Regular issues
Monographs
Special Issues



Southeastern Naturalist
    SENA Home
    Range and Scope
    Board of Editors
    Staff
    Editorial Workflow
    Publication Charges
    Subscriptions

Other EH Journals
    Northeastern Naturalist
    Caribbean Naturalist
    Urban Naturalist
    Eastern Paleontologist
    Eastern Biologist
    Journal of the North Atlantic

EH Natural History Home

Diversity and Community Similarity of Arthropods in Response to the Restoration of Former Pine Plantations
John C. Burkhalter, Daniel C. Moon, and Anthony M. Rossi

Southeastern Naturalist, Volume 12, Issue 1 (2013): 121–136

Full-text pdf (Accessible only to subscribers.To subscribe click here.)

 

Site by Bennett Web & Design Co.
2013 SOUTHEASTERN NATURALIST 12(1):121–136 Diversity and Community Similarity of Arthropods in Response to the Restoration of Former Pine Plantations John C. Burkhalter1, Daniel C. Moon1, and Anthony M. Rossi1,* Abstract - Ecological restoration is becoming an increasingly important tool in humanity’s attempt to manage, conserve, and repair the world’s ecosystems. In the current study, the objective was to compare the effects of two restoration methods on arthropod biodiversity and community composition in two former pine plantations; these treatments included both intensive restoration effort (= cleared) and moderate restoration effort (= thinned). For the cleared treatment, vegetation was clear-cut to the soil surface, and all vegetation was removed from the plots, while the thinned treatment consisted of reducing the Pinus elliotii (Slash Pine) density to that of a native ecosystem and removing of all exotic plants from the plots as well. Arthropods were sampled by employing pitfall traps, sticky traps, and sweep netting and identified to family and morphospecies; species richness, diversity, and community similarity were compared between treatments and sites. Experimental treatments quickly reached or exceeded arthropod diversity and richness of an unmanipulated control treatment; however, the two sites produced non-overlapping ordination plots, suggesting that the diversity of the two sites are either compositionally different (alpha diversity) or community assemblage is incomplete and overall regional (beta) diversity has not reached an equilibrium across sites. Additional long-term data should reveal if these plots are proceeding along different successional trajectories in terms of community species composition, or whether treatments, while having similar richness, support different communities because the three types of plots used in this study (control, thinned, and cleared) represent various successional stages which affect arthropod species identity, but not overall richness. Introduction Human activities are impairing the normal operation of ecosystem services and functions on a large scale (Slobodkin 2000). This impairment has led to an increase in the frequency and importance of attempts to restore degraded ecosystems. The restoration of biodiversity in an environment can be a critical aspect of improving ecosystem processes and functions, and the regeneration of community structure is a vital part of the re-establishment process (Redi et al. 2005, Zerbe and Kreyer 2006). For instance, managed forests, one type of anthropogenically disturbed ecosystem, are often viewed negatively from a conservation standpoint largely because they exhibit reduced biodiversity, and they may develop altered community composition relative to natural forests (Freedman et al. 1996, Friend 1982). Managed pine plantations are now the most extensive ecosystem in North Florida, comprising approximately 70% of the forested landscape (Clark et al. 2004). As researchers and land managers strive to understand and ameliorate the effects of anthropogenic disturbances on ecosystem function and biodiversity, 1Department of Biology, University of North Florida, 1UNF Drive, Jacksonville, FL 32224. *Corresponding author - arossi@unf.edu. 122 Southeastern Naturalist Vol. 12, No. 1 there is the need to make informed decisions in a timely and cost-effective manner. This study seeks to provide land managers with a simultaneous comparison of multiple restoration techniques, using biologically relevant indicators of ecosystem function in the form of biodiversity and community composition in a former pine plantation. To quantify biodiversity, a number of different groups of organisms have been monitored including microbes, plants, and animals (Gaines et al. 2007, Kulmatiski 2011, York et al. 2011). Traditionally, restoration studies have focused on establishing plant communities (Wheater et al. 2000), but it is also important to examine effects on various fauna to determine the overall success of restoration (Longcore 2003). Restoration of natural communities relies on the assumption that with the reestablishment of natural vegetation, reestablishment of the fauna normally associated with the habitat will follow (Gratton and Denno 2005). However, little information is available on how arthropod assemblages are affected by the reintroduction of native flora (Gratton and Denno 2005). It is assumed that increased plant diversity is likely to result in higher herbivore diversity (Siemann et al. 1999), but the reality may not be that simple. Insect life-history strategies are closely attuned to vegetational characteristics of the habitat (Brown 1985), but there could be a number of factors controlling arthropod community recovery, e.g., plant-community structural complexity and life-history attributes of the various insects of the communities (Lawton 1978, Siemann et al. 1998, Steffan- Dewenter and Tscharntke 1997). Any one or a combination of these factors could hinder an increase in arthropod diversity in spite of an increase in plant diversity in a given habitat. Assessing the progress of restoration projects with arthropods has many advantages (Finnamore 1996, Kremen et al. 1993). For instance, the short generation times of most arthropods make them ideal to track year-to-year changes at a site, while their small size and relatively large population sizes make them efficient indicators of subtle yet important variations that may influence the quality of a habitat (Longcore 2003). Arthropod species diversity is a useful metric of restoration success because of the important roles they fill, such as herbivores (or plant regulators), pollinators, detritivores, predators, and parasites; they also serve as important prey sources for reptiles, birds, and mammals (Gardner-Gee and Beggs 2010, Price 1984, Ruzicka et al. 2010). Insects play a crucial role in the decomposition of leaf litter, fallen trees, and standing deadwood, making insects critical for nutrient cycling. Therefore, reduction in insect species abundance and richness has important functional implications. (Kattan et al. 2006, Spence 2001). In addition, insects can alter relationships between plant diversity and ecosystem processes by altering relative abundances of plant species (Mulder et al. 1999), particularly through pollination services. A critical review by Maleque et al. (2009) provides a more exhaustive and thorough explanation of the benefits of using arthropods as indicators of restoration su ccess. Although species richness is an important attribute of an ecosystem, environmental variation can produce disparate community composition, even when species richness is similar. Community similarity measures can indicate what 2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 123 effect restoration has on the community on a broader scale. For instance, local (alpha) diversity, in which habitats are more likely to be homogenous, is typically more similar than regional (beta) diversity. As a result, alpha and gamma diversity may vary substantially between these scales (Ricklefs and Miller 2000). Thus, measures of community similarity provide a metric by which we can determine how resources are being utilized between sites based upon species composition and their relative abundances in various communities or treatment groups. Tree plantations often provide habitat characterized by a closed canopy and tree densities that are greatly elevated above natural levels; these high densities may affect underlying community composition and/or resource quality (Kattan et al. 2006). For example, Zhu et al. (2007) found that plantation densities of Pinus resinosa Aiton (Red Pine), which range from 89–332/ha, reduced average tree ring width by approximately 50% in comparison to natural stands, thereby reducing the growth of trees within plantations. In canopies of lower density, the environment may be made more hospitable for understory species. Mixed species stands are also likely to use nutrients more efficiently compared to pure stands such as pine plantations because of differences among species in a number of factors such as mycorrhizal associations, shade-tolerance, growth rate, form, nutrient demands, and abilities to fix nitrogen. In addition, combinations of species that differ in height, growth form, shade-tolerance, and phenology are also likely to increase site productivity (Hartley 2002). A common restoration method utilized to restore higher levels of biodiversity in managed forest ecosystems is tree density reduction (= thinning). Thinning has been shown to promote a more balanced stand structure where there is an excess of smaller trees (Edminster and Olson 1996). Intermittent thinning can also preserve tree and stand vigor while still maintaining structure (Edminster and Olson 1996). A study by Feeney et al. (1998) demonstrated that thinning produced improved resource uptake, growth, and insect resistance in stands of Pinus ponderosa Lawson (Ponderosa Pine). Thinning has also been shown to have positive effects for species that benefit from burning such as Pinus elliotii Englem (Slash Pine; Wilson and Watts 1999) and may be a viable alternative when burning is not an option. More intensive methods of restoration are sometimes used, ones in which silviculturists or researchers do not simply remove certain species or decrease density within the environment, but instead conduct an extensive removal of all vegetation to restore ecosystem processes (Hobbs 2007). This treatment may be especially useful for restoring commercially viable pine plantations because it returns the site to an earlier successional stage and allows the community to regenerate via secondary succession. The path of succession will be influenced by a number of factors including seed-bank sources, as well as seed dispersal from neighboring areas and competitive interactions among and between native and invasive plants, resulting in ecosystem dynamics that are extremely complex (Wallington et al. 2005). Lastly, the least expensive method of “restoration” is passive in which researchers or land managers do nothing to the site and allow existing successional 124 Southeastern Naturalist Vol. 12, No. 1 processes to occur. While this method of restoration is preferred by some because it is both inexpensive and easy, it has some disadvantages (Morrison and Lindell 2011). Passive restoration often takes a very long time, and the outcome is relatively uncertain due to the fact that it is not directed; thus, the resulting community may not return to its previous successional trajectory due to factors such as dispersal limitation (Battaglia et al. 2008). Although factors such as dispersal could also be an issue with active restoration, the ability of researchers to “disperse” individuals via intentional planting, could obviate these issues. The precise objective of the current study was to compare two active techniques and one passive technique of restoring a former Slash Pine plantation, and the resultant effects on abundance, diversity, and community similarity of arthropod assemblages. As stated previously, managed ecosystems such as pine plantations are often lower in diversity than unmanaged forests, and higher diversity within the treated areas could indicate levels of diversity closer to a more natural forest. Study Sites Slash Pine is native to the southeastern US, but its density on former pine plantations is significantly higher than that in undisturbed ecosystems. In native ecosystems of North Florida, densities of Slash Pines in pure stands are typically 2–3 trees per 100 m2, but in the managed ecosystems of a pine plantation densities can average 20 trees per 100 m2 (J.C. Burkhalter, unpubl. data). The current study utilized two former pine plantations, McGirt’s Creek (30.26107°N, 81.81843°W) in southwestern Duval County, FL and Tiger Point (30.5007941°N, 81.4937016°W) in northeastern Duval County, FL, which have been acquired for conservation by the city of Jacksonville, FL. Both sites were heavily planted with a Slash Pine density varying from 20 to 50 trees per 100 m2, and both sites had low herbaceous growth due to large amounts of leaf litter (for a more thorough description, see Rossi et al. 2011). McGirt’s Creek is embedded in a mosaic of mostly residential developments, and Tiger Point is surrounded primarily by a more or less contiguous landscape of managed or formerly managed pine plantations. Methods We employed two experimental restoration treatments per site in this study in addition to controls. One experimental treatment entailed removing all exotic plant species and thinning the pine trees that previously made up a large amount of the vegetation in the experimental plots. Each Slash Pine tree within a plot was assigned a number, and then trees were randomly removed. Trees were removed until stand density was reduced to 2–3 trees per 100 m2, with all remaining trees being of similar size and condition so as to not introduce any confounding effects of tree size. In the second experimental treatment, all vegetation was clear-cut manually with chainsaws and loppers and then removed from the plot. Although all vegetation was removed to the surface of the substrate, no below-ground biomass was removed from the plots. Control plots were not manipulated and contained a high density (approximately 20 trees/100 m2) of Slash Pine, as well 2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 125 as native oaks and other small vegetation. All plots were 10 m x 10 m. Few, if any, undisturbed reference sites remain in Florida and none near the study sites, and thus we cannot directly compare the diversity and community similarity of our restored sites to a pristine reference site. Due to the relatively homogenous biotic and abiotic conditions within the study plots prior to restoration, it is assumed that there were no differences in either invertebrate diversity or community similarity. For each treatment group as well as the controls, there were five replicate plots at McGirt’s Creek (15 total plots) and three replicates of each at Tiger Point (9 total plots). All experimental manipulation was completed in February of 2006, and afterwards plots underwent natural succession; invertebrate biodiversity was compared approximately two years after the study was initiated. All arthropod sampling was conducted by the same person throughout the entirety of the monitoring phase of the experiment. For both the pitfall traps and the sticky traps, one trap was used per plot. Sampling periods lasted for a month in the spring of 2008 and 2009, alternating between the two restoration sites each week, allowing for two weeks of sampling at each site per sampling period. During the sampling period at each site, collection of the pitfall traps occurred every other day, while sweep sampling and collection of the sticky traps was performed once per week. Pit fall traps Ground-dwelling insects were captured using pitfall traps constructed from small buckets that were approximately 15.2 cm in both diameter and depth. Pitfall trap covers, constructed from pieces of plywood approximately 1.3 cm thick, were used to exclude debris from the traps. The covers were approximately 20 cm x 20 cm and were supported by two short blocks of 2.5-cm-thick wood that had been screwed to the trap cover surface on opposing corners using wood screws. Pitfall traps were placed in holes that were dug as close as possible to the center of the plot with the lip of the trap being level with the surrounding ground, with one pitfall trap per plot. Sticky traps Flying insects were sampled using commercially prepared yellow sticky traps (Sticky Strips Insect Traps, Olson Products) measuring 7.6 cm x 12.7 cm. Sticky traps were hung from 1.3-cm-diameter PVC piping via a long binder clip (2.5 cm) that was attached to the PVC piping using 1.3-cm screws approximately 2.5 cm from the top of the PVC piping. The PVC piping was hammered into the ground to a depth of 0.30 m, leaving approximately 1 m of piping exposed. The sticky trap apparatus was placed approximately 2.5 cm to the right (if standing with your back to the access road) of the pitfall trap in every sampling plot, and the yellow sticky trap was oriented in the same direction (i.e., towards the access road) on every trap. Sweep sampling Sweep netting was conducted by starting in one corner of the experimental plot and walking diagonally through the center of the plot to the opposing corner 126 Southeastern Naturalist Vol. 12, No. 1 once. A fine mesh net was swept in a horizontal motion at chest height while walking through the plot so as to capture any arthropods that might be resting upon vegetation. Sweeps were made in each plot, and the contents were emptied into vials (large insects such as butterflies were removed prior to collection in vials). All samples were returned to the lab for identification. Statistical analyses In addition to mean species richness, diversity of treatment and control plots were assessed using the Shannon index of diversity (H'). (See a complete species list in Appendix 1). Furthermore, a community similarity value for each pair-wise comparison of the different treatment groups and sites was measured using a Bray-Curtis similarity matrix. Due to the fact that a large amount of data was generated by the Bray-Curtis similarity matrix, the mean similarity value of each cross treatment/site comparison (Table 1) was used to construct a non-metric MDS ordination plot, so as to provide a concise 2-D representation of the community assemblage groupings. To analyze for differences between treatment groups in terms of diversity, a two-way ANOVA was utilized using both treatment and site as the fixed factors. To meet homogeneity assumptions of ANOVA, diversity data for the spring of 2008 was ℮(n)-transformed for the Shannon index. ANOVA was conducted using SPSS v.15, and the Bray-Curtis and the NMDS ordination analyses were conducted using PRIMER v. 5. Results For both sites and survey years, mean species richness was not significantly higher in the experimental plots compared to control plots. For each site, mean species richness in cleared and thinned plots were 36% and 15% higher Table 1. Average Bray-Curtis index values for cross treatment comparisons at McGirt’s Creek and Tiger Point, Spring 2008. Con = Control, Clr = Clear, Thin = Thinned, Tig = Tiger Point, and McG = McGirt’s Creek. Sites Con McG Clr McG Thin McG Con Tig Clr Tig 2008 Con McG Clr McG 46.236560 Thin McG 45.317220 49.09561 Con Tig 22.477060 28.66109 24.930130 Clr Tig 15.436490 21.21960 14.075170 41.84704 Thin Tig 8.826945 13.84275 9.974205 38.67925 69.50629 2009 Con McG Clr McG 9.523809524 Thin McG 28.571428570 10.344827590 Con Tig 5.504587156 21.656050960 10.19108280 Clr Tig 18.487394960 7.185628743 16.76646707 0.0 Thin Tig 14.516129030 5.813953488 11.62790698 6.451612903 6.060606061 2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 127 than control plots, respectively (Figs. 1, 2) (F1,24 = 0.062; P α=0.05 = 0.807 and F1,24= 2.379; P α=0.05 = 0.140, 2008 and 2009, respectively). Additionally, there was no apparent effect on species richness due to treatments, although both Figure 1. Mean species richness (± 1 SE) across all treatment groups at McGirt’s Creek, Spring 2008 and 2009. Figure 2. Mean species richness (±1 SE) across all treatment groups at Tiger Point, Spring 2008 and 2009. 128 Southeastern Naturalist Vol. 12, No. 1 cleared and thinned plots equaled or exceeded the number of species in control plots after two years; once again these differences were not statistically significant (F2,24 = 0.017 and Pα =0.05 = 0.983, and F2,24, = 0.884 and Pα =0.05 = Figure 3. Mean Shannon diversity value (±1 SE) across all treatment groups at McGirt’s Creek, Spring 2008 and 2009. Figure 4. Mean Shannon diversity value (±1 SE) across all treatment groups at Tiger Creek, Spring 2008 and 2009. 2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 129 0.430 for 2008 and 2009, respectively; for a complete species list for both years, see Appendix 1). Richness and diversity trends for spring 2009 were similar to data from spring 2008, and treatment produced no significant difference compared to controls (F2,24 = 0.174; P α=0.05 = 0.842 and F2,24,= 1.162; P α=0.05 = 0.335 for 2008 and 2009, respectively). Lastly, species diversity (H') was initially approximately 30% higher at McGirt’s Creek in 2008 than in comparison to 2009, a difference that was statistically significant (F1,24 = 4.668; Pα =0.05 = 0.044); this trend was not the same for Tiger Point, with the difference not reaching significance (F2,24 = 3.358 Pα =0.05 = 0.083) (Figs. 3, 4). No significant interaction was found between site and treatment for any dependent variable for either year. The mean similarity value of each cross treatment/site comparison and the resulting ordination plot revealed that there were two main groupings separated by site, with subsequent subgroupings, consisting of the cleared and thinned plots grouping closer together in 2008 (Table 1; Figs. 5, 6). This high degree of similarity between the cleared plots and the thinned plots was to be expected due to the similar vegetation structure in these two experimental treatment groups in comparison to the control replicates two years after the study began. These mean similarity values were fairly disparate, even within sites (Figs. 5, 6). Figure 5. Non-metric MDS ordination plot of community assemblages, Spring 2008. M = McGirt’s Creek, T = Tiger Point. 130 Southeastern Naturalist Vol. 12, No. 1 Discussion In the current study, arthropod diversity was not significantly affected by restoration treatments. However, both treatment groups (thinned and cleared), which experienced biodiversity reduction through experimental treatments, exceeded some measures of biodiversity (i.e., the Shannon index) compared to control plots by the second year of the study at both sites. These results suggest that insufficient time has passed to assess long-term (greater than 5 years) effects of the treatments on arthropod biodiversity or that they have not yet reached an equilibrium. Several reasons may account for differences in biodiversity between treatment groups; certainly one possible reason is that, thus far, only short-term responses have been assessed. Experimental manipulations were undertaken in 2006, and data collection began approximately 2 years later; it is possible that an insufficient period of time elapsed to allow arthropod communities to reach their maximum diversity. A series of studies concerned primarily with arthropod community restoration have shown that recovery may take longer than expected. For instance, a study investigating terrestrial arthropod recolonization following riparian restoration found that abundance of all arthropods was lower at the restored sites in comparison to reference sites, but all sampling was done within two years of restoration, much like our study, in which there may not have been enough time to see appreciable effects (Williams 2006). Blake et al. (1996) found that five years after restoration, the carabid beetle fauna associated with Figure 6. Non-metric MDS ordination plot of community assemblages, Spring 2009. M = McGirt’s Creek, T = Tiger Point. 2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 131 wildflower meadows exhibited fewer species and lower diversity than undisturbed sites. Additionally, arthropod communities following mine reclamation exhibited lower species richness, diversity, and evenness up to six years following revegetation in comparison to control plots (Parmenter and MacMahon 1987, Parmenter et al. 1991). More recently, moth species richness following prairie restoration was found to be lower with the newer restorations (1–3 years) compared to older restorations (7–10 years) and control prairie remnants (Summerville et al. 2007). While trends in overall species richness and diversity in cleared and thinned plots is similar to the control plots at each site, community similarity between sites is relatively low. These results seem to suggest that the fixed size of the plots used in the study support a similar equilibrium species richness (due to simple species-area relationships, similar number of niches, etc.)—thus, plots may have reached an equilibrium for local diversity that varies slightly between our sites, or the plots may contain subsets of the larger overall regional diversity. However, these studies, as with the results reported here, demonstrate that increased diversity or species richness are not likely to be an immediate result of restoration. In a similar study, Rossi et al. (2011) measured the plant community response to the restoration techniques used at McGirt’s Creek and Tiger Point. They demonstrated that plant species diversity did increase as a result of restoration at McGirt’s Creek, but the increase was recent; thus, too little time may have elapsed for vegetation to fully affect the arthropod community. Although treatments did not differ in species diversity within our study, they did differ in terms of community similarity. This result raises the question of whether or not community similarity is a useful, if not better, measure of restoration success. Communities are considered important biological entities in their own right, and conserving representative samples of communities is often viewed as an efficient way of maintaining high levels of diversity (Hunter et al. 1988). Conservation of communities is considered to be a coarse-filter approach as opposed to the fine-filter approach that attempts to conserve a single species (Hunter et al. 1988). In the context of coarse-filter conservation, community similarity measures may be more appropriate than measures of species richness (Su et al. 2004). For instance, the restored treatments represent different successional stages of the ecosystem and they support communities containing different assemblages of species and thus contain a greater proportion of the species pool for the area. In this study, different treatment groups support different arthropod species assemblages as measured by the various community similarity indices, and by taking this into consideration, it might be appropriate to implement multiple restoration techniques so as to conserve the greatest species reservoir at the two experimental sites. Conversely, similarity will shed some insight into species identity and how it compares and relates to ecosystem function in other communities. If one community contains key species, such as keystone species or ecosystem engineers, similarity measures can allow researchers to compare its abundance and proportion to other communities and then provide researchers possible insights into what may drive the presence or absence of such key species. 132 Southeastern Naturalist Vol. 12, No. 1 Acknowledgments The authors would like to thank the following organizations for providing funding for the project: The Nature Conservancy, United States Fish and Wildlife Service, Florida Department of Environmental Protection, and Preservation North Florida. Additional support was provided by the University of North Florida. We also thank Dale Casamatta, Courtney Hackney, Julie Lockwood, and especially Keith Stokes, whose comments improved the manuscript. Literature Cited Battaglia, L., D. Pritchett, and P. Minchin 2008. Evaluating dispersal limitation in passive bottomland forest restoration. Restoration Ecology 16:417–424. Blake, S., G. Foster, G. Fisher, and G. Ligertwood. 1996. Effects of management practices on the carabid faunas of newly established wildflower meadows in Southern Scotland. Annales Zoologici Fennici 33:139–147. Brown, V. 1985. Insect herbivores and plant succession. Oikos 44:17–22. Clark, K., H. Gholz, and M. Castro. 2004. Carbon dynamics along a chronosequence of Slash Pine plantations in North Florida. Ecological Applications 14:1154–1171. Edminster, C.B., and W.K. Olson. 1996. Thinning as a tool in restoring and maintaining diverse structure in stand of southwestern Ponderosa Pine. USDA Forest Service General Technical Report RM-GTR-278. Pp. 61–68. Feeney, S., T. Kolb, W. Covington, and M. Wagner. 1998. Influence of thinning and burning restoration treatments on presettlement Ponderosa Pines at the Gus Pearson Natural Area. Canadian Journal of Forest Research 28:1295–1306. Finnamore, A.T. 2006. The advantages of using arthropods in ecosystem management. Biological survey of Canada (Terrestrial Arthropods) for Canadian Museum of Nature and Entomological Society of Canada, Ottawa, ON, Canada. Freedman, B., V. Zelazny, D. Beaudette, T. Fleming, G. Johnson, S. Flemming, J. Gerrow, G. Forbes, and S. Woodley. 1996. Biodiversity implications of changes in the quantity of dead organic matter in managed forests. Environmental Review 4:238–265. Friend, G.R. 1982. Bird populations in exotic pine plantations and indigenous eucalypt forests in Gippsland, Victoria. Emu 82:80–91. Gaines, W., M. Haggard, J. Lehmkuhl, A. Lyons, and R. Harrod. 2007. Short-term response of land birds to Ponderosa Pine restoration. Restoration Ecology 15:670–678. Gardner-Gee, R., and J. Beggs. 2010. Challenges in food web restoration: An assessment of the restoration requirements of a Honeydew-Gecko trophic interaction in the Auckland Region, New Zealand. Restoration Ecology 18:295–303. Gratton, C., and R. Denno. 2005. Restoration of arthropod assemblages in a Spartina salt marsh following removal of the invasive plant Phragmites australis. Restoration Ecology 13:358–372. Hartley, M.J. 2002. Rationale and methods for conserving biodiversity in plantation forests. Forest Ecology and Management 155:81–95. Hobbs, R. 2007. Setting effective and realistic restoration goals: Key directions for research. Restoration Ecology 15:354–357. Hunter, M.L., G. Jacobson, and T. Webb. 1988. Paleoecology and the coarse-filter approach to maintaining biological diversity. Conservation Biology 2:375–385. Kattan, G., D. Correa, F. Escobar, and C. Medina. 2006. Leaf-litter arthropods in restored forests in the Columbian Andes: A comparison between secondary forest and tree plantations. Restoration Ecology 14:95–102. 2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 133 Kremen, C., R. Colwell, T. Erwin, D. Murphy, R. Noss, and M. Sanjayan. 1993. Terrestrial arthropod assemblages: Their use in conservation planning. Conservation Biology 7:796–808. Kulmatiski, A. 2011. Changing soils to manage plants communities: Activated carbon as a restoration tool in ex-arable fields. Restoration Ecology 19:1 02–110. Lawton, J.H. 1978. Host plant influences on insect diversity: The effects of time and space. In L.A. Mound and N. Waloff (Eds.). Diversity of Insect Faunas. Symposium of the Royal Entomological Society of London 9:105–125. Longcore, T. 2003. Terrestrial arthropods as indicators of ecological restoration success in coastal sage scrub (California, USA). Restoration Ecology 11:397–409. Maleque, M.A., K. Maeto, and, H.T. Ishii. 2009. Arthropods as bioindicators of sustainable forest management, with a focus on plantation forests. Applied Entomology and Zoology 44:1–11. Morrison, E., and C. Lindell. 2011. Active or passive forest restoration? Assessing restoration alternatives with avian foraging behavior. Restoration Ecology 19:170–177. Mulder, C., J. Koricheva., H. Huss-Dannell, P. Hogberg, and J. Joshi. 1999. Insects affect relationships between plant species richness and ecosystem processes. Ecology Letters 2:237–246. Parmenter, R.R., and J.A. MacMahon. 1987. Early successional patterns of arthropod recolonization on reclaimed strip mines in southwestern Wyoming: The grounddwelling beetle fauna (Coleoptera). Environmental Entomology 16:168–177. Parmenter, R.R., J.A. MacMahon, and C. Gilbert. 1991. Early successional patterns of arthropod recolonization on reclaimed Wyoming strip mines: The grasshoppers (Orthoptera: Acrididae) and allied faunas (Orthoptera: Gryllacrididae, Tettigoniidae). Environmental Entomology 20:135–142. Price, P.W. 1984. Insect Ecology. John Wiley and Sons, New York, NY. Redi, B., R. van Aarde, and T. Waasenaar. 2005. Coastal dune forest development and the regeneration of millipede communities. Restoration Ecology 13:284–291. Ricklefs, R., and G. Miller. 2000. Ecology. 4th Edition. W.H. Freeman, New York, NY. Rossi, A.M., R.C. Meyer, D.C. Moon, and K. Stokes. 2011. Effects of thinning and clearing on plant abundance, diversity, and community composition in former pine tree (Pinus elliottii) farms in northeast Florida. Southeastern Naturalist 10:741–750. Ruzicka, K., J. Groninger, and J. Zaczek. 2010. Deer browsing, forest edge effects, and vegetation dynamics following bottomland forest restoration. Restoration Ecology 18:702–710. Siemann, E., D. Tilman, and J. Haarstad. 1998. Experimental tests of the dependence of arthropod diversity on plant diversity. The American Naturalist 152:738–750. Siemann, E., J. Haarstad, and D. Tilman. 1999. Dynamics of plant and arthropod diversity during old-field succession. Ecography 22:406–414. Slobodkin, L. 2000. Proclaiming a new ecological subdiscipline. Bulletin of the Ecological Society of America 81:223–226. Spence, J.R. 2001. The new boreal forestry: Adjusting timber management to accommodate biodiversity. Trends in Ecology and Evolution 16:591–593. Steffan-Dewenter, I., and T. Tscharntke. 1997. Early succession of butterfly and plant communities on set-aside fields. Oecologia 109:294–302. Su, J.C., D. Debinski, M. Jakubauskas, and K. Kindscher. 2004. Beyond species richness: Community similarity as a measure of cross-taxon congruence for coarse-filter conservation. Conservation Biology 18:167–173. 134 Southeastern Naturalist Vol. 12, No. 1 Summerville, K., A. Bonte, and L. Fox. 2007. Short-term temporal effects on community structure of Lepidoptera in restored and remnant tallgrass prairies. Restoration Ecology 15:179–188. Wallington, T., R. Hobbs, and S. Moore. 2005. Implications of current ecological thinking for biodiversity conservation: A review of the salient issues. Ecology and Society 10:15. Wheater, C.P., W.R. Cullen, and J.R. Bell. 2000. Spider communities as tools in monitoring reclaimed limestone quarries. Landscape Ecology 15: 401–406. Williams, K. 2006. Use of terrestrial arthropods to evaluate restored riparian woodlands. Restoration Ecology 1:107–116. Wilson, M.D., and B.D. Watts. 1999. Response of Brown-headed Nuthatches to thinning of pine plantations. Wilson Bulletin 111:56–60. York, R., J. Battles, A. Eschtruth and F. Schurr. 2011. Giant Sequoia regeneration in experimental canopy gaps. Restoration Ecology 19:14–23. Zerbe, S., and D. Kreyer. 2006. Introduction to special section on “Ecosystem restoration and biodiversity: How to assess and measure biological diversity?” Restoration Ecology 14:103–104. Zhu, J., C. Scott, K. Scallon, and G. Myers. 2007. Effects of plantation density on wood density and anatomical properties of Red Pine. Wood and Fiber Science 39:502–512. 2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 135 Appendix 1. Complete morphospecies list for both sites, 2008 and 2009. Gryllus carolina Johansson Chaoboridae Sp. 1 Eurytomidae Sp. 1 Lycosid Sp. 9 Reduviidae Sp. 2 Gyrllus fasciata Haan Chrysomelidae Sp. 1 Flatidae Sp. 1 Lyssomanes viridesWakkenaer Rhapidophoridae Sp. 1 Acrididae Sp.1 Chrysomelidae Sp. 2 Forminicae Sp. 1 Mantidae Sp. 1 Salticidae Sp. 1 Acrididae Sp.3 Cicadellidae Sp. 1 Gastercantha cancriformis (L.) Meloidae Sp. 1 Salticidae Sp. 2 Actinae Sp.1 Cicadellidae Sp. 2 Geometridae Sp. 1 Meloidae Sp. 2 Salticidae Sp. 3 Aedes Sp.1 Cicadellidae Sp. 3 Geometridae Sp. 2 Meloidae Sp. 3 Salticidae Sp. 4 Agelenidae Sp.1 Cicadellidae Sp. 4 Geometridae Sp. 3 Mermeria bivittata Serville Salticidae Sp. 5 Alticinae Sp.1 Cicadellidae Sp. 5 Geometridae Sp. 4 Miridae Sp. 1 Salticidae Sp. 6 Anobiidae Sp.1 Cicadellidae Sp. 6 Geometridae Sp. 5 Miridae Sp. 2 Salticidae Sp. 7 Anopheles Sp.1 Cicadellidae Sp. 7 Gnaphosid Sp. 1 Miridae Sp. 3 Salticidae Sp. 8 Anthocoridae Sp.1 Cicujidae Sp. 1 Gnaphosid Sp. 2 Miridae Sp. 4 Salticidae Sp. 9 Anthribidae Sp.1 Cisseps fulvicollis (Hübner) Gnaphosid Sp. 3 Mogoplistidae Sp. 1 Scaphididae Sp. 1 Anthritidae Sp.1 Cleridae Sp. 1 Gnaphosid Sp. 4 Mordellidae Sp. 1 Scarabidae Sp. 1 Aphididae Sp.1 Cocinellidae Sp. 1 Graphocephala coccinea Forster Muscid Sp. 1 Scarabidae Sp. 2 Aphididae Sp.2 Coelopidae Sp. 1 Griburius larvatus (Newman) Muttlilidae Sp. 1 Scatopsidae Sp. 1 Aphididae Sp.3 Coenagrionidae Sp. 1 Gryllidae Sp. 1 Mycetophilidae Sp. 1 Sciaridae Sp. 1 Aradidae Sp.1 Coenagrionidae Sp. 2 Gryllidae Sp. 2 Mymaridae Sp. 1 Sciaridae Sp. 2 Araenidae Sp. 2 Coleoptera Sp. 1 Gryllus Sp. 1 Mymircinae Sp. 1 Sciomyzidae Sp. 1 Araenidae Sp.1 Colias eurytheme Boisduval Hemisphaerota cyanea (Say) Myridae Sp. 1 Sepsidae Sp. 1 Araenidae Sp.3 Colias philodice Godart Ichneumonidae Sp. 1 Myridae Sp. 2 Simullidae Sp. 1 Araenidae Sp.4 Collembola Sp. 1 Ichneumonidae Sp. 3 Myridae Sp. 3 Solenopsis invicta Buren Araenidae Sp.5 Collembola Sp. 2 Ixodidae Sp. 1 Myrmicinae Sp. 2 Specidae Sp. 1 Araenidae Sp.6 Colydidae Sp. 1 Ixodidae Sp. 2 Nocturridae Sp. 1 Staphylinidae Sp. 1 Asilidae Sp.1 Crematogaster Sp. 1 Largus succintus L. Ochertidae Sp. 1 Staphylinidae Sp. 2 Bibionidae Sp.1 Crytocephalus Sp. 1 Largidae Sp. 1 Oedemeridae Sp. 1 Stratiomyidae Sp. 1 Blattidae Sp.1 Culicidae Sp. 1 Largidae Sp. 2 Oniscidae Sp. 1 Symphyta Sp. 1 Blattidae Sp.2 Curculionidae Sp. 1 Lauxaniidae Sp. 1 Oniscus asellus L. Tabanidae Sp. 1 Buprestidae Sp.1 Curculionidae Sp. 2 Lepidoptera Sp. 1 Opiliones Sp. 1 Tachinidae Sp. 1 Buprestidae Sp.2 Curculionidae Sp. 3 Leptura Sp. 1 Orchesella hexfasciata Harvey Temnothorax Sp. 1 Buprestidae Sp.3 Curculionidae Sp. 4 Lespidae Sp. 1 Orgyia leucostigma Smith Tenebrionidae Sp. 1 Buprestidae Sp.4 Curculionidae Sp. 5 Linphyiidae Sp. 1 Orthoptera Sp. 1 Tenebrionidae Sp. 2 Calliphoridae Sp.1 Curculionidae Sp. 6 Lycaenidae Sp. 1 Orthoptera Sp. 2 Tenebrionidae Sp. 3 136 Southeastern Naturalist Vol. 12, No. 1 Camponotus Sp.1 Delphacidae Sp. 1 Lycidae Sp. 1 Oxypidae Sp. 1 Tephritidae Sp. 1 Cantharidae Sp.1 Delphacidae Sp. 2 Lycosid punctulata Hentz Pardosa Sp. 1 Tetragnathidae Sp. 1 Carabidae Sp.1 Diptera Sp. 1 Lycosid Sp. 14 Pardosa Sp. 1 Tettigonidae Sp. 1 Carabidae Sp.2 Dolichopodidae Sp. 1 Lycosid Sp. 15 Pardosa Sp. 2 Theatops Sp. 1 Carabidae Sp.3 Dolichopodidae Sp. 2 Lycosid Sp. 16 Phaeothripidae Sp. 1 Theatops Sp. 2 Carabidae Sp.4 Dolichopodidae Sp. 2 Lycosid Sp. 1 Phaeothripidae Sp. 2 Thomisidae Sp. 1 Cassidinae Sp.1 Dolichopodidae Sp. 3 Lycosid Sp. 10 Phalacridae Sp. 1 Tipullidae Sp. 1 Catharidae Sp.1 Dolichopodidae Sp. 4 Lycosid Sp. 11 Philodromidae Sp. 1 Tortricidae Sp. 1 Cecidomyiidae Sp.1 Elateridae Sp. 1 Lycosid Sp. 12 Phoridae Sp. 1 Trochosa Sp. 1 Centruroides hentzi Banks Elateridae Sp. 2 Lycosid Sp. 13 Pipuncullidae Sp. 1 Trombididae Sp. 1 Cerambycidae Sp.3 Endomychidae Sp. 1 Lycosid Sp. 2 Pirata Sp. 1 Uca Sp. 1 Ceratopogonidae Sp.1 Epicauta floridensis Werner Lycosid Sp. 3 Pirata Sp. 2 Ulidiidae Sp. 1 Chaecididae Sp.2 Eucnemidae Sp. 1 Lycosid Sp. 4 Psyllobora Sp. 1 Zoraptidae Sp. 1 Chalcididae Sp.1 Eulophidae Sp. 1 Lycosid Sp. 7 Pteromalidae Sp. 1 Chaloididae Sp.1 Eurytides marcellus (Cramer) Lycosid Sp. 8 Reduviidae Sp. 1