Post-hurricane vegetation recovery in an urban forest

Available online at www.sciencedirect.com

Landscape and Urban Planning 85 (2008) 111–122

Post-hurricane vegetation recovery in an urban forest Scott Burley, Sarah L. Robinson, Jeremy T. Lundholm ∗ Biology Department/Environmental Studies Program, Saint Mary’s University, 923 Robie Street, Halifax, Nova Scotia B3H 3C3, Canada Received 26 January 2007; received in revised form 29 August 2007; accepted 16 October 2007 Available online 3 December 2007

Abstract Urban forests are increasingly vulnerable to catastrophic disturbance, and their isolation and human use may challenge the ability of vegetation to recover spontaneously. We examined vegetation responses to recent hurricane disturbance in a temperate mixedwood urban forest: Point Pleasant Park in Halifax, Nova Scotia, which suffered over 70% canopy loss during Hurricane Juan in fall 2003. In 2005 we surveyed 30 paired plots with disturbed and intact tree canopies to assess early regeneration patterns and seed banks. Native early successional tree species dominated seed bank and seedling layers. Soil properties were similar between intact and disturbed urban plots and local reference forests, thus long-term woody debris removal, hurricane disturbance and subsequent clean-up activities have not caused substantial soil degradation. Non-native species were not abundant throughout the park but were concentrated at the park boundary adjacent to residential neighbourhoods. The results of this study suggest that urban forests can show natural successional trajectories after catastrophic disturbance, and management is probably not necessary for forest recovery in Point Pleasant Park. Conversely, intervention to speed up regeneration of shade-tolerant canopy species may be desired by local citizens, so managers will have to balance conflicting values in developing a restoration plan for the park. © 2007 Elsevier B.V. All rights reserved. Keywords: Seed bank; Regeneration; Disturbance; Soil; Succession; Management

1. Introduction Severe windstorms can greatly alter forest dynamics through vegetation removal and soil disturbance (Foster and Boose, 1992; Catovsky and Bazzaz, 2000; Roberts, 2004). Hurricanes and insect outbreaks are the most important stand-replacing disturbances for forests in Atlantic Canada, but these are very infrequent: most disturbances in the region simply create small gaps (Mosseler et al., 2003). There has been a 100year trend toward increasing hurricane frequency in the region (Environment Canada, 2003) thus changes in forest structure in the region can be anticipated. Knowledge of urban forest responses to intense, infrequent disturbances would greatly benefit urban forest managers as these forests may be subjected to numerous anthropogenic disturbances that may influence vegetation recovery following the disturbance. Vegetation recovery after a disturbance is a function of seed dispersal from neighbouring or distant plants, re-sprouting from root or stump sources, emergence from seed banks, and the

∗

Corresponding author. Tel.: +1 902 420 5506; fax: +1 902 496 8014. E-mail address: [email protected] (J.T. Lundholm).

0169-2046/$ – see front matter © 2007 Elsevier B.V. All rights reserved. doi:10.1016/j.landurbplan.2007.10.003

presence of surviving individuals within the disturbed area (advanced regeneration) (Frelich, 2002; Leck and Sch¨utz, 2005). Frequent and small patch disturbances can result in the maintenance of existing plant assemblages since propagules originate from the surrounding vegetation or local seed bank (Goldblum, 1997). Coarser scale disturbances, however, can contribute to the alteration of vegetation composition by removing local sources for recolonization: tree species are differentially susceptible to windfall (Foster, 1988a). Hurricanes typically remove large proportions of tree canopies while having a low impact on the ground vegetation layer. However, intense storms can disrupt understory vegetation by uprooting trees or causing landslides which can remove the seed bank and reduce the amount of advanced regeneration present in portions of the disturbed area (Roberts, 2004). In these areas, dispersed seeds from colonizers adjacent to the disturbed area as well as species that colonize through root and stump sprouts will dictate the successional dynamics of the disturbed area (Greene et al., 1999). Several studies have shown the importance of buried propagules and seed rain for vegetation regeneration following gap disturbances (Jonsson, 1993; Kalamees and Zobel, 2002; Foster and Tilman, 2003; Clarke and Davison, 2004; Pakeman and Small, 2005). Small, isolated forest fragments may be at greater risk from

112

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

wind disturbance due to lowered protective effects of surrounding canopy and potential absence of seed sources (e.g. Williams et al., 2005) to provide for post-storm re-vegetation, although the role of isolation in the loss of species diversity due to urbanization is complex (Honnay et al., 1999a,b). Two main possibilities for post-hurricane succession are that early successional species already present in the forest can be removed by the disturbance leading to accelerated growth of shade-tolerant species or succession can be re-set by the rapid colonization of shadeintolerant pioneer species (Battaglia et al., 1999). ‘Urban forest’ refers to the tree populations that make up street trees, small woodlots and larger forested areas (often parks) within cities (McPherson et al., 1997), but much of urban forest management literature deals with street trees (McPherson et al., 1999; Maco and McPherson, 2002). Urban forests are subject to a number of human disturbances in addition to the natural disturbance regime of the region. Fragmentation, non-native species, trails, nutrient enrichment, and active management (including the removal or relocation of biomass) can all alter forest properties in urban areas (Sanders, 1984; Nowak, 1994; McPherson et al., 1997; Dwyer et al., 2003). Disturbances in urban forests tend to be small in extent, but occur at high frequencies over a long time (Rudnicky and McDonnell, 1989). Urban influences altering ecosystem properties such as litter accumulation can directly affect regeneration in forests (Kostel-Hughes et al., 1998). Point Pleasant Park (PPP) is located in Halifax, Nova Scotia, the largest urban region in Atlantic Canada (>350,000 people). After the arrival of European colonists in 1749, Point Pleasant Park was cleared of its original forest vegetation in preparation for settlement (Kitz and Castle, 1999). Fortifications and military accommodations were constructed on site and vegetation was regularly removed for defensive purposes. Eventually military use dwindled and residents increasingly used the trails and roads of the peninsula for leisure. By 1866, usage shifted to exclusively recreational and the area was granted park status. As a result, vegetation was allowed to regenerate to the point where prior to 2003, PPP was dominated by long-lived tree species typical of the common mixedwood Acadian forest type of the region (Mosseler et al., 2003). The forest stand structure was primarily even-aged, with a canopy layer and little regeneration of tree species below (Jotcham et al., 1992). An extensive trail system was developed within the park, fragmenting the forest with patches of early successional tree species. Past management recommendations included selective thinning to diversify the canopy and transform forest structure toward a more uneven-aged stand. Public pressure prevented this thinning, although a large number of trees were removed in the early 2000s after an insect outbreak damaged many of the dominant red spruce (Picea rubens) in the canopy. It was predicted that the even age structure of the forest would render the area highly vulnerable to catastrophic windfall (Jotcham et al., 1992). In 2003 Hurricane Juan made landfall on the eastern shore of Nova Scotia, removing over 70% of the tree canopy in PPP. Three years after the catastrophic disturbance to Point Pleasant Park in Nova Scotia, unusually severe windstorms caused massive damage to trees in another prominent urban forest park in Vancouver, British Columbia. Stanley Park is the largest urban

park in Canada and was subject to an extratropical cyclone in December 2006 which removed up to 60% of the trees from the western side of the park. Climate change and extreme weather events will likely increasingly focus attention on highly valued fragmented urban forests. While some natural forests undergo periodic renewal as a result of stand-level disturbances, the additional disturbances associated with urban forests complicate the prediction of postdisturbance recovery. The urban forest studied here (PPP) is isolated from potential seed sources, being surrounded on three sides by salt water and bordering urban residential areas on the fourth. Park managers have also periodically removed snags and coarse woody debris over the last 100 years in order to mitigate fire hazards and enhance public safety (Kitz and Castle, 1999). It is widely believed that the long history of woody debris removal has impoverished the soils in the park, leading to poor opportunities for the forest to replace itself after the hurricane (Neily et al., 2004), but soil nutrient data are lacking. It is also believed by many that the hurricane disturbance may favour the regeneration of non-native species already present in the park. Other studies in eastern North American forests show that canopy structure in 50 years is determined by the species that establish within 4 years of gap formation (McClure et al., 2000) so our study examines early regeneration patterns to assess the potential future forest composition. Increasing urbanization will place more pressure on urban habitats, as the growing size of cities makes forest biodiversity more valuable to the human population. This paper examines the post-hurricane recovery of a large forest ecosystem that has retained some historical and ecological continuity with the typical natural forests of the region. Understanding the effects of large-scale disturbances will be essential in learning to manage such fragments for maximum sustainability. The goals of the current study were to (1) evaluate the potential for natural regeneration in the post-hurricane forest vegetation, including seedling regeneration and seed banks; (2) evaluate the threat of non-native species colonization post-hurricane and (3) compare urban forest soil properties with soils in other local forests with lower rates of ongoing disturbance and coarse woody debris removal. Overall, we wish to address the known management concerns of park staff, local officials and concerned citizens to assess the need for restoration planting, control of non-native species and/or soil fertilization. 2. Methods 2.1. Study site Point Pleasant Park is a 75 ha forested park, located at the south end of the Halifax Peninsula, Nova Scotia, Canada (44◦ 39 N; 63◦ 36 W). The underlying geology of this area consists of pyritic slate, schist, and migmatite rock types (AGS, 1994). Soils are podzolic, brown shaley loam (MacDougall and Cann, 1963). Before the 2003 hurricane, most of the park had mature, even-aged forest dominated by Acadian forest species: red spruce (Picea rubens) and white pine (Pinus strobus) with scattered red maple (Acer rubrum), balsam fir (Abies balsamea)

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

and paper birch (Betula papyrifera) (Jotcham et al., 1992; Kottegoda, 2001; Neily et al., 2004). The Acadian forest occurs within ecoclimatic zones considered to be cool temperate or boreal, with more boreal environments found closer to the coast (Weber and Flannigan, 1997). This forest type has large floristic overlap with the more widespread boreal forest, including many shared tree, shrub and herbaceous understorey species. There was no quantitative information on soils or other environmental factors in the park prior to the hurricane. In September 2003 a class 2 hurricane destroyed over 70% of the PPP forest canopy (Environment Canada, 2003). Damage to trees included stem breakage, uprooting and removal of limbs. Stem breakage was the most frequent type of damage; little uprooting was seen which is consistent with other hurricane disturbances where high winds were not associated with large amounts of rain to loosen soil (Foster, 1988a). Gaps created were heterogeneously distributed throughout the park, ranging in size from 300 to 10,000 m2 . Larger blowdown areas were concentrated at the south-east, south-facing low elevation area and the least disturbed areas were in the north-west, north-facing area (map: Fig. 1). The coarse-scale pattern of damage is consistent with the main wind direction from the south during the storm but the proximity of damaged and relatively intact stands indicates that discrete downbursts also caused much of the canopy loss (Greenberg and McNab, 1998) leading to a more random distribution of disturbed patches over the landscape (Boutet and Weishampel, 2003). Fallen trees and coarse woody debris were removed from the park as part of clean up after the hurricane damage in 2003, some of which was chipped and left on site or applied to trails. 2.2. Sampling design In spring 2004 (the first growing season after the hurricane) PPP consisted of forested habitat in various successional stages that were dissected by an extensive trail system creating forest blocks ranging from 0.5 to 5.5 ha. To compare the vegetation in areas heavily damaged by the hurricane with relatively intact areas, we found 30 blocks defined by paths or topographic features that contained both intact and disturbed areas, but were similar in terms of elevation, topography, aspect and pre-hurricane species composition (as assessed by stumps, coarse woody debris and extant vegetation). From these 30 forest blocks, 15 were randomly chosen and patches within each block were identified as disturbed (<20% canopy cover) or intact (>80% canopy cover). A single 1 m × 1 m permanently marked plot was established at a random location within each of the 15 disturbed and intact canopy patches, at least 20 m from any road or trail, and not containing trees over 2 m. A 10 m × 10 m plot was also marked with the 1 m × 1 m plot at its center in each location. Plants were classified as seedlings or adults and identified to species where possible using Roland and Smith (1963), Gleason and Cronquist (1991) and Zinck (1998) with nomenclature and native/non-native status following the Provincial Species List and Ranks for Nova Scotia (ACCDC, 2005). Plots were visited several times over the growing season (May–September) to find all species.

113

In summer 2005, percent cover of each species present within each 10 m × 10 m plot was visually classified as <5, 5–25, 25–50, 50–75, or 75–100% (classes: 0.1, 1, 2, 3, 4). This vegetation data was separated into layers based on height: seedling (<5 cm), herbaceous (5–15 cm), shrub (15–2 m), lower canopy (2–10 m), mid-canopy (10–20 m) and canopy (>20 m). The 1 m × 1 m plots were used for counts of seedlings and yearlings (individuals germinated in 2004) which were used to calculate the seedling and yearling density of each species. This was only done in the 1 m × 1 m plots due to time constraints. Mean densities of seedlings and yearlings in 2005 were compared using paired t-tests after assessment of normality and homogeneity of variance. Relative cover of non-native species was calculated as the sum of cover classes of non-native species divided by the total sum of all species cover classes (×100%). 2.3. Seed bank analyses In 2005, soil was harvested over 1 week in early May (before leaf-out for the dominant tree species and germination of most species) (Leck et al., 1989; Stark et al., 2003) from each 10 m × 10 m plot. Soil for seed bank analysis was harvested from six 10 cm3 subplots at designated points. The soil from all subplot sample locations was pooled for analysis at the 10 m × 10 m plot level, thus there were 30 samples total. Soil samples were stored in a 4 ◦ C refrigerator after excavation until all samples were collected, then concentrated using the methods of ter Heerdt et al. (1996). The concentrated soil was spread (1 cm thick) in trays containing a 1 cm layer of sand and kept moist throughout the experiment. The trays were kept outside on the roof of a campus building. As seedlings emerged, they were identified and removed, or transplanted and grown until identification was possible. The study was terminated after 8 weeks when seeds had stopped germinating. This method allows an estimation of non-dormant viable seed banks. Seed densities in plots were converted to density/m2 to allow comparison with other studies. Seed bank density, richness and percent nonnative seeds were compared between intact and disturbed plots using paired t-tests after assessment of normality and homogeneity of variance. Overall species composition was compared between intact and disturbed plots using analysis of similarity (ANOSIM), using Bray Curtis distances (Primer 6.1: PrimerE Ltd., Plymouth, UK; Clarke and Gorley, 2001). The effect size R indicates the magnitude of difference between groups (0–1.0 with 0 indicating complete similarity and 1.0 complete difference). 2.4. Soil nutrients Soil nutrients were sampled at four predetermined points within the 10 m × 10 m plots: 4.2 m on the diagonal from the SE and NW corners and at the midpoint of the north and south sides of the 1 m × 1 m plots in the center of the larger plots. For chemical analysis, at least 250 ml of soil was collected from the B horizon (when possible) from the four sampling points within each plot at the same time as the soil for the seedbank analysis. Organic matter content was determined by loss on ignition after

114

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

Fig. 1. Location of 30 plots in Point Pleasant Park, Halifax, Nova Scotia, Canada. U indicates an intact plot (>80% canopy cover in 2004); D indicates a disturbed plot (<20% canopy cover in 2004). “OUT” identifies one outlier plot (see text for details).

1 h at 450 ◦ C. Soil pH was determined following the AdamsEvans buffer method (COEC, 1992) and a pH meter (Accumet AR25: Fisher Scientific, Ottawa, Canada). To quantify the soil content of P, K, Ca, Mg, Na, Mn, Cu and Zn, Mehlich 3 extraction was used, followed by the inductively coupled argon plasma method. Cation exchange capacity (CEC) was determined by calculating the sum of the milliequivalents of sodium, calcium, potassium, magnesium, and hydrogen per 100 g of soil (Baird, 1999). Nitrate was extracted from 10 g of soil using a dilute salt solution and measured with a specific ion electrode and a double reference electrode. All analyses were conducted by Nova Scotia Agricultural College Lab Services (Truro, Nova Scotia). Three forested areas, generally less damaged by the hurricane, located outside the park were chosen in 2005 for comparison of soil nutrient content. These reference sites were chosen for their proximity to PPP, similarity in vegetation types

and relative lack of intensive management (Frog Pond, 38 ha: 44◦ 37 N, 63◦ 40 W; Hemlock Ravine, 80 ha: 44◦ 41 N, 63◦ 42 W; and Long Lake Provincial Park, 2103 ha: 44◦ 37 , 63◦ 42 ). Three sample sites corresponding to intact canopy plots (>80% canopy cover) were randomly chosen within the three locations and soil was sampled and analysed as above. Overall differences in soil chemical properties were compared using ANOSIM, using Euclidian distances (Clarke and Gorley, 2001). 3. Results 3.1. Species composition The most frequent tree species encountered within the canopy layer (occurring in more than twenty 10 m × 10 m plots) were red maple (Acer rubrum), paper birch (Betula papyrifera),

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

Fig. 2. Seedling and yearling density comparisons between intact forest plots and disturbed plots located within Point Pleasant Park (means ± 1S.E. calculated for fifteen 1 m × 1 m plots in each category). Groups were not significantly different (see text).

American mountain ash (Sorbus americana), white pine (Pinus strobus) and red spruce (Picea rubens) (Appendix A). Within the Shrub Layer, the most frequent species (occurring in 16 or more plots) included red maple (Acer rubrum), American mountain ash (S. americana) and staghorn sumac (Rhus typhina). The most frequent herbaceous layer species (occurring in 15 or more plots) included Maianthemum canadensis, Aralia nudicaulis, Clintonia borealis, Cornus canadensis, Gaultheria procumbens, Trientalis borealis and Vaccinium angustifolium. All thirty plots contained at least one seedling with the most frequent species (occurring in more than 10 plots) being Acer rubrum, Betula papyrifera, S. americana, and Pinus strobus. Total vegetation cover in all plots ranged from 50 to 100%. An average of 312 seeds/m2 germinated from the soil seed banks, with a range of 0 to ∼4000 seeds/m2 . No seeds germinated from the soil collected at four plots. A total of 563 viable seeds germinated from the seed bank samples, representing 35 species.

115

Fig. 4. Seed bank composition by plant type for 15 intact and 15 disturbed 10 m × 10 m plots in Point Pleasant Park.

plots (ANOSIM: R = −0.037; P = 0.450). Seed banks were three times denser in intact vs. disturbed plots, but this difference was not significant (t14 = −1.577, P = 0.137) (Fig. 3). Seed bank richness and percent non-native seeds were also not significantly different between treatments (richness: t14 = −1.486, P = 0.159; percent non-native: t14 = −0.837, P = 0.417)(Fig. 3). Intact plots had a mean of 58% non-native seeds, compared to 5% for disturbed plots, but this declined to 1% when a single outlier plot, dominated by a locally common non-native annual (G. tetrahit) was removed from the analysis (Fig. 3). This plot was removed as it was an outlier for nitrate content (mean of 2.7 ± 0.5 ppm compared with 0.2 ± 0.05 ppm mean for all other plots) and had the greatest relative non-native species cover above-ground (62%). The majority of the seed bank in both plot classes (once the outlier plot was removed) is comprised of (native) tree seeds (Fig. 4). 3.3. Non-native species

3.2. Disturbed versus intact seedling regeneration There were no differences in seedling or yearling density between intact and disturbed plots (seedlings: t14 = 1.824, P = 0.090; yearlings: t14 = 0.730, P = 0.478), but yearlings were more abundant than seedlings in the disturbed treatment only (t14 = −2.345, P = 0.034) (Fig. 2). The most abundant species in the disturbed plot seed banks were Betula papyrifera (31% total seed density), Carex spp. (15%), and Picea spp. (2.3%). In the intact seed bank, the most abundant species were Galeopsis tetrahit (34%), Betula papyrifera (31%), and Carex spp. (15%). Despite these differences, there was no overall difference in seedbank species composition between disturbed and intact

Sixteen out of 105 species encountered in the vegetation are considered non-native to the region (ACCDC, 2005). Hieracium maculatum and Quercus robur were the most frequent nonnatives, found in the vegetation of eight and seven plots respectively. One intact canopy plot had 11 of the 16 non-native species; a second had 5 species. Ten disturbed and nine intact plots had at least one non-native species. Non-native seedlings were observed in six of the thirty plots but total non-native seedlings made up less than 5% ground cover per plot. Relative cover of non-natives was not significantly different between intact and disturbed plots (paired t-test: t14 = 0.895, P = 0.386, intact mean ± S.E.: 6.1 ± 4.2%; disturbed: 2.6 ± 0.8%). When

Fig. 3. Seedbank density (a), species richness (b) and percent non-native seeds (c) for intact, disturbed and intact with one outlier omitted (see methods). Data for thirty 10 m × 10 m plots in Point Pleasant Park. Groups were not significantly different.

116

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

the outlier plot was removed, the mean relative non-native cover in the intact plots dropped to 2.6 ± 1.4%. Five non-native species were present in the seed bank (Acer pseudo-platanus, Achillea millefolium, Fraxinus excelsior, G. tetrahit, Plantago major and Solanum dulcamara), accounting for 38% of the total seed density but each occurred at two or fewer plots. Achillea, Fraxinus and Plantago were not present in the plot vegetation and accounted for less than 1% of the total seed density in all plots. The plots with non-native species in the seed bank contained 86, 25, 12.5, 4.3, 3.7 and 3.2% non-native seeds. One plot contained all five non-native species in the seed bank. The outlier plot with the greatest above-ground richness and cover of non-natives also had the greatest proportion of non-native seeds in the seed bank (86%). 3.4. Soil nutrient comparison Overall, soil chemical properties were similar between reference and park plots and between disturbed and intact plots. While the global test (ANOSIM) that determines whether there are any significant differences among any of the groups was statistically significant (global test: R = 0.068, P = 0.023), Rvalues of less than 0.20 are typically considered not biologically meaningful (Clarke and Gorley, 2001). The pairwise comparisons tell the same story: differences are either not statistically significant or so small as to be uninterpretable (intact vs. reference: R = 0.162, P = 0.108; disturbed vs. reference: R = 0.199, P = 0.066; intact vs. disturbed: R = 0.032, P = 0.018). Examination of differences in means and overlap of error bars shows that PPP soils contained more than twice the nitrate content, more organic matter, potassium, calcium, sodium, sulphur and half the available phosphorus content of the reference soils (Fig. 5). Available P was highly spatially variable within PPP, with individual samples ranging from 0.5 to 70 mg/kg. 4. Discussion 4.1. Post-hurricane forest communities

Fig. 5. Mean values (±1S.E.) for chemical soil parameters in 15 intact, 15 disturbed and 9 reference 10 m × 10 m forest plots.

Overall similarities in species composition between intact and disturbed areas suggest that the severity of the disturbance was not sufficient to remove enough of the pre-disturbance vegetation required to alter the post-disturbance community structure (Castelli et al., 1999). The dominant tree species in the study plots are pioneer species such as red maple (Acer rubrum), paper birch (Betula papyrifera), and American mountain ash (S. americana), indicating an early successional state following the hurricane as these species were found to dominate some areas classified as having intact canopies as well as disturbed areas. Many understorey species can tolerate a range of soil moisture and light conditions that enable them to persist after removal of the canopy causes fluctuations in these levels (Frelich et al., 2003). Species such as bluebead lily (C. borealis) are latesuccessional understory species that have persisted at least 2 years after the hurricane and are common in the park. Other studies have shown that hurricane disturbances, whether natural or simulated often result in ground vegetation being retained

or rapidly recovered, despite the physical impacts of the disturbance (Cooper-Ellis et al., 1999). In general, the hurricane seems to have re-set succession in the park by favouring early successional, shade-intolerant pioneer tree species (Battaglia et al., 1999). While it might be expected that urban forests show different successional trajectories compared with rural forests due to the presence of stressors associated with cities, for example, larger squirrel population densities (Rudnicky and McDonnell, 1989), our example in a relatively small city appears to be following typical natural successional patterns. The visual impacts of the hurricane damage are disproportionally greater than the ecological damage to forest structure as found in other studies (Cooper-Ellis et al., 1999). This highlights the potential for conflict between the findings of ecological studies and the impact on the public of the visual assessment of damage, and likely will shape recovery planning for PPP. The disturbance pattern in the park suggests that even-aged succession from heavily disturbed canopy areas will occur,

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

117

matching patterns common to coastal forest types in New England (Foster, 1988b) but not typical of Acadian forests which are presumed to show uneven-age successional trajectories based on small-scale tree fall gaps (Mosseler et al., 2003). Past recommendations to thin the even-aged forests of the park might have resulted in a forest where fewer trees would have been lost to the hurricane, but while even-aged overstoreys in PPP were most likely a result of intentional clearcutting, similar structures can result from catastrophic natural disturbance (Foster, 1988b). Increased hurricane frequency in the region suggests that catastrophic disturbances may increasingly shape forests in Atlantic Canada. Fast-growing pioneer tree species are now dominant in PPP and will likely follow a thinning kind of succession (Foster, 1988b) in which mortality will thin the stands, allowing more shade-tolerant species such as Picea rubra and Tsuga canadensis to reach the canopy. Park managers could intervene in this successional process by removing fast-growing trees such as Acer rubrum or Betula papyrifera and open up the new canopy and/or planting species typical of the predisturbance canopy such as Pinus strobus or Picea rubra. While the original composition of the forests of PPP included many white pine (Pinus strobus), if tree planting is adopted as a restoration strategy for the park, managers will have to decide whether to replant this windthrow-susceptible species (Foster, 1988a) or to emphasize less susceptible species in order to foster greater resistance to disturbance in the future forest. This represents one of the key dilemmas in managing this and other urban forests: do we prioritize natural processes and patterns (an ecological approach) or emphasize competing views more closely linked with preferences of the public, such as resilience in the face of future disturbance? Likewise, the decision to manage the forest or leave it alone to recover naturally pits worldviews and value systems against one another.

(Pinus strobus), with a variety of tree, shrub, and herbaceous species, such as yellow birch (Betula alleghaniensis), paper birch (Betula papyrifera), and trembling aspen (Populus tremuloides). Although white pine was the most frequent tree canopy species, occurring in 20 of 30 plots, no white pine seedlings germinated in the seed bank study. In the above-ground sampling, white pine was found in all layers including the seedling layer and it was the fourth most common species of seedling observed. The absence of white pine seedlings in the seed bank may be explained by a poor year for seed production in 2004, as good seed years are thought to occur only every 3–5 years and seeds are not long-lived in the soil (Burns and Honkala, 1990). It is also possible that some unknown difference between conditions in our ex situ germination area and PPP could have prevented our detection of viable white pine seeds in the soil. There were no overall differences between intact and disturbed plots in the density, species richness or species composition of the seeds that germinated from the seed bank when one outlier plot was removed. Disturbed sites did contain an average of four times fewer seeds per m2 , yet the difference was not statistically significant: the soil seed banks were heterogeneously distributed between different areas of the park, leading to high variances in seed bank density and richness. Though hurricane damage in more disturbed sites may provide suitable seedbeds for regeneration by exposing mineral soil in treefall pits (Peterson and Pickett, 1990; Neily et al., 2004), areas of bare mineral soil or treefall pits were infrequent in the park and were not encountered in the seed bank sampling subplots. Our results suggest that the hurricane disturbance neither destroyed seed banks nor provided greater opportunities for recruitment from the seedbank.

4.2. Regeneration and seed banks

Urban parks with remnant forests often contain non-native species (Turner et al., 2005) that are a concern for park managers: non-native species may inhibit the establishment and growth of native plant species in post-disturbance communities. Stand initiating events in Nova Scotian Acadian forests such as clearcutting have been shown to present opportunities for exotic taxa to invade and persist during secondary succession (Moola and Vasseur, 2004). The presence of invasive plant species coupled with the fragmentation of forest stands by trails may proliferate the spread of exotics into the interior of the park. Many if not all of the non-native species sampled during this study were present in the park before the hurricane disturbance and comprised a small portion of the advanced regeneration following the disturbance. The greatest non-native cover was in an intact plot with very high levels of soil nitrate. The underlying geology of Point Pleasant Park consists of pyritic slate which creates soils that have a relatively low pH. This may help to explain the low cover of non-native species in emerging species assemblages following the disturbance since native species, better suited to the soil conditions of the area, may hold a competitive advantage over non-native species that may favour less stressed environments.

The park is isolated from the dispersal of propagules from natural forest. Fragmented, anthropogenically disturbed forests are slower to recover following gap-creating events than forests in closer proximity to relict old forests (Singleton et al., 2001; Moola and Vasseur, 2004) thus it might be predicted that seed limitations would impede natural regeneration in PPP. The seed bank seed densities in this study (from 0 to 4000 seeds/m2 ) are in the range of findings reported in other Acadian forest sites dominated by peat bog and conifers (from 0 to 30 seeds/m2 ) and deciduous trees (>100 seeds/m2 ) (Moore and Wein, 1977). PPP seed banks contained a high proportion of tree seeds relative to other reports from other boreal, mixed and deciduous forests (Qi and Scarratt, 1998; Leckie et al., 2000; Landenberger and McGraw, 2004). Vegetation removal, human influence and increasing distance from old forests can have negative consequences for regeneration of late-successional forest plants (Moola and Vasseur, 2004). However, seed banks and seedling layer regeneration show that forests in PPP are undergoing succession along a trajectory typical for this forest region. Neily et al. (2004) predicted that the majority of natural regeneration following Hurricane Juan would consist of white pine

4.3. Non-native plants

118

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

Perhaps because of the general absence of soil exposure after the hurricane, there is little evidence from the current study to suggest that non-natives are spreading into areas damaged by the hurricane: there is no difference in non-native cover or seed bank density between intact and disturbed plots. The plot with the highest density of non-native seeds in the seed bank was an intact canopy plot. Most of these seeds were accounted for by a single non-native species, G. tetrahit. This plot also contained the most non-native species, probably due to its proximity to human residential development. Areas near trails and road edges were not sampled in this study, but many non-native species are found in these areas. The species inhabiting trail and road edges are not considered invasive for the most part, but represent ruderal species common to urbanized, disturbed environments. There is also a disproportionate absence of non-natives in the seedling layer: 19 plots had non-natives in the vegetation but only six had non-native seedlings. Only one of the species in the park, Norway maple (Acer plantanoides) is considered to be invasive (Catling and Mitrow, 2005). This species occurred in only four plots in this study, but no seedlings were encountered. Mature trees, saplings and seedlings of Norway maple, however, are frequently found along the north and west sides of the park where street trees act as seed sources. An unpublished survey shows that Norway Maple size and abundance, declines rapidly with distance from edges close to street trees and trails in PPP. In terms of the overall regeneration of the interior forest patches of the park, this study does show that the contribution of non-native species is minimal and species assemblages emerging following the hurricane disturbance are primarily composed of native plant species. Non-native species do not seem to be hindering the establishment and growth of native plant species. Nevertheless, non-native species are present in the park and should be monitored to determine their abundance as post-hurricane forest recovery proceeds. Our results suggest that the recent hurricane damage and the long-term impact of trail disturbance as a vector for non-natives are independent problems for park managers to solve. 4.4. Soil properties and management history Decaying trees and woody debris release nutrients into the forest floor, thus long-term removal of biomass may cause poor soil development and nutrient accumulation. Our study area has a long history of woody debris removal coinciding with its status as a park. Removal of woody debris has the most immediate impact on the organic layer of the soil horizon, while having little to no effect on the long-term carbon and nitrogen contents of mineral soil (Knoepp and Swank, 1997; Johnson and Curtis, 2001). On the other hand, the accumulation of organic material on the forest floor, which may be limited by recent removal of coarse woody debris, may have more effect on microsite availability and soil moisture. Based on these studies it may be inferred that the PPP soils may still retain adequate nutrient levels regardless of the removal of large amounts of debris as long as finer litter was allowed to accumulate, however this activity may reduce the amount of organic content contained within the O soil hori-

zon (Laiho and Prescott, 2004), which was not sampled in this study. Reports on PPP soils have attributed a loss of soil fertility to compaction and past management activities that removed coarse woody debris and reduced hardwood percentages (Neily et al., 2004), but the only evidence that might suggest soil degradation are average phosphorus contents approximately half that at reference sites which have not been subjected to a long history of coarse woody debris removal. Mean extractable phosphorus contents in PPP soils were double that in coastal forest soils in British Columbia considered to be limiting in available P for coniferous tree growth (Preston and Trofymow, 2000) and double that of uncut boreal forest soils (Simard et al., 2001). The difference in available P content between PPP and the reference sites is not likely due to increased uptake by postdisturbance vegetation as intact canopy and disturbed plots within PPP showed similar phosphorus levels in the mineral soil (Fig. 5). Likewise, the overall impact of biomass removal on soil phosphorus levels is estimated to be only about 3% of total ecosystem P stores for whole-tree harvesting (Yanai, 1998). Past management in PPP has removed primarily large treefalls and branches and does not correspond to whole-tree harvesting. Nevertheless, biomass and forest floor layers cycle P more rapidly hence represent a more important source of P for a regenerating forest. Greater nitrate levels in the mineral soil of PPP compared with the reference sites could be due to increased levels of microbial activity and nitrogen mineralization after canopy opening (Hope et al., 2003), but there was no significant difference between intact and disturbed canopy plots within PPP and this issue is debated by forest ecologists. 5. Conclusions and recommendations Despite negative environmental impacts associated with the various anthropogenic activities conducted within and around Point Pleasant Park, this study found that natural regeneration is occurring following the disturbance caused by Hurricane Juan. The regenerating vegetation consists mainly of native species assemblages emerging from local seed and propagule sources. The presence of non-native species within core forest patches of PPP is minimal but should be monitored as forest regeneration continues. Soils of the park were not infertile overall, although there may be local phosphorus limitation. Long-term removal of woody debris may not necessarily compromise urban forest soils. If regeneration is found to continue along the successional trajectory apparent in this study, passive restoration techniques such as erecting fencing or closing certain areas to allow for natural regeneration may be all that is necessary to ensure the development of a viable forest ecosystem. Although natural regeneration is occurring within the park, it may be desired to increase the rate of recovery by supplemental planting of large trees. Management decisions will be based on economic and social issues as well as ecological science and will be a product of the dominant values held by community members and municipal officials. While urban forests in large urban centers

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

may have special barriers to forest regeneration following disturbance, this study in a metropolitan area that is small but relatively large regionally, shows that natural succession can occur in urban forests, and suggests that managers should not make assumptions about forest health in the absence of local ecological data on soil nutrients and non-native species. While received wisdom and local knowledge of urban forests should be taken seriously, it may be prudent to subject popular claims to scientific scrutiny before designing policy or management strategies. Acknowledgements We are grateful to Andrea Coombs, Jill DiPenta, Erica Oberndorfer and Erinn O’Toole who helped with field work. Greg Baker provided valuable help with GIS and field surveying. We are grateful for the comments of three anonymous reviewers for improving the manuscript. We also thank Kevin Keys and Peter Neily for critical discussion of our results. Funding for Species Latin name

Abies balsamea (L.) Mill. Acer pensylvanicum L. Acer platanoides L. Acer pseudoplatanus L. Acer rubrum L. Acer saccharinum L. Achillea millefolium Aesculus hippocastanum L. Agrostis capillaris L. Alnus incana (L.) Moench. Amelanchier sp. Aralia hispida Vent. Aralia nudicaulis L. Betula alleghaniensis Britt. Betula papyrifera Marsh. Bromus erectus Huds. Calluna vulgaris (L.) Hull Carex communis L. Bailey Carex conoidea Willd. Carex debilis Michx. Carex deflexa Hornem. Carex folliculata L. Carex ovalis Good. Carex rosea Willd. Carex sp. 1 Carex sp. 2 Carex sp. 3 Carex sp. 4 Carex trisperma Dewey Carex umbellata Schkuhr. Clintonia borealis (Aiton.) Raf. Comptonia perigrina (L.) Coult. Coptis trifolia (L.) Salisb. Cornus canadensis L. Cypripedium acaule Aiton. Danthonia spicata (L.) Beauv. Dennstaedtia punctilobula (Michx.) Moore Deschampsia flexuosa (L.) Trin. Dichanthelium acuminatum var. fasciculatum (Torr.) Fern.

Species common name

119

the research was provided by the Halifax Regional Municipality (HRM), Saint Mary’s University, and the Natural Sciences and Engineering Research Council of Canada (NSERC). Appendix A A.1. List of species and frequencies for all native and non-native species encountered in Point Pleasant Park Species list for all native and non-native species sampled in Point Pleasant Park. The S rank indicates the Nova Scotia rarity status for each species sampled where; S2—rare, 6–20 or fewer occurrences; S3—uncommon, 21–100 occurrences; S4—widespread, fairly common, >100 occurrences; S5—abundant, demonstrably widespread; SE—exotic established in the province; SU—unrankable, more information required (ACCDC, 2005). Frequency refers to the number of plots each species was reported in.

Native/non-native

S rank

Frequency above-ground

Frequency seed bank

Disturbed

Undisturbed

Disturbed

Undisturbed

7 2 3 1 14 3 0 1 1 2 9 0 8 0 9 0 1 1 0 2 5 1 0 1 0 0 0 0 2 0 6 0 6 9 1 2 4

0 0 0 1 2 0 1 0 0 0 0 0 0 0 9 1 0 0 0 0 0 0 0 0 0 3 2 1 0 1 0 0 0 0 0 1 0

0 0 0 0 2 0 0 0 0 0 0 1 0 0 11 1 0 0 0 0 0 0 0 0 0 5 5 2 0 6 0 0 0 0 0 1 0

8 0

1 0

2 0

Balsam fir Moose maple Norway maple Sycamore maple Red maple Sugar maple Yarrow Horsechestnut Colonial bentgrass Speckled alder Serviceberry species Bristly sarsaparilla Wild sarsaparilla Yellow birch White birch Meadow brome Heather Fibrous-root sedge Field sedge White-edge sedge Short-stemmed sedge Long sedge Oval sedge Rosy sedge Sedge species Sedge species Sedge species Sedge species Three-seed sedge Hidden sedge Bluebead lily Sweet fern Gold thread Bunch berry Pink lady’s slipper Poverty oat-grass Hay-scented fern

Native Native Non-native Non-native Native Native Non-native Non-native Non-native Native Native Native Native Native Native Non-native Non-native Native Native Native Native Native Non-native Native

S5 S5 SE SE S5 S5 SE SUSE SE S5 S? S5 S5 S5 S5 SE SE S5 S4? S5 S4 S5 SE S3?

Native Native Native Native Native Native Native Native Native

S5 S4 S5 S5 S5 S5 S5 S5 S5

8 0 1 0 14 2 0 0 0 1 9 9 12 1 12 0 0 2 1 1 8 0 2 0 1 0 0 0 2 3 13 1 8 10 1 1 4

Crinkled hairgrass Western witchgrass

Native Native

S5 S5

11 1

120

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

Appendix A (Continued ) Species Latin name

Dichanthelium acuminatum (Gould.) Clark. Diervilla lonicera P. Mill. Dryopteris carthusiana (Vill.) Fuchs. Epilobium angustifolium L. Erechtites hieracifolia (L.) Raf. Fagus grandifolia Ehreh. Fragaria vesca L. Fraxinus excelsior Galeopsis tetrahit L. Gaultheria hispidula (L.) Muhl. Gaultheria procumbens L. Gaylussacia baccata (Wang.) K. Koch. Gramineae sp. Hamamelis virginiana L. Hedyotis caerulea (L.) Hook. Heiracium sp. Hieracium kalmii L. Hieracium maculatum C. Gmelin. Hieracium × flagellare Willd. Ilex verticillata (L.) A. Gray. Iris versicolor L. Juncus tenuis Willd. Juniperus communis L. Kalmia angustifolia L. Linnaea borealis L. Luzula multiflora (Retz.) Lejeune Lycopus uniflorus Maianthemum canadensis Desf. Matteuccia struthiopteris (L.) Todaro Medeola virginiana L. Melampyrum lineare Desr. Mitchella repens L. Monotropa uniflora L. Nemopanthus mucronata (L.) Trel. Oclemena accuminatus (Michx.) Greene Oenothera parviflora L. Oryzopsis asperifolia Michx. Osmunda cinnamomea L. Osmunda regalis L. Picea abies (L.) Karst. Picea glauca (Moench.) Voss. Picea rubens Sarg. Pinus strobus L. Pinus sylvestris L. Plantago major Potentilla simplex Michx. Prenanthes trifoliolata (Cass.) Fern. Prunus pensylvanica L. f. Prunus serotina Ehrh. Pteridium aquilinum (L.) Kuhn. Quercus robur L. Quercus rubra L. Ranunculus acris L. Rhus typhina L. Rubus allegheniensis Porter Rubus pubescens Raf.

Species common name

Native/non-native

S rank

Frequency above-ground

Frequency seed bank

Disturbed

Disturbed

Undisturbed

Undisturbed

Panic grass

Native

S5

0

1

1

2

Bush honeysuckle Spinulose wood fern

Native Native

S5 S5

1 0

1 1

0 0

0 0

Fire weed or large willow herb Fireweed American beech Strawberry European ash Hemp-nettle Creeping snowberry Wintergreen Huckleberry

Native Native Native Native Non-native Non-native Native Native Native

S5 S5 S5 S4 SE SE S5 S5 S5

1 2 0 1 0 0 1 9 4

0 1 1 0 0 1 1 11 6

0 2 0 1 0 0 0 0 0

0 2 0 0 2 1 0 0 0

Native Native

S5 S5

Native Non-native Non-native Native Native Native Native Native Native Native Native Native Native

S2? SE SE S5 S5 S5 S5 S5 S5 S5 S5 S5 S5

0 4 1 0 1 5 1 1 0 1 0 8 1 2 0 15 0

1 2 0 0 3 3 0 0 1 0 1 3 0 0 0 14 1

0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 1 0

0 0 0 1 0 2 0 0 0 0 0 0 0 0 1 0 0

Indian cucumber root Cow wheat Partridge-berry Indian pipe False holly Whorled wood aster

Native Native Native Native Native Native

S5 S5 S5 S5 S5 S5

1 1 6 0 5 6

0 0 4 1 2 6

0 0 0 0 0 0

0 0 0 0 0 2

Small-flowered evening primrose White-grained mountain ricegrass Cinnamon fern Royal fern Norway spruce White spruce Red spruce White pine Scots pine Common plantain Old-field cinquefoil Three-leaved rattlesnake-root

Native Native Native Native Non-native Native Native Native Native Non-native Native Native

S4? S5 S5 S5 SE S5 S5 S5 SE SE S5 S5

0 0 0 0 3 7 9 10 2 0 1 1

1 1 2 1 2 7 12 11 4 0 0 4

0 0 0 0 0 0 1 0 0 1 0 0

0 0 0 0 0 1 2 0 0 0 0 3

Pin cherry Black cherry Bracken fern English oak Red oak Buttercup Staghorn sumac Blackberry Dewberry

Native Native Native Non-native Native Non-native Native Native Native

S5 S5 S5 SE S5 SE S4 S5 S5 S5

9 1 7 5 10 1 12 5 0

2 0 6 2 19 1 4 4 1

1 0 0 0 0 0 0 1 0

0 0 0 0 0 0 0 3 0

Grass species Witch hazel Bluets Hawkweed species Kalm’s hawkweed Common hawkweed Whiplash hawkweed Winterberry Blueflag iris Slender rush Common juniper Sheep laurel Twin flower Common woodrush Northern bugleweed Canada mayflower Ostrich fern

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

121

Appendix A (Continued ) Species Latin name

Rubus strigosus Michx. Scirpus sp. Solanum duclamara L. Solidago puberula Nutt. Solidago rugosa Ait. Sorbus americana Marsh. Symphyotrichum lateriflorum (L.) Love Symphyotrichum novi-belgii (L.) Nesom. Taraxacum officinale Weber Taxus canadensis Marsh. Torreyochloa pallida (Torr.) G.L. Church Trientalis borealis Raf. Tsuga canadensis (L.) Carr. Ulmus americana L. Ulmus glabra Huds. Vaccinium angustifolium Ait. Vaccinium myrtilloides Michx. Vaccinium vitis-idaea L. Viburnum nudum L. Viola cucullata Aiton.

Species common name

Native/non-native

S rank

Wild raspberry Bullrush species Bittersweet nightshade Downy goldenrod Rough-leaved goldenrod American mountain ash Farewell-summer aster

Native Native Non-native Native Native Native Native

S5

New York aster

Frequency above-ground

Frequency seed bank

Disturbed

Disturbed

Undisturbed

Undisturbed

SE S5 S5 S5 S5

5 0 2 3 1 12 1

3 0 2 2 3 9 0

2 1 0 0 2 0 1

1 0 1 0 1 0 0

Native

S5

1

2

0

0

Dandelion Yew Pale manna grass

Non-native Non-native Native

SE SE S4 S5

0 0 0

1 1 1

0 0 0

0 0 0

Star flower Eastern hemlock American elm Scotch elm Lowbush blueberry Canada blueberry Mountian cranberry Wild raisin Blue violet

Native Native Native Non-native Native Native Native Native Native

S5 S4 S5 S4 SE S5 S5 S5 S5 S5

8 2 0 0 11 6 1 10 2

12 1 1 1 11 7 3 10 4

0 0 0 0 0 0 0 0 0

0 0 0 0 0 0 0 0 1

References Atlantic Canada Conservation Data Center (ACCDC), 2005. Provincial Lists and Ranks. Retrieved January 11, 2006 from: http://www.accdc.com/ products/lists/. Atlantic Geoscience Society (AGS), 1994. Geology Map of Nova Scotia. Land Registration and Information Services & Province of Nova Scotia Department of Supply and Services. Nova Scotia, Canada. Baird, C., 1999. Environmental Chemistry. W.H. Freeman and Company, New York. Battaglia, L.L., Sharitz, R.R., Minchin, P.R., 1999. Patterns of seedling and overstory composition along a gradient of hurricane disturbance in an old-growth bottomland hardwood community. Can. J. For. Res. 29, 144– 156. Boutet Jr., J.C., Weishampel, J.F., 2003. Spatial pattern analysis of pre- and posthurricane forest canopy structure in North Carolina, USA. Landscape Ecol. 18, 553–559. Burns, R.M., Honkala, B.H., 1990. Silvics of North America. 1: Conifers; 2: Hardwoods. Agriculture Handbook 654. U.S. Department of Agriculture, Forest Service, Washington, DC. Castelli, J.P., Casper, B.B., Sullivan, J.J., Latham, R.E., 1999. Early understory succession following catastrophic wind damage in a deciduous forest. Can. J. For. Res. 29, 1997–2002. Catling, P., Mitrow, G., 2005. A prioritized list of the invasive alien plants of natural habitats in Canada. Can. Bot. Assoc. Bull. 34, 55– 57. Catovsky, S., Bazzaz, F.A., 2000. The role of resource interactions and seedling regeneration in maintaining a positive feedback in hemlock stands. J. Ecol. 88, 100–112. Clarke, K.R., Gorley, R.N., 2001. PRIMER v5: User Manual/Tutorial. PRIMERE Ltd., Plymouth, UK. Clarke, P.J., Davison, E.A., 2004. Emergence and survival of herbaceous seedlings in temperate grassy woodlands: recruitment limitations and regeneration niche. Austral Ecol. 29, 320–331. Cooper-Ellis, S., Foster, D.R., Carlton, G., Lezberg, A., 1999. Forest response to catastrophic wind: results from an experimental hurricane. Ecology 80, 2683–2696.

Council Officers Executive Committee (COEC), 1992. Handbook on Reference Material for Soil Analysis. Soil and Plant Analysis Council, Inc., Georgia, USA. Dwyer, J.F., Nowak, D.J., Noble, M.H., 2003. Sustaining urban forests. J. Arboric. 29, 49–55. Environment Canada, 2003. A Climatology of Hurricanes for Canada: Improving Our Awareness of the Threat. Hurricane Juan 2003. Retrieved November 2, 2005 from: http://www.atl.ec.gc.ca/weather/hurricane/climatology/ preview e.html. Foster, D.R., 1988a. Species and stand response to catastrophic wind in Central New England, USA. J. Ecol. 76, 135–151. Foster, D.R., 1988b. Disturbance history, community organization and vegetation dynamics of the old-growth Pisgah Forest, south-western New Hampshire, USA. J. Ecol. 76, 105–134. Foster, D.R., Boose, E.R., 1992. Patterns of forest damage from catastrophic wind in Central New England, USA. Ecology 80, 79– 98. Foster, B.L., Tilman, D., 2003. Seed limitation and the regulation of community structure in oak savanna grassland. J. Ecol. 91, 999– 1007. Frelich, L.E., 2002. Forest Dynamics and Disturbance Regimes. Cambridge University Press, Cambridge. Frelich, L.E., Machado, J.L., Reich, P.B., 2003. Fine scale environmental variation and structure of understory plant communities in two old-growth pine forests. J. Ecol. 91, 283–293. Gleason, H.A., Cronquist, A., 1991. Manual of Vascular Plants of Northeastern United States and Adjacent Canada, 2nd ed. New York Botanical Garden, NY, USA. Goldblum, D., 1997. The effect of treefall gaps on understory vegetation in New York State. J. Veg. Sci. 8, 125–137. Greenberg, C.H., McNab, W.H., 1998. Forest disturbance in hurricane-related downbursts in the Appalachian mountains of North Carolina. For. Ecol. Manage. 104, 179–191. Greene, D.F., Zasada, J.C., Sirois, L., Kneeshaw, D., Morin, H., Charron, I., Simard, M.J., 1999. A review of the regeneration dynamics of North American boreal forest tree species. Can. J. For. Res. 29, 824– 839.

122

S. Burley et al. / Landscape and Urban Planning 85 (2008) 111–122

Honnay, O., Hermy, M., Coppin, P., 1999a. Effects of area, age and diversity of forest patches in Belgium on plant species richness, and implications for conservation and reforestation. Biol. Conserv. 87, 73–84. Honnay, O., Endels, P., Vereecken, H., Hermy, M., 1999b. The role of patch area and habitat diversity in explaining native plant species richness in disturbed suburban forest patches in northern Belgium. Divers. Distributions 5, 129–141. Hope, G.D., Prescott, C.E., Blevins, L.L., 2003. Responses of available soil nitrogen and litter decomposition to openings of different sizes in interior Douglas-fir forests in British Columbia. For. Ecol. Manage. 186, 33–46. Johnson, D.W., Curtis, P.S., 2001. Effects of forest management on soil C and N storage: meta analysis. For. Ecol. Manage. 140, 227–238. Jonsson, B.G., 1993. The bryophyte diaspore bank and its role after small-scale disturbance in a boreal forest. J. Veg. Sci. 4, 819–826. Jotcham, J.R., Strong, K.W., Marvin, T.K., 1992. An Ecological Survey of Point Pleasant Park. Marbicon Inc., Maritime Testing Ltd., Halifax, Nova Scotia. Kalamees, R., Zobel, M., 2002. The role of seed bank in gap regeneration in a calcareous grassland community. Ecology 83, 1017–1025. Kitz, J., Castle, G., 1999. Point Pleasant Park An Illustrated History. Pleasant Point Publishing, Halifax, NS. Knoepp, J.D., Swank, W.T., 1997. Forest management effects on surface soil carbon and nitrogen. Soil Sci. Soc. Am. J. 61, 928–935. Kostel-Hughes, F., Young, T.P., Carreiro, M.M., 1998. Forest leaf litter quantity and seedling occurrence along an urban-rural gradient. Urban Ecosys. 2, 263–278. Kottegoda, I., 2001. A vegetation survey and tree regeneration assessment in Point Pleasant Park, Halifax, Nova Scotia. Honour’s Thesis. Environmental Studies Program, Faculty of Science, Saint Mary’s University. Halifax, Nova Scotia. Laiho, R., Prescott, C.E., 2004. Decay and nutrient dynamics of coarse woody debris in northern coniferous forests: a synthesis. Can. J. For. Res. 34, 763–777. Landenberger, R.E., McGraw, J.B., 2004. Seed-bank characteristics in mixedmesophytic forest clearcuts and edges: does “edge effect” extend to the seed bank? Can. J. Bot. 82, 992–1000. Leck, M.A., Parker, V.T., Simpson, R.L. (Eds.), 1989. Ecology of Soil Seed Banks. Academic Press, San Diego, pp. 123–147. Leck, M.A., Sch¨utz, W., 2005. Regeneration of Cyperaceae, with particular reference to seed ecology and seed banks. Perspect. Plant Ecol. Evol. Syst. 7, 95–133. Leckie, S., Vellend, M., Bell, G., Waterway, M.J., Lechowicz, M.J., 2000. The seed bank in an old-growth, temperate deciduous forest. Can. J. Bot. 78, 181–192. MacDougall, J.I., Cann, D.B., 1963. Soil Survey of Halifax County, Nova Scotia. Report No. 13, N.S. Soil Survey, Truro, Nova Scotia. Maco, S.E., McPherson, E.G., 2002. Assessing canopy cover over streets and sidewalks in street tree populations. J. Arboric. 28, 270–276. McClure, J.W., Lee, T.D., Leak, W.B., 2000. Gap capture in northern hardwoods: patterns of establishment and height growth in four species. For. Ecol. Manage. 127, 181–189. McPherson, E.G., Nowak, D., Heisler, G., Grimmond, S., Souch, C., Grant, R., Rowntree, R., 1997. Quantifying urban forest structure, function, and value: the Chicago Urban Forest Climate Project. Urban Ecosys. 1, 49–61. McPherson, E.G., Simpson, J.R., Peper, P.J., Xiao, Q., 1999. Benefit-cost analysis of Modesto’s municipal urban forest. J. Arboric. 25, 235– 248.

Moola, F.M., Vasseur, L., 2004. Recovery of late-seral vascular plants in a chronosequence of post-clearcut forest stands in coastal Nova Scotia, Canada. Plant Ecol. 172, 183–197. Moore, J.M., Wein, R.W., 1977. Viable seed populations by soil depth and potential site recolonization after disturbance. Can. J. Bot. 55, 2408–2412. Mosseler, A., Lynds, J.A., Major, J.E., 2003. Old-growth forests of the Acadian forest region. Environ. Rev. 11 (Suppl. 1), S47–S77. Neily, P., Keys, K., Quigley, E., 2004. Forest Ecosystems of Point Pleasant Park. Ecosystem Management Group. Nova Scotia Department of Natural Resources, Truro, Nova Scotia. Nowak, D.J., 1994. Understanding the structure of urban forests. J. For. 92, 36–41. Qi, M., Scarratt, J.B., 1998. Effect of harvesting method on seed bank dynamics in a boreal mixedwood forest in northwestern Ontario. Can. J. Bot. 76, 872–883. Pakeman, R.J., Small, J.L., 2005. The role of seed bank, seed rain and the timing of disturbance in gap regeneration. J. Veg. Sci. 16, 121–130. Peterson, C.J., Pickett, S.T.A., 1990. Microsite and elevational influences on early forest regeneration after catastrophic windthrow. J. Veg. Sci. 1, 657–662. Preston, C.M., Trofymow, J.A., 2000. Characterization of soil P in coastal forest chronosequences of southern Vancouver Island: effects of climate and harvesting disturbance. Can. J. Soil Sci. 80, 633–647. Roberts, M.R., 2004. Response of the herbaceous layer to natural disturbance in North American forests. Can. J. Bot. 82, 1273–1283. Roland, A.E., Smith, E.C., 1963. The Flora of Nova Scotia. Nova Scotian Institute of Science, Halifax, Nova Scotia. Rudnicky, J.L., McDonnell, M.J., 1989. Forty-eight years of canopy change in a hardwood-hemlock forest in New York City. Bull. Torr. Bot. Club 116, 52–64. Sanders, R.A., 1984. Some determinants of urban forest structure. Urban Ecol. 8, 13–28. Simard, D.G., Fyles, J.W., Par´e, D., Nguyen, T., 2001. Impacts of clearcut harvesting and wildfire on soil nutrient status in the Quebec boreal forest. Can. J. Soil Sci. 81, 229–237. Singleton, R., Gardescu, S., Marks, P.L., Gerber, M.A., 2001. Forest herb colonization of postagricultural forests in central New York State, USA. Ecology 89, 325–338. Stark, K.E., Lundholm, J.T., Larson, D.W., 2003. Relationships between seed banks and spatial heterogeneity of North American alvar vegetation. J. Veg. Sci. 14, 205–212. ter Heerdt, G.N.J., Verweij, G.L., Bekker, R.M., Bakker, J.P., 1996. An improved method for seed-bank analysis: seedling emergence after removing the soil by sieving. Funct. Ecol. 10, 144–151. Turner, K., Lefler, B., Freedman, B., 2005. Plant communities of selected urbanized areas of Halifax, Nova Scotia, Canada. Landscape Urban Plan. 71, 191–206. Weber, M.G., Flannigan, M.D., 1997. Canadian boreal forest ecosystem structure and function in a changing climate: impact on fire regimes. Environ. Rev. 5, 145–166. Williams, N.S.G., Morgan, J.W., McDonnell, M.J., McCarthy, M.A., 2005. Plant traits and local extinctions in natural grasslands along an urban–rural gradient. J. Ecol. 93, 1203–1213. Yanai, R.D., 1998. The effect of whole-tree harvest on phosphorus cycling in a northern hardwood forest. For. Ecol. Manage. 104, 281–295. Zinck, M., 1998. Roland’s Flora of Nova Scotia. Nova Scotia Museum and Nimbus Publishing, Halifax, Nova Scotia.