The Andean bear, or Spectacled bear, (Tremarctos ornatus) is the only extant ursid in South America, endemic to the tropical Andes. The Andean bear is a medium-sized ursid, with males generally weighing between 140 and 175 kg (Peyton 1999). The Andean bear is perhaps the most herbivorous of the ursids, with a diet consisting primarily of bromeliad hearts (a family of monocot flowering plants found in the tropics) and only 3.3% meat (Peyton 1980, Peyton 1999). They are also known to consume cacti pulp and fruit, various tree fruits, berries, rodents, carrion, and insects (Peyton 1980, Peyton 1999). The Andean bear range spans five countries, from southern Bolivia to northwestern Venezuela. Andean bears occur on all three ranges of the Andes mountains, from elevations as low as 250 m to as high as 4,750 m. The eastern slope of the Oriental Andes (which contains the highest proportion of cloud forest) is suspected to contain the majority of the Andean bear population (app. 85%) (Peyton et al. 1998, Peyton 1999).
Andean bears cover extensive terrain in search of seasonal food sources, from cloud forest to paramo (high mountain shrub biome) to grasslands, with some research even suggesting “continuous movements . . . within an altitudinal gradient” (Cuesta et al. 2003), however, these movements are not well-documented and have only been studied in detail in five parks (Peyton 1999, Kattan 2004). Similar to most large mammals, they are suspected to be particularly vulnerable to the insidious effects of habitat loss and fragmentation, meaning there is good reason to believe that landscape conversion is impeding Andean bear movements, including these altitudinal migrations, as well as natal dispersal (Kattan 2004). There is approximately 260,000 km2 remaining Andean bear habitat; however, this area is comprised of over 100 distinct fragments, most of which are small (between 42 and 56% of these are less than 500 km2) (Kattan et al. 2004). Only 48,000 km2 of Andean bear habitat is under protective management (within 58 separate parks/reserves), and only 15 of these parks exceed 1900 km2 (Yerena 1998), which is considered the minimum patch size required to sustain a viable population of Andean bears (Peyton 1999).
Compared to other charismatic megafauna, relatively little is known about the Andean bear. For example, there are no empirically derived estimates of Andean bear densities, and thus studies have substituted the average density of north American black bears to obtain estimates of remaining Andean bear populations based on remaining habitat (about 20,000 individuals) (Yerena and Torres 1992, Peyton 1998, Kattan et al. 2004, Goldstein et al. 2008). The lack of knowledge about the Andean bear is precisely one of the things that makes their conservation particularly challenging (Goldstein et al. 2008). According to Peyton (1999), “there is no population level management being implemented for spectacled bears in the Andes that has an empirical foundation.” Because basic data (such as litter sizes in the wild and age of first reproduction) are lacking, researchers are unable to effectively model and predict how Andean bear populations are likely changing over time (Peyton 1999), though we do know they are declining (Goldstein et al. 2008).
The rate of agricultural expansion in the tropical Andes is the primary concern for many ecologists, and some suspect that this trend in landscape conversion alone will put the Andean bear at risk for extinction (Goldstein et al. 2008). Historically, the general inaccessibility of cloud forest served to protect core Andean bear habitat, but as more roads are built, there is greater intrusion into these strongholds (Peyton 1999).
Roads are also believed to substantially impede the movements of Andean bears, with an impact width of 2 km (Peyton 1999).
Unfortunately, habitat loss and fragmentation are not the only conservation concerns for Andean bears. Poaching is considered one of the primary threats to Andean bear persistence (Peyton 1999). Once revered in many parts of their range, Andean bears are now often considered a pest species by locals. Andean bears are known to depredate nearly 20% of corn fields near forest boundaries, sometimes inflicting catastrophic losses on vulnerable farmers (Peyton 1980). Peyton (1980) surveyed 25 depredated cornfields; twenty percent had lost about half of the crop, while 3 of the 25 sites had lost the entire crop in a single incident. Andean bears are also implicated in cattle depredations, though their carrion-eating habits likely cause more cattle losses to be attributed to them than which they are actually responsible (Peyton 1999). In some places, depredatory retaliation by farmers and ranchers is thought to contribute as much to the decline of the Andean bear as habitat loss (Yerena 1998).
Andean bears are classified as “vulnerable” by the IUCN, and are covered by CITES in all five countries where they occur. However, most of these countries lack the resources to actually enforce the laws enacted for their protection (Peyton 1999, Goldstein et al. 2008).
Furthermore, any attempts to move forward with conserving or connecting primary Andean bear habitat are hindered by the lack of hard data. Cloud forests are certainly important for Andean bears, but in what proportion should other habitat types be protected to ensure that Andean bears can perform the altitudinal migrations that grant them access to food resources all year round?
Habitat Fragmentation, Landscape Connectivity, and Corridors
Habitat loss and fragmentation continue to drive worldwide declines in biodiversity, despite ongoing efforts to curb these processes and their inimical effects on organisms (Harrison and Bruna 1999, Rands et al. 2010). Habitat loss and the effects associated with it are conceptually simple to understand with broad-scale negative impacts consistently measured and accepted by biologists across the board. Fragmentation, however, is substantially less straightforward, with varying definitions found throughout the literature, often with researchers confusing fragmentation directly with habitat loss. Fahrig (2003) performed a literature review of fragmentation studies, finding that most researchers had defined fragmentation as “a landscape-scale process involving both habitat loss and the breaking apart of habitat.” Apparent in this definition is the unity of these two processes, for as once contiguous landscapes become honeycombed with various other land uses and varying degrees of anthropological influences, not only does there exist the net loss of habitat, but there follows the increased separation of remaining habitat patches from each other, between which organisms were once able to move freely. Therefore, the effects of habitat loss and fragmentation are so entwined that the effects of fragmentation per se are often obfuscated.
Though definitions of fragmentation are sometimes ambiguous, the impediment of organismal movement via fragmentation is generally believed to decrease population viability within the isolated fragments of habitat (however, it should be noted that fragmentation does not necessarily negatively impact all species, especially those that are known to do well in edge habitats) (Laurance 2008). There are a variety of mechanisms through which population viability may be impacted.
Landscape connectivity is defined as “the degree to which the landscape facilitates or impedes movement among resource patches” (Taylor et al. 1993, Tischendorf and Fahrig 2000). Thus, as fragmentation increases, landscape connectivity decreases because habitat (or “resource”) patches become ever more isolated within potentially inhospitable matrices. There are two primary measures of connectivity: structural and functional. Structural connectivity is “equated with habitat contiguity and is measured by analyzing landscape structure, independent of any attributes of the organism(s) of interest” (Tischendorf and Fahrig 2000). Conversely, functional connectivity is directly concerned with the attributes of the organisms and how these traits and behaviors impact the way it perceives and responds to landscape attributes as it moves through the environment. Functional connectivity is more in keeping with the original concept of connectivity. Though these terms may be synonymous, they may also differ. More on this later.
Prior to the introduction of island biogeography theory, fragmentation and connectivity received little attention from ecologists, but island biogeography and the metapopulation theory that followed, drove many ecologists to start considering how to mitigate fragmentation and maintain connectivity in the face of extensive habitat conversion, even in mainland habitats (Laurance 2008). The theory of island biogeography asserted that “the spatial configuration of habitats [has] an important influence on populations and communities” (Harrison and Bruna 1999). Therefore, island biogeography has had a good deal of influence on core principles of reserve design, though
Fragmentation studies now number in the thousands, and enhancing connectivity has become a priority driving a considerable investment of resources (Laurance 2008, McRae et al. 2008).
Measuring connectivity is no straightforward task, as numerous methods exist to accomplish this.
Percolation-based connectivity
One technique to preserve connectivity that has gained traction in conservation biology is either the maintenance or establishment of habitat corridors. Though these, too, are not homogenously defined in the literature, they are generally recognized as linear habitats that connect two or more core habitat areas, and, most importantly, facilitate movement of animals between them (Beier and Noss 1998, Chetkiewicz et al. 2006). Though this goal appears objective enough, there are a myriad of techniques available to identify or predict where such corridors actually appear in the landscape, and depending on which technique is used, the location of corridors may shift.