Introduced woody plants pose significant challenges to the conservation of both forest and grassland ecosystems world-wide (Myers 1983; Braithwaite, Lonsdale & Estbergs 1989; Vitousek & Walker 1989; Wyckoff & Webb 1996; Hutchinson & Vankat 1997; Grice, Radford & Abbot 2000; Lepßet al. 2002; Barton et al. 2004). Woody invaders often form persistent patches that significantly alter the structure and composition of native plant communities, and may ultimately disrupt ecosystem function. For example, the conversion of sedgeland communities in Australia to shrublands by the invasive shrub Mimosa pigra L. threatens several resident and migratory bird species that rely on these sedge-dominated communities (Braithwaite, Lonsdale & Estbergs 1989). Similarly, in eastern North America shade-tolerant woody invaders that form monotypic stands (e.g. Acer spp., Berberis thunbergii DC, Lonicera spp. and Rhamnus spp.) have been shown to reduce tree species diversity by inhibiting the recruitment of native species (Woods 1993; Hutchinson & Vankat 1997; Martin 1999; Silander & Klepeis 1999; Frappier, Eckert & Lee 2003). Nevertheless, most woody invaders in North America and elsewhere were intentional introductions and many are still widely planted as ornamental or commercial species (Nowak & Rowntree 1990; Silander & Klepeis 1999; Rouget et al. 2002).
In order to better understand which factors govern the invasion process following the successful establishment of an introduced species, numerous mathematical models have been developed to characterize the spread of invasive organisms (Higgins & Richardson 1996; With 2002). These models range from reaction–diffusion (Skellam 1951) and stratified–diffusion (Shigesada, Kawasaki & Takeda 1995) models to complex integrodifference equation (Neubert & Caswell 2000) and neutral landscape (Gardner et al. 1987; With 2002) models. The invasion process can also be described in terms of a simple logistic growth curve, representing the three phases of the invasion process: introduction, colonization and naturalization (Radosevich, Stubbs & Ghersa 2003). These phases roughly parallel the phases of stratified diffusion: establishment, expansion and saturation (Shigesada, Kawasaki & Takeda 1995). The expansion phase in stratified diffusion is theorized to reflect one of three distinct forms: (i) a constant linear expansion; (ii) a biphasic expansion with an initial slow linear expansion followed by a more rapid linear expansion; or (iii) an exponential increase (types 1–3 stratified diffusion, respectively; Shigesada, Kawasaki & Takeda 1995). The saturation phase is reached when the population reaches a quasi-threshold density where population growth stabilizes and the geographical extent of the invasion remains approximately constant. This threshold is reached when niche occupancy and available resources limit the rate of spread (Radosevich, Stubbs & Ghersa 2003) or the population has reached a geographical limit to expansion (Shigesada, Kawasaki & Takeda 1995).
The expansion phase, which is typically of great interest to practitioners and policy makers, is modulated by the life-history characteristics of the invader as well as native plant community and landscape structure. For example, forest fragmentation may facilitate the spread of ruderal exotic invaders, but the isolation of forest remnants may slow the invasion of shade-tolerant perennial invaders (Janzen 1983; Brothers & Spingarn 1992). Likewise, roads and water courses may provide corridors for invasion (Parendes & Jones 2000). Buckley et al. (2003) found that logging roads serve as the primary conduits of dispersal for invasive plants into the interior of managed forest stands. The distance that these vectors allow propagules to spread can be more important than the frequency with which they carry propagules (Neubert & Caswell 2000). Consequently, a species’ capacity for short- vs. long-distance dispersal is an important determinant of the shape of the invasion curve during the expansion phase. For example, species that establish primarily within the range of the parent population expand the occupied area from the periphery at a relatively constant rate resulting in a type 1 shape (Shigesada, Kawasaki & Takeda 1995). Types 2 and 3 shapes result from mixtures of long- and short-distance dispersal and are strongly influenced by where long- and short-distance dispersers originate along the invasion front and how they are dispersed (Shigesada, Kawasaki & Takeda 1995). Consequently, by examining the spread of invasive organisms across heterogeneous landscapes, we may be able to elucidate the comparative importance of long- and short-distance dispersal events, and identify landscape features that act as either vectors, which accelerate and/or facilitate invasion, or barriers to invasion (With 2002). This knowledge could be used to design more effective control strategies that can be applied during the expansion phase, which is all too often when the negative impacts of invaders are first recognized.
Because of their longevity and size, the movement, growth and reproduction of introduced woody plants can be relatively slow, and impacts may take a relatively long time to materialize compared with herbaceous and animal invaders with short generation times (Frappier et al. 2003). For example, Norway maple Acer platanoides L. was introduced into eastern North America from Europe during the mid-1700s as an ornamental shade tree and then widely planted during the latter half of the 20th century to replace Ulmus americana L. lost to Dutch elm disease. In spite of its long history in North America, A. platanoides has only recently been recognized as a serious threat to native forest plant communities (Martin 1999; Webb et al. 2000). In North America, this species is moderately long lived (100–150 years; Nowak & Rowntree 1990), and usually does not produce viable seed until 25–30 years of age (Gordon & Rowe 1982). Acer platanoides is shade-tolerant (Kloeppel & Abrams 1995) and can be an aggressive colonizer of forest understoreys (Webb & Kaunzinger 1993) that can persist as advance regeneration for extended periods of time (Webster, Nelson & Wangen 2005). Long time lags between introduction and range expansion may mask the invasive potential of many common ornamentals (Frappier et al. 2003). Conversely, changes in extrinsic factors, such as land use and disturbance regime, may facilitate rapid range expansion by previously non-invasive species.
In order to identify landscape features that influence the invasion of A. platanoides and to clarify the potential for long or multiple time lags during invasions of introduced trees, we reconstructed the spread of A. platanoides on a temperate forested island. The island setting allowed us to reconstruct the invasion process across a relatively large finite landscape while maintaining a high degree of resolution (the location of every A. platanoides≥ 0·5 m tall was mapped). Detailed tree-ring analysis enabled us to backdate the invasion to the first successful establishment within the island's forest and follow the process forward to the present. Additionally, the presence of a minimum of 3 km of lake in any direction helped to isolate the island from outside influences, such as additional unaccounted for seed sources. There are, however, limitations to this approach because the island setting provides a limited range of landscape features (the island is mostly forested). Nevertheless, this setting provided an unprecedented opportunity to improve our understanding of the invasion process as it relates to shade-tolerant exotic trees.