2010 NORTHEASTERN NATURALIST 17(2):273–284 Effect of Forest Harvest on the Vegetation of an Urban Park Jack T. Tessier* Abstract - Forested parks are an urban oasis, and forest management plays an important part in their maintenance. I used a systematic sampling technique to quantify the vegetation and propagule bank in a small urban park prior to and following a forest management event. The number of non-forest species and the abundance of Toxicodendron radicans (Poison Ivy) increased after forest harvest. Distance from an edge did not affect the change in vegetation. In the propagule bank, loss of basal area in a plot was positively correlated with an increase in the percent of species that were non-forest and non-native. The Sørenson coefficient of community similarity comparing the species composition of the vegetation before and after harvest was 0.769 (out of 1.0), but that for the propagule bank was 0.308. Forest management practices in small urban parks should be designed with extreme caution due to the volatile settings of these refuges. Introduction Forested urban parks are used by local residents for recreation, nature study, and relaxation. Given the importance of urban forests, community residents often take an interest in urban forestry and the general management of their forested parks (Kuhns et al. 2005, Ricard and McDonough 2007, Shin et al. 2005, Treiman and Gartner 2004, Zhang et al. 2007a). Forested urban parks are also important sinks for atmospheric carbon and air pollutants (Escobedo et al. 2008, Kielbaso 1990, Myeong et al. 2006, Nowak et al. 2007) and provide habitat for wildlife (Atchison and Rodewald 2006, Leston and Rodewald 2006, Morrison and Chapman 2005). These multiple qualities exemplify the critical importance of forested urban parks for people, wildlife, and ecosystem function. Forested urban parks, however, can be fragile ecosystems as a result of their landscape context. The history of a particular location and its adjacent lands affect the ecological patterns that are present (Agrawal et al. 2007, Bertin et al. 2006, Jim and Liu 2001), and these impacts may last for centuries (Foster and Aber 2004, Goodale and Aber 2001, Goodale et al. 2000). Development continually threatens forested urban parks (Zérah 2007), and increased urbanization alters community composition and biogeochemical cycling (Groffman et al. 2006, 2007). Forested urban parks tend to differ compositionally from rural forests (Ochimaru and Fukuda 2007), and are particularly susceptible to the influx of invasive species (Heckmann et al. 2008, Szlavecz et al. 2006) with extensive effort required to remove them *Division of Liberal Arts and Sciences, SUNY Delhi, 722 Evenden Tower, Delhi, NY, 13753; tessiejt@delhi.edu. 274 Northeastern Naturalist Vol. 17, No. 2 once established (Vidra et al. 2007). Encroachment of invasive species is often associated with tree harvest (Brown and Gurevitch, 2004, Zettler et al. 2004), and the abundance of such species increases with proximity to edges (Guirado et al. 2006, Pauchard and Alaback 2004) and as forest patch size decreases (Guirado et al. 2006). The critical challenges that invasive species pose in forest management (Chorensky et al. 2005, Lovett et al. 2006) may therefore be exacerbated in forested urban parks. Invasive species are of particular concern due to their ability to alter community composition (Csuzdi and Szlávecz 2003, Prati and Bossdorf 2004), influence biogeochemical cycling (Ehrenfeld et al. 2001, Yorks et al. 2003), modify trophic interactions (Davis 2003), and outcompete native vegetation (Daehler 2003). Maintaining the ecological integrity of forested urban parks depends on proper management (Heckmann et al. 2008), but making forested urban parks ecologically sustainable is a significant challenge (Elmendorf et al. 2003). Urban forest management is an integrated discipline, requiring an understanding of the relevance of factors including urbanization, invasive species, recreation needs, and ecosystem services (Anderson 2003, Dwyer et al. 2003). Because of these myriad factors, cohesive management plans are needed (Zhang et al. 2007b) that are site-specific (Dwyer et al. 2003). Having an individual or department in charge of forested urban park management within a municipality is critical (Lewis and Boulahanis 2008), but many cities lack the expertise (Treiman and Gartner 2004), funds (Elmendorf et al. 2003, Grado et al. 2006, Kielbaso 1990, Lewis and Boulahanis 2008), and awareness (Grado et al. 2006) to develop and maintain these challenging management plans. The provision of data documenting the response of the vegetation in forested urban parks to management activities is therefore important in helping with the development of suitable approaches to the management of these valuable and imperiled ecosystems. In the fall of 2004, a canopy-thinning forest harvest was conducted in Stanley Quarter Park, in New Britain, CT. The objective of this study was to assess immediate changes in the composition of the vegetation and propagule bank associated with this harvest operation. The goal was to determine how this harvest would alter the vegetation in regard to non-native (those not native to the northeast United States as described in Gleason and Cronquist [1991]) and non-forest (those not associated with closed-canopy forests as described in Gleason and Cronquist [1991]) plant species. Methods Study site Stanley Quarter Park (41°41′30″N, 72°46′57″W; approximately 26 ha including ball fields, pond, and forested areas) is a forested park surrounded by residential urban neighborhoods in central Connecticut. The presence of wolf trees (larger trees than those surrounding them, with evidence of large lower branches indicative of previous high-light conditions) suggests previous pasturing of the area, but its recent history (≈100–200 yrs) has been that 2010 J.T. Tessier 275 of a forested urban park. The dominant tree species (in order of decreasing basal area as quantified in the initial vegetation sampling; see below) are Quercus rubra L. (Northern Red Oak), Acer saccharum Marshall (Sugar Maple), Liriodendron tulipifera L. (Tulip Tree), and Carya cordiformis (Wangenh.) K. Koch (Bitternut Hickory). Data collection I sampled the vegetation of Stanley Quarter Park using a systematic sampling scheme to assess the plant community of the park. The starting point of the first transect was placed to avoid edge effect from trails and roads. Three transects were placed 20 m apart, and square-shaped 100-m² study plots (see details by vegetation layer below) were located 50 m apart (from plot center to plot center) along each transect. To accommodate the shape of the forested portion of the Park, two of the transects had four plots and one transect had two plots. Therefore, ten plots were established in an approximately 7-ha area of the space harvested in the forestry operation. While this plot size, spatial spread, and relatively small sample size (10) led to high error rates (see Results), it was necessary for representative sampling of the whole area of interest, while preventing the re-sampling of individual canopy trees in adjacent plots. Because other areas of the Park included Pinus resinosa Aiton (Red Pine) plantations and a former ski slope; in an effort to maximize qualitative similarity and reduce background variation among plots, these areas were not sampled. Also, because all of the forested area of the Park was harvested, there was not an available unharvested control area. Vegetation sampling was conducted during the summers of 2004 (prior to the harvest) and 2006 (two years after the harvest). In 2005, overstory vegetation (trees >10 cm diameter) alone was sampled to assess changes as a result of the harvest. Overstory vegetation was sampled in the 100-m² plots to collect data on species composition, density, and diameter of individuals. Sapling (trees <10 cm diameter) species composition and diameter of individuals were determined in 25-m² plots in the northwest corner of each overstory plot. Species-specific ground cover of understory vegetation (vascular plants below 1 m in height) was measured in 1-m² plots in each corner of the overstory plots (total of 40 understory plots). Data from understory plots within a single overstory plot were averaged by species for data analysis. A presence-absence census of all vascular species in each overstory plot (including saplings and understory vegetation) was conducted to make a record of species that may have been present in the overstory plots but not sampled in the smaller plots. Finally, because edges can serve as sources of propagules, the distance from the plot center to the nearest edge (walking trail, skid road, or city street) was measured to the nearest cm. I conducted a propagule-bank study to assess the species composition of the propagules in the soil of Stanley Quarter Park. From the center of each overstory plot, 2000 cm³ (20 cm on a side and 5 cm deep) of soil were removed, in July of 2004 and 2006, and spread over washed sand in plastic trays in a greenhouse. Soil for the 2006 propagule-bank study was taken from 276 Northeastern Naturalist Vol. 17, No. 2 an area adjacent to that taken for the 2004 propagule-bank study. A control tray of only washed sand was used to screen for greenhouse weeds (though none grew in the control tray). Propagule bank samples were monitored for germinants weekly until no further germinants had appeared for two months (March of the following year for each year of study). Germinants were allowed to grow until identification was possible and then they were removed. Neither propagule stratification nor scarification was employed. Data from the propagule bank were accompanied by data from the vegetation sampling, which included seeds that germinated in background ecosystem conditions. Therefore, propagules that germinated as a result of background stratification (seasonal thermal changes) and scarification (ingestion and defecation) were already documented in the vegetation sampling. Overstory harvest The harvest of canopy trees (trees above the sapling layer) occurred during the fall of 2004. A forest management company was hired by the City of New Britain Department of Parks and Recreation to employ a single-treeselection system thinning to reduce basal area in the stand by 20 to 30%. The stand was marked by a Connecticut Certified Forester and was subsequently harvested. The pre-harvest basal area of the stand was 38.27 ± 8.18 m²/ha (mean ± standard error), based on the initial sampling. Data analyses I conducted all statistical analyses using Minitab version 15 (Minitab Inc., State College, PA) and α = 0.05. The within-plot change (pre- vs. postharvest) in the number of non-forest and non-native species in the existing plant community vs. diameter loss in canopy trees (total within-plot diameter of harvested trees) and distance to the nearest edge were individually compared using linear regression. The within-plot change in the percent of species that were non-forest and non-native in the propagule bank vs. diameter loss in canopy trees was also compared using linear regression. A paired t-test (using plots as replicates) was used to compare the cover of non-forest species, non-native species, and Toxicodendron radicans L. Kuntze (Poison Ivy) before and after the harvest. I calculated the Sørenson’s coefficient of community similarity (Brower et al. 1990) to compare the pre- and post-harvest plant community (in all strata) and propagule bank. Lastly, I tabulated species that were detected in the Park and propagule bank only before the harvest and those that were detected in the Park and propagule bank only after the harvest, making note of non-forest and non-native species in each case. Results Based on data from the vegetation sampling, the post-harvest basal area was 29.77 ± 7.47 m²/ha (± s.e.), a reduction of 8.51 ± 4.01 m²/ha or 22.03 ± 10.62%, in alignment with the goal of a 20% to 30% reduction. High values of standard error are the result of the inherently small 2010 J.T. Tessier 277 sample and plot size in studying this small forested park (see Methods). Mean tree density per hectare was reduced from 310 ± 37.86 before the harvest to 260 ± 42.69 after the harvest. An average of 0.5 ± 0.22 trees was removed from the sample plots. The species of canopy trees that were harvested from the plots were Northern Red Oak and Sugar Maple (though not all of the individuals of these species were harvested). These species are valuable timber species, and the individuals harvested included the second largest for each species within the plots (diameter at breast height [1.5 m off the ground] of 62.8 cm for Northern Red Oak and 45.7 cm for Sugar Maple). Canopy tree species not removed from sample plots were Tulip Tree, Ulmus americana L. (American Elm), Bitternut Hickory, Carya ovata (Miller) K. Koch (Shagbark Hickory), Ostrya virginiana (Miller) K. Koch (Hop-hornbeam), Quercus alba L. (White Oak), Acer rubrum L. (Red Maple), and Fraxinus americana L. (White Ash). The increase in the number of non-forest species following the harvest was positively correlated with diameter loss in sample plots (Table 1). Diameter loss as a result of the harvest, however, was not significantly related to the increase in the number of non-native species within a plot (Table 1). Increasing diameter removal, therefore, was correlated with an increase in the presence of non-forest species but not non-native species. There was no significant relationship between the change in the number of non-forest and non-native species in the vegetation relative to the distance of the plot from the closest walking trail, skid road, or city street (Table 1). The average distance to the nearest edge was 15.37 ± 9.37 m and ranged from 0 to 96.5 m. Nine of the ten plots were closer to a skid road than any other edge (including three plots whose center was crossed by a skid road), and the other plot was closest to a city street and was the most distant plot from any other edge. The distance to a potential source of propagules was therefore not important to the ability of non-forest and non-native species to establish in the short-term post-harvest plant community. Table 1. Regression relationships associated with a forest harvest event in a small forested urban park in New Britain, CT. Diameter loss is the summed diameter of all species removed from a plot. An edge in this study is a city street or a skid road. X variable Y variable Line equation r² P Diameter loss Change in number of y = 0.0668x + 0.055 0.791 0.001 non-forest species in vegetation Diameter loss Change in number of y = 0.0058x + 0.374 0.029 0.640 non-native species in vegetation Distance to edge Change in number of y = -0.0292x + 1.92 0.138 0.291 non-forest species in vegetation Distance to edge Change in number of y = 0.0003x + 0.49 0.000 0.980 non-native species in vegetation Diameter loss Change in percent of species that y = 0.011x + 0.063 0.511 0.020 were non-forest in propagule bank Diameter loss Change in percent of species that y = 0.285x + 0.59 0.580 0.011 were non-native in propagule bank 278 Northeastern Naturalist Vol. 17, No. 2 Within the propagule bank, the percentage of species that were non-forest or non-native both increased with increasing tree diameter loss within a plot (Table 1). Because the propagule-bank samples from pre- and post-harvest were treated identically in the greenhouse, the difference in the species found in the propagule bank represented those species that were added during the harvest period. Therefore, the species with potential to grow in Stanley Quarter Park following the harvest were increasingly non-forest and non-native species as diameter loss increased. While the percent cover of non-native species did not significantly increase after the harvest, the percent cover of Poison Ivy did increase significantly, and the increase in percent cover of non-forest species was marginally significant (Fig. 1; based on cover from understory samples: non-native P = 0.69, Poison Ivy P = 0.05, non-forest P = 0.07). While the actual cover of non-forest and non-native species remained small despite their increasing presence with higher diameter loss, the cover of Poison Ivy nearly doubled after the harvest and was distinctly greater than that of nonforest and non-native species before and after the harvest. The Sørenson coefficient of community similarity comparing the preharvest and post-harvest plant community in all strata was 0.769, and that for the propagule bank was 0.308. This result equates to roughly a 23% change in the species composition of the forest community and a 69% change in the species composition of the propagule bank as a result of the harvest. The tree Figure 1. Mean cover of nonforest species, n o n - n a t i v e species, and Poison Ivy before and after a forest harvest event in a forested urban park in New Britain, CT. P-values are based on dependent measures t-tests c o m p a r i n g cover before and after the harvest. Error bars represent one standard error above the mean. 2010 J.T. Tessier 279 harvest did not factor directly into these changes because none of the tree species removed in the harvest was completely lost from the collection of sample plots. Several species disappeared from the forest and the propagule bank as a result of the harvest (Tables 2 and 3). All but one of the eight species that were absent following the harvest were native, forest species. The Cornus florida (Flowering Dogwood) and Sorbus americana (American Mountain Ash) that were lost were present as seedlings in the initial sample and were probably lost due to direct damage in the harvest. Several other species appeared in the forest community and the propagule bank after the harvest (Tables 2 and 3). Of the 14 species detected in the forest community after the harvest, two were non-forest species, two were non-native, and five were both non-forest and non-native (Table 2). Of the six species detected in the propagule bank after the harvest, one was a non-forest species and three Table 3. Plant species present in the propagule bank of a forested urban park in New Britain, CT only before or only after a forest harvest event. * indicates a species that does not typically grow in closed forest settings (non-forest species). ^ indicates a non-native species. Species only present before harvest Species only present after harvest Aster acuminatus Michx. *^Celastrus orbiculatus Thunb. (Oriental Bittersweet) (Whorled Aster) *^Lonicera japonica Thunb. Dennstaedtia punctilobula (Michx.) Moore (Japanese Honeysuckle) (Hay-scented Fern) *Phytolacca americana L. (American Pokeweed) Thelypteris noveboracensis (L.) Nieuwl. (New York Fern) *^Tussilago farfara L. (Coltsfoot) *^Veronica officinalis L. (Common Speedwell) Table 2. Plant species present in the vegetation of a forested urban park in New Britain, CT only before or only after a forest harvest event. * indicates a species that does not typically grow in closed forest settings (non-forest species). ^ indicates a non-native species. Species only present before harvest Species only present after harvest Cornus florida L. (Flowering Dogwood) ^Acer platanoides L. (Norway Maple) Onoclea sensibilis L. (Sensitive Fern) Athyrium filix-femina (L.) Roth (Lady Fern) Polystichum acrostichoides (Michx.) *^Cardamine pratensis L. (Cuckoo Flower) Schott. (Christmas Fern) Prenanthes alba L. (Wild White Lettuce) Dryopteris intermedia (Muhl.) A. Gray (Common Wood Fern) Sorbus americana Marshall ^Euonymus alatus (Thunb.) Siebold (American Mountain Ash) (Winged Burning Bush) Thelypteris noveboracensis (L.) Nieuwl. Mitchella repens L. (Partridgeberry) (New York Fern) *^Oxalis stricta L. (Common Yellow Wood Sorrel) *Phytolacca americana L. (American Pokeweed) Prunus pennsylvanica L.f. (Pin Cherry) *^Rosa multiflora Thunb. (Multiflora Rose) *Rubus occidentalis L. (Black Raspberry) *^Solanum dulcamara L. (Bittersweet) Solidago caesia L. (Blue-stemmed Goldenrod) *^Verbascum thapsus L. (Common Mullein) 280 Northeastern Naturalist Vol. 17, No. 2 were both non-forest and non-native (Table 3). Collectively, the change in forest and propagule bank composition represented a decrease in the number of native forest species and an increase in the number of non-forest and/or non-native species. Discussion Forest harvest in Stanley Quarter Park generally promoted the presence of non-forest species, non-native species, and Poison Ivy (Tables 1–3, Fig. 1). Of the eleven measures of the plant community and propagule bank, eight (number of non-forest species in the vegetation, percent of both non-forest and non-native species in the propagule bank, cover of Poison ivy, and both loss and gain of both non-forest and non-native species in the vegetation and propagule bank) indicate an immediate shift toward more non-forest species, non-native species, and Poison Ivy. The influx of A. platanoides (Norway Maple), Euonymus alatus (Winged Burning Bush), and Rosa multiflora (Multiflora Rose) are of the greatest concern among the new non-native species because they are invasive (Invasive Plant Atlas of New England 2009), and invasive species are known to pose a threat to ecosystem structure and function (e.g., Csuzdi and Szlávecz 2003, Daehler 2003, Ehrenfeld et al. 2001). These localized results corroborate previous findings of an increase in non-forest and non-native species in association with disturbance brought on by forest harvest (Brown and Gurevitch 2004, Guirado et al. 2006, Zettler et al. 2004). Forested urban parks, because of their small size and proximity to roads and other edges, should therefore be considered vulnerable forests in terms of encroachment of species not representative of the region’s forests (Guirado et al. 2006, Heckmann et al. 2008, Pauchard and Alaback 2004, Szlavecz et al. 2006). In fact, this forested urban park is so small, that the distance from a plot to an edge (an anticipated source of propagules) did not affect the ability of non-forest and non-native species to establish in the post-harvest plant community (Table 1). Because historical events can have long lasting impacts on a plant community (Agrawal et al. 2007, Bertin et al. 2006, Foster and Aber 2004, Goodale and Aber 2001), the ecological integrity of urban forests should be given utmost consideration as managers seek to stem the threat posed to forests by invasive species (Chorensky et al. 2005, Lovett et al. 2006). In the longer term, some of these changes in the vegetation and propagule bank may prove to be transient. Non-forest and non-native species likely came into the forest from the neighboring non-forest ecosystem (Brothers and Spingarn 1992, Cadenasso and Pickett 2001). The opening of the canopy has likely encouraged their success. As the canopy re-closes with the growth of remnant trees, some of the non-forest and non-native species may be extirpated. Other species in these categories, particularly the invasive species, may well persist and become a long-term part of the local flora. Additionally, T. radicans can persist under a closed canopy as evidenced by its presence 2010 J.T. Tessier 281 prior to the harvest. Additional samplings are planned in order to document the longer term (10+ years) impact of the harvest. The threat posed by increases in non-forest species, non-native species, and Poison Ivy can have far-reaching consequences. Local residents value urban forests and their management (Kuhns et al. 2005, Ricard and McDonough 2007, Shin et al. 2005, Treiman and Gartner 2004, Zhang et al. 2007a). Preharvest uses of Stanley Quarter Park included cross-country races attended by secondary schools and colleges, bird watching and other nature observation by the local residents, and visitation by elementary and middle school groups led by college faculty. As the abundance of Poison Ivy increases, local residents and schools may be less likely to venture into the forest for fear of contact with the plant and its associated dermatitis. This fear may further discourage youth from seeking outdoor exercise and entertainment, a trend that has caused great concern in society (Louv 2005, Pergams and Zaradic 2008). Additionally, an increase in non-forest and non-native species in forested urban parks could cause more urban residents to think of forests as “weedy” places without any of the grandeur placed upon them by conservationists, foresters, and ecologists. Thus, the result could be less public support for such parks and less appreciation of their importance for habitat (Atchison and Rodewald 2006, Leston and Rodewald 2006, Morrison and Chapman 2005), air pollution abatement (Escobedo et al. 2008, Nowak et al. 2007), and carbon uptake (Kielbaso 1990, Myeong et al. 2006, Nowak et al. 2007). Infrastructural decay in cities themselves results in increased health problems and fears of crime within the area (Kruger 2008), effectively adding to the urban decay. Degradation in the status of and opportunities provided by forested urban parks in the eyes of urban residents may similarly erode urban dwellers’ support for the conservation of forests and other wild habitats on a global basis (Kareiva 2008), thus reducing humanity’s capacity to conserve important habitats. Conclusion In managing forested urban parks of the northeast United States, managers should carefully consider the ecological context of these vulnerable forests (Agrawal et al. 2007, Heckmann et al. 2008) and take appropriate action in order to limit the encroachment of non-forest and non-native species and the growth of Poison Ivy (Tables 1–3, Fig.1). Based on these data, managers should limit forestry operations in small urban parks to hazard tree removal and other small operations that are necessary to protect the human visitors and the ecological integrity of the park (including habitat maintenance for biodiversity). 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