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Science of the Total Environment 631-632 (2018) 429-438 


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Increased anthropogenic disturbance and aridity reduce phylogenetic 

Check for 

and functional diversity of ant communities in Caatinga dry forest 

Xavier Arnan a,b '*, Gabriela B. Arcoverde c , Marcio R. Pie d , Jose D. Ribeiro-Neto a,e , Inara R. Leal 1 

a Programa de Pos-Cradua(ao em Biologia Vegetal, Universidade Federal de Pernambuco, Av. Prof. Moraes Rego s/no, Recife, PE 50670-901, Brazil 
b CREAF, Cerdanyola del Valles, ES-08193, Catalunya, Spain 

c Research School of Environment and Livelihoods, Charles Darwin University, Darwin, NT 0909, Australia 
d Departamento de Zoologia, Universidade Federal do Parana, Caixa Postal 19020, Curitiba, PR 81531-980, Brazil 

e Departamento de Fitotecnia e Ciencias Ambientais, Centro de Ciencias Agrarias, Universidade Federal da Paraiba, Rodovia PB-079,58397-000 Areia, PB, Brazil 
f Departamento de Botanica, Universidade Federal de Pernambuco, Av. Prof. Moraes Rego s/no, Recife, PE 50670-901, Brazil 



Ant biodiversity patterns under global 
change drivers are assessed in Caatinga. 
Functional and phylogenetic diversity 
decrease with aridity and human distur¬ 

Human disturbance and aridity interact 
in complex ways to endanger biodiver¬ 

Aridity can intensify the negative effects 
of disturbance on biodiversity. 

Concerns about the future of biodiver¬ 
sity in neotropical semi-arid regions. 

Miscellaneous resource use Miscellaneous resource use 

Livestock grazing 

Miscellaneous resource use 



Article history: 

Received 16 November 2017 
Received in revised form 3 March 2018 
Accepted 4 March 2018 
Available online xxxx 

Editor: Yolanda Pico 


Climate change 
Environmental filtering 
Functional diversity 
Anthropogenic disturbance 
Morphological traits 
Phylogenetic diversity 

Anthropogenic disturbance and climate change are major threats to biodiversity. The Brazilian Caatinga is the 
world's largest and most diverse type of seasonally dry tropical forest. It is also one of the most threatened, but re¬ 
mains poorly studied. Here, we analyzed the individual and combined effects of anthropogenic disturbance (three 
types: livestock grazing, wood extraction, and miscellaneous use of forest resources) and increasing aridity on tax¬ 
onomic, phylogenetic and functional ant diversity in the Caatinga. We found no aridity and disturbance effects on 
taxonomic diversity. In spite of this, functional diversity, and to a lesser extent phylogenetic diversity, decreased 
with increased levels of disturbance and aridity. These effects depended on disturbance type: livestock grazing 
and miscellaneous resource use, but not wood extraction, deterministically filtered both components of diversity. 
Interestingly, disturbance and aridity interacted to shape biodiversity responses. While aridity sometimes intensified 
the negative effects of disturbance, the greatest declines in biodiversity were in the wettest areas. Our results imply 
that anthropogenic disturbance and aridity interact in complex ways to endanger biodiversity in seasonally dry trop¬ 
ical forests. Given global climate change, neotropical semi-arid areas are habitats of concern, and our findings sug¬ 
gest Caatinga conservation policies must prioritize protection of the wettest areas, where biodiversity loss stands 
to be the greatest. Given the major ecological relevance of ants, declines in both ant phylogenetic and functional di¬ 
versity might have downstream effects on ecosystem processes, insect populations, and plant populations. 

© 2018 Elsevier B.V. All rights reserved. 

* Corresponding author at: CREAF, Campus UAB, 08193 Cerdanyola del Valles, Spain. 
E-mail address: (X. Arnan). 

https://d 0 i. 0 rg/l 0.1016/j.scitotenv.2018.03.037 
0048-9697/© 2018 Elsevier B.V. All rights reserved. 


X. Arrian et al. / Science of the Total Environment 631-632 (2018) 429-438 

1. Introduction 

Anthropogenic disturbance and global climate change are key 
threats to biodiversity (Bellard et al., 2012) because they have signifi¬ 
cant impacts on biological populations and community organization. 
This is especially true in seasonally diy tropical forests (SDTFs), which 
are experiencing increased rates of both acute and chronic disturbance. 
Major acute disturbances include habitat loss and fragmentation (Miles 
et al., 2006). Major chronic disturbances (hereafter referred to as CADs— 
chronic anthropogenic disturbances [sensu Singh, 1998]) include live¬ 
stock grazing, wood extraction, and the exploitation of miscellaneous 
forest resources. These activities all result in the removal of significant 
amounts of biomass. In STDFs, few efforts have been made to under¬ 
stand the impacts of CADs on biodiversity, but negative impacts have 
been described in communities of both plants (Sagar et al., 2003; 
Ribeiro et al., 2015, 2016; Rito et al., 2017) and animals (Ribeiro-Neto 
et al., 2016; Oliveira et al„ 2017). Climate change also threatens SDTFs 
(Dirzo et al., 2011) and might even exacerbate the effects of anthropo¬ 
genic disturbances (Hirota et al., 2011; Ponce-Reyes et al., 2013; Gibb 
et al., 2015a; Frishkoff et al., 2016). Hot and arid environments are likely 
at the greatest risk (Anderson-Teixeira et al., 2013; Gibb et al., 2015a). 

Studies of biological communities have generally focused on pat¬ 
terns of species diversity, which are often quantified using species rich¬ 
ness and/or composition (Pavoine and Bonsall, 2011). However, new 
diversity metrics that incorporate information about phylogenetic di¬ 
versity (PD) and functional diversity (FD) can reveal more information 
about community organization in different contexts (Faith, 1992; 
Webb et al., 2002; Petchey and Gaston, 2006; Swenson, 2014), including 
those marked by anthropogenic disturbance and climate change 
(Mouillot et al., 2013). While phylogenetic diversity reflects the accu¬ 
mulated evolutionary history of a community (Webb et al., 2002), func¬ 
tional diversity reflects the diversity of morphological, physiological, 
and ecological traits found therein (Petchey and Gaston, 2006). It is gen¬ 
erally accepted that PD and FD can increase with taxonomic diversity by 
chance, since the presence of more species should mean that more lin¬ 
eages and functions are represented. However, these relationships are 
not always linear. Two communities with equal taxonomic diversity 
might greatly differ in PD and FD (Petchey and Gaston, 2006; Safi 
et al., 2011 ; Arnan et al„ 2015,2017) due to different levels of functional 
redundancy, different evolutionary histories, and/or contrasting envi¬ 
ronmental conditions. However, a strong correlation between FD and 
PD would be expected if the functional traits that allow species to per¬ 
sist in the environment are evolutionarily conserved, that is to say, 
they display phylogenetic signals (Webb et al., 2002; Cavender-Bares 
et al., 2009). Remarkably, very little is known about how PD and the 
functional composition of animal communities change in response to 
disturbance and environmental conditions, especially in SDTFs. 

When examining biodiversity patterns, ants are a good study group 
—they are among the most diverse and abundant terrestrial organisms 
on earth and they are highly sensitive to environmental change 
(Holldobler and Wilson, 1990). Moreover, ants play an important role 
in many basic ecosystem services (Bihn et al., 2010, Del Toro et al., 
2012). In particular, ants are crucial contributors to soil cycling and aer¬ 
ation, organic matter decomposition, seed dispersal, and plant protec¬ 
tion (Del Toro et al., 2012). Ants are extremely phylogenetically 
diverse, especially in the tropics (Holldobler and Wilson, 1990), and 
ant morphological traits have frequently been used to infer ecosystem 
services (Weiser and Kaspari, 2006; Gibb et al., 2015b; Parr et al., 
2017; Salas-Lopez, 2017). 

In this study, we analyzed the effects of CADs and climate change, 
notably increasing aridity, on the phylogenetic and functional diversity 
of ants in the Brazilian Caatinga, the largest and most diverse of the 
world's SDTFs (Leal et al., 2005). The Caatinga is the third most- 
threatened Brazilian ecosystem and yet is the most poorly studied and 
understood (Overbeck et al., 2015; Oliveira and Bernard, 2017). The 
27 million people living in the Caatinga are highly dependent on its 

natural resources for their livelihoods, which has resulted in its slow 
degradation over time (Leal et al., 2005; Ribeiro et al., 2015). Moreover, 
the Caatinga is one of the six ecosystems with the greatest intrinsic vul¬ 
nerability to climate variability (Seddon et al., 2016); climate models 
consistently predict a reduction in rainfall levels (22%) and an increase 
in temperature (3-6 °C) (Magrin et al., 2014). 

In this context, the Caatinga is a good model system with which to 
investigate the effects of anthropogenic disturbance and climate change 
(i.e., increased aridity) on the biological communities of dry forests. It 
can also be used to characterize changes in community organization 
arising from transformations in SDTFs. Previous studies in the Caatinga 
found no or small differences in ant species richness along CAD gradi¬ 
ents; however, large changes in species composition were observed 
(Ribeiro-Neto et al., 2016; Oliveira et al., 2017). This finding suggests 
shifts in phylogenetic and functional diversity along CAD gradients. 
We therefore first corroborated that species diversity is not modulated 
by CAD and aridity gradients, and hypothesize the following: (a) PD 
and FD will decrease as anthropogenic disturbance and aridity increase; 
(b) PD and FD will decrease even more sharply in areas that are both 
highly disturbed and arid; and (c) PD and FD patterns along gradients 
of disturbance and aridity will be driven by deterministic processes 
rather than by stochasticity. 

2. Materials and methods 

2.1. Study area 

This study was conducted in Catimbau National Park (8°24'00" and 
8°36'35" S; 37°0'30" and 37°1'40" W, state of Pernambuco, Brazil), 
which cover an area of 607 km 2 of Caatinga vegetation (Sociedade 
Nordestina de Ecologia, 2002). The climate is hot. Mean annual temper¬ 
ature is 25 °C, and mean annual rainfall ranges between 1100 mm in the 
southeast to 480 mm in the northwest (Rito et al., 2017). However, the 
park experiences substantial interannual and spatial variability in con¬ 
ditions (Sociedade Nordestina de Ecologia, 2002). Most of the park has 
quartzolic sandy soils (70%), but planosols (15%) and lithosols (15%) 
are also present (Sociedade Nordestina de Ecologia, 2002). The domi¬ 
nant families of woody plants are Fabaceae, Euphorbiaceae, and 
Boraginaceae; on the surface of the forest floor, Cactaceae, 
Bromeliaceae, Malvaceae, Asteraceae, and Fabaceae dominate (Rito 
et al., 2017). 

The park was established in 2002 (Sociedade Nordestina de 
Ecologia, 2002), but its original human inhabitants remain; they 
continue to hunt, graze livestock, extract timber, collect firewood, 
and harvest other plant resources (Rito et al., 2017). Their historical 
presence has resulted in an extensive mosaic of differential land use 
and anthropogenic pressure on biota. This fact means Catimbau rep¬ 
resents an excellent opportunity for examining how anthropogenic 
disturbance (e.g., farming, livestock grazing, extraction of timber, 
firewood gathering, and hunting) affects the biota of the Caatinga. 
Also, the considerable variation in precipitation within the park 
(100%) can help reveal whether high levels of aridity can intensify 
the negative effects of human disturbance. 

2.2. Characterization of disturbance and aridity gradients 

We sampled 20 0.1-ha plots (20 x 50 m; separated by at least 
2 km) located within areas dominated by old-growth vegetation; 
the plots occurred along an aridity gradient and experienced varying 
degrees of CAD (Fig. 1). Thanks to aerial photographs and 
preliminary interviews with locals, we could confirm that the plots 
had not experienced any acute disturbances over the past 80 years. 
All plots were located in areas with the same soil type (sandy soil), 
slope (flat terrain), and vegetation type (dry forest with short- 
stature trees) (Rito et al., 2017). 

X.Aman et al. / Science of the Total Environment 631-632 (2018) 429-438 


We characterized disturbance intensity by calculating three different 
indices that corresponded to the main CADs affecting the Caatinga in 
general and Catimbau in particular: (1) livestock grazing (LG)—con¬ 
sumption of vegetation, trampling, and other physical damage caused 
by cattle and goats; (2) wood extraction (WE)—the extraction of dead 
and live wood for fuel, fence construction, and artisanal production; 
and (3) miscellaneous resource use (MU)—use of non-wood resources 
by humans (e.g., food and medicinal plants, hunting). Index values 

were calculated using the following formula: / = Sfeji 3 ' 1 x 

100, where / is disturbance intensity; y,- is the observed value for a given 
disturbance metric in plot i; y min is the minimum observed value for the 
disturbance metric across all plots; y max is the maximum observed value 
for the disturbance metric across all plots; and n is the number of indi¬ 
vidual disturbance metrics incorporated in the index. This formula thus 
standardizes the metrics (sometimes of different units) to take on a 
value between 0 and 1, allowing them to be combined in the same 
index. Index values ranged from 0 to 100 (from no disturbance to 
maximum-intensity disturbance). Both the LG and WE indices quanti¬ 
fied disturbances that were directly measured in the field. For the LG 
index, we estimated grazing levels by measuring the length of goat trails 
and the frequency of cattle and goat dung (see Appendix SI for details). 
Then, we combined the two estimates of goat grazing (trail length and 

dung frequency) by means of principal component analysis (PCA). 
Both measures were highly positively correlated (r > 0.90) with the 
first PCA axis, which explained 88% of variance. We therefore used its 
coordinates to obtain a single measure of goat grazing. The LG index 
was then calculated by inputting measures of goat grazing and cattle 
dung frequency into the formula above. For the WE index, we estimated 
the extraction of live wood and the collection of firewood (Appendix 
SI) and plugged them directly into the formula above. Finally, the MU 
index was determined using three indirect variables that are proxies 
for local anthropogenic pressure and habitat accessibility. More specifi¬ 
cally, we estimated two relevant geographic distances—plot proximity 
to the nearest house and plot proximity to the nearest road (using sat¬ 
ellite imagery and ArcGis f 0.1 software). We also used a socioecological 
variable—the number of people living in the area that influence the plot 
(Appendix SI). Then, the values of these metrics were inputted into the 
formula above to obtain the MU index. The three disturbance indices 
displayed a wide range of values (min-max for LG: 0-60, WE: 0-100, 
and MU: 5-63; Fig. 1 and Appendix S2) and were not highly correlated 
(LG vs. WE: r = 0.05, LG vs. MU: r = 0.61, and MU vs. WE: r = -0.10). 
This result underscores that the indices are quite independent and mea¬ 
sure different forms of anthropogenic disturbance. 

Field-based research into temporal evolutionary change typically re¬ 
quires long-term data that are unavailable for most systems. 

Fig. 1 . Location of the study region (in dark gray; within Brazil) (A), Catimbau National Park (white box; within Pernambuco) (B), and the study plots (bar graphs; within Catimbau) (C). 
The bar graphs depict the intensity of each anthropogenic disturbance (wood extraction, WE; livestock grazing, LG; miscellaneous resource use, MU) on each plot. The color scale depicts 
aridity (i.e„ climatic water deficit). 


X. A man et al. / Science of the Total Environment 631-632 (2018) 429-438 

Furthermore, carrying out climate change studies in the laboratory is 
difficult, especially if the aim is to address issues at the community 
level. At present, the only effective means of exploring changes in biodi¬ 
versity due to rapid climatic shifts is the space-for-time approach, 
where space acts as a substitute for time (Blois et al„ 2013). However, 
historical and evolutionary processes might not act similarly along tem¬ 
poral versus spatial gradients. Consequently, inferences based on cer¬ 
tain climatic conditions can, but do not always, reflect the responses 
expected over the timeline of future climate change (Bellard et al., 
2012). Here, we explored the potential effects of declining precipitation 
and increasing temperature in the Caatinga by analyzing changes in 
aridity along a spatial gradient. Aridity was estimated using mean an¬ 
nual climatic water deficit, which is the difference between potential 
evapotranspiration (PET) and actual evapotranspiration (AET; based 
on biologically usable energy and water) (Lutz et al., 2010). Climatic 
water deficit was calculated using 30-arc-second (1-km) resolution 
maps of long-term mean annual PET and AET (CGIAR-CSl's Global Arid¬ 
ity and PET Database and Global High-Resolution Soil-Water Balance 
Database;[2009]). These maps were generated 
using temperature and precipitation data from the WorldClim global 
climate data repository ( For each plot, the differ¬ 
ence between annual PET and AET was calculated to obtain a climatic 
water deficit value. All calculations were performed using ArcGIS 10.1 
software. Climatic water deficit values ranged from 658 mm (minimum 
aridity) to 1086 mm (maximum aridity), and was not correlated to any 
of the disturbance indices (r= —0.14, —0.18 and —0.06 for LG, WE and 
MU, respectively). 

2.3. Sampling ant communities 

Each plot contained 20 pitfall traps (4x5 grid; separated by 5 m). 
The traps were 4.5-cm diameter plastic containers partially filled with 
a mixture of alcohol, ethylene glycol, and soap. Traps were left open 
for a single 48-h period in March 2015, at the beginning of the rainy sea¬ 
son. All the ants collected were sorted to morphospecies following 
Baccaro et al. (2015). They were identified when possible; unidentifi¬ 
able species were assigned a code. Identifications were verified by R. 
Feitosa (Laboratorio de Sistematica e Biologia de Formigas, 
Universidade Federal do Parana). Vouchers of all species are available 
at the Universidade Federal Pernambuco in Recife and the Universidade 
Federal do Parana in Curitiba. 

2.4. Characterization of ant phylogenetic relationships 

At present, there is no complete, species-level ant phylogeny. We 
therefore used an approach that incorporated as much information as 
possible given our current understanding of ant relationships while 

simultaneously accounting for existing phylogenetic uncertainty. We 
began by using a backbone tree derived from a time-calibrated, genus- 
level phylogeny (Moreau and Bell, 2013) ; however, the phylogenetic re¬ 
lationships within Myrmicinae were taken from Ward et al. (2015). This 
phylogeny was then pruned to keep a single species per genus and thus 
generate a genus-level phylogeny. We subsequently used the list of spe¬ 
cies in our dataset (Appendix S3) to simulate 1000 species-level phylog- 
enies. Species relationships within genera were obtained from a Yule 
(pure-birth) process using the'ee function in the 
phytools package (Revell, 2012) in R (R Development Core Team, 
2016). However, given that one genus ( Mycetophylax ) was missing, 
we randomly added this lineage to our genus-level tree as a sister 
genus of Kalathomyrmex (Klingenberg and Brandao, 2009) prior to the 
addition of species in each iteration. The entire process was repeated 
1000 times to account for phylogenetic uncertainty in later analyses. 

2.5. Characterization of ant functional traits 

We quantified functional diversity using a suite of morphological 
traits that reflect body size, foraging capacity, foraging period, and re¬ 
source acquisition mode (Bihn et al., 2010; Parr et al., 201 7). These traits 
serve as proxies for the impact a species might have on ecosystem pro¬ 
cesses related to resource use. By focusing on morphological traits, we 
obviously ignored some aspects of the ants' ecology; however, direct 
links between morphological traits and functional roles have been ob¬ 
served in ants (Table 1 ). For each ant species, we determined body 
size (Weber's length), relative eye length, relative scape length, relative 
mandible length, relative clypeus length, and relative leg length 
(Table 1). We standardized all trait measurements (except Weber’s 
length) by dividing each by the Weber's length to limit correlations 
with body size. The trait values were then log-transformed to achieve 
normality. We measured approximately six randomly selected workers, 
and the mean measurements were used as the species-specific values 
for monomorphic and polymorphic species. In species with distinct 
minor and major worker castes, only minor workers were employed. 
In total, 503 workers representing 69 ground-foraging ant species 
were measured (number of individuals measured per species: mean 
± SE = 7.4 ± 0.16, median = 6.4, min = 1 and max = 70). 

2.6. Estimating taxonomic, phylogenetic and functional diversity 

For each plot, taxonomic diversity was characterized as species rich¬ 
ness (S, the number of ant species in each plot) and species diversity 
(Shannon diversity index, H, which accounts for both species richness 
and evenness). 

We measured five complementary metrics of phylogenetic diversity 
and functional diversity (Swenson, 2014). Phylogenetic diversity was 

Table 1 

Morphometric traits of workers used to characterize the functional diversity of Caatinga ant communities (Catimbau National Park, Pernambuco state, NE Brazil). 

Trait Functional significance 

Reference Definition 

Body size 
Relative eye 

Relative scape 
Relative leg 

Strongly correlated with many physiological, ecological, and life-history 
traits, including resource use 

Likely correlated with main foraging period (day vs. night) 

Possibly correlated with ability to receive chemosensory information. Ants 
with long scapes may be more sensitive to pheromone trails. 

Possible indicator of predatory lifestyle and thus types of resources consumed 

Correlated with sucking ability and liquid-feeding behavior 

Possibly correlated with resource acquisition mode and foraging efficiency, as 
well as with the ability to cope with the foraging surface temperature 

Kaspari and Weiser, 
1999, Bihn et al., 

Bihn et al., 2010 

Weiser and Kaspari, 

Weiser and Kaspari, 

Davidson et al., 2004 

Kaspari and Weiser, 
1999, Bihn et al., 

Maximum longitudinal length from the most anterior 
part of the clypeus to the occipital margin, in full face 

Ratio of eye length to mesosoma length 
Ratio of scape length to mesosoma length 
Ratio of mandible length to mesosoma length 

Ratio of clypeus length to mesosoma length 

Ratio of leg length (combined length of femur and tibia) 
to mesosoma length 

X.Aman et al. / Science of the Total Environment 631-632 (2018) 429-438 


estimated using the following indices (Table 2): (a) Faith's phylogenetic 
diversity (Faith's PD); (b) mean pairwise distance (MPD); (c) mean 
nearest-taxon distance (MNTD); (d) the net relatedness index (NRI); 
and (e) the nearest taxon index (NTI). Faith's PD is widely used in con¬ 
servation research (Forest et al., 2007; Morion et al., 2011) and, here, 
was the total branch length (divergence time) of the phylogenetic tree 
linking all the species represented in the community (Faith, 1992). 
MPD was the mean distance (in millions of years) between two ran¬ 
domly selected individuals within a specific plot (considering conspe- 
cifics), while MNTD was the mean distance separating each individual 
in the community from its closest heterospecific relative (Webb et al., 
2002). Thus, when MNTD is more strongly correlated with environmen¬ 
tal gradients than is MPD, it indicates that the environment has a stron¬ 
ger effect on terminal than basal community phylogenetic composition. 
Since both metrics might depend on species richness, we also measured 
the standardized effect size (SES) for MPD and MNTD (i.e., NRI and NTI, 
respectively) by comparing observed phylogenetic relatedness to ex¬ 
pected phylogenetic relatedness in null communities generated at ran¬ 
dom. Random communities were generated by randomizing the 
community data matrix using the independent swap algorithm 1000 
times. Then, we computed the SES of MPD and the SES of MNTD by tak¬ 
ing the difference between the mean phylogenetic distances in the ob¬ 
served communities versus in the null communities, standardized by 
the standard deviation of the phylogenetic distances in the null data 
(SES = (mean ob s — mean nu ii)/sd n uii) (Webb et al., 2002; Swenson, 
2014). We then multiplied the SES of MPD and the SES of MNTD by — 
1, obtaining NRI and NTI, respectively. The NRI and NTI values indicated 
whether taxa in the community were more closely related (positive 
values) or less closely related (negative values) than expected by 
chance. However, the two indices differ in phylogenetic scale. NRI re¬ 
flects information about whole phytogenies, while NTI reflects informa¬ 
tion about branch tips. 

The phylogenetic indices were calculated using the pd. (PD), ses.mpd 
(MPD and NRI), and ses.mntd (MNTD and NTI) functions in the picante 
package in R. For each index, 1000 phylogenetic trees were simulated, 
and the mean value was retained for use in further analyses. 

To characterize functional diversity, we calculated Petchey and 
Gaston's FD (hereafter PG-FD; Petchey and Gaston, 2002), as well as 
the functional equivalents of MPD, MNTD, NRI, and NTI (hereafter, FD- 
MPD, FD-MNTD, FD-NRI, and FD-NTI) (Table 2). PG-FD is the total 
branch length of the functional dendrogram that results when species 

are clustered in trait space (Petchey and Gaston, 2002). Here, the func¬ 
tional dendrogram was created by generating a Euclidean distance ma¬ 
trix from the z-standardized trait values of the different species and by 
clustering the species using the unweighted pair group method with ar¬ 
ithmetic mean (UPGMA). PG-FD was calculated using the alpha function 
in the BAT package in R. FD-MPD, FD-NRI, FD-MNTD, and FD-NTI were 
calculated as described above, except that trait-based Euclidean dis¬ 
tances rather than phylogenetic distances (i.e., divergence times) were 
used. Note that while Faith's PD and PG-FD are not abundance- 
weighted indices, all the remaining indices are abundance-weighted, 
i.e. they reflect trends in both abundance and evenness. 

2.7. Statistical analyses 

We used general linear models (GLMs) with a Gaussian distribution 
error and “identity" link to analyze the relationships between the taxo¬ 
nomic, phylogenetic and functional diversity indices and the distur¬ 
bance and aridity gradients. We used a separate model for each 
response variable (S, H, Faith's PD, MPD, MNTD, NRI, NTI, PG-FD, FD- 
MPD, FD-MNTD, FD-NRI, and FD-NTI); the explanatory variables were 
climatic water deficit, the LG index, the WE index, and the MU index. In¬ 
teractions between the disturbance indices and water deficit were also 
included. We employed Akaike's information criterion with a correction 
for finite sample sizes (AICc) to select the best-supported models; this 
approach reduces the problems associated with multiple testing, co¬ 
linearity of explanatory variables, and small sample sizes (Burnham 
and Anderson, 2002). All the initial models were full models. The best- 
supported models were selected based on their AICc weights, which re¬ 
veal the relative likelihood of a given model—based on the data and the 
fit—scaled to one; thus, models with a delta (AICc difference) of <2 were 
selected (Burnham and Anderson, 2002). The relevant variables were 
those that were retained in the best-supported models (except, obvi¬ 
ously, when the best-supported model consisted only of the intercept). 
Model selection was carried out using the dredge function in the MuMIn 
package in R. 

2.8. Phylogenetic signals in ant functional traits 

To test for phylogenetic signals (i.e., the degree of phylogenetic con¬ 
straint in species resemblance) in the six morphological traits, we com¬ 
puted Pagel's \ (Pagel, 1999). This index compares the observed 

Table 2 

List of selected phylogenetic and functional diversity metrics, their description and references, and their range of values in this study. 

Diversity metric 




Phylogenetic diversity 

Faith's PD 

Total branch length (divergence time) of the phylogenetic tree linking all the species represented in the 
community. Not abundance-weighted. 

Faith, 1992 


Mean pairwise distance 

Mean phylogenetic distance (in millions of years) between two randomly selected individuals within a 
specific plot (considering conspecifics). Abundance-weighted. 

Webb et al., 2002, 
Swenson, 2014 


Mean nearest-taxon 
distance (MNTD) 

Mean phylogenetic distance (in millions of years) separating each individual in the community from its 
closest heterospecific relative. Abundance-weighted. 

Webb et al., 2002, 
Swenson, 2014 


Net relatedness (NRI) 

Quantifies the structure of a sample phylogeny derived from the mean phylogenetic distance, consequently 
capturing the degree of clustering of the phylogeny from root to terminal leaves. Abundance-weighted. 

Webb et al., 2002, 
Swenson, 2014 


Nearest taxon index 

Quantifies the terminal structure of the sample phylogeny, hence only captures the clustering of the terminal 
nodes in the tree. Abundance-weighted. 

Webb et al., 2002, 
Swenson, 2014 


Functional diversity 

Petchey and Gaston's FD 

Total branch length of the functional dendrogram that results when species are clustered in trait space. Not 

Petchey and Gaston, 


FD-MPD (MPD with 
functional data) 

Mean functional distance between two randomly selected individuals within a specific plot (considering 
conspecifics). Abundance weighted. 

Webb et al., 2002, 
Swenson, 2014 


functional data) 

Mean functional distance separating each individual in the community from its closest co-occurring relative. 
Abundance weighted. 

Webb et al., 2002, 
Swenson, 2014 


FD-NRI (NRI with 
functional data) 

Quantifies the structure of a sample functional dendrogram derived from the mean functional distance, 
consequently capturing the degree of clustering of the functional dendrogram from root to terminal leaves. 
Abundance weighted. 

Webb et al., 2002, 
Swenson, 2014 


FD-NTI (NTI with 
functional data) 

Quantifies the terminal structure of the sample functional dendrogram, hence only captures the clustering of 
the terminal nodes in the functional dendrogram. Abundance weighted. 

Webb et al., 2002, 
Swenson, 2014 



X. A man et al. / Science of the Total Environment 631 -632 (2018) 429-438 

distribution of traits with the expected distribution of traits based on a 
Brownian motion model of evolution. Values of 0 and 1 indicate the ab¬ 
sence and presence, respectively, of a phylogenetic signal under such 
conditions. We computed \ values for each trait and for each of the 
1000 simulated trees using the phylosig function in the phytools pack¬ 
age (Revell, 2012) in R. To conservatively test signal significance, we 
used a likelihood ratio test based on the minimum values to estimate 
the probability that the observed \ differed from the null \ value of 0. 
It is important to note that, given that the simulated tree might lead to 
an underestimation of phylogenetic signal, our estimates of lambda 
are probably conservative. 

3. Results 

3.1. Ant communities 

Our traps captured representatives of 71 ant species belonging to 23 
genera and 7 subfamilies (Appendix S4). Myrmicinae was by far the 
most species-rich subfamily (39 species), followed by Dolichoderinae 
and Formicinae (12 and 9 species, respectively). The most species-rich 
genus was Pheidole (14 species), followed by Dorymyrmex (9 species) 
and Solenopsis (8 species); Camponotus, Cephalotes, and Pseudomyrmex 
were represented by 6 species each. The most frequently occurring spe¬ 
cies were Ectatomma muticum (61.1% of all traps), Solenopsis virulens 
(44.8% of all traps), and Dinoponera quadriceps (41.2% of all traps). 
Dinoponera quadriceps occurred on all 20 plots, while Ectatomma 
muticum and Solenopsis virulens occurred on 19 plots. Nineteen and 18 
species occurred on one and two plots, respectively. We found between 
15 and 26 species per plot. Species richness and species diversity were 
not modulated by the aridity and disturbance gradients (Appendix S5). 

3.2. Phylogenetic and functional diversity along disturbance and aridity 

The best-supported models retained climatic water deficit (called 
aridity hereafter), the LG index, and the MU index, but not the WE 
index (Table 3, Appendix S5). Faith's PD decreased along the LG gradient 
(Table 3, Fig. 2a, R 2 = 0.31). The NRI decreased along the aridity, LG 
(Fig. 2b, R 2 = 0.02), and MU gradients (Table 3): communities 
transitioned from being more closely related to less closely related. 
This result indicates that ants coexisting under the same aridity or dis¬ 
turbance conditions are non-randomly phylogenetically clustered. Fur¬ 
thermore, the NRI was influenced by an interaction between aridity 
and MU (Fig. 3a, R 2 = 0.64). More specifically, the decrease in the NRI 
along the aridity gradient was much stronger in areas with lower MU, 
and the negative relationship between the NRI and MU became positive 
when aridity was greater. None of the variables considered here helped 
explain MPD, MNTD, or NTI values. 

PG-FD decreased as aridity and LG increased (Table 3, Fig. 2c,d, R 2 = 
0.11 and R 2 = 0.22, respectively). FD-MPD decreased along the aridity 
gradient (Table 3, Fig. 2e, R 2 = 0.16). The FD-NRI was not associated 
with any of the explanatory variables (Table 3). FD-MNTD generally de¬ 
creased along the aridity, LG, and MU gradients (Table 3), but there 
were also complex aridity-by-disturbance interactions. More specifi¬ 
cally, when aridity was lower, FD-MNTD increased faster as MU de¬ 
clined, and FD-MNTD only decreased along the aridity gradient when 
MU was lower (Fig. 3b, R 2 = 0.20). At the same time, FD-MNTD de¬ 
creased and increased with LG at low and high levels of aridity, respec¬ 
tively. It also increased and decreased with aridity at low and high levels 
of LG (Fig. 3c, R 2 = 0.22), respectively. The FD-NTI generally increased 
along the aridity and MU gradients (Table 3), meaning that communi¬ 
ties transitioned from being functionally overdispersed to functionally 
clustered. However, there was once again an interaction between arid¬ 
ity and disturbance. The positive relationship between disturbance and 
the FD-NTI became negative in the most arid areas, and the positive re¬ 
lationship between aridity and the FD-NTI became negative in the most 
disturbed areas (Table 3, Fig. 3d, R 2 = 0.56). 

3.3. Phylogenetic signals in ant functional traits 

Strong, significant (p < 0.0001) phylogenetic signals were present in 
all traits, indicating that they were phylogenetically conserved. The only 
exception was relative clypeus length (p = 0.334) (Appendix S6). 

4. Discussion 

Our results highlight how ant phylogenetic diversity and functional 
diversity changed, despite unmodified taxonomic diversity, along CAD 
and aridity gradients in the Caatinga, a type of SDTF found exclusively 
in Brazil. In the 21st century, increased CAD and aridity are the main 
threats faced by biota in dry tropical regions (Dirzo et al., 2011 ). Overall, 
we found that functional diversity and, to a lesser extent, phylogenetic 
diversity, decreased with increasing aridity. Both types of diversity 
also decreased as anthropogenic disturbance increased; livestock graz¬ 
ing and miscellaneous resource use had an influence while wood ex¬ 
traction did not. Remarkably, we also found that anthropogenic 
disturbance and aridity interacted to shape biodiversity responses. We 
discovered clear evidence that the main mechanism involved was hab¬ 
itat filtering. 

We found support for our first hypothesis: in general, increased 
levels of disturbance and aridity decreased ant phylogenetic diversity 
and functional diversity. However, these relationships were somewhat 
complex because different patterns were observed in different diversity 
metrics. Also, environmental gradients interacted in different ways. The 
patterns were especially complex in the case of phylogenetic diversity. 
For example, increased levels of livestock grazing reduced the amount 

Table 3 

Variables retained in the best-supported models analyzing how ant phylogenetic diversity and functional diversity change along aridity (climatic water deficit) and anthropogenic distur¬ 
bance gradients in the Caatinga (Catimbau National Park, Pernambuco, NE Brazil). The positive and negative signs depict the direction of the relationship; an “X” signals that an interaction 
was present. The absence of any sign means there was no relationship between the response variable and the explanatory variable. 

Response variable Water deficit (WD) 

Miscellaneous resource use (MU) 

Livestock grazing (LG) Wood extraction (WE) MU x WD 

Interactions WE x WD 

Phylogenetic diversity 







Functional diversity 










X. Arrian et al./Science of the Total Environment 631-632 (2018) 429-438 


Fig. 2. Linear relationships between ant phylogenetic diversity and functional diversity metrics and the explanatory variables retained in the best-supported models: (a) Faith's PD and 
livestock grazing; (b) NR1 and livestock grazing; (c) PG-FD and climatic water deficit; (d) PG-FD and livestock grazing; and (e) FD-MPD and climatic water deficit. 

of evolutionary history shared by ants in the community (decrease in 
Faith's PD), but increased (although weekly) the non-random mean 
phylogenetic distance between ant species (decrease in the NRI). 
These results suggest that livestock grazing might act as an important 
filter, removing phylogenetically distant groups. Indeed, previous stud¬ 
ies in the same area (Arcoverde et al., unpublished data; Oliveira et al„ 
unpublished data) and in other Caatinga areas (Oliveira et al., 2017) 
found that populations of Dinoponera quadriceps (subfamily Ponerinae) 
decreased as disturbance increased. Alternatively, increased livestock 
grazing might non-randomly remove species from the most species- 
rich lineages or add species from relatively distant lineages (although 
not as distant as Ponerinae). Meanwhile, aridity and miscellaneous re¬ 
source use appeared to have positive effects on phylogenetic diversity: 
as both pressures increased, the NRI decreased, which meant there 
was more phylogenetic overdispersion. Thus, these environmental gra¬ 
dients are important abiotic filters that may structure ant communities 
by removing species from different lineages in the most humid or least 
disturbed areas. Conversely, they may supply species from different lin¬ 
eages when disturbance or aridity increase. However, in the most arid 
areas, the NRI increased as disturbance increased, meaning that ant 
communities in areas experiencing high levels of both disturbance and 
aridity were more phylogenetically clustered. These results provide 
support for our second hypothesis—that disturbance and aridity can in¬ 
tensify each other's negative impacts on diversity. They also draw atten¬ 
tion to the synergistic effects that anthropogenic disturbance and 

climate change may have on biodiversity (Hirota et al., 2011; Ponce- 
Reyes et al., 2013; Gibb et al., 2015a; Frishkoff et al., 2016). 

Disturbance and aridity play strong and consistent roles in reducing 
ant functional diversity, as evidenced by the functional responses that 
we observed along the disturbance and aridity gradients. Interestingly, 
these patterns cannot be explained by chance alone, since the effects 
were significant for the FD-NTI. Thus, aridity and certain forms of an¬ 
thropogenic disturbance (i.e., livestock grazing and miscellaneous re¬ 
source use) act as important abiotic filters of ant functional diversity 
in the Caatinga. More specifically, greater aridity and disturbance pro¬ 
duced a major drop in the diversity of functions related to food acquisi¬ 
tion and foraging habits. Interestingly, however, they did so in an 
interactive way. While miscellaneous resource use had very weak ef¬ 
fects on FD-MNTD and the FD-NTI in the most arid areas of the park, 
its effects were quite strong in the less arid areas, which contradicts 
our second hypothesis. Given that functional diversity is already 
impoverished in highly arid areas, there might not be much room for 
further loss. In the less arid and more diverse areas of the park, however, 
disturbance can exert a much stronger influence. The effects of livestock 
grazing on FD-MNTD were also modulated by aridity: positive effects 
were found in the most arid areas, while negative effects were found 
in the least arid areas. Taken together, these findings present new evi¬ 
dence that anthropogenic disturbance and climate change, acting in 
tandem, can have complex effects on biodiversity (Travis, 2003; 
Ponce-Reyes et al., 2013; Garcia-Valdes et al., 2015; Rito et al., 2017). 


X. Arnan et al. / Science of the Total Environment 631-632 (2018) 429-438 




















n i i r 
10 20 30 40 50 60 

Miscellaneous resource use 


























1 20 

10 20 30 40 50 60 

Miscellaneous resource use 

10 20 30 40 50 

Livestock grazing 







£ 1000 




900 - 

800 - 

£ 700 - 








- 1.0 


20 30 40 50 60 

Miscellaneous resource use 

Fig. 3. Contour plots showing model results for the interactive effects of (a) climatic water deficit and miscellaneous resource use on the NRI; (b) climatic water deficit and miscellaneous 
resource use on FD-MNTD; (c) climatic water deficit and livestock grazing on FD-MNTD; and (d) climatic water deficit and miscellaneous resource use on the FD-NTI. 

Our results do not match those of previous studies that analyzed the 
effects of aridity or precipitation (here, precipitation and water deficit 
were highly correlated: r = 0.98) on ant phylogenetic diversity and 
functional diversity. For instance, Arnan et al. (2015) examined geo¬ 
graphical gradients in central and western Europe, Machac and collabo¬ 
rators (2011) examined three altitudinal gradients in the USA, and 
Smith (2015) examined several altitudinal gradients worldwide. All 
three found that mean precipitation had a weak to non-existent influ¬ 
ence on ant phylogenetic diversity. Instead, patterns of phylogenetic di¬ 
versity were primarily driven by mean temperature. In the case of 
functional diversity, no effects of mean precipitation were found along 
elevational gradients in northwestern Patagonia (Argentina) 
(Werenkraut et al., 2015) or across different vegetation types in the cen¬ 
tral North Kimberley region of Australia's seasonal tropics (Cross et al., 
2016). In western and central European ant communities, functional di¬ 
versity was found to be shaped by mean precipitation but was lowest in 
the wettest areas (Arnan et al., 2015). This pattern was attributed to re¬ 
laxed local competition in areas with high levels of primary productivity 
and resource availability (Pavoine and Bonsall, 2011). Although mean 
precipitation does not generally seem to be an important driver of ant 
community structure at large spatial scales (Dunn et al., 2009; 
Andersen et al., 2015), our results suggest that water availability 
might significantly shape community structure in some regions or bi- 
omes, such as semi-arid areas where water is scarce (Parr et al., 2004). 

In contrast, our discovery that disturbance reduced ant phylogenetic 
diversity and functional diversity corresponds to what has been found 
in many other studies (e.g., Bihn et al., 2010; Arnan et al., 2013, 2015; 
Liu et al., 2016). Interestingly, these studies explored the effects of 
acute disturbances—mainly land-use changes—which significantly 
drive down biodiversity (Sala et al., 2000). Other studies have also re¬ 
ported that intense grazing has negative effects on ant functional diver¬ 
sity in semi-arid areas (e.g., Chillo et al., 2017; Oliveira et al., 2017), but 
effects on invertebrate phylogenetic diversity have never before been 
reported. Our study is the first to highlight that small but constant bio¬ 
mass removal (i.e„ chronic disturbance) can have similar deleterious ef¬ 
fects on phylogenetic diversity and functional diversity as large, sudden 

biomass removal (i.e., acute disturbance). These findings are notewor¬ 
thy given concerns about the functional consequences of current biodi¬ 
versity losses (Bellard et al., 2012) that result from acute and chronic 
disturbances alike (Barlow et al., 2016). 

We found strong phylogenetic signals in all the functional traits we 
measured (with the exception of relative clypeus length), indicating 
that more closely related ant species share more similar functional 
traits. These results agree with those of other studies that found signif¬ 
icant and, frequently, strong phylogenetic signals in ant morphological 
traits (Machac et al., 2011; Donoso, 2014; Arnan et al., 2017). They 
also suggest that the functional morphological traits of ant species in 
the Caatinga are evolutionarily conserved, and consequently, a strong 
correlation between PD and FD patterns is to be expected (Webb 
et al., 2002; Cavender-Bares et al., 2009). However, we found that the 
phylogenetic diversity and functional diversity indices did not respond 
in the same way to disturbance and aridity gradients (Table 1 ). Other 
studies across very different taxonomic groups have observed similar 
mismatches, even when strong phylogenetic signals exist 
(e.g., Purschke et al., 2013 for plants; Devictor et al., 2010 for birds; 
Safi et al., 2011 for mammals; Arnan et al., 2015, 2017 and Liu et al., 
2016 for ants). Our results thus lend further support to the idea that 
the environment may strongly condition covariation between different 
diversity components via differential filtering (Safi et al., 2011; Arnan 
et al., 2015, 2017). 

From a conservation perspective, our results echo recent work dem¬ 
onstrating that CADs and aridity are immediate threats to biodiversity in 
SDTFs (e.g., Ribeiro et al., 2015, 2016; Ribeiro-Neto et al., 2016; Oliveira 
et al., 2017; Rito et al., 2017). However, this study is the first to describe 
these detrimental effects in an animal taxon in a species-rich SDTF such 
as the Caatinga. Ants provide a variety of key ecosystem services and 
disservices in most terrestrial ecosystems (Del Del Toro et al., 2012); 
these services are largely related to species dietaiy ecology. It is thus 
likely that a decline in both ant phylogenetic diversity and functional di¬ 
versity (but especially in the latter) could have downstream effects on 
ecosystem processes, plant populations, and non-ant insect 

X. Arrian et al. / Science of the Total Environment 631-632 (2018) 429-438 


5. Conclusions 

We need a clear understanding of the main factors threatening bio¬ 
diversity in SDTFs. Conducted in Catimbau National Park, our study pro¬ 
vides evidence that ant phylogenetic diversity and functional diversity 
can be deterministically impoverished due to increased anthropogenic 
disturbance and aridity, even if absolute levels of ant species diversity 
remain unchanged. More alarmingly, aridity can intensify the negative 
effects of disturbance. Taken together, our results underscore concerns 
about what will happen under future global change scenarios in neo¬ 
tropical semi-arid regions. These regions are already facing major de¬ 
clines in precipitation and constant-to-increasing anthropogenic 
exploitation of forest resources (Magrin et al., 2014). However, we ob¬ 
served that anthropogenic disturbance had most negative impacts in 
the wettest areas, which contain the highest levels of phylogenetic di¬ 
versity and functional diversity. Caatinga conservation policies must 
thus give special priority to the wettest areas, where biodiversity loss 
could be the highest. Finally, our findings strongly suggest that studies 
in the Caatinga must address functional and phylogenetic diversity in 
addition to species richness if they wish to uncover how ant communi¬ 
ties are reorganized after disturbance and how climate change modu¬ 
lates this process. 


We are very grateful to Rodrigo Feitosa for helping to identify the 
ants, to Davi Jamelli for providing Fig. 1, and to Jessica Pearce-Duvet 
for editing the manuscript's English. This study was funded by the Foun¬ 
dation for Science and Technology Support of the State of Pernambuco 
(FACEPE; APQ 06012.05/15, APQ 0738-2.05/12, and PRONEX 0138- 
2.05/14), the Brazilian National Council for Scientific and Technological 
Development (CNPq; PELD 403770/2012-2, Universal 470480/2013-0), 
and the Rufford Small Grants Foundation (RSG 17372-1). CNPq receives 
thanks from XA for his postdoctoral grants (PDS-167533/2013-4 and 
PDS-165623/2015-2), from GBA for her scholarship (236918/2012-5), 
and from IRL for her research grants (Produtividade 305611/2014-3). 

Appendices 1-6. Supplementary data 

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