Verges provide critical habitat and linkage corridors between remaining patches of native habitat for dispersal of native flora and fauna. These functions help to sustain biodiversity and to mitigate against fragmentation of populations, genetic isolation and ultimately, population collapse and potentially local extinctions (New et al., 2021). However, if occupying a verge limits reproduction or dispersal, or impacts survivorship, and ultimately population persistence, then these habitats could represent ecological traps that act as population sinks. Also, due to compositional shifts, verge habitats could trigger changes in species interactions and co-evolutionary processes (Gardiner et al., 2018; Lazaro-Lobo and Ervin, 2019).

Verge vegetation can be highly diverse and can harbour remnant populations of rare and threatened flora and fauna as well as species excluded from the matrix, however this does not mean to say that such habitats can support viable populations (Roy and de Blois, 2006; Gardiner et al., 2018; New et al., 2021). The seed bank within verges is supportive of revegetation, and vegetated slopes contribute to the control of erosion (Lazaro-Lobo and Ervin, 2019). Verges are frequently prevented from progressing into mature forest by management interventions. These “early successional states” are open habitats supportive of many biomes and high biodiversity. Edge habitat conditions predominate which favour some plant communities over others due to altered biotic and abiotic conditions (Roy and de Blois, 2006; Suarez-Esteban et al., 2016). Vegetation management practices within verges also influence species composition.

Verges are key to maintaining native plant biodiversity in human modified landscapes (Lazaro-Lobo and Ervin, 2019), however, plant diversity is generally lower in verges within a natural landscape matrix (Suarez-Esteban et al., 2016). Verges can be unsupportive to the establishment of forest herbs which can be disturbance intolerant (Gardiner et al., 2018). Contiguous hedgerows generally contain fewer early flowering herbs and have a higher likelihood of, and susceptibility to, frost than native forest environments (Roy and de Blois, 2006). Generalists, ruderal, wind dispersed, high-seed-output, insect pollinated and plants which reproduce by vegetative propagation predominate in these locations (Roy and de Blois, 2006; de la Riva et al., 2011; Ranta et al., 2015).

The verge environment is associated with anthropogenic factors such as noise levels, lighting, and traffic density, which can have a detrimental impact on animal behaviour, resulting in higher energy expenditure and reduced fitness. Similarly, chemical pollutants can cause harm to natural functioning (Trombulak and Frissell, 2000; Shier et al., 2012; Gardiner et al., 2018; Barrientos et al., 2019; Ouedraogo et al., 2020). The introduction of artificial elements within the verge can have a detrimental effect on biodiversity through habitat loss and barrier effects leading to, amongst other things, broken food web linkages (Gardiner et al., 2018). Imbalance in community compositions and environmental factors may be detrimental to prolonged occupation of verges and combined with increased mortality, these areas may become ecological traps (Gardiner et al., 2018).

The high suitability of the verge environment for habitat generalist, open-specialist, early successional, disturbance tolerant or opportunistic animal and plant species (including invasive species) is well documented, as is the limiting effect on many habitat specialist species (Koivula, 2005; Van Rossum, 2009; Schaffers et al., 2012; Knapp et al., 2013; Bochet and Garcia-Fayos, 2015; Ranta et al., 2015; Gardiner et al., 2018; Villemey et al., 2018; Barrientos et al., 2019; Lazaro-Lobo and Ervin, 2019). Verge vegetation composition and condition is a key factor for invertebrate diversity and abundance, influencing movement and sometimes supporting life-processes, even reproduction and overwintering. Rare and functionally important insect species (e.g., pollinators or pest predators) can gain an essential foothold in remnant vegetation in an otherwise hostile landscape matrix (Noordijk et al., 2011; Schaffers et al., 2012; Gardiner et al., 2018; New et al., 2021).

In general, for human-altered environments, invertebrate diversity indicators were higher within verges compared to surrounding environments, even accounting for mortality from vehicles and pollution (Villemey et al., 2018). For some invertebrates, studies confirm that verges can support the full range of life processes (Koivula, 2005; Noordijk et al., 2011; Schaffers et al., 2012). Local factors (e.g., High nectar plants, tall vegetation, percentage litter) and regional factors (e.g., road density) were found to be key drivers for grassland pollinator species compositions (Saarinen et al., 2005; Davis et al., 2007; Phillips et al., 2020). Additionally, the presence of dead wood in the verge is a crucial resource for saproxylic and hole-nesting invertebrates (New et al., 2021).

Vertebrate presence in verges is influenced by the infrastructure type, vegetation cover and the surrounding matrix. Mammal predators such as foxes, badgers and snakes tend to be negatively affected by road verges. Studies indicate that small mammal species may therefore utilise verges as refugia (habitats where prey is relatively safe from predators) due partly to release from predation (Redon et al., 2015; Ouedraogo et al., 2020). A 2015 examination of the factors influencing the presence of bird groups in Natura 2000 sites in Cyprus found that road networks have negative effects on four of the five bird categories (Mammides et al., 2015).

Road noise is thought to have a detrimental effect on the activities of some birds by interrupting acoustic signals, causing stress, and impacting foraging activity (Ortega, 2012; Gardiner et al., 2018). Similarly, a study of the Stephen’s kangaroo rat found that road noise can lead to false responses that may be energetically costly (Shier et al., 2012). Artificial lighting is thought to have an overall negative impact on bat reproduction and commuting behaviour despite local benefits due to insects attracted to lights (Ouedraogo et al., 2020).

It is thought that railway verges in particular promote bird species richness due to the availability of remnants of diverse vegetation and the availability of food, shelter, perching, nesting, and resting (Ouedraogo et al., 2020; Dylewski and Tobolka, 2022). In addition, the presence of culverts and other transportation structures can also offer nesting opportunities (Dean et al., 2018).

Plant abundance and diversity within verges is generally higher compared to adjacent areas when embedded in a human-altered landscape but lower compared to natural areas, although verges in natural areas are more biodiverse than verges in human altered areas (Suarez-Esteban et al., 2016; Lazaro-Lobo and Ervin, 2019). Verges can provide better conditions than adjacent habitats for plant dispersal and post-dispersal processes which contribute to success such as providing light, animal dispersers, high seedbank counts and less seed predation (Suarez-Esteban et al., 2016; Lazaro-Lobo and Ervin, 2019). For example, light favouring plant species can establish within verges, raising overall plant diversity in nearby forest areas (Avon et al., 2010).

In most instances invertebrate abundance is generally higher within the verge, compared with adjacent areas of similar composition. This is especially true amongst pollinators on railway embankments (Villemey et al., 2018). Verges in rural areas can provide shelter for pollinators and invertebrate predators of crop pests as well as resources and refuge habitats that are no longer available in the wider landscape. However they are susceptible to overspray from agricultural chemicals (New et al., 2021). A 2013 study of roadside invertebrates found that habitat generalists and open habitat specialists appeared to benefit from road proximity in both forest and open habitats, suggesting that roads could offer habitat and corridors for the benefit of non-specialist species (Knapp et al., 2013).

Vertebrate biodiversity within verges is dependent on the surrounding matrix. Urban areas tend to be detrimental to amphibian populations, shrubland birds and bats, whereas verges in natural areas have similar biodiversity to similar habitat in adjacent land (Ouedraogo et al., 2020). For fragmentation-intolerant birds such as passerine species (true perching birds, with three toes directed forward and one backward) as well as larger mammals, road verge occupancy is lower than surrounding areas. However, raptor biodiversity indicators were higher, suggesting that they can exploit the verge environment for hunting due to increased food availability (Gardiner et al., 2018; Ouedraogo et al., 2020). Studies suggest that in rural landscapes, roadside vegetation can attract birds for breeding, foraging, or resting, leading to collisions or conditions which are detrimental to survival (Gardiner et al., 2018).

Transportation corridor verges tend to increase movement and dispersal of plants, vertebrates, and invertebrates between habitat patches, but overall effectiveness varies among taxa (Ouedraogo et al., 2020). Verges appear to assist in the dispersal of some moth species, but vegetation can hinder the movement of open-ground specialists such as carabid beetles (Butterweck and Ce, 2000; Do et al., 2017; Villemey et al., 2018; Elo et al., 2021). Many foodplants for invertebrates exist in verge environments, promoting dispersal over a prolonged period (Brunzel et al., 2004). Modelling has confirmed that highway verges played an important role in A. albopictus (mosquito) range expansion in mainland France, consistent with observed findings (Roques and Bonnefon, 2016).

Mammal movement along verges was found to be equivalent or higher to similar habitat elsewhere, and may function as both habitat and movement corridor (Ouedraogo et al., 2020). Meso-predators, including foxes, were found to frequently utilise road corridors within their home range for hunting (Frey and Conover, 2006). Also, hedgehogs were observed to stay close to roads when commuting (Doncaster et al., 2001). Similar can be inferred from genetic modelling for cottontail rabbits and a study on Virginia opossums (Beatty et al., 2014; Amaral et al., 2016). A study carried out in Melbourne found that bandicoots had longer home range areas in narrow strips than wider remnants (Maclagan et al., 2020). Using landscape genetic analysis, wood mice were found not to utilise road environs for dispersal (Wilson et al., 2016).

Studies into bat foraging behaviour around roadside verges found that traffic density was a determining factor, with detrimental effects for frequencies above 10 vehicles every 5 min and barrier effects at frequencies over 200 vehicles every 5 min (Bennett et al., 2013). Another study found that foraging near artificial lighting occurred predominantly around water bodies and well vegetated areas, but not around roads (Voigt et al., 2020). Studies also found that commuting bats favoured dark corridors and flew parallel to the edges of linear landscape features and that feeding along linear features is low, whilst other studies report no negative impact (Kalcounis-Rueppell et al., 2013; Vandevelde et al., 2014; Voigt et al., 2020). Additionally, it has been found that barn owls do not favour linear features when selecting roost sites (Ruiz et al., 2021).

Studies on reptile movements showed that remnant shape influenced ranges, with linear remnants being least preferred (Driscoll, 2004). One study in South Australia found that small linear fragments provided poor connectivity over larger distances (Williams et al., 2012). Alpine newts were observed to utilise road corridors in areas of adjacent low quality (Prunier et al., 2014).

Pollinated and early successional plants were found to benefit from open, linear landscape elements, which promoted long distance dispersal (Van Rossum, 2009). Seed dispersing animals also tended to move along verges promoting seed establishment (Suarez-Esteban et al., 2016). Forest herbs appear to benefit from the shelter and habitat provided by linear hedgerow features as well as facilitating dispersal (Roy and de Blois, 2006). Stepping-stone habitats (see Figure 1 for illustration) can offer sufficient conditions for dispersal, seed establishment and regeneration to disperse over multiple generations (Travers et al., 2021). Also, grassland ecosystems have been found to propagate effectively via transportation verges (Tikka et al., 2001).

When considering animal and plant dispersal perpendicular to transportation corridors, barrier effects can be considerable leading to genetic isolation of populations over multiple generations and an overall fragmenting impact on biodiversity (Holderegger and Di Giulio, 2010; Barrientos et al., 2019; Ouedraogo et al., 2020). It is shown below that the magnitude of the barrier effect is dependent upon physiological and behavioural aspects as well as the composition and layout of the verge environment and surrounding land use. Mortality due to crossing attempts can also impact on population size to significant levels (Barrientos et al., 2019).

Research on animal mortality around transport networks focusses mainly on vehicular collisions, although electrocution, drowning and catenary line collisions have been studied (Holderegger and Di Giulio, 2010; Barrientos et al., 2019; Ouedraogo et al., 2020). A 2019 study suggested that current low mortality figures in some areas could also be indicative of massive past mortality, impacting populations and behaviours (Ascensao et al., 2019).

Areas of high functional connectivity, high traffic density and high road sinuosity (hence low visibility) were modelled to be high risk factors for animal mortality, as were road junctions in fenced highways and areas where roads bi-sect favourable cover, foraging, or breeding habitat (Grilo et al., 2011; Cserkesz et al., 2013). Gunson et al. (2011) provides an informative table of predictors for WVCs grouped by ungulates, carnivores, small-medium vertebrates, birds, and amphibians and reptiles (Gunson et al., 2011). The relatively infrequent traffic seen on railways (when compared with roads) is considered to introduce an element of unpredictability which can influence mortality risk (Dorsey et al., 2015).

Insect mortality is higher in grassland areas than forests and pollinator crossing likelihood appears to be linked to the availability of resources on the near verge (Daniel-Ferreira et al., 2022). Behavioural studies of two butterfly species revealed a reluctance of one species to cross habitat edges, suggesting that edge tolerant species are more likely to emigrate successfully across highly fragmented landscapes (Ries and Debinski, 2001). Reporting on invertebrate road mortality data, the UK based survey, “Buglife” has revealed a 63.7% decrease in “insect splat” events over 18 years up to 2022, suggesting marked declines in insect populations, which is consistent with a 2017 study in Germany which reported a 76% drop in flying insect biomass over 27 years (Ball, 2022).

Amphibian mortality is higher after rain events, and as well as crossing to reach favourable habitat, road crossing could serve other purposes including foraging, dispersal and regulation of body temperature (Timm and McGarigal, 2014). Studies on dormice and newts revealed no road barrier effect (Prunier et al., 2014; Friebe et al., 2018). A study in the Czech Republic found that forested areas appear to increase road mortality risk for medium sized carnivores such as the otter, marten, polecat, fox and badger (Cervinka et al., 2015). Genetic studies on wolverines demonstrated sex-bias in road crossing behaviours which may lead to isolation of populations and demographic fragmentation (Sawaya et al., 2019).

Over and under corridor crossing structures are most cited interventions to mitigate against Wildlife-vehicle collisions, but exclusion fencing, audible alarms, repellent odours, tree screening are also promoted in appropriate circumstances. Purpose built wildlife overpasses were found to be used significantly more than those shared with humans, and natural materials were preferred (Cserkesz et al., 2013; Denneboom et al., 2021).

Short under-road crossings are favoured over long ones and some studies found that central reservation, or median strips of vegetated land linking crossings are an effective remedy. Preferences for sizes and entry characteristics as well as spacing were different across taxa. The characteristic of adjacent vegetation is also important (Clevenger et al., 2001; McDonald and St Clair, 2004; Taylor and Goldingay, 2014).

Connectivity models on the impact of tree loss to ecological connectivity demonstrate that roadside trees provide a disproportionately high contribution to landscape-scale dispersal ability (Henry et al., 2017). Spatial dynamic modelling of land use change models found that development resulting in 10% habitat losses could half populations of some species within the landscape (Aurambout et al., 2005). A study in Poland of non-specialised pollinators (butterflies, bees, hoverflies) along railway verges found that community composition on embankments were spatially dependant, indicating that railways potentially perform a dispersal corridor function. (Moron et al., 2017).

A 2011 study in Texas on small mammal movements across and along highways found that small mammals have been observed to utilise unfarmed crop corners, and so these features act as remnant habitat when linking up with transportation verges for habitat connectivity (Kuykendall and Keller, 2011). The State Acres for Wildlife Enhancement programme for small mammal conservation in United States of America, reported Vole abundance during the first year following restoration depended on spatial connectivity provided by linear habitats (roadside ditches and grass waterways) (Mulligan et al., 2013). In an urban setting, public green spaces were found to increase “urban-adapted” bird connectivity, although the cumulative effect of urban barriers, traffic volume, distance between vegetation and amount of canopy cover can impact this significantly (Tremblay and St Clair, 2009; 2011; Rycken et al., 2022). A study on possum populations in Australia found that canopy connectivity is an important factor as crossing at floor level does not occur (Wilson et al., 2007).

Management actions may increase plant diversity by emulating natural disturbance regimes (Suarez-Esteban et al., 2016). However, this can create pronounced fluctuations in environmental conditions, including access to light and nutrients, which increases susceptibility to invasion by exotic species. Maintaining a perennial vegetation cover reduces this threat as well as enhancing nutrient cycling, reducing erosion and controlling noise. However, hydroseeding practices using non-native seed mixtures can displace native flora (Milton et al., 2015; Lazaro-Lobo and Ervin, 2019). Topsoil spreading to control erosion was found to accelerate establishment of a dense plant cover and ecosystem functioning (Lazaro-Lobo and Ervin, 2019). Embankments planted with woody vegetation did not promote natural plant colonisation and plant colonization occurs spontaneously in road verges, dependant on embankment age and surrounding vegetation (Arenas et al., 2015). A study of abiotic and biotic drivers in forest edges along railway lines in Sweden found that early successional productivity was influenced by the herb layer, soil moisture, humidity, and altitude, whilst late successional productivity was influenced by forest edge profile, canopy stratification and cover, forest continuity and the shape and contrast of edges in the surrounding landscape. (Wistrom and Nielsen, 2016).

A study into hedgerow planting found that a mosaic of hedgerows and grassland habitats increased spider and plant biodiversity at a landscape scale (Le Viol et al., 2008). Studies in Iowa which restored prairie habitat with the aim of increasing butterfly biodiversity found that the species richness of plants in flower had the strongest effect on richness and abundance of both generalist and habitat specialist butterfly species (Ries et al., 2001). Bee and dragonfly biodiversity was found to benefit from reduced mowing frequency, reduced artificial lighting, and replacement of invasive species with native nectar sources (Villemey et al., 2018; Phillips et al., 2020). Also, maintaining high flower densities and diversity in verges were found to reduce road crossing mortality of flower-visiting invertebrates (Daniel-Ferreira et al., 2022). Planting the larval foodplant in verges with many nectar plants was found to encourage the Oregon silver spot butterfly to remain and lay eggs on one side (Zielin et al., 2016).

Invertebrate mortality under differing mowing regimes were found to vary across taxa but was optimally two or less per annum, avoiding key life stage windows (Halbritter et al., 2015; Villemey et al., 2018; Parile et al., 2021). A study on meadow butterflies found that mid-summer mowing led to lower population density and delayed return to the area until the vegetation had regenerated (Valtonen and Saarinen, 2005). A study on monarch butterflies found that egg abundance was higher in mowed plots, and best results were seen in plots mowed once before egg-laying peaks (Knight et al., 2019). The use of eco-friendly mowing heads was found to reduce invertebrate mortality but effectiveness varied across taxa, with full mitigation for butterflies and moths and poor results for bees, wasps and beetles (Steidle et al., 2022).

Bat response to artificial lighting is species dependent. One review concluded that bat activity was highest in white light and lowest with orange (Ouedraogo et al., 2020). Bats appear to have an acoustic disturbance threshold, avoiding roads when noise levels reached around 88 dB for the studied species (Bennett and Zurcher, 2013).

Studies indicate that vegetated overpass crossing structures can benefit a broad range of taxonomic groups, including birds, amphibians, and reptiles (Jones and Pickvance, 2013; McGregor et al., 2015; Colley et al., 2017). A study in Brisbane found that forest specialist birds were also found to use wildlife overpasses designed for ground dwelling animals (Jones and Pickvance, 2013). Abundant vegetation and habitat restoration in the locality of crossing structures was found to increase the likelihood of use (Clevenger et al., 2001; Ng et al., 2004). Fencing has also been promoted in some locations where favourable habitat is located on both sides of the corridor (Ng et al., 2004), as is insulation of exposed electrical equipment for the protection of birds and arboreal animals (Laidlaw et al., 2021).

A pilot scheme in Switzerland installed “Amphibian plates” to the inside and outside of rails to facilitate crossing of rail tracks with some success (Chang et al., 2014; UIC, 2023). A study in Quebec found that a vegetated median strip improved the use of under-road crossings by providing foraging opportunities for some small mammals between culverts and by reducing the overall length of culvert (Martinig and McLaren, 2019). A pilot scheme was implemented in the Netherlands which utilised a modified hollow rail sleeper to enable small animals such as frogs, snakes and toads to safely cross the network below the tracks (UIC, 2023).

Whilst the loss of total habitat area is understood to consistently reduce biodiversity and negatively impact populations, research reveals that verge environments which are functionally similar to adjacent habitat patches can increase the overall regional habitat area. In this capacity, verges may function as stepping-stone habitats. The extensive reach of transportation verges throughout the landscape provides an opportunity to utilise these features for biodiversity gains. Also, studies indicate that overall biodiversity can be increased by eco-sensitive management interventions of the verge where it is an ecotone between open and dense habitats. Ecotones have been described in literature as having potential to support greater biodiversity (Perry et al., 2008).

We have seen that habitat specialists suffer greatest harm from habitat loss and fragmentation, and therefore stand to gain more from increased efforts to provide suitable habitat conditions. There is therefore a risk that a management focus on biodiversity indicators of richness and abundance alone would fail to address the needs of habitat specialists. It would therefore be beneficial to consider habitat specialist species within the list of target species for habitat restoration. A review of the IUCN Red List across a range of taxa for the region under consideration could yield candidate species for special consideration.

In human-altered environments, we have seen that verges offer significant opportunities to maintain biodiversity by providing suitable conditions for plant and animal refuge and dispersal between habitat patches, which may be prohibitive in adjacent areas. It follows that verges can provide a valuable ecosystem function at landscape scale. Improvements in verge condition and vegetation continuity between habitat patches will therefore promote ecological recovery. Additionally, other valuable ecosystem services such as pollination and crop pest control can result from these improvements. Therefore, local context should be considered when making decisions on verge management practices.

We have seen that the physiological characteristics of an organism greatly influence dispersal ability. This organism-specific response precludes a “one size fits all” approach to optimising the verge environment for ecological connectivity. Therefore assessment of interventions consider representative communities rather than individual species.

Also, when considering dispersal and range-expansion a key consideration is the rate of movement. In many animals this may be expressed by speed across the terrain, but for smaller and less mobile organisms such as plants it is often more appropriate to consider dispersal over generations (Beier et al., 2008). In the latter case, the verge environment would need to support all life cycle stages and is therefore likely to resemble native habitat for slow-dispersing species.

The network of roads and railways bisecting the landscape is a significant fragmentation factor and studies suggest that this can create physiological and behavioural barriers to movement, potentially leading to genetic isolation and eventually population decline. A key question is how to facilitate ecological connectivity to beneficial levels whilst at the same time managing the infrastructure for safe and cost-effective operations. This would involve raising awareness within all the business functions able to contribute and having a structure in place to coordinate, implement, monitor, and evaluate such activity.

We can draw a general conclusion that mortality risk is highest where roads bisect areas of natural functional connectivity (Grilo et al., 2011). This indicates that an efficient strategy to both reduce losses and to facilitate functional connectivity is to prioritise crossing safety at these locations.

Crossing structures can range from simple pipes within the ground through to elaborate vegetated overbridges. However, studies suggest that installing culverts and pipes of varying sizes at frequent intervals at crossing locations would benefit a large range of taxa from invertebrates up to small mammals (Clevenger et al., 2001). For larger animals, crossing prevention may help to limit mortality.

Climate change and land use change will influence the effectiveness of connectivity measures over time due to shifts in vegetation assemblies (IPCC, 2001). Therefore, measures to maintain and improve connectivity along verges should allow for consideration of these temporal factors on species distributions and the effectiveness of proposals over time. Successful implementation at landscape scale depends upon effective collaboration with stakeholders and coordinated design (Weber and Allen, 2010; Apfelbaum et al., 2012).

Research indicates that, where multiple habitat types co-exist in an area, verges containing stepping-stone patches of habitat at suitable intervals for the focal species will provide permeability for multiple habitat ecosystems. Also, since studies confirm the benefits of transportation verges in functional connectivity for pollinators and woodland/forest species it seems reasonable to adopt a strategy for protection and restoration of these habitat types where they exist along verges today.

Verge management practices need to consider safety aspects related to the operational infrastructure, such as route visibility or windthrow, and some treatments may be prohibitive due to regular access requirements (Milton et al., 2015; Villemey et al., 2018). The methods used would also need to consider risks and hazards to all focal species (Beier et al., 2008). The challenge for management is therefore to manage vegetation to enhance the positive impacts and reduce their negative effects (Milton et al., 2015).

Given the reach of road and railway networks throughout the landscape, it would be prudent to identify and prioritise locations on the network which are best able to support the function of habitat connectivity at a landscape scale. With this knowledge, management action can be coordinated in such a way as to promote biodiversity and habitat connectivity at the same time as prioritising for transportation delivery. A body of knowledge should be collated on vegetation management methods and interventions which promote biodiversity and reduce harm within given habitat types and given taxa. Whilst research is available to support this, we have seen that there are still many gaps across taxa.

Appropriate siting and design of crossing structures will be influenced by habitat distribution across the landscape and the taxa under consideration as well as commercial and operational constraints. Priority is likely to be given to underpass rather than overpass designs where appropriate due to cost and operational factors, especially in railway environments. When designing crossings, careful consideration is to be given towards vegetation composition around crossing points as research indicates that composition and layout can both help and hinder safe crossing (Cserkesz et al., 2013; Nevrelova et al., 2022). Culverts should be frequently spaced and of mixed sized openings to benefit a wide range of taxa (Clevenger et al., 2001; Jones and Pickvance, 2013). Similarly, livestock/wildlife fencing design requires a careful balance of preventing collisions and promoting dispersal (Ng et al., 2004; Holderegger and Di Giulio, 2010; Cserkesz et al., 2013; Ouedraogo et al., 2020).