Introduction

Micromammals are notoriously difficult to observe because of their small size, and their cryptic, and often nocturnal lifestyle (Halle and Stenseth, 2001). Therefore, live trapping methods are commonly employed to determine distribution, estimate population densities, and infer animal activity (Gentry et al., 1965; Vickery and Rivest, 1991; Atsalis, 2000; Caro et al., 2001; Radespiel et al., 2001; Schwab and Ganzhorn, 2004; Halle 2006; Prugh and Brashares, 2010). Because trapping is an indirect method of studying animal behavior, results should be interpreted cautiously (Prugh and Brashares, 2010), and should take into account a number of biotic and abiotic factors (Prugh and Golden, 2014). Factors that can affect activity and therefore capture rates of small mammals include predator behavior (Hilton et al., 1999; Michalski and Norris, 2011; Cozzi et al., 2012; Falcy and Danielson, 2013), and the temporal and spatial availability of food and shelter (Kotler et al., 1993; Caro et al., 2001; Davidson and Morris, 2001; Kelt et al., 2004). In addition, possible affects of season, temperature, rainfall, and the lunar cycle must be considered (Kotler et al., 1991; Daly et al., 1992; Sutherland and Predavec, 1999; Stokes et al., 2001; Michalski and Norris, 2011; Maestro and Marinho,2014; Prugh and Golden, 2014).

Small mammals may reduce activity during extreme temperatures (either hot or cold) due to thermoregulatory demands (Gentry and Odum, 1957; Vickery and Bider, 1981; Falcy and Danielson, 2013). Cold is especially dangerous to small mammals because of their susceptibility to hypothermia. This might explain why old-field mice (Peromyscus polionotus) were less frequently captured during cold than warm nights (Gentry and Odum, 1957), and why common voles (Microtus arvalis) reduced activity with decreasing temperature (Lehman and Sommersberg, 1980). Some small mammals (Lyman, 1983), including small primates such as mouse lemurs (Microcebus) and dwarf lemurs (Cheirogaleus), may become completely inactive for periods of time, undergoing hibernation (dwarf lemurs) or periods of torpor (Genin et al., 2005; Blanco and Godfrey, 2014). Mouse lemurs may be in an inactive, torpid state for days, weeks, or even months during the austral winter (Ortmann et al., 1997; Schmid, 2000).

Introduction

Species divergence is driven by multiple factors, obfuscating the speciation mechanisms responsible for the diversity of life. The speciation mechanisms that dominate the literature are biogeographic models emphasizing allopatric speciation, or adaptive models invoking ecological adaptation and sexual selection (Coyne and Orr, 2004; Nosil, 2012). In the classic biogeographic model, vicariance is the dominant mechanism of speciation, in which previously interbreeding populations are sundered by a dispersal barrier (Ronquist, 1997; Ree and Smith, 2008; Ree and Sanmartín, 2009). The subdivided populations accumulate genetic differences due to drift, leading to reproductive isolation via genetic incompatibilities (Dobzhansky, 1937; Mayr, 1947; Orr, 1996). In the strict vicariant model, no divergent adaptation is necessary and the daughter species retain the niche of the ancestor, because their divergence was due to some physical dispersal barrier only (phylogenetic niche conservatism; Wiens et al., 2010). If the dispersal barriers disappear, the newly formed species may meet at contact zones, but species boundaries are maintained and hybridization is limited by genetic incompatibilities (Coyne and Orr, 1998).

In contrast, adaptive models, especially ecological speciation, posit that species diverge via local adaptation to different habitats (Nosil, 2012). Contact zones exist at parapatric species boundaries, such as in ecotones of the different habitats (Schluter, 2001). Hybridization may be limited by lower fitness of immigrants into suboptimal habitats (prezygotic), or by lower fitness of hybrids because they are maladapted to the preferred habitat of either parent species (postzygotic; Nosil et al., 2005). Immigrants into ecotones and suboptimal habitats may have lower fitness due to competition with neighboring, closely related species that are well adapted to that habitat (Bridle and Vines, 2007). Alternatively, immigrants into the neighbor species' habitats may exhibit ecological character displacement, altering their phenotypes and ecology to avoid competition and persist (Schluter, 2000; Case et al., 2005; Grant and Grant, 2006). The dynamics in contact zones offer an excellent opportunity to tease apart the subtle differences in geographic speciation mechanisms.

Speciation mechanisms related to behavior, especially sexual selection, are less frequently cited (Panhuis et al., 2001). Reinforcement is the positive selection for trait divergence driven by sexual selection and accentuated when closely related species are found in sympatry compared to in allopatry (Noor, 1999).

Introduction

Because Madagascar has been isolated from other continents for more than 150 million years (Jernvall & Wright, 1998; Kremen et al., 2008), it is characterized today by an extraordinary biodiversity that is unique to the island. Over the past 1000 years, 90% of Madagascar’s natural habitat has been destroyed, likely due primarily to human impact (Green & Sussman, 1990; Perez et al., 2005). Today many of its endemic species of plants and animals are threatened or endangered due to a combination of habitat destruction and hunting (Banks et al., 2007; Mayor et al., 2004; Wright et al., 2008). In very recent years tourism may have become an additional factor affecting Malagasy biodiversity. Madagascar’s diverse ecosystems, including spiny desert, subtropical dry forest, and rainforests, are all now contained in protected areas visited by tourists (Garbutt, 2009; Mittermeier et al., 2010).

Before the national park system was organized, tourism in Madagascar was primarily limited to beach resorts. A 15-year Environmental Action Plan was started in 1990 through which the national protected area system was established and the Association Nationale pour la Gestion des Aires Protégées (ANGAP) was organized to manage it (Wright & Andriamihaja, 2002). In 1990 there were two national parks in Madagascar, but by 2008, 18 had been established (Figure 7.1). Today the majority of ecologically minded tourists visit three of those parks: Mantadia, Isalo, and Ranomafana. In 2008, ANGAP was renamed Madagascar National Parks (MNP).

INTRODUCTION

Madagascar, “La Grande Ile” off the coast of south-east Africa, is the fourth largest island on earth. Its 587000 km2 are topped only by the islands of Greenland, New Guinea and Borneo. The island broke off from Africa some 150 to 160 million and from India some 88 to 95 million years ago. Although Madagascar is separated from Africa only by the Mozambique Channel, no more than about 300 to 450 km wide, the prevailing winds and ocean currents were and still are unfavorable for repeated colonization of the island (Krause et al., 1997). Due to the long isolation and low rate of colonization events, the flora and fauna of Madagascar underwent impressive adaptive radiations, resulting in one of the world's most diverse arrays of endemic plants and animals (Myers, 1986; Mittermeier, 1988; Table 4.1).

Based on phytogeographic criteria, the evergreen forests of eastern Madagascar are distinguished from the deciduous formations of the west and south (Du Puy & Moat, 1996; Lowry et al., 1997; Fig. 4.1). The evergreen rainforests of the east receive between 1500 and more than 3000 mm of rain per year. The deciduous forests of the west and extreme north of Madagascar are subject to a distinct dry season of four to eight months without rain and annual precipitation of 500 to 2000 mm. In both vegetation types, annual rainfall decreases from the north to the south. Parts of the south and south-west of the island receive less than 500 mm of rain per year at irregular intervals with an extended dry season for more than eight months. Here, habitats are represented by dry deciduous, riverine and spiny forest characterized by Didieraceae and other succulent plants.

INTRODUCTION

In widely different habitats and among taxonomically diverse groups, recent extinction rates are 100–1000 times their pre-human levels, with regions rich in endemics especially at risk (Wilson, 1988; Pimm et al., 1995). There is a strong recognition that many primate species are in jeopardy of extinction in the near future (Mittermeier, 1988; Rowe, 1996). Half of the 250 primate species are considered to be of conservation concern according to the Primate Specialist Group of the Species Survival Commission of the World Conservation Union (IUCN, 1996). It is an achievement for primate conservationists that we have not lost any species in this millennium, as other groups such as rodents, birds and reptiles have (Mittermeier, 1996). And primatologists should be congratulated that within the last decade more than 15 new primate species have been discovered or rediscovered (Meier et al., 1987; Simons, 1988; Hershkovitz, 1987, 1990; Meier & Albignac, 1991; Mittermeier et al., 1992; Ferrari & Lopes, 1992; Queiroz, 1992; Silva & Noronha, in press). But it is a sobering fact that 96 primate species are in the Critically Endangered or Endangered category and could disappear within the next 100 years (Rowe, 1996).

The inventory data and categories of threat developed by the Conservation Monitoring Center (IUCN) and The United States Endangered Species Act (USESA), (World Conservation Monitoring Centre, 1992; Rowe, 1996) are very useful in targeting primate species with high conservation priority. But the lists of endangered species alone have limited use in estimating causes and consequences of primate extinctions. Indeed, a major limitation of species counting in conservation biology is that ecological measures of biodiversity are not reducible to taxonomic measures (e.g., Fig. 18.1, Jernvall & Wright, 1998).