Urban Ecology

Urban Ecology and Special Features of Urban EcosystemsAuthor(s): Franz RebeleSource: Global Ecology and Biogeography Letters, Vol. 4, No. 6 (Nov., 1994), pp. 173-187Published by: WileyStable URL: http://www.jstor.org/stable/2997649 .

Accessed: 23/09/2014 21:13

Your use of the JSTOR archive indicates your acceptance of the Terms & Conditions of Use, available at .http://www.jstor.org/page/info/about/policies/terms.jsp

.JSTOR is a not-for-profit service that helps scholars, researchers, and students discover, use, and build upon a wide range ofcontent in a trusted digital archive. We use information technology and tools to increase productivity and facilitate new formsof scholarship. For more information about JSTOR, please contact [email protected].

.

Wiley is collaborating with JSTOR to digitize, preserve and extend access to Global Ecology andBiogeography Letters.

http://www.jstor.org

This content downloaded from 141.213.236.110 on Tue, 23 Sep 2014 21:13:11 PMAll use subject to JSTOR Terms and Conditions

http://www.jstor.org/action/showPublisher?publisherCode=black

http://www.jstor.org/stable/2997649?origin=JSTOR-pdf

http://www.jstor.org/page/info/about/policies/terms.jsp


Global Ecology and Biogeography Letters (1994) 4, 173-187

A *1;;';U1 i I*I'I

Urban ecology and special features of urban ecosystems F R A N Z R E B E L E Institute of Ecology, Technical University of Berlin, Schmidt-Ott-Str. 1, D-12165 Berlin, Germany

Abstract. The paper deals with urban ecology as a biological science and applies some of the topics of general importance in ecology to the special conditions found in towns and cities. I consider whether cities should be treated as one integrated ecosystem, or as an assemblage of various ecosystems. In contrast to the holistic, organismic concept of the ecosystem as a new hierarchical level of organization and as an evolving whole which guides the development of the species, I follow the methodological definition of Tansley (1935), who defined ecosystems as 'mental isolates' for 'the purpose of study'. According to Evans (1956) ecosystems can be defined at every level of the biological organization, at the level of the organisms, populations or communities.

The introduction of species from other biogeographical regions is a worldwide phenomenon, but the proportion of successfully established introduced species is higher in cities than in rural or forest areas. This is due to numerous colonizing species which fit the anthropogenous habitats. Due to unequal rates of immigration and extinction of species, urban habitats show an imbalanced turnover of species.

Another special feature of urban ecology is man-in- duced disturbance, which initiates the colonization of disturbed or newly created habitats. According to the

type of substrate and the availability of diaspores there may be both primary, secondary or intermediate types of succession. Besides disturbance, the main component for structuring communities is biological interactions. In this paper I discuss some aspects of competition, predation and mutualism.

The special feature of higher species' richness of cities compared with ecosystems in the countryside can be explained by the high habitat diversity of urban and industrial areas. Although some components which contribute to the complexity of communities, such as competition, are of minor importance in various urban habitats, there may be communities of high complexity.

I also consider community characteristics such as stability and productivity. Since most urban communities are in a state of inequilibrium, theories of stability based on equilibrium are inadequate for urban ecosystems. The productivity of the 'ecosystem city' mainly depends on the area of unsealed open space and the successional stage of the plant communities of the various habitats.

Key words. Ecosystem city, biological invasions, colonization, diversity, stability.

INTRODUCTION

Urban ecology has become an expanding field of research during the last two decades. Various studies were carried out on urban climate, soils, flora, fauna, urban habitats and green space of cities (Berry & Kasarda, 1977; Bomkamm, Lee & Seaward, 1982; Duvigneaud & Denayer-de Smet, 1977; Exline, Peters & Larkin, 1982; Gill & Bonnett, 1973; Gilbert, 1989; Hollis, 1991; Kieran, 1982; Klausnitzer, 1993; Laurie, 1979; Sukopp, 1990; Sukopp & Hejny, 1990; Sukopp

& Wittig, 1993; Wittig, 1991; for reference see also Sukopp & Werner, 1982; Sukopp et al., 1986; Dawe, 1990). Most of this work has been done to solve practical problems of urban planning, environmental protection and urban wildlife conservation. On the other hand, theoretical and conceptual work in urban ecology was hardly stressed (Trepl, 1994).

Urban ecology is both a practical science dealing with the environment of people living in towns and cities, and the associated 'environmental problems' such as water, air and soil pollution, extrac-

173



174 Franz Rebele

tion of drinking water, transport planning, noise, etc., and a biological science as well. As a sub-discipline of ecology, urban ecology is concerned with the distribution and abundance of plants and animals in towns and cities. As with other ecological disciplines it is possible to distinguish between organizational levels of individual organisms, populations and communities.

The practical aspect of urban ecology is not dealt with here. In the following, I discuss some ecological concepts and principles as they apply to urban ecosystems.

ECOSYSTEM CITY AND URBAN ECOSYSTEMS

A question often asked in urban ecology is whether towns and cities can be viewed as 'ecosystems'. In 'The use and abuse of vegetational concepts and terms', Tansley (1935) defines an ecosystem as a system which takes account both of the organisms and the entire complex of physical environmental factors. Ecosystems are 'mental isolates', set up for 'the purpose of study' (Tansley, 1935, p. 300). This is primarily a methodological definition, and the limits of the ecosystem are drawn by the scientists involved; it can form both a part of larger systems, can overlap with other ecosystems or can interact with these.

According to Evans (1956), the definition of ecosystems as the basic unit in ecology does not mean its definition as a new hierarchical level in the organization of communities. Ecosystems can be defined at every level of biological organization, at the level of the organisms, populations or communities, and he recommends that the level be indicated by an appropriate combination (community ecosystem, population ecosystem).

Begon, Harper & Townsend (1990) also argue that no ecological system, whether individual, population or community, can be studied in isolation from the environment in which it exists. Therefore they do not distinguish any separate ecosystem level of organization. Domains of ecosystem researchers such as nutrient dynamics and energetics are dealt with in connection with the structure and function of communities. This approach takes due account of the fact that ecology as a science investigates the interactions between organisms and their environment, and also of the fact that an environment considered in isolation from the organisms whose environment it fornms is an empty concept.

In contrast to the view that communities or (community) ecosystems can be described by the sum of their parts (i.e. the population of species) plus their interactions, the ecologists of the holistic school see ecosystems as an important hierarchical level of organization of life on earth (c.f. Odum, 1971; Pomeroy & Alberts, 1988.)

This view of ecosystems is in the tradition of Clements' organismic concept of the community against which Tansley (1935) originally wrote. Ecosystems are seen as evolving entities, which guide the development of the species and which in principle are independent of the species combination. These ecosystems are attributed properties similar to those of organisms, such as self-regulation. At the core of this direction of ecosystem research is the energetics of ecosystems. (On the discussion of this, see e.g. McIn- tosh, 1980; O'Neill et al., 1986, Trepl, 1988).

Looking at the parameters discussed in ecosystem research, there are none which cannot be explained at the organizational level of the communities and populations forming them, e.g. species richness, diversity, productivity, energetics of the system, stability, resilience). The level of 'self-regulation' (if any) can be explained by the activity of the organisms and by the various forms of biological interaction in communities.

The numerous types of interaction within a community and their physical environment mean that there is positive and negative feedback in (community) ecosystems, which can have a different specific form in each community. However, self-regulation as a sort of steering force does not exist; this is an idea which arises from an organismic view of ecosystems. The dilemma of holistic ecosystems research is clearly demonstrated in cities, since most urban ecosystems, in particular the 'ecosystem city', are not self-regulating, but are 'regulated' by humans. Some people reject the ecosystem concept completely because it is so closely associated with the holistic approach to ecosystem research.

In terms of Tansley's original definition it is possible in the case of cities to define both individual urban 'ecosystems' (e.g. the centre, parks, wasteland) as well as 'the ecosystem' of the entire city (e.g. Brussels). The important thing is not the mere use of the term ecosystem, but the way in which it is used. The 'ecosystem' of a specific city is in fact a complex, consisting of various communities which can overlap and interact to a greater or lesser extent. A city park itself can be divided into various sub-communities



Urban ecology 175

(e.g. lawns, meadows, woodlands, streams, ponds), which can in turn interact.

The variety of urban activities and land uses means that the boundaries between various sub-habitats are particularly marked, but individual populations and communities can stretch over several of these sub- habitats. The boundaries of a survey area ('an ecosystem') are laid down by the scientist more or less arbitrarily, depending on the topic under study and on practical considerations. Thus the investigation of the energy balance of an entire city will make use of the politically defined city boundaries, even if this en- closes non-urban settlements. In other cases, such as the comparison of species distribution between built- up and rural areas, other boundaries are ecologically more appropriate.

MIGRATION, DISPERSAL AND EXTINCTION OF SPECIES

A significant and far-reaching aspect of global ecological change is the elimination of biogeographical barriers to the dispersal of species as a result of human activities (Elton, 1958; Drake et al., 1989; D'Antonio & Vitousek, 1992). The result of this is that species are migrating widely between continents and islands, meeting species they would otherwise never have encountered. This applies in particular to habitats which have been considerably altered, or indeed created by human activity. Thus it is much more common in cities than in natural habitats to find communities consisting of species without a common evolutionary past which have not previously been found elsewhere in such a combination.

A relatively small proportion of the species inten- tionally or unintentionally introduced by humans from other geographic regions successfully disperse and establish themselves in relatively natural communities (see Lohmeyer & Sukopp (1992) for agriophytes in Central Europe, D'Antonio & Vitousek (1992) for grasses globally).

In addition to overcoming geographical barriers, anthropochorous dispersal can also overcome ecological barriers when species meet under new conditions. Thus Urbanska (1992) reports about the two thistles Carduus acanthoides L. and Carduus nutans L. in Ontario, which differ in their ecological demands. C. acanthoides prefers in Ontario skelletal, permeable soils, whereas C. nutans prefers moister sites. Anthro- pogeneous dispersal meant that both species, which

had been separated by ecological preferences, met on a site created by human activity - a roadside embank- ment - where they crossed. By back-crossing, the tol- erance range of C. acanthoides was considerably increased.

In the literature on introduced species and their distribution, the terms 'invasion' and 'colonization' are often used synonymously (e.g. Ehrlich, 1986, 1989). With regards to both the anthropochorous species from other biogeographic regions, and also to the process of colonization in primary successions (and in part in secondary successions) it is, however, sensible to distinguish between the terms, or at least to differ- entiate between the processes covered by the term 'biological invasion'.

For plants, Bazzaz (1986) distinguishes between .colonizers', 'immigrants', and 'invaders'. Colonizers settle open sites in primary succession, and in part in secondary succession. A common feature of all colonizers is that the site they settle is not yet occupied by other species, and that there is initially no competition. If there are repeated disturbances it may not come to competition between species at all. This is a common situation in urban and industrial areas.

'Immigrants' are species which do not significantly impede populations already to be found on a site, and which integrate themselves into the existing community. 'Invaders' are species which penetrate natural, intact phytocoenoses, and either dominate these or displace other species. Such situations are very im- probable in urban habitats since these are either influenced or created by human activity.

The characteristics described by Bazzaz (1986) for immigrants and invaders corresponded more or less to the processes of 'integration' and 'displacement', or the behaviour of agriophytes in 'unsaturated' and ,saturated' plant communities as described by Sukopp (1962). It is, however, appropriate to examine sepa- rately the process described as colonization, and to separate this clearly from the processes of 'integration' and 'displacement'.

Of the many species introduced by human activity from other biogeographic regions, only some are successful (see Lohmeyer & Sukopp, 1992 for ferns and flowering plants in Central Europe, Simberloff, 1986 for insects). According to Lohmeyer & Sukopp (1992), in near-natural vegetation in Central Europe less than 2% of the introduced plant species have established themselves. (On the question of 'Resist- ance of plant communities to biological invasion' see also Trepl, 1990, 1994; Trepl & Sukopp, 1994.)



176 Franz Rebele

In cities, the proportion of successfully established introduced species in the flora is higher than in near- natural habitats (Kowarik, 1991). These belong mainly to the group of colonizers, or are integrated in plant communities which have been newly formed under urban conditions (many ruderal plant communities).

Not only the introduction and dispersal of species which were not previously found in an area (including not only 'alien' species but all synanthropic species), but also the local extinction of species is typical for towns and cities. In particular this affects species from communities on natural or near-natural sites, i.e. habitats which are classically non-urban.

The reasons for this are closely linked to urban development. Construction work directly destroys habitats, eliminating local populations (see Rebele, 1991). Other processes, such as the lowering of ground water levels or eutrophication, can change site factors and thus lead to changes in the abundances of populations, even to the extent of elimination (see Sukopp, 198 1).

However, where there are Ino alterations of the habitat, direct displacement by invasive competitors leading to local extermination of species is very un- likely in cities. Firstly, in cities there are a large number of habitats in which resources are not limited, or in which the low population density means that there is no competition for limited resources. This is the case for a number of plant species which settle open habitats (colonizers, see above), but also for the large group of phytophagous insects. On the other hand, a precondition for the exclusion of competition is stable conditions, which are rarely, if ever, achieved in urban habitats (see also stability and equilibrium, p. 182).

According to MacArthur & Wilson (1967), islands have a natural turnover of species. The number of species on islands is determined by an equilibrium between immigration and extinction. For cities or individual urban communities the turnover of species ought to be particularly high at least for certain groups of organisms, since the conditions exist for both immigration (in the wider sense) and local extermination. The processes underlying the turnover of species, however, are not determined primarily by the ability of the species to disperse and by biological interactions, but by introduction, dispersal and local extermination by human activity. In contrast to the assumption of stability in the island model of MacArthur & Wilson (1967), there is not generally a

balance of species richness in cities. The numbers of species can either increase or decrease.

In Zurich the number of flowering plant species increased from 970 to 1200 species over the last 150 years. One hundred and thirty species died out, but about 350 introduced species became established (Landolt, 1992). A decline in the absolute number of species may occur, if besides the process of extermination of species in near-natural habitats habitat diversity there is diminished through enforced sealing of wasteland or levelling of sites (Rebele, 1991). For Bochum (Ruhrgebiet) a decline of the flora of vascu- lar plants from 600 species in 1887 (Hamann, 1976) to about 580 in 1985 (Schulte, 1985) can be recog- nized.

HABITATS, ENVIRONMENTAL FACTORS AND RESOURCES

In urban and industrial areas it is possible not only to find many 'exotic' species, but also substrates and habitats which are alien to the original states. The introduced substrates can be both natural substrates from other geographical and geological regions (e.g. railway ballast, mine spoil), or artificial substrates (e.g. slag), or a mixture of artificial and natural substrates (e.g. slag/clay). In extreme cases earth can be transported even between continents (Lindroth, 1957).

A large number of urban habitats such as buildings, roofs, underground pipes are specifically related to human habitation. Natural soils are usually considerably altered by human settlement activities and production (soil is moved, substrates are mixed, thre is compaction and sealing over). An overview of the properties and the classification of soils in cities is given by Hollis (1991). Various technological substrates of urban and industrial sites are listed by Meuser (1993).

A feature of many cities is the wide variety of environmental conditions. Soils in cities can be very poor in nutrients, or may be highly enriched (e.g. former sewage farms). Toxicity can play a role in places, due to deposits of toxic substances and substrates, or as a result of pollution. There are open sites with high relative irradiance and very dark habitats (artificial caves). Habitats can also be anything from very dry to wet, although dry sites are more common in urban habitats. For plants the relationship between light and soil resources is of vital importance. Sub- strates with low levels of available nutrients tend to



Urban ecology 177

offer ample light, whereas if soil resources are plenti- ful there is usually a lack of light at ground level (Tilman, 1986).

The urban climate is generally warmer than that in the surrounding countryside. Even more significant for organisms than the annual mean temperature, however, are microclimate variations. Temperatures, for example, can vary extremely widely over short distances. Abiotic site factors can be relatively homogeneous over areas (e.g. large excavations or embank- ments), but also very heterogeneous (e.g. industrial wasteland).

Extreme abiotic site conditions (e.g. lack of light or water) are an important selection factor for the establishment of many plant and animal species. Begon et al. (1990), however, rightly point out that 'extreme' sites are only 'extreme' to the outside observer. Lim- ited resources can be a stress factor for some species, leading to reduced productivity. For tolerant species, on the other hand, the conditions can be near ideal, since they do not have to face the competition of other species for resources.

DISTURBANCE

According to White & Pickett (1985), a disturbance is a relatively discrete event which suddenly disrupts the structure of an ecosystem, community or population, changing either the availability of resources or the physical environment. This is a very broad definition, but it says nothing about the cause or the nature of the disturbance.

Disturbances, can have not only physical but also biological causes and generally lead to the death or reduction in abundance of a species, which can in turn favour other species. Thus Sousa (1984) speaks of 'discrete, punctuated killing, displacement, or damag- ing of one or more individuals (or colonies) that directly or indirectly creates an opportunity for new individuals (or colonies) to become established' (p. 356).

However, disturbances in ecosystems do not necessarily lead to the death of organisms. Soil disturbances, for example, can improve the conditions for germination, or increase the availability of nutrients. Local eutrophication can also represent a sudden alteration in the availability of resources, although not if this is a continual state. As a rule, however, both physical and biological disturbances involve a de- struction of biomass, the physical generally being

non-specific, but the biological often being specific, namely in the form of predation.

Natural disturbances include storms, fires and land- slides, but also the impact of grazing animals. In towns and cities there are a number of additional disturbances, particularly those linked with construction activities, gardening and recreation. Anthropoge- neous disturbances are usually of much greater importance in urban habitats than natural disturbance factors.

Disturbed areas in cities can vary in size from a few square centimetres to square kilometres. It is particularly common for construction work to destroy communities completely. Other disturbances are localized, and less severe. The extent and the intensity of disturbances has an important influence on its biological effectiveness. Localized disturbances can increase the heterogeneity of the environment, creating a larger number of safe sites for plant species (Harper, 1977; Grubb, 1977). More wide-spread disturbances, such as the removal of top soil prior to construction, can lead to a reduction in the heterogeneity. Frequently recur- ring disturbances can hinder the growth of perennial plant species, and thus also reduce heterogeneity.

The regional frequency of anthropogenous disturbances depends primarily on economic, political and social factors, which means that in cities they are very difficult to predict, since the economy does not develop linearly, social habits and fashions change con- stantly, and political processes are often chaotic.

Various statements have been made, for example, about the lifetime of buildings. 'Turnover rates', however, cannot be generalized for all cities, and can even vary considerably within a city, as well as depending on economic and political developments. It is generally true for urban ecosystems that disturbances do not usually occur in phases, but displaced, and thus various stages of succession exist side-by-side. An important feature of species in urban habitats is that they are able either to regenerate themselves after disturbances or to settle in newly created gaps.

Here, typical urban-industrial ecosystems are considered. Special consideration is given to the numerous disturbances to residual near-natural sites within and directly around cities.

COLONIZATION OF OPEN HABITATS

Previously, the process of colonization was considered in connection with introduced plant species.



178 Franz Rebele

However, some of the most abundant and successful colonizers are native species, though these have usually developed their characteristics under different evolutionary circumstances. As a rule these are species from naturally disturbed habitats, e.g. river banks.

In the process of colonization, newly created habitats are settled, or sites are re-settled following a disturbance. It is not usual in the case of a secondary succession to distinguish between a settlement by migrant diaspores from elsewhere and the recruitment of plant populations from the diaspore bank. Thus for Grubb (1987), for example, colonizers are simply 'the first organisms to become established in a succession' (p. 81).

The new colonization of open sites is primarily dependent on the number of diaspores reaching a site or already to be found there, and on the physical- chemical environmental factors. For example, plant populations must be able to germinate and establish themselves. Competition within and between species does not occur as long as relevant population densities are not exceeded. Other forms of biological interaction can play a role, such as predation by seed-eating birds or insects. In addition to the establishment of new organisms from seed etc., some plant species can also settle sites by vegetative spread.

In cities, open habitats (those not already occupied by other species) are relatively common. This is due to the creation of new habitats and the disturbances already considered in the previous two sections. The colonizers of open sites use up the existing soil nutrients and thus create zones of resource depletion. On the other hand they also produce organic debris which returns nutrients to the soil. In this way they produce a spatial heterogeneity of plant nutrients even in relatively homogeneous substrates.

Animals (including e.g. ants or earth wasps) can also contribute significantly to the restructuring of the micro-habitats of open sites such as sand heaps.

SUCCESSIONS

Successions are directed changes of the species composition over time. Vegetation successions are generally also linked to a clear change of the vegetation structure. A distinction is usually made between primary and secondary successions.

Primary successions occur naturally on sand dunes, glacial debris or volcanic ash. Due to man's industrial activities, primary successions can also occur in cities.

They occur wherever natural soil is excavated or a substrate is spread which did not previously bear any plant growth and thus is devoid of any diaspore bank. The content of organic substances and nitrogen is initially very low in a primary succession. Other nutrients are not limited in many substrates (Tilman, 1986). In the course of primary succession, organic matter and nitrogen accumulates. The level of available phosphates, on the other hand, can decline (Vi- tousek & Walker, 1987).

There is usually an increase in the number of species in the first years of primary vegetation succession, which then levels off in the course of succession (Crawley, 1986). In cities the settlement of primary habitats by plants is often more rapid than in isolated natural primary successions, since there are usually large numbers of diaspores of colonizing species in the near vicinity. Birds perching in adjacent gardens, parks, road sides or wasteland can also accelerate the colonization process (van der Pijl, 1982; McDonnell & Stiles, 1983).

Secondary successions occur when sites which already bore vegetation cover are disturbed (storm dam- age in forests), or when cultivated land is left to lie fallow. In cities, secondary successions are particular frequent, since both disturbances and also temporary fallow phases are very common. Features of secondary successions are the presence of organic matter at the start and an increased release of nutrient after a disturbance of the vegetation cover (or direct enrich- ment with dead organic matter).

The vegetation development on deposited top-soils and landfill-soils in cities usually has the character of a secondary succession, or a mixed-form of secondary and primary successions. This is because substrates are often used which have already had vegetation cover (soil recycling), or which are enriched with organic matter and diaspores (e.g. compost enriched top-soils).

A further important feature of secondary successions is the 'initial floristic composition factor' (Egler, 1954). Most species which play a part in the course of succession (annuals, short- and long-lived perennials), are already present at the start of the succession, or migrate to the site from adjacent areas soon after the disturbance or the deposition of the substrate. De- pending on the nutrient content of the soil, the vegetation develops faster or slower, or one can recognize successional stages (dominance of annuals, mono- carpic perennials, polycarpic perennials). Secondary succession on urban deposit soils usually have the



Urban ecology 179

greatest number of species at the beginning of the succession. The higher the level of nitrogen in the applied substrate the more rapid is the decline in the species density, since in highly productive plant communities the lack of light at the soil surface hinders both the early successional species as well as the woody perennials (Rebele, 1992).

In urban and industrial areas, the frequency of disturbances means that early- and mid-successional stages are common. Late successional stages, on the other hand, are very rare (e.g. birch forests on mine spoil heaps in the Ruhrgebiet, or some 'urban forests' on ruderal sites in Berlin).

BIOLOGICAL INTERACTION

Most communities are organized, according to Sousa (1984), by biological interactions such as competition and predation in the wider sense of the term, as well as by disturbances (see also Begon et al., 1990). Mutualism can also be important. The relative importance of the various forms of biological interaction can, however, undergo a systematic change, with competition and predation being less important in mechanically disturbed habitats. Disturbances have already been considered, see page 177. Some forms of biological interaction are considered here in more detail.

(a) Competition

There are several different kinds of competition. Re- source or exploitative competition, which is particularly important for plants, requires niche differentiation if it is not to lead to competitive exclusion. Interference competition and competition for space, on the other hand, do not require niche differentiation (Yodzis, 1986).

Exploitative competition occurs at the same trophic level and the same resources demands. According to the classical competition theory, two species with exactly the same resource requirements cannot co-exist (Gause, 1934). The co-existence of species at the same trophic levels can be explained by niche differentiation, which is either produced by current competition, or already exists because of evolutionary avoidance of competition in the past ('ghost of competition past', Connell, 1980). A feature of co-existing species in the case of niche differentiation is that the intra-specific competition is more important for the

dominant species than the inter-specific competition. Some urban habitats demonstrate considerable het-

erogeneity over time and space. For autotrophic plants, light and nitrogen, for example, are the two crucial factors which can be at a minimum and which are the subject of competition. Even within relatively homogeneous old fields the nitrogen content in an area of 0.14 ha can vary be a factor of 3-6.7 between samples. During secondary succession, the nitrogen content can also vary by a factor of 6 within 120 years (Tilman, 1986).

The species which grow together on fallow land need not form part of a converging community. Each species can successfully compete with other species along a nutrient/light gradient (Tilman, 1986). Even when there is exploitative competition, the spatial and temporal environment heterogeneity can ensure the coexistence of competing species.

Niche differentiation in the narrow sense of the term is, however, more of an ideal case, and requires a relatively stable environment and a long process of joint evolution, or evolution under similar environmental conditions (Crawley, 1987). Even under natural conditions, without human influence, this is not always so, or is controversial (see Hubbell & Foster, 1986). Under urban conditions it is even less the case, since unforeseeable disturbances make the environment extremely unstable.

In the case of competition which requires no niche differentiation or environmental heterogeneity, the initial density of a species can be of significance, or first occupancy of the site (founder effect, see e.g. Schmidt, 1981; Rebele, 1994). Since gaps are repeat- edly being created in urban communities by physical disturbances (in part also by predation), there are always competition-free situations which allow weaker species to co-exist with superior competitors.

(b) Predation

Of the various forms of predation, herbivory is considered here.

There are numerous examples of 'empty niches' for herbivores in urban habitats in particular. Competition between herbivores generally plays a less important role, particularly in the case of phytophagous insects (Hairston, Smith & Slobodkin, 1960; Slobodkin, Smith & Hairston, 1967; Strong, Lawton & South- wood, 1984; Jermy, 1985). Investigations of herbivorous insects on bracken (Pteridium aquilinum (L.) Kuhn) on three continents have shown that some



180 Franz Rebele

feeding sites of this wide-spread, cosmopolitan species are exploited in one part of the world but not in another, since the insect fauna differs (Lawton, 1984).

Comparisons are frequently made between the phytophagous fauna of native and exotic species. It is known that neophytic species in their anthropochoric areas usually have a lower number of herbivorous insects than in their area of origin (Lohmeyer & Sukopp, 1992). It has also been established for entire plant groups, such as woody perennials, that exotic species represent a poorer source of nutrition for animals than native species (Southwood, 1961; Kennedy & Southwood, 1984; Kowarik, 1986). How- ever, the difference is probably not as large as is generally supposed (see, e.g. Glauche, 1991). Thus Owen (1991), after observing her garden in Leicester for 15 years, concluded that only 33.7% of all native plant species served as a source of food for moth species, compared with 42.4% of all exotic species. However, of a total of sixty-eight moth species, forty- six fed on native species compared with thirty-eight on exotic species.

There are a number of possible reasons for the difference in the phytophagous fauna, and urban habitats offer ample opportunity to investigate these. The feeding on exotic plant species by phytophagous insects can depend, for example, on the length of their presence in the new area, the distance to the area of autochtonous distribution, the frequency of the species in their area of origin as well as in their anthropogenous area, the total size of the area, the taxonomic proximity of the species to native species, or the taxonomic status of the species (families with numerous species also have a relatively large number of phytophages). One of the crucial factors is probably the existence of transport and trade routes, and the frequency and volume of the exchange of goods (see e.g. Simberloff, 1989).

It is certainly misleading to compare native plant species with rich phytophagous insect fauna with exotic species which are poor in phytophagous fauna. There are also large differences between native species as far as the numbers of phytophagous insects is concerned (e.g. Quercus species/Taxus baccata L. in Central Europe). Some introduced species have low numbers of phytophagous insects in their place of origin also (e.g. Ailanthus altissima (Mill.) Swingle, after Rohricht, 1991). As the example of Pteridium aquilinium shows, the use of plant species can also vary widely over its area of natural distribution.

Of the other herbivores, common urban species are of particular interest, such as rabbits in many central European cities. Herbivores cause biological 'disturbances' (see above) amongst the plant species, and can thus affect biological interactions such as competition. It is known that rabbits can have a considerable influence on the species diversity of chalk grassland (Tansley & Adamson, 1925).

Many carnivorous species change their feeding habits in towns and cities. A well-known example is for fox populations in Britain (Harris, 1986). But population densities and life expectancy can also be affected, and with them the age structure of populations (Harris, 1986).

(c) Mutualism

The majority of the earth's biomass consists of organisms which have mutualist relationships to others (Begon et al., 1990). Well-known examples of mutualism are the symbiosis of legumes with nitrogen- fixing bacteria. Legumes also play an important role in urban phytocoenoses, e.g. on N-poor sandy soils. They are not always among the first colonizers, but they are among the species which form the dense ground cover.

In natural ecosystems, mycorrhiza is the rule rather than the exception. In some families, e.g. Brassi- caceae, Carvophvllacea and Chenopodiaceae, mycorrhiza is rarer, but in the case of woody species, grasses and ferns it is widespread. The VA-mycorrhiza common in some families serves mainly the uptake of phosphates (Harley & Harley, 1987). Phos- phates often act as a constraint for later successional stages (see above).

Annuals are frequently without mycorrhiza. This is true both for annuals on ruderal sites rich in nitrogen, as well as for therophytes on dry sandy lawns and spoil heaps (e.g. Cerastium semidecandrum L., Chae- narrhinum minus (L.) Lange, Minuartia verna (L.) Hiern (see Harley & Harley, 1987). It is still question- able whether mycorrhiza plays a role right from the start in the colonization of urban habitats by grasses and woody perennials.

Where soil is removed and stored, the microbial activity and the proportion of VA-mycorrhiza declines and this also delays the development of mycorrhiza in grasses, or reduces its level (Rimmer, 1991). Reeves et al. (1979) report that the level of mycorrhiza on



Urban ecology 181

severely disturbed sites is lower than on natural, undisturbed sites.

Overall, it seems that symbiosis with nitrogen- fixing bacteria plays an important role for urban habitats in view of the generally low levels of nitrogen, whereas mycorrhiza plays a less important role than, for example, in forests.

(d) Interactions between interactions

Thompson (1988) has pointed out that there are no sharp distinctions between the various forms of biological interaction, that competition and mutualistic interactions overlap. Thus, if competing plants blos- som at the same time they will attract more pollina- tors, to the advantage of both species. The symbiosis between legumes and N-fixing bacteria is only of advantage for legumes on N-poor soils, but not on N-rich soils. The results of interaction can also be dependent on environmental factors. Several references have already been made to predation as a biological 'disturbance' which reduces competition.

SPECIES RICHNESS AND DIVERSITY

Species richness can be taken to mean the number of species per unit area (species density), or in a specific location (numerical species richness). When calculat- ing diversity, the evenness plays a role, and it can be expressed using various indices (Hurlbert, 1971; Peet, 1974). After Whittaker (1965), the diversity can be divided into several components. Alpha-diversity refers to the diversity within communities, beta-diversity to the rate of change in the species cornposition along a habitat gradient (diversity between communities), and the gamma-diversity to the overall diversity of the landscape (see also Peet, Glenn-Lewin & Walker Wolf, 1983).

The alpha-diversity in urban phytocoenoses can vary from very low to very high. Examples of plant communities with low diversity can be found in extreme soil conditions such as very low or very high pH, on toxic substrates or also under eutrophic conditions (e.g. dominant stands of Calamagrostis epigejos (L.) Roth or Urtica dioica L. Examples of high alpha- diversity are to be found in some ruderal plant communities with a species density of more than 40 species per square metre with a low dominance of individual species. If one compares various phyto-

coenoses in urban habitats, then it is not rare to find plant communities with very low diversity as close neighbours of communities with high diversity.

The beta-diversity can be very high in cities, since very pronounced gradients are frequently formed there. There are, however, also some very uniform areas. The fact that cities are frequently richer in species than the surrounding areas (at least as far as plant species and some groups of animals are concerned, see Klausnitzer, 1993; Landolt, 1992; Reich- holf, 1989; Sukopp & Trepl, 1993, Wittig, 1991) can be attributed above all to the great habitat diversity. The relationship between introduced species which are newly establishing themselves to species facing local extermination also plays a role.

With reference to both the habitat diversity and also the species turnover in cities, there are not only general characteristics, but also geographical and historical components. Scandinavian cities differ from central European cities, and the latter from Mediter- ranean cities, not only in their architecture, but also in the species composition. Within central Europe, cities of comparable size can also differ in their species abundance, with historical components (urban development, trade connections, industrial development etc.) providing a degree of 'local character' (Gilbert, 1989).

In addition to (spatial) habitat diversity, the time factor also plays an important role, with various successional stages side-by--side. References were made to vegetation succession earlier (p. 178). Little is known as yet about succession of soil communities and animal communities in urban ecosystems.

For some animal groups it is not the species richness of plant communities which is most important, but the structural diversity of vegetation, and/or the spatial heterogeneity of the non-living environment (e.g. for birds or lizards).

It should be noted that the floral diversity can either increase or decrease in the course of successions, depending on the type of successions. Regarding soil resources, the highest diversity is generally found on sites of moderate fertility (Tilman, 1982). The size and nature of disturbances also have an influence on species diversity, with the highest levels frequently being found where disturbances are intermediate, an observation not only made in towns and cities ('intermediate disturbance hypothesis', Connell, 1978). Ob- jections have, however, been voiced against making generalizations in this context (Yodzis, 1986; Leigh, 1 990).



182 Franz Rebele

COMPLEXITY

Tansley (1935) pointed out that the degree of integration of ecosystems can vary. This depends in the first place on the nature and extent of interactions within the community. According to Pimm (1984) a community has the following components.

1. The species richness. 2. The connectance (the number of inter-specific

interactions as a fraction of all possible inter- specific interactions).

3. The interaction strength (i.e. the effect of the density of a species on the growth rate of another species).

4. The evenness within a community.

Little work has as yet been done on the complexity of urban communities. However, it is not possible to set up a simple rule, for example that urban communities are less complex than non-urban ones. Rather, the few investigations in the literature suggest that there is a whole range from very low to high complexity (Briand, 1983, example of aspen parkland), as does the rich fauna and flora of much wasteland.

On average, the complexity of urban communities is probably less than various non-urban ecosystems (e.g. forests), since biological interactions, especially inter-specific competition becomes less significant given the frequent disturbances.

STABILITY AND EQUILIBRIUM

Tansley (1935) also pointed out that the ecosystems are less stable than, for example, chemical elements, since the ecosystems consist of components which are themselves unstable to a lesser or greater extent - the climate, the soil, and the organisms. The climate changes with time, soils develop, and the composition of organisms can also change, even in natural ecosystems undisturbed by humans. Since, by definition, ecosystems are not enclosed systems, components from other systems can penetrate them and cause them to disintegrate. Under natural conditions a dy- namic equilibrium may be established over decades or centuries in the sequence of successions, but there are also forces which work against this.

A major problem in the discussion of stability is that this was originally defined in terms of natural ecosystems in equilibrium. According to Hurd et al. (1971), stability is a concept adopted fom thermody- namics, and defined as 'the ability of a system to

maintain or return to its ground state after external perturbation' (p. 1134). The definition of Holling (1973) is similar, but using 'equilibrium state' and 'temporary disturbance', and omitting the aspect of maintaining its state: 'Stability, ..., is the ability of a system to return to an equilibrium state after a temporary disturbance. The more rapidly it returns, and with the least fluctuation, the more stable it is.' (p. 17)

Pimm (1984), corresponding essentially with May (1973), defines stability more precisely: 'A system is deemed stable if and only if the variables all return to the initial equilibrium following their being perturbed from it. A system is locally stable if this return is known to apply only certainly for small perturbations and globally stable if the system returns from all possible perturbations'. 'Resilience', for Pimm (1984) expresses, 'how fast the variables return towards their equilibrium following a perturbation', and is therefore defined only for a stable system, but not for unstable ones. 'Resistance', on the other hand, is 'the degree to which a variable is changed following a perturbation' (p. 322).

The variables which play a role in ecosystems are, for example, the number of species, the abundance of the various species, the productivity, element concen- trations or energy levels. Common to the definitions presented here is the assumption of a ground state or equilibrium. For living communities, an equilibrium means that the net growth rate of populations (but also other parameters such as biomass and element con- centration) are zero.

Because of disturbances and climate fluctuations, natural communities are seldom in a state of equilibrium, except perhaps if viewed on a wider spatial level. Chesson & Case (1986) point out that many inequilibrium theories are compatible with the theory of global stability of a community. Local inequilibria can lead, over the average of all local communities to a global stable equilibrium. This also corresponds to the concept of 'shifting mosaic steady-state' presented for natural forest ecosystems by Bormann & Likens (1979).

When considered over larger areas, however, natural systems are also frequently in a state of transition (depending on the extent and the intensity of the disturbance). Connell (1978) gives examples of natural ecosystems (tropical rainforests, coral reefs), which are in states of transition after a disturbance.

Turning to urban ecosystems, this means that states of equilibrium are virtually impossible, since it is highly probable that a perturbed system will be dis-



Urban ecology 183

turbed again before the state of equilibrium has been reached. It will also rarely be possible to speak of a system returning to equilibrium, since the systems disturbed in a town or city will rarely be in a state of equilibrium to start with, but in a state of transition or a state maintained by human intervention (such as regular mowing and fertilizing of lawns).

It follows from this that when assessing populations in urban communities, and biological interactions between them, theories and models which are based on equilibria and stable environments will be less relevant than for systems with relatively constant environments and little disturbance.

A definition of stability which also includes states maintained by humans is given by Gigon (1984): 'Ecological stability is the persistence of an ecological system and its ability to return after alteration to its initial position' (p. 14).

In this definition, the term 'equilibrium' is replaced by 'initial position', so that the return to a state maintained by human activity can be termed 'stability'. According to Gigon (1984), if a meadow was no longer mown then 'instability' would result. However, the succession from a fallow meadow to a forest can be regarded as the tendency towards a natural point of equilibrium. It is therefore essential when discussing stability, to explain what one means by the stability of an ecological system, or which stable state is envis- aged (e.g. the meadow or the 'natural equilibrium').

The question of equilibrium is usually dealt with in conjunction with the complexity and diversity of communities. As has been noted, high diversity is found more frequently in instable communities (Connell, 1978), and complex communities are not necessarily more stable than less complex ones.

PRODUCTIVITY

The productivity of urban ecosystems can vary from very low to very high depending on soil parameters. Low productivity is found, e.g. in plant communities on slag heaps, gravel, or other dry substrates poor in nutrients. Very high productivity can be found in early to mid-successional stages on eutrophic soil deposits. In Brussels, ruderal phytocoenoses such as Tanaceto- Artemisietum vulgaris or Solidaginetum giganteae achieve a yearly productivity of 14-20 tonnes/hectare (Duvigneaud, 1975).

If a city is examined as a whole, then the productiv-

ity depends primarily on the degree of land sealing, or put another way, on the proportion of area which can be settled by green plants. In some cities the unsealed area can be as much as 50% of the total surface area, e.g. Brussels (Duvigneaud & Denayer-de Smet, 1977) or Halle-Neustadt (Frotscher, 1990). Within the cities there are usually different zones for the level of sealing. The city centre is usually more sealed than the outer districts (see e.g. B6cker, 1985).

The standing biomass is not only dependent on the proportion of the area available to plants for settlements, but also on the age of the successional stage. The older the successional stage, then the more biomass is available. Little data has been published about the overall phytomass in cities. The figure of 750,000 tonnes dry matter given for Brussels (Duvi- gneaud & Denayer-de Smet, 1977) corresponds over the entire area of 162 km2 approximately to a phytomass of 46 tonnes/hectare.

DECOMPOSERS AND DETRITIVORES

Decomposers and detritivores are very important for the formation of soils, and they play an even more important role in cities. A special feature of urban ecosystems is that in addition to the production of dead organic matter by the urban organism communities, organic waste is also produced by humans (usually imported from outside). To some extent this is removed again by waste disposal, emissions, sewage systems and run-off into running waters, but there is also an accumulation of nutrients in certain places in the cities from antropogenic sources.

CONCLUSIONS

The topics discussed in this paper are questions of general importance in ecology. Nevertheless, there are some special features of urban ecosystems like mosaic phenomena, specific disturbance regimes, the processes of species invasions and extinctions, which influence the structure and dynamics of plant and animal populations, the organization and characteristics of biotic communities and the landscape pattern as well in a different manner compared with natural ecosystems. On behalf of the ongoing urbanization process, urban ecosystems should attract increasing



184 Franz Rebele

attention by ecologists, not only to solve practical problems, but also to use the opportunity for the study of fundamental questions in ecology.

ACKNOWLEDGMENTS

I would like to thank Prof. Dr Ludwig Trepl (Munich) for his valuable comments on the manuscript, and Richard Holmes for help with the English.

REFERENCES

Bazzaz, F.A. (1986) Life history of colonizing plants: some demographic, genetic, and physical features. E- cology of biological invasions of North America and Hawaii (ed. by H.A. Mooney and J.A. Drake), pp. 96-110. Springer-Verlag, New York.

Begon, M., Harper, J.L. & Townsend, C.R. (1990) Ecol- ogy - individuals, populations, communities, 2nd edn, p. 945. Blackwell Scientific Publications, Oxford.

Berry, B.J.L. & Kasarda, J.D. (1977) Contemporary urban ecology. Collier Macmillan Publishers, London.

B6cker, R. (1985) Bodenversiegelung. Karte zur Oko- logie des Stadtgebietes Berlin (West). Umweltatlas Bd. I (ed. by Senator fur Stadtentwicklung und Umweltschutz). Berlin (West).

Bormann, F.H. & Likens G.E. (1979) Pattern and process in a forested ecosystem. Springer-Verlag, New York.

Bornkamm, R., Lee, J.A. & Seaward, M.R.D. (eds) (1982) Urban ecology. The Second European Ecologi- cal Symposium. Berlini, Septenmber 1980. Blackwell Scientific Publications, Oxford.

Briand, F. (1983) Environmental control of food web structure. Ecology, 64, 253-263.

Chesson, P.L. & Case, T.J. (1986) Overview: nonequi- librium community theories, chance, variability, history and coexistence. Community ecology (ed. by J. Diamond and T.J. Case)., pp. 229-239. Harper & Row, New York.

Connell, J.H. (1978) Diversity in tropical rainforests and coral reefs. Science, 199, 1302-1310.

Connell, J.H. (1980) Diversity and the coevolution of competitors, or the ghost of competition past. Oikos, 35, 131-138.

Crawley, M.J. (1986) The structure of plant communities. Plant ecology (ed. by M.J. Crawley), pp. 1-50. Black- well Scientific Publications, Oxford.

Crawley, M.J. (1987) What makes a community invasi- ble? Colonization, succession and stability (ed. by A.J. Gray, M.J. Crawley and P.J. Edwards), pp. 429-453. Blackwell Scientific Publications, Oxford.

D'Antonio, C.M. & Vitousek, P. (1992) Biological invasions by exotic grasses, the grass/fire cycle, and global change. Ann. Rev. Ecol. Syst. 23, 63-87.

Dawe, G. (ed.) (1990) The urban environment: a source-

book for the 1990s. The Centre for Urban Ecology, Birmingham.

Drake, J.A., Mooney, H.A., di Castri, F., Groves, R.H., Kruger, F.J., Rejmanek, M. & Williamson, M. (1989) Biological invasions. A global perspective. SCOPE 37. John Wiley & Sons, Chichester.

Duvigneaud, P. (1975) Etudes 6cologiques de l'6cosysteme urbain Bruxellois. Structure, biomasses, min6ralomasses, productivit6 et cation du plomb dans quelques associations ruderales. Bull. Soc. roy. Bot. Belg. 108, 93-128.

Duvigneaud, P. & Denayer-de Smet, S. (1977) L'6cosystem urbs. L'6cosystem urbain bruxellois. Pro- ductivite biologique en Belgique (ed. by P. Duvi- gneaud and P. Kestemont), pp. 581-599. SCOPE. Editions Duculot, Paris-Gembloux.

Egler, F.E. (1954) Vegetation science concepts. I. Initial floristic composition, a factor in old-field vegetation development. Vegetatio, 4, 412-417.

Ehrlich, P.R (1986) Which animal will invade? Ecology of biological invasions of North America and Hawaii (ed. by H.A. Mooney and J.A. Drake), pp. 79-95. Springer-Verlag, New York.

Ehrlich, P.R. (1989) Attributes of invaders and the invad- ing processes: Vertebrates. Biological invasions. A global perspective (ed. by J.A. Drake, H.A. Mooney, F. di Castri, R.H. Groves, F.J. Kruger, M. Rejmanek and M. Williamson), pp. 315-328. SCOPE 37. John Wiley & Sons, Chichester.

Elton, C.S. (1958) The ecology of invasions by animals and plants. Methuen, London.

Evans, F.C. (1956) Ecosystem as the basic unit in ecology. Science, 123, 1127-1128.

Exline, C.H., Peters, G.L. & Larkin, R.P. (1982) The City. Patterns and processes in the urban ecosystem. Westview Press, Boulder, Colorado.

Frotscher, W. (1990) Verfahren der Luftbildinterpretation zur Unterstutzung urbanokologischer Untersuchungen in der Stadtregion Halle. Dissertation A. Sekt. Geogra- phie, Univ. Halle. (Cited from Rohricht 1991).

Gause, G.F. (1934) The struggle for existence. Williams & Wilkins, Baltimore.

Gigon, A. (1984) Typologie und Erfassung der 6kologi- schen Stabilitat und Instabilitat mit Beispielen aus Gebirgs6kosystemen. Verh. Ges. Okologie, 12, 13-29.

Gilbert, O.L. (1989) The ecology of urban habitcats. Chapman & Hall, London.

Gill, D. & Bonnett, P. (1973) Nature in the urban landscape: a study of city ecosystems. York Press, Baltimore.

Glauche, M. (1991) Bedeutung neophytischer Geh6lze fur den Artenreichtum stadtischer und siedlungsnaher Biozonosen. Berliner Naturschutzblatter, 35, 5-16.

Grubb, P. (1977) The maintenance of species richness in plant communities: the importance of the regeneration niche. Biol. Rev. 53, 107-145.

Grubb, P. (1987) Some generalizing ideas about colonization and succession in green plants and fungi.



Urban ecology 185

Colonization, succession and stability (ed. by A.J. Gray, M.J. Crawley and P.J. Edwards), pp. 81-102. Blackwell Scientific Publications, Oxford.

Hairston, N.G., Smith, F.E. & Slobodkin, L.B. (1960) Community structure, population control, and competition. Am. Nat. 44, 421-425.

Hamann, U. (1976) Uber Veranderungen der Flora von Bochum in den letzten 90 Jahren. Abhandl. Landes- mus. Naturkunde Munster, H. 1, 38. Jg., 15-25.

Harley, J.L. & Harley, E.L. (1987) A check-list of mycorrhiza in the British flora. New Phytol. (Suppl.) 105, 1-102.

Harper, J.L. (1977) The population biology of plants. Academic Press, London.

Harris, S. (1986) Urban foxes. Whittet Books, London. (Cited from Gilbert, 1989).

Holling, C.S. (1973) Resilience and stability of ecological systems. Ann. Rev. Ecol. Syst. 4, 1-23.

Hollis, J.M. (1991) The classification of soils in urban areas. Soils in the urban environment (ed. by P. Bul- lock and P.J. Gregory), pp. 5-27. Blackwell Scientific Publications, Oxford.

Hubbell, S.P. & Foster, R.B. (1986) Biology, chance, and history and the structure of tropical rain forest tree communities. Community ecology (ed. by J. Diamond and T.J. Case), pp. 314-329. Harper & Row, New York.

Hurd, L.E., Mellinger, M.V., Wolf, L.L. & McNaughton, S.J. (1971) Stability and diversity at three trophic levels in terrestrial successional ecosystems. Science, 173, 1134-1136.

Hurlbert, S.H. (1971) The nonconcept of species diversity: a critique and alternative parameters. Ecology, 52, 577-586.

Jermy, T. (1985) Is there competition between phytophagous insects? Z. zool. Syst. Evolut. forsch. 23, 275-285.

Kennedy, C.E.J. & Southwood, T.R.E. (1984) The number of species of insects associated with British trees: a re-analysis. J. anim. Ecol. 53, 455-478.

Kieran, J. (1982) A natural history of New York City. Fordham University Press. New York.

Klausnitzer, B. (1993) Okologie der GroJ3stadtfauna, 2nd ed. Fischer Verlag, Jena.

Kowarik, I. (1986) Okosystemorientierte Geholzarten- wahl fur Grunflachen. Das Gartenamt, 35, 524-532.

Kowarik, I. (1991) Beruicksichtigung anthropogener Standort- und Florenveranderungen bei der Aufstel- lung Roter Listen. Rote Listen der gefdhrdeten Pflanzen und Tiere in Berlin (ed. by A. Auhagen, R. Platen and H. Sukopp), pp. 25-56. Landschaftsent- wicklung und Umweltforschung S 6. Technische Uni- versitat Berlin.

Landolt, E. (1992) Veranderungen der Flora der Stadt Zurich in den letzten 150 Jahren. Bauhinia. 10, 149- 164.

Laurie, I.C. (ed.) (1979) Nature in cities. John Wiley, Chichester.

Lawton, J.H. (1984) Non-competitive populations, non- convergent communities, and vacant niches: The herbivores of bracken. Ecological communities: conceptual issues and the evidence (ed. by D.R. Strong, D. Simberloff, L.G. Abele and A.B. Thistle). Princeton University Press, Princeton, NJ. (Cited from Begon et al., 1990).

Leigh, E.G. (1990) Community diversity and environmental stability: a re-examination. TREE, 5, 340-344.

Lindroth, C.H. (1957) The faunal connections between Europe and North America. John Wiley, New York. (Cited from Simberloff, 1989).

Lohmeyer, W. & Sukopp, H. (1992) Agriophyten in der Vegetation Mitteleuropas. Schriftenreihe fur Vegeta- tionskunde 25. Bonn-Bad-Godesberg.

MacArthur, R.H. & Wilson, E.O. (1967) The theory of island biogeography. Monographs in Population Bi- ology 1. Princeton University Press, Princeton, NJ.

May, R.M. (1973) Stability and complexity in model ecosystems. Monographs in population biology 6. Princeton University Press, Princeton, NJ.

McDonnell, M.J. & Stiles, E.W. (1983) The structural complexity of old field vegetation and the recruitment of bird-dispersed plant species. Oecologia, 56, 109- 116.

McIntosh, R.P. (1980) The background and some current problems of theoretical ecology. Synthese, 43, 195- 255.

Meuser, H. (1993) Technogene Substrate in Stadtboden des Ruhrgebietes. Z. Pflanzenerndhr. Bodenk. 156, 137-142.

Odum, E.P. (1971) Fundamentals of ecology. Saunders, Philadelphia.

O'Neill, R.V., DeAngelis, D.L. Waide, J.B. & Allen, T.F.H. (1986) A hierarchical concept of ecosystems. Monographs in population biology 23. Princeton Uni- versity Press, Princeton, NJ.

Owen, J. (1991) The ecology of a garden. The first fifteen years. Cambridge University Press, Cambridge.

Peet, R.K. (1974) The measurement of species diversity. Ann. Rev. Ecol. Syst. 5, 285-307.

Peet, R.K., Glenn-Lewin, D.C. & Walker Wolf, J. (1983) Prediction of man's impact on plant species diversity. Man's impact on vegetation (ed. by W. Holzner, M.J.A. Werger and I. Ikusima), pp. 41-54. Dr W. Junk Publ., The Hague.

Pijl, L. van der (1982) Principles of dispersal in higher plants, 3rd edn, p. 214. Springer-Verlag, Berlin.

Pimm, S.L. (1984) The complexity and stability- of ecosystems. Nature, 307, 321-326.

Pomeroy, L.R. & Alberts, J.J. (1988) Concepts of ecosystem ecology. Ecol. Studies, 67, Springer-Verlag, New York.

Rebele, F. (1991) Gewerbegebiete - Refugien fur bedro- hte Pflanzenarten? NNA-Berichte, 4/1, 68-74.

Rebele, F. (1992) Colonization and early succession on anthropogenic soils. JVS, 3, 201-208.

Rebele, F. (1994) Konkurrenz und Koexistenz von Tanacetum vulgare L., Solidago canadensis L. und



186 Franz Rebele

Calamagrostis epigejos (L.) Roth auf Aufschuittungs- boden unterschiedlichen Nahrstoffgehalts. Habilita- tionsschrift, Technische Universitat Berlin.

Reeves, F.B., Wagner, D. Moorman, T. & Kiel, J. (1979) The role of endomycorrhizae in revegetation practices in the semi-arid west. I. A comparison of incidence of mycorrhizae in severely disturbed vs. natural environments. Am. J. Bot. 66, 6-13.

Reichholf, J. (1989) Siedlungsraum. Zur Okologie von Dorf Stadt und StraJ3e. Mosaik Verlag, Muinchen.

Rimmer, D.L. (1991) Soil storage and handling. Soils in the urban environment (ed. by P. Bullock and P.J. Gregory), pp. 76-86. Blackwell Scientific Publica- tions, Oxford.

Rohricht, W. (1991) Okosystemanalysen in und um Halle (Saale). Untersuchengen zur Einordnung ausgewahlter phytosuger und phytophager Insekten in die Griinanla- genstruktur von Halle-Neustadt. Diplomarbeit Martin- Luther-Universitat Halle-Wittenberg.

Schmidt, W. (1981) Uber das Konkurrenzverhalten von Solidago canadensis und Urtica dioica. Verh. Ges. Okologie, 9, 173-188.

Schulte, W. (1985) Florenanalyse und Raumbewertung im Bochumer Stadtbereich. Materialien zur Raumord- nung Bd, 30. Ruhr-Universitat Bochum. Bochum.

Simberloff, D. (1986) Introduced insects: A biogeographic and systematic perspective. Ecology of biological invasions of North America and Hawaii (ed. by H.A. Mooney and J.A. Drake), pp. 3-26. Ecological Studies 58. Springer-Verlag, New York.

Simberloff, D. (1989) Which insect introductions suc- ceed and which fail? Biological invasions. A global perspective (ed. by J.A. Drake, H.A. Mooney, F. di Castri, R.H. Groves, F.J. Kruger, M. Rejmanek and M. Williamson), pp. 61-75. SCOPE 37. John Wiley, Chichester.

Slobodkin, L.B., Smith, F.E. & Hairston, N.G. (1967) Regulation of terrestrial ecosystems, and the implied balance of nature. Am. Nat. 101, 109-124.

Sousa, W.P. (1984) The role of disturbance in natural communities. Ann. Rev. Ecol. Syst. 15, 353-391.

Southwood, T.R.E. (1961) The number of species of insect associated with various trees. J. anim. Ecol. 30, 1-8.

Strong, D.R., Lawton, J.H. & Southwood, T.R.E. (1984) Insects on plants: community patterns and mechanisms. Blackwell Scientific Publications, Oxford.

Sukopp, H. (1962) Neophyten in natuirlichen Pflanzengesellschaften Mitteleuropas. Ber. Dt. Bot. Ges. 75, 193-205.

Sukopp, H. (Hrsg.) (1990) Stadtokologie. Das Beispiel Berlin. Dietrich Reimer Verlag, Berlin.

Sukopp, H. (1981) Grundwasserabsenkungen - Ursachen und Auswirkungen auf Natur und Landschaft Berlins. Wasser-Berlin 1981, Bd. 1. Die technisch-wis- senschaftlichen Vortrage auf dem Kongress Wasser 1981, pp. 239-272. Berlin (West).

Sukopp, H. & Hejny S. (eds) (1990) Urban ecology.

Plants and plant communities in urban environments. SPB Academic Publishing, The Hague.

Sukopp, H. & Trepl, L. (1993) Stadtokologie. Handbuch zur Okologie (ed. by W. Kuttler), pp. 391-396. Ana- lytica, Berlin.

Sukopp, H. & Werner, P. (1982) Nature in cities. A report and review of studies and experiments concern- ing ecology, wildlife and nature conservation in urban and suburban areas. Council of Europe Nature and Environment Series, Strasbourg.

Sukopp, H. & Wittig, R. (eds) (1993) Stadtokologie. Fischer Verlag, Stuttgart.

Sukopp, H., Werner, P., Schulte, W. & Fliieck, R. (1986) Untersuchungen zu Naturschutz und Landschaftspflege im besiedelten Bereich. Dokumentation fur Umweltschutz und Landespflege 26, Sonderheft 7, Bib- liographie 51. Bonn.

Tansley, A.G. (1935) The use and abuse of vegetational concepts and terms. Ecology, 16, 284-307.

Tansley, A.G. & Adamson, R.S. (1925) Studies on the vegetation of the English chalk. ILL. The chalk grass- lands of the Hampshire-Sussex border. J. Ecol. 13, 177-223.

Thompson, J.N. (1988) Variation in interspecific interactions. Ann. Rev. Ecol. Syst. 19, 65-87.

Tilman, D. (1982) Resource competition and community structure. Monographs in population biology, 17. Princeton University Press, Princeton, NJ.

Tilman, D. (1986) Evolution and differentiation in terrestrial plant communities: the importance of the soil resource: light gradient. Community ecology (ed. by J. Diamond and T.J. Case), pp. 359-380. Harper & Row, New York.

Trepl, L. (1988) Gibt es Okosysteme? Land- schaft + Stadt, 20, 176-185.

Trepl, L. (1990) Zum Problem der Resistenz von Pflanzengesellschaften gegen biologische Invasionen. Verh. Berl. Bot. Ver. 8, 195-230.

Trepl, L. (1994) Towards a theory of urban bio- coenoses - some hypotheses and research questions. Urban ecology as the basis of urban planning (ed. by H. Sukopp and M. Numata). SPB Academic Publish- ing, The Hague (in press).

Trepl, L. & Sukopp, H. (1994) Zur Bedeutung der Introduktion und Naturalisation von Pflanzen und Tieren fur die Zukunft der Artenvielfalt. Dynamik von Flora und Fauna Artenvielfalt und ihre Erhaltung (ed. by Bayerische Akademie der Wissenschaften). Verlag Pfeil, Munchen (in press).

Urbanska, K.M. (1992) Populationsbiologie der Pflanzen. Fischer Verlag, Stuttgart.

Vitousek, P.M. & Walker, L.R. (1987) Colonization, succession and resource availability: ecosystem- level interactions. Colonization, succession and stability (ed. by A.J. Gray, M.J. Crawley and P.J. Edwards), pp. 207-223. Blackwell Scientific Publications, Oxford.

White, P.S. & Pickett, S.T.A. (1985) Natural disturbance and patch dynamics: an introduction. The ecology of



Urban ecology 187

natural disturbance and patch dynamics (ed. by S.T.A. Pickett and P.S. White), pp. 3-13. Academic Press, New York.

Whittaker, R.H. (1965) Dominance and diversity in land plant communities. Science, 147, 250-260.

Wittig, R. (1991) Okologie der GroJ3stadtflora. Fischer Verlag, Stuttgart.

Yodzis, P. (1986) Competition, mortality and community structure. Community ecology (ed. by J. Diamond and T.J. Case), pp. 480-491. Harper & Row, New York.



Urban Ecology

Documents

Transcript of Urban Ecology