Instituto Nacional de Pesquisas da Amazônia – INPA
Programa de Pós-Graduação em Biologia (Ecologia)
Perda de habitat e efeitos de borda na comunidade de aves de sub-bosque em florestas
de areia branca na Amazônia Central
Pâmela Vanessa Friedemann Tavares
Manaus, Amazonas
Abril, 2018
Pâmela Vanessa Friedemann Tavares
Perda de habitat e efeitos de borda na comunidade de aves de sub-bosque em florestas
de areia branca na Amazônia Central
Orientadora: Dra. Cintia Cornelius
Coorientadora: Dra. Marina Corrêa Côrtes
Manaus, Amazonas
Abril, 2018
Dissertação apresentada ao
Instituto Nacional de
Pesquisas da Amazônia
como parte dos requisitos
para obtenção do titulo de
Mestre em Biologia
(Ecologia)
Sinopse: analisamos como o desmatamento, a distância à borda e a
distância às estradas afetam o número de espécies, a abundância, a
composição e a β-diversidade de aves de sub-bosque em uma região de
campinarana.
Palavras-chave: desmatamento, distância à borda, turnover, campinarana
T231 Tavares, Pâmela Vanessa Friedemann
Perda de habitat e efeitos de borda na comunidade de aves de sub-
bosque em florestas de areia branca na Amazônia Central / Pâmela
Vanessa Friedemann Tavares. - Manaus: [s.n.], 2018.
30 f. : il. color.
Dissertação (Mestrado) - INPA, Manaus, 2018.
Orientadora : Cintia Cornelius
Coorientadora : Marina Corrêa Côrtes
Programa: Ecologia
1. Aves. 2. Desmatamento. I. Título.
CDD 598
iii
Agradecimentos
A minha orientadora e co-orientadora, Cintia Cornelius e Marina Côrtes, duas cientistas
incríveis que fizeram com que fosse possível eu executar o mestrado;
A Capes pela bolsa concedida;
A todos os voluntários e ao mateiro que me ajudaram muito em campo e sem os quais seria
impossível realizar as coletas de dados: Dirley, Yuri, Paula, Victor, Jadson, André, Gisiane,
Radleigh, Giselle, Sascha, Sarah, Pilar, Kumara, Lizzie, Hannah, Rémi, Laura, Fernando e
Gabriel;
Às proprietárias (os) das casas onde fiquei durante minhas saídas de campo, agradeço
imensamente por todo suporte e disposição! Obrigada dona Lucimar, Daniel, Narcilene, Dona
Laura, Seu Bernardo, Aline (e aos meninos!), Seu Calisto e Seu Carlos;
Ao Mario Conh-Haft, por resolver todas as minhas dúvidas de identificação de aves e fazer
comentários entusiasmados sobre os passarinhos;
Aos meus colegas da turma de Ecologia 2016 pela ajuda antes da aula de qualificação, nosso
primeiro “grande momento” no INPA e pelas boas risadas no início de tudo;
A todos os professores do INPA que me ensinaram muito sobre Ecologia e sobre a Amazônia,
esse bioma maravilhoso e indescritível;
A minha família, pelas boas risadas e lembranças nos jantares de final de ano;
A minha mãe, por ser meu maior exemplo de força e pelo apoio incondicional;
Ao Gabriel, meu companheiro em Manaus e no mundo afora, que se tornou essencial na
minha vida e que esteve ao meu lado em praticamente todos os momentos mais estressantes
durante o mestrado. Juntamente com a Tapioca!
iv
Perda de habitat e efeitos de borda na comunidade de aves de sub-bosque em florestas
de areia branca na Amazônia Central
Resumo
A perda de habitat e o efeito de borda estão entre as principais ações humanas
afetando a biodiversidade. A região amazônica tem perdido uma grande extensão de habitat
pelo desmatamento que ocorre principalmente pela prática de corte raso ou queimadas.
Impactos humanos, entretanto, têm sido estudados em florestas de terra firme (contínuas), mas
habitats naturalmente distribuídos em mancha e pobre em nutrientes podem responder
diferentemente à mudança da paisagem. Ecossistemas de areia branca cobrem uma pequena
porção da Amazônia, tem baixa resiliência após distúrbios e são pouco conhecidos. Aqui, nós
reportamos os efeitos do desmatamento, do efeito da borda florestal e das estradas na riqueza,
abundância, composição e β-diversidade de aves de sub-bosque em uma paisagem modificado
por ação humana de florestas de areia branca. Nós também avaliamos a resposta de aves
insetívoras e frugívoras separadamente. Usando redes de neblina, nós amostramos 14
paisagens (em sete dias não-consecutivos) e registramos 82 espécies e 703 indivíduos. Nós
encontramos o processo de turnover controlando composição da comunidade ao longo do
gradiente de distúrbio, indicando que espécies não são perdidas, mas substituídas pelas
paisagens, mas não encontramos evidência de compensação de densidade por causa da
chegada de espécies tolerantes em sítios com maior distúrbio. O desmatamento e a
proximidade da borda florestal tiveram um efeito negativo na abundância de aves que foi
carregada principalmente pelas espécies insetívoras. As aves frugívoras, entretanto, não foram
fortemente afetadas pelas métricas da paisagem. Mudanças na composição de espécies foram
controladas pelo desmatamento. Dada as peculiares caraterísticas das florestas de areia
branca, nós sugerimos que a quantidade mínima de cobertura deste habitat deveria fica acima
dos limiares comuns, para evitar o declínio drástico da população, que eventualmente pode
levar a extinção local de espécies.
Palavras-chave: campinarana, desmatamento, distância à borda, efeito da estrada, turnover de
espécies
v
Habitat loss and edge effects on the understory bird community of white-sand forests in
Central Amazonia
Abstract
Habitat loss and edge effects are the most common human-driven processes affecting
biodiversity. The Amazon has lost a large extension of habitat through deforestation, which
occurs mostly by clear-cutting and burning practices. Human impacts, however, have mostly
been studied in originally continuous terra firme forests. Naturally patchy and poor-nutrient
environments may respond differently to landscape change. White-sand ecosystems cover a
small proportion of the Amazon, have low resilience after disturbances, and are poorly
understood. Here we report on the effects of deforestation, edge effects and road effects on
richness, abundance, composition and β-diversity of understory bird communities in a white-
sand forest human-modified landscape. We also evaluated the response of insectivorous and
frugivorous feeding guilds separately. Using mist-nets, we sampled 14 landscapes (with seven
non-consecutive surveys per landscape) and recorded 82 species and 703 individuals. We
found a turn-over driven community composition along the disturbance gradient, indicating
that species are not lost, but replaced across landscapes, but found no evidence for density
compensation because of the arrival of generalist species into more disturbed sites.
Deforestation and distance to the edge had a negative effect on bird abundance which was
mostly driven by forest insectivore species. Frugivore birds, however, were not strongly
affected by landscape metrics. Changes in species composition were mostly driven by
deforestation. Given the peculiar characteristics of white-sand forests, we suggest that the
minimum amount of habitat cover should be above usual thresholds, to avoid drastic
population declines, which can eventually lead to local species extinction.
Keywords: campinarana, deforestation, edge distance, road effect, species turnover
vi
Sumário
Introdução...................................................................................................................1
Objetivos.....................................................................................................................4
Capítulo 1
Cover page........................................................................................................6
Funding..................................................................................................7
Acknowledgments...................................................................................8
Introduction…………………………………………………………………………..9
Material and methods........................................................................................10
Study area…………………………………………………………………….10
Landscape metrics……………………………………………………………11
Bird sampling…………………………………………………………………11
Data analysis………………………………………………………………….11
Results.............................................................................................................12
Number of species…………………………………………………………….12
Abundance of individuals…………………………………………………….13
Functional groups responses………………………………………………….13
Species composition and β-diversity………………………………………….13
Discussion.........................................................................................................14
Deforestation and bird community…………………………………………...14
Edge effect and bird community……………………………………………...14
Road and bird community…………………………………………………….15
References............................................................................................................16
Figures and tables……………………………………………………………………..20
Supplementary material..........................................................................................24
Conclusões.......................................................................................................................30
1
1. Introdução
Comunidades de animais são fortemente afetadas pelas características do
habitat e seus arredores, como a idade florestal (Thompson & Donnelly 2018),
desmatamento (Solar et al. 2016), fragmentação (Ahumada et al. 2011), proximidade
da borda florestal, e perda da conectividade estrutural (Laurance et al. 2002).
Geralmente, distúrbio da paisagem podem negativamente afetar diversidade de
espécies (Durães et al. 2013; Murphy & Romanuk 2014). Entretanto, não é incomum a
diversidade aumentar com distúrbios, em geral pela chegada da espécies tolerantes a
impactos (Menke et al. 2012; Rotholz & Mandelik 2013) e espécies exóticas
(Blackburn et al. 2004). Estes efeitos em comunidades de plantas (Powell et al. 2013)
e animais podem variar dependendo do grupo, região e histórico de distúrbio.
Estudos de modelagem têm feito predições sobre a biodiversidade
considerando mudanças climáticas e mudanças de usos da terra, e o resultado são
geralmente similares: a biodiversidade continuará em declínio no século XXI (Pereira
et al. 2010). A agricultura (Balmford et al. 2012) e a pecuária (Bowman et al. 2012)
são duas das principais causas da perda de habitat nos trópicos. São esperados grandes
números de extinções se o histórico de desmatamento for mantido (Strassburg et al.
2012). Para as aves, o efeito da perda de habitat tem sido amplamente documentado.
Especificamente, o desmatamento pode afetar negativamente espécies especialistas de
interior de florestas e positivamente espécies generalistas em diferentes escalas
(Carrara et al. 2015). Na Amazônia, o desmatamento é um distúrbio crescente
(Nepstad et al. 1999) e as práticas de corte raso e queimadas para o prepare do uso da
terra é comum, causando estragos severos (Barlow et al. 2016).
Outras perturbações que aumentam a presença de assentamentos humanos,
como a construção de estradas, também contribuem para o crescente desmatamento
em determinadas regiões (Pfaff et al. 2018). Há evidência de diversos efeitos
negativos de estradas sobre as comunidades de animais (Canaday 1997), devido aos
ruídos do tráfego (Shannon et al. 2014) e a colisão animal-veículo (Freitas et al. 2017).
Aves também podem ser afetadas pela proximidade da estrada, que tende a diminuir a
movimentação entre manchas de florestas (Laurance et al. 2004). A borda florestal
tem sido reportada afetando diferentemente as espécies; por exemplo, bordas podem
ter efeito positivo na riqueza de espécies quando algumas espécies são capazes de
2
tolerar as novas condições em áreas impactadas (Lacasella et al. 2015). Na Amazônia,
aves onívoras são s grupos mais comuns nas bordas florestais (Dario et al., 2017). Em
contraste, aves que são habitantes do interior de floresta são negativamente pelas
bordas (Schneider et al. 2015). Em geral, aves de sub-bosque parecem ser afetadas
negativamente por qualquer tipo de distúrbio, como o corte de árvores e fragmentação
do habitat (Johns 1991; Bierregaard & Lovejoy 1989).
A maioria dos estudos na Amazônia considerando desmatamento e mudança
na estrutura da comunidade tem sido conduzida em floresta de terra firme (Prist et al.
2012; Mokross et al. 2014). Ecossistemas de areia branca são distribuídos em um
padrão de manchas por toda bacia Amazônica e com alta taxa de endemismo, mas são
muito menos estudos do que florestas de terra firme (Adeney et al., 2016, Vicentini
2016). Estes habitats são caracterizados pelas florestas de areia brancas
(campinaranas) e um ambiente arbustivo (campinas), e ambos pobres em nutrientes e
com solos arenosos (Alencar 1990). Já que muitas das pequenas manchas de
campinaranas (<1 km²) estão imersas em matriz de floresta de terra firme, a
visualização em imagens de satélite é dificultada, e como resultado, este ecossistema
não está completamente mapeado, mas é estimado que este ecossistema cobre ~ 5% da
região Amazônica (Adeney et al. 2016). Apesar de sua relevância biológica, estes
ecossistemas estão sob pressões antropogênicas devido à extração de areia para
construções (de estrada principalmente) e desmatamento, principalmente por
queimadas, para o preparo da terra para uso humano (Nascimento 2009; Matos et al.
2009). Áreas queimadas em ecossistemas de solos arenosos raramente suportam o
recrescimento da mesma comunidade de planta ou a ocupação pela mesma
comunidade animal. Sendo assim, as alterações na cobertura da terra têm um impacto
sobre a composição de espécies e sobre o funcionamento do ecossistema em curto e
longo prazo (Adeney et al. 2016).
A estrutura da comunidade de aves de ecossistemas de areia branca é
determinada pelo contexto biogeográfico e pela estrutura da paisagem local (Borges et
al., 2016). Entretanto, amostragens sistemáticas da comunidade de aves neste tipo
florestal (campinaranas) inexistem e não se conhece a resposta de espécies de aves às
mudanças da paisagem induzidas por atividades humanas nestes ambientes. Para
compreender os efeitos da paisagem nas comunidades de aves é importante avaliar
3
como as diferentes guildas são afetadas pelos distúrbios. Por exemplo, aves insetívoras
tendem a ser muito afetadas pela perda de habitat e pelas bordas (Morante-Filho et al.
2015; Laurance et al. 2004). Em contraste, aves que se alimentam de frutos são
frequentemente mais tolerantes a distúrbios e a abundância e riqueza de espécies desta
guilda podem aumentar em áreas perturbadas (Saavedra et al. 2014). Além disso, as
aves frugívoras podem ser menos afetadas pela presença de estradas e assentamento
humanos do que outras guildas (Laurance et al. 2004; Lim & Sodhi 2004). Tendências
opostas têm sido registradas: a diversidade de aves frugívoras pode reduzir em áreas
impactadas (Vollstädt et al. 2017). Também, as aves frugívoras de grande porte
parecem ser mais afetadas pela perda de habitat em comparação aves frugívoras de
pequeno porte (Bregman et al. 2014; Bovo et al. 2017).
Aqui nosso objetivo é determinar como a perda de habitat (pelo
desmatamento), a borda da floresta e as estradas afetam a riqueza de espécies,
abundância, a composição de espécies e a β-diversidade da comunidade de aves de
sub-bosque em uma paisagem modificada por ação humana em florestas de solos
arenosos na Amazônia Central. Para cada métrica ecológica nós temos uma ou duas
previsões associadas: (1) Nós esperamos uma maior riqueza de espécies em sítios
menos desmatados e distantes da borda florestal e das estradas por causa do provável
desaparecimento de espécies sensíveis a distúrbios. Para abundância total (2) nós
esperamos maior número de indivíduos em sítios menos desmatados e distantes das
bordas e estradas, ou a abundância poderia permanecer similar nos sítios,
independente dos distúrbios, devido à compensação de densidade pelos indivíduos
tolerantes em sítios perturbados. (3) Nós esperamos que a estrutura da paisagem
exerça um efeito negativo mais forte sobre as aves insetívoras do que sobre as aves
generalistas frugívoras, que parecem ser mais tolerantes a distúrbios. (4) Nós
esperamos que a β-diversidade será controlada pelo processo de turnover mais do que
pela simples perda de espécies (como no processo de aninhamento), que poderia ser
resultado da substituição de espécies em áreas impactadas. (5) Nós esperamos que as
paisagens com cobertura florestal similar e configuração similar (distância à borda e à
estrada) apresentarão composição de espécies similar.
4
2. Objetivos
Geral
Avaliar como o desmatamento, a distância da borda florestal e a distância da estrada afetam a
estrutura da comunidade de aves de sub-bosque em uma região de campinarana.
Específicos
1. Analisar a influência das métricas da paisagem na riqueza total de aves de sub-bosque.
2. Avaliar como a abundância total de indivíduos é afetada pelas métricas da paisagem.
3. Verificar como a riqueza e a abundância de as aves insetívoras e frugívoras respondem
às métricas da paisagem.
4. Testar como a composição de espécies da comunidade é influenciada pelo
desmatamento, distância da borda e da estrada.
5. Analisar como as métricas da paisagem afetam a β-diversidade na comunidade de aves
de sub-bosque.
Capítulo 1
__________________________________________________
Friedemann, P., Côrtes, M. & Cornelius, C. 2018. Habitat loss and edge effects on the
understory bird community of white-sand forests in Central Amazonia. Manuscrito em
preparação para Biological Conservation.
6
Habitat loss and edge effects on the understory bird community of white-sand forests in
Central Amazonia
Pâmela Friedemann a*
, Marina Corrêa Côrtes b & Cintia Cornelius
c
a Departamento de Ecologia, Instituto Nacional de Pesquisas da Amazônia (INPA), 69067-375 Manaus,
Amazonas, Brazil
b Departamento de Ecologia, Instituto de Biociências, Universidade Estadual Paulista (UNESP), 13506-900 Rio
Claro, São Paulo, Brazil
c Instituto de Ciências Biológicas, Universidade Federal do Amazonas (UFAM), 69077-000 Manaus, Amazonas,
Brazil
* Corresponding author: [email protected]
7
Funding
This work was supported by the Brazilian Ministry of Education (CAPES, fellowship to
Pâmela Friedemann).
8
Acknowledgements
We are grateful to all field assistants: D. Rodrigues, Y. Martins, G. Lima, P. Gomez, V.
Gomes, J. Viana, A. Girão, R. Herschel, G. Owen, S. Turisini, P. Braga, S. Friedlander, K.
MacLeod, E. Forrester, R. Bigonneau, H. Wheatley, L. Kahane, F. Andriolli and G.
Stefanelli-Silva. We also thank the owners of the properties where we carried out the field
research. We thank Mario Cohn-Haft for ornithological assistance; A. Vicentini and P.
Campos for facilitating the field trips; P. Gomez, H. Wheatley, E. Forrester and G. Stefanelli-
Silva for reviewing the manuscript. P. Friedemann received a Capes fellowship.
9
1. Introduction
Animal communities are strongly affected by habitat characteristics and their
surroundings, such as forest age (Thompson & Donnelly 2018), deforestation (Solar et al.
2016), fragmentation (Ahumada et al. 2011), edge proximity, and loss of structural
connectivity (Laurance et al. 2002). In general, landscape disturbance can negatively affect
species diversity (Durães et al. 2013; Murphy & Romanuk 2014). However, it is not
uncommon for diversity to increase with disturbance, in general by the arrival of tolerant
(Menke et al. 2012; Rotholz & Mandelik 2013) and exotic species (Blackburn et al. 2004).
These effects on plant (Powell et al. 2013) and animal communities may vary depending on
the group, region and disturbance history.
Modelling studies have attempted to predict biodiversity considering climate change
and land-use practices, and results are generally similar: biodiversity will continue to decline
in the 21th
century (Pereira et al. 2010). Among land-use practices, agriculture (Balmford et
al. 2012) and cattle ranching (Bowman et al. 2012) are two of the main causes of habitat loss
in the tropics. Very high levels of species extinction are expected if historical levels of
deforestation are maintained (Strassburg et al. 2012). For birds, the effect of habitat loss has
been widely documented. Specifically, deforestation may negatively affect specialist species
and positively affect generalist species at different scales (Carrara et al. 2015). In the Amazon,
deforestation is a pervasive disturbance (Nepstad et al. 1999) and the practice of clear-cutting
and burning to prepare the land for pasture and crop is common, causing severe damage
(Barlow et al. 2016).
Other disturbances that increase the presence of human settlements, such as road-
building, also contribute to deforestation in these areas (Pfaff et al. 2018). There is widespread
evidence of the several negative effects of roads on animal communities (Canaday 1997), due
to traffic noise (Shannon et al. 2014) and animal-vehicle collision (Freitas et al. 2017). Birds
can also be affected by road proximity, which tends to decrease movement between patches
(Laurance et al. 2004). Edge effects have been reported to affect species differently; for
instance, edges can have a positive effect on species richness when some species are capable
of tolerating the new conditions in impacted areas (Lacasella et al. 2015). In the Amazon,
omnivorous birds are the most common group in forest edges (Dario et al., 2017). In contrast,
birds that are forest interior dwellers are negatively affected by edges (Schneider et al. 2015).
In general, understory birds seem to be negatively affected by any kind of disturbance, such as
tree logging and habitat fragmentation (Johns 1991; Bierregaard & Lovejoy 1989).
The majority of studies in the Amazon linking deforestation and changes in animal
diversity have been conducted in lowland terra firme forests (Prist et al. 2012; Mokross et al.
2014). White-sand ecosystems (WSE) are patchily distributed across de Amazon and present
high endemism, but are far less studied than terra firme forests (Adeney et al., 2016, Vicentini
2016). These habitats are characterized by white-sand forests (campinaranas) and white-sand
scrublands (campinas) and both occur on nutrient-poor, sandy soils (Alencar 1990). Since
many of the small patches of white-sand forests (<1 km²) are immersed in a matrix of non-
flooded terra firme forest, it is difficult to identify WSE in satellite images. As a result, WSE
are not fully mapped, but are estimated to cover ~5% of the Amazon (Adeney et al. 2016).
Despite their biological relevance, WSE are under anthropogenic pressure due to sand
extraction for construction (road-building specially) and deforestation, mostly by fire, to
prepare the land for human use (Nascimento 2009; Matos et al. 2009). Burned areas of WSE
seldom support regrowth of the same plant community or “re-occupation” by the same animal
10
community. Therefore land cover transformations impact species composition and ecosystem
functioning in the short and long term (Adeney et al. 2016).
The structure of bird communities of WSE is determined by both the biogeographic
context and by local landscape structure (Borges et al., 2016). However, systematic sampling
of bird communities in the forested type of WSE (i.e. campinaranas) are absent and nothing
is known concerning how these species respond to landscape changes induced by human
activities. To better understand the effects of landscape structure on bird communities it is
important to evaluate how different guilds are affected by disturbances. For example,
insectivorous birds tend to be very affected by habitat loss and edges (Morante-Filho et al.
2015; Laurance et al. 2004). In contrast, fruit-eating birds are often more tolerant to
disturbances and their abundance and richness can actually increase in impacted areas
(Saavedra et al. 2014). Also, frugivorous birds may be less affected by the presence of roads
and human settlements than other guilds (Laurance et al. 2004; Lim & Sodhi 2004). Opposite
trends have also been recorded: frugivore diversity may be reduced in impacted areas
(Vollstädt et al. 2017). Also, large frugivores seems to be more affected by habitat loss, than
small frugivores (Bregman et al. 2014; Bovo et al. 2017).
Here we aim to determine if habitat loss (through deforestation), forest edges and
roads influence species richness, abundance, composition and β-diversity of the understory
bird community across human-modified landscapes of white-sand forests in Central
Amazonia. For each ecological metric we have one or two associated predictions: (1) We
expect higher species richness in sites with low deforestation and farther from forest edges
and roads because of the disappearance of species more sensitive to disturbance. For total
abundance (2) we expect abundance will be higher in less deforested sites located farther from
edges and roads, or that abundance will remain similar across patches regardless of
disturbances due to density compensation by individuals of tolerant species in disturbed sites.
(3) We expect a stronger negative effect of landscape structure on insectivorous birds than on
generalist frugivores, which seem to be more tolerant to disturbances. (4) We expect that β-
diversity will be driven by a turnover process rather than by the simple loss of species (i.e.
nestedness), which would be the result of the replacement of species in impacted areas. (5)
We expect that landscapes sharing similar composition (forest cover) and configuration
(distance to forest edge and road) will present similar species composition.
2. Material and Methods
2.1 Study area
We conducted our study in a rural region ca. 30 km from the urban center of Manaus,
Amazonas, Brazil (02°51’28.0” S/060°13’28.5” W). The rural population in that region has
grown rapidly over the last 10 years, mostly facilitated by road construction and settlement
programs carried out by the National Institute of Colonization and Agricultural Land Reform
(Nascimento 2009). These settlements have intensified small scale agriculture (mostly palm
tree açaí, banana and cassava), road building and sand extraction activities. Also, there has
been more silting and pollution of river courses as a consequence of riparian forest removal
for the construction of water recreation areas (Nascimento 2009; Matos et al. 2009).
The region is dominated by patches of white-sand forests immersed in a matrix of
continuous terra firme forest, with the main differences between these types of forest being:
sandy soil in white-sand forest and clayey, with higher canopies and an overall higher
11
biodiversity in terra firme forests (Anderson et al. 1975; Alencar 1990). Small to medium size
forest clearings opened for human use are spread throughout the landscape along secondary
access roads. We delimited 14 sampling landscapes with 2-50% deforestation within a 500-m
radius and established a sampling site in each landscape (Fig. 1). Sampling sites have similar
vegetation structure and composition (i.e. campinarana forest), although some variability is
unavoidable, with canopy height varying from 12 to 20 meters.
2.2 Landscape metrics
We defined each landscape as a circle of 500-m radius centered in sampling sites,
which were transects for bird capture (Fig. 1). To select landscapes, we used a 2014, 30-m
resolution Landsat 8 image of the study region. To calculate the percentage of deforestation
we used the Semi-Automatic Classification plug-in in QGIS 2.18.2 software (QGIS
Development Team 2015) and performed a classification using the minimum distance
algorithm to produce a categorical map of two classes: forest and deforested area (i.e. forest
clearings, roads, houses, fish ponds and exposed soil; Fig. 1; Table 1). We chose a 500-m
radius to evaluate the direct effect of the anthropogenic activities at a relevant scale for the
target organisms. It is important to mention that very little deforestation occurs above 1 or 2
km from our sites. White-sand forest patches extend over a small area (2-3 km wide) with
small forest clearings along roads. To obtain the distance to the forest edge and roads in each
landscape we set up six points distant 50 m from each other along the bird-sampling transect
and used the mean distance from the six points to the nearest forest edge and/or road.
2.3 Bird sampling
In each sampling site we set up 20 mist nets along the transect. Mist nets were
operated from 06:00 am to 11:00 am, with a total of seven non-consecutive days of sampling
per site between April and October 2017. Fieldwork was carried out during the dry season to
avoid logistic constraints due to flooding in the rainy season. Sampling interval for each site
was between one and two months. Total sampling effort was 9083 net-hours (1 net open for 1
hour=1 net-hour).
Captured individuals were identified, banded and released. Individuals were identified
following the current classification of the bird species of South America (Remsen et al. 2018)
and individuals were marked with numbered metallic bands. We excluded all canopy-species
from the analysis, which constituted occasional captures (n=10, 1.16% of captured
individuals).
2.4 Data analysis
All statistical analyses were performed in R (R Core Team, 2017). To verify if
community sampling was representative, we constructed species accumulation curves of
captured birds in each of the 14 landscapes based on sampling days (Gotelli & Colwell 2001).
Curves were developed using the ‘vegan’ package (Oksanen et al. 2015). We excluded
recaptures from the abundance analysis, as well as hummingbirds because they were not
banded and, therefore, not individualized. Our response variables were number of species and
total bird abundance per landscape. To assess the influence of landscape metrics on the
response variables we performed generalized linear models (GLM) (Nelder & Wedderburn
1972) with Poisson distribution (Stigler 1982). Landscape metrics were centered and scaled
using the scale function. For each response variable we set up models with each isolated
predictor variable (e.g. deforestation, distance to edges and roads). We also constructed
additive and interaction models, in which deforestation was combined with distance to forest
12
edges or distance to roads, because we wanted to analyze the combined effect of deforestation
and these two types of disturbance. We also included a null model into the candidate model
set. Thus, for each response variable we contrasted eight models. We used Akaike's weights
(wAICc) and the delta AICc to rank models according to the candidate set, using the function
ICtab from the ‘bbmle’ package (Bolker, 2017). We considered that models with delta
AICc<2 were equally plausible and had similar strength of evidence (Burnham & Anderson
2002). We found no co-linearity between predictor variables (Pearson r<0.7 in all cases).
We analyzed separately the insectivore and frugivore assemblages to determine if
landscape metrics affected particular functional groups differently. We classified each
recorded species as omnivore, carnivore, frugivore, nectarivore or insectivore according to the
literature (del Hoyo et al., 2017; Wilman et al., 2014). Guilds were defined by evaluating the
proportion of one type of food in the diet. If one type of food (e.g. fruit) comprised more than
60% of the bird’s diet, the species was considered, for example, frugivore. We performed
GLMs with Poisson distribution to assess the influence of landscape metrics on species
richness and abundance of insectivores and frugivorous, two of the most dominant guilds in
WSE. We constructed candidate model sets for frugivorous and for insectivorous birds
following the same structure as for the total assemblage.
To evaluate (dis)similarities between communities we estimated the components of β-
diversity. If driven by turnover, species are replaced by others, whereas if nestedness controls
the process, disappearing species are not replaced. We estimated the components of β-
diversity by calculating the Sørensen dissimilarity index in the “beta.part” package (Baselga
et al. 2017). Finally, we used multiple regressions on distance matrices (MRM) to test if β-
diversity is explained by the geographic distance and dissimilarity in landscape metric values
between the sites, which was measured using the Mahalanobis distance method, in which of
predictors variables are made orthogonal, uncorrelated, as the response matrix (R
documentation, 2018).
We performed a redundancy analysis (RDA) to visualize and test if species
composition was associated with the landscape metrics. The RDA summarizes the variation of
response variables that can be explained by a set of explanatory variables (Legendre &
Legendre 1998). Since we considered the total number of captured species we used a
presence/absence matrix to perform the RDA analysis, which also included hummingbirds.
Landscape variables were standardized using the Hellinger transformation, which takes the
square root of values.
3. Results
3.1 Number of species
We recorded a total of 82 understory bird species belonging to 25 families. The most
representative families were Thamnophilidae (Antbirds, n=20), Furnariidae (Ovenbirds,
n=10) and Trochilidae (Hummingbirds, n=9). Fifteen species were captured only once in the
study. The mean number of species per landscape was 25 (min=17; max=31; Table 1).
Although none of the species accumulation curves reached a full asymptote, the number of
species of the majority of the landscapes was representative (Fig. S1 – Supplementary
material).
13
None of the models were considered more plausible than the null model according to
the information criteria (with dAICc<2; Table 2), indicating that the considered landscape
metrics have a weak effect on the species richness.
3.2 Abundance of individuals
We recorded 703 individuals, excluding hummingbirds (n=187) and recaptures
(n=141). The mean abundance per landscape was 50 (min=26; max=68; Table 1). The best
model was the additive model including deforestation and distance to edge forest as predictor
variables (Table 2). Models including each variable separately were also plausible (dAICc<2,
Fig. 2). In addition, the additive combination of deforestation and road distance was also
considered plausible, but was possibly driven by deforestation, since the estimated coefficient
of road was not statistically different from zero (Fig. 2). The cumulative weight of models
with dAICc<2 was 0.85, indicating stronger evidence than the null model (w=0.004; Table 2).
Deforestation and distance to edges had a negative and positive effect, respectively, on total
abundance (Fig. 2).
3.3 Functional group responses
We classified 11 species within the frugivorous and 53 species within the
insectivorous functional group (13% and 64% of the total assemblage, respectively). None of
the landscape metrics had any effect on the species richness or abundance of frugivorous
birds, since the null model was among the ones with dAICc<2 (Table S2). As for frugivores,
landscape metrics were not related to the number of insectivores species, since the null model
had a dAICc<2. But, similar to the pattern for total abundance, two best models explained the
abundance of insectivorous birds, one including deforestation (dAICc=0) and one including
an additive effect of deforestation and distance to edge (dAICc=0), which together had a 70%
probability of best explaining abundance relative to the competing models (Table S3, Fig. S2
- Supplementary material). The estimated coefficient of distance to the edge, however, had a
95% confidence interval that included zero (Fig. S2), indicating that deforestation may be the
main force affecting abundance of insectivorous.
3.4 β-diversity and species composition
The β-diversity (βsor) was mostly explained by turnover (βsim=~95%) and weakly
explained by nestedness (βnes=~5%), indicating a trend of species replacement between sites.
The turnover and nestedness values differed from the expected distribution based on the null
model, with βsim=0.76, C.I. 95%=(0.781, 0.782) and βnes=0.04, C.I. 95%=(0.027, 0.028). β-
diversity was not significantly associated with either the dissimilarity in the landscape metrics
or the geographic distance between sites (Table S4).
The only species common to all 14 sites was the hummingbird Phaethornis
superciliosus. We found that a significant portion of the variation in avian composition was
accounted for by landscape metrics (R²=0.26, p=0.049). A large proportion of the total
variance, however, was due to unconstrained variation (0.73) compared to constrained
variance (0.26). This means that just a small part of variation in the response variables is
accounted by the predictors and therefore represented in the RDA (Fig. 3).
4. Discussion
14
We found that deforestation was the main disturbance affecting abundance and species
composition, but not species richness, of the understory bird community in the white-sand
forest landscapes. The results were mostly driven by the effects on insectivores. Distance to
the forest edge also emerged as a factor positively associated with bird abundance. To our
knowledge, this is the first study to evaluate the effects of anthropogenic actions on birds in
the understudied and threatened Amazonian white-sand forests.
4.1 Deforestation and the bird community
Contrary to our expectation, deforestation was not related to total or functional groups
species richness. The extinction threshold hypothesis (Fahrig 2001; Fahrig 2002) posits that a
certain amount (threshold) of habitat loss is needed before extinctions rise drastically, but this
threshold depends on the organisms and the region, and may vary from 50% (Morante-Filho
et al. 2015) to 90% of habitat loss (Radford et al. 2005). Therefore, it is possible that
deforestation of white-sand forests in the region has not yet reached critical values for causing
significant species extirpation. Deforestation, however, did explain the variation in species
abundance and composition across landscapes. The bird community in these forests seems to
be affected when 30% is deforested at our scale. Although species have not been locally
extinct in disturbed sites yet, the decline in local abundance may lead to an ecological
extinction, that is, when the abundance reduces to a point in which it can no longer interact
significantly with other species, even if it is still present in the community (Estes et al. 1989).
Ecological extinction causing loss of ecological interactions, disrupts species functionality
and ecosystem services at a faster rate than species extinctions (Valiente-Banuet et al. 2015;
Säterberg et al. 2013). Therefore, population declines have to be considered in conservation of
ecosystems and species (Redford 1992).
Insectivorous birds were the more representative guild in this study, followed by
frugivores. These groups responded differently to landscape change in regards to species
abundance. As predicted, the abundance of insectivores was affected negatively by
deforestation. Insectivorous birds are commonly adapted to forest interior and considered
sensitive to disturbances at different scales (Stratford & Stouffer 2015). Frugivorous birds
may respond negatively to anthropogenic changes (Fontúrbel et al. 2015), but we found no
evidence for an effect on frugivorous. Generally the response is weaker when compared to
insectivores, since frugivores can find resources and persist in secondary forests or degraded
landscapes (Cleary et al. 2007; Barlow et al. 2007). Large frugivorous birds can be negatively
affected by disturbances (Bregman et al. 2014), but the generalist species, such as the ones in
the understory of white-sand forests, may not be affected by deforestation (Carlo & Morales
2016; Palacio et al. 2016).
4.3 Edge effects on bird communities
Edge distance had a positive effect on the abundance but not on the species richness.
The edge effect is suggested as one of the main drivers affecting biodiversity (Banks-Leite et
al. 2010). It is undeniable that edges affect diversity in contexts such as fragmented
landscapes (Laurance et al. 2002) or land-use changes (Lyra-Jorge et al. 2010). The prevalent
role of turnover along our disturbance gradient may explain the lack of edge distance effects
on species richness. It is possible that tolerant species occupy sites near edges compensating
for the species in the forest interior, keeping species richness somewhat constant with regards
to edges.
15
On the other hand, individual abundance was affected by edge distance and habitat
loss. This means that the higher the deforestation and the closer to the edge, the lower the
abundance of birds. Thus, we did not observe a density compensation pattern; abundance of
tolerant species was overall low and did not compensate for loss of individuals from forest
specialist species. Considering that white-sand ecosystems (WSE) are nutrient-poor habitats
(Adeney et al. 2016) we may presume the edge of this habitat is also nutrient-poor, with
scarce resources to support a high abundance (Borges 2013) even of tolerant species.
4.3 Road distance and the bird community
Roads design linear clearings that can act as barriers impeding movement of sensitive
species (Laurance et al. 2004), and changing the number of species and abundance of some
guilds (Canaday 1997). We expected to find an effect of roads on insectivorous birds rather
than on frugivores, since these seem more tolerant to road clearings and edges (Laurance et al.
2009). Roads in the landscape studied are, however, secondary access roads, unpaved and
with low traffic intensity, which might explain why they have the least prevalent effect in our
study system. Therefore, the most prevalent impact might come in fact from road edges rather
than roads per se, even for the insectivores.
5. Conclusions
Here we show evidence for changes in species composition and reduction on the
abundance in impacted landscapes (i.e. sites with more deforestation and closer to the forest
edge). In these landscapes the reduction in abundance indicates that birds are probably less
likely to interact in the same way with other individuals (conspecifics and heterospecifics) and
their environment, possibly leading to demographic erosion and ecological extinctions. The
negative effect on abundance was mostly driven by insectivores, which are a sensitive guild
only abundant in well conserved primary forests (Powell et al. 2015). Likewise, here we show
that this guild is also the most sensitive in these forests. Moreover, since WSE have low
resilience, even after vegetation regrowth, populations may not re-occupy in the same way as
they did before the disturbance (Burivalova et al. 2015; Adeney et al. 2016).
Some species were never recorded in sites with deforestation higher than 15%, such as
all three species within the Myrmotherula genus and Formicarius colma (specialized
insectivores), and Neopelma chrysocephalum (frugivore). None of the species that we
registered are highly threatened, but we captured two Near Threatened species, Hypocnemis
cantator and Epinecrophylla gutturalis, which are close to qualifying for a threatened
category in the near future (IUCN, 2017). The tolerant and generalist species we recorded
more frequently in disturbed sites were Cyanoloxia rothschildii, Coereba flaveola and
Momotus momota.
Generally, studies aim to identify the critical amount of forest cover ( Morante-Filho
et al. 2015), fragmentation (Taubert et al. 2018) and edge (Laurance et al. 2009) relevant for
the conservation of species considered as specialists. However, because responses to
disturbance are different among species and habitats, we need more information on the
minimum amount and quality of the habitat necessary to ensure persistence of sensitive
species. This is even more important for WSE, which are a fragile and naturally patchy
system surrounded by different types of habitats, with several peculiarities (Vicentini 2016;
Adeney et al. 2016). Here we suggest that these forests should be preserved above the
estimated thresholds disturbances in which individuals and populations begin to disappear.
16
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Amazon. Herpetological Bulletin, 126, pp.14–24.
Wilman, H. et al., 2014. EltonTraits 1.0: Species-level foraging attributes of the world’s birds
and mammals. Ecology, 95(7), p.2027.
20
Fig. 1 – On the left: Study area location and overview, with sites displayed as black empty circles with
respective identification. Opaque blue represents white-sand forests, green is terra firme forest, pink
are roads and deforested areas and dark blue is water. On the right: Each of the 14 study landscapes
categorized in forest (green) and non-forest (yellow). Orange lines are roads.
1 2 3
4 5 6
7 8 9
10 11 12
13 14
21
Table 1 – Percentage of deforestation, average distance to the edge and average distance to the
road in each of the 14 landscapes. Landscape Deforestation
(%)
Edge
distance (m)
Road
distance (m)
Total
species
richness
Total
individuals
abundance
1 2 94.6 420.5 28 55
2 5 161.8 399.2 30 66
3 8 176.2 1191.8 29 58
4 9 300.1 1156.6 31 61
5 14 163.8 382.7 29 68
6 15 227.5 361.5 27 50
7 18 88.9 114.2 18 29
8 22 134.4 136.9 17 26
9 27 178.7 477.2 29 62
10 32 98.6 104 23 53
11 35 145.7 588.7 27 50
12 38 87.7 194.3 25 47
13 45 107.8 163.6 22 44
14 50 104.4 564.2 18 34
Table 2 – Candidate models of the number of species and number of individuals as a function
of landscape metrics in the 14 landscapes. The dAICc is the difference in Akaike value with
respect to the best model among the set, df is the degrees of freedom of the models and weight
is the probability of best explaining the observed pattern giving all competing models. Models
with dAICc<2 are in bold.
Response variable Model dAICc df weight
Number of species Edge distance 0.0 2 0.260
Deforestation 0.5 2 0.207
Road distance 0.6 2 0.197
Null 1.5 1 0.121
Deforestation+Edge 2.0 3 0.097
Deforestation+Road 2.1 3 0.092
Deforestation*Edge 5.9 4 0.014
Deforestation*Road 6.1 4 0.012
Number of individuals Deforestation+Edge 0.0 3 0.262
Edge distance 0.6 2 0.197
Deforestation+Road 0.6 3 0.196
Deforestation 0.6 2 0.196
Road distance 2.5 2 0.076
Deforestation*Edge 3.8 4 0.039
Deforestation*Road 4.6 4 0.026
Null 8.0 1 0.004
22
Fig. 2 – Parameter estimates and their 95% confidence interval values for the four selected
models (with dAICc<2; Table 2) regarding the effects of landscape metrics on the total
abundance of birds. Negative effects of landscape metrics are indicated on the left side of the
graph and the positive effects on the right side.
23
Fig. 3 - Species composition of sites and the effect of landscape metrics based on a RDA
ordination. The green polygon indicates sites with less than 30% of deforestation; the orange
polygon indicates sites with deforestation higher than 30%.
24
Supplementary material
Table S1 – Number of species and abundance of insectivorous and frugivorous species
Landscape
Number of
insectivorous
species
Number of
insectivorous
individuals
Number of
frugivorous
species
Number of
frugivorous
individuals
1 17 43 6 12
2 20 47 4 17
3 17 34 5 21
4 21 47 4 12
5 20 54 3 14
6 16 36 5 11
7 13 24 2 4
8 8 18 4 6
9 18 40 8 24
10 11 25 4 23
11 16 36 3 8
12 16 33 3 11
13 13 32 3 8
14 11 21 2 9
Total 53 (64%)* 490 (69%)* 11 (13%)* 180 (25%)*
* Percentage of species or individuals of the total species/individuals
Site 1 Site 2
25
Site 3 Site 4
Site 5 Site 6
Site 7 Site 8
26
Fig. S1 – Rarefied species accumulation curve for each of the 14 assemblages, based on the sampling
days. Gray envelope represents the 95% CI. Text inset: total observed species richness of the
respective site (in bold) and richness estimated by the Chao index with the standard errors estimation
(±).
Site 9 Site 10
Site 11 Site 12
Site 13 Site 14
27
Table S3 - Candidate models of the number of frugivorous species and individuals as a function of
landscape metrics in the 14 landscapes. The dAICc is the difference in Akaike value with respect to the
best model among the set, df is the degrees of freedom and weight is the probability of best explaining
the observed pattern giving all competing models. Models with dAICc < 2 are in bold.
Response variable Model dAICc df weight
Number of frugivorous
species
Null 0.0 1 0.399
Deforestation 1.3 2 0.211
Edge distance 1.9 2 0.155 Road distance 2.4 2 0.122
Deforestation + Edge 4.4 3 0.044
Deforestation + Road 4.5 3 0.041
Deforestation * Edge 6.0 4 0.020
Deforestation * Road 8.6 4 0.005
Number of frugivorous
individuals
Road distance 0.0 2 0.289
Null 0.3 1 0.255
Deforestation percentage 1.1 2 0.168
Edge distance 1.4 2 0.144
Deforestation + Road 2.7 3 0.076
Deforestation + Edge 3.9 3 0.042
Deforestation * Road 5.7 4 0.017
Deforestation * Edge 7.5 4 0.006
Table S4 - Candidate models of the number of insectivorous species and individuals as a function of
landscape metrics in the 14 landscapes. The dAICc is the difference in Akaike value with respect to the
best model among the set, df is the degrees of freedom and weight is the probability of best explaining
the observed pattern giving all competing models. Models with dAICc < 2 are in bold.
Response variable Model dAICc df weight
Number of insectivores
species
Edge distance 0.0 2 0.247
Deforestation 0.1 2 0.235
Road distance 0.7 2 0.175
Null 1.3 1 0.127
Deforestation + Edge 1.8 3 0.100
Deforestation + Road 2.0 3 0.091
Deforestation * Edge 5.8 4 0.013
Deforestation * Road 5.9 4 0.013
Number of insectivores
individuals
Deforestation 0.0 2 0.355
Deforestation + Edge 0.0 3 0.348
Deforestation + Road 2.1 3 0.125
Edge distance 2.9 2 0.085
Deforestation * Edge 3.7 4 0.056
Deforestation * Road 5.8 4 0.019
Road distance 7.6 2 0.007
Null 10.5 1 0.001
28
Fig. S1 – Parameter estimates and their 95% confidence interval values for selected models (with
dAICc <2) considering the effects of landscape metrics on the abundance of insectivorous birds.
Negative effects of landscape metrics are indicated in the left side of the graph and the positive effects
on the right side.
29
Table S5 – Influence of landscape metrics on β-diversity total (βsor). None of the landscape metrics
significantly affected the β-diversity, since all R² were low and the p-values > 0.05
Landscape metrics R² p
βsor Deforestation 0.02 0.33
Edge distance 0.01 0.43
Road distance 0.02 0.38
Geographic distance 0.02 0.24
30
Conclusões
Nosso estudo indica um efeito da perda de habitat como um dos principais distúrbios
afetando a abundância de indivíduos e a composição, mas não tendo efeito consistente na
riqueza de espécies nas aves de sub-bosque na região de campinarana que amostramos. Estes
padrões foram principalmente controlados pelo efeito sobre as aves insetívoras, que foram as
mais representativas e tiveram uma abundância menor em sítios com maior distúrbio. A
distância da borda também emergiu como um fator associado positivamente com a
abundância de aves.
O efeito no número de indivíduos pode indicar uma possível futura extinção de
espécies, já que os indivíduos parecem estar menos presentes nestes sítios a ponto de não
serem mais registrados. O efeito negativo foi controlado pelos insetívoros, que são
considerados aves sensíveis e pode se mostrar abundante somente em florestas primárias bem
conservadas (Powell et al., 2015), e para algumas espécies de aves, mesmo após os distúrbios
cessarem, os indivíduos podem não se recuperar e ocupar novamente uma determinada área
da mesma maneira (Burivalova et al., 2015). O declínio local da abundância pode levar a
extinção ecológicas, que ocorre quando a abundância reduz a ponto de os indivíduos não
interagirem significativamente com outras espécies, mesmo que ainda presente na
comunidade (Estes et al. 1989). Extinção ecológica causa a perda de interações, da
funcionalidade das espécies e dos serviços ecossistêmicos a uma taxa mais alta do que a
extinção de espécies (Valiente-Banuet et al. 2015; Säterberg et al. 2013). Sendo assim, o
declínio de populações tem de ser considerados na conservação de ecossistemas e espécies
(Redford 1992).
Geralmente os estudos são capazes de identificar a quantidade de distúrbio para
cobertura florestal (Lindenmayer e Luck, 2005; Morante-Filho et al., 2015), fragmentação
(Banks-Leite et al., 2010; Taubert et al., 2018) e borda (Canaday, 1996; Laurance et al., 2004,
2009), em que as espécies consideradas especialistas não estão mais presentes. Entretanto, as
respostas a distúrbios podem ser diferentes entre espécies e habitats. E isto é ainda mais
importante para um ambiente como as campinaranas, que são frágeis e ocorrem naturalmente
em pequenas manchas inseridas em diferentes habitats, com muitas peculiaridades e espécies
endêmicas (Adeney et al., 2016; Vicentini, 2016). Aqui, sugerimos que estas florestas
deveriam ser preservadas acima dos níveis usuais de distúrbio em que a população diminui e
indivíduos e espécies começam a desaparecer.