Impacts of roads on wildlife
Barrier effects
Roads can significantly impact the permeability of landscapes, reducing the ability of some species to migrate or disperse between habitat patches (Vos et al. 2001). This can lead to habitat fragmentation, reduced gene flow between occupied habitat patches, increased genetic drift and reduced heterozygosity, thus threatening the persistence of populations (Puky 2006). Indeed, road construction is thought to be one of the main causes of local extinctions of hazel dormice (Bright et al. 1994). Dormouse populations need at least 20ha of continuous habitat in order to persist, but may also persist if smaller, high-quality patches are connected by tree lined corridors or hedgerows (Bright et al. 1994). For this reason, dormouse presence is often used as an indicator of landscape-scale habitat integrity, and was integral to the inclusion of the first habitat bridge in the UK, over the Lamberhurst Bypass in south-east England. This and other studies have concluded that well-planned and executed road mitigation measures for dormice can successfully conserve their populations (Cresswell and Wray 2006), although species presence does not equate to long-term population survival.
For bats, movement ability is species-specific, being a function of behaviour and flight style. Altringham and Berthinussen (2012) found that bat abundance (principally Pipistrellus species) declined with distance from a major road, which is consistent with roads acting as a barrier to bats, but also with roads being attractive to bats. In contrast, Waters et al. (1999) found that Leisler’s bat (Nyctalus leisleri) crossed major roads regularly. Kerth and Melber (2009) compared the effects of a busy German motorway on foraging and movement patterns of barbastelle bats (Barbastella barbastellus), which forage in open space, and Bechstein’s bats (Myotis bechsteinii), which glean prey from vegetation. In total, seven species of bat were found to fly through underpasses (tunnels) to cross the motorway, supporting the importance of these mitigation features for bats. However, while a greater proportion (five of six monitored) of barbastelle bats crossed the motorway either by flying over it or through underpasses, all Bechstein’s bats that crossed the motorway (three of 34) flew through the underpass. Thus, motorways may present less of a barrier to barbastelle bats than Bechstein’s. However, no account was taken of differences in mortality between species crossing the motorway.
The consequences of road-induced habitat fragmentation for strictly terrestrial species, including amphibians may be greater than for bats. Many amphibian species persist within meta-populations and fragmentation may reduce migration rates between habitat patches, thus reducing meta-population viability. If populations become isolated from each other so that migration between them is no longer possible, they become smaller and more vulnerable to extinction (Sjögren 1991). The increased mortality of amphibian populations that reside near roads further contributes to the likelihood of local extinctions and reduces the possibility of recolonization (Fahrig et al. 1995; Vos and Chardon 1998).
Amphibians may be particularly vulnerable to direct mortality and barrier effects of roads because their life histories often involve migration between wetland and terrestrial habitats, regular movements within a fixed home-range (DeMaynadier and Hunter 2000) and individuals tend to be inconspicuous and slow moving (Ashley and Robinson 1996). Vos et al. (2001) found that genetic differentiation among moor frogs (Rana arvalis) was positively correlated with geographic distance between frogs (range 0.5km to 7.6km), and further explanatory power was yielded with the addition of data on habitat resistance (permeability), including factors describing road type, railway lines, traffic volume, and unsuitable habitat-types. Among these, road type and railways were most strongly associated with genetic differentiation, implying that these features acted as barriers to frog movement.
Amphibian behavioural response to traffic is another factor contributing to their vulnerability on roads. Many species of frog, toad and salamander react by staying immobile in response to the light and sound associated with traffic (Mazerolle et al. 2005). It has been suggested that this might be due to amphibian pupils taking some time to recover after exposure to bright light (Cornelle and Hailman 1984; Mazerolle et al. 2005). Alternative explanations are that animals become physiologically arrested by debilitating compounds associated with de-icing salts, or they linger on the relatively warm road surface for the purposes of thermoregulation (Gibbs and Shriver 2005).
Mortality impacts
The rapid increase over the past century in roads and traffic volume is implicated in global amphibian population declines (Puky 2006) and the negative impact of road mortality on populations has been demonstrated for several species (Hels and Buchwals 2001; Cooke and Sparks 2004; Mazerolle 2004; Pellet et al. 2004; Gibbs and Shriver 2005). Fahrig et al. (1995) identified a higher proportion of dead frogs and toads on roads with high traffic volumes and lower abundance of amphibians on land surrounding them. However, large, high-traffic volume roads do not necessarily cause the greatest impacts for all taxa. Indeed, Clevenger et al. (2003) found more intense mortality hotspots for birds and terrestrial mammals on minor roads that passed through park lands in comparison with a major motorway. Raised roads, little roadside vegetation and proximity to road crossing structures were also associated with lower roadkill frequencies.
Puky (2006) reviewed a number of factors associated with amphibian road mortality, and concluded that amphibian-specific factors (such as movement characteristics and behaviours), local habitat structure, stochastic weather effects, and road characteristics (such as traffic intensity and vehicle speed), all influenced mortality risks.
A variety of factors have also been associated with mortality of bats on roads. Lesiński (2007) identified bat age (young of the year were killed more frequently than adults), road proximity to bat fly-ways (especially near forest edges), and adjacent habitat type (proximity to forests and tree stands) were associated with higher rates of bat mortality on roads in Poland. Impacts may also be temporally variable. Time of day is likely to be important for species with diurnally variable activity patterns, and season may also influence mortality rates (Lesiński 2007).
Secondary effects
As well as direct mortality and barrier effects directly caused by roads, there are a number of additional impacts that roads can have on wildlife due to road-associated activities (English Nature 1996). For example, a primary de-icing agent used on roads is sodium chloride (common salt), which is highly corrosive and can be toxic to many aquatic organisms including GCN (Duff et al. 2011). Calcium magnesium acetate (CMA) is an alternative de-icer that is less corrosive, less mobile in soil, biodegradable and less toxic to aquatic organisms (National Research Council 1991; Amrhein 1992; Ostendorf 1993). The reason it is not used as standard is cost, as it can be many times more expensive than common salt.
Road traffic can also create chemical pollution, and noise pollution as well (Coffin 2007). Traffic noise can affect basic behaviours of some taxa (great tits, Parus major, sang at higher pitches in noisier environments in the Netherlands, Slabbekoorn and Peet 2003), potentially enhancing mortality impacts due to inappropriate behaviours, such as remaining stationary in response to approaching loud noises (Mazerolle et al. 2005).
Road construction can significantly alter the hydrodynamics of associated water bodies (Coffin 2007). Their path may cross that of a watercourse that feeds a breeding pond for amphibians, thus potentially reducing inflows. Alternatively, road run-off may increase inflows, but not necessarily with water of sufficient quality to sustain amphibian populations. The integrity of a breeding pond’s watershed can be partially maintained by allowing natural hillside drainage to cross under new roads, and keeping this separate from road run-off using under-drains. Road run-off should be directed for discharge downstream of breeding ponds (Merrow 2007).
The presence of lights along roads can repel or attract some animals to roads. Some bats are attracted by aggregations of insects around street lights (Rydell and Racey 1995), potentially increasing mortality by enhancing the risk of collisions with road traffic. Conversely, others, such as the lesser horseshoe bat (Rhinolophus hipposideros) avoid street lighting, altering flight paths (Stone et al. 2009) and potentially exacerbating the barrier effects of roads. Effects of street lights on dormice were not found during this review, but lights may affect the behaviour of amphibians. For example Alpine newts (Mesotriton alpestris) mate throughout the diurnal period, but there may be adaptive benefits to investing in mating displays during the hours of darkness, not least to reduce the risk of predation by visual predators (Denoël and Doellen 2010). The presence of lights along roads that are associated with populations of newts may therefore influence mating patterns, genetic structure and predation mortality, although this hypothesis has yet to be demonstrated empirically to our knowledge.
Sustainable Urban Drainage Schemes (SUDS)
English Nature (2001) reviewed research demonstrating that some drainage systems can result in high GCN mortality, and that upright kerbs and gully pots have the same mode of action as a drift fence and pitfall system, but without the possibility of newt retrieval. These effects can be avoidable with careful design which should be employed when GCN are anticipated to come into contact with roads.
SUDS are becoming increasingly popular for a variety of environmental reasons. They aim to control and treat drainage at its source by using a range of features such as porous surfaces, grassy ditches (swales), buffer strips and filter beds. The design of SUDS can be adapted to enhance or create amphibian habitat while avoiding trap-like drains. English Nature (2001) recommended that SUDS should be implemented wherever road-related GCN mortality is indicated, but lamented the fact that there was an apparent reluctance among some bodies to adopt road construction practices involving SUDS. If the use of gully pots is genuinely unavoidable it may be necessary to exclude GCNs from the area using permanent fencing. English Nature (2001) offered a variety of options for the installation of exclusion fencing and stressed that fencing should be situation-specific. However, it may often not be desirable to prevent GCN from crossing roads as this may interfere with migration or dispersal routes thus disrupting meta-population dynamics. Reduction of the likelihood of GCN becoming trapped in drains by making them more newt-friendly may offer a solution. Possible options to achieve this are to use sloping kerbs near to drainage holes, kerbs with special design features to allow amphibian passage (e.g. or to eliminate kerbs altogether.
Amphibian ladders and netting or mesh over drain grates have been offered as mitigation options, but these were not recommended by English Nature (2001) as they may be ineffective, maintenance heavy and are not supported by some local authorities because they can interfere with the drain efficiency and cause problems during cleaning. English Nature (2001) advocated preventing GCN falling into drains in the first place rather than focussing on ways of allowing them to escape.
Road effect zone
It is important to consider adjoining habitat deterioration as well as habitat loss when investigating ecological fragmentation and barrier effects caused by roads. A more complete appreciation of the ecological impacts of road construction is evident when the consequences of habitat alteration in terrestrial and aquatic ecosystems adjacent to roads are considered simultaneously. For example, road effects in the UK have typically been assumed to extend 5m on either side of the road (Underhill and Angold 2000), which may preclude consideration of potential impacts on many ponds. However, amphibians may disperse much greater distances than 5m; Dodd and Cade (1998) recorded striped newts (Notophthalmus perstriatus) dispersing up to 709m from the nearest pond and in non-random directions. Impacts of roads on aquatic environments may occur indirectly as a result of direct impacts on terrestrial environments. Moreover, management of the environment adjacent to the road, such as verges and adjoining land can also influence the road effect zone. For example, upgrade work to a road in Pennsylvania USA discovered to be associated with high bat (Myotis lucifugus and M. sodalis) mortality, included plans to manage vegetation on the road verges. Telemetry studies on the bats revealed that they were less likely to cross the road at locations where the tree canopy was absent, but crossed closer to traffic at locations of reduced canopy height (Russell et al. 2009). Clearly, changes in roadside vegetation management can stimulate complex responses in both the barrier effect and mortality risks posed by roads.
The area over which significant ecological effects occur extends outwards from the road itself (Smith 1999). The road surface presents an environment that is far from conducive to the persistence of individuals, and can impact movements within and between populations, as discussed above. Roadside verges also constitute a heavily impacted environment due to the immediate effects of their proximity to the road, such as sediment and pollution runoff (Gjessing et al. 1984), but also due to mechanical disturbance such as routine vegetation management. However, verges also represent valuable semi-natural habitats for a variety of species, particularly in highly modified landscapes such as those of Great Britain, where road verges constitute approximately 1% of the land area (Underhill and Angold 2000). Determining this ‘road effect zone’ is central to evaluating the overall area affected by roads and their associated infrastructure (Forman and Alexander 1998; Forman and Deblinger 2000). Forman and Deblinger (2000) focussed on factors that were likely to extend for the greatest distance outwards from the road surface and immediate surroundings. They found that the impacts of roads on amphibian migrations can extend for several hundred metres on both sides of the road surface if it is built within that distance of a breeding pond. Furthermore, the effect of road salt (Forman and Deblinger 2000) or fine sediments, nutrients and contaminants (Gjessing et al. 1984) on aquatic systems may be noticeable for up to 1500m from the road. Road construction may also affect runoff/recharge ratios, water temperature, and other factors affecting water quantity and quality, which may also affect amphibian population processes.
It has previously been assumed that highly vagile species are more resistant to habitat fragmentation due to their ability to migrate between more dispersed habitats (Carr and Fahrig 2001). However, vagility also affects the size of the road effects zone. More vagile species may be susceptible to enhanced mortality on roads due to the greater frequency of road crossings (Carr and Fahrig 2001) and this effect can be so strong that land between habitats can become a demographic drain for many amphibians, potentially leading to local extinction (Gibbs 1998).
Current status of selected European Protected Species
GCN are widely distributed throughout western Europe. It is the largest newt species present in Great Britain, but is also the least common. It suffered marked declines in abundance during the 20th century, but is now believed to be slowly declining, with the national population estimated to be approximately 18,300 (JNCC 2010a). It is classed as least concern by the World Conservation Union (IUCN) and the British population is classed as favourable with regards to its range, but inadequate and deteriorating with regards to its population size. Habitat loss (mainly ponds) through changes to land use and introduction of fish to traditional breeding ponds are thought to be the main threats to GCN (JNCC 2010a).
The hazel dormouse is sparsely distributed throughout southern Britain, mostly to the south of England. It is a small mammal, weighing approximately 17g when adult (Bright et al. 2006). The national population is estimated at roughly 45,000 and the species continues to decline (JNCC 2010b). It is classed as least concern by the IUCN and the British population is classed as favourable with regards to its range, but bad and deteriorating with regards to its population size. Loss of ancient semi-natural and coppiced hazel woodland and climate change are considered to be the main threats to dormice (JNCC 2010b).
Among the 17 species of bat that breed in Great Britain, seven (Barbastelle Barbastella barbastellus, Bechstein’s bat Myotis bechsteinii, noctule Nyctalus noctula, soprano pipistrelle Pipistrellus pygmaeus, brown long-eared bat Plecotus auritus, greater horseshoe Rhinolophus ferrumequinum, lesser horseshoe bat R. hipposideros) are listed as priority species on the UK Biodiversity Action Plan and five of these have their own Species Action Plan (barbastelle, Bechstein’s bat, soprano pipistrelle, greater and lesser horseshoe bats). National populations range from approximately 1500 Bechstein’s bats to more than 700,000 soprano pipistrelles. All are either rare or declining, largely due to habitat loss, but populations of all listed species in Britain are considered favourable with regards to their range. However, while lesser horseshoe bats, brown long-eared bats and noctules are considered favourable with regards to their population sizes, the greater horseshoe and Bechstein’s bats are considered inadequate and the population status of barbastelles and soprano pipistrelles is unknown ( accessed 19/09/2014).