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Managing invasive species in cities: A framework from Cape Town, South Africa
Landscape and Urban Planning 151 (2016) 1–9
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Landscape and Urban Planning journal homepage: www.elsevier.com/locate/landurbplan
Perspective essay
Managing invasive species in cities: A framework from Cape Town, South Africa Mirijam Gaertner a,b,∗ , Brendon M.H. Larson a,c , Ulrike M. Irlich a,b , Patricia M. Holmes a,b , Louise Stafford b , Brian W. van Wilgen a , David M. Richardson a a
Centre for Invasion Biology, Department of Botany and Zoology, Stellenbosch University, Private Bag X1, Matieland 7602, South Africa Environmental Resource Management Department (ERMD), City of Cape Town, Westlake Conservation Office, Ou Kaapse Weg, Tokai, 7966, Cape Town, South Africa c School of Environment, Resources and Sustainability, University of Waterloo, Waterloo, Ontario N2L3G1, Canada b
h i g h l i g h t s • • • •
Existing invasive species management frameworks are inadequate in urban areas. Urban stakeholders often hold conflicting views and are critical of management. Divergent stakeholder perceptions need to be considered explicitly and transparently. Urban management frameworks should allow for acceptance of some invasive species.
a r t i c l e
i n f o
Article history: Received 24 June 2015 Received in revised form 11 March 2016 Accepted 17 March 2016 Available online 28 March 2016 Keywords: Biological invasions Non-native species Urban invasive species management Table Mountain National Park Urban ecology
a b s t r a c t Invasive non-native species are often more prevalent in cities than in rural areas because of numerous environmental disturbances and higher propagule pressure. Attempts to manage invasive species in cities are often controversial because of the diversity of stakeholder views. Until now, however, environmental managers in cities have managed invasive species using approaches and paradigms developed for a rural context, despite the radically different socio-environmental conditions that prevail in cities. We examine the case of Cape Town, South Africa, a rapidly growing metropolitan centre within a global biodiversity hotspot and a developing country, to underline the considerable challenges and complexities of managing invasive species in cities. We argue that traditional management approaches need to be supplemented by consideration of stakeholder views and the social consequences of management actions. We present a framework for selecting appropriate goals for the management of invasive species, ranging from eradication to acceptance. © 2016 Elsevier B.V. All rights reserved.
1. Introduction Invasive non-native species (sensu Richardson, Pyˇsek, & Carlton, 2011; hereafter ‘invasive species’) are often abundant in cities (Kowarik, 2011). Cities contain a high density of people and they are hubs of human-mediated movement of commodities. Transport linkages (e.g., airports and harbours) facilitate the introduction
∗ Corresponding author at: Centre of Excellence for Invasion Biology, Department of Botany and Zoology, Natural Sciences Building, Private Bag X1, University of Stellenbosch, Matieland 7602, South Africa. E-mail addresses: [email protected] (M. Gaertner), [email protected] (B.M.H. Larson), [email protected] (U.M. Irlich), [email protected] (P.M. Holmes), [email protected] (L. Stafford), [email protected] (B.W. van Wilgen), [email protected] (D.M. Richardson). http://dx.doi.org/10.1016/j.landurbplan.2016.03.010 0169-2046/© 2016 Elsevier B.V. All rights reserved.
and dissemination of non-native species through dispersal pathways such as trade, tourism, and horticulture (Dehnen-Schmutz, Touza, Perrings, & Williamson, 2007); such activities release high numbers of individuals into a region (high ‘propagule pressure,’ see Lockwood, Cassey, & Blackburn, 2005). In cities these non-native species encounter habitats, soils, climatic conditions and hydrology that have been profoundly changed by human activity and that can promote their spread if they are pre-adapted to similar conditions in their region of origin (Pickett et al., 2001; Kowarik, 2011). Urban heat-island effects, for example, may facilitate the spread of invasive species (Nobis, Jaeger, & Zimmermann, 2009). Also, typical urban conditions such as fragmented habitats and altered disturbance regimes often favour non-native species (Cilliers, Williams, & Barnard, 2008; Zisenis, 2015). In Central Europe, most urbanophilic non-native plant species can be found in inner city areas, which
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provide suitable conditions for plant species that tolerate or even thrive when temperatures are warmer and disturbances more frequent (Klotz & Kühn, 2010). In short, many non-native species are more prolific in cities because long histories of human dispersal, disturbance and habitat modification enhance their opportunities for establishment, proliferation and spread. Invasions pose well-documented risks in both natural and seminatural habitats and in both protected areas and public open spaces. Invasive species may displace native species and contribute to homogenization of habitats within cities (Burton, Samuelson, & Pan, 2005; Kühn & Klotz, 2006; McKinney, 2006; Trentanovi et al., 2013). Ornamental invasive plant species in gardens act as significant sources of non-native propagules (Alston & Richardson, 2006; Bowers, Bean, & Turner, 2006), and non-native animals kept as pets can establish and become invasive (van Wilgen & Richardson, 2012). Invasive species in cities may also disrupt important ecosystem services such as water filtration, flood attenuation and coastal protection. Invasive plants can clog streams and canals, resulting in flooding, and they may also increase fire severity and soil erosion risk (van Wilgen & Scott, 2001). Although some invasive species were introduced to provide particular ecosystem services, such as trees for timber production or erosion control, their subsequent spread and proliferation may ultimately have a net detrimental effect (e.g. van Wilgen, Reyers, Le Maitre, Richardson, & Schonegevel, 2008; Vilà et al., 2009). Invasive species may also affect urban biodiversity in ways that reduce human well-being more directly, for example by changing the aesthetics of the environment (Fuller, Irvine, Devine-Wright, Warren, & Gaston, 2007; Kowarik, 2011). For all of these reasons, policies are in place in many parts of the world to manage invasive species in parks, public areas and other urban zones. Despite growing concerns about invasive species, some authors argue that we may, in some instances, have to learn to “cohabitate” with them in cities (e.g., Foster & Sandberg, 2004). Even if an invasive species disrupts an ecosystem service, it may also be beneficial. Studies in urban forests in Florida, for example, have shown that invasive trees were most successful in sequestering CO2 (Escobedo, Varela, Zhao, Wagner, & Zipperer, 2010). As another example, studies from northern parts of South Africa show that invasive Eucalyptus trees are used extensively as roosting sites for the vulnerable Lesser kestrel (Falco naumannii) and as breeding sites for the African fish eagle (Haliaeetus vocifer) (Cilliers & Siebert, 2012). Managing invasive species in urban areas is sometimes further complicated by the cultural connections that people have forged with them. Just like native species, invasive species can become associated with a place and be regarded as culturally important by some city inhabitants (Warren, 2007). Stakeholders in cities often have strongly divergent views about the impacts and benefits of particular invasive species, so conflicts over the management of invasive species are emerging (e.g., trees, Dickie et al., 2014). In particular, invasive species may provide provisioning ecosystem services, but at the expense of various elements of biodiversity, which can lead to conflicts over whether to manage for the former or the latter. Such conflicts exemplify the extent to which invasive species management, especially in human-dominated areas, is increasingly viewed as a “wicked problem” (Rittel & Webber, 1973) because sometimes there are no straightforward “win-win” solutions. A particular problem is that even if an invasive species is “accepted” (i.e., not regulated or targeted for containment or control) within a city environment, such acceptance may pave the way for its spread into adjacent rural landscapes where it may have substantial negative impacts (Botham et al., 2009; Moreira-Arce, de la Barrera, & Bustamante, 2014). In this paper, we explore the challenges and complexities of managing invasive species in cities by examining the exemplary case of Cape Town, South Africa. Cape Town is a rich case study
for elucidating the complexity of managing invasive species in cities because it highlights several interwoven social and ecological dimensions. We use this city to highlight challenges that will be faced by an increasing number of cities given ongoing urbanization and growing human populations. Our objectives are to utilize this case study: (1) to review the challenges faced by managers who seek to control invasive species in an urban environment; and (2) to develop a framework to assist environmental managers globally as they seek to integrate a range of management options for invasive species in urban systems to deal with diverse and often conflicting views of what is appropriate.
2. The case of Cape Town, South Africa Cape Town’s conservation significance derives from its location in the Cape Floristic Region, a global centre of plant endemism (Cowling, Rundel, Lamont, Arroyo, & Arianoutsou, 1996). The city (2445 km2 ) includes Table Mountain National Park (221 km2 ), as well as 17 smaller nature reserves and 500 biodiversity network sites that together cover 270 km2 . It has a population of 3.8 million people and is growing more rapidly than any other southern African metropolis on a per capita basis (Boraine et al., 2006), especially within its poorer suburbs (“townships”) which have experienced an influx of mainly Black citizens following the collapse of apartheid. At present, 26% of Cape Town is urban, 35% is agricultural, and 39% is natural and semi-natural vegetation concentrated in mountainous areas (mainly within Table Mountain National Park) (Fig. 1). Many lowland areas have been transformed, with remnants being highly threatened and thus having become a priority for conservation (Rebelo, Holmes, Dorse, & Wood, 2011). Cape Town remains a focal point of the national economy and international tourism, and thus faces ever-growing needs for housing, transport networks and trade. The fact that Cape Town has high levels of human population growth, unemployment and crime adds dimensions of complexity. Nonetheless, many of Cape Town’s citizens are involved in conservation initiatives, such as “Friends” groups for nature reserves and various conservation stewardship and citizen science initiatives (e.g., spotter networks for emerging invasive species, see http://www.capetowninvasives.org.za). Cape Town has a long history of European colonization and the associated introductions of non-native species present a significant challenge to people and landscapes (van Wilgen, 2012) (Fig. 2). For example, invasive tree species such as pines (Pinus species), grown in plantations, and Australian wattles (Acacia species), planted mainly along the coast for dune stabilization, have spread widely into natural vegetation. Aquatic invasive species such as Water hyacinth (Eichhornia crassipes) block waterways and affect water quality (Richardson & van Wilgen, 2004). The trade in ornamental plants and pets, and other enterprises that rely on non-native taxa, continue to introduce new species into the city; many of these remain undetected and/or unregulated. Some invasive plant species pose serious risks to humans; for example, invasive pines and wattles increase the severity of wild fires near residential areas (Fig. 1) (van Wilgen & Scott, 2001). The Department of Environmental Affairs is responsible for the overall administration of the National Environmental Management: Biodiversity Act (NEMBA), which places obligations on all landowners and all organs of state, including the City of Cape Town, with regard to the management of invasive species. Invasive species control programs date back to the 1940s (Macdonald, Clark, & Taylor, 1989), and in 2008 the city established an Invasive Species Management Unit, with an annual budget of about 84,000 USD and one semi-skilled team. Since then the program has grown to include areas managed by multiple departments within the city.
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Fig. 1. The City of Cape Town is situated around Table Mountain National Park. Lowland areas (red polygon) are transformed, and remnants of semi-natural vegetation are highly threatened and a priority for conservation. The spread of ornamental plants across the city threatens biodiversity in the national park. Photo: Bruce Sutherland, City of Cape Town. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)
Although originally funded by the city’s ratepayers, other sources (notably the government-sponsored Working for Water programme which aims to provide employment to poor people, see van Wilgen & Wannenburgh, 2016) have since provided additional and substantial funding, increasing the available funds to 1.9 million USD in 2014. The city now manages 55 teams averaging 10 workers each; their mandates include early detection and rapid response programs for emerging invasive species and the control of
more established ones. Besides these control operations, the city is piloting the promotion of indigenous plant species over non-native ones to provide the same benefits and reduce negative perceptions of control operations. The city spends 70% of its invasive species funds on controlling terrestrial invasive plants in nature reserves, targeting Australian Acacia species (e.g., Acacia cyclops, Acacia mearnsii, and Acacia saligna), Australian myrtle (Leptospermum laevigatum), and pines
Fig. 2. Examples where the management of invasive plant species in the City of Cape Town has been challenging and contentious: (a) Utilization of invasive Acacia cyclops stands for firewood; b) Invasive squirrels are being ‘protected’ by municipalities; Photo: Sophia Turner; (c) Clearing waterways choked with water hyacinth with heavy machinery; (d) Tuart trees have been planted along many roads to provide shade; Photo: Emma Booysen Photography; (e) Invasive trees as fire hazard, Photo: Chris Cloete; (f) Himalayan thars on Table Mountain; Photo: Tony Rebelo; (g) Recreational activities under pine plantations; Photo: Lisel Mc Gregor.
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(Pinus halepensis, Pinus pinaster and Pinus radiata). The bulk of the budget for aquatic weeds is spent on Water hyacinth and Mexican water lily (Nymphaea mexicana). Priority is given to nature reserves, followed by areas where invasive plants are a fire hazard or provide shelter for criminals. Areas that have been cleared of non-native plants are prioritized for follow-up operations to prevent re-invasion. The objective is to restore native biodiversity where possible or to at least ensure adequate levels of ecosystem functioning through either active or passive restoration. Other organizations involved in control of non-native species are South African National Parks (e.g. Foxcroft & Freitag-Ronaldson, 2007, the South African National Biodiversity Institute (Wilson, Ivey, Manyama, & Nänni, 2013) and the Working for Water Programme (van Wilgen & Wannenburgh, 2016). Control efforts are often compromised because of significant conflicts over the management of invasive species. These conflicts are particularly severe in a city environment because of a diversity of stakeholder views. For example, conservationists and citizen science groups support the control of invasive species (mainly to protect biodiversity and to decrease negative impacts on ecosystem services), but control operations are met with strong resistance from commercial foresters, in the case of pine plantations, and also from citizens who have concerns over loss of forested space in urban environments (van Wilgen, 2012). 3. A general framework for invasive species management in cities Management of invasive species worldwide is usually directed towards one or more goals. These typically include the prevention of spread to non-invaded areas, local eradication of isolated populations, containment where local eradication is not possible, and the protection of selected assets or areas where the invasive species are so widespread that containment is no longer feasible (Grice, Clarkson, & Calvert, 2011; Auld & Johnson, 2014). Such goals are simpler to incorporate in decision-making protocols in protected areas and rural situations than in cities where divergent stakeholder views and resulting conflicts of interest (for example relating to biodiversity conservation and the provision of ecosystem services) often complicate decisions. Managers of invasive species in cities need a framework for making decisions on how to deal with individual species in a way that will lead to acceptable and effective control. In some cases, it may be acceptable (or justifiable) to proceed with control, while in others it may be better to tolerate a species. Whether a given species should be tolerated or not is highly context-specific and depends on its potential for causing significant impact (now and in the future) and on its perceived value. It is useful to envisage the positioning of species along two axes: the first considers the real or perceived benefits associated with the species, while the other considers its potential to generate significant negative impacts (Fig. 3). This provides a framework in which species can be positioned along these two axes, giving rise to three categories. (1) Species with a low potential for negative impact, regardless of the benefits they deliver, would fall into a “tolerance” category. These species would be generally acceptable to society and to conservationists, requiring very little management. For some species no action will be needed (those with high perceived benefits and low perceived impacts), but some may need to be monitored (species with low perceived impacts, but that may become more problematic with time, e.g., emerging weeds). (2) Species that both provide benefits and have negative impacts fall into an “active engagement” category. Their management
will be more complex because the optimum course of action is less clear, and any option is likely to be controversial. Here the appropriate course of action would depend both on the societal response and the political consequences of deciding to control or tolerate the species. In some cases, the ongoing presence of harmful species might have to be accepted (where there is strong societal opposition to control and/or insufficient evidence to demonstrate widespread impacts or justify expensive intervention). (3) Species with a relatively high potential for impact and that deliver relatively low benefits would fall into a “control priority” category. The “control priority” category will include species to be eradicated (where eradication is possible) or controlled (where eradication is not feasible, e.g., widespread plant species with a long-lived seedbank). Depending on the circumstances, control could aim to reduce the extent of the invasion and the density of populations, prevent further expansion, or (in cases where the species is very widespread) focus on protecting particular assets (e.g., where a species poses a high risk in terms of fire hazard or to endemic species or ecosystems).
Determining the real (or perceived) benefits and negative impacts of a species can be difficult to quantify and it is often contentious. While some information can be obtained from the literature, or through research if funding and time allows, the full range of understanding and perceptions need to be explored through stakeholder engagement. It is challenging to design appropriate stakeholder engagement that develops inter-party trust, not least because this is time-consuming, and time may not be available in the context of a ‘rapid-response’ program (see Mackenzie & Larson, 2010). For assigning species to the management categories presented in Fig. 3, we propose a simple decision tree (Fig. 4). Unlike traditional approaches to invasive species management, our proposed decision tree explicitly considers the potential societal reactions to invasive species control operations in cities, which would become apparent in consultations with affected local stakeholders. In our experience (as elsewhere, e.g., Dickie et al., 2014), societal opposition to the control of particular types of invasive species will at best add substantial costs and delays to control projects, and at worst prevent them from proceeding at all (and potentially affect management initiatives on other invasive species). It is thus clearly prudent to anticipate and plan for such possibilities, and to accept that, in some cases, traditional goals need to be replaced with partial or complete tolerance of invasive species. We note, however, that there are very real uncertainties about both the ecological and social status of many species, especially in the case of emerging invasive species, which may necessitate research to facilitate rapid decision-making (Mackenzie & Larson, 2010). Rapid decision-making to deal with emerging invaders is crucial; initiation of management efforts at an early stage of invasion is the most cost-efficient (Mgidi et al., 2007). Another advantage of managing species at an early stage of invasion is that cultural connections with the non-native species have not yet been established and opposition to management is likely to be more limited. The decision of whether or not, and if so where and how, to manage invasive species, is also influenced by factors unrelated to impacts or benefits. Even for damaging invasive species, if no effective control methods exist it may be prudent to tolerate the species, and to deploy resources on other species that can more effectively be controlled (Hobbs et al., 2014). In degraded areas, where the costs of rehabilitation would exceed the benefits of control, an invasive species may also be tolerated, although steps should be taken to prevent spread to less degraded areas.
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Fig. 3. Proposed categorization of invasive species according to perceived benefits and potential negative effects (invasiveness, ecosystem and social impacts). Three management approaches are recognized: (1) tolerance; (2) active engagement; and (3) control priority. Shading reflects the probability of conflicts of interest (light grey to dark grey = low to high probability).
Experience with invasive species in Cape Town allows us to use our decision tree (Fig. 4) to categorize species from a wide range of taxa within this framework (Table 1; Fig. 3). We describe several examples below, with the ‘steps’ of the decision tree (Fig. 4) included in brackets. 4. Application of the framework, with examples from Cape Town We begin with several examples from the genus Eucalyptus. Karri gum (Eucalyptus diversicolor), for example, occurs in plantations in Table Mountain National Park that were planted over a century ago and have grown to an impressive size. Karri gum plantations are popular with hikers, cyclists and tree enthusiasts (van Wilgen, 2012) (perceived benefits are high). Although these gum trees impact water resources, they are not highly invasive (impact low). Furthermore, proposals to clear Karri gum plantations and to restore those areas to fynbos shrublands were vigorously opposed. Because of the political ramifications of going against public opinion, especially in the case of a species that is not aggressively invasive, it was decided to tolerate and monitor the species and to retain the plantations (monitor) (van Wilgen, 2012). Similarly, Tuart (E. gomphocephala), which is endangered in its native Western Australia, has been widely planted in Cape Town’s poorer suburbs and along many roads where it provides shade and does not spread. This species has and should be tolerated. River red gum (E. camaldulensis), on the other hand, is both attractive (high benefit) and highly invasive (high impact), especially along river courses (Forsyth, Richardson, Brown, & van Wilgen, 2004), which suggests that there will be social opposition to clearing it (active engagement). An appropriate response here would be to tolerate large specimens in parks and gardens, but to remove plants from protected areas and river courses, and to prohibit further sale and planting anywhere. Here our decision tree allows for “partial tolerance” as a possible outcome. Pine trees introduced from Europe and North America also can be classified using the framework. Cluster pine (P. pinaster) and Aleppo pine (P. halepensis) have negligible commercial value and
relatively little aesthetic appeal to local people (low benefit). They impact water resources and biodiversity, and increase fire severity (Richardson & Higgins, 1998) (high impact). Furthermore, effective control methods exist, which places them in the control priority category (impact reduction and containment). In contrast, stone pines (P. pinea) are often viewed as “heritage trees” on Table Mountain and elsewhere (van Wilgen, 2012), (high perceived benefit). The species is widespread, but not highly invasive (low impact) and should hence be monitored and tolerated where appropriate. Monterey pine (P. radiata), on the other hand, occurs in commercial plantations that are popular for the same reasons as karri plantations (high perceived benefit), but they are highly invasive (Richardson & Brown, 1986) and pose a substantial threat to the biodiversity of Table Mountain National Park (Richardson et al., 1996) (high impact). Monterey pines thus require active engagement, and there are plans to systematically remove them from plantations and protected areas. Because there has been public resistance to these removals (van Wilgen, 2012), it is important to actively engage with the public in stakeholder workshops about the proposed removal (and reasons for it) and a potential plan for replacing them. In Cape Town, opposition to the removal of pine plantations was countered by support from the national and provincial conservation agencies, NGOs such as WWF (South Africa), and leading academics (equal opposition and support). In addition, the control and rehabilitation operations provided opportunities for employment, which was politically attractive and could offset jobs lost in forestry-related activities (political consequences). Control should therefore be continued and awareness for the species impacts is being raised. Attempts to control most Australian wattles (Acacia species) have not invoked the same level of negative reaction as pines and eucalypts (low perceived benefit), and most species are known for their high impacts. Acacia species can hence be placed within the category of control priority. An exception is Pepper tree wattle (Acacia elata), which is valued as a shade and amenity tree, especially of golf courses; management of this invasive species requires active engagement (Donaldson et al., 2014).
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Fig. 4. Assigning species to management categories within the boundaries of the City of Cape Town. Tolerance category (highlighted green), control priority category (highlighted orange) and active engagement category (highlighted blue) are described in detail in the text. The position of a selection of invasive species in the three “management approach” zones is shown in Fig. 3. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)
Some non-native aquatic plant species, such as Water hyacinth, are highly invasive and have significant negative impacts but no benefits (Richardson & van Wilgen, 2004), so they are placed into the control priority category. Managing damaging invasive animals is often controversial in cities, where residents challenge the ethics of killing or removing animals, and point to the perceived cruelty of these operations. For example, a decision was taken to attempt to eradicate an isolated but thriving population of introduced Himalayan thars (Hemitragus jemlahicus) on Table Mountain, because of the damage they did through over-grazing (high impact), and to make the re-introduction of historically extirpated Klipspringer antelope (Oreotragus oreotragus) possible. Public reaction was substantial, and the campaign had to proceed in the face of strong resistance (active engagement category). However, the campaign received strong backing from conservation authorities and NGOs, making it feasible to continue (equal opposition and support). A similar campaign to remove European mallards (Anas platyrhynchos), which interbreed with the indigenous Yellow-billed duck (Anas undulata) (high impact), was also met with resistance. Management efforts were effectively halted because the arguments for the campaign (genetic contamination of a single indigenous species) were less convincing to the public than arguments for the widespread ecolog-
ical impacts of more damaging invasive species (strong opposition and relatively low perceived impact). Similarly, any attempt to control Eastern grey squirrels (Sciurus carolinensis) to reduce limited impacts on biodiversity would almost certainly meet with strong resistance from some quarters. Interestingly, a campaign to remove a large and growing population of Indian house crows (Corvus splendens) met with little resistance, and has proceeded without substantial opposition. Several lessons have been learned from managing invasive animal and plant species in the city of Cape Town. Experience has shown that we need to use our framework to decide on a course of action when any new invasive species becomes a potential target for control. In order to make logical and defensible choices, it would further be essential to gauge the views of stakeholders. Where significant numbers of people are likely to oppose management, the potential political consequences of different courses of action, must be explicitly considered. It also has to be accepted that conventional goals (such as eradication or reductions in populations) will need to be replaced with partial or complete tolerance of invasive species in some cases. A potential critique of our framework is that ‘perceived impacts’ and ‘perceived benefits’, the core steps in our framework, are context specific, difficult to quantify, and subjective. Such con-
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Table 1 Arguments for and against the control of selected non-native species mentioned in the text, and the rationale for their placement into management categories (Fig. 3). PA = protected area (including Cape Towns’ critical biodiversity areas and the Biodiversity Network). Species
Characteristics
Negative effects
Positive effects
Management solution
Acacia cyclops (Red eye)
Tree or shrub, ecosystem transformer
Impacts on biodiversity, water, sand movement and fire intensity
Important source of firewood
Control priority, as impacts outweigh benefits. Remove from PAs. Contain elsewhere.
Acacia elata (Peppertree wattle)
Tree, ecosystem transformer
Impacts on biodiversity
Aesthetic value
Active engagement needed to garner support for control.
Acacia mearnsii (Black wattle)
Tree, ecosystem transformer
Impacts on biodiversity and water
Source of firewood
Control priority, as impacts outweigh benefits. Remove from PAs. Contain elsewhere.
Acacia paradoxa (Kangaroo thorn)
Shrub, potential ecosystem transformer
Impacts on biodiversity
None
Control priority. Distribution restricted, so target for eradication.
Acacia saligna (Port Jackson willow)
Tree, ecosystem transformer
Impacts on biodiversity and water
Source of firewood and fodder
Control priority, as impacts outweigh benefits. Remove from PAs. Contain elsewhere.
Ailanthus altissima (Tree of heaven)
Tree, ecosystem transformer
Impacts on urban infrastructure and potentially on biodiversity
Aesthetic value
Active engagement needed to garner support for control.
Eichhornia crassipes (Water hyacinth)
Aquatic weed, ecosystem transformer
Impacts on water quality, water distribution networks and biodiversity
Aesthetic appeal (limited)
Control priority. Remove from water bodies.
Eucalyptus camaldulensis (River red gum)
Tree, ecosystem transformer
Impacts on biodiversity and water
Heritage tree status, aesthetic value
Active engagement needed to garner support for control along rivers. Tolerate elsewhere, but phase out over time.
Eucalyptus diversicolor (Karri gum)
Tree, naturalized
Impacts on biodiversity and water
Heritage tree status, aesthetic value
Species to be tolerated, as not aggressively invasive.
Eucalyptus gomphocephala (Tuart)
Tree, naturalized
Impacts on groundwater
Aesthetic value, shade in townships
Species to be tolerated, as not aggressively invasive.
Nymphaea mexicana (Mexican water lily)
Aquatic plant, invasive
Potential impacts on aquatic biodiversity
Aesthetic appeal
Control priority. Remove from water bodies.
Pinus halepensis (Aleppo pine)
Tree, ecosystem transformer
Impacts on biodiversity, water and fire intensity
Previously valued for timber
Control priority, as impacts outweigh benefits. Remove from PAs. Contain elsewhere.
Pinus pinaster (Cluster pine)
Tree, ecosystem transformer
Impacts on biodiversity, water and fire intensity
Previously valued for timber
Control priority, as impacts outweigh benefits. Remove from PAs. Contain elsewhere.
Pinus pinea (Stone pine)
Tree, naturalized
Not native to protected areas, local biodiversity impacts
Heritage tree status, aesthetic value
Species to be tolerated, as not aggressively invasive. Limited and ongoing removal from PAs.
Pinus radiata (Monterey pine)
Tree, ecosystem transformer
Impacts on biodiversity, water and fire intensity
Source of timber, aesthetic value
Active engagement needed to garner support for removal of plantations. Ongoing removal from PAs.
Sambucus canadensis (Canadian elder)
Shrub, invasive
Potential impacts on biodiversity
Popular garden plant
Species to be tolerated and monitored until effective control methods can be found.
Anas platyrhynchos (European mallard)
Aquatic bird, naturalized
Hybridization with native ducks
Ethical (animal rights), aesthetic
Species to be tolerated, as perceived benefits outweigh impacts.
Corvus splendens (Indian house crow)
Bird, invasive
Impacts on biodiversity and health risks; nuisance value
Ethical (animal rights)
Control priority, as impacts outweigh benefits. Target for local eradication.
Hemitragus jemlahicus (Himalayan thar)
Large mammalian herbivore, invasive
Impacts on biodiversity
Ethical (species threatened in native range; animal rights)
Active engagement needed to offset opposition to control. Target for eradication.
Sciurus carolinensis (Eastern grey squirrel)
Rodent, naturalized (dependent on non-native trees)
Minor impacts on biodiversity
Ethical (animal rights), aesthetic (cuteness)
Species to be tolerated, as perceived benefits outweigh impacts.
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cerns should be openly debated as part of stakeholder engagement to ensure that all parties agree on such evaluations.
5. Conclusions Managing invasive species in cities provides additional challenges to those traditionally faced in agricultural areas or natural ecosystems, so existing frameworks for guiding management are increasingly being found inadequate in this context. The problem becomes acute where urban areas abut or enclose regions of high conservation value. Cape Town is an exemplar, so we draw upon insights from recent attempts to manage different types of invasive species in this city to highlight the need for a new framework to guide invasive species management in cities more generally. Two fundamental recommendations emerge from our analysis. First, divergent perceptions need to be given explicit and transparent consideration. Most people do not fully understand issues related to biodiversity and conservation, or the negative economic impacts of invasive species, but a platform for airing their views must be provided. Experience has shown that failure to do so often results in resistance that may derail long-term plans (see Mackenzie & Larson, 2010 and references therein). Second, we may need to tolerate some widespread invasive species within cities for a combination of social and pragmatic reasons. This acknowledgement parallels the recognition that some ecosystems have been so transformed that they cannot practically be restored to their historic condition (Richardson & Gaertner, 2013; Hobbs et al., 2014). However, we caution that where some form of cohabitation is agreed upon, approaches should be sought for minimizing potential negative impacts of retaining areas with invasive species. Potential pro-active avenues to consider include biological control, ecological restoration and gradual replacement with alternative non-invasive species. Our framework provides a workable solution to the failures of implementing rural management approaches to invasive species within urban areas.
Acknowledgements We acknowledge funding from the DST-NRF Centre of Excellence for Invasion Biology (M.G., B.L., B.v.W. and D.M.R), a Standard Research Grant from the Social Science and Humanities Research Council of Canada (B.L.), and the National Research Foundation, South Africa (D.M.R.; grant 85417; B.W.v.W.; grant 87550). This work was also supported by the National Socio-Environmental Synthesis Center (SESYNC) under funding received from the National Science Foundation DBI-1052875, and by sDiv, the Synthesis Centre of the German Centre for Integrative Biodiversity Research (iDiv) Halle-Jena-Leipzig (German Research Foundation DFG FZT 118).
References Alston, K. P., & Richardson, D. M. (2006). The roles of habitat features, disturbance, and distance from putative source populations in structuring alien plant invasions at the urban/wildland interface on the Cape Peninsula, South Africa. Biological Conservation, 132(2), 183–198. http://dx.doi.org/10.1016/j.biocon. 2006.03.023 Auld, B. A., & Johnson, S. B. (2014). Invasive alien plant management. CAB Review, 9(37), 1–12. http://dx.doi.org/10.1079/PAVSNNR20149037 Boraine, A., Crankshaw, O., Engelbrecht, C., Gotz, G., Mbanga, S., Narsoo, N., et al. (2006). The state of South African cities a decade after democracy. Urban Studies, 43(2), 259–284. http://dx.doi.org/10.1080/00420980500416990 Botham, M. S., Rothery, P., Hulme, P. E., Hill, M. O., Preston, C. D., & Roy, D. B. (2009). Do urban areas act as foci for the spread of alien plant species? An assessment of temporal trends in the UK. Diversity and Distributions, 15(2), 338–345. http://dx.doi.org/10.1111/j.1472-4642.2008.00539.x Bowers, J. E., Bean, T. M., & Turner, R. M. (2006). Two decades of change in distribution of exotic plants at the desert laboratory, Tucson. Arizona. Madro˜ no, 53(3), 252–263. http://www.jstor.org/stable/41425651
Burton, M., Samuelson, L., & Pan, S. (2005). Riparian woody plant diversity and forest structure along an urban-rural gradient. Urban Ecosystems, 8(1), 93–106. http://dx.doi.org/10.1007/s11252-005-1421-6 Cilliers, S., Williams, N. G., & Barnard, F. (2008). Patterns of exotic plant invasions in fragmented urban and rural grasslands across continents. Landscape Ecology, 23(10), 1243–1256. http://dx.doi.org/10.1007/s10980-008-9295-7 Cilliers, S. S., & Siebert, S. (2012). Urban ecology in cape town: south african comparisons and reflections. Ecology and Society, 17(3), 33. http://dx.doi.org/ 10.5751/ES-05146-170333 Cowling, R. M., Rundel, P. W., Lamont, B. B., Arroyo, M. K., & Arianoutsou, M. (1996). Plant diversity in Mediterranean-climate regions. Trends in Ecology and Evolution, 11(9), 362–366. http://dx.doi.org/10.1016/0169-5347(96)10044-6 Dehnen-Schmutz, K., Touza, J., Perrings, C., & Williamson, M. (2007). A century of the ornamental plant trade and its impact on invasion success. Diversity and Distributions, 13(5), 527–534. http://dx.doi.org/10.1111/j.1472-4642.2007. 00359.x ˜ Dickie, I., Bennett, B., Burrows, L., Nunez, M. A., Peltzer, D. A., Porté, A., et al. (2014). Conflicting values: ecosystem services and invasive tree management. Biological Invasions, 16(3), 705–719. http://dx.doi.org/10.1007/s10530-0130609-6 Donaldson, J. E., Hui, C., Richardson, D. M., Robertson, M. P., Webber, B. L., & Wilson, J. R. U. (2014). Invasion trajectory of alien trees: the role of introduction pathway and planting history. Global Change Biology, 20(5), 1527–1537. http:// dx.doi.org/10.1111/gcb.12486 Escobedo, F., Varela, S., Zhao, M., Wagner, J. E., & Zipperer, W. (2010). Analyzing the efficacy of subtropical urban forests in offsetting carbon emissions from cities. Environmental Science and Policy, 13(5), 362–372. http://dx.doi.org/10.1016/j. envsci.2010.03.009 Forsyth, G. G., Richardson, D. M., Brown, P. J., & van Wilgen, B. W. (2004). A rapid assessment of the invasive status of Eucalyptus species in two South African provinces. South African Journal of Science, 100(1/2), 75–77. Foster, J., & Sandberg, L. A. (2004). Friends or Foe? Invasive species and public green space in Toronto. Geographical Review, 94(2), 178–198. http://dx.doi.org/ 10.1111/j.1931-0846.2004.tb00166.x Foxcroft, L. C., & Freitag-Ronaldson, S. (2007). Seven decades of institutional learning: managing alien plant invasions in the Kruger National Park, South Africa. Oryx, 41(2), 160–167. http://dx.doi.org/10.1017/S0030605307001871 Fuller, R. A., Irvine, K. N., Devine-Wright, P., Warren, P. H., & Gaston, K. J. (2007). Psychological benefits of greenspace increase with biodiversity. Biology Letters, 3(4), 390–394. http://dx.doi.org/10.1098/rsbl.2007.0149 Grice, A. C., Clarkson, J. R., & Calvert, M. (2011). Geographic differentiation of management objectives for invasive species: a case study of Hymenachne amplexicaulis in Australia. Environmental Science and Policy, 14(8), 986–997. http://dx.doi.org/10.1016/j.envsci.2011.07.006 Hobbs, R. J., Higgs, E. S., Hall, C. M., Bridgewater, P., Chapin, F. S., Ellis, E. C., et al. (2014). Managing the whole landscape: historical, hybrid and novel ecosystems. Frontiers in Ecology and the Environment, 12(10), 557–564. http:// dx.doi.org/10.1890/130300 Klotz, S., & Kühn, I. (2010). Urbanisation and alien invasion. In K. J. Gaston (Ed.), Urban ecology (pp. 120–133). Cambridge: Cambridge University Press. Kowarik, I. (2011). Novel urban ecosystems, biodiversity, and conservation. Environmental Pollution, 159(8/9), 1974–1983. http://dx.doi.org/10.1016/j. envpol.2011.02.022 Kühn, I., & Klotz, S. (2006). Urbanization and homogenization—comparing the floras of urban and rural areas in Germany. Biological Conservation, 127(3), 292–300. http://dx.doi.org/10.1016/j.biocon.2005.06.033 Lockwood, J. L., Cassey, P., & Blackburn, T. (2005). The role of propagule pressure in explaining species invasions. Trends in Ecology and Evolution, 20(5), 223–228. http://dx.doi.org/10.1016/j.tree.2005.02.004 Mackenzie, B. F., & Larson, B. M. H. (2010). Participation under time constraints: landowner perceptions of rapid response to the emerald ash borer. Society and Natural Resources, 23(10), 1013–1022. http://dx.doi.org/10.1080/ 08941920903339707 Macdonald, I. A. W., Clark, D. L., & Taylor, H. C. (1989). The history and effects of alien plant control in the Cape of Good Hope Nature Reserve: 1941–1987. South African Journal of Botany, 55, 56–75. McKinney, M. L. (2006). Urbanization as a major cause of biotic homogenization. Biological Conservation, 127(3), 247–260. http://dx.doi.org/10.1016/j.biocon. 2005.09.005 Mgidi, T. N., Le Maître, D. C., Schonegevel, L., Nel, J. L., Rouget, M., & Richardson, D. M. (2007). Alien plant invasions—incorporating emerging invaders in regional prioritization: a pragmatic approach for Southern Africa. Journal of Environmental Management, 84(2), 173–187. http://dx.doi.org/10.1016/j. jenvman.2006.05.018 Moreira-Arce, D., de la Barrera, F., & Bustamante, R. O. (2014). Distance to suburban/wildland border interacts with habitat type for structuring exotic plant communities in a natural area surrounding a metropolitan area in central Chile. Plant Ecology and Diversity, 8(3), 363–370. http://dx.doi.org/10.1080/ 17550874.2014.983201 Nobis, M. P., Jaeger, J. A. G., & Zimmermann, N. E. (2009). Neophyte species richness at the landscape scale under urban sprawl and climate warming. Diversity and Distributions, 15(6), 928–939. http://dx.doi.org/10.1111/j.1472-4642.2009. 00610.x Pickett, S. T. A., Cadenasso, M. L., Grove, J. M., Nilon, C. H., Pouyat, R. V., Zipperer, W. C., et al. (2001). Urban ecological systems: linking terrestrial ecological, physical, and socioeconomic components of metropolitan areas. Annual Review
M. Gaertner et al. / Landscape and Urban Planning 151 (2016) 1–9 of Ecology and Systematics, 32, 127–157. http://dx.doi.org/10.1146/annurev. ecolsys.32.081501.114012 Rebelo, A. G., Holmes, P. M., Dorse, C., & Wood, J. (2011). Impacts of urbanization in a biodiversity hotspot: conservation challenges in Metropolitan Cape Town. South African Journal of Botany, 77(1), 20–35. http://dx.doi.org/10.1016/j.sajb. 2010.04.006 Richardson, D. M., & Brown, P. J. (1986). Invasion of mesic mountain fynbos by Pinus radiata. South African Journal of Botany, 52, 529–536. Richardson, D. M., & Gaertner, M. (2013). Plant invasions as builders and shapers of novel ecosystems. In R. J. Hobbs, E. S. Higgs, & C. M. Hall (Eds.), Novel ecosystems: intervening the new ecological world order (pp. 102–113). Chichester: John Wiley & Sons. Richardson, D. M., & Higgins, S. I. (1998). Pines as invaders in the southern hemisphere. In D. M. Richardson (Ed.), Ecology and biogeography of Pinus (pp. 450–473). Cambridge: Cambridge University Press. Richardson, D. M., Pyˇsek, P., & Carlton, J. T. (2011). A compendium of essential concepts and terminology in invasion ecology. In D. M. Richardson (Ed.), Fifty years of invasion ecology. The legacy of Charles Elton. Oxford: Wiley-Blackwell, pp. 409–435. Richardson, D. M., & van Wilgen, B. W. (2004). Invasive alien plants in South Africa: how well do we understand the ecological impacts? South African Journal of Science, 100(1/2), 45–52. Richardson, D. M., van Wilgen, B. W., Higgins, S. I., Trinder-Smith, T. H., Cowling, R. M., & McKelly, D. H. (1996). Current and future threats to biodiversity on the Cape Peninsula. Biodiversity and Conservation, 5(5) http://dx.doi.org/10.1007/ BF00137612 Rittel, H. W. J., & Webber, M. M. (1973). Dilemmas in a general theory of planning. Policy Sciences, 4, 155–169. Trentanovi, G., von der Lippe, M., Sitzia, T., Ziechmann, U., Kowarik, I., & Cierjacks, A. (2013). Biotic homogenization at the community scale: disentangling the roles of urbanization and plant invasion. Diversity and Distributions, 19(7), 738–748. http://dx.doi.org/10.1111/ddi.12028
9
van Wilgen, B. W. (2012). Evidence, perceptions, and trade-offs associated with invasive plant control in Table Mountain National Park, South Africa. Ecology and Society, 17(2), 23. http://dx.doi.org/10.5751/ES-04590-170223 van Wilgen, B. W., Reyers, B., Le Maitre, D. C., Richardson, D. M., & Schonegevel, L. (2008). A biome-scale assessment of the impact of invasive plants on ecosystem services in South Africa. Journal of Environmental Management, 89(4), 336–349. http://dx.doi.org/10.1016/j.jenvman.2007.06.015 van Wilgen, N. J., & Richardson, D. M. (2012). The roles of climate, phylogenetic relatedness, introduction effort, and reproductive traits in the establishment of non-native reptiles and amphibians. Conservation Biology, 26(2), 267–277. http://dx.doi.org/10.1111/j.1523-1739.2011.01804.x van Wilgen, B. W., & Scott, D. F. (2001). Managing fires on the Cape Peninsula: dealing with the inevitable. Journal of Mediterranean Ecology, 2, 197–208. van Wilgen, B. W., & Wannenburgh, A. (2016). Co-facilitating invasive species control, water conservation and poverty relief: achievements and challenges in South Africa’s Working for Water programme. Current Opinion in Environmental Sustainability, 19, 7–17. http://dx.doi.org/10.1016/j.cosust.2015.08.012 Vilà, M., Basnou, C., Pyˇsek, P., Josefsson, M., Genovesi, P., Gollash, S., et al. (2009). How well do we understand the impacts of alien species on ecosystem services? A pan-European, cross-taxa assessment. Frontiers in Ecology and the Environment, 8(3), 135–144. http://dx.doi.org/10.1890/080083 Warren, C. R. (2007). Perspectives on the ‘alien’ versus ‘native’ species debate: a critique of concepts, language and practice. Progress in Human Geography, 31(4), 427–446. http://dx.doi.org/10.1177/0309132507079499 Wilson, J. R. U., Ivey, P., Manyama, P., & Nänni, I. (2013). A new national unit for invasive species detection, assessment and eradication planning. South African Journal of Science, 109(5/6), 01–13. http://dx.doi.org/10.1590/sajs.2013/ 20120111 Zisenis, M. (2015). Alien plant species: a real fear for urban ecosystems in Europe? Urban Ecosystems, 18(2), 355–370. http://dx.doi.org/10.1007/s11252-0140400-1
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Landscape and Urban Planning journal homepage: www.elsevier.com/locate/landurbplan
Perspective essay
Managing invasive species in cities: A framework from Cape Town, South Africa Mirijam Gaertner a,b,∗ , Brendon M.H. Larson a,c , Ulrike M. Irlich a,b , Patricia M. Holmes a,b , Louise Stafford b , Brian W. van Wilgen a , David M. Richardson a a
Centre for Invasion Biology, Department of Botany and Zoology, Stellenbosch University, Private Bag X1, Matieland 7602, South Africa Environmental Resource Management Department (ERMD), City of Cape Town, Westlake Conservation Office, Ou Kaapse Weg, Tokai, 7966, Cape Town, South Africa c School of Environment, Resources and Sustainability, University of Waterloo, Waterloo, Ontario N2L3G1, Canada b
h i g h l i g h t s • • • •
Existing invasive species management frameworks are inadequate in urban areas. Urban stakeholders often hold conflicting views and are critical of management. Divergent stakeholder perceptions need to be considered explicitly and transparently. Urban management frameworks should allow for acceptance of some invasive species.
a r t i c l e
i n f o
Article history: Received 24 June 2015 Received in revised form 11 March 2016 Accepted 17 March 2016 Available online 28 March 2016 Keywords: Biological invasions Non-native species Urban invasive species management Table Mountain National Park Urban ecology
a b s t r a c t Invasive non-native species are often more prevalent in cities than in rural areas because of numerous environmental disturbances and higher propagule pressure. Attempts to manage invasive species in cities are often controversial because of the diversity of stakeholder views. Until now, however, environmental managers in cities have managed invasive species using approaches and paradigms developed for a rural context, despite the radically different socio-environmental conditions that prevail in cities. We examine the case of Cape Town, South Africa, a rapidly growing metropolitan centre within a global biodiversity hotspot and a developing country, to underline the considerable challenges and complexities of managing invasive species in cities. We argue that traditional management approaches need to be supplemented by consideration of stakeholder views and the social consequences of management actions. We present a framework for selecting appropriate goals for the management of invasive species, ranging from eradication to acceptance. © 2016 Elsevier B.V. All rights reserved.
1. Introduction Invasive non-native species (sensu Richardson, Pyˇsek, & Carlton, 2011; hereafter ‘invasive species’) are often abundant in cities (Kowarik, 2011). Cities contain a high density of people and they are hubs of human-mediated movement of commodities. Transport linkages (e.g., airports and harbours) facilitate the introduction
∗ Corresponding author at: Centre of Excellence for Invasion Biology, Department of Botany and Zoology, Natural Sciences Building, Private Bag X1, University of Stellenbosch, Matieland 7602, South Africa. E-mail addresses: [email protected] (M. Gaertner), [email protected] (B.M.H. Larson), [email protected] (U.M. Irlich), [email protected] (P.M. Holmes), [email protected] (L. Stafford), [email protected] (B.W. van Wilgen), [email protected] (D.M. Richardson). http://dx.doi.org/10.1016/j.landurbplan.2016.03.010 0169-2046/© 2016 Elsevier B.V. All rights reserved.
and dissemination of non-native species through dispersal pathways such as trade, tourism, and horticulture (Dehnen-Schmutz, Touza, Perrings, & Williamson, 2007); such activities release high numbers of individuals into a region (high ‘propagule pressure,’ see Lockwood, Cassey, & Blackburn, 2005). In cities these non-native species encounter habitats, soils, climatic conditions and hydrology that have been profoundly changed by human activity and that can promote their spread if they are pre-adapted to similar conditions in their region of origin (Pickett et al., 2001; Kowarik, 2011). Urban heat-island effects, for example, may facilitate the spread of invasive species (Nobis, Jaeger, & Zimmermann, 2009). Also, typical urban conditions such as fragmented habitats and altered disturbance regimes often favour non-native species (Cilliers, Williams, & Barnard, 2008; Zisenis, 2015). In Central Europe, most urbanophilic non-native plant species can be found in inner city areas, which
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provide suitable conditions for plant species that tolerate or even thrive when temperatures are warmer and disturbances more frequent (Klotz & Kühn, 2010). In short, many non-native species are more prolific in cities because long histories of human dispersal, disturbance and habitat modification enhance their opportunities for establishment, proliferation and spread. Invasions pose well-documented risks in both natural and seminatural habitats and in both protected areas and public open spaces. Invasive species may displace native species and contribute to homogenization of habitats within cities (Burton, Samuelson, & Pan, 2005; Kühn & Klotz, 2006; McKinney, 2006; Trentanovi et al., 2013). Ornamental invasive plant species in gardens act as significant sources of non-native propagules (Alston & Richardson, 2006; Bowers, Bean, & Turner, 2006), and non-native animals kept as pets can establish and become invasive (van Wilgen & Richardson, 2012). Invasive species in cities may also disrupt important ecosystem services such as water filtration, flood attenuation and coastal protection. Invasive plants can clog streams and canals, resulting in flooding, and they may also increase fire severity and soil erosion risk (van Wilgen & Scott, 2001). Although some invasive species were introduced to provide particular ecosystem services, such as trees for timber production or erosion control, their subsequent spread and proliferation may ultimately have a net detrimental effect (e.g. van Wilgen, Reyers, Le Maitre, Richardson, & Schonegevel, 2008; Vilà et al., 2009). Invasive species may also affect urban biodiversity in ways that reduce human well-being more directly, for example by changing the aesthetics of the environment (Fuller, Irvine, Devine-Wright, Warren, & Gaston, 2007; Kowarik, 2011). For all of these reasons, policies are in place in many parts of the world to manage invasive species in parks, public areas and other urban zones. Despite growing concerns about invasive species, some authors argue that we may, in some instances, have to learn to “cohabitate” with them in cities (e.g., Foster & Sandberg, 2004). Even if an invasive species disrupts an ecosystem service, it may also be beneficial. Studies in urban forests in Florida, for example, have shown that invasive trees were most successful in sequestering CO2 (Escobedo, Varela, Zhao, Wagner, & Zipperer, 2010). As another example, studies from northern parts of South Africa show that invasive Eucalyptus trees are used extensively as roosting sites for the vulnerable Lesser kestrel (Falco naumannii) and as breeding sites for the African fish eagle (Haliaeetus vocifer) (Cilliers & Siebert, 2012). Managing invasive species in urban areas is sometimes further complicated by the cultural connections that people have forged with them. Just like native species, invasive species can become associated with a place and be regarded as culturally important by some city inhabitants (Warren, 2007). Stakeholders in cities often have strongly divergent views about the impacts and benefits of particular invasive species, so conflicts over the management of invasive species are emerging (e.g., trees, Dickie et al., 2014). In particular, invasive species may provide provisioning ecosystem services, but at the expense of various elements of biodiversity, which can lead to conflicts over whether to manage for the former or the latter. Such conflicts exemplify the extent to which invasive species management, especially in human-dominated areas, is increasingly viewed as a “wicked problem” (Rittel & Webber, 1973) because sometimes there are no straightforward “win-win” solutions. A particular problem is that even if an invasive species is “accepted” (i.e., not regulated or targeted for containment or control) within a city environment, such acceptance may pave the way for its spread into adjacent rural landscapes where it may have substantial negative impacts (Botham et al., 2009; Moreira-Arce, de la Barrera, & Bustamante, 2014). In this paper, we explore the challenges and complexities of managing invasive species in cities by examining the exemplary case of Cape Town, South Africa. Cape Town is a rich case study
for elucidating the complexity of managing invasive species in cities because it highlights several interwoven social and ecological dimensions. We use this city to highlight challenges that will be faced by an increasing number of cities given ongoing urbanization and growing human populations. Our objectives are to utilize this case study: (1) to review the challenges faced by managers who seek to control invasive species in an urban environment; and (2) to develop a framework to assist environmental managers globally as they seek to integrate a range of management options for invasive species in urban systems to deal with diverse and often conflicting views of what is appropriate.
2. The case of Cape Town, South Africa Cape Town’s conservation significance derives from its location in the Cape Floristic Region, a global centre of plant endemism (Cowling, Rundel, Lamont, Arroyo, & Arianoutsou, 1996). The city (2445 km2 ) includes Table Mountain National Park (221 km2 ), as well as 17 smaller nature reserves and 500 biodiversity network sites that together cover 270 km2 . It has a population of 3.8 million people and is growing more rapidly than any other southern African metropolis on a per capita basis (Boraine et al., 2006), especially within its poorer suburbs (“townships”) which have experienced an influx of mainly Black citizens following the collapse of apartheid. At present, 26% of Cape Town is urban, 35% is agricultural, and 39% is natural and semi-natural vegetation concentrated in mountainous areas (mainly within Table Mountain National Park) (Fig. 1). Many lowland areas have been transformed, with remnants being highly threatened and thus having become a priority for conservation (Rebelo, Holmes, Dorse, & Wood, 2011). Cape Town remains a focal point of the national economy and international tourism, and thus faces ever-growing needs for housing, transport networks and trade. The fact that Cape Town has high levels of human population growth, unemployment and crime adds dimensions of complexity. Nonetheless, many of Cape Town’s citizens are involved in conservation initiatives, such as “Friends” groups for nature reserves and various conservation stewardship and citizen science initiatives (e.g., spotter networks for emerging invasive species, see http://www.capetowninvasives.org.za). Cape Town has a long history of European colonization and the associated introductions of non-native species present a significant challenge to people and landscapes (van Wilgen, 2012) (Fig. 2). For example, invasive tree species such as pines (Pinus species), grown in plantations, and Australian wattles (Acacia species), planted mainly along the coast for dune stabilization, have spread widely into natural vegetation. Aquatic invasive species such as Water hyacinth (Eichhornia crassipes) block waterways and affect water quality (Richardson & van Wilgen, 2004). The trade in ornamental plants and pets, and other enterprises that rely on non-native taxa, continue to introduce new species into the city; many of these remain undetected and/or unregulated. Some invasive plant species pose serious risks to humans; for example, invasive pines and wattles increase the severity of wild fires near residential areas (Fig. 1) (van Wilgen & Scott, 2001). The Department of Environmental Affairs is responsible for the overall administration of the National Environmental Management: Biodiversity Act (NEMBA), which places obligations on all landowners and all organs of state, including the City of Cape Town, with regard to the management of invasive species. Invasive species control programs date back to the 1940s (Macdonald, Clark, & Taylor, 1989), and in 2008 the city established an Invasive Species Management Unit, with an annual budget of about 84,000 USD and one semi-skilled team. Since then the program has grown to include areas managed by multiple departments within the city.
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Fig. 1. The City of Cape Town is situated around Table Mountain National Park. Lowland areas (red polygon) are transformed, and remnants of semi-natural vegetation are highly threatened and a priority for conservation. The spread of ornamental plants across the city threatens biodiversity in the national park. Photo: Bruce Sutherland, City of Cape Town. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)
Although originally funded by the city’s ratepayers, other sources (notably the government-sponsored Working for Water programme which aims to provide employment to poor people, see van Wilgen & Wannenburgh, 2016) have since provided additional and substantial funding, increasing the available funds to 1.9 million USD in 2014. The city now manages 55 teams averaging 10 workers each; their mandates include early detection and rapid response programs for emerging invasive species and the control of
more established ones. Besides these control operations, the city is piloting the promotion of indigenous plant species over non-native ones to provide the same benefits and reduce negative perceptions of control operations. The city spends 70% of its invasive species funds on controlling terrestrial invasive plants in nature reserves, targeting Australian Acacia species (e.g., Acacia cyclops, Acacia mearnsii, and Acacia saligna), Australian myrtle (Leptospermum laevigatum), and pines
Fig. 2. Examples where the management of invasive plant species in the City of Cape Town has been challenging and contentious: (a) Utilization of invasive Acacia cyclops stands for firewood; b) Invasive squirrels are being ‘protected’ by municipalities; Photo: Sophia Turner; (c) Clearing waterways choked with water hyacinth with heavy machinery; (d) Tuart trees have been planted along many roads to provide shade; Photo: Emma Booysen Photography; (e) Invasive trees as fire hazard, Photo: Chris Cloete; (f) Himalayan thars on Table Mountain; Photo: Tony Rebelo; (g) Recreational activities under pine plantations; Photo: Lisel Mc Gregor.
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(Pinus halepensis, Pinus pinaster and Pinus radiata). The bulk of the budget for aquatic weeds is spent on Water hyacinth and Mexican water lily (Nymphaea mexicana). Priority is given to nature reserves, followed by areas where invasive plants are a fire hazard or provide shelter for criminals. Areas that have been cleared of non-native plants are prioritized for follow-up operations to prevent re-invasion. The objective is to restore native biodiversity where possible or to at least ensure adequate levels of ecosystem functioning through either active or passive restoration. Other organizations involved in control of non-native species are South African National Parks (e.g. Foxcroft & Freitag-Ronaldson, 2007, the South African National Biodiversity Institute (Wilson, Ivey, Manyama, & Nänni, 2013) and the Working for Water Programme (van Wilgen & Wannenburgh, 2016). Control efforts are often compromised because of significant conflicts over the management of invasive species. These conflicts are particularly severe in a city environment because of a diversity of stakeholder views. For example, conservationists and citizen science groups support the control of invasive species (mainly to protect biodiversity and to decrease negative impacts on ecosystem services), but control operations are met with strong resistance from commercial foresters, in the case of pine plantations, and also from citizens who have concerns over loss of forested space in urban environments (van Wilgen, 2012). 3. A general framework for invasive species management in cities Management of invasive species worldwide is usually directed towards one or more goals. These typically include the prevention of spread to non-invaded areas, local eradication of isolated populations, containment where local eradication is not possible, and the protection of selected assets or areas where the invasive species are so widespread that containment is no longer feasible (Grice, Clarkson, & Calvert, 2011; Auld & Johnson, 2014). Such goals are simpler to incorporate in decision-making protocols in protected areas and rural situations than in cities where divergent stakeholder views and resulting conflicts of interest (for example relating to biodiversity conservation and the provision of ecosystem services) often complicate decisions. Managers of invasive species in cities need a framework for making decisions on how to deal with individual species in a way that will lead to acceptable and effective control. In some cases, it may be acceptable (or justifiable) to proceed with control, while in others it may be better to tolerate a species. Whether a given species should be tolerated or not is highly context-specific and depends on its potential for causing significant impact (now and in the future) and on its perceived value. It is useful to envisage the positioning of species along two axes: the first considers the real or perceived benefits associated with the species, while the other considers its potential to generate significant negative impacts (Fig. 3). This provides a framework in which species can be positioned along these two axes, giving rise to three categories. (1) Species with a low potential for negative impact, regardless of the benefits they deliver, would fall into a “tolerance” category. These species would be generally acceptable to society and to conservationists, requiring very little management. For some species no action will be needed (those with high perceived benefits and low perceived impacts), but some may need to be monitored (species with low perceived impacts, but that may become more problematic with time, e.g., emerging weeds). (2) Species that both provide benefits and have negative impacts fall into an “active engagement” category. Their management
will be more complex because the optimum course of action is less clear, and any option is likely to be controversial. Here the appropriate course of action would depend both on the societal response and the political consequences of deciding to control or tolerate the species. In some cases, the ongoing presence of harmful species might have to be accepted (where there is strong societal opposition to control and/or insufficient evidence to demonstrate widespread impacts or justify expensive intervention). (3) Species with a relatively high potential for impact and that deliver relatively low benefits would fall into a “control priority” category. The “control priority” category will include species to be eradicated (where eradication is possible) or controlled (where eradication is not feasible, e.g., widespread plant species with a long-lived seedbank). Depending on the circumstances, control could aim to reduce the extent of the invasion and the density of populations, prevent further expansion, or (in cases where the species is very widespread) focus on protecting particular assets (e.g., where a species poses a high risk in terms of fire hazard or to endemic species or ecosystems).
Determining the real (or perceived) benefits and negative impacts of a species can be difficult to quantify and it is often contentious. While some information can be obtained from the literature, or through research if funding and time allows, the full range of understanding and perceptions need to be explored through stakeholder engagement. It is challenging to design appropriate stakeholder engagement that develops inter-party trust, not least because this is time-consuming, and time may not be available in the context of a ‘rapid-response’ program (see Mackenzie & Larson, 2010). For assigning species to the management categories presented in Fig. 3, we propose a simple decision tree (Fig. 4). Unlike traditional approaches to invasive species management, our proposed decision tree explicitly considers the potential societal reactions to invasive species control operations in cities, which would become apparent in consultations with affected local stakeholders. In our experience (as elsewhere, e.g., Dickie et al., 2014), societal opposition to the control of particular types of invasive species will at best add substantial costs and delays to control projects, and at worst prevent them from proceeding at all (and potentially affect management initiatives on other invasive species). It is thus clearly prudent to anticipate and plan for such possibilities, and to accept that, in some cases, traditional goals need to be replaced with partial or complete tolerance of invasive species. We note, however, that there are very real uncertainties about both the ecological and social status of many species, especially in the case of emerging invasive species, which may necessitate research to facilitate rapid decision-making (Mackenzie & Larson, 2010). Rapid decision-making to deal with emerging invaders is crucial; initiation of management efforts at an early stage of invasion is the most cost-efficient (Mgidi et al., 2007). Another advantage of managing species at an early stage of invasion is that cultural connections with the non-native species have not yet been established and opposition to management is likely to be more limited. The decision of whether or not, and if so where and how, to manage invasive species, is also influenced by factors unrelated to impacts or benefits. Even for damaging invasive species, if no effective control methods exist it may be prudent to tolerate the species, and to deploy resources on other species that can more effectively be controlled (Hobbs et al., 2014). In degraded areas, where the costs of rehabilitation would exceed the benefits of control, an invasive species may also be tolerated, although steps should be taken to prevent spread to less degraded areas.
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Fig. 3. Proposed categorization of invasive species according to perceived benefits and potential negative effects (invasiveness, ecosystem and social impacts). Three management approaches are recognized: (1) tolerance; (2) active engagement; and (3) control priority. Shading reflects the probability of conflicts of interest (light grey to dark grey = low to high probability).
Experience with invasive species in Cape Town allows us to use our decision tree (Fig. 4) to categorize species from a wide range of taxa within this framework (Table 1; Fig. 3). We describe several examples below, with the ‘steps’ of the decision tree (Fig. 4) included in brackets. 4. Application of the framework, with examples from Cape Town We begin with several examples from the genus Eucalyptus. Karri gum (Eucalyptus diversicolor), for example, occurs in plantations in Table Mountain National Park that were planted over a century ago and have grown to an impressive size. Karri gum plantations are popular with hikers, cyclists and tree enthusiasts (van Wilgen, 2012) (perceived benefits are high). Although these gum trees impact water resources, they are not highly invasive (impact low). Furthermore, proposals to clear Karri gum plantations and to restore those areas to fynbos shrublands were vigorously opposed. Because of the political ramifications of going against public opinion, especially in the case of a species that is not aggressively invasive, it was decided to tolerate and monitor the species and to retain the plantations (monitor) (van Wilgen, 2012). Similarly, Tuart (E. gomphocephala), which is endangered in its native Western Australia, has been widely planted in Cape Town’s poorer suburbs and along many roads where it provides shade and does not spread. This species has and should be tolerated. River red gum (E. camaldulensis), on the other hand, is both attractive (high benefit) and highly invasive (high impact), especially along river courses (Forsyth, Richardson, Brown, & van Wilgen, 2004), which suggests that there will be social opposition to clearing it (active engagement). An appropriate response here would be to tolerate large specimens in parks and gardens, but to remove plants from protected areas and river courses, and to prohibit further sale and planting anywhere. Here our decision tree allows for “partial tolerance” as a possible outcome. Pine trees introduced from Europe and North America also can be classified using the framework. Cluster pine (P. pinaster) and Aleppo pine (P. halepensis) have negligible commercial value and
relatively little aesthetic appeal to local people (low benefit). They impact water resources and biodiversity, and increase fire severity (Richardson & Higgins, 1998) (high impact). Furthermore, effective control methods exist, which places them in the control priority category (impact reduction and containment). In contrast, stone pines (P. pinea) are often viewed as “heritage trees” on Table Mountain and elsewhere (van Wilgen, 2012), (high perceived benefit). The species is widespread, but not highly invasive (low impact) and should hence be monitored and tolerated where appropriate. Monterey pine (P. radiata), on the other hand, occurs in commercial plantations that are popular for the same reasons as karri plantations (high perceived benefit), but they are highly invasive (Richardson & Brown, 1986) and pose a substantial threat to the biodiversity of Table Mountain National Park (Richardson et al., 1996) (high impact). Monterey pines thus require active engagement, and there are plans to systematically remove them from plantations and protected areas. Because there has been public resistance to these removals (van Wilgen, 2012), it is important to actively engage with the public in stakeholder workshops about the proposed removal (and reasons for it) and a potential plan for replacing them. In Cape Town, opposition to the removal of pine plantations was countered by support from the national and provincial conservation agencies, NGOs such as WWF (South Africa), and leading academics (equal opposition and support). In addition, the control and rehabilitation operations provided opportunities for employment, which was politically attractive and could offset jobs lost in forestry-related activities (political consequences). Control should therefore be continued and awareness for the species impacts is being raised. Attempts to control most Australian wattles (Acacia species) have not invoked the same level of negative reaction as pines and eucalypts (low perceived benefit), and most species are known for their high impacts. Acacia species can hence be placed within the category of control priority. An exception is Pepper tree wattle (Acacia elata), which is valued as a shade and amenity tree, especially of golf courses; management of this invasive species requires active engagement (Donaldson et al., 2014).
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Fig. 4. Assigning species to management categories within the boundaries of the City of Cape Town. Tolerance category (highlighted green), control priority category (highlighted orange) and active engagement category (highlighted blue) are described in detail in the text. The position of a selection of invasive species in the three “management approach” zones is shown in Fig. 3. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)
Some non-native aquatic plant species, such as Water hyacinth, are highly invasive and have significant negative impacts but no benefits (Richardson & van Wilgen, 2004), so they are placed into the control priority category. Managing damaging invasive animals is often controversial in cities, where residents challenge the ethics of killing or removing animals, and point to the perceived cruelty of these operations. For example, a decision was taken to attempt to eradicate an isolated but thriving population of introduced Himalayan thars (Hemitragus jemlahicus) on Table Mountain, because of the damage they did through over-grazing (high impact), and to make the re-introduction of historically extirpated Klipspringer antelope (Oreotragus oreotragus) possible. Public reaction was substantial, and the campaign had to proceed in the face of strong resistance (active engagement category). However, the campaign received strong backing from conservation authorities and NGOs, making it feasible to continue (equal opposition and support). A similar campaign to remove European mallards (Anas platyrhynchos), which interbreed with the indigenous Yellow-billed duck (Anas undulata) (high impact), was also met with resistance. Management efforts were effectively halted because the arguments for the campaign (genetic contamination of a single indigenous species) were less convincing to the public than arguments for the widespread ecolog-
ical impacts of more damaging invasive species (strong opposition and relatively low perceived impact). Similarly, any attempt to control Eastern grey squirrels (Sciurus carolinensis) to reduce limited impacts on biodiversity would almost certainly meet with strong resistance from some quarters. Interestingly, a campaign to remove a large and growing population of Indian house crows (Corvus splendens) met with little resistance, and has proceeded without substantial opposition. Several lessons have been learned from managing invasive animal and plant species in the city of Cape Town. Experience has shown that we need to use our framework to decide on a course of action when any new invasive species becomes a potential target for control. In order to make logical and defensible choices, it would further be essential to gauge the views of stakeholders. Where significant numbers of people are likely to oppose management, the potential political consequences of different courses of action, must be explicitly considered. It also has to be accepted that conventional goals (such as eradication or reductions in populations) will need to be replaced with partial or complete tolerance of invasive species in some cases. A potential critique of our framework is that ‘perceived impacts’ and ‘perceived benefits’, the core steps in our framework, are context specific, difficult to quantify, and subjective. Such con-
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Table 1 Arguments for and against the control of selected non-native species mentioned in the text, and the rationale for their placement into management categories (Fig. 3). PA = protected area (including Cape Towns’ critical biodiversity areas and the Biodiversity Network). Species
Characteristics
Negative effects
Positive effects
Management solution
Acacia cyclops (Red eye)
Tree or shrub, ecosystem transformer
Impacts on biodiversity, water, sand movement and fire intensity
Important source of firewood
Control priority, as impacts outweigh benefits. Remove from PAs. Contain elsewhere.
Acacia elata (Peppertree wattle)
Tree, ecosystem transformer
Impacts on biodiversity
Aesthetic value
Active engagement needed to garner support for control.
Acacia mearnsii (Black wattle)
Tree, ecosystem transformer
Impacts on biodiversity and water
Source of firewood
Control priority, as impacts outweigh benefits. Remove from PAs. Contain elsewhere.
Acacia paradoxa (Kangaroo thorn)
Shrub, potential ecosystem transformer
Impacts on biodiversity
None
Control priority. Distribution restricted, so target for eradication.
Acacia saligna (Port Jackson willow)
Tree, ecosystem transformer
Impacts on biodiversity and water
Source of firewood and fodder
Control priority, as impacts outweigh benefits. Remove from PAs. Contain elsewhere.
Ailanthus altissima (Tree of heaven)
Tree, ecosystem transformer
Impacts on urban infrastructure and potentially on biodiversity
Aesthetic value
Active engagement needed to garner support for control.
Eichhornia crassipes (Water hyacinth)
Aquatic weed, ecosystem transformer
Impacts on water quality, water distribution networks and biodiversity
Aesthetic appeal (limited)
Control priority. Remove from water bodies.
Eucalyptus camaldulensis (River red gum)
Tree, ecosystem transformer
Impacts on biodiversity and water
Heritage tree status, aesthetic value
Active engagement needed to garner support for control along rivers. Tolerate elsewhere, but phase out over time.
Eucalyptus diversicolor (Karri gum)
Tree, naturalized
Impacts on biodiversity and water
Heritage tree status, aesthetic value
Species to be tolerated, as not aggressively invasive.
Eucalyptus gomphocephala (Tuart)
Tree, naturalized
Impacts on groundwater
Aesthetic value, shade in townships
Species to be tolerated, as not aggressively invasive.
Nymphaea mexicana (Mexican water lily)
Aquatic plant, invasive
Potential impacts on aquatic biodiversity
Aesthetic appeal
Control priority. Remove from water bodies.
Pinus halepensis (Aleppo pine)
Tree, ecosystem transformer
Impacts on biodiversity, water and fire intensity
Previously valued for timber
Control priority, as impacts outweigh benefits. Remove from PAs. Contain elsewhere.
Pinus pinaster (Cluster pine)
Tree, ecosystem transformer
Impacts on biodiversity, water and fire intensity
Previously valued for timber
Control priority, as impacts outweigh benefits. Remove from PAs. Contain elsewhere.
Pinus pinea (Stone pine)
Tree, naturalized
Not native to protected areas, local biodiversity impacts
Heritage tree status, aesthetic value
Species to be tolerated, as not aggressively invasive. Limited and ongoing removal from PAs.
Pinus radiata (Monterey pine)
Tree, ecosystem transformer
Impacts on biodiversity, water and fire intensity
Source of timber, aesthetic value
Active engagement needed to garner support for removal of plantations. Ongoing removal from PAs.
Sambucus canadensis (Canadian elder)
Shrub, invasive
Potential impacts on biodiversity
Popular garden plant
Species to be tolerated and monitored until effective control methods can be found.
Anas platyrhynchos (European mallard)
Aquatic bird, naturalized
Hybridization with native ducks
Ethical (animal rights), aesthetic
Species to be tolerated, as perceived benefits outweigh impacts.
Corvus splendens (Indian house crow)
Bird, invasive
Impacts on biodiversity and health risks; nuisance value
Ethical (animal rights)
Control priority, as impacts outweigh benefits. Target for local eradication.
Hemitragus jemlahicus (Himalayan thar)
Large mammalian herbivore, invasive
Impacts on biodiversity
Ethical (species threatened in native range; animal rights)
Active engagement needed to offset opposition to control. Target for eradication.
Sciurus carolinensis (Eastern grey squirrel)
Rodent, naturalized (dependent on non-native trees)
Minor impacts on biodiversity
Ethical (animal rights), aesthetic (cuteness)
Species to be tolerated, as perceived benefits outweigh impacts.
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cerns should be openly debated as part of stakeholder engagement to ensure that all parties agree on such evaluations.
5. Conclusions Managing invasive species in cities provides additional challenges to those traditionally faced in agricultural areas or natural ecosystems, so existing frameworks for guiding management are increasingly being found inadequate in this context. The problem becomes acute where urban areas abut or enclose regions of high conservation value. Cape Town is an exemplar, so we draw upon insights from recent attempts to manage different types of invasive species in this city to highlight the need for a new framework to guide invasive species management in cities more generally. Two fundamental recommendations emerge from our analysis. First, divergent perceptions need to be given explicit and transparent consideration. Most people do not fully understand issues related to biodiversity and conservation, or the negative economic impacts of invasive species, but a platform for airing their views must be provided. Experience has shown that failure to do so often results in resistance that may derail long-term plans (see Mackenzie & Larson, 2010 and references therein). Second, we may need to tolerate some widespread invasive species within cities for a combination of social and pragmatic reasons. This acknowledgement parallels the recognition that some ecosystems have been so transformed that they cannot practically be restored to their historic condition (Richardson & Gaertner, 2013; Hobbs et al., 2014). However, we caution that where some form of cohabitation is agreed upon, approaches should be sought for minimizing potential negative impacts of retaining areas with invasive species. Potential pro-active avenues to consider include biological control, ecological restoration and gradual replacement with alternative non-invasive species. Our framework provides a workable solution to the failures of implementing rural management approaches to invasive species within urban areas.
Acknowledgements We acknowledge funding from the DST-NRF Centre of Excellence for Invasion Biology (M.G., B.L., B.v.W. and D.M.R), a Standard Research Grant from the Social Science and Humanities Research Council of Canada (B.L.), and the National Research Foundation, South Africa (D.M.R.; grant 85417; B.W.v.W.; grant 87550). This work was also supported by the National Socio-Environmental Synthesis Center (SESYNC) under funding received from the National Science Foundation DBI-1052875, and by sDiv, the Synthesis Centre of the German Centre for Integrative Biodiversity Research (iDiv) Halle-Jena-Leipzig (German Research Foundation DFG FZT 118).
References Alston, K. P., & Richardson, D. M. (2006). The roles of habitat features, disturbance, and distance from putative source populations in structuring alien plant invasions at the urban/wildland interface on the Cape Peninsula, South Africa. Biological Conservation, 132(2), 183–198. http://dx.doi.org/10.1016/j.biocon. 2006.03.023 Auld, B. A., & Johnson, S. B. (2014). Invasive alien plant management. CAB Review, 9(37), 1–12. http://dx.doi.org/10.1079/PAVSNNR20149037 Boraine, A., Crankshaw, O., Engelbrecht, C., Gotz, G., Mbanga, S., Narsoo, N., et al. (2006). The state of South African cities a decade after democracy. Urban Studies, 43(2), 259–284. http://dx.doi.org/10.1080/00420980500416990 Botham, M. S., Rothery, P., Hulme, P. E., Hill, M. O., Preston, C. D., & Roy, D. B. (2009). Do urban areas act as foci for the spread of alien plant species? An assessment of temporal trends in the UK. Diversity and Distributions, 15(2), 338–345. http://dx.doi.org/10.1111/j.1472-4642.2008.00539.x Bowers, J. E., Bean, T. M., & Turner, R. M. (2006). Two decades of change in distribution of exotic plants at the desert laboratory, Tucson. Arizona. Madro˜ no, 53(3), 252–263. http://www.jstor.org/stable/41425651
Burton, M., Samuelson, L., & Pan, S. (2005). Riparian woody plant diversity and forest structure along an urban-rural gradient. Urban Ecosystems, 8(1), 93–106. http://dx.doi.org/10.1007/s11252-005-1421-6 Cilliers, S., Williams, N. G., & Barnard, F. (2008). Patterns of exotic plant invasions in fragmented urban and rural grasslands across continents. Landscape Ecology, 23(10), 1243–1256. http://dx.doi.org/10.1007/s10980-008-9295-7 Cilliers, S. S., & Siebert, S. (2012). Urban ecology in cape town: south african comparisons and reflections. Ecology and Society, 17(3), 33. http://dx.doi.org/ 10.5751/ES-05146-170333 Cowling, R. M., Rundel, P. W., Lamont, B. B., Arroyo, M. K., & Arianoutsou, M. (1996). Plant diversity in Mediterranean-climate regions. Trends in Ecology and Evolution, 11(9), 362–366. http://dx.doi.org/10.1016/0169-5347(96)10044-6 Dehnen-Schmutz, K., Touza, J., Perrings, C., & Williamson, M. (2007). A century of the ornamental plant trade and its impact on invasion success. Diversity and Distributions, 13(5), 527–534. http://dx.doi.org/10.1111/j.1472-4642.2007. 00359.x ˜ Dickie, I., Bennett, B., Burrows, L., Nunez, M. A., Peltzer, D. A., Porté, A., et al. (2014). Conflicting values: ecosystem services and invasive tree management. Biological Invasions, 16(3), 705–719. http://dx.doi.org/10.1007/s10530-0130609-6 Donaldson, J. E., Hui, C., Richardson, D. M., Robertson, M. P., Webber, B. L., & Wilson, J. R. U. (2014). Invasion trajectory of alien trees: the role of introduction pathway and planting history. Global Change Biology, 20(5), 1527–1537. http:// dx.doi.org/10.1111/gcb.12486 Escobedo, F., Varela, S., Zhao, M., Wagner, J. E., & Zipperer, W. (2010). Analyzing the efficacy of subtropical urban forests in offsetting carbon emissions from cities. Environmental Science and Policy, 13(5), 362–372. http://dx.doi.org/10.1016/j. envsci.2010.03.009 Forsyth, G. G., Richardson, D. M., Brown, P. J., & van Wilgen, B. W. (2004). A rapid assessment of the invasive status of Eucalyptus species in two South African provinces. South African Journal of Science, 100(1/2), 75–77. Foster, J., & Sandberg, L. A. (2004). Friends or Foe? Invasive species and public green space in Toronto. Geographical Review, 94(2), 178–198. http://dx.doi.org/ 10.1111/j.1931-0846.2004.tb00166.x Foxcroft, L. C., & Freitag-Ronaldson, S. (2007). Seven decades of institutional learning: managing alien plant invasions in the Kruger National Park, South Africa. Oryx, 41(2), 160–167. http://dx.doi.org/10.1017/S0030605307001871 Fuller, R. A., Irvine, K. N., Devine-Wright, P., Warren, P. H., & Gaston, K. J. (2007). Psychological benefits of greenspace increase with biodiversity. Biology Letters, 3(4), 390–394. http://dx.doi.org/10.1098/rsbl.2007.0149 Grice, A. C., Clarkson, J. R., & Calvert, M. (2011). Geographic differentiation of management objectives for invasive species: a case study of Hymenachne amplexicaulis in Australia. Environmental Science and Policy, 14(8), 986–997. http://dx.doi.org/10.1016/j.envsci.2011.07.006 Hobbs, R. J., Higgs, E. S., Hall, C. M., Bridgewater, P., Chapin, F. S., Ellis, E. C., et al. (2014). Managing the whole landscape: historical, hybrid and novel ecosystems. Frontiers in Ecology and the Environment, 12(10), 557–564. http:// dx.doi.org/10.1890/130300 Klotz, S., & Kühn, I. (2010). Urbanisation and alien invasion. In K. J. Gaston (Ed.), Urban ecology (pp. 120–133). Cambridge: Cambridge University Press. Kowarik, I. (2011). Novel urban ecosystems, biodiversity, and conservation. Environmental Pollution, 159(8/9), 1974–1983. http://dx.doi.org/10.1016/j. envpol.2011.02.022 Kühn, I., & Klotz, S. (2006). Urbanization and homogenization—comparing the floras of urban and rural areas in Germany. Biological Conservation, 127(3), 292–300. http://dx.doi.org/10.1016/j.biocon.2005.06.033 Lockwood, J. L., Cassey, P., & Blackburn, T. (2005). The role of propagule pressure in explaining species invasions. Trends in Ecology and Evolution, 20(5), 223–228. http://dx.doi.org/10.1016/j.tree.2005.02.004 Mackenzie, B. F., & Larson, B. M. H. (2010). Participation under time constraints: landowner perceptions of rapid response to the emerald ash borer. Society and Natural Resources, 23(10), 1013–1022. http://dx.doi.org/10.1080/ 08941920903339707 Macdonald, I. A. W., Clark, D. L., & Taylor, H. C. (1989). The history and effects of alien plant control in the Cape of Good Hope Nature Reserve: 1941–1987. South African Journal of Botany, 55, 56–75. McKinney, M. L. (2006). Urbanization as a major cause of biotic homogenization. Biological Conservation, 127(3), 247–260. http://dx.doi.org/10.1016/j.biocon. 2005.09.005 Mgidi, T. N., Le Maître, D. C., Schonegevel, L., Nel, J. L., Rouget, M., & Richardson, D. M. (2007). Alien plant invasions—incorporating emerging invaders in regional prioritization: a pragmatic approach for Southern Africa. Journal of Environmental Management, 84(2), 173–187. http://dx.doi.org/10.1016/j. jenvman.2006.05.018 Moreira-Arce, D., de la Barrera, F., & Bustamante, R. O. (2014). Distance to suburban/wildland border interacts with habitat type for structuring exotic plant communities in a natural area surrounding a metropolitan area in central Chile. Plant Ecology and Diversity, 8(3), 363–370. http://dx.doi.org/10.1080/ 17550874.2014.983201 Nobis, M. P., Jaeger, J. A. G., & Zimmermann, N. E. (2009). Neophyte species richness at the landscape scale under urban sprawl and climate warming. Diversity and Distributions, 15(6), 928–939. http://dx.doi.org/10.1111/j.1472-4642.2009. 00610.x Pickett, S. T. A., Cadenasso, M. L., Grove, J. M., Nilon, C. H., Pouyat, R. V., Zipperer, W. C., et al. (2001). Urban ecological systems: linking terrestrial ecological, physical, and socioeconomic components of metropolitan areas. Annual Review
M. Gaertner et al. / Landscape and Urban Planning 151 (2016) 1–9 of Ecology and Systematics, 32, 127–157. http://dx.doi.org/10.1146/annurev. ecolsys.32.081501.114012 Rebelo, A. G., Holmes, P. M., Dorse, C., & Wood, J. (2011). Impacts of urbanization in a biodiversity hotspot: conservation challenges in Metropolitan Cape Town. South African Journal of Botany, 77(1), 20–35. http://dx.doi.org/10.1016/j.sajb. 2010.04.006 Richardson, D. M., & Brown, P. J. (1986). Invasion of mesic mountain fynbos by Pinus radiata. South African Journal of Botany, 52, 529–536. Richardson, D. M., & Gaertner, M. (2013). Plant invasions as builders and shapers of novel ecosystems. In R. J. Hobbs, E. S. Higgs, & C. M. Hall (Eds.), Novel ecosystems: intervening the new ecological world order (pp. 102–113). Chichester: John Wiley & Sons. Richardson, D. M., & Higgins, S. I. (1998). Pines as invaders in the southern hemisphere. In D. M. Richardson (Ed.), Ecology and biogeography of Pinus (pp. 450–473). Cambridge: Cambridge University Press. Richardson, D. M., Pyˇsek, P., & Carlton, J. T. (2011). A compendium of essential concepts and terminology in invasion ecology. In D. M. Richardson (Ed.), Fifty years of invasion ecology. The legacy of Charles Elton. Oxford: Wiley-Blackwell, pp. 409–435. Richardson, D. M., & van Wilgen, B. W. (2004). Invasive alien plants in South Africa: how well do we understand the ecological impacts? South African Journal of Science, 100(1/2), 45–52. Richardson, D. M., van Wilgen, B. W., Higgins, S. I., Trinder-Smith, T. H., Cowling, R. M., & McKelly, D. H. (1996). Current and future threats to biodiversity on the Cape Peninsula. Biodiversity and Conservation, 5(5) http://dx.doi.org/10.1007/ BF00137612 Rittel, H. W. J., & Webber, M. M. (1973). Dilemmas in a general theory of planning. Policy Sciences, 4, 155–169. Trentanovi, G., von der Lippe, M., Sitzia, T., Ziechmann, U., Kowarik, I., & Cierjacks, A. (2013). Biotic homogenization at the community scale: disentangling the roles of urbanization and plant invasion. Diversity and Distributions, 19(7), 738–748. http://dx.doi.org/10.1111/ddi.12028
9
van Wilgen, B. W. (2012). Evidence, perceptions, and trade-offs associated with invasive plant control in Table Mountain National Park, South Africa. Ecology and Society, 17(2), 23. http://dx.doi.org/10.5751/ES-04590-170223 van Wilgen, B. W., Reyers, B., Le Maitre, D. C., Richardson, D. M., & Schonegevel, L. (2008). A biome-scale assessment of the impact of invasive plants on ecosystem services in South Africa. Journal of Environmental Management, 89(4), 336–349. http://dx.doi.org/10.1016/j.jenvman.2007.06.015 van Wilgen, N. J., & Richardson, D. M. (2012). The roles of climate, phylogenetic relatedness, introduction effort, and reproductive traits in the establishment of non-native reptiles and amphibians. Conservation Biology, 26(2), 267–277. http://dx.doi.org/10.1111/j.1523-1739.2011.01804.x van Wilgen, B. W., & Scott, D. F. (2001). Managing fires on the Cape Peninsula: dealing with the inevitable. Journal of Mediterranean Ecology, 2, 197–208. van Wilgen, B. W., & Wannenburgh, A. (2016). Co-facilitating invasive species control, water conservation and poverty relief: achievements and challenges in South Africa’s Working for Water programme. Current Opinion in Environmental Sustainability, 19, 7–17. http://dx.doi.org/10.1016/j.cosust.2015.08.012 Vilà, M., Basnou, C., Pyˇsek, P., Josefsson, M., Genovesi, P., Gollash, S., et al. (2009). How well do we understand the impacts of alien species on ecosystem services? A pan-European, cross-taxa assessment. Frontiers in Ecology and the Environment, 8(3), 135–144. http://dx.doi.org/10.1890/080083 Warren, C. R. (2007). Perspectives on the ‘alien’ versus ‘native’ species debate: a critique of concepts, language and practice. Progress in Human Geography, 31(4), 427–446. http://dx.doi.org/10.1177/0309132507079499 Wilson, J. R. U., Ivey, P., Manyama, P., & Nänni, I. (2013). A new national unit for invasive species detection, assessment and eradication planning. South African Journal of Science, 109(5/6), 01–13. http://dx.doi.org/10.1590/sajs.2013/ 20120111 Zisenis, M. (2015). Alien plant species: a real fear for urban ecosystems in Europe? Urban Ecosystems, 18(2), 355–370. http://dx.doi.org/10.1007/s11252-0140400-1