perda de habitat e efeitos de borda na comunidade …...pâmela vanessa friedemann tavares perda de...

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



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



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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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*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.


Number of frugivorous

species

Null 0.0 1 0.399


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


Null 0.3 1 0.255

Deforestation percentage 1.1 2 0.168






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.


Number of insectivores

species




Null 1.3 1 0.127





Number of insectivores

individuals








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.

perda de habitat e efeitos de borda na comunidade …...pâmela vanessa friedemann tavares perda de...

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