- Top of page
- Materials and methods
- Supplementary material
- Supporting Information
More than a decade ago Bradshaw (1987) suggested that restoration could serve as an ‘acid test’ for our understanding of ecological processes. Although small-scale experiments are increasingly used to elucidate factors influencing forest recovery (Cabin et al. 2002; Hooper, Condit & Legendre 2002; Sweeney, Czapka & Yerkes 2002), multiple restoration sites can be valuable as experiments to evaluate factors that influence forest restoration or to test the applicability of general predictions of ecological theory to management questions. These opportunities have rarely been realized (Holl, Crone & Schultz 2003).
Although landscape patterns (the spatial relationship of ecosystems) and processes (the flow of genes, individuals, materials and energy across large areas) are important to ecosystem recovery (Forman & Godron 1986; Bell, Fonseca & Motten 1997; Holl, Crone & Schultz 2003), the relative importance of local- vs. landscape-level patterns and processes on ecosystem restoration has rarely been measured in the field (Holl, Crone & Schultz 2003). Where similar restoration techniques have been repeated at multiple sites that differ with respect to local biotic and abiotic conditions, as well as surrounding land use, community composition of these sites can be used to assess the importance of local vs. landscape factors for ecosystem recovery. In previous work in a range of managed landscapes, some studies have highlighted the importance of landscape patterns for community composition (Saab 1999; Mitchell, Lancia & Gerwin 2001; Luck & Daily 2003). A few have suggested that local parameters are more important than the surrounding landscape (Brose 2001; Clergeau, Jokimaki & Savard 2001; Graham & Blake 2001) and most have shown that patterns and processes at multiple scales affect community composition (Mörtberg 2001; Verheyen & Hermy 2001; Fisher, Suarez & Case 2002; Fletcher & Koford 2002; Lee et al. 2002).
For plants, the relative importance of local vs. landscape patterns for community composition is likely to vary with dispersal ability (Brose 2001; Verheyen & Hermy 2001; Campbell, Rochefort & Lavoie 2003). For example, in abandoned agricultural lands forest recovery is often limited by proximity to seed sources (a landscape factor). Restoration efforts that increase local woody vegetation cover are likely to enhance seed dispersal of animal- but not wind- or gravity-dispersed species (Robinson & Handel 2000; Harvey 2000; Holl 2002), thereby lowering the relative importance of landscape patterns for certain dispersal guilds.
Restored sites that differ in size and isolation from natural vegetation could also serve as a test of specific predictions from island biogeography theory (MacArthur & Wilson 1967) for restoration planning. In most restoration projects a few species are planted, with the expectation that others will colonize naturally once suitable site conditions (e.g. light and nutrient levels, safe sites for germination, mutualistic species) are available. If extinction–colonization dynamics are important for natural colonization of restored sites, island biogeography theory predicts that few species would be found in smaller and/or more isolated restoration sites. However, most restoration efforts, including our study system, differ from controlled manipulation of patch size and isolation (Cairns et al. 1969; Simberloff & Wilson 1969; Molles 1978; Dickerson & Robinson 1985) in several important ways. First, restored sites are seldom, if ever, true experimental replicates, identical in all ways except size and isolation. In restoration sites the effects of patch size and isolation may be small, relative to heterogeneity in abiotic and biotic conditions within sites and the vegetation matrix separating them (Lomolino & Perault 2001; Ricketts 2001; Fleishman et al. 2002). Thus, the community composition of restored sites suggests whether patch size and isolation are important variables for land managers to consider relative to other sources of variation, not whether size and isolation would be important, all else being equal. Secondly, if restoration ‘success’ is evaluated, it typically occurs within 5–10 years of implementation (Holl & Cairns 2002), whereas species richness may take decades to reach equilibrium. Thus, colonization patterns of restored sites tell us whether patch size and isolation matter over typical restoration and management time frames, not whether they determine equilibrium patterns.
We analysed large-scale riparian forest restoration at multiple sites along the upper Sacramento River (California, USA). This river, the largest in California, has been dammed and leveed for flood control and irrigation since the late 1800s. The Sacramento River riparian ecosystem was heavily deforested for fuelwood in the second half of the 19th century, with additional clearing in the 1950s and 1960s for conversion to orchards and row crops (California Resources Agency 2000). By the late 1970s only 5·5% of the original riparian forest cover remained (Greco 1999). The remaining forest is highly fragmented and impacted by altered hydrology and invasion by exotic species.
Several private and public agencies are working together to acquire lands and alter flow regimes to maintain and restore both hydrological processes and riparian habitat, under the mandate of California Senate Bill 1086 and the CALFED Bay Delta Program (CALFED Bay-Delta Program; California Resources Agency 2000). Thus far, most restoration efforts along the Sacramento River have focused on replanting orchards with native trees and shrubs, although geomorphological and hydrological processes are an increasing focus of restoration efforts. The Nature Conservancy (TNC) and other organizations aim to purchase properties within the 2·5-year floodplain along the 160 km of river between Red Bluff and Colusa (Griggs 1993). To date these organizations have planted approximately 2000 ha of riparian land with native tree species, with the hope that over time additional native flora and fauna will colonize the sites. Although there have been no previous systematic surveys or mechanistic studies of the understorey flora, observations suggest that restored sites are often dominated by aggressive exotic species, such as Centaurea solstitialis, Lolium perenne and Sorghum halepense (TNC, unpublished restoration reports). These observations further suggest that the assumption that native understorey communities will naturally recover has not yet been met, and that establishment is limited by factors at the local and/or landscape scale.
Our goal was to assess the relative importance of several local- and landscape-level variables on natural establishment of understorey vegetation at 15 sites planted with trees to restore forests in four sections of the upper Sacramento River (Fig. 1). Based on predictions of island biogeography theory and field observations, we hypothesized that (i) native species richness and cover would be higher in sites that were larger, were closer to forest and had lower exotic cover; and (ii) species richness and cover of exotic species would be higher in sites surrounded by a high proportion of fallow lands, where herbicides are not used to control exotic species of agricultural concern. To quantify different scales at which the surrounding landscape might affect restoration success, we summed the amount of surrounding forest and fallow lands over different distances. In addition, we compared the proportion of within- and among-site variance in native and exotic species richness and cover explained by landscape effects, relative to probable sources of biotic and abiotic heterogeneity. Specifically, we included elevation and distance to the river (as indicators of flood interval and moisture availability), soil texture and past land use as abiotic factors that are typically considered important in restoration planning.
Figure 1. Map of study sites and the surrounding landscape. Older restoration sites include the 15 sites sampled. Other restoration sites include more recently restored sites and older sites that were not sampled. Agriculture/other land includes orchard and row crops and other land cover, such as gravel and developed land.
Download figure to PowerPoint
- Top of page
- Materials and methods
- Supplementary material
- Supporting Information
We recorded 40 native and 58 exotic species, in addition to 11 plant genera or families for which we could not identify the origin. All of the common overstorey species, such as Acer negundo, Populus fremontii, Quercus lobata and Salix spp., were native and were planted in restored sites. As expected, average overstorey cover was higher in reference sites, although two of the older restored sites had > 50% overstorey cover, approaching the 74–90% range of remnant forests (Table 1).
Table 1. Vegetation species richness and cover in newly restored (1 year old, n= 3), older restored (5–12 years old, n= 15) and reference sites (n = 5). Values are means per site (minimum–maximum). Means with the same letter are not significantly different (P < 0·05) across habitat type based on Tukey's LSD
| ||Newly restored||Older restored||Reference|
|Overstorey cover|| 0a||29·2 (3·7–54·4)b||82·0 (73·5–90·0)c|
|Understorey native richness*|| 3·3 (2–5)a|| 5·1 (2–8)a||11·8 (11–13)b|
|Understorey native cover*|| 1·5 (0·3–3·3)a|| 9·7 (1·3–35·7)a||50·7 (30·8–66·7)b|
|Understorey exotic richness||17·3 (14–20)a||15·4 (8–30)a||10·2 (4–14)a|
|Understorey exotic cover||42·0 (35·2–46·2)a||40·0 (10·2–66·6)a||21·3 (4·0–44·7)a|
|Understorey total cover||40·3 (35·0–45·8)a||50·1 (30·7–78·8)a||73·3 (64·2–79·8)b|
Native understorey cover in older restored sites was intermediate between reference forests and newly restored sites (Table 1). Native understorey species richness was much higher in reference forests compared with older restored sites, despite the fact that in many cases larger areas were surveyed in older restored sites (Table 1). Understorey exotic cover and species richness were not significantly different in reference and restored sites. Variance in native and exotic cover and species richness was high among sites, particularly the older restored sites (Table 1). The exotic grasses Bromus spp., Cynodon dactylon, Lolium multiflorium, Sorghum halepense and Vulpia spp., as well as Brassica spp., dominated understorey cover in older restored sites (Appendix S2 in Supplementary material). In these sites, only two native understorey species, Artemisia douglasiana and Galium aparine, were common. A number of other native species common in the remnant forest understorey, such as Aristolochia californica, Carex barbarae, Rubus ursinus and Vitis californica, were much less common in or absent from restored sites (Appendix S2 in Supplementary material). Thus, restored riparian forests appeared to be recruiting some native plant species, but the composition of older restored sites still differed substantially from remnant forest.
In general, biotic and abiotic variables, rather than the amount of surrounding fallow land, explained substantially more of the variance in both exotic species richness and cover at older restored sites (Fig. 2). When data were analysed by quadrat, exotic species richness and cover differed strongly among sites (Table 2). Both were higher with lower overstorey cover and higher elevation (Table 2). Exotic species richness was significantly higher near the river, although only 1–2% of the variance was explained (Table 2); surprisingly, distance to the river was weakly negatively correlated with elevation (R = −0·16, P= 0·0012). The small amounts (c. 1%) of the variance in exotic species richness (bare ground only) and cover explained by bare ground and percentage fallow land at 50 m were not significant after Bonferroni correction (Table 2). At the site level, none of the 11 variables tested was significantly related to exotic species cover or richness, even before Bonferroni correction (Table 3).
Figure 2. Proportion of variance explained by biotic and abiotic local variables (overstorey cover, exotic cover, bare ground, elevation, soil texture), surrounding landscape (distance to forest, distance to Sacramento River, percentage surrounding forest or fallow land) and patch size and site history (age, past land use) for (a) exotic and (b) native species. Quadrat- and site-level results were combined by multiplying the site-level coefficients by the proportion of variance explained by site in the quadrat analysis, and summing these with quadrat results.
Download figure to PowerPoint
Table 2. Stepwise forward regression of abiotic, biotic and landscape variables on native and exotic understorey cover and species richness at quadrats (1 × 1 m)
|Exotic species richness, n= 538 (all quadrats)|
| ||Sites||14|| ||0·22|| 10·5||< 0·0001*||Linear|
| ||Overstorey cover|| 1||−0·316||0·09|| 63·9||< 0·0001*|| |
| ||Distance to river|| 1||−0·207||0·02|| 11·8|| 0·0006*|| |
| ||Bare ground|| 1|| 0·006||0·01|| 5·7|| 0·0172|| |
|Exotic species richness, n= 394 (subset of quadrats with elevation data)|
| ||Sites||10|| ||0·14|| 6·3||< 0·0001*||Linear|
| ||Overstorey cover|| 1||−0·317||0·09|| 42·4||< 0·0001*|| |
| ||Elevation|| 1|| 0·300||0·04|| 18·5||< 0·0001*|| |
| ||Distance to river|| 1||−0·191||0·01|| 7·2|| 0·0071|| |
| ||Bare ground|| 1|| 0·128||0·01|| 5·4|| 0·0201|| |
|Exotic species cover, n= 538 (all quadrats)|
| ||Sites||14|| ||0·17|| 7·6||< 0·0001*||Linear|
| ||Overstorey cover|| 1||−0·367||0·11|| 76·9||< 0·0001*|| |
| ||Bare ground|| 1||−0·121||0·01|| 8·9|| 0·0030|| |
| ||Fallow land at 50 m|| 1|| 0·093||0·01|| 5·2|| 0·0231|| |
|Exotic species cover, n= 394 (subset of quadrats with elevation data)|
| ||Sites||10|| ||0·10|| 4·3||< 0·0001*||Linear|
| ||Overstorey cover|| 1||−0·43||0·13|| 67·4||< 0·0001*|| |
| ||Elevation|| 1|| 0·41||0·03|| 15·2 ||< 0·0001*|| |
| ||Bare ground|| 1||−0·10||0·01|| 4·5|| 0·0350|| |
|Native species richness, n= 538|
| ||Sites||14|| ||0·21||109·4 ||< 0·0001*||Logistic|
| ||Exotic cover|| 1||−0·18||0·03|| 16·7||< 0·0001*|| |
| ||Distance to forest|| 1||−0·23||0·01|| 7·5|| 0·0063|| |
|Native species cover, n= 538|
| ||Sites||14|| ||0·21||118·4||< 0·0001*||Logistic|
| ||Exotic cover|| 1||−0·18||0·07|| 37·8||< 0·0001*|| |
| ||Distance to forest|| 1||−0·48||0·02|| 14·6 ||< 0·0001*|| |
Table 3. Stepwise forward regression of local and landscape variables on native and exotic cover and species richness at the site level
|Exotic species richness|
|Native species richness|
|Native cover||Forest at 1000 m|| 0·63||1||0·31|| 5·9||0·0306|
|Exotic cover||−0·50||1||0·24|| 6·6||0·0246|
|Native wind-dispersed||Forest at 1000 m|| 0·70||1||0·49||12·4||0·0037*|
|Native externally dispersed||Past land-use fallow||−0·70||1||0·49||12·5||0·0037*|
|Native gravity- or water-dispersed||Distance to river||−0·73||1||0·54||15·2||0·0018*|
Native species composition at older restored sites was explained by biotic interactions, in concert with connectedness with forest (Fig. 2). Like exotic species, at the quadrat level native species richness and cover differed strongly among sites (Table 2). In addition, native species richness and cover decreased somewhat (3–7% of the total variance) with increasing exotic cover (Table 2). We found more native species in quadrats closer to forest, although this difference was small relative to the total variation and significant only for cover after Bonferroni correction (Table 2). Including elevation in native species cover and richness models did not explain additional variance. At the site level, average native cover was positively related to percentage forest at 1000 m surrounding the site, and negatively related to average exotic cover; although these relationships explained a substantial amount of variance in native cover (31% and 24% of the variance, respectively), they were not significant after Bonferroni correction given the low number of sites (Table 3). Contrary to our expectations, none of the independent variables in the site-level model, including site area and time since restoration, explained a significant amount of variation in native species richness.
Separating native species by dispersal strategy, we found higher than average cover of wind-dispersed species, the most common of which was Artemisia douglasiana, in regions with high forest cover (Table 3 and Fig. 3a). Not surprisingly, gravity- or water-dispersed species (e.g. Carex barbarae and Urtica dioica) were more likely to be found nearer the river (Table 3 and Fig. 3b). Cover of native externally animal-dispersed species (primarily Galium aparine) was always low at sites that had been left fallow for a few years prior to restoration, whereas in sites that were utilized for orchard or row crops immediately prior to restoration, cover was variable but on average higher (Table 3 and Fig. 3c). The presence of internally animal-dispersed species was sufficiently low that it was not possible to draw conclusions about distribution patterns. Interestingly, sites that were located in close proximity to one another (indicated by the same two-letter symbol on Fig. 3) varied substantially in native species composition, exotic species composition (data not shown) and most landscape and local variables.
Figure 3. Significant explanatory variables of native plant cover at the site level by dispersal strategy (a) Wind-dispersed; (b) gravity- or water-dispersed; (c) externally animal-dispersed. Site locations are indicated by two letter acronyms (see Appendix S1 for full names). Note differences in y-axis scales.
Download figure to PowerPoint
- Top of page
- Materials and methods
- Supplementary material
- Supporting Information
The colonization of restored sites by native species appeared to be limited by the presence of exotic understorey species and lack of connectivity with remnant forest, whereas cover of exotic species was primarily associated with low percentage overstorey cover and, to a lesser degree, high floodplain position (Tables 2 and 3). Dispersal limitation may be important for particular exotic species during the invasion process. We suspect, however, that the exotic species that establish in an area at high abundance are likely to have a nearly ubiquitous seed distribution in landscapes dominated by disturbed lands. Therefore, native species, which primarily occur in habitat remnants, are more likely to be dispersal limited. None the less, in our system, native species distributions were more negatively related to presence of exotic species than to isolation.
Based on our results, the best way to ensure successful establishment of native understorey species in this system may be first to choose sites with low elevation relative to river base flow and near remnant forests, and then tend planted overstorey species so that the canopy closes quickly and exotic understorey species are shaded out. Seeding or planting species after establishing an overstorey cover would almost certainly increase establishment rates, but would cost a great deal and would require a longer term commitment than is currently typical (3 year) for these projects.
Species’ distributions in restored sites support the importance of both local- and landscape-scale factors, both within and among sites, although local abiotic and biotic factors explained a larger proportion of the total variance (Fig. 2). Exotic species cover was most affected by overstorey cover of planted native species, which is probably determined by local conditions, but also by landscape position relative to the floodplain (Alpert, Griggs & Peterson 1999). Native understorey species’ distributions were negatively related to cover of exotic understorey species. Although statistically significant, landscape position effects were smaller than has been found in other studies of vegetation distribution in floodplains (Rot, Naiman & Bilby 2000; van Coller, Rogers & Heritage 2000; Drezner, Fall & Stromberg 2001). This may be because TNC primarily conducts horticultural restoration in the 2·5-year floodplain, so sites are relatively uniform in elevation and proximity to the river. It also may be because many studies of riparian communities focus on overstorey, rather than understorey, plant species (Scott, Friedman & Auble 1996; Shafroth, Stromberg & Patten 2000; van Coller, Rogers & Heritage 2000).
Native species with different dispersal mechanisms were affected differentially by local and landscape factors. Wind-dispersed species were most abundant when the surrounding landscape within 1 km was more than 20% remnant forest (Fig. 3). Externally animal-dispersed species were best explained by past land use, a site feature, whereas gravity- and water-dispersed species were most abundant within 250 m of the Sacramento River main channel (Fig. 3).
As a test of the relevance of island biogeography theory to restoration, native species distribution in restored forests provides weak support for effects of isolation. Native species richness and cover were significantly higher near remnant forests, but these effects explained only about 7% of the total variation in native cover and about 1% of the total variation in species richness (Table 1). In part, this relationship might be weakened by analysis at the community level. For wind-dispersed species cover, percentage forest in the surrounding landscape explained about 50% of the among-site variance in cover, which would be approximately 10–12% of the total variance.
Our sites violate a number of assumptions of island biogeography theory. Some native species are present in areas other than remnant forests, and not all remnant forests contain all native species. Our predictive power would almost certainly increase if we could map all potential source populations of all native species, and relate species-specific colonization probabilities to isolation (Bastin & Thomas 1999). In addition, while we assume that few seeds of native species were present in the seed bank due to > 30 years of intense agricultural use, including regular herbicide treatments of herbaceous species, past studies in long-used agricultural lands in the temperate zone (Hutchings & Booth 1996; Bekker et al. 1997; Middleton 2003) suggest that some seeds of some native understorey plants may persist under such conditions. However, these violations of the assumptions of island biogeography theory are common, if not universal, to restored sites in complex landscape mosaics. Therefore, they do not undermine our ability to test whether effects of patch size and proximity to remnant habitat, drawn from island biogeography theory and commonly recommended as considerations in restoration planning (e.g. Sauer 1998; Hobbs 2002), are important predictors of restoration success.
Our results provide no support for effects of patch size on species richness. Although species–area relationships are widely documented in natural habitat remnants (Diamond 1972; Freemark & Merriam 1986; Laurance et al. 2002), they may not be applicable to restored systems approximately a decade after establishment, because they refer to long-term equilibria. Plant communities in the restored forests we surveyed were intermediate between newly planted restoration sites and reference forests (Table 1), suggesting ongoing succession. Furthermore, a number of recent studies point to the importance and potential confounding effects of among-site variation in patch quality (Brose 2001; Foster 2001; Verheyen & Hermy 2001; Fleishman et al. 2002). In our sites, among-site variation in biotic and abiotic factors was high, which is typical for restored sites.
A few unexpected trends resulting from our analyses bear further discussion. First, we were surprised that site age (i.e. time since restoration) did not explain a significant amount of variance in native or exotic species cover or richness. A common assumption in restoration is that restored sites will follow a successional trajectory towards a reference system (Bradshaw 1984; MacMahon 1987). The probable explanation for this result is that restoration methodologies have improved over time, resulting in more rapid establishment of overstorey cover in more recently restored sites. Over time, species have increasingly been selected for sites and locations within sites based on improved knowledge of adaptations to soil type and depth to groundwater. In addition, irrigation and control of exotic species of agricultural concern, through a combination of herbicides and mowing, have become more systematic. Alternatively, the 1997 flood could have generated a single, overriding colonization pulse.
Secondly, the negative correlation between elevation relative to river base flow and distance to river seems counterintuitive. It is important to note, however, that all these sites have been levelled for agriculture and many have levees at the river edge, which has altered their natural topography. As a result, a number of the sites slope slightly upwards in the direction of the river. The significant effect of elevation above river base flow highlights the important effect of even small difference in depth to the water table for plant communities in these arid systems (Hupp & Osterkamp 1996; Goodwin, Hawkins & Kershner 1997; Shafroth, Stromberg & Patten 2000).
Thirdly, although soil texture has often been demonstrated to strongly influence riparian plant communities (Hosner & Minkler 1963; Johnson, Burgess & Keammerer 1976; Alpert, Griggs & Peterson 1999), we did not find significant soil texture effects. This lack of significance might be due to the fact that we only surveyed the soil texture at the surface, which may not be a good indicator of soil texture at greater depths (Alpert, Griggs & Peterson 1999). Alpert, Griggs & Peterson (1999), working on a subset of these sites, reported higher growth of tree species on soils that were deeper and finer grained in the top 1·5 m. Augering to obtain detailed soil profiles was not feasible given the large number of quadrats we sampled.
In closing, we revisit the applicability of ideas from landscape ecology and island biogeography theory to restoration planning. As in many landscape studies, both local and landscape factors appeared to influence plant communities, supporting the general idea that community dynamics operate across multiple scales. Patch size and isolation, the key factors in island biogeography theory, explained only a small amount of variance in species richness and cover. Our data also demonstrate strong heterogeneity among sites and ongoing change in plant communities, c. 10 years after planting. Thus, from a theoretical perspective, restoration at multiple sites appears to be a poor test of island biogeography theory because the sites do not meet theoretical assumptions. From the restorationists’ perspective, however, we can conclude that patch size and, to a lesser extent, isolation are relatively unimportant predictors of understorey plant species richness and cover in this system, after a time period greater than that used to evaluate the success of most restoration projects. Therefore, in order to restore native understorey plant communities in this highly fragmented landscape mosaic, managers should focus on local-scale restoration methodologies, such as efforts to increase native overstorey cover and reduce exotic plant cover, and place less emphasis on choosing sites near remnant forest.