Overabundant wild ungulate populations in Europe: management with consideration of socio-ecological consequences

Introduction

In the last few decades, throughout Europe, ungulates have experienced a dramatic increase in numbers and geographic range size (Apollonio et al. 2010a). This increase is mainly linked to profound human socio-demographic changes, including the abandonment of land and traditional agricultural practices, the consequent habitat re-naturalisation (resulting in increased food availability for ungulates), more restricted hunting legislation, and the reduction of native predators in some regions, for example the wolf Canis lupus in the Iberian Peninsula (López-Bao et al. 2015, Torres & Fonseca 2016, but see Hindrikson et al. 2017). Moreover, the extended anthropogenic impact (with contrasting consequences; Tucker et al. 2018), global warming (Melis et al. 2006, Mysterud & Sæther 2010), ungulate biology including some species-specific traits, such as high plasticity (Laddomada 2000), and high reproductive rates (Fonseca et al. 2011) have contributed to increasing numbers and range sizes of ungulates. This rise is further aggravated by the general European trend of declining numbers and increasing age of hunters (Massei et al. 2015, Carpio et al. 2017), and by local reintroductions (Apollonio et al. 2014). Although ungulates can bring advantages to ecosystems and society, for example through financial means owing to hunting and ecotourism, or through their natural role in the environment, we are focusing here on some of their negative consequences.

There are already scenarios of ungulate overabundance in some European countries, but what is overabundance? Overabundance may be defined rigorously as too many animals, but the rigour ends there, according to Caughley (1981). Overabundance situations were identified first in the USA (Leopold 1943, where ‘deer would rather starve than move’), and later in Europe (Gossow 1986, although some concerns were raised earlier by Mitchell et al. 1977). Since then, a lot of work has been done trying to define and predict these situations (Côté et al. 2004, see also Carpio et al. 2017). Early overabundance studies were focused on the description of the impacts of overabundant animal populations on the ecosystem (Caughley 1981) in order to produce specific management suggestions (Mysterud 2010, Boadella et al. 2012a, Sorensen et al. 2014). Overabundance of animal populations can be defined in three different ways. A population can be considered overabundant when: (1) it conflicts with human populations (Caughley 1981, Côté et al. 2004); (2) its abundance leads to declines in populations of other species (Carpio et al. 2014) or causes damage to agriculture or forestry (Reimoser & Reimoser 2010); or (3) there are ‘too many for their own good’ (Caughley 1981), which can result in, for example, periurban populations, particularly with wild boar Sus scrofa (Jansen et al. 2007 in Germany; López et al. 2010 in Spain; but see Licoppe et al. 2013 for detailed data).

Periurban wild mammal populations (stimulated by anthropogenic food availability; Duarte et al. 2015) may increase human–wildlife conflicts (e.g. in residential areas in the Costa do Sol, Spain, there are conflicts involving red deer Cervus elaphus; Duarte et al. 2015) and are drivers of numerous problems, including (but not limited to) the spread of infectious diseases (Vicente et al. 2007, Gortázar et al. 2012). Wild ungulates play a tremendous role as reservoirs and dispersers of infectious diseases (Ruiz-Fons et al. 2008, Boadella et al. 2012b). Likewise, there has been a general increase in wild ungulate-vehicle collisions (UVCs) in Europe especially since the beginning of the 21st Century (Mysterud 2004, Marques et al. 2010). Although the problems that have been identified constitute major threats, management of ungulates for hunting purposes can sometimes increase the carrying capacity of the ecosystem, hiding the damage or just postponing it. These problems are predicted to increase in the coming years, and there is a need for standardised ecological indicators on a European scale, that can capture ungulate population trends (Morellet et al. 2007). Density and abundance can be useful tools in the characterisation of overabundance scenarios, when coupled with other ecological indicators (Morellet et al. 2007). In an attempt to standardise methods and gather data to develop European wildlife disease surveillance programmes, the APHAEA (harmonised Approaches in monitoring wildlife Population Health, And Ecology and Abundance – https://www.aphaea.eu/) project was developed, as well as ENETWILD consortium et al. (2018), which gave researchers valuable guidelines for the implementation of the recommended methods to asses wild boar density. Similarly, with the purpose of easily collecting wild boar and red deer data at a European scale, the EuroBoar (http://euroboar.org/) and EuroDeer (http://eurodeer.org/) projects were developed (for more details about density values and sampling methodology see Appendices S1 and S2).

With an holistic perspective, it is possible to see how human populations can shape populations of wild mammals and that humans can have a role in ecosystem evolution; therefore, understanding human attitudes and motivation is essential for integrated management (Røskaft et al. 2007, Delibes-Mateos 2015). Alongside their role as the main drivers of the expansion and increase in ungulate populations, humans are also victims of this increase, so it is essential to assess and understand human perception of the consequences of ungulates and their benefits. Human–ungulate interactions constitutes a major gap in ungulate studies in Europe, and it can be considered a proxy for density and for the damage caused by wild ungulate populations.

This review concerns ungulate species in Europe, with special focus on red deer and wild boar, due to their wide distribution in Europe and their increasing densities, along with the possible density-dependence scenario (however, examples relate to other species – namely roe deer Capreolus capreolus – whenever necessary). The consequences of overabundance are used here as examples, namely damage to agriculture and forestry, the prevalence and intensity of infectious diseases and UVCs. These consequences were chosen due to their wide use among European researchers and the availability of data on a large temporal scale, enabling us to perceive potential fluctuations over time. The sociological view of abundance was added, due to the emerging importance of this topic in ecology and the recognition of the consequent gap in wildlife studies. We aim to contribute to the development of what is expected to be a well-studied field in future, and to help decision makers adjust wild ungulate management practices according to people’s expectations and in accordance with animal and ecosystem needs.

Methods

The literature search was conducted using the web browser Google Scholar. Articles were searched for using combinations of the keywords: ‘mammals’, ‘overabundance’, ‘ungulates’, ‘attitudes’, ‘diseases’, ‘infection’, ‘density’, ‘damages’, ‘agriculture’, ‘forestry’, ‘traffic’ ‘ungulate vehicle collisions’ and ‘Europe’. To maximise completeness, we also searched for key authors who have published extensively on ungulates in Europe (including, C. Gortázar, A. Mysterud and J. Langbein), and for references cited in the articles we selected. We generally restricted our search to peer-reviewed papers and book chapters; however, exceptionally we included reports and official websites that provided important information. All query results were verified manually.

Damage to agriculture and forestry

Damage to agriculture and forestry by wild ungulates has been acknowledged in Europe for decades (since the 1940s; Klemm 1948). The increase in ungulate population density (especially red deer and wild boar) in the last decades has greatly exacerbated the problem.

Regarding agricultural damage due to wild ungulates, wild boar can be considered the main problematic species (Remoiser & Putman 2011), probably because of their dietary preferences (Fonseca et al. 2007), but also due to their increasing densities (Bleier et al. 2012, Cutini et al. 2014). Wild boar are considered responsible for 95% of agricultural damage caused by wild ungulates in Croatia (Kusak & Krapinec 2010), for 90% in Italy (Apollonio et al. 2010b), for 87% in France (Maillard et al. 2010), and for 60% in Slovenia (Adamic & Jerina 2010). Damage by wild boar has great economic impacts, costing about 80 million Euros (ME) each year in Europe (Reimoser & Putman 2011). The density-dependence of damage to agriculture and forestry is still a debatable topic; some authors question the real impact of ungulate population size on the amount of damage caused (e.g. Remoiser & Putman 2010), while others argue that high density contributes towards increased damage (e.g. Bleier et al. 2012, Cutini et al. 2014). In countries where payments are made to compensate for damage, these payments have increased along with increases in ungulate density. For example in Hungary, the total compensation payment for crop damage in 1997 was 2.8 ME; in 2008, it was 6 ME (Bleier et al. 2012). In Slovakia, compensation payments totalled 130000 Euros in 2001-2003 and 320000 Euros in 2005 (Findo & Skuban 2010). In forestry, damage mainly caused through browsing and bark stripping (also density-dependent; Verheyden et al. 2006) is of great concern (Putman et al. 2011) and is a source of conflict, mostly due to economic losses (Linnell et al. 2020). Damage to forestry can also have concerning costs, such as 585000 Euros per year in Hungary (Csányi & Lehoczki 2010), 1.5 ME per year in the Czech Republic (Bartoš et al. 2010), 1.49 ME per year in Slovakia (Findo & Skuban 2010), and 3.2 ME per year in Finland (Ruusila & Kojola 2010).

Although damage to agriculture and forestry can have ecological, economic and social impacts, monitoring systems at a national scale are very scarce (Reimoser & Putman 2011), and direct compensations are not always available. Compensation payments to people affected by damage are made in four countries for agriculture and forestry, in two countries only for agriculture, and in another two only for forestry (Cutini et al. 2014), but such payments can be effective to minimise human–wildlife (or human–human) conflicts. Thus, it is important to perform compensatory payments coupled with other management approaches, such as hunting (although this requires further study; Novosel et al. 2012) or electric fencing (this is proven to be efficient; Geisser & Reyer 2004).

Besides damaging forestry, overabundant ungulate populations can have an impact on other species, such as small mammal species (Dolman et al. 2010), invertebrates (Gobbi et al. 2018), and birds (Putman et al. 2011, Newson et al. 2012), mainly through their effects on the woodland. This reinforces the need for integrative management of the ecosystem as a whole.

Ungulate-vehicle collisions

Ungulates are the main species involved in wildlife-related road traffic accidents in Europe (Groot-Bruinderink & Hazebroek 1996, Langbein et al. 2011). The increase in ungulates throughout Europe, and the increased road network (with consequent fragmentation of habitats; Lagos et al. 2012), has enhanced the rise of UVCs (Seiler 2005, Langbein 2006, Langbein et al. 2011, Saint-Andrieux et al. 2020). For example in Spain, there have been increases of 3600% and 300% in UVCs with wild boar and red deer respectively from 2006 to 2012 (Sáenz-de-Santa-Maria & Tellería 2015), and in Sweden, there has been an increase of almost 600% in UVCs with wild boar from 2003 to 2012 (Savberger 2010 in Thurfjell et al. 2015). UVCs also include train collisions, which are affected by ungulate density (Jasińska et al. 2019) and are increasing in some countries (e.g. Poland, Czech Republic, Hungary), although few studies have been performed regarding this subject (but see Cserkész & Farkas 2013, Kušta et al. 2014, Krauze-Gryz et al. 2017). UVCs are now considered the main cause of ungulate mortality, apart from hunting (Langbein 2006, Apollonio et al. 2014, Hothorn et al. 2015).

Analysis of the incidence of UVCs is a complex topic which comprises several factors, including the amount of traffic, speed of the vehicles, vegetation near the roadside, weather conditions, and possible mitigation measures (e.g. under- or overpasses, road fencing, road signs; Langbein et al. 2011, Seiler et al. 2016). UVCs represent indeed a risk for human safety: in the UK, around 12 human fatalities occur each year in consequence of UVCs, and 100 accidents cause serious injuries and 450 slight injuries to humans (Langbein & Putman 2005). On average, there were an estimated 30000 human injuries as a result of UVCs in Europe in 1996 (Groot Bruinderink & Hazebroek 1996); this number has now potentially risen. These collisions almost always lead to the death of the ungulate (Almkvist et al. 1980 in Seiler 2005, Niemi et al. 2015). UVCs also have economic consequences (via, e.g. damage to vehicles, hunting profit diminution, medical treatments of injured persons, deficit of working capability of employees; Pokorny 2006, Langbein et al. 2011), causing a monetary loss of 15 ME per year in Slovenia (Pokorny 2006), 100 ME per year in Sweden and France, and 447 ME per year in Germany (Apollonio et al. 2010a).

Several projects have been developed in attempts to collect and analyse UVC information in Europe and in individual countries (e.g. for Europe ‘Ungulate-vehicle collisions in Europe: identify spatial risk factors and predict high accident likelihood’ – https://www.researchgate.net/project/UNGULATE-VEHICLE-COLLISIONS-IN-EUROPE-identify-spatial-risk-factors-and-predict-high-accident-likelihood, and for the Czech Republic: www.srazenazver.cz/)). Understanding both temporal and spatial patterns of UVCs is essential (Rolandsen et al. 2011, Jakubas et al. 2018), and would be easier with an available temporal and spatial frame. Such information will help decision-makers decide on mitigation measures adapted to each scenario, with a greater potential of success (Langbein 2006). UVCs are not spatially or temporally random (Haikonen & Summala 2001, Lagos et al. 2012, Seiler et al. 2016); there are peaks in UVCs at sunrise and sunset, when ungulate species are most active (e.g. Haikonen & Summala 2001 in Finland; Pokorny 2006 in Slovenia; Lagos et al. 2012 and Rodríguez-Morales et al. 2013 in Spain; Morelle et al. 2013 in Belgium; Steiner et al. 2014 in Europe; Hothorn et al. 2015 in Germany; Thurfjell et al. 2015 in Sweden; Seiler et al. 2016 in Sweden, Norway and Catalonia; Kušta et al. 2017 in the Czech Republic). Most UVCs occur during the rut (Groot Bruinderink & Hazebroek 1996, Lagos et al. 2012, Rodríguez-Morales et al. 2013, Thurfjell et al. 2015). In red and roe deer, this is reinforced with the existence of newborn fawns, and fawns that become independent and wander erratically (Langbein & Putman 2005, Hothorn et al. 2015).

Evaluation and implementation of mitigation measures that can successfully reduce the emergent problem of UVCs is urgent. With this purpose, some investigations have been carried in recent years (Steiner et al. 2014, Niemi et al. 2015). The main conclusions suggest that fencing roads and providing underpasses and overpasses for ungulates is the most successful strategy (Seiler 2004). Many other approaches have been taken, including the use of whistles or scenting devices in cars to repel mammals (not proven to be effective; Lutz 1994 in Groot Bruinderink & Hazebroek 1996), and the reinforcement of road signs, lighting of roads, presence of road reflectors, roadside vegetation clearance, and educational programmes for drivers (limited success, still undergoing research; Seiler 2004, Steiner et al. 2014).

It is predicted that numbers of UVCs will continue to increase in Europe as ungulate populations are also expanding (but see Appendix S3). Further research is needed to develop mitigation measures that can reduce numbers of UVCs, preventing at the same time human fatalities and economic losses.

Infectious diseases in wild ungulates

Emerging diseases are a major concern for wildlife biologists and veterinarians, and their incidence and prevalence have been increasing along with increasing ungulate numbers (Côté et al. 2004, Gortázar et al. 2008). In fact, these are not independent events: wild ungulate population density has a determining role in the spread and maintenance of most of the diseases that affect ungulates (including domestic ones; Ruiz-Fons et al. 2014).

Among others, animal tuberculosis (Gortázar et al. 2006, Gortázar et al. 2008), Aujeszky’s disease (Boadella et al. 2012b, Pannwitz et al. 2012), and classical swine fever (Artois et al. 2002, Ruiz-Fons et al. 2008) are the most concerning diseases affecting wild ungulates.

Tuberculosis is a chronic infectious disease caused by the Mycobacterium tuberculosis complex (Vicente et al. 2007, Boadella et al. 2012a) which affects both wild and domestic ungulates and is known to be favoured by increasing wild population density, and also by animal aggregation. Aggregation can be linked to habitat fragmentation (Gortázar et al. 2006), but is mainly related to management strategies (Gortázar et al. 2006, Acevedo et al. 2007) such as supplementary feeding and watering for intensive hunting (Gortázar et al. 2012, Boadella et al. 2012a). Tuberculosis in wild ungulates is spread throughout Europe – with wild boar and red deer as reservoirs – but its peak occurs in Mediterranean countries, especially in the Iberian Peninsula (prevalence is 52% in wild boar and 27% in red deer in Spain – Gortázar et al. 2008; 6–46% in wild boar and 6–38% in red deer in Portugal – Santos et al. 2009, Matos et al. 2016, Madeira et al. 2017; 29–54% in wild boar and 13–33% in red deer in France – Hars et al. 2010, Zanella et al. 2008; and 6.8% in wild boar in Italy – EFSA 2009). Tuberculosis constitutes a source of human–wildlife conflict due to the economic losses caused by possible transmission to livestock and the risk to human health (Gortázar et al. 2008, Santos et al. 2009).

Aujeszky’s disease, also known as pseudorabies (Lari et al 2006, Boadella et al. 2012b), also reaches concerning prevalence values in Europe with the increase in wild boar density and geographic range size (Laddomada 2000, EFSA 2018). Aujeszky’s disease affects mostly suids, but can also affect other mammals (ungulates, carnivores, lagomorphs and rodents) which represent dead-end hosts and die from the infection (Mettenleiter 2000, Meier et al. 2015), and has an economic impact in the livestock industry (Ruiz-Fons et al. 2008, Boadella et al. 2012b). This infectious disease is widespread throughout Europe (prevalence among wild boar is 11–61% in Spain – Vicente et al. 2002, Ruiz-Fons et al. 2006; 26–31% in Slovenia – Vengust et al. 2005, Vengust et al. 2006; 55% in Romania – Vută et al. 2009; 55% in Croatia – Gagrcin et al. 1989, Zupancić et al. 2002; 39% in Russia – Shcherbakov et al. 2007; 16% in Germany – Pannwitz et al. 2012; 30–51% in Italy – Lari et al. 2006, Montagnaro et al. 2010; and 1.6% in Switzerland – Meier et al. 2015; the lowest Aujeszky’s disease prevalence in Europe). Surveillance and control of Aujeszky’s disease is required for animal safety, and essential to avoid financial and social costs associated with disease outbreaks.

Classical swine fever affects both domestic and wild ungulates (pigs Sus scrofa domesticus and wild boar) and is caused by a highly contagious porcine pestivirus (Ruiz-Fons et al. 2008, Postel et al. 2018). Although classical swine fever is now absent in most of Western Europe, it is still present in a limited number of areas in several Central and Eastern European countries (Artois et al. 2002, Ruiz-Fons et al. 2008). In past decades, classical swine fever outbreaks have occurred, for example in the Netherlands, Germany, Spain, Italy and Belgium in 1997 (Edwards et al. 2000, Fritzemeier et al., 2000, Laddomada 2000, Fernández-Carrion et al. 2016), in Spain in 2001 (Fernández-Carrion et al. 2016), in Latvia between 2012 and 2015, in France in 2003, and in Germany in 2006 (Postel et al. 2018). These outbreaks caused major economic losses in several countries (e.g. in Spain, 60 ME in 1997 and 48 ME in 2001; Fernández-Carrion et al. 2016, and in the Netherlands, about 1800 ME in 1997/1998; Postel et al. 2018), and calls for urgent action to eradicate classical swine fever have been made, alongside the implementation of effective control and surveillance strategies (Moennig & Becher 2015). Although there is some debate (see Aubert et al. 1994), the wild boar is believed to be a classical swine fever reservoir (Ruiz-Fons et al. 2008); thus, management of wild boar has a major role in disease control and spread, and measures can potentially hasten outbreak episodes. In fact, practices such as artificial feeding and watering increase the ecosystem’s carrying capacity, and promote high densities and animal aggregations, known to increase the incidence and prevalence of classical swine fever (Laddomada 2000, Ruiz-Fons et al. 2008).

Other infectious diseases affecting wild ungulates include African swine fever (EFSA 2018) detected in Poland, Estonia, Czech Republic, Belarus and Lithuania, with a recent outbreak in Belgium in 2018; porcine circovirus type 2 (Vicente et al. 2004); porcine parvovirus (Vengust et al. 2006); and various tick-borne diseases (Ruiz-Fons et al. 2006, Gortázar et al. 2015) including bluetongue disease (Ruiz-Fons et al. 2014). Hence, in order to maintain healthy wild and domestic ungulate populations and prevent economic effects due to infected livestock, the need for surveillance and control schemes adapted to each case or scenario should be assessed (Laddomada 2000, Ruiz-Fons et al. 2008). Various mitigation measures are required, depending on factors such as disease epidemiology, host condition, host density (high densities are related to high risk of disease spread) and management practices (Edwards et al. 2000, Laddomada 2000, Gortázar et al. 2015). Numerous techniques have been employed to reduce or eradicate infectious diseases. Culling and vaccination are valid alternatives, but sanitary programmes and biosafety are considered the most suitable measures (Artois et al. 2002, Boadella et al. 2012a). In fact, the success of management strategies is largely related to the economic investment that the countries or governments undertake, also considering the economic impact that taking no action can have (Gortázar et al. 2011, Linnel et al. 2020). It is important to alert human populations and stakeholders (meaning people or groups who will be affected by or will affect wildlife or wildlife management; Decker et al. 1996) to the need for preventive management strategies for infectious diseases. Positive advances have undoubtedly been made in communication, especially in Western Europe where wild ungulate densities are increasing (Artois et al. 2002, Gortázar et al. 2006, Ruiz-Fons et al. 2008). Indeed, it is necessary to go even further: long-term funding is essential for projects on infectious diseases, not only in domestic animals, but also in wildlife, due to their role as reservoirs (Ruiz-Fons et al. 2008, Gortázar et al. 2015).

A sociological view of overabundance

The ungulate overabundance scenarios documented throughout Europe and their consequences show the need to establish multidisciplinary teams in order to understand human dimensions of ungulate overabundance (Riley et al. 2002, Riley et al. 2003, Madden 2004, Brown & Harris 2005, Marshall et al. 2007, Gerner et al. 2012). As Røskaft et al. (2007) said, humans are part of nature; therefore, their behaviour, or at least their attitudes (Delibes-Mateos 2014), can and should be regularly investigated and analysed. In this context, several studies have been performed to give wildlife managers tools to minimise the controversial (and costly) problems, to agree management methods and to establish a threshold (a maximum acceptable animal density) for each stakeholder group (West & Parkhurst 2002, Treves et al. 2006, Liordos et al. 2017, Storie & Bell 2017). In fact, it is essential to involve people in management decisions, and the human dimension of the problem is particularly important when dealing with species, such as ungulates, that can be part of a human-dominated landscape (Behr et al. 2017). Furthermore, without knowledge of what the stakeholders think, wildlife managers might not be aware of people’s support for or opposition to certain mitigation measures that are needed (Brown & Harris 2005); this lack of knowledge may threaten the success of any management programme.

In the USA, socio-ecological studies in relation to ungulate populations started long ago (e.g. Connelly et al. 1987, Decker & Gavin 1987, Decker & Purdy 1988, Morgan et al. 1992, Cornicelli et al. 1993) and revealed that increases in ungulate populations have early consequences, stimulating conflicts between human populations (rather than human–wildlife conflicts as described by Delibes-Mateos 2017). Madden (2004) argues that human–wildlife conflicts occur when the needs and behaviour of wildlife impact negatively on the goals of humans, or when the goals of humans negatively impact the needs of wildlife. In fact, human–wildlife conflicts sometimes manifest themselves as human–human conflicts, when hostilities often arise between people who have different goals, attitudes, values or levels of empowerment (Madden 2004, Marshall et al. 2007, Gerner et al. 2012). Ezebilo et al. (2012) provides an example of a human–human conflict in Sweden, where game hunters and forest owners have different views about wildlife management and conservation. Delibes-Mateos (2015) stated that, for example: ‘…conservationists may aim to reduce the abundance of a particular species, because it detrimentally affects other species/habitats, but this reduction clashes with the interests of other stakeholders (e.g. hunters), who promote an increase in the species’ numbers’. Conservation conflicts arise mostly from the impacts caused by wildlife (Young et al. 2005, 2007) and from different perspectives of the damage that wildlife can cause and of how to solve or manage these conservation issues (Delibes-Mateos 2015).

In Europe, the human dimension of wildlife is a newer topic, and most studies are recent (e.g. Keuling et al. 2016, Liordos et al. 2017, Storie & Bell 2017, Prager et al. 2018). The lack of studies regarding ungulates (but see Echegaray 2004) may be related to a greater interest in conflicts involving predators, such as the wolf and lynx Lynx lynx, on which several studies have been focused (e.g. Naughton-Treves et al. 2003, Heberlein & Ericsson 2005, Behr et al. 2017). There is clearly a gap in our information regarding the human dimension of overabundance of ungulates that urgently needs to be filled. Integrated actions can increase the success rate of measures to mitigate and prevent conflict, so it is essential to continue multidisciplinary studies in Europe. For example, in the Iberian Peninsula where red deer and wild boar densities are achieving concerning values (see Appendix S1), it is especially urgent to include humans' perceptions and attitudes in management plans. Multidisciplinary research is, in fact, an emerging topic; we believe that the human dimension should be included as an important ecological indicator in research studies, alongside ecological factors relating to the non-human species. Only with the cooperation of the people living close to wild ungulate populations is it possible to conduct (or attempt) successful management strategies.

Wild ungulate management - future steps and recommendations

Management of wild ungulates depends on various factors, beginning with the main reasons for management, alongside the preventive actions that are required. The problems reported in this review require quick answers and close monitoring. Several techniques have been employed to manage troubling situations. The ecosystem must be assessed as a whole, as well as the solutions for the problems that arise as ungulate populations increase in abundance and range. Some factors need to be considered as contributing towards ungulate overabundance, such as the decline in numbers of hunters and the lack of new hunters. This can also constitute an obstacle when trying to adopt management measures involving removal of ungulates from the population at certain percentages (e.g. EFSA 2018).

Increasing populations of ungulates have created several problems that should not be addressed as separate and independent events, as they are not. While mitigation measures can be beneficial, they may also have deleterious effects. For example, road fencing (even with underpasses and overpasses) can prevent animals from crossing roads (Seiler 2004, Rolandsen et al. 2011), but may also promote animal aggregation that is known to be related to the transmission of infectious diseases (Ruiz-Fons et al. 2008, Gortázar et al. 2015). The same problems occur with fencing of some areas to avoid silvicultural or agricultural damage (Côté et al. 2004, Geisser & Reyer 2004). Moreover, fencing may stimulate conflicts between people with different interests or goals (e.g. see Baltzinger et al. 2018). The management of infectious disease depends mostly on the epidemiology of the disease; the most commonly adopted measures are sanitary programmes and biosafety (Gortázar et al. 2015), culling (random or specific for some gender or age groups), vaccination, and combinations of methods (Boadella et al. 2012a, Postel et al. 2018). There are a few examples of successful approaches using vaccination of wild boar (Ballesteros et al. 2011), and some advances are expected in the near future (Beltrán-Beck et al. 2012). Vaccination seems to work in red deer for some diseases such as bluetongue disease and bovine tuberculosis, if a certain percentage of animals in the population can be immunised (Falconi et al. 2011); this involves financial costs that countries are not always inclined to spend.

The use of these methods comes with drawbacks that need to be weighed up. Culling is the most widely accepted method to reduce population size (Sorensen et al. 2014), but it can provoke the dispersal of individuals, potentially promoting rapid spread of infectious diseases (Ruiz-Fons et al. 2008, Gortázar et al. 2015), so culling should not be considered in isolation (as documented in Poland; Vicente et al. 2019). To overcome this potential drawback, culling should take place simultaneously with vaccination in some cases (Gortázar et al. 2015). The compensatory reproductive strategy adopted by wild boar (Artois et al. 2002, Fonseca et al. 2011) may limit management success in this species. Supplementary feeding is sometimes used to increase the ecosystem’s carrying capacity or to minimise problematic crop damage (Mysterud 2010, Sorensen et al. 2014), but feeding may promote aggregation and increase the incidence of infectious diseases (Sorensen et al. 2014, Gortázar et al. 2015). Some animals become habituated to supplementary feed (Putman & Staines 2004), which constitutes a problem when it is no longer available and the animals’ body condition tends to decline. Some authors provide evidence for the effectiveness of supplementary feeding to control crop damage (Mysterud 2010), but not all agree; Milner et al. (2014) stated that supplementary feeding often results in unintended effects and can lead to additional risks and indirect costs, though they do not reject the contribution of this practice to reducing UVCs. This subject requires further study, and the creation of a European database of road traffic accidents involving wildlife would be helpful. This would allow researchers to establish more accurate patterns that could be used as a baseline for mitigation measures adapted to each country and to each scenario.

Alternatively, some novel management measures should be considered, such as the creation of a supply chain for the meat produced by hunting (Gaviglio et al. 2017), as already exists in some parts of Europe (e.g. France, Austria). This could have positive economic, social, and environmental impacts (Gaviglio et al. 2017) and is a fair way to take advantage of potential ungulate overabundance throughout Europe.

Convergence is found in the basic principle that each case of ungulate overabundance should be studied individually, in order to take the appropriate measures by analysing cost-performance balance. Financial support should be claimed from governments to carry out successful management strategies that end up being investments rather than expenses. The need to assess several ecological indicators (e.g. Morellet et al. 2007), as well as having suitable and robust population size estimation methods, is crucial in a holistic and adaptive management approach (Kenward & Putman 2011, Santos et al. 2018). With this purpose, multidisciplinary teams should be built, with biologists, veterinarians, stakeholders, sociologists, and others who are interested in wildlife management (e.g. naturalists or naturalist associations). These teams should consider past lessons and important information already available, bearing in mind the knowledge that each situation is unique and should be addressed as such. Ungulate populations can bring advantages when the principles of adaptive management are met (Kenward & Putman 2011, Apollonio et al. 2017), and benefits could be attained through ecotourism, via hunting, and from an ecological point of view. However, researchers should be aware of the drawbacks of increases in ungulate populations and should have all the necessary information to deal with this potentially alarming scenario throughout Europe.