The maintenance of the evolutionary and ecological processes that guarantee the sustainability of biological diversity has been identified as the main goal of any action directed to the protection of natural environments (Purvis et al. 2005). In the face of limiting resources, the need to prioritize efforts in order to guarantee efficient conservation actions is of great importance. The identification of those taxa and areas that should be conservation priorities has been long known as the “agony of choice” (Vane-Wright et al. 1991) or the “conservation resource allocation problem” (Wilson et al. 2006). One of the characteristics of current times that complicates the classification and management of taxa and ecosystems is the uncertainty as to the conditions, especially climatic, that taxa will face in the mid and long term. Global change processes have sped up with respect to historic registries, and several possible scenarios of climatic change are considered that taxa and ecosystems will have to face in the future (Midgley et al. 2002). Several measures have been developed to quantitatively estimate the “value” of taxa and ecosystems. Currently, traditional richness data coupled with information on evolutionary patterns have been suggested as a meaningful measure of biodiversity aimed at preserving the maximum amount of character diversity, in the face of an uncertain future (Vane-Wright et al. 1991; Sechrest et al. 2002; Fisher and Owens 2004; Faith and Baker 2006; Forest et al. 2007; Pio et al. 2011). Given limited resources, conservation efforts should arguably be directed toward taxa and/or areas that represent the greatest amount of unique evolutionary history (Collen et al. 2011).
The most commonly used evolutionary measure given a phylogeny with branch lengths is phylogenetic diversity (PD; Faith 1992). Basically, PD measures the accumulation of attributes or adaptations in a taxon or a group of taxa and provides a quantitative idea of how much evolutionary history would be lost if those taxa were not preserved (Faith 1992; Purvis 2000). PD is calculated by summing the branches that connect the target taxa to either the root of the subtree, PDNODE (Faith 1992) or the root of the whole phylogeny, PDROOT (Rodrigues and Gaston 2002). The PD index is particularly significant as it considers the accumulated evolution of a set of taxa and hence its evolutionary potential (Forest et al. 2007; Potter 2008), but see Winter et al. (2013) for discussion on this topic. It has also been shown that PD of a community is positively correlated with ecosystems primary productivity (Cadotte et al. 2009). Whether or not species richness is a good surrogate for PD has been argued, and it is currently recognized that this is not always the case, given the multiple processes that affect speciation, extinction, and radiation in an area (Rodrigues and Gaston 2002; Tôrres and Diniz-filho 2004; Forest et al. 2007; McGoogan et al. 2007; Pio et al. 2011; Scherson et al. 2012). In addition to PD, phylogenetic endemism, meaning the evolutionary history that is unique to an area (Faith et al. 2004), can help improve conservation planning by identifying areas with high or unique levels of evolutionary history (Faith et al. 2004; Faith 1992; Winter et al. 2013).
Another widely used measure is phylogenetic structure (PS), which measures how dispersed or clustered is a community with respect to the tree of life (represented by a larger phylogeny that contains the taxa of interest), than expected by chance (Webb 2000). This is a topology-based measure, which relies on counting mean pairwise nodal distances between taxa in a community (Webb 2000). A suite of null models is used to estimate whether taxa in the community are more or less clustered than expected by chance, providing an idea of the resilience of a given set of taxa. A community or group of taxa with a higher than expected PS, indicating dispersion on the tree of life, has a larger evolutionary diversity and could therefore contain a higher potential to recover from stress or capacity for adaptation (Potter 2008).
In the face of global change, areas of high endemism are particularly sensitive due to the valuable and irreplaceable set of taxa that they host (Margules and Pressey 2000). A given territory is considered an area of endemism when it harbors at least two endemic non-related taxa (Harold and Mooi 1994). At a global scale, endemism of endangered plant taxa is a very important matter to consider. Over 90% of the IUCN, threatened plant species are endemic to a single country (Pittman and Jorgensen 2002) meaning that their preservation depends on a single government conservation policy. The flora of Chile is remarkable in that it harbors the highest percentage of endemicity in South America: four endemic families and 83 endemic genera (Moreira-Muñoz 2011), 67 genera in continental Chile and 16 in the islands. Argentina, for example, has one endemic family and four endemic genera (Zuloaga et al. 1999); Perú, with a flora of over 17,000 vascular plant species, more than three times the flora of Chile, has 51 endemic genera and no endemic family (Brako and Zarucchi 1993). In Ecuador, there are approximately 2110 genera, and only 23 of them are endemic (Jorgensen and León-Yáñez 1999). In an area of similar biogeographic composition such as New Zealand, there are 48 endemic genera (reviewed by Moreira-Muñoz 2011).
Highest levels of floral endemism concentrate in central Chile (Moreira-Muñoz 2011). This is not surprising, given the long-known importance of this area in terms of biodiversity, largely coincident with the location of the “Chilean winter rainfall–Valdivian forest” biodiversity hotspot, with priority for conservation (Armesto et al. 1998; Myers et al. 2000; Sechrest et al. 2002; Arroyo et al. 2004). For example, the families Aextoxicaceae, Gomortegaceae, and Lactoridaceae are restricted to this hotspot (Arroyo et al. 2008). Chile is a very centralized country, with 80% of the population concentrating in central Chile, as well as the main centers for agricultural, industrial, and services activities (INE 2002; CAPP 2008). Despite this, a very small percentage of its area, 5.5% in the north an only 1.7% in the central-southern area, is protected by the National System of Protected Wild Areas (CONAMA 2008). Regarding islands, the flora of the Juan Fernández archipelago, for example, is one of the most vulnerable in Chile and worldwide, mainly due to introduction of alien invading species from the continent, resulting in more than 75% being highly threatened (Swenson et al. 1997). Given the importance of the endemic component and vulnerability of Chilean flora, measures of its evolutionary value become instrumental for aiding conservation efforts.
The explosive growth of bioinformatics and genomics technology provides an unprecedented availability of information that can be used in evolutionary conservation (Roquet et al. 2013). The main objective of this study was to quantify the evolutionary potential of Chilean endemic genera of vascular plants, and its geographic patterns, using as a backbone a previously published vascular plant supertree (Thuiller et al. 2011), and newly added Chilean genera. In addition, the study focused on the relationship between PD and richness at different geographic levels.