Research articleUngulate management in European national parks: Why a more integrated European policy is needed
Introduction
Protected areas are a cornerstone of both national and international conservation strategies for preserving the functioning of natural ecosystems and to halt the loss of biodiversity (Dudley, 2008; Leroux et al., 2010). Their establishment has increased during the twentieth century because of concern over environmental degradation (Dudley, 2008; Watson et al., 2014). Despite this trend, species extinctions caused by human activities continue at an alarming rate (Convention on Biological Diversity, n.d.).
Most conservationists advocate greater attention to the protection of biodiversity and increased protected area coverage. However, biodiversity loss is likely to increase unless the effectiveness of protected area management is improved (Chape et al., 2005; Laurance et al., 2012). To increase its effectiveness, the underlying mechanisms causing biodiversity loss must be understood, management objectives should be clearly defined, and management practices should be tailored to the prevailing biological and social context within protected areas. Furthermore, appropriate governance systems and resources are required to be able to achieve conservation objectives (Chape et al., 2005). This goal is complicated by competing management objectives of the different stakeholders, which often results in complex management plans (Dupke et al., 2019; Leroux et al., 2010; Naughton-Treves et al., 2005; Watson et al., 2014). To overcome these complications and to ensure long-term nature conservation, the International Union for Conservation of Nature (IUCN) has developed standardized guidelines for six protected area categories, classifying protected areas for purposes of planning, setting regulations, and negotiating land and water uses (Ia Strict Natures Reserve, Ib Wilderness Area, II National Park, III Natural Monument or Feature, IV Habitat/Species Management Area, V Protected Landscape/Seascape, VI Protected Area with Sustainable Use of Natural Resources) (Dudley, 2008). Although the six categories are recognized as the global standard by the United Nations and many national governments, they are not consistently implemented and there is no consistent use of the terminology. For example, in the United Kingdom, several protected areas use the term national park, but more correctly fit the definition of Category V Protected Landscape, due to their cultural value and continuous human intervention (Department for Environment Food and Rural Affairs, 2016; National Parks UK, 2018). Hence, decisions on proper management practices not only depend on guidelines set by the IUCN, but also on the cultural and political context of the particular area, which determines the legislation (Theuerkauf and Rouys, 2008). Consequently, management practices of different areas within the same IUCN category can differ.
Management of national parks and other protected areas is further complicated by different opinions on the primary objectives because of concerns over the impact of wildlife on lower trophic levels and cascading effects on vegetation (Côté et al., 2004; Demarais et al., 2012), animal population declines due to human–wildlife conflicts (Woodroffe and Ginsberg, 1998) and the spread of diseases (Gortázar et al., 2007; Putman et al., 2011). Within the primary objectives of national parks, the protection of biodiversity and natural processes are emphasized (Dudley, 2008). Other objectives defined are, among other things, to maintain viable and ecologically functional populations, to take the needs of indigenous people and local communities into account and to contribute to local economies through tourism (Dudley, 2008). However, management intervention and other human influences are only allowed at a level that will not cause significant biological or ecological degradation (Dudley, 2008). As a result, ungulate populations within national parks should ideally be regulated by food availability, interspecific competition and predation by large carnivores (Sinclair, 1998).
The persecution of large carnivores in the past (Ripple et al., 2014) has led to their disappearance in many parts of Europe and North America. In the last decades, large carnivores are successfully recovering in large parts of Europe (Chapron et al., 2014). However, most of the continent supports low carnivore densities, and the proportion of their geographical range covered by European national parks is small, as they require extensive areas of habitat to maintain viable populations (Soulé et al., 2003; Woodroffe and Ginsberg, 1998). Furthermore, large carnivores within European protected areas (Rauset et al., 2016) or at their edges are prone to legal hunting and poaching, which will consequently reduce their densities and affects their functional role (Kowalczyk et al., 2015; Müller et al., 2014; Woodroffe and Ginsberg, 1998). In protected areas where large carnivores are absent or occur in low densities and in protected areas located within highly productive environments, ungulate populations can be high (Hansen and DeFries, 2007; Melis et al., 2009). Populations of large ungulates occurring at high densities can severely affect plant communities through extensive browsing and consequently change vegetation structure and impact natural processes and subsequently local biodiversity (Apollonio et al., 2017; Demarais et al., 2012; Fuller and Gill, 2001; Gill, 1992; Kuijper et al., 2009). These direct and indirect impacts have triggered much debate about appropriate management of large ungulates (Demarais et al., 2012). Human intervention in the form of culling or artificial feeding to reduce the pressure on the natural vegetation, to compensate for lost wintering grounds or food shortage are often considered necessary (Conover, 2001; Möst et al., 2015; Putman and Staines, 2004). In Europe, hunting is currently the main cause of mortality for ungulate populations. Thus, park authorities have to balance the objectives of national parks to protect biodiversity and to keep human interference to a minimum to determine appropriate management practices.
The naturalness concept has proven to be a valid management tool for quantifying the intactness or integrity of ecosystems (Anderson, 1991; Cole et al., 2008; Steinhoff, 2012; Winter 2012; Winter et al., 2010). Although a commonly accepted definition of naturalness is lacking, it has been proposed that naturalness is a non-binary variable that can be described relative to ecosystem structure and human activity (Anderson, 1991; Günther and Heurich, 2013; Leroux et al., 2010; Winter 2012; Winter et al., 2010) or that naturalness or a natural system is a self-regulating ecosystem that should be free of any human influence (Anderson, 1991; Cole et al., 2008). The naturalness concept has been used descriptively to provide a conceptual framework for the evaluation of ecosystems (Anderson, 1991) or quantitatively in terms of defined naturalness indicators (Winter 2012). Such naturalness indicators have to be able to detect differences between variables measured in the field and a reference system, which in general should represent the most natural state (Winter 2012). Some studies associate this with the state before human colonization, or subject to limited human intervention (i.e. Pleistocene, early Holocene) (Anderson, 1991; Winter et al., 2010). The assessment of how natural a system is provides a relevant framework for ecosystem maintenance and restoration (Winter 2012). However, the feasibility of using the naturalness concept solely to define natural ecosystem integrity has been questioned (Porter and Underwood, 1999; Winter et al., 2010). First, because completely natural ecosystems no longer exist, due to the global effects of human activities (Cole et al., 2008; Winter 2012). Second, because management decisions should not only focus on the desired state of naturalness but also describe which natural processes should be preserved (Cole et al., 2008; Theuerkauf and Rouys, 2008). In previous studies the naturalness concept was applied to describe the current status of (forest) ecosystems using a reference system (for an overview, see Anderson, 1991) to influence forestry management practices or to enable legislation and management policy for the preservation of protected areas (Cole et al., 2008; Steinhoff, 2012; Winter 2012; Winter et al., 2010). In contrast, we used the naturalness concept as a tool to rank ungulate management in terms of the natural state (species composition) and the processes influencing the natural state (assessment of human impact on the system). So far, only one study has focused on the relationship between ungulate management and the naturalness concept, where a red deer management naturalness index was calculated for twenty national parks using five pre-defined naturalness indicators (Günther and Heurich, 2013). Our study aims at filling this gap by assessing the current status of ungulate management, evaluating the diversity of ungulate management within European national parks using the naturalness concept and to analyse which variables influence the naturalness of ungulate management. We believe that a better understanding of ungulates and their management in Europe is necessary to develop ungulate management strategies in accordance with the objectives defined for protected areas.
Section snippets
Methods
We collected data on management objectives and practices within national parks across Europe by means of an electronic questionnaire survey. The survey (Fig. 1) was sent to experts in wildlife management in each European country, who subsequently distributed it to local wildlife management authorities. Each participating national park submitted one completed questionnaire for our analyses. Similar to Winter et al. (2010), we focused on national parks as these represent the areas in Europe with
General characteristics of national parks
Of the 335 national parks from 31 countries across Europe contacted, managers of 209 national parks from 29 countries responded (Fig. 1; Table 1). Although all of these parks identified themselves as a national park, only 87.5% correspond to IUCN category II and consequently are managed following the IUCN guidelines. Of the national parks in the study, 88.5% were established between 1951 and 2016, of which most were established between 1991 and 2000 (23.4%; Table 1). The national parks
Discussion
Our study provides new insights on the current status of the naturalness of ungulate management in European national parks and identifies the explanatory variables that influence the naturalness of ungulate management. Our results prove that the cultural and political context of the country more strongly influences ungulate management in European national parks compared to the other variables tested. Park size, the human influence index and the percentage of government-owned land also
Conclusions
The results of our study reveal the usefulness of the naturalness score as a tool to assess the current status of ungulate management and the diversity of the management, and to determine which variables influence the naturalness of ungulate management. Our study shows that despite IUCN categorization of protected areas and the standardized guidelines set for protected area management, ungulate management greatly varied among European countries and that most European national parks do not
Author contribution
Suzanne T.S. van Beeck Calkoen; Conceptualization, Methodology, Formal analysis, Investigation, Writing – Original Draft, Visualization, Project administration. Lisa Mühlbauer; Conseptualization, Methodology, Resources. Henrik Andrén; Resources, Validation, Writing – Review&Editing. Marco Apollonio; Resources, Validation, Writing – Review&Editing. Linas Balčiauskas; Resources, Validation, Writing – Review&Editing. Elisa Belotti; Resources, Validation, Writing – Review&Editing. Juan Carranza;
Acknowledgements
We thank the staff of all European national parks for kindly responding to our questionnaire survey and we would like to thank all others involved for their help in the data collection. Further, we would like to thank University of Aveiro, Portugal and FCT/MEC for the financial support to CESAM RU (UID/AMB/50017) through national funds and co-financed by the FEDER, within the PT2020 Partnership Agreement as well as the Instituto da Conservação da Natureza e das Florestas (ICNF) for the data
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