REVIEW: Managing urban ecosystems for goods and services

THE UK NATIONAL ECOSYSTEM ASSESSMENT

Managing urban ecosystems for goods and services

Kevin J. Gaston1,*, Mar�ıa L. �Avila-Jim�enez1 and Jill L. Edmondson2

1Environment and Sustainability Institute, University of Exeter, Penryn, Cornwall TR10 9EZ, UK; and 2Department of

Animal and Plant Sciences, University of Sheffield, Sheffield S10 2TN, UK

Summary

1. Concomitant with the rise in the proportion of the global human population that resides

in urban areas has been growth in awareness of the importance of the provision of ecosystem

goods and services to those people. Urban areas are themselves of significance in this regard

because of their areal extent, and hence the quantity of services falling within their bounds,

and because of the need for local provision of services to urban residents.

2. Here, we review key challenges to the effective management of ecosystem goods and

services within urban areas.

3. These challenges include the structure of green space, its temporal dynamics, spatial con-

straint on ecosystem service flows, occurrence of novel forms of flows, large numbers of land

managers, conflicting management goals, possible differences between perceptions of urban

dwellers and the reality of the distribution and flow of ecosystem services, and the ‘wicked’

nature of the problem of ecosystem service management.

4. Synthesis and applications. Urban areas present very particular combinations of challenges

and opportunities for the management of ecosystem goods and services. The spatial and

temporal heterogeneity of green spaces greatly complicates the maintenance and improvement

in service provision as well as dramatically inflating costs. Spatial constraints on ecosystem

service flows mean that these can be highly dependent on the maintenance of particular areas

of connectivity, but also that provision of additional key points of connectivity may be

disproportionately beneficial to those flows. The existence of novel forms of flows of ecosys-

tem services in urban areas offers means of overcoming spatial constraints on more natural

flows, but will require the development of new kinds of ecosystem process models to inform

their design and management. The large numbers of land managers, conflicts between the best

approaches for managing for different goods and services, and frequent differences between

the perceptions of urban dwellers and the reality of urban landscapes create a complex

management context. The management of ecosystem goods and services is closely allied to

the challenges of conventional urban planning. However, applied ecology has a broad range

of tools available to assist in determining solutions, including the use of high-resolution

remote sensing techniques, landscape ecology principles and theory (e.g. patch and matrix

frameworks, meta-population models), and systematic conservation planning approaches.

Key-words: ecosystem services, ecosystems, flows, land-use change, planning, urban ecology,

urbanisation

Introduction

The individuals of a species are seldom randomly distrib-

uted in space (Gaston 2003). Rather, they tend to be

highly aggregated, with most individuals occurring in the

close proximity of many others. This is true of humans,

with the present centres of aggregation being the towns,

cities and conurbations in which the majority of us live

and work. Indeed, in much of the world, an increasing

proportion of people occur in such urban foci (United

Nations 2008).

Unlike the aggregated populations of the majority of

other species (with obvious exceptions such as seabird and

bat colonies), human urban populations typically obtain

most of their ecosystem resources from sources that are

distributed over a substantially larger area – although*Correspondence author. E-mail: [email protected]

© 2013 The Authors. Journal of Applied Ecology © 2013 British Ecological Society

Journal of Applied Ecology 2013 doi: 10.1111/1365-2664.12087

there is much variation, their ecological footprint is often

one to two orders of magnitude greater than the area

occupied by the population itself (Rees 1992, 1999;

Wackernagel et al. 2006). Acknowledging that urban

areas can make more efficient use of some resources than

more dispersed populations (Bettencourt et al. 2007), a

vital issue in limiting human impacts on the environment

at large is how those footprints can most effectively be

reduced (particularly as demand outstrips supply; e.g.

McDonald et al. 2011). Potential solutions include reduc-

ing the overall demand for resources (e.g. controlling pop-

ulation growth, promoting more sustainable resource use),

increasing their supply (e.g. increasing the flow of food

and energy from within parts of the existing footprint)

and increasing the intensity (decreasing the area require-

ment) of this supply (e.g. more intensive agriculture) (e.g.

Newman & Jennings 2008; Nelson et al. 2010; Phalan

et al. 2011; Sulston et al. 2012). However, a key challenge

to the second two approaches is that, to date, increases in

yield and intensity in modern agriculture have largely

been achieved by unsustainable means; there is a global

need to reduce the environmental impact of the agricul-

tural system (Godfray et al. 2010; Sulston et al. 2012).

Whilst approaches to increasing the supply and the inten-

sity of supply of resources focus on nonurban areas,

approaches to reducing the overall demand for resources

focus on the urban areas themselves. This has tended to

foster a belief that urban areas have little role to play in

the direct provision of ecosystem goods and services, and

such possibilities are commonly ignored both in discus-

sions of the distribution of those goods and services and

in global and regional accounting procedures for them

(e.g. Haines-Young 2009; Harrison et al. 2010). Instead,

attention is focussed on the influence of urban areas on

the destruction of potential regional provision and alter-

ation of natural regional patterns of flow.

Conversely, a growing body of research is demonstrat-

ing that urban areas can themselves be vitally important

for the provision of ecosystem goods and services. This

occurs in two ways. First, as urban areas increase in

extent, the ecosystem goods and services provided within

their bounds will inevitably constitute a growing propor-

tion of their regional and global provision. This is partic-

ularly so given that no such areas are entirely covered by

impermeable surfaces, indeed in many cities and towns

green spaces contribute a significant proportion of total

urban land cover (e.g. see Churkina, Brown & Keoleian

2010; Davies et al. 2011b). The global coverage of urban

areas remains relatively small, with a figure of 2–3% of

land (excluding permanent ice cover) commonly quoted

(e.g. Millennium Ecosystem Assessment 2005). However,

regional coverage may be substantially larger; figures for

165 countries vary from close to zero to 32% (World

Resources Institute 2007). At least for more temperate

zones, this means that urban areas can make substantial

contributions to ecosystem service stocks and flows,

particularly where these have been heavily depleted from

nonurban areas and where urban areas tend to be devel-

oped in zones that are rich in these resources (Nowak &

Crane 2002; Gaston 2005; Pataki et al. 2006; Pouyat,

Yesilonis & Nowak 2006; Davies et al. 2009, 2011a,a,b;

O’Neill & Abson 2009; Hutyra, Yoon & Alberti 2011). At

the very least, this may often be true to the point where it

is necessary to ensure the inclusion of urban areas when

determining regional baselines and conducting regional

accounting for ecosystem goods and services.

Second, and arguably much more significantly, urban

areas can themselves be key in the local provision of eco-

system goods and services to their occupants. Here, the

contribution to global stocks and flows of goods and ser-

vices is of far less importance than is their spatial and

temporal coincidence with these centres of human popula-

tion. Examples of such local provision arise amongst each

of the major groups of ecosystem goods and services: sup-

porting (e.g. soil formation and nutrient cycling), provi-

sioning (e.g. urban food production), regulating (e.g. local

climate and flood regulation) and cultural (e.g. aesthetic,

sense of place and health benefits of green space and wild-

life; Davies et al. 2011a). In most cases, the actual level

and pattern of provision remain to be well documented,

at least beyond a few exemplar case studies. However, it

is clear that existing local provision, and the improve-

ments that can be attained, makes urban areas both prac-

tically much more functional and more pleasant places to

live for their occupants. Indeed, a growing number of

papers have highlighted the significance of the levels of

provision and of the inequalities in their availability and/

or how they are accessed by different socioeconomic

groups (e.g. Jo & McPherson 2001; Hope et al. 2003;

Kinzig et al. 2005; Grove et al. 2006; Barbosa et al. 2007;

Tratalos et al. 2007; Alexandri & Jones 2008; Davies et al.

2011a,b, 2012; Fuller et al. 2012). The consequences of

compromising the provision of particular ecosystem ser-

vices in urban areas can in some cases be extreme, as evi-

denced for example by the rise in human mortality rates

associated with urban heat island effects (e.g. Goggins

et al. 2012).

Accepting that the provision of ecosystem goods and

services within urban areas is an important issue, then so

is their effective management; the term ‘management’ has

been used in a variety of ways in the context of urban

ecosystems (Jansson & Lindgren 2012), but here we

employ it broadly to refer to any change that positively

influences the availability and/or provision of an ecosys-

tem good or service (and ideally multiple goods and ser-

vices). Indeed, substantial public and private sums are

already spent annually in many urban areas on environ-

mental management actions, at a variety of spatial scales,

that intentionally or unintentionally have impacts on eco-

system goods and services. These include the creation of

green spaces (e.g. new public spaces, landscaping develop-

ments), hard and soft landscaping, managing vegetation

in existing or newly created green spaces (e.g. grazing,

mowing, tree planting and surgery, coppicing, growing

© 2013 The Authors. Journal of Applied Ecology © 2013 British Ecological Society, Journal of Applied Ecology

2 K. J. Gaston, M. L. Avila-Jim�enez & J. L. Edmondson

fruit and vegetables), managing green waste (e.g. compo-

sting, wood chipping), installation of green roofs and

walls, and ‘wildlife gardening’ (e.g. provision of nectar-

rich plants, bird feeders, nest boxes, bat boxes, ponds)

(Snep & Opdam 2010; Douglas & Ravetz 2011; Rowe

2011; Sadler et al. 2011; Hale & Sadler 2012).

There are several overarching challenges to the manage-

ment of ecosystem goods and services in urban areas,

including the structure of green space, its temporal

dynamics, the spatial constraint on service flows, the

occurrence of novel forms of flows, the large numbers of

land managers, conflicting management goals, the possible

differences between the perceptions of urban dwellers and

the reality of urban landscapes, and the ‘wicked’ nature of

the management problems. In this paper, we examine

each of these challenges in turn. We draw largely on

examples from the UK. In keeping with much of Europe,

many UK cities have a long history of human settlement

and a compact urban form, in contrast to the shorter his-

tory and sprawling urban form that typifies many cities

in, for example, the USA and Australia (Gaston 2010).

However, most of the key points generalise widely.

Green space structure

Although often not of itself sufficient, key to the provi-

sion of many, and perhaps the majority, of ecosystem

goods and services within urban areas is the provision of

green space (here used broadly to mean any area of land

not covered by impermeable surface, including remnant

vegetation, public parks, public and private landscaping,

domestic gardens, sports and playing fields, allotments,

cemeteries, nature reserves, and derelict lands). Such space

can vary greatly in its overall extent within any given

town, city or conurbation (Fuller & Gaston 2009; Gaston

2010). However, almost invariably this overall area is spa-

tially distributed in complex ways. Of particular relevance

to the provision of ecosystem goods and services are that

the majority of green spaces tend to be small (Fig. 1) and

that habitat heterogeneity is commonly high amongst

spaces. In some cases, the level of fragmentation is such

that cumulatively the smaller patches (albeit not the

smallest) comprise a large proportion of green space

(Fig. 2; Gaston et al. 2005; Fuller et al. 2010). The spatial

distribution of patches of different sizes tends to be highly

variable amongst different urban areas; however, large

green spaces are often more peripheral, whether these

urban areas are best described as monocentric or polycen-

tric in their built structure.

The habitat heterogeneity of green spaces is influenced

foremost by the different uses to which they are put (see

above). However, within these categories, there can

remain substantial additional heterogeneity, with fine-scale

patchworks of habitat types typifying some land uses,

such as public parks and domestic gardens (e.g. Smith

et al. 2005; Loram, Warren & Gaston 2008). This

heterogeneity is such that it may only become apparent

when urban green spaces are mapped at spatial resolu-

tions that are much finer than commonly employed when

studying rural landscapes, and necessitating high-resolution

remote sensing data (Fuller et al. 2010). A comparison

between two different areas of the city of Leicester, UK,

the first displaying nondomestically owned green space

dominated by a public park (Fig. 3a) and the second dis-

playing domestically owned or rented land within the

city’s suburbs (Fig. 3b), illustrates the heterogeneity of

green space within urban areas. For example, the cover of

trees within the area of the city dominated by park was

40% of total nondomestic green space, dropping to only

13% in the domestically owned or rented green space

within the city’s suburbs. However, the most stark

(a) (b)

(c) (d)

Fig. 1. Patch size frequency histogram, representing the number

of green spaces with a given areal extent (in km2), within New-

castle-upon-Tyne, UK, urban boundary: (a) histogram including

all green space patches; (b) histogram of green spaces patch size

excluding those with an area smaller than 0�001 km2; (c) histo-

gram of all green spaces with patch sizes exceeding 0�01 km2; and

(d) histogram of those green spaces exceeding 0�1 km2. Data

derived from MasterMap.

Fig. 2. Cumulative contribution of different-sized green space

patches to the overall areal extent of green space within Newcastle-

upon-Tyne, UK. The patches of green space are ranked in order of

size (smallest to largest), and the cumulative size (km2) is plotted

against the rank. Data derived from MasterMap.


Urban ecosystems for goods and services 3

contrast between these two areas of the city was the patch

size of the different vegetation land cover classes. In total,

the domestic green space within the suburban area was

comprised of over 6200 patches (Fig. 3b), dropping to

c. 2200 in the nondomestic green space (Fig. 3a). Mean

patch size ranged from 93 m2 for domestic green space to

204 m2 for nondomestic, with median values of 64 m2

and 81 m2, respectively.

This size and habitat structure of green spaces present a

major challenge to the management of the ecosystem

goods and services that they provide. First, they make

determining the levels of those goods and services diffi-

cult. Carefully structured survey programmes need to be

employed to ensure adequate coverage of the full breadth

of types and sizes of green spaces (e.g. Nowak et al.

2008a,b; Davies et al. 2011b). Complex techniques may

often also be required to extrapolate local observations

more widely. For example, whilst estimating the potential

impacts of individual trees on building energy use is

relatively straightforward, mapping these effects across an

entire city is not simple (Fahmy & Sharples 2009).

Second, the structure of green space is such that the

provision of many ecosystem goods and services scales,

typically positively, with patch size (e.g. Fuller et al. 2010;

Su et al. 2012). This makes an understanding of these

scaling relationships, many of which are markedly nonlin-

ear, and thus the potential consequences of increasing or

decreasing green space patch sizes, key to the manage-

ment of those goods and services. Third, the heterogeneity

of urban green spaces, with complex mosaics of habitats

(Davies et al. 2011b), means that management practices

are likely to have to be varied both within and between

them, which can render these practices costly, with econo-

mies of scale readily being lost. Fourth, the manner by

which ecosystem service provision is best maintained may

vary with green space size. For example, heavy manage-

ment may be most cost-effective for smaller (but probably

not the smallest) patches, but it may be better to rely

more fully on more natural processes in larger patches

both because these processes are themselves likely to be

more functional and because the scale of more interven-

tionist approaches may be impractical. Fifth, the cumula-

tive extent of smaller green spaces is significant because

strategic urban planning, and indeed much green space

planning per se, tends to focus almost exclusively on the

larger patches and typically ignores the smaller ones.

However, it is important to note that management to

maintain existing urban green spaces should be used pru-

dently as certain techniques may alter the ecosystem ser-

vice balance in these areas; for example, the use of fossil

fuel-powered machinery to prune urban trees will result in

a net reduction of carbon sequestration (Nowak et al.

2002; Davies et al. 2011b).

The consequences of differentially managing green

spaces of different sizes can be profound. In Table 1, in a

purely illustrative exercise which could be extended to

other ecosystem goods and services, we estimate the con-

sequences for above-ground carbon storage of managing

the green space stock of Newcastle-Upon-Tyne in the fol-

lowing ways – Model 1: all patches bigger than 0�1 km2

are converted into woodland, any smaller patch is man-

aged as a domestic garden; Model 2: all patches bigger

than 0�01 km2 are converted into woodland, any smaller

patch is managed as a domestic garden; Model 3: medium

size patches (>0�01 km2 and <0�1 km2) are converted into

woodland, smaller patches are managed as domestic gar-

dens, and larger patches are devoted to agriculture; and

Model 4: all patches managed similarly, with 50% wood-

land cover and 50% herbaceous cover. Carbon storage

potential can double from model 1 to model 3 and triple

from model 1 to model 2, whereas an intermediate

increase in storage could be achieved with model 4.

Temporal dynamics

Urban areas are typically very dynamic and responsive to

history, policy and other drivers, meaning that through

time a proportion of green spaces will turn over and

(a) (b)

Fig. 3. A comparison of the spatial heterogeneity of green space in the city of Leicester between; (a) nondomestically owned green space

in an area of the city dominated by a public park and (b) domestically owned or rented green space within a suburban area of the city.

Herbaceous vegetation (light green); shrubs, tall shrubs and trees (dark green); open water (blue); artificial surfaces (grey); buildings

(black). Within (a) white areas are domestically owned or rented green space, and (b) white areas are nondomestically owned green

space. Data derived from LandBase and MasterMap.



change in size and shape (e.g. Pauleit, Ennos & Golding

2005; Uy & Nakagoshi 2007; Kattwinkel, Biedermann &

Kleyer 2011; Zhou et al. 2011; Gillespie et al. 2012). For

example, Fuller et al. (2010) show how the proportion of

green space in 250 9 250 m grid cells across Sheffield,

UK, declines with the time (from 1860) since each cell

became predominantly urbanised; once urbanised, green

space in a cell has become progressively further eroded.

On a shorter time horizon, Dallimer et al. (2011) docu-

ment a net increase in the extent of green space in all but

one of 13 UK cities between 1991 and 2006. However,

this gain in the main occurred before 2001, since when

green space declined in nine of the cities, following policy

reforms towards greater housing densification in 2000.

Likewise, across the early years of the present century,

tree cover decreased significantly in 17 of 20 US cities

(Nowak & Greenfield 2012).

Particularly when combined with urban expansion, such

dynamics typically lead to mosaics of green space of

different ages, potentially including remnants of original

vegetation, heavily contaminated lands, well-established

stands of new vegetation, through to areas of fresh bare

ground and pioneer communities. The ages of green

spaces, or of the buildings with which they are associated,

can be correlated with their levels of vegetation cover (e.g.

Smith et al. 2005; Kendal, Williams & Williams 2012),

species richness and composition (e.g. Smith et al. 2006;

Luck & Smallbone 2010), and organic carbon storage

(e.g. Golubiewski 2006; Smetak, Johnson-Maynard &

Lloyd 2007). This dynamism of green spaces suggests that

the provision of ecosystem goods and services in urban

areas may need to be viewed in something akin to one of

the ‘patch’ frameworks of population ecology. By anal-

ogy, one can envisage a range of patterns of provision of

goods and services, spanning the equivalents of single

population, classical metapopulation, mainland-island

metapopulation, patchy population and non-equilibrium

metapopulation spatial structures (Harrison 1994). This

will be challenging, particularly because management

frameworks often do not match the spatial and temporal

scale of ecosystem processes (Borgstr€om et al. 2006).

Much of the implementation of management actions will

need to be resourced and conducted on the understanding

that its consequences will not be realised potentially for

many years or even decades.

Spatial constraint on flows

The spatial flows of ecosystem goods and services are

highly constrained (canalised) across urban landscapes.

That is, the built infrastructure (Fig. 4a and b) imposes

major limitations on how goods and services can pass

across these landscapes, whether those flows deliver the

benefits locally, regionally (e.g. to rural communities or

other urban ones) or beyond. It does so both at a gross

level in terms of the pattern of buildings and transport

networks (Fig. 4a and c) and at a finer resolution in terms

of the position of, for example, individual roads and

waterways (Fig. 4c and d). This constraint has perhaps

been best characterised for organismal movement

(Andrieu et al. 2009; Shanahan et al. 2011; Tremblay &

St. Clair 2011; Hale et al. 2012; Vergnes, Le Viol &

Clergeau 2012) and hence the services which wildlife provide

or to which they contribute. However, it generalises much

more widely. Indeed, it seems likely that the majority of

ecosystem goods and services show much more spatially

constrained flows than in many other environments.

From a management perspective, such spatial constraint

means that, on the one hand, flows of ecosystem goods

and services across urban landscapes may be especially

vulnerable to the severance of particular corridors of flow.

On the other hand, there may be considerable opportuni-

ties for disproportionately improving flows by the creation

of new corridors. Such concerns have been widely recog-

nised in plans to protect and develop the blue and green

infrastructures of many urban areas, with riverine corri-

dors having been a heavy focus of attention (e.g. Gledhill,

James & Davies 2008; Kazmierczak & Carter 2010; Kelly,

Luke & Lima 2011). A persistent issue, however, has been

that such planning tends also to concentrate, sometimes

exclusively, on the role of large unified ‘green/blue corri-

dors’ through urban landscapes, and to ignore the poten-

tial roles of the huge numbers of small patches of

vegetation (and perhaps water). This matrix within which

Table 1. Potential carbon stored above-ground within the urban

green space (87 km2) of Newcastle-upon-Tyne urban boundary

on different management scenarios, based on average above-

ground carbon quantifications by Davies et al. (2011a,b)

(0�76 kg C m�2 stored in domestic gardens, 0�14 kg C m�2 stored

in land covered by herbaceous vegetation, and 28�86 kg C m�2

associated with tree covered land): (i) potential amount of carbon

stored within the Newcastle-upon-Tyne urban boundary if all the

green spaces were covered by trees; (ii) if all the green spaces

were managed as domestic gardens; (iii) if all the green spaces

were covered by herbaceous vegetation; (iv) Model 1: if all the

large green spaces(area >0�1 km2) were tree covered and all the

smaller patches were managed as domestic gardens; (v) Model 2:

if all large and medium size green spaces (area >0�01 km2) are

tree covered and all the smaller patches were managed as domes-

tic gardens; (vi) Model 3: if all medium size green spaces

(<0�1 km2 to >0�01 km2 area) are tree covered, smaller patches

are managed as domestic gardens, and all large green spaces are

devoted to agriculture (with herbaceous cover); (vii) Model 4:

each greenspace is 50% tree cover and 50% herbaceous cover.

Distribution of green spaces derived from MasterMap

Scenario Potential C storage (tonnes)

All green spaces tree covered 2 508 301

All green spaces managed as

domestic gardens

66 054

All green spaces covered by

herbaceous species

12 168

Model 1 524 508

Model 2 1 477 882

Model 3 1 035 098

Model 4 1 260 234



the larger corridors are set will often in practice be of

great (and perhaps greater) significance to the flows of

ecosystem goods and services, which may exhibit patterns

of constraint across this matrix that need critically to be

identified and appropriately managed.

Key to planning for the management of urban ecosys-

tem services will be the use of spatial planning tools that

prioritise areas in terms of their realised or potential

importance for spatial flows and that can be used to

explore the consequences of different planning decisions.

Although the general principles have long been promoted

(McHarg 1992), lessons need to be learnt from those tools

that have been developed for biodiversity conservation,

which faces many closely related issues, and which can in

some cases be adapted to this end (e.g. Margules & Pressey

2000; Moilanen et al. 2012; Moilanen, Wilson & Poss-

ingham 2009; for application to ecosystem services see

Moilanen et al. 2011); the use of these tools for species

conservation in urban areas has already been highlighted

(e.g. Gordon et al. 2009).

Novel flows

As well as often being strongly spatially constrained, eco-

system goods and services can also flow across urban

landscapes in ways that do not occur, or are typically

much less significant, in other landscape types. In particu-

lar, humans directly move material around urban land-

scapes on a massive scale. This includes soil (e.g. for

landscaping), vegetation (e.g. through waste collection sys-

tems and urban horticulture) and water (e.g. through

drainage systems). For example, in a survey of 575

residents in the city of Leicester, UK (Living in Leicester;

Lomas et al. 2010), respondents reported annually remov-

ing in total the equivalent of nearly 6000 bin bags of

green waste from their gardens. Nearly 30% of this waste

was hedge and shrub clippings and tree prunings; this can

be converted to c. 630 kg organic carbon removed from

the urban green space (estimated using analysis of organic

carbon concentration and measurement of the dry weight

of full bin bags of seven common urban tree species and

11 common urban shrub species). To contextualise the

scale of movement of waste across the city, the total gar-

den area of survey respondents was c. 8 ha, which was

<1% of total garden area in Leicester. Removal from gar-

dens was greatly in excess of the flow of organic matter

into gardens, with respondents reporting annually adding

the equivalent of an estimated 1200 bin bags of material

to their gardens (including manure, commercial and own

produced compost, bark and tree chippings, straw and

topsoil). However, this represents an addition of organic

carbon to urban gardens above the natural inputs from

vegetation.

Such novel flows often remain rather poorly character-

ised and thus constitute major unknowns in documenting

the dynamics of ecosystem goods and services in urban

areas and formalising these in ecosystem models. None-

theless, they do offer potential avenues for managing

those goods and services that may be easier to alter than

might be the case with more natural patterns of flow.

Multiple managers

Urban ecosystems are unlike almost any others in the

exceedingly large numbers of land managers present.

Whilst substantial tracts of land may be under the control

(a) (b)

(c) (d) Fig. 4. Spatial structure of green spaces in

Newcastle-upon-Tyne, UK, and spatial

structure of identified barriers to and

channels for the flow of ecosystem ser-

vices. (a) Green spaces exceeding 0�1 km2

(in green) and built infrastructure (in

grey); (b) all green spaces within the urban

area (in green) and their physical bound-

aries (in grey); (c) transport infrastructure

(in grey: minor roads; in red: major roads;

in black: railway lines); (d) surface water

(rivers, streams, open canals, ponds and

lakes) (in blue). Data derived from

MasterMap.



of a few governmental or nongovernmental organisations

(private individuals, companies or charities), typically

much is divided amongst a vastly greater number of

individual homeowners or tenants, each of whom may be

responsible for an area of the order of tens to hundreds

of square metres. For example, in Leicester, green

space covers 56% of the urban area (73 km2), 80% of

which is privately managed. Green space associated

with c. 123 000 households, and therefore individual

managers, within the city constitutes 40% of the total,

with a further 40% of private land managed nondomesti-

cally (Fig. 5). A smaller proportion of the city’s green

space is owned and managed by the local authority

(20%); however, much of this land is held in larger

patches such as urban parks.

Most obviously, the large numbers of land owners and

tenants in urban areas make coordination of management

activities extremely difficult. Some level of control can be

exerted through legislation; however, this provides a

rather blunt instrument. The converse argument is that

because of the large numbers of land managers, only a

relatively small proportion may need to carry out a par-

ticular management action for a wide impact to result.

For example, the establishment of new ponds in just 10%

of the domestic gardens in the urban area of Sheffield

would result in the addition of 17 500 such habitat

patches (albeit typically small ones), at a density of c. 120

per km2 (Gaston et al. 2005). In Leicester, the urban

green space managed by 575 survey respondents (Living

in Leicester survey; Lomas et al. 2010) was c. 8 ha in

extent (<1% citywide garden area), of which nearly 50%

was covered by herbaceous vegetation. If 10% of this

herbaceous area was converted to land dedicated to

own-grown food, potential yield could exceed 10 tonnes

per annum (based on UK agricultural potato yields and

unpublished data from the Royal Horticultural Society

allotment yield trials in 1974; Tompkins 2006; Supit et al.

2010). A further 20% of the respondents’ gardens were

capped by artificial surfaces (e.g. driveways, patios and

footpaths; all common features of UK gardens; Loram,

Warren & Gaston 2008). Excavation of surface soil is rou-

tine during the construction of artificial surface; on aver-

age, the top 15 cm of soil is lost in domestic gardens and

consequently 6�7 kg m-2 organic carbon (Edmondson

et al. 2012). If 10% of the artificial surface in these

domestic gardens was removed and converted to lawn, soil

organic carbon storage, to 1 m depth, could increase by

13 tonnes, from 26 to 39 tonnes (data from Edmondson

et al. 2012). Only 20% of the same respondents used a

compost heap, with a smaller number adding household

fruit and vegetable waste, and any increase in the number

of people using this method could significantly reduce the

resource demand associated with waste management in

our urban areas (Gaston et al. 2005).

Other challenges that result from the large numbers of

managers include that (i) lots of managers are not manag-

ing for ecosystem service provision but for alternative,

and often conflicting, goals; (ii) different groups of man-

agers may have different perceptions as to what changes

are most desirable (e.g. Hofmann et al. 2012); (iii) there is

a loss of benefits of scale of management costs, both

financial and environmental (e.g. need for physical tools

for management, and the gas emissions that those tools

give rise to); and (iv) universal management actions to

target-specific environmental problems across whole urban

ecosystems, such as maximisation of green space flood

mitigation or organic carbon storage potential, are diffi-

cult to achieve.

Conflicting management goals

Inevitably, different ecosystem goods and services in

urban areas require different management approaches,

and in some cases, these will conflict. Thus, for example,

(i) increasing carbon sequestration and reducing summer

temperatures will often involve retaining or planting trees,

but this may result in an increase in emissions of biogenic

volatile organic compounds which are hazardous to

human health (Leung et al. 2010); and (ii) increasing

urban food production will reduce land available for

growing trees, and may involve use of chemicals that

impact on water quality. Given the large number and

small size of many green spaces, such conflicts are proba-

bly best handled by managing for different goods and ser-

vices in different patches, rather than in different parts of

the same patch. Unfortunately, this may sometimes be at

Fig. 5. The distribution of residential and nonresidential spaces

in Leicester, UK; nonresidential green space (light green), nonres-

idential artificial surface (pale grey), residential artificial surface

(dark grey), residential green space (dark green), buildings (white)

and inland water (blue). Distribution derived from the Master-

Map and LandBase GIS data sets.



odds with the goals of the owners or tenants of individual

green space patches. For example, in the UK, urban gar-

dens are often managed so as to maintain high levels of

habitat heterogeneity, such that although the number of

habitat features increases with garden size, many features

are present in a high proportion of gardens with their

extent scaling strongly with garden size (Smith et al. 2005;

Loram, Warren & Gaston 2008).

Perception and reality

Throughout, we have focussed on the challenges for the

management of ecosystem goods and services in urban

areas posed by their actual distribution in space and time.

Under some circumstances, there may also be important

challenges that result from differences between their

actual distributions and flows, and those that urban dwell-

ers perceive to prevail. For example, Dallimer et al.

(2012) have shown that whilst there are no consistent rela-

tionships between the psychological well-being of urban

green space visitors and the species richness of various

groups of organisms, there are positive relationships

with the richness that those users perceive to be present.

This is potentially highly problematic, as management

for actual levels of biodiversity may well conflict with

management that would enhance perceived levels of

biodiversity.

Appreciation of the value of an ecosystem service may

also vary with the role and perception of the stakeholder

(i.e. beneficiary, manager, policymaker) and the ecosystem

service considered. For example, beneficiaries may place

high value on local resources (e.g. parks, allotments) that

in broader planning terms are insignificant. Furthermore,

the value of the ecosystem service can be expressed in dif-

ferent ways (e.g. economic, ecological or social value),

and these valuations are often not comparable (de Groot

et al. 2010).

Wicked problems

From a more over-arching perspective, The Royal Com-

mission on Environmental Pollution (2007) argues that

urban environmental management presents a classic case

of a ‘wicked problem’; ‘wicked’ in the sense of nasty or

vicious, rather than an ethical judgement. The mainte-

nance of and improvement in the provision of ecosystem

goods and services is increasingly seen as a substantive

component of this management, and the same conclusion

might justifiably be reached in this regard. The notion of

wicked problems derives from a treatise by Rittel & Webber

(1973), who observed that the kinds of societal problems

that planners deal with are ill-defined and cannot be

definitively solved and thus are intrinsically different from

archetypal problems in science. Their characterisation of

such problems, which derive in large part from ‘the inter-

dependencies and complexities of living together without

a shared set of values and views’ (Roberts 2000), is

rephrased here in the context of the management of eco-

system goods and services in urban systems:

1. There is no definitive formulation of an ecosystem ser-

vice management problem – The process of describing the

problem, say the need to improve urban food production,

and of solving it are essentially the same. This can serve

to fuel disagreement as to what the ‘problem’ actually is

and lead to a framing of the problem in a manner that

more readily connects it with the solution preferred by a

particular stakeholder (Roberts 2000).

2. Ecosystem management problems have no stopping

rule – Because there is no definitive formulation of the

problem, there is no point at which the solution has been

found. There is, for example, no point at which an

improvement in urban climate regulation would defini-

tively be sufficient.

3. Solutions to ecosystem service management problems

are not true-or-false, but good-or-bad – The nature of a

particular solution is likely to depend on who provides it,

with, for example, local residents and regional govern-

ment likely to manage a given green space in a different

way (the former tending to focus on their own needs, the

latter on standardising practices across a regional portfo-

lio of green spaces).

4. There is no immediate and no ultimate test of a solu-

tion to an ecosystem service management problem – A

given solution is likely to have many consequences, some

unintended and unexpected, and these may play out over

long periods. This highlights a particular need for multi-

ple studies of the outcomes of ecosystem service manage-

ment actions in urban areas.

5. Every solution to an ecosystem service management

problem is a ‘one-shot operation’; because there is no

opportunity to learn by trial-and-error, every attempt

counts significantly – This is because management actions

are seldom entirely reversible but have many conse-

quences. It is impossible, for example, accurately to pre-

dict all the ecosystem service consequences of establishing

a new green space in an urban area, and once established

it would be impossible simply to reverse some of those

consequences.

6. Ecosystem management problems do not have an enu-

merable (or an exhaustively describable) set of potential

solutions, nor is there a well-described set of permissible

operations that may be incorporated into the solutions –

The set of potential solutions and the extent to which they

are permissible will depend on who provides them. Some

have argued that in a complex world, what is required are

‘clumsy solutions’, which combine alternative (and some-

times conflicting) ways of perceiving and organising

answers (Verweij et al. 2006).

7. Every ecosystem service management problem is essen-

tially unique – There are always particularities to a prob-

lem that may override the commonalities with other

problems. Thus, for example, the detail of approaches to

improving the human–wildlife interactions that benefit

people’s well-being will vary amongst cities, amongst



communities within a city, and at different times for any

given community.

8. Each ecosystem service management problem can be

considered a symptom of another problem – For example,

poor local climate regulation might follow from a lack of

carbon storage and sequestration, as a consequence of

poor management of vegetation cover.

9. The existence of a discrepancy representing an ecosys-

tem service management problem can be explained in

numerous ways. The choice of explanation determines the

nature of the problem’s resolution.

10. The ecosystem service manager has no right to be

wrong – The objective is to improve a situation, and there

are plenty of tools available to assist a manager to that

end.

Roberts (2000) distinguishes three generic strategies for

identifying wicked problems and their solutions. If power

is concentrated amongst a small number of stakeholders,

or it is placed in their hands by consent, then authorita-

tive strategies can be employed. This can greatly reduce

the complexity of the process, but can also result in prob-

lems being too narrowly or incorrectly characterised, and

in other parties being ill-informed and unengaged. If

power is dispersed and contested, then competitive strate-

gies can be employed, encouraging different sets of stake-

holders to garner sufficient power to define the wicked

problem and its solution. This has the advantage of

encouraging innovation, but can result in protracted,

distracting and costly battles for that power, which in

the extreme can prevent any practical progress being

achieved. Finally, if power is dispersed but not contested,

then collaborative strategies can be used to define the

problem and solution. This can be efficient, in spreading

costs, and increasing ‘weight of numbers’ and the breadth

of ‘solution space’ that can be explored. However, effec-

tive collaboration can be challenging to achieve, and

costly in the effort required. Across the breadth of wicked

problems posed by ecosystem service management in

urban areas, no one of these three strategies will be ade-

quate in itself. For some issues, power is highly concen-

trated (e.g. water flows), for others it is highly dispersed

(e.g. vegetation management).

CONCLUSION

The local provision of ecosystem goods and services in

urban areas is essential to the populations that benefit

from them and will help reduce the regional and global

footprint of cities and towns. The management of these

goods and services poses a number of substantive chal-

lenges, including the structure of green space, its temporal

dynamics, the spatial constraint on ecosystem service

flows, the occurrence of novel forms of those flows, the

large numbers of land managers, conflicting management

goals, the possible differences between the perceptions of

urban dwellers as to the distribution and flow of ecosys-

tem services and the reality of that distribution and flow,

and the ‘wicked’ nature of ecosystem service management

in urban landscapes. However, there is also clearly a

broad range of tools available from applied ecology to

assist in their resolution. These include the use of high-

resolution remote sensing techniques, landscape ecology

principles and theory (e.g. patch and matrix frameworks,

meta-population models) and systematic conservation

planning approaches. These will need to be employed

within a broader transdisciplinary framework (Ervin et al.

2012) to address the interactions between natural and

human systems that are arguably at their most complex in

urban ecosystems.

Acknowledgements

This work was supported by EPSRC grant EP/I002154/1 SECURE: Self

conserving urban environments (a consortium of Loughborough Univ.,

Newcastle Univ., Univ. Exeter and Univ. Sheffield); EPSRC grant

EP/F007604/1 4M: Measurement, Modelling, Mapping and Management

(a consortium of Loughborough Univ., De Montfort Univ., Newcastle

Univ., Univ. Sheffield and Univ. Exeter); and NERC grant NE/J015237/1

Fragments, functions and flows (a consortium of British Trust for

Ornithology, Cranfield Univ., Univ. Exeter and Univ. Sheffield). Infoterra

provided access to LandBase; MasterMap data were supplied by Ordnance

Survey. We are grateful to S. Gaston, Z. Grabowski, J. Jones and an

anonymous reviewer for comments.

References

Alexandri, E. & Jones, P. (2008) Temperature decreases in an urban can-

yon due to green walls and green roofs in diverse climates. Building and

Environment, 43, 480–493.Andrieu, E., Dornier, A., Rouifed, S., Schatz, B. & Cheptou, P.-O. (2009)

The town Crepis and the country Crepis: how does fragmentation affect

a plant-pollinator interaction? Acta Oecologica, 35, 1–7.Barbosa, O., Tratalos, J.A., Armsworth, P.R., Davies, R.G., Fuller, R.A.,

Johnson, P. & Gaston, K.J. (2007) Who benefits from access to green

space? A case study from Sheffield, UK. Landscape and Urban Planning,

83, 187–195.Bettencourt, L.M.A., Lobo, J., Helbing, D., K€uhnert, C. & West, G.B.

(2007) Growth, innovation, scaling, and the pace of life in cities.

Proceedings of the National Academy of Sciences of the USA, 104,

7301–7306.Borgstr€om, S.T., Elmqvist, T., Angelstam, P. & Alfsen-Norodom, C.

(2006) Scale mismatches in management of urban landscapes. Ecology

and Society, 11, 16.

Churkina, G., Brown, D.G. & Keoleian, G. (2010) Carbon stored in

human settlements: the conterminous United States. Global Change

Biology, 16, 135–143.Dallimer, M., Tang, Z., Bibby, P.R., Brindley, P., Gaston, K.J. & Davies,

Z.G. (2011) Temporal changes in greenspace in a highly urbanised

region. Biology Letters, 7, 763–766.Dallimer, M., Irvine, K.N., Skinner, A.M.J., Davies, Z.G., Rouquette,

J.R., Armsworth, P.R., Maltby, L., Warren, P.H. & Gaston, K.J.

(2012) Biodiversity and the feel-good factor: understanding associations

between self-reported human well-being and species richness. BioScience,

62, 46–55.Davies, Z.G., Fuller, R.A., Loram, A., Irvine, K.N., Sims, V. &

Gaston, K.J. (2009) A national scale inventory of resource provision

for biodiversity within domestic gardens. Biological Conservation, 142,

761–771.Davies, L., Kwiatkowski, L., Gaston, K.J., Beck, H., Brett, H., Batty,

M., Scholes, L., Wade, R., Sheate, W.R., Sadler, J., Perino, G.,

Andrews, B., Kontoleon, A., Bateman, I. & Harris, J.A. (2011a)

Urban. UK National Ecosystem Assessment: Technical Report, pp.

361–410. UNEP-WCMC, Cambridge.

Davies, Z.G., Edmondson, J.L., Heinemeyer, A., Leake, J.R. & Gaston,

K.J. (2011b) Mapping an urban ecosystem service: quantifying above-

ground carbon storage at a city-wide scale. Journal of Applied Ecology,

48, 1125–1134.



Davies, Z.G., Fuller, R.A., Dallimer, M., Loram, A. & Gaston, K.J.

(2012) Household factors influencing participation in bird feeding activ-

ity: a national scale analysis. PLoS ONE, 7, e39692.

Douglas, I. & Ravetz, J. (2011) Urban ecology – the bigger picture. Urban

Ecology: Patterns, Processes, and Applications (ed. J. Niemel€a), pp. 246–262. Oxford University Press, Oxford.

Edmondson, J.L., Davies, Z.G., McHugh, N., Gaston, K.J. & Leake, J.R.

(2012) Organic carbon hidden in urban ecosystems. Scientific Reports, 2,

963.

Ervin, D., Brown, D., Chang, H., Dujon, V., Granek, E., Shandas, V. &

Yeakley, A. (2012) Growing cities depend on ecosystem services. Solu-

tions, 2, 74–86.Fahmy, M. & Sharples, S. (2009) On the development of an urban passive

thermal comfort system in Cairo, Egypt. Building and Environment, 44,

1907–1916.Fuller, R.A. & Gaston, K.J. (2009) The scaling of green space coverage in

European cities. Biology Letters, 5, 352–355.Fuller, R.A., Tratalos, J., Warren, P.H., Davies, R.G., Pezpkowska, A. &

Gaston, K.J. (2010) Environment and biodiversity. Dimensions of the

Sustainable City (eds M. Jenks & C. Jones), pp. 75–103. Springer

Science, Dordrecht.

Fuller, R.A., Irvine, K.N., Davies, Z.G., Armsworth, P.R. & Gaston, K.J.

(2012) Interactions between people and birds in urban landscapes.

Studies in Avian Biology, 45, 249–266.Gaston, K.J. (2003) The Structure and Dynamics of Geographic Ranges.

Oxford University Press, Oxford.

Gaston, K.J. (2005) Biodiversity and extinction: species and people.

Progress in Physical Geography, 29, 239–247.Gaston, K.J. (2010) Urbanisation. Urban Ecology (ed. K.J. Gaston),

pp. 10–34. Cambridge University Press, Cambridge.

Gaston, K.J., Warren, P.H., Thompson, K. & Smith, R.M. (2005) Urban

domestic gardens (IV): the extent of the resource and its associated fea-

tures. Biodiversity and Conservation, 14, 3327–3349.Gillespie, T.W., Pincetl, S., Brossard, S., Smith, J., Saatchi, S., Pataki, D.

& Saphores, J.D. (2012) A time series of urban forestry in Los Angeles.

Urban Ecosystems, 15, 233–246.Gledhill, D.G., James, P. & Davies, D.H. (2008) Pond density as a deter-

minant of aquatic species richness in an urban landscape. Landscape

Ecology, 23, 1219–1230.Godfray, H.J.C., Beddington, J.R., Crute, I.R., Haddad, L., Lawrence,

D., Muir, J.F., Pretty, J., Robinson, S., Thomas, S.M. & Toulman, C.

(2010) Food security: the challenge of feeding 9 billion people. Science,

327, 812–818.Goggins, W.B., Chan, E.Y.Y., Ng, E., Ren, C. & Chen, L. (2012) Effect

modification of the association between short-term meteorological fac-

tors and mortality by urban heat islands in Hong Kong. PLoS ONE, 7,

e38551.

Golubiewski, N.E. (2006) Urbanization increases grassland carbon pools:

effects of landscaping in Colorado’s Front Range. Ecological Applica-

tions, 16, 555–571.Gordon, A., Simondson, D., White, M., Moilanen, A. & Bekessy, S.A.

(2009) Integrating conservation planning and landuse planning in urban

landscapes. Landscape and Urban Planning, 91, 183–194.de Groot, R.S., Alkemade, R., Braat, L., Hein, L. & Willemen, L. (2010)

Challenges in integrating the concept of ecosystem services and values

in landscape planning, management and decision making. Ecological

Complexity, 7, 260–272.Grove, J.M., Troy, A.R., O’Neill-Dunne, J.P.M., Burch, W.R., Cadenas-

so, M.L. & Pickett, S.T.A. (2006) Characterization of households and

its implications for the vegetation of urban ecosystems. Ecosystems, 9,

578–597.Haines-Young, R. (2009) Land use and biodiversity relationships. Land

Use Policy, 26, S178–S186.Hale, J.D. & Sadler, J. (2012) Resilient ecological solutions for urban

regeneration. Proceedings of the ICE – Engineering Sustainability, 165,

59–68.Hale, J.D., Fairbrass, A.J., Matthews, T.J. & Sadler, J.P. (2012) Habitat

composition and connectivity predicts bat presence and activity at

foraging sites in a large UK conurbation. PLoS ONE, 7, e33300.

Harrison, S. (1994) Metapopulations and conservation. Large-scale Ecol-

ogy and Conservation Biology (eds P.J. Edwards, R.M. May & N.R.

Webb), pp. 111–128. Blackwell Scientific Publications, Oxford.

Harrison, P.A., Vandewalle, M., Sykes, M.T., Berry, P.M., Bugter, R.,

de Bello, F., Feld, C.K., Grandin, U., Harrington, R., Haslett, J.R.,

Jongman, R.H.G., Luck, G.W., da Silva, P.M., Moora, M., Settele, J.,

Sousa, J.P. & Zobel, M. (2010) Identifying and prioritising services in

European terrestrial and freshwater ecosystems. Biodiversity and Conser-

vation, 19, 2791–2821.Hofmann, M., Westermann, J.R., Kowarik, I. & van der Meer, E. (2012)

Perceptions of parks and urban derelict land by landscape planners and

residents. Urban Forestry and Urban Greening, 11, 303–312.Hope, D., Gries, C., Zhu, W., Fagan, W.F., Redman, C.L., Grimm, N.B.,

Nelson, A.L., Martin, C. & Kinzig, A. (2003) Socioeconomics drive

urban plant diversity. Proceedings of the National Academy of Sciences

of the USA, 100, 8788–8792.Hutyra, L.R., Yoon, B. & Alberti, M. (2011) Terrestrial carbon stocks

across a gradient of urbanization: a study of Seattle, WA region. Global

Change Biology, 17, 783–797.Jansson, M. & Lindgren, T. (2012) A review of the concept ‘management’

in relation to urban landscapes and green spaces: toward a holistic

understanding. Urban Forestry and Urban Greening, 11, 139–145.Jo, H.-K. & McPherson, E.G. (2001) Indirect carbon reduction by residen-

tial vegetation and planting strategies in Chicago, USA. Journal of

Environmental Management, 61, 165–177.Kattwinkel, M., Biedermann, R. & Kleyer, M. (2011) Temporary conser-

vation for urban biodiversity. Biological Conservation, 144, 2335–2343.Kazmierczak, A. & Carter, J. (2010) Adaptation to climate change using

green and blue infrastructure. A database of case studies. Interreg IVC

Green and blue space adaptation for urban areas and eco towns

(GRaBS). Manchester, UK. Available at: http://www.grabs-eu.org/

(accessed 5 July 2012).

Kelly, S., Luke, A. & Lima, M. (2011) Developing Urban Blue Corridors.

Scoping study. DEFRA, London Borough of Croydon, UK.

Kendal, D., Williams, N.S.G. & Williams, K.J.H. (2012) Drivers of diver-

sity and tree cover in gardens, parks and streetscapes in an Australian

city. Urban Forestry and Urban Greening, 11, 257–265.Kinzig, A.P., Warren, P., Martin, C., Hope, D. & Katti, M. (2005) The

effects of human socioeconomic status and cultural characteristics on

urban patterns of biodiversity. Ecology and Society, 10, 23.

Leung, D.Y., Wong, P., Cheung, B.K. & Guenther, A.B. (2010) Improved

land cover and emission factors for modeling biogenic volatile organic

compounds emissions from Hong Kong. Atmospheric Environment, 44,

1456–1468.Lomas, K.J., Bell, M., Firth, S.K., Gaston, K.J., Goodman, P., Leake,

J.R., Namdeo, A., Rylatt, M., Allinson, D., Davies, Z.G., Edmondson,

J.L., Galatioto, F., Guo, L., Hill, G., Irvine, K., Taylor, S.C. & Tiwary,

A. (2010) 4M: Measurement; modelling; mapping and management –the carbon footprint of UK cities. ISOCARP Low Carbon Cities.

Isocarp, The Hague, pp 168–191.Loram, A., Warren, P.H. & Gaston, K.J. (2008) Urban domestic gardens

(XIV): the characteristics of gardens in five cities. Environmental

Management, 42, 361–376.Luck, G.W. & Smallbone, L.T. (2010) Species diversity and urbanisation:

patterns, drivers and implications. Urban Ecology (ed K.J. Gaston),

pp. 88–119. Cambridge University Press, Cambridge.

Margules, C.R. & Pressey, R.L. (2000) Systematic conservation planning.

Nature, 405, 243–253.McDonald, R.I., Green, P., Balk, D., Fekete, B.M., Revenga, C., Todd,

M. & Montgomery, M. (2011) Urban growth, climate change, and

freshwater availability. Proceedings of the National Academy of Sciences

of the USA, 108, 6312–6317.McHarg, I.L. (1992) Design in Nature. Wiley, New York.

Millennium Ecosystem Assessment (2005) Ecosystems and Human Well-

Being: Current State and Trends. Vol. I. Island Press, Washington, DC.

Moilanen, A., Wilson, K.A. & Possingham, H. (eds.) (2009) Spatial Con-

servation Prioritization: Quantitative Methods and Computational Tools.

Oxford University Press, Oxford.

Moilanen, A., Anderson, B.J., Eigenbrod, F., Heinemeyer, A., Roy, D.B.,

Gillings, S., Armsworth, P.R., Gaston, K.J. & Thomas, C.D. (2011)

Balancing alternative land uses in conservation prioritization. Ecological

Applications, 21, 1419–1426.Moilanen, A., Meller, L., Leppanen, J., Montesino-Puzols, F., Arponen, A.

& Kujala, H. (2012) Spatial Conservation Planning Framework and Soft-

ware: Zonation User Manual (v 3.1). University of Helsinki, Helsinki.

Nelson, E., Sander, H., Hawthorne, P., Conte, M., Ennaanay, D., Wolny,

S., Manson, S. & Plasky, S. (2010) Projecting global land-use change

and its effect on ecosystem service provision and biodiversity with

simple models. PLoS ONE, 5, e14327.

Newman, P. & Jennings, I. (2008) Cities as Sustainable Ecosystems: Princi-

ples and Practices. Island Press, Washington, DC.



Nowak, D.J. & Crane, D.E. (2002) Carbon storage and sequestration by

urban trees in the USA. Environmental Pollution, 116, 381–389.Nowak, D.J. & Greenfield, E.J. (2012) Tree and impervious cover change

in U.S. cities. Urban Forestry and Urban Greening, 11, 21–30.Nowak, D.J., Stevens, J.C., Sisinni, S.M. & Luley, C.J. (2002) Effects of

urban tree management and species selection on atmospheric carbon

dioxide. Journal of Arboriculture, 28, 113–122.Nowak, D.J., Walton, J.T., Stevens, J.C., Crane, D.E. & Hoehn, R.E.

(2008a) Effect of plot and sample size on timing and precision of urban

forest assessments. Aboriculture and Urban Forestry, 34, 386–390.Nowak, D.J., Crane, D.E., Stevens, J.C., Hoehn, R.E., Walton, J.T. &

Bond, J. (2008b) A ground-based method of assessing urban forest

structure and ecosystem services. Arboriculture and Urban Forestry, 34,

347–358.O’Neill, D.W. & Abson, D.J. (2009) To settle or protect? A global analysis

of net primary productivity in parks and urban areas. Ecological

Economics, 69, 319–327.Pataki, D.E., Alig, R.J., Fung, A.S., Golubiewski, N.E., Kennedy, C.A.,

McPherson, E.G., Nowak, D.J., Pouyat, R.V. & Lankao, P.R. (2006)

Urban ecosystems and the North American carbon cycle. Global Change

Biology, 12, 209–2102.Pauleit, S., Ennos, R. & Golding, Y. (2005) Modeling the environmental

impacts of urban land use and land cover change – a study in Mersey-

side, UK. Landscape and Urban Planning, 71, 295–310.Phalan, B., Onial, M., Balmford, A. & Green, R.E. (2011) Reconciling

food production and biodiversity conservation: land sharing and land

sparing compared. Science, 333, 1289–1291.Pouyat, R.V., Yesilonis, I.D. & Nowak, D.J. (2006) Carbon storage by

urban soils in the United States. Journal of Environmental Quality, 35,

1566–1575.Rees, W.E. (1992) Ecological footprints and appropriated carrying capac-

ity: what urban economics leaves out. Environment and Urbanization, 4,

121–130.Rees, W.E. (1999) The built environment and the ecosphere: a global

perspective. Building Research and Information, 27, 206–220.Rittel, H.W.J. & Webber, M.M. (1973) Dilemmas in a general theory of

planning. Policy Sciences, 4, 155–169.Roberts, N. (2000) Wicked problems and network approaches to resolu-

tion. International Public Management Review, 1, 1–19.Rowe, D.B. (2011) Green roofs as a means of pollution abatement.

Environmental Pollution, 159, 2100–2110.Sadler, J., Bates, A., Donovan, R. & Bodnar, S. (2011) Building for biodi-

versity: accommodating people and wildlife in cities. Urban Ecology:

Patterns, Processes, and Applications (ed. J. Niemel€a) pp. 286–297.Oxford University Press, Oxford.

Shanahan, D.F., Miller, C., Possingham, H.P. & Fuller, R.A. (2011) The

influence of patch area and connectivity on avian communities in urban

revegetation. Biological Conservation, 144, 722–729.Smetak, K.M., Johnson-Maynard, J.L. & Lloyd, J.E. (2007) Earthworm

population density and diversity in different-aged urban systems.

Applied Soil Ecology, 37, 161–168.Smith, R.M., Gaston, K.J., Warren, P.H. & Thompson, K. (2005) Urban

domestic gardens (V): relationships between landcover composition,

housing and landscape. Landscape Ecology, 20, 235–253.

Smith, R.M., Warren, P.H., Thompson, K. & Gaston, K.J. (2006) Urban

domestic gardens (VI): environmental correlates of invertebrate species

richness. Biodiversity and Conservation, 15, 2415–2438.Snep, R. & Opdam, P. (2010) Integrating nature values in urban planning

and design. Urban Ecology (ed K.J. Gaston), pp. 261–286. Cambridge

University Press, Cambridge.

Su, S., Xiao, R., Jiang, Z. & Zhang, Y. (2012) Characterizing landscape

pattern and ecosystem service value changes for urbanization impacts at

an eco-regional scale. Applied Geography, 34, 295–305.Sulston, J., Bateson, P., Biggar, N., Fang, C., Cavenaghi, S., Cleland, J.,

Cohen, J., Dasgupta, P., Eloundou-Enyegue, P.M., Fitter, A., Habte,

D., Harper, S., Jackson, T., Mace, G., Owens, S., Porritt, J., Potts, M.,

Pretty, J., Ram, F., Short, R., Spencer, S., Xiaoying, Z. & Zulu, E.

(2012) People and the Planet: The Royal Society Science Policy Centre

Report 01/12. The Royal Society, London.

Supit, I., van Diepen, C.A., de Wit, A.J.W., Kabat, P., Baruth, B. &

Ludwig, F. (2010) Recent changes in the climatic yield potential of

various crops in Europe. Agricultural Systems, 103, 683–694.The Royal Commission on Environmental Pollution (2007) The Urban

Environment. The Stationery Office, London.

Tompkins, M. (2006) The edible urban landscape. MSc thesis Available at:

http://www.cityfarmer.org/MikeyTomkins_UA_thesis.pdf (accessed July

2012).

Tratalos, J., Fuller, R.A., Warren, P.H., Davies, R.G. & Gaston, K.J.

(2007) Urban form, biodiversity potential and ecosystem services.

Landscape and Urban Planning, 83, 308–317.Tremblay, M.A. & St. Clair, C.C. (2011) Permeability of a heterogeneous

urban landscape to the movements of forest songbirds. Journal of

Applied Ecology, 48, 679–688.United Nations (2008) World Urbanization Prospects. The 2007 Revision.

United Nations, New York.

Uy, P.D. & Nakagoshi, N. (2007) Analyzing urban green space pattern

and eco-network in Hanoi, Vietnam. Landscape and Ecological

Engineering, 3, 143–157.Vergnes, A., Le Viol, I. & Clergeau, P. (2012) Green corridors in urban

landscapes affect the arthropod communities of domestic gardens.

Biological Conservation, 145, 171–178.Verweij, M., Douglas, M., Ellis, R., Engel, C., Hendriks, F., Lohmann, S.,

Ney, S., Rayner, S. & Thompson, M. (2006) Clumsy solutions for a

complex world: the case of climate change. Public Administration, 84,

817–843.Wackernagel, M., Kitzes, J., Moran, D., Goldfinger, S. & Thomas, M.

(2006) The ecological footprint of cities and regions: comparing resource

availability with resource demand. Environment and Urbanization, 18,

103–112.World Resources Institute (2007) Available at: http://earthtrends.wri.org/

index.php

Zhou, W., Huang, G., Pickett, S.T.A. & Cadenasso, M.L. (2011) 90 years

of forest cover change in an urbanizing watershed: spatial and temporal

dynamics. Landscape Ecology, 26, 645–659.

Received 29 November 2012; accepted 11 March 2013

Handling Editor: Julia Jones



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