In the face of increasing human impacts, a major goal for conservation biology is to provide principles by which biological diversity can be preserved. The revisitation methodology employed in this study enabled us to observe three major outcomes of the effects of fragmentation over the last three decades. First, despite >50 years since agricultural intensification of the district in the 1950s, which likely precipitated the first wave of local population extinction, native species populations continue to be lost from remnant woodlands. Second, extrinsic factors (remnant area and remnant shape) provided low predictive power for understanding these ongoing effects, whereas some intrinsic factors (e.g. plant height, seed dispersal) had greater explanatory value. Hence, fragmentation effects have both a spatial and biological component. Finally, the population abundance of species in 1975 was a good predictor of local extinction by 2006, indicating that small populations are more likely to disappear over time. Hence, species at low abundance are most prone to local extirpation due to demographic and environmental stochasticity.
Site area and shape did not substantially explain the probability of local population extirpation of native plant species in this fragmented landscape, contrary to theoretical expectations (Fahrig 2003; Ewers & Didham 2006; Lindenmayer & Fischer 2006). Typically, species in the smallest remnants with the highest edge to area ratio are considered to be at most risk of extinction. However, Bayesian logistic regression suggests that these factors have rather weak explanatory value (and wide variability) in grassy woodlands. Perhaps this is because all the sites are small (<40 ha in area) and mostly linear, hence obscuring the true effects of these factors. Additionally, site isolation, rather than size may be a more important determinant of species persistence (Thomas et al. 2001). However, isolation distances between remnants was not quantified in this study as all sites occur in a matrix of agriculture with scattered paddock trees, making estimates of true isolation difficult.
Alternatively, the lack of distinct effects of remnant size and shape on persistence of native species may be because each of the sites are under different pressures in addition to fragmentation, that act in a non-uniform way to affect habitat quality (Thomas et al. 2001; Ewers & Didham 2006; Brook et al. 2008). These pressures include inter-specific competition, decreased endogenous disturbances, increased exogenous disturbances, edge effects and invasions of both plant competitors and animal herbivores (McIntyre et al. 1995; Davies & Margules 1998; Hobbs & Yates 2003; Lunt & Spooner 2005; Williams et al. 2005). Habitat deterioration can affect the population viability and persistence of plants (Brys et al. 2005) and, in most cases, degradation of habitat quality occurs due to the impact of the surrounding landscape matrix on habitat patches.
Habitat deterioration is likely to have played a role in the loss of local populations in this study. Although detailed site management histories were not available for any of the sites, information given by local residents is illuminating. For example, the site with the lowest proportional loss of species (Lawson's Lane; 13%) was reported to have been annually burnt until approximately 8 years before the 2006 surveys were conducted (unnamed local farmer, pers. comm., Nov 2006). In contrast, the site with the largest proportion of species lost (Balmoral; 44%) used to be burnt regularly (approximately annually) when the railway line was in use (Agnes Coxon, pers. comm., Jan 2007). However, as the railway line became unused three to four decades ago, burning also ceased at that time (Agnes Coxon, pers. comm., Jan 2007). This has resulted in many species being lost from the site, as has been found by Williams et al. (2006) for formerly burnt but now long-unburned grasslands. Many ‘new’ species to this site since 1975, including native shrubs (e.g. Astroloma conostephioides, Grevillea aquifolium) are now encroaching, which also suggests a more infrequent burning regime (Roques et al. 2001; Heisler et al. 2003). These changes are unlikely to represent natural succession because most new shrubs observed are not considered a component of the native vegetation (using the descriptions of Blackburn & Gibbons (1956), Gibbons & Downes (1964) and Willis (1971) as a guide to the pre-fragmented woodland state), but rather have encroached from nearby windbreaks and apparently via seed dispersal by roadside maintenance machinery. Their establishment in the absence of fire, due to land use change, might therefore be considered opportunistic. Hence, within-site habitat management could play an under-appreciated role in maintaining native species in this fragmented landscape, regardless of the size or spatial context of remnants. This is because of its direct impacts on habitat quality and, presumably, habitat heterogeneity. Despite the apparent importance of maintaining habitat quality for species persistence in fragmented landscapes, it remains poorly recognized relative to the fragmentation paradigm that focuses on the spatial patterns of habitat vs. non-habitat (Harrison & Bruna 1999; Thomas et al. 2001; Radford & Bennett 2007).
No life form appeared to predispose plant species to local population extinction due to habitat fragmentation in this grassy woodland landscape. Chamaephytes, geophytes and hemicryptophytes, whether they were clonal or not, all appeared equally capable of persistence in the small remnants. However, shorter species, in general, had a much higher probability of local extirpation than taller species. Taller species generally have a longer life span than shorter species (Duncan & Young 2000) and hence do not need to successfully reproduce as frequently. In addition, taller species are generally more competitive (Duncan & Young 2000; Kolb & Diekmann 2005). It is likely, however, that the lower probability of local extinction for taller species (i.e. trees) may simply reflect a time-lag in response to fragmentation (i.e. the extinction debt sensu Tilman et al. 1994), where long-lived adult individuals persist but conditions are no longer suitable for their recruitment.
Species with wind- and mobile-dispersed seeds had a higher probability of local extirpation, a finding at odds with much fragmentation literature. The ability to persist in a fragmented landscape has been considered to, in part, depend on the dispersal abilities of species (e.g. Hoehn et al. 2007). Those species that disperse readily and frequently will more likely persist (as populations rather than individuals) in habitat remnants though time than species that do not. For plants, however, dispersal by wind may ‘waste’ many propagules as they are more likely to be lost in the relatively unfavourable matrix of agricultural lands (Holt et al. 1995; Roy & de Blois 2006). Alternatively, the very small seeds of wind-dispersed species may render them poor competitors in the recruitment phase of their life cycle (Ehrlen & van Groenendael 1998; Soons & Heil 2002), making them susceptible to changes in the landscape caused by non-native plant invasions and changed disturbance regimes. Hence, it would appear that population persistence may not necessarily be enhanced by seed dispersal between remnants but rather, by other plant traits such as an ability to maintain site occupancy by longevity or a persistent soil seed bank, or site factors themselves. From a conservation perspective, differences in the colonization capacity among species imply that maintenance of plant diversity must not only focus on conditions within patches, but also consider the spatial arrangement of patches in order to enable plants to bridge gaps in time and space (Ehrlen & van Groenendael 1998).
By contrast, the probabilities of species with animal-dispersed seeds to go locally extinct were lower. This may be because sheep, cattle and other introduced animals (e.g. rabbits, cats, foxes) act as vectors for dispersal of some species (Hobbs & Yates 2003) or because some fauna may be able to cross the rural matrix, increasing the biotic dispersal between remnants.
Despite its theoretical significance, very few studies compare the past abundance of plant populations to their present abundance and how this may affect the probability of local extinction (but see Robinson & Quinn 1988; Fischer & Stocklin 1997; Davies & Margules 1998; Davies et al. 2000; Matthies et al. 2004; van Calster et al. 2008). Those that have used this approach tend to focus on rare species, neglecting the more abundant populations and species. Yet, it is implicit in MacArthur & Wilson's (1967) theory of island biogeography that extinction rates due to fluctuations in populations should decline with increasing population size, regardless of their ecological status. In this study, the past abundance (1975) of a species was a good predictor of local extinction risk for small populations when populations were revisited 30 years later. Small populations tended to have an increased risk of local extirpation in comparison to populations classed as ‘Very Common’ in 1975. Disappearance events in grassland annuals have previously been observed to be determined by population size in the previous year; small-sized populations were more prone to be missing (Dostal 2005), as have small populations elsewhere (Fischer & Stöcklin 1997; Duncan & Young 2000; Matthies et al. 2004; Walker & Preston 2006; Wiegmann & Waller 2006; van Calster et al. 2008). This is likely to be a result of their increased vulnerability to localized catastrophic events and genetic erosion caused by a reduced gene pool and inbreeding (Fischer & Stöcklin 1997; Matthies et al. 2004). In addition, their ability to recruit may be inhibited due to strong competition for space or perhaps seed limitation due to lack of seed production (Hobbs & Yates 2003). Interestingly, some populations classed as ‘Very Rare’ in 1975 became ‘Common’ by 2006, hinting that populations can be very dynamic in remnant woodlands and this may be influenced by temporal rescue effects (Piessens et al. 2004), climatic variability (Levine et al. 2008) or relaxation of previous fire regimes (Ross et al. 2002; Briggs et al. 2005).
Populations with a large past abundance varied in their response to fragmentation. Most initially large populations remained abundant, hence supporting the few studies that show common species have a decreased risk of extinction (e.g. Fischer & Stöcklin 1997; Duncan & Young 2000; Matthies et al. 2004; Walker & Preston 2006; Wiegmann & Waller 2006; van Calster et al. 2008). Importantly, not all large populations in 1975 were immune from local extinction by 2006; this has rarely been reported in the literature over such short timescales (Hooftman & Diemer 2002). Hence, this suggests that species can be at risk in fragmented landscapes for reasons other than having small populations.
Three potential explanations might be invoked for the observed declines in formerly common populations. First, woodlands in southern Australia are exposed to annual summer drought, the severity of which is linked to El Niño Southern Oscillation events. Severe drought can cause high species-specific mortality in grasslands (Williams 1968; Austin et al. 1981) and woodlands (Pook et al. 1966; Ashton & Spalding 2001) and hence, infrequent but catastrophic declines of common or dominant species due to environmental stochasticity are probably ‘normal’ in such ecosystems at decadal to centennial timescales (Nicholls 1991; Lande 1993; Brook et al. 2008).
Second, it is very likely that woodland habitat quality changed in different or non-uniform ways across the 12 sites over the 31 years. It is possible that what was initially good habitat that supported large populations of plants has become poorer habitat in some, but not all cases. Although there is empirical evidence that habitat modification and degradation negatively affect the biota of fragmented landscapes (Harrison & Bruna 1999), there is no theoretical framework for evaluating the effects of habitat modification and how this interacts with population dynamics (Davies et al. 2000). Clearly, it is important to determine what drives habitat quality. This will likely be edge effects, disturbance and invasions by non-native species.
Finally, another potentially important explanation for rapid and profound population loss and decline is considered to involve recent changes in disturbance regimes. Competition from dominant native tussock grasses and exotic species is known to underpin patterns of diversity in grassy ecosystems in temperate Australia (e.g. Stuwe & Parsons 1977) and, where there is a lack of disturbance by grazing or fire (Stuwe & Parsons 1977; Williams et al. 2006), these processes are likely to be intense. In the absence of fire, habitat heterogeneity can also decline, leading to reductions in species richness of plant communities (Collins 1992; Ross et al. 2002). Similarly, Leach and Givnish (1996) postulated that the changes in fire regimes that accompany habitat subdivision of tallgrass prairie were responsible for substantial losses of plant species. This suggests that fragmentation might often promote plant population loss in flammable ecosystems that are no longer exposed to fire, sentiments echoed more recently by Briggs et al. (2005).