Euphytica (2006) 148: 97–109 DOI: 10.1007/s10681-006-5944-6
C
Springer 2006
The importance of population growth, seed dispersal and habitat suitability in determining plant invasiveness David A. Bass1 , Neville D. Crossman1,2,∗ , Susan L. Lawrie1 & Mark R. Lethbridge1 1
Environmental Weeds Group, School of Geography, Population and Environmental Management, Flinders University, GPO Box 2100, Adelaide, SA, 5001, Australia; 2 Policy and Economic Research Unit, CSIRO Land and Water, PMB 2, Glen Osmond, SA, 5064, Australia (∗ author for correspondence: e-mail:
[email protected])
Key words: geographic information systems, GIS, habitat suitability, invasiveness, population growth, seed dispersal
Summary This paper examines the roles of plant demography, seed dispersal ecology and habitat suitability in influencing invasiveness of horticulturally important species. Section one investigates the relative invasiveness of two woody species, Crataegus monogyna and Prunus mahaleb, and concentrates on differences in demographic and dispersal traits. The second section delineates the invasion of two Asparagus spp. and concentrates on differences in seed dispersal ecology. Section three reports the use of a geographical information system analysis to determine whether habitat suitability, seed dispersal or land management is more important in determining threat of invasion by adventive Olea europaea. C. monogyma, P. mahaleb are closely related with similar habits and overlapping home ranges in Europe. Crataegus monogyna is very invasive in northern New South Wales, having spread rapidly and conspicuously throughout the region and elsewhere in southern Australia at rates of 80–120 m yr−1 . Prunus mahaleb is far less invasive, being restricted to a small population, which is expanding at 20 m yr−1 . Demographic analysis showed that potential growth rates of P. mahaleb (1.713–1.490) are greater than those for C. monogyna (1.138–1.103). Assessment of the seed dispersal ecology of both species revealed that C. monogyna had seeds dispersed by one bird and three mammals over many kilometers. P. mahaleb had seeds dispersed by six birds and four mammals over distances generally 1 m tall and 4680 seedlings and 232 P. mahaleb plants >1 m tall and 490,281 seedlings. Based on both the height and basal circumference models the projected rates of population growth indicate that both species have expanding populations. The growth rate of C. monogyna was 1.103, lower than 1.494 for P. mahaleb. A population growth model based on six height classes also yielded similar results. Seed dispersal ecology Both species produce fleshy fruits and have seeds dispersed by vertebrates. Crataegus monogyna is an autumn–winter fruiting species and P. mahaleb a summer fruiting species. Detailed observations of seed disperser behaviour and collection of faeces containing seeds were conducted over a 3-year period. Bass (1994) can be consulted for detailed descriptions of the methodology. The seed dispersal ecologies differed greatly (Table 2). C. monogyna had a seed dispersal system dominated by Pied Currawongs that were very abundant in the area during autumn and winter (Bass, 1995). Mammals, principally brush-tailed possums, foxes and macropods also dispersed seeds. Pied Currawongs dispersed seeds up to 10–15 km as they flew from feeding areas within rural townships to overnight
0–100 m
roosts in the surrounding area. C. monogyna seeds were not well represented in Pied Currawong diets, as birds tended to feed more on Pyracantha and Ligustrum. Prior to Pyracantha and Ligustrum introductions C. monogyna was likely to have been more important in Pied Currawong diets and hence more widely dispersed (Bass, 1995). Pied Currawongs were observed feeding on fruit and subsequently flying to overnight roosts up to 10 km distant where they would void seeds. Prunus mahaleb on the other hand had a seed dispersal system with at least six major bird vectors, foxes, rabbits and brush-tailed possums. However seed dispersal of P. mahaleb only occurred over relatively short distances (350 mm) regions of South Australia, Western Australia, Victoria, Norfolk Island and Lord Howe Island with localised populations in Victoria and Tasmania. The distribution of A. declinatus is currently limited to discrete populations in South Australia (lower Eyre, Yorke and Fleurieu Peninsulas and Kangaroo Island) and Western Australia (Perth and Bunbury). Asparagus asparagoides successfully competes with native vegetation, climbing and covering native species. Forming a dense canopy and underground root
mass, A. asparagoides inhibits germination and growth of native plants, including several endangered species, and reduces overall biodiversity. Although less widely studied, similar impacts are associated with A. declinatus. Leah (2001) found A. declinatus alters species composition of ground cover stratum, reduces overall species richness and diversity and likely inhibits native overstory vegetation recruitment and regeneration. Longer-term impacts may involve a vegetation succession from woodland to A. declinatus meadow (Leah, 2001) and associated changes to vertebrate and invertebrate fauna. Biology of Asparagus asparagoides and A. declinatus Asparagus asparagoides and A. declinatus are geophytic-terrestrial plants with perenniating buds below ground (Raymond, 1996). During summer the aboveground biomass dies off, regrowing in winter from underground storage organs. Both produce greenish white flowers during winter with fruit forming in early spring and ripening through to early summer. The fruit characteristics of A. asparagoides and A. declinatus are summarised in Table 3. Asparagus asparagoides fruit ripen to red and A. declinatus fruit are green and ripen to a pale translucent white. Asparagus declinatus produces relatively large fruit with an average diameter of 9–12 mm (based on sample of one season’s fruit crop) compared with a 5–10 mm diameter for A. asparagoides (NWS, 2000). Both fruits contain black, globose seeds ranging from 2.5–3.5 mm. A. declinatus fruits, however, contain 5–8 seeds compared to 2–3 seeds for A. asparagoides (Clifford & Conran, 1987). Fruit production for A. asparagoides varies considerably but can be as high as 1000 berries/m2 (NWS, 2000). While no quantitative data exists on fruit production rates for A. declinatus, preliminary field trials suggest fruit production reaches similar levels.
102 Seed dispersal ecology of Asparagus asparagoides and A. declinatus Native and introduced animals consume fruit and disperse the seeds of Asparagus species (Stansbury, 1999). Most vertebrate seed dispersal is restricted to less than 100 m. Rare long distance dispersals up to several kilometres have been reported for A. asparagoides (Stansbury, 2001) and for other environmental weeds (Bass, 1994). Potential seed dispersers are in part determined by fruit characteristics such as nutritional quality, fruit phenotype and accessibility (Stansbury, 1999). Red or black bicoloured fruits are more attractive to birds than other coloured fruit and smaller fruit and seed sizes allow a broader array of dispersers. Different disperser suites have unique seed dispersal shadows depending on dispersers’ behaviour, flight patterns and gut-passage times. Hence, invasion differs as a result of variation in fruit characteristics. The main dispersers of A. asparagoides are Zosterops lateralis (Silvereye) and Turdus merula, (Blackbird) (Stansbury, 2001). Larger birds have also been observed feeding on A. asparagoides. Preliminary seed dispersal observations of A. declinatus in South Australia indicate the main dispersers are medium to large gregarious birds such as Strepera versicolor (grey currawong), Gymnorhina tibicen (Australian magpie) and Anthochaera carunculata (red wattlebird). Arboreal mammals, such as Trichosurus vulpecula (brush-tailed possum) and Pseudocheirus peregrinus (common ringtail possum) and small rodents, (as evident by fruit damage) also disperse A. declinatus seeds. There is also the potential of Tiliqua rugosa (Sleepy Lizard) to disperse A. declinatus as white or pale coloured fruit is associated with frugivory by lizards (Lord & Marshall, 2001). In comparison then, A. asparagoides appears to attract a broader array of dispersers particularly in terms of the size and species of bird dispersers. Current distribution of A. asparagoides and A. declinatus is restricted to beneath tree canopies with seeds dispersed to conservation areas and roadside remnant vegetation corridors. Asparagus species do not occur in ‘open’ environments. Waite (2001) confirms this relationship between A. declinatus and surrounding landscape variables, concluding that A. declinatus was present only in areas displaying vegetation canopy. Shading experiments, however, show that A. declinatus grows well in both 70–90% shade and open sun (Craigie, 2002, personal communication). Since in the field A. declinatus only occurs under tree canopies
(Waite, 2001), seed dispersal is clearly influential in determining the distribution of A. declinatus and course of invasion. Stansbury (2001) makes similar findings in a study of A. asparagoides in Western Australia concluding that distribution is influenced by the preferred habitat and behaviour of the main dispersal agent, Z. lateralis. Avian dispersers utilise trees for perching and digesting fruit and then void seeds. This explains the relationship between distribution and tree canopy. Shaded and more mesic conditions and, in the case of A. asparagoides, structural support may also make areas under tree canopies more suitable for establishment (Stansbury, 1999). Certainly, the distribution of Asparagus species is closely linked to the habitat and post-feeding behaviour of dispersal agents. Remnant vegetation, including road reserves, appears to have the greatest risk of invasion. A. asparagoides has fruit available higher above the ground and has smaller fruit. This translates to a larger suite of seed dispersers and would contribute to a greater seed dispersal shadow. This would account for its greater invasiveness. The course of invasion by both species is determined by seed dispersal, especially birds, to areas of native vegetation with an overstory. The greater invasiveness of A. asparagoides is attributed to a combination of small fruit, bicoloured display and the greater height above the ground that the fruit is displayed. Human factors in population growth and dispersal The human factor is very important in shaping the current and future distribution of A. declinatus. In South Australia, historical plantings of A. declinatus have largely determined its present distribution. One of the earliest known plantings was on Stranraer homestead, Kangaroo Island in 1926 (Maguire, 2001 personal communication; Figure 1). The Kangaroo Island population was first considered naturalised in 1954 (Wiedenbach, 1994) and is now considered one of the worst infestations in South Australia. In Yorke Peninsula a similar scenario has occurred, where in 1913 the historic town of Inneston was settled and A. declinatus was introduced as stock feed (Pickburn & Sutcliffe, 2001). The establishment of Innes National Park in 1970 incorporated the township and surrounding region. All stock was removed and A. declinatus flourished and spread, highlighting the influence of a change in land management practices in determining invasiveness. Many other South Australian populations can be traced back to its use as an ornamental plant in early homesteads or as part of Botanic Garden plantings. Subsequent
103 invasion of roadside corridors may also be a result of road maintenance activities. The anthropogenic influence on A. declinatus distribution will continue and is likely to vary over time. A consequence of rural decline and changes in land use on Kangaroo Island (and other regions across South Australia) will promote a higher population growth rate of A. declinatus. This involves a shift from intensive grazing to smaller holdings under multiple use regimes (e.g., hobby farms) that incorporate tree revegetation. It is likely that these revegetation areas will attract birds that disperse Asparagus seeds and therefore produce new infestations. The importance of changes in land management to levels of invasiveness is the central theme of the next section.
Olea europaea Olea europaea (European olive) is a major weed in parts of southern South Australia and has the potential to become so in other parts of southern temperate Australia. These naturalised O. europaea (i.e. self-reproducing trees not planted for domestic or commercial use) reduce native plant diversity by crowding out and preventing indigenous species from regenerating. Diversity of native flora was found to be at least 50% lower in eucalypt woodland heavily invaded by O. europaea, when compared to similar woodland relatively free of O. europaea (Crossman, 2002). Olea europaea are highly fecund, produce fruit highly attractive to avian and mammalian dispersal agents, have few predators, and evolved in a climate similar to that of southern temperate Australia. These characteristics alone lend the olive to be characterised as highly invasive. However, dispersal by both anthropogenic and non-anthropogenic vectors contributes significantly to its invasiveness. Cultivated O. europaea is recorded as one of the earliest plant introductions into Australia. Reichelt and Burr (1997) suggest that O. europaea were first introduced into Sydney in 1800. Since then O. europaea have been introduced to Australia on several occasions, and are now cultivated in most states (Spennemann & Allen, 2000). Early attempts to establish olive industries in Victoria, New South Wales and South Australia were greeted with mixed success. Only in South Australia was early olive production profitable. In the 1870s olive production in South Australia was considered to be a more viable enterprise than wine production. This led to the establishment of a 1.2 ha
orchard in 1874, which was expanded to over 40.5 ha by 1882 (Reichelt & Burr, 1997). By the early 1900s, however, the olive industry in South Australia was in decline. Interest in olive products increased during the late 1940s and 1950s, coinciding with increased immigration from the Mediterranean region. By 1959, 2929 ha were under cultivation in Australia (Hartmann, 1962). The 1990s has seen a revival in the olive industry. By 1998 plantings of O. europaea had increased to more than 5000 ha, with plants equivalent to another 7000 ha of trees on order (Spennemann & Allen, 2000). Olive orchards are now found in all six Australian states. Distribution of naturalised O. europaea Olea europaea has naturalized across a wide range of habitats in South Australia. Occurrences are predominantly within the 400–600 mm median annual rainfall range (Crossman, 2002). The highest concentrations occur on the western foothills of the southern and central Mount Lofty Ranges (Figure 1). These populations are likely to be the direct descendants of the orchards planted in the late-1800s/early-1900s. The naturalisation of O. europaea in the southern and central Mount Lofty Ranges is not a recent event. As early as the 1920s, Adamson and Osborne (1924) noticed the abundance of O. europaea on the dry and rockier slopes of the western foothills. Specht and Perry (1948) and Specht and Cleland (1961) made similar observations. Other large naturalized populations are found in the Barossa and Clare valleys, 60 and 120 km north of Adelaide, respectively. Populations of naturalised O. europaea are also found in Western Australia (Hussey et al., 1997), particularly in bushland near Perth (Sandy Lloyd, 2000 pers. comm.); Victoria (Carr et al., 1992); New South Wales (Dellow et al., 1987; Parsons & Cuthbertson, 2001); and south-east Queensland (Csurhes & Edwards, 1998). O. europaea dispersal Dispersal of O. europaea seed is almost wholly dependent on birds. Jupp et al., (1999) found that 17 species of birds included the fruit of O. europaea in their diet, although only seven dispersed the seed. The Common Starling (Sturnus vulgaris L.) is the most prevalent disperser: a common sized flock of 100 could disperse thousands of seeds in a day (Mladovan, 1998). They usually swallowed O. europaea fruits whole, regurgitating the stones 20–50 min later. Most seeds were likely to be dispersed within 100 m of the parent trees,
104 as starlings moved to nearby perch sites to consume the fruit. The exception was at dusk when starlings returned to nocturnal roosts, so that some seeds could be dispersed kilometres away from parent trees. Starlings had difficulty in handling and swallowing large fruits, and showed a strong preference for intermediate sized fruits. There is evidence to suggest that Emus (Dromaius novaehollandiae) disperse seeds in this manner (Jupp et al., 1999). The European Fox (Vulpus vulpus) consumes O. europaea fruit and disperses seed in the southern Mount Lofty Ranges with seeds frequently found in fox scats during the winter months (Paton et al., 1988). Spennemann and Allen (2000) suggest that foxes may disperse O. europaea seed 40–50 km. The nursery trade and expanding olive industry are responsible for long distance dispersal >100 km. Large olive orchards have recently been planted in southeast Queensland, Larenta (100 km north of Perth) and Coonalpyn and Pinaroo (180 km southeast and 250 km east of Adelaide respectively) (Crabb, 1999; Haran, 1999; McIlwraith, 1999). These areas were, until recently, relatively free of O. europaea. Predicting the threat of O. europaea invasion Discussion so far has mainly centred on the relationship between dispersal and invasiveness, with dispersal rates determined by the presence of suitable dispersal vectors and the population growth rate of the target species. These two variables, i.e. population growth and dispersal have been used extensively in predictive spatiotemporal models of invasive plant spread (cf. Higgins & Richardson, 1996 for a detailed review). Another key component of plant invasiveness is the state of the receiving environment. The habitat into which a dispersed seed arrives has a strong bearing on the growth rates of the potential population. Habitat suitability is, in turn, made up of two components: (1) climatic and edaphic variables, and (2) land management variables. Many predictions of plant invasiveness have included at least some measurement of component 1, based on the axiom that plant distribution is primarily constrained by climate (Woodward, 1987). However, component two is rarely included in estimates of invasiveness. This shortcoming is unlikely to be due to an oversight, as it has been recognised for several decades that land management practices are highly influential in the invasion process (Auld & Coote, 1980), particularly the positive relationship between the levels of anthropogenic disturbance and invasiveness (Hobbs, 1991; Hobbs & Huenneke, 1992). The omission of land man-
agement variables in estimates of plant invasiveness is more likely a result of the resolution at which most predictive models are constructed. Many predictions are made on a 0.5◦ grid (Kriticos & Randall, 2001), making it impractical to include land management variables, which exhibit considerable variation at this scale. We present the outputs of a high resolution, broad scale invasive threat model for naturalised O. europaea in the central Mount Lofty Ranges, South Australia (Figure 1). The model considers land management in addition to dispersal, climatic and edaphic variables. The aim is to confirm that a change in land management plays a significant role in O. europaea invasion. Overview of model construction We present only a brief outline of model construction (for detailed methods, see Crossman et al., 2002). The model consists of three spatial surfaces: (1) probability of O. europaea establishment based on habitat suitability (p1 ); (2) probability of dispersal to the nearest remnant patch of native vegetation based on a simple dispersal curve (p2 ); and, (3) probability of O. europaea establishment based on land management (p3 ). All modeling was undertaken in a geographic information systems (GIS) environment using the cell-based GRID module of ESRI’s Arc/INFO software (ESRI, 2001). Habitat suitability was calculated by regressing a series of climatic and edaphic variables using a regression tree (Brieman et al., 1984). The dependent variable is the distribution and density of naturalised O. europaea canopy across a western portion of the central Mount Lofty Ranges, adjacent to, but separate from the study area. This data was collected as part of an unrelated exercise (cf. Crossman & Bass, 2002). Those cells exhibiting habitat highly suitable to O. europaea establishment and growth were assigned a high p-value, those with low habitat suitability, a low p-value. Probability of dispersal was determined by applying the negative exponential dispersal kernel to the distance to native vegetation. The negative exponential is one of the simplest and widely used descriptors of plant dispersal (Willson, 1993). Cells close to native vegetation received a high p-value, those more distant, a low p-value. Probabilities for the third surface were assigned based on the intensity of land use at each cell. High intensity land uses, such as cropping and grazing, were assigned a low p-value, as the likelihood of O. europaea establishment is low. Conversely, land not used for productive purposes, e.g., abandoned or protected for conservation, was assigned
105 a high p-value. The final threat surface was calculated by multiplying surfaces one to three.
Equation (2) now becomes 2 2 2 + p12 p32 σ p2 + p12 p22 σ p3 σ P2 = p22 p32 σ p1
(3)
The role of land management A series of steps were followed to assess the outputs of the model given a change in land management. The first step involved the creation of three scenarios, based on hypothetical alterations to land use conditions (Table 4). The next step involved running a sensitivity analysis over the input surfaces within each scenario. The sensitivity analysis, in greater detail, is as follows. Assuming the cell values in the three surfaces (p1 , p2 and p3 ) depict independent variables, a crude sensitivity analysis can be conducted using propagation of error theory. These surfaces are probabilistic and are hence scaled on a 0–1 scale. While error propagation theory does assume the data are normally distributed, this is not a serious violation and normalising the data potentially dilutes secondary interactive effects. Using the following multiplicative model where the total threat is given by P = p1 × p2 × p3
(1)
And based upon error propagation theory (Rainsford, 1957). σ P2 =
∂P ∂ p1
2
2 σ p1 +
∂P ∂ p3
2
2 σ p2 +
∂P ∂ p3
2 2 σ p3
(2) Here σ P2 is the variance in P and ∂∂pfi are the partial differentials of the ith surface with respect to Equation (1).
It can therefore be seen that the contribution of error or variation in P contributed by p1 is p22 p32 , p2 is p12 p32 and p3 is p12 p22 . For each scenario, a principal components analysis (PCA) was performed on the products of each pair of squared surfaces: p12 × p22
(4)
×
(5)
p12 p22
×
p32 p32
(6)
The principal components were then correlated back to the three surfaces ( p12 × p22 , p12 × p32 and p22 × p32 ) to determine which surface explained the most variance. The cross-correlations (Table 5) were examined and comparisons were made between each scenario. The strongest correlations between each surface and the three principal components differ from scenario 1 to scenarios 2 and 3 (Table 5). In scenario 1, a hypothetical landscape consisting of extensive and intensive levels of disturbance, habitat suitability and dispersal drive the model. This is evidenced by the strong correlation between PC1, which explains the most variation in the data, and the product of the habitat suitability and dispersal surfaces. However, a change in management to a landscape consisting of more land not frequently disturbed, as described in scenarios 2 and 3, alters the driving forces of the model. In both of these scenarios the probability of O. europaea establishment based on land management has a greater influence on the overall model. There exist strong correlations to
Table 4. Description of land use scenarios Scenario
Description
p-value
% of study area
1
Probabilities heavily skewed to unfavourable land use for Olea europaea establishment (i.e. most land under agricultural production)
Low Medium High
68.9 31.1 0.0
2
Relatively equal distribution of probabilities (i.e. the real-world situation)
Low Medium High
29.4 39.5 31.1
3
Probabilities heavily skewed to favourable land use for O. europaea establishment (i.e. many abandoned farms and/or protected lands)
Low Medium High
0.0 29.4 70.6
Note. Probabilities for dispersal and habitat suitability remain constant.
106 Table 5. Correlation matrix of principal components against the products of each pair of squared surfaces Scenario
Surfacea
PC1
PC2
PC3
1
p12 × p22 p12 × p32 p22 × p32
0.994 0.508
−0.108 0.654
0.008 −0.560
0.429
0.889
0.162
2
p12 p12 p22
×
p22 p32 p32
0.587
0.794
0.161
0.815
0.105
−0.569
0.951
−0.276
0.142
p12 p12 p22
×
p22 p32 p32
0.721
0.642
0.262
0.753
0.270
−0.599
0.916
−0.392
0.083
3
× × × ×
Note. Correlations greater than ±0.7 are displayed in bold. a p : probability of Olea europaea establishment based on habitat suitability; p : probability of dispersal to the nearest remnant 1 2 patch of native vegetation; p3 : probability of O. europaea establishment based on land management.
PC1 due to the greater variability in land management across the landscape. These findings confirm that land management plays an important role in naturalised O. europaea invasions, and that a change in this variable can alter overall invasiveness.
Discussion More invasive species should have a faster rate of spread, a higher population growth rate, and a better developed (or more complex) seed dispersal system. We have also demonstrated that land management, as a factor of habitat suitability, is also a strong determinant of invasiveness. In this study the more invasive C. monogyna had rates of spread four to five times greater than P. mahaleb. However the projected rate of population growth of C. monogyna is less that P. mahaleb. Also, contrary to expectations, C. monogyna has a more simple seed dispersal system dominated by Pied Currawongs. Prunus mahaleb has, by comparison, a relatively more complex seed dispersal system which is on a par with the seed dispersal systems developed in its home range in southern Europe (Herrera & Jordano, 1981; Guitian et al., 1992; Jordano, 1994) with similar numbers of dispersal agents and dispersal distances. The differences in dispersal are a function of season. In autumn and winter the dominant Pied Currawong disperses seeds of at least twenty introduced plants as it concentrates on fruit during times of lower invertebrate abundance. At this time flocks of pied currawongs congregate in large (>500 birds) feeding flocks near urban
centres to feed on fruit of ornamental species (Bass, 1995). In summer birds have dispersed into breeding territories and switched to invertebrates for food. The disparity between expected attributes and relative invasiveness is counted by the influence of humans and the effectiveness of seed dispersal. Crataegus monogyna has been widely planted thereby providing a large number of invasion foci, which under a similar seed dispersal system will occupy new territories much faster than a single invasion point (Mack, 1985). Prunus mahaleb on the other hand has a single point of introduction. Despite having many seed dispersal vectors, seed dispersal only occurs over short distances of
