a. platanoidespopulation growth and spread
In 1965, approximately 25 years after the first successful establishment of A. platanoides in the forest reserve on Mackinac Island and approximately 40 years since the first known introduction (Webster, Nelson & Wangen 2005), only scattered clumps of A. platanoides were present along the periphery of developed areas on the southern end of the island. At that time, it would have been difficult to find an A. platanoides in the forest, much less perceive the rapid population growth and range expansion that would ensue over the course of the next three decades. This prolonged lag phase highlights the dilemma posed by long-lived exotic woody invaders (Frappier et al. 2003). In the case of A. platanoides, one of the ‘most popular’ and widely planted urban street trees in eastern North America is belatedly coming into focus as a serious invader of native forests (Nowak & Rowntree 1990; Martin 1999; Webb et al. 2000). Similarly, well-documented lag phases between wide-scale planting and invasive spread have been reported for commercially important exotic Pinus spp. in the southern hemisphere (Richardson, Williams & Hobbs 1994; Rouget et al. 2004).
The population of A. platanoides on the island entered the expansion phase around 1955, as new individuals and colonies expanded the area occupied. The linear increase in area occupied and range radius of the population during the expansion phase are suggestive of type 1 stratified diffusion (Shigesada, Kawasaki & Takeda 1995). This pattern of range expansion is typically generated when offspring primarily establish within the neighbourhood of the parent population via short-distance dispersal events (Shigesada, Kawasaki & Takeda 1995). During the first phase of expansion on the island, most new colonies and recruits were proximate to developed areas that hosted source populations. However, by 1975 numerous satellite populations resulting from long-distance dispersal events were evident. Given the time required for these satellite populations to reach reproductive age (25–30 years for open grown trees; Gordon & Rowe 1982), they still contributed relatively little to range expansion and population growth. The apparent decline in the rate of population growth and spread of the contemporary population, even though large areas of forest remain uninvaded, suggests that most of the easily invaded sites near parent populations are moderately well saturated. This slow down could represent an additional lag phase in population growth and range expansion. Nevertheless, as satellite populations come of age, their expansion may become the primary driver of population growth and spread (Moody & Mack 1988; Neubert & Caswell 2000). This demographic shift could greatly accelerate the invasion and might best be described by a range expansion pattern similar to type 2 stratified diffusion. Type 2 stratified diffusion is characterized by a biphasic expansion phase with initially slow growth followed by a higher constant rate, which could reflect an increase in the importance of long-distance dispersals (Shigesada, Kawasaki & Takeda 1995). Nevertheless, multiple lag phases and/or shifts in expansion type are not features of contemporary stratified diffusion theory, but may represent logical theoretical extensions. At a metapopulation scale, the expansion of an invasive tree, such as A. platanoides, may again slow as older satellite populations converge with the advancing front and new non-reproductively active satellite populations are established by chance long-distance dispersal events. In other words, exotic species with long generation times may expand their ranges in relatively discrete pulses with multiple lag phases. Eventually, once a critical density of parent populations are established, range expansion would become continuous until all suitable sites have been occupied.
Given that A. platanoides can aggressively colonize (Wyckoff & Webb 1996; Martin 1999; Webb et al. 2000) and persist in shaded understoreys (Kloeppel & Abrams 1995; Sanford, Harrington & Fownes 2003), it can build up a large pool of advance regeneration in relatively undisturbed native forests. These understorey saplings are typically non-reproductive and may persist for upwards of 30 years (Bertin et al. 2005; Webster, Nelson & Wangen 2005). Consequently, the reproductively active portion of the population on Mackinac Island, during any given time period, was probably less than would be predicted based on tree age alone. However, saplings growing in canopy gaps or under high shade may be able to recruit into the overstorey nearly a decade before native shade-tolerant species (Webster, Nelson & Wangen 2005). Highly variable understorey residence times prior to canopy recruitment and subsequent seed production make it difficult to accurately model the invasion dynamics of shade-tolerant woody perennials in forested systems. Following successful introduction and establishment, these species may appear harmless for decades before being released by natural or anthropogenic disturbance (e.g. overstorey tree-fall gaps or logging). Nevertheless, canopy turnover rates in most forests would still result in the gradual conversion of native forests to exotic species through natural gap dynamics. Large-scale anthropogenic disturbances or catastrophic natural disturbances that remove most of the native canopy during a single event, however, could result in a rapid shift towards exotic dominance in forests with well-developed seedling layers of A. platanoides or other perennial woody invaders tolerant of shade (Hutchinson & Vankat 1997; Webb et al. 2000; Collier, Vankat & Hughes 2002; Yates, Levia & Williams 2004). Other exotic tree species that are intolerant of shade, such as Ailanthus altissima (Miller) Swingle, may be able to gain a foothold in large tree fall gaps and then invade rapidly following more intense disturbance (Knapp & Canham 2000; Call & Nilsen 2003). Consequently, it may be difficult to demonstrate the risk posed by exotic, woody perennials, especially if they are commercially important species, given that widespread range expansion and deleterious impacts may lag decades behind widespread introduction.
The potential for multiple lag phases during the expansion phase of population spread may require somewhat more complex invasion models, which incorporate population age structure, the temporal and spatial availability of habitat, and changes in dispersal mode as the population expands (Higgins, Richardson & Cowling 2001). For example, the theory of stratified diffusion could be expanded to include a punctuated or stepped expansion phase that is sensitive to both positive and negative feedbacks. In the case of shade-tolerant exotic trees, lags during the expansion phase may be strongly influenced by the long time needed to reach reproductive maturity, which is also influenced by the frequency of disturbances that facilitate canopy recruitment. For example, the availability of canopy gaps proximate to parent populations or probable long-distance dispersal sites may be an important determinant of lag behaviour because it influences the time needed for individuals to reach reproductive maturity (as is the case with suppressed understorey saplings awaiting the formation of a canopy gap). Similarly, spread may slow as the invasion radiates away from dispersal corridors and more readily invaded habitats into a less easily invaded matrix. All of these phenomena are scale dependent and may become obscured at increasingly large temporal and spatial scales. Additionally, at the metapopulation scale non-synchronous introductions could conceivably ‘average out’ lags displayed by individual populations. Nevertheless, lag behaviour during the expansion phase may have important implications for control and risk assessment efforts at local and regional scales.
vectors and barriers to establishment
While several studies have demonstrated that A. platanoides does not require disturbed edge environments for establishment or survival (Webb et al. 2000; Sanford, Harrington & Fownes 2003), road corridors appear to facilitate the dispersal of this species at least in forested landscapes (Anderson 1999). Most studies of invasive spread along roads have concluded that roads act as disturbances that facilitate the establishment of exotic species (Greenberg, Crownover & Gordon 1997; Parendes & Jones 2000; Silveri, Dunwiddie & Michaels 2001; Call & Nilsen 2003; Rentch et al. 2005), but the role of roads as corridors for the flow of propagules has been less clear (cf. Harrison, Hohn & Ratay 2002; Buckley et al. 2003). Our results suggest that roads may be associated with long-distance dispersal events that move A. platanoides well beyond the interface between developed areas and native forests. During each time period examined, the furthest dispersals from developed areas were along road and trail corridors. This association is probably the result of both standard and non-standard dispersal mechanisms. In developed areas, A. platanoides is commonly planted as a street tree and rains a tremendous amount of seed onto the road surface and passing vehicles. Consequently, in addition to samaras floating down road corridors that act as wind tunnels or simply blowing along the road surface, A. platanoides may hitchhike on passing vehicles. Transportation on the island is restricted to foot traffic, bikes, horses and horse-drawn wagons and carriages, all of which could effectively transport seeds varying distances. Horses may be especially influential in the transportation of seed as mixtures of seed and manure may also adhere to hooves and cart wheels. However, the consolidation of manure and a discontinued practice of transporting manure into the forest for composting (J. Dykehouse, Mackinac Island State Park, personal communication) are probably responsible for at least the early association between roads and A. platanoides observed in our reconstruction. Contrary to expectations, roads still appear to be one of the primary vectors for the dispersal of A. platanoides into the native forests on Mackinac Island. This result highlights the importance of human-mediated jump dispersal in the spread of invading organisms (Suarez, Holway & Case 2001).
Another possible interaction between horses and A. platanoides is that A. platanoides may benefit from a fertilization effect associated with the deposition of urine and manure along roads and trails. Nutrient additions in conjunction with disturbance may interact synergistically to promote invasions by exotic species (Kercher & Zedler 2004; Perry, Galatowitsch & Rosen 2004; Rickey & Anderson 2004). However, given the high nutrient use efficiency of A. platanoides compared with shade-tolerant native species (A. saccharum; Kloeppel & Abrams 1995) it is unclear how these factors may influence successful establishment of this species. Additionally, a strong affinity between A. platanoides and trail edges has been documented in areas without horse traffic (Anderson 1999). Nevertheless, high nutrient and resource levels may allow A. platanoides to compete against more aggressive native species or persist in otherwise hostile environments, and warrants further study.