Conservation biologists usually argue for a proactive approach to species conservation – making targeted investments before a species is endangered and under substantial risk of extinction (Drechsler et al., 2011, Benson, 2012 and Polasky, 2012). But management to abate conservation threats can represent significant investments; globally, annual cost to reduce extinction risk of threatened species was estimated at US$76 billion (McCarthy et al., 2012), and in the U.S., annual cost to protect endangered species from two conservation threats was estimated at US$32 – 42 million (Wilcove and Chen, 1998). Consequently, sufficient action to abate threats starts only when species are under mandated statutory protection to prevent extinction, despite the fact that costs associated with such a reactive delay-and-repair policy may be higher than those of a proactive policy (Scott et al., 2010 and Drechsler et al., 2011). Changing policies that direct species conservation from reactive to proactive processes will be one of the major challenges for the conservation community in the coming decades.
In the United States, the Endangered Species Act (ESA) of 1973 is considered as one of the world’s strongest legislation providing protection for species of conservation concern (Czech and Krausman, 2001, Taylor et al., 2005, Schwartz, 2008 and Harris et al., 2011). Like other conservation policies, the ESA is largely a reactive process. On the eve of its 40th anniversary, over 1400 wildlife and plant species were listed as threatened and endangered, and an additional 185 species were designated as candidate for listing (U.S. Fish and Wildlife Service (USFWS), 2013). Candidate status implies there is enough information to warrant protection under the ESA, but listing is precluded because other species are in greater conservation need and therefore receive a higher listing priority (Harris et al., 2011). While candidate species receive no immediate statutory protection, they can provide a unique opportunity to implement proactive management to avoid future listing and prevent them from becoming conservation-reliant species (i.e., requiring continued intervention to maintain viable populations; Scott et al., 2010 and Goble et al., 2012).
The greater sage-grouse (Centrocercus urophasianus; hereafter sage-grouse) is a year-round sagebrush (Artemisia spp.) community obligate whose populations have been declining primarily due to habitat loss and fragmentation, which prompted its candidate species designation in 2010 (USFWS, 2010). Key threats leading to sagebrush habitat loss and fragmentation include urbanization and energy development, conversion to croplands, invasion of exotic grasses, large-scale wildfires, and encroachment of conifer species (Knick et al., 2013a). It is estimated that as much as 90% of conifer encroachment in the western U.S. is occurring in sagebrush habitats ( Davies et al., 2011 and Miller et al., 2011). In its early stages (successional Phase I; Miller et al., 2005), conifer encroachment into sagebrush communities reduces shrub and herbaceous species diversity and increases bare ground ( Knapp and Soulé, 1998 and Miller et al., 2000). Overtime, trees become co-dominant (Phase II) resulting in the modification of community processes ( Miller et al., 2005 and Peterson and Stringham, 2008); sagebrush eventually lose vigor and decline in canopy cover, and conifers become the dominant species (Phase III; Miller et al., 2000 and Knapp and Soulé, 1998). Miller et al. (2000) documented non-linear declines in sagebrush to approximately 20% of its maximum cover when conifers reached 50% canopy cover. Such losses of sagebrush habitat to conifer encroachment can be detrimental to sagebrush obligate wildlife species, especially those which are already of conservation concern such as the sage-grouse ( Knick et al., 2013b, Rowland et al., 2006 and Davies et al., 2011).
Previous studies have identified the negative effects of conifer encroachment on sage-grouse by empirically sampling characteristics of used sites (e.g., Freese, 2009, Casazza et al., 2011 and Knick et al., 2013a), or by modeling habitat use using the percentage of conifer cover as a covariate (e.g., Doherty et al., 2008, Atamian et al., 2010 and Doherty et al., 2010a; but see Casazza et al., 2011). However, there is large variability in stand characteristics as they relate to successional phases after stand establishment (Miller et al., 2005), and understanding how those characteristics affect sage-grouse demographics is essential to target proactive management that is already underway. Launched on the heels of the ESA candidate designation, the Sage Grouse Initiative (SGI) is a collaborative effort between federal and state agencies, non-governmental conservation organizations, and private landowners, to increase ecological understanding, identify critical management needs, and reduce threats to sage-grouse through proactive habitat management (Natural Resources Conservation Service (NRCS) 2013). The SGI implements habitat improvement programs that include acquisition of permanent conservation easements, promotion of sustainable grazing practices, and removal of encroaching conifers (NRCS, 2012), and in the first 2 years of its existence, SGI invested over US$92 million in sage-grouse habitat management. Given such large-scale investments and the immense conservation task at hand, it is important to target SGI’s actions to maximize conservation return for every dollar spent.
In this paper we modeled sage-grouse demographics as a function of conifer stand characteristics in eastern Oregon. We demonstrated the application of such analyses to conservation planning by using modeling results to identify areas with high prevention and restoration management potential and to estimate the costs to apply such management. Overall we sought to better understand how conifer stand characteristics relate to sage-grouse demographics to provide guidance for the proactive conservation of this candidate species.
