Isabelle Leviol
email: ileviol@mnhn.fr

Romain Julliard
email: julliard@mnhn.fr

Christian Kerbiriou
email: kerbiriou@mnhn.fr

Louis De Redon
email: redon@mnhn.fr

Nathalie Carnino
email: ncarnino@club-internet.fr

Nathalie Machon
email: machon@mnhn.fr

PLANT AND SPIDER COMMUNITIES BENEFIT DIFFERENTLY FROM THE PRESENCE OF PLANTED HEDGEROWS IN HIGHWAY VERGES

Keywords: α and β-diversity, Biodiversity, Functional traits, Roadside management, Revegetation, RLQ analysis

Road verges should play a crucial role as a refuge for native flora and fauna in human dominated landscapes. However, the influence of construction choices, such as plantation of woody species, on the biodiversity supported by roadsides has received little attention, although the presence of hedgerows in roadsides is likely to enhance their role as a refuge, notably for woodland species. Using standardised methods, we assessed the impact of planted hedgerows on two taxonomic groups (plants and spiders) inhabiting highway verges within an intensive agricultural landscape. We examined community richness, taxonomic and functional composition in sites with and without planted hedgerows. At the site level, the response of plant and spider communities to the presence of planted hedgerows differed markedly: hedgerows were associated with significantly higher plant richness (higher α-diversity), but similar spider richness. Plant communities in sites without hedgerows appeared as a subset of communities in sites with hedgerows, whereas spider communities in non-planted sites were complementary to that of planted sites (increased β-diversity). The presence of planted hedgerows was also associated with increased taxonomic and functional trait diversity at the landscape level (γ-diversity), through an increased β-diversity in both plants and spiders. Our results thus suggest that a mosaic of planted hedgerows and grassland habitats is crucial for the maintenance of biodiversity at a landscape scale. By providing information for road practitioners and policy makers regarding their potential impact on biodiversity, these results have important direct implications for the management of road networks.

Introduction

Road networks, which have expanded over large areas with human population growth [START_REF] Watts | Roadless Space of the Conterminous United States[END_REF], are known to influence landscape structure and have major negative impacts on ecosystems dynamics (Trombulak and Frissel, 2000). However, the potential biological value of road verges has also long been recognised [START_REF] Way | Roadside verges and conservation in Britain: A review[END_REF] with possible contribution to the conservation of indigenous flora [START_REF] Spooner | Effects of soil disturbance from roadworks on roadside shrubs in a fragmented agricultural landscape[END_REF][START_REF] O'farrell | Road verge and rangeland plant communities in the southern Karoo: exploring what influences diversity, dominance and cover[END_REF] and fauna [START_REF] Meunier | Use of roadsides by diurnal raptors in agricultural landscapes[END_REF][START_REF] Ries | Conservation Value of Roadside Prairie Restoration to Butterfly Communities[END_REF]. This role as a refuge depends on the surrounding landscape: in natural habitats, generally supporting a high species diversity, road verges may not serve as a refuge [START_REF] O'farrell | Road verge and rangeland plant communities in the southern Karoo: exploring what influences diversity, dominance and cover[END_REF] and even have negative effects, notably by promoting invasion by non-indigenous species [START_REF] Hansen | The influence of disturbance and habitat on the presence of non-native plant species along transport corridors[END_REF][START_REF] Brown | Toad on the road: Use of roads as dispersal corridors by cane toads (Bufo marinus) at an invasion front in tropical Australia[END_REF]. In contrast, in human dominated areas such as intensive agricultural landscapes, where nonagricultural habitats (e.g. edges) are critical to the conservation of biological diversity and ecological processes [START_REF] Burel | Hedgerows and their role in agricultural landscapes[END_REF], road verges should play a crucial role as a refuge and as ecological corridors [START_REF] Tikka | Road and railway verges serve as dispersal corridors for grassland plants[END_REF][START_REF] Smart | Spatial relationships between intensive land cover and residual plant species diversity in temperate farmed landscapes[END_REF].

Road verges support intensively managed habitats, which are artificially created on bare soil after road construction (restoration) and subsequently experience frequent mowing. This has direct consequences on the communities they can support and there is a need to identify and promote biodiversity-friendly construction and management practices, in order to optimise the role of road verges as a refuge for wild flora and fauna. As in any habitat, community composition in road verges is controlled by an array of distinct factors acting as filters at both landscape and local scales [START_REF] Bochet | Road slope revegetation in semiarid mediterranean environments. Part I: Seed dispersal and spontaneous colonization[END_REF], few of which can easily be manipulated to optimise biodiversity conservation.

There have been extensive studies addressing the influence of management on roadside biodiversity, notably mowing frequency and timing [START_REF] Schaffers | Soil, biomass, and management of semi-natural vegetation -Part II. Factors controlling species diversity[END_REF][START_REF] Jantunen | Flowering and seed production success along roads with different mowing regimes[END_REF]. In contrast, construction choices have received much less attention, with the exception of revegetation with herbaceous species [START_REF] Matesanz | Early dynamics of plant communities on revegetated motorway slopes from southern Spain: Is hydroseeding always needed?[END_REF][START_REF] Leps | Long-term effectiveness of sowing high and low diversity seed mixtures to enhance plant community development on ex-arable fields[END_REF]. In particular, the role of planted trees in the maintenance of diversity in road verges was never considered to our knowledge, even though the presence of hedgerows along roadsides is likely to enhance their role as a refuge, by increasing the diversity of microhabitats. This role of hedgerows as a refuge however depends on their ability to receive a significant proportion of the local species pool and of the functional diversity they shelter (Roy and de Blois, 2006). Outside of roadsides, hedgerows appear to be a sustainable habitat and even functional corridors (flora: Corbit et al., 1999, Roy and[START_REF] Roy | Evaluating hedgerow corridors for the conservation of native forest herb diversity[END_REF]fauna: Maudsley, 2000[START_REF] Griffiths | The representation and functional composition of carabid and staphylinid beetles in different field boundary types at a farm-scale[END_REF][START_REF] Davies | Are hedgerows effective corridors between fragments of woodland habitat? An evidence-based approach[END_REF] for a review) even though other authors consider hedgerows as sink habitats [START_REF] Mccollin | Hedgerows as habitat for woodland plants[END_REF][START_REF] Sitzia | Hedgerows as corridors for woodland plants: a test on the Po Plain, northern Italy[END_REF].

The aim of the present study is to evaluate the effect of hedgerows on the surrounding biodiversity of highway verges at different spatial scales, and on the distribution of biodiversity within and among sites in the context of an intensive agricultural landscape. Highway verges offer ideal settings to study the influence of hedgerows, because, in contrast to other habitats, the influence of history on community composition is strongly reduced, due to common history of all study sites. To address the impact of hedgerows on biodiversity, we compared the community richness and composition in sites with and without planted hedgerows for two taxonomic groups (plants and spiders), because no single taxon can inform on the ecological behaviour of other groups (Lowell et al., 2007); In addition, in the context of the search for biodiversity indicators, it is of great importance to better assess the characteristics of different taxa [START_REF] Pearman | Common species determine richness patterns in biodiversity indicator taxa[END_REF]. These two groups differ in their ecological requirements and dispersal capabilities and are likely differentially affected by the presence of hedgerows: plant community composition is expected to be controlled mostly by soil and light characteristics [START_REF] Schaffers | Soil, biomass, and management of semi-natural vegetation -Part II. Factors controlling species diversity[END_REF] whereas spider communities are known to be highly sensitive to habitat structure and microclimatic conditions [START_REF] Marc | Spiders (Araneae) useful for pest limitation and bioindication[END_REF][START_REF] Entling | Niche properties of Central European spiders: shading, moisture and the evolution of the habitat niche[END_REF]; see also [START_REF] Beals | Understanding community structure: a data-driven multivariate approach[END_REF] and may exhibit variation on much smaller time scales than plants. By comparing similar grassland habitats in sites with and without planted hedgerows, we specifically examined (1) whether the presence of planted hedgerows influenced the surrounding specific and functional trait diversity,

(2) whether plant and spider communities exhibited different environmental patterns, and (3) whether artificial hedgerows resulted in a higher overall biodiversity supported by highway verges.

Methods

Study area

This study was conducted within a 50 km section of highway A11 (1°39'30 Lotus corniculatus, Trifolium arvense and Achillea millefolium), along the whole highway section studied, and were planted with trees at discrete locations (Cofiroute construction company, personal communication). In the following, we define hedgerows as linear patches of planted shrubs and trees longer than 150 m, following the definition of [START_REF] Baudry | Hedgerows: An international perspective on their origin, function and management[END_REF]. Sites with and without planted hedgerows are subsequently referred to as "WT" sites and "WO" sites, respectively. Management practices of highway verges depend on distance from the carriageway edge: all verges receive a visibility cut up to 6 m from the road edge twice a year, but areas behind this safety zone are only cut once a year, whereas vegetation growing under planted trees is seldom, if ever, cut.

Field sampling and data collection

Field sampling

We selected 25 field sites (Fig. 1) based on the following structural and biological criteria: highway sections were "roadcuts" (resulting from excavation), with edges of the same shape (width = 16.1 m ± 0.7 SE, height = 4 m ± 0.3 SE), slope (30° ± 2.4 SE)

and aspect, and were bordered by crop fields. Within the 50 km study area, we selected 11 sites with planted hedgerows (WT) along these criteria, using geomorphic maps, aerial photographs, and field surveys. Hedgerows consisted of mature deciduous trees in the canopy layer, and were on average above 3 meters tall and 5 meters wide (5.2 ± 0.5 SE, varying between 3 and 8 meters). When possible, one site without hedgerows (WO) was chosen within 1.5 km of a given site with hedgerows (WT). We thus selected 7 pairs of geographically close sites with and without hedgerows and completed this sample with isolated WT (n = 4) and WO sites (n =7), for a total of 25 sampled sites (11 WT -14 WO). WO sites were always located at least 50 m from the closest tree.

Data collection

All inventories were performed through a 60 m transect running parallel to the carriageway, located at least 7.5 m from the carriageway edge and at 8.2 ± 0.6 SE from the neighbouring crop field. Such location corresponded to extensive management practices. In WT sites, transects were also located 0.5 to 1 m from the canopy of planted trees (vertical projection of lower branches), and started and ended at least 50 m away from the end of the hedgerow, to avoid border effects.

For flora, sites were surveyed from May 10th to June 10th 2006. In each site, all vascular plant species were inventoried once in five 1 m² quadrats regularly distributed over the 60 m transect. Almost all taxa were identified to species level according to the International Plant Names Index. A few species that were difficult to distinguish when not in flower or fruit (Vicia gr tetrasperma-hirsuta for example) were grouped.

Spiders were sampled using five pitfall traps per site. Traps were placed within 30 cm of each vegetation quadrat on the 60 m transect and were left in place from May 10 to June 10 2006. Each trap was 8.5 cm wide in diameter and 10 cm deep, and contained ethylene glycol, a non-attractive preservative. Two pitfall traps were destroyed and were omitted in the analyses (n = 123). Adult spiders were identified to species level following the nomenclature of [START_REF] Platnick | The World Spider catalog, Version 8[END_REF]. We failed to identify three individuals to species level; nevertheless, because they were obviously different from other species, they were conserved for analyses. Spider juveniles were generally not identifiable to species level and were not considered.

Sites characteristics: sampling design was defined to test impact of hedgerows, so that the variation of other environmental variables was minimised. We recorded sites characteristics and checked the absence of correlation between presence/absence of hedgerows and other environmental factors using pearson correlations, to check that the effects of hedgerows were not confounded with other environmental factors. For each site, we recorded geographic coordinates, verge width (final variation: 10-20 m), adjacent land use type (crop vs. temporary fallow), slope, distance from crop edge, aspect, and soil nutrient status. The latter was characterised via determination of soil pH, as well as phosphorus (P2O5, Olsen %), salt (Na2O), and total nitrogen (%) concentration in a bulked representative soil sample, taken to a depth of 10 cm in each of five plots and subsequently mixed. For each site, we also recorded the distance from the nearest woodland patch (411.6 m ± 70.4 SE), as well as the total woodland area within 250 m, 500 m and 1 000 m radius buffers, using aerial photographs and the Geographical Information System package ARCGIS 8 (ESRI, 2000). In this open-field system, woodland patches can be considered as the main natural habitat in Beauce.

Consequently, we considered that they could be important reservoirs of species compared to others habitats, providing a pool of both plants and spiders from which species were able to disperse and potentially colonise highway verges.

Species attributes

To address the influence of hedgerows on community functional composition, we characterised species based on readily available functional traits. As such data were analyzed using multivariate methods that are highly sensitive to "rare" species or classes, we chose to group some classes to minimize the resulting bias (see caption of fig. 2). For plants, these functional traits include Raunkier types, maximum adult height, and seed dispersal mode, scored according to a literature review (supplementary material 1). For spiders, we retained hunting strategies, body size and aerial-dispersal behaviour (ballooning vs. non-ballooning species). Hunting strategies (diurnal cursorial hunters, nocturnal cursorial hunters, ambush hunters, and web-builders) and speciesspecific body sizes were derived from [START_REF] Roberts | The Spiders of Great Britain and Ireland[END_REF][START_REF] Roberts | The Spiders of Great Britain and Ireland[END_REF]. As body sizes usually differ between sexes, the mean value of male and female median sizes was retained.

Species exhibiting aerial dispersal using silk (ballooning) are considered species with great dispersal capabilities [START_REF] Bell | Ballooning dispersal using silk: world fauna, phylogenies, genetics and models[END_REF] for review) and were defined using the world species list of ballooning Araneomorphae established by [START_REF] Bell | Ballooning dispersal using silk: world fauna, phylogenies, genetics and models[END_REF]. All species that did not appear in this list were considered "non ballooners", which may be incorrect because the list is not exhaustive. However, as we worked with common species frequently found in agricultural landscapes and whose biology is generally wellknown, we expect that this bias should be small. For both plants and spiders, a given taxon was assigned attributes only when it was identified to species level and observed in more than four instances (quadrats or traps), which yielded a total of 46 plant species (100% of species observed in more than four quadrats) and 56 spider species (84% of species detected in more than four pitfalls) for which all traits were scored.

Data analysis

To remove the obvious increase in plant species richness due to the simple presence of planted species and examine solely indirect effects of planted hedgerows on plant community composition, we chose to work with spontaneous plant species only. Hence, unless otherwise stated, shrub or tree species that were observed in the herbaceous strata in a site where they had been planted (as inferred from historical plantation map data, Cofiroute -Vegetude) were discarded in subsequent analyses, i.e. they were excluded of this particular site; not however that these species were retained in other sites. In addition, for spider species, we chose to work with presence/absence data only, so that a range of different analyses, some of which valid with presence/absence data only, could be used on the same dataset. With few exceptions, including abundance data yielded very similar results. All analyses were conducted with R (R Development Core Team, 2007).

Species richness

To assess the effect of hedgerows at different spatial scales, we considered species richness at the quadrat scale (Slocal, plants only), at the site scale (Ssite, in a transect of 60 m, plants and spiders) and at the highway section scale, within each type of site (Stype, within WT or WO types of sites, plants and spiders). For plants, we considered local species richness (Slocal) within quadrats because we assumed that all species were detected, whereas for spiders, pitfall trapping is known to be influenced by the activitydensity of species [START_REF] Melbourne | Bias in the effect of habitat structure on pitfall traps: An experimental evaluation[END_REF]. In addition, analysis on local spider richness showed no effect of any environmental variable. Comparing observed species richness across sites and habitats might not be valid in case of unequal species detection probabilities [START_REF] Boulinier | Higher temporal variability of forest breeding bird communities in fragmented landscapes[END_REF][START_REF] Nichols | Inference methods for spatial variation in species richness and community composition when not all species are detected[END_REF]. We therefore used statistical methods derived from capture-recapture approaches in population or community dynamics to estimate species richness and changes in community composition As in recent studies [START_REF] Lekve | Spatio-temporal dynamics of species richness in coastal fish communities[END_REF][START_REF] Doherty | Sexual selection affects local extinction and turnover in bird communities[END_REF][START_REF] Selmi | Does time of season influence bird species number determined from point-count data? A capture-recapture approach[END_REF][START_REF] Devictor | Functional Homogenization Effect of Urbanization on Bird Communities[END_REF][START_REF] Kerbiriou | The impact of human frequentation on coastal vegetation in a Biosphere Reserve[END_REF] addressing richness estimation and differences in community composition from species count data, we used the program COMDYN [START_REF] Hines | COMDYN: software to study the dynamics of animal communities using a capture-recapture approach[END_REF], based on the jack-knife estimator of [START_REF] Burnham | Robust estimation of population-size when capture probabilities vary among animals[END_REF] model including all environmental variables likely to influence species richness (listed above : see "site characteristics"). Plant richness was included as an additional environmental variable for spiders. Data were log-transformed when necessary to meet normality assumptions; the effects of variables retained by the stepwise algorithm were tested adjusted to others and the nested structure of the data was taken into account (local richness per quadrat nested within site: lme, library nlme in R).

Species composition

We investigated a possible spatial autocorrelation in community composition by performing mantel tests (mantel.randtest, library ade4 in R) to assess the correlation between matrices of geographic vs. specific distances among sites. For plants and spiders, matrices of specific distances across sites were computed with the Jaccard index on presence/absence matrices. This effect was also tested separately within both types of edges (WT and WO sites). For these analyses, all species were retained.

We also examined the effect of hedgerows on the composition of plant and spider communities in the herbaceous strata using two types of multivariate analyses:

Constrained Analysis of Principal Coordinates (CAP) and partial Constrained

Correspondence Analysis (pCCA), followed-up by ANOVA-like permutation tests (n = 999) to assess significance of effects. CAP (capscale, library vegan in R, see [START_REF] Anderson | Canonical analysis of principal coordinates: A useful method of constrained ordination for ecology[END_REF] is an ordination method similar to Redundancy Analysis (RDA), i.e.

exploring the relationship between two sets of variables, but allowing non-Euclidean dissimilarity indices. It was used to examine whether species similarity among sites depended on the presence/absence of hedgerows, and on the distance to nearest woodland patch, which was used as a covariable. It was performed on presence/absence matrices across sites, using the Jaccard index. pCCA (cca, library vegan in R, ter [START_REF] Ter Braak | Canonical Correspondence-Analysis -a New Eigenvector Technique for Multivariate Direct Gradient Analysis[END_REF]) is an ordination method based on CCA, a constrained ordination method developed to relate community composition to known variation in the environment. It is considered as a good choice if the user has clear and strong a priori hypotheses on constraints. Here, we used pCCA to test whether the presence of some species was associated with the presence or absence of planted hedgerows. We chose pCCA instead of a simple CCA to remove the effect of the distance to the nearest woodland patch, used as a covariable, as in CAP. As both analyses are sensitive to rare species, we only retained species observed in more than four samples, i.e. 46 plant species and 67 spider species.

Third, we examined the complementarity of communities [START_REF] Nichols | Inference methods for spatial variation in species richness and community composition when not all species are detected[END_REF] in each type of site (WO type and WT type) using COMDYN with sites as replicates (n = 11). In addition to estimating richness within each type of site (Stype within WO and WT type of sites), we obtained the following estimators: complement of "extinction probability" between the two types, Φ (i.e. proportion of WT species present in WO), estimated complement of species "turnover", γ (i.e. proportion of WO species present in WT), and "estimated local colonising" species, B (i.e. number of species not present in WT but present in WO). As for richness, we compared the mean and standard error (over 15 re-sampling events) of these estimators between WT and WO types.

Functional trait composition

We investigated whether some species attributes were more likely associated with one type of site and distance to neighbouring woody patch (environmental variables). To this end, we used a multivariate ordination analysis, RLQ, which is an extension of coinertia analysis. RLQ performs a double inertia analysis of two arrays (R, here the environmental array and Q, here the trait array), i.e. measures the adequacy between the two arrays, with a link expressed by a contingency table (L, here the species array: matrix of presence-absence of species across sites) (rlq, library ade4 in R, complete description: see [START_REF] Doledec | Matching species traits to environmental variables: A new three-table ordination method[END_REF]; for an example: [START_REF] Ribera | Effect of land disturbance and stress on species traits of ground beetle assemblages[END_REF]. The general significance of the relationship between environmental variables and species traits was tested using random permutations test (n = 999). We also used only species observed in more than four samples to limit the potentially strong effects of rare species.

Results

Across 25 highway verge sites, we observed a total of 85 plant species and 123 different spider species. For spiders, we captured 8 300 individuals, 7 862 of which were adults that could be identified to species level. Only eight of the 85 plant species (A.

campestre, A. pseudoplatanus, B. pendula, F. excelsior, P. avium, P. padus, P. spinosa, V. opulus) were identical to those planted in the immediately neighbouring hedgerow in WT sites and were excluded from most analyses; note however that including them had no effect on the following results.

Species richness

Influence of planted hedgerows on plant communities

We observed a significant positive effect of the presence of hedgerows (Table 1) on plant richness at the local scale (quadrat level, n = 125, SlocalWT = 10.20 ± 0.46 SE; SlocalWO = 8.20 ± 0.47 SE, F1,20 = 5.79, p = 0.026). We also found a negative effect of distance from crop edge (F1,20 = 5.75, p = 0.026), as well as tendencies for effects of adjacent land use (higher richness with temporary fallow, F1,20 = 4.15, p = 0.055) and phosphorous rate (negative : F1,20 = 4.00, p = 0.059) on local plant richness. In contrast, no effect of hedgerows or any environmental factor was detected on plant richness at the site level (n = 25, SsiteWT = 25.45 ± 1.66 SE; SsiteWO = 23.75 ± 3.12 SE) and there was no difference in detection probabilities between the two types of sites. However, at the section level (n =11), the estimated plant richness was 139.16 over all WT sites (StypeWT, ± 29.20 SE; detection probability dp = 0.47 ± 12.42 SE) vs. 87.51 over all WO sites (StypeWO, ± 19.43 SE; detection probability dp = 0.68 ± SE = 0.12).

Influence of planted hedgerows on spider communities

We found no effect of the presence of planted hedgerows on spider richness at the site level. Spider richness was significantly affected by distance to the nearest woodland patch (negative effect, F1,18 = 5.94, p = 0.025), site slope (positive effect, F1,18 = 4.65, p = 0.044) and plant richness (positive effect, F1,18 = 6.81, p = 0.018). There was no difference in detection probabilities between the two types of sites. At the section level, (n = 11), estimated spider richness was 119.3 over all WT sites (StypeWT, ± 19.6 SE; detection probability dp =0.80 ± 0.10 SE) and 142.0 over all WO sites (StypeWO, ± 27.15 SE; detection probability dp = 0.64 ± 0.11 SE).

Community composition

Spatial autocorrelation

Using Mantel tests on the full dataset, we found significant spatial autocorrelation in plant communities (p = 0.001), but not in spider communities (p = 0.066). In contrast, within types of site, spatial autocorrelation was observed in plants (WT: p = 0.006; WO: p = 0.001) and spiders (WT: p = 0.007; WO: p = 0.028). However, as WO and WT sites were randomly distributed in space (Fig. 1), autocorrelation should not interfere with our main purpose, i.e. to assess the influence of hedgerows on community composition.

The presence of planted hedgerows influence community composition

Sites with and without planted hedgerows supported significantly different plant and spider communities, as shown by CAP (ANOVA-like permutation tests: plants: F1,22 = 2.26, p < 0.001, inertia explained by hedgerows = 9% and by distance to the nearest woodland patch = 4%; spiders: F1,22 = 1.90, p = 0.006, inertia explained = 7% and 6%, respectively). This effect of hedgerows on plant and spider community composition was confirmed by the pCCA analysis (ANOVA-like permutation tests: plants: F1,22 = 2.13, p < 0.01, inertia explained = 8% and 4% respectively; spiders : F1,22 = 1.81, p < 0.005, inertia explained = 7% and 5% respectively). This method also identified species that seemed preferably associated with WT sites. For plants, these were A. campestre, B.

pendula, Galium mollugo, Centaurium erythraea, Trifolium repens, and for spiders, Panamomops sulcifrons, Maso sundevalli, Walckenaera dysderoides, Ceratinella brevis, and Drassodes cupreus. In contrast, the spider species Tegenaria picta, Ero furcata, and Palliduphantes ericaeus, were preferably associated with WO sites.

Nested vs. complementary communities

We found that the plant community observed in WO sites was a subset of the plant community in WT sites (Table 1). WO sites did not exhibit any private species: we found 99% of WO sites species in WT sites, so that all WO species were expected in WT sites (COMDYN estimator B = 0). The reverse was not true: WT exhibited several private species (14%), with only 86% of WT sites species also observed in WO sites. In contrast, spider communities appeared complementary. Respectively 83% (WT) and 77% (WO) of spider species were present in the other type of site, so that each type of sites exhibited private species.

Relationship between species attributes and presence of hedgerows

For both plants and spiders, there was a significant relationship between species attributes and environmental variables (Fig. 2; RLQ randtest, p = 0.001 and p = 0.001).

For the two taxa, the presence/absence of planted hedgerows was associated with the first axis, which explained respectively 81% and 80% of the total inertia, whereas the distance to the nearest woodland patch tended to be associated with axis 2 (respectively 19% and 20% of inertia). In spontaneous plant species, Raunkier type was influenced by the presence of hedgerows, with Phanerophytes and Chamaephytes (Rank: P) positively associated with WT sites (Fig. 2). In contrast, adult height was not significantly associated with site type. Finally, species with no specific means of dispersal (barochory, autochory) tended to be more present in WT sites than in WO sites, whereas animal-dispersed species (zoochory) were more common in WO than in WT sites;

wind-dispersed species were not affected by the presence of hedgerows. In spiders, the mode of dispersal (ballooning vs. non ballooning) was strongly affected by the presence of hedgerows: ballooning species were more likely found in WO sites, and nonballooning species in WT sites. In addition, web-builders were more common in WT than in WO sites, and diurnal wandering hunters in WO sites. Finally, larger species tended to be more common in WT sites than in WO sites.

Discussion

We addressed the effect of the presence of planted hedgerows on neighbouring plant and spider communities of highway verges. We showed that (1) planted hedgerows were associated with higher species richness, i.e. a higher α-diversity (αWO < αWT), for plants, but not for spiders. This was due to the fact that plant communities in sites without hedgerows (WO) were a subset of communities in sites with hedgerows (WT), whereas spider communities in WO sites were complementary to communities in WT sites. Below, we argue that such differences are attributable to different ecological characteristics of species, which is supported by the distribution of functional traits, and derive the consequences in terms of roadside management.

The influence of propagule source vs. local conditions on plant and spider community composition

Following the environmental filter model [START_REF] Keddy | Assembly and Response Rules -2 Goals for Predictive Community Ecology[END_REF], two non-exclusive hypotheses may explain the effect of hedgerows on community composition, depending on taxonomic group. Hedgerows could either be a source of dispersers ("source" hypothesis) or could provide microclimatic conditions that affect the presence of some species due to their environmental requirements ("local conditions" hypothesis).

"Source" hypothesis and the effect of hedgerows on plant communities

The dispersal of some species from neighbouring hedgerows into a suboptimal habitat (grassland) would affect community composition in the suboptimal habitat. This phenomenon, known as the edge effect, has been observed in several taxa: contrasted adjacent habitats supporting distinct species assemblages often mix over a few meters on either side of the interface (see for example [START_REF] Bedford | Distribution of arthropod species across the margins of farm woodlands[END_REF]. For example, [START_REF] Downie | Distribution and dynamics of surface-dwelling spiders across a pasture-plantation ecotone[END_REF] studying spider community composition across a plantationpasture ecotone, found that most of the community (70% of species) was formed by the overlap of both pasture and plantation species encroaching across the interface through dispersal or foraging processes. Interface-specialist species were present but contributed less to the interface community content than the other species categories (23% of individuals and 25% of species).

Under the "source" hypothesis, communities are therefore expected to exhibit higher species richness in WT vs. WO sites. This was the case for plant communities (αwo < αwt), at the quadrat scale, but not at the site level. Another prediction of the source hypothesis is that WO communities should be a subset of WT communities. This was true again for plants: in WT sites, we observed 'private' species which were not supported by WO sites, whereas the reverse was not true. Finally, in the case of a higher effect of dispersal vs. local conditions on species distribution within similar habitats, a distance effect can be expected independently of presence/absence of hedgerows. In plant communities, spatial autocorrelation was in fact detected regardless of the presence of hedgerows. These three observations converge to suggest that dispersal is a major driver of community composition in this group.

The distribution of plant functional traits is consistent with this role of dispersal. We found that Phanerophytes and Chamaephytes were more represented in WT sites, corresponding mostly to the occurrence of the following species: A. campestre, B. pendula, C. scoparius, F. excelsior and P. spinosa. These species observed in the herbaceous strata most likely originated from seed-bearer trees that grow spontaneously within the otherwise planted hedgerow. Thus, hedgerows offer a sustainable habitat where particular species, and notably woody plants, can grow and reproduce, whereas such species cannot reach maturity within WO sites, due to frequent mowing. The fact that zoochory was more represented in WO sites is however not easily interpreted in this framework.

Local conditions hypothesis and the effect of hedgerows on spider communities

Microclimatic conditions (influenced, among others, by the neighbouring hedgerows) might be responsible for observed community composition, especially in spiders. For example, spider richness within sites was positively correlated with local plant richness;

plant community composition and consequently the architecture of habitat are wellknown predictors of spider community composition (e.g. [START_REF] Beals | Understanding community structure: a data-driven multivariate approach[END_REF]. Moreover, woody hedgerows are known to provide particular microclimate conditions in their neighbourhood [START_REF] Forman | Hedgerows and hedgerow networks in landscape ecology[END_REF] and most authors agree that those are a major factor controlling spider distribution (see [START_REF] Marc | Spiders (Araneae) useful for pest limitation and bioindication[END_REF]. For example, [START_REF] Downie | Distribution and dynamics of surface-dwelling spiders across a pasture-plantation ecotone[END_REF], examining spider community composition, proposed that close to an interface, the role of vegetation structure is low compared to other factors, including changes in microclimate. Finally, perennial non-crop habitats are known to be used by spiders as overwintering sites and refuges [START_REF] Oberg | Landscape effects on recolonisation patterns of spiders in arable fields[END_REF]. Grassland habitats located next to hedgerows maybe therefore provide better conditions than sites without hedgerows for overwintering, with direct consequences on spider community composition. Thus, neighbouring hedgerows may have an environmental filter effect, which leads to the selection of species with particular ecological requirements.

Under the 'local conditions' hypothesis, we expect that sites with and without planted hedgerows support different species. This was the case with spiders: WT and WO communities were complementary instead of nested, thus yielding similar species richness with or without hedgerows. This implies that some spider species occurring in WO sites were absent from planted sites. As the two types of sites offered the same habitat (grassland with similar percentage cover and dominance of graminoïds), the observed differences are likely be explained by specific microclimatic conditions created by the neighbouring hedgerows. Finally, the role of hedgerow-induced local conditions in spider community composition is also supported by patterns of spatial autocorrelation: the dominant impact of neighbouring hedgerows on spider community composition removed spatial autocorrelation patterns across all sites, whereas spatial autocorrelation remained significant within WT and WO sites.

Despite the significant role of local conditions, spider community composition is also likely affected by dispersal. For example, spider richness within sites was negatively correlated with the distance to the nearest woodland patch. Such effect of distance is well known from the theory of island biogeography (Mac-Arthur and Wilson, 1967) and consistent with results from other field studies: spider species richness in wheat fields has been shown to increase with the proportion of non-crop habitats in the surrounding landscape [START_REF] Schmidt | The role of perennial habitats for Central European farmland spiders[END_REF]; see also [START_REF] Bianchi | Sustainable pest regulation in agricultural landscapes: a review on landscape composition, biodiversity and natural pest control[END_REF] for other groups).

Although the observed distribution of species between site types suggests a major role of local conditions, trait analyses outlined that aerial dispersal capabilities are an important factor explaining the interaction between the presence of hedgerows and spider community composition: species exhibiting ballooning behaviour were more represented in WO sites and the reverse was true for non-ballooning species. This result is comparable to that of [START_REF] Bonte | Effects of aerial dispersal, habitat specialisation, and landscape structure on spider distribution across fragmented grey dunes[END_REF], who found that the distribution of 29 spider species inhabiting fragmented grey dunes depended on their aerial dispersal potential, as well as on the interaction between patch connectivity and area. Ballooning is thought to be more common in species from unstable and ephemeral habitats [START_REF] Bonte | Low propensity for aerial dispersal in specialist spiders from fragmented landscapes[END_REF][START_REF] Bell | Ballooning dispersal using silk: world fauna, phylogenies, genetics and models[END_REF]; moreover, high dispersal capabilities are generally dominant in disturbed habitats (e.g. Nyffeler and Sunderland, 2003 for spiders;[START_REF] Ribera | Effect of land disturbance and stress on species traits of ground beetle assemblages[END_REF] for beetles), allowing the maintenance of populations in ephemeral agricultural ecosystems [START_REF] Weyman | Do seasonal changes in numbers of aerially dispersing spiders reflect population density on the ground or variation in ballooning motivation?[END_REF]. We therefore suggest that the significantly higher proportion of ballooning species in WO sites may be due to the absence of a neighbouring hedgerow, i.e. of a stable habitat. In contrast, species that do not exhibit this behaviour should be more common in WT sites because hedgerow likely provides more stable conditions, both in terms of management (no mowing) and microclimatic variables. Hedgerows would thus be overwintering refuges (see also [START_REF] Pywell | Determinants of overwintering habitat quality for beetles and spiders on arable farmland[END_REF], from which species colonise the surrounding habitats in spring. Sites without planted hedgerows, which receive a cut, generally in autumn, would be poor refuges that are partly recolonised from more distant sources in spring. Hence, they support a higher proportion of species exhibiting large dispersal capabilities, i.e. "ballooners".

Dispersal and local conditions therefore appear to drive spider community composition at different temporal scales.

Finally, note that taxonomic differences in composition of spider communities were also associated with differences in hunting strategies: web-builder were more represented in WT sites possibly due to the greatest heterogeneity of support provided by the presence of neighbouring hedgerow and the higher richness of plant whereas diurnal wandering hunter were more abundant in WO sites.

Conclusions

Planting generally increased the γ-diversity of highway verges, via an increase in both taxonomic and functional trait β-diversity between sites with and without hedgerows, but this positive effect was different in the two taxa considered. For plants, it was mostly due to an increase in taxonomic α-diversity within sites with hedgerows, whereas for spiders it was related to a direct increase of taxonomic β-diversity between the two types of sites. Note that this partition of diversity is likely to vary through time as well (e.g. due to overwintering dispersal patterns in spiders), with different temporal scales in the two taxa, but this was not the focus of the present study. From a theoretical viewpoint, the fact that the partitioning of diversity between α and β diversity differed between the two taxa is probably due to the different spatial and temporal scales at which the heterogeneity of interactions between organisms and their environment can be found (see [START_REF] Loreau | Are communities saturated? On the relationship between alpha, beta and gamma diversity[END_REF]. According to [START_REF] Loreau | Are communities saturated? On the relationship between alpha, beta and gamma diversity[END_REF], "the challenge now is to understand the relationship between α and β diversity at multiple scales, and the processes that determine it". In our opinion, comparing the partitioning of diversity among several taxa and functional groups at different spatial scales is therefore a promising way to explore the processes that control the distribution of biodiversity.

Consequences for the management of highway verges

Our findings have direct consequences for the management and conservation of highway verges in agricultural landscapes, where most biodiversity is aggregated in more stable non-crop habitats and field edges [START_REF] Bianchi | Sustainable pest regulation in agricultural landscapes: a review on landscape composition, biodiversity and natural pest control[END_REF]. Increasing the botanical and structural heterogeneity of hedgerows by planting is likely to have significant beneficial effects on native flora and fauna [START_REF] Maudsley | A review of the ecology and conservation of hedgerow invertebrates in Britain[END_REF]. Obviously, plantation increases species richness both directly (introduction of new species) and indirectly, by providing particular environmental conditions within woody hedgerows and thus contributing to the maintenance of some species (e.g. woody plants) in landscapes where they are poorly represented. But our results suggest that planted hedges may also affect community composition of nearby herbaceous strata, enhancing both taxonomic and functional trait γ-diversity along highway verges.

In practice, our results highlight the importance of creating a mosaic of planted sites alternating with non planted sites in highly disturbed landscapes. Of course, "natural" dynamics could yield the same results and probably a greater biodiversity as demonstrated by [START_REF] Boutin | Plant diversity in three types of hedgerows adjacent to cropfields[END_REF], who compared natural vs. planted hedgerows.

However, in the case of highway verges, natural hedgerows would only arise by recolonisation of the newly created habitats, which involves successional dynamics with intermediate states, and the presence of pioneer species often considered as weeds. In the context of intensive agricultural landscape, where farmers wish to limit weeds in field boundaries as much as possible, planting trees along highways verges appears as a compromise between biodiversity conservation and anthropogenic pressures. The issues of the identity of planted species (which are usually exotic species), of management of hedgerows [START_REF] Croxton | The restoration of neglected hedges: a comparison of management treatments[END_REF]) and of their spatial arrangement along highway verges (to provide effective corridors) remain to be addressed. Given the urge to conserve biodiversity, especially in the context of climate change, road practitioners should consider plantation not only for esthetical reasons but also as a possibility to increase the role of verges as a refuge and, consequently, landscape connectivity.

Captions

Table 1 -Changes in species composition in the herbaceous strata of highway verges in sites without planted hedgerows (WO) vs. sites with planted hedgerows (WT). The analysis is based on presence/absence of species with known attributes that were observed in more than four samples (46 plant species and 56 spider species). 1) Environmental parameters. WT: sites with planted hedgerows, WO: sites without hedgerows; dist: distance to the distance to the nearest woodland patch. 2) Plant traits (A). rank: raunkier type: H: Hemicryptophytes and Geophytes (28 species), P: Phanerophytes and Chamaephytes (6), T: Therophytes (12); heigh: maximum adult height: S: small (0 to 45 cm: 9), M: medium (46 to 100 cm: 19), H: high (100 and more: 18); seed: Seed dispersal: A : anemochory (30), B: barochory and autochory (9), Z: zoochory (7). 2) Spider traits (B). ballo: ballooning behaviour: Y: Yes (30), N: No (26); size; hunt: hunting strategies: W: diurnal wandering (18); N: nocturnal wandering (10); W: web-builders (21); A: ambush hunters (7).

Fig. 1 -

 1 Fig. 1-Location of the 25 study sites.

Fig. 2 -

 2 Fig. 2 -RLQ analysis ordination biplot illustrating relationships among environmental

Fig. 1 -Fig. 2 -

 12 Fig. 1-Location of the 25 study sites.

  . To estimate species richness at the site level (Ssite), we used quadrat or pitfall captures as replicates. To estimate richness at type of site level (Stype within WT or WO type of sites), we used sites as replicates (n = 11): analyses were based on presence-absence data at the site level, pooled across the five quadrats for plants and the five pitfall traps for spiders. As 11 replicates only were available for WT sites, we sampled randomly without replacement 11 sites out of 14 in the WO group; this sampling procedure was performed 15 times. Each of 15 pairs of (1) 11 sampled WO sites plus (2) the 11 WT sites was input into program COMDYN to obtain estimated richness per type of site. The estimated richness in each type of site was then characterised using the mean and standard error of the 15 outputs.

	We analysed the combined influence of hedgerows and environmental variables
	on both local species richness for plants (Slocal) and estimated richness per site for
	plants and spiders (Ssite), using linear models and analyses of variance, as follows: for

each taxonomic group (plants and spiders) and dependent variable (Slocal and Ssite), we selected the most informative model using a stepwise algorithm based on Akaike Information Criterion (StepAIC, library MASS in R) starting from the most complete

Acknowledgements:

This work was financially supported by the Direction of Construction of Cofiroute.

Emmanuelle Porcher was partly funded by Agence Nationale de la Recherche (ANR grant #2006-JCJC-0032). We are grateful to Cofiroute for providing access to Highway verges and to Vegetude for information about revegetation. We thank Coralie Dedeystere for assistance in the field, Sophie Gachet for help with plants traits, Sandrine Pavoine for precious help with statistical analysis and Didier Bas and Vincent Devictor for software facilities (COMDYN).

____________________________________________________________


Audrey Muratet

Nathalie Machon

Frédéric Jiguet

Jacques Moret

Emmanuelle Porcher

teaching and research institutions in France or abroad, or from public or private research centers.

Introduction

Conservation policies in urban zones usually target relicts of natural landscapes, such as wetlands or woodlands. In contrast, artificial urban habitats are not considered interesting green zones by most city-dwellers [START_REF] Sukopp | Développement de la faune et de la flore en territoire urbain[END_REF]. In particular, wastelands, defined as abandoned lands where plant species grow without human control, are frequently neglected in conservation plans. However, despite their often negative image, they can be of importance due to their high floristic diversity (Maurer and others 2000;Zerbe and others 2003;[START_REF] Herbst | The development of an evaluation method using a geographic information system to determine the importance of wasteland sites as urban wildlife areas[END_REF]. The presence of wastelands in cities may thus contribute significantly to the total biodiversity in urban areas, although the extent of their role is poorly documented. Thus, studying wastelands is of primary concern to assess their role in maintaining urban biodiversity and to define effective urban management plans.

Although the high floristic diversity of wastelands is now relatively well documented, very little is known regarding the mechanisms that control the distribution of flora in this habitat; this fundamental information should be accounted for in conservation strategies. As in many human-disturbed habitats, species composition in urban wastelands is likely influenced by (1) biotic and abiotic environmental conditions (soil, light, humidity, interspecific competition…),

(2) the pool of species able to colonize from neighboring (wasteland or non-wasteland) sites,

(3) the barriers preventing pollen/seed flow among sites, and (4) the frequency and intensity of human-induced disturbance. Wastelands are subject to various development pressures [START_REF] Harrison | Conserving biodiversity that matters: practitioners' perspectives on brownfield development and urban nature conservation in London[END_REF], and their life span is highly variable but generally short. In the short to mid-term, wastelands are partly protected from human-induced disturbance by their negative image, which may favor biodiversity. However, wastelands can also be short-lived habitats whose life span is defined by the necessity to construct new buildings or parks. The succession of local extinctions and colonizations of newly vacant sites from neighboring sites may thus create metacommunity dynamics in urban wastelands. Depending on which factors exert the strongest influence, plant species composition can be spatially autocorrelated (e.g. if among-site dispersal is possible but limited in space) or completely patchy (e.g. with little among-sites dispersal, strong effect of local environments, or influence of local human pressure only). Unraveling the relative importance of biological interactions, dispersal limitation, and environmental determinism in inducing differences in species composition among sites is a central issue in plant ecology [START_REF] Hurtt | The consequences of recruitment limitation: reconciling chance, history and competitive differences between plants[END_REF]Whitfield 2002).

In the present study, we quantify floristic diversity in ca. 100 wastelands of the Greater Paris area, and for the first time explore the role of some of the human and environmental factors controlling their floristic composition, to evaluate the consequences for the conservation of urban biodiversity. Wasteland floristic diversity was measured via species richness, species rarity and indigeneity. We described the characteristics of wasteland sites (spatial distribution, floristic diversity, age, and area) in one of the most urbanized zones of France and used data from the Land Use Pattern to address the following questions:

(1) What is the influence of the abiotic environment on the floristic diversity of wastelands? To address this, we compared matrices of floristic distances to matrices of environmental distances between wasteland sites.

(2) Does migration among wastelands play a role in maintaining floristic diversity? To address this, we compared matrices of floristic distances to matrices of geographic distances between wasteland sites.

(3) Do urban structures surrounding wasteland sites impact floristic diversity? To address this, we studied the relationship between the floristic diversity of wasteland sites and the presence of various urban structures around the sites, as inferred from the Land Use Pattern.

Material and methods

Study area

Wasteland floristic diversity was studied in Hauts-de-Seine (France), a crescent-shaped administrative department (48°50' N; 2°14' E) located in the heart of the Greater Paris area (which represents a total of 2720 km 2 ). This department is composed of 36 boroughs that cover a total area of 176 km² (Figure 1). The climate is oceanic with continental trends: the mean annual temperature is 11.7°C with a thermal amplitude of 16°C and an average annual rainfall of 641 mm. The department is located between the Beauce Plateau and the Seine River Valley, comprising of a plain, small hills and a plateau. It is one of the most densely inhabited zones of France, with a human density of 8118 inhabit./km² (INSEE 1999) as opposed to 95 inhabit./km² on average in France. The successive Land Use Patterns from 1982 to 2003 were provided by [START_REF] Iaurif | Institute for Planning and Development of the Paris Ile-de-France Region[END_REF]. They show, in particular, that built zones have covered about 70% of the territory for more than 20 years (IAURIF 2003 see Figure 1).

Sampling design and inventories

The flora of Hauts-de-Seine was inventoried over all the department, on a total of 986 sites, using a stratified sampling strategy, i.e. the surface of inventoried sites in a given habitat was proportional to the total area occupied by this habitat in the department (see [START_REF] Muratet | Diversité végétale en milieu urbain[END_REF] for details on site location and sampling methods). Each site was inventoried once between 2001 and 2005, in spring, summer or autumn. During an inventory, we recorded all wild vascular plant species over the whole site area ("global method"), i.e. we dismissed cultivated and ornamental species. Species were classified as "indigenous" or "naturalized" according to a list compiled by professional botanists of the National Botanical Conservatory of the Parisian Region (CBNBP). Naturalized species are non-indigenous species that now behave like indigenous species (Richardson and others 2000). Several environmental parameters were also recorded: light exposition, slope, type of parent rock, and site geomorphology.

Among the 986 inventoried sites, 98 were wastelands with areas ranging from 12 to 18 263 m².

As mentioned above, wastelands were defined as sites where vegetation management has been more or less temporarily abandoned. In cities, four types of wastelands can be found (1) Wastelands located in building sites and vacant urban areas, corresponding to transitions in land use, (2) Wastelands located in open urban areas, corresponding to irregularly managed plots of gardens, parks, golf courses and sport fields, (3) Wastelands located in built-up areas (dwellings, activities, facilities), corresponding to abandoned industrial plots or to small unused interstices in the built matrix, (4) Wastelands located in subnatural environments (e.g. woods or river banks), corresponding to the disturbed edges of natural zones.

In 13 of these sites, chosen for their accessibility, an alternative method of inventory ("quadrat method") was used in addition to the "global" method to test the accuracy of the latter at producing unbiased estimates of local plant diversity. In the quadrat method, we performed five replicate inventories within each site, using 3 m x 3 m quadrats that were placed systematically in the field to avoid an observer bias during the choice of quadrat location. We used the CAPTURE program in the Mark software [START_REF] Rexstad | User's guide for interactive program CAPTURE. Abundance estimation of closed animal populations[END_REF] to obtain species richness estimates. The model including heterogeneity effect on species detectability only, M(h), yielded the best fit of data and was retained to analyze the "quadrat method". We found no significant difference between the richness observed via the "global" method and the richness estimated by capture-recapture via the "quadrat" method (Wilcoxon signed-rank test, P = 0.097) although observed richness was generally lower than estimated richness. Therefore, the global method seems to provide an accurate and confidently unbiased measure of species richness, and the results we present in the following are derived from this method only.

Geographic information system

We mapped the distribution of wastelands in Hauts-de-Seine (Figure 1) and calculated their area using a Geographic Information System (MapInfo software, MapInfo Corporation 2004).

Additional information regarding wastelands was retrieved from six updates of the Land Use Pattern (LUP), performed in 1982[START_REF] Sukopp | Développement de la faune et de la flore en territoire urbain[END_REF][START_REF] Grumbine | Viable Populations, Reserve Size, and Federal Lands Management: A Critique[END_REF][START_REF] Rebele | Urban ecology and special features of urban ecosystems[END_REF][START_REF] Hurtt | The consequences of recruitment limitation: reconciling chance, history and competitive differences between plants[END_REF], and 2003and provided by IAURIF (2003).

The LUP contains 83 different classes that were grouped into nine major classes for this study (Appendix A). We identified the LUP classes in which wastelands were found. By comparing the different updates of the LUP (Table 1), we could estimate the age of each wasteland site, defined as the number of years during which the site LUP class remained the same. We also identified the LUP classes around each wasteland, to characterize their urban environment.

Data analysis

Indices of floristic interest

To study the impact of the urban environment on the floristic interest of wastelands, we used three parameters of floristic interest for a given site: (1) species richness (Rich), the total number of species observed in the site, (2) indigeneity (Ind), the proportion of indigenous (versus naturalized) species and (3) rarity (Rar). For a given species, rarity (Rar sp ) is the fraction of sites (including all habitats) in the department in which the species was not observed. Rar sp varies between 0 (the species was observed in all sites) and 985/986 (the species was observed in a single site in the department). At the site level, rarity (Rar) is the mean of specific rarity indices over all species observed in the site.

Relationship among floristic, environmental and geographic distances

As site area has a major influence on species richness (see Results), sites were grouped into four quartile area classes to avoid confounding area effects when analyzing the effect of geographic and environmental distances. The geographic distance between two sites was defined as the Euclidean distance between their barycentres. The floristic distance between two sites, d f , measuring the difference in species composition, was calculated from the global inventory data via the Jaccard similarity index s f with the R software (R Development Core Team 2005), using the ADE4 package (Thioulouse and others 1997) as follows:

1 f f d s = -
where s f , the fraction of total species found in both sites. The environmental distance between two sites was calculated via the Gower index s e , which includes a correction factor for the missing values, using the ADE4 package (Thioulouse and others 1997).

s e = number of environmental parameters with identical values in both sites/(total number of values for all parameters + number of parameters values not found in both sites)

Correlations between floristic distance and geographic or environmental distances were tested with simple and partial Mantel tests based on Pearson correlations using the vegan package from the R software (Oksanen and others 2005).

Impact of urban environment on floristic interest

We analyzed the variation of floristic interest indices using three ANCOVAs where site richness, indigeneity or rarity was a function of (1) the LUP class in which sites were located,

(2) distance from the center of Paris, a good proxy for urbanization intensity (the proportion of built-up areas strongly depends on the distance to the center of Paris, linear regression, proportion = 1.18 -0.05 x distance, r 2 = 0.83, P < 0.0001), (3) site age, (4) season of inventory, (5) all pairwise interactions among these four variables, and (6) simple effects of ecological parameters: parent rock, geomorphology, slope, and exposition. Site area was added as a quantitative covariate in the model. Using a stepwise AIC method, this full model was simplified to retain the most informative model.

We assessed the impact of urbanization patterns around wastelands on their floristic interest by studying the relationships between (1) floristic interest indices (Rich, Ind and Rar) of a wasteland site and (2) variables with a significant effect retained from the ANCOVA above, plus the proportion of area occupied by the 9 LUP classes within a 200 m buffer around the site center. The LUP class area proportions were obtained using the MapInfo software (MapInfo Corporation 2004) and Arcsin-root transformed to approach normality. The 200 m radius was identified in a former study as the best radius fitting floristic variation in Hauts-de-Seine to the proportion of the different LUP classes [START_REF] Muratet | Diversité végétale en milieu urbain[END_REF]. Hierarchical partitioning (Walsh and McNally 2004) were conducted on r² values to isolate parameters with a significant independent effect on the site floristic interest and to correct for multicollinearity [START_REF] Mcnally | Multiple regression and inference in ecology and conservation biology: further comments on identifying important predictor variables[END_REF]. A randomization (500 times) permitted the segregation of the significant independent effects on the Rich, Ind and Rar indices.

Results

Description of studied wastelands

The 98 studied wastelands were distributed over the whole area of the Hauts-de-Seine department (Figure 1) and were observed in each of the nine LUP classes (Figure 2). However, they were more frequently located in the "building and vacant area" (BUILVAC) class (28%), corresponding to transitions in land use yielding temporary abandoned plots and in the "open and rural areas" (OPENRUR) class (25%), i.e. irregularly managed plots of horticulture, orchards, gardens, parks, golf courses and sport fields. Wastelands in the built (ACTI, COLL, FACI, INDI and TRAN) classes corresponded to abandoned industrial plots or to small unused interstices in the built matrix. There were very few wastelands in « natural environments » (WATE and WOOD) (2%).

In contrast to what we expected for frequently disturbed habitats, wastelands were generally old, with 66% of sites older than 21 years. The oldest (> 21 years) wastelands were significantly smaller than sites of intermediate age (between 4 and 21 years) (r² = 0.21, P = 0.0002, Table 1).

Floristic results

We recorded a total of 365 vascular plant species in the Hauts-de-Seine wastelands, which corresponds to 58% of the total species richness recorded over the whole department. The mean richness was 39 species per site, varying between 5 and 92 species.

Among the 365 species, nine were found in more than half of the sites (Artemisia vulgaris L., Cirsium arvense (L.) Scop., Plantago lanceolata L., Picris hieracioides L., Urtica dioica L., Taraxacum campylodes G.E.Haglund , Calystegia sepium (L.) R.Br., Rumex obtusifolius L., and Buddleja davidii Franch.). All are very common species that are also frequently found in other habitats of the study area. These species are perennial species that prefer the nutrient rich soils generally found in urban zones.

A total of 109 species among 365 were observed only once in wastelands but none of them is protected. These species could be rare species at the department scale (e.g. Lythrum hyssopifolia L., Euphorbia cyparissias L., and Verbascum lychnitis L., which were observed in two sites only over the entire department) or not adapted to wasteland habitat (e.g. Ranunculus ficaria L., Veronica hederifolia L., and Carpinus betulus L, which were found in numerous sites (>100) in non-wasteland habitats). In fact, only four species observed in wastelands are considered endangered at a regional scale (Cardamine impatiens L.) or of floristic interest in the region (Nepeta cataria L., Chondrilla juncea L., and Torilis nodosa (L.) Gaertn.) (CSRPN and DIREN Idf 2002), but all four were observed in at least three sites (up to 10 sites for

Torilis nodosa).

Finally, 73 (20%) observed species were naturalized species. Within sites, the fraction of naturalized species varied between 0 and 46%, with a mean of 15%. This proportion was not significantly dependent on wasteland age (r² = 0.009, P = 0.86). Each naturalized species remained somewhat rare, being found on average at 7.5 of the 98 sites (vs. 11.3 for indigenous species); in addition, 27 of the 73 naturalized species were found in a single site. However, a number of naturalized species that are considered invasive by managers of green spaces were also found in a large number of wasteland sites. Five of the main invasive species in the Greater Paris area, namely Buddleja davidii Franch., Solidago canadensis L., Ailanthus altissima (Mill.) Swingle, Robinia pseudoacacia L., and Reynoutria japonica Houtt. occurred respectively in 49, 21, 20, 20 and 18 of the 98 sites we inventoried. Senecio inaequidens DC., which is also considered invasive, had a lower occurrence (6/98), but was found almost exclusively in wastelands.

Relationship between floristic, environmental, and geographic distances

Regardless of site area, floristic distances were never significantly correlated with environmental distances (results not shown). In contrast, in the largest sites only (> 2500 m 2 ), floristic similarity among sites was significantly negatively correlated with geographic distance (Mantel statistics r = 0.214, P < 0.001, Figure 3); this remained true when the environmental distances among sites were accounted for (Mantel statistics r = 0.182, P = 0.003).

Impact of urban structures on floristic interest

We examined the relationship between (1) the floristic interest of sites (Rich, Ind and Rar) and

(2) the significant variables of the ANCOVA and the proportion of each urbanization class within a radius of 200 meter around the sites. The results of the hierarchical partition are given Table 2.

Species richness

Site species richness depended exclusively on site area and age: large wastelands or wastelands of intermediate age (4 to 13 years) had a higher richness than small or old (>21 years) wastelands (Figure 4 and5).

Rarity

Site rarity was influenced by season of inventory (rarity was highest in summer during which most of the plants flower, thus their detectability is at the maximum) and Land Use Pattern in the site (e.g. rarer species on average in "open zones" than in "built zones" wastelands, Appendix A). The presence of collective or individual dwellings within a radius of 200 m around a site had a significant negative independent effect on species rarity. Inversely, the presence of water (WATE) within 200 m was significantly positively correlated with higher species rarity.

Indigeneity

The indigeneity of a site depended on its distance from the center of Paris, with higher indigeneity away from Paris. In contrast to rarity, the presence of individual dwellings (INDI) within 200 m of a site was positively correlated with the proportion of indigenous species.

Discussion

We studied 98 wastelands sites in the Hauts-de-Seine department, one of the most densely populated areas in France, to document the significance of wastelands as reservoirs of floristic diversity in urban zones and to explore the factors controlling the diversity of wastelands. In the following, we review the main characteristics of wasteland floristic interest in Hauts-de-Seine and discuss the possible mechanisms underlying the observed patterns as well as the potential role of wastelands in the conservation of urban biodiversity.

Characteristics of plant communities in urban wastelands

Wastelands seem to be a relatively rich habitat of urban zones: their floristic richness represents 58% of the total richness observed in the whole study area, which makes wastelands the richest habitat in the department (365 species among the 626 species recorded), followed by urban lawns (303 species) [START_REF] Muratet | Diversité végétale en milieu urbain[END_REF]). This value is somewhat similar to that observed in a comparable study in Düsseldorf, where wasteland richness represented 51% of the total richness (277 among 550 species recorded in the whole city, Gödde and others 1995). Despite high species richness, a substantial fraction of the observed species consisted of naturalized species (20%), which is more than the department average of 17% (Student's t-Test = -2.75, pvalue = 0.006). This may have important consequences in terms of conservation policies. These species are characterized by (1) perenniality, (2) high dispersal capacity and seed production, and (3) high nitrophily (average N Ellenberg index of 7, [START_REF] Ellenberg | Indicator values of vascular plants in Central Europe[END_REF]. Cirsium arvense, Rumex obtusifolius and Buddleja davidii have also been mentioned by Godefroid and others (2006) among the most frequent species of wastelands in Brussels. On average, species that are considered invasive by managers were more frequent in wastelands habitat (e.g. 76% of sites harbored at least one of the six main invasive species (B. davidii, S. canadensis, A. altissima, R. pseudoacacia, R. japonica, and S. inaequidens) than in most other habitats of Hauts-de-Seine (on average 35% only of other sites were "invaded"). Wastelands have a similar level of invasibility than other highly disturbed habitats (i.e. edges of running water (65%), wet thickets (74%) or railway slopes (78%)) and could, similarly, play a major role in invasion processes.

Wastelands also exhibited a large number of locally very rare species (17), i.e. species that were observed only once in the department. These species, because they are locally very rare and occupy temporary habitats, owe their maintenance through time to the ability to migrate among sites. Their maintenance in the department relies on the existence of networks of wastelands, i.e. groups of geographically close wastelands that are connected through the existence of urban structures enabling gene flow (streets, parks, river banks…). In the wastelands of Hauts-de-Seine, we observed a single regionally protected species. The presence of such a rare species was unexpected because protected species generally grow in rare or declining habitats (such as marshes, calcareous lawns…) and are rarely observed in common, human-made habitats.

Factors influencing the floristic interest of wasteland sites in a densely populated urban area

Anthropogenic parameters

We defined wastelands as sites where vegetation management had been temporarily abandoned, but not long enough for an identified natural habitat (e.g. a forest) to settle. Therefore, a large number of wastelands arise following destruction of buildings (BUILVAC) and are temporarily abandoned until a new project is undertaken. This transition state of wastelands in building and vacant areas makes them highly vulnerable and, in contrast to other wastelands in the department, they are generally rather young because of the rapid turn-over of such habitats within cities (70% of wastelands in BUILVAC were younger than 13 years).

Wastelands in building sites and vacant areas were also the largest wasteland types (4494 m² on average).

A significant number of wasteland sites, ranging from 12 m² to 18260 m², have also been observed in other classes of the land use pattern. Wastelands observed in open urban and rural areas (OPENRUR) corresponded to green spaces that were less regularly managed than lawns.

They were older on average than sites in building and vacant areas (44% older than 21 years) and their floristic diversity was more likely to depend strongly on the intensity of management.

Wasteland sites observed in the different types of built urban areas (activities, individual and collective dwellings, facilities, transports) corresponded to abandoned industrial plots or to small interstices in the built matrix on which no urban projects were developed. These sites were significantly the smallest ones (808 m² on average) and were highly isolated (because they were surrounded by construction), but they were stable in the long term (92% were older than 21 years). Finally, old wastelands (all older than 21 years) could also be observed in natural plots (water and wood), but they represented a minor fraction (2%) of wasteland sites. These wastelands were observed in the disturbed edge of natural zones. Site age and area, which can be considered anthropogenic parameters because they are predominantly controlled by urban management, had a significant impact on the biodiversity of wastelands. The largest sites were also the richest ones (Figure 4), as found in similar studies [START_REF] Davis | Urban ecosystems and island biogeography[END_REF] and may be more resistant to potential disturbance than small sites [START_REF] Grumbine | Viable Populations, Reserve Size, and Federal Lands Management: A Critique[END_REF]). We also found a non-monotonous relationship between site age and richness, with a maximum floristic richness observed in sites of intermediate age (4 to 13 years, Figure 5). Old wastelands were significantly less diverse than sites of intermediate age.

This observed maximum richness at intermediate age could be attributable to classical patterns of ecological succession, with transient high levels of diversity but establishment of a smaller number of species in the stable communities resulting from the initial colonization/competition processes [START_REF] Horn | The ecology of secondary succession[END_REF]. Alternatively, as most (70%) old wastelands were situated in built urban areas, the urban matrix may isolate them from each other, thus preventing exchanges of species (via seeds for example) and resulting in lower levels of floristic diversity.

The proportion of indigenous species was found to be positively correlated with the distance from the center of Paris, a correlate of urbanization intensity. This pattern was already observed in other cities [START_REF] Mcdonnell | Ecosystem structure and function along urban-rural gradients: An unexploited opportunity for ecology[END_REF][START_REF] Mcdonnell | Ecosystem structure and function along urban-rural gradients: An unexploited opportunity for ecology[END_REF][START_REF] Maurer | The flora of selected urban land-use types in Berlin and Potsdam with regard to nature conservation in cities[END_REF] and highlights the role of major communication routes and perturbations in the establishment of non-native species. The average percentage of aliens in central European cities is 13.7% higher than their representation in the total species pool available in the region, indicating a remarkable concentration of aliens in urban areas [START_REF] Pysek | Alien and native species in Central European urban floras: A quantitative comparison[END_REF]. The proportion of non-native species rises from 6% in nature preserves outside the city of Berlin to 25% in the suburbs to 54% in the most intensively urbanized central areas [START_REF] Kowarik | Time lags in biological invasions with regard to the success and failure of alien species[END_REF].

Environmental parameters

Unexpectedly, we did not find evidence of any effect of the environmental parameters (parent rock, exposition, slope, or geomorphology) on wasteland floristic interest and, accordingly, the floristic distance among sites was not significantly correlated with their environmental distance. This likely does not reflect an actual absence of environmental effects in urban areas, as these are unavoidable and have been observed by Godefroid and others (2006) in Brussels, but may rather be due to a limited accuracy of measurements or a lack of statistical power. For example, the use of parent rock (not retained in the final model of ANOVA) to describe site environment might be inefficient because, in urban and industrial areas, the ground is usually covered with alien substrates from other geographical regions (e.g. railway ballast, mine spoil), with artificial substrates (e.g. slag), or with a mixture of artificial and natural substrates (e.g. slag/clay) [START_REF] Rebele | Urban ecology and special features of urban ecosystems[END_REF]). In addition, site slope and light exposition were very homogeneous across studied wastelands, with more than 75% of sites having a slope < 1 (i.e. no specific exposition). This is the result of strong urbanization, which tends to level all surfaces. Site geomorphology (hill, plain, valley etc…) may not have any effect on floristic interest in sites of such a fragmented landscape and may be relevant only in large natural landscapes. Finally, soil type and especially soil fertilization frequently have a conspicuous role in floristic composition and diversity in urban context (Peet and others 1983). However, this parameter was so heterogeneous within a site that a much larger number of samples would have been required to adequately explore its influence.

Isolation by distance, migration among wasteland and non-wasteland sites

In large sites, we detected significant isolation by distance for floristic diversity, i.e. large neighboring sites were often floristically more similar than distant sites. This, together with the observed absence of environmental effects on floristic interest, suggests that migration of species among large sites might have a significant impact on local floristic composition, which is influenced by the pool of available species in the neighborhood and their capacity to disperse/colonize. In contrast, small sites, which have a lower probability to receive propagules by chance and are often isolated by the matrix of buildings, are less influenced by the composition or surrounding sites.

The supply of species from non-wasteland sites likely influenced floristic interest in wastelands as well. We have shown that the presence of water (rivers and ponds notably) In contrast, the presence of collective and individual dwellings around a site had a significant negative influence on floristic rarity. These findings have major consequences for management plans. They indicate, for example, that the establishment of dwelling districts will certainly reduce the quality of neighboring wastelands and more generally of green lands.

However, the presence of individual dwellings seemed to have a positive impact on the proportion of indigenous species in the wastelands around. This result is somewhat surprising, as small gardens surrounding houses were expected to host many exotic species for ornamental purpose. It may be that these cultivated exotic species remain "trapped" in the gardens, and that meticulous gardening practices prevent the dispersal and establishment of exotic species considered as weeds (e.g. Conyza canadensis and C. sumatrensis, Artemisia annua …)

Conclusions

Our study confirms that urban wastelands are a habitat of crucial importance in urban areas.

Thanks to their large and homogeneous distribution in the department, they host a substantial proportion of floristic diversity of cities and certainly play a key role in favoring exchanges among urban habitats. Thus, although the preservation of wastelands has little consequence for the conservation of rare species in the present case, our study shows that their maintenance is necessary to protect urban biodiversity as a whole because they favor connectivity among sites and provide a source of species to colonize vacant sites. Unfortunately, wastelands are one of the more threatened urban habitats due to urban intensification: for example, the total surface of building and vacant areas, which contain a large number of wastelands (Figure 2), has decreased by 24% in the last 20 years in Hauts-de-Seine.

In this work, we could not study whether the presence of wastelands has a positive impact on the floristic diversity of other habitats in the department because we did not locate all wastelands and could not correlate the diversity at a given site with the presence/absence of wastelands in the neighborhood. However, we have shown that humid zones have a positive impact on the mean species rarity of wastelands, and it is highly likely that, in a symmetrical way, wastelands are beneficial to other habitats. Assessing the actual impact of wasteland sites on the floristic diversity of other urban habitats will require further investigation; nonetheless our results suggest a number of simple actions that could be taken to favor the preservation of the floristic diversity in the Hauts-de-Seine. For example, the maintenance and the creation of large sites (> 2500 m²) could favor species exchange among wastelands, lower the extinction risk of plant populations and provide seeds to colonize new sites. Note however that invasive species were frequently found in wastelands, so highly connected sites could also act as relays to favor the spread of these species. The role of wastelands, a habitat submitted to frequent disturbance, in species invasion remains unclear. Additional information regarding abundance within sites is required to understand the dynamics of invasive species in these habitats.

Finally, we also suggest that too rapid turnover should be avoided: waiting around ten years before construction of new buildings would leave enough time for installation of highly diverse plant communities and for a maximum contribution to the floristic diversity of surrounding sites. Table 1. Age classes inferred from the Land Use Pattern (LUP) updates; number and average area of wasteland sites inventoried in each class. Significant differences in area (P ≤ 0.05, Tukey's test) are indicated by different letters. Two groupings were performed so that each age class contained at least five inventories. Age could be inferred for 88 of 98 sites only, because LUP type was not available for 10 wasteland sites. 5

Table 2. Results of a hierarchical partition on r 2 for independent contributions from 1000 randomizations of data matrices for potential explanation of the floristic interest of sites (Rich, Ind and Rar). Stars indicate significant contribution at the upper 95% confidence limit, The plus (respectively minus) sign indicates a positive (respectively negative) relationship. The explanatory variables include the significant variables of the ANCOVA for each floristic 10 interest parameter, plus the proportion of each urbanization class (Appendix A) within a radius of 200 meter around the sites.

  Audrey MURATET 1,* -Nathalie MACHON 1,2 -Frédéric JIGUET 2 -Jacques MORET 1 -Emmanuelle PORCHER 1,2 1 Conservatoire botanique national du Bassin parisien, UMS Inventaire et Suivi de la Biodiversité, Muséum national d'Histoire naturelle, 61 rue Buffon F-75005 Paris, France; 2 UMR 5173, Conservation des Espèces, Restauration et Suivi des Populations, Muséum national d'Histoire naturelle, 61 rue Buffon F-75005 Paris, France; * Corresponding author. Tel: +33 1 40 79 35 45; fax: +33 1 40 79 35 53. E-mail address: muratet@mnhn.fr

A

  large number of species observed in a majority of the wastelands were generally temperate, almost cosmopolitan species: Artemisia vulgaris L. (found in 81% of the sites), Cirsium arvense (L.) Scop. (72%), Plantago lanceolata L. (65%), Picris hieracioides L. (63%), Urtica dioica L. (59%), Taraxacum campylodes G.E. Haglund (52%), Calystegia sepium (L.) R.Br. (51%), Rumex obtusifolius L. (50%), and Buddleja davidii Franch. (50%).

  within a radius of 200 m had a positive impact on the floristic rarity of a wasteland. Humid habitats are generally highly diverse and harbor many rare species of high conservation interest. In wastelands situated near water sites, we found Alnus incana (L.) Moench (that was observed in only four sites in the department (986Sites), Alopecurus pratensis Mattf. (four sites), Cardamine impatiens L. (four sites), Equisetum telmateia Ehrh. (three sites) and Stachys palustris L. (five sites).

Figure 1 .

 1 Figure 1. Location of the study area, the French department of Hauts-de-Seine (Built zones are

Figure 2 .

 2 Figure 2. Percentage of wastelands in each LUP type. Open urban areas are in grey, natural

Figure 3 .

 3 Figure 3. Floristic similarity (Jaccard index = fraction of total species found in both sites)

Figure 4 .

 4 Figure 4. Relationship between site richness and site area.

Figure 5 .

 5 Figure 5. Average richness as a function of wasteland age. Error bars indicate standard

Table 1

 1

	AGE CLASSES (years)	LUP CONDITIONS	NUMBER OF WASTELANDS	MEAN AREA (m²)
	Younger than 4	In the same LUP class since 2003	10	2027(BC)
	Between 4 and 13	In the same LUP class from 1994 to 1999	15	4053 (AB)
	Between 13 and 21	In the same LUP class from 1987 to 1990	5	6743 (A)
	Older than 21	In the same LUP class since 1982	58	1261 (C)

Table 2

 2

	RICHNESS	RARITY	INDIGENEITY

Acknowledgments

Funding for this research was provided by the Conseil Général des Hauts-de-Seine. We thank Yoann Faivre for his help on the field, Gérard Arnal, Sébastien Filoche and Fabrice Perriat for their help in plant identification and three anonymous reviewers for helpful comments on an earlier version of the manuscript.

Appendix A. Abbreviations for Land Use Pattern classes [START_REF] Iaurif | Institute for Planning and Development of the Paris Ile-de-France Region[END_REF].

LAND USE PATTERNS DEFINITIONS

ACTI