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The definition of habitat fragmentation for this research is the partitioning of areas with natural land cover into smaller, separate and isolated patches (within a matrix of dissimilar habitats) through urbanisation, which is the replacement of vegetated land cover by impermeable surfaces (Collinge, 1996; Fahrig, 2003; Fischer &

Lindenmayer, 2007; Douglas & James, 2015; Mitchell et al., 2015).

Habitat fragmentation is a major problem worldwide, leading to the decrease of habitat area, isolation of populations, alteration of the conditions at habitat edges, and loss of biodiversity (Bierregaard Jr et al., 1992; Collinge, 1996; Millennium Ecosystem Assessment, 2005; Cain et al., 2011; Joint Nature Conservation Committee, 2014a).

There are two types of approach to the study of habitat fragmentation: the species- oriented and pattern-oriented approaches (Fischer & Lindenmayer, 2007). The study of habitat fragmentation (the pattern-oriented approach) started with the island biogeography theory (MacArthur & Wilson, 1967). This influenced seminal

ecological projects such as the fragmentation of tropical rainforest project started by Thomas Lovejoy in 1979. This project examined the biodiversity within intentionally divided and separated small patches of forests (Bierregaard Jr et al., 1992), and documented the effects of the distance between fragments, the size of fragments, and the physical and biological effects at fragment edges. This and other equally important research contributed to habitat fragmentation becoming a major research theme in conservation and restoration ecology (Bierregaard Jr et al., 1992; Fischer & Lindenmayer, 2007; Cain et al., 2011; Mitchell et al., 2015).

A critical review of the literature indicates that habitat fragmentation due to

urbanisation can have the following detrimental effects on biodiversity (Bierregaard Jr et al., 1992; Millennium Ecosystem Assessment, 2005; Laurance, 2008; Cain et al., 2011; Joint Nature Conservation Committee, 2014a; Mitchell et al., 2015):

• Loss of habitat area;

• Loss of top predators, which indirectly results in the loss of a population control mechanism;

• Small patches of habitats are susceptible to edge effects, which lead to the increase of invasive species abundance, increased rates of inbreeding and genetic drift;

• Spatial isolation of population, which increases the risk of local native species extinction.

Habitat loss can have a variety of negative effects on humans (Millennium

Ecosystem Assessment, 2005; Cardinale et al., 2012; Gottdenker et al., 2014). For example, Lyme disease carrying ticks (Acarina) population can expand due to loss of predators (Sol et al., 2013; Uspensky, 2014; Mitchell et al., 2015). The replacement of natural habitat with impervious surfaces can increase damage caused by flooding due to the loss of vegetation which previously slowed down stormwater runoff rate (Millennium Ecosystem Assessment, 2005; Woods-Ballard et al., 2007). The loss of vegetation coverage also increases air and water pollution (Maiti & Agrawal, 2005; Millennium Ecosystem Assessment, 2005; Qadir et al., 2013; Räsänen et al., 2013; Cohen et al., 2014; Baró et al., 2015; Paul & Nagendra, 2015). Our mental health is negatively affected due to loss of natural green spaces (Millennium Ecosystem

Assessment, 2005; Tzoulas et al., 2007; Croucher et al., 2008; Barton & Pretty, 2010; Konijnendijk, 2012; Maruthaveeran & Konijnendijk van den Bosch, 2014; Sandifer et al., 2015).

Habitat fragmentation can be reversed by employing techniques from ecological restoration, the study of which was founded by Anthony Bradshaw (Bradshaw, 1987 in Bradshaw, 1996). However, the term “ecological restoration” was first introduced in the late 1980s by John Aber and William Jordan (Douglas & James, 2015). It stems from conservation ecology, and its main aim is to reverse the effects of

fragmentation by increasing habitat connectivity (Hilderbrand et al., 2005; Vaughn et al., 2010; Cain et al., 2011; Sudduth et al., 2011; Douglas & James, 2015). The ecological concepts concerning restoration ecology are disturbance, genetic

diversity, succession (biological community composition recovers over time following a disturbance event), community assembly theory (similar sites can develop different biological communities depending on order of arrival of different species) and habitat fragmentation (Vaughn et al., 2010; Douglas & James, 2015). Finally, ecological restoration has a botanical bias (Douglas & James, 2015). Therefore, planting vegetation to create links (habitat corridors) and buffer zones is an effective technique that connects and protects fragmented habitats, thus allowing opportunities for wildlife to move around and utilise previously isolated, broken habitats.

Habitat corridors can help maintain biodiversity in a fragmented landscape. They are areas that connect two or more separated habitats, thus allowing organisms and matter to move around (Cain et al., 2011; Mitchell et al., 2015). Buffer zones (areas with less stringent controls on land use, yet which are at least partially compatible with many species resource requirements) can also help maintain biodiversity in a fragmented landscape (Marshall & Moonen, 2002; Cain et al., 2011; Mitchell et al., 2015; Street et al., 2015). Creating habitat corridors and buffer zones is part of the overall ecological restoration solution in urban areas (Collinge, 1996; Vaughn et al., 2010; Sudduth et al., 2011; Douglas & James, 2015)

Throughout the majority of the last century there was a significant decline in pond and wetland numbers in the UK, caused by urbanisation (Brown et al., 2010; UK

National Ecosystem Assessment, 2011a; Janse et al., 2015), which led to habitat loss and fragmentation for many of the UK’s aquatic ecosystems.

SuDS schemes have the potential to support and enhance freshwater biodiversity in urban areas. For example, in Dunfermline, Scotland, research found that SuDS ponds can support up to 47 invertebrate species (Briers, 2014). Jackson & Boutle (2008) showed that colonisation by aquatic fauna occurred at newly constructed SuDS swales and ponds at Upton, Northampton, UK. Therefore, detached River Nene Valley aquatic and semi-aquatic species can use these new SuDS features as places of refuge (Jackson & Boutle, 2008). Viol et al. (2009) observed that similarly rich and varied aquatic macroinvertebrate communities (displaying comparable composition and structure at the family level) can be supported by highway stormwater ponds, despite their poor water quality due to their pollutant retention function, compared with surrounding natural ponds (Viol et al., 2009). This makes the highway stormwater ponds being studied ideal wildlife refuges and connections to fragmented aquatic habitats (Viol et al., 2009). Moore and Hunt (2012) examined the richness and diversity of vegetated and aquatic macroinvertebrate communities in stormwater wetlands and ponds in the US. They found more than 50 vegetation species and 31 macroinvertebrate families are present in the stormwater ponds and wetlands surveyed (Moore & Hunt, 2012). They also noted that emergent vegetation plays a vital role in attracting some insect families (Odonatae) and provide a link to the vegetation at the littoral zones (or fringed wetlands), so that more diverse groups of macroinvertebrates can colonise, which helps provide different trophic functions to the ecosystem (Moore & Hunt, 2012).

Aside from ground level, vegetated SuDS systems, green roofs can also act as connections to fragmented habitats. For example, Kim (2004) examined how the Ecosystem Approach green roofs can reconnect fragmented habitats by studying the case of using different green roof designs (wetland, meadow, scrub, woodland, vegetable field) to form an urban eco-network in Seoul. Oberndorfer et al. (2007) reviewed ecosystem services that can be provided by green roofs, and found that they can support a variety of invertebrate and avian communities in several

countries. Rare and uncommon species of insects such as beetles, ants, bugs, flies,

bees, spiders, and leafhoppers have been recorded on green roofs around the world, which are positively linked to vegetation species richness (Brenneisen, 2006;

Oberndorfer et al., 2007). Bates et al. (2013) observed that green roofs with a range of substrate types can support a variety of species because they can act as different types of microhabitats, offering “disturbance refugia” for challenging environmental conditions, such as droughts (Bates et al., 2013). When looking at habitat

connectivity in terms of pollination, green roofs are comparable with ground-based green infrastructures, such as parks and prairies (Tonietto et al., 2011), even though they contain smaller and less diverse pollinators (Ksiazek et al., 2012).

Finally, various studies have shown the SuDS approach (especially when using vegetated SuDS techniques) can contribute to reversing habitat fragmentation by acting as wildlife corridors and buffer zones to connect and protect separated and isolated habitats due to urbanisation (Kim, 2004; Brenneisen, 2006; Oberndorfer et al., 2007; Jackson & Boutle, 2008; Viol et al., 2009; Tonietto et al., 2011; Ksiazek et al., 2012; Moore & Hunt, 2012; Bates et al., 2013; Briers, 2014). Nevertheless, stand-alone SuDS systems are not adequate in contributing to the efforts to reverse habitat fragmentation. SuDS sites (in particular sites that contain “micro-and meso- vegetative SuDS systems”) should work with existing urban green infrastructures (parks and gardens) as together they can provide not only a fully sustainable surface water management system, but also connectivity to habitats fragmented by

urbanisation, in order to recover previously lost urban biodiversity (Ashton et al., 2010; Wise et al., 2010; Natural England, 2011; Ellis, 2013; Graham et al., 2013).