Although it is now axiomatic that global biodiversity is threatened and that species are going extinct at an accelerating rate (Ceballos et al. 2015), remarkably little is known about the distribution dynamics of many threatened taxa (Brook et al. 2006; Fahrig 2003; Grenouillet & Comte 2014; Guisan et al. 2013). This conservation crisis has spurred the development of new fields of applied research, such as conservation biogeography. Conservation biogeography applies novel conceptual approaches to classic biogeographic models to determine how environmental and anthropogenic changes influence biodiversity (Whittaker et al. 2005). At its core, conservation biogeography focuses on the theory and statistical analyses of spatial dynamics of taxa within a changing environment (Franklin 2010). These dynamic processes include critical questions on how plants and animals respond to changes in habitat availability and use, due to the effects of global warming, forest loss and fragmentation, and anthropogenic pressures (Malcolm et al. 2006; Riitters et al. 2000; Woodroffe & Ginsberg 1998).

Introduction

Seed dispersal by animals

Plants vary in the mechanisms by which they disperse their seeds. Plant dispersal mechanisms are either abiotic (wind, water and ballistic dispersal), or biotic (endozoochory and epizoochory) (Howe and Smallwood, 1982). Epizoochory refers to dispersal of seeds via adhesion to animal fur, while endozoochory refers to the consumption of fleshy fruits and seeds by frugivores, and their subsequent deposition, through either spitting or defecation (Couvreur et al., 2005). While the latter mechanism is relatively rare in temperate forests, up to 90% of tropical tree species depend on endozoochory, relying on animals as dispersal vectors (Howe and Smallwood, 1982). Different animal dispersers affect seed dispersal and seed viability via seed handling (Lambert, 1999), gut passage (Knogge et al., 2003; Valenta and Fedigan, 2009) and the spatial location of seed deposition (Valenta and Fedigan, 2010).

The effect of gut passage on seeds is a potentially crucial determinant of plant reproductive success. In some cases, studies have shown that gut passage neither helps nor hinders a seed's chance of germinating (Smith, 2004). However, the majority of research on the effect of gut passage suggests that it increases the potential for germination for numerous species (Knogge et al., 2003). For some plant species, passage through a specific animal dispersal vector is necessary for successful seed germination (Temple, 1977; Chapman et al., 1992).

Other processes influencing the success of seed dispersal are secondary dispersal and predation (Nathan and Mueller-Landau, 2000; Valenta and Fedigan, 2009). The deposition of a seed is not necessarily the final step in the plant–animal interaction, because the forest floor is home to invertebrates and vertebrates that variously prey on, or safely redeposit, dispersed seeds. One study of primate-dispersed seeds in Costa Rica found that approximately 80% (N = 793) of seeds were removed from animal feces a mere week after initial deposition (Chapman, 1989). A study of howler monkey (Alouatta palliata)-dispersed seeds in Mexico found that a majority of the dung beetle species in the area were overwhelmingly attracted to howler dung, steering clear of the feces of other, more folivorous, herbivores, resulting in the massive redeposition of defecated seeds (Estrada and Coates-Estrada, 1991).

Introduction

Anthropogenic disturbance of forests leads to fragmentation and habitat loss, which strongly influence genetic diversity, extirpations, and extinction of many species (e.g., Fahrig and Merriam, 1994; Gibson et al., 2013; Radespiel and Bruford, 2014). One of the most significant consequences of fragmentation and habitat loss is an increase in the amount of edge effects and edge habitat (Chen et al., 1992). Edge effects represent the penetration, to varying depths and intensities, of biotic and abiotic conditions from the surrounding environment (matrix) into the forest interior (Malcolm, 1994). Murcia (1995) described three ecological consequences of edge effects: (1) abiotic effects, (2) direct biological effects, and (3) indirect biological effects. Abiotic effects occur as the result of the penetration of environmental factors, such as temperature and light levels, from the matrix into the forest interior. Direct biological effects are changes in the abundance and distribution of organisms due to physical conditions near the forest edge. For example, Laurance et al. (1997) documented that wind damage led to a drastic drop in tree biomass within 100 m of the forest edge in South America. Indirect biological effects involve changes in species interactions (e.g., herbivory, frugivory), such that species that avoid the edge may be doing so due to the scarcity of preferred food items in these stochastic habitats (Mills, 1995). The combination of these three ecological processes results in dynamic, multidimensional edge habitats that differ from the matrix and forest interior. Although edge effects have been shown to alter forest structure and animal communities (e.g., Lovejoy et al., 1986; Laurance and Yensen, 1991; Laurance et al., 1997), they have rarely been studied directly in regards to primate ecology and biogeography.

Edge effects are particularly relevant to ecological and biogeographic studies in Madagascar. Madagascar has an unrivaled level of plant endemicity and diversity: of the estimated 12,000 species of plants in Madagascar, 81% are endemic to this country (e.g., Phillipson, 1994, 1996; Du Puy et al., 1999). This plant diversity is remarkable given that much of the original forest cover has been lost to human perturbations, such as slash-and-burn agriculture, logging, mining, and associated erosion (Green and Sussman, 1990; Du Puy and Moat, 1998).

Introduction

A fundamental issue in conservation biogeography is determining how and why population density (the number of individuals per unit area) and abundance (number of individual animals) vary across the geographic range of plant and animal species (Whittaker et al., 2005). Understanding these relationships is critical because they provide information on population dynamics and extinction probabilities (Wilcox and Murphy, 1985; Lima and Zollner, 1996). For example, Davidson et al. (2009) conducted a meta-analysis of 4500 mammal species and found that population density was one of the main predictors of extinction risk. Consequently, researchers have explored numerous covariates to primate density and abundance, including food quality and amount (Stevenson, 2001; Chapman et al., 2004), temporal distribution of key food resources (Terborgh, 1986), habitat structure such as basal area, stem density, tree height and size (Rovero and Struhsaker, 2007; Pozo-Montuy et al., 2011; Grow et al., 2013), habitat quality such as plant productivity (Janson and Chapman, 1999), rainfall and temperature (Pinto et al., 2009), distribution (Harcourt and Doherty, 2005), fragment area and isolation (Anzures-Dadda and Manson, 2007), and anthropogenic disturbance (Peres, 1990; Ganzhorn and Schmid, 1998; Anzures-Dadda and Manson, 2007; Peres and Palacios, 2007). Despite the wide range of species and habitats studied, few consistent patterns have emerged that explain spatial variations in primate density and abundance. Moreover, those patterns that have found statistical support tend to have been focused on large-bodied, diurnal species, leaving many unanswered questions on the conservation biogeography of small-bodied, nocturnal taxa (McGoogan et al., 2007).

Food availability

Food availability typically refers to temporal and spatial variations in food quantity and quality, and has long been considered one of the main drivers of primate abundance and density. Despite extensive investigations into the relationship between food quality and amount and primate density and abundance (Hanya and Chapman, 2012), it is often difficult to determine quantitatively rigorous measures of food availability. This issue arises due to the high plant diversity at many primate research sites in the tropics, as well as stochastic variations in the temporal availability of food resources. Many researchers have employed plant dendrometrics as proxies for food availability, such as diameter at breast height (DBH), basal area, and stem density (e.g., Wieczkowski, 2004; Anzures-Dadda and Manson, 2007; Grow et al., 2013).

Introduction

Many conservation biologists consider the lemurs of Madagascar to be one of the world's highest conservation priorities (Rylands and Konstant, 2000; Allnutt et al., 2008; Kremen et al., 2008). Of the 103 extant lemurs species monitored under the International Union for Conservation of Nature (IUCN) Red List, 70.8% (N = 73 species) are endangered or critically endangered. Moreover, 6 lemurs are on the most recent list of the 25 most endangered primates: Eulemur flavifrons, Lepilemur septentrionalis, Propithecus candidus, Microcebus berthae, Varecia rubra, and Indri indri (Mittermeier et al., 2012). Considerable research and conservation efforts have been applied to studies of some threatened lemur species and their forest habitats (e.g., Kremen et al., 1999; Seddon et al., 2000; Andrianandrasana et al., 2005). As a result of these efforts, there is a rich literature on the status of some rare lemur species and the need for concerted efforts to integrate local communities with conservation biology programs in the establishment and maintenance of protected areas (Schwitzer et al., 2014). However, most of these efforts have been applied to diurnal, large-bodied taxa, such as species in the genus Propithecus (Banks et al., 2007; Patel and Andrianandrasana, 2008; Irwin et al., 2009; Quéméré et al., 2010).

Although the Cheirogaleidae represents the most speciose lemur family, with 35 species described to date (Chapter 2; IUCN), there are relatively few data on the conservation assessments and threats experienced by taxa in this family compared to large-bodied, diurnal taxa in Madagascar. This lack of conservation attention arises, in part, from suggestions that small-bodied, nocturnal taxa have relatively low extinction risks, are ubiquitous in their geographic distribution, suffer least from anthropogenic disturbance (e.g., hunting and deforestation), and have physiological adaptations such as torpor and hibernation in many species that mitigate extinction threats (Cardillo, 2003; Liow et al., 2009; Hanna and Cardillo, 2014). The most endangered cheirogaleids have not benefitted from conservation campaigns focusing on charismatic species, due largely to their nocturnality and concomitant difficulties in visual access for tourists. Recent assessments of extinction risks in lemurs indicate increased conservation concerns for many cheirogaleids, particularly in light of studies indicating the critical role of some species as seed dispersers for endemic fruiting plants (Valenta et al., 2013), and their high levels of phylogenetic diversity (Lehman, 2006).