RECOVERY OF INSECTIVOROUS BIRD ECOLOGICAL FUNCTION IN TROPICAL FOREST RESTORATIONS By Steven M ichael Roels A DISSERTATION Submitted to Michigan State University in partial fulfillment of the requirements for the degree of Integra tive Biology Doctor of Philosophy Environmental Science and Policy Dual Major Ecology, Evolutionary Biology and Behavior Dual Major 2018 ABSTRACT RECOVERY OF INSECTIVOROUS BIRD ECOLOGICAL FUNCTION IN TROPICAL FOREST RESTORATIONS By Steven M ichae l Roels Recovery of animal - dependent ecosystem functions is a key component of ecological restoration on degraded tropical forest lands. One of these fun ctions, regulation of herbivorous insects by bird s, shapes tropical communities through trophic cascad es that protect trees from insect damage . Un derstanding how bird - driven trophic cascades var y with environmental context , re foresta tion strategy , and management would guide restoration practitioners seeking to facilitate beneficial relationships between bi rds and plants . My dissertation examines the speed of bird community recovery following restoration action , the effects of climate dynamism on bird - d riven trophic cascades, the mechanisms by which bird - driven trophic cascades occur, and insectivorous bird ecological function in forest restorations relative to other land cover types. In Chapter 1, I studi ed temporal trends and spatial patterns in bird a ctivity and diversity over a five - year period at a forest restoration in Panama. Four years after tree plan ting, bird activity had increased three - fold and species richness eleven - fold, compared to pre - planting. However, species richness differed strongly between experimental plots within the field site. Using a multi - species hierarchical occurrence model, I co nsidered possible reasons for this variation. Features within plots, such as the number of tree species planted, did not explain differences in the n umber of species observed. However, neighborhood context did explain differences; there were more species i n plots with more adjacent woodland and farther from pasture. My results demonstrate that native tree planting can generate rapid responses from trop ical bird communities but the surrounding matrix will mediate bird responses to restoration. In Chapter 2, I considered the ecological significance of insectivorous bird and bat activity in tropical forest restorations. I conducted bird and bat exclosure e xperiment s in a Panamanian forest restoration during a typical wet season, dry season, and wet season with an El Ni ñ o drought. A trophic cascade was present on one of two tree species studied but only during a wet season with normal rainfall. These results highlight the importance of resident bird species for regulation of damaging insects in forest restoration s and also point to the ability of regular seasonality and irregular El Ni ñ o events to a lter trophic relationships in restoration ecosystems. In Chapter 3, I more closely inspected the mechanisms that produced the trophic cascade described in Chapter 2. O n the tree species featuring a trophic cascade, predation by birds and bats during the n ormal wet season reduced biomass for five folivorous insect orders and reduced numbers of individuals for four of the five orders. Birds and bat effects were especially noticeable for larger - bodied arthropods. These patterns were weaker or absent on the tr ee species where a trophic cascade was not observed. Facilitation of bird - driven t rophic cascades is potentially a usef ul tool for tropical forest restoration but the presence of a substantial trophic cascade cannot be assumed or taken for granted. I recommend that land managers adopt a holistic approach to reforestation t hat restores bird habitat and encourages recovery o f animal - dependent ecosystem functions. iv This dissertation is dedicated to my parents, who encouraged me to play in the dirt, catch bugs, and chase birds, and to my wife, who still does. v ACKNOWLEDG E MENTS vi vii TABLE OF CONTENTS LIST OF TABLE S ................................ ................................ ................................ ..................... i x LIST OF FIGURES ................................ ................................ ................................ ................... x i KEY TO ABBREVIATIONS ................................ ................................ ................................ . xii i CHAPTER 1 : Recovery of B ird A ctivity and S pecies R ic hness in an E arly - stage T ropical F orest R estoration ................................ ................................ ................................ ....................... 1 ABSTRACT ................................ ................................ ................................ .......................... 2 I NTRODUCTION ................................ ................................ ................................ ................. 3 M ETHODS ................................ ................................ ................................ ........................... 8 Study site ................................ ................................ ................................ ......................... 8 Bird surveys ................................ ................................ ................................ ..................... 9 2014 plot content and context measurements ................................ ................................ . 11 Bird guild assignment ................................ ................................ ................................ .... 12 Statistical analyses ................................ ................................ ................................ ........ 13 R ESULTS ................................ ................................ ................................ ........................... 16 Multi - year tren ds in site - scale Activity and Species Richness ................................ ......... 16 Plot - scale Activity and Species Use four years after planting ................................ ......... 18 D ISCUSSION ................................ ................................ ................................ ..................... 21 Trends and turnover ................................ ................................ ................................ ...... 21 Site - scale Species Richness estimates four years post - planting ................................ ...... 23 Effects of content and contex t variables on birds ................................ ........................... 23 Occurrence model potential and limitations ................................ ................................ ... 25 Implications for bird conservation, ecological function, and forest restoration .............. 25 ACKNOWLEDGEMENTS ................................ ................................ ................................ . 27 APPENDIX ................................ ................................ ................................ .............................. 28 LITERATURE CITED ................................ ................................ ................................ ............. 34 CHAPTER 2: El Niño Disrupts Bird - and Bat - Dri ven Trophic Cascade in a Tropical Forest Restoration ................................ ................................ ................................ ..................... 42 ABSTRACT ................................ ................................ ................................ ........................ 43 INTRODUCTION ................................ ................................ ................................ ............... 44 METHODS ................................ ................................ ................................ ......................... 50 Study site and species ................................ ................................ ................................ .... 50 Field seasons ................................ ................................ ................................ ................. 51 Experimental protocol ................................ ................................ ................................ ... 52 Statistical analyses ................................ ................................ ................................ ........ 57 RESULTS ................................ ................................ ................................ ........................... 60 Leaf production ................................ ................................ ................................ ............. 60 Herbivory ................................ ................................ ................................ ...................... 60 Folivor ous insect biomass ................................ ................................ .............................. 63 Insect - eating bird surveys ................................ ................................ .............................. 65 viii DISCUSSION ................................ ................................ ................................ ..................... 66 ACKNOWLEDGEMENTS ................................ ................................ ................................ . 71 APPENDIX ................................ ................................ ................................ .............................. 73 LITERATURE CITED ................................ ................................ ................................ ............. 80 CHAPTER 3: Tropical Trophic Cascade Linked to Reduction in Large - bodied Insects by Birds and Bats ................................ ................................ ................................ ........................... 87 ABSTRACT ................................ ................................ ................................ ........................ 88 INTRODUCTION ................................ ................................ ................................ ............... 88 METHODS ................................ ................................ ................................ ......................... 92 Study site and species ................................ ................................ ................................ .... 92 Experimental protocol ................................ ................................ ................................ ... 92 RESULTS ................................ ................................ ................................ ........................... 94 DISCUSSION ................................ ................................ ................................ ................... 102 ACKNOWLEDGEMENTS ................................ ................................ ............................... 107 APPENDIX ................................ ................................ ................................ ............................ 108 LITERATURE CITED ................................ ................................ ................................ ........... 110 CHAPTER 4: Predation P ressure by B irds and A rthropods on H erbivorous I nsects A ffected by T ropical F orest R estoration S trategy ................................ ................................ .... 116 ABSTRACT ................................ ................................ ................................ ...................... 117 INTRODUCTION ................................ ................................ ................................ ............. 118 METHODS ................................ ................................ ................................ ....................... 122 Study site ................................ ................................ ................................ ..................... 122 Experimental land covers ................................ ................................ ............................ 123 Caterpillar construction ................................ ................................ .............................. 124 Caterpillar placement criteria ................................ ................................ ..................... 125 Exposure season, exposure duration, and model recovery ................................ ............ 126 Bird s urveys ................................ ................................ ................................ ................. 127 Statistical methods ................................ ................................ ................................ ....... 128 RESULTS ................................ ................................ ................................ ......................... 130 Caterpillars exposed, recovered, and attacked ................................ ............................. 130 Reliability of observers ................................ ................................ ................................ 131 Effect of shape and posture on attack probability ................................ ......................... 131 Land cover and attack probability ................................ ................................ ............... 131 Evidence for spillover pre dation ................................ ................................ .................. 132 Bird surveys ................................ ................................ ................................ ................. 133 DISCUSSION ................................ ................................ ................................ ................... 134 Implications for ecosystem management and restoration ................................ ............. 13 6 Methodological considerations ................................ ................................ .................... 137 Conclusion ................................ ................................ ................................ .................. 138 ACKNOWLEDGEMENTS ................................ ................................ ............................... 138 APPENDIX ................................ ................................ ................................ ............................ 140 LITERATURE CITED ................................ ................................ ................................ ........... 144 ix LIST OF TABLES Table 1.1 . Tree species planted by experimental plot ................................ ................................ . 29 Table 1.2 . Results from five years of bird surveys at a forest restoration in the Mamoní Valley, Panama ................................ ................................ ................................ ................................ ..... 30 Table 2.1 . Bird survey results by field season ................................ ................................ ............ 66 Table 2. 2 . Qualitative summary of environmental conditions and results for Ochroma pyramidale (O.p.) and Terminalia amazonia (T.a.) across three seasons ................................ .... 66 Table 2. 3 . Biometric equations for arthropod length and dry biomass ................................ ........ 74 Table 2. 4 . Effects of predator exclosures and seasonality on herbivory for fully flushed O . pyramidale leaves ................................ ................................ ................................ ..................... 75 Table 2. 5 . Effects of predator exclosures and seasonality on herbivory for still - expanding O. pyramidale leaves ................................ ................................ ................................ ..................... 75 Table 2. 6 . Effects of predator exclosures on herbivory for T . amazonia leaves in the 2015 drought ................................ ................................ ................................ ................................ ..... 75 Table 2. 7 . Eff ects of predator exclosures on her bivory for T . amazonia leaves in the 2016 wet season ................................ ................................ ................................ ................................ ....... 75 Table 2. 8 . Seasonal differences in folivorous insect biomass on O . pyramidale ......................... 76 Table 2. 9 . Seasonal differences in folivorous insect biomass on T . amazonia ............................ 76 Tab le 2. 10 . P redator exclosure effects on folivorous insect biomass on O . pyramidale in the 2015 drought ................................ ................................ ................................ ................................ ..... 76 Table 2. 11 . P redator exclosure effects on folivorous insect biomass on O . pyramidale in the 2016 dry season ................................ ................................ ................................ ................................ . 76 Table 2. 1 2 . P redator exclosure effects on folivor ous insect biomass on O . pyramidale in the 2016 wet season ................................ ................................ ................................ ................................ . 7 7 Table 2. 1 3 . P redator exclosure effects on folivorous insect biomass on T . amazonia in the 2015 drought ................................ ................................ ................................ ................................ ..... 77 x Table 2.1 4 . P redator exclosure effects on folivorous insect biomass on T . a mazonia in the 2016 dry season ................................ ................................ ................................ ................................ . 77 Table 2.1 5 . P redator exclosure effects on folivorous insect biomass on T . amazonia in the 2016 wet season ................................ ................................ ................................ ................................ . 7 8 Table 3.1 . Biomass (mg) of arthropod orders per m 2 leaf area during pre - and post - experiment surveys on control (Ctr) and exclosure (Exc) branches of O. pyramidale ................................ ... 96 Table 3.2 . Census counts of arthropod orders (individuals per m 2 leaf area) during pre - and post - e xperiment surveys on control (Ctr) and exclosure (Exc) branches of O. pyramidale ................. 96 Table 3.3 . Bioma ss (mg) of arthropod orders per m 2 leaf area during pre - and post - experiment surveys on control (Ctr) and exclosure (Exc) branches of T. amazonia ................................ ...... 98 Table 3.4 . Census counts of arthropod orders (individuals per m 2 leaf area) during pre - and post - experimen t surveys on control (Ctr) and exclosure (Exc) branches of T. amazonia .................... 99 Table 4.1 . Caterpillars and controls recovered by land cover and distance - t o - forest treatment ................................ ................................ ................................ ................................ . 129 Table 4 . 2 . Generalized linear model for effects of s hape (caterpillar vs. control ) and p osture (straight - bodied vs. i nchworm) on attack p robability. ................................ ................................ .................. 141 Table 4.3 . Generalized linear model for e ffects of p roximity to natural forest (gallery or national park) on attack p robability ................................ ................................ ................................ ....................... 141 xi LIST OF FIGURES Figure 1. 1 . Multi - year trends in site - scale Activity for four habitat guilds following tree planting in a forest restoration in the Mamoní Valley, Panama ................................ ................................ 17 Figure 1. 2 . Estimates of site - scale bird Species Richness following tree planting ...................... 18 Figure 1.3 . Plot - scale bird Activity (uppe r panel) and observed Species Use (lower) by plot in 2014 Single - year survey ................................ ................................ ................................ ............ 19 Figure 1.4 . Effect of plot variables on Species Use ................................ ................................ .... 20 Fi gure 1.5 . Relationships between plot context variables, Adjacent woodland and Pasture distance, and plot use by bird species ................................ ................................ ........................ 20 Figure 1.6 . Posterior probabilities for effects of plot variables for individual species, sorted by habitat guild. ................................ ................................ ................................ ............................. 32 Figure 2.1 . Illustration of trophic cascade predictions ................................ ................................ 49 Figure 2.2 . Location of the Agua Salud Project (star), Colón Prov ince, Panamá ) ................................ ................................ ................................ ........................... 50 Figure 2.3 . Exclosures on O. pyramidale (left) and T. amazonia (right) ................................ ..... 52 Figure 2. 4 . Leaf produ ctivity by tree species and season ................................ ............................ 60 Figure 2. 5 . Herbivory on fully flushed leaves of O. pyramidale by season an d treatment ........... 61 Figure 2. 6 . Herbivory on still - expanding leaves of O. pyramidale by season and treatment ....... 62 Figure 2. 7 . Herbivory on T. amazonia by season and treatment ................................ ................. 63 Figure 2. 8 . Total biomass of folivorous insect orders by season on O. pyramidale ..................... 6 4 Figure 2. 9 . Total biomass of folivorous insect orders by season on T. amazoni a ........................ 64 Figure 2. 1 0. Biomass of folivorous insects on O. pyramidale in the 2016 wet season ................ 65 Figure 2. 11 . Linear regression models for the effect of start time on number of bir ds counted ... 79 Figure 3. 1 . Arthropod biomass by order on O. pyramidale before (Pre) and after (Post) the exclosure experiment ................................ ................................ ................................ ................ 95 xii Figure 3. 2 . Census counts of arthropod orders on O. pyramidale before (Pre) and after (Post) the exclosure experiment ................................ ................................ ................................ ................ 97 Figure 3. 3 . Arthropod biomass by order on T. amazonia before (Pre) and after (Post) the exclosure experiment ................................ ................................ ................................ ................ 9 8 Figure 3. 4 . Census counts of arthropod orders on T. amazonia before (Pre) and after (Post) the exclosure experiment ................................ ................................ ................................ ................ 99 Figure 3. 5 . Distribution of arthropod sizes on O. pyramidale for control branches (top panel) and exclosure branches (bottom panel) before (Pre) and after (Post) the exclosure experiment ....... 100 Figure 3. 6 . Distribution of arthropod sizes on T. amazonia for control branches (top panel) and exclosure branches (bottom panel) before (Pre) and after (Post) the exclosure experiment ....... 101 Figure 3. 7 . Arthropod biomass on O . pyramidale by order before (Pre) and after (Post) the exclosure experiment without arthrop ods >90 mg removed ................................ ..................... 109 Figure 3. 8 . Arthropod biomass on T. amazonia by order before (Pre) and after (Post) the exclosure experiment without arthropods >90 mg removed ................................ ..................... 109 Figure 4. 1 . Caterpillar models wired onto branch tips ................................ .............................. 125 Figure 4. 2 . Attacks on models by shape (caterpillar vs. control ball) and p redator class ........... 130 Figure 4. 3 . Daily attack rates by bird and arthropods on model caterpillars vary by land cover ................................ ................................ ................................ ................................ ....... 132 Figure 4. 4 . Insect - eating bird abundance in three land covers ................................ .................. 133 Figure 4. 5 . Agua Salud Project location ................................ ................................ ...................... 142 Figure 4. 6 . Agua Salud Project m ap of land cover types, Area A ... 142 Figure 4. 7 . Agua Salud Project map of land cover types, Area B (9°1 1 49 4 5 143 xiii KEY TO ABBREVIATIONS ASP Agua Salud Project cm centimeter hr ho ur m meter mg milligram min minute O.p. Ochroma pyramidale STRI Smithsonian Tropical Research Institute T.a. Terminalia amazonia 1 CHAPTER 1 : Recovery of Bird A ctivity and Species Richness i n a n Early - stage T ropical F orest R estoration Steven M. Ro els 1,2,3 , Melissa B. Hannay 1,3 and Catherine A. Lindell 1,3,4 1 Department of Integrative Biology 2 Environmental Science and Policy Program 3 Ecology, Evolutionary Biology, and Behavior Michigan State University, 288 Farm Lane, East Lansing, Michigan, US A 44824 4 Center for Global Change and Earth Observations Michigan State University, 1405 S. Harrison Road, East Lansing, Michigan, USA 48823 2 ABSTRACT Creation of bird habitat is often a goal of tropical forest restoration because bird - driven ecosystem services can accelerate forest recovery. However, resident tropical bird responses are not well - characterized in the earliest years following restoration action. During a five - year study of the bird community in an experimental tropical forest restoration , we exam ined temporal trends in bird activity and diversity and the effects of habitat variables on the distribution of bird species within the site. Our site consisted of 16 replicate plots with 1, 2, 4, or 8 native tree species planted into former heavi ly - grazed pasture. Four years after tree planting, we observed a three - fold increase in bird activity and eleven - fold increase in species richness compared to pre - planting. We also found changes in proportions of habitat guilds, with marked declines in ope n country birds and increases in birds associated with brushy, early secondary growth, and forest edge habitats. Number of bird species observed differed strongly between experimental plots four years post - planting. Using a Bayesian multi - species hierarchi cal occur rence model, we considered possible reasons for these differences related to plot content and context. Content species identity, and presence of legacy trees, d id not explain differences in number of bird species observed, potentially because of small plot size relative to bird mobility. Neighborhood with more adja cent woodland and farther from actively grazed pasture. Our results demonstrate that planting native tree species in highly degraded sites can generate rapid, positive responses from tropical bird communities. These responses are likely mediated b y surroun ding habitat matrix, which influences rates of bird community recovery. Considering site context can improve predictions of fine - scale distribution of bird activity and diversity within restoration sites. 3 INTRODUCTION Interest in forest restoratio n is increasing in tropical regions because forests provide significant ecosystem services such as carbon sequestration, timber production, recreational opportunities, and wildlife habitat (Lamb et al. 2005, Chazdon 2008, Rodri gues et al. 2011, Suding et a l. 2015). Although disturbed tropical forests can recover via natural regeneration (Aide et al. 2000), native tree planting may accelerate forest and ecosystem service recovery (Parrotta et al. 1997, Carnevale and Montagnini 20 02, Fink et al. 2009, Holl et al. 2016). Recovery of wildlife habitat is often a rationale for forest restoration (Miller and Hobbs 2007) but how wildlife populations respond to different restoration techniques is often uncertain. Furthermore, wildlife - pla nt interactions can influence long - term restoration outcomes via mutualisms, herbivory, and trophic cascades so an improved understanding of wildlife activity during the early stages of forest restoration is important for restoration planning (Fraser et al . 2015, McAlpine et al. 2016) . Among wildlife taxa, birds are conspicuous early responders to forest restoration (Dunn 2004, MacGregor - Fors et al. 2010, Lindell et al. 2012, Rolo et al. 2017). Due to their diverse life history strategies including wide va riation in diet, specialized foraging strategies, and micro - niche preferences, birds can act as an indicator taxon for recovering ecological complexity during forest restoration (Da Silva and Vickery 2002). In addition, birds provide ecosystem functions th at facilitate forest recovery including seed dispersal, pollination, and herbivorous insect reduction (Sekercioglu 2006, Morrison and Lindell 2012, Lindell et al. 2013, Frick et al. 2014, Carlo and Morales 2016). Bird - driven ecosystem functions are tied to activity levels and composit ion of the local - scale bird community and should be closely related to habitat affiliations of species using restoration sites. For example, bird species that regularly use trees 4 are more likely to carry seeds of woody species than open - country species (Li ndell et al. 2013). Forest restoration sites experience rapid changes in the bird community during the early stages of forest recovery, potentially affecting bird ecosystem function (Gould and Mackey 2015, Lindenmayer et al. 20 16). The return of forest - as sociated birds to forest restorations is a metric of restoration progress (Twedt et al. 2002, Nichols and Nichols 2003, Rolo et al. 2017). Increasing compositional overlap with forest bird communities at less disturbed referenc e sites indicates progress to wards biodiversity conservation goals (Catterall et al. 2012, Rolo et al. 2017). In restorations intended as wildlife corridors between forest fragments, forest - associated birds demonstrate restorations are functioning as plann ed (Jansen 2005). Finally, fo rest - associated species can carry seeds and pollen from forest plant species into restorations, encouraging the re - establishment of a diverse native plant community (Wunderle 1997, Frick et al. 2014, Carlo and Morales 2016). Pr ior studies of regenerating t ropical forests have found bird activity and species richness rapidly increase in the first decade or two, although a return to compositional similarity with primary forest bird communities may take over a century (Dunn 2004, C atterall et al. 2012, Paxton et al. 2018). Greater insight into bird responses to fine - scale habitat characteristics in forest restorations will aid restoration practitioners seeking to create wildlife habitat or harness bird ecosystem services to acceler ate the restoration process. Choices made at project initiation, such as the identity and number of tree species planted, affect development of characteristics like vegetation density and canopy cover (Holl et al. 2013, Rolo et al. 2016). Such structural c haracteristics can, in turn, 5 affect site attractiveness to birds several years after tree planting (Fink et al. 2009, Lindenmayer et al. 2010, Lindell et al. 2012, Lindell et al. 2013). Other fine - scale habitat characteristics may be less easy to alter, su ch as legacy features (e.g., old snags or land use history) or land use on adjacent parcels with different ownerships, but are also potentially relevant to birds and thus restoration planning. Studies of bird recovery in forest restorations frequently sub stitute space for time, using chronosequences to assess trajectories in bird abundance and diversity (e.g., Catterall et al. 2012, Rolo et al. 2017). However, chronosequences with limited temporal resolution may not provide insights into the pace of bird r ecovery during earliest years , when change should be most dynamic (Jansen 2005, Paxton et al. 2018). Setting expectations for the earliest years, such as when the first forest - associated birds will appear, is especially important for restoration projects t hat plan to use an adaptive m anagement framework with ongoing monitoring (Murray and Marmorek 2003). As such, longitudinal data on bird responses to restoration, starting at project initiation, are useful. We conducted a five - year study of the bird commun ity in a plantation - style for est restoration in Panama, beginning at the initiation of restoration efforts. Prior research in bird responses to forest restoration strategies has examined effects of planting configuration (Fink et al. 2009, Lindenmayer et a l. 2016), understory enrichme nt plantings (MacGregor - Fors et al. 2010), and fast - growing nurse trees (Hamel 2003). Our study site featured replicate plots to compare bird responses to restoration planting regimes with different numbers and mixes of tree sp ecies. Tropical forest restor ation efforts are challenged by extremely high diversity of tree species 6 present in natural forests. In central Panama, a single wet lowland forest site can harbor over 300 tree species (Condit et al. 1992). Restoration with a large component of the native tree diversity (e.g. Rodrigues et al. 2009) is ideal but practical considerations, such as seed collection, tree nursery space, and labor costs, mean that plantation - style tropical forest restorations typically occur with far less than a full complement o f old - growth forest species. Nevertheless, we hypothesized that planting a few tree species would still be sufficient to quickly induce a strong response from the bird community. A previous restoration study in Central America showed that only 2 4 years of growth by four planted tree species was enough to attract a diverse bird community to heavily degraded areas (Fink et al. 2009; Lindell et al. 2012). At the scale of our entire site (all plots aggregated), we predicted bird a ctivity and species richness would increase quickly after restoration initiation because high plant productivity in the tropics supports rapid development of vegetative structure, long known to be a determinant of bird diversity (MacArthur and MacArthur 19 61, Karr 1968). We also antic ipated vegetative development would result in species turnover as species associated with earlier seral stages abandoned the site (Twedt et al. 2002, Gould and Mackey 2015). At the scale of individual plots, we hypothesized th at plot planting regime would influence the number of bird species recorded. Specifically, we expected the number of bird species in a plot to be positively associated with greater canopy cover, more planted tree species, and the presence of a particular p lanted tree genus, Inga . Cano py cover and number of tree species have previously been associated with greater bird species diversity in neotropical coffee plantations (Van Bael et al. 2007, Philpott et al. 2008). Many tropical bird species rely on the cano py layer 7 as foraging, resting , and breeding habitat. Higher tree species diversity could affect bird diversity by providing complementary resources, such as variable foraging substrates, host - specific arthropod prey, open branch structures for foraging, an d dense branch structures for nesting and predator protection. However, the original authors who emphasized the importance of vegetative structure to bird also regarded plant species diversity per se as having little effect on bird diversity (MacArthur and MacArthur 1961, Karr 1968). The tree genus, Inga (Family: Fabaceae), has a bushy growth form and tends to produce canopy cover more rapidly than other kinds of trees planted at our site. Inga trees have previously been shown to be an attractive habitat fe ature for birds (Fink et al. 2009). In addition to aspects of our planting regimes, we expected a positive response of bird diversity to the presence of legacy trees that remained in the former pasture; such trees can draw a variety of woodland species int o otherwise open country (Fis cher and Lindenmayer 2002). During our study, we noticed strong spatial patterns in the distribution of bird activity and number of species observed within plots that appeared unrelated to planting regime or legacy trees. We collected data on the distance to nearest pasture and amount of woodland adjacent to our plots to examine the hypothesis that habitat features external to plots influenced number of species observed. Understanding the relative importance of resto features neighborhood) is key to effective restoration planning for faunal recovery (Lindenmayer et al. 2010, Reid et al. 2014, Gould and M ackey 2015). 8 METHODS Study site We con small agricultural plots, secondar y growth of various ages, tree plantatio ns, and primary forest fragments. In July 2010, sixteen 50 m x 50 m forest restoration plots were established on 4 ha of recently abandoned heavily - grazed pasture featuring only a few, scattered trees. Plots were arr anged in a loose grid with 12 - m minimum buffers between plots, avoiding steep slopes and excessively wet soils that could negatively affect the survival and performance of planted trees. Each plot was manually cleared of herbaceous vegetation and planted with seedlings of native trees grown fr om locally collected seed. Seedlings were planted in a grid with 3 m between rows, resulting in 256 seedlings per plot. Plots were planted with 1, 2, 4, or 8 species with equal shares in multi - species plots (s ee Appendix, Table 1 .1 for planting mixes). Tre e species were selected based on a variety of characters including ease of seed germination, ability to act as a wildlife resource (e.g. nectar producing flowers, edible fruits), potential for eventual timber harvest, and performance in plantation contexts . Planting mixes were chosen based on seedling availability and also to include at least one nitrogen - fixing legume (Family Fabaceae). One tree type, Inga , was discovered post - planting to be a mix of several I nga species. The majority were I. thibaudiana , but I. stenophylla , I. ruiziana , and I. sapindoides were also planted (Rolando Perez, personal communication). Because these Inga species have similar growth forms, we grouped Inga ses. The seedlings of one species ( Minquartia g uianensis ) in the plots planted with 8 species died and were replaced in 2011, primarily with a new species 9 ( Hieronyma alchorneoides ) and secondarily with other species already present in those plots. To reduc e competition for the young trees, non - planted vegetation in plots and buffers was manually cleared multiple times in the first two years and again in the fifth year. Although we did not measure tree growth, we observed that canopy height four years after planting was typically 3 5 m, with some tall tr ees exceeding 6 m. Bird surveys Activity can increase due to 1) more individuals using a survey location; 2) more ti me by the same individuals at a survey location ; or 3) a combination of 1 and 2. We did not attempt to distinguish between these possibilities because our objectives were to characterize general habitat suitability for birds and potential changes in bird e cological function. To distinguish between numb - - f unique species across all counts for all plot Richness four years post - planting (see Analyses) but did not do so for Activity due to data limitations. Multi - year survey protocol To measure changes in Activity and Species Richness through time, we conducted annual bird surveys mid - rainy season (July - - baseline survey prior to tree planting in 2 010. In each plot, we conducted 12 - min area cou nts, 10 recording all birds present by sight and sound while taking care to avoid double - counting of individuals within counts. We used two to four vantage points in each plot to limit the effect of obstructive v egetation and terrain on bird detection. Counts were conducted four times in each plot within a single day each year, with a minimum of 15 min between counts. All counts were conducted between 0630 1100 hr. Daily temperatures were fairly consistent (low: ~ 23°C, high: ~32°C) but morning fog, cloud cover , and scattered rain showers were highly variable. To reduce detection bias, counts were not conducted during rain, fog, or high winds. However, other weather variables such as cloud cover may have affected de tection. Birds flying overhead or flying throug h plots without landing or foraging were not included in the final data set. 2014 Single - year survey protocol and mist netting In July Single - year arding plot content and context effects on Species Use. Repeated 10 - min area counts were made in each plot except Plot 16, which was excluded due to difficult access. The change from 12 - min (multi - year survey protocol) to 10 - min counts was due to the need to accommodate travel time between plots while still completing surveys in the morning hours. All plots were surveyed once each day and visited in a set sequence with a randomized starting plot. Count protocol was otherwise similar to the annual protocol. Ten surveys were conducted over a 20 - day period, resulting in a Single - year data set containing 10 replicates for each plot. As a supplement to our surveys in 2014, we operated mist nets at ten locations acros s the site for total of 180 net hours from May July 2014. Mist net sampling is a complementary approach to point counts that can effectively detect taxa often undersampled by point counts in tropical forests (e.g. Blake and Loiselle 2001). Net operations w ere conducted 11 under fair weather conditions bet ween 0600 1130 hr and nets were checked every half hour. Captures were identified to species and were banded or had a tail feather trimmed to identify re - captures. 2014 plot content and context measurements In 2014, we collected data for six variables tha t described the plot (content variables) or surrounding neighborhood (context variables). Content variables were: 1) Percent canopy amount of canopy cover in a plot; 2) Tree species the numbe r of tree species planted in a plot; 3) Inga whether Inga s pp. were planted in a plot; and 4) Residual trees whether legacy trees greater than 5 m tall were present in a plot prior to planting. Context variables were: 5) Adjacent woodland percentage of land adjacent to plot that was forest restoration or fores t fragment; and 6) Pasture distance distance from plot center to nearest actively grazed pasture. To quantify canopy cover, we divided each plot into quadrants and took c anopy measurements wi th a spherical densiometer at 3 random points per quadrant (1 2 points per plot). Measurement points were between rows of planted trees to avoid inflating canopy cover values by standing directly next to tree trunks. Densiometer readings were taken in the f our cardinal directions and then averaged to create the perce nt canopy estimate for an individual point. The mean of all original plot design and were confirmed with tree survivorshi 12 r at a set distance of 25 m from each of the four plot edges. We defined woodland as any area of forest restoration or secondary forest. Forest restoration included an adjacent restoration planting with a similar tree species mix initiated one year prior t o our study. Non - woodland vegetation included active pasture, weedy fields, marsh, and brushy areas. To create a single value for each plot, we lated in QGIS 2.18 (QGIS Development Team 2017) using the str aight - line distance between plot center points and the nearest pasture. Bird guild assignment We assigned all recorded bird species to habitat guilds based on published natural history descriptions (Ridgely 1981, Stiles and Skutch 1989, Angehr and Dean 20 10, Neotropical Birds 2018) and personal experience ( Appendix, Table 1. 2). We placed species in to one of four guilds: Open Country, Brushy, Early Secondary, and Forest Edge. Open Country habitats are characterized by a lack of woody vegetation, although pe rches such as isolated tall trees may be present. Brushy habitats are weedy and have woody grow th less than 2 m tall. Early Secondary habitats feature dense, woody growth greater than 2 m tall. Forest Edge habitats are established woodland/forest adjacent to other, less heavily wooded areas. Individual birds not identified to species level were assi gned to guilds based on partial identifications if possible (e.g. all members of the genus were in the same guild). 13 Statistical a nalyses Changes in Activity a nd Species Richness through time All analyses were conducted in R (version 3.4.3; R Development Core Team, 2017). Activity and Species Use were low in individual plots during the first three years of the study (many plots with 1 or 0 species recorded) so w e investigated trends in the Multi - year data for the restoration site as a whole, ra ther than conduct plot - scale analyses. To facilitate analyses of Multi - year data, we pooled all plot - scale surveys within a year into a single site - scale data set. Thus, an nual site - scale Activity was the total number of birds recorded each year. Annual si te - scale observed Species Richness was the total of unique species recorded each year. To estimate actual Species Richness (including hypothetical species present but unobs erved) for a given year, we used the first - order jackknife estimator (Walther and Mo - order jackknife is a resampling technique that estimates undersampling bias using the numb only once in the data s et. As applied to species richness estimation, the first - order jackknife is calculated as: S actual = S observed + ( n - 1 / n , where S is the number of species, n is the sampl e size, 1 is the number of singletons. To compare survey methods (Multi - year vs. Single - year) and richness estimation techniques, we also produced Species Richness estimates for 2014 from the Single - year data set using the first - order jackknife and ou r occurrence model (see below). Species Use responses to plot content and context variables To examine effects of content and context variables on use of plots by birds, we analyzed the Single - year (2014) data set collected by SR in a multi - species hierar chical occurrence model (Dorazio et al. 2006, Zipkin et al. 2010). Thi s class of models uses aggregate occurrence data for 14 all species and survey replicates to improve parameter estimates for rare and unobserved species and also accounts for imperfect dete ction of individuals inherent in avian surveys. In our context, with p lot sizes smaller than the typical home range of most bird species, we interpret the model as estimating the effect of variables on probability that a given species will use habitat in a plot rather than site occupancy per se . Our full model had two compo nents: an occurrence model and a detection model. The occurrence model assesses effects of six plot variables on plot use by birds, which is assumed to be the outcome of a Bernoulli rand om variable for each species. This distribution is specified as: z_( i, j )~Bern( ;_( i,j ), where z is the actual use, i is the species, and j is the plot. Continuous variables (Percent canopy, Tree species, Adjacent woodland, and Pasture distance) were stand ardized by subtracting the mean and dividing by one standard deviation . Categorical variables (Inga and Residual trees) were treated as binary (0 or 1). To guard against problems with collinearity, we confirmed all pairwise correlations of variables had r < |0.7| (Dormann et al. 2013). The occurrence model for species i at p lot j is represented as: i,j i i PercentCanopy j i TreeSpecies j i Inga j + i ResidualTrees j i AdjacentWoodland j i PastureDistance j i represents the probability of plot use (on a logit scale) by species i given mean values for continuous variabl es and zero values for binary variables (i.e. Inga and Residual trees n plot use by species i . Effects are either linear (continuous variables) or the effect of changing states (bina ry variables). 15 The detection model assesses the effect of survey Date and Time on the detection of birds. Like species occurrence, species dete ction is assumed to be the outcome of a Bernoulli random variable for each species. This distribution is specifi ed as: y_( i,j,k )~Bern(p_( i,j,k )*z_( i,j )), where y is the observed use, p is the probability of detection, i is the species, j is the plot, and k is the replicate. Thus, our observed use is the product of the probability of detection (p) and actual use (z). Both predictor variables were standardized by subtracting the mean and dividing by one standard deviation. The detection model for species i at plot j in replicate k is represented as: logit( p i,j,k i i Date j i Time j i represents the probability of detection (on a logit scale) for species i given mean i i represent the linear effects of each survey variable on detection of species i . We analyzed our model in a Bayesian framework in R and JAGS data with 40 addition assessed effects of model variables on plot use by the bird community as a whole by examining posterior estimates for parameters at the community level of the hierarchical model. We used - and varianc e=2.70, as proposed by Lunn et al. (2012). The shape and scale parameters of the gamma priors for the variance parameters were set to 0.1. The h yper - parameters govern estimation of species - level occurrence and detection parameters, which we assume are draw n from the corresponding community - level distributions (Zipkin et al. 2010). We ran three MCMC (Markov chain Monte Carlo) chains for 275,000 ite rations with a burn - in of 225,000 (10,000 in 16 the adaptive phase). Posterior chains were thinned by 5, yielding a total of 30,000 estimates for each model parameter. We used the R - hat statistic to confirm model convergence (Gelman and Hill 2007). RESULTS Over the Multi - year survey (five years), we recorded 40 bird species with 607 occurrence records ( Appendix, Tabl e 1. 2). In the Single - year survey of 2014, four years post - planting, we recorded 525 birds representing 41 species ( Appendix, Table 1. 2). That same year, our mist nets captured 252 birds representing 47 species ( Appendix, Table 1. 2). Multi - year trends in site - scale Activity and Species Richness Site - scale Activity and observed Species Richness were strongly correlated post - plant ing (Multi - year data: 2011 2014, r = 0.988). Activity prior to tree planting was moderate and initially declined after planting be fore a sustained three - year increase (Fig. 1. 1). Four years after tree planting, our Multi - year survey protocol recorded over three times as much Activity as before planting (Fig. 1 .1 ). Open Country Activity decreased rapidly following tree planting and re mained at low levels throughout the study. One Open Country species, Eastern Meadowlark ( Sturnella magna ), represented 89.9% o f Activity (62 of 69 birds identified) pre - planting and was not recorded after one year post - planting. All other habitat guilds sh owed year - to - year gains in Activity. 17 Figure 1 .1 . Multi - year trends in site - scale Activity for four habitat guilds following tree planting in a forest restoration in the Mamoní Valley, Panama . Some birds could not be assigned to a guild because of insu fficient taxonomic identification during field surveys . Site - scale Species Richness was low prior to tree planting and did not significantly change until two years post - planting (Fig. 1. 2). Four years after tree planting, we observed eleven times as many species as before planting (Fig. 1. 2). The actual diff erence in richness may be even greater than directly observed; estimates of true richness (total observed and unobserved species) increased at a faster rate than observed species. First - order jackknife estimates of Species Richness four years after plantin g were similar between Multi - year (46.1 species, 95% confidence interval: 34.9 - 57.4) and Single - year data sets (51.3 species, 95% confidence interval: 41.6 61.0) (Fig. 1. 2). The Single - year occurrence m odel estimate (56.1 species, 95% credible interval: 46 .0 74.0) was also similar to the Single - year jackknife estimate (Fig. 1. 2). All habitat guilds except Open Country demonstrated a general pattern of increasing Species Richness from two years after plan ting onward ( Appendix, Table 1. 2). 18 Figure 1. 2. Estimates of site - scale bird Species Richness following tree planting. For the Multi - year survey data set, we present observed species and the first - order jackknife estimate. For the Single - year data set (2 014), we present observed species, the first - order jackknife estimate, and the occurrence model estimate. Error bars are 95% confidence or credible intervals for the jackknife and occurrence estimates, respectively. Plot - scale Activity and Species Use fou r years after planting In the Single - year survey of 2014, plot - scale Activity and Species Use were strongly correlated (r = 0.935), with a tendency toward higher values in southeastern plots and lower values in northern plots (Fig. 1. 3). At the community level (i.e. hyper - parame ters) of our occurrence model, the context variables Adjacent woodland and Pasture distance were significant (95% credible intervals did not overlap 0) while all content variables were not (Fig. 1. 4). Accounting for unobserved speci es in our model made the univariate relationships between context variables and plot - scale Species Use more evident (Fig. 1. 5). The regression slope for the relationship between estimated Species Use and Adjacent woodland predicts one additional species fo r every ~2.8% increase i n Adjacent woodland (Species = 16.67 + 0.359*Adjacent woodland, r 2 = 0.431). For the relationship between the estimated Species Use and Pasture distance, the model predicts 19 one additional species for each additional ~10.8 m from the nearest pasture (Specie s = 15.72 + 0.0929*Pasture distance, r 2 = 0.745). In the detection component of our model, Date was a significant variable, with higher detection on later dates (mean: 0.35, 95% credible interval: 0.14 0.56), while Time was not sign ificant (mean: - 0.04, 95 % credible interval: - 0.19 0.11). 20 Figure 1. 4. Effects of plot variables on Species Use. 95% credible intervals are presented for posterior distributions of six plot variables. Variables are: 1) Percent canopy amount of c anopy cover in plot; 2) Tree species the number of tree species planted in plot; 3) Inga Inga spp. were planted in plot; 4) Residual trees legacy trees (alive or dead) were present in plot; 5) Adjacent woodland percentage of land adjacent to plot that was forest restoration or forest fragment (see text for explanation); and 6) Pasture distance distance from plot center to nearest actively grazed pasture. Credible intervals not over lapping zero are deemed significant variables and are denoted with a *. Figure 1. 5. Relationships between plot context variables, Adjacent woodland and Pasture distance, and plot use by bird species. The bird community at the restoration site contained many species that irregularly used plots, leading to low estima tes of detection. Accounting for detection with a multi - species occurrence model, yields estimates for the true number of species using eac h plot higher than was actually observed. Estimating true use may more effectively recover relationships between plot variables and use by birds. 21 At the species level, our occurrence model produced posterior distributions with wide credible intervals and point estimates that rarely deviated significantly from those of community - level hyper - parameters ( Appendix, Fig. 1 .6 ). Low detection of Species Use, likely due to single plots being smaller than typical home range sizes, limited our ability to recover evi dence of species - specific responses to plot variables. For example, several species demonstrated use of a plot (i.e. at least one record) but were only recorded once out of ten replicate surveys making it difficult to e plots where they were not recorded. We were also unable to discern any patterns in responses to plot variables when species were grouped by guild ( Appendix, Fig. 1.6 ). DISCUSSION Trends and turnover Over four years, we documented an increase in Activity and Species Richness at the restoration site, supporting our hypothesis that planting even a limited number of tree species can elicit a strong, rapid response from a tropical bird community. However, there was also an unexpected two - year time lag post - planting before increases in Activity and Species Richness were realized, which included a temporary drop in Activity o ne year post - planting. The few studies that have conducted annual monitoring starting with restoration initiation show varying results for the speed of avian response, possibly due to landscape context. Paxton et al. (2018) reported low bird densities in a Hawaiian montane forest region during the first five years post - planting, with more substantial gains occurring in the following twenty years. Jansen (2005) observed a strong response by birds in the first three years, including forest - associated species, to tree planting in an Australian rainforest region. Such differences are likely in part due to variable distance to 22 more than approximately 200 m. Our site was approximately 500 m from extensive forest but with smaller fragments at shorter distances. Different survey methodologies and community sizes make comparisons between studies difficult, but we regard the speed of bird response in our study to be intermediate to rates observed by Paxton et al. (20 18) and Jansen (2005). A meta - analysis of tropical forest bird community recovery found that species richness, standardized to that of mature forest, typically recovers a fter 20 years of forest regeneration (Dunn 2004). Exhaustive sampling of wet lowland forest in central Panama by Robinson et al. (2000) yielded average species densities of 111 resident species per 2 ha plot. Although our study site was slightly larger (4 ha), this suggests that species richness at our site had returned to approxim ately half of what would be expected in mature forest only four years after native tree planting. The decline in Open Country guild dominance presents a case of rapid faunal turn over during the transition from pasture to young restored forest. This shift was essentially due to disappearance of Eastern Meadowlark, the most abundant species prior to restoration. Unlike continued substantial increases in Activity by Early Secondary a nd Forest Edge guilds, Brushy guild Activity leveled off from three to four y ears post - planting (Fig. 1 .1 ). This may indicate a peak for this guild, which we expect will eventually decline as the site matures. In an Australian tropical forest, turnover of non - woodland species occurred in the first 10 15 years while woodland - associa ted species demonstrated a steady accumulation of species richness for at least 23 the first 20 years after revegetation (Gould and Mackey 2015). If our site follows similar patterns , and eventually reaches a species density similar to mature forest as predic ted by the model of Dunn (2004), then species accumulation rather than species turnover will be the primary driver of community changes at the site in the coming decade or two. S ite - scale Species Richness estimates four years post - planting Regardless of t he survey data set or estimation technique, we calculated actual Species Richness considerably higher than observed (Fig. 1. 2), underscoring the difficulty of exhaustively samplin g tropical bird communities and the importance of accounting for undetected s pecies. Combining all observations from formal plot surveys, mist netting, and incidental encounters in 2014 yields a total of 67 species ( Appendix, Table 1. 2). This total is beyo nd the 95% confidence interval limits given by the jackknife estimates but wi thin the 95% credible interval given by the occurrence model. Half of the species not recorded on formal plot surveys were hummingbirds, a group that is difficult to effectively s urvey with area counts because their behavior and size reduce surveyor abilit y to make species - level identifications. Forest birds with only mist net records may have been flying through the site but not using it for foraging or other activities. Even so, this would indicate the site is becoming a functional corridor between nearby forest fragments. Effects of content and context variables on birds Contrary to our hypothesis that plot planting regime would influence the number of bird species recorded, non e of our content variables showed a significant relationship with plot - scale Species Use, suggesting these variables do not strongly differentiate plot attractiveness within a single 24 year. From an avian perspective, differences between plot content may not have been as great as we perceived; most of the species using our site have fairly broad habitat tolerances and may only respond to coarser habitat differences. However, the site - scale multi - year trends of increasing Activity and Species Richness demonstr ate vegetation structure (which developed substantially over the study) does matter. The significance of vegetation structure and unimportance of number of tree species planted are consistent with previous research (MacArthur and MacArthur 1961, Karr 1968) . Many species using our site have home ranges larger than our plots, or even our entire study area, so within - year content effects may only become apparent at a larger scale. Finally, content effects may have been obscured by the overwhelming influence of context effects on Species Use. We caution our results only indicate content variables did not strongly affect Species Use of plots during our study, in an early stage of restoration. Effects of plot planting design may become more important over time as differences in tree growth habits manifest more strongly and trees become rep roductively mature. Inga , for example, will not grow as tall as other species but will produce large amounts of flowers and fruit attractive to birds (Johnson 2000). Like previo us studies, we found restoration site context significantly influenced plot - scale Species Use (Lindenmayer et al. 2010, Reid et al. 2014). At our site, there were more species using plots near other woodland habitat and farther from pasture. Increasing Spe cies Use with distance to pasture could be aversion to pastu re or because plots nearest pasture represented poorly connected habitat not part of efficient foraging routes. We find the second interpretation more likely because we commonly observed substanti al bird activity at forest - pasture interfaces in the Mamoní Valley, provided the forest area was large. Other adjacent land cover types also 25 may have influenced Species Use but could not be included in our model because of small sample size. For example, b irds were rarely observed in a dense fern - covered wetland th at was within 25 m of some of the northernmost plots. The plot with highest Activity and observed Species Use was adjacent to a small fragment of riparian woodland. Riparian woodland was pooled in side the Adjacent woodland category but may have unique char acteristics that make it especially attractive to birds. Occurrence model potential and limitations For tropical communities, where species richness is high and many species are rare, acquiring s ufficient sample sizes for individual species is challenging (Gotelli and Colwell 2001; Herzog et al. 2002). In this context, multi - species occurrence modeling, which can assess both community - and species - level responses to environmental variables, may be particularly useful. The results of our model need to be in terpreted in light of characteristics of our field site and the plot variables we included in the model. The proximity of our plots to each other means that plot - scale Species Use is likely not in dependent as birds move about the site. It is unclear why su rvey date was a significant detection model variable over the brief period when our Single - year survey was conducted (3 weeks). Possibilities include acclimation of birds to observer presence, cha nges in bird space use due to breeding cycle, or changes in local food resources. Implications for bird conservation, ecological function, and forest restoration Most species we recorded are typical of disturbed areas and not commonly found in primary or old secondary forest. Of nearly 300 resident species our res earch group has recorded in the Mamoní Valley, almost 200 occur in intact forest and only a few typical forest species, such as 26 Blue Dacnis ( Dacnis cayana ) and Red - legged Honeycreeper ( Cyanerpes c yaneus ), occurred at our restoration site (S. Roels, unpubli shed data). This lack of overlap is unsurprising given many studies have found bird species richness recovers more quickly than community composition (Dunn 2004, Catterall et al. 2012, Gould and M ackey 2015). Major groups characteristic of wet lowland fore st in central Panama that were absent or nearly absent from our restoration site include antbirds (Thamnophilidae), woodcreepers (Dendrocolaptinae), trogons (Trogonidae), toucans (Ramphastidae), a nd parrots (Psittacidae). These groups correspond to foragin g guilds that are often absent in young restorations and successional areas: understory insectivores (antbirds and woodcreepers) and large frugivores (trogons, toucans, and parrots) (Powell et al. 2015, Rolo et al. 2017). Many species within these groups a re of conservation concern (Powell et al. 2015, Strahl and Grajal 1991). Presence of large frugivores in forest restorations is especially important for ecological function because they disperse l arge - seeded trees that are otherwise unlikely to colonize re storation areas (Wunderle 1997). Restoration of habitat for these groups and return of unique ecosystem services associated with them is a long - term prospect, even when native trees are planted in an effort to accelerate forest recovery. At finer scales, within restoration sites, we should not expect uniform spatial distribution of bird - driven ecological functions. This is true despite early users of restored tropical forest habitat frequently bei ng highly mobile species (e.g., tanager species in our study ). Our study suggests that fine - scale patterns of habitat use, and thus ecological functioning, within a restoration site are influenced by neighborhood - scale features like adjacent habitat and co nnectivity. Relationships between neighborhood - scale habitat features and fine - scale spatial variation in tropical bird ecological function have been documented for seed dispersal and herbivorous insect 27 control (Wenny and Levey 1998, Karp et al. 2013, Maas et al. 2015, Roels et al. 2018). Our study implies that rec overy of bird - driven ecosystem functions may be reduced near pastures, even if vegetative structure is no different than areas farther from pasture that are heavily used by birds. For forest resto ration projects that rely on bird - driven ecosystem functions like seed dispersal, areas near pasture and other matrix land covers unfriendly to birds may warrant additional monitoring and investment to ensure restoration goals are met. A complementary stra tegy would be encouraging silvopastoral techniques in pastur e areas adjacent to forest restorations (Murgueitio et al. 2011). Adding native trees and shrubs to active pasture would encourage bird use of habitat at the edges of forest restorations by creati ng additional foraging habitat and facilitating bird movemen t along restoration - pasture interfaces. ACKNOWLEDGEMENTS We thank P. Ostrom, G. Roloff, S. Bodbyl, the Lindell Lab Group, and two anonymous reviewers for valuable comments on the manuscript. S. W illiams contributed to the Multi - year survey. A. Dennhardt a nd A. Wright provided essential statistical guidance. E. Zipkin provided R code for the multi - species occurrence model in Zipkin et al. (2010). Funding provided by rtment of Integrative Biology, Ecology, Evolutionary Biology , and Behavior, Caribbean and Latin American Studies Program, and School of Graduate Studies; and Daystar Research. Access and support in the Mamoní Valley provided by the Earth Train Foundation. 28 APPENDIX 29 APPENDIX Table 1. 1. Tree species planted by experimental plot. Sixteen experimental plots (50 m x 50 m) were planted with 1, 2, 4, or 8 native tree species. Seedlings of Minquartia guianensis died and were replaced by other spec ies, primarily Hyeronima alchorneoides . 30 Multi - year survey Single - year survey Mist net Guild (Common Name) Latin Name 2010 2011 2012 2013 2014 2014 2014 Open Country Ruddy Ground - Dove Columbina talpacoti - - - - - - 1 Black - throated Mango Anthracothorax nigricollis - 1 - - - - 5 American Kestrel Falco sparverius - - - - - 1 - Tropical Kingbird Tyrannus melancholicus - - 4 6 7 6 - Fork - tailed Flycatcher* Tyrannus savana - - - - - - - Tropical Mockingbird Mimus gilvus 3 - 2 1 1 2 1 Eastern Meadowlark Sturnella magna 62 13 - - - - - Shiny Cowbird Molothrus bonariensis - - - - 3 3 - Yellow - faced Grassquit* Tiaris olivaceus - - 1 - - - - Yellow - bellied Seedeater Sporophila nigricollis - - - - - - 2 Open Country sub - total 65 14 7 7 11 12 9 Brushy Smooth - billed Ani Crotophaga ani 4 1 - 8 18 23 5 Pale - breasted Spinetail Synallaxis albescens - - 5 13 5 9 4 Cantorchilus sp. 1 Cantorchilus sp. - - 2 5 2 8 2 Blue - black Grassquit Volatinia jacarina - - - 7 - 12 4 Thick - billed Seedfinch Sporophila funerea - - - 8 13 42 22 Variable Seedeater Sporophila corvina - - 3 18 25 39 10 Seedeater/Seedfinch 2 Sporophila sp. - - 1 - 1 16 - Brushy sub - total 4 1 11 59 64 149 47 Early Secondary White - tipped Dove Leptotila verreauxi - - - 1 1 22 1 Striped Cuckoo Tapera naevia - - - 2 2 5 1 Common Pauraque Nyctidromus albicollis - - - - - 1 - Long - billed Starthroat Heliomaster longirostris - - - - - - 2 Garden Emerald Chlorostilbon assim ilis - - - 3 6 2 17 Scaly - breasted Hummingbird Phaeochroa cuvierii - - - - - - 7 Snowy - bellied Hummingbird Amazilia edward - 1 - - - 1 15 Rufous - tailed Hummingbird Amazilia tzacatl - - - 4 1 13 26 Striped Owl Asio clamator - - - - - 1 - Barred Antshri ke Thamnophilus doliatus - - - 1 - 2 1 Southern Beardless Tyrannulet Camptostoma obsoletum - - - - - - 2 Yellow Tyrannulet Capsiempis flaveola - - - 3 11 33 6 Yellow - bellied Elaenia Elaenia flavogaster - - 3 2 7 20 13 Lesser Elaenia Elaenia chiriquensi s - - 3 9 4 5 20 Elaenia sp. 3 Elaenia sp. - - - - 2 1 - Bran - colored Flycatcher Myiophobus fasciatus - - - - - 1 - Black - striped Sparrow Arremenops conirostris - - 4 9 10 12 1 Palm Tanager Thraupis palmarum - - - - - - 2 White - lined Tanager Tachyphon us rufus - - - 6 16 22 7 Flame - rumped Tanager Ramphocelus flammigerus - - 2 7 5 6 - Crimson - backed Tanager Ramphocelus dimidiatus - - 6 17 34 49 10 Early Secondary sub - total 0 1 18 64 99 196 131 Table 1. 2 . Results from five years of bird surveys at a forest restoration in the Mamoní Valley, Panama . Counts should not be directly compared across survey types because of different methodologies (see text) and sampling effort. The Single - year survey was approximately double the effort (total observation tim e) of the annual effort in the Multi - year survey. 31 Table 1. 2 Multi - year survey Single - year survey Mist net Guild (Species) 2010 2011 2012 2013 2014 2014 2014 Forest Edge Scaled Pigeon Patagioenas speciosa - - - 1 2 5 - White - necked Jacobin Florisuga mellivora - - - - - - 2 Rufous - breasted Hermit Glaucis hirsutus - - - - - - 2 Long - billed Hermit* Phaethornis longirostris - - - - - - - Stripe - throated Hermit Phaethornis striigularis - - - - - - 2 Violet - headed Humming bird Klais guimeti - - - - - - 1 Rufous - crested Coquette Lophornis delattrei - - - - - - 1 Crowned Woodnymph Thalurania colombica - - - - - - 1 Olivaceous Piculet Picumnus olivaceus - - - - - - 2 Red - crowned Woodpecker Melanerpes rubricapillus - - - - - 5 - Lineated Woodpecker Dryocopus lineatus - - - - 1 - 1 Yellow - crowned Tyrannulet Tyrannulus elatus - - - - 10 10 9 Ochre - bellied Flycatcher Mionectes oleagineus - - 1 - - - - Paltry Tyrannulet Zimmerius vilissimus - - - - - - 1 Common Tody - Flycatc her Todirostrum cinereum - - - 10 7 32 8 Great Kiskadee Pitangus sulphuratus - - - - 1 - - Panama Flycatcher Myiarchus panamensis - - - - 1 - 1 Rusty - margined Flycatcher Myiozetetes cayanensis - - - - - 2 - Lesser Greenlet Pachysylvia decurtata - - - - - - 1 Yellow - green Vireo Vireo flavoviridis - - - - 1 3 7 Tropical Gnatcatcher Polioptila plumbea - - - - - 1 - Clay - colored Thrush Turdus grayi - - - 1 5 2 7 Thick - billed Euphonia Euphonia laniirostris - - - - 1 - - Blue - gray Tanager Thraupis episco pus - - 8 9 8 15 9 Golden - hooded Tanager Tangara larvata - - - - 21 - - Plain - colored Tanager Tangara inornata - - 3 3 1 6 - Red - legged Honeycreeper Cyanerpes cyaneus - - - - - 2 - Blue Dacnis Dacnis cayana - - - - - 2 - Bananaquit Coereba flaveola - - - - - 1 2 Buff - throated Saltator Saltator maximus - - - 2 1 4 2 Streaked Saltator Saltator striatipectus - - 1 3 1 8 4 Forest Edge sub - total 0 0 13 29 61 98 63 Forest Interior Black - crowned Antshrike Thamnophilus atrinu cha - - - - - - 1 Olivaceous Flatbill Rhynchocyclus olivaceus - - - - - - 1 Forest Interior sub - total 0 0 0 0 0 0 2 No Habitat Guild Hummingbird sp. - - 4 12 1 20 - Unidentified bird 2 - 16 30 14 50 - Total Bird Activity 71 16 69 201 250 525 252 * Not recorded during formal surveys or mist netting in 2014 but observed during other field work 1 - Isthmian Wren ( Cantorchilus elutus )/Buff - breasted Wren ( Cantorchilus leucotis ) 2 - Variable Seedeater/Thick - billed Seedfinch 3 - Lesser Elaenia/Yellow - bellied Elaenia 32 Code Species Code Species Code Species AMKE American Kestrel FRTA Flame - rumped Tanager STOW Striped Owl BAAN Barred Antshrike GAEM Garden Emerald STSA Striped Saltator BANA Bananaquit LEEL Lesser Elaenia TBSF Thick - billed Seedfinch BBGR Blue - black Grassquit PBSP Pale - breasted Spinetail TRGN Tropical Gnatcatcher BCFL Bran - colored Flycatcher PCTA Plain - colored Tanager TRKI Tropical Kingbird BGTA Blue - gray Tana ger RCWO Red - crowned Woodpecker TRMO Tropical Mockingbird BLDA Blue Dacnis RLHO Red - legged Honeycreeper VASE Variable Seedeater BSSP Black - striped Sparrow RMFL Rusty - margined Flycatcher WLTA White - lined Tanager BTSA Buff - throated Saltator RTHU Rufous - ta iled Hummingbird WTDO White - tipped Dove CANT Cantorchilus wren sp. SBAN Smooth - billed Ani YBEL Yellow - bellied Elaenia CBTA Crimson - backed Tanager SBHU Snowy - bellied Hummingbird YCTY Yellow - crowned Tyrannulet CCTH Clay - colored Thrush SCPI Scaled Pigeon Y ETY Yellow Tyrannulet COPA Common Paraque SHCO Shiny Cowbird YGVI Yellow - green Vireo COTF Common Tody - Flycatcher STCU Striped Cuckoo Figure 1.6 . Posterior probabilities for effects of plot variables for individual species, sorted by habitat gu ild. The community - level estimates (hyper - parameters) are in white, Open Country species in tan, Brushy species in yellow, Early Secondary in ligh t green, and Forest Edge in dark green. Species abbreviations in alphabetical order are in the first panel. 33 Figure 34 LITERATURE CITED 35 LITERATURE CITED regeneration in a chronosequence of tropical abandoned pastures: implications for restoratio n ecology. Restoration Ecology 8:328 338. http://doi.org/10.1046/j.1526 - 100x.20 00.80048.x Angehr, G., and R. Dean. 2010. The birds of Panama. 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Harrison Road, East Lansing, Michigan, USA 48823 43 ABSTRACT Terrestrial t rophic cascades are widespread but their interactions with other dynamic ecological processes, including seasonal variation in precipitation, migration of predators, and irregular global climate cycles, have uncertain effects on trophic cascade strength . In tropical ecosystems, bird - and ba t - driven cascades influence ecological function, agricultural production, and ecosystem restoration. E cological theory predicts that top - down forces on food webs increase with higher ecosystem productivity. In Neotropical regions with highly seasonal rainf all, productivity peaks in the wet season, suggesting cascades will be most evident then. However, insectivorous birds, key contributors to tropical cas cades, are often most abundant during the dry season when migrants are overwintering. This misalignment between peaks in plant productivity and predator abundance generates competing predictions regarding when cascades should be strongest in tropical syste ms. We conducted bird and bat exclosure experiments on two tree species in a plantation - style forest res toration in Panama. We measured leaf production, herbivory rates, and abundance of arthropods and insect - eating birds during a typical wet season, a dry season, and an aberrant wet season featuring a severe El Ni ñ o - induced drought. Leaf production rates we re lower during the dry season than either the typical or aberrant wet season, which were similar. H erbivory and folivorous insect biomass were higher d uring a typical wet season than in the dry season or aberrant wet season. Insect - eating bird abundance w as highest during the dry season, when overwintering migrants augmented the resident bird population. During the typical wet season, exclosures impeding bird and bat foraging increased folivorous insect biomass on one of the two focal tree species, causing 38.0 58.2% more herbivory. No exclosure effects were detectable during the dry season or the aberrant wet season. Our results from the typical wet and dry seasons are consistent with the hypothesis that trophic cascade 44 strength increases with ecosystem pr oductivity. However, the El Ni ñ o - induced drought apparently disrupted the cascade in the aberrant wet season by uncoupling the correlation between plant productivity and herbivorous insect abundance; productivity was similar to the typical wet season but i nsect abundance was more lik e the dry season. Predicted climate change in the Neotropics has comparable effects to El Niño events, so the shift in trophic relationships we observed may preview consequences of planetary warming. INTRODUCTION Bird - and bat - driven trophic cascades , whe re predators indirectly benefit plants through reductions in herbivorous insects, are now recognized to be widespread in a variety of tropical ecosystems (Van Bael et al. 2008 ; Mäntylä et al. 2011; Maas et al. 2016 ). However, we have limited understanding of how trophic cascade strength varies within dynamic ecological contexts , such as intra - annual precipitation cycles, multi - year global climate cycles, o r long - distance migration of predators ( Pace et al. 1999; Meserve et al. 20 03; Van Bael et al. 2008). I nsights into mechanisms underlying variations in trophic cascade strength have implications for agricultural production (Van Bael et al. 2008; Maas et al. 2016) and ecosystem functioning and restoration (Holmgren et al. 2001; Mo rrison and Lindell 2012). T heoretical ecologists suggest that top - down forces in food webs should be strongest where plant productivity is high because high productivity ecosystems feature sufficiently large herbivore populations to support stable carnivo re populations (Oksanen et a l. 1981). Consequently, population sizes of herbivores in high productivity systems will be regulated by predators, instead of by limited plant resources as in low productivity systems (Dyer and Coley 2008). 45 Maintenance of herbi vore populations at levels l ower than could be supported by available plant productivity results in lower herbivory in the presence of predators. In this manner, top - down predators indirectly reduce herbivory via cascading effects through the food web. Se veral studies of bird - driven trophic cascades in both tropical and temperate ecosystems support the idea that top - down forces are more evident with higher productivity (Van Bael et al. 2003; Van Bael and Brawn 2005; Bridgeland et al. 2010; Garibaldi et al. 2010). Plant productivity i n tropic al regions often features significant seasonal variation , with major leaf flushing events typically occurring just before or at the onset of wet seasons (van Schaik et al. 1993). Wet and dry seasons also cause arthropod abundance s to vary, with abu ndance generally higher during wet season s (Robinson and Robinson 1970; Buskirk and Buskirk 1976; Wolda 1978). Similarly, insec tivorous bird and bat breeding activity in many tropical locations coincides with wet seasons ( Flemin g et al. 1972; Karr 1976; Po ulin et al. 1992) ; the increased metabolic demand of insectivore breeding activity may increase predation pressure on herbivorous insects . These seasonal patterns at three trophic levels (producer, herbivore, predator) may stren gthen trophic cascades durin g wet seasons. However, d uring the dry season many tropical regions receive an influx of insect ivorous migratory birds escaping temperate winters, potentially increasing strength of top - down forces. In tropical Central America, abundance of insectivorous birds in some land cover types , such as shade coffee plantations, more than doubles with the arrival of overwintering migrants (Greenberg et al. 2000). Species richness of wintering m igrato ry bird s is positively associated with reductions in insect abundan ce in tropical agroforests (Van Bael et al. 2008) , and wintering 46 migratory birds have been credited with economically significant crop protection (Kellermann et al. 2008; Johnson et al. 2010) . Migratory bird - friendly vegetation management is recommended to maximize this ecosystem service provided by wintering insectivorous birds (e.g. Van Bael et al. 2007a,b; Kellermann et al. 2008; Karp et al. 2013). Similarly, tropical forest restoration goals may be advanced through encourage ment of foraging activity by insectivorous birds and bats. The science and practice of forest restoration in tropical regions has progressed rapidly in recent decades but more research is needed on ecosystem functions of animals in restorations (Lindell 20 08; Fras er et al. 2015; McAlpine et al. 2016). In Central America, resident bird species can rapidly recolonize forest restorations (Roels et al. unpublished manuscript) and wintering m igra tory birds also readily use habitat created by forest restoration a ctivitie s (Lindell et al. 2012) . Morrison and Lindell ( 2012) found birds and bats reduced insect damage to trees in a Costa Rican forest restoration but strength of the effect varied by tree planting strategy . Their study was conducted during the wet seaso n and no comparable study has yet investigated trophic cascades in forest restorations over multiple seasons. A seasonal comparison of bird - driven trophic cascade s in a mature Panamanian forest found s ignificant cascade effects during the wet season but not the dry season (Van Bael and Brawn 2005). Whether this pattern would hold in a forest restoration is uncertain. Although plantation - style forest restorations in Central America often include tree species representative of mature forests, they have vege tation s tructure and dry season bird communities more similar to th ose of agroforests . Canopy height and vegetation density in young plantation restorations and agroforests is comparable. In both land cover types, overwintering migrants are a substantial 47 p roportio n of birds present in the dry season (S. Roels, personal observations) . In contrast, in the mature forest where Van Bael and Brawn (2005) did not find a dry season trophic cascade, overwintering migra nts comprised only 3% of foliage - gleaning birds in the d ry season (Van Bael 2003) . Thus, strength of top - down forces during the dry season in young forest restorations may be greater than those observed in mature forests. Interpretation of ecological experiments is often challenged by interannual varia bility. We began our study during the 2015 wet season, during what became one of the strongest El Niño events ever recorded (Paek et al. 2017). El Niño events in the Neotropics are in some ways analogous to current climate change patterns (Meehl and Washin gton 199 6) and may be considered a preview of biological responses in the tropics to ongoing planetary warming (Condit 1998). Climate records for central Panama, where our study was conducted, indicate a long - term drying trend and also show that strong El Niño eve nts are associated with major droughts, as was the case during the 2015 wet season (Condit 1998; Condit et al. 2004). We responded to the aberrant wet s eason, w hen El Niño - Southern Oscillation (ENSO) conditions had returned to neutral and wet season rainfall was more typical (NOAA 2018; STRI 2018). Prior studies have demonstrated how ENSO - caused shifts in precipitation and temperature regimes can dramat ically a lter trophic interactions via both bottom - up and top - down effects (Holmgren et al. 2001). For example, in a semi - arid ecosystem in Chile, low precipitation normally limits the strength of trophic interactions, as primary productivity is insufficien t to sup port large populations at higher trophic levels. However, El Niño years bring above - average 48 rainfall, boosting plant productivity and also intensifying top - down forces as populations of herbivorous small mammals and their predators increase dramati cally in response to the pulse of resources (Meserve et al. 2003). In a tropical forest in Panama, the return of rainfall following an El Niño - induced drought caused an outbreak of folivorous insects and a more than three - fold increase in herbivory (Van Ba el et al . 2004). If trophic cascade strength is regulated by productivity, itself regulated by precipitation, cascades should become stronger in locations where El Niño events increase rainfall and weaker where El Niño events cause droughts. Studies to da te in tr opical ecosystems have produced inconsistent results regarding seasonal differences in arthropod abundance and the relationship between trophic cascade strength and ecosystem productivity (Maas et al. 2016). Despite the general presumption that art hropods are more abundant in the wet season, some studies in coffee plantations have found equal (Jedlicka et al. 2006) or higher (Williams - Guillén et al. 2008) arthropod abundance during the dry season. These observations may be related to coffee phenolog y, which can influence arthropod population cycles (Karp et al. 2013). Although numerous studies comparing two contexts have demonstrated stronger trophic cascades in the higher productivity context ( Van Bael et al. 2003; Van Bael and Brawn 2005; Bridgelan d et al. 2010; Garibaldi et al. 2010) , some found either no productivity - based differences (Philpott et al. 2009; Mooney et al. 2010) or stronger cascades in what was presumed to be the lower productivity context (Greenberg and Ortiz 1994; Williams - Guillén et al. 2008). We conducted a predator exclosure experiment on two native tree species during a typical wet season (2016), a dry season (2016), and a drought (2015) in a tropical forest restoration in central 49 Panama. Following other studies, we expected h igher insect bio mass associated with increased rainfall and plant productivity during the wet season would amplify top - down forces of birds and bats relative to the dry season. We anticipated a greater absolute decrease in leaf damage in the presence of bi rds and bat s dur ing the wet season than the dry season . However, we also predicted that birds and bats would cause a greater proportional reduction in insect biomass and leaf damage during the dry season, due to the addition of wintering migratory birds to the insectivore community (Fig. 2. 1) . We expected l ower biomass of i nsects during the dry season w ould result in less total leaf damage, such that the absolute change in leaf damage between exclosure and control treatments w ould be less than during the we t season. Follow ing the drought in 2015, we also predicted that trophic interactions during a wet season with an El Niño - induced drought would be more similar to a dry season than to a typical wet season because of reduced plant productivity during the dro ught. Figure 2.1 . Illustration of trophic cascade predictions. We predicted the absolute decrease in leaf damage associated with bird and bat predation would be greater in the wet season than the dry season (illustrated difference of 1.0% damage in wet s eason an d 0.5% damage in dry season). However, we also predicted the proportional decrease would be greater in the dry season (damaged reduced by one - half in the dry season and one - third in the wet season). 50 METHODS Study s ite and species We conducted our study a t 300 m a.s.l. , Fig. 2.2 ). The ASP contains a variety of experimental land covers , including a plantation - style native species forest restoration. The p lantation was established in 2008 and contains replicate plots with native species monocultures or polycultures (Mayoral et al. 2017) . Figure 2.2 . Location of the Agua Salud Project (star), N, 7 Non - plante d understory vegetation is cleared approximately quarterly. We selected two plantation tree species for our experiments that are photosynthetically active year - round in central Panama (so we were able to conduct experiments in both the wet and dry seasons) . The first species, Ochroma pyramidale (Cav. ex Lam.) Urb. (Family: Malvaceae ), is a short - lived pioneer species common in recently disturbed areas (Condit et al. 2011). There is a commercial 51 market for the timber of this species, with large plantations i n the Neotropics (Stilwell et al. 2014). The second, Terminalia amazonia (J.F. Gmel.) Exell (Family: Combretaceae ), is common in secondary forests and regularly planted in native species plantations in Central America due its reliable performance and qualit y timber (Condit et al. 2011; Carpenter et al. 2004). Field s easons Central P anama has pronounced seasonality with a wet season from approximately May December and a dry season from approximately January April. We conducted our experiments during three fi eld seasons, corresponding to portions of the 2015 drought (9 July 17 August), 2016 dry season (24 January 4 March), and 2016 wet season (11 June 27 July). In 2015, June August rainfall recorded at the ASP meteorological station was 426 mm ; typical June Au gust rainfall at this station is 818 mm (average of 2009 2017, STRI 2018). In the 2016 wet season, ENSO conditions had returned to neutral and rainfall at the ASP station was 743 mm (STRI 201 8 ) . In addition to lower rainfall, the June August period in 2015 featured higher mean temperatures and wind speeds and lower relative humidity than the same period in 2016 (STRI 2018). A longer term data set from the weather station on nearby Barro Colorado Island provides greater context for severity of the drought; y ear - to - date rainfall through August 2015 ranked as the lowest ever recorded in 90 years (Steve Paton, personal communication). Notably, the ASP received no rainfall during a two - week period in late June 2015, just prior to the start of our experiment. In 2 016, January March rainfall at the ASP was 32.3 mm; typical January March rain fall is 103 mm (average of 2009 2018, STRI 2018). 52 Experimental protocol Exclosure material and design We constructed predator e xclosure s out of light - weight plastic netting (St andard BirdNet ; Bird - X, Chicago, IL, USA), with mesh large enough (2 x 2 cm) to permit passage of large arthropods but not foraging birds or bats. Netting was cut into 80 x 160 cm pieces that were folded in half around branches, creating bubble - shaped excl osures with ma ximum dimensions approximately 70 x 70 x 35 cm. Exclosures were closed with either twist ties ( O. pyramidale ) or binder clips ( T. amazonia ) but a small gap (~10 cm) was left around the proximal end of the branch, allowing passage of large art hropods crawli ng along the branch toward the tip. The terminal end (~ 50 cm length) of a single branch was placed in each exclosure (Fig. 2.3 ). Figure 2. 3. Exclosures on O. pyramidale (left) and T. amazonia (right). Branches of T. amazonia were assigne d to one of three treatments: control, 24 - hr exclosures, and daytime exclosures. The daytime exclosure (with no exclosure overnight) was intended to separate effects of diurnally foraging birds from nocturnally foraging bats. For exclosure treatmen ts on T. amazonia , netting was stretched over a frame of wooden dowels that was supported by a plastic pole staked into the ground (Fig. 2.3 ). In this manner, the netting formed a 53 bubble that limited contact with the branch contained inside (Fig. 2.3 ). For 24 - hr e xclosures, this netting was left in place for the duration of the experiment in each season. For daytime exclosures, netting was removed at dusk and replaced at dawn each day of the experiment. The dowel frame facilitated net opening and closing by reducin g tangling of netting in branch tips. The daytime exclosure treatment was dropped from the experimental design during the 2016 wet season because of time constraints. Control branches had dummy frames with a pole and dowels but no netting. Branches of O. pyramidale were assigned to one of two treatments: control or 24 - hr exclosure. Logistical constraints prevented a daytime exclosure treatment for O. pyramidale . For the 24 - hr exclosure treatment on O. pyramidale , netting was placed directly on branches (Fi g. 2.3 ). We did not use dowel frames with support poles for exclosures on O. pyramidale because most branches were too far above the ground for poles and branch architecture did not allow for attaching frames di rectly onto branches. However, branch archite cture naturally limited contact between the net and the actively flushing leaves at the branch tip. Control branches were left unaltered. For each tree species, we created pairs or groups of neighboring trees w ithin plantation plots and randomly assigned treatments within pairs or groups. The majority of trees selected during the first field season were used in subsequent field seasons although some replacements were necessary due to tree death and light competi tion from neighboring trees. For O. pyramidal e there were 38, 37, and 36 pairs of trees in the 2015 drought, 2016 dry season, and 2016 wet season, respectively. For T. amazonia there were 40, 38, and 41 groups in the 2015 drought, 2016 dry season, and 2016 wet season, respectively. On each tree, we s elected a single branch 54 that showed evidence of active growth (newly flushed leaves). Branches ranged in height above the ground from 1.0 4.0 m for O. pyramidale and 0.5 2.0 m for T. amazonia . Exclosures remaine d in place on O. pyramidale for 28 32 days in the 2015 drought, 31 32 days in the 2016 dry season, and 30 31 days in the 2016 wet season. Exclosures remained in place on T . amazonia for 34 35 days in all field seasons. Arthropod abundance surveys A single observer, S. Roels, surveyed arthropod abund ance just prior to the placement of exclosures and again at the end of the experiment. During each survey, S. Roels visually searched for arthropods on all leaves of focal branches , inspecting both upper and lower leaf surfaces . All arthropods were identif ied to order and categorize d into the following size classes : 4 mm, >4 6 mm, >6 mm . Arthropods greater than 6 mm were measured with a hand ruler. For O. pyramidale , we recorded the length of ea ch leaf searched and converted leaf length to leaf ar ea using season - specific leaf length:area power functions fit to leaf area data produced with ImageJ (see below) . For T . amazonia , measurements of every leaf were not feasible (often over 100 leaves per branch). We timed arthropod surveys, using time neede d to search the entire branch as a proxy for leaf area. To convert search time into estimated leaf area, we conducted time trials where we surveyed test branches, harvested all the leaves, estimated tota l leaf area with ImageJ, and calculated the mean leaf area surveyed per second. The duration of each experimental branch survey was then multiplied by this rate to give a leaf area estimate. We converted arthropod lengths into total dry arthropod biomass per branch using previously published length:mass reg ression relationships specific to taxonomic order (Schoener 1980; 55 Sample et al. 1993; Johnson and Strong 2000; Wardhaugh 2013; Appendix, Table 2. 3 ). Because length:mass relationships vary by latitude and climate, we used equations from Schoener (1980) if t hey were available for an order. Schoener (1980) produced separate equations for dry and wet forests in Costa Rica, the location with published allometric equations nearest our study site. Our study site receives annual rainfall intermediate to the two for est types in Schoener (1980) so we and wet forest equations. For larval Lepidoptera, we took the average o f biomass estimates produced from equations in Sample et al. (1993) and Wardhaugh (2013). For both tree species, we report a rthropod biomass per m 2 of leaf area. The biomass of arthropods placed into body mm for bio mass calculations). Arthropods in the < 2 mm class were treated as having a body length of 1.5 mm because arthropods smaller than 1 mm were unlikely to be counted on visual surveys. We conducted analyses using only orders with folivorous (i.e. chewing) spe cies that could be responsible for removal of leaf area ( Blattodea, Coleoptera, Lepidoptera, O rthoptera, and Phasmida). Productivity and h erbivory measurements A ll leaves flushed during the experimental period were harvested and either digitally photograp hed on a white background ( O. pyramidale ) or scanned with a flatbed scanner ( T. amazonia ). We distinguished new leaves from pre - existing leaves by marking leaves near the branch tip with a permanent marker at the start of the experiment. Leaves of O. pyram idale still in the leaf expansion phase (youngest of all new leaves) were distinguishable from fully flushed leaves by color and texture. We recorded leaf age (still expanding or fully flushed) for O. 56 pyramidale during the photography phase. Leaf age disti nctions were not made for leaves of T. a mazonia as differences in leaf age were less clear in this species. We calculated branch productivity and herbivory by processing leaf photos in ImageJ (version 1.50; Schneider et al. 2012) to measur e total leaf area produced and leaf area removed. Insect - eating bird surveys To measure relative bird abundance during each field season, S. Roels conduct ed fixed - radius point count s through out the native tree plantation . Fifteen points were arranged to maximize coverage of plots containing experimental trees while mi nimizing overlap between points. There was a minimum distance of 85 m between points (mean distance - to - nearest - neighboring point = 121 m, max = 176 m) . Each point was surveyed four times while the exclosure ex periments were active , with counts at a given point separated by at least 3 days. Count duration was 10 min and all birds detected by sight and sound within 25 m were recorded. We pruned the data by removing birds that flew overhead or through the count ar ea without stopping an d birds that do not eat insects. To characterize bird species diets, we reviewed published diet descriptions ( Ridgely 1981 ; Stiles and Skutch 1989 ). Surveys occurred during the following times: 2015 drought, 0800 1030 hrs; 2016 dry se ason, 0800 1100 hrs; 2 016 wet season, 0630 0900 hrs. Earlier survey times were possible in the 2016 wet season because we were no longer managing a daytime exclosure treatment for T. amazonia . We tested for effects of these different time periods on result s. 57 Statistical anal yses Plant productivity All statistical analyses were conducted in R (version 3.4 .4; R Development Core Team, 201 8 ). We used = 0.05 for all hypothesis testing. We calculated productivity on a per branch basis using the summed area of all leaves flushed by the branch during the experiment. Branch productivity was approximately normally distributed so we used a standard ANOVA t o test for differences in mean branch productivity between seasons. H erbivory models We evaluated effects of our exclosure treatments on herbivory using mixed - effects regression models with a logit link created using the glmmTMB package (Brooks et al. 201 7). Our herbivory data best fit a beta distribution, a continuous probability distribution appropriate for overdispersed pr oportion data where 0 < y < 1 . To include leaves with no herbivory (a zero value), we used the transformation recommended by Smithson and Verkuilen (2006): y y ( N 1) + 0.5 ]/ N where N is the sample size. For O. pyramidale , each beta regression model contained treatment, season, and their interaction as fixed effects and experimental branch within season as a random effect. During the exploratory phase of data analysis, we noticed that herbivory on still - expanding O. pyrami dale leaves was substantially lower than on fully flushed leaves. This is not surprising given that still - expanding leaves would be younger than fully flushed lea ves and, thus, exposed to herbivores for less time. We collected few still - expanding leaves in the 2015 drought, possibly because trees were becoming water stressed near the end of the experiment. Because of differences in 58 herbivory between still - expanding and fully flushed leaves and variabl e ratios of still - expanding:fully flushed leaves across seasons, we decided to model fully flushed and still - expanding leaves separately. The 2015 drought is not included in the model for still - expanding leaves. For T. amazonia , each beta regression model contained treatment as a fixed effect and experimental branch as a random effect. W e were unable to collect sufficient herbivory data for T. amazonia during the 2016 dry season due to lack of leaf growth. It is likely that the 2015 drought compounded the normal water stress trees experience during the dry season and caused T. amazonia to stop flushing leaves in the 2016 dry season. During pilot studies the previous year, T. amazonia was flushing leaves during the dry se ason, so dry season herbivory measure ments would be possible in a more typical year. We analyzed T. amazonia data from the 2015 drought and 2016 wet season in separate models because they had different treatment groups (the daytime exclosure treatment was dropped in the 2016 wet season). Analysis of model residuals for O. pyramidale revealed some extreme data points associated with exceptionally high herbivory; these values represent less than two percent of leaves in either O. pyramidale data set. There w as little difference between model results with and without these apparent outliers. Thus, we present models without the extreme values. Some T. amazonia leaves had similarly high herbivory but model residuals were more uniformly distributed than for O. py ramidale so no leaves were removed from the T. amazonia data set. To translate herbivory model coefficients into biologically interpretable values, we calculated the antilogit of model coefficients. 59 Folivorous insect biomass models To evaluate the effect o f our ex closure treatments on folivorous insect biomass, we created mixed - effect models for each tree species and season with treatment and survey period (pre - and post - experiment) as fixed effects and experimental branch as a random effect. The interactio n term i n these models characterizes differences in responses of experimental groups to treatment. The model reference group was control branches during the pre - experiment survey. Since our insect biomass data were derived from count - based surveys, models assumed a negative binomial distribution with a log link. Data were transformed to integers by rounding to nearest whole number to meet assumptions of a negative binomial distribution. To estimate seasonal differences in folivorous insect biomass, we poole d data f rom all treatments within seasons and created mixed - effect models with a fixed effect of season and random effect of experimental branch within season. For T. amazonia , the daytime exclosure treatment data was not used for this analysis because it did not occur in all seasons. Bird abundance models The average survey time of day was different between field seasons. We created linear regression models for each field season to estimate time of day effects on bird counts at survey points. While this e ffect would best be modeled over an entire day with a non - linear function, we believe a linear approximation is sufficient when modeling within our three - hour survey windows and avoiding extrapolation beyond the limits of the data. Models were used to pred ict the number of birds on a hypothetical count occur ring at 0800 hrs, a time of day within the range of survey data for each field season. 60 RESULTS Leaf production Leaf production by both tree species was significantly higher during the 2016 wet season th an the 2016 dry season (Fig. 2 .4 ) . Total leaf area fl ushed by O. pyramidale during the 2016 wet season was over double the leaf area flushed during the 2016 dry season. Total leaf area flushed by T. amazonia was much greater than area flushed during the 20 16 dry season, when growth was almost zero (Fig. 2 .4 ) . Unexpectedly , leaf production for both tree species during the 2015 drought was not significantly different from the 2016 wet season (Fig. 2 .4 ). Figure 2.4. Leaf productivity by tree species and seas on. Plot depicts mean leaf production per branch with 50% (boxes) and 95% (whiskers) confidence intervals. Data from experimental treatment groups (control and exclosure branches) were pooled within seasons. Terminalia amazonia leaf production during the 2 016 dry season was negligible. Herbivory Relative to the 2016 wet season, herbivory on fully flushed O. pyramidale leaves was significantly lower in both the 2016 dry season and 2015 drought, with estimated reductions in herbivory at 34.9% and 26.1%, resp ect ively (Fig. 2.5 ). During the 2016 wet season, herbivory on 61 fully flushed O. pyramidale leaves was significantly higher (38%) in exclosures that restricted bird and bat access, compared to leaves on control branches (Fig. 2.5 ; Appendix, Table 2 . 4 ). Howev er, this effect was absent during the 2016 dry season and 2015 drought (Fig. 2.5 ). Figure 2.5 . Herbivory on fully flushed leaves of O. pyramidale by season and treatment. Plot depicts herbivory estimates with 50% (boxes) and 95% (whiskers) confi dence in tervals. Sample size (number of leaves) is stated below each group. Letters denote statistically significant groups (p < 0.05). Herbivory on still - expanding O. pyramidale leaves was not significantly different between the 2016 wet and 2016 dry sea sons (Fi g. 2.6 ). As was the case for fully flushed leaves, exclosures significantly increased herbivory on still - expanding leaves during the 2016 wet season, but not during the 2016 dry season (Fig. 2.6 ; Appendix, Table 2. 5 ). Our model estimates that herbivory on still - expanding O. pyramidale was 58.2% higher when bird and bat access was restricted during the 2016 wet season. 62 Figure 2.6 . Herbivory on still - expanding leaves of O. pyramidale by season and treatment. Plot de picts herbivory estimates with 50% (boxes) and 95% (whiskers) confidence intervals. Sample size (number of leaves) is stated below each group. Letters denote statistically significant groups (p < 0.05). Relative to the 2016 wet season, herbivory on T. amazonia was significantly lower in the 2015 drought, with an estimated reduction of approximately 56% (Fig. 2.7 ). Herbivory on T. amazonia during the 2015 drought was not significantly different between con trol, daytime exclosure, and 24 - hr exclosure treatments (Fig. 2.7 ; Appendix, Table 2. 6 ). Durin g the 2016 wet season, herbivory on T. amazonia leaves was not significantly different between control and 24 - hr exclosure treatments (Fig. 2.7 ; Appendix, Table 2. 7 ). 63 Figure 2.7 . Herbivory on T. amazonia by season and treatment. Plot depicts herbivory es timates with 50% (boxes) and 95% (whiskers) con fidence intervals. Sample size (number of leaves) is stated below each group. Herbivory estimates for each season were generated from separate models. Letters denote statistically significant groups (p < 0.05) . Folivorous insect biomass Biomass of folivorous insects on O. pyramidale was not significantly different between seasons (Fig. 2.8 ; Appendix, Table 2. 8 ). On T. amazonia , folivorous insect biomass was significantly lower in both the 2015 drought and 2016 dry season compared to the 2 016 wet season (Fig. 2.9 ; Appendix, Table 2. 9 ). Predator exclosures on O. pyramidale significantly increased folivorous insect biomass in the 2016 wet season but not in the 2016 dry season or 2015 drought (Fig. 2.10 ; Appendix, Tables 2. 10 2. 1 2 ). On T . amazo nia , n either exclosure treatment (24 - hr or daytime exclosure) significantly changed folivorous insect biomass in any season, compared to control treatments ( Appendix, Tables 2. 1 3 2. 1 5 ). 64 Figure 2.8 . Total biomass of folivorous insect orders by season on O. pyramidale . Data from all treatment groups were used. Plot depicts biomass estimates with 50% (boxes) and 95% (whiskers) confidence intervals. Letters denote statistically significant groups (p < 0.05). Figure 2.9. Total b iomass of folivorous insect orders by season on T. amazonia . Data from all treatment groups were used. Plot depicts biomass estimates with 50% (boxes) and 95% (whiskers) confidence intervals. Letters denote statistically significant groups (p < 0.05). 65 F igure 2.10 . Biomass of folivorous insects on O. pyramidale i n the 2016 wet season. Plot depicts biomass estimates with 50% (boxes) and 95% (whiskers) confidence intervals. Letters denote statistically significant groups (p < 0.05). Y - axis on log scale. In sect - eating bird surveys Insect - eating bird abundance in the native tree plantation was low, with only one or two birds recorded at a typical point (Table 2.1 ). Abundance of resident insect - eating birds was stable between seasons; regression models ( Appen dix, Fig. 2.11 ) estimated that a point surveyed at 0800 hrs in all seasons would yield between 1.15 1.26 resident birds. However, overwintering migrants boosted the population of insect - eating birds during the 2016 dry season by an estimated 42.3%, accordi ng to the models adjusting for count time. 66 Season Total Birds Total Insect - eating Birds Predicted (0800 hrs) Individuals Species Individuals Species Individuals Per count 2016 Wet 87 27 71 21 70.2 1.17 2016 Dry 80 26 52 (68) 15 (22) 75.4 (107.4) 1.26 (1.79) 2015 Dro ught 55 20 48 18 69.2 1.15 Table 2.1 . Bird survey results by field season. Fifteen points were surveyed four times per season, yielding 60 point counts. Dry season numbers in p arentheses are results including overwintering migrant birds. Predicted survey results derive from season - specific linear regression models (Appendix, Fig. 2.11). DISCUSSION Our results show that bird - and bat - driven trophic cascades in the tropics a re mediated not only by seasonality but also by irregular climatic fluctuations of the ENSO cycle (see Table 2 . 2 for summary of results). Predator exclosures increased biomass of folivorous insects and herbivory on O. pyramidale during a typical wet seaso n, when leaf production was high, but not during the dry season, when leaf production was low. Table 2. 2 . Qualitative summary of environmental conditions and results for Ochroma pyramidale (O.p.) and Terminalia amazonia (T.a.) across three seasons. Rai nfall values are compared across seasons; rainfall totals in the 2016 wet season and 2016 dry season were near historical averages. Leaf production, herbivory, and folivorous insects (unmanipulated branches) are compared across seasons within t ree species and should not be directly compared across species. Insect - eating bird abundance is compared across seasons but also to a priori knowledge of bird abundance in the broader landscape. 67 This result supports our prediction that the absolute trophi c effect of birds and bats on herbivory is greater during the wet season than the dry season and is also consistent with ecological theory and prior empirical evidence that trophic cascade strength increases with higher ecosystem productivity (Oksanen et a l. 1981; Van Bael and Brawn 2005). We did not find evidence for our second prediction that increased abundance of insectivorous birds during the dry season would lead to a greater proportional trophic effect compared to the wet season. Although we document ed the expec ted increase in insectivorous bird abundance during the dry season, exclosures did not affect folivorous insect biomass or herbivory on either O. pyramidale or T. amazonia during this season. Our third prediction was that trophic interactions in an aberra nt wet season with drought conditions would be more similar to a dry season than to a typical wet season because of similarly low productivity. We found some support for the first part of our prediction. First, herbivory on fully flushed O. pyr amidale and on T. amazonia was significantly lower during the 2015 drought compared to the 2016 wet season. For fully flushed O. pyramidale , herbivory during the 2015 drought was comparable to the 2016 dry season. Second, on T. amazonia , folivorous insect biomass was higher in the 2016 wet season than either the 2016 dry season or 2015 drought, which were not different from each other. Third, the trophic cascade effect on O. pyramidale present during the 2016 wet season was not evident during the 2016 dry s eason or 201 5 drought. However, patterns observed in herbivory and folivorous insect biomass did not closely align with our measure of plant productivity. Leaf production rates for both species during the 2015 drought 68 were not distinguishable from the 20 16 wet seaso n and much higher than during the 2016 dry season. This suggests that low folivorous insect biomass during the 2015 drought was the result of unfavorable abiotic conditions for arthropods (lower precipitation and humidity, higher temperatures a nd winds), r ather than being regulated by plant productivity. Why leaf production did not decline on either tree species during the 2015 drought relative to the 2016 wet season is unclear, given that physiological responses by many tropical trees to El Niñ o mirror the ir dry season responses (Detto et al. 2018). It may be the case that water is not a limiting resource for O. pyramidale and T. amazonia in typical wet seasons and that rainfall during the 2015 drought was still sufficient for these species to a void water s tress. Alternatively, trees could have been depleting water stored in their root systems to compensate for the lack of rainfall and did not show signs of water stress until after July 2015. Differences in amount of herbivory (greater on T. amazonia than O . pyramidale ), arthropod biomass (higher on T. amazonia ) and trophic cascade strength (cascade only evident on 2016 wet season O. pyramidale ) are likely driven by species - specific traits. Leaves of O. pyramidale are widely spaced and rarely touch each othe r, which reduces shelter for herbivorous insects from herbivores insects more exposed to foliage - gleaning predators. In contrast, foliage on T. amazoni a branches is f requently dense with many overlapping leaves that may create more favorable microclimates for arthropods. The dense foliage may also make it more difficult for vertebrate predators to glean arthropods from T. amazonia leaves. In addition, th e life history strategies of our two species are different. Ochroma pyramidale may be more vulnerable to generalist herbivores which, in turn, may be more vulnerable to vertebrate predators (Singer et 69 al. 2014; Bosc et al. 2018), because it is a short - live d pioneer that may not invest heavily in mechanical or chemical leaf defenses (but see Coley 1983). Although T. amazonia had higher herbivory rates than O. pyramidale , it may be less palatable to generalist herbivores because it is a longer - lived species c ommon in well - d eveloped secondary forests and has relatively tough leaves (Paul et al. 2011). We found a demonstrable trophic cascade in one context, despite low predator abundance in the ASP native tree plantation. Insect - eating bird abundance was consid erably less tha n in other land cover types in central Panama, including residential areas, gallery forest, and old - growth forest (Petit et al. 1999; Roels et al. 2018) and a less heavily managed forest restoration of comparable age ( see Chapter 1 of this d issertation). A separate study of bird predation pressure in the broader ASP landscape found that bird predation activity in the native tree plantation was second - lowest of any land cover type surveyed, including non - native teak plantation (the lowest), un managed seconda ry growth, gallery forest, old - growth forest, and countryside hedgerows (Roels et al. 2018). Although we did not collect data on bat abundance, other researchers have documented low bat densities in the ASP native tree plantation relative to nearby forests ( S. Brändel , personal communication). Since exclosures on O. pyramidale were in place for 24 hours a day, we cannot quantify the proportional contributions of birds and bats to the trophic cascade we observed in the 2016 wet season. We suspect birds were the primary drivers of the trophic cascade because bat species that cons ume folivorous insects are unlikely to be present in the ASP native tree plantation in substantial numbers. Phyllostomid bats, a major component of Neotropical bat communities, 70 genera lly avoid areas where understory vegetation has been removed (García - Mora les et al. 2013). Furthermore, within the Phyllostomidae, species that are specifically described as foliage - gleaning insectivores (subfamily Phyllostominae) are especially sensitive to forest disturbance (Medellín et al. 2000). However, another phyllostom id subfamily, the Carolliinae, is positively associated with forest disturbance (Medellín et al. 2000) and some species in this generally frugivorous group may consume more insects th an has traditionally been assumed (York and Billings 2009). Additional st udy of bat ecology in tropical forest restorations is clearly needed. Our results only represent 4 5 weeks of herbivory on newly flushed leaves; measures of standing crop herbivory a re likely to be higher. Leaf life span is approximately 100 days for O. p yramidale (Selaya et al. 2008) and 200 days or more for T. amazonia (Kitajima et al. 2013; K. Sinacore, personal communication). If damage continues to accumulate over leaf life spans , longer experiments may have found greater differences in herbivory betw een more of the season - treatment comparisons. The difference in magnitude of the exclosure effect on still - expanding and fully flushed O. pyramidale leaves in the 2016 wet season also supports the idea that longer experiments may have found more evidence f or trophic cascades. The proportional increase in herbivory caused by exclosures was greater for still - expanding leaves than for fully flushed leaves (58.2% vs. 38.0%), despite less h erbivory overall on the younger leaves. If the increase in folivorous art hropod biomass on exclosure branches developed over time (i.e. not instantaneously), leaves began flushing near the end of the experiment would have experienced the greatest divergenc e in folivorous insect biomass between treatment groups, leading to the g reater exclosure effect we observed on still - expanding leaves. Given these observations, we recommend that exclosure experiments be run as long as logistically feasible. 71 Native speci es timber plantations are becoming increasingly common in tropical regions as human societies seek to re store ecosystem services and reduce logging pressure on dwindling tracts of primary forest. Low abundances of birds and bats at our site, and lack of tr ophic cascade effects on the tree/season combination with the highest level of herbivory (2016 wet season T. amazonia ) suggest a missed opportunity to take advantage of a potentially beneficial tr ophic cascade. Bird - and bat - friendly management of these ne w forests, such as less aggressive management of understory vegetation and placement of artificial nesting or roosting structures (Lindell et al. 2018), may increase benefits provided by insectivo rous wildlife to plantation trees. However, w e still have mu ch to learn about multi - trophic level interactions with global - scale environmental patterns like the ENSO cycle before we can manage or manipulate them with confidence. Shifting global climate may change seasonal patterns in rainfall, primary productivity, and wildlife communities, altering trophic cascades in ways that will be difficult to predict with our currently limited understanding of their operation. A CKNOWLEDGEMENTS We thank P . Ostrom, G. Roloff, T. Getty, S . Bodbyl , and M. B. Hannay for valuable comments on the manuscript. Jade Porter provided essential assistance in the field each season. A. Dennhardt provided critical statistical guidance. Funding provided by the Smithsonian Tropical Re search Institute (STRI); Michigan State University (MSU) Dep artment of Integrative Biology ; MSU Ecology, Evolutionary Biology, and Behavior Program; MSU Caribbean and Latin American Studies Program ; and MSU School of Graduate Studies . We thank STRI and Fed erico Davis for access and logistical support at the Agua Sa lud Project (ASP). The ASP is a STRI 72 Smart Reforestation® site in collaboration with the Panama Canal Authority (ACP), the Ministry of the Environment of Panama, and other partners. The ASP is par t of ForestGEO and the TreeDivNet network. Plantation manage ment is supported by the ACP. Funding for the ASP Competitive Grants for Science and Grand Challenge grant to BiodiversiTREE, the Heising - Simons Foundation, and the Nati onal Science Foundation (NSF grant EAR - 1360391). Meteorological d ata provided by Physical Monitoring Program. 73 APPENDIX 74 APPENDIX Taxon Authority Location a b Habitat Aranae Johnson & Strong 2000 Jamaica 0.1253 2.039 Various Blattod ea Wardhaugh 2013 Australia 0.0187 2.760 Rainforest Coleoptera* Schoener 1980 Costa Rica 0.1170 2.110 Dry forest Schoener 1980 Costa Rica 0.1260 1.910 Wet forest Diptera* Schoener 1980 Costa Rica 0.0740 1.640 Dry forest Schoener 1980 Costa Rica 0.068 0 1.590 Wet forest Hemiptera* Schoener 1980 Costa Rica 0.0350 2.480 Dry forest Schoener 1980 Costa Rica 0.0270 2.280 Wet forest Homoptera* Schoener 1980 Costa Rica 0.0230 2.650 Dry forest Schoener 1980 Costa Rica 0.0300 2.230 Wet forest Hymenoptera (Formicidae)* Schoener 1980 Costa Rica 0.0120 2.720 Dry forest Schoener 1980 Costa Rica 0.0210 2.310 Wet forest Hymenoptera (non - Formicidae)* Schoener 1980 Costa Rica 0.0430 2.070 Dry forest Schoener 1980 Costa Rica 0.0220 2.290 Wet forest Lepid optera (Larval)* Sample et al. 1993 West Virginia 0.0027 2.959 Forest Wardhaugh 2013 Australia 0.0043 2.550 Rainforest Mantodea Wardhaugh 2013 Australia 0.0018 3.010 Rainforest Neuroptera Sample et al. 1993 West Virginia 0.0113 2.570 Forest Orthoptera * Schoener 1980 Costa Rica 0.2220 1.810 Dry forest Schoener 1980 Costa Rica 0.0660 2.100 Wet forest Phasmatodea Wardhaugh 2013 Australia 0.0027 2.310 Rainforest Table 2. 3 . Biometric equations for arthropod length and dry biomass. Arthropod length:mass equations take the form: mass (mg) = a*L^b where a and b are empirically derived constants and L is length in mm. Modern taxonomy regards Homoptera as a clade within Hemiptera. Asterisks denote where the average of two equations was used. Approximate annu al t forest sites are 1800 mm and 5500 mm, respectively. Approximate annual rainfall at the Agua Salud Project is 2700 mm. 75 Estimate Std. Error z value p Intercept - 3.684 0.114 - 32.21 <0.001 * Season(2016 Dry ) - 0.430 0.183 - 2.35 0.019* Season(2015 Drought ) - 0.302 0.154 - 1.96 0.0 50 * Treatment( 24 - hr Exclosur e) 0.322 0.152 2.12 0.034* Treatment( 24 - hr Exclosure)* Season(2016 Dry ) - 0.325 0.249 - 1.31 0.19 2 Treatment( 24 - hr Exclosure)* Season(2015 Drough t) - 0.535 0.216 - 2.48 0.013* Table 2. 4 . Effects of predator exclosures and seasonality on herbivory for fully flushed O. pyramidale leaves. Leaves in the control group during the 2016 wet season are the reference group (intercept). Statist ically significance differences (p< 0.05) denoted by *. Estimate Std. Error z value p Intercept - 4.167 0.1 45 - 28.756 <0.001 * Season(Dry 2016) - 0. 188 0.1 87 - 1.008 0. 313 Treatment( 24 - hr Exclosure) 0. 459 0. 180 2.546 0.0 11 * Treatment( 24 - hr Exclosure)* Season(Dry 2016) - 0.248 0.266 - 0.93 4 0.351 Table 2. 5 . Effects of predator exclosures and seasonality on herbivory for still - expanding O. pyramidale leaves. Leaves in the control group during the 2016 wet season are the reference group (intercept). Statistically significa nce differences (p<0.05) denoted by *. Estimate Std. Error z value p Intercept - 3.713 0. 046 - 81.45 <0.001 * T reatment( D aytime Exclosure) 0.037 0. 049 0.77 0. 4 4 0 Treatment(24 - hr Exclosure) - 0.026 0.054 - 0.49 0.624 Table 2. 6 . Effects of predator exclosures on herbivory for T. amazonia leaves in the 2015 drought. Leaves in the control group are the reference group (intercept). Statistically significance differences (p<0.05) denoted by *. Estimate Std. Error z value p Intercept - 2.800 0. 062 - 45.44 <0.001 * Treatment(24 - hr Exclosure) - 0.124 0.077 - 1.61 0.108 Table 2. 7 . Effects of predator exclosures on herbivory for T. amazonia leaves in the 2016 wet season. Leaves in the control group are the reference group (intercept). Statistically significance differences (p<0.05) denoted by *. 76 Folivorous Insect Biomass Estimate Std. Error z value p Intercept 1.499 0.636 2.356 0. 019* Season(2015 Drought) 0.257 0.765 0.336 0.737 Season(2016 Wet) 0.860 0.783 1.098 0.272 Table 2. 8 . Seasonal differences in folivorous insect biomass on O. pyramidale . Branches during the 2016 dry season are the reference group (intercept). Statistically significance differences (p<0.05) denoted by *. Folivorous Insect Biomass Estimate Std. Error z value p Intercept 2.282 0.402 5.672 <0.001* Season (2015 Drought) 0.194 0.511 0.379 0.705 Season(2016 Wet) 1.510 0.507 2.977 0.003* Table 2. 9 . Seasonal differences in folivorous insect biomass on T. amazonia . Branches during the 2016 dry season a re the reference group (intercept). Statistically significance differences (p<0.05) denoted by *. Folivorous Insect Biomass Estimate Std. Error z value p Intercept 1.807 1.185 1.525 0.127 Period(Post - experim ent) 0.361 1.459 0.248 0.804 Treatment( 24 - hr Exclosure) - 1 . 782 1.534 - 1.161 0. 245 Period(Post - experiment) *Treatment(24 - hr Exclosure) 1.131 2.051 0.552 0.581 Table 2. 10 . Predator exclosure effects on folivorous insect biomass on O. pyramidale in the 2015 drought. Branches in the control group during the pre - experiment survey period are the reference group (intercept). Statistically significance differences (p< 0.05) denoted by *. Folivorous Insect Biomass Estimate Std. Error z value p Intercept - 0.182 0.879 - 0.208 0.836 Period(Post - experiment) 2.022 1.248 1.620 0. 105 Treatment( 24 - hr Exclosure) 2 . 373 1.230 1.929 0. 054 Period(Post - experiment) *Treatment(24 - hr Exclosure) - 0.991 1.750 - 0.566 0.571 Table 2. 11 . Predator exclosur e effects on folivorous insect biomass on O. pyramidale in the 2016 dry season. Branches in the control group during the pre - experiment survey period are the reference group (intercept). Statistically significance differences (p<0.05) deno ted by *. 77 Folivorous Insect Biomass Estimate Std. Error z value p Intercept 0.649 0.582 1.115 0.265 Period(Post - experiment) 0.687 0.813 0.845 0.398 Treatment( 24 - hr Exclosure) - 0 . 226 0.820 - 0.275 0. 783 Period(Post - experiment) *Treatment(24 - hr Exclosure) 3.025 1.144 2.643 0.008* Table 2. 1 2 . Predator exclosure effects on folivorous insect biomass on O. pyramidale in the 2016 wet season. Branches in the control group during the pre - experiment survey period are the reference group (intercept). Statistically significance differences (p<0.05) denoted by *. Folivorous Insect Biomass Estimate Std. Error z value p Intercept 2.483 0.699 3.553 <0.001* Period(Post - experiment) - 0.103 0.988 - 0.105 0.917 Treatment( 24 - hr Exclosure) 0.624 0.988 0.632 0.528 Treatment( Daytime Exclosure) - 0.272 0.988 - 0.275 0.783 Period(Post - experiment) *Treatment(24 - hr Exclosure) - 0.433 1.397 - 0.310 0.757 Period(Post - experiment) * Treatment(Daytime Exclosure) - 0.304 1.399 - 0 .217 0.828 Table 2. 1 3 . Predator exclosure effects on folivorous insect biomass on T. amazonia in the 2015 drought. Branches in the control group during the pre - experiment survey period are the reference group (intercept). Statistically significance diff erences (p<0.05) denoted by *. Folivorous Insect Biomass Estimate Std. Error z value p Intercep t 2.209 0.731 3.022 0.003* Period(Post - experiment) 0.102 1.034 0.098 0.922 Treatmen t( 24 - hr Exclosure) 0.800 1.033 0.774 0.439 Treatment( Daytime Exclosure) 0.436 1.055 0.414 0.679 Period(Post - experiment) *Treatment(24 - hr Exclosure) - 0.659 1.461 - 0.451 0.652 Period (Post - experiment) *Treatment(Daytime Exclosure) - 1.738 1.495 - 1.163 0.245 Tab le 2.1 4 . Predator exclosure effects on folivorous insect biomass on T. amazonia in the 2016 dry season. Branches in the control group during the pre - experiment survey period are the reference group (intercept). Statistically significance differences (p<0.0 5) denoted by *. 78 Folivorous Insect Biomass Estimate Std. Error z value p Intercept 3.824 0.539 7.099 <0.001* Period(Post - experiment) 0.058 0.762 0.077 0.939 Treatment( 24 - hr Exclosure) - 0.952 0.762 - 1.248 0.212 Period(Post - experiment) *Treatment(24 - hr Exclosure) 1.645 1.078 1.526 0.127 Table 2.1 5 . Predator exclosure effects on folivorous insect biom ass on T. amazonia in the 2016 wet season. Branches in the control group during the pre - experiment survey period are the reference group (intercept). Statistically significance differences (p<0.05) denoted by *. 79 Figure 2.11 . Linear regression models for the effect of start time on number of birds counted. y = - 0.66x + 1.39 R² = 0.00 0 1 2 3 4 5 6 7 8 9 06:00 07:00 08:00 09:00 Birds counted Count start time 2016 Wet Season y = - 5.84x + 3.10 R² = 0.01 0 1 2 3 4 5 6 7 8 9 10 11 08:00 09:00 10:00 11:00 Count start time 2015 Drought y = - 11.05x + 5.47 R² = 0.03 0 1 2 3 4 5 6 7 8 9 08:00 09:00 10:00 11:00 Birds counted Count start time 2016 Dry Season y = - 6.58x + 3.45 R² = 0.01 0 1 2 3 4 5 6 7 8 9 08:00 09:00 10:00 11:00 Count start time 2016 Dry Season Resident Birds Only 80 LITERATURE CITED 81 LITERATURE CITED Bosc, C., F. Roets, C. Hui, and A. P auw . 2018. 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Billings. 2009. Stable - isotope analysis of diets of short - tailed fruit bats (Chiroptera: Phyllostomidae: Carollia). Journal of Mammalogy 90:1469 1477. 87 CHAPTER 3 : Tropical Trophic Cascade Linked to Reduction in Large - bodied Insect s by Birds and Bats Steven M. Roels 1,2,3 and Jefferson S. Hall 4 1 Department of Integrative Biology 2 Environmental Science and Policy Program 3 Ecology, Evolutionary Biology, and Behavior Michigan State University, 288 Farm Lane, East Lansing, Michig an, USA 44824 4 Smithsonian Tropical Research Institute Luis Clement Avenue, Bldg. 401 Tupper, Balboa Ancon, Panama, Republic of Panama 88 ABSTRACT I n many tropical ecosystems, predation pressure from birds and bats regulates populations of folivorous inse cts and decreases leaf damage in a trophic cascade that benefits trees. Detailed information regarding bird and bat predation pressure on individual arthropod taxa is necessary for understanding mechanisms underlying these trophic cascades. We conducted a bird and bat exclosure experiment on two tree species, Ochroma pyramidale and Terminalia amazonia , in a tropical forest restoration. There was a sign ificant trophic cascade on O. pyramidale , but not T. amazonia . We report the effects of bird and bat exclos ures on taxonomic composition and distribution of biomass within arthropod communities on both tree species. On O. pyramidale , predation by birds and bats reduced biomass for five folivorous insect orders (Coleoptera, Lepidoptera, Phasmida, Blattodea, and Orthoptera) and numbers of individuals for four of the five orders. Spider (Araneae) biomass and numbers on O. pyramidale wer e also reduced by bird and bat predation. These patterns were noticeably weaker or absent on T. amazonia , where a trophic cascade w as not observed. We did not find evidence that intraguild predation by birds and bats on spiders dampened trophic cascade eff ects. Bird - and bat - driven trophic cascades have been found in natural, agricultural, and restoration ecosystems, demonstrating tha t conservation of bird and bat populations is a critical component of maintaining trophic relationships and ecological functi on. INTRODUCTION Ecological effects of herbivorous insects in the tropics are pervasive. Insects are regarded as drivers of tree c ommunity structure and diversity (Janzen 1970; Connell 1971), plant speciation rates (Coley and Kursar 2014), and evolution o f plant defenses (Coley and Barone 1996). 89 Arboreal insects are the primary herbivores of tropical foliage (Coley and Barone 1996), and serve as prey for hundreds of species in two of the most speciose groups of tropical vertebrates; birds and bats. In many tropical ecosystems, predation pressure from birds and bats regulates populations of folivorous (i.e., chewing) insects and decrea ses leaf damage in a trophic cascade that benefits trees ( Van Bael et al. 2008 ; Mäntylä et al. 2011; Maas et al. 2016). Deta iled information regarding bird/bat predation pressure on individual arthropod taxa is necessary for understanding mechanisms under lying bird/bat - insect - plant trophic cascades. On trees in temperate regions, indirect reductions in plant damage by birds are often attributed to high rates of caterpillar consumption (Atlegrim 1989; Marquis and Whelan 1994; Strong et al. 2000; Mols and Vi sser 2002; Singer et al. 2014). In tropical forests, Lepidoptera are also important herbivores (Van Bael et al. 2004; Dyer et al. 2007) but studies of arthropod community structure suggest the relative importance of other insect taxa with herbivorous speci es, such as Coleoptera, Blattodea, Phasmida, and Orthoptera, may be greater than in temperate regions (tropical studies: Stor k 1988; Basset 2001; Kitching et al. 2001; Ellwood and Foster 2004, temperate studies: Trieff 2002; Southwood et al. 2004). Predato ry arthropods, especially spiders, are also common in tropical arthropod communities and may consume folivorous insects. High intraguild predation by vertebrates on arthropod predators is counterintuitively associated with strong trophic cascades because o f positive correlations between vertebrate predation on arthropod predators and herbivores (Mooney et al. 2010). However, the re are circumstances where intraguild predation dampens cascade effects (Karp and Daily 2014; Bosc et al. 2018). 90 Studies on tropic al trophic cascades with bird/bat exclosures have typically reported exclosure effects on arthropod density (individuals/unit area; Van Bael et al. 2003; Van Bael and Brawn 2005; Kalka et al. 2008; Williams - Guillén et al. 2008; Maas et al. 2013). Some stud ies reporting arthropods (large va riably defined as body lengths >3 or >5 mm), based on the presumption that larger arthropods are preferred prey items and would, th erefore, be more likely to show effects of bird/bat predation (Philpott et al. 2004; Jedlicka et al. 2006; Van Bael et al. 20 07; Karp and Daily 2014). While census - based approaches to characterizing arthropods in trophic cascade studies are straightforward , they rely on the assumption that herbivory is proportional to number of individual arthropods, despite the wide range of ar thropod body sizes. Correlations between aggregate biomass and census counts vary widely among arthropod communities and metrics of biomass have been recommended over census counts when ecological functions and trophic interactions are of interest (Saint - G ermain et al. 2007). Allometric scaling of arthropod body length to body mass causes a two - fold increase in body length to result i n a four - fold or greater increase in body mass (Schoener 1980; Sample et al. 1993), with concomitant increases in metabolic d emands (Brown et al. 2004). The amount of herbivory caused by 50 mm caterpillar, for example, likely dwarfs that caused by a 10 mm caterpillar because the longer caterpillar may have a mass 60 or even 120 times greater (Sample et al. 1993; Wardhaugh 2013). Exclosure studies that have accounted for differences in arthropod size by converting census data into total arthropod or folivor ous insect biomass are uncommon (Bridgeland et al. 2010; Morrison and Lindell 2012). Bridgeland et al. (2010) reported both counts of individuals and biomass per unit area; one of their research sites showed stronger evidence for exclosure effects 91 on bioma ss than counts and the other site showed the reverse (Bridgeland et al. 2010: Fig. 4). Effects of predator ex closures will be more evident for biomass instead of count data in cases where predators disproportionately consume large arthropods. Accounting fo r arthropod size is especially important in tropical studies, where the range of arthropod body size is great er than temperate regions due to the presence of extremely large taxa (Makarieva et al. 2005). There is some evidence that tropical birds preferen tially forage on arthropods larger than 5 mm (Philpott et al. 2004; Jedlicka et al. 2006; Van Bael et al. 200 7) and may especially target large prey items during the breeding season for efficient food delivery to nestlings (Greenberg 1995). Preference for large arthropods as prey items is likely also true for tropical foliage - gleaning bats. Kalka and Kalko (2006) studied the diet of the smallest of nine foliage - gleaning bat species occurring on Barro Colorado Island, Panama and found it primarily consumed a rthropods larger than 10 mm. This bat species has a body mass (5 7 g, Kalka and Kalko 2006) similar to that o f a small insectivorous bird, such as a gnatcatcher (Polioptilidae; Dunning 2008). Gnatcatchers do not usually consume arthropods greater than 10 m m (Burger et al. 1999), suggesting that foliage - gleaning bats in general may be even more focused on large ar thropod prey than birds. We conducted a multi - season bird/bat exclosure experiment on two tree species in a tropical forest restoration to measure trophic cascade strength and evaluate the effects of bird/bat predation on tropical arthropod communities. A measurable trophic cascad e was only detectable during the 2016 wet season, which had typical rainfall (see Chapter 2 of this dissertation). One tr ee species showed greater leaf herbivory in the exclosure treatment (i.e., birds/bats absent) while the other species showed no differe nce between exclosure and control treatments. Here, we 92 report the effects of bird/bat exclosures on the taxonomic composi tion and distribution of biomass within the arthropod communities on both tree species during the 2016 wet season. METHODS Study s ite and species We conducted our study in a plantation - style native species forest restoration at the Agua Salud Project (ASP 300 m a.s.l.). The plantation was established in 2008 with plots c ontaining native species monocultures or polycultures (see Mayoral et al. 2017 for details) . We selected two tree species for our study, Terminalia amazonia (J.F.Gmel.) Exell (Family: Combretaceae ) and Ochroma pyramidale (Cav. ex Lam.) Urb. (Family: Malvaceae). Our study occurred in the middle portion of the 2016 wet season (June 11 July 27). Rainfall during June and July wa s 502 mm, similar to the average of 549 mm (2009 2017, STRI 2018). Experimental protocol Exclosure material and design We constructed predator e xclosure s out of light - weight plastic netting (Standard BirdNet; Bird - X, Chicago, IL, USA), with mesh large eno ugh (2 x 2 cm) to permit passage of large arthropods but not foraging birds or bats. For each tree species, we paired n eighboring trees and randomly assigned trees within pairs to control and exclosure treatments. There were 36 pairs of O. pyramidale and 4 1 pairs of T. amazonia . On each tree, we selected a single branch (~50 cm length) that showed evidence of active growth so we could measure herbivory on leaves flushed during the experiment. Exclosures remained in place for 30 31 days on O. pyramidale and 34 93 35 days on T . amazo nia . Additional details of exclosure construction are provided in Chapter 2 of this dissertation. Arthropod surveys A single observer, S. Roels, censused arthropods just prior to placement of exclosures and again at the end of the ex periment. During each survey, S. Roels visually searched for arthropods on all leaves of focal branches , inspecting both upper and lower leaf surfaces . Arthropods were identif ied to order and categorize d into the following body length classes 4 mm, >4 6 mm, >6 mm . Ar thropods greater than 6 mm were measured with a hand ruler. We did not attempt to count ants (Hymenoptera: Formicidae) due to difficulty of censusing individuals on trees with colonies. We report arthropod counts and biomass per m 2 of leaf area (see Chapte r 2 of this dissertation for leaf area calculation methods) . Arthropod b iomass estimate s We converted arthropod lengths into dry arthropod biomass using published length:mass regression relationships specific to taxonomic order (Scho ener 1980; Sample et al. 1993; Johnson and Strong 2000; Wardhaugh 2013). The biomass of arthropods placed into body length classes was calculated using the midpoint of the range (e.g., 2 4 mm was treated as 3 mm for biomas s calculations). Arthropods in the < 2 mm class were treated as having a body length of 1.5 mm because arthropods smaller than 1 mm were unlikely to be counted on visual surveys. Additional details of biomass calculations are provided in Chapter 2 of this dissertation (Methods and Table 2. 1). 94 Preliminary review of the distributions of individual arthropod biomass values on each tree species revealed clear discontinuities at the high end of the distributions. A few extremely large arthropods (>90 mg) on eac h tree species were over twice as massive, or more, as the next largest individual. We regarded these arthropods as outliers and removed two individuals from the O. pyramidale and four from the T. amazonia biomass data sets. Biomass estimates with th e appendix . Sample sizes for individual arthropod orders sometimes represented large portions of total biomass for their taxa w ithin a survey period. For these r easons, we qualitatively describe the effects of exclosures on the biomass of individual taxa. RESULTS We counted 357 arthropods on O. pyramidale and 558 on T. amazonia . We recorded arthropods in three non - insect orders ( spiders: Araneae, scorpions: Scorp iones, pillbugs: Isopoda) and ten insect orders (roaches: Blattodea, beetles: Coleoptera, flies: Diptera, true bugs: Hemiptera, bees and wasps: Hymenoptera, butterflies and moths: Lepidoptera, mantises: Mantodea, lacewings : Neuroptera, grasshoppers and all ies: Orthoptera, and stick insects: Phasmida). Diversity within these taxa appeared to be high; although we only identified arthropods to the order level, we did not notice any particular morpho - species that were dominant within their order. The biomass an d census count of many orders increased from our pre - to post - experiment surveys on both control and exclosure branches, suggesting a trend driven by environmental conditions unrelated to treatment group (Tables 3. 1 3. 4). 95 On O. pyramidale , our previous st udy found a significant increase in folivorous insect biomass on exclosure, but not control, branches ( see Chapter 2 of this dissertation : Fig. 2.1 ). On exclosure branches, biomass increased for each of the five primary folivorous orders (Table 3. 1; Fig. 3 .1 ). Figure 3.1. Arthropod biomass by order on O. pyramidale before (Pre) and after (Post) the exclosure experiment. Two individual arthropods with mass greater than 90 mg were regarded as outliers and removed from the data set. Brackets denote folivorou s orders. On control branches, the direction of change for folivorous orders was not consistent (three increases, one without change, one decrease) and the magnitudes of any positive changes were smaller than for exclosure branches (Table 3. 1). Spider bio mass increased in both treatment groups but the increase on control branches was less than half that observed on exclosure branches (4 .80 mg/m 2 leaf area vs. 10.47 mg/m 2 leaf area) (Table 3. 1). The effect of exclosures on biomass may actually be underestim ated in Fig. 3.1 data set were found on exclosure branches in the post - experiment survey (Table 3. 1; Appendix, Fig. 3. 7 ). 96 Taxon Ctr - Pre Ctr - Post Ctr Exc - Pre Exc - Post Exc Coleoptera 0.62 1.10 0.48 0.96 3.88 2.92 Lepidoptera 0.25 1.47 1.22 0.00 5.02 (18.14) 5.02 (18.14) Phasmida 0.00 0.00 0.00 0.00 2.08 2.08 Blattodea 0.04 0.84 0.80 0.38 4.14 3.76 Orthoptera 1.62 0.89 - 0.73 0.00 2.64 (40.43) 2.64 (40.43) Hemiptera 9.20 6.52 - 2.68 8.55 1.79 - 6.76 Araneae 1.36 6.16 4.80 4.28 14.75 10.47 Other 0.43 0.75 0.32 0.66 2.03 1.37 Folivorous 2.53 4.30 1.77 1.34 17.76 (68.67) 16.42 (67.33) All Arthropods 13.53 17.73 4.20 14.83 36.33 (87.24) 21.50 (72.41) Table 3.1. Biomass (mg) of arthropod ord ers per m 2 leaf area during pre - and post - experiment surveys on control (Ctr) and exclosure (Exc) branches of O. pyramidale . total of Coleoptera, L epidoptera, P differences between pre - and post - experiment surveys within treatment group. Exclosure effects were less obvious using count data because the total number of arthropods increased by over double i n both treatment groups between pre - and post - experiment surveys (Table 3. 2; Fig. 3 . 2 ). However, 93.1% of the increase in arthropod number s on exclosure branches was due to folivorous insects and spiders while 61% of the increase on control branches was du e to Hemipterans (especially small Auchenorrhynch 3. 2). Taxon Ctr - Pre Ctr - Post Ctr Exc - Pre Exc - Post xc Coleoptera 0.19 0.86 0.67 0.28 0.77 0.49 Lepidoptera 0.10 0.31 0.22 0.00 1.08 1.08 Phasmida 0.00 0.00 0.00 0.00 0.23 0.23 Blattodea 0.10 0.08 - 0.02 0.09 0.85 0.75 Orthoptera 0.19 0.08 - 0.12 0.00 0.31 0.31 Hemi ptera 1.07 3.76 2.69 1.50 2.01 0.51 Araneae 0.68 1.80 1.12 1.12 4.01 2.89 Other 0.39 0.23 - 0.15 0.47 0.39 - 0.08 Folivorous 0.58 1.33 0.75 0.37 3.24 2.87 All Arthropods 2.72 7.12 4.40 3.46 9.64 6.18 Table 3.2. Census counts of arthropod orde rs (individuals per m 2 leaf area) during pre - and post - experiment surveys on control (Ctr) and exclosure (Exc) branches of O. pyramidale . ferences between pre - and post - experiment surveys within treatment group. 97 Figure 3. 2 . Census counts of arthropod orders on O. pyramidal e before (Pre) and after (Post) the exclosure experiment. Brackets denote folivorous orders. On T. amazonia , our previous study did not find significant exclosure effects on folivorous insect biomass. There were no consistent patterns in the direction of change for individual order biomass in either treatment group (Table 3. 3; Fig. 3. 3 included suggests a greater increase in folivore biomass on exclosure than control branches (exclosure: 79.89 mg/m 2 leaf area vs. control: 14.73 mg/m 2 leaf area) but the large change on exclosure branches was due to a single roach and single caterpillar (T able 3 .3 ; Appendix, Fig. 3. 8 ). Without those two individuals, the change in folivore biomass is lower on exclosure than control br anches (exclosure: 6.12 mg/m 2 leaf area vs. control: 14.73 mg/m 2 leaf area) (Table 3 .3 ). 98 Taxon Ctr - Pre Ctr - Post Exc - Pre Exc - Post Coleoptera 7.99 5.59 - 2.40 4.09 2.54 - 1.55 Lepidoptera 1.05 (32.40) 9.65 8.60 ( - 22.75) 4.12 4.33 (26.91) 0.21 (22.79) Phasmida 0.00 0.00 0.00 0.00 8.20 8.20 Blattodea 0.06 16.20 (36.56) 16.20 (36.56) 6.83 9.28 (60 .46) 2.44 (53.62) Orthoptera 8.47 0.86 - 7.60 3.35 0.17 - 3.17 Hemiptera 9.83 1.38 - 8.45 1.48 5.78 4.30 Araneae 4.20 11.28 7.08 10.63 21.59 10.96 Other 0.97 0.14 - 0.83 0.36 0.06 - 0.30 Folivorous 17.57 (48.91) 32.30 (52.66) 14.73 (3.75) 18.39 2 4.51 (98.28) 6.12 (79.89) All Arthropods 32.57 (63.91) 45.10 (65.46) 12.53 (1.55) 30.85 51.94 (125.71) 21.09 (94.85) Table 3.3 . Biomass (mg) of arthropod orders per m 2 leaf area during pre - and post - experiment surveys on control (Ctr) and exclosure (E xc) branches of T. amazonia . Numbers Coleoptera, Lepidoptera, Phasmida, Bla differences between pre - and post - ex periment surveys within treatment group. Figure 3. 3 . Arthropod biomass by order on T. amazonia before (Pre) and after (Post) the exclosure experiment. Four individual arthropods with mass greater than 90 mg were regarded as outliers and removed from the data set. Brackets denote folivorous orders. Examining count data, there is also a lack of evidence that exclosures affected the arthropod community (Table 3 . 4; Fig. 3. 4 ). Changes in census counts for individual orders were largely parallel between contr ol and exclosure branches. For example, the increases in Blattodea and 99 spiders on exclosure branches were similar to increases observed on control branches, su ggesting dynamic populations of these taxa but not exclosure effects (Table 3. 4; Fig. 3. 4 ). Taxon Ctr - Pre Ctr - Post Ctr Exc - Pre Exc - Post xc Coleoptera 1.97 1.51 - 0.46 1.70 1.05 - 0.65 Lepidoptera 0.66 0.60 - 0.05 0.46 0.60 0.14 Phasmida 0.00 0.00 0.00 0.00 0.30 0.30 Blattodea 0.16 1.96 1.80 0.62 2.25 1.63 Orthoptera 0.49 0.15 - 0.34 0.15 0.15 0.00 Hemiptera 2.46 0.76 - 1.70 1.08 2.40 1.31 Araneae 2.46 6.34 3.88 4.80 8.10 3.30 Other 2.30 0.76 - 1.54 1.55 0.60 - 0.95 Folivorous 3.28 4.23 0.95 2.94 4.35 1.41 All Arthropods 10.49 12.08 1.59 10.38 15.44 5.07 Table 3.4. Censu s counts of arthropod orders (individuals per m 2 leaf area) during pre - and post - experiment surveys on control (Ctr) and exclosure (Exc) branches of T. amazonia . tr - an d post - experiment surveys within treatment group. Figure 3. 4 . Census counts of arthropod orders on T. amazonia before (Pre) and after (Post) the exclosure experiment. Brackets denote folivorous orders. 100 Examining the distribution of body sizes for O. py ramidale , we found that individuals with small body size (square root of biomass (mg) < 2) increased for both con trol and exclosure groups between survey periods, indicating a temporal effect unrelated to our experimental treatments (Fig. 3. 5 ). For arthropods with larger body sizes (square root of biomass (mg) > 1.5), exclosure branches showed clearly increased numbe rs in the post - experiment period while control branches demonstrated only a weak positive trend (Fig. 3. 5 ). Figure 3. 5 . Distribution of arthropod sizes on O. pyramidale for control branches (top panel) and exclosure branches (bottom panel) before (Pre ) and after (Post) the exclosure experiment. Two individual arthropods with mass greater than 90 mg were regarded as outliers and removed from the data set. Bins are centered on whole numbers (e.g. bin 2 includes values from 1.5 to < 2.5). The x - axis is on the square root scale. 101 Like O. pyramidale , distributions of body sizes on T. amazonia (Fig. 3. 6 ) also showed sli ght increases in the number of small - bodied individuals in post - experiment surveys for both treatment groups. Unlike O. pyramidale , there was n ot a clear signal of exclosure effects on the number of larger - bodied individuals as neither treatment group demo nstrated substantial changes in numbers between survey periods (Fig. 3. 6 ). Figure 3. 6 . Distribution of arthropod sizes on T. amazonia for c ontrol branches (top panel) and exclosure branches (bottom panel) before (Pre) and after (Post) the exclosure experiment. Four individual arthropods with mass greater than 90 mg were regarded as outliers and removed from the data set. Bins are center ed on whole numbers (e.g. bin 2 includes values from 1.5 to < 2.5). The x - axis is on the square root scale. 102 DISCUSSION Predation by birds/bats appeared to reduce biomass for each of five folivorous insect orders (Blattodea, Coleoptera, Lepidoptera, Phasmi da, Or thoptera) on O. pyramidale , the tree species where we detected a trophic cascade. Birds and bats also reduced number of individuals for four of the five orders (the positive change in Coleoptera on exclosure branches was not more than observed on con trol b ranches). Others have also suggested these taxa are regular prey items of tropical birds and bats (Greenberg 1995; Kalka and Kalko 2006) and exclosure studies have found these taxa are likely to be affected by bird/bat predation (Van Bael et al. 2003 ; Morr ison and Lindell 2012; Maas et al. 2013). These patterns were noticeably weaker or absent on T. amazonia , where no trophic cascade was present. Possible reasons for differences in bird/bat effects between tree species are discussed in Chapter 2 of this dis sertation. A study in the broader ASP landscape found a positive correlation between insectivorous bird abundance and attack rates by birds on artificial caterpillars (Roels et al. 2018). Out of six land cover types studied, attack rates in the nat ive spec ies plantation were second - lowest (Roels et al. 2018). This suggests the magnitude and pervasiveness of vertebrate predator effects on arboreal folivorous insects in central Panama may generally be higher than what we observed in this study, in a c ontext w ith low bird abundance. Phloem - feeding (also called sap - sucking) herbivores (Hemiptera) were one of the more common herbivorous taxa on both tree species but assessing their effect on plants is difficult since they do not leave obvious damage. Lei gh (1999 ) speculated that, given their abundance and metabolisms, phloem - feeders in tropical forests may actually remove more plant biomass than chewing insects. Other authors agree that the functional significance of phloem - feeders has been 103 overlooked (Co ley and Barone 1996). In a temperate forest, presence of birds had a significant, but weaker, negative effect on phloem - feeders than on folivorous insects (Bridgeland et al. 2010). In our study, phloem - feeders did not appear to increase in bird/bat absence , matchi ng prior results from central Panama (Van Bael et al. 2003). High biomass and counts of Hemiptera on O. pyramidale relative to folivorous orders may be related to the life history strategy of the tree. Schowalter (1994) proposed that higher abundan ces of p hloem - feeders and lower abundances of chewing insects may be a characteristic of disturbed areas in tropical forests. Ochroma pyramidale is regarded as a short - lived pioneer species that specializes on disturbed areas. However, it is unclear if the associa tion between phloem - feeders and disturbance is a function of abiotic conditions in disturbed areas or characteristics of host plants that colonize such areas (Schowalter 1994). Like folivorous insect orders, spiders demonstrated positive exclosure effects on O. pyramidale (increased biomass and numbers relative to changes on control) while exclosure effects on T. amazonia were equivocal (slightly greater change in biomass but smaller change in numbers than control). Intraguild predation of intermed iate pre dators, especially spiders, by birds/bats has been cited as a mechanism that potentially negates top - down trophic cascades on herbivorous insects (Mooney et al. 2010; Karp and Daily 2014; Bosc et al. 2018). Our finding that bird/bat predation on sp iders wa s associated with reductions in both folivorous insects and herbivory aligns with the general conclusions of Mooney et al. (2010), who found that high rates of vertebrate insectivore predation on predatory arthropods were positively correlated with vertebr ate insectivore predation on herbivorous arthropods. Thus, intraguild predation did not dampen trophic cascade strength (Mooney et al. 2010). Unlike Bosc et al. (2018), we did not notice a 104 decline in numbers of phloem - feeders on exclosure branches in respo nse to an increase in intermediate predator numbers. We did not census ants during our surveys but multiple species were regularly observed. A small portion of our selected trees, mostly T. a mazonia , were colonized by a tree - nesting species ( Aztec a sp.) a nd dozens of individual ants could sometimes be observed on study branches. We are uncertain regarding the ecology of those ants, but ants in general are strong interactors in tropical food webs that can act as predators or tenders of herbivorous i nsects ( Floren et al. 2002; Davidson et al. 2003; Philpott et al. 2004; Gras et al. 2016). Intermediate predators other than ants and spiders on our trees were uncommon, including Hymenoptera (Vespidae), Hemiptera (Reduviidae) and possibly a few Coleoptera (e.g., Lampyridae) and Neuroptera ( Chrysopidae ). prey items for birds have delineated body length categories at either 3 or 5 mm (Philpott et al. 2004; Jedlicka et al. 2006; Van Bael et al. 2007; Karp and Daily 2014). Dry biomass of 1 3 mg is approximately equivalent to a 5 mm body length, although there is wide variation by taxon due to differences in stereotypical body shape (e.g., long and thin for Phasmida) a nd water content (e.g., high in Lepidopteran larvae). In our study, exclosure effects on O. pyramidale were most evident on arthropods larger than 2.25 mg (square root > 1.5 on Fig. 3. 5 , lower panel), equivalent to individuals with body lengths longer than 5 11 mm , depending on order. Birds and bats likely select prey based on a number of characteristics, including body size, conspicuousness, nutritional content, and palatability, so it is unclear whether distinctions made using body length or biomass would be more relevant when seeking to characterize availability of preferred prey 105 items. However, we find that using biomass > 2.25 mg as the mark of preferred insectivorous bird prey is fu nctionally similar to using > 5 mm. We concur with Saint - Germain et al. (2007) that studies focused on ecological function of arthropod feeding guilds or higher - level taxa like orders should consider using biomass values in addition to census counts. This approach is commonly taken by studies investigating general arthropod communit y structure in tropical forests (Basset 2001; Ellwood and Foster 2004; Dial et al. 2006). Presenting biomass per unit of leaf area accounts for arthropod size and would allow fo r more direct comparisons between trophic cascade studies occurring in location s with different mean arthropod sizes (e.g., temperate vs. tropical regions, dry vs. wet forests). Published allometric equations are available for spiders and most insect taxa but researchers should take care to use equations from latitudes and climates s imilar to their field site due to variation in coefficient values between ecosystems (Schoener 1980). The extremely large (> 90 mg) arthropods we treated as biomass outliers we re rare but may be functionally significant components of folivorous insect com munities on the tree species studied. One exceptional insect (Orthoptera: Prosco piidae ) found on O. pyramidale had a body length of 70 mm. Its estimated mass w as nearly equivalent to the combined mass of all other folivorous insects found on O. pyramidale exclosure branches during the post - experiment survey period. Two other individua ls of this species were encountered on O. pyramidale during related field work, including one measuring 130 mm; it is possible this species is an important herbivore of O. pyramidale , despite low population density. e 106 sufficiently large to reduce predation likelihood is uncertain. Consumption b y smaller species of insectivorous bird seems improbable but foliage - gleaning bats demonstrate a remarkable ability to consume very large prey items relative to their own body s ize (Kalka and Kalko 2006). Studies seeking to accurately quantify effects of e cological experiments on arthropod biomass in tropical regions should expect to be challenged by the presence of extremely large, but rare, individuals. Bird and bat effects ar e confounded in our study because we did not use diurnal and nocturnal exclosur es, but kept exclosures in place continuously. Thus, patterns we report characterize the net effects of both predator groups. Prior studies have found mixed results regarding wh ether bird/bat effects on arthropod communities are additive or functionally di stinct (Williams - Guill én et al. 2008; Morrison and Lindell 2012; Maas et al. 2013; Karp and Daily 2014). Our arthropod surveys occurred during the daytime hours, when temperatur es are higher and humidity is lower. It is possible that some arthropod species present on our study trees take refuge from abiotic conditions and risk of bird predation during the day and are more active at night. Nocturnal sampling may have provided a di fferent perspective on relative abundance of each arthropod order or on their c ontributions to overall arthropod biomass, especially for Blattodea and Orthoptera, which are often nocturnal (Novotny et al. 1999; Mass et al. 2013). Precise mechanisms of bird/bat - driven trophic cascades are variable between ecosystems yet the benefici al relationship between vertebrate predators and trees has been revealed at a variety of latitudes on at least five continents. These cascades have been found in natural (Marquis and Whelan 1994; Van Bael et al. 2003; Bridgeland et al. 2010), agricultural (Van Bael et al. 2007; 107 Williams - Guillén et al. 2008; Maas et al. 2013) and restoration (Morrison and Lindell 2012, this study) ecosystems, demonstrating that conservation of bird and bat populations is a critical component of maintaining trophic relationsh ips and ecological function. A CKNOWLEDGEMENTS We thank C. Lindell, P . Ostrom, G. Roloff, T. Getty, M. B. Hannay, and S . Bodbyl for valuable comments on the manuscript. Jade Porter provided essential assistance and humor in the field. Funding provided by t he Smithsonian Tropical Research Institute (STRI); Michigan State University (MSU) Department of Integrative Biology ; MSU Ecolo gy, Evolutionary Biology, and Behavior Program; MSU Caribbean and Latin American Studies Program ; and MSU School of Graduate Stud ies . We thank STRI and Federico Davis for access and logistical support at the Agua Salud Project (ASP). The ASP is a STRI Smar t Reforestation® site in collaboration with the Panama Canal Authority (ACP), the Ministry of the Environment of Panama, and othe r partners. The ASP is part of ForestGEO and the TreeDivNet network. Plantation management is supported by the ACP. Funding for the ASP comes from Stanley Motta, the Silicon Valley d Grand Challenge grant to BiodiversiTREE, the Heising - Simons Foundation, and the National Science Foundation (NSF grant EAR - 13 60391). 108 APPENDIX 109 APPENDIX Figure 3. 7 . Arthropod biomass on O. pyramidale by order before (Pre) and after (Post ) the exclosure experiment without arthropods >90 mg removed. Brackets denote folivorous orders. Figure 3. 8 . Arthropod biomass on T. amazonia by order before (Pre) and after (Post) the exclosure experiment without arthropods >90 mg removed. Brackets den ote folivorous orders. 110 LITERATURE CITED 111 LITERATURE CITED Atlegrim, O. 1989. Exclusion of birds from bilberry stands: impact on insect larval density and damage to the bilberry. Oecologia 79:136 139. Basset, Y. 2001. Invertebrates in the ca nopy of tropical rain forests How much do we really know? Plant Ecology 153:87 107. Bosc, C., F. Roets, C. Hui, and A. Pauw. 2018. 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The direct and indirect effects of insectivory by birds in two contrasting Neotropical forests. Oecologia 145:658 668. Van Bael, S. A. Van, J. D. Brawn, and S. K. Robinson. 2003. Birds defend trees from herbivores 115 in a Neotropical forest canopy. Proceedings of the National Academy of Sciences 100:8304 8307. Van Bael, S. A., S. M. Philpott, R. S. Greenberg, P. Bichier, N. A. Barber, K. A. Mooney, and D. S. Gruner. 2008. Birds as predators in tropical agroforestry systems. Ecology 89:928 934. Wardhaugh, C. 2013. Estimation of biomass from body length and width for tropical rainforest canopy invertebrates. Australian Journal of Entomology 52:291 298. Williams - Guillén, K., I. Perfecto, and J. Vandermeer . 2008. Bats limit insects in a neotropical agroforestry system. Science 320:70. 116 Steven M. Roels 1,2 , Jade L. Porter 3 , Catherine A. Lindell 1,4 1 Department of Integrative Biology, Michigan State University, 288 Farm Lane, East Lansing, Michigan, USA 44824 2 Address correspondence to S. M. Roels, email roelsste@msu.edu 3 Forsite Consultants Ltd, Salmon Arm, British Columbia, Canada V1E 2Y9 4 Center for Global C hange and Earth Observations, Michigan State University, East Lansing, Michigan, USA 48824 117 118 How tropical forest restoration s trategy a ff ect s development of animal - dependent ecosystem functions is not well - studied. A nimals drive important ecologi cal function s in restorations like pollination, seed dispersal, and regulation of herbivorous insects (Lindell 2008). H erbivore s and their predator s shape tropical communities (Floren et al. 2002; Coley & Kursar 2014); herbivory 119 can affect survival and gro wth of young tropical trees ( C lark & Clark 1985; Plath et al . 2011; Riedel et al. 2013 ) and predators of herbivores can create top - down trophic cascades that benefit plants in tropical ecosystems ( reviewed in Pace et al. 1999; Van Bael et al. 2008 ). While a number of studies have demon strated the important ecological function s of insect - eating birds and bats in tropical agro forest ecosystems ( reviewed by Maas et al. 2016 ), there has been little evaluation of predation pressure on herbivores in communities a long the restoration staircase (Morrison & Lindell 2012). P lantation - style forest restorations are in many ways analogous to agro forest ecosystems like shade coffee and cacao ; both feature reduced ecological complexity relative to natural forest and have s pecific management goals ( e.g . crop or timber production). 120 121 122 123 124 125 126 127 128 129 130 0% 10% 20% 30% 40% 50% 60% Bird Attack Arthropod Attack Any Attack Proportion Attacked Predator Category Caterpillar Control Ball ns ns * 131 ± 132 133 ns 134 135 136 137 138 139 140 141 142 143 Figure 4. 7 . W). Land cover types in and near Area B are non - native plantation (NON , brown ), native plantation (NAT , blue ), old - growth (OGU) mixed with other secondary forest ( both are gray ), secondary gallery forest (GAL , green ) , and Soberanía National Park (forest, dark gray). Red stars denote general areas where caterpillars were placed. 144 145 C oley P . D . , and T. A. Kursar . 20 07 . On Tropical Forests and Their Pests. Science 343:35 36 . 146 147 148 149 150 the r est oration of d egraded a reas in the t ropics . Restoration Ecology 13:92 102 .