Response of edge communities to aging
We found a negative influence of edge age on species richness at both the plot (α) and transect (γ) scales, suggesting that few species are able to survive the environmental conditions that progressively develop as the edge ages, especially severe shading and leaf litter deposition (see Figs 1 and 2). This contrasts with the species accumulation predicted by the species–time relationship (STR; Rosenzweig 1995) and observed in the early stages of forest succession (Brunet 2007). The distance from the outer limit of the edge is the most important factor explaining species richness, with a clear gradient of decreasing richness from the edge to the forest interior, consistent with the well-known ‘edge effect’ (Saunders et al. 1991; Murcia 1995). As expected, this gradient becomes more pronounced as time since edge formation increases (Matlack 1993; Hermy et al. 1999; Honnay et al. 2002a; Brunet 2007). The strongest decrease in species richness was observed within the first 3 m. Similar or even lower values are reported in the literature (Brothers & Spingarn 1992; Honnay et al. 2002a); moreover, the few penetrating species have sharply decreasing cover values when expanding towards the forest interior (Honnay et al. 2002a).
Interestingly, the reverse trend was found for relative light intensity, with the outer plots receiving more light than the inner ones in recent edges (26 yr), and the inner plots receiving more light than the outer ones in ancient edges (300 yr; Fig. 1a). These light patterns contrast with those usually expected (e.g. Hermy et al. 1999; Brunet 2007), but are consistent with the competition-induced wave of biomass (Reichman et al. 1993): the high light availability at the beginning of the transect creates a peak of leaf density that in turn casts a shadow just behind it before light availability increases to intermediate values further behind. As a result, edges are likely to become less permeable to species coming from outside, which are predominantly light-demanding species, an effect already reported for ancient forest edges (Henry & Aarssen 1997; Honnay et al. 2002a; Marchand & Houle 2006); hence the penetration distance is expected to decrease as edge age increases (Meiners & Pickett 1999).
The increased litter thickness of aging edges represents a supplementary physical and chemical (low pH) barrier to many non-forest species. It is expected to promote a limited set of specialized species (Sydes & Grime 1981; Decocq & Hermy 2003), as well as tall competitors that use vegetative propagation, such as Pteridium aquilinum and Rubus fruticosus coll. (both associated with the oldest forests in the RDA), which may accelerated the loss of smaller species as the edge ages.
The RDA confirms that edge age is the most influential factor on species composition, with a clear shift along the first axis, which is associated with increasing litter thickness and decreasing shrub and herb layer cover: light- and/or nitrogen-demanding species (e.g. Geum urbanum, Poa trivialis, Galium aparine) are progressively replaced by shade-tolerant, mesotrophic species (e.g. Poa nemoralis, Hyacinthoides non-scripta, Anemone nemorosa). Remarkably, this species shift equally affects the plots of a given transect, since edge age did not affect species turnover along the transect (β-diversity), and plot position was marginally significant in the RDA. No niche partitioning effect was thus evidenced, contrary to our expectations and the theoretical models of edge structure (Matlack 1994; Mikk & Mander 1995; Murcia 1995; but see Alignier & Deconchat 2011). Instead, our results suggest a specialization of the flora, which primarily benefits to the so-called ‘ancient forest species’ (Peterken & Game 1984; Hermy et al. 1999). The latter are indeed known for not being suppressed by edge effects (Fraver 1994; Honnay et al. 2002a) and for accumulating over time as, due to their dispersal and recruitment limitations, they are poor colonizers (Verheyen & Hermy 2001). The positive correlation between edge age and patch area (see Fig. 3) suggests that larger forest patches may keep their edges supplied with more diaspores of ancient forest species compared to younger and smaller patches.
We conclude that an edge–forest interior environmental gradient progressively develops as an edge ages (H1); however it poorly translates into species richness and not at all into species composition (no age effect on β-diversity), consistent with the idea that the depth of influence of species richness differs from the depth of influence of microclimatic variables (Gehlhausen et al. 2000; Alignier & Deconchat 2011).
Response of edge communities to adjacent land management
Our findings highlight the key impact of matrix management on species richness and composition of forest edges. As management intensity (quantified by the variable LAND in this study) increased, species richness decreased at both the plot (α) and transect (γ) scales without influencing species turnover (β). At the same time, communities shifted from mesophilous, calciphilous (in edges of open fields) to more hygrophilous, acidiphilous (in edges in bocage) assemblages (see RDA axis 2). These differences in species composition may be partly explained by the fact that land use and management decisions depend on environmental conditions, especially soil quality. In our study area, substrates that were nutrient-poor and/or waterlogged were not suitable for intensive agriculture and thus were more prone to be patterned as ‘bocage-like’ landscapes. Hence, forest edge communities are expected to be more acidiphilous and/or hygrophilous in bocages than in open fields. However, three types of factor associated with agricultural practices can also explain the observed patterns.
First, mechanical disturbances are more common in open field landscapes. Ploughing and mowing operations conducted in croplands often expand to the field margins, including herbaceous fringes at the field–forest boundary, and may directly reduce species richness. Moreover, forest edges are maintained by thinning so that the outer branches of trees and shrubs bordering cultivated fields do not overhang the cropland, to minimize crop shading and allow agriculture vehicles to access the field outer limit. This is well reflected by the high relative light intensity in the first two plots; as a result, a few shrub and lower shrub species (e.g. Crataegus monogyna, Ulmus minor, Rubus fruticosus coll.) often form a dense curtain which shades out many other species (Honnay et al. 2002a; Alignier & Deconchat 2011), explaining the strong decrease in species richness in the first four plots. In contrast, in bocage landscapes forest patches are usually adjacent to pastured meadows and, due to grazing pressure, forest edges are typically cantilevered edges, characterized by an overhanging canopy of branches that grow above the adjacent meadow. Such a structure acts as a shelter that shades the edge understorey, buffering it from desiccation and light penetration (Murcia 1995), as reflected by the reduced relative light intensity in the first two plots of the transect. Highly competitive species such as Rubus fruticosus coll. and Urtica dioica may be hampered from becoming dominant, hence space and resources are likely available for smaller-sized, less light-demanding species (Endels et al. 2004), potentially explaining why species richness peaked 4 m inside the forest.
Second, agrochemicals are much more heavily used in open fields compared to bocage landscapes (Marshall & Moonen 2002). In open fields fertilizer input from neighbouring arable land is likely to favour nitrophilous species along edges (e.g. Alliaria petiolata, Ulmus minor), to lead to the dominance of competitive-ruderal species (e.g. Rubus fruticosus coll., Urtica dioica), and hence to reduce species diversity (Marrs 1993; Kleijn & Snoeijing 1997; Honnay et al. 2002a). Moreover, nutrient cycling processes may be affected in the first plots through increased solar radiation, which in turn increases the activity of soil microorganisms and invertebrates (Klein 1989; Parker 1989) and thus accelerates litter decomposition and nutrient release. Similarly, lime drift is likely to increase soil pH along forest edges, promoting calciphilous species such as e.g. Viburnum lantana, Orchis purpurea and Melica uniflora (Honnay et al. 2002a). The drift of other agrochemicals, particularly herbicides, may directly contribute to reduce species diversity (Kleijn & Snoeijing 1997).
Third, the composition of the landscape matrix can have a strong influence on the seed rain entering edges, as most species found in small forest patches are non-forest species (Jamoneau et al. 2011). As bocage landscapes consist of a mosaic of grasslands, small fields and hedgerows, both density and diversity of diaspore sources are expected to be much higher than for open fields. Moreover, hedgerows that connect forest patches may act as ecological corridors, allowing plant species and their animal vectors to migrate along them (Corbit et al. 1999; Damschen et al. 2006); the movements of these vectors (e.g. birds, rodents, ants) are also likely to be facilitated by the lighter use of biocides (Marshall & Moonen 2002; Murphy & Lovett-Doust 2004). This increasing permeability of the landscape matrix with decreasing management intensity is likely to increase propagule pressure on forest edges (Marshall & Moonen 2002; Jamoneau et al. 2011, 2012), and may explain both the higher species richness and the presence of typically weak dispersers (e.g. Viola reichenbachiana, Lamium galeobdolon) in forest edges of bocages compared to open fields.
It should be noted that landscape type and edge age are not fully independent in our study, as indicated by the significant interaction term in ANOVAs and collinearities among explanatory variables in the RDA. This was expected as forest fragments have been reported to be larger and older in open field landscapes than in bocages. Similarly, the negative correlation between slope and age reflects the fact that most slopes became afforested after the agriculture mechanization of the 1960s, when vehicles could no longer drive where animals were once used (Jamoneau et al. 2010, 2011, 2012).
We conclude that the edge–forest interior transition zone tends to be shortened as the management intensity of adjacent lands increases, with more abrupt changes in environmental conditions, species richness and floristic composition (H2); hence forest edges resemble more a sharp transition zone (i.e. an ecotone sensu Van der Maarel 1990) as proximal disturbances increase in intensity.