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In-Depth Information
2003b; Myers et al., 2000) or bird (Stattersi eld et al., 1998) endemism, often
supported by terrestrial vertebrate endemism overall (Mittermeier et al.,
1997, 2003b; Myers et al., 2000). The logic for this is that the more endem-
ics a region holds, the more biodiversity is lost if that region is lost (even
if anywhere holding even one endemic is irreplaceable in a strict sense). In
addition to numbers of endemic species, other aspects of irreplaceability
have been proposed including taxonomic uniqueness, unusual phenomena
and global rarity of major habitat types (Olson and Dinerstein, 1998), but
these remain dii cult to quantify. Despite the fact that species richness
within a given area is sometimes assumed to be important in prioritization
(Prendergast et al., 1993), none of the approaches rely on species richness
alone. This is because species richness is driven by common, widespread
species, thus strategies focused on species richness tend to miss exactly
those biodiversity features most in need of conservation (Orme et al., 2005;
Possingham and Wilson, 2005; Lamoreux et al., 2006). Three approaches
do not incorporate irreplaceability (Bryant et al., 1997; Sanderson et al.,
2002; Hoekstra et al., 2005).
The choice of measures of irreplaceability is to some degree subjective,
in that data limitations currently preclude the measurement of biodiver-
sity wholesale. Further, these same data constraints have meant that, with
the exception of endemic bird areas (Stattersi eld et al., 1998), the meas-
ures of irreplaceability used in global conservation prioritization have
necessarily been derived from specialist opinion. Subsequent tests of plant
endemism estimates (Krupnick and Kress, 2003) have proven this expert
opinion to be quite accurate. However, reliance on specialist opinion
means that results cannot be replicated, raising questions concerning the
transparency of the approaches (Humphries, 2000; Mace et al., 2000). It
also prevents formal measurement of irreplaceability, which requires the
identities of individual biodiversity features, such as species names, rather
than just estimates of their magnitude expressed as a number (Balmford,
et al., 2000; Humphries, 2000; Mace et al., 2000; Brummitt and Lughadha,
2003).
Measures of vulnerability
Five of the templates of global conservation priority incorporate
vulnerability - measures of temporal conservation options (Margules and
Pressey, 2000; Pressey and Taf s, 2001). A recent classii cation of vulner-
ability (Wilson et al., 2005) recognizes four types of measures based on:
environmental and spatial variables; land tenure; threatened species; and
expert opinion. Of these, environmental and spatial variables have been
used most frequently in global conservation prioritization, measured as
proportionate habitat loss (Myers et al., 2000; Sanderson et al., 2002;
