity (g/cm3)

Mean Wood SpecificGravity (g/cm3) 0.1 0.0 0.2 0.4 0.6 0.8 1

Hunted Mixed Unhunted 0.0 Bat Bird L.Bird Explo Mam Water Wind Dispersal Mode Class Dispersal Agent

FIGURE 2.6. (A) Species with unhunted dispersal agents have lower WSGs. (B) Wind dispersed species have significantly lower mean wood density than both mammal and large bird dispersed species. Bars with same letter were not significantly different in post hoc comparisons.

Discussion

Hunting, vertebrate abundance, and plant functional diversity

This study demonstrates that hunting effects change in the seedling community of tropical forests primarily by perturbing biotic seed dispersal and vertebrate seed predation. Increases in lianas and reductions in species with hunted biotic dispersal agents were observed in the hunting experiment, where primary dispersers were reduced (Wright et al. 2007b), but not in the LTEE, where primary dispersers were not manipulated (Figs. 2 and 3). Activity of seed predators was dramatically reduced in the LTEE, and here we observed increases in median seed mass (Fig. 4). This result is 46 CHAPTER 2: CONSEQUENCES OF HUNTING FOR SEED DISPERSAL AND WOOD DENSITY consistent with larger-seed species experiencing an escape from vertebrate seed predation (Dirzo et al., 2007), and lends strong support to the view that changes in seed predator abundance are responsible for the increases in median seed mass observed in hunted forests relative to protected forests (Wright et al. 2007b). Leaf traits showed no differences in both the LTEE and the hunting experiment. This suggests that the reduction of vertebrate herbivores by hunting does not have a strong impact on changes in seedling community composition at early stages of plant recruitment. Wood density responses to herbivore reduction differed between the LTEE and hunting experiment. The likeliest explanation for the discrepancy is that, as with differences in dispersal mode and growth form responses between our experiments, differences in wood density responses are dispersal-driven. To test this post hoc hypothesis, we evaluated whether dispersal mode classes differ in their mean wood densities. Consistent with our hypothesis that experimental differences in wood density responses are due to differences in dispersal, wood density was significantly lower for species with unhunted dispersal agents than for species having either hunted dispersal agents, or a combination of the two (Fig. 6). In contrast to the hunting experiment, the LTEE shows higher spiny rat activity in the defaunated treatment. As spiny rats are known to be significant seedling predators in the absence of other competitors (Asquith et al. 1997), it is also possible that they may be compensating for the reduction in other seedling predators inside the exclosures. However several lines of evidence contradict this putative explanation. First, if spiny rats were compensating for other herbivores inside the exclosure as in the study of Asquith et al. (1997), we would expect overall seedling densities to be similar in open and exclosure plots, but instead, seedling densities in exclosures are about 2-fold higher than in open plots (data not shown). The herbivory study on islands which indicated that spiny rats alone can achieve seedling predation rates as high as communities with relatively intact mammal fauna were conducted on islands in which spiny rat densities are among the highest rodent densities ever documented (Adler 1996), but our study does not replicate these high densities. In addition, 47 herbivory by spiny rats cannot explain why wood density responses differ between experiments, but leaf traits did not.

Our evidence for a link between seed dispersal and community wood density is an interesting result because Brodie and Gibbs (2009) recently proposed that reductions in seed dispersal as a consequence of defaunation may result in shifts in mean wood density at the community level. However, the mechanism that they proposed assumed that seed mass was positively correlated with wood density across species in tropical forests, which is true neither generally, nor for our site specifically (Wright et al. 2010). Here we show that dispersal likely mediates changes in community wood density distributions via an alternative mechanism, namely that low and high wood density plant species are disportionately associated with dispersal mode agents that are favored and disfavored respectively by hunting.

Implications of defaunation for carbon sequestration in tropical forests.

Our results suggest that there may be important links between hunting, frugivore abundance, and the ability of future tropical forests to sequester carbon. Disruption of seed dispersal by game animals has the potential to alter the ability of forests to store carbon by (1) decreasing the abundance of trees relative to lianas (Wright et al. 2007b), and (2) shifting wood density distributions in the community toward lower wood density species (Fig 5a). Higher liana relative abundance may reduce carbon storage at the ecosystem level in two ways. Lianas compete with trees for light, displace tree leaf mass (Kira & Ogawa 1971), and lower tree growth. One recent study from Amazonia estimated that above ground competition from lianas was responsible for reductions in tree growth that reduced carbon sequestration by the equivalent of 0.25 Mg C ha -1 yr -1 (van der Heijden & Phillips, 2009). Growth of the lianas themselves amounted to 0.07 Mg C ha -1 yr -1, which did not compensate for reductions in tree growth. The authors 48 CHAPTER 2: CONSEQUENCES OF HUNTING FOR SEED DISPERSAL AND WOOD DENSITY therefore estimated that competition from lianas reduced carbon sequestration at the ecosystem level by net amount of 0.18 Mg C ha -1 yr -1 or 71%. The consequences of lowering community mean wood density for carbon storage depend on the relationship between wood density and tree volume in a particular forest. Three Panamanian forests in the region, including BCI, exhibit a weak, but negative relationship between plot mean wood density and plot biomass (Stegen et al. 2009). This suggests that the reduction in community wood density in our study site may actually lead to higher forest biomass, and partially compensate for any reduction resulting from the relative reduction in free-standing stems. However, forests exhibiting positive wood density-biomass relationships could experience a decrease in carbon sequestration potential under the same scenario of reduced plot wood density. The phenomenon of over-hunting in tropical forests is widespread and severe. If the changes in carbon storage potential represented by shifting seedling communities in this study are as wide-spread as over-hunting itself, the potential impact may be great. However, many questions must be answered before we can understand the magnitude of the impact that hunting may have on the ability of tropical forests to sequester carbon. First, to our knowledge, this is the first study demonstrating that tropical tree wood densities are associated with dispersal mode classes, and in a way that would respond to hunting. Studies investigating the generality of this association would represent the first step in predicting whether the changes observed in our site would apply pan-tropically. Secondly, here we report changes in the seedling community, but better models of forest dynamics are needed to understand the extent to which those changes will propagate to the adult community. Third, how lianas alter carbon storage at the ecosystem level is poorly understood, and in particular we do not understand how liana competition with trees varies as a function of liana abundance. Finally, as noted by Stegen et al. (2009), there are many unanswered questions pertaining to how changes in wood density distributions translate to changes in biomass at the plot level. 49

Hunting is not the only cause of defaunation. Frugivores, granivores, and herbivores may also decrease in abundance as a result of habitat fragmentation and loss (Asquith et al. 1999, Galetti et al. 2006, Terborgh et al. 2006, Cramer et al. 2007). As with hunting, decreases in abundance with intensity of habitat loss disproportionately affect larger-bodied animals. Therefore, even in the absence of hunting, other sources of defaunation may be expected to have similar indirect effects on plant community functional trait composition and ecosystem services in tropical forests.

Acknowledgements

We extend our gratitude to the Autoridad Nacional del Ambiente (ANAM) for permitting us to conduct research within the Parque Nacional Soberanía. David Brassfield, Didimo Ureña, and Joana Balbuena conducted the seedling censuses in the hunting experiment. A competitive research grant from the Smithsonian Tropical Research Institute funded seedling censuses in hunted and protected forests. The Smithsonian Tropical Research Institute, the U.S. Department of Defence Legacy Fund, and the U.S. Agency for International Development funded canopy tree censuses at hunted and protected forests. NSF grant DEB 9527729 funded the LTEE. NSF DEB-0808338, the Theresa Heinz Environmental Scholars program, and the STRI short-term fellowship program provided support to E. Kurten.

50 CHAPTER 2: CONSEQUENCES OF HUNTING FOR SEED DISPERSAL AND WOOD DENSITY

Chapter 3

Terrestrial mammalian herbivores influence the distribution of defense and nutrient traits in a Neotropical forest

Erin. L. Kurten, Walter P. Carson

ABSTRACT

Mammalian herbivores are known to be important regulators of plant community composition in tropical savannas and temperate systems. However, their impact on tropical communities beyond their impact as seedling predators has remained largely unexplored, perhaps because of the assumed importance of insect herbivory for tropical plants beyond the seedling stage. In this study, we censused 35,069 saplings in a long-term terrestrial mammal exclosure experiment in Central Panama to determine whether mammalian herbivores had the ability to regulate the abundances of species differing in wood density, specific leaf area (SLA), leaf toughness, and leaf nitrogen. We found that community mean leaf toughness decreased in exclosures over time, while community mean leaf nitrogen increased. These differences were primarily due to an increased abundance of species possessing favored traits, rather than species turnover as a consequence of herbivore exclusion. No community-level differences were observed for mean wood density or specific leaf area. We also found 52 CHAPTER 3: EFFECTS OF MAMMALIAN HERBIVORES ON SAPLING TRAITS that intraspecific leaf toughness was lower in exclosures. Our results suggest that mammalian herbivores are important for shaping plant functional trait distributions well beyond the seedling stage, and that their effects on the sapling community are distinct from their effects on the seedling community.

INTRODUCTION

Mammalian herbivores are known to be important top-down regulators of plant community composition in many ecosystems (Olff & Ritchie 1998, Parker et al. 2006). In tropical forests, mammals are recognized as significant seedling predators (Barone and Coley 1996). However their importance as herbivores in later stages of plant recruitment has been less studied. One study in peninsular Malaysia found that native pigs ( Sus scrofa) uproot or snap 500 saplings per female pig in their reproductive nest building (Ickes et al. 2001). Experimental exclusion of pigs resulted in sapling density increasing by 32 percent, relative to open controls in just two years (Ickes et al. 2001). Lizcano (2006) excluded large mammalian herbivores, including mountain tapir ( Tapirus pinchaque ) and brocket deer ( Mazama spp.), in montane tropical forests in Columbia and discovered a 2-fold increase in understory plant biomass in two years. These two studies found higher species richness in exclosures, as may be expected with higher stem densities, but no differences in species diversity indices. The discrepancy between species richness and species diversity indices is explained by the fact that species dominance often increases in mammal exclosures as well. The positive and negative influences of increased species richness and dominance respectively on the diversity index of an exclosure community can cancel each other out, leading to the diversity index of an exclosure community being similar to that of a control community. In other ecosystems, herbivore exclosure not only results in changes in plant density and diversity, but also in plant functional trait composition. This is because, although mammals are generalist herbivores relative to many insects, they still exhibit diet preferences. In one temperate forest, browsers selected species lower in defense 53 compounds (e.g. leaf phenolics, condensed tannins, fiber, lignin) (Wardle et al. 2002), and in a sub-humid grassland, grazers preferred taller-growing species (Diaz et al. 2001). However, at the community level, evolutionary history and present day intensity of herbivory interact with diet preferences to produce mixed effects on functional trait composition. In cases where species have evolved naïve of mammalian herbivores, introduced browsers change species composition in favor of species with lower average tissue quality (Wardle et al . 2002). The same occurs in intensely grazed grasslands (Chaneton et al . 1988). In cases where plant communities have evolved with mammalian herbivores, herbivory results in communities of species with higher average tissue quality, both in the presence of native herbivores (McNaughton 1985), as well as domestic herbivores that are not grazing at intensive levels (Cingolani et al . 2002). In contrast to other ecosystems, mammals in tropical forests are perceived to be relatively unimportant herbivores. Mammals are responsible for about one quarter of all herbivory in tropical forests. As generalists, they are perceived to have less of a selective influence in shaping plant community composition and plant anti-herbivore defense than specialist invertebrate herbivores (Coley and Barone 2003). However, tropical forest mammals are important seedling predators. Therefore, the primary aim of this study was to test whether mammalian herbivores could indeed have effects on plant functional trait composition in tropical forests similar to those seen in temperate forests by influencing the composition of the younger cohorts of the plant community. We conducted this study in the Barro Colorado National Monument in central Panama, where a terrestrial mammal exclosure experiment has been ongoing for the past fifteen years (Royo & Carson 2005). We predicted that if mammalian herbivores were important for determining sapling functional trait distributions, exclosure would favor species that are faster growing (typically species with low wood density and high specific leaf area, or SLA), less defended (low leaf toughness) and which have higher nutrient leaves (higher leaf nitrogen). Our second goal was to evaluate the hypothesis that differences in traits between treatments developed both as a 54 CHAPTER 3: EFFECTS OF MAMMALIAN HERBIVORES ON SAPLING TRAITS consequence of species turnover as well as increases in abundance of favored species already common to both treatments. A final goal of this study was to examine whether the traits of the species themselves changed intraspecifically with mammal exclusion. In temperate systems, changes in mammalian herbivory induce changes in photosynthetic rates (Oleksyn et al . 1998, Nabeshima et al. 2001), photosynthetic efficiency (Retuerto et al. 2004), leaf toughness (Shimazaki & Miyashita 2002). In tropical savanna, megaherbivores influence the concentrations of condensed tannins in Acacia (Ward & Young 2002). Though herbivory rates in tropical forests are much higher overall in comparison to temperate forests (Coley & Barone 1996), suggesting that perhaps constitutive anti- herbivore defenses should be ubiquitous. However, several tropical species alter their defense in response to herbivory, and it has been suggested that the lack of more tropical examples of plasticity in defense are due to lack of study rather than biological differences between temperate and tropical plants (Karban & Baldwin 1997). We therefore examined whether the defense trait we examined, leaf toughness, may decrease intraspecifically in response to mammal exclusion.

METHODS

Study sites. The forest in Central Panama surrounding Lake Gatun is semi-deciduous moist forest. We conducted our experiment on Barro Colorado Island (BCI) and on the Gigante Peninsula, within the Barro Colorado National Monument (BCNM). The BCI has relatively high abundances of vertebrate herbivores affected by mammal exclosure, compared to other Neotropical forests, including deer (Mazama Americana and Odocoileus virginianus ), collared peccary (Tayassu tajacu ), agoutis (Dasyprocta punctata ), pacas (Agouti paca ), rabbits (Silvilagus brasiliensis ). This is due to a combination of factors, including a lack of resident large felines on the island, protection from poaching, and other environmental factors. Tapir (Tapirus bairdii ) are also present on BCI, but are rare.

Exclosure experimental design . Between late-1993 and mid-1994, eight pairs of fenced, exclosure plots and open, control plots were established within the BCNM (Royo & Carson 2005). Four pairs of plots are on (BCI) and the remaining four pairs are on the mainland Gigante Penninsula. Fenced and unfenced plots are 30 m x 45 m and are approximately 5 m apart. Exclosure fences are constructed of 12.7 x 12.7 cm galvanized steel fencing 2.2 m tall and buried 0.25 m deep. A secondary 1.3 x 1.3 cm mesh surrounds the lower 70 cm and also extends 0.25 m below ground. A 5-7 m buffer zone at the plot edge, and bisecting the plot, allowed for access and the avoidance of fence effects. Exclosures excluded the species mentioned above, but did not exclude climbing mammals such as squirrels and spiny rats, nor did they exclude arboreal mammals such as monkeys (Fig. 2.1). Monitoring data suggest that birds were not altered by the exclosures, with the possible exception of tinamous (Fig. 2.1). Though tortoises occur on our study site and are likely to be excluded by the fences, they are small and rare, and generally not considered to have an effect on our study results.

Plant censuses. In 1994 and 2003, all woody juveniles (“saplings”) taller than 40 cm or larger than 3 mm diameter at 20 cm in height and not in the buffer zone were identified, measured and marked.

Plant traits . We sampled leaves from up to six individual saplings of each tree species in each of the eight plots occurring on BCI. Specific leaf area was measured on shade leaves of saplings as the fresh lamina area divided by the mass of a whole leaf with petiole after oven drying to constant mass at 60ºC (Cornelissen et al. 2003). Leaf samples were dried at 65 ºC, ground using a ball grinder, and analyzed on an NA1500 elemental analyzer (Carlo Erba Instruments, Italy) for total N and C. Leaf toughness was measured as the force necessary to punch a 3mm diameter hole the leaf lamina with a force gauge, avoiding leaf veins to the extent possible (Sagers &Coley 1995). Wright and colleagues (2007b) determined dry seed mass and adult tree wood specific gravity (2010) for both free standing species and lianas in central Panama.

56 CHAPTER 3: EFFECTS OF MAMMALIAN HERBIVORES ON SAPLING TRAITS

Analyses . Trait correlations were conducted to assess whether traits in this study were orthogonal to one another, and thus likely to respond independently to herbivory. Community mean trait values were calculated in three ways. First, we calculated the grand means of the traits at the species level, and used these species means to determine the average trait value of the species occurring in each plot. We next repeated this analysis, but weighted the species’ trait means by species’ abundances in the plots. Finally, we calculated species trait means separately for exclosure and open communities, and recalculated the abundance-weighted plot means using these treatment-specific species trait values. Linear mixed models were used to evaluate whether changes in community mean trait values over time differed between treatments. Exclosure and time were treated as fixed factors and the exclosure-control pair were treated as a random factor nested within time. Our experiment tested for an exclosure-time interaction, and significance values reported here refer to that interaction. All analyses were conducted in R (R Development Core Team, 2008).

RESULTS

SPECIES TRAITS Functional traits were measured for 38.5-76.4 percent of species in the experiment and 58.8-91.6 percent of stems in the experiment (Table 3.1). Traits correlations were largely not significant, with the exceptions of SLA-leaf toughness and SLA-leaf nitrogen (Table 3.2). However, the correlation coefficients for these two relationships were still low (r 2 = 0.325 and r2 = 0.231 respectively), and our results did not show these trait pairs responding in a coordinated fashion (Fig. 3.1 & 3.2).

TABLE 3.1. The number of species and stems for which traits were measured.

No. Spp % Spp No. Stems % Stems Total 369 - 35,069 - Wood specific gravity 282 76.4 32,113 91.6 Specific Leaf Area 154 41.7 23,342 66.6 Leaf Toughness 152 41.2 23,322 66.5 Leaf Nitrogen 142 38.5 20,633 58.8

TABLE 3.2. R2 values for pair-wise correlations between species mean trait values evaluated in this study. The sign in parenthesis denotes whether the relationship was positive or negative. All correlations are significant ( p < 0.001).

Wood specific gravity SLA Toughness SLA 0.031 (-) Toughness 0.004 (-) 0.325 (-) Nitrogen 0.11 (-) 0.231 (+) 0.081 (-)

EFFECTS OF SPECIES OCCURRENCE In 1993, approximately 87 percent of species occurring in the experiment were present in both treatments. In 2003, 75 percent of species occurring in the experiment were present in both treatments. Therefore, when we averaged the trait means of the species occurring in each plot, we observed that the changes in plot trait means over time did not differ between treatments for any of the traits measured (Fig. 3.1). 58 CHAPTER 3: EFFECTS OF MAMMALIAN HERBIVORES ON SAPLING TRAITS

0.60 22 1994 2003 1994 2003 0.59 21

0.58 20

0.57 19

0.56 18

0.55 MeanSLA (mg/mm2) 17

WoodSpecific Gravity (g/cm3) 0.54 16 OpenExcl OpenExcl OpenExcl OpenExcl A Treatment B Treatment

34 1994 2003 2.6 1994 2003 33 2.4 32 2.2 31 2.0 30 1.8 29 MeanLeaf % Nitrogen 1.6 28 MeanLeaf Toughness (g/mm2) OpenExcl OpenExcl OpenExcl OpenExcl C Treatment D Treatment

FIGURE 3.1. Changes in plot-level trait means, not weighted by species abundance, did not differ significantly between treatments for (A) wood density, (B) SLA, (C) leaf toughness or (D) leaf nitrogen.

EFFECTS OF SPECIES ABUNDANCE Stem densities increased by 46% in exclosures from 1994 to 2003, but did not increase in open plots (data not shown). When we averaged the trait means of the species occurring in each plot, and weighted traits by species abundance, we did observe differences between treatments for some traits. We predicted that exclosures would favor species with lower wood density, lower SLA, lower leaf toughness, and higher leaf nitrogen. We did not see significant differences between treatments in wood density (Fig. 3.2A) or SLA (Fig. 3.2B). However, changes in leaf toughness (Fig. 3.2C) and leaf nitrogen (Fig. 3.D) over time were significantly different between treatments. Consistent with our expectations, exclosures had lower community mean leaf toughness and higher community mean 59 leaf nitrogen in 2003, relative to open plots. However these differences did not exist when the experiment was initiated in 1994 (Fig. 3.2 C & D).

0.60 24 1994 2003 1994 2003 0.59 22

0.58 20

0.57 18

0.56 16

0.55 MeanSLA (mg/mm2) 14

WoodSpecific Gravity (g/cm3) 0.54 12 OpenExcl OpenExcl OpenExcl OpenExcl A Treatment B Treatment

38 2.6 1994 2003 1994 2003 36 2.4 * 34 * 2.2 32 2.0 30 1.8

28 1.6 MeanLeaf % Nitrogen 26 1.4 MeanLeaf Toughness (g/mm3) OpenExcl OpenExcl OpenExcl OpenExcl C Treatment D Treatment

FIGURE 3.2. Changes in abundance-weighted, plot-level trait means did not differ significantly between treatments for (A) wood density or (B) SLA, but did differ significantly between treatments for (C) leaf toughness and (D) leaf nitrogen. (* p = 0.06 , time x treatment interaction)

INTRASPECIFIC DIFFERENCES IN LEAF TRAITS Because we sampled leaf traits in both open and exclosure plots, we could examine how intraspecific differences in trait means contributed to the differences between treatments at the community level. For one trait, leaf toughness, we found that determining species trait means separately for open and exclosure treatments, and using those treatment-specific species trait means to calculate plot-level means, resulted in an enhanced difference between treatments at 60 CHAPTER 3: EFFECTS OF MAMMALIAN HERBIVORES ON SAPLING TRAITS the plot-level in the later time point (Fig. 3.3), relative to using grand trait means for species (Fig. 3.2C).

38 1994 2003 36 * 34 32 30 28

MeanLeaf Toughness (g/mm2) 26 Open Excl Open Excl Treatment

FIGURE 3.3. Intraspecific differences in species’ mean leaf toughness between open and exclosure plots increase the difference in treatment-level mean leaf toughness in 2003. (* p = 0.004, time x treatment interaction)

For species having at least two leaf samples from each treatment type, 61.8% of species had lower leaf toughnesses in exclosures relative to controls (binomial test, N = 76, p = 0.025). This bias was even stronger when the sample was limited to species with treatment mean toughnesses based on at least eight leaf samples, with 73.9% of species having lower leaf toughnesses in exclosures (binomial test, N = 23, p = 0.017). Because sample sizes among species and between treatments was unbalanced, a MANOVA approach could not be taken to assess whether treatment means within a species were significantly different. Therefore, using the latter subset of species above, a two-way ANOVA was applied. The association of exclosure treatment with lower mean leaf toughness with species was also significant using this approach (Table 3.3).

TABLE 3.3. Effect of species and exclosure on variation in leaf toughness. N = 23 species, each of which had at least eight leaf toughness samples from each of the two treatment types.

Effect df SS MS F p Species 22 207.4 9.425 165.2 < 0.001 Exclosure 1 0.305 0.305 5.352 0.021 Species x Exclosure 22 1.292 0.059 1.029 0.425 Residuals 604 34.467 0.057

DISCUSSION

Regulation of plant functional trait composition by mammalian herbivores In this study, we demonstrate that terrestrial mammals influence understory functional trait distributions in tropical forests. They do so primarily by suppressing populations of species with traits that are correlated with high nutrition and low defense. Consistent with our predictions that reduced herbivory should favor species with low defense and high nutrient content, community mean leaf toughness decreased over time in exclosures over time (Fig. 3.2C), and community mean leaf nitrogen increased in exclosures over time (Fig 3.2D), while open control plots did not show equivalent changes. Wood density is associated with seedling survival in the presence of herbivores in this forest (Alvarez-Claire & Kitajima 2007, 2009). Browsers and grazers have also been shown to select species with higher SLA in other systems (Wardle et al. 2002, Cingolani et al. 2002). However, neither of these traits changed in response to mammal exclosure. This could be because, at the sapling stage in tropical forests, the low light conditions of the forest understory also favor low SLA, high wood density species and perhaps this environmental factor constrains the community response to herbivory with respect to those traits. We did not find that changes in functional traits at the community-level were the result of divergences in species presence-absence between exclosure and open plots. The sapling community established prior to the imposition of the exclosure treatments, and therefore two treatments started out with very similar communities. 62 CHAPTER 3: EFFECTS OF MAMMALIAN HERBIVORES ON SAPLING TRAITS

Over longer time scales, competitively dominant species may outcompete less competitive species in the exclosure treatment, resulting in lower species richness, particularly of species which are better defended but slower growing. However, we did not observe decreases in species richness in the time-frame of this experiment (data not shown). Seed and seedling predators are also excluded from the exclosure communities. Therefore, it is reasonable to question whether the effects we observed are due exclusively to herbivore exclusion. Two lines of evidence allow us to attribute our results to herbivores. First, long-term studies of seedling growth in central Panama suggest that all individuals in this size class established before the experimental treatments were imposed (S.J. Wright, pers. comm .). Therefore, saplings in both treatments should have been exposed to the same seed and seedling predation agents. Consistent with the assumption that saplings established prior to exclusion of seed and seedling predators, community mean seed mass does not differ between treatments for the sapling community (data not shown). This is in contrast to the seedlings in the exclosure communities, which did establish under conditions of seed and seedling predator exclusion, and as a result have higher community mean seed mass than seedlings in open control communities (Fig. 2.4). These two lines of evidence suggest that exclusion of seed and seedling predators are not responsible for the exclosure effects observed in the sapling community, leaving only herbivory as a possible explanation.

Intraspecific differences in leaf toughness The majority of species in the experiment showed lower mean leaf toughness inside the exclosure, relative to open plots. When these intraspecific differences in species leaf toughness were accounted for (Fig. 3.3), treatment differences in plot level leaf toughness in 2003 were 66 percent larger than when a grand mean for the species leaf toughness was applied (Fig. 3.2C). This result suggests that in the case of leaf toughness, the trait values measured for juvenile tropical plants may vary with the degree of herbivory experienced. 63

There are several possible explanations for the intraspecific divergence in leaf toughness. First, it is possible that, within species, genotype or environment produces variation in leaf toughness, and individuals with lower leaf toughness survive at a lower rate in open plots than in exclosures . Support for this mechanism can be found in studies of insect herbivory on plant genotypes which differ in their expression of qualitative defense traits (Kessler & Baldwin, 2001, Silfer et al. 2009). An alternative explanation is that plants are relaxing their investment in mechanical defenses when they no longer experience mammalian herbivory. Hundreds of plant species are known to increase investment in defense when exposed to herbivory, and subsequently attenuate that defense when herbivory levels are lower (Karban & Baldwin 1997). However few studies have documented reduced investment in defense upon reduction of mammalian herbivory. In Acacia drepanolobium decreases in spine length or spinescence are detectable after 22 months (Young & Okello 1998), decreases in condensed tannin concentrations are detectable after 24-36 months (Ward & Young 2002) and reduced investment in indirect defenses occur within several years (Huntzinger et al . 2004). Regardless of the mechanism, the fact that mammalian herbivores may play a role in altering leaf toughness intraspecifically in tropical plants is unexpected. Insects are thought to be responsible for three quarters of herbivory in tropical forests, and also the source of the strongest selection pressures plants receive to develop anti- herbivore defenses in this system (Barone & Coley 2002). The experimental plots in which we measured leaf toughness did not manipulate insects, and it is plausible that plants in both treatments could have experience high rates of invertebrate herbivory, and therefore shown no attenuation in defense traits. Yet, it appears that insect herbivores could not compensate for mammalian herbivores in maintaining high intraspecific leaf toughness when mammals were excluded.

Implications for tropical defaunation In many temperate forests, populations of native or introduced large mammalian herbivores are not regulated in a top-down fashion by predators. As a 64 CHAPTER 3: EFFECTS OF MAMMALIAN HERBIVORES ON SAPLING TRAITS result, these communities currently experience intense levels of herbivory, low rates of plant recruitment and often decreases in species richness and diversity (Wardle et al. 2002, Rooney 2001). However, tropical forests are undergoing the opposite perturbation. Throughout the tropics, mammalian herbivores are being unsustainably hunted (Peres & Palacios 2007, Corlett 2007, Fa & Brown 2009). Preferred game species of animals are often large-bodied herbivores, such as elephants, tapir, pigs, peccaries and deer. Understanding how and why such perturbations in herbivore abundance alter plant species diversity and functional composition will help conservation biologists identify potential interventions to conserve plant species threatened indirectly by defaunation. However, because most studies examining changes in functional diversity have been conducted in temperate forests, where the perturbation experienced by plant communities is increased rates of herbivory, little empirical data exists with which to evaluate how tropical forests may respond to decreased rates of herbivory by native herbivores as a consequence of defaunation. The biomass of terrestrial mammalian herbivores in tropical forests is quite low, relative to other ecosystems (Bodmer 1989, Leigh et al. 1982). In addition, insects are thought to account for approximately 75 percent of overall herbivory in tropical forests. Taken together, these data might suggest that observed reductions in terrestrial mammalian herbivore populations in tropical forests may not have large impacts on sapling community composition. However, this study demonstrates reduced herbivory as a consequence of tropical defaunation is likely to cause an increase in dominance of sapling species with higher leaf nitrogen and lower leaf toughness. Leaves with higher leaf nitrogen and lower concentrations of defense compounds typically have higher decomposition rates. Therefore, it is conceivable that if changes in the sapling community propagate to alter canopy composition, defaunation may alter nutrient cycling in tropical forests.

ACKNOWLEDGEMENTS

D. Ackerly, L. Curran, R. Dirzo and P. Vitousek provided constructive criticism which improved this work. The Smithsonian Tropical Research Institute granted permission to conduct this work, and provided logistical support. D. Turner provided assistance with leaf nitrogen analysis. J. Wright provided seed mass and wood density data. Many thanks go to L. Jimenez, C. Sarmiento, S. Rebellon, A. Calderón, R. Bethancourt for help in the field and lab. A. Hernandez and O. Valdes provided indispensible help with species identification. NSF DEB-0808338, the Theresa Heinz Environmental Scholars program, and the STRI short-term fellowship program provided support to E. Kurten.

66 CHAPTER 3: EFFECTS OF MAMMALIAN HERBIVORES ON SAPLING TRAITS

Chapter 4

Hunting does not alter seed predation rates as a function of seed size in a Neotropical forest

Erin L. Kurten

ABSTRACT

This study tested the hypothesis that vertebrate seed predation rates vary with seed size and with hunting intensity in tropical forests, corresponding to direct impacts of hunting on vertebrate species of differing size classes. Seed fate of seeds from species with mean fresh seed masses ranging from 0.002 to 62.4 g was compared between a protected and nearby hunted forest in central Panama. Transect survey data verified that higher hunting intensity corresponded with lower abundances of key mammalian seed predators. Seed arrays of a subset of species were monitored with camera traps to verify the identities of animals responsible for seed removal. In this system, there was only weak support for a relationship between seed predation rates and seed size, and no evidence that this relationship changed at high hunting intensity. There was also no evidence that seed predator body size varied with seed size, as agoutis (Dasyprocta punctata ) were responsible for most of the verified removal for five plant species examined, ranging from 1.95 to 62.4 g in fresh seed mass.

68 CHAPTER 4: SEED PREDATION RATES AS A FUNCTION OF SEED SIZE AND DEFAUNATION

INTRODUCTION

Overharvesting of bushmeat is a pervasive threat to biodiversity across the tropics (Fa & Peres 2001, Peres & Palacios 2007, Corlett 2007). This defaunation not only directly lowers vertebrate abundance in hunted forests, but initiates trophic cascades which perturb the plant, insect, and pathogen communities as well. Preferred game species of animals are usually large-bodied vertebrates, such as ateline monkeys, tapir, peccary and deer, as well as large birds, and in some cases reptiles, in the context of Neotropical forests. Understanding how and why such cascading perturbations alter species diversity and functional composition in the plant community is important for understanding the role mammals play in maintaining the diversity of tropical forests, as well as identify potential interventions to conserve plant species threatened by the negative indirect effects of defaunation. Several studies have documented reduced seed dispersal as a consequence of hunting in tropical forests, particularly for larger-seed species (e.g. Wright et al . 2000, Guariguata et al. 2000, Guariguata et al . 2002, Galetti et al . 2006, Wang et al . 2007, Brodie et al. 2009, Fadini et al . 2009, Holbrook & Loiselle 2009). These studies suggest that reduced dispersal should result in reduced recruitment of large-seeded species, in particular those that rely on large-bodied mammals vulnerable to hunting for their seed dispersal (Peres & van Roosmalen 2002, Nuñez-Iturri et al . 2008, Terborgh et al. 2008). However, in most cases, reduced dispersal due to hunting has not been explicitly linked to subsequent reductions in seedling recruitment. While one study modeled the potential reduction in seedling recruitment, based on estimates of reduced dispersal to and differential survival in different microhabitats (Brodie et al. 2009), only one set of studies of which I am aware measured seed dispersal, seed predation, and seedling recruitment for the same species in both protected and hunted sites (Wright et al . 2000, Wright & Duber 2001). Wright and colleagues (2000, 2001) focused on two large-seeded palm species, members of genera which are widespread in the Neotropics. They found seed dispersal to be indeed reduced, and seedling densities near parent trees to be higher, as one would predict to be associated with 69 hunting. However, contrary to what has been hypothesized in the seed dispersal literature, recruitment of Attalea and Astrocaryum palms was actually higher in the hunted sites (Wright et al . 2000, 20001). The reason for this unexpected increase in recruitment was that hunting reduced the intensity of vertebrate seed predation, and this reduction was not fully compensated for by invertebrates. (Cramer et al. 2003) suggested a similar reduction in seed predation was responsible for higher recruitment of the African species Balanites wilsoniana in defaunated sites, despite lower dispersal by elephants, however seed predation rates were never measured (Babweteera et al. 2007). It is clear from this example that a better understanding of the consequences of hunting for seed predation is necessary for providing context to the studies documenting reduced seed dispersal. In studies which have measured seed predation in sites with differing intensities of hunting, rates of vertebrate seed predation have often been lower in hunted sites (Terborgh & Wright 1994, Wright et al . 2000, Wright & Duber 2001, Guariguata 2000, Beckman & Muller-Landau 2007, but see Roldán & Simonetti 2001). Even less clear is how perturbations in seed predation rates as a consequence of hunting may vary with seed size. Because larger-seeded species are expected to disproportionately experience reduced seed dispersal, understanding how seed predation changes with hunting across a range of seed sizes is important for understanding whether or not the potentially negative demographic consequences of reduced seed dispersal of larger-seeded species is intensified, maintained, or compensated for by alterations in seed predation across species (Stoner et al. 2007). However, most studies which have measured seed predation rates under various hunting regimes have focused on one or two plant species (e.g. Terborgh & Wright 1994, Wright et al . 2000, Roldán & Simonetti 2001, Wright & Duber 2001, Galetti et al . 2006, Beckman & Muller-Landau 2007, Donatti et al. 2009). Due to the variation in location, mammal community composition, and degree of defaunation in the sites studied, these data cannot be directly compared to assess how seed predation rates change with defaunation intensity as a function of seed size. Guariguata and 70 CHAPTER 4: SEED PREDATION RATES AS A FUNCTION OF SEED SIZE AND DEFAUNATION colleagues (2000) measured seed predation rates of seven species, but focused primarily on small-seeded timber species. . In 2007, Dirzo and colleagues published a model of how vertebrate seed predation could be expected to vary with seed size and intensity of vertebrate defaunation in tropical forests (Fig. 4.1). At low levels of hunting, vulnerable, large- bodied animals drastically decrease in abundance or become locally extirpated. In response to this competitive release, as well as a likely reduction in predation by felines, medium-sized seed predators initially become more abundant, until they become the primary game species at moderate levels of hunting. At the highest intensities of defaunation, even medium-sized seed predators have become rare. Small rodents increase steadily with increasing hunting intensity, as they experience a reduction in competition and predation. In this model, predation intensity of larger-seeded species mirrors the changes in abundance of medium-bodied seed predators (e.g. 4-20 kg body mass in Neotropical communities), while predation of smaller-seeded species increases as small bodied seed predator (< 1 kg) biomass increases. As a result, the relative proportion of larger and smaller seeds comprising the seedling bank is expected to shift as defaunation intensity increases. This model was important for clarifying how seed predation is likely to vary by seed-size, and how this relationship is dependent upon the defaunation context.

Large vertebrates A Mid-sized vertbrates Small rodents BCI PNS C No Def. High Def. Seed predator Biomass Seed predator

Large Seeds B Small Seeds Seed predation Intensity Seed predation

Small La rge Seed Size Seed predation Intensity Seed predation

None Moderate High Defaunation Intensity FIGURE 4.1. Model of how seed predation should vary with seed mass as a function of defaunation intensity in a Neotropical forest (modified from Dirzo et al . 2007). (A) Changes in large, medium and small vertebrate seed predator abundance with increasing defaunation intensity. The approximate positions of BCI and PNS along the defaunation gradient shown here are based on mammal transect surveys. (B) Seed predation intensity of large and small seeded species on a biomass basis is correlated with the abundance of medium- and small-bodied seed predators respectively. (C) As defaunation intensifies, the relationship between seed predation rates and seed size goes from positive to negative.

With this study, I aimed to test several assumptions and hypotheses inherent in this model. The first assumption I tested was that larger-seeded species experience higher seed predation intensity than smaller seeded species at all but the highest levels of defaunation intensity. The second assumption of the model I tested was that larger- seeded species tend to be consumed by large- and medium-bodied vertebrates. Likewise, smaller-seeded species tend to be consumed by smaller-bodied vertebrates. A third hypothesis implicit in the model is that at high levels of defaunation intensity, where the number of large- and medium-bodied vertebrates is highly reduced, larger- seeded species experience reduced seed predation, relative to less defaunated sites. A 72 CHAPTER 4: SEED PREDATION RATES AS A FUNCTION OF SEED SIZE AND DEFAUNATION fourth prediction of the model is that smaller-seeded species experience increased seed predation as defaunation intensity increases, as populations of small bodied vertebrates increase in response to competitive release and reduced predation. There is a great interest in determining the extent to which other vertebrates or natural enemies can compensate for the loss of a primary seed predator. To test these hypotheses, seeds of 13 species spanning the range of 0.002 to 64.2 g mean fresh seed mass were monitored in arrays in both a protected forest and a nearby hunted forest. Fate of seeds was monitored for four weeks. To examine whether larger- and smaller-seeded species indeed had larger- and smaller-bodied seed predators respectively, the seed arrays on BCI were monitored with camera traps for five species ranging from 1.95 to 64.2 g fresh mean seed mass for one week, and the the agents of seed removal were recorded. To assess whether other vertebrates might compensate for the loss of a principle seed predator, I monitored two large-seeded palm species with camera traps in both sites. This allowed me to determine if and how the species of vertebrate seed disperser differed between the protected and hunted sites, in particular if smaller vertebrates compensated for the removal of larger vertebrates. I also recorded the source of mortality for all species when seeds were killed by invertebrates and pathogens, to examine whether those agents could also compensate for the loss of larger bodied seed predators.

METHODS

Study sites . The forest in Central Panama surrounding Lake Gatun is semi-deciduous moist forest. I examined the effects of defaunation by hunting by comparing protected and hunted forests in this area (Fig. 4.2). Barro Colorado Island (BCI) was chosen as the protected forest site, and could be regarded as a moderately defaunated site. Due to its size (15 km2), the island no longer supports white-lipped peccary, and has visiting, rather than resident, jaguar ( Panthera onca ) and puma ( Puma concolor ). Other large mammals, such as tapir ( Tapirus bairdii ) and spider monkeys ( Ateles 73 geoffroyi ), were locally extirpated and historically reintroduced (Enders 1939, Terwilliger 1978, Milton & Hopkins 2005). However today, activity of poachers is virtually nonexistent on BCI due to the monitoring by forest guards and Panamanian police. The last confirmed incident of poaching on BCI occurred in 1989 (Wright et al. 2007). Consistent with an absence of poachers, researchers on BCI do not encounter evidence of hunting, and wildlife is common to see. Adjacent to the Barro Colorado National Monument (BCNM) is the Parque Nacional Soberanía (PNS). PNS is a protected area, however the region of the park in which the study was conducted is not actively patrolled. This is due to both fewer forest guards at the park, a lack of vehicles to access the study area, which is more than 20 km from the park headquarters. Consistent with a lack of protection, various evidence of hunting activity was encountered while conducting this study, including spent shells, campfires, litter, gun shots, hunting dogs, experimental vandalism, and in one case, encountering a hunting party.

BCNM PNS trails

N 0 5 km

FIGURE 4.2. Map of Lake Gatun study area in central Panama. Seed predation experiments were conducted in areas demarcated by trails.

74 CHAPTER 4: SEED PREDATION RATES AS A FUNCTION OF SEED SIZE AND DEFAUNATION

Mammal census. Diurnal mammal transect surveys were conducted from mid-August to mid-December in 2008. Once per week in each site, an observer surveyed a 5 km transect between the hours of 6:30 AM and 11:30 AM, walking at a speed of 1 km/hr. Surveys were not conducted during rain. Observers recorded the species and age (adult or juvenile) of all mammals sighted, and measured initial detection distance with a Nikon ProStaff 550 range finder (Nikon Inc., Melville, NY). Because of the time and difficulty involved in counting individuals in troupes of primates, these species were recorded in units of groups. Because most species had few observations in PNS, and the parameters that density estimators require could not be accurately estimated for that site, I report animal abundances (individuals/km or groups/km), rather than densities.

TABLE 4.1. Mean fresh seed masses of study species. BCI and PNS columns denotes species used in each site. Most species are free standing growth forms. C. turczanninowii and T. richardii are lianas. Species Family Seed mass (g) BCI PNS Apeiba tibourbou Tiliaceae 0.0057 X X Protium tenuifolium Burseraceae 0.18 X Cupania latifolia Sapindaceae 0.19 X Chamaedorea tepejilote Arecaceae 0.49 X X Connorus turczanninowii Conneraceae 0.49 X Thevetia ahouai Apocynaceae 1.11 X Oneocarpus mapora Arecaceae 1.95 X Virola surinamensis Myristicaeae 2.43 X Calophyllum longifolium Clusiaceae 5.48 X Astrocaryum standleyanum Arecaceae 9.84 X X Gustavia superb Lecythidaceae 16.0 X Attalea butyracea Arecaceae 16.2 X X Tontelea richardii Celastraceae 64.2 X

Seed preparation . Thirteen woody plant species ranging from 0.002 to 64.2 g mean fresh seed mass were selected for this study (Table 4.1), representing 76 % of the range in fresh seed mass for this area on a log scale. While these species are common to both BCI and PNS, heavy insect and/or pathogen infestation of some species in PNS prevented collection of viable seeds in this site. Because of protected area rules, seeds could not be moved between sites. Therefore, eleven species were examined on BCI and six species in PNS, with four species common to both sites, spanning almost the entire range of seed size in the study. Seeds appearing rotten, insect-damaged, or otherwise unviable were not included in the experiment. To remove confounding effects of differing seed pulps, all seeds were cleaned of pulp or mucilage. After pulp removal, mass, width and length were measured for seeds of all species but Apeiba tibourbou . Apeiba seed mass for BCI was previously recorded (Wright et al. 2007). A thread was attached to all but the smallest seeds (A. tibourbou and P. tenuifolium) , so that I could recover the seeds and record their fate.

76 CHAPTER 4: SEED PREDATION RATES AS A FUNCTION OF SEED SIZE AND DEFAUNATION

Seed predation experiment . A randomized block design was used to compare seed predation rates in the two sites. Within each forest, six blocks were delineated. Within each block, each species was placed on the forest floor in an array of 6-8 seeds in a different random location. For G. superba on BCI and C. tepelijote in PNS, only four arrays were used, as few seeds unattacked by insects were available. Because A. tibourbou and P. tenuifolium , were small and not marked with thread, these species were set out in shallow mesh trays with 6-8 seeds per tray. Data reported here is from a total of 732 seeds. Experiments were conducted between mid-July and mid-September 2009. Seeds were monitored for four weeks, during which seed fate and agent of mortality were recorded. Classes of seed fate were: unmoved (intact), cached, moved but not buried (intact), predated by vertebrates, predated by insects, and killed by pathogens. Seeds which were removed and only the thread recovered were presumed predated. .

Assessment of seed predator identity . The arrays of 5 species with masses ~2 g or larger were monitored with Reconyx RC-55 infrared cameras on BCI for the first week. Only half of the Virola arrays were monitored due to camera limitation. In addition, the arrays of two of the larger-seeded species, A. standleyanum and A. butyracea were monitored in PNS. More species were not monitored in PNS due to concerns over equipment loss. As a conservative assignment of seed predator identity, the species of seed predator was only recorded when the removal of a marked seed was clearly visible in the image. Instances of probable seed removal, but in which camera angle, position of the animal, or clarity of the image prevented a clear determination of the event, were not assigned to a particular animal species.

Analyses . The relationships between mean seed mass and mean seed mortality in the two sites was assessed via an ANCOVA, with site as the fixed factor and log- transformed mean seed mass as the covariate. Differences in seed caching rates between sites were evaluated with a t-test for both palm species. All analyses were conducted in R (R Development Core Team, 2008). 77

RESULTS

Mammal census. Diurnal mammal transect surveys, conducted from mid-August to mid-December in 2008, verified differences in degree of defaunation between BCI and PNS, which approximately correspond to moderate and intense degrees of defaunation respectively (Fig. 4.3.)

Medium-bodied Dasyprocta punctata Cebus capuchinus Alouatta palliata Large-bodied Ateles geoffroyi ‡ Species Mazama americana Odocoileus virginianus * Pecari tajacu BCI Tapirus bairdii ‡ PNS

0.0 0.2 0.4 0.6 0.8 1.0 Individuals pe r km

FIGURE 4.3. Animal abundances in BCI and PNS as assessed by trail census in 2008. Above are species of medium body size known to maintain or increase their population density with moderate hunting, but decrease with intense hunting (Peres and Palacios 2007). Below are large-bodied animals that have been historically present in the sites and which are known to be most vulnerable to hunting. At both BCI and PNS, most large-bodied mamals are rare or locally extinct. However, medium-sized mammals are highly abundant in BCI, whereas they are also quite rare in PNS, indicating the two sites are moderately and intensely hunted, respectively. † locally extirpated in PNS; ‡ locally extirpated in BCI, reintroduced, and currently rare; *present and rare at both sites, not observed in this census.

Seed predation intensity . Relationships between seed mortality and seed size were evaluated by ANCOVA, with site being a fixed factor. Seed mortality over all species, when all sources of mortality were pooled, was not correlated with seed mass

(F1,13 = 1.47, p = 0.25), and this did not differ between sites (F1,13 = 0.69, p = 0.42) 78 CHAPTER 4: SEED PREDATION RATES AS A FUNCTION OF SEED SIZE AND DEFAUNATION

(Fig. 4.4a). For the subset of species with verified vertebrate seed predation , there was not significant (F1,12 = 1.45, p = 0.25) (Fig 4.4b) Apeiba , which appeared to have no vertebrate seed predators at either site. was excluded from the model. ). Also, contrary to the model (Fig. 4.1), the effect of site was not signficant (F1,13 = 0.11, p = 0.74) (Fig. 4.4b). Mortality caused by invertebrates and fungal pathogens was negatively correlated with seed mass across all species (F1,13 = 21.1, p > 0.001).

Again, this relationship did not differ by site ( F1,13 = 0.19, p = 0.67). Mortality by All Sources Vertebrates Invertebrates & pathogens 1.0 A B C 0.8

0.6

0.4

0.2 Proportion seeds killed 0.0 -6 -4 -2 0 2 4 -6 -4 -2 0 2 4 -6 -4 -2 0 2 4 BCI Log(Seed mass) PNS FIGURE 4.4. Seed predation rates as a function of seed size on BCI and in PNS. (A) Across all species there was a negative, but non-significant, relationship between seed mass and seed mortality. (B) Seed mass and vertebrate seed predation were not correlated when analysis was restricted to species with verified vertebrate seed predators. (Points for Apeiba slightly offset for visualization.) (C) Seed mass and seed mortality by invertebrates and fungal pathogens were negatively correlated (includes Apeiba ). No differences were observed between sites.

Scatter-hoarding rodents do not just predate seeds. For two of the large seeded palms, seed caching by scatter-hoarding rodents is critical for escaping seed predation by bruchid beetles. I predicted that caching of Attelea and Astrocaryum would be lower in the hunted site, as the primary seed caching agent, the agouti (Dasyprocta punctata ), was less abundant at that site (Fig. 4.3), unless compensated for by squirrels. Contrary to this prediction, the number of seeds cached did not differ between sites for either species ( Attalea : t = -0.129, p = 0.55, one-tailed test; Astrocaryum : t = -0.412, p = 0.66, one-tailed test). (Fig. 4.5). This equivalence 79 between sites appears not to be due compensation by squirrels, in as far as every seed caching event for which the agent could be verified by camera trap was committed by an agouti (data not shown).

2.5

2.0

1.5

1.0 Seeds Cached Seeds

0.5

0.0 BCI PNS BCI PNS Attalea Astrocaryum Sites by Species FIGURE 4.5. Number of seeds cached per array of eight seeds for two large seeded palms did not differ between BCI and PNS.

Seed predator identity . Camera trap monitoring of seeds in the first week on BCI revealed D. punctata , to be the primary agent of seed removal for all species (Fig. 4.6). Seed removal by spiny rats ( Proechimys steerei ) and squirrels (Sciurus granatensis ) accounted for little of the seed predation and varied by plant species (Fig. 4.6). Data from BCI do not appear support the hypothesis that seed predator size is correlated with seed size. 80 CHAPTER 4: SEED PREDATION RATES AS A FUNCTION OF SEED SIZE AND DEFAUNATION

50 50 Not taken Unknown Squirrel Spiny rat 40 40 Agouti

30 30

20 20 Number Number of seeds Number of seeds 10 10

0 0 BCI PNS BCI PNS Oneocarpus Virola Astrocaryum Attalea Tontelea Attalea Astrocaryum Species ranked by mass Sites by Species

FIGURE 4.6. Identity of species removing seeds during first week of experiment. (Left) On BCI five species ranging in mass from 1.95 g (Oneocarpus ) to 64.2 g ( Tontelea ) were primarily removed by agoutis, and showed no relationship between seed size and seed predator body size. (Right) Seed removal of large-seeded palms at BCI and PNS. Removal of Attelea was higher in PNS, but Astrocaryum removal did not differ between sites. Seed predation by small mammals was higher in PNS for Attalea , and higher in BCI for Astrocaryum .

A comparison of palm seed dispersal in BCI and PNS revealed that the seed removal of Attelea in PNS was more than 4-fold higher in the hunted site, contrary to expectation. Verified Attelea predation by D. punctata was similar in both sites, while seed predation by small mammals, primarily squirrels, was much higher in the hunted site. Astrocaryum removal in the first week was 100% in both sites, with confirmed removals by D. punctata being slightly higher in the hunted site, and removals by P. steerei being slightly higher in the protected site (Fig. 4.6).

DISCUSSION

Hunting is a pervasive threat to biodiversity in tropical forests. Decreases in plant diversity as high as 66% have been documented as an indirect consequence of hunting. Yet the mechanisms by which that diversity is lost are poorly understood, as is their context dependence. This study aimed to test some basic assumptions about plant- seed predator interactions in the context of seed size, as well as hypotheses as to how those interactions may change as a consequence of defaunation. This information is useful for understanding how vertebrate seed predators shape community level distributions of seed size, and to predict how and why plant community composition may change with increasing hunting intensity. This discussion focuses primarily on plant-seed predator interactions. However, as several of the animals relevant to this discussion are also recognized to be secondary seed dispersers, much of the discussion which follows is equally relevant to seed dispersal.

Seed size and vertebrate seed predation intensity . Overall, this study found little support for the model assumption that vertebrate seed predation rates are correlated with seed mass, either positively in moderately defaunated sites, or negatively in highly defaunated sites. Other seed traits, as well as natural history, may be more important for determining seed predation rates than seed mass per se . For instance, of the three of largest seed-species in this study, the Astrocaryum and Attelea exhibited the highest rates of seed removal, whereas Gustavia exhibited one of the lowest. Relative to Gustavia , the palm seeds have an endosperm that is high in fat content, and therefore energy rich for their weight. The rodents are well adapted to the primary defense of these palms against seed predators: a woody endocarp. While the thin seed coat of Gustavia would appear to make it more vulnerable to vertebrate seed predators, it may exhibit chemical defenses that make them less desirable to rodents(e.g. Forget 1992). Indeed, it was not uncommon on BCI and PNS to see bits of Gustavia pulp with a pile of intact seeds of the forest floor, the result of agoutis consuming the fruit flesh and discarding the seeds (E. Kurten, pers. obs.). However, 82 CHAPTER 4: SEED PREDATION RATES AS A FUNCTION OF SEED SIZE AND DEFAUNATION there are circumstances in which Gustavia seed predation has been higher (Sork 1987, Forget 1992). The natural history of the two palm species also make them preferable for hoarding. A. standleyanum and A. butyracea germinate after 8 and 12 months respectively. Therefore, they are likely to persist intact as hoarded seeds until the period of low fruit availability from November through February (Foster 1982). In contrast, Gustavia germinates within 1-2 weeks of fruit fall (E. Kurten, pers. obs.). While the seeds remain attached to the establishing seedling for many more months, seeds and cotyledons are vulnerable to consumption by other seedling predators, and seed energy stores become reduced over time. By November, Gustavia seedlings are no longer of interest to agoutis on BCI (Forget 1992). Therefore, in contrast to the palms, Gustavia is not a species useful for provisioning rodents through the period of low fruit availability. Comparison of the large-seeded palm species and Gustavia illustrates how variability in seed consumption by vertebrates among seeds of similar size is likely to be introduced by a variety of differences in nutritive value, defense traits, and germination strategy. Variation in these traits will likely cause deviations in consumption rates from what may be predicted by seed size and optimal foraging theory alone.

Seed predator identity . Camera-based observations for five species preyed upon by vertebrates showed little support for the model assumptions that the seed predator body size is correlated with the size of the seed. While enclosure and exclosure experiments have demonstrated that small rodents prefer small seeded species to taxonomically related larger seeds, and can exert a large, and disproportionate seed predation pressure on smaller seeded species (Dirzo et al . 2007, Mendoza & Dirzo 2007), this study suggests that when seeds are available to the entire community of potential seed predators, the importance of smaller mammals, even for relatively smaller seeds, is diminished. In particular, agoutis appear to be a highly important seed predator and secondary seed disperser in this system. 83

There may be other circumstances, however, in which the relationship between seed size and seed predator biomass may be stronger. It is notable that, despite the fact that peccaries are well-documented seed predators of palms such as Astrocaryum and Attelea in many parts of the Neotropics (Beck 2006), no observations of predation of these two species by peccaries were observed in this study. This is likely due to the fact that larger-bodied seed predators such as peccaries and brocket deer tend to consume fruits and seeds under fruiting trees, where fruit and seed densities are highest and foraging most efficient. In contrast, the seed arrays in this experiment best mimicked the situation in which a few seeds have already been dispersed away from the parent tree. Rodents such as agoutis and squirrels, will also forage under fruiting trees. However, as scatter-hoarding species, they are behaviorally more likely than larger seed predators to search for, find and consume small, isolated patches of seeds on the forest floor. In addition to dispersal context, the composition of the mammal community is likely influential in determining the strength of any relationship between seed size and seed disperser size. For example, in forests such as Cocha Cashu, Peru, agoutis are much less abundant (Terborgh & Wright 1994), and therefore less likely to demonstrate a high importance as seed predators and seed dispersers across a range of species, as seen on BCI. BCI is also lacking an abundance of terrestrial vertebrates at the two extremes of seed predator body size. White-lipped peccaries ( T. pecari ) have been locally extinct for more than half a century on BCI, and tapirs are rare. At the other extreme, in parts of Mexico and Belize, mice such as Peromyscus and Heteromys can be significant seed predators of some species (Coates-Estrada & Estrada 1988, Klinger and Rejmanek 2009), whereas rodents smaller than spiny rats (P. steerii ) are rare on BCI.

Seed predation and defaunation . Despite large differences in abundances of key seed predators such as collared peccaries and agoutis, decreased vertebrate seed predation rates of larger-seeded species were not generally observed between BCI and PNS, nor did smaller-seeded species generally show increases in seed predation rates in the 84 CHAPTER 4: SEED PREDATION RATES AS A FUNCTION OF SEED SIZE AND DEFAUNATION hunted site relative to the protected site. Rates of seed caching of large-seeded palms, critical for seedling recruitment in those species, also did not differ between sites. These data corroborate the work of Wright and colleagues (2001) with respect to large-seeded palms, and extend it to a broader spectrum of species. Rather than seed predation, and secondary seed dispersal, varying linearly with agouti abundance in this system, it may be useful to consider the possibility of seed predator thresholds. In particular, lower populations of agoutis as a consequence of hunting may serve to decrease competition for fruit and seed resources among agoutis locally, but within certain limits, may not interfere with seed dispersal and seed predation rates from the perspective of the plant. However, if hunting reduces agouti populations below a particular threshold in abundance, there are no longer enough animals to adequately maintain levels of seed predation and dispersal. Conceptualizing the disruption of plant-seed predator interactions in terms of thresholds is similar to the model proposed by Galetti and colleagues (2006), in which Astrocaryum seed predation rates are a sigmoidal or logistic, rather than linear, function of agouti abundance. Data from two defaunation gradients suggest this may be the case (Fig. 4.7). In Panama, sites with agouti densities 60%-80% lower than BCI still had high rodent seed predation rates equivalent to or higher than those observed on BCI, but where agouti densities were further reduced, seed predation rates dropped dramatically (Wright et al . 2001). Likewise in Brazil, sites with agouti abundances 80% lower than the maximum density observed still maintained agouti seed predation rates approximately equal to those observed in the sites with the highest agouti abundance (Donatti et al. 2009). 85

100 Above threshold Below threshold

20 Rodent seed predation (%) predation Rodent seed 0 A B C D E Literature Study

FIGURE 4.7. Published studies examining seed predation of large-seeded palms by rodents, primarily agoutis, across defaunation gradients consistently find that high rates of seed predation are maintained despite reductions in agouti densities up to 80% of maximum observed densities. Below this threshold, seed predation rates drop 48-96%. (A) Galetti et al . (2006), Astrocaryum aculeatissimum , under parent trees; (B) Donatti et al . (2009), A. aculeatissimum , controlled experiment; (C) Galetti et al .(2006), ), A. aculeatissimum , controlled experiment; (D) Wright et al . (2000), Astrocaryum standleyanum , “dispersed” seeds; (E) Wright et al . (2000), Attalea butyracea , “dispersed” seeds. At a community level, evidence has been found both for large, mammal- or primate-dispersed species decreasing in abundance in association with hunting (Nunez-Iturri et al . 2008, Terborgh et al . 2008), and community mean seed mass increasing in association with hunting (Wright et al . 2007, Ch. 2). Differences in the mammalian community context, in the relative levels of defaunation being compared, and differences in the traits and natural histories of plant species driving the community level patterns may all be contributing to such discrepancies in community level response to hunting. 86 CHAPTER 4: SEED PREDATION RATES AS A FUNCTION OF SEED SIZE AND DEFAUNATION

CONCLUSIONS

Overall, this empirical test of a theoretical model (Dirzo et al. 2007) of how seed predation rates should vary as a function of seed size and defaunation intensity found little general support for the model. In this system, there was only weak support for a relationship between seed mass and vertebrate seed predation rate, and no evidence that seeds were increasingly predated upon by larger-bodied seed predators as seed size increased by species. Furthermore, despite large differences in the abundances of key seed predators between a hunted and protected site, no differences were seen in vertebrate seed predation, or seed caching by agoutis. This work highlights the importance of considering both plant natural history and mammal community context when trying to predict the indirect effects of hunting on plant recruitment. It also is consistent with accumulating evidence that the functional roles that seed predators and seed dispersers play in tropical forests may not be linearly correlated with their abundance in a system, but rather, may be threshold-dependent.

ACKNOWLEDGEMENTS

The author is grateful to the Autoridad Nacional del Ambiente (ANAM) of Panama for permission to conduct work in Parque Nacional Soberanía, and to the Smithsonian Tropical Research Institute (STRI) for permission to conduct this work on BCI, and for providing logistical support. I would like to thank J. Wright for seed mass data and advice about sites. N. Beckman also provided useful advice in the planning stages of this experiment. The Fondo Peregrino–Panamá, in particular A. Muela, provided logistical support at the site in the Parque Nacional Soberanía. Many thanks go to C. Sherman, R. Bethancourt and R. Acosta for their hard work in the field and the lab, and to S. Rebellon for help with pilot work. D. Brassfield and O. Calderón provided help with species identification. I would like to thank O. Arosemena at STRI for her help with permitting, and all the staff on BCI whose help made conducting this work easier. R. Dirzo and C. Donatti provided comments which improved the manuscript. This project was made possible with funding from NSF DEB-0808338. 87

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