Using null models to identify under-represented species in protected areas: A case study using European amphibians and reptiles
Introduction
Europe is one of the most intensively studied areas of the world with respect to its biota (Hawksworth and Bull, 2008). Biodiversity of Europe includes about 1500 vertebrate species, 20–25,000 species of vascular plants and well over 100,000 invertebrate species (IUCN European Red List, 2010). Apart from this high diversity of flora and fauna, a large proportion of European species can be considered as endemic to the region: from 27% of terrestrial mammal species to 88% of freshwater molluscs (IUCN European Red List, 2010). However, it must be also considered that Europe is a continent where landscapes have been subjected to an almost unparalleled amount of human disturbance and industrial development (Hawksworth and Bull, 2008).
Biodiversity loss is one of the main environmental challenges facing the planet (Sala et al., 2000). The establishment of protected area (PA) systems is the most common approach to prevent this loss of biodiversity, and is the mainstay of current conservation policies (Gaston et al., 2008, Le Saout et al., 2013). Consequently, one of the main issues in conservation biology is assessing how much biodiversity is currently represented by PAs using various methods (Cabeza and Moilanen, 2001, Pressey, 1994, Rodrigues et al., 2004b). One of the most common is ‘gap analysis’, a planning approach based on the assessment of the comprehensiveness of existing protected area networks (Jennings, 2000, Scott et al., 1993) by overlapping the distribution maps of individual species, habitats or other attributes of biodiversity with the PA network (Campos et al., 2013, D’Amen et al., 2011, e Silva et al., 2014, Rodrigues et al., 2004a, Vasconcelos et al., 2012). This allows exploration of the amount and nature of biodiversity falling within or outside these PAs and the identification of elements that need further protection through systematic conservation planning (Margules and Pressey, 2000).
However, this approach involves two major methodological problems when deciding if a species is “represented” in a PA. Firstly, the resolution of species distribution data is often coarser than that of PAs. Reliable species distribution data are usually available as regular grids on maps (e.g., distribution atlases), especially in broad scale studies, while reserves are represented as vectorial data, making it difficult to know whether species recorded in a particular grid cell (or spatial unit) actually occur within PAs. Thus, matching PAs with coarse planning units for species requires the definition of arbitrary rules to assign PAs to these planning units (Araújo, 2004), and subjectivity in the choice of a specific threshold to consider a cell as part of the PA would affect the results of species representation (see Abellán and Sánchez-Fernández, 2015, Araújo, 2004). The second uncertainty lies in the choice of a representation target to consider a species as represented by a PA network (Vimal et al., 2011). Frequently, a species is considered as represented or covered in gap analyses when it is present in at least in one, or at most a few, of the cells identified as a PA (e.g. Araújo et al., 2007, Rodrigues et al., 2004b). Although this is the most traditional measure of effectiveness (Rodrigues et al., 1999), it may be misleading, especially if the species are only represented at sites that are inadequate to ensure their persistence (Carvalho et al., 2011a, Gaston et al., 2002, Sánchez-Fernández et al., 2013). Hence, recent papers claim for the need to assess sensitivity of reserve effectiveness assessments to data and decision-rules (e.g. see Abellán and Sánchez-Fernández (2015) for an example using the same dataset than here), as the obtained results could be completely different depending on (i) the threshold used to solve the mismatch between the spatial resolution of data for species and PAs (Araújo, 2004) and (ii) the different conservation targets used to consider a species as represented by a PA network (Vimal et al., 2011). Furthermore, such targets, criteria and thresholds usually focus on just a part of the entire distribution of a species, are arbitrary and sometimes difficult to defend, especially when conservation priorities are appraised against competing land uses (Young et al., 2005).
A complementary form of evaluation that avoids fixed targets would be to explore the representation of the entire distribution of a given species in a PA network. Then, null models could be used to test the departure of the species’ representation in PAs from that expected by chance (based on the same number of planning units with occurrences of the species taken at random), which would allow identifying if the species is over and under-represented by the PA network. Although null models have been used to assess the effectiveness of protected areas, this was always done with the purpose of estimating the number of species that would be expected to be represented by chance given a determined number of cells, using the target and threshold as explained above (e.g. Araújo et al., 2007, Popescu et al., 2013).
We present here such an alternative and complementary approach to identify under or over-represented species by protected areas, avoiding traditional methodological uncertainties and providing a value of significance using null models. As an example, using this approach, we determined how well amphibian and terrestrial reptile species in Europe were represented by two protected area networks: nationally designated protected areas (NPAs) and the Natura 2000 network (N2000). We also tested to see if there were any differences in species representation depending on their conservation status, range size, type of distribution and if they were endemic to Europe or not. We expected that threatened endemic species with narrow distribution in Europe would be better represented in PAs than the remaining species and also than random species with the same number of occurrences. We focused on amphibians and reptiles because (i) they are two of the groups with the most comprehensive and updated information on their distribution in Europe (Sillero et al., 2014); and (ii) most of their species are threatened worldwide (e.g. Alford and Richards, 1999, Böhm et al., 2013, Gibbons et al., 2000, Houlahan et al., 2000), and especially in highly human-transformed landscapes such as European territory (IUCN European Red List, 2010).
Section snippets
Species data and study area
Species distribution data for both reptiles and amphibians were obtained from the recently available New Atlas of European Amphibians and Reptiles (NA2RE; Sillero et al., 2014; see also http://na2re.ismai.pt) which covers all European herpetofauna. This study provides taxonomical information and maps species distributions on an equal-area grid with cells of 50 × 50 km based on the Universal Transverse Mercator (UTM) projection and the Military Grid Reference System (MGRS). We used part of NA2RE
Results
The MPO between amphibians and both NPAs and N2000 ranged from 0.8% to 54.5% (mean 16.95 ± SD 9.73%) and from 13.3% to 47.4% (21.95 ± 7.42%), respectively (Table 1). Values for reptiles were quite similar, ranging from 2.9% to 60.6% (mean 16.99 ± 10.55%) and from 14.4% to 79.5% (27.84 ± 10.12%) for NPAs and N2000, respectively (Table 1). We found that only 19 (29.7%) and 22 (34.4%) out of the 64 amphibian species included in the analyses were better represented within the NPAs and N2000, respectively,
Discussion
The approach presented here offers a complementary view to traditional gap analysis in order to assess how well biodiversity is represented by PA systems. One of the main advantages of our approach over traditional gap analyses is that it allows obtaining a continuous value of representativeness for each one of the species from the degree of spatial overlap between its entire distribution in the study area and the protected area system, thus avoiding the need to select an arbitrary
Concluding remarks
Our approach avoids two major typical methodological uncertainties usually involved in gap analysis assessments; namely, the definition of rules (i) to assign PAs to PU, and (ii) to identify species as represented within a given PAs network. The method presented here provides complementary information to traditional assessments of protected areas’ effectiveness or gap analyses, and it is particularly useful especially when the aim is to explore differences in the effectiveness of PA networks in
Acknowledgments
We thank Salvador Carranza for his interesting comments on previous versions of the manuscript and Melissa Crim and Ana María Soria for checking the English. We also thank two anonymous referees for their constructive comments. D.S.-F. was supported by a postdoctoral grant (Juan de la Cierva program) from the Spanish Ministry of Economy and Competitiveness.
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