Download habitat connectivity for carnivores in mediterranean landscapes

UNIVERSIDADE DE LISBOA
FACULDADE DE CIÊNCIAS
DEPARTAMENTO DE BIOLOGIA ANIMAL
HABITAT CONNECTIVITY FOR CARNIVORES
IN MEDITERRANEAN LANDSCAPES
IMPLICATIONS FOR CONSERVATION PLANNING
CLARA BENTES GRILO
DOUTORAMENTO EM BIOLOGIA
(Biologia da Conservação)
2009
UNIVERSIDADE DE LISBOA
FACULDADE DE CIÊNCIAS
DEPARTAMENTO DE BIOLOGIA ANIMAL
HABITAT CONNECTIVITY FOR CARNIVORES
IN MEDITERRANEAN LANDSCAPES
IMPLICATIONS FOR CONSERVATION PLANNING
CLARA BENTES GRILO
Tese co-orientada por:
Professora Doutora Margarida Santos-Reis
Professor Doutor John A. Bissonette
DOUTORAMENTO EM BIOLOGIA
(Biologia da Conservação)
2009
This research was funded by Fundação para a Ciência e Tecnologia
(Project
POCTI/MGS/47435/2002
and
a
PhD
grant
SFRH/BD/10600/2002) and additional support was provided by Brisa
Auto-Estradas de Portugal S.A.
This dissertation should be cited as:
Grilo C (2009) Habitat Connectivity for Carnivores in Mediterranean
Landscapes. Phd Thesis, University of Lisbon, Portugal
NOTA PRÉVIA
Na elaboração desta dissertação foram usados resultados de trabalhos já publicados ou
submetidos para publicação em revistas científicas internacionais indexadas. Uma vez que estes
trabalhos foram realizados em colaboração com outros investigadores, e de acordo com o
disposto nos termos do Nº 1 do Artigo 40, Capitulo V, do Regulamento de estudos PósGraduados da Universidade de Lisboa, publicado no Diário da República – II Série Nº 153, de 5
de Julho de 2003, esclareço que participei integralmente na concepção, obtenção, análise e
discussão dos resultados, bem como na redacção dos manuscritos dos artigos I a IV.
Devido à integração de vários artigos científicos o padrão de formatação apresentado é variável
conforme as normas de cada revista de publicação.
Lisboa, 26 de Fevereiro de 2009
Clara Bentes Grilo
i
ACKNOWLEDGMENTS /AGRADECIMENTOS
À Professora Doutora Margarida Santos-Reis pela confiança que depositou em mim ao
aceitar a orientação da tese de doutoramento, por todo o apoio e incentivo ao longo deste
processo, e sobretudo, pelos ensinamentos e oportunidades que me proporcionou e que
contribuíram para a minha formação cientifica e pessoal nos últimos 13 anos. Obrigada!
To Professor John A. Bissonette that kindly accepted to be my co-adviser which gave me a
great opportunity to make contact with new ideas and procedures to my work. Thank you for
all the support, guidance and encouragement even at distance.
À Fundação para a Ciência e Tecnologia pela concessão da bolsa de doutoramento
(SFRH/BD/10600/2002) e do estágio nos EUA bem como do financiamento do projecto
POCTI/MGS/47435/2002. À Brisa Auto-Estradas de Portugal S.A., pelo apoio logístico que
tornou possível a realização deste estudo, nomeadamente ao Engº Victor Santiago e Engª
Margarida Braga.
Às minhas musas inspiradoras: ao Joaquim Pedro Ferreira pelas longas e intermináveis
conversas sobre Conservação, e sobretudo, pelas utopias! Tu foste o meu maior apoio, o grande
camarada, também o mais crítico, mas sempre aquele que mais acreditou em mim e no meu
trabalho; ao Fernando Ascensão pela troca e partilha de ideias, pela colaboração constante nas
diferentes fases da tese; tu foste a pessoa que mais compreendeu este estudo e sem ti, não me
teria dado tanto gozo a concretizá-lo; ao Carlos Fernandes pelo seu calibre científico e pela sua
energia positiva que me contagiou e me deu ânimo para acabar este trabalho e sonhar com
novas áreas e projectos de investigação.
A todos os que me ajudaram no campo e que proporcionaram momentos inesquecíveis,
sobretudo à Marta Cruz por me ter aturado e por ter sido uma camarada ao longo de seis meses
intensivos de trabalho de campo. À Ana Marta Serronha e Ana Rita Mateus, que apesar de não
terem contribuído directamente neste estudo, a vossa participação na 2ª fase do estudo das
passagens hidráulicas, permitiu que fossem atingidos os objectivos do protocolo com a Brisa, e
libertar-me para a escrita da tese. Aos que me ajudaram a obter os dados de atropelamentos,
que sob o tema de estágio ou mestrado me substituíram por uns meses na monitorização das
estradas: a Carla Baltazar, o Luís Gomes e, em particular, a Clara Silva pela sua energia e
profissionalismo, e sobretudo, por fazer do trabalho de campo uma festa constante!
Aos funcionários dos Centros Operacionais de Estremoz, Grândola e Vendas Novas da
Brisa Auto-Estradas de Portugal pelo registo e recolha dos carnívoros atropelados; à central
betuminosa Tecnobet (Évora) pela cedência de pó de filler.
Aos donos das propriedades e caseiros que permitiram que as suas terras fossem invadidas
por máquinas fotográficas e estações de cheiro e, acima de tudo, por terem dado a conhecer, no
terreno, a realidade sobre a conservação de carnívoros: Sr. Joaquim Carriço da Sociedade
Agrícola das Silveiras, Sr. Tavares Silva, Sr. Almeida Martins, Sr. José Laboreiro, Sr. Manuel
Bruno, Sr. Picolino, Sr. Marcolino da Sociedade Agrícola de Vale do Cão, Sr. Luís, Sr. Arqtº
Aires Mateus, Sr. Eng. Alfredo Araújo.
Ao Professores Doutores Henrique Cabral, Rui Rebelo e Francisco Petrucci-Fonseca pela
revisão dos artigos e ainda aos colegas e amigos Fernando Ascensão, Joaquim Pedro Ferreira e
Joana Roque de Pinho pelos comentários e sugestões ao longo das diferentes fases deste estudo.
iii
To Dr. Susan Durham for the valuable statistical support, Dr. Eric Geese, Dr. JP Metzger,
Dr. Chip Leslie and anonymous revisores that helped with valuable comments to improve the
manuscripts.
A todos os colegas voluntários que participaram no trabalho campo e que contribuíram
para que este fosse bem mais animado: à Ana Filipa Ferreira, Dulce, Marta Pimentel, Ana Paula
Rosa, Joaquim Pedro Ferreira, Soraia Oliveira, Marco Lusquiños, João Rosário, Sílvia Rosa,
Hugo Tavares, Hugo Baptista, Dulce Ferreira, Mónica Rodrigues, Nuno Negrões, Joana
Bernardino, Joana Roque de Pinho, Jorge Román, Sara Cortez e Luís Almeida.
A toda a equipa dos carnívoros do CBA pelo apoio bibliográfico, pela partilha do material e
equipamento de campo; e em especial, ao Carlos Fernandes e Mónica Rodrigues pelo apoio nas
necrópsias.
Ao Miguel Caldeira Pais, à Helena Rio-Maior, ao Rogério Cangarato, à Sofia Alexandre, ao
Fernando Ascensão, à Catarina Ferreira, à Sílvia Rosa pela contribuição com informação de
dados de atropelamentos e bibliografia. À Ana Galantinho e Filipe Carvalho da Universidade
de Évora pelos contactos dos proprietários dos terrenos de Monfurado e dicas para as estações
de cheiro. Ao Paulo Cardoso pelo apoio nos SIGs e ao Ricardo Pita pelos esclarecimentos nas
dúvidas de estatística (já está!).
Ao recente mini-mega Grupo de Trabalho de Ecologia de Estradas (GTEE) – Fernando
Ascensão, Dyana Reto, Joel Filipe, Joana Sousa, Hugo Matos, Filipe Dias, Ana Rita Mateus e
Ana Marta Serronha, Inês Leitão, pela paciência e compreensão durante a minha ausência nos
últimos tempos, em particular à Dyana Reto por me ter apoiado e substituído em várias tarefas
do projecto nos últimos dois meses de escrita da tese.
Aos jornalistas Paulo Caetano e Carla Castelo pela divulgação dos trabalhos desenvolvidos
no âmbito desta tese na comunicação social, que é um importante contributo para a conservação
dos carnívoros em Portugal.
To John and Mary Bissonette that received me with open arms and share your home and
tasty food and all the US friends: TammyWilson and Josh, Marla Kirschbaum, Magathi Jayaram
and Jerry W., Chris Winstead, Richard Willis, Shauna Leavitt, Patty Cramer for the three kick ass
months! Also, to Ryan Lockwood (my roomy that made my stay easier in Logan), thank you so
much!
Aos amigos do meu coração pela força e estímulo e por me fazerem esquecer as partes mais
difíceis deste trabalho: à Joana Roque Pinho, ao Joaquim Pedro Ferreira, à Catarina Pereira, à
Vanda Lobo, e à Ana Henriques (Náná).
À minha família pelo exemplo e pelo apoio constante.
Aos meus pais por tudo.
iv
ABSTRACT
The goal of this thesis is to provide sound information related to humaninduced landscape changes on the Portuguese carnivore guild. More specifically, it
addresses at what scale(s) extent(s) did carnivores appear to respond to fragmentation;
which are the patterns of carnivore-vehicle collisions and their causes; how carnivores
respond to highways culverts and underpasses; and which is the relationship between
“montado” connectivity and carnivore mortality.
Stone martens were the most sensitive species to “montado” attributes and
fragmentation, although mongoose movement may be limited to shorter distances
when landscape connectivity is low. Roads, on the other hand, appear to be selectively
permeable and may act as filters to movement. On average, 47 carnivores/100
km/year were road-killed, with red fox suffering the highest mortality. Some species
were more vulnerable during specific life history periods (e.g., provisioning young,
dispersal, and breeding) and the major contributions to the deadliest segment roads
were road sinuosity, suitable habitat, and low human disturbance. This study
documented the regular use by most carnivores (0.7 ind./crossing structure/day)
except weasels, polecats and wildcats that appeared to avoid passages. In general,
carnivores preferred larger passages with vegetation close to the entrances, favorable
habitat in the vicinity, and low human disturbance. This study also showed that the
likelihood of stone marten occupancy decreased with distance to source areas and
increased with patch size. We found an increased cost attributable to open areas and
distance to riparian areas.
Results presented show that common and wide-ranging species also suffer the
effects of habitat fragmentation, and “montado” restoration efforts, as well as, road
mitigation measures are required to restore habitat connectivity in the area and ensure
long-term sustainability of a diverse carnivore guild. Further research is however
needed to clarify the regional status of less common species and to disentangle the
cumulative effects of population density, demography and behavior.
Palavras-chave – "montado", habitat fragmentation, road permeability, animal-vehicle
collisions, ecological modeling, habitat restoration.
v
RESUMO
Ao longo do último século a paisagem tem sofrido alterações à larga escala, e de
forma acelerada, devido essencialmente às actividades humanas. O crescimento
exponencial da população humana tem conduzido a um aumento da exploração dos
recursos naturais que se traduz na modificação da paisagem associada à perda e
fragmentação do habitat de muitas espécies. A destruição e fragmentação do coberto
vegetal original, a par da expansão da rede viária relacionada com o crescente número
de áreas urbanas, constituem exemplos de alterações significativas da paisagem que
podem resultar na ocorrência de populações animais cada vez mais pequenas e
isoladas e, deste modo, mais vulneráveis à extinção local. A necessidade de manter o
fluxo de indivíduos entre populações, e deste modo um fluxo genético que assegure a
viabilidade das populações a longo prazo, tornou-se uma questão chave para a
conservação das espécies só sendo realizada através da preservação e/ou recuperação
da conectividade dos seus habitats.
Os ecossistemas mediterrânicos, embora fortemente moldados pela actividade
humana e mantidos ao longo de séculos de práticas tradicionais de uso do solo, têm-se
mostrado vulneráveis à degradação, com a consequente perda de espécies, devido a
uma série de actividades humanas associadas à industrialização da agricultura,
urbanização, recreação, silvicultura extensiva e introdução de espécies exóticas. Um
exemplo, é o montado de sobro (Quercus suber) e/ou de azinho (Q. ilex) que domina a
paisagem do sul da Península Ibérica. Este sistema agro-silvo-pastoril constitui um
interessante equilíbrio entre a exploração dos recursos e a conservação da
biodiversidade, nomeadamente de pequenos e médios carnívoros (raposa Vulpes
vulpes, doninha Mustela nivalis, toirão Mustela putorius, fuinha Martes foina, texugo
Meles meles, lontra Lutra lutra, geneta Genetta genetta, sacarrabos Herpestes ichneumon,
gato-bravo Felis silvestris, e lince Ibérico Lynx pardinus). Embora actualmente a maioria
não seja alvo de preocupação em termos de conservação, excepção feita ao toirão e às
duas espécies de felinos, as alterações que se verificaram na paisagem nas últimas
décadas no sul de Portugal, colocam novas ameaças à conservação das espécies
consideradas comuns e de distribuição generalizada.
vii
A presente tese constitui uma importante contribuição para a avaliação dos
impactes da fragmentação do montado e da expansão da rede viária nos carnívoros do
sul de Portugal. Mais especificamente, neste trabalho pretendeu-se: 1) perceber qual a
sensibilidade dos carnívoros à fragmentação do montado e a que escala os atributos do
montado se expressam; 2) descrever os padrões temporais e espaciais de mortalidade
rodoviária e as suas causas, 3) analisar a resposta dos carnívoros à presença de
passagens hidráulicas e inferiores ao longo das auto-estradas; e 4) avaliar a
conectividade da paisagem para a espécie mais sensível à fragmentação do montado e
as implicações da mesma para o risco de mortalidade por atropelamento. Para atingir
estes objectivos recorreu-se a diferentes abordagens metodológicas. A avaliação da
ocupação das manchas pelos diferentes carnívoros foi realizada com recurso a estações
de cheiro, a mortalidade por atropelamento foi avaliada através da monitorização
regular das estradas e das auto-estradas, e a utilização das passagens hidráulicas e
inferiores das auto-estradas pelos carnívoros foi avaliada com câmaras fotográficas
digitais e pó de pedra colocado no interior das mesmas. A modelação ecológica a
várias escalas (local, área vital e paisagem) permitiu determinar os factores que
influenciam as respostas por parte dos carnívoros às alterações da paisagem e, a
aplicação de modelos espacialmente explícitos permitiu compreender como, a espécie
identificada como mais sensível à fragmentação, interage com a composição e a
configuração da paisagem.
Com este estudo, que obteve dados significativos relativamente a um conjunto
de cinco espécies (raposa, fuinha, texugo, geneta e sacarrabos), comprovou-se que a
raposa não aparenta ser sensível à fragmentação do montado, ao contrário da fuinha e
do sacarrabos que responderam negativamente à diminuição da dimensão das
manchas de montado e ao isolamento. Os modelos multi-escala revelaram que a fuinha
é a espécie mais sensível às características do montado e à sua fragmentação; a
qualidade do habitat à escala local (maturidade do montado e ausência de pastoreio
intensivo) mostrou-se ser o parâmetro mais importante para explicar a ocupação das
áreas de montado pela fuinha, seguido do efeito conjunto do isolamento e área de
montado à escala da área vital e da paisagem, respectivamente. O sacarrabos mostrouse bastante sensível à percentagem de coberto vegetal o que se traduziu na ocupação
de manchas com uma reduzida distância entre elas e na preferência pela ocupação de
áreas com elevada densidade de árvores e de sub-coberto. No entanto, a maior
contribuição para explicar os padrões de ocupação das áreas de montado pelo
viii
sacarrabos foi dada pela interacção das variáveis à escala local, da área vital e da
paisagem, sugerindo que a conectividade do habitat se expressa pelo efeito cumulativo
dos factores às três escalas.
Por outro lado, as estradas revelaram-se selectivamente permeáveis à
comunidade de carnívoros, parecendo actuar como filtros ao movimento de algumas
espécies. Nos 574 km de estradas e auto-estradas monitorizadas detectaram-se em
média 47 cadáveres de carnívoros em cada 100 km por ano, tendo sido a raposa, a
espécie com maior número de atropelamentos (42%), seguida da fuinha (17%),
sacarrabos (12%), texugo (11%) e geneta (11%). A lontra, o toirão e a doninha foram as
espécies menos atropeladas. Os cinco carnívoros vítimas de um maior número de
atropelamentos, com excepção da geneta, mostraram-se vulneráveis ao tráfego
rodoviário em determinados períodos biológicos: a época da alimentação das crias para
a raposa e a fuinha, a dispersão para o texugo, e a época do cio para o sacarrabos, o que
a longo prazo pode implicar alterações na estrutura das populações e na densidade
populacional.
Nas auto-estradas, as passagens hidráulicas e inferiores de maiores dimensões,
integradas em habitat favorável e com reduzida perturbação humana foram as mais
frequentemente
utilizadas
pelos
carnívoros.
Em
média
detectou-se
0,7
carnívoros/passagem/dia, o que significa que estas estruturas contribuem para
manutenção da permeabilidade das auto-estradas e deste modo para a conectividade
do habitat dos carnívoros, com excepção da doninha, toirão e gato-bravo cuja
utilização regular não foi detectada; a aparente reduzida densidade destas espécies na
região (que também não foram detectadas nas estações de cheiro) pode, no entanto,
contribuir para este resultado. A modelação da distribuição dos atropelamentos de
cada espécie revelou que em geral, a presença de habitats adequados à sua ocorrência,
a proximidade a curvas e a reduzida perturbação humana foram os factores que mais
contribuíram para a mortalidade nas estradas. O modelo de superfície de custo obtido
para a fuinha sugere que a probabilidade de ocupação das manchas de montado
diminui à medida que aumenta a distância às áreas-fonte, com um incremento máximo
de custo nas áreas abertas e distantes de linhas de água, e está positivamente associada
à dimensão da mancha. A análise conjunta da composição e configuração da paisagem
e da incidência de atropelamentos revelou que a área de estudo ainda parece
apresentar um nível de conectividade do montado pouco preocupante para a fuinha,
ix
mas a incidência de atropelamentos desta espécie nos segmentos de estrada nas áreas
contínuas de montado, em particular, naqueles com um maior número de curvas,
significa que essas áreas apresentam um maior risco de mortalidade e deste modo,
constituem uma barreira aos movimentos da espécie.
Os resultados deste estudo indicam que os esforços de conservação devem ser
conduzidos com vista à melhoria da qualidade do habitat para a fuinha e o sacarrabos,
tendo em conta a conectividade ao nível das áreas vitais. O melhoramento das galerias
ripícolas e o aumento das áreas de matos, mostram-se cruciais para a recuperação da
qualidade e da conectividade não só do montado mas também da matriz cerealífera.
No que diz respeito à minimização dos efeitos negativos da rede viária, o seu
planeamento deve ter em conta: 1) o espaçamento entre as passagens de forma a
garantir a permeabilidade das estradas a toda a comunidade de carnívoros, 2) a
adaptação de passagens com elementos naturais à entrada e no seu interior para
promover a utilização das espécies menos familiarizadas com estruturas artificiais, 3) a
colocação de redes de malha fina e enterradas em passagens que se encontram em
segmentos de estrada com habitat favorável para as espécies e 4) o corte regular de
vegetação nas bermas de curvas para aumentar a visibilidade do condutor e do
individuo. Os resultados apresentados nesta tese mostram como os carnívoros comuns
e de ampla distribuição podem estar ameaçados com as alterações da paisagem
induzidas pelo homem se, persistirem estas alterações, e se, não forem tomadas
medidas para minimizar os efeitos negativos da fragmentação do montado e da
expansão da rede viária. Contudo, mais investigação é necessária para clarificar o
estatuto local das espécies menos comuns e isolar os efeitos cumulativos da densidade
populacional, demografia e comportamento.
Palavras-chave – montado, fragmentação do habitat, permeabilidade das estradas,
mortalidade por atropelamento, modelação ecológica, restauração do habitat.
x
CONTENTS
Agradecimentos…………………………………………………………………………………………………………
iii
Abstract…………………………………………………………………………………………………………………..
v
Resumo…………………………………………………………………………………………………………………...
vii
CHAPTER 1. GENERAL INTRODUCTION……………………………………………………….
1
1.1 OVERVIEW…………………………………………………………………………………………..
1
1.1.1 Conceptualizing Human effects on landscapes and species………….....................................
3
1.1.1.1 Landscape models and patterns……………………………………….…….………………...
3
1.1.1.2 Effects of landscape changes on species…………………………………………..…………
5
1.1.1.3 Habitat connectivity………………………………………………….…………………………
7
1.1.2 History of the Mediterranean landscape…………………………..………………………….
8
1.1.2.1 Evolution of land-use………………………………………….………………………………..
9
1.1.2.2 “Montado” - one example of a changing landscape……………..…………………..……..
10
1.1.3 Carnivores as a focal group……………………………………………….…………………..
11
1.1.3.1 Carnivores as a conservation interest taxa…………………………………...……………...
11
1.1.3.2 Mediterranean carnivore guild……………………………………………….……………….
13
1.1.3.3 Habitat fragmentation and road impacts on Mediterranean carnivores …………..…….
18
1.2 AIMS AND THESIS STRUCTURE…………….…………………………..…………………………...
20
1.3 REFERENCES………………………………………………………………………………………..
22
CHAPTER 2. STUDY AREA…………………….....…..………………………………………..
37
2.1 STUDY AREA………………………………………………………………………………………...
39
2.2 REFERENCES………………………………………………………………………………………...
42
CHAPTER 3. RESPONSE OF MEDITERRANEAN CARNIVORES TO “MONTADO”
FRAGMENTATION……………………………………………………………………………………..
44
3.1. PAPER I
Scale-dependent effects of Mediterranean landscapes on meso-carnivore occupancy patterns. Grilo C,
Cruz M, Bissonette JA, Santos-Reis M. Landscape Ecology (submitted)…………………………………
49
CHAPTER 4. ROAD EFFECTS ON MEDITERRANEAN CARNIVORES…………………
71
4.1. PAPER II
Spatial and temporal patterns in Mediterranean carnivore road casualties: consequences for mitigation.
Grilo C, Bissonette JA, Santos-Reis M. Biological Conservation (2009) 142 301-313……………….
75
4.2. PAPER III
Response of carnivores to existing highway culverts and underpasses: implications for road planning and
mitigation. Grilo C., Bissonette JA, Santos-Reis M. Biodiversity and Conservation (2008) 17, 16851699…………………………………………………………………………………………………………
91
CHAPTER 5. MODELLING STRUCTURAL CONNECTIVITY FOR STONE MARTEN
IN SOUTHERN PORTUGAL……………………………………………………………….…………
107
5.1 PAPER IV
Structural connectivity in a Mediterranean landscape and its consequences for road mortality risk of
stone marten. Grilo C, Ascensão F, Bissonette JA, Santos-Reis M. Landscape and Urban Planning
(submitted)………………………………………………………………………………………………..
111
CHAPTER 6. GENERAL DISCUSSION…………………………….………………………………
131
6.1 OVERVIEW OF MAIN FINDINGS…………..…………………..…………………………………….
133
6.1.1 Are common and wide-ranging carnivores vulnerable to “montado” fragmentation?.............................
133
6.1.2 Do roads have impact on the small- and medium-sized carnivores?.........................................................
135
6.1.3 Is Alentejo structurally connected for carnivore species?..........................................................................
136
6.2 CHALLENGES AND PRIORITIES IN CARNIVORE CONSERVATION IN ALENTEJO…...……..………..
138
6.2.1 Improve knowledge on the impacts of human induced changes on carnivores guild……………............
139
6.2.2 Focus conservation action on restoring habitats…………………………………………………...…….
140
6.2.3 Mitigate the effects of road construction and their associated impacts……………….........................…
142
6.3 REFERENCES………………………………………………………………………………………...
143
xii
CHAPTER 1 - GENERAL INTRODUCTION
CHAPTER 1 GENERAL INTRODUCTION
1.1 OVERVIEW
There is a growing consensus within the scientific community that the extinction
of many species is largely unprecedented and is taking place on a short time scale (e.g.,
Pimm and Raven 2000, Novacek 2001, Thomas et al. 2004). Human population growth
and increasing natural resources demands have lead to landscapes modification with
associated habitat destruction and fragmentation for many species. As a consequence,
extinction rates have risen to levels rivalling the geologic history of mass extinctions
(Chapin III et al. 2000, Pimm et al. 2006) causing widespread changes in the global
species distributions and in the ecosystems they belong to (e.g., Soulé 1996, Ceballos
and Helrich 2002).
Humans have transformed almost 50% of the ice-free land surface, changing
forests, prairies and wetlands into agricultural and urban systems in nearly all regions
of the world (Daily 1999). Several dramatic examples show that expansion of forestry
and agricultural activities, combined with inappropriate management practices, have
led to the destruction of huge areas of natural habitats (Skole and Tucker 1993). In
tropical forests, the richest habitat for biodiversity, a large part of the original extent
may have been lost to deforestation in the last two decades, primarily as a result of
agricultural expansion (e.g., Ferraz et al. 2005). Annually tropical deforestation in the
Brazilian Amazon ranged from 100 000 to 165 000 km2 in the late 1980s; currently 50 to
70% of this region has been deforested (Skole and Tucker 1993). Similar patterns of
deforestation, and its consequent environmental degradation, have occurred in many
regions of the world: Europe, south Asia, Australia, temperate North and South
America (see Lindenmayer and Fisher 2006). Today the forests of Europe's are some of
the most degraded in the world. Although forests extend over some 30% of the
Europe’s surface (half the original maximum extent of forest cover), only about 2% is
classified as virgin forest or old growth forest remnants (Colchester 1998). In Australia,
1
General Introduction
93% of the western native vegetation was removed and converted to wheat over the
past 100 years, much of it disappearing within the last 50 years (Saunders et al. 1993). In
United States, during the 18th and 19th centuries much of the native forest and prairie
ecosystems of the Northeastern and Midwestern were completely transformed to
pasture and cropland (Hessburg and Agee 1993, August et al. 2002).
The development and expansion of urban areas and associated infrastructures
also reduced natural habitats, resulting in further losses (Forman et al. 2003). Roads
occupy a considerable area of land and form extensive longitudinal obstacles. In the
European Union, the traffic system has developed into a fine-meshed network with an
average ‘grid’ density up to 1.3 km.km-2 (source: European Road statistics 2007).
Estimate of the length of the road network is about 6.4 million kilometres in United
States (source: U.S. Department of Transportation 2007). This network has
fundamentally altered or destroyed millions of hectares of land and water bodies
(Trombulak and Frissel 2000). Roads are thought to be particularly strong barriers or
filters for many animal species because of dissection of their habitat, mortality caused
by animal-vehicle crashes, and avoidance behaviour due to traffic volume (see reviews
in Forman and Alexander 1998, Trombulak and Frissel 2000, Forman et al. 2003). These
species include a wide variety of taxa such as anuran (e.g., Eigenbrod et al. 2008), turtles
(e.g., Beaudry et al. 2008), birds (e.g., Erritzoe et al. 2003), European hares (Roedenbeck
and Voser 2008), badgers (Clarke et al. 1998), bobcats and coyotes (Riley et al. 2006),
Iberian lynx (Ferreras et al. 1992), Florida panther (Taylor et al. 2002), wolves (Paquet
1993) and bighorn sheep (Epps et al. 2005).
Not surprisingly, intervention by humans is required for any possibility of shortterm recovery and maintenance of the biota. Issues associated with landscape changes
by humans have rapidly become a central issue in conservation policy (Donald et al.
2007, Fontaine, et al. 2007, Gaston et al. 2008), as evidenced by the signing and
ratification of international conventions (e.g., Bern Convention (1982), Habitat directive
92/43/EEC) and the development of regulatory actions (e.g., Iberian lynx Action Plan Delibes et al. 2000) aimed to mitigate human impacts on biodiversity. Meanwhile, those
impacts and associated mitigations are becoming a major focus of research in
conservation biology, namely in significant and emerging fields such as landscape and
road ecology (e.g., Wilcox and Murphy 1985, Malanson and Cramer 1999, Forman et al.
2
General Introduction
2003, Turner et al. 2003, Lindenmayer and Fisher 2006). Nevertheless, many current
recovery efforts still have deficiencies, including insufficient information on the
diversity and distribution of species, ecological processes, and the magnitude and
interaction of threats to biodiversity. This is particularly true in Europe, and has
prevented the accomplishment of the 2010 target of decreasing or at least halting the
loss of biodiversity (Countdown 2010 Initiative - http://www.countdown2010.net).
1.1.1 CONCEPTUALIZING HUMAN EFFECTS ON LANDSCAPES AND SPECIES
1.1.1.1 Landscape models and patterns
Various theories and landscape models make predictions to determine and
understand how landscape alteration affects organisms. For example, MacArthur and
Wilson (1967) developed the island biogeography theory (IBT) to explain how
distance and size combine to regulate the balance between immigration and extinction
in island populations. This theory suggests that fragmented ecosystems have lower
species richness per unit of area compared with contiguous habitats. Further, species
survival is dependent on fragment size and isolation. The limitations of the IBT become
obvious when the theory fails to consider crucial issues including: 1) how community
composition in fragments changes over time, 2) which species are most vulnerable to
fragmentation, 3) what are the effects of barriers, edges, and landscape heterogeneity
and fragmentation in ecosystem functioning, and 4) what is the role of anthropogenic
changes on biodiversity (Brown and Lomolino 2000, Laurance 2008). Recent studies
(e.g., Brotons et al. 2003) have provided new insights in fragmentation studies (Forman
and Godron 1986, McIntyre and Barret 1992). For example, the matrix is not necessarily
hostile to native species as assumed by the IBT (McIntyre and Barrett 1992). Forman
(1995) argued that functional flows and movements through the landscape are
determined by the arrangement and structural pattern of three landscape components
(patches, corridors and matrix) and change over time (patch-corridor-matrix model).
Patches are relatively homogeneous non-linear areas that differ from the surroundings;
they are dynamic and occur on a variety of spatial and temporal scales according to
species perception (Wiens 1976, Wiens and Milne 1989). Small patches may serve as
stepping stones for species dispersal or re-colonization, providing heterogeneity in the
3
General Introduction
matrix (Forman 1995). Corridors are non-matrix strips that connect two or more
patches. As a consequence of their form and context, structural corridors may function
as habitat, dispersal conduits, or even barriers. The matrix is the most extensive and
most connected landscape element type, and therefore plays the dominant role in the
functioning of the landscape (Forman and Godron 1986, Lindenmayer and Fisher
2006). The variegation model is a modification of the IBT and patch-corridor-matrix
models, according to which the landscapes formerly dominated by the original habitats
have been variously modified rather than extensively destroyed. The result is in the
differential filtration of species across the elements of the landscape (McIntyre and
Barret 1992). The matrix ceases to be an impermeable barrier.
According to Forman (1995, see also Lindenmayer and Fisher 2006),
anthropogenic changes have modified landscapes around the world to varying
patterns including: 1) perforation (the process of making ‘holes’ in a landscape, e.g.,
mine sites in remote areas, disturbance patches in a remnant forest), 2) dissection (the
subdivision of an area, e.g., by roads, power lines), 3) fragmentation (the changing of a
contiguous landscape into patches, e.g., remnant vegetation in grazing lands), 4)
shrinkage (a decrease in size, e.g., patch size reduction due to realignment of fences
around pastures), and 5) attrition (the disappearance of landscape patches, e.g., lowest
productivity patches cleared in heavily exploited landscapes). McIntyre and Hobbs
(1999) suggested another classification of spatial patterns resulting from humaninduced landscape changes that include a temporal component. As the human
modification increases and habitat amount decreases, landscapes can be intact,
variegated, fragmented or relictual, depending on the extent (90%, 60-90%, 10-60% or
10% of cover, respectively) of habitat destruction and associated land uses. These
classes represent segments along a continuum of destruction of a previously
continuous habitat matrix. In intact and variegated landscapes, habitat still forms the
matrix, whereas in fragmented landscapes the matrix is comprised of altered habitat
(e.g., Dunning et al. 1995). Both classifications recognize that landscape changes often
result in a decrease in the average size of the remaining vegetation patches, an increase
in the average distance between these patches, and, as a consequence, a decrease in
landscape connectivity between patches (Fahrig 2003).
4
General Introduction
1.1.1.2 Effects of landscape changes on species
Research shows that species respond strongly to the different spatial patterns
created by different kinds of landscape change (Hames et al. 2001), and responses may
not be in accordance with human visual perception of the landscape pattern. The key
process that is associated with landscape change is habitat loss and it is the primary
driver of biodiversity decline (Fahrig and Merriam 1985, Pimm and Askins 1995). In
general, habitat loss is typically accompanied with habitat degradation, subdivision,
and isolation, meaning limitation on food, shelter, and space availability, all
fundamental to species survival. This can lead to smaller local populations, which then
face a greater risk of extinction due to demographic or environmental stochastic events
(Belovsky et al. 1994, Bascompte et al. 2002, Donovan and Flather 2002).
Fragmentation may also have significant consequences for population genetic
structure. Distance and physical barriers may reduce exchanges of individuals, and
therefore, decrease the gene flow (Wright 1943, Hedrick and Gilpin 1996, Gerlach and
Musolf 2000, Hale et al. 2001). Loss of genetic variation will decrease the ability of
individuals to respond to changing environmental conditions and may even result in
genetic fixation of deleterious alleles as well inbreeding in local populations that can
lead to reduced individual longevity and fecundity (Frankham et al. 2002).
Taken together, the environmental, demographic, and genetic effects of habitat
fragmentation can accelerate the decline and extinction of populations and species
(Gilpin and Soulé 1986). In a recent review, Fahrig (2003) demonstrated that the habitat
fragmentation impacts community structure including the strength of species
interactions, such as competition, predation, parasitism, herbivory (e.g., Tilman et al.
1994 Taylor and Merriam 1995, Chapin III 2000, Hames et al. 2001, Lin et al. 2005,
Pakkala et al. 2006), patterns of dispersal (e.g., Cale 2003) and trophic chain length in
food webs (e.g., Kareiva 1987, Ewers and Didham 2006), as well as population
distribution and abundance (e.g., Mbora and Meikle 2004), sociality (e.g., Wolff 1999),
breeding success (e.g., Kurki et al. 2000, Donovan and Flather 2002, Horn et al. 2005),
foraging success (e.g., Mahan and Yahner 1999, Ewers and Didham 2006) and genetic
diversity (e.g., Banks et al. 2005). Often there is a threshold below which the population
cannot sustain itself (extinction threshold). For example, small populations will be at a
higher risk of extinction due to stochastic events; similar risks hold for species that are
5
General Introduction
habitat specialist, species at limits of their geographic range or species with low
mobility or dispersal power (e.g., Swihart et al. 2003). Complex biological processes,
e.g.; reproductive potential, annual survival, sociality, trophic position and species
interactions may also increase sensitivity to landscape change (Henle et al. 2004).
Nevertheless, the critical values of the landscape changes affecting species persistence
have been poorly studied and are a controversial issue. For example, Andrén (1994)
found significant effects of fragmentation for birds and mammals in approximately 2030% of vegetation cover, while Bissonette et al. (1997) found that after ~30% logging on
intact forest, American marten populations declined in Newfoundland. However, With
and King (1999) argued that the level extinction threshold varied from 5 to 90% of
habitat loss in the landscape, depending on both species response to fragmentation and
spatial arrangement of the remaining habitat.
Moreover, the scale of analysis also plays an important role in understanding
the organism’s responses to landscape structure (e.g., Gehring and Swihart 2003,
Holland et al. 2004, Swihart et al. 2006). Differences in scales at which different species
perceive landscape are largely dictated by physiological and morphological features
such as body size, mobility, and behavioural characteristics (e.g., Morris 1992). As a
consequence, several studies highlight the need to evaluate factors simultaneously at
different scales as well the interactions between them (e.g., Zabala et al. 2005, AnzuresDadda and Manson 2007). For instance, a multiple scale study of temperate zone
rodents (Morris 1987) revealed that microhabitat plays a small role and large scale
efforts are needed to enhance rodent density. In contrast, Cushman and McGarigal
(2004) found that micro-scale features were better predictors of Oregon Coast Range
bird community structure than patch- or landscape-level features. Hence, multiple
scale analysis can have major implications in the findings of fragmentation studies and
enhance the effectiveness of conservation strategies (Bissonette and Storch 2002);
otherwise, management actions implemented to ameliorate the negative effects of
fragmentation might be less effective or fail (Orrock et al. 2000, Gehring and Swihart
2003).
6
General Introduction
1.1.1.3 Habitat Connectivity
As a natural consequence of the concern about habitat fragmentation, the
maintenance and establishment of connectivity is a relatively recent concept in
landscape change and planning analyses (Fahrig and Merriam 1985, Saunders et al.
1991, Chetkiewicz et al. 2006). Because connectivity facilitates biological fluxes such as
animal dispersal, genetic flow and multiple other ecological functions, it has become a
key conservation issue to mitigate the effects of habitat loss and fragmentation (Hale et
al. 2001, Moilanen and Nieminen 2002, Goodwin and Fahrig 2002). Bennett (1999)
considered two kinds of connectivity: 1) structural - the spatial arrangement of
different types of habitat such as density and complexity of corridors, distance between
patches and inter-habitat matrix permeability, and 2) functional (or behavioural) - the
interaction between behavioural response of individuals, species, or ecological
processes and the physical structure of the landscape. Connectivity depends not only
on the spatial pattern of the landscape elements, but also on the interactions between
this pattern and the biological characteristics of the species such as, their ability to use
and move through the landscape and their risk of mortality in different landscape
elements (Uezu et al. 2005, FitzGibbon et al. 2007). The last definition emphasizes that
the type, composition and configuration of habitat patches on the landscape influence
movement, and ultimately population dynamics and community structure (Taylor et al.
2006, Fahrig 2007). Further, structurally connected landscapes may be functionally
connected for some species and not for others (e.g., Uezu et al. 2005). One example is
the study reported by Wiegand et al. (2005) showing that functionally connected
landscapes are represented by abundant dispersal habitat and sufficiently large
dispersal potential despite the level of fragmentation.
Increasing knowledge on the species ecology combined with the use of new
technologies (e.g., geographic information systems) and analytical tools (e.g., cost
surface analysis) has enabled the development of spatially explicit models to
understand how organisms interact with landscape pattern (e.g., Hunter et al. 2003,
Jepsen et al. 2005, Drielsma et al. 2007, Larue and Nielsen 2008, Rodriguez-Freire and
Crecente-Mesada 2008). One of the most widely used methods is the cost surface
analysis (CSA), which links animal behaviour and landscape structure by modelling
how individuals of a species perceive the permeability of a landscape (e.g., Ferreras et
7
General Introduction
al. 2001). The method assumes variable travel costs for different features of the
landscape (e.g., LaRue and Nielsen 2008). The cost of each landscape element reflects
the degree to which that element facilitates or impedes movements of the studied
species (Adriaensen et al. 2003). Data to define the cost values are based on the location
and characteristics of each landscape element and on the biological aspects of the
species being evaluated, namely the movement ability in the landscape elements
(Baguette and Dyck 2007). With no prior information on species movement, studies
used cost surface simulations with different combinations of cost values (e.g., Walters
2007) or weight the cost of each landscape element by using habitat suitability models
(e.g., Wang et al. 2008). The habitat suitability map gives the probability for each patch
to be used by a given species and therefore, provides information to generate cost
layers for CSA. Recently, researchers went further by validating the CSA with
population- or individual-based genetic analysis to assess functional landscape
connectivity (e.g., Walker et al. 2007, Wang et al. 2008).
Habitat connectivity is also a growing concern in roaded landscapes (Iuell et al.
2003). In fact, even in well structurally connected areas, the incidence of road-kills or
the avoidance effect of roads may bisect suitable areas that became gradually more
fragmented (Forman et al. 2003). Hence, the connectivity assessment needs to be
associated with road network characteristics (e.g., Eigenbrod et al. 2008). A growing
number of monitoring surveys are providing new and important insights on the
relationship between landscape permeability and individual movements (e.g., Yanes et
al. 1995, Clevenger et al. 2001, Ramp et al. 2005, Riley et al. 2006). Accurate information
on the spatial patterns of mortality and on the use of road crossing structures by fauna
may help in the assessment of road permeability, and therefore, help to identify
adequate mitigation measures to maximize connectivity across roads to wildlife.
1.1.2 HISTORY OF THE MEDITERRANEAN LANDSCAPE
Mediterranean ecosystems have been intensively managed by human
populations (Naveh 1993, Blondel and Aronson 1999). Although shaped by human
activity and maintained by centuries of traditional practices of land use, Mediterranean
ecosystems are now vulnerable to degradation and species loss due to a range of
human activities associated with agriculture intensification, recreation, forestry,
8
General Introduction
urbanization, industrialization, resource exploitation and introduction of exotic species
(Allen 2003, Blondel 2006, Vogiatzakis et al. 2006).
The conservation value of the Mediterranean basin is recognized worldwide
(e.g., Olson and Dinnerstein 1998, Atauri and Lucio 2001), and although occupying
only 1.2% of the Earth’s surface (Di Castri, 1981) represents one of the 25 world
terrestrial biodiversity hotspots (Myers et al. 2000) as a result of an exceptionally high
floristic richness and a high proportion of endemic species (~50%). The extremely
complex geological and geomorphological history and the climatic variations that
occur within the Mediterranean Basin account for the remarkable variety of species
and habitats (Blondel and Aranson 1999).
1.1.2.1 Evolution of land-use
The environmental history of the Mediterranean region is described in detail by
Allen (2003). Briefly, the Mediterranean region has been characterized by shifts in
vegetation communities over the last three million years. Higher latitudes experienced
major advances of ice sheets and glaciers interspersed with warm interglacial stages;
meanwhile Mediterranean Basin suffered deep falling and rising temperatures on a
similar cyclical basis (Mannion 2008), with a continual reassembly of the vegetation
cover as temperatures changed. Equally, Meditterranean landscapes have experienced
a long and intense history of human influence (Blondel 2006). According to Naveh and
Dan (1973) humans have exerted effects on Mediterranean ecosystems for over 50 000
years, but the major impact began about c. 10 000 years ago when modern humans
(Homo sapiens sapiens) colonized the region and began the domestication of plants and
animals. By far the most obvious consequence of humans in the Mediterranean Basin
was forest destruction (Blondel and Aranson 1999). Quézel and Médail (2003) estimate
that no more than 15% of the pristine Mediterranean forest vegetation remains today,
the remaining being in a more or less advanced stages of degradation or deforestation.
Over time the Mediterranean Basin experienced a progressive replacement of
deciduous broad-leaved forests by evergreen sclerophyllous oak forests (Quercus sp.)
and scrublands, which increased patchiness, and affected both the distribution of
populations and their genetic diversity. This co-evolution shaped the interactions
9
General Introduction
between the ecosystems and humans, through constantly evolving land use practices
and their associated intermediate disturbance regimes due to agriculture and grazing
(Blondel 2006). Today, the dominant vegetation community is the evergreen
sclerophyllos shrubland, comprised by evergreen oaks, olives (Olea europaea) shrubs
(Ceratonia sp., Pistacia sp., Cistus sp.) and diverse herbaceous plants (Allen 2003) which
express the environmental gradients, such as moisture availability, nutrient
availability, and temperature (Vogiatzakis et al. 2006).
1.1.2.2 “Montado” - one example of a changing landscape
Nowadays, a highly specific agro-silvo-pastoral system named “montado” in
Portugal and “dehesa” in Spain dominates the landscape of the Iberian Peninsula
(Schnabel and Ferreira 2004). In the southwestern Iberian Peninsula, this system is
characterised by the presence of a savannah-like open tree layer, mainly dominated by
Mediterranean evergreen oaks – cork oak (Quercus suber) and holm oak (Q. ilex) – and
to a lesser extent by the deciduous Q. pyrenaica and Q. faginea associated with
livestocking and cultivation. Pinto-Correia (1993) described the “montado” as an
almost closed system: trees furnish cork and wood, as well shelter and fruits, which
constitutes an important complement to the diet of animals; and grazing animals
provide soil fertilization and contribute to field clearance. Additionally, this traditional
land use system shows an interesting balance between exploitation and conservation of
biodiversity. It is characterized by a mosaic of crops, grasslands, scrublands and
scattered trees that furnish a large diversity of products and ecological conditions to
sustain a large number of wildlife species (e.g., Santos-Reis et al. 2004, Blondel 2006,
Rosalino et al. 2008).
Nevertheless, in last few decades the Mediterranean rural landscape is
undergoing major changes. As rural economies improve, additional anthropogenic
activity is impacting biological diversity (Plieninger 2003, Vogiatzakis et al. 2006). In
Portugal, during what was known as the "Wheat Campaign" in the first half of the 20th
century, the “montado” forests were reduced, and mechanization and fertilization were
intensified to increase wheat production. As a result, large areas of forest/shrubby
cover were cleared, soil erosion increased, the fragile relationships of the “montado”
system were disrupted, and resources deteriorated (Pinto Correia 1993). As a
10
General Introduction
consequence, fields were abandoned; the land no longer had adequate production or
regeneration potential. Recently, as a result of the European Union Common
Agricultural Policy the tendency to abandon less productive lands accelerated and the
exploitation of the more profitable ones was intensified (Gómez-Limón and Fernadez
1999). In the abandoned areas, cultivation and often grazing disappeared and the
density of undercover vegetation increased, resulting in the development of the
maquis, a new type of thick scrubby evergreen underbrush. This represented the new
wilderness areas, providing high cover and shelter for some fauna, especially forestdwelling species (Mangas et al. 2008), but resulted in habitat reduction for species
requiring open habitats. However, the lack of light and space led gradually to the
almost total suppression of the herbaceous understorey in these new maquis areas
(Naveh 1993), in turn are the source food for many prey species in which carnivores
and birds of prey depend (Goméz-Limón and Fernandez 1999).
Current land use pattern and the continuing changes seen today are the result
of a long and ever changing history of political influence and economic factors that are
related to national and international policies. Consequently, the current situation offers
a series of opportunities for “montado” conservation and should receive considerable
attention from scientists, land managers, and policy makers. A detail analysis of
species requirements and the scale at which they perceive the landscape changes is
critical in helping to define an integrated “montado” management for the maintenance
of its conservation value.
1.1.3 CARNIVORES AS A FOCAL GROUP
1.1.3.1 Carnivores as a conservation interest taxa
Many carnivores have suffered dramatic declines over the past few hundred
years (e.g., Ceballos and Ehrlich 2002, Rodriguez and Delibes 2003, Naves et al. 2003).
Habitat loss, depletion of their prey, exploitation for fur, and direct human persecution
have all contributed to the past carnivore declines (Woodroffe 2001). Today,
persecution and loss of prey are the most immediate short-term threats to carnivores,
but continued land conversion and the recent additional mortality caused by traffic are
11
General Introduction
probably the greatest long-term threat to their persistence (Ginsberg 2001, Kerley et al.
2002, Burkey and Reed 2006).
Fortunately, the recent recognition of the ecological importance of top predators
and of the lack of information to support recovery strategies has led to increased
concern over the effects of the human-induced changes on them (Miller et al. 2001,
Sergio et al. 2008). The removal of a top carnivore from an ecosystem can have an
impact on the relative abundance of prey species and within the guild (e.g., Palomares
et al. 1996, Delibes et al. 2008), causing effects that cascade down a food chain to plants,
affecting the interactions among species as well as the structure of ecological
communities, and altering basic ecosystem processes (e.g., Estes and Palmisano, 1974;
Crooks and Soulé 1999, Duffy et al., 2007, Delibes-Mateus et al. 2008). One example is
the decline of the cougar (Puma concolor) in Zion Canyon (Utah, USA) that led to higher
mule deer densities, higher browsing intensities, reduced recruitment of riparian
cottonwood trees, increased bank erosion and reductions in both terrestrial and aquatic
species abundance (Ripple and Beschta 2006). Another relevant example of their
importance is embodied in the “mesopredator release” hypothesis, according to which
the removal of a top predator causes an increase in meso-predators with detrimental
effects on prey populations (e.g., Gehrt and Prange 2006). In Doñana National Park
(Spain), Palomares et al. (1998) reported that Egyptian mongoose (Herpestes ichneumon)
densities were directly controlled by Iberian lynx (Lynx pardinus); these authors also
found that in areas without Iberian lynx, mongooses took up to 10 times more
European rabbits (Oryctolagus cuniculus) than in areas with lynxes where rabbits on
average were three times more abundant.
In general, carnivores are especially and particularly vulnerable to threatening
processes, namely changes in landscape composition and patterns and also road
network expansion (Sunquist and Sunquist 2001). Many predators occupy small
geographic ranges, have small populations, and occur in low population densities.
Most show low reproductive output and relatively long gestation length and show
territorial behaviour. Others have larger home ranges and do not tolerate large urban
areas or high human activity (Cardillo et al. 2004). One example of extreme range
contraction over the last few is the Iberian lynx. There is evidence indicating a
continuous decline due to severe depletion of its primary prey (European rabbit) by
diseases and over-hunting (Rodriguez and Delibes 1992). Other human activities have
12
General Introduction
further compounded the problems of prey scarcity. Reduction and fragmentation of
the Mediterranean scrublands on which the lynx and its prey depend have also
negatively impacted the species (Rodriguez and Delibes 1992). Nowadays, declining
lynx populations persist under serious human pressure, including forestation with
non-native species and scrub clearance schemes, direct persecution, road building,
dam construction, and urbanization (Fernandes 2000, Rodriguez and Delibes 2004,
Santos et al. 2007). Growing vehicle traffic volume is also responsible for additional
mortality on lynx; each year juveniles die when they disperse and search for new areas;
this has been documented particularly in the surroundings of Doñana National Park
(Ferreras et al. 1992). In this context, reducing mortality and the risk of extinction for
Iberian lynx involves the conservation and improvement of habitat connectivity and
the removal of human-caused mortality (Ferreras et al. 2001, Palomares 2001).
Many studies have reported that single species are often poor indicators of
habitat quality or community level biodiversity (e.g., Chase et al. 2000). While studies
of this type are helpful for the conservation of specific species, they tend to contribute
little to our understanding of the collective habitat relationships of a community
(Manley et al. 2004). Since not all carnivores respond in the same way to landscape
pattern due to a wide range of morphological, ecological and behavioural adaptations
(Gittleman 1989, Carroll et al. 2001, Crooks 2002, Virgós et al. 2002, Mortelliti and
Boitani 2008), carnivores are expected to show marked differences in their response to
landscape changes. Hence, a multi-species approach to address questions related to
human landscape disturbance is needed to provide important key insights to create a
cohesive strategy to ensure the coexistence of multiple species.
1.1.3.3 Mediterranean carnivore guild
In the Mediterrenean Basin, there is a high and diverse predator community
comprised of 23 small-, intermediate- and large-sized carnivores species, most native
(golden jackal Canis aureus, wolf Canis lupus, red fox Vulpes vulpes, brown bear Ursus
arctos, stoat Mustela erminea, steppe polecat Mustela eversmanii, European mink Mustela
lutreola, weasel Mustela nivalis, Western polecat Mustela putorius, marbled polecat
Vormela perergusna, stone marten Martes foina, pine marten Martes martes, Eurasian
badger Meles meles, Eurasian otter Lutra lutra, European wildcat Felis silvestris, Lynx
13
General Introduction
Lynx lynx, Iberian lynx) and a few exotic (racoon dog Nyctereutes procyonoides, racoon
Procyon lotor, American mink Mustela vison, common genet Genetta genetta, small Indian
mongoose Herpestes javanicus, Egyptian mongoose, Indian grey mongoose Herpestes
edwardsi) (Mitchell-Jones et al. 1999). With the exception of the wolf, the European
mink, the marbled and the Western polecats, the Eurasian otter and the Iberian lynx,
the remaining species are not considered threatened and are classed as of ‘Least
Concern’ (IUCN 2008 Red List http://www.iucnredlist.org).
These carnivores exhibit different life-history characteristics and strategies
including habitat generalists and specialists, opportunistic vs. specialist foragers, as
well as solitary and group-living species.
Among solitary living carnivores, the red fox is the most widespread, habitat
and food generalist carnivore living in the Mediterranean (Mitchell-Jones et al. 1999). It
has a high ecological resilience due to its mobility and reproductive capacity (Blanco
1998, Piñero 2001). The spatial behaviour of red foxes shows great flexibility; they are
found in a great variety of habitats that provide food and cover (Fedriani 1996, Piñero
2002). As opportunistic predators, red foxes consume a wide variety of food resources
and feed according to the availability of prey (Carvalho and Gomes 2001, Cavallini and
Lovari 1991, Delibes-Mateos et al. 2008).
Stoats and weasels are also widespread, although the first is restricted to the
northern latitudes (Blanco 1998, Mitchell-Jones et al. 1999). Stoats show a preference for
areas with high prey density and are stereotypic in their habitat selection (e.g.,
Aunapuu and Oksanen 2003), while weasels were more habitat generalists.
Nevertheless, it is believed that habitat quality and selection for weasels is driven by
the abundance and distribution of prey and this largely influences their demography
with population fluctuations that in some cases may lead to almost local extinctions
(Santos-Reis 1989, Klemola 1999). Although they have a shared preference for rodents
(Elmeros 2006), their use of alternative preys allows stoats to coexist with weasels,
which are more efficient predators of rodents (Brandt and Lambin 2005, Hellstedt et al.
2006).
14
General Introduction
The steppe polecat is represented in Europe by two restricted populations
divided by the Carpathian mountains. They show a preference for steppes, in
particular dry grasslands as well as cultivated fields and population density depends
strongly on prey (primarily small mammals) availability (e.g., Lanszki and Heltai
2007). Similarly, marbled polecats are generally found in the dryer areas and
grasslands of south-eastern Europe, but they are threatened mainly due to habitat
destruction through extensive cultivation, hunting and reduction of available prey
(Ben-David 1991, Gorsuch and Larivière 2005).
The Western Polecat show association with wetlands (e.g., Mestre et al. 2007)
where they find high diversity and availability of prey (e.g., rodents, amphibians and
crayfish), although it is a wide ranging species that uses a high diversity of land uses
with cover (Virgós 2003, Zabala et al. 2005). Also showing a strong association with
water bodies is the European mink, a semi-aquatic species that is mainly found along
small and medium sized rivers and seldom leaves the corridor of the riparian forest
and their resting places above ground (under herbs or bushes) or in cavities between
tree roots, mainly in flooded environments (Palazon and Ruiz-Olmo 1998, Lodé et al.
2001, Fournier et al. 2007). It is endemic in Europe and has suffered from a major
decline since mid-19th century becoming extinct in most European countries. The
Eurasian otter is a unique water-obligate species with distribution is restricted to
aquatic environments, from coastal waters, to freshwater streams and dams with
available shelter (Trindade et al. 1998, Barbosa et al. 2003); although the species
underwent a widespread decline during the 20th century, there are evidences of
recovery in Europe (Mitchell-Jones et al. 1999); its. Water quality and human density
were considered to be relevant causes of changes in otter distribution (e.g., Cortés et al.
1998).
Brown bear and lynxes (except the Iberian lynx as described in chapter 1.1.3.1)
are now restricted to mountain ranges where they find forest cover and low human
presence (e.g., Breitenmoser et al. 2000, Klober and Adamic 2000). Human-caused
mortality and habitat destruction are the main factors driving the decline of their
populations. Stone martens, pine martens, European wildcats are wide-ranging and
forest-dwelling species (e.g., Blanco 1998, Santos-Reis & Petrucci-Fonseca 1999, SantosReis et al. 2004), that show sensitivity to forest fragmentation in different degrees.
15
General Introduction
European wildcats suffered a significant decline mainly due to habitat destruction,
combined with prey decline (mostly rabbits in the south), human persecution and, to a
small extent, hybridization with feral cats (Stahl and Artois 1991, Oliveira et al. 2007).
Wildcats are found in a scrub–pastureland mosaic landscape with abundant
watercourses and prey and high shrub cover at a microhabitat scale in the
Mediterranean region (Lozano et al. 2003). The wildcat is a facultative specialist; rabbits
are preferred when present, but rodents constitute most of their diet when rabbits are
scarce or absent (Moleón and Gil-Sanchez 2003, Lozano et al. 2006, Sarmento 1996).
Similarly, stone and pine martens are generalists predators distributed across much of
Europe and habitat selection depends on prey availability and cover (e.g., Clevenger
1994, Virgós and Garcia 2002, Santos-Reis et al. 2004, Mortelliti and Boitani 2008), but a
combination of factors, such as persecution, trapping, and habitat loss have led to
sharp declines in the pine marten numbers and range.
Although habitat and food generalist, the golden jackal is, also, particularly
tolerant of dry habitats and has a mixed social organization (mated pair but sometimes
develop social group of adult offspring of both sexes). Once widespread the European
population declined dramatically over the first half of the 20th century, due to habitat
fragmentation and intensive hunting pressure. Thanks to effective conservation efforts,
the density of the Balkan population increased at the end of the last century, especially
in Bulgaria and Greece (Giannatos 2004). Another canid, the wolf, is a truly social
species. In southern Europe, due to intensive direct persecution and major habitat
changes, particularly in forest cover and natural prey availability (e.g., roe deer
Capreolus capreolus, red deer Cervus elaphus), they are nowadays restricted to mountain
areas (e.g. Blanco and Cortés 2002, Pimenta et al. 2005). Eurasian badgers also live in
social groups (e.g., Revilla and Palomares 2005). The availability of suitable sites for
setts is the primary limiting habitat factor for Eurasian badgers in “montado”
landscapes (Rosalino et al. 2005). In Southern Portugal, Rosalino et al. (2008) also
reported that oak woodland with a brushy understorey was the most used habitat at
local (1 km2) and landscape scales (100 km2).
A wish of new fur animal or the need for biological control of pest animal are
the most common reasons behind introductions of carnivores in the Mediterranean
basin (Kauhala 1996). The introduction and spread of the exotic species is a problem to
16
General Introduction
the conservation of native species, because exotic compete with them and may limit
their ability to persist in humanized landscapes. The raccoon dog is an East Asian
species introduced into the European part of the former USSR in the first half of the
20th century because it was considered a valuable fur animal, whereas the North
American raccoon, was introduced in the same period in Germany (Mitchell-Jones et al.
1999). Both species are habitat and feeding generalists but prefer the proximity of water
and are often found along lakes and streams (Sutor 2008). The American mink is native
to north America and was introduced by escapes from fur farms; it is a semi-aquatic
and generalist predator and is still spreading and colonising new areas (e.g., Melero et
al. 2007). The numbers of American minks and raccoon dogs increased rapidly and
today are widely distributed in Europe. Introductions of mongooses and genet,
originally from Africa, are traced to historic times when they were brought by Romans
as domestic animals to help in eliminating rodents and snakes in anthropic settlements
(Riquelme-Cantal et al. 2008). The Egyptian mongoose is still expanding towards north
(Borralho et al. 1996). They forage in group showing a wide food spectrum diet that
includes rabbits, rodents, birds, reptiles and carrion (Palomares 1993, Zapata et al.,
2007). Two other mongooses originally found across Southeast Asia, the small Indian
mongoose and the Indian grey mongoose (both solitary species), have being recently
introduced (1920’s and 1930’s, respectively) in former Yugoslavia and Italy,
respectively (Macdonald and Barret 1993). Because of their diurnal character and
human persecution, mongooses avoid open areas and select abundant vegetation cover
areas with high density of prey and shelter (e.g., Palomares and Delibes 1993). As in
the case of Egyptian mongoose, genets were introduced in historical times (RiquelmeCantal et al. 2008); they are generalist foragers and habitat selection also depends on
cover (Rosalino and Santos-Reis 2002, Galantinho and Mira 2008). In the southwestern
Europe, genets overlap extensively with stone martens in their distribution ranges and
resource preferences (Santos-Reis et al. 2004), although genets are associated with
warmer areas and therefore, mostly found in lower altitudes (Virgós et al. 2001). The
tolerance between the species is however greater than between conspecifics and there
is no evidence of intensive exploitation of the same key resources (Santos-Reis et al.
2004, Barrientos and Virgós 2006).
Although the majority of carnivores are widespread in Mediterranean Europe,
some occur in small and fragmented populations (e.g., brown bear, wildcat), or others
17
General Introduction
in low densities (e.g., wildcat) or are specialized in habitat (e.g., pine marten) and most
live in low densities due to higher spatial requirements and dependence of fluctuating
resources (e.g., Sunquist and Sunquist 2001, Loureiro et al. in press). The habitat loss
and fragmentation are therefore major causes of carnivore population decline, and
hence, addressing the vulnerability of the most common and wide-ranging species in
the face of growing human pressure that causes loss of landscapes connectivity is a
critical need.
1.1.3.2 Habitat fragmentation and road impacts on Mediterranean carnivores
Several studies have demonstrated the impact of landscape pattern and
composition on carnivores’ distribution or abundance in Mediterranean landscapes
(e.g., Virgós 2001, Virgós 2002, Virgós and Garcia 2002, Virgós et al. 2002, Zabala et al.
2005, Galantinho and Mira 2008, Mortelliti and Boitani 2008). In those studies, patch
use was affected by vegetation type, size, isolation, and geographic location, at
different scale extents ranging from local (e.g., shelter, refuge and prey availability,
inter-specific competition) to regional (e.g., distance between patches, matrix
composition). Virgós et al. (2002) in North and Central Spain reported that five
carnivore species (red fox, stone marten, Eurasian badger, common genet and
European wildcat) responded more favorably to fragments larger than 300ha than to
smaller fragments but their sensitivity varied among the species: Eurasian badger and
European wildcat seemed to have a clear response to area reduction and red fox was
the species least affected by fragment size. In a different analysis, Virgós and Garcia
(2002) documented that both patch size and isolation affected the stone marten
distribution. Additionally, Mortelliti and Boitani (2008) showed that when resource
amount was increased, suitability of smaller and isolated patches increased for stone
marten.
Equally, there is evidence of carnivore vulnerability to road network
expansions; both mortality and barrier effects were reported (Ferreras et al. 1992,
Forman et al. 2003, Iuell et al. 2003). In Portugal, vehicle collisions accounted for 40% of
Iberian wolf mortality (Pimenta et al. 2005) and accounts for ~10% of the endangered
Iberian wolf population in the southern Douro river area (Petrucci-Fonseca 1990, Grilo
18
General Introduction
et al. 2004). Ferreras et al. (1992) reported that road-kills accounted for 17% of mortality
of Iberian lynx in Southern Spain; with increases in road expansion and traffic
intensity, accidental deaths caused by vehicles are expected to increase the lynx
mortality on the road. In Britain, traffic is believed to kill more than 40% of the
estimated adult Eurasian badger population annually (Clarke et al. 1998). Over the past
decades, a growing number of studies have described small- and medium-sized
carnivore road mortality in Portugal but most data is just available in unpublished
reports or was presented in scientific meetings, and are restricted in time and space, or
focus on a few species (e.g., Marques 1994, Franco 2000, Ascensão 2001, Ascensão 2005,
Ascensão and Mira 2005, Petronilho and Dias 2005). Nevertheless, a large number of
road mortality studies worldwide have already stressed the importance of several
temporal (e.g., Orlowski and Nowak 2006) and spatial descriptors (e.g., Clevenger et al.
2003, Malo et al. 2004, Ramp et al. 2005) to explain road mortality rates. Differences in
species behavior and activity patterns, related to biological periods, account for
temporal variations in road-kills (e.g., Saeki and Macdonald 2004) while population
density, habitat and landscape structure, species biology, as well as road and traffic
characteristics, are the descriptors that seem to influence the spatial pattern of
mortality (Clevenger et al. 2003, Malo et al. 2004), suggesting that not all carnivores
respond equally to these factors.
Although road mortality is the most visible impact, less is known about the
extent of the barrier effect caused by road avoidance (McLellan and Shackleton 1988,
Lovallo and Anderson 1996, Whittington et al. 2004, Alexander et al. 2005; Jaeger et al.
2005). In a few mammal species (e.g., desert bighorn sheep Ovis canadensis nelsoni,
bobcat Lynx rufus and coyote Canis latrans), there is evidence that highways have
blocked movement of individuals between road sides affecting reproductive ability
and accelerating the loss of genetic diversity (Epps et al. 2005, Riley et al. 2006,
Strasburg 2006). In the Iberian Peninsula, the knowledge of barrier effect of roads on
carnivores is scarce and controversial; road density impacted fox and stone marten
movements (Blanco 1988, Lopéz-Martin et al. 1997) resulting in home-range changes,
but in contrast, Blanco et al (2005), suggest that highways were not an important barrier
for wolves in North-Central Spain. This evidence was not confirmed over an 18 km
highway section that transverse the Iberian wolf population living south to Douro river
19
General Introduction
(Roque et al. 2005): only once over five months a wolf used one of the 10 under- and
overpasses surveyed.
Existing crossing structures under and over roads and highways have been
extensively studied to evaluate their effectiveness in restoring landscape permeability
for some small and medium-sized carnivores (e.g., Yanes et al. 1995, Rosell et al. 1997,
Rodríguez et al. 1997, Mata et al. 2005, Ascensão and Mira 2007, Mata et al. 2008). All
studies documented that carnivores focused their crossing activity in passages that
varied in structure and surrounding habitat characteristics, road-related attributes,
human disturbance levels, and differences in use may be related with life-traits and
behavioral patterns (e.g., Rodriguez et al. 1997, Mata et al. 2005, Ascensão and Mira
2007, Mata et al. 2008). However, there is also little consensus on whether the type of
passage significantly influences crossing rates. Rosell et al. (1997) found no significant
differences related to size of passage and Mata et al. (2005) reported that the majority of
carnivores preferred larger structures. Yanes et al. (1997) and Ascensão and Mira
(2007), however, showed evidences of carnivore species using small culverts regularly.
2. AIMS AND THESIS STRUCTURE
Although the majority of carnivores inhabiting southern Portugal are of no
conservation concern, the recent increase in human presence in the area poses new
threats that should be evaluated in order to minimize those effects and promote longterm populations viability. Many relevant questions related with human-induced
changes on the carnivore guild remain unanswered, namely the response of these new
top carnivores to habitat loss and the expansion of the road network are still poorly
understood.
In this thesis, I provide a general description of human impacts on the smalland medium sized Mediterranean carnivore guild. More specifically, I address the
effect of scale on carnivore sensitivity to fragmentation; describe the patterns of road
mortality and their causes; analyse the response of the carnivores to the culverts and
underpasses on the highways; and assess the habitat connectivity for the most
vulnerable species to fragmentation.
20
General Introduction
To accomplish these goals I used scent stations to evaluate the “montado”
patch/site occupancy by carnivores, monitored roads and highways to estimate
carnivore road mortality and used infra-red digital cameras and marble dust to
evaluate underpasses and culverts use by carnivores. To analyse the data I applied
statistical modelling and developed spatially explicit models to understand how
species interact with landscape composition and configuration.
I organized the thesis as four independent papers (Chapters 3 to 5) plus this general
introduction (Chapter 1), the study area description (Chapter 2) and a summary of the
main findings of the present study supporting the discussion of its implications to
conservation planning (Chapter 6).
In Paper I, I analyzed the sensitivity of the five common meso-carnivores to
“montado” fragmentation and evaluate at what spatial scale extent the
environmental variables are expressed (Grilo C, Cruz M, Bissonette JA,
Santos-Reis M (submitted) Scale-dependent effects of Mediterranean
landscapes on meso-carnivore occupancy patterns. Landscape Ecology).
Paper II provides a general description of the temporal and spatial patterns of
each carnivore species road-kills (Grilo C, Bissonette JA, Santos-Reis M (2008)
Spatial-temporal
patterns
in
Mediterranean
carnivore
casualties:
consequences for mitigation. Biological Conservation, 142, 301-313).
In Paper III, I determined how often carnivores used crossing structures and
assessed whether differences in passage use were related to structure size or to
other structural and environmental variables associated with each structure
(Grilo C, Bissonette JA, Santos-Reis M (2008) Response of carnivores to
existing highway culverts and underpasses: implications for road planning
and mitigation. Biodiversity and Conservation, 17, 7, 1685-1699).
In Paper IV, I examine the extent to which different levels of landscape
connectivity influence stone marten occupancy of “montado” areas and the
relation between “montado” connectivity and road mortality risk (Grilo C,
21
21
General Introduction
Ascensão F, Bissonette JA, Santos-Reis M (submitted). Structural connectivity
in a Mediterranean landscape and its consequences for road mortality risk of
stone marten. Landscape and Urban Planning).
3. REFERENCES
Adriaensen F, Chardon JP, De Blust G, Swinnen E, Villalba S, Gulinck H, Matthysen E (2003) The
application of 'least-cost' modelling as a functional landscape model. Landscape and Urban
Planning, 64, 233-247.
Alexander SM, Waters NM, Paquet PC (2005) Traffic volume and highway permeability for a
mammalian community in the Canadian Rocky Mountains. Canadian Geographer, 49, 321-331.
Allen HD (2003) Response of past and present Mediterranean ecosystems to environmental change
Progress in Physical Geography, 27, 359-377.
Andrén H (1994) Effects of Habitat Fragmentation on Birds and Mammals in Landscapes with
Different Proportions of Suitable Habitat: A Review. Oikos, 71, 355-366
Anzures-Dadda A, Manson RH (2007) Patch- and landscape-scale effects on howler monkey
distribution and abundance in rainforest fragments. Animal Conservation, 10, 69–76.
Ascensão F (2001) Mortalidade de vertebrados por atropelamento em estradas do Alto Alentejo.
Relatório de estágio para obtenção da licenciatura em Biologia. FCUL. Lisboa.
Ascensão F (2005) Análise de estudos sobre mortalidade e vertebrados por atropelamento e o uso
de passagens hidráulicas por vertebrados. Tese de Mestrado. Universidade de Évora.
Ascensão F, Mira A (2005) Spatial patterns of road kills: a case study in Southern Portugal In: Irwin
CL, Garrett P, McDermott KP. (Eds.) Proceedings of the 2005 ICOET Center for
Transportation and the Environment, North Carolina State University, Raleigh, NC.
Ascensão F, Mira A (2007) Factors affecting culvert use by vertebrates along two stretches of road
in southern Portugal. Ecological Research, 22, 57-66.
Atauri J, Lucio JV (2001) The role of landscape structure in species richness distribution of birds,
amphibians, reptiles and lepidopterans in Mediterranean landscapes. Landscape Ecology, 16,
147-159.
August P, Iverson L, Nugranad J (2002) Habitat conversion of terrestrial habitats. In: Gutzwiller KJ
(Ed.). Applying Landscape Ecology in Biological Conservation. Springer.
Aunapuu M, Oksanen T (2003) Habitat selection of coexisting competitors: a study of small
mustelids in northern Norway. Evolutionary Ecology, 17, 371–392.
Baguette M, Dyck H (2007) Landscape connectivity and animal behavior: functional grain as a key
determinant for dispersal. Landscape Ecology, 22, 1117-1129.
Banks SC, Lindenmayer DB, Ward SJ, Taylor AC (2005) The effects of habitat fragmentation via
forestry plantation establishment on spatial genotypic structure in the small marsupial
carnivore, Antechinus agilis Molecular Ecology, 14, 1667–1680.
22
22
General Introduction
Barbosa AM, Real R, Olivero J, Vargas JM (2003) Otter (Lutra lutra) distribution modeling at two
resolution scales suited to conservation planning in the Iberian Peninsula. Biological
Conservation, 114, 377–387.
Barrientos R, Virgós E (2006) Reduction of potential food interference in two sympatric carnivores
by sequential use of shared resources. Acta Oecologia, 30, 107-116.
Bascompte J, Possingham H, Roughgarden J (2002) Patchy Populations in Stochastic Environments:
Critical Number of Patches for Persistence. The American Naturalist, 159, 128-137.
Beaudry F, deMaynadier PG, Hunter Jr. ML (2008) Identifying road mortality at multiple spatial
scales for semi-aquatic turtles. Biological Conservation, 141, 2550-2563.
Belovsky GE, Bissonette JA, Dueser RD, Edwards Jr TC, Luecke CM, Ritchie ME, Slade JB, Wagner
FH (1994) Management of small populations: concepts affecting the recovery of endangered
species. Wildlife Society Bulletin, 22, 307-316.
Ben-David M, Pellis SM, Pellis VC (1991) Feeding Habits and Predatory Behaviour in the Marbled
Polecat (Vormela Peregusna Syriaca): I. Killing Methods in Relation To Prey Size and Prey
Behaviour. Behaviour, 118, 127-143.
Bennett AF (1999) Linkages in the Landscape: The Role of Corridors and Connectivity in Wildlife
Conservation. IUCN, Gland, Switzerland and Cambridge.
Bissonette JA, Harrison DJ, Hargis CD, Chapin TG (1997) Scale-sensitive properties influence
marten demographics. Chapter 15 In: Bissonette JA (Ed.). Wildlife and landscape ecology: effects
of pattern and scale. Springer-Verlag, New York.
Bissonette JA, Storch I (Eds.) 2003. Landscape Ecology and Resource Management. Island Press,
Washington, D.C., USA.
Blanco JC (1988) Estudio ecologico del zorro, Vulpes vulpes (L, 1758), en la sierra de Guadarrama. Tese
doctoral, Universidade Oviedo.
Blanco JC (1998) Mamíferos de España. Vol. I. GeoPlaneta, Barcelona.
Blanco JC, Cortés Y (2002) Ecología, censos, percepción y evolución del lobo en España: análisis de un
conflicto. Sociedad Española para la Conservación y Estudio de lo Mamíferos, Málaga, Spain.
Blanco JC, Cortés Y, Virgós E (2005) Wolf response to two kinds of barriers in an agricultural
habitat in Spain. Canadian Journal of Zoology, 83, 312-323.
Blondel J (2006) The Design of Mediterranean Landscapes: A Millennial Story of Humans and
Ecological Systems during the Historic Period. Human Ecology, 34, 713-729.
Blondel J, Aronson J (1999) Biology and Wildlife of the Mediterranean region. Oxford University Press:
Oxford.
Borralho R, Rego F, Palomares F, Hora A (1996) The distribution of the Egyptian Mongoose
Herpestes ichneumon (L.) in Portugal. Mammal Review, 26, 11-8.
23
23
General Introduction
Brandt MJ, Lambin X (2005) Summertime activity patterns of common weasels Mustela nivalis
vulgaris under differing prey abundances in grassland habitats. Acta Theriologica, 50, 67-79.
Breitenmoser U, Breitenmoser -Wursten CH, Okarma H, Kaphegyi T, Kaphegyi -Wallmann U,
Muller U (2000) Action Plan for the Conservation of the Eurasian lynx in Europe (Lynx lynx).
Nature and environment No. 112, Council of Europe Publishing, Strasbourg: 1-83.
Brotons L, Mönkkönen M, Martin JL (2003) Are Fragments Islands? Landscape Context and
Density-Area Relationships in Boreal Forest Birds. The American Naturalist, 162, 343–357.
Brown JH, Lomolino MV (2000) Concluding remarks: historical perspective and future of island
biogeography theory. Global Ecology and Biogeography, 9, 87-92.
Burkey TV, Reed DH (2006) The effects of habitat fragmentation on extinction risk: Mechanisms
and synthesis. Songklanakarin. Journal of Science and Technology, 28, 9-37.
Cale P (2003) The influence of social behaviour, dispersal and landscape fragmentation on
population structure in a sedentary bird. Biological Conservation, 109, 237-248.
Cardillo M, Purvis A, Sechrest W, Gittleman JL, Bielby J, Mace GM (2004) Human Population
Density and Extinction Risk in the World’s Carnivores. PLoS Biology, 2, 909-914.
Carroll C, Noss RF, Paquet PC (2001) Carnivores as focal species for conservation planning in the
Rocky mountain region. Ecological Applications, 11, 961-980.
Carvalho JC, Gomes P (2001) Food habits and trophic niche overlap of the red fox, european wild
cat and common genet in the Peneda-Gerês National Park. Galemys, 13, 39-48.
Cavallini P, Lovari S (1991) Environmental Factors influencing the use of habitat in the Red Fox,
Vulpes vulpes (L., 1758). Journal of Zoology, 223, 323–339.
Ceballos G, Helrich P (2002) Mammal Population Losses and the Extinction. Science, 296, 904-907.
Chapin III FS, Zaveleta ES, Eviner VT, Naylor RL, Vitousek PM, Lavorel S, Reynolds HL, Hooper
DU, Sala OE, Hobbie SE, Mack MC, Diaz S (2000) Consequences of changing biotic
diversity. Nature, 405, 234-242.
Chase M, Kristan K, William B, Lynam A J, Price MV, Rotenberry JT (2000) Single Species as
Indicators of Species Richness and Composition in California Coastal Sage Scrub Birds and
Small Mammals Conservation Biology, 14, 474-487.
Chetkiewicz CB, Cassady St. Clair C, Boyce MS (2006) Corridors for Conservation: Integrating
Pattern and Process. Annual Review of Ecology, Evolution, and Systematics, 37, 317-342.
Clarke GP, White PCL, Harris S (1998) Effects of roads on badger Meles meles populations in southwest England. Biological Conservation, 86, 117-124.
Clevenger AP (1994) Habitat characteristics of Eurasian pine martens Martes martes in an insular
Mediterranean environment. Ecography, 17, 257-263.
Clevenger AP, Chruszcz B, Gunson K (2001) Drainage culverts as habitat linkages and factors
affecting passage by mammals. Journal of Applied Ecology, 38, 1340-1349.
Clevenger AP, Chruszcz B, Gunson KE (2003) Spatial patterns and factors influencing small
vertebrate fauna vehicle-kill aggregations. Biological Conservation, 109, 15-26.
24
24
General Introduction
Colchester M (1998) Europe and the world’s forests (1998) Underlying Causes of Deforestation and Forest
Degradation. Synthesis Report of the European Regional Meeting (Bonn 28-29 October 1998).
Cortés Y, Fernández-Salvador R, García FJ, Virgós E, Llorente M (1998) Changes in otter (Lutra
lutra) distribution in central Spain during the 1984-1995 period. Biological Conservation, 86,
179-183.
Crooks KR, Soulé ME (1999) Mesopredator release and avifaunal extinctions in a fragmented
system. Nature, 400, 563-566.
Crooks KR (2002) Relative Sensitivities of Mammalian Carnivores to Habitat Fragmentation.
Conservation Biology, 16, 488-502.
Cushman SA, McGarigal K (2004) Hierarchical analysis of forest bird species-environment
relationships in the Oregon Coast Range. Ecological Applications, 14, 1090-1105.
Daily G (1999) Developing a Scientific Basis for Managing Earth's Life Support Systems. Ecology
and Society, 3, 2-14.
Delibes M, Rodriguez A, Ferreras P (2000) Action Plan for the conservation of the Iberian Lynx (Lynx
pardinus) in Europe. Convention on the Conservation of European Wildlife and Natural
Habitats (Bern Convention) Nature and environment, No. 111.
Delibes-Mateos M, Delibes M, Ferreras P, Villafuerte R (2008) Key Role of European Rabbits in the
Conservation of the Western Mediterranean Basin Hotspot. Conservation Biology, 22, 1106–
1117.
Delibes-Mateos M, Fernandez de Simon J, Villafuerte R, Ferreras P (2008) Feeding responses of the
red fox (Vulpes vulpes) to different wild rabbit (Oryctolagus cuniculus) densities: a regional
approach. European Journal of Wildlife Research, 54, 71–78.
Di Castri F (1981) Mediterranean-type shrublands of the world. In: Di Castri F, Goodall W, Specht
RI (Eds.), Ecosystems of the World II: Mediterranean-type Shrublands. Elsevier. New York.
Donald PF, Sanderson FJ, Burfield IJ, Bierman SM, Gregory RD, Waliczky Z (2007) International
Conservation Policy Delivers Benefits for Birds in Europe. Science, 317, 5839, 810 – 813.
Donovan, TM, Flather CH (2002) Relationships among North American songbird trends, habitat
fragmentation, and landscape occupancy. Ecological applications, 12, 364-374.
Drielsma M, Manion G, Ferrier S (2007) The spatial links tool: automated mapping of habitat
linkages in variegated landscapes. Ecological Modelling, 200, 403-411.
Duffy JE, Cardinale BJ, France KE, McIntyre PB, Thébault E, Loreau M (2007) The functional role of
biodiversity in food webs: Incorporating trophic complexity. Ecology Letters, 10, 522-538.
Dunning JB Jr, Borgella JR, Clements K, Meffe GK (1995) Patch isolation, corridor effects, and
colonization by a resident sparrow in a managed pine woodland. Conservation Biology, 9,
542–550.
Elmeros M (2006) Food habits of stoats Mustela erminea and weasels Mustela nivalis in Denmark.
Acta Theriologica, 51, 179-186.
Eigenbrod F, Hecnar SJ, Fahrig L (2008) Accessible habitat: an improved measure of the effects of
habitat loss and roads on wildlife populations. Landscape Ecology, 23, 159-168.
25
25
General Introduction
Epps CW, Palsbøll P, Wehausen JD, Roderick G, Ramey II R., McCullough DR (2005) Highways
block gene flow and cause a rapid decline in genetic diversity of desert bighorn sheep.
Ecology Letters, 8, 1029–1038.
Erritzoe J, Mazgajski TD, Rejtz L (2003). Bird casualties on European roads — a review. Acta
Ornithologica, 38, 77-93.
Estes JA, Palmisano JF (1974) Sea otters: their role in the structuring nearshore communities.
Science, 185, 1058-1060.
Ewers RM, Didham RK (2006) Confounding factors in the detection of species responses to habitat
fragmentation. Biological Reviews, 81, 117-142.
Fahrig L (2003) Effects of habitat fragmentation on biodiversity. Annual Review of Ecology, Evolution,
and Systematics, 34, 487-515.
Fahrig L (2007) Non-optimal animal movement in human-altered landscapes. Functional Ecology,
21, 1003-1015.
Fahrig L, Merriam G (1985) Habitat Patch Connectivity and Population Survival. Ecology, 66, 17621768.
Fedriani JM (1996) Dieta anual del zorro, Vulpes vulpes, en dos hábitats del Parque Nacional de
Doñana. Doñana Acta Vertebrata, 23, 143-152.
Fernandes JP (2000) Landscape ecology and conservation management – evaluation of alternatives
in a highway EIA process. Environmental Impact Assessment Review, 20, 665-680.
Ferraz SFB, Vettorazzi CA, Theobald DM, Ballester MVR (2005) Landscape dynamics of
Amazonian deforestation between 1984 and 2002 in central Rondônia, Brazil: assessment
and future scenarios. Forest Ecology and Management, 204, 67-83.
Ferreras P, Aldama JJ, Beltrán JF, Delibes M (1992) Rates and causes of mortality in a fragmented
population of Iberian lynx Felis pardina Temminck, 1824. Biological Conservation, 61, 197-202
Ferreras P, Gaona P, Palomares F, Delibes M (2001) Restore habitat or reduce mortality?
Implications from a population viability analysis of the Iberian lynx. Animal Conservation, 4,
265-274.
FitzGibbon SI, Putland DA, Goldizen AW (2007) The importance of functional connectivity in the
conservation of a ground-dwelling mammal in an urban Australian landscape. Landscape
Ecology, 22, 1513-1525.
Fontaine B, Bouchet P, Achterberg KV, Alonso-Zarazaga MA, Araujo R et al. (2007) The European
union’s 2010 target: Putting rare species in focus. Biological Conservation, 139, 167-185
Forman RTT (1995) Some general principles of landscape and regional ecology. Landscape Ecology,
10, 133-142.
Forman RTT, Alexander LE (1998) Roads and their major ecological effects. Annual Review of
Ecology and Systematics, 29, 207–231.
Forman RTT, Godron M (1986) Landscape ecology. John Wiley, New York, New York, USA.
Forman RTT, Sperling D, Bissonette JA, Clevenger A, Cutshall C, Dale V, Fahrig L, France R,
Goldman C, Heanue K, Jones J, Swanson F, Turrentine T, Winter T (2003) Road Ecology:
Science and Solutions. Island Press, Washington.
26
26
General Introduction
Fournier P, Maizeret C, Jimenez D, Chusseau JP, Aulagnier S, Spitz F (2007) Habitat utilization by
sympatric european mink Mustela lutreola and polecats Mustela putorius in south-western
France. Acta Theriologica, 52, 1-12.
Franco I (2000) Aspectos ecológicos da mortalidade de vertebrados em rodovias do interior
Alentejano. Relatório de estágio para obtenção da licenciatura em Biologia. FCUL. Lisboa.
Frankham R, Ballou JD, Briscoe DA (2002) Introduction to Conservation Genetics. Cambridge
University Press.
Galantinho A, Mira A (2008) The influence of human, livestock, and ecological features on the
occurrence of genet (Genetta genetta): a case study on Mediterranean farmland. Ecological
Research DOI 10.1007/s11284-008-0538-5
Gaston KJ, Jackson SF, Nagy A, Cantú-Salazar L, Johnson M (2008) Protected Areas in Europe:
Principle and Practice. Annals of New York Academy of Science, 1134, 97–119.
Gehring TM, Swihart RK (2003) Body size, niche breadth, and ecologically scaled responses to
habitat fragmentation: mammalian predators in an agricultural landscape. Biological
Conservation, 109, 283-295.
Gehrt SD, Prange S (2006) Interference competition between coyotes and raccoons: a test of the
mesopredator release hypothesis. Behavioral Ecology, 18, 204-214.
Gerlach G, Musolf K (2000) Fragmentation of landscape as a cause for genetic subdivision in Bank
Voles. Conservation Biology, 14, 1066-1074.
Giannatos G (2004) Conservation Action Plan for the Golden Jackal Canis aureus L. in Greece. WWF
Greece, Athens.
Gilpin ME, Soulé ME (1986) Minimum viable populations: processes of species extinction. In: Soulé
ME (Ed.) Conservation biology: the science of scarcity and diversity. Sinauer Associates Inc.,
Sunderland, Massachusetts.
Ginsberg JR (2001) Setting Priorities for Carnivore Conservation: What Makes Carnivores
Different? In: Macdonald D, Gittleman J, Wayne R (Eds) Carnivore Conservation. Cambridge
University Press.
Gittleman JL (1989) Carnivore behaviour, ecology, and evolution. London Chapman and Hall.
Goodwin BJ, Fahrig L (2002) How does landscape structure influence landscape connectivity?
Oikos, 99, 552-570.
Goméz-Limón J, Fernandez JV (1999) Changes in use and landscape changes preferences on the
agricultural-livestock landscapes of the central Iberian Península (Madrid, Spain). Landscape
and Urban Planning, 44, 165-175.
Gorsuch WA, Larivière S (2005) Vormela peregusn. Mammalian Species 779, 1-5.
Grilo C, Roque S, Rio-Maior H, Petrucci-Fonseca F (2004) The Isolated Wolf Population South of
Douro River: status and action lines for its recovery. WolfPrint 20, 13-15.
Hale ML, Lurz PWW, Shirley MDF, Rushton S, Fuller RM, Wolff K (2001) Impact of Landscape
Management on the Genetic Structure of Red Squirrel Populations. Science, 293, 5538, 2246–
2248.
27
27
General Introduction
Hames RS, Rosenberg KV, Lowe JD, Dhondt AA (2001) Site Reoccupation in Fragmented
Landscapes: Testing Predictions of Metapopulation Theory. Journal of Animal Ecology, 70,
182-190.
Hedrick, PW, Gilpin ME (1996) Genetic effective size of a metapopulation. In: Hanski I, Gilpin ME
(Eds.) Metapopulation biology: ecology, genetics and evolution. Academic Press, New York.
Hellstedt P, Sundell J, Helle P, Henttonen H (2006) Large-scale spatial and temporal patterns in
population dynamics of the stoat, Mustela erminea, and the least weasel, M. nivalis in Finland.
Oikos, 115, 286-298.
Henle K, Davies KF, Kleyer M, Margules C, Settele J (2004) Predictors of Species Sensitivity to
Fragmentation. Biodiversity and Conservation, 13, 207-251.
Hessburg PF, Agee JK (1993) An environmental narrative of Inland Northwest United States
forests, 1800–2000. Forest Ecology and Management, 178, 23-59.
Holland JD, Bert DG, Fahrig L (2004) Determining the Spatial Scale of Species' Response to Habitat.
Bioscience, 54, 3 227–233.
Horn DJ, Phillips ML, Koford RR, Clark WR, Sovada MA, Greenwood RJ (2005) Landscape
composition, patch size, and distance to edges: interactions affecting duck reproductive
success. Ecological Applications, 15, 1367–1376.
Hunter R, Fisher R, Crooks KR (2003) Landscape-level connectivity in coastal southern California
as assessed by carnivore habitat suitability. Natural Areas Journal, 23, 302-314.
Iuell B, Bekker GJ, Cuperus R, Dufek J, Fry G, Hicks C, HlavÁč V, Keller VB, Rossel C, Sangwine T,
Torslov N, Wandall BM (Eds.) (2003) Wildlife and Traffic: An European Handbook for Identifying
Conflicts and Designing Solutions. KNNV publishers, UK.
Jaeger JAG, Bowman J, Brennan J, Fahrig L, Bert D, Bouchard J, Charbonneau N, Frank K, Gruber
B, Toschanowitz KT (2005) Predicting when animal populations are at risk from roads: an
interactive model of road avoidance behaviour. Ecological Modelling, 185, 329–348.
Jepsen JU, Baveco JM, Topping CJ, Verboom J, Vós CC (2005) Evaluating the effect of corridors and
landscape heterogeneity on dispersal probability: a comparison of three spatially explicit
modelling approaches. Ecological Modelling, 181, 445-459.
Kauhala K (1996) Introduced carnivores in Europe with special references to central and northern
Europe. Wildlife Biology, 2, 197-204.
Kareiva P (1987) Habitat fragmentation and the stability of predator-prey interactions. Nature, 326,
388-390.
Kerley LL, Goodrich JM, Miquelle DG, Smirnov EN, Quigley HB, Hornocker MG (2002) Effects of
roads and human disturbance on Amur tigers. Conservation Biology, 16, 97 – 108.
Klemola T, Korpimaki E, Norrdahl K, Tanhuanpaa M, Koivula M (1999) Mobility and habitat
utilization of small mustelids in relation to cyclically fluctuating prey abundances. Annales
Zoologici Fennici, 36, 75-82.
Klober A, Adamic M (2000) Identifying brown bear habitat by a combined GIS and machine
learning method Ecological Modelling, 135, 291-300.
Kurki S, Nikula A, Helle P, Lindén H (2000) Landscape Fragmentation and Forest Composition
Effects on Grouse Breeding Success in Boreal Forests. Ecology, 81, 1985-1997.
28
28
General Introduction
Lanszki J, Heltai M (2007) Diet of the European polecat and the steppe polecat in Hungary.
Mammalian Biology, 72, 49-53.
LaRue MA, Nielsen CK (2008) Modelling potential dispersal corridors for cougars in midwestern
North America using least-cost path methods least-cost path methods. Ecological Modelling,
212, 372-381.
Laurance WF (2008) Theory meets reality: How habitat fragmentation research has transcended
island biogeographic theory. Biological Conservation, 141, 1731-1744.
Lin Z-S., Qia X-Z, Lib B-L (2005) Can best competitors avoid extinction as habitat destruction?
Ecological Modelling, 182, 107-112.
Lindenmayer DB, J Fisher (2006) Habitat fragmentation and landscape change – an ecological and
conservation synthesis. Island Press.
Lodé T, Cormier J-P, Jacques D (2001) Decline in endangered species as an indication of anthropic
pressures: the case of European mink Mustela lutreola western population. Environmental
Management, 28, 4 727-735.
Lopéz-Martin JM, Lampreave G, Ruiz-Olmo J (1997) Selección de habitat de la garduña (Martes
foina): importancia de las islas de vegetación en ecosistemas mediterráneos alterados. In:
Abstracts of the III Jornadas SECEM – I Jornadas Ibéricas sobre la Nutria. Castelló de
Empúries, Girona.
Lovallo MJ, Anderson EM (1996) Bobcat movements and home ranges relative to roads in
Wisconsin. Wildlife Society Bulletin, 24, 71-76.
Lozano J, Virgós E, Malo AF, Huertas DL, Casanovas JG (2003) Importance of scrub–pastureland
mosaics for wild living cats occurrence in a Mediterranean area: implications for the
conservation of the wildcat (Felis silvestris). Biodiversity and Conservation, 12, 921–935.
Lozano J, Moleón M, Virgós E (2006) Biogeographical patterns in the diet of the wildcat, Felis
silvestris Schreber, in Eurasia: factors affecting the trophic diversity. Journal of Biogeography,
33, 1076–1085.
MacArthur RH, Wilson EO (1967) The Theory of Island Biogeography. Princeton University Press.
Princeton, NJ.
Macdonald D, Barret P (1993) Mammals of Britain & Europe Collins Field Guide. Harper Collins
Publishers., UK
Mahan CG, Yahner RH (1999). Effects of forest fragmentation on behaviour patterns in the eastern
chipmunk (Tamias striatus). Canadian Journal of Zoology, 77, 1991–1997.
Malanson GP, Cramer BE (1999) Landscape Heterogeneity, Connectivity, and Critical Landscapes
for Conservation. Diversity and Distributions, 5, 27-39.
Malo, JE, Suárez F, Díez A (2004) Can we mitigate animal–vehicle accidents using predictive
models? Journal of Applied Ecology, 41, 701–710.
Mangas JG, Lozano J, Cabezas-Díaz S, Virgós E (2008) The priority value of scrubland habitats for
carnivore conservation in Mediterranean ecosystems. Biodiversity and Conservation, 17, 43-51.
Manley PN, Zielinski WJ, Schlesinsger MD, SR Mori (2004) Evaluation of a multiple-species
approach to monitoring species at the ecoregional scale. Ecological Applications, 14, 296–310.
29
29
General Introduction
Mannion AM (2008) The Tertiary–Quaternary Environmental History of the Mediterranean Basin:
The Background to Mediterranean Island Environments In: Vogiatzakis I, Pungetti G,
Mannion AM (eds) Mediterranean Island Landscapes 9 Springer Netherlands.
Mannion AM (2008) The Tertiary–Quaternary Environmental History of the Mediterranean Basin:
The Background to Mediterranean Island Environments. Mediterranean Island Landscapes, 9,
15-35.
Marques JS (1994) Vertebrados mortos por atropelamento EN118. In: Proceedings of Seminário sobre a
avaliação do impacte ambiental de projectos rodoviários. Espinho.
Mata C, Hervás I, Herranz J, Suárez F, Malo JE (2005) Complementary use by vertebrates of
crossing structures along a fenced Spanish motorway. Biological Conservation, 124, 397-405.
Mata C, Hervás I, Herranz J, Suárez F, Malo JE (2008) Are motorway wildlife passages worth
building? Vertebrate use of road-crossing structures on a Spanish motorway. Journal of
Environmental Management, 88, 407–415.
Mbora DNM, Meikle DB (2004) Forest fragmentation and the distribution, abundance and
conservation of the Tana river red colobus (Procolobus rufomitratus). Biological Conservation,
168, 67-77.
McLellan BN, Shackleton DM (1988) Grizzly bears and resource extraction industries: effects of
roads on behavior, habitat use, and demography. Journal of Applied Ecology, 25, 451-460.
McIntyre S, Barrett GW (1992) Habitat variegation, an alternative to fragmentation. Conservation
Biology, 6, 146-147.
McIntyre S, Hobbs RJ (1999) A framework for conceptualizing human impacts on landscapes and
its relevance to management and research models. Conservation Biology, 13, 1282-1292.
Melero Y, Palazón S, Gosàlbez J (2007) Morphological adaptation of an invasive American mink
population in Mediterranean areas of Spain. Acta Zoologica, 89 1, 47–51.
Mestre FM, Ferreira JP, Mira A (2007) Modelling the distribution of the European Polecat Mustela
putorius in a Mediterranean agricultural landscape. Revue d'Ecologie (La Terre et la Vie), 62,
35-47.
Miller B, Dugelby B, Foreman D, Martinez del Río C, R Noss, M Phillips, R Reading, ME Soulé, J
Terborgh, L Willcox (2001) The Importance of Large Carnivores to Healthy Ecosystems.
Endangered Species UPDATE, 18, 202-210.
Mitchell-Jones A J, Amori G, Bogdanowicz W, Krystufek B, Reijnders PJH, Spitzenberger E, Stubbe
M, Thissen JBM, Vohralik V, Zima J (1999) The Atlas of European Mammals. Academic Press,
San Diego, CA
Moilanen A, Nieminen M (2002) Simple connectivity measures in spatial ecology. Ecology 83, 11311145.
Moleon M, JM Gil-Sanchez (2003) Food habits of the wild cat (Felis silvestris) in a peculiar habitat:
the Mediterranean high. mountain. Journal of Zoology, 260, 17-22.
Morris DW (1987) Ecological Scale and Habitat Use. Ecology, 68, 362-369.
30
30
General Introduction
Morris DW (1992) Scales and costs of habitat selection in heterogeneous landscapes. Evolutionary
Ecology, 6, 412-432.
Mortelliti A, Boitani L (2008) Interaction of food resources and landscape structure in determining
the probability of patch use by carnivores in fragmented landscapes. Landscape Ecology, 23,
285-298.
Myers N, Mittermeier RA, Mittermeier CG, Fonseca GAB, Kent J (2000) Biodiversity hotspots for
conservation priorities. Nature, 403, 853-858.
Naveh Z, Dan J (1973) The human degradation of Mediterranean landscape in Israel. In: Di Castri
F, Mooney HA (Eds.) Mediterranean type ecosystems. Origin and structures Springer-Verlag.
Naveh Z (1993) Red books for threatened Mediterranean landscapes as an innovative tool for
holistic landscape conservation. Introduction to the western Crete Red Book case study.
Landscape and Urban Planning, 24 241-247.
Naves J, Wiegand T, Revilla E, Delibes M. (2003) Endangered Species Constrained by Natural and
Human Factors: the Case of Brown Bears in Northern Spain. Conservation Biology, 17, 12761289.
Nepstead D C, Verssimo A, Alencar A, Nobre C, Lima E, Lefebvre P, Schlesinger P, Potter C,
Moutinho P, Mendoza E, Cochrane M, Brooks V (1999) Large-scale impoverishment of
Amazonian forests by logging and fires. Nature, 398, 505–508.
Novacek MJ (Ed.) (2001) The Biodiversity Crisis - Losing What Counts. An American Museum of
Natural History.
Oliveira R, Godinho R, Randi E, Ferrand N, Alves PC, (2007) Molecular analysis of hybridisation
between wild and domestic cats (Felis silvestris) in Portugal: implications for conservation.
Conservation Genetics, 9, 1-11.
Olson D, Dinerstein E (1998) The Global 200: a representation approach to conserving the Earth’s
most biologically valuable ecoregions. Conservation Biology, 12, 502-515.
Orlowski G, Nowak L (2006) Factors influencing mammal road-kill in the agricultural landscape of
south-western Poland. Polish. Journal of Ecology 54, 283-294.
Orrock JL, Pagels JF, McShea WJ, Harper EK (2000) Predicting presence and abundance of a small
mammal species: the effect of scale and resolution. Ecological Applications, 10, 1356–1366.
Pakkala T, Kouki J, Tiainen J (2006) Top predator and interference competition modify the
occurrence and breeding success of a specialist species in a structurally complex forest
environment. Annales zoologici Fennici, 43, 137-164.
Palazon, S, Ruiz-Olmo J (1998) A preliminary study of the behaviour of the European mink Mustela
lutreola in Spain, by means of radiotracking. Symp. Zool. Soc. London 71, 93-105.
Palomares F (1993) Opportunistic feeding of the Egyptian mongoose, Herpestes ichneumon, in
southwestern Spain. Revue d'Ecologie (Terre and Vie), 48, 295-304.
Palomares F (2001) Vegetation structure and prey abundance requirement of the Iberian lynx:
implications for the design of reserves and corridors. Journal of Applied Ecology, 38, 9-18.
Palomares F, Delibes M (1993) Key habitats for Egyptian mongooses in Doñana National Park,
southwestern Spain. Journal of Applied Ecology, 30, 752-758.
31
31
General Introduction
Palomares F, Ferreras P, Fedriani JM, Delibes M (1996) Spatial relationships between lynx and
other carnivores in an area of south –western Spain. Journal of Applied Ecology, 33, 5-13.
Palomares F, Ferreras P, Travaini A, Delibes M (1998) Co-existence between Iberian lynx and
Egyptian mongooses: estimating interaction strength by structural equation modelling and
testing by an observational study. Journal of Animal Ecology, 67, 967-978.
Paquet PC (1993) Summary reference document - Ecological studies of recolonizing wolves in the central
Canadian Rocky Mountains. Unpublished report by John/Paul and Associates for the
Canadian Parks Service, Banff, Alberta.
Petronilho JMS, Dias CMM (2005) Impact of two forest roads upon wildlife after a road pavement
change in a coastal area in the Center of Portugal. Wildlife Biology in Practice, 1, 128-139.
Petrucci-Fonseca F (1990) O lobo (Canis lupus signatus Cabrera, 1907) em Portugal. Problemática da
sua conservação. Tese de Doutoramento. Universidade de Lisboa.
Plieninger T, Pulido FJ, Konold W (2003) Effects of land-use history on size structure of holm oak
stands in Spanish dehesas: implications for conservation and restoration. Environmental
Conservation, 30, 61-70.
Pimenta V, Barroso I, Álvares F, Correia J, Ferrão da Costa G, Moreira L, Nascimento J, PetrucciFonseca F, Roque S, Santos E (2005) Situação populacional do lobo em Portugal, resultados do
Censo Nacional 2002/2003. Relatório Técnico. Instituto da Conservação da Natureza/Grupo
Lobo.
Pimm SL, Askins RA (1995) Forest losses predict bird extinctions in eastern North America.
Proceedings of the National Academy of Sciences of the United States of America, 92, 9343–9347.
Pimm SL, Raven P (2000) Extinction by numbers. Nature, 403, 843-845.
Pimm S, Raven P, Peterson A, Şekercioğlu ÇH, Ehrlich PR (2006) Human impacts on the rates of
recent, present, and future bird extinctions. PNAS 103, 10941-10946.
Piñero JR (2002) Mamíferos Carnívoros Ibéricos (Iberian Mammalian Carnivores). Lynx Edicions.
Pinto-Correia T (1993) Threatened landscape in Alentejo, Portugal: the “montado” and other
“agro-silvo-pastoral” systems. Landscape and Urban Planning, 24, 43–48.
Quézel P, Médail F (2003) Ecologie et biogéographie des forêts du bassin méditerranéen, Elsevier,
France.
Ramp D, Caldwell J, Edwards KA, Warton D, Croft DB (2005) Croft a Modelling of wildlife fatality
hotspots along the Snowy Mountain Highway in New South Wales, Australia. Biological
Conservation, 126, 474–490.
Revilla E, Palomares F (2005) Patrones generales de organización social en el tejón eurasiático. In:
Virgós E, Revilla E, mangas JG, Domingo-Roura X (Eds.) Ecología y conservación del tejón en
ecosistemas mediterráneos. Sociedad Española para la Conservación y Estudio de los
Mamíferos SECEM.
Riley SPD, Pollinger JP, Sauvajot RM, York EC, Bromley C, Fuller TK, Wayne RK (2006) A southern
California freeway is a physical and social barrier to gene flow in carnivores. Molecular
Ecology, 15, 1733–1741.
32
32
General Introduction
Ripple WJ, Beschta RL (2006) Linking a cougar decline, trophic cascade, and catastrophic regime
shift in Zion National Park. Biological Conservation, 133, 397-408.
Riquelme-Cantal JA, Simón-Vallejo MD, Palmqvist P, Cortés-Sánchez M (2008) The oldest
mongoose of Europe. Journal of Archaeological Science, 35, 2471–2473
Rodriguez A, Andren H (1999) Comparison of Eurasian Red Squirrel Distribution in Different
Fragmented Landscapes. Journal of Applied Ecology, 36, 649-662.
Rodriguez A, Crema G, Delibes M (1997) Factors affecting crossing of red foxes and wildcats
through non-wildlife passages across a high-speed railway. Ecography, 20, 287-294
Rodriguez A, Delibes M (1992) Current range and status of the Iberian lynx Felis pardina
Temminck, 1824 in Spain. Biological Conservation, 61, 189-196.
Rodriguez A, Delibes M (2003) Population fragmentation and extinction in the Iberian lynx.
Biological Conservation, 109, 321-331.
Rodríguez A, Delibes M (2004) Patterns and causes of non-mortality in the Iberian lynx during a
40-year period of range contraction. Biological Conservation, 118, 151–161.
Rodriguez-Freire M, Crecente-Maseda R (2008) Directional connectivity of wolf (Canis lupus)
populations in Northwest Spain and anthropogenic effects on dispersal patterns.
Environmental Modeling and Assessment, 13, 35-51.
Roedenbeck I, Voser P (2008) Effects of roads on spatial distribution, abundance and mortality of
brown hare (Lepus europaeus) in Switzerland. European Journal of Wildlife Research, 54, 425-437.
Roque S, Espírito Santo C, Grilo C, Rio-Maior H, Petrucci-Fonseca F (2005) A população lupina a sul
do Rio Douro: análise temporal, atitudes publicas e corredores ecológicos. Relatório Final. Centro de
Biologia Ambiental, Faculdade de Ciências da Universidade de Lisboa.
Rosalino LM, Santos-Reis M (2002) Feeding habits of the common genet Genetta genetta (carnívora
Viverridae) in a semi-natural landscape of central Portugal. Mammalia, 65, 195-205.
Rosalino LM, Macdonald DW, Santos-Reis M (2005) Resource dispersion and badger population
density in Mediterranean woodlands: is food, water or geology the limiting factor? Oikos,
110, 441-452.
Rosalino LM, Santos MJ, Beier P, Santos-Reis M (2008) Eurasian badger habitat selection in
Mediterranean environments: does scale really matter? Mammalian Biology, 73, 189-198.
Rosell C, Parpal J, Campeny R, Jové S, Pasquina A, Velasco JM (1997) Mitigation of barrier effect of
linear infrastructures on wildlife. In: Canters K. (Ed.) Proceedings of the International
Conference on Habitat Fragmentation, Infrastructure and the Role of Ecological Engineering,
Maastricht and the Hague
Saeki M, Macdonald DW (2004) The effects of traffic on the raccoon dog (Nyctereutes procyonoides
viverrinus) and other mammals in Japan. Biological Conservation, 118, 559–571.
Santos MJ, Pedroso N, Ferreira JP, Matos H, Sales-Luís T, Pereira Í, Baltazar C, Grilo C, Cândido
AT, Sousa I, Santos-Reis M (2007) Assessing dam implementation impact on threatened
carnivores: the case of Alqueva in SE Portugal. Environmental Monitoring and Assessment, 142,
47-54.
33
33
General Introduction
Santos-Reis M (1989) As doninhas ibéricas (Carnivora: Mustela) um estudo taxonómico e ecológico. Tese
de Doutoramento, Universidade de Lisboa.
Santos-Reis M, Petrucci-Fonseca F (1999) Carnívoros. In: Mathias ML (Ed.) Mamíferos terrestres de
Portugal Continental Madeira e Açores ICN/CBA, Lisboa, Portugal.
Santos-Reis M, Santos MJ, Lourenço S, Marques T, Pereira I, Pinto B (2004) Chapter 7.
Relationships between stone martens, genets and cork oak woodlands in Portugal. In:
Harrison DJ, Fuller AK, Proulx G (Eds.) Marten and fishers in human altered landscapes. An
International perspective, Kluwer Academic Publishers, Massachusetts.
Sarmento P (1996) Feeding ecology of the European wildcat Felis silvestris in Portugal. Acta
Theriologica 41, 409-414.
Saunders DA, Hobbs R, Margules C (1991) Biological consequences of ecosystem fragmentation: A
review. Conservation Biology, 5, 18-32.
Saunders DA, Hobbs RJ, Arnold GW (1993) The Kellerberrin project on fragmented landscapes: A
review of current information. Biological Conservation, 64, 185-192.
Schnabel S, Ferreira A (Eds.) (2004) Sustainability of Agrosilvopastoral Systems – Dehesas,
“montado” Advances in GeoEcology 37. IUSS.
Sergio F, Caro T, Brown D, Clucas B, Hunter J, Ketchum J, McHugh K, Hiraldo F (2008) Top
Predators as Conservation Tools: Ecological Rationale, Assumptions, and Efficacy. Annual
Review of Ecology, Evolution, and Systematics, 39, 1–19.
Skole D, Tucker C (1993) Tropical deforestation and habitat fragmentation in the Amazon: Satellite
data from 1978 to 1988. Science, 260, 1905-09.
Soulé ME (1996) Conservation Biology: the science of scarcity and diversity. Sinauer Sunderland,
Massachusetts.
Stahl, P, Artois M (1991) Status and conservation of the wildcat (Felis silvestris) in Europe and around
the Mediterranean Rim. Nature and Environment 69.
Strasburg JL (2006) Roads and genetic connectivity. Nature, 440, 875-876.
Sunquist ME, F Sunquist (2001) Changing landscapes: consequences for carnivores In: Gittleman
JL, Funk SM, MacDonald DW, Wayne RK (Eds) Carnivore Conservation. Conservation
Biology 5. Cambridge.
Sutor A (2008) Dispersal of the alien raccoon dog Nyctereutes procyonoides in Southern
Brandenburg, Germany. European Journal Wildlife Research, 54, 321–326.
Swihart RK., Atwood TC, Goheen JR, Scheiman DM, Munroe KE, Gehring TM (2003) Patch
occupancy of North American mammals: is patchiness in the eye of the beholder? Journal of
Biogeography, 30 8, 1259-1279.
Swihart RK, Lusk JJ, Duchamp JE, Rizkalla CE, Moore JE (2006) The roles of landscape context,
niche breadth, and range boundaries in predicting species responses to habitat alteration.
Diversity and Distributions, 12, 277–287.
Taylor SK, Buergelt CD, Roelke-Parker ME, Homer BL, Rotstein DS (2002) Causes of mortality of
free-ranging Florida panthers. Journal of Wildlife Diseases, 38, 107–114.
34
34
General Introduction
Taylor PD, Merriam G (1995) Habitat fragmentation and parasitism of a forest damselfly. Landscape
Ecology, 11, 181-189
Taylor, PD, Fahrig L, With K (2006) Landscape connectivity: a return to the basics. In: Crooks KR,
Sanjayan M (Eds.) Connectivity Conservation. Cambridge, Cambridge University Press.
Thomas CD, Cameron A, Green RE, Bakkenes M, Beaumont LJ, Collingham YC, Erasmus BFN, de
Siqueira MF, Grainger A, Hannah L, Hughes L, Huntley B, van Jaarsveld AS, Midgley GF,
Miles L, Ortega-Huerta, MA, Peterson AT, Phillips OL, Williams SE (2004) Extinction risk
from climate change. Nature, 427, 145–48.
Tilman D, May RM, Lehman CL, MA Nowak (1994) Habitat destruction and the extinction debt.
Nature, 371, 65- 66.
Trindade A, Farinha N, Florêncio E (1998) A distribuição da lontra Lutra lutra em Portugal—situação
em 1995. ICN, Lisboa.
Trombulak SC, Frissel CA (2000) Review of ecological effects of roads on terrestrial and aquatic
communities. Conservation Biology, 14, 18-30.
Turner MG, Gardner RH, O'Neill RV (2003) Landscape Ecology in Theory and Practice: Pattern and
Process. Springer.
Uezu A, Metzger JP, Viellard JME (2005) Effects of structural and functional connectivity and patch
size on the abundance of seven Atlantic Forest bird species. Biological Conservation, 123, 507519.
Virgós E (2001) Role of isolation and habitat quality in shaping species abundance: a test with
badgers (Meles meles L.) in a gradient of forest fragmentation. Journal of Biogeography, 28, 381389.
Virgós E (2002) Are habitat generalists affected by forest fragmentation? A test with Eurasian
badgers (Meles meles) in coarse-grained fragmented landscapes of central Spain. Journal of
Zoology 258, 313–318.
Virgós E, Romero T, Mangas JG (2001) Factors determining “gaps” in the distribution of a small
carnivore, the common genet (Genetta genetta), in central Spain. Canadian Journal of Zoology,
79, 1544-1551.
Virgós E., Garcia FJ (2002) Patch occupancy by stone martens Martes foina in fragmented
landscapes of central Spain: the role of fragmented size, isolation and habitat structure. Acta
Oecologica, 23, 231-237.
Virgós E, Telleria JL, Santos T (2002) A comparison on the response to forest fragmentation by
medium-sized Iberian carnivores in central Spain. Biodiversity and Conservation, 11, 10631079.
Virgós E (2003) Association of the polecat Mustela putorius in eastern Spain with montane pine
forests. Oryx, 37, 4:484-487.
Vogiatzakis IN, Mannion AM, Griffiths GH (2006) Mediterranean Ecosystems: problems and tools
for
conservation. Progress in Physical Geography, 30, 175–200.
Walker RS, Andrés J, Novaro L., Branch C (2007) Functional connectivity defined through costdistance and genetic analyses: a case study for the rock-dwelling mountain vizcacha
(Lagidium viscacia) in Patagonia, Argentina. Landscape Ecology, 22, 1303-1314.
35
35
General Introduction
Walters S (2007) Modeling scale-dependent landscape pattern, dispersal and connectivity from the
perspective of the organism. Landscape Ecology, 22, 867-881.
Wang Y-H, Yang K-C, Bridgman CL, Lin L-K (2008) Habitat suitability modelling to correlate gene
flow with landscape connectivity. Landscape Ecology, 23, 989-1000.
Whittington J, St. Clair CC, Mercer G (2004) Path tortuosity and the permeability of roads and trails
to wolf movement. Ecology and Society, 9, 4.
Wiens JA (1976) Population responses to patchy environments. Annual Review of Ecology and
Systematics, 7, 81-120.
Wiens JA, Milne BT (1989) Scaling of "landscapes" in landscape ecology, or, landscape ecology
from a beetle's perspective. Landscape Ecology, 3, 87-96.
Wiegand T, Revilla E, Moloney KA (2005) Effects of habitat loss and fragmentation on population
dynamics. Conservation Biology, 19, 108-121.
Wiens JA (2002) Central concepts and issues in landscape ecology. In Gutzwiller KJ (Ed), Applying
Landscape Ecology in Biological Conservation, Springer-Verlag.
Wilcox BA, Murphy DD (1985) Conservation strategy: the effects of fragmentation on extinction.
The American Naturalist, 125, 879-887.
With KA, King AW (1999) Extinction thresholds for species in fractal landscapes. Conservation
Biology, 13, 314-326.
Wolff JO (1999) Behavioral Model Systems. In: Barrett GW, Peles JD (Eds.) Landscape Ecology of
Small Mammals. Springer-Verlag, New York.
Woodroffe R (2001) Strategies for carnivore conservation: lessons for contemporary extinctions. In
Gittleman JL, Funk SM, MacDonald DW, Wayne RK (Eds.) Carnivore Conservation.
Conservation Biology 5. Cambridge.
Wright S (1943) Isolation by distance. Genetics, 28, 139–156.
Yanes M, Velasco JM, Suaréz F (1995) Permeability of roads and railways to vertebrates: the
importance of culverts. Biological Conservation, 71, 217-222.
Zabala J, Zuberogoitia I, Martinez-Climent JA (2005) Site and landscape features ruling the habitat
use and occupancy of the polecat (Mustela putorius) in a low density area: a multiscale
approach. Wildlife Research, 51, 157-162.
Zapata SC, Travaini A, Ferreras P, Delibes M (2007) Analysis of trophic structure of two carnivore
assemblages by means of guild identification. European. Journal of Wildlife Research, 53, 276–
286.
36
36
CHAPTER 2
STUDY AREA
CHAPTER 2 - STUDY AREA
2.1 STUDY AREA
The study area was located within the Alentejo province (37º54’N, 8º06’ W) in
southern Portugal (Figure 1). Most of the Alentejo is under the influence of
Mediterranean climate. It is therefore characterized by long, dry and very hot summers
(stretching from May/June to September/October), where the temperatures often
attain 30-40ºC, and high annual temperature amplitude. The climate is also,
characterized by irregular distribution of rainfall over the wet season, as well between
years, with river flows showing very intense flood peaks (mainly between October and
March) and frequent drought periods in summer and early autumn (source: Instituto
de Metereologia).
Figure 1 – Land use in the study area.
39
39
Study area
This region occupies the south western part of the large Iberian structural unit
called Maciço Antigo (ancient massif), where metamorphic Palaeozoic formations
dominate. Characterised by extensive plains, elevation ranges from approximately 200
m to a maximum of 500 m (source: Atlas do Ambiente). Agricultural landscape
dominates the region and comprises a mosaic of “montado” (savannah-like woods of
cork oak Quercus suber and holm oak Q. ilex) patches of different range sizes associated
with extensive animal grazing of natural pastures and intermittent cereal cultivation
(oats, barley, and wheat) (Blondel 2006). Although scrublands are scarce, they are
composed by scrub oaks (Q. coccifera, Q. lusitanicus) and many other Mediterranean
shrubs (Arbutus unedo, Cistus sp., Lavandula stoechas, Rosmarinus officinalis, Ulex sp.) as
well as a wide diversity of herbaceous plants (Pinto Correia 1993).
Primary human activities include cork extraction (in the cork oak “montado”),
cattle husbandry (cattle, swine, sheep and goats), intermittent and extensive cereal
cultivation (oats, barley and wheat), intensive agriculture related with olive groves and
vineyards, and hunting (mainly wild rabbit, partridge Alectoris rufa and wild boar Sus
scrofa). The human population density (excluding the two major cities, Évora and Beja,
with 1980 inhabitants.km-2) is 21 inhabitants.km-2 and the road network density
averages 0.25km.km-2, about half of the mean national road network density of 0.49
km.km-2 (source: IgeoE).
Six Natura 2000 sites are included, at least partially, within the limits of the
survey area (Sado, Alvito, Cabrela, Monfurado, Caia and Moura-Barrancos - Figure 2)
but the lack of management strategies may limit their future conservation value.
Nevertheless, this region supports a large variety of wild fauna including some rare
and threatened species such as: birds (e.g., black stork Ciconia nigra, booted eagle
Hieraetus pennatus, montagu’s harrier Circus pygargus, goshawk Accipiter gentiles,
golden eagle Aquila crysaetus), bats (e.g., Mehely’s horseshoe bat Rhinolophus mehelyi)
and rodents (e.g., Cabrera’s vole Microtus cabrerae, garden dormouse Elyomis quercinus).
Out of the 23 carnivores living in the Mediterranean region (see Chapter
1.1.3.3), and of the 14 reported in Portugal (wolf, red fox, weasel, stoat, polecat,
American mink, pine marten, stone marten, badger, otter, genet, Egyptian mongoose,
wildcat and Iberian lynx), nine can be found in the study area (Santos-Reis and
Petrucci-Fonseca 1999). Exceptions are the wolf, regionally extinct in the 1970's 40
40
Study area
Petrucci-Fonseca (1990), and the stoat and the pine marten, two Eurosiberian species
that show a preference for colder regions and that were never reported south to river
Tejo (Santos-Reis 1985, Matos and Santos-Reis 2004); the American mink is a recent
invader in the NW of Portugal, having escaped from a mink farm close to river Minho
(Vidal-Figueroa 1987), and apparently stil did not reached Douro River.
Figure 2 – Sampling units with scent-stations and roads surveyed.
Of the species ocurring in the study area, the Iberian lynx is known to be in the
verge of extinction (Sarmento et al. 2008) and the wildcat shows a declining trend and
fragmented pattern of distribution (Cabral et al. 2005). The lack of information about
the polecat distribution, abundance and ecology in Portugal and the decreasing
number of observations, suggest a declining trend and promoted its classification as
potentially threatened in the last Red Book revision (Data Deficient - Cabral et al. 2005).
For the remaining, although known to be wideranging they are suspected to show
different ocupancy patterns and different sensitivities to “montado” and road-related
fragmentation.
41
41
Study area
Paper I and IV focused in a rectangular area of 7680 km2 which include the
sampling units with scent-stations and comprises 39% of “montado” contiguous areas
(n=3 with >35 000ha) and “montado” patches of different range sizes (n=1948; from 1
ha to 15 000 ha) interspersed within a cropland matrix (44%) (Figure 2). The remaining
area is occupied by production forest (8%), intensive vineyards and olive groves (7%)
and urban areas (2%).
As for papers II and III, the study was conducted along sections of two
highways under private concession of BRISA, Auto-Estradas de Portugal, S.A. (HW)
(A2 and A6 - 260km in total) and six segments of national roads (NR) administered by
the National Road Institute - IEP (N4/N114, N370, N380, IC1, IP2, IP8 - 314km in total)
(Figure 2). Both HWs had four-lanes with a median strip approximately 7m wide, and
with livestock exclusion fencing along both sides. NRs had one-lane in each direction
without a median strip and only a few sections are fenced for livestock exclusion.
Annual average daily traffic ranged between 330-2494 vehicles/night with a speed
limit of 120km.h-1 on the HWs, and between 500-2161 vehicles/night and with a speed
limit of 100km.h-1 on NRs. For approximately 100km, HWs and NRs run parallel to
each other and share the same adjacent vegetative cover.
2.2 REFERENCES
Blondel J (2006) The Design of Mediterranean Landscapes: A Millennial Story of Humans and
Ecological Systems during the Historic Period. Human Ecology, 34, 713-729.
Cabral MJ, Almeida J, Almeida PR, Dellinger T, Ferrand de Almeida N, Oliveira ME, Palmeirim JM,
Queiroz AI, Rogado L, Santos-Reis M (2005) Livro Vermelho dos Vertebrados de Portugal.
Instituto da Conservação da Natureza, Lisboa.
Matos H, Santos-Reis M (2006) Distribution and status of the pine marten in Portugal. In: Martes in
carnivore communities (Santos-Reis M, Birks JDS, O’Doherty EC, Proulx G (Eds.) Alpha
Wildlife Publications, Sherwood Park, Alberta, Canada. Pp 43-57.
Petrucci-Fonseca F (1990) O lobo (Canis lupus signatus Cabrera, 1907) em Portugal. Problemática da sua
conservação. Tese de Doutoramento. Universidade de Lisboa.
Santos-Reis M (1985) Mustela erminea Linnaeus, 1758: a new mustelid to Portugal. Mammalia, 49, 136138.
Santos-Reis M, Petrucci-Fonseca F (1999) Carnívoros. In: Mathias ML (Ed.), Mamíferos terrestres de
Portugal Continental Madeira e Açores ICN/CBA, Lisboa, Portugal.
42
42
Study area
Sarmento P, Cruz J, Monterroso P, Tarroso P, Ferreira C, Negrões N, Eira C (2008) Status survey of
the critically endangered Iberian lynx Lynx pardinus in Portugal. European Journal of Wildlife
Research DOI 10.1007/s10344-008-0240-5.
Vidal-Figueroa T, Delibes M (1987) Primeros datos sobre el visón Americano en el suroeste de
Galizia y Noroeste de Portugal. Ecologia, 1, 145-152.
43
43
CHAPTER 3
RESPONSE OF MEDITERRANEAN CARNIVORES TO
“MONTADO” FRAGMENTATION
Paper I
Scale-dependent effects of Mediterranean landscapes on mesocarnivore occupancy patterns.
Clara Grilo, Marta Cruz, John A. Bissonette, Margarida Santos-Reis
This paper was submitted to the journal Landscape Ecology on February 22, 2009.
Scale-dependent effects of Mediterranean landscapes on mesocarnivore occupancy patterns
Clara Grilo1, Marta Cruz2, John A. Bissonette3, Margarida Santos-Reis1
1University
of Lisbon, Environmental Biology Center / Departament of Animal Biology, Faculty of Sciences C2 5º floor
1749-016 Lisbon, Portugal. E-mail address: [email protected], [email protected]
2Rua
de Guerra Junqueiro, 474 4º dto 4150 Porto, Portugal. E-mail address: [email protected]
3U.S.
Geological Survey, Utah Cooperative Fish and Wildlife Research Unit, Department of Wildland Resources, College
of Natural Resources, Utah State University, USA, 84322-5290. E-mail address: [email protected]
Abstract
Our goal was to assess the effects of “montado” fragmentation on five meso-carnivores
(red fox Vulpes vulpes, stone marten Martes foina, Eurasian badger Meles meles, common genet
Genetta genetta and Egyptian mongoose Herpestes ichneumon) by addressing the following
questions: 1) Are these meso-carnivores sensitive to “montado” fragmentation? 2) Where does
the threshold effects in species occupancy are likely to occur? 3) Which environmental variables
explain occupancy at plot-(50 m), patch-(mean home range diameter) and landscape scale
extents (median dispersal distance)? and 4) Are there interactions between the contributions of
the explanatory variables at the three scale extents? Scent-station surveys were run to evaluate
occupancy of “montado” areas. Low detection probabilities were obtained for the Eurasian
badger and common genet and hence they were precluded from the analysis. We found that red
foxes were not sensitive to “montado” fragmentation, but stone marten and Egyptian mongoose
responded negatively to loss of vegetation cover and isolation. The multi-scale models revealed
that stone marten were the most sensitive species to “montado” attributes and fragmentation.
Stone marten models indicated a dominant independent contribution of habitat quality at plot
level followed by the joint effect of isolation and amount of habitat at patch and landscape level,
respectively. The major amount of variation in the Egyptian mongoose occupancy patterns was
provided by the joint of variables at three scales. We conclude that conservation efforts aimed at
improving local habitat quality for both species but with distinct levels of connectivity at patch
scale extent.
Keywords - “montado” woodlands, patch occupancy, red fox, stone marten and Egyptian
mongoose, landscape ecology, variance partitioning.
49
49
Response of mediterranean carnivores to "montado" fragmentation
Introduction
Landscapes of the Mediterranean Basin are a fragmented mixture of natural
and anthropogenically changed habitats (Plieninger 2006). One example is the
“montado” (“dehesa” in Spain), a landscape comprised of cork (Quercus suber) and
holm (Q. ilex) oak woodland patches interspersed with agricultural crops, fallow lands,
and pastures. This complex and dynamic system supports a diverse assemblage of
plants and animals, and is found exclusively in the western part of the Mediterranean
Basin Iberian Peninsula (Blondel 2006). Over the last century, the “montado” has
experienced
significant large-scale land
cover
changes,
including
landscape
modification, vegetation loss and deterioration, increased edge effects, and loss of
landscape connectivity (Plieninger 2006). In southern Portugal, during what was
known as the "Wheat Campaign" in the first half of the 20th century, very large areas of
forest and shrub cover were cleared, with about half of the region changed to a
cropland matrix (Pinto-Correia 1993). In the following decades, the Portuguese
government promoted reforestation programs that transformed some areas of seminatural marginal woodlands into artificial plantations populated by non-native conifer
and eucalyptus species.
Landscape fragmentation has been visualized as a process with four implied
effects: a reduction in the amount of habitat, an increase in the number of habitat
patches, a decrease in the sizes of habitat patches, and an increase in the isolation of
patches (e.g., Fahrig 2003). All fragmentation effects arguably are a major threat to
biodiversity and are often associated with population declines and increased risk of
extinction (e.g., Bascompte et al 2002). Nevertheless, the critical values of the amount of
habitat area and connectivity depends on the landscape pattern and on the processes
by which organisms move (With and Crist 1995), and these still are poorly studied and
a controversial issue (e.g., Andrén 1994; Hargis et al 1999).
Mammalian carnivores usually possess general characteristics (e.g., low
densities, large home ranges) that may make them particularly vulnerable to effects of
fragmentation (Sunquist and Sunquist 2001). The influence of landscape pattern and
composition on carnivores distribution or abundance has been demonstrated
elsewhere (e.g., Oehler and Litvaitis 1996; Crooks 2002; Wiegand et al 2005) and more
recently, in Mediterranean landscapes (e.g., Virgós et al 2002). Even though carnivores
50
50
Response of mediterranean carnivores to "montado" fragmentation
typically use a wide range of habitats, western Mediterranean carnivore species are
predominantly associated with “montado” and scrubland habitats where abundant
food sources and shelter are provided (Mangas et al 2008; Rosalino et al 2008). One
dramatic example of sensitivity to habitat fragmentation is the extreme range
contraction over the last decades of the critically endangered Iberian lynx (Lynx
pardinus), a highly specialized carnivore that is linked almost exclusively to the wild
rabbit (Oryctolagus cuniculus) as its main prey (Rodriguez and Delibes 2002). Iberian
lynx habitat selection is strongly influenced by the density of wild rabbit which
originated from the Iberian Peninsula and thrive best in Mediterranean scrubland
(Palomares et al 2001). Nevertheless, Virgós (2002) found that generalist species (e.g.,
badgers) may be affected in the similar ways as specialist species. Hence, in
fragmented landscapes which have lost larger carnivore species in the last decades (e.g.
Iberian lynx (pre-scenario extinction) and Iberian wolf Canis lupus) (Petrucci-Fonseca
1990; Sarmento et al 2008), the conservation of the new top carnivores such as smalland medium- sized carnivores is not necessarily assured.
Recent research has shown that patch occupancy depends on the scale at which
organisms interact with landscape pattern operating from local (e.g., shelter, refuge
and food availability, inter-specific competition) to regional scale (e.g., distance
between patches, matrix composition) (Dunning et al 1992; Bowman et al 2001).
Differential species responses are largely related to morphological, physiological, and
behavioural characteristics, including body size, mobility, diet requirements, resource
specialization, and social structure (e.g., Levin 1992; Gehring and Swihart 2003;
Holland et al 2004; Swihart et al 2006). For example, Zabala et al (2005) reported that
western polecat (Mustela putorius) survival depended on small scale habitat changes
because this species intensively exploited areas where resources were locally abundant.
In contrast, distance between forested patches and dispersal habitat were major factors
allowing connectivity for Iberian lynx populations in Doñana (Ferreras 2001). We know
of no study that has addressed the relative impact of forest structural attributes
measured at different scale extents (i.e. the total domain under consideration) on
distribution of a community of common meso-carnivore species. Consideration of
multiple scale responses can have major implications for the findings of fragmentation
studies and enhance the effectiveness of conservation strategies (Bissonette 2003);
51
51
Response of mediterranean carnivores to "montado" fragmentation
otherwise, management actions implemented to ameliorate the negative effects of
fragmentation might be less effective or fail (Orrock et al 2000).
The main goal of our research was to assess the effects of fragmentation of
Mediterranean “montado” landscapes on the most common meso-carnivores in
Portugal by addressing the following questions: 1) Are these meso-carnivores sensitive
to “montado” fragmentation? 2) Where does the threshold effects in species occupancy
are likely to occur? 3) Which environmental variables explain their occupancy at plot-,
patch and landscape scale extents? and 4)What are the independent and/or joint
contributions of the explanatory variables at the three scale extents? We initially
focused on the effect of patch size and isolation on species presence and then examined
which key “montado” attributes were related to differences in carnivore occupancy at
multiple spatial scales: the plot (50m), the patch (mean home range diameter) and the
landscape (median dispersal distance). The five most common carnivore species
present in our study area (Santos-Reis and Petrucci-Fonseca 1999), were: red fox Vulpes
vulpes, stone marten Martes foina, Eurasian badger Meles meles, common genet Genetta
genetta and Egyptian mongoose Herpestes ichneumon. These species have different
evolutionary histories, life-history traits and ecological requirements. We therefore
expected different responses to fragmentation. We predicted that eurytypic species
should be less adversely affected than more stenotypic species (Swihart et al. 2003). We
also predicted that specific behavioral and ecological attributes of species would be
important. For example, forest-dwelling species like stone martens and common genets
may be more affected by variables expressed at patch and landscape extents, whereas
and the remaining species may be more affected by environmental variables expressed
locally. We also expected that highly social species, e.g., Eurasian badgers and
Egyptian mongooses, would be negatively affected by fragmentation because area
effects limit carrying capacity of a patch and may result in a lower threshold of patch
sizes to sustain populations.
52
52
Response of mediterranean carnivores to "montado" fragmentation
Methods
Study area
The study area was located within the Alentejo province (37º54’N, 8º06’ W) of
Portugal (Figure 1) and is characterized by its low elevation (~ 200 m to 500 m) and
extensive plains and a Mediterranean climate.
Figure 1 - Study area in Alentejo Province, Portugal. Sampling sites are shown in “montado”
patches (triangles) as well in the continuous landscape (stars).
The landscape is highly fragmented, with about 44% of the area occupied by a
cropland matrix, followed by “montado” (39%) and, to a much lesser extent, by
production forest (8%) and intensive vineyards and olive groves (7%). Primary human
activities are predominately extensive and intensive agriculture related with olive
groves and vineyards, cork extraction, cattle husbandry and hunting (mainly wild
rabbit and partridge Alectoris rufa).
The human population density (excluding the two major cities, Évora and Beja,
with ~1980 inhabitants.km-2) is 21 inhabitants.km-2 and the road network density
53
Response of mediterranean carnivores to "montado" fragmentation
averages 0.25 km.km-2, about half of the mean national road network density of 0.49
km km-2 (source - IgeoE).
Data collection
Because scent stations are widely used and are considered one of the most
efficient techniques for surveying carnivores (e.g., Conner et al 1983; Zielinski and
Stauffer 1996; Wilson and Delahay 2001), we used them to detect the target carnivores.
We assumed that species identification based on footprint size and shape is reliable
(Piñero 2002); previous studies have successfully identified these species by their tracks
and sign (e.g. Sargeant et al 2003). Carnivore surveys were conducted between May
and September of 2005 in 44 sampling units (SU): 25 were located in different sized
“montado” patches (ranging from 55 to 526 ha), and 19 in continuous “montado” areas
(11668 - 60175 ha). Both were included in a matrix characterized mainly by agricultural
croplands. The sampling protocol is given in figure 2: at each SU we placed 7 to 17
scent stations (depending on the extent of the site) in a grid with 300 m spacing. Intersampling unit distance was at least 4 km to insure independence and avoid pseudoreplication (mean distance = 6838 ± 2249 m).
Figure 2 - Example of the sampling protocol used to analyze the effect of habitat descriptors at
the plot (50m radius), patch (home range diameter) and landscape (dispersal ability) extent on
carnivore occupancy.
A scent-station was composed of one square plate (1 m x 1 m) covered with
marble dust to detect tracks. Both bait (chicken wings) and lure (Cavens GUSTO Lure)
were placed in the centre of the plate to attract individuals. Each scent station was
checked for tracks every other day for six consecutive days (three visits), and tracks
identified and registered. During each visit we replaced scents and bait and smoothed
the surface of the scent-station. In the analysis we used only presence/absence data for
54
54
Response of mediterranean carnivores to "montado" fragmentation
each SU to avoid counting multiple visits by the same individual in the same scentstation or in a group of scent-stations.
Data analysis
To assure the reliability of the data to model the carnivores occupancy patterns,
we estimated the detection probability of each SU using the model likelihood for the
set of detections by scent-stations developed by Mackenzie et al (2002) with the
software
program
PRESENCE
(http://www.mbr-pwrc.usgs.gov/software/
doc/presence/presence.html). We assumed that detection is constant across both space
and time and varies among carnivore species and hence, we estimated the detectability
of continuous “montado” areas for each carnivore species, where they are known to be
present. When detection probability is 0.3 or greater, the occupancy rate is unbiased
(Mackenzie et al 2002) and data can be used to evaluate the site characteristics and
environmental variables that influence the species occupancy patterns. Additionally,
detection probability values were converted into estimates of the sampling effort
(number of scent stations needed to assure of its absence rather than non-detected)
using the procedure suggested by Reed (1996): N=ln (α level)/ln(1-p), α was fixed at
0.05 and p is the detection probability.
We tested differences in the proportion of sites occupied by meso-carnivores in
relation to patch size and distance to large patches. The homogeneity of proportions
was evaluated in relation to three non-overlapping size classes of “montado” areas (a < 60% of a home range; b ≥60% to 100% of a home range, and c - > than ~10 home
ranges for continuous “montado” areas) and three distance classes to nearest large
patch, here defined as ≥4km2 (a - 0m; b - 150-1500m; c - ≥1500m) using the GLIMMIX
procedure in SAS 9.1.3 (SAS Institute, Inc., Cary, North Carolina) software package
Twenty-six descriptors were recorded to investigate species-environment
relationships at three spatial scales: plot-, patch- and landscape-extent (Table 1, Figure
2). In a 50-m radius circular area centered on each scent–station, we recorded variables
describing “montado” vegetation structure (density and maturity), understory height
and density, presence of riparian vegetation and evidence of livestock use. We then
calculate an average of descriptors from the replicate scent-stations to characterize the
55
Response of mediterranean carnivores to "montado" fragmentation
plot level in the sampling unit. For the patch level we measured a series of variables
including: patch size, size of the nearest patch, distance to the nearest patch, distance
to the nearest large patch (>4km2), and distance to the nearest continuous “montado”
area. In order to extract the area of surrounding habitat of the patch we created a
buffer zone based on the estimated home range diameter extent for medium-sized
carnivores. Based on literature we defined a mean home range as 300 ha for stone
marten, common genet, and Egyptian mongoose, and 500 ha for red fox and Eurasian
badger (see Palomares and Delibes, 1993; Santos-Reis et al 2004; Rosalino et al 2005).
Using the mean diameter of the species home range (1,954 and 2,522 m radius,
respectively), we created a circle around the centroid of the scent stations group.
Table 1. Variables considered at each scale for habitat description.
Scale Variables
“Montado” density
PATCH
PLOT
“Montado” maturity
Under storey density
Under storey height
Cows
Sheep
Streams
Patch/site area
Area of the nearest patch
Distance to the nearest patch
Distance do nearest large patch
Distance do nearest continuous
patch
Distance to the stream
Range values
1-3
68-191
0-88
0-112
0-1
0-1
0-1
54-625
0.17-394
150-3400
150-6000
Distance to nearest continuous patch (>40km2) (km)
1.5-27
0-2900
Extensive agriculture*
Intensive agriculture*
Forest production*
Distance to the stream (m)
Linkage to other patches by riparian vegetation (0=no
link; 1=link)
Area of extensive agriculture (ha)
Area of intensive agriculture (ha)
Area of forest production (ha)
Extensive agriculture area
Area of extensive agriculture (ha)
Intensive agriculture area
Area of intensive agriculture (ha)
Forest production area
Area of forest production (ha)
“montado” area
Area of “montado” (ha)
AVDP
MPS
ED
Average distance among “montado” patches (m)
Mean “montado” patch size (ha)
“montado” edge density (km)
MPAR
“montado” Mean Perimeter-Area Ratio
SHDI
Shannon’s Diversity Index
Linkage by riparian vegetation
LANDSCAPE**
Definition
Oak woodland density (1= sparse, 2= dense, 3=very
dense)
Age of trees measured by tree perimeter (surrogate of
maturity) (cm)
Under storey density (%)
Under storey height (cm)
0=absent; 1=present
0=absent; 1=present
Streams (0=absent; 1=present)
Patch/site size (ha)
Area of the nearest patch (ha)
Distance to the nearest patch (m)
Distance to nearest patch >4km2 (m)
0-1
99-794; 188-1281
0-318; 0-591
0-190; 0-299
6105-37777; 1173258601
921-9796;129010019
101-13906; 29925935
3871-31533;968549252
58-1690; 72-1526
31-1660; 39-860
0.2-1.95; 0.36-1.82
113-10141; 1958040
0.60-2.33
*radius for the average diameter of each carnivore home range (HR) (HR 300 ha for stone marten and Egyptian mongoose; HR
500 ha for red fox and badger); ** radius of the average dispersal capacity for medium-sized carnivores (12124 m for stone
marten, genet and Egyptian mongoose; 15652 m for red fox and Eurasian badger).
For the landscape analyses, we calculated the median dispersal distance for
medium-sized carnivores based on the equation 7*√home range size (ha) from
Bowman et al (2002). Therefore, a buffer zone of 12124 m radius was placed around
56
56
Response of mediterranean carnivores to "montado" fragmentation
each sampling unit center for stone marten, genet and Egyptian mongoose, and a
similar procedure for red fox and European badger (buffer zone of 15652 m radius).
We measured nine variables (Table 1) related to landscape composition and
configuration: percent of each most representative type of land use (i.e., extensive and
intensive agriculture, forest production, “montado”), average distance among patches,
mean patch size, edge density, mean perimeter-area ratio, and Shannon’s diversity
index.
The spatial descriptors at the patch and landscape scale extents were assessed using
digital information available in a geographic information system: a land use map (COS
90, 1990) and a riparian vegetation map (SPOT satellite image from 2000). Spatialpattern indices were calculated using FRAGSTATS (McGarigal and Marks 1995) in the
Patch Analyst extension (Rempel and Carr 2003) for ArcView 3.1 (ESRI 1996).
Variance partitioning was used to decompose the variation explained by variables
in the occupancy of target species (see Lawler and Edwards 2006) at each of the three
spatial extents under analysis (Figure 3). Variation in the occupancy data was
partitioned using a series of logistic regression models produced for each target
species. We started by fitting one model for each spatial scale extent (plot, patch and
landscape) per species by applying stepwise selection procedure to retain the variables
(using default variable entry/removal criteria of p = 0.05 and 0.10, respectively) with
higher explanatory power. Prior to statistical analysis, the indices and proportions
were square-root transformed to meet normality and homocedasticity assumptions
(Chaterjee and Price 1991).
The presence/absence of each species in each sampling unit was used as the
dependent (response) variable for the three scale extents. We then used the variables
included in the three models to build a full model for each species and then built three
additional models for each species using: 1) plot and patch variables, 2) plot and
landscape variables, 3) patch and landscape variables. An adaptation of the procedure
suggested by Borcard et al (1992) (Table 2) was used to partition the total deviance
explained ([Deviance null-Deviance residual]/Deviance null) by the models (see
Figure 3, Table 2). All statistical analyses were carried out using R freeware package
(available at www.r-project.org/).
57
Response of mediterranean carnivores to "montado" fragmentation
Figure 3 - Conceptual model (Venn diagrams) underpinning the individual and combined
variance components. Circles correspond to the total species variance accounted by each scale
level. The numbered areas correspond to the individual variance components as described in
table 2 (adapted from Cushman and McGarigal 2002).
Moran’s Index was measured for the residuals of each model to check the
existence of spatial autocorrelation using a Monte Carlo randomization test (999
simulations - CrimeStats III software, Ned Levine 2007). The Moran’s function
calculates a Z score value that indicates the significance of the spatial autocorrelation.
A ‘Z’ score would have to be less than - 1.96 or greater than 1.96 to be statistically
significant at the level of 0.05.
Table 2. Steps to decompose the variance into individual
variance components.
Steps
(1) Plot
Description
Best Model with plot variables
(2) Patch
Best Model with patch variables
(3) Landscape
Best Model with landscape variables
(4) Plot+Patch
Model with both variables from (1) and (2)
(5) Plot+Landscape
Model with both variables from (1) and (3)
(6) Patch+Landscape
Model with both variables from (2) and (3)
(7) Full
(8) Pure Plot
Model with all variables from (1), (2) and
(3)
(7)-(6)
(9) Pure Patch
(7)-(5)
(10) Pure Landscape
(7)-(4)
(11) Shared Plot+Patch
(7)-(3)-(8)-(9)
(12) Shared Plot+Landscape
(7)-(2)-(9)-(10)
(13) Shared Patch+Landscape
(7)-(1)-(9)-(10)
(14) Shared Plot+Patch+Landscape (7)-(8)-(9)-(10)-(11)-(12)-(13)
1-(7)
(15) Unexplained variance
58
58
Response of mediterranean carnivores to "montado" fragmentation
Results
A total of 799 carnivore visits were recorded in 2902 scent stations-nights
established over the 44 sampling units. We found red fox in 35 SU (78% of SU), stone
marten in 31 SU (70%), Egyptian mongoose in 26 SU (59%), Eurasian badger 23 (52%),
and common genet 22 (50%) (Table 3).
Table 3. Summary of the occupancy of red fox, stone marten, Eurasian badger,
common genet and Egyptian mongoose in different sized sampling units.
Sites #
1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
Size (ha)
55
59
69
75
91
97
104
110
117
122
126
131
147
148
149
164
173
223
229
233
267
288
301
425
526
11669
11669
13602
39466
39466
39466
39630
39630
39630
39630
39630
39630
39630
60175
60175
60175
60175
60175
60175
Red fox
1
1
0
1
0
1
0
1
0
1
0
1
1
1
1
1
0
0
1
1
1
1
1
1
1
1
1
0
1
1
1
1
1
1
1
1
1
1
1
0
1
1
1
1
Stone marten
0
1
0
1
1
0
0
0
1
0
0
0
1
1
0
0
1
0
1
0
1
1
1
1
1
1
1
1
1
1
1
1
1
0
1
1
1
1
1
1
1
1
1
1
Eurasian badger
1
0
0
0
0
0
1
1
1
0
1
0
0
1
1
1
1
0
0
0
1
0
1
1
1
0
0
0
1
1
0
1
0
1
0
1
1
1
1
1
0
1
0
0
Common genet
0
1
0
0
0
0
0
0
1
1
1
0
1
0
0
1
0
0
0
1
0
1
0
0
0
1
0
1
1
0
1
1
0
0
1
1
1
1
1
0
1
1
1
1
Egyptian mongoose
0
0
1
1
1
1
0
0
0
0
1
0
1
0
0
0
0
0
0
0
0
1
1
1
0
1
1
0
1
1
1
1
1
0
1
1
1
1
1
1
1
1
1
0
The five species co-occurred only in five continuous “montado” sites. Likelihood
models revealed that detection probability had a negligible effect on red fox, stone
59
Response of mediterranean carnivores to "montado" fragmentation
marten and Egyptian mongoose distribution estimates (p>0.3), so occupancy patterns
for these species are reliable (Table 4).
Low detection probabilities were obtained for the Eurasian badger (p=0.136) and
genet p=0.24); 22 and 11 scent-stations were necessary in each site, respectively to
assure their absence (Table 4). Hence, we preclude them from the following analysis.
Red fox appeared to be unaffected by fragmentation; non-significant differences
in proportion of sites with the species presence were found for patch size and degree of
isolation. Stone marten and Egyptian mongoose were sensitive to vegetation loss and
isolation, but with distinct thresholds. The proportion of sites with species presence
was
significantly
higher
in
continuous
“montado”
areas
than
in
patches
(independently of the size) (Figure 4).
1.0
0.9
“montado” patch size less 60% of homerange (n=16*; n=22**)
“montado” patch size from 60% to 100% of homerange (n=8)
“montado” continuous area (n=19)
b
a
b
0.8
a
0.7
a
a
0.6
0.5
a
a
0.4
0.3
0.2
0.1
b
0.0
red fox
stone marten
Egyptian mongoose
Figure 4 – Proportions of sites with species presence as a function of the amount of “montado”
vegetation in their home range between pairs of “montado” size classes (* average home range
for stone marten and Egyptian mongoose;** average home range for red fox). Same letter are
not statistically different (p>0.05).
60
60
Response of mediterranean carnivores to "montado" fragmentation
A few stone marten were further than 1500m from large patches, but Egyptian
mongoose revealed a lower threshold for patch isolation, with significantly fewer
presences in patches between 150-1500m from larger patches (Figure 5).
0m (n=19)
1.0
0.9
0.8
a
a
0.7
150m and 1500m (n=13)
a
a
a
c
a
0.6
more than 1500m (n=12)
0.5
b
0.4
0.3
b
c
0.2
0.1
b
0.0
red fox
stone marten
Egyptian mongoose
Figure 5 – Estimated the proportions of sites with species presence as a function of the distance
to the closest large patch size (≥4km2). Same letter indicates not statistic differences (p>0.05).
When we ran the multi-scaled analysis, only stone marten and Egyptian
mongoose responded significantly (p>0.05) to the “montado” attributes at different
scale extents (Table 5).
Table 5. Results of the regression models performed for each species on the
three scale extents. Given are the selected variables, coefficients (β), standard
error (SE), and z test values (z value), p-values (p) and % of deviance explained
(D).
Variables
β
SE
z value
p
constant
“montado” maturity
Sheep
constant
“montado” density
Understory density
-6.020
0.799
-2.242
-11.007
6.229
0.591
3.293
0.361
0.894
4.082
2.601
0.192
-1.828
2.214
-2.508
-2.695
2.394
3.091
0.068
0.027
0.012
0.007
0.010
0.002
constant
Distance to nearest large patch
constant
Patch/site area
2.963
-0.069
-0.764
0.143
0.830
0.200
0.448
0.045
3.568
-3.403
-1.707
3.115
0.001
0.001
0.087
0.001
constant
“montado” area
constant
Extensive agriculture area
-5.818
0.057
6.366
-0.039
1.938
0.017
2.227
0.014
-3.001
3.356
2.858
-2.798
0.003
0.001
0.004
0.005
D
PLOT
stone marten
Egyptian mongoose
25
39
PATCH
stone marten
Egyptian mongoose
34
24
LANDSCAPE
stone marten
Egyptian mongoose
35
18
No significant spatial autocorrelation was found on the residuals of the three
models performed for stone marten (Z score (p>0.05) = -0.56(plot), 0.15(patch); 1.64
(landscape)), and for Egyptian mongoose (Z score (p>0.05) = -0.04(plot), -0.03(patch), -
61
Response of mediterranean carnivores to "montado" fragmentation
0.013 (landscape)). At the plot extent, understory and “montado” density had a
positive impact on Egyptian mongoose, whereas sheep absence and “montado”
maturity influenced positively stone marten site use. At patch extent, only distance to
the nearest large patch influenced negatively stone marten’s use, while the “montado”
area influenced Egyptian mongoose’s use and the effect was positive. At the landscape
extent, the area of “montado” had the highest positive influence in stone marten’s use,
and the area in extensive agriculture influenced negatively Egyptian mongoose.
The percent of deviance explained for the stone marten and mongoose full
models was 72% and 42%, respectively. Within the deviance explained, the largest
proportion of the deviance in stone marten occupancy was accounted by the single
effects of plot (44%) and the joint effect of patch and landscape variables (34%) (Figure
6).
100%
90%
80%
70%
60%
50%
40%
30%
20%
10%
0%
-10%
-20%
Shared Plot+Patch+Landscape
Shared Patch+Landscape
Shared Plot+Landscape
Shared Plot+Patch
Pure Landscape
Pure Patch
Pure Plot
stone marten
Egyptian mongoose
Figure 6 - Results of variation partitioning for the stone marten and Egyptian mongoose
regression models at the three scales.
The joint effect of the three spatial scales and the single effects of plot variables
explained the major proportion of the deviance in Egyptian mongoose occupancy
(49%) followed by the single effect of the plot variables (43%) (Figure 6). The negative
shared contribution of plot/patch (-15%) and plot/landscape (-0.5%) variables for
stone marten occupancy indicated that the relationship between the variables were
suppressive and not additive (Chevan and Sutherland 1991).
62
62
Response of mediterranean carnivores to "montado" fragmentation
Discussion
Our findings revealed that common and wide-ranging Mediterranean
carnivores show sensitivity to fragmentation and the response is species-specific and
expressed at different scale extents. The highest variation explained by the stone
marten models suggested that this species is the most sensitive to “montado” attributes
at three scales while red fox appear to have no sensitivity to “montado” fragmentation.
This result is unsurprising for red foxes, referred as generalist and opportunistic
predators (e.g. Dell'Arte et al 2007). Likewise, habitat composition seems also to have
no effect on their occupancy patterns (e.g. Lucherini and Lovari 1996), even to
vegetation cover losses of ~40% of their home range.
For the remaining species (stone marten and Egyptian mongoose), we found
significant differences in their occupancy when the “montado” become patchy but the
low thresholds of patch isolation for Egyptian mongoose demonstrated that this
species is clearly sensitive to cover, and movement may be limited at shorter distances.
Palomares and Delibes (1993) documented that Egyptian mongoose show avoidance of
open areas and preferred the use of dense vegetation which also provides refuges for
the wild rabbit (Lombardi et al 2007), their main prey when available (Delibes et al
1984). Also, since mongooses are diurnal, social, and are subject to control by humans
(Borralho et al 1996), cover seems to be critical for its survival.
As a typical nocturnal, solitary and forest-dwelling species (Santos-Reis et al
2004), stone marten show high sensitivity to isolation, responding positively to 33% of
patches that were separated by 1500 m or more from other large patches, in contrast to
their occupancy in 88% of patches separated by less of 1500 m. Virgós and Garcia
(2002), however, have documented distances up to 5000m; because these authors
analyzed distances ranging from 5 km to 20 km, the differences in response appears to
be related to the different scale extent analyses used. Regarding to patch size, we only
found significant occupancy differences between “montado” patches (even with 100%
cover within average home-range size) and continuous areas, a finding which did not
agree with either Hargis et al (1999) or Virgós and Garcia (2002), who reported that the
critical proportion of mature forest for martens ranged between 50 to 70%, and even
63
Response of mediterranean carnivores to "montado" fragmentation
less with those mentioned by Andrén (1994), Fahrig (1997) and Flather and Beevers
(2002), who detected, in general, significant effects of fragmentation in approximately
20-30% levels of vegetation cover. Indeed, complex interactions of site-specific factors
may explain these different findings. It is important to keep in mind that when
interpreting these results, the impact of vegetation loss may be confounded by other
contributing factors such as prey density, human disturbance, predator removal
activities and climate conditions (Carbone and Gittleman 2002). We were unable to
measure these variables in this study but Mortelliti and Boitani (2008) reported that by
increasing resource availability, stone marten occupancy increased even in relatively
small and isolated patches. These results show an interesting interaction between
resource availability and occupancy rates that appears to reduce the effects of patch
size and isolation.
Although patch area and isolation appear to influence carnivore occupancy
patterns (e.g., Hargis et al 1999; Crooks 2002; Virgós et al 2002), we found that other
factors at different scale extents emerged as explanatory variables for carnivores
occupancy. The importance of both habitat quality and quantity on carnivores patterns
is well documented by, e.g., Crooks (2002), Virgós and Garcia (2002), Mortelliti and
Boitani (2008), but the contributions of environmental variables expressed at different
scale extents has been poorly understood and usually reported for other species (e.g.,
Crozier and Niemi 2003). Studies of Mediterranean carnivores have more often relied
on traditional methods where species-environment relationships were examined with
the local descriptors to evaluate habitat selection (e.g. Palomares and Delibes 1993;
Santos-Reis et al 2004) or landscape variables influencing its distribution (e.g., Borralho
et al 1996; Virgós et al 2002). Although informative, these approaches may provide
insufficient information concerning the relative importance of variables if they are
significantly correlated (Mac Nally 2000).
The variation partitioning method used in this study provides new insights into
species-environmental relationships because it measures the independent and joint
explanatory capacities of variables obtained in the different scale extent models. The
method allows an examination of both joint and single variable effects in explaining
species occupancy patterns.
64
64
Response of mediterranean carnivores to "montado" fragmentation
For stone marten, we observed a dominant independent contribution of
descriptors at the fine scale, whereas the joint contribution of the plot-, patch- and
landscape-scales contributed most to explain the Egyptian mongoose occupancy
patterns. The presence of stone marten seems to depend highly on the presence of
older-growth forests where it can find arboreal cavities as resting and denning sites
that provide security from enemies and thermal insulation which is in accordance to
the findings of Rondinini and Boitani (2002) and Santos-Reis et al (2004). The joint
effects of continuous or functionally connected “montado” areas also show a high
correlation with stone marten occupancy. These results support the findings of Virgós
and Garcia (2002) who evaluated the role of patch size and isolation on the stone
marten distribution patterns and found evidence, although poorly predicted by the
models, of the role that isolation plays in stone marten occupancy. Nevertheless, the
importance of distance to large patches and the detection of stone marten in very small
patches, suggests that patch size may have little or no effect on occupancy patterns
when located in the surroundings of continuous “montado” areas. Virgós and Garcia
(2002) argued that continuous areas may act as source of individuals (Pulliam 1988),
that help maintain healthy populations in small patches. However, individuals far
from continuous areas may use and move between small patches to supplement food
resources. This behavior has been reported by Palomares and Delibes (1994) with
radio-tracked genets in sub-optimal habitats in Northern Doñana National Park; genets
used the same patch for several days and then switched to another patches to access
resources.
Habitat quality at plot extent (“montado” and understorey density) is more
highly correlated with Egyptian mongoose occupancy rates than variables expressed at
patch or landscape extent. Nevertheless, the combined effects of habitat quality and
quantity have higher importance than when acting independently. These findings
provide a new perspective of mongoose-environment relationships; even during
population expansion (Borralho et al 1996), their occurrence may be only assured when
the connectivity among cover areas is established. Finally, the amount of habitat in
surrounding patches appeared to have distinct effects that minimized the adverse
effect of fragmentation for the stone marten and the Egyptian mongoose. Our results
demonstrate that fragmentation of high habitat quality might have little effect on
65
Response of mediterranean carnivores to "montado" fragmentation
Egyptian mongoose if the matrix facilitates the exchange of individuals between
patches (Wiegand et al 2005).
Due to their wide ranging habits, carnivores response to landscape
fragmentation effects is challenging to assess. Distinct responses of carnivores can be
attributable to phenomenon that may be operating at each particular site, potentially
confounding the analysis. For example, the detection of a species in a patch does not
necessarily imply that the site can support resident animals but rather that it may be
visited occasionally as one of several patches required for existence (sensu Dunning et
al 1992; Andrén 1994). Alternatively, patches surrounded by a hostile matrix may be
more intensively used resulting in ephemerally high population densities (e.g.
Anderson 1997).
Our results reinforce the view that a multiple species approach at multiple
spatial scales is crucial to understand the factors affecting patterns of species
distribution, particularly when measures of habitat quality and quantity and their
interactions are involved. Management policies will be more effective if they take into
account these considerations. In conclusion, in highly altered landscapes such as those
in the Mediterranean, this study suggests that conservation efforts should focus
primarily on the local habitat quality for the maintenance of stone marten and
Egyptian mongoose with distinct levels of connectivity at patch scale extent.
Acknowledgments
This study was supported by a research project (POCTI/MGS/47435/2002) from the Fundação para
a Ciência e a Tecnologia. C. Grilo was supported by a Ph.D. Grant (SFRH/BD/10600/2002). We also
would like to thank Clara Silva, Sara Cortez and Luis Almeida for the fieldwork assistance and also
Fernando Ascensão, Ricardo Pita, Darryl Mackenzie, Sarah Durham, Joaquim Pedro Ferreira for
statistical support. We also would like to thank Rui Rebelo and Joaquim Pedro Ferreira for the early
critical comments.
References
Andrén H (1994) Effects of habitat fragmentation on birds and mammals in landscapes with different
proportions of suitable habitat: a review. Oikos 71:355-366
Anderson DR (1997) Corridor use, feeding ecology, and habitat relationships of black bears in a
fragmented landscape in Louisiana. Dissertation, University of Tennessee, Knoxville
Bascompte J, Possgham H, Roughgarden J (2002) Patchy populations in stochastic environments:
critical numbers of patches for persistence. Am Nat 159:128-137
66
66
Response of mediterranean carnivores to "montado" fragmentation
Bissonette JA (2003) Linking landscape patterns to biological reality. In: Bissonette JA Storch I, (Eds)
Landscape Theory and Resource Management: Linking Theory to Practice. Island Press,
Covelo, California, pp15-34
Blondel J (2006) The 'Design' of Mediterranean Landscapes: a Millennial story of humans and
ecological systems during the Historic Period. Hum Ecol 34:713-729
Borcard D, Legendre P, Drapeau P (1992) Partialling out the spatial component of ecological
variation. Ecology 73:1045-1055
Borralho R, Rego F, Palomares F, Hora A (1996). The distribution of the Egyptian mongoose
Herpestes ichneumon (L.) in Portugal. Mammal Rev 26:1-8
Bowman J, Forbes GJ, Dilworth TG (2001) The spatial component of variation in small-mammal
abundance measured at three scales. Can J Zool 79:137-144.
Bowman J, Jaeger JAG, Fahrig L (2002) Dispersal distance of mammals is proportional to home range
size. Ecology 83: 2049-2055.
Carbone C, Gittleman JL (2002) A common rule for the scaling of carnivore density. Science 295,
2273-2276.
Chaterjee S, Price B (1991) Regression analysis by example. Wiley, New York.
Chevan A, Sutherland M (1991) Hierarchical partitioning. Am Stat 45:90-96
Conner MC, Labisky RF, Progulske Jr RD (1983) Scent stations as measures of population abundance
for bobcats, raccoons, gray foxes and opossums. Wildlife Soc B 11:146-152
Crozier GE, GJ, Niemi (2003) Using patch and landscape variables to model bird abundance in a
naturally heterogeneous landscape. Can J Zool 81: 441-452
Crooks KR (2002) Relative sensitivities of mammalian carnivores to habitat fragmentation. Conserv
Biol 16:488-502
Cushman SA, McGarigal K (2002) Hierarchical, multi-scale decomposition of species-environment
relationships. Landscape Ecol 17:637-646
Delibes M, Aymerich M, Cuesta L (1984) Feeding habits of the Egyptian mongoose or ichneumon in
Spain. Acta Theriol 29:205-218
Dell'Arte GL, Laaksonena T, Norrdahla K, Korpimäki E (2007) Variation in the diet composition of a
generalist predator, the red fox, in relation to season and density of main prey. Acta Oecol
31, 276-281
Dunning JB, Danielson BJ, Pulliam HR (1992) Ecological processes that affect populations in complex
landscapes. Oikos 65:169-175
ESRI (1996) ArcView GIS - User’s guide version 3.1. Environmental Systems Research Institute, Inc.
USA
Fahrig L (1997) Relative effects of habitat loss and fragmentation on population extinction. J. Wildlife
Manage 61:603-610
Fahrig L (2003) Effects of habitat fragmentation on biodiversity. Ann Rev Ecol Evol S 34, 487-515
Ferreras P (2001) Landscape structure and asymetrical interpatch connectivity in a metapopulation
of the endangered Iberian lynx. Biol Conserv 100:125-136
67
Response of mediterranean carnivores to "montado" fragmentation
Flather CH, Bevers M (2002) Patchy reaction–diffusion and population abundance: the relative
importance of habitat amount and arrangement. Am Nat 159: 40–56
Gehring TM, Swihart RK (2003) Body size, niche breadth, and ecologically scaled responses to habitat
fragmentation: mammalian predators in an agricultural landscape. Biol Conserv 109: 283-295
Hargis CD, Bissonette JA, Turner DL (1999) The influence of forest fragmentation and landscape
pattern on American martens. J Appl Ecol 36: 157-172
Holland JD, Bert DG, Fahrig L (2004) Determining the spatial scale of species' response to habitat.
BioScience 54: 227
Kurki S, Nikula A, Helle P, Lindén H (1999) Abundances of red fox and pine marten in relation to the
composition of boreal forest landscapes. J Appl Ecol 67:874-886
Lawler JJ, Edwards Jr TC (2006) A variance decomposition approach to investigating multiscale
habitat associations. The Condor 108: 47-58
Levin SA (1992) The problem of pattern and scale in ecology. Ecology 73:1943-1967
Lombardi L, Fernández N, Moreno S (2007) Habitat use and spatial behaviour in the European rabbit
in three Mediterranean environments. Basic Appl Ecol 8: 453-463
Lucherini M, Lovari S (1996) Habitat richness affects home range size in the red fox Vulpes vulpes.
Behav Process 36: 103-105
MacKenzie DI, Nichols JD, Lachman GB, Droege S, Royle JA, Langtimm CA (2002) Estimating site
occupancy rates when detection probabilities are less than one. Ecology 83: 2248-2255
Mac Nally R (2000) Regression and model-building in conservation biology, biogeography and
ecology: the distinction between—and reconciliation of— ‘predictive’ and ‘explanatory’
models. Biodiversity Conserv 9:655-671
Mangas JG, Lozano J, Cabezas-Díaz S, Virgós E. (2008) The priority value of scrubland habitats for
carnivore conservation in Mediterranean ecosystems. Biodivers Conserv 17: 43–51
McGarigal K, Marks BJ (1995) FRAGSTATS: spatial pattern analysis program for quantifying
landscape structure. Gen. Tech. Rep. PNW-GTR-351. Portland, OR: US
Mortelliti A, Boitani L (2008) Interaction of food resources and landscape structure in determining
the probability of patch use by carnivores in fragmented landscapes. Landscape Ecol 23: 285298
Ned Levine (2007) CrimeStat: A Spatial Statistics Program for the Analysis of Crime Incident
Locations (v 3.1). Ned Levine & Associates, Houston, TX, and the National Institute of
Justice, Washington, DC
Oehler JD, Litvaitis JA (1996) The role of spatial scale in understanding responses of medium-sized
carnivores to forest fragmentation. Can J Zool 74: 2070-2079
Orrock JL, Pagels JF, Mcshea WJ, Harper EK (2000) Predicting presence and abundance of a small
mammal species: the effect of scale and resolution. Ecol Appl 10:1356-1366
Palomares F. Delibes M (1993) Key Habitats for Egyptian Mongooses in Donana National Park,
South-Western Spain. J Appl Ecol 30: 752-758
68
68
Response of mediterranean carnivores to "montado" fragmentation
Palomares F, Delibes M (1994) Spatio-temporal ecology and behaviour of European genets in
southwestern Spain. J Mammal 75:714-724
Petrucci-Fonseca F (1990) O lobo (Canis lupus signatus Cabrera, 1907) em Portugal. Problemática da
sua conservação. Tese de Doutoramento. Universidade de Lisboa
Piñero JR (2002) Mamíferos Carnívoros Ibéricos. Lynx Edicions
Pinto-Correia T (1993) Threatened landscape in Alentejo, Portugal: the “montado” and other
“agro-silvo-pastoral” systems. Landscape and Urban Plan 24:43–48
Plieninger T (2006) Habitat loss, Fragmentation, and Alteration – Quantifying the Impact of Land-use
Changes on a Spanish Dehesa Landscape by Use of Aerial Photography and GIS. Landscape
Ecol 21:91-105
Pulliam HR (1988) Sources, sinks, and population regulation. Amer Nat 132:652-661
Reed JM (1996) Using statistical probability to increase confidence of inferring species extinction.
Conserv Biol 10:1283-1285
Rempel
RS, Carr AP (2003) Patch Analyst extension
http://flash.lakeheadu.ca/~rrempel/patch/index.html
for
ArcView:
version
3.
Rodríguez A, Delibes M (2002) Internal structure and patterns of contraction in the geographic range
of the Iberian lynx. Ecography 25:314–328
Rondinini C, Boitani L (2002) Habitat use by beech martens in a fragmented landscape. Ecography
25: 257–264
Rosalino LM, Macdonald DW, Santos-Reis M (2005) Resource dispersion and badger population
density in Mediterranean woodlands: is food, water or geology the limiting factor? Oikos
110: 441-452
Rosalino LM, Santos, MJ, Beier P, Santos-Reis M (2008) Eurasian badger habitat selection in
Mediterranean environments: does scale really matter? Mammal Biol 73:189-198.
Santos-Reis M, Petrucci-Fonseca F (1999). Carnívoros. In: Mathias ML (Ed) Mamíferos terrestres de
Portugal Continental Madeira e Açores. ICN/CBA, Lisboa, Portugal
Santos-Reis M, Santos MJ, Lourenço S, Marques T, Pereira I, Pinto B (2004)Chapter 7. Relationships
between stone martens, genets and cork oak woodlands in Portugal. In: Harrison DJ, Fuller
AK, Proulx G (Eds) Marten and fishers in human altered landscapes. An international
perspective, Kluwer Academic Publishers, Massachusetts, pp147-172
Sargeant GA, Johnson DH, William EB (2003) Sampling designs for carnivore scent-station surveys. J
Wildlife Manage 67: 289-299
Sarmento P, Cruz J, Monterroso P, Tarroso P, Ferreira C, Negrões N, Eira C (2008) Status survey of
the critically endangered Iberian lynx Lynx pardinus in Portugal. Eur J Wildl Res DOI
10.1007/s10344-008-0240-5
Sunquist ME, Sunquist FC (2001). Changing landscapes: Consequences for carnivores. In Gittleman
JL, Funk SM, MacDonald DW, Wayne RK (Eds) Carnivore Conservation. Conservation
Biology 5. Cambridge University Press
Swihart RK, Todd CA, Goheen JR, Scheiman DM, Munroe KE, Gehring TM (2003) Patch occupancy
of North American mammals: is patchiness in the eye of the beholder? J Biogeogr 30:1259–
1279
69
Response of mediterranean carnivores to "montado" fragmentation
Swihart RK, Lusk JJ, Duchamp JE, Rizkalla CE, Moore JE (2006) The roles of landscape context, niche
breadth, and range boundaries in predicting species responses to habitat alteration. Divers
Distrib 12: 277-287
Virgós E (2002) Are habitat generalists affected by forest fragmentation? A test with Eurasian
badgers (Meles meles) in coarse-grained fragmented landscapes of central Spain. J Zool
258:313–318
Virgós E, Garcia FJ (2002) Patch occupancy by stone martens Martes foina in fragmented landscapes
of central Spain: the role of fragmented size, isolation and habitat structure. Acta Oecol 23,
231-237
Virgós E, Telleria JL, Santos T (2002) A comparison on the response to forest fragmentation by
medium-sized Iberian carnivores in central Spain. Biodivers Conserv 11:1063-1079
Wiegand T, Revilla E, Moloney K (2005) Effects of habitat loss and fragmentation on population
dynamics. Conserv Biol 19:108-121
Wilson GJ, Delahay RJ (2001) A review of methods to estimate the abundance of terrestrial carnivores
using field signs and observation. Wildl Res 28 151-164.
With KA, Crist TO (1995) Critical thresholds in species' responses to landscape structure. Ecology
76:2446-2459
Zabala J, Zuberogoitia I, Martinez-Climent JA (2005) Site and landscape features ruling the habitat
use and occupancy of the polecat (Mustela putorius) in a low density area: a multiscale
approach. Wildl Res 51, 157-162
Zielinski WJ, Stauffer HB (1996) Monitoring Martes populations in California: Survey design and
power analysis. Ecol Appl 6:1254-1267
70
70
CHAPTER 4
ROAD EFFECTS ON MEDITERRANEAN CARNIVORES
Paper II
Spatial and temporal patterns in Mediterranean carnivore road casualties:
consequences for mitigation.
Clara Grilo, John A. Bissonette, Margarida Santos-Reis
This paper is published in Biological Conservation (2009) 142: 301-313
B I O L O G I C A L C O N S E RVAT I O N
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
available at www.sciencedirect.com
journal homepage: www.elsevier.com/locate/biocon
Spatial–temporal patterns in Mediterranean carnivore
road casualties: Consequences for mitigation
Clara Griloa,*, John A. Bissonetteb, Margarida Santos-Reisa
a
Universidade de Lisboa, Centro de Biologia Ambiental/Departamento de Biologia Animal, Faculdade de Ciências,
C2 5° 1749-016 Lisboa, Portugal
b
US Geological Survey Utah Cooperative Fish and Wildlife Research Unit, Department of Wildland Resources,
College of Natural Resources, Utah State University, UT 84322-5290, USA
A R T I C L E I N F O
A B S T R A C T
Article history:
Many carnivores have been seriously impacted by the expansion of transportation systems
Received 16 December 2007
and networks; however we know little about carnivore response to the extent and magni-
Received in revised form
tude of road mortality, or which age classes may be disproportionately impacted. Recent
8 October 2008
research has demonstrated that wildlife–vehicle-collisions (WVC) involving carnivores
Accepted 19 October 2008
are modulated by temporal and spatial factors. Thus, we investigated road mortality on a
Available online 3 December 2008
guild of small and medium-sized carnivores in southern Portugal using road-kill data
obtained from a systematic 36 months monitoring period along highways (260 km) and
Keywords:
national roads (314 km) by addressing the following questions: (a) which species and age
Meso-carnivores
class are most vulnerable to WVC? (b) are there temporal and/or spatial patterns in road-
Road-kill patterns
kill? and (c) which life-history and/or spatial factors influence the likelihood of collisions?
‘‘Montado’’ lands
We recorded a total of 806 carnivore casualties, which represented an average of 47 ind./
Modeling
100 km/year. Red fox and stone marten had the highest mortality rates. Our findings high-
Road ecology
light three key messages: (1) the majority of road-killed individuals were adults of common
Mitigation
species; (2) all carnivores, except genets, were more vulnerable during specific life-history
phenological periods: higher casualties were observed when red fox and stone marten were
provisioning young, Eurasian badger casualties occurred more frequently during dispersal,
and higher Egyptian mongoose mortality occurred during the breeding period; and (3) modeling demonstrated that favorable habitat, curves in the road, and low human disturbance
were major contributors to the deadliest road segments. Red fox carcasses were more likely
to be found on road sections with passages distant from urban areas. Conversely, stone
marten mortalities were found more often on national roads with high of cork oak woodland cover; Egyptian mongoose and genet road-kills were found more often on road segments close to curves. Based on our results, two key mitigation measures should help to
reduce WVC in Portugal. The first involves the improvement of existing crossings with buried and small mesh size fence to guide the individuals towards to the passages, in road segments with high traffic volume (>1200 vehicles/night) and located in preferred carnivore
habitats. The second mitigation involves cutting or removal of dense vegetation in verges
of road segments with curves to aid motorists in seeing animals about to cross.
Ó 2008 Elsevier Ltd. All rights reserved.
* Corresponding author: Tel.: +351 217500000; fax: +351 217500028.
E-mail addresses: [email protected] (C. Grilo), [email protected] (J.A. Bissonette), [email protected] (M. Santos-Reis).
0006-3207/$ - see front matter Ó 2008 Elsevier Ltd. All rights reserved.
doi:10.1016/j.biocon.2008.10.026
75
Road effects on Mediterranean carnivores
302
1.
B I O L O G I C A L C O N S E RVAT I O N
Introduction
As roads are built and upgraded to accommodate greater traffic densities in rural areas, the rate of wildlife–vehicle collisions (WVC) has increased significantly (Forman et al.,
2003). Although WVC involving ungulates and other larger
animals have been well documented in North America
(Hansen, 1983; Conover et al., 1995; Romin and Bissonette,
1996; Hubard et al., 2000; Joyce and Mahoney, 2001; Biggs
et al., 2004) and in Europe (Bruinderink and Hazenbroek,
1996; Seiler, 2005), much less information is available for the
suspected high number of small fauna–vehicle collisions
(e.g., Huijser and Bergers, 2000; Ramp et al., 2006).
WVC involving small fauna can as well cause human death
and injury (Williams and Wells, 2005) and are increasingly relevant in the context of biodiversity conservation. It is well
known that some carnivores have been severely affected by
the expansion of transportation structures (Ferreras et al.,
1992; Van der Zee et al., 1992; Caro et al., 2000; Taylor et al.,
2002; Forman et al., 2003; Iuell et al., 2003), however we know
little about carnivore response to the extent and magnitude of
road mortality, or which age classes may be disproportionally
impacted (Noss et al., 1996; Ferreras et al., 2001; Blanco et al.,
2005). Forman and Alexander (1998) suggested that population
effects of road mortality on wildlife may be restricted primarily to a few rare species, but clearly collision records for all
wildlife species are underestimated; further, data on the impact of roads on population stability is scarce. In general, carnivores are particularly vulnerable to WVC and barrier effects
because of their life-history characteristics; e.g., low population density, often low fecundity, and relatively large home
ranges (Spellerberg, 1998; Trombulak and Frissel, 2000; Sunquist and Sunquist, 2001; Forman et al., 2003). In fact, for some
carnivores, road mortality has been identified as a serious
source of mortality; e.g., the Florida panther (Taylor et al.,
2002), wolves in Canada (Paquet, 1993); and Iberian lynx in
Spain (Ferreras et al., 1992). In Britain, traffic is believed to kill
more than 40% of the estimated adult Eurasian badger population annually (Clarke et al., 1998) and mortality on roads accounts for 10% of the endangered Iberian wolf population
south of the Douro river in Portugal (Petrucci-Fonseca, 1990;
Grilo et al., 2004). Although road mortality is the most visible
impact, less is known about the barrier effect caused by road
avoidance (McLellan and Shackleton, 1988; Lovallo and
Anderson, 1996; Whittington et al., 2004; Alexander et al.,
2005; Blanco et al., 2005; Jaeger et al., 2005). In a few mammal
species (e.g., desert bighorn sheep (Ovis canadensis nelsoni),
bobcat (Lynx rufus) and coyote (Canis latrans)), there is evidence
that highways have blocked movement of individuals between
road sides affecting reproductive ability and accelerating the
loss of genetic diversity (Epps et al., 2005; Riley et al., 2006;
Strasburg, 2006). These impacts impose serious concerns
about stability and sustainability of wildlife populations in
roaded landscapes (Lodé, 2000; Ramp et al., 2005).
Despite the large number of road mortality studies (e.g.,
Mech et al., 1988; Brody and Pelton, 1989; Dodd et al., 1989;
Bruinderink and Hazenbroek, 1996; Hubard et al., 2000; Carr
and Fahrig, 2001; Ramsden, 2003; Kolowski and Nielsen,
2008), only a few have focused on a group of species (Clevenger
76
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
et al., 2003; Malo et al., 2004; Ramp et al., 2006; Eigenbrod et al.,
2008) and no study has ever focused on the small and medium-sized carnivore guild of a country. To our knowledge, only
a few papers have attempted to summarize and interpret the
patterns of some carnivores in relation to roads (Clarke et al.,
1998; Philcox et al., 1999; Cain et al., 2003; Seiler, 2003; Saeki
and Macdonald, 2004; Baker et al., 2007). Recent research has
demonstrated that WVC involving carnivores are modulated
by temporal (Philcox et al., 1999; Guter et al., 2005; Orlowski
and Nowak, 2006) and spatial factors (Clarke et al., 1998; Caro
et al., 2000; Cain et al., 2003; Clevenger et al., 2003; Malo et al.,
2004; Ramp et al., 2005). Temporal variations in road-kill are
related to differences in species behavior and activity, e.g., foraging and resting, mating, and dispersal of juveniles (Saeki
and Macdonald, 2004). Additionally, WVC show significant
spatial clustering and appear to depend on population density,
species biology, habitat and landscape structure, as well as
road and traffic characteristics (Clevenger et al., 2003; Malo
et al., 2004). Vehicle speed (Jaarsma et al., 2006), adjacent vegetative cover (Ramp et al., 2005), roadside topography
(Clevenger et al., 2003), and type of nearby passages (Clevenger
et al., 2003; Malo et al., 2004) are road-related features that
have been shown to have varying effects on road-crossing by
wildlife. Moreover, the influence of traffic volume both on road
mortality and on the barrier effect has long been recognized
(e.g., Clarke et al., 1998; Alexander et al., 2005). Up to a threshold volume of traffic, WVC increase with increasing traffic;
above that threshold, some species are discouraged from
crossing roads (Forman et al., 2003; Seiler, 2003; Alexander
et al., 2005). Thus, road managers need reliable data to identify
when and where particular species are susceptible to high
road-kill rates in order to implement mitigation measures during the road design planning and/or exploration stage. Given
the complexity of factors contributing to fatalities and the
specific life-history characteristics of individual species, a
fruitful approach involves identifying explanatory variables
that influence the probability of collisions.
In this paper we provide an assessment of road impacts on
the guild of small and medium-sized carnivores in Portugal by
addressing the following questions: (a) which species and age
classes are most vulnerable to WVC? (b) are there consistent
temporal and/or spatial patterns in road-kill? and (c) are there
life-history and/or spatial factors that influence the likelihood
of collisions? In this context, we: (i) quantified the number of
fatalities of each species, (ii) calculated the proportion of juveniles vs. adults fatalities, (iii) compared the mean number of
fatalities among each known biological period for distinct
species, and (iv) developed predictive models for each species
to identify explanatory variables that may influence the probability of road mortality. We further discuss potential preventive measures.
2.
Methods
2.1.
Study area
In Portugal, from the 1990s onwards, the road network
increased rapidly, mainly because of high speed road construction (source: Portuguese Roads Institute (IEP), BRISA
Road effects on Mediterranean carnivores
B I O L O G I C A L C O N S E RVAT I O N
Auto-Estradas). Two basic categories of roads can be found in
Portugal: national roads (NR) administered by the IEP and
highways (HW) under private concession (e.g., BRISA, AutoEstradas de Portugal, S.A.). Our study was conducted along
sections of two BRISA HW (A2 and A6 – 260 km) and six segments of NR (N4, N114, N370, N380, IP2, IP8 – 314 km) in the
southern region of the country (Fig. 1). Both HW consist of a
total of four-lanes with a median strip of approximately 7 m
wide, and with livestock exclusion fencing along both sides.
Annual average daily traffic (AADT) ranges between 330 and
2494 vehicles/night with a speed limit of 120 km/h. NR consist
of one-lane in each direction without a median strip and only
a few sections are fenced for livestock exclusion. AADT
ranges between 500 and 2161 vehicles/night and with a speed
limit of 100 km/h. Although both type of roads have a similar
range of traffic volumes, HW have a higher proportion of road
segments with higher traffic (+7%) and in several locations
(100 km) HW and NR run parallel to each other sharing the
same adjacent vegetative cover (Fig. 1).
The study area was originally covered by Mediterranean
woodlands and scrublands, but since the 1930s has been
highly modified. The resulting landscape is highly fragmented
with about half of the area occupied by a cropland matrix.
The remaining land (so-called ‘‘montado’’ lands) is dominated
mainly by cork oak (Quercus suber) woods in the west and by
holm oak (Quercus ilex) woods in the east. Human activity is
303
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
predominately cork extraction, cattle husbandry, agriculture
and hunting. Elevation ranges from approximately 200 to
500 m, including four mountains: Adiça, Monfurado, Portel,
and Grândola. The climate is Mediterranean with an Atlantic
influence in the western part of the study area. Five Natura
2000 Network sites are included, at least partially, within
the limits of the study area: Sado, Alvito, Cabrela, Monfurado
and Caia (Fig. 1).
The small and medium-sized carnivore community in the
study area is diverse and comprised of nine species (SantosReis and Petrucci-Fonseca, 1999): red fox (Vulpes vulpes), weasel (Mustela nivalis), Western polecat (Mustela putorius), stone
marten (Martes foina), Eurasian badger (Meles meles), otter (Lutra lutra), common genet (Genetta genetta), Egyptian mongoose
(Herpestes ichneumon) and European wildcat (Felis silvestris).
Only Western polecats and European wildcats are considered
threatened, being the remaining classed as least concern species in view of its widespread distribution in the country
(Trindade et al., 1998; Cabral et al., 2005; Santos-Reis et al.,
2005). Nevertheless, detailed data on the distribution, abundance and habitat requirements of most species is still scarce.
Weasels are rodent-specialists and this largely influences
their demography with populations showing strong fluctuations that in some cases may lead to almost local extinctions
(Santos-Reis, 1989). Although otters are wide ranging, they are
a water-obligate species with a distribution restricted to aqua-
Fig. 1 – Study area and the location of the highways and national roads surveyed (Alentejo Province, Portugal).
77
Road effects on Mediterranean carnivores
304
B I O L O G I C A L C O N S E RVAT I O N
tic environments, namely freshwater streams and dams
(Trindade et al., 1998; Barbosa et al., 2003). Western polecats,
although not strictly water dependent, show a high association with aquatic environments (Mestre et al., 2007) where
they find high diversity and availability of prey (e.g., amphibians and crayfish). Stone martens, genets and European wildcats are forest-dwelling species (Santos-Reis et al., 2004), but
the European wildcat suffered a significant decline mainly
due to habitat destruction, prey decline (mostly rabbits in
the south), human persecution and, to a small extent, hybridization with feral cats (Stahl and Artois, 1991; Lozano et al.,
2007; Oliveira et al., 2007). They exist now in small and fragmented populations (Cabral et al., 2005).
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
2.2.
Road-kill surveys
Surveys on NR were conducted by one observer by car at low
speed (30 km h 1), twice a month from January 2004 to
December 2006. The observer drove along the roadside in both
directions searching the road and its verge for carnivore fatalities. For each observed casualty we recorded species, age and
its geographic coordinates using a global positioning system.
Road-kills on HW were obtained from the BRISA database for
the same period. HW observations were recorded by BRISA
staff during routine road checks, usually carried out several
times a day. Information recorded included date of collection,
species, and location (at a scale resolution of 100 m). When
Table 1 – Definition and description of the variables used in the modeling.
Variable name
Definition
Road related variables
T_ROAD
ROAD_VEG
ROAD_TOPO
Mean (min.–máx.)
Type of road: 1- HW 2- NR
Roadside vegetation: 1 – presence; 0 - absence
Roadside topography measured as the distance from the road
pavement until the roadside vegetation (m) limit: ( ) buried road or (+)
raised road.
Distance to the nearest curve (m)
Number passages (above and below grade)
Daily traffic volume at night (number vehicles/day)
Daily traffic volume at daya (number vehicles/day)
D_CURVE
NUMBER PASSAGES
TRAFFIC
Landscape related variables
ELEVATION
URBAN
INT_AGRI
EXT_AGRI
OAKWOOD
PFOREST
SHRUB
D_STREAM
D_ROADS
D_TOWN
1.55 (1–2)
0.31 (0–1)
0.40 ( 100 to 10)
1896 (0–17 796)
0.88 (0–6)
7000 (2712–15 586)
977 (330–2494)
Mean altitude (m)
% Habitat type within 500 m radius:
Urban
Intensive agriculture
Extensive agriculture
Cork oak woodland
Production forest
Shrub cover measured within a 500 m radius (%)
Minimum distance to nearest water (m)
Minimum distance to other roads (m)
Distance to nearest town (m)
88 (0–400)
4.86 (0–100)
10.85 (0–100)
27.3 (0–100)
12.55 (0–100)
14.42 (0–100)
880 (0–10 406)
1056 (0–6664)
4085 (50–11 159)
Road kills/100km/year
a we used traffic at day time for the Egyptian mongoose model due to its diurnal activity (Palomares and Delibes, 1992) unlike all other
carnivores that are nocturnal (Santos-Reis and Petrucci-Fonseca, 1999).
25
20
15
10
5
0
red fox
weasel
Western
polecat
stone marten
Eurasian
badger
otter
genet
Egyptian
mongoose
Fig. 2 – Frequency of carcasses and the respective standard error found on the national roads and highways surveyed (January
2004 to December 2006).
78
Road effects on Mediterranean carnivores
806
8
68
8
67
7
75
7
45
7
59
8
82
6
110
6
89
7
61
8
31
7
Total number of road-kills
Total number of species
57
7
2
9
17
5
5
10
1
5
5
2
4
4
62
8
8
10
8
13
6
8
1
4
4
4
9
10
22
11
1
6
3
2
9
2
1
2
4
1
7
7
5
7
9
343
12
20
132
81
28
80
110
25
1
1
10
11
6
8
6
31
43
25
33
2
2
9
0
1
4
8
9
10
Fox
Weasel
Western polecat
Stone marten
Eurasian badger
Otter
Genet
Egyptian mongoose
11
3
3
14
7
3
11
10
10
1
5
12
6
1
8
18
33
1
1
25
8
55
4
58
September
August
July
January
February
March
April
May
June
Month
Species
Table 2 – Monthly distribution of carnivore road-kills in the study area (January 2004 and December 2006).
October
November
December
All casualties of given
species
B I O L O G I C A L C O N S E RVAT I O N
305
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
animal carcasses were in good condition, they were collected
and frozen to allow age identification by the authors, as well
as the validation of the species identification. We used the
information provided by the teeth wear to identify the
approximate age of an animal when it died and classed specimens as adults – definitive and completely developed teeth,
or juveniles – teeth with very little wear and pointed incisors
(Harris, 1978). Non-identifiable individuals were pooled as
‘‘other carnivores (including domestic animals)’’ in the original data set and not used in the statistical analyses.
2.3.
Temporal analysis
For the most frequently observed species, we tested the
hypothesis that the general behavior of activity and mobility
related to phenological biological periods might be related
to road mortality. We considered five biological periods: birth,
provisioning of young, mother-young hunting, dispersal, and
breeding (Blanco, 1998). Each represents different behavioral
scenarios with putatively different levels of mobility. For
example, just before birth females tend to den and move less,
whereas when young are being provisioned adult female
movements increase (see Blanco, 1998; Piñero, 2002). We used
Chi-square tests to determine if differences among observed
counts of road casualties in different biological periods were
statistically significant (p < 0.05). Exact binomial tests were
used to determine if there were significant differences in
the occurred number of accidents among biological periods
(Quinn and Keough, 2002).
2.4.
Spatial analysis
We used logistic regression to identify which environmental
and road related variables best predicted the likelihood of
casualties for the carnivore guild, and for each species in particular, following Hosmer and Lemeshow (1989) procedures.
We classed the HW and NR into 500 m segments (n = 1125)
and plotted road-kill data onto the road network. Each segment was coded either 1 (with road-kills) or 0 (without roadkills) and described in terms of sixteen explanatory variables
(Table 1) obtained from field measurements or from a geographic information system. We also, created a surface of distances among the objects (e.g., curves, landmarks such as
towns, roads and streams) and then overlaid the mid-point
of road segments to get the average distances to the objects.
Pearson rank correlations were run to test for relationships
among explanatory variables (Quinn and Keough, 2002). We
built regression models only for species with more than 70 segments coded as 1. First, we examined the ability of each variable to predict mortality using univariate logistic regression.
We selected only variables for the model when the significance
of the Wald test was <0.1. Next, we used stepwise forward
regression procedures to build the models, leaving out 20% of
the original data for cross-validation procedures in order to
evaluate the explanatory power of each model. Model robustness was assessed by comparing the cross-validation classification accuracies between the modeling sample and the
validation sub-sample. Residuals were tested for spatial autocorrelation using Moran I test implemented in the CrimeStats
III (Ned Levine, 2007) using a Monte Carlo randomization test
79
Road effects on Mediterranean carnivores
306
B I O L O G I C A L C O N S E RVAT I O N
(999 simulations). All statistical calculations were carried out
using SPSS 15.0 for Windows software (2006 SPSS, Inc.).
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
(18%), while the proportion of juveniles for the remaining carnivore species varied between 2% and 10% of records.
3.2.
Results
3.1.
Species and age composition of road-kills
A total of 806 carnivores road-kills were recorded over the
three year period, and included eight carnivore species; only
European wildcats were not recorded. On average we found
47 ind./100 km/year killed on the road, with red fox suffering
the highest mortality (n = 343, 20 ind./100 km/year), followed
by stone marten (n = 132, 8 ind./100 km/year), Egyptian mongoose (n = 110, 6 ind./100 km/year) and both Eurasian badger
and genet (n = 81, n = 80, 5 ind./100 km/year) (Fig. 2). Otters
(n = 28), Western polecats (n = 20), and weasels (n = 12) were
less frequently killed. We were able to assess age of 241
(30%) of all observations. The majority of carcasses were
adults; red foxes had the highest proportion of juveniles
50
Chi-square test=22 p<0.05-05
b
b
40
b
30
20
10
c
a
0
birth
feeding
young
hunting with dispersal
mother
casualties/period/year
Eurasian badger
15
10
c
a
c
b
5
0
hunting with
mother
dispersal
casualties/period/year
birth
b
20
a
15
10
c
5
20
birth
c
hunting with dispersal
mother
breeding
Chi-square
Chi-sq
e test=2.76 p>0.05
20
15
10
5
0
birth
feeding
young
hunting with dispersal
mother
breeding
Chi-square test=13,2
t st 13,2 p<0.05
a
c
15
10
feeding
young
genet
breeding
Egyptian mongoose
c
0
Chi-square test=10.44 p<0.05
20
Chi-square test=18.03p<0.05
stone marten
25
breeding
casualties/period/year
casualties/period/year
red fox
Temporal pattern of road-kills
The number of road casualties varied throughout the year
(Table 2). The highest mortality of the carnivore guild as a
whole was recorded in late spring (May: n = 89 and June:
n = 110). Spring and early summer was a critical high mortality period for red fox (42% of casualties between May and July)
and stone marten (36% between May and June). For badger,
genet and Egyptian mongoose, road-kills were concentrated
earlier in the year, corresponding to 21% (February), 24%
(March and April) and 28% (April and May) of annual casualties, respectively (Table 2).
The number of road-kills differed significantly among biological periods (p < 0.05) for all species except for genet (n = 80,
p > 0.05) (Fig. 3). Higher casualties were observed when red fox
and stone marten were feeding young (on average, 44 and 24
road-kills per month over the 3 year period, respectively)
(Fig. 3). Nevertheless, while we found significant differences
casualties/period/year
3.
a
a
b
5
0
birth
hunting with
mother
dispersal
breeding
Fig. 3 – Mean road mortality rate per biological period (road-kills/month) throughout the three years for red fox, stone marten,
Eurasian badger, genet and Egyptian mongoose. Same letters indicate no significant differences (p > 0.05) and different letters
indicate significant differences (p < 0.05) between pairs of life periods in exact binomial test.
80
Road effects on Mediterranean carnivores
1
0.229(**)
D_ROADS
1
0.045
0.018
D_STREAMS
3.3.
1
0.035
0.078(**)
0.067(*)
SHRUBS
1
0.361(**)
0.015
0.075(*)
0.117(**)
PFORES
1
0.283(**)
0.236(**)
0.015
0.032
0.091(**)
OAKWOOD
1
0.155(**)
0.381(**)
0.162(**)
0.130(**)
0.027
0.02
0.146(**)
1
0.125(**)
0.140(**)
.249(**)
0.072(*)
0.117(**)
0.042
0.042
0.163(**)
1
0.003
0.052
0.068(*)
0.002
0.108(**)
0.047
0.106(**)
0.127(**)
0.225(**)
1
0.322(**)
0.033
0.003
0.122(**)
0.016
0.163(**)
0.110(**)
0.220(**)
0.198(**)
0.318(**)
0.316(**)
0.002
0.001
0.106(**)
0.024
.098(**)
0.065(*)
0.246(**)
0.199(**)
0.309(**)
1
(***)
1
0.005
0.017
0.092(**)
0.137(**)
0.034
0.087(**)
0.112(**)
0.021
0.035
0.088(**)
0.021
0.072(*)
1
0.230(**)
0.27 6(**)
0.23 2(**)
0.124(**)
0.134(**)
0.08 4(**)
0.007
0.01
0.134(**)
0.024
0.355(**)
0.114(**)
0.062(*)
1
0.070(*)
0.121(**)
0.058
0.046
0.016
0.01
0.017
0.04
0.043
0.032
0.031
0.045
0.04
0.025
1
0.085(**)
0.052
0.094(**)
0.033
0.052
0.057
0.094(**)
0.005
0.107(**)
0.005
0.056
0.181(**)
0.035
0.016
0.033
(*) Correlation is significant at the 0.01 level.
(**) Correlation is significant at the 0.05 level.
(***) Were not used in the same model: TRAFFIC_D used for diurnal species and TRAFFIC_N used for nocturnal species.
1
0.488(**)
0.309(**)
0.403(**)
0.046
0.011
0.036
EXT_AGRI
INT_AGRI
URB
ELEVATION
TRAFFIC_N
TRAFFIC_D
N_PASSAGES
VISIBILITY
ROAD TOPO
ROAD VEG
T ROADS
1
0.129(**)
0.190(**)
0.446(**)
0.430(**)
0.096(**)
0.026
0.02
0.234(**)
0.077(*)
0.03
0.112(**)
0.048
0.055
0.238(**)
0.099(**)
0.122(**)
T_ROADS
ROAD_VEG
ROAD_TOPO
VISIBILITY
N_PASSAGES
TRAFFIC_D
TRAFFIC_N
ELEVATION
URB
INT_AGRI
EXT_AGRI
OAKWOOD
PFORES
SHRUBS
D_STREAMS
D_ROADS
D_TOWNS
Table 3 – Pearson correlation coefficient between pairs of explanatory variables.
307
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
for stone marten between the period when young were being
fed and other periods, we did not find significant differences
in red fox road mortalities when females were feeding the
cubs, hunting with the young, or during the juvenile dispersal
period. Eurasian badger casualties occurred more frequently
during the dispersal and breeding periods and the birth and
hunting periods were significantly different from the other
periods. Egyptian mongooses were killed on the road more
frequently during the breeding period.
1
D_TOWNS
B I O L O G I C A L C O N S E RVAT I O N
Spatial pattern of road-kills
The Pearson correlation matrix did not show any significant
correlation (r > 0.7) among the explanatory variables (Table
3). Overall, the models moderately fit the data (Table 4). No
significant spatial autocorrelation was found on the residuals
of the models, except for red fox (Fig. 4). To correct the spatial
autocorrelation we only used road segments with red fox
road-kill that were at least 1000 m apart using the remaining
segments for validation (Fig. 4). The correct classification
rates were reasonable (>65%) for the stone marten model
(n = 194, 69%) and Eurasian badger (n = 124, 68%). The red
fox (n = 400, 60%), genet (n = 134, 59%) and Egyptian mongoose
(n = 164, 59%) showed lower correct classification rates. For
genet and Egyptian mongoose the results were due primarily
to the high proportion of false negative prediction errors (Type
II Error), suggesting that these models may only have utility
for explaining the presence of road-kill. The validation sample showed the same general overall classification rate but
the red fox and stone marten models showed a weaker performance indicating a limited capability for model
extrapolation.
Habitat and human disturbance variables as well as visibility and type of road explained the occurrence of the carnivore
guild fatalities (Table 4). The mortality of red fox, in particular,
was related to the number of passages and the longer distance to urban areas. NR and habitat cover (% cork oak woodland) were important variables related to stone marten
mortalities on the road. The likelihood of Eurasian badgers
being hit by vehicles was low in areas with a high human
presence (urban areas and presence of other roads on the
vicinity). Egyptian mongoose and genet road-kill were found
in areas with low visibility, i.e., close to curves.
Traffic did not enter in the models but it appears that the
higher traffic volume may have acted as a barrier to crossing.
Traffic data was classified into three classes according with
Jenk’s optimization algorithm (ESRI, 1996) that minimizes
the variation within each class. We reported mortality in
54% of road segments with low traffic (330–763 vehicles/
night), in 53% with intermediate traffic volumes (764–
1271 vehicles/night) and only in 35% of high traffic (1272–
2494 vehicles/night) road segments.
Although the distribution of each road-kill was not evenly
distributed on roadways (from 9% to 26% of road segments –
Table 5), the carnivore guild casualties was widely distributed
(47% of the road segments), which complicates the implementation of mitigation measures. Additionally, our models
indicated that a high proportion of road segments (57–87%)
show >50% likelihood for a carnivore vehicle collision (Table
5 and Fig. 5).
81
Road effects on Mediterranean carnivores
308
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
B I O L O G I C A L C O N S E RVAT I O N
Table 4 – Logistic regression of five species models with the selected variables, coefficients (b) standard error (SE(b)), Wald
test and p-value, model accuracy and validation.
Species
b
Variables
SE
Model accuracy (%)a
Wald p-Value
%ccp
Validation (%)a
%cca
%tcc
%ccp
%cca
%tcc
Red fox
URBAN
N_PASSAGES
Constant
0.022
0.373
0.278
0.0110
0.0980
0.14d2
4.154
14.370
3.818
0.042
0.000
0.051
60
60
60
94
18
56
Stone marten
NR
OAKWOOD
Constant
1.256
0.012
1.417
0.004
0.333
0.353
8.447
14.275
16.108
0.000
0.004
0.000
64
73
69
64
67
65
Eurasian badger
URBAN
D_ROAD
Constant
0.062
0.001
0.367
0.0300
0.0000
0.2870
4.371
8.019
1.636
0.037
0.005
0.201
68
68
68
67
73
70
Genet
D_CURVE
Constant
0.0003
0.373
0.0001
0.2110
7.917
3.127
0.005
0.077
78
40
59
81
41
62
Egyptian mongoose
D_CURVE
Constant
0.0002
0.302
0.0008
0.1940
6.073
2.433
0.014
0.119
76
42
59
80
45
63
a %ccp – Correct classification of presences, %cca – correct classification of absences, %tcc – total correct classification.
re fox
red
stone
st
marten
0.0
1.0
2000
4000
4
6000
8000
8
1000
1
12000
1
Moran I
Moran I
0
-0.5
0.5
0.0
-1.0
0
distance lag (m)
2000
4000
4
6000
8000
8
10000
1
12000
1
distance lag (m)
Eur
Eurasian
badger
genet
1.0
0.5
0.0
0
2000
4000
6000
8000
10000 12000
Moran I
Moran I
1.0
0.5
0.0
0
-0.5
2000
4000
6000
8000
10000 12000
-1.0
distance lag (m)
distance lag (m)
Egypt
Egyptian mongoose
Moran I
1.0
0.5
0.0
0
2000
4000
6000
8000
10000
12000
distance lag (m)
Fig. 4 – Moran I correlogram of residuals from the logistic models for each species. None of the values was statistically
different from 0 (p > 0.05). The red fox model correlogram of residual after the spatial autocorrelation correction.
82
Road effects on Mediterranean carnivores
B I O L O G I C A L C O N S E RVAT I O N
309
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
Table 5 – Proportion of road segments with road-kills and number of road-kills from observed data and predicted by
logistic regression models (50% and 80% of probability of being killed).
Species
Observations
Logistic regression models
50%c
% Road segmentsa
Red fox
stone marten
Eurasian badger
Genet
Egyptian mongoose
a
b
c
d
e
4.
26
11
7
7
9
80%c
Road-killb
% Road segmentsd
Road-kille
% Road segmentsd
Road-kille
1–6
1–6
1–2
1–3
1–3
47
87
57
63
67
60
95
76
50
79
0.3
30
21
–
–
0.3
61
30
–
–
% Of 500 m road segments with road-kill (n = 1125).
Range of road-kill per 500 m road segment.
Likelihood of being killed from the logistic model.
% Of road segments with likelihood of being killed by logistic regression models.
% Of road-kill included in road segments with likelihood of being killed by logistic regression models.
Discussion
Identifying key factors related to carnivore casualties through
spatial and temporal patterns assessment provides a useful
tool not only for road managers charged with road safety
but also is valuable for environmental managers who try to
minimize this additional source of mortality in the context
of a comprehensive wildlife conservation strategy. Because
Fig. 5 – Road casualties probabilities for each carnivore species given by logistic regression models on the monitored roads.
83
Road effects on Mediterranean carnivores
310
B I O L O G I C A L C O N S E RVAT I O N
the guild of small and medium-sized carnivores surveyed in
this study is composed mainly by wide-ranging and habitat
generalist species (Sunquist and Sunquist, 2001), this complicates attempts to find a simple solution to reduce road mortality. Nevertheless, our findings highlight three simple but
important key mitigation messages: (1) adults of the most
common species tend to be the most commonly hit on the
road; (2) with the exception of genets, all carnivores were
more vulnerable during specific life-history phenological periods, and (3) habitat, human made structures or the proximity
to human habitation and road related features have variable
impacts on the likelihood of carnivores vehicle collisions.
Overall, the high rate of carnivore casualties we reported
here may be explained by their abundance patterns, as seen
in other studies (Baker et al., 2004; Guter et al., 2005). Indeed,
the five most frequently killed carnivores are the most known
common species statewide (Santos-Reis and Petrucci-Fonseca, 1999). Although weasels and Western polecats are also
widely distributed in Portugal (Santos-Reis and Mathias,
1996), the lower number of casualties may suggest that potential mortalities were not detected by this sampling method
(freshly-killed corpses are often taken by birds of prey or other
larger predators) (Love et al., 2000) or perhaps more reasonably, that the species abundance is local and low in the region.
Alternatively, weasels and Western polecats may be less susceptible to road mortality because of their natural history
traits; a previous study (Grilo et al., 2008) showed that both
species avoided highway underpasses and culverts, strongly
suggesting road avoidance.
Road-kills showed seasonal patterns that were consistent
with individual carnivore life-history phenologies. The higher
incidence of road casualties may be related to the higher
mobility periods, e.g., dispersal and breeding. As expected,
in these biological periods, Eurasian badgers and Egyptian
mongooses were more susceptible to road traffic. Because
individuals need to move from their territory through landscape to find new places to establish or mates for reproduction, they might have been unfamiliar with roads.
Interestingly, feeding young was the most vulnerable period
for red fox and stone marten parents. According to Blanco
(1998) they visit the breeding sites several times to feed their
cubs and as the cubs grew, the number of visits increase to as
many as 10 per night. Female mortality is particularly critical
during this period because cubs are totally dependent. In contrast, genets did not appear to be more vulnerable during any
particular biological period; they do not have a restricted
breeding period as other species and therefore, differences
in their movement patterns are difficult to detect throughout
the year (Blanco, 1998).
Clear spatial patterns emerged in the road-kill data; in general, we observed a close relationship between preferred carnivore habitats and low human disturbance with carnivore
fatalities, and the vulnerability of some species to particular
features of roads. Habitat type was an important contributor
only for stone marten model, realistically reflecting the habitat
preferences of this species (Virgós and Garcia, 2002; SantosReis et al., 2004). Even though considered a generalist species
using a wide range of habitats and food items, stone marten
are predominantly associated with cork oak woodlands,
shrublands, and riparian galleries (Santos-Reis et al., 2004).
84
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
We also observed that a high degree of disturbance (urban
areas and the presence of other roads) was an important factor related to the absence of red fox and Eurasian badger
fatalities. These results suggest that the cumulative effect of
human activities may influence dispersal and daily movements and may therefore have impact on the wildlife abundance. Contrary to our expectations, traffic volume did not
significantly influence carnivore mortality on the road. This
issue has been well documented in previous studies of mammals (Clarke et al., 1998; Alexander et al., 2005; Orlowski and
Nowak, 2006). To our knowledge, no studies have examined
how traffic volume affects carnivores when they are active.
In this study, we recorded a higher mortality in low and intermediate traffic volume road segments, suggesting the avoidance of high traffic volume (1200 vehicles/night) roads by
carnivores.
Interestingly, our stone marten mortality data show significant associations with national roads. Most national roads
are narrow and at the same level as the surrounding area,
hence appear integrated into the landscape. In contrast, highways are wider and mostly elevated or below the level of the
surrounding landscape. It is unknown if this configuration
has a barrier effect on stone marten. This observation, in
combination with traffic volume, width of highways, and regularly managed highway verges may discourage stone marten
from crossing highways. Surprisingly, we found that the passages were related with red fox road mortality which are commonly suggested as a mitigation measure to prevent road
casualties (Forman et al., 2003; Iuell et al., 2003; Bissonette
and Adair, 2008). Moreover, a previous study revealed that
red foxes regularly used culverts and passages (Grilo et al.,
2008). As shown in Table 3, the number of passages was significantly related with the presence of favorable habitat,
streams, and high elevations, where we expected to have a
higher abundance of carnivores. It may be that number of
passages per se is not directly related to road-kill but rather
that their proximity to good habitat pre-disposes animals
using these habitats to a higher potential for a vehicle
collision.
When fatalities are widespread over the length of roads,
modeling using binary data (presence/absence of fatalities)
shows poor ability to identify small areas with a higher probability of collisions. However, our results provide evidence
that two key mitigation measures should help to reduce
WVC in Portugal. The first involves the improvement of existing crossings with buried and small mesh size fence to guide
the individuals towards to the passages, in road segments
with high traffic volume (>1200 vehicles/night) and located
in preferred carnivore habitats. The second mitigation involves cutting or removal of dense vegetation in verges of
road segments with curves to aid motorists in seeing animals
about to cross.
A number of important questions remain to be answered.
Even with common species, the effects of the roaded landscape on population viability still need clarification. No doubt,
this will vary by location and species. Further work on barrier
effects and mortality, with emphasis on systematic surveys of
population sizes near and away from roads, behaviour, and
gene flow (e.g. Riley et al., 2006; Baker et al., 2007; Beaudry
et al., 2008), especially for less vagile species, would help refine
Road effects on Mediterranean carnivores
B I O L O G I C A L C O N S E RVAT I O N
our knowledge of road effects. Long term, consistently collected data is not commonly collected for most species, but
would aid immensely in distinguishing the spatial and temporal patterns of WVC hotspots. Modeling WVC of less common
species, assessing the relationship between density and the
likelihood of road mortality, as well addressing the relationship between WVC of carnivores, prey abundance, and vegetation composition along verges would certainly provide key
insights to assist the development of mitigation priorities that
can minimize the adverse effects of roads on wildlife.
Acknowledgments
C. Grilo was supported by a Ph.D. Grant (SFRH/BD/10600/2002)
from the Fundação para a Ciência e a Tecnologia. Additional
funding was provided by BRISA Auto-Estradas de Portugal
S.A. This work was only possible due to the help of Clara Silva,
Carla Baltazar and Luı́s Gomes for carnivore road-kills records
and also all FCUL students that join us in the field work. Many
thanks to Susan Durham for statistical support. Thanks are
due to Fernando Ascensão and Francisco Petrucci-Fonseca,
for suggestions on previous drafts, and to J.P. Metzger and
two anonymous reviewers for insightful and helpful comments on the final manuscript.
R E F E R E N C E S
Alexander, S.M., Waters, N.M., Paquet, P.C., 2005. Traffic volume
and highway permeability for a mammalian community in the
Canadian Rocky Mountains. Canadian Geographer 49, 321–331.
Baker, P.J., Harris, S., Robertson, C.P.J., Saunders, G., White, P.C.L.,
2004. Is it possible to monitor mammal population changes
from counts of road traffic casualties? An analysis using
Bristol’s red foxes Vulpes vulpes as an example. Mammal
Review 34, 115–130.
Baker, P.J., Dowding, C.V., Molony, S.E., White, P.C.L., Harris, S.,
2007. Activity patterns of urban red foxes (Vulpes vulpes)
reduce the risk of traffic-induced mortality. Behavioural
Ecology 18, 716–724.
Barbosa, A.M., Real, R., Olivero, J., Vargas, J.M., 2003. Otter (Lutra
lutra) distribution modeling at two resolution scales suited to
conservation planning in the Iberian Peninsula. Biological
Conservation 114, 377–387.
Beaudry, F., deMaynadier, P.G., Hunter Jr., M.L., 2008. Identifying
road mortality threat at multiple spatial scales for semiaquatic turtles. Biological Conservation. doi:10.1016/
j.biocon.2008.07.016.
Biggs, J., Sherwood, S., Michalak, S., Hansen, L., Bare, C., 2004.
Animal related vehicle accidents at the Alamos National
Laboratory, New Mexico. The Southwestern Naturalist 49, 384–
394.
Bissonette, J.A., Adair, W., 2008. Restoring habitat permeability to
roaded landscapes with isometrically-scaled wildlife
crossings. Biological Conservation 141, 482–488.
Blanco, J.C., 1998. In: Mamı́feros de España, vol. I. GeoPlaneta,
Barcelona.
Blanco, J.C., Cortés, Y., Virgós, E., 2005. Wolf response to two kinds
of barriers in an agricultural habitat in Spain. Canadian
Journal of Zoology 83, 312–323.
Brody, A.J., Pelton, M.P., 1989. Effects of roads on black bear
movements in western North Carolina. Wildlife Society
Bulletin 17, 5–10.
311
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
Bruinderink, G.W.T.A., Hazenbroek, E., 1996. Ungulate
traffic collisions in Europe. Conservation Biology 10,
1059–1067.
Cabral, M.J., Almeida, J., Almeida, P.R., Dellinger, T., Ferrand de
Almeida, N., Oliveira, M.E., Palmeirim, J.M., Queiroz, A.I.,
Rogado, L., Santos-Reis, M., 2005. Livro Vermelho dos
Vertebrados de Portugal. Instituto da Conservação da
Natureza, Lisboa.
Cain, A.T., Tuovila, V.R., Hewitt, D.G., Tewes, M.E., 2003. Effects of a
highway and mitigation projects on bobcats in Southern
Texas. Biological Conservation 114, 189–197.
Caro, T.M., Shargei, J.A., Stoner, G.J., 2000. Frequency of mediumsized mammal road-kill in an agricultural landscape in
California. American Midland Naturalist 11, 362–369.
Carr, L.W., Fahrig, L., 2001. Effect of road traffic on two amphibian
species differing vagility. Conservation Biology 15,
1071–1078.
Clarke, G.P., White, P.C.L., Harris, S., 1998. Effects of roads on
badger Meles meles populations in south-west England.
Biological Conservation 86, 117–124.
Clevenger, A.P., Chruszcz, B., Gunson, K.E., 2003. Spatial patterns
and factors influencing small vertebrate fauna vehicle-kill
aggregations. Biological Conservation 109, 15–26.
Conover, M.R., Pitt, W.C., Kessler, K.K., DuBow, T.J., Sanborn, W.A.,
1995. Review of human, injuries, illnesses and economic
losses caused by wildlife in the United States. Wildlife Society
Bulletin 23, 407–414.
Dodd, C.K., Enge, K.M., Stuart, J.N., 1989. Reptiles on highways in
North-Central Alabama, USA. Journal of Herpetology 23, 197–
200.
Eigenbrod, F., Hecnar, S.J., Fahrig, L., 2008. The relative effects of
road traffic and forest cover on anuran populations. Biological
Conservation 141, 35–46.
Epps, C.W., Palsbøll, P., Wehausen, J.D., Roderick, G., Ramey II, R.,
McCullough, D.R., 2005. Highways block gene flow and cause a
rapid decline in genetic diversity of desert bighorn sheep.
Ecology Letters 8, 1029–1038.
ESRI, 1996. ArcView GIS – User’s Guide Version 3.1. Environmental
Systems Research Institute, Inc., USA. 148pp.
Ferreras, P., Aldama, J.J., Beltrán, J.F., Delibes, M., 1992. Rates and
causes of mortality in a fragmented population of Iberian lynx
Felis pardina Temminck 1824. Biological Conservation 61, 197–
202.
Ferreras, P., Gaona, P., Palomares, F., Delibes, M., 2001. Increase
numbers or reduce mortality? Implications from a population
viability analysis of the Iberian lynx. Animal Conservation 4,
265–274.
Forman, R.T.T., Alexander, L.E., 1998. Roads and their major
ecological effects. Annual Review of Ecology and Systematics
29, 207–231.
Forman, R.T.T., Sperling, D., Bissonette, J.A., Clevenger, A.,
Cutshall, C., Dale, V., Fahrig, L., France, R., Goldman, C.,
Heanue, K., Jones, J., Swanson, F., Turrentine, T., Winter, T.,
2003. Road Ecology: Science and Solutions. Island Press,
Washington.
Grilo, C., Roque, S., Rio-Maior, H., Petrucci-Fonseca, F., 2004. The
isolated wolf population south of Douro River: status and
action lines for its recovery. WolfPrint 20, 13–15.
Grilo, C., Bissonette, J.A., Santos-Reis, M., 2008. Response of
carnivores to existing highway culverts and underpasses:
implications for road planning and mitigation. Biodiversity
and Conservation 17, 1685–1699.
Guter, A., Dolev, A., Saltz, D., Kronfeld-Schor, N., 2005. Temporal
and spatial influences on road mortality in otters:
conservation implications. Israel Journal of Zoology 51, 199–
207.
Hansen, C.S., 1983. Costs of deer–vehicle accidents in Michigan.
Wildlife Society Bulletin 11, 161–164.
85
Road effects on Mediterranean carnivores
312
B I O L O G I C A L C O N S E RVAT I O N
Harris, S., 1978. Age determination in the red fox (Vulpes vulpes) –
an evaluation of technique efficiency as applied to a sample of
suburban foxes. Journal of Zoology 184, 91–117.
Hosmer, J.R., Lemeshow, D.W., 1989. Applied Logistic Regression.
John Wiley & Sons, New York.
Hubard, M.W., Danielson, B.J., Schmitz, R.A., 2000. Factors
influencing the location of deer–vehicle accidents in Iowa.
Journal of Wildlife Management 64, 707–713.
Huijser, M.P., Bergers, P.J.M., 2000. The effect of roads and traffic
on hedgehog (Erinaceus europaeus) populations. Biological
Conservation 95, 111–116.
Iuell, B., Bekker, G.J., Cuperus, R., Dufek, J., Fry, G., Hicks, C.,
HlavÁč, V., Keller, V.B., Rossel, C., Sangwine, T., Torslov, N.,
Wandall, B.M. (Eds.), 2003. Wildlife and Traffic: An European
Handbook for Identifying Conflicts and Designing Solutions.
KNNV Publishers, UK.
Jaarsma, C., van Langevelde, F., Botma, H., 2006. Flattened fauna
and mitigation: traffic victims related to road, traffic, vehicle
and species characteristics. Transportation Research, Part D
11, 264–276.
Jaeger, J.A.G., Bowman, J., Brennan, J., Fahrig, L., Bert, D.,
Bouchard, J., Charbonneau, N., Frank, K., Gruber, B.,
Toschanowitz, K.T., 2005. Predicting when animal populations
are at risk from roads: an interactive model of road avoidance
behavior. Ecological Modelling 185, 329–348.
Joyce, T.L., Mahoney, S.P., 2001. Spatial and temporal distributions
of moose–vehicle collisions in Newfoundland. Wildlife Society
Bulletin 29, 281–291.
Kolowski, J.M., Nielsen, C.K., 2008. Using Penrose distance to
identify potential risk of wildlife–vehicle collisions. Biological
Conservation 141, 1119–1128.
Lodé, T., 2000. Effect of a motorway on mortality and isolation of
wildlife populations. Ambio 29, 163–166.
Lovallo, M.J., Anderson, E.M., 1996. Bobcat movements and home
ranges relative to roads in Wisconsin. Wildlife Society Bulletin
24, 71–76.
Love, R.A., Webon, C., Glue, D.E., Harris, S., 2000. Changes in the
food of British barn owls Tyto alba between 1974 and 1997.
Mammal Review 30, 107–129.
Lozano, J., Virgós, E., Cabezas-Dı́az, S., Mangas, J.G., 2007. Increase
of large game species in Mediterranean areas: is the European
wildcat (Felis silvestris) facing a new threat? Biological
Conservation 138, 321–329.
Malo, J.E., Suárez, F., Dı́ez, A., 2004. Can we mitigate animal–
vehicle accidents using predictive models? Journal of Applied
Ecology 41, 701–710.
McLellan, B.N., Shackleton, D.M., 1988. Grizzly bears and resource
extraction industries: effects of roads on behavior, habitat use,
and demography. Journal of Applied Ecology 25,
451–460.
Mech, L.D., Fritts, S.H., Radde, G.L., Paul, W.J., 1988. Wolf
distribution and road density in Minnesota. Wildlife Society
Bulletin 16, 85–87.
Mestre, F.M., Ferreira, J.P., Mira, A., 2007. Modelling the
distribution of the European Polecat Mustela putorius in a
Mediterranean agricultural landscape. Revue d’Ecologie (Terre
Vie) 62, 35–47.
Ned Levine, 2007. CrimeStat: A Spatial Statistics Program for the
Analysis of Crime Incident Locations (v 3.1). Ned Levine &
Associates/ The National Institute of Justice, Houston, TX/
Washington, DC (March).
Noss, R.F., Quigley, H.B., Hornocker, M.G., Merrill, T., Paquet, P.C.,
1996. Conservation biology and carnivore conservation in the
Rocky Mountains. Conservation Biology 10, 949–963.
Oliveira, R., Godinho, R., Randi, E., Ferrand, N., Alves, P.C., 2007.
Molecular analysis of hybridisation between wild and
domestic cats (Felis silvestris) in Portugal: implications for
conservation. Conservation Genetics 9, 1–11.
86
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
Orlowski, G., Nowak, L., 2006. Factors influencing mammal roadkill in the agricultural landscape of south-western Poland.
Polish Journal of Ecology 54, 283–294.
Palomares, F., Delibes, M., 1992. Circadian activity patterns of
free-ranging large gray mongooses, Herpestes ichneumon, in
Southwestern Spain. Journal of Mammalogy 73, 173–177.
Paquet, P.C., 1993. Summary reference document – ecological
studies of recolonizing wolves in the central Canadian Rocky
Mountains. Unpublished report by John/Paul and Associates
for the Canadian Parks Service, Banff, Alberta.
Petrucci-Fonseca, F., 1990. O lobo (Canis lupus signatus Cabrera,
1907) em Portugal. Problemática da sua conservação. Ph.D.
Thesis, Faculdade de Ciências da Universidade de Lisboa.
Philcox, C.K., Grogan, A.L., MacDonald, D.W., 1999. Patterns of
otter Lutra lutra road mortality in Britain. Journal of Applied
Ecology 36, 748–762.
Piñero, J.R., 2002. Mamı́feros carnı́voros ibéricos. Lynx Edicions,
Barcelona.
Quinn, G.P., Keough, M.J., 2002. Experimental Design and Data
Analysis for Biologists. Cambridge University Press.
Ramp, D., Caldwell, J., Edwards, K.A., Warton, D., Croft, D.B., 2005.
Modelling of wildlife fatality hotpsots along the Snowy
Mountain Highway in New South Wales, Australia. Biological
Conservation 126, 474–490.
Ramp, D., Vanessa, K.W., Croft, D.B., 2006. Assessing the impacts
of roads in peri-urban reserves: road-based fatalities and road
usage by wildlife in the Royal National Park, New South Wales,
Australia. Biological Conservation 129, 348–359.
Ramsden, D.J., 2003. Barn owls and major roads: results and
recommendations from a 15-year research project. Barn Owl
Trust, Ashburton, Devon, UK.
Riley, S.P.D., Pollinger, J.P., Sauvajot, R.M., York, E.C., Bromley, C.,
Fuller, T.K., Wayne, R.K., 2006. A southern California freeway is
a physical and social barrier to gene flow in carnivores.
Molecular Ecology 15, 1733–1741.
Romin, L.A., Bissonette, J.A., 1996. Deer–vehicle collisions:
nationwide status of state monitoring activities and
mitigation efforts. Wildlife Society Bulletin 24, 276–283.
Saeki, M., Macdonald, D.W., 2004. The effects of traffic on the
raccoon dog (Nyctereutes procyonoides viverrinus) and other
mammals in Japan. Biological Conservation 118, 559–571.
Santos-Reis, M., 1989. As doninhas ibéricas (Carnivora: Mustela)
um estudo taxonómico e ecológico. Ph.D. Thesis, University of
Lisbon.
Santos-Reis, M., Mathias, M.L., 1996. The historical and recent
distribution and status of mammals in Portugal. Hystrix 8, 75–89.
Santos-Reis, M., Petrucci-Fonseca, F., 1999. Carnivora. In: Mathias,
M.L. et al. (Eds.), Mamı́feros Terrestres de Portugal
Continental, Açores e Madeira. ICN, Lisboa, pp. 135–165.
Santos-Reis, M., Santos, M.J., Lourenço, S., Marques, T., Pereira, I.,
Pinto, B., 2004. Relationships between stone martens, genets
and cork oak woodlands in Portugal. In: Harrison, D.J., Fuller,
A.K., Proulx, G. (Eds.), Marten and Fishers in Human Altered
Landscapes. An International Perspective. Kluwer Academic
Publishers, Massachusetts (Chapter 7).
Santos-Reis, M., Rosalino, L.M., Loureiro, F., Santos, M.J., 2005.
Capitulo XII Los tejones en Portugal: distribuición, estatus y
conservación. In: Virgós, E., Revilla, E., Mangas, J.G., DomingoRoura, X. (Eds.), Ecologia y conservación del tejón en
ecossistemas mediterráneos. SECEM, Málaga.
Seiler, A., 2003. The toll of the automobile: wildlife and roads in
Sweden. Ph.D. Thesis, Swedish University of Agricultural
Sciences.
Seiler, A., 2005. Predicting locations of moose–vehicle collisions.
Sweden. Journal of Applied Ecology 42, 371–382.
Spellerberg, I.F., 1998. Ecological effects of roads and traffic: a
literature review. Global Ecology & Biogeography Letters 7,
317–333.
Road effects on Mediterranean carnivores
B I O L O G I C A L C O N S E RVAT I O N
Stahl, P., Artois, M., 1991. Status and Conservation of the Wild Cat
(Felis silvestris) in Europe and Around the Mediterranean Rim.
Council of Europe, Strasbourg.
Strasburg, J.L., 2006. Roads and genetic connectivity. Nature 440,
875–876.
Sunquist, M.E., Sunquist, F., 2001. Changing landscapes:
consequences for carnivores. In: Gittleman, J.L., Funk, S.,
Macdonald, D., Wayne, R. (Eds.), Carnivore Conservation,
Conservation Biology, vol. 5. The Zoological Society of London,
Cambridge, pp. 399–418.
Taylor, S.K., Buergelt, C.D., Roelke-Parker, M.E., Homer, B.L.,
Rotstein, D.S., 2002. Causes of mortality of free-ranging Florida
panthers. Journal of Wildlife Diseases 38, 107–114.
Trindade, A., Farinha, N., Florêncio, E., 1998. A distribuição da
Lontra Lutra lutra em Portugal – situação em 1995. Estudos de
Biologia e Conservação da Natureza, vol. 28. Instituto da
Conservação da Natureza, Lisboa.
313
1 4 2 ( 2 0 0 9 ) 3 0 1 –3 1 3
Trombulak, S.C., Frissel, C.A., 2000. Review of ecological effects of
roads on terrestrial and aquatic communities. Conservation
Biology 14, 18–30.
Van der Zee, F.F., Wiertz, J., Ter Braak, C.J.F., Apeldoorn, R.C., 1992.
Landscape change as a possible cause of the badger Meles
meles L. Decline in the Netherlands. Biological Conservation
61, 17–22.
Virgós, E., Garcia, F.J., 2002. Patch occupancy by stone martens
Martes foina in fragmented landscapes of central Spain: the
role of fragment size, isolation and habitat structure. Acta
Oecologica 23, 231–237.
Whittington, J., St. Clair, C.C., Mercer, G., 2004. Path tortuosity and
the permeability of roads and trails to wolf movement. Ecology
and Society 9, 4.
Williams, A.F., Wells, J.K., 2005. Characteristics of vehicle–animal
crashes in which vehicle occupants are killed. Traffic Injury
Prevention 6, 56–59.
87
Road effects on Mediterranean carnivores
Paper III
Response of carnivores to existing highway culverts and underpasses:
implications for road planning and mitigation.
Clara Grilo, John A. Bissonette, Margarida Santos-Reis
This paper is published in Biodiversity and Conservation (2008) 17, 1685-1699
89
Biodivers Conserv (2008) 17:1685–1699
DOI 10.1007/s10531-008-9374-8
ORIGINAL PAPER
Response of carnivores to existing highway culverts
and underpasses: implications for road planning
and mitigation
Clara Grilo Æ John A. Bissonette Æ Margarida Santos-Reis
Received: 14 June 2007 / Accepted: 13 March 2008 / Published online: 26 March 2008
Ó Springer Science+Business Media B.V. 2008
Abstract Roads with high traffic volumes are a source of animal mortality, can disrupt
normal animal movements and dispersal, and may represent a potentially serious threat to
wildlife population stability and viability. Retrofitting existing structures built for other
purposes (e.g., drainage culverts or small below-grade access roads) to facilitate wildlife
crossing by animals and to reduce mortality may be expensive if modifications to the
existing structures themselves were involved. However, the environmental context surrounding these structures may influence the willingness of animals to cross, and
management of some of these attributes may enhance the attractiveness of these structures.
Culverts and underpasses are two common structures along roads in Portugal. We quantified the response of small and medium-sized carnivores to the presence of both types of
existing passages by determining: (1) frequency of use; (2) whether use differed by type of
passage, and if so; (3) by examining if associated environmental attributes might explain
the differences observed. We surveyed 57 different passages along 252 km of highway
with a total sampling effort of 2,330 passage trap-days. The mean passage rate for carnivores combined was 0.7 complete passages per crossing structure per day. Crossings by
weasel, polecat, otter, and wildcat were infrequent or absent. Red fox, badger, genet and
Egyptian mongoose used the crossing structures regularly and without obvious preference;
stone marten preferred underpasses. Regression analyses showed the frequency of use by
carnivores varied with structural, landscape, road-related features, and human disturbance
with 17 of 26 (65%) attributes being significant. Larger passages with vegetation close to
the passage entrances, favorable habitat in the surrounding area, and low disturbance by
humans were important key features to regular use by the guild of species studied. Mitigation planning in areas with ecological significance for carnivores will be beneficial.
Structural attributes and human disturbances are more difficult or expensive to change,
C. Grilo (&) M. Santos-Reis
Departamento de Biologia Animal/Centro de Biologia Ambiental, Faculdade de Ciências,
Universidade de Lisboa, C2 5° Piso, 1749-016 Lisboa, Portugal
e-mail: [email protected]
J. A. Bissonette
USGS Utah Cooperative Fish and Wildlife Research Unit, Department of Wildland Resources, College
of Natural Resources, Utah State University, Logan, UT 84322-5290, USA
123
91
Road effects on Mediterranean carnivores
1686
Biodivers Conserv (2008) 17:1685–1699
even though related significantly to crossing use. Management of vegetation at passage
entrances and restricting human use near passages in carnivore suitable areas may substantially improve crossing attractiveness for the guild of carnivore species.
Keywords Carnivores Conservation Habitat connectivity Non-wildlife passages Road design
Introduction
Increasing road density has emerged recently as a major conservation issue (Noss et al.
1996; Trombulak and Frissel 2000; Forman et al. 2003). Wide roads with high traffic
volumes represent a serious threat to wildlife by introducing an additional source of
mortality (Forman and Alexander 1998; Gerlach and Musolf 2000; Trombulak and Frissel
2000; Iuell et al. 2003; Epps et al. 2005). Furthermore, roads can disrupt animal movement
and dispersal, creating effective barriers and resulting in the loss of landscape connectivity.
Highways, in particular, often are avoided by wildlife as a consequence of increasing
urbanization (Ruediger 1998; Hlaváč and Andĕl 2002; Whittington et al. 2005). In some
mammalian carnivores there is evidence of vulnerability to road network expansions
(Ferreras et al. 1992; Haas 2000; Cain et al. 2003; Percy 2003; Kramer-Schadt et al. 2004)
because of their relatively large ranges and low densities (Sunquist and Sunquist 2001).
Some carnivores face threats of extinction in some regions. For example, Van der Zee
et al. (1992) explicitly linked badger population declines in the Netherlands during the late
1980s to high road density; Thiel (1985) found that wolves disappeared when road densities exceeded 0.6 km km-2. According to Chruszcz et al. (2003), in Banff National Park,
Canada, grizzly bears were found closer to roads with low traffic volumes. In Spain, road
density impacted fox and stone marten movements (Blanco 1988; Lopéz-Martin et al.
1997) resulting in home range changes. Similarly, Whittington et al. (2004) reported that
roads increased the tortuosity of wolf movement pathways in areas of high road density.
Given these data, it is clear that barrier effects in particular, have emerged as a significant
threat to the viability of some carnivore populations world-wide. Thus, an opportunity to
minimize barrier effects includes the adaptation of existing transportation infrastructures to
facilitate passage and restore large-scale habitat connectivity.
Passages designed specifically to increase road permeability have been constructed for
some groups of vertebrates (Foster and Humphrey 1995; Keller and Pfister 1997; Clevenger and Waltho 2000) but they are not common in Portugal. However, because existing
structures built for other purposes (e.g., drainage culverts, and below-grade local access
roads) are numerous and may often coincide with preferential crossing locations, they have
received increasing attention as potential valuable alternatives to mitigate barrier effects
(Hunt et al. 1987; Yanes et al. 1995; Rodriguez et al. 1996; Clevenger et al. 2001; Cain
et al. 2003; Mata et al. 2005). The use of existing passages by carnivores is relatively welldescribed (Foster and Humphrey 1995; Clevenger et al. 2001; Gloyne and Clevenger 2001;
Cain et al. 2003; Dodd et al. 2004; Clevenger and Waltho 2005). There is evidence that
carnivores do not cross highways at random, but rather focus their crossing activity in
locations that vary with passage characteristics, road-related attributes, surrounding habitat
characteristics, and human disturbance levels (Barnum 2003; Rodriquez et al. 1997; Clevenger and Waltho 2000; Mata et al. 2005). Therefore, managing the habitat surrounding
in existing crossing structures may well be a cost effective way of increasing their use
(Iuell et al. 2003).
123
92
Road effects on Mediterranean carnivores
Biodivers Conserv (2008) 17:1685–1699
1687
In the Mediterranean region, a world biodiversity hotspot (Myers et al. 2000), some
studies have investigated the role of existing structures and their importance in restoring
landscape permeability for some small and medium-sized carnivores (Yanes et al. 1995;
Rosell et al. 1997; Rodriguez et al. 1997; Mata et al. 2005; Ascensão and Mira 2007).
However, there is little consensus on whether the type of passage significantly influences
crossing rates. Rosell et al. (1997) found no significant differences related to size of
passage; Mata et al. (2005) reported that most carnivores in Spain (except weasels) preferred larger structures; Yanes et al. (1995) and Ascensão and Mira (2007) found some
carnivore species used smaller culverts regularly. To examine the importance of passage
size on restoring carnivore habitat connectivity, we collected quantitative data on the
response of small and medium-sized carnivores to the most common existing structures
along selected highways in Portugal: culverts and underpasses. Culverts are small-sized
circular or rectangular tubes that allow for water flow from surrounding drainages, whereas
underpasses are larger passages usually built for passage by farm vehicles. We first sought
to determine how often carnivores used crossing structures. Second, we were interested in
whether differences in passage use were related to structure size, or to measurable environmental variables associated with each structure. We expected a regular use of existing
passages by carnivores but different crossing rates for each type of passage. Specifically,
we predicted that both fossorial (fox and badger) and ground-dwelling (mongoose) species
would use the culverts more often whereas arboreal (stone marten and genet) species
would use the larger underpasses. Additionally, we expected that cover would explain
some of the variance in passages use.
Study area and methods
Study area
The study was conducted along 252 km of two highways (A2—94 km and A6—158 km)
in Alentejo province, southern Portugal between March and September 2004. Both highways are four-lane roads and are bordered on both sides by unburied livestock exclusion
fences (1.5 m high, 10 cm mesh size). Between 1998 and 2001, reconstructed stretches of
both highways were opened to traffic. Currently A2 and A6 have an average daily traffic
volume of *13,000 and *6,000 vehicles/day, respectively. The densities of the different
crossing structures were similar for both highways and averaged: 1.22 circular culverts/km
(1 9 1 and 1.5 9 1.5 m), 0.34 box culverts/km (2 9 2, 3 9 3, 4 9 4 and 5 9 5 m), and
0.47 underpasses/km (5 m height and 8 m width). Mean distance between crossing
structures was *390 m, and ranged from 5 to 1,566 m.
The study area was characterized by its vast plains, with elevations ranging from *200
to 500 m, and by its Mediterranean climate. The region was originally covered by
woodlands and scrublands, but has been modified by human activity for centuries, and
especially since the 1930s. The resulting landscape is highly fragmented, with about half of
the area occupied by cropland. The remaining area is dominated by cork oak (Quercus
suber) woods in the west and holm oak (Quercus ilex) woods in the east. Primary human
activities include cork extraction, livestock, agriculture, and hunting. Three Natura 2000
sites (Sado, Cabrela, and Monfurado) are located in the province. Human population
density is *43 inhabitants km-2 (INE 2002) and the road network density averages
0.25 km km-2, about half of the mean national road network density of 0.49 km km-2
(IgeoE 1999).
123
93
Road effects on Mediterranean carnivores
1688
Biodivers Conserv (2008) 17:1685–1699
The carnivore community in the study area is diverse and comprised of nine species
(Santos-Reis and Petrucci-Fonseca 1999): fox (Vulpes vulpes), weasel (Mustela nivalis),
polecat (Mustela putorius), stone marten (Martes foina), badger (Meles meles), otter (Lutra
lutra), genet (Genetta genetta), Egyptian mongoose (Herpestes ichneumon) and wildcat
(Felis silvestris). The wildcat is legally classified as threatened (Vulnerable), while the
polecat is classified as Data Deficient (Cabral et al. 2005). The otter, although classed as
Least Concern in Portugal (Cabral et al. 2005) is of conservation concern in Europe (Near
Threatened, IUCN 2006).
Crossing structures and monitoring
A sample of 57 crossing structures was selected for monitoring of carnivore use along the
A2 and A6 highways (Fig. 1), and included two structural designs: (1) 44 circular culverts
(1 and 1.5 m of diameter) and 13 underpasses (5 m high, 8 m wide). On average, the
minimum distance between crossings selected for this study was 2 km (roughly equivalent
to the average diameter of carnivore home range, as documented by Palomares and Delibes
1994; Rosalino et al. 2004; Santos-Reis et al. 2004; Rondinini et al. 2006) in order to
minimize the effects of spatial autocorrelation and help insure independence of observations (Guisan and Zimmermann 2000).
Field work took place during the driest period of the year when passages were least
likely to be flooded. We used marble dust to detect tracks during 20 consecutive days in
spring and 20 days in summer, 2004. A dust plot 1 m wide and 3–10 mm deep was put on
the ground at both ends of each culvert and in the middle of the underpass, checked every
5 days, and carnivore tracks identified (Lawrence and Brown 1973; Blanco 1998; Piñero
2002). On average, we checked each crossing eight times during the monitoring periods.
Adverse weather conditions and livestock and human activities occasionally prevented
clear identification of tracks, hence we only recorded data when a correct assessment of
Fig. 1 Study area and passages location along the A2 and A6 highways
123
94
Road effects on Mediterranean carnivores
Biodivers Conserv (2008) 17:1685–1699
1689
animal sign was possible. Each record included species ID, number of tracks, and whether
the tracks completely or partially crossed through the passage). In selected passages,
infra-red digital CamtrackÒ cameras were used at both entrances to validate track
identifications.
Data analysis
To calculate the crossing rate, we summed the number of times a given species used the
passage (complete passage) and divided by the number of operative days the passage was
sampled for each season. We also quantified partial passage, which occurred when the
tracks were not found in both passage entrances in the same direction and infra-red
cameras detect enter and exit of the same species in a short period of time.
In general, crossing rates tended to be higher in spring than in other seasons, with
significant differences between spring and summer (Wilcoxon test, P \ 0.05), hence for
the analysis, we compared the mean crossing rates between the two seasons.
To evaluate the effect of passage size on number of crossings, we removed the confounding effects of the surrounding landscape by using two approaches: (1) we analyzed 26
crossing structures (13 circular culverts plus 13 underpasses) placed along both highways;
and (2) we compared six paired culverts and underpasses. Mann–Whitney U-tests
(P \ 0.05) were conducted to test for significant differences in use of passages based on
size and Wilcoxon signed rank tests for differences in use for six paired passages.
To assess the influence of environmental variables on passage use, we used all sampled
crossing structures (n = 57); each was characterized according to 32 independent variables
(Table 1). Structural variables included crossing structure dimensions: width, height, and
an openness ratio (Yanes et al. 1995), distance from highway to passage, and year of
construction. Landscape attributes at a micro-scale included percentage of five land use
types measured within a 500 m radius of the crossing structure: urban, intensive agriculture, extensive agriculture, oak woodlands, and production forest (Pinus spp.; Eucalyptus
spp.). In order to evaluate the influence of landscape variables on the passages use, we also
estimated the proportion of the five habitat cover types mentioned above within a circle
with a radius equivalent to the carnivores’ home range around each passage. We used the
home range size of 300 ha for stone marten, genet and mongoose data (see Palomares and
Delibes 1991, 1993; Santos-Reis et al. 2004), and 500 ha for fox and badgers (see Rosalino
et al. 2004). Distances from passage to the nearest forest cover, as well as the proportion
and average height of shrubs within 10 m were also measured. In order to evaluate the role
of streams and associated vegetation, streams were characterized by width, direction, and
distance to the highway. Road-related attributes included: number of additional existing
structures in a 1,000 m buffer around each passage, distance to the next nearest crossing
structure, and average distance from each structure to the nearest two structures. Human
disturbance attributes included 2004 daily average traffic volume (BRISA, unpublished
data), intensity of human activity, minimum distance to other roads and average distance to
urban areas.
Because of the proportional nature of the response variable (species crossing rates), we
assumed that errors followed a binomial distribution (Zuur et al. 2007). Therefore, generalized linear models with a logarithmic function were fitted to the data to investigate the
influence of environmental attributes near passages on species crossing rates (Faraway
2004). Only variables that yielded significant results (P \ 0.10) for species-specific
crossing rates in the bivariate regression analysis were used as predictors for subsequent
123
95
Road effects on Mediterranean carnivores
1690
Biodivers Conserv (2008) 17:1685–1699
Table 1 Names, definitions, mode and range of values of variables used in the analysis
Variables
Definition
Mode
Range values
Structural
Width
Crossing structure width (m)
1
1–9
Height
Crossing structure height (m)
1
1–6.2
Openness
Crossing structure openness
(= width 9 length/height)
0.03
0.03–1.33
Landscape
Urban_500 m
Urban within 500 m radius (%)
0
0–10
Intensive agriculture_500 m
Intensive agriculture within 500 m radius
(%)
0
0–50
Extensive agriculture_500 m
Extensive agriculture within 500 m radius
(%)
0
0–100
Oak woodland_500 m
Oak woodland within 500 m radius (%)
0
0–100
Production forest_500 m
Production forest within 500 m radius (%)
0
0–20
Urban_hr
Urban within home range areaa (%)
0
0–7
Intensive agriculture_hr
Intensive agriculture within home range
areaa (%)
0
0–100
Extensive agriculture_hr
Extensive agriculture within home range
areaa (%)
0
0–95
Oak woodland_hr
Oak woodland within home range areaa
(%)
0
0–98
Production forest_hr
Production forest within home range areaa
(%)
0
0–7
% Shrub cover_500 m
Shrub cover measured within 500 m radius
(%)
0
0–100
Distance to forest cover
Minimum distance from the passage to
forest cover (m)
0
0–2,000
Density of vegetation
at the entrance
0 = absent; 1 = sparse; 2 = dense
0
0–2
Height vegetation
Average height of vegetation within 10 m
of both entrances (m)
0
0–1.25
Stream width
Stream width (m)
0
0–8
Distance to streams
Distance to streams (m)
Stream direction
0 = absent; 1 = parallel,
2 = perpendicular
0
0–2
Type of riparian vegetation
0 = absent, 1 = shrubs, 2 = gallery
0
0–2
Structure of riparian vegetation
0 = absent, 1 = open, 2 = closed
0
0–2
Highway—passage distance
Distance from the passage entrance to
highway edge (m)
7
1–21
Year of construction
Year of construction of the highway (year)
1998
Distance to the nearest passage
Distance to the nearest crossing structure
(m)
100
1–580
Distance to the two nearest
passages
Average distance to the nearest crossing
structure at both sides of the passage (m)
250
75–875
Number passages
Number passages in within 1,000 m radius
from the passage
6
2,000
0–2,000
Road-related
123
96
1995–2001
1–12
Road effects on Mediterranean carnivores
Biodivers Conserv (2008) 17:1685–1699
1691
Table 1 continued
Variables
Definition
Mode
Range values
Traffic volume
Mean 2004 average daily traffic volume
(no vehicles/day)
9,404
3,000–18,000
Human use
0 = absent; 1 = irregular; 2 = regular
Distance to roads
Minimum distance to linear infra
structures (other roads) (m)
349
Distance to cities
Average distance to cities within 1,000 m
radius (m)
250
0–4,031
Road density_hr
Sum of road length within home range
areaa (m)
2,514
2,247–8,591
Human disturbance
a
0
0–2
109–519
300 ha for stone marten, genet and mongoose and 500 ha for fox and badger
multiple regression analysis. We applied a stepwise selection procedure to retain significant variables and their interactions. Suitable transformations on the dependent and
independent variables were conducted to reduce the effect of outliers (Chaterjee and Price
1991). In addition, we constructed a correlation matrix and excluded correlated variables to
minimize multicollinearity. All statistical analyses were conducted using the Brodgar 2.5.1
(Zuur et al. 2007) and SPSS 14 v. (SPSS 2003) software programs.
Results
Which species used the crossings?
Eight carnivore species used the crossing structures over 2,330 trap-days: red fox, weasel,
polecat, stone marten, badger, otter, genet, and Egyptian mongoose. Although present in
the study area (Pinto and Fernandes 2001; Fernandes 2004), no wildcats were recorded. A
total of 1,940 carnivore tracks were recorded; 1,649 (85%) were complete passages. The
average crossing rate was 0.7 complete passages per structure per day (Table 2). Egyptian
mongoose (28%) and badger (27%) crossed most frequently, followed by fox (18%), stone
marten (12%), genet (12%) and otter (2%). Weasels (n = 2) and polecats (n = 1) crossed
least frequently. The average number of species detected at a crossing structure was 3.32
(SD = 1.17). Absence of wildcat records and the small number of weasel, polecat, and
otter records precluded statistical analysis. We focused on complete crossings by the five
species that used the passages the most (red fox, stone marten, badger, genet, and Egyptian
mongoose).
Did passage size make a difference?
Even though there was a tendency to use the larger underpasses (Fig. 2), only stone marten
showed a statistically significant preference (U = 40.5, 0.01 \ P \ 0.05). When use of paired
culverts and underpasses was conducted, we found similar results: selection was random
except for stone marten, which preferred underpasses (Z = -1.782, 0.01 \ P \ 0.05).
123
97
123
98
a
0.014
0.015
0.000
0.000
0.000
0.001
PP
CP, complete passage; PP, partial passage
0.085
0.151
Summer
CP
CP
PP
Weasel
Fox
Spring
Typea
0.001
0.001
CP
Polecat
0.000
0.000
PP
0.053
0.117
CP
0.017
0.028
PP
Stone marten
Table 2 Movements through crossing structures expressed in number of tracks per day
0.162
0.212
CP
Badger
0.015
0.025
PP
0.013
0.013
CP
Otter
0.003
0.001
PP
0.065
0.104
CP
Genet
0.006
0.021
PP
0.207
0.188
CP
0.043
0.060
PP
Mongoose
Road effects on Mediterranean carnivores
1692
Biodivers Conserv (2008) 17:1685–1699
Road effects on Mediterranean carnivores
Biodivers Conserv (2008) 17:1685–1699
1693
0.8
number of passages/day
0.7
underpasses
circular culverts
0.6
0.5
0.4
0.3
0.2
0.1
0.0
fox
stone marten
badger
genet
mongoose
Fig. 2 Mean crossing rates for the culverts and underpasses and standard deviation respectively (n = 26)
Table 3 Deviance explained with their slope (± slope value) and level of significance and numbers (1–9)
indicating variable importance rank for each species
Variables
Fox
Stone marten
Badger
Genet
+0.100** 5
+0.06*
Mongoose
Structural
Width
+0.09** 1 +0.22**
1
5 ns
Height
+0.08** 2 +0.18**
2
+0.080** 7
+0.04*
6 ns
Openness
+0.07** 3 +0.16**
4
0.060**
9
+0.06*
5 ns
ns
6
-0.18**
1
-0.09*
2 ns
+0.11**
4
Landscape
Extensive agriculture_500 m
0.08**
Oak woodland_500 m
+0.08** 2 +0.17**
% Shrub cover_500 m
ns
ns
Intensive agriculture_hr
ns
-0.17*
3
3
+0.07** 4 +0.04*
5
ns
ns
+0.07** 4
ns
ns
ns
Oak woodland_hr
ns
0.12*
5
0.07*
8
ns
Distance to forest cover
ns
-0.08**
7
-0.17**
2
-0.06*
Height of vegetation
+0.07** 3 ns
ns
ns
ns
5 ns
+0.14** 1
Stream direction
ns
ns
+0.04*
10 ns
+0.11** 2
Type of riparian vegetation
ns
ns
+0.09**
6
ns
+0.07** 4
Structure of riparian vegetation ns
ns
ns
ns
+0.07** 4
Road-related
Distance to the nearest passage ns
-0.09**
6
ns
ns
-0.08** 3
Highway—passage distance
+0.08*
7
ns
ns
ns
ns
Human disturbance
Traffic volume
ns
ns
+0.06**
8
ns
Human use
ns
ns
-0.08**
7
-0.08*
Distance to roads
+0.07** 3 +0.06**
+0.12**
3
+0.10** 1 ns
10 ns
8
Distance to cities
ns
ns
+0.04*
Road density_hr
ns
ns
ns
ns
3 ns
ns
-0.09*
ns—not significant; * 0.05 [ P \ 0.10; ** 0.00 [ P \ 0.05
123
99
Road effects on Mediterranean carnivores
1694
Biodivers Conserv (2008) 17:1685–1699
What environmental variables influenced passage use?
Regression analyses showed that the frequency of use by carnivores varied with structural,
landscape, road-related features, and human disturbance variables (Table 3); 19 of 32
(59%) attributes were significant. Structural attributes (width, height, and openness) had
high explanatory power; with the exception of mongoose, carnivores were less likely to use
smaller passages. Landscape attributes also influenced crossing rate. All carnivores tended
to use passages within oak woodland areas. High vegetation height at crossing entrances
was important for fox and mongoose; distance to forest cover was important for all carnivores. Mongoose crossings were correlated with stream direction, as well as the type and
structure of riparian vegetation. Stone marten and mongoose passage use was negatively
correlated with distance to the next nearest existing structure. Intensive agriculture and
forest production, density of vegetation at the entrance, stream width, and distance to
streams, as well as road related descriptors such as the number of passages in a 1,000 m
buffer and year of highway construction, were not significantly related to crossing use.
Five species-specific models were developed (Table 4). The proportion of variance (R2)
explained by the models developed for each carnivore had values that ranged from 0.15 to
0.40, suggesting other unmeasured variables were responsible for crossing behavior. Some
structural attributes, habitat features, road-related attributes, and human disturbance were
Table 4 Variables retained in the stepwise regression model for each carnivore species, coefficients (B),
standard error, significances, P-value (t test) and deviance explained (D)
Species
Foxa
Variables
B
(Constant)
Oak woodland_500 m
Height vegetation
Stone martena
a
Badger
0.037
0.000
0.146
0.01
0.004
0.010
0.09
0.030
0.501
0.000
Width (1)
0.09
0.430
0.040
Width (2)
1.2
0.380
0.002
Oak woodland_500 m
0.01
0.005
0.010
-2.61
0.640
0.001
Oak woodland_500 m
0.01
0.003
0.010
Distance to roads
0.00
0.001
0.003
(Constant)
-0.31
0.120
0.010
0.78
0.260
0.005
0.16
0.300
0.590
-3.11
0.770
0.000
Distance to roads
0.01
0.001
0.010
Oak woodland_500 m
0.00
0.004
0.010
Human use (1)
-0.69
0.330
0.040
Human use (2)
-0.83
0.450
0.070
(Constant)
-1.20
0.230
0.000
% Shrub cover_500 m
0.01
0.003
0.008
Stream direction (1)
0.26
0.320
0.410
Stream direction (2)
0.90
0.290
0.003
(Constant)
The data used for these species were square root transformed
123
100
-2.03
0.04
Type of riparian vegetation (2)
a
D
-3.04
Cover distance
Mongoosea
P-value
(Constant)
Type of riparian vegetation (1)
Geneta
Std Error
0.311
0.395
0.245
0.195
Road effects on Mediterranean carnivores
Biodivers Conserv (2008) 17:1685–1699
1695
important model components for carnivores, although their influence varied by species.
Except for mongoose, the presence of oak woodland forest was the most important variable
in the models. Distance to other roads was positively related to badger and genet crossing
rates, whereas human use was negatively correlated with genet crossings. Passage width
was a predictor of stone marten crossings. Vegetation height and % shrub cover helped to
explain fox and mongoose crossing rates, respectively.
Discussion
The most common habitat-generalists (fox, stone marten, badger, genet and mongoose)
used large and small passages regularly, but specific attributes influenced species differently in determining the effectiveness of a passage. Our findings suggest that structure size
is important especially for the arboreal stone marten which preferred larger passages. In
general, crossing rates were two times higher in passages 1.5 m wide or larger. These
results differ from those reported by Rodriguez et al. (1997) who did not find any preference or avoidance by fox or wildcat; however, their passages ranged from 1.2 to 3.5 m
while our passages ranged from 1 to 8 m, suggesting a threshold response.
In general, larger passages with vegetation close to the passage entrances, favorable
habitat in the surroundings, and low disturbance by humans were important key attributes.
The presence of vegetation [0.5 m high at the passages entrances was associated with
higher fox, stone marten, and mongoose crossing rates. When the proportion of oak
woodland forest cover was [75%, we found higher values of crossing rates for all carnivores. Both features contributed to masking passage structure and provide greater
protection and security for animals. Human disturbances in the vicinity of passages limited
their use (Clevenger and Waltho 2000; Ng et al. 2004; Ascensão and Mira 2007). We
observed a significant increase in fox, stone marten, badger, and genet crossing rates when
distance to other roads was 500 m or greater. Where we found absence of human activity,
the crossing rates doubled for badger and genet. When the distance to urban centers was
greater than 2,000 m, badger crossing rates increased significantly. We documented that
when stream flow passed through a passage, mongoose crossing rates were two times
higher than when the stream paralleled the crossing. Streams with riparian vegetation
provide travel corridors providing shelter and food, and anti-predator cover (Simberloff
and Cox 1987; Hobbs 1992; Palomares and Delibes 1993; Virgós 2001), and influence the
moving pattern of the mongoose, a largely diurnal species that avoids open areas (Palomares and Delibes 1991).
Both culverts and underpasses play an important role in maintaining landscape connectivity for carnivores in general, but the existing structures appear to be selectively
permeable. Species-specific habitat preferences contribute to the observed differences in
permeability and may act cumulatively. Otters, polecats and wildcats are habitat specialists. Otters are water-obligate (Beja 1992), and because we specifically chose the driest
season to avoid water in the passages, their lower frequency of crossing does not reflect
their high abundance in Portugal (Trindade et al. 1998; Cabral et al. 2005). Polecats,
although not strictly water dependent, show a high association with aquatic environments
(Lode 1994; Zabala et al. 2005) for the higher availability and diversity of prey (e.g.,
amphibians and crayfish). For wildcats, as forest-dependent and highly persecuted species
(Fernandes 2004), the need for undisturbed old growth forest patches may not apply to the
road-related environments, at least at the scale of analysis of this study. Weasels, on the
contrary, do not show strict habitat preference (King 1975; Santos-Reis 1989; Blanco
123
101
Road effects on Mediterranean carnivores
1696
Biodivers Conserv (2008) 17:1685–1699
1998; Klemola et al. 1999) but are rodent-specialists and this largely influences their
demography with populations showing strong fluctuations that in some cases may lead to
almost local extinctions (Santos-Reis 1989). The 2 year drought preceding and during the
study was the driest period during the last 75 years and influenced plant productivity and
rodent numbers, and weasel populations were very low. Regardless, weasels avoid open
areas and use linear corridors such as stone walls or vegetation strips (King 1975);
structures not normally present in the passages.
The variability found on passages use frequency was also related to differences in
population densities. The occurrence of a particular species in the vicinity of a passage and
fluctuations in its local abundance and activity patterns may explain the large amount of
crossing rate variance (Yanes et al. 1995; Rodriguez et al. 1996; Clevenger et al. 2001;
Hlaváč and Andĕl 2002). Rodriguez et al. (1997) found that for foxes and wildcats the
variance explained by passage attributes was much lower (0.06–0.16) than that explained
by patterns of the species abundance (0.23–0.53). Undoubtedly, shelter and food resources
exert an important role. We found an increase in crossing rates in oak woodland where
relative abundances were higher.
Overall, our results indicate that culverts and underpasses facilitate the crossing of
highways by carnivores. These structures appear to be important for landscape connectivity. To increase the likelihood of passage use by the carnivores guild we suggest some
guidelines for highway managers: (1) promote the construction of large passages; (2)
prioritize mitigative measures in areas with ecological significance for carnivores, e.g.,
forested areas of natural woods and where streams associated with riparian vegetation run
through or close to the passages; (3) plant vegetation at passage entrances to guide animals
towards the existing structure; (4) restrict human use of passages and (5) maintenance of
natural vegetation structure inside the passage (soil, logs, rocks, woody debris) will help
encourage use by more sensitive species (weasel and polecat) to artificial structures.
An integrated approach using combined mitigation measures should reduce the impact
of roads. Additional work on other methodologies that account for seasonal and yearly
variations as well as effective monitoring to evaluate the effectiveness of these measures
will be beneficial. Further information is needed to clarify the relationship between species
density and crossing rates.
Less common species are important components of the ecosystem and their inclusion
will be important. Consideration of adequate gene flow and connected populations will
enable decisions regarding the effectiveness of passages and the restoration and maintenance of basic ecological processes and functions.
Acknowledgments C. Grilo was supported by a Ph.D. grant (SFRH/BD/10600/2002) from the Fundação
para a Ciência e a Tecnologia. Additional funding was provided by Brisa Auto-Estradas de Portugal S.A.
We also would like to thank Joana Roque de Pinho and Fernando Ascensão for valuable comments on the
earlier drafts of the manuscript and Eric Gese and Henrique Cabral for the final critical comments on the
manuscript. We also thank one anonymous reviewer for their helpful comments on this manuscript.
References
Ascensão F, Mira A (2007) Factors affecting culvert use by vertebrates along two stretches of road in
Southern Portugal. Ecol Res 22:57–66
Barnum SA (2003) Identifying the best locations along highways to provide safe crossing opportunities for
wildlife. Final report, Colorado Department of Transportation Research Branch
Beja P (1992) Effects of freshwater availability on the summer distribution of otters Lutra lutra on the
Southwest Coast of Portugal. Ecography 15:273–278
123
102
Road effects on Mediterranean carnivores
Biodivers Conserv (2008) 17:1685–1699
1697
Blanco JC (1988) Estudio ecológico del zorro (Vulpes vulpes L. 1758) en la Sierra de Gaudarrama. Ph.D.
Thesis, University of Oviedo
Blanco JC (1998) Mamı́feros de España. Insectivoros, Quirópteros, Primates, Carnı́voros de la Peninsula
Iberica, Baleares y Canarias. Geoplaneta, Barcelona
Cabral MJ, Almeida J, Almeida PR, Dellinger T, Ferrand de Almeida N, Oliveira ME, Palmeirim JM,
Queiroz AI, Rogado L, Santos-Reis M (eds) (2005) Livro Vermelho dos Vertebrados de Portugal.
Instituto da Conservação da Natureza, Lisboa
Cain AT, Tuovila VR, Hewitt DG, Tewes ME (2003) Effects of a highway and mitigation projects on
bobcats in Southern Texas. Biol Conserv 114:189–197
Chaterjee S, Price B (1991) Regression analysis by examples. Wiley, New York
Chruszcz B, Clevenger A, Gunson K, Gibeau M (2003) Relationships among grizzly bears, highways and
habitat in the Banff-Bow Valley, Alberta, Canada. Can J Zool 81:1378–1391
Clevenger AP, Waltho N (2000) Factors influencing the effectiveness of wildlife underpasses in Banff
National Park, Alberta, Canada. Conserv Biol 14(1):47–56
Clevenger AP, Waltho N (2005) Performance indices to identify attributes of highway crossing structures
facilitating movement of large mammals. Biol Conserv 121:453–464
Clevenger AP, Chruszcz B, Gunnison K (2001) Drainage culverts as habitat linkages and factors affecting
passage by mammals. J Appl Ecol 38:1340–1349
Dodd K, Barichivich W, Smith L (2004) Effectiveness of a barrier wall and culverts in reducing wildlife
mortality on a heavily travelled highway in Florida. Biol Conserv 118:619–631
Epps CW, Palsbøll P, Wehausen JD, Roderick G, Ramey R II, McCullough DR (2005) Highways block
gene flow and cause a rapid decline in genetic diversity of desert bighorn sheep. Ecol Lett 8:1029–1038
Faraway JJ (2004) Linear models with R. Chapman & Hall/CRC, USA
Fernandes ML (2004) Gato-bravo no nordeste transmontano. Património Natural Transmontano. Roger
Lopes, Mirandela
Ferreras P, Aldama JJ, Beltrán JF, Delibes M (1992) Rates and causes of mortality in a fragmented
population of Iberian lynx Felis pardina Temminck 1824. Biol Conserv 61:197–202
Forman RTT, Alexander LE (1998) Roads and their major ecological effects. Annu Rev Ecol Syst 29:207–
231
Forman RTT, Sperling D, Bissonette J, Clevenger A, Cutshall C, Dale V, Fahrig L, France R, Goldman C,
Heanue K, Jones J, Swanson F, Turrentine T, Winter T (2003) Road ecology: science and solutions.
Island Press, Washington
Foster ML, Humphrey SR (1995) Use of highway underpasses by Florida panthers and other wildlife. Wildl
Soc Bull 23:95–100
Gerlach G, Musolf K (2000) Fragmentation of landscape as a cause for genetic subdivision in bank voles.
Conserv Biol 14(4):1066–1074
Gloyne CC, Clevenger AP (2001) Cougar Puma concolor use of wildlife crossing structures on the TransCanada highway in Banff National Park, Alberta. Wildl Biol 7:117–124
Guisan A, Zimmermann N (2000) Predictive habitat distribution models in ecology. Ecol Modell 135:147–
186
Haas CD (2000) Distribution, relative abundance, and roadway underpass responses of carnivores
throughout the Puente-Chino Hills. Master Thesis, Faculty of California Polytechnic University
Hlaváč V, Andĕl P (2002) On the permeability of the roads for wildlife. A handbook. Agency for Nature
Conservation and Landscape Protection of Czech Republic, Praha
Hobbs RJ (1992) The role of corridors in conservation: solution or bandwagon? Trends Ecol Evol 7:389–392
Hunt A, Dickens HJ, Whelan RJ (1987) Movements of mammal through tunnels under railway lines. Aust
Zool 24:89–93
IgeoE (1999) Carta militar itinerária de Portugal 1/500 000. Instituto Geográfico do Exército, Lisboa
INE (2002) Censos 2001—Resultados definitivos—Portugal. Instituto Nacional de Estatı́stica, Lisboa
IUCN (2006) 2006 IUCN red list of threatened species. http://www.iucnredlist.org
Iuell B, Bekker GJ, Cuperus R, Dufek J, Fry G, Hicks C, HlavÁč V, Keller VB, Rossel C, Sangwine T,
Torslov N, le Maire Wandall B (eds) (2003) Wildlife and traffic: an European handbook for identifying
conflicts and designing solutions. KNNV publishers, UK
Keller V, Pfister HR (1997) Wildlife passages as a means of mitigating effects of habitat fragmentation by
roads and railway lines. In: Canters K (ed) Habitat fragmentation and infrastructure, Maastricht, The
Hague, Ministry of Transport, Public Works and Water Management, Delft, pp 17–21
King CM (1975) The home range of the weasel (Mustela nivalis) in an English woodland. J Anim Ecol
44(2):639–668
Klemola T, Korpimaki E, Norrdahl K, Tanhuanpaa M, Koivula M (1999) Mobility and habitat utilization of
small mustelids in relation to cyclically fluctuating prey abundances. Ann Zool Fenn 36:75–82
123
103
Road effects on Mediterranean carnivores
1698
Biodivers Conserv (2008) 17:1685–1699
Kramer-Schadt S, Revilla E, Wiegand T, Breitenmoser U (2004) Fragmented landscapes, road mortality and
patch connectivity: modelling influences on the dispersal of Eurasian lynx. J Appl Ecol 41:711–723
Lawrence MJ, Brown RW (1973) Mammals of Britain their tracks, trails and signs. Blandford Press, London
Lode T (1994) Environmental factors influencing habitat exploitation by the polecat Mustela putorius in
western France. J Zool Lond 234:75–88
Lopéz-Martin JM, Lampreave G, Ruiz-Olmo J (1997) Selección de habitat de la garduña (Martes foina):
importancia de las islas de vegetación en ecosistemas mediterráneos alterados. In: Abstracts of the III
Jornadas SECEM—I Jornadas Ibéricas sobre la Nutria. Castelló de Empúries, Girona, 5–7 December
1997
Mata C, Hervás I, Herranz J, Suárez F, Malo JE (2005) Complementary use by vertebrates of crossing
structures along a fenced Spanish motorway. Biol Conserv 124:397–405
Myers N, Mittermeier RA, Mittermeier CG, Fonseca GAB, Kent J (2000) Biodiversity hotspots for conservation priorities. Nature 403:853–858
Ng SJ, Dale JW, Sauvajot RM, Riley SPD, Valone T (2004) Use of highway undercrossings by wildlife in
Southern California. Biol Conserv 115:499–507
Noss RF, Quigley HB, Hornocker MG, Merril T, Paquet P (1996) Conservation biology and carnivore
conservation in the Rocky Mountains. Conserv Biol 10:949–963
Palomares F, Delibes M (1991) Assessing three methods to estimate daily activity patterns in radio-trakked
mongooses. J Wildl Manage 55(4):698–700
Palomares F, Delibes M (1993) Key habitats for Egyptian mongooses in Doňana National Park, southwestern Spain. J Appl Ecol 30:752–758
Palomares F, Delibes M (1994) Spatial-temporal ecology and behaviour of European genets in southwestern
Spain. J Mammal 75(3):714–724
Percy MP (2003) Spatio-temporal movement and road crossing patterns of wolves, black bears and grizzly
bears in the bow river valley of Banff National Park. M.Sc. Thesis, University of Alberta
Piñero JR (2002) Mamı́feros carnı́voros ibéricos. Lynx edicions, Barcelona
Pinto B, Fernandes M (2001) Abordagem preliminar à distribuição do gato-bravo em Portugal. DHE/ICN,
Lisboa
Rodriguez A, Crema G, Delibes M (1996) Use of non-wildlife passages across a high speed railway by
terrestrial vertebrates. J Appl Ecol 33:1527–1540
Rodriguez A, Crema G, Delibes M (1997) Factors affecting crossing of red foxes and wildcats through nonwildlife passages across a high-speed railway. Ecography 20:287–294
Rondinini C, Ercoli V, Boitani L (2006) Habitat use and preference by polecats (Mustela putorius L.) in a
Mediterranean agricultural landscape. J Zool 269:213–219
Rosalino LM, Macdonald DW, Santos-Reis M (2004) Spatial structure and land-cover use in a low-density
Mediterranean population of Eurasian badgers. Can J Zool 82(9):1493–1502
Rosell C, Parpal J, Campeny R, Jové S, Pasquina A, Velasco JM (1997) Mitigation of barrier effect of linear
infrastructures on wildlife. In: Canters K (ed) Proceedings of the international conference on habitat
fragmentation, infrastructure and the role of ecological engineering, Maastricht and the Hague
Ruediger B (1998) Rare carnivores and highways—moving into the 21st century. In: Evink GL, Garret P,
Zeigler D, Berry J (eds) Proceedings of the international conference on wildlife ecology and transportation, Tallahasse, pp 10–16
Santos-Reis M (1989) As doninhas ibéricas (Carnivora: Mustela) um estudo taxonómico e ecológico. Ph.D.
Thesis, University of Lisbon
Santos-Reis M, Petrucci-Fonseca F (1999) Carnı́voros. In: ICN/CBA (eds) Mamı́feros terrestres de Portugal
Continental, Madeira e Açores. Instituto de Conservação da Natureza/Centro de Biologia Ambiental,
Lisboa
Santos-Reis M, Santos MJ, Lourenço S, Marques T, Pereira I, Pinto B (2004) Relationships between stone
martens, genets and cork oak woodlands in Portugal. In: Harrison DJ, Fuller AK, Proulx G (eds)
Marten and fishers in human altered landscapes: an international perspective. Kluwer Academic
Publishers, Massachusetts
Simberloff D, Cox J (1987) Consequences and costs of conservation corridors. Conserv Biol 1:63–71
SPSS (2003) SPSS statistical software—version 12.0 for windows. SPSS Inc., Chicago
Sunquist ME, Sunquist FC (2001) Changing landscapes: consequences for carnivores. In: Gittleman JL,
Funk SM, MacDonald DW, Wayne RK (eds) Carnivore conservation. Conservation biology 5.
Cambridge University Press, pp 399–418
Thiel RP (1985) Relationship between road densities and wolf habitat suitability in Wisconsin. Am Midl Nat
113(2):404–407
Trindade A, Farinha N, Florêncio E (1998) A Distribuição da Lontra Lutra lutra em Portugal—situação em
1995. Estudos de Biologia e Conservação da Natureza, 28. ICN, Lisboa
123
104
Road effects on Mediterranean carnivores
Biodivers Conserv (2008) 17:1685–1699
1699
Trombulak SC, Frissel CA (2000) Review of ecological effects of roads on terrestrial and aquatic communities. Conserv Biol 14(1):18–30
Van der Zee FF, Wiertz J, Ter Braak CJF, Apeldoorn RC (1992) Landscape change as a possible cause of
the badger Meles meles L. decline in the Netherlands. Biol Conserv 61:17–22
Virgós E (2001) Relative value of riparian woodlands in landscapes with different forest cover for mediumsized Iberian carnivores. Biodivers Conserv 10:1039–1049
Whittington J, St Clair CC, Mercer G (2004) Path tortuosity and the permeability of roads and trails to wolf
movement. Ecol Soc 9(1):4. http://www.ecologyandsociety.org/vol9/iss1/art4/
Whittington J, St Clair CC, Mercer G (2005) Spatial responses of wolves to roads and trails in mountain
valleys. Ecol Appl 15(2):543–553
Yanes M, Velasco J, Suaréz F (1995) Permeability of roads and railways to vertebrates: the importance of
culverts. Biol Conserv 71:217–222
Zabala J, Zuberogoitia ÆI, Martı́nez-Climent ÆJA (2005) Site and landscape features ruling the habitat use
and occupancy of the polecat (Mustela putorius) in a low density area: a multiscale approach. Eur J
Wildl Res 51:157–162
Zuur A, Ieno E, Graham S (2007) Analysing ecological data. Springer, UK
123
105
CHAPTER 5
MODELLING STRUCTURAL CONNECTIVITY FOR
STONE MARTEN IN SOUTHERN PORTUGAL
Paper IV
Structural connectivity in a Mediterranean landscape and its
consequences for road mortality risk of stone marten.
Clara Grilo, Fernando Ascensão, John A. Bissonette, Margarida Santos-Reis
This paper was submitted to the journal Landscape and Urban Planning on February
23, 2009
Modelling structural connectivity for stone marten in southern Portugal
Structural connectivity in a Mediterranean landscape and its
consequences for road mortality risk of stone marten
Clara Grilo1, Fernando Ascensão1 John A. Bissonette2 Margarida Santos-Reis1
1Universidade
de Lisboa, Centro de Biologia Ambiental/Departamento de Biologia Animal, Faculdade de Ciências, C2
5º 1749-016 Lisboa, Portugal (E-mail address: [email protected], [email protected], [email protected])
2U.
S. Geological Survey Utah Cooperative Fish and Wildlife Research Unit, Department of Wildland Resources, College
of Natural Resources, Utah State University, USA, 84322-5290 (E-mail address: [email protected])
Abstract
Mediterranean basin landscapes have been shaped historically by human-nature
interactions that have lead to the emergence of the unique “montado” ecosystem now
recognized with a high conservation value. Recently, “montado” has experienced significant
large-scale changes, including agriculture intensification and road network expansion. Some
wide-ranging and forest-dwelling carnivores have already shown a particular vulnerability to
these developments. For example, although stone marten Martes foina ecology is reasonably
well known, its response to the loss of spatial connectivity of populations is poorly understood.
In this paper, we measured the structural connectivity of the landscape in southern Portugal
and examined how it may influence road mortality risk. We addressed three questions: 1) how
do different levels of landscape connectivity influence stone marten occupancy of “montado”
areas? 2) is there any relation between “montado” connectivity and road mortality risk? 3) if so,
which road-related features might be responsible for high road mortality? We developed a
series of 12 connectivity models using cost surface scenarios on the basis of differential weights
given to each vegetation cover type to reflect different resistance to species movement. Our
findings show that stone marten occupancy likelihood decreases with distance to source areas
(with an increased cost in open areas and distance to riparian areas) and is positively affected
by patch size. The high road-kill incidence was detected when roads traversed well connected
areas. Road sinuosity was the most important factor influencing the mortality incidence in those
areas. Recommendations are provided to restore “montado” structural connectivity and
decrease road mortality.
Keywords – habitat fragmentation, “montado”, carnivores, road-kill, hierarchical partitioning,
cost surface scenarios.
111
111
Modelling structural connectivity for stone marten in southern Portugal
1. Introduction
Landscapes within the Mediterranean basin have been historically shaped by
human-nature interactions. This has lead into the emergence of the cork and holm oak
woodlands (hereafter “montado”) ecosystem (Blondel, 2006), an agro-silvo-pastoral
system characterised by a mosaic of crops, grasslands, scrublands, and forest patches,
with variable densities of cork (Quercus suber) and/or holm (Q. ilex) oaks, associated
with animal grazing and cultivation (Pinto Correia, 1993). These complex and dynamic
“montado” landscapes evolved together with numerous species being considered to
hold high levels of biodiversity (Plieninger and Wilbrand, 2001). In spite of coevolutionary processes that allowed species to adapt to small-scale habitat change over
time (Blondel, 2006), Mediterranean landscapes in the Iberian Peninsula, and
particularly in southern Portugal, have experienced significant large-scale changes
over the past century, including agriculture intensification and significant road
network expansion. Free-range grazing at high stocking levels also has extensively
cleared scrubland dominated areas. Today, large areas of the region are dominated by
a cropland matrix where woodlands patches with low understory cover are mixed
with non-native plantations of conifer (Pinus sp.) and eucalyptus (Eucalyptus sp.)
species (Plieninger, 2006). In addition, there has been significant road network
expansion since the integration of Portugal into the European Union; two new
highways were built in this region, adding ca. 500km to the existing road network, and
several other roads were improved. As the road network extends across the landscape
and traffic volumes intensify, well connected areas become increasingly fragmented.
Because high traffic volume roads are a source of animal mortality and can disrupt
animal movements (Forman et al., 2003), they represent a serious threat to long-term
wildlife population stability and viability (e.g., Dixon et al., 2006).
Despite the cumulative effects of habitat fragmentation and road mortality on
biodiversity conservation, these are often evaluated separately. Landscape connectivity
assessments generally take into account only the amount of habitat, and generally
exclude human-made features (e.g., Mortelliti and Boitani, 2008). Similarly, the effects
of roads are generally evaluated using mortality risk or road permeability, lacking the
112
112
Modelling structural connectivity for stone marten in southern Portugal
landscape perspective (e.g., Ramp et al. 2005). We argue that wildlife conservation
must account for the effects of both factors in order to capture the responses of
organisms to different levels of landscape connectivity and to provide guidelines to
minimize the negative effects of fragmentation and road mortality.
Recent research has shown that some endangered mammalian carnivores are
particularly vulnerable to habitat fragmentation and road network expansion
(Sunquist and Sunquist, 2001; Crooks, 2002; McRae et al. 2005) and are facing threats of
extinction in some regions (e.g., Ferreras, 2001; Naves et al. 2003). Surprisingly, little is
known about the vulnerability of the more widespread and common species to habitat
fragmentation. One example is the stone marten Martes foina, recently documented to
be particularly sensitive to forest fragmentation and vulnerable to roads (e.g., Mortelitti
and Boitani, 2008; Grilo et al. 2009). Several studies have show them to be highly
dependent on the presence of contiguous forested areas; e.g., cork woodlands (Virgós
and Garcia, 2002; Santos-Reis et al. 2004; Mortelliti and Boitani, 2008), and avoids open
areas and arable lands (Rondinini and Boitani, 2002). Previous research in southern
Portugal also revealed that stone marten is highly vulnerable to traffic, being the
second most killed carnivore species in the region, with casualties occurring mainly
when the species are feeding young (Grilo et al., 2009), compromising therefore the
breeding success.
We evaluate the “montado” structural connectivity for stone marten in southern
Portugal and examined how it influences road mortality risk by addressing three
questions: 1) how does landscape permeability influence the occupancy by stone
marten of “montado” landscapes? 2) is there any relation between landscape
connectivity and stone marten road mortality? and 3) if so, which road-related features
are responsible for high road mortality in connected areas? Because little is known
about how land use patterns influence movements by stone marten, we developed a
series of 12 connectivity models using cost surface scenarios on the basis of differential
weights assigned to each vegetation cover type (e.g., Ray et al., 2002) to reflect different
resistance to stone marten movement. Then, we examined how changing the cost
surface values of land use affected the fit of models relating to stone marten
“montado” occupancy. We used the best scenario to evaluate what role connected
areas played in the mortality of stone marten.
113
113
Modelling structural connectivity for stone marten in southern Portugal
2. Material and methods
2.1 Study area
The study area comprises 7680 km2 in the Alentejo province of southern
Portugal (Figure 1). This region is characterized by its vast plains, with elevations
ranging from 200 m to 500 m a.s.l. The climate is Mediterranean, with mild winters and
hot and dry summers. Mean annual temperature is 15.6 °C and mean annual
precipitation is 500 mm.
Figure 1 – Study area in southern Portugal with the location of the sampling units to assess
stone marten patch occupancy.
The landscape is dominated by a cropland matrix (44% of the study area),
“montado” lands (39%), an agro-silvo-pastoral mix that includes both cork and holm
oaks, followed by conifer and eucalyptus plantations (8%) and orchard patches/olive
yards (7%). The study area´s mean human population (excluding the two major cities,
Évora and Beja, with 1980 inhabitants.km-2) is 21 inhabitants.km-2 (source: INE 2002).
Mean road density is 0.27 km.km-2, including 8% of highways (ca. 160km), with traffic
volumes ranging between 6 000 and 13 000 vehicles per day.
114
114
Modelling structural connectivity for stone marten in southern Portugal
2.2 Land use map
Prior to field work we created a land use map using a geographic information
system. We used the COS’90 land cover (a vectorial information produced by photo
interpretation of 1990s aerial photographs; source: IGP) and a georeferenced 1998
SPOT imagery with a spatial resolution of 20m (source: IgeoE). Many of the original
630 land cover classifications were represented by such a small fraction of total area
that we could ignore their presence. We aggregated the data into four classifications.
These included: “montado” (MNT), open areas (pastures, meadows and agriculture
areas) (OPEN), riparian areas (RIP), and other (OTHER).
2.3 Stone marten survey
From May to September 2005, we surveyed 44 sampling units (SU) in 25
different sized “montado” patches (ranging from 55 to 526 ha), and in 19 sites in
contiguous “montado” areas (11 668-60 175 ha) (Figure 1). Sampling units were located
at least four kilometers apart to prevent sampling repeated individuals (mean distance
= 6 838 ± 2 249 m). In each SU we placed 7 to 17 scent-stations based on area. A scentstation was composed of one square plate (1 m x 1 m) covered with marble dust to
detect tracks and were placed in a 300m interval grid and baited with chicken parts
and lure (Cavens GUSTO Lure) (e.g., Sargeant et al., 2003). Scent-stations were checked
for stone marten signs every second day for six consecutive days. For each SU, results
were coded as either present (if detected at least once), or absent.
Stone marten road-kills were obtained from August 2003 to December 2007 in
two types of roads: national roads and highways (Figure 1). National roads consist of
one-lane in each direction with an annual average daily traffic ranging from 500 to 2161
vehicles/night and a speed limit of 90/100 km.h-1. Highways consist of a total of fourlanes with a 7 m wide median strip, an average of 330 to 2494 vehicles/night, and with
a speed limit of 120 km.h-1. We conducted road-kill surveys at low speed (ca. 50 km.h-1)
along six segments of national roads (302 km, Figure 1), in a fortnightly basis (106
surveys) and casualty locations were recorded using a handheld GPS (maximum error
5m). Road-kill data from highways (166 km, Figure 1) was obtained from the BRISA
115
115
Modelling structural connectivity for stone marten in southern Portugal
(highway private concession) database being surveys performed by the BRISA staff on
a daily basis. Locations were recorded using nearest hectometre mark (maximum error
50 m).
2.4 Data analysis
2.4.1 Landscape connectivity models
A series of landscape connectivity models (LCM) was developed using cost
surface scenarios with different resistance weights to represent different movement
costs (e.g.; Ray et al., 2002; Adriaensen et al., 2003; Gonzales and Gergel, 2007). The
LCM were created on the basis of two assumptions: 1) over time the occupancy of
smaller and more isolated “montado” patches is to great extent influenced by the
effective distance to “montado” source areas, and 2) effective distances are dependent
on different land use patterns with different connectivity values for stone marten,
given their characteristic movement abilities. We assumed that different vegetation
cover types influenced stone marten movements and we considered stone marten
source areas as those sampling units with an occupancy probability higher than 90%.
In this context, we measured distance to source areas using several scenarios in order
to obtain the best scores that fit with stone marten occupancy data. These scenarios
were built by creating cost weight grids and changing the relative weights of each
vegetation cover. The selection of the stone marten source areas was supported by
comparing the probability of detection between “montado” contiguous areas vs.
patches, following the MacKenzie et al. (2002) procedure. Analyses were conducted
using
Presence
software
(Hines,
2006)
available
at
http://www.mbr-
pwrc.usgs.gov/software/doc/presence/presence.html.
Cost weight grids were obtained by two sets of scenarios (Table 1). For the first
one we took into account different movement resistance scores related to MNT, OPEN,
riparian areas within “montado” (RIP_MNT), and riparian areas within open areas
(RIP_OPEN). Eight cost distance grids were derived: A – equal resistance for all
vegetation cover types (Euclidian distance); B, C and D – increasing resistance for
OPEN, no effect of riparian areas presence; E, F, G and H – lower resistance of riparian
areas (lower value of RIP_MNT and RIP_OPEN), varying the resistance of MNT and
116
116
Modelling structural connectivity for stone marten in southern Portugal
OPEN. A second set of four scenarios (I, J, K and L) resulted from the addition of cost
weight grids derived from scenarios A to D with the grid containing the distance to
riparian areas (Dist_Rip_Areas).
Table 1 – Scenarios with values of relative connectivity for
land uses: “montado” (MNT), open areas (OPEN),
riparian areas within “montado” (RIP MNT) and riparian
areas within open areas (RIP OPEN) (A to H), and sum of
A to D scenarios with distance to riparian areas
(Dist_Rip_Areas) (I to L).
Scenario
MNT
OPEN
A
1
1
RIP
MNT
1
B
1
2
1
2
C
1
5
1
5
D
1
10
1
10
E
2
5
1
1
F
1
5
1
1
G
2
10
1
1
H
1
10
1
1
I
RIP OPEN
1
Scenario A + Dist_Rip_Areas
J
Scenario B + Dist_Rip_Areas
K
Scenario C + Dist_Rip_Areas
L
Scenario D + Dist_Rip_Areas
This enabled us to evaluate the influence of riparian area proximity while
varying the connectivity of the vegetation cover. Mean cost weight values of all
scenarios were then assigned for each patch. All grids were built using a cell size of
200m. For each patch the total area (Area) was also measured. The cost distance
analysis was performed with tools from the Spatial Analyst and Patch Analyst
extensions (http://flash.lakeheadu.ca/~rrempel/patch/) of ArcView 3.2 (ESRI, 1999).
We applied GLM logistic regression for patch data to identify which
connectivity model best fitted the likelihood of stone marten occupancy. All variables
were log transformed to achieve normality assumptions (Quinn and Keough, 2002).
Models were built using all possible combinations of non-correlated variables (r < 0.70).
Basically they were built by using patch area or scenarios alone, or by including one
distance measure at a time along with Area. For each model, the Nagelkerke R2, and
the area under the receiver operating characteristic (ROC) curve (AUC) were calculated
for the assessment and comparison of the effectiveness of the models to predict stone
marten patch occupancy. Models were ranked according to AIC scores and ∆AIC was
determined. We considered that a value of ∆AIC higher than 7 indicated that a model
117
117
Modelling structural connectivity for stone marten in southern Portugal
had a poor fit relative to the best model, whereas a value less than 3 indicated that a
model was equivalent (i.e., equally informative) to the minimum AIC model (Burnham
and Anderson, 2002). To test the accuracy of the best model, we applied the K-fold
cross-validation prediction error for generalized linear models (Manly et al., 2002). This
function calculates the mean sum square of the differences between the observations
and the predicted value by using R package boot.
2.4.2 Road kills analyses
We used the Linear Nearest Neighbor Analysis (Levine, 2004) to evaluate
whether stone marten casualties were clustered or dispersed along road surveyed
segments. The linear nearest neighbor index (NNI) is a ratio between the average
distance of the nearest neighbor and the expected random distance. If the observed
mean distance is the same as the mean random distance, then the ratio will be 1.0. If the
observed distance is smaller then the random, the NNI is less than 1.0 (clustered). The
locations are dispersed when NNI is larger than 1.0.
Road-kill hotspots were defined according to the Malo et al. (2004) procedure.
We compared the observed number of collisions for each 1000m of road segment; i.e.,
the mean diameter of the stone marten home range (Santos-Reis et al., 2004), with the
number of collisions expected in a random situation. If random, the likelihood of
collisions would show a Poisson distribution and we can identify a hotspot by the
minimum number of collisions in a road segment that is extremely unlikely (<5%) to
occur. By using nearest neighbor hierarchical clustering (Levine, 2004), we identified
the spatial location of these hotspots along the studied roads. We placed buffers with a
1000m radius around each hotspot centroid.
We selected the same number of buffers in road segments with no road kills
(non-hotspots) with the Hawth’s Analysis Tools extension ArcGis 9.2 (ESRI, 2006) and
compared the means of cost distance surfaces between hotspots and non hotspots with
a Mann-Whitney test (Dytham, 2003). If we found significant differences between
hotspots and non-hotspots, we identified other features related to mortality risk. In this
way, we were able to predict the locations of other road sections that are in need of
mitigation measures. The presence or absence of hotspots and five predictor roadrelated variables were then assessed using GLM logistic regression (Table 2).
118
118
Modelling structural connectivity for stone marten in southern Portugal
Table 2 – Road related variables used in road-kill hotspots analyses.
Variables
Definition
Type
Range
Nominal
1;2
Road type
1 - national road 2 - highway
Permeability_n
Number of passages
Scale
0-9
Permeability_wt
Sum of passages width (overpasses, underpasses, culverts, viaducts)
Scale
0-142
Sinuosity_c
Number of curves
Scale
0-7
Sinuosity_index
total length of the road / distance between the start and finish locations
Scale
1.00-1.12
Fences
Presence/absence of fences
Nominal
0;1
Traffic
Daily traffic volume at night in 2005 (number of vehicles/night)
Scale
351-2494
Variance partitioning procedures (e.g., Chevan and Sutherland, 1991, Mac
Nally, 2000) were used to assess the independent and joint effect of each road related
variable on hotspots buffers. This protocol employs goodness-of-fit measures for all 2N
possible models for N predictors instead of identifying a single best model (Mac Nally,
2002). Variances are partitioned so that the total independent (I’s) and joint (J’s)
contributions of explanatory variables can be estimated (Mac Nally, 2000). This
procedure allowed the identification of those variables that are likely to have the
greatest influence in explaining the locality of mortality hotspots. We used ‘log
likelihood’ as a goodness-of-fit measure. To assess the statistical significance of
variables, the data matrix was subjected to 1000 randomizations to compute I’s
distributions for each predictor. Following Mac Nally (2002), if the observed value of
Iobs for each variable was extreme (>95 percentile) relative to the generated distribution,
then we retained the variable as potentially important for explaining hotspots
occurrence. Results are expressed as Z-scores, and the statistical significance based on
the upper 95% confidence limit (Z>1.65) (Mac Nally, 2002). All statistical analyses were
carried out using the program R (http://www.r-project.org).
3. Results
3.1 Landscape connectivity models
Stone marten presence was detected in 13 “montado” patches (52%) and in 18
contiguous areas (95%). The similarity for observed (N) and estimated presence
119
119
Modelling structural connectivity for stone marten in southern Portugal
probabilities (Psi) (N - 0.52 and Psi - 0.56 for patches and N – 0.95 and Psi – 0.95 for
contiguous areas) demonstrated that our sampling method detected stone marten
presence quite well. The probability of presence of stone marten in contiguous areas
was higher than 90%, while for the patches it decreased to 56%. This supported our
assumption that contiguous areas can be regarded as source areas.
Eight connectivity models fitted quite well to the data (Table 3) with similar
AIC values, suggesting that they are equivalent. Overall results show that stone marten
patch occupancy was positively influenced by smaller distances to source areas, with
both open areas and riparian areas within those areas having a lower connectivity for
movements. Moreover, the non-Euclidian (fractal) distance with differential resistance
scores performed better in describing patch occupancy than did Euclidian distance.
The best landscape connectivity model found that the stone marten occupancy was
related to scenario L (β= -3.272; SE= 1.810) and Area (β= 3.927; SE= 2.331) (Table 3),
suggesting that the probability of occupancy decreases as the cost to source areas
increases in open areas, far from riparian areas and in smaller patches.
Table 3 – Summary of parameters estimated for the landscape connectivity models. R2:
Nagelkerke R-square; AUC: Area Under the ROC Curve; AICc: Aikaike Information Criteria
adjusted for small sample size; ∆AICc: difference between AICc of each model and the minimum
AIC found for all the models compared; w is the probability that each model is the true model.
Only null model and 3 <∆AICc<7 models are presented.
R2
AUC
AICc
∆AICc
w
Scenario L + Area
0.338
0.840
28.440
6.351
0.161
Scenario D + Area
0.317
0.833
28.972
5.820
0.124
Scenario K + Area
0.315
0.821
29.038
5.753
0.120
Scenario C + Area
0.290
0.801
29.635
5.156
0.089
Scenario J+ Area
0.260
0.763
30.344
4.448
0.062
Scenario B + Area
0.239
0.750
30.821
3.970
0.049
Scenario I + Area
0.228
0.744
31.067
3.724
0.043
Scenario L
0.189
0.7633
31.344
3.447
0.038
-
-
34.791
-
-
Model
Null
The model had a correct classification of presences and absences of 77% and
75%, respectively. The area under the ROC curve was 0.84 and appeared to be a good
predictor of stone marten patch occupancy. Nagelkerke R2 was 0.34 which suggests
that a large part of patch occupancy is explained by landscape connectivity. The
prediction error estimated by cross-validation was 0.24 which means that the model fit
120
120
Modelling structural connectivity for stone marten in southern Portugal
well to the data. When extrapolating the model to entire study area, 92% of “montado”
is likely (≥75% probability) to be occupied by stone marten being the southeastern the
least occupied area (Figure 2).
Figure 2 – Stone marten occupancy likelihood (low < 75% (light gray); high > 75% (medium
gray), source areas (dark gray) and road-kill hotspots locations (+). Dark lines are the surveyed
roads.
3.2 Road kill analyses
We documented 182 stone marten fatalities on 468 km of national roads and
highways in Alentejo. Although not significant, the nearest neighbor index revealed
that stone marten casualties were clustered (NNI=0.94). According to Malo et al. (2004)
procedure we defined as hotspots the road sections of 1000m length that had at least
four road kills over the study period. We identified 15 road kill hotspots (Figure 2) and
randomly selected another 15 sites with non-road kills. We placed buffers with a
1000m radius around each and found significant differences between road kill hotspots
and non-road kill segments for stone marten occupancy likelihood (Mann-Whitney
test: U= 44.5. n0= 15. n1= 15; p < 0.05). On average, we found a 53% likelihood of stone
marten occupancy for road-kill clusters and 23% for non-road kill segments (Figure 3).
121
121
Modelling structural connectivity for stone marten in southern Portugal
Figure 3 – Box plots for the hotspots and non-hotspots locations in relation to the likelihood of
stone marten occupancy.
We then used the hotspot analysis and selected other buffer areas without stone
marten road kills but in road stretches with a high likelihood of stone presence to
identify which road features may influence road mortality incidence. Variance
partitioning analyses suggest that road sinuosity was the most important factor
influencing the hotspot occurrences and it is the important one for management
purposes (Figure 4).
Figure 4 - Hierarchical partitioning of the proportions of explained variance of road related
variables influence on roadkill hotspots. Black bars show the variance exclusively explained by
the particular variable (I’s), open bars show the variance explained jointly by the respective
122
122
Modelling structural connectivity for stone marten in southern Portugal
variable and others. Star indicates significant result (being outside the 0.95 confidence range)
after 1000 randomizations.
Road kill data also supported the LCM model results: 87% (n=158) were within
“montado” areas or at up to 500 m distance, 10% (n=18) up to 1000m and only 3%
(n=6) ranged from 1000 m to 3220 m.
4. Discussion
Our study emphasized the value of distance from contiguous forested areas on
stone marten patch occupancy and the importance of roads as a disruptive factor in the
directionality of their movement. Moreover, the inclusion of the matrix characteristics
and the non-Euclidian distance in the models may strongly improve the understanding
of the structural connectivity of the habitat for the species, as also reported in a
growing number of studies (e.g. Hunter et al. 2003; Larue and Nielsen, 2008).
The clear species response to distance to source areas and landscape
permeability provided good insights and supported the hypothesis that patch
occupancy is species-specific and influenced by the nature of the matrix, as
documented elsewhere (e.g. With et al., 1997; Tischendorf and Fahrig, 2000; D’Eon et
al., 2002). Additionally, mobility is highly related with the species capability for
moving across landscape features with different competition and predation exposure,
energy costs, and availability of food and shelter to individuals (e.g. Johnson and
Gaines, 1985), and will affect access to and occupancy of patches.
In this study, stone martens were more likely to occur in contiguous “montado”
areas and the observed patch occupancy pattern suggests that open areas may act as a
filter to movement. Although the avoidance effect of the open areas is well
documented (Rondinini and Boitani, 2002; Virgós and García, 2002; Goszczyński et al.,
2007; Mortelliti and Boitani, 2008), little is known about the average distance across
open areas that this species will cross. Unsurprisingly, the best landscape connectivity
model had the highest resistance scores for open areas, and road kill data suggested
that crossings may be limited to distances shorter than 5000 m (Virgós and Garcia,
2002). These data suggest that species vulnerability to fragmentation may depend also
123
123
Modelling structural connectivity for stone marten in southern Portugal
on landscape type and its associated regional context, an often unappreciated
relationship.
Additionally, riparian areas appear to be of little importance as potential
corridors in promoting connectivity among the “montado” patches. This is in contrast
to what has been recently reported, especially for carnivores in fragmented
Mediterranean ecosystems (e.g. Virgós, 2001; Rondinini and Boitani 2002; Matos et al.,
2008). In fact, the riparian areas found in our study area are in poor condition, with
only a few meters width of shrubs (Matos et al., 2008), which may hamper carnivore
mobility. Nevertheless, the distance to streams combined with the proximity to
contiguous forested areas was an important parameter accounting for the variation in
patch occupancy. This finding suggests that riparian areas in the proximity of
“montado” patches might favor carnivore presence by providing additional resources.
Rondinini et al. (2002) emphasized the wide variety of fruits and plants, as well the
high density of rodents in the riparian areas which are important items in stone
marten’s diet (Genovesi et al., 1996), particularly in the dry season when their
availability apparently exceeds the surrounding matrix, even if the matrix is
“montado” (Matos et al. 2008).
Large patch areas were also highly correlated with stone marten occupancy,
which has been widely reported as one of the most important features influencing
species occupancy (e.g. Michalski and Peres, 2005), particularly with carnivores (e.g.,
Crooks, 2002; Virgós et al., 2002). Although stone martens are considered to be more
habitat generalist than other conspecifics, e.g., American marten Martes americana (e.g.,
Hargis et al., 1999) or pine marten Martes martes (e.g., Goszczyński et al., 2007), food
availability (e.g., Genovesi et al., 1996) and adequate shelter and anti-predator cover
appear to strongly influence habitat occupancy (Mortelliti and Boitani, 2008). The
detection of stone marten presence in smaller patches stressed that individuals were
able to move across the study area and use several smaller patches to supplement
resources (Dunning et al., 1992; Rödel and Stubbe, 2006). This behavior was also
reported by Palomares and Delibes (1994) with genets (Genetta genetta) in a sub-optimal
fragmented habitat in Northern Doñana National Park where radio-tracked
individuals used the same patch for several days and then changed patches to access
resources.
124
124
Modelling structural connectivity for stone marten in southern Portugal
Although some studies suggest that inter-patch movements may increase with
habitat fragmentation and therefore may also increase road mortality risk (Ferreras,
2001; Adriaensen et al., 2003), we did not find a higher frequency of road casualties in
more fragmented areas. Conversely, the incidence of road-kills was significantly higher
in road segments traversing contiguous habitat supporting the idea that well
connected landscapes promote stone martens movement and extended networks of
roads in those areas may represent a serious threat to long-term population stability
and viability (Dixon et al., 2006). Given our results, particular attention should be paid
to road segments with higher sinuosity. In highly sinuous road sections, both vehicle
drivers and animals have little time to be aware of each other’s proximity, resulting in
a shorter response time to avoid collision. Similar results were also obtained for genet
and Egyptian mongoose (Herpestes ichneumon) in a previous study (Grilo et al., 2009),
which emphasized the role of road sinuosity in the incidence of mortality, particularly
for carnivores.
The results presented here provide a regional image of the habitat connectivity
in Alentejo for stone marten and show that the low degree of landscape connectivity in
southeastern part of the study area may be critical to the persistence of the species in
this region. Riparian restoration efforts may increase the potential use as habitat or
movement corridor for this species and therefore, would enhance the landscape
connectivity of the region. In connected areas, cutting or removal of the vegetation
along road verges of highly sinuous road sections may reduce suitability for the
species, both in terms of cover and prey (e.g., Bellamy et al. 2000; Dean and Milton,
2003), that might prevent accidents by improving the visibility.
In this study, we argued that stone marten forest patch occupancy is a function
of the connectivity in Mediterranean farmland landscapes. Although the high
detectability in the source areas gives us some support in the reliability of the results,
data reported here may underestimated patch use because surveys were restricted to
dry season in a short period of time. The challenge remains to explore which processes
underlie the complex relationship between mobility, “montado” spatial patterns and
riparian areas, and examine at what distance threshold affects the long term
population viability. We believe that the large-scale perspective we present here can
125
125
Modelling structural connectivity for stone marten in southern Portugal
prove useful in guiding local conservation efforts, but future refinements are needed to
examine site-specific issues at a scale appropriate for site planning (see Jepsen et al.,
2005; Chetkiewicz et al., 2006). For example, movement patterns of adults and
dispersers with radio–tracking or non-invasive methods may provide good insights to
assess the relative resistance of different vegetation cover types and assess the viability
of population at long-term (e.g. Hunter et al., 2003; Broquet et al., 2006; Walker et al.,
2007).
5. References
Adriaensen, F., Chardon, J.P., De Blust, G., Swinnen, E., Villalba, S., Gulinck, H., Matthysen, E., 2003.
The application of 'least-cost' modelling as a functional landscape model. Landscape Urban
Plan. 64, 233-247.
Bellamy, P.E., Shore, R.F., Ardeshir, D., Treweek, J.R., Sparks, T.H., 2000. Road verges as habitat for
small mammals in Britain. Mammal Rev. 30, 131-139.
Blondel, J. 2006. The ‘Design’ of Mediterranean Landscapes: A Millennial Story of Humans and
Ecological Systems during the Historic Period. Hum. Ecol. 34, 713-729.
Broquet ,T., Ray, N., Petit, E., Fryxell, J.M., Burel, F., 2006. Genetic isolation by distance and
landscape connectivity in the American marten (Martes americana). Landscape Ecol. 21, 877–
889.
Burnham, K. P., Anderson, D.R., 2002. Model selection and multi-model inference: a practical
information-theoretic approach. New York: Springer-Verlag.
Chetkiewicz, C.B., St Clair, C.C., Boyce M.S., 2006. Corridors for conservation: integrating pattern
and process. Annu. Rev. Ecol. Syst. 37, 317-342.
Chevan, A., Sutherland, M., 1991. Hierarchical partitioning. Amer. Statistician 45, 90–96.
Crooks, K.R., 2002. Relative Sensitivities of Mammalian Carnivores to Habitat Fragmentation.
Conserv. Biol. 16, 488-502.
Dean, W.R.J., Milton, S.J., 2003. The importance of roads and road verges for raptors and crows in the
Succulent and Nama-Karoo, South Africa. J. African Ornithology 74, 181-186.
D'Eon, R.G., Glenn, S.M., Parfitt I., Fortin, M., 2002. Landscape connectivity as a function of scale and
organism vagility in a real forested landscape. Conserv. Ecol. 6 2 10.
Dixon, J.D., Oli., M.K., Wooten, M.C., Eason, T.H., McCown J.W., Paetkau, D., 2006. Effectiveness of a
Regional Corridor in Connecting Two Florida Black Bear Populations. Conserv. Biol. 20, 155162.
Dunning, J.B., Danielson B.J., Pulliam H.R., 1992. Ecological processes that affect populations in
complex landscapes. Oikos 65, 169-175.
Dytham, C., 2003. Choosing and Using Statistics: A Biologist's Guide, second ed. Blackwell
Publishing.
126
126
Modelling structural connectivity for stone marten in southern Portugal
ESRI, 1999. ArcView GIS 3.2. Environmental Systems Research Institute, Redlands, California.
ESRI, 2006. ArcGIS 9.2. Environmental Systems Research Institute Ltd., USA
Ferreras, P. 2001. Landscape structure and asymmetrical inter-patch connectivity in a
metapopulation of the endangered Iberian lynx. Biol. Conserv. 100, 125-136.
Forman, R.T.T., Sperling, D., Bissonette, J.A., Clevenger, A., Cutshall, C., Dale, V., Fahrig, L., France,
R., Goldman, C., Heanue, K., Jones, J., Swanson, F., Turrentine, T., Winter, T., 2003. Road
Ecology: Science and Solutions. Washington: Island Press.
Genovesi, P., Secchi M., Boitani L., 1996. Diet of stone marten an example of ecological flexibility. J.
Zool. 238, 545-555.
Gonzales E. K., Gergel S.E., 2007. Testing assumptions of cost surface analysis - a tool for invasive
species management. Landscape Ecol. 22, 1155-1168.
Goszczyński, J., Posłuszny, M., Pilot, M., Gralak, B., 2007. Patterns of winter locomotion and foraging
in two sympatric marten species: Martes martes and Martes foina. Can. J. Zool. 85, 239-249.
Grilo, C., Bissonette, J.A., Santos-Reis, M., 2009. Spatial-temporal patterns in Mediterranean
carnivore road casualties: consequences for mitigation. Biol. Conserv. 142, 301-313.
Hargis, C., Bissonette, J.A., Turner, D., 1999. The influence of forest fragmentation and landscape
pattern on American martens. J App. Ecol. 36, 157-172.
Hines, J., 2006. PRESENCE 2 - Software to estimate patch occupancy and related parameters. USGSPWRC.
Hunter, R.D., Fisher R.N., Crooks, K.R., 2003. Landscape-level connectivity in southern California as
assessed through carnivore habitat suitability. Nat. Areas J. 23 302-314.
Jepsen, J.U., Baveco, J.M., Topping, C.J., Verboom, J., Vos, C.C., 2005. Evaluating the effect of
corridors and landscape heterogeneity on dispersal probability: a comparison of three
spatially explicit modelling approaches. Ecol. Model. 181, 445-459.
Johnson, M.L., Gaines, M.S., 1985. Selective Basis For Emigration of the Prairie Vole, Microtus
ochrogaster: Open Field Experiment. J. Anim. Ecol. 54, 399-410.
Larue M.A., Nielsen, C.K., 2008. Modelling potential dispersal corridors for cougars in midwestern
North America using least-cost path methods. Ecol. Model. 212, 372-381.
Levine, N., 2004. CrimeStat III: A Spatial Statistics Program for the Analysis of Crime Incident
Locations (v. 3.0). Ned Levine & Associates: Houston, TX/ National Institute of Justice:
Washington, DC.
Mac Nally, R., 2000. Regression and model-building in conservation biology, biogeography and
ecology: The distinction between – and reconciliation of – ‘predictive’ and ‘explanatory’
models. Biodiversity Conserv. 9, 655-671.
Mac Nally, R., 2002. Multiple regression and inference in ecology and conservation biology: further
comments on identifying important predictor variables. Biodiversity Conserv. 11, 1397-1401.
MacKenzie, D.I., Nichols, J.D., Lachman, G.B., Droege, S., Andrew Royle, J., Langtimm, C.A., 2002.
Estimating site occupancy rates when detection probabilities are less than one. Ecology 83,
2248-2255.
127
127
Modelling structural connectivity for stone marten in southern Portugal
McRae, B.H., Beier, P., Dewald, L.E., Huynh, L.Y., & Keim, P., 2005. Habitat barriers limit gene flow
and illuminate historical events in a wide-ranging carnivore, the American puma. Mol. Ecol.
14, 1965-1977.
Malo, J.E., Suarez, F., Diez, A., 2004. Can we mitigate animal-vehicle accidents using predictive
models? J. App. Ecol. 41, 701-710.
Manly, B.F.J, McDonald L.L., Thomas, DL., 2002. Resource selection by animals: statistical design and
analysis for field studies, second ed. New York: Chapman & Hall
Matos, H.M., Santos, M.J., Palomares, F., Santos-Reis, M., 2008. Does riparian habitat condition
influence mammalian carnivore abundance in Mediterranean ecosystems? Biodiversity
Conserv 18, 373-386.
Michalski, F., Peres, C.A., 2005. Anthropogenic determinants of primate and carnivore local
extinctions in a fragmented forest landscape of southern Amazonia. Biol. Conserv. 124, 383396.
Mortelliti, A., Boitani, L., 2008. Interaction of food resources and landscape structure in determining
the probability of patch use by carnivores in fragmented landscapes. Landscape Ecol. 23,
285-298.
Naves, J., Wiegand, T., Revilla, E., Delibes, M., 2003. Endangered Species Constrained by Natural
and Human Factors: the Case of Brown Bears in Northern Spain. Conserv. Biol., 17, 12761289.
Palomares, F., Delibes M., 1994. Spatio-temporal ecology and behaviour of European genets in
southwestern Spain. J. Mammal. 75, 714-724.
Pinto-Correia, T., 1993. Threatened landscape in Alentejo, Portugal: the “montado” and other
“agro-silvo-pastoral” systems. Landscape Urban Plan. 24, 43–48.
Plieninger, T., 2006. Habitat loss, Fragmentation, and Alteration – Quantifying the Impactof Landuse Changes on a Spanish Dehesa Landscape by Use of Aerial Photography and GIS.
Landscape Ecol. 21, 91-105.
Plieninger, T., Wilbrand, C., 2001. Land use, biodiversity conservation, and rural development in the
dehesas of Cuatro Lugares, Spain. Agroforest. Syst. 51, 23-34.
Quinn G.P., Keough, M.J., 2002. Experimental Design and Data Analysis for Biologists. Cambridge
University Press.
Ramp, D., Caldwell, J., Edwards, K.A., Warton, D., Croft, D.B., 2005. Modelling of wildlife fatality
hotspots along the Snowy Mountain Highway in New South Wales, Australia. Biol. Cons.
126, 474-490.
Ray N., Lehmann A., Joly P., 2002. Modeling spatial distribution of amphibian populations: a GIS
approach based on habitat matrix permeability. Biodiversity Conserv. 11, 2143-2165.
Rödel H., Stubbe M., 2006. Shifts in food availability and associated shifts in space use and diet in
stone marten. Lutra 49, 67-72.
Rondinini, C., Boitani, L., 2002. Habitat use by beech martens in a fragmented landscape. Ecography
25, 257-264.
Santos-Reis, M., Santos, M., Lourenço, S., Marques, J., Pereira, I., Pinto, B., 2004. Relationships
between Stone Martens, Genets and Cork Oak Woodlands in Portugal, in: D.J. Harrison,
A.K. Fuller & G. Proulx (Eds.). Martens and Fishers (Martes) in Human-Altered
Environments Springer-Verlag, pp. 147-172.
128
128
Modelling structural connectivity for stone marten in southern Portugal
Sargeant, G.A., Johnson, D.H., William, E.B., 2003. Sampling designs for carnivore scent station
surveys. J. Wildl. Manage. 67, 289-299.
Sunquist, M., Sunquist, F., 2001. Changing landscapes: Consequences for carnivores, in: Gittleman J.,
Funk S., MacDonald D., Wayne R. (Eds.). Carnivore Conservation. Cambridge: Cambridge
University Press, pp. 399-418.
Tischendorf L., Fahrig, L., 2000. How should we measure landscape connectivity? Landscape Ecol.
15, 633–641.
Virgós, E., 2001. Relative value of riparian woodlands in landscapes with different forest cover for
medium-sized Iberian carnivores. Biodiversity Conserv. 10 1039-1049.
Virgós, E., García, F.J., 2002. Patch occupancy by stone martens Martes foina in fragmented
landscapes of central Spain: the role of fragment size, isolation and habitat structure. Acta
Oecol. 23, 231-237.
Walker, R.S., Novaro, A.J., & Branch, L.C., 2007. Functional connectivity defined through costdistance and genetic analysis: a case study for the rock-dwelling mountain vizcacha
(Lagidium viscacia) in Patagonia, Argentina. Landscape Ecol. 22, 1303-1314.
With, K.A., Gardner, R.H., Turner, M.G., 1997. Landscape connectivity and population distributions
in heterogeneous environments. Oikos 78, 151-169.
129
129
CHAPTER 6
GENERAL DISCUSSION
CHAPTER 6 - GENERAL DISCUSSION
The results presented in this thesis provided important insights to create a
cohesive landscape strategy for Mediterranean carnivore conservation. With a multiple
species and a multi-scale approach, this study addressed habitat connectivity for
carnivores in Mediterranean landscapes taking into account “montado” fragmentation
and road network expansion. Study results suggest that common and wide-ranging
carnivores could become threatened in the future if landscape changes persist and no
measures are taken to minimize their effects.
6.1 OVERVIEW OF MAIN FINDINGS
6.1.1 ARE COMMON AND WIDE-RANGING
“MONTADO” FRAGMENTATION?
CARNIVORES VULNERABLE TO
Not all common and wide-ranging meso-carnivore species responded equally
to “montado” fragmentation and the association between occupancy patterns and the
“montado” attributes were scale-dependent. Stone marten and Egyptian mongoose
occupancy patterns were significantly affected by “montado” fragmentation and
according to multi-scale analyses (paper I), stone marten appear to be the most
sensitive species to changes in “montado” attributes. We observed a high preference
for older-growth forests where stone marten can find arboreal cavities for rest and den
sites. This result has also been reported by Rondinini and Boitani (2002) and SantosReis et al. (2004).
The vulnerability of stone marten to fragmentation has been previously found
but with a different sampling schemes in what concerns the size range of investigated
patches. For example, Virgós and Garcia (2002) surveyed a range of patches from <1 to
>100 ha and observed that the stone marten showed low inclination to use small (<50
ha) and isolated patches. This study considered “montado” forest areas ranging from
55 to 60175 ha, and showed that stone marten responded negatively when the montado
become patchy independently from the patch size, suggesting that their sensitivity may
be higher than reported by the previous study. We found high explanatory power for
133
133
General Discussion
the joint effect of “montado” isolation and area at patch and landscape extents,
respectively. The joint effect was higher than the independent effect of those variables
at the patch scale extent as reported by Virgós and Garcia (2002). Similar results were
obtained by Mortelitti and Boitani (2007) but their analysis was made only at patch
extent. The importance of distance to large patches and the detection of stone marten in
very small patches supported the argument of Virgós and Garcia (2002) that
continuous areas, by providing healthy populations, may act as source of individuals
for small patches located in the surrounding areas. Another analysis using cost surface
scenarios (paper IV) as the basis for the differential weights assigned to each vegetation
cover type show clearly that matrix permeability is also important for stone marten
well being in Mediterranean habitats. Although distance to contiguous “montado”
areas decrease the likelihood of stone marten occupancy, the highest resistance scores
for open areas indicated that this matrix may act as a filter to movement given certain
distances between forested areas. Unsurprisingly, scent stations (paper I) and road kill
data (paper IV) show that isolation values (distance between forested areas) higher
than 1000/1500 m may decrease the likelihood of stone marten presence in the region.
Nevertheless, Virgós and Garcia (2002) observed that stone marten were scarce only
when fragments were farther than 5000 m from continuous forest areas, suggesting
that species response to fragmentation also depends on landscape type and its
associated regional context.
In contrast, Egyptian mongoose seemed to have a lower connectivity threshold
than stone marten. They were highly sensitive to understorey and “montado” density,
which had a large effect on the local occupancy model. The combined effects of local
habitat quality (montado and understorey density), patch size, and proportion of lowquality habitat in the surrounding matrix accounted for the highest variation in
mongoose occupancy. Mongoose is diurnal and a species persecuted by humans
(Palomares and Delibes 1993) being therefore understandable their needs in terms of
cover.
The detection of stone marten and Egyptian mongoose in very small patches
relatively far from large ones indicated that they may use several small patches and
moved between them to supplement food resources (see Dunning et al. 1992). Another
explanation for this finding, was documented by Mortelliti and Boitani (2008) that even
134
134
General Discussion
in smaller and relatively isolated patches, carnivores may occupy those areas if there is
food availability. Hence, within a certain cost/benefit trade-off, by improving resource
availability, it is possible to decrease the negative effect of smaller and isolated patches
and hence the negative impact of fragmentation on carnivores survival.
6.1.2 DO
ROADS HAVE IMPACT ON THE SMALL- AND MEDIUM-SIZED
CARNIVORES?
Based on our results, roads may act as filters to movement and appear to be
selectively permeable (papers II and III). Road mortality results indicated on average
47 ind./100 km/year were killed on the road, with red fox suffering the highest
mortality (42%), followed by stone marten (17%), Egyptian mongoose (12%), and both
Eurasian badger and genet (11%) (paper II). Otters, Western polecats and weasels were
less frequently killed, the first due to their dependence of water lines (Blanco 1998) and
the later due to their probable low density in the roads vicinity (e.g., Zabala et al. 2005).
Assuming that the carnivores we studied included roads in their territories, we can
predict the annual mortality rate due to collisions with vehicles. Considering a territory
of ~5km2 for Mediterranean red foxes and badgers (e.g. Cavallini and Lovari 1994,
Rosalino et al. 2005) we estimate that 50% of the red foxes that live close to the roads
die each year and the same happens to 12% of the badgers. A similar approach was
done for stone marten, genet, and Egyptian mongoose assuming that their mean
territory size was ~3km2 (Palomares and Delibes 1993, Santos-Reis et al. 2004). We
estimated that 16%, 10%, and 12% died due to vehicle collisions, respectively.
Additionally, this study provided evidence that road mortality is not randomly
distributed in time and may limit specific life-history movements with potential
consequences for population demography (e.g. Baker et al. 2007). Road mortality may
compromise breeding success for red fox, stone marten, and Egyptian mongoose and
dispersal for Eurasian badger. High red fox and stone marten casualties during the
period when provisioning young may place them at greater risk. The high number of
road casualties during the breeding and dispersal period when individuals were likely
to be moving through unfamiliar terrain and would have to cross roads may also result
in unbiased sex-ratios and age distributions with implications on the population
structure and density.
135
135
General Discussion
On the other hand, culverts and underpasses facilitate the crossing of highways
by carnivores, and appear to be important to minimize the potential barrier effect of
the highways. In general, carnivores preferred larger passages with vegetation close to
the passage entrances, favourable habitat in the surrounding area, and low disturbance
by humans. The mean passage rate for carnivores over the two studied highways
sections was 0.7 ind./passages/day (paper III). Egyptian mongoose (28%) and badger
(27%) crossed most frequently, followed by fox (18%), stone marten (12%), genet (12%)
and otter (2%); however, weasels, Western polecats, and wildcats were not detected in
the passages, suggesting that highways might be selectively permeable. Low densities
near the highway and the avoidance of artificial since they are used to move along
linear corridors such as walls and riparian habitats may explain the results referring to
these last species (e.g. Santos-Reis 1989, Rondinini et al. 2006, Mestre et al. 2007, Klar et
al. 2008). The rate of passage by some carnivores suggests that the road permeability is
not a problem. However, daily movement does not necessarily include the regular use
of passages. We suspect that individuals persisted in crossing over the road directly,
neglecting the presence of passages. Low traffic volume at night (<2500 vehicles), once
most species are restricted to night activity (Blanco 1998), may have promoted animals
to cross over the road and not use crossing structures as regularly as expected (Seiler
2003). This may be the case for individuals that incorporated roads within their
territory, as documented in an ongoing radio-tracking study with stone martens (Grilo
et al. 2008).
6.1.3 IS ALENTEJO STRUCTURALLY CONNECTED FOR CARNIVORE SPECIES?
The high representativeness of “montado” areas (papers I and IV), the low
human population (~21 inhabitants.km-2 excluding major cities) and the average road
network density (half of the mean national road network density) holds significant
potential value for carnivore guild conservation at Alentejo region. Nevertheless,
recent human induced landscape changes, mainly due to agriculture intensification
and road network expansion, may threaten carnivores species including some of those
considered common and wide-ranging (papers I, II, III and IV). In spite of Alentejo
136
136
General Discussion
being originally entirely covered by “montado” this forest is presently reduced to less
than 40% of the land cover and the quality and the configuration of the “montado”
may limit carnivore persistence. The results presented here show only a low degree of
“montado” connectivity in southeastern part of the study area for stone marten (paper
IV), the species most affected by “montado” fragmentation (paper I). But the evidence
of low distance movements (paper I) may increase the concern for the degree of
“montado” connectivity in the region for mongooses. Surprisingly, riparian areas
appeared to play only a small role as potential corridors to promote connectivity
among the “montado” patches, neither for stone martens nor for mongooses in spite of
the fact that Matos et al. (2008) have previously shown a marked association of
carnivores with riparian habitats. Only in the cost surface analysis (paper IV), did the
variable distance to streams appear to influence stone marten “montado” patches
occupancy. Therefore, the 0.45 km.km-2 of streams and the associated vegetation in the
study area seems to have little importance as habitat or provide long movement
corridor for carnivores, but rather as a landscape feature that provides shelter and
cover as well foraging opportunities (Virgos 2001, Matos et al. 2008).
The permeability of the road network is, also, a key factor for the ability of
carnivores to move and disperse. Data from 2004 show that the study area has a daily
traffic of 9 357±5 018 vehicles in the A2 and A6 and 8 341 ± 3 262 in the national roads
surveyed; this is significantly lower than the average estimates for the 1 014km of
BRISA highways (25 244 ± 23 402). The average daily values at night decrease
significantly to 1 944 ± 1 152 in A2 and A6 and 1 034 ± 456 in the national roads.
According to the conceptual model showed by Seiler (2003), these values suggest low
mortality risk to carnivore and a low avoidance effect because, apparently, with less
than 2 500 vehicles/day, animals appear to successfully cross the roads with only few
collisions with vehicles while attempting to cross. Nevertheless, the death of the
individuals even in road sections with high crossing success may not assure their
persistence close to the roads. For example, it is not yet known whether the high
number of red fox casualties in this study (paper II) is additive or compensatory to
other mortality factors. Equally, the incidence of stone marten road-kills (paper IV) in
road segments traversing well connected landscapes may represent a serious threat to
long-term population stability and viability (Dixon et al. 2006).
137
137
General Discussion
The road permeability however may be aided by the use of crossing structures.
We showed the regular use of these structures by the majority of the carnivore species.
Crossing structure availability and spacing are also indicators of road permeability
level (Bissonette and Adair 2008). On average we found 2.5 crossing structures/km
(culverts, viaducts, under- and overpasses) on the A2 and A6 highway and a lower
number (1.02 crossing structures/km) on the main national roads (314km) of the study
area, which means that the average spacing between crossings structures was 400 m
and 980 m, respectively. According to Bissonette and Adair (2008) spacing of crossings
can be determined by the scaling properties of species movement, and therefore, by the
isometric measure that express the daily (√home-range) and median dispersal distance
(7*√home-range). For species with home-ranges ranging from ~100 to ~500ha (SantosReis 1989, Santos-Reis et al. 2004, Rosalino et al. 2005), we conclude that permeability
may be assured in the region for the daily (1 000 – 2 236 m) and dispersal movement
for the carnivores guild (7 000 – 15 652 m).
6.2 CHALLENGES
AND PRIORITIES IN CARNIVORE CONSERVATION IN
ALENTEJO
The species-specific and the broad-scale analysis of the anthropogenically
changed landscape reveal uneven conservation portrait on the carnivore guild in
southern Portugal. The geographic focus of research on carnivores does not necessarily
represent areas where carnivore communities have been most adversely affected but
rather have a great conservation potential. Although some of these species have
benefited from research attention (e.g. Virgós 2001, Virgós and Garcia 2002, Mortelitti
and Boitani 2008, Rosalino et al. 2008), none of them have receive conservation efforts
or legal protection (except Western polecat, Eurasian otter and wildcat, Cabral et al.
2005).
The results presented here suggest increased attention to species that display
unexpected vulnerability, but for which the state of knowledge is still insufficient to
quantify the increasing threats for population viability over the long-term. The
emergence of small- and medium- sized carnivores as the new top carnivores in these
138
138
General Discussion
impacted landscapes is not necessarily assured. The opportunity exists in human
transformed landscapes to prevent further losses of carnivore species but this will
require additional focused research and conservation attention towards species, as well
as incorporating adequate mitigation measures in future land-use and roads
construction.
This thesis has tried to illuminate conservation priorities focused on the
carnivore guild in Alentejo. While the recommendations are not exhaustive, they
highlight important and achievable actions in the region. Therefore, we hope the
recommendations provided here inform ongoing efforts to conserve these species and
stimulate the development of new initiatives.
6.2.1 IMPROVE
KNOWLEDGE ON THE IMPACTS OF HUMAN INDUCED
CHANGES ON CARNIVORES GUILD
Among the well-studied carnivores in this study, stone marten, rise to the top in
priority for research attention. Apparently, “montado” seems to be structurally
connected for stone marten in the majority of the study area, but the functional
connectivity is only assessed by genetic analysis (e.g. Wang et al. 2008). Further,
“montado” connectivity still remains unknown for species with apparent lower
mobility, e.g., Egyptian mongoose. Hence, demonstrations of the genetic consequences
of fragmentation of habitat by anthropogenic activities, including roads, are needed to
develop more accurate plans to mitigate their effects (e.g. Riley et al. 2006, Millions and
Swanson 2007). Several other factors that are much more difficult to measure should be
included in future studies if possible. They include assessments of population
abundance, prey availability, relevant species interactions (e.g., predation), and other
factors including predator control (e.g. Patten and Bolger 2003, Tanner and Perry 2007,
Barrientos and Bolonio 2008, Mortelitti and Boitani 2008). Additional knowledge of
these variables will help to determine the direction and magnitude of the human
induced effects. For example, selective mortality effects on the demography and
population genetics of selected populations would provide most useful information. A
recent study suggest that road mortality has been implicated as a significant
demographic
force
in
turtles,
particularly
for
females,
which
are
killed
disproportionately during overland nesting movements (Gibbs and Steen 2005).
139
139
General Discussion
Therefore, long-term road surveys would help us understand if the road mortality may
induce significant changes in the sex-ratio and age-structure of populations.
When considered independently, human induced changes may have relatively
minor impacts on the wildlife conservation. However, if considered collectively, the
impacts can be devastating. A comprehensive understanding of additive or
compensatory effects of incremental human-induced changes would inform
management. Such an approach ideally should include the assessment of the habitat
suitability, characteristics of the road network, and wildlife connectivity, as they affect
populations at multiple scales – from smaller movement patterns to habitat selection at
landscape scale (e.g. Beaudry et al. 2008, Roger and Ramp 2009). It is important to
understand if and how species respond to effects expressed at finer scale extents,
because this is the scale at which management decisions are typically made to conserve
local populations. Therefore, the behavioral response of individuals towards roads
through the movement patterns analysis is required using more precise techniques
(e.g., radio-tracking) and may clarify the effectiveness of mitigation measures. Finally,
this study also enhanced an apparently unfavorable status of some species represented
in the landscape, namely the weasel and the western polecat, together with the
threatened wildcat (Cabral et al. 2005), and these deserve special attention. These
species need target studies of their basic ecology, to identify their population status,
trend and ecological requirements, and therefore the putative impacts of human
disturbance.
6.2.2 FOCUS CONSERVATION ACTION ON RESTORING HABITATS
Management policies may be more effective if they take into account speciesspecific life-history characteristics and are conducted at multiple scales (Wu 2007).
Given the need of carnivores to access resources in highly humanized landscapes such
as those in the Mediterranean, high habitat quality and connectivity to large habitat
patches appears to be key attributes for the persistence of the carnivore guild (paper I
and IV).
“Montado” conservation
and
restoration
would
benefit
by receiving
considerably more attention from scientists, land managers and policy makers.
140
140
General Discussion
Nowadays, the “montado” is a result of cleared large areas of forest/shrubby cover,
soil erosion where the resources were deteriorated, due to intensive live stocking and
cultivation (Pinto Correia 1993). Despite its current conservation value for the
carnivore guild, “montado” restoration efforts may increase its potential value as
habitat for them. Restoration is most effectively accomplished when it encompasses the
improvement of other associated landscape elements; e.g., riparian areas and
scrublands. Several studies have reported higher species richness and abundance of
most species in the riparian areas than in the surrounding matrix (Hilty and
Merenlender 2004, Lees and Peres 2008, Matos et al. 2008), suggesting a higher
importance for carnivore conservation in the most deforested and impacted areas
(Virgós 2001). Equally, the importance of riparian areas as movement corridors has
been acknowledged elsewhere (e.g. Simberloff and Cox 1987, Hobbs 1992), but we
detected no such evidence in this study. Hence, riparian habitat restoration initiatives
could enhance their role as potential corridors.
A growing number of studies have already highlighted the importance of
scrublands in providing cover and shelter for predatory activities, or minimizing the
predation risk, being highly beneficial for many carnivore species (e.g. Virgós and
Casanovas 1997, Revilla et al. 2001, Mangas et al. 2008). In the region, we found some
areas covered by scrublands as a consequence of progressive rural abandonment.
Nevertheless, large areas of scrubland are not recommended because they may
decrease the availability of herbaceous areas that provide food resources for prey and
upon which carnivores depend. Additionally, scrublands appear to promote strong
and large standing fires fire events during dry summers. In order to improve the
conservation value of the “montado” as well as the matrix quality, we suggest
managing landscapes that combine scrublands and pasture lands in order to develop a
mosaic of scrub-pastureland (Lozano et al. 2003). The objective is to create a mosaic of
vegetation types that mimic historical landscapes that were produced by the traditional
practices implemented over decades, where people conducted small shrub removal by
cattle grazing.
141
141
General Discussion
6.2.3 MITIGATE
THE EFFECTS OF ROAD CONSTRUCTION AND THEIR
ASSOCIATED IMPACTS
Existing structures built for other purposes (e.g., drainage culverts, and belowgrade local access roads) that are numerous over the roads and highways, have
documented to be crucial to mitigate the barrier effect of the road network (paper III).
The incorporation of crossing structures in the construction of major roads through
suitable habitat areas for species concerned and far from human disturbance, may
improve functional habitat connectivity and may also greatly reduced the number of
road-related mortalities (Clevenger and Waltho 2005, Mata et al. 2008). The
construction of large passages enhances permeability, because it enhances the
likelihood of a structure being used (paper III). Natural substrate on the floor of the
crossing structures appears to improve the natural continuity of the habitat and
therefore, enhances their use by animals (Yanes et al. 1995). Plant vegetation at passage
entrances to guide animals towards the existing structure; and the maintenance of
natural material inside the passage (soil, logs, rocks, woody debris) will help encourage
use by species that tend to avoid open areas and use linear corridors (e.g. weasel and
Western polecat). In contrast, frequent verge mowing in road segments with curves to
enhance visual quality are recommended; they would aid motorists in seeing animals
about to cross.
Although the number of passages needed was not addressed in this study, it
seems to be crucial that the greater the opportunities available for crossings, the lower
the possible impacts on the population demographics (Olsson et al. 2008). Bissonette
and Adair (2008) suggested that the placement of passages should be based on the
HR0.5 metric to re-establish roads permeability, but the number and spacing of
crossings alone may not prevent population isolation and decline. For example, Riley et
al. (2006), found that bobcats and coyotes that cross the freeway rarely reproduced, and
hence experienced a decreased genetically effective migration.
Currently the use of fences is a subject of great controversy among planners and
nature conservationists. Fences can reduce mortality but may increase population
fragmentation if animals cannot or do not use crossing structures (Boarman et al. 1997).
In a recent study, Jaeger and Fahrig (2004) recommended that fences should only be
142
142
General Discussion
placed in combination with crossing structures when there is a low degree of road
avoidance and a high probability of an animal being killed on the road. Although we
still don’t know for sure the mortality likelihood for carnivores, buried and small mesh
size fence to guide the individuals towards to the passages would be helpful near
passages with associated traffic volumes higher than 1200 vehicles/night (paper II),
and in preferred carnivore habitats (forested areas of natural woods and where streams
associated with riparian vegetation run through or close to the passages).
6.3 REFERENCES
Baker PJ, dowding CV, Molony SE, White PCL, Harris S (2007) Activity patterns of urban red foxes
(Vulpes vulpes) reduce the risk of traffic-induced mortality. Behavioural Ecology, 18, 716-724
Barrientos R., Bolonio L (2008) The presence of rabbits adjacent to roads increases polecat road
mortality. Biodiversity and Conservation, 18, 405-418
Beaudry F, deMaynadier PG, Hunter Jr. ML (2008) Identifying road mortality at multiple spatial
scales for semi-aquatic turtles. Biological Conservation, 141, 2550-2563
Bissonette JA, Adair W (2008) Restoring habitat permeability to roaded landscapes with
isometrically-scaled wildlife crossings. Biological Conservation 141, 482-488
Blanco JC (1998). Mamíferos de España. Vol. I. GeoPlaneta, Barcelona
Boarman WI, SazakiI M, WB Jenning (1997) The Effect of Roads, Barrier Fences, and Culverts on
Desert Tortoise Populations in California, USA. Proceedings: Conservation, Restoration, and
Management of Tortoises and Turtles—An International Conference 54–58
Cabral MJ, Almeida J, Almeida PR, Dellinger T, Ferrand de Almeida N, Oliveira ME, Palmeirim JM,
Queiroz AI, Rogado L, Santos-Reis M (2005) Livro Vermelho dos Vertebrados de Portugal.
Instituto da Conservação da Natureza, Lisboa
Cavallini P, Lovari S (1994) Home range, habitat selection and activity of the red fox in a
Mediterranean coastal ecotone. Acta Theriologica, 39, 279-287
Clevenger AP, Waltho N (2005) Performance indices t identify attributes of highway crossing
structures facilitating movement of large mammals. Biological Conservation, 121, 453-464
Dixon JD, Oli MK, Wooten MC, Eason TH, McCown JW, Paetkau D (2006) Effectiveness of a
Regional Corridor in Connecting Two Florida Black Bear Populations. Conservation Biology,
20, 155-162
Dunning JB, Danielson BJ, Pulliam HR (1992) Ecological processes that affect populations in complex
landscapes. Oikos, 65, 169-175
Forman RTT, Sperling D, Bissonette JA, Clevenger A, Cutshall C, Dale V, Fahrig L, France R,
Goldman C, Heanue K, Jones J, Swanson F, Turrentine T, Winter T (2003) Road Ecology:
Science and Solutions. Island Press, Washington
Gibbs JP, Steen DA (2005) Trends in sex ratios of turtles in the United States: implications of road
mortality. Conservation Biology, 19, 552-556
143
143
General Discussion
Grilo C, Ascensão F, Sousa J, Matos H, Reto D, Filipe J, Serronha AM, Mateus AR, Leitão I, SantosReis M (2008) Eficácia de medidas de minimização dos efeitos negativos das auto-estradas. I
Relatório de Progresso CBA/FCUL.
Hilty JA, Merenlender AM (2004) Use of riparian corridors and vineyards by mammalian predators
in Northern California. Conservation Biology, 18, 126-135
Hobbs RJ (1992) The role of corridors in conservation: solution or bandwagon? Trends in Ecology and
Evolution, 7, 389–392
Jaeger AG, Fahrig L (2004) Effects of road fencing on population persistence. Conservation Biology, 18,
1651-1657
Klar N, Fernández N, Kramer-Schadt S., Herrmann M, Trinzen M., Büttner I, Niemitz C (2008)
Habitat selection models for European wildcat conservation. Biological Conservation, 141, 308319
Lees AC, Peres CA (2008) Conservation value of remnant riparian forest corridors of varying quality
for Amazonian birds and mammals. Conservation Biology, 22, 439–449
Lozano J, Virgós E, Malo AF, Huertas DL, Casanovas JG (2003) Importance of scrub-pastureland
mosaics for wild-living cats occurrence in a Mediterranean area: implications for the
conservation of the wildcat (Felis silvestris). Biodiversity and Conservation, 12, 921-935
Loureiro F, Bissonette JA, Macdonald DW, Santos-Reis M (in press) Temporal variation in availability
of Mediterranean food resources: Do badgers track them? Wildlife Biology
Mata C, Hervás I, Herranz J, Suárez F, Malo JE (2008) Are motorway wildlife passages worth
building? Vertebrate use of road-crossing structures on a Spanish motorway. Journal of
Environmental Management, 88, 407–415
Mangas JG, Lozano J, Cabezas-Díaz S, Virgós E (2008) The priority value of scrubland habitats for
carnivore conservation in Mediterranean ecosystems. Biodiversity and Conservation, 17, 43-51
Matos HM, Santos MJ, Palomares F, Santos-Reis M, (2008). Does riparian habitat condition influence
mammalian carnivore abundance in Mediterranean ecosystems? Biodiversity and
Conservation, 18, 373-386
Mestre FM, Ferreira JP, Mira A (2007) Modelling the distribution of the European Polecat Mustela
putorius in a Mediterranean agricultural landscape. Revue d'Ecologie (Terre Vie) 62, 35-47
Millions DG, Swanson BJ (2007) Impact of natural and artificial barriers to dispersal on the
population structure of bobcats. Journal of Wildlife Management, 71, 96–102
Mortelliti A, Boitani L (2008) Interaction of food resources and landscape structure in determining
the probability of patch use by carnivores in fragmented landscapes. Landscape Ecology, 23,
285-298
Olsson MPO, Widén P, Larkin JL (2008) Effectiveness of a highway overpass to promote landscape
connectivity and movement of moose and roe deer in Sweden. Landscape and Urban Planning,
85, 133-139
Palomares F, Delibes M (1993) Key habitats for Egyptian mongooses in Doñana National Park,
southwestern Spain. Journal of Applied Ecology, 30, 752-758
Patten MA, Bolger DT (2003) Variation in top-down control of avian reproductive success across a
fragmentation gradient. Oikos 101, 479-488
144
144
General Discussion
Pinto-Correia T (1993) Threatened landscape in Alentejo, Portugal: the “montado” and other
“agro-silvo-pastoral” systems. Landscape and Urban Planning, 24, 43–48.
Revilla E, Palomares F, Fernández N (2001) Characteristics, location, and selection of diurnal resting
dens by Eurasian badgers (Meles meles) in a low density area. Journal of Zoology, 255, 291–
299
Riley SPD, Pollinger JP, Sauvajot RM, York EC, Bromley C, Fuller TK, Wayne RK (2006) A southern
California freeway is a physical and social barrier to gene flow in carnivores. Molecular
Ecology, 15, 1733–1741
Roger E, Ramp D (2009) Incorporating habitat use in models of fauna fatalities on roads. Diversity and
Distributions, 15, 222–231
Rondinini C, Boitani L (2002) Habitat use by beech martens in a fragmented landscape. Ecography, 25,
257–264
Rondinini C, Ercoli V, Boitani L (2006) Habitat use and preference by polecats (Mustela putorius L.) in
a Mediterranean agricultural landscape. Journal of Zoology, 269, 213-219
Rosalino LM, Macdonald DW, Santos-Reis M (2005) Resource dispersion and badger population
density in Mediterranean woodlands: is food, water or geology the limiting factor? Oikos,
110, 441-452
Rosalino LM, Santos MJ, Beier P, Santos-Reis M (2008) Eurasian badger habitat selection in
Mediterranean environments: does scale really matter. Mammalian Biology 73, 189-198
Santos-Reis M (1989) As doninhas ibéricas (Carnivora: Mustela) um estudo taxonómico e ecológico. Tese de
Doutoramento. Universidade de Lisboa.
Santos-Reis M, Santos MJ, Lourenço S, Marques T, Pereira I, Pinto B (2004) Chapter 7. Relationships
between stone martens, genets and cork oak woodlands in Portugal. In: Harrison DJ, Fuller
AK, Proulx G (Eds.). Marten and fishers in human altered landscapes. An international
perspective, Kluwer Academic Publishers, Massachusetts. Pp 147-172.
Seiler A (2003) The toll of the automobile: Wildlife and roads in Sweden. Ph.D. Thesis, University of
Uppsala, Sweden
Simberloff D, Cox J (1987) Consequences and costs of conservation corridors. Conservation Biology, 1,
63–71
Tanner D, Perry J (2007) Road effects on abundance and fitness of Galápagos lava lizards
(Microlophus albemarlensis). Journal of Environmental Management 85, 270-278
Virgós E (2001) Role of isolation and habitat quality in shaping species abundance: a test with
badgers (Meles meles L.) in a gradient of forest fragmentation. Journal of Biogeography, 28,
381-389.
Virgós E, Casanova JG (1997) Habitat selection of genet Genetta genatta in the mountains of central
Spain. Acta Theriologica, 42, 169-177
Virgós E, Garcia FJ (2002) Patch occupancy by stone martens Martes foina in fragmented landscapes
of central Spain: the role of fragmented size, isolation and habitat structure. Acta Oecologica,
23, 231-237
Wang Y, Yang K, Bridgman CL, Lin L (2008) Habitat suitability modelling to correlate gene flow
with landscape connectivity. Landscape Ecology 23, 989–1000
1
145
General Discussion
Wu J (2007) Scale and scaling: a cross disciplinary perspective. In: Wu J, Hobbs RJ (Eds.) Key topics in
landscape ecology. Cambridge University Press. Pp115-142
Yanes M, Velasco JMF Suaréz (1995) Permeability of roads and railways to vertebrates: the
importance of culverts. Biological Conservation, 71, 217-222.
Zabala J, Zuberogoitia I, Martínez-Climent JA (2005) Site and landscape features ruling the habitat
use and occupancy of the polecat (Mustela putorius) in a low density area: a multiscale
approach. European Journal of Wildlife Research, 51, 157–162
146
146