DOI: 10.5061/dryad.gb5mkkws7
title: Reduced predation by arthropods and higher herbivory in burned Amazonian forests

description: Biodiversity losses have increased in tropical forests due to fire-related
 disturbances. As landscape fragmentation and climate change increase,
 fires will become more frequent and widespread across tropical rain
 forests worldwide, with important implications for forest dynamics by
 altering plant-animal interactions. Here we tested the hypothesis that
 recurrent fires in tropical rain forests change bottom-up and top-down
 forces controlling the abundance of insect herbivores, which in turn
 increases herbivory. To quantify herbivory, we collected 50 leaves per
 tree of five species in burned and unburned experimental plots (N = 75) in
 southeastern Amazonian forests. We measured leaf nitrogen content and leaf
 thickness of tree leaves as bottom-up factors that could explain
 differences in herbivory; we measured predation pressure on model
 caterpillars and estimated the abundance of predatory ants as top-down
 factors. We found higher herbivory in burned than in unburned forests, as
 well as lower predator attacks in caterpillar models and lower abundance
 of predatory ants. Leaf nitrogen content did not vary across treatments.
 Birds attacked model caterpillars more frequently in burned than in
 unburned forests, and leaf thickness was higher in burned forests, but
 these factors together were not enough to offset the higher herbivory in
 burned plots. Fire degrades tropical forests not only by killing trees and
 altering their structure and community dynamics, but also by reducing
 predatory arthropods and disrupting predator-prey interactions, which
 triggers increased herbivory. These indirect impacts of recurrent fires
 probably contribute to further alter forest structure, functioning, and to
 decrease forest regeneration in Amazonian forests.
description: Information from our manuscript: Study
 site The study was established at “Fazenda Tanguro”, in
 the state of Mato Grosso, Brazil, in the southern portion of the Amazon
 basin (13º 04’ S, 52º 23’ W). The climate is tropical humid, and the
 transitional forest between the Cerrado (Brazilian savannas) and the
 southeast Amazon forest composes the vegetation of our experimental plots.
 Expanding agriculture threatens the region (Brando et al. 2013; Marques et
 al. 2020), ultimately altering regional climate and increasing fire
 occurrence (Brando et al. 2013, 2020). Experimental
 fires were set in three large and adjacent 50-ha (500 × 1000 m) forest
 plots edging a soybean field. One plot was burned annually from 2004 to
 2010, except in 2008 (Burn 1yr); a second one was burned triennially, in
 2004, 2007 and 2010 (Burn 3yr); and a third plot was not burned (Control).
 Burnings were conducted at the end of the dry season between August and
 September, with kerosene drip torches. Fires were more intense and severe
 in the triennially burned plot because of higher litter accumulation
 (Balch et al. 2015), especially after the extreme drought events of 2007
 and 2010 (Brando et al. 2014). Even eight years after fires ceased, these
 burned forests continued to degrade, with an increased mortality of large
 trees, and decreased above-ground biomass and canopy cover (Brando et al.
 2019b). Our plots were not replicated due to logistical, ethical and legal
 limitations. However, the plots did not differ regarding several
 vegetation and microclimatic variables before fire (Balch et al., 2008),
 suggesting that observed differences can be attributed to the treatment
 rather than to site effects. For more details of the fire experiments,
 including pictures of burned and control sites, see Balch et al.
 (2015). Sampling design We selected
 five individuals from five tree species in each plot treatment (N = 75).
 These individuals were located across two linear transects, distant 500
 and 750 m from forest plot edges, and were spaced at least 10 m from each
 other. The selection of tree species was based upon the following
 criteria: absences of extrafloral nectaries and commonness across the
 study area, according to previous inventories (e.g., Brando et al. 2019b).
 The selected species were Myrcia multiflora (Lam.) DC. (Myrtaceae),
 Micropholis egensis (A.DC.) Pierre (Sapotaceae), Sacoglottis guianensis
 Benth. (Humiriaceae), Sloanea sinemariensis Aubl. (Elaeocarpaceae) and
 Tapirira guianensis Aubl. (Anacardiaceae). Herbivory
 assessment We quantified leaf area loss by chewing
 insects (hereafter herbivory) between March and April 2018 by sampling 50
 leaves from each tree – and a total of 1,250 leaves per plot. Sampled
 trees were at least 4m high, so sampled branches were chosen from a
 distance to avoid selection bias. We sampled branches from the outer part
 of the crown of all species; for that, we used a slingshot-like apparatus
 to throw a rope in the selected branch and then bring it down for
 sampling.  We collected the first 50 leaves from the base to the apex of
 each branch (according to the standardized protocol of Mendes et al.
 2021). Leaves were taken to the laboratory, scanned and had their margins
 or limb completed to account for leaf area loss to herbivores using the
 software GIMP. Measurements of leaf area and leaf area lost to herbivory
 were performed with the R package ‘EBImage’ (Pau et al. 2010). We
 calculated herbivory in each tree by averaging the percentage of leaf area
 loss across all 50 leaves sampled. Plant
 palatability Leaf nitrogen content indicates the plant
 nutritional value to insects (Mattson 1980) and it was estimated through
 chlorophyll content using a SPAD-502 (Spectrum Technologies, Inc.,
 Plainfield, IL, USA), as leaf chlorophyll concentration and leaf nitrogen
 content are linearly correlated (Loh et al. 2002). One intact leaf was
 randomly sampled from each tree (different from those which we measured
 herbivory), and measurements were taken at the lamina midpoint close to
 the midrib of the leaves. We measured leaf thickness as
 a proxy for sclerophylly. We sampled 10 intact leaves of each individual
 and measured leaf thickness with a Digital Micrometer (0-25 mm/0.001 mm
 MDC-Lite 293-821-30). Measurements were taken around the midpoint in both
 sides of the midrib of the leaves. We calculated mean leaf thickness for
 each leaf and then for each individual plant, using leaves as replicates
 per species. Predation pressure To
 estimate predation pressure, we first established a field experiment using
 artificial model prey – which represent herbivores as caterpillars and
 other herbivores that do not have anti-predation behavioral forms and
 strategies (Dáttilo et al. 2016; Roslin et al. 2017). Artificial
 caterpillars were made with oil-based, odorless, non-toxic soft modeling
 clay, and were ~ 3 cm long and 5 mm in diameter. All models were green and
 had a wire passing through their center to attach them to the branch
 (Figure S1). We established the experiment in May of 2018, at 500 m from
 the forest plots edges, in different plants than those used for the
 herbivory experiment, irrespectively of plant species. For each treatment
 plot we set 15 sampling points across a previously established trail (2 m
 wide), which were located 10 m apart from each other. Each sampling point
 consisted of two plants, located at least two meters apart and on opposite
 sides of the trail. Plants were between 1.5 and 2 m in height, and were
 chosen arbitrarily. One model prey was fixed distally in an outer branch
 at ~ 1.5 m height of each plant, totaling 30 artificial prey per plot -
 each sampling point had a pair of plants with an artificial prey in each.
 From the total artificial prey, four were colonized by termites and were
 discarded from our analyses. All model prey remained in the field for four
 days, and afterwards each model was examined for predation marks, and
 categorized to major groups of predators (birds, arthropods, mammals)
 according to Low et al. (2014). As an additional proxy
 to our estimation of predation pressure by arthropods we compared the
 abundances of predatory ants, which are the main arthropod predators in
 the tropics (Zvereva et al. 2019), across treatment plots. For that, we
 sampled ant communities in the same study area approximately one year
 (February 2019) after vegetation and predation experiment samplings. We
 established six sampling sites in each treatment plot (N = 18); all
 sampling sites were located at least 250 m from the forest edge, and at
 least 170 m apart from each other. Each sampling site consisted of a 3 × 3
 grid of nine sampling stations located 10 m apart. Each sampling station
 consisted of unbaited pitfall traps – one in the epigaeic and the other in
 the arboreal stratum. All pitfall traps were 5 cm in diameter, were partly
 filled with a salt solution and detergent, and were left open for 48 h.
 The epigaeic traps were buried with their rims flush to the soil surface,
 and arboreal traps were tied at a height of 2 m to the trunk of the
 nearest tree with dbh ≥10 cm. We sorted ants to species, and assigned
 morphospecies when species identification was not possible. Voucher specimens are held at the Laboratório de
 Ecologia de Comunidades e Ecossistemas Tropicais (EcoTrop), Universidade
 Federal de Viçosa (UFV), Brazil.<br clear="all"> Data
 analysis We adjusted a mixed beta regression model with
 logit link to test for the effects of fire on herbivory, with tree species
 as random variable. Results of beta regression modeling are shown in the
 text and figure as percentage rather than proportion for ease of
 interpretation. To test whether leaf nitrogen content and leaf thickness
 varied across treatment plots, we built two generalized linear mixed
 models, one with leaf nitrogen content and the other with leaf thickness
 as response variable, and fire treatments as explanatory variable. Both
 models had Gaussian error distribution, and tree species as a random
 effect. We compared predation pressure across treatment
 plots by considering arthropod and bird marks separately, as these two
 groups of predators were the most abundant in our study area. We built two
 generalized linear models using the proportion of predation marks in each
 sampling point (i.e., a pair of plants with a caterpillar model in each)
 as response variable - i.e., zero, 0.5, or 1 attack – and fire treatment
 as explanatory variable. We used binomial error distribution in both
 models, with a ‘quasi’ correction for overdispersion for the arthropods
 model. Finally, we classified sampled ants (both from epigaeic and
 arboreal strata) as predatory or non-predatory (according to Baccaro et
 al., 2015 and with help of a taxonomist; Julio M.C. Chaul), and used only
 predatory ants for analyses. We built a generalized linear model with
 Poisson error distribution using predatory ant abundance (epigaeic and
 arboreal ants combined) as response variable, and fire treatment as
 explanatory variable. Ant abundance in each transect was estimated as the
 sum of occurrence of predatory species in all nine traps of the transect
 combined. We conducted all analyses using the software
 R (R Core Team 2019); we used the glmmTMB package v1.0.1 (Brooks et al.
 2017) to adjust mixed beta regression models and the package lme4 to
 adjust glmm models. We analyzed the residuals to check for distribution
 suitability and fit in all models. To evaluate differences between fire
 plot treatments, pairwise contrasts analyses were made between control and
 two burned treatments combined, which were the most similar levels
 (Crawley 2013).