Forest Community Structure Differs, but not ...

2014

Forest Impacts and Ecosystem Effects of the Hemlock Woolly Adelgid in the Eastern US 13(Special Issue 6):61–87 Southeastern Naturalist

Forest Community Structure Differs, but not Ecosystem Processes, 25 Years after Eastern Hemlock Removal in an Accidental Experiment Jenna M. Zukswert1,2,*, Jesse Bellemare1, Amy L. Rhodes3, Theo Sweezy3, Meredith Gallogly1, Stephanie Acevedo1, and Rebecca S. Taylor1 Abstract - The spread of Adelges tsugae (Hemlock Woolly Adelgid) directly threatens the survival of Tsuga canadensis (Eastern Hemlock) and has also triggered pre-emptive and salvage logging. In this study, we took advantage of a 25-year-old accidental experiment involving Eastern Hemlock removal by logging at Smith College’s MacLeish Field Station, in western Massachusetts, to investigate how microclimate, ecosystem processes, and forest-floor animal communities might change in the decades following Eastern Hemlock loss. On average, mean understory light levels in summer were 68% higher under young Black Birch (Betula lenta) canopies as compared to adjacent mature Eastern Hemlock forest. Mean daily air temperature, relative humidity, soil temperature, and organic-layer moisture content were similar between young Black Birch and mature Eastern Hemlock plots, although some of these factors were significantly more variable in the former. The soil organic horizon was significantly thicker in Eastern Hemlock plots, but net nitrification rates did not differ substantially between young Black Birch and mature Eastern Hemlock forest plots. We detected significantly greater densities of microarthropods (e.g., mites, collembolans) in the forest floors of Eastern Hemlock plots, possibly linked to the thicker organic horizon. Our results indicate substantial changes in forest structure and microarthropod communities with Eastern Hemlock removal, but little evidence of large changes in key ecosystem processes, like nitrogen cycling. Other sites that represent similar accidental experiments with Eastern Hemlock removal due to past human disturbance likely exist and should be studied before intact reference stands are lost to Hemlock Woolly Adelgid or preemptive salvage logging.

Introduction The decline of Tsuga canadensis (L.) Carriere (Eastern Hemlock, hereafter Hemlock) in eastern North America due to the spread of the exotic pest Adelges tsugae Annand (Hemlock Woolly Adelgid [HWA]) is expected to significantly influence ecosystem processes and forest communities due to Hemlock’s role as a foundation tree species (Ellison et al. 2005). Accidentally introduced to the United States in the 1950s, HWA is an invasive insect that feeds on xylem ray parenchyma cells of Hemlocks, often causing tree mortality within as few as 4 years (McClure 1991, Young et al. 1995). Beyond the direct threat of HWA, the spread Department of Biological Sciences, Smith College, Northampton, MA 01063. 2Faculty of Forestry, University of British Columbia, Vancouver, BC V6T 1Z4, Canada. 3Department of Geosciences, Smith College, Northampton, MA 01063. *Corresponding author [email protected]. 1

Manuscript Editor: Roland de Gouvenain 61

2014 Vol. 13, Special Issue 6 Southeastern Naturalist J.M. Zukswert, J. Bellemare, A.L. Rhodes, T. Sweezy, M. Gallogly, S. Acevedo, and R.S. Taylor

of this invasive insect has also led to pre-emptive and salvage logging efforts (Foster and Orwig 2006, Kizlinski et al. 2002), despite Hemlock’s limited value as a lumber species. Whether due to HWA-induced mortality, pre-emptive cuts, or salvage logging, the community structure and function of the Hemlock-dominated forests in eastern North America is predicted to change dramatically with the loss of this important foundation species in coming decades (Ellison et al. 2005, Orwig and Foster 1998). In areas where Hemlock trees have died from HWA infestation or have been logged, deciduous tree species rather than other conifer species (Orwig and Foster 1998) have typically emerged in their place. The tree species most commonly seen replacing Hemlock in the northeastern United States is Betula lenta L. (Black Birch; Orwig and Foster 1998). In the long term, forest succession resulting from the loss of Hemlocks is predicted to result in mature mixed deciduous forests, and Hemlocks are unlikely to re-colonize these forests for many decades, if ever (Ellison et al. 2005, Orwig and Foster 1998). This transition from evergreen, conifer-dominated ecosystems to deciduous, angiosperm-dominated ecosystems is expected to produce significant changes in associated plant and animal communities and in key biogeochemical processes (Cobb 2010, Ellison et al. 2005, Orwig and Foster 1998). As a foundation species, Hemlock strongly influences forest community structure by creating a unique understory environment that is often cooler, darker, moister, and characterized by more acidic soils than typically seen in deciduous forests (Ellison et al. 2005). The changes in understory microclimate accompanying Hemlock loss as well as changes in leaf-litter inputs, forest-floor composition, and decomposition could alter net rates of nitrogen transformation in soil, stimulating increases in net nitrogen mineralization and nitrification (Cobb 2010, Jenkins et al. 1999, Orwig et al. 2008, 2013). Beyond changes in ecosystem processes, changes are also predicted for plant and animal communities in forests that experience Hemlock loss, including shifts in the composition, abundance, or diversity of bird (Tingley et al. 2002), arthropod (Rohr et al. 2009), salamander (Mathewson 2009), and bryophyte assemblages (Cleavitt et al. 2008). Numerous studies have been launched in the past 10–15 years to examine the ecological changes that might result from the loss of Hemlock due to HWA or logging (Kizlinski et al. 2002, Orwig and Foster 1998). The majority of these studies have been observational or descriptive, involving longitudinal studies of sites succumbing to HWA (e.g., Cleavitt et al. 2008, Eschtruth et al. 2013), or employing space-for-time substitutions along regional gradients of HWA infestation (e.g., Jenkins et al. 1999, Orwig and Foster 1998, Orwig et al. 2008). These studies have provided valuable insights into the environmental changes occurring in the wake of Hemlock loss; however, some longitudinal studies could be confounded by interannual variability in climate, and space-for-time substitutions might be influenced by underlying differences in soils, geology, and physiography among sites. To address some of these limitations, experimental manipulations have recently been implemented in which intact Hemlock control stands are compared to nearby plots in which the species has either been girdled to simulate slow death 62

2014 Vol. 13, Special Issue 6 Southeastern Naturalist J.M. Zukswert, J. Bellemare, A.L. Rhodes, T. Sweezy, M. Gallogly, S. Acevedo, and R.S. Taylor

by HWA or harvested to simulate salvage or pre-emptive logging (e.g., Ellison et al. 2010, Knoepp et al. 2011). For example, large-scale removal experiments have been launched at the Coweeta Hydrologic Laboratory in North Carolina and at Harvard Forest in Massachusetts within the past decade. Ideally, these experiments allow side-by-side comparison of intact Eastern Hemlock stands with forested areas where removal of the species has triggered replacement by deciduous tree species like Black Birch (e.g., Orwig et al. 2013). Although these planned experimental manipulations are already providing important new insight into the early stages of forest succession following Hemlock removal, their relatively recent initiation might confound the specific effects of Hemlock loss with general disturbance effects and also limit the information available on later successional stages. Unfortunately, the rapid spread of HWA also threatens to compromise intact Hemlock control stands before these experiments have had a chance to fully simulate the decades following Hemlock loss. To obtain a longer-term view of the effects of Hemlock removal while still retaining many of the desirable aspects of an experimental manipulation, the present study at Smith College’s MacLeish Field Station in Whately, MA takes advantage of a 25-year-old accidental experiment involving partial removal of Hemlock from a Hemlock-northern hardwoods forest by small-scale logging in the late 1980s. This logging activity created a patchwork of forest stands, with young Black Birch-dominated stands embedded within a more mature Hemlock-dominated forest matrix. This study system allows for documentation of a later stage in the successional trajectory following Hemlock removal, while still minimizing differences in underlying environmental conditions through close spatial proximity of Hemlock-dominated areas and young Black Birch-dominated patches generated by logging. The objective of this study was to collect baseline ecological data from forest stands at Smith College’s MacLeish Field Station in order to investigate the longer-term effects of Hemlock removal on a subset of ecosystem processes and forest-floor community components. In particular, we collected microclimatic and biogeochemical data to observe how understory environmental conditions and forest-floor ecosystem processes, particularly nitrogen cycling, differed between mature Hemlock-dominated forest and young Black Birch-dominated forest ≈25 years after Hemlock removal. We predicted that Hemlock stands would be cooler and darker than young Black Birch stands, and hypothesized that soils beneath the young Black Birch canopy would have thinner organic horizons and exhibit higher net nitrogen transformation rates than soils beneath the Hemlock canopy (cf. Finzi et al. 1998a, b; Lovett et al. 2004). We also investigated forest-floor animal communities to test for differences in the abundance of a key functional group involved in decomposition: forest-floor microarthropods (e.g., mites and collembolans; Coleman et al. 2004). We hypothesized that these forest-floor mesofauna would be more abundant under Black Birch canopies because the higher nutrient content and lower C:N of Black Birch leaf litter (e.g., Cobb 2010) might be expected to stimulate more productive detritivore communities. 63

2014 Vol. 13, Special Issue 6 Southeastern Naturalist J.M. Zukswert, J. Bellemare, A.L. Rhodes, T. Sweezy, M. Gallogly, S. Acevedo, and R.S. Taylor

Field-Site Description The study site for this project was located at Smith College’s 98-hectare MacLeish Field Station in Whately, MA (42°27'N, 72°40'W), in the foothills of the Berkshire Plateau (270 m elevation). Soils at the field station are predominantly inceptisols, specifically classified as extremely rocky loam of the Westminster Series (Mott and Fuller 1967), and tend to be acidic (pH ≈4.1–4.7; Zukswert 2013). The glacial till-derived soil is shallow, well drained, and typically has a persistent organic horizon. The mineral subsoil extends ≈45 cm until reaching Devonian gray mica schist and quartz bedrock of the Waits River Formation, a bedrock type that extends through the foothills of the Berkshire Plateau in Massachusetts into Vermont and includes occasional beds of impure calcitic marble (USGS 2013, Willard 1956, Zen et al. 1983). The natural vegetation of the study area is northern hardwoods-Hemlock-Pinus strobus L. (Eastern White Pine) forest (Westveld 1956). At the MacLeish Field Station, forests are dominated by Hemlock, Black Birch, and Quercus rubra (L.) (Red Oak); canopy trees range from ≈90 to 110 years in age (J. Bellemare, unpubl. data). The land was cleared for farming during the late 18th and early to mid-19th centuries (Crafts and Temple 1899), but then largely abandoned to secondary forest succession in the late 19th to early 20th centuries, as has occurred with many upland areas in southern and central New England (Bellemare et al. 2002, Foster et al. 1998). The forests at the MacLeish Field Station appear to have experienced occasional selective cutting in the early to mid-20th century while under private ownership, with a more substantial, 36-acre commercial cut in 1988 managed by Smith College. In the latter operation, many large Red Oak trees were selectively cut and small patches of forest (≈20 × 20 m) were clear-cut; the rationale for the small intensive cuts is not known because the supervising forester is now deceased. About 94,540 board feet were cut in 1988, with Hemlock representing 35.8% of the total board feet removed (Davies 1988). Black Birch has regenerated vigorously and now forms dense stands of 20–25 year-old trees in the gaps created by this logging operation, surrounded by a matrix forest of mature Hemlock, Red Oak, and Black Birch (Fig. 1). Methods Plot establishment To investigate changes in ecosystem processes and forest-floor communities associated with Hemlock removal, we established a series of research plots situated in adjacent patches of mature Hemlock-dominated forest and young Black Birch stands generated by the 1988 logging operation. Specifically, in 2010, we delineated 4 adjacent pairs of 10 × 15 m young Black Birch and mature Hemlock forest plots, as well as 3 additional Hemlock forest plots (n = 11 plots total, Appendix A). The plot size of 10 x 15 m was selected to reduce edge effects in the ≈20 x 20 m Black Birch gap areas. In the context of the accidental experiment initiated by 1988 logging, we presume that the forest vegetation prior to this cut 64

2014 Vol. 13, Special Issue 6 Southeastern Naturalist J.M. Zukswert, J. Bellemare, A.L. Rhodes, T. Sweezy, M. Gallogly, S. Acevedo, and R.S. Taylor

Figure 1. One of the young Black Birch forest plots situated within an Eastern Hemlockdominated forest at the MacLeish Field Station in Whately, MA. These former logging gaps measure ≈20 × 20 m, but our study plots within them measure 10 × 15 m. Wooden boards on the forest floor are being used as artificial cover objects in an ongoing study of forestfloor animal populations. 65

2014 Vol. 13, Special Issue 6 Southeastern Naturalist J.M. Zukswert, J. Bellemare, A.L. Rhodes, T. Sweezy, M. Gallogly, S. Acevedo, and R.S. Taylor

was relatively homogenous across the entire study site (i.e., similar to the mature Hemlock-Red Oak-Black Birch matrix forest surrounding the young Black Birchdominated gaps today) and that any differences currently observed in ecosystem processes and forest-floor ecology between the two forest types trace to their divergent histories and differing canopy compositions since 1988. The presence of Hemlock stumps in the young Black Birch plots confirms that these areas supported Hemlock prior to logging. In our baseline tree survey in Fall 2012, the 7 Hemlock plots investigated averaged 51.7 m2 ha-1 total basal area and 1489 tree stems per hectare, with Hemlock trees constituting 56% of total basal area, on average (Fig. 2). The young Black Birch plots averaged 14.0 m2 ha-1 total basal area and 7822 tree stems per hectare, with Black Birch trees constituting 84% of the basal area, on average (Fig. 2). We also delineated one mature deciduous forest plot near the study area (≈50 m away), dominated by mature Black Birch of similar age to the mature Hemlock forest, to serve as a proxy for the type of forest into which the young Black Birch stands might develop in ≈50–100 years (Fig. 2). Black Birch has the potential to dominate secondary forests in this area for many decades (e.g., Bellemare et al. 2002).

Figure 2. Mean basal area (m 2 ha-1 ± SE) and stem density (x = # trees ha-1) of Black Birch, Eastern Hemlock and other trees species in the young Black Birch, mature Eastern Hemlock, and mature deciduous forest types.

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2014 Vol. 13, Special Issue 6 Southeastern Naturalist J.M. Zukswert, J. Bellemare, A.L. Rhodes, T. Sweezy, M. Gallogly, S. Acevedo, and R.S. Taylor

Microclimate characterization In order to investigate microclimatic differences among Black Birch and Hemlock plots, we measured understory environmental conditions for a 30-day period from 16 July through 15 August 2010. This time period was selected to provide a snapshot of understory conditions in mid-summer when influences of Hemlock in the canopy might be most important for moderating understory conditions. We installed HOBO Micro Station data loggers (Onset Computer Corporation, Bourne, MA) in 2 pairs of adjacent Hemlock and Black Birch plots (n = 4 data loggers total). Each data logger had an air temperature and relative humidity smart-sensor probe (model #S-THB-M00x) and a 12-bit temperature smart-sensor probe (model #S-TMB-M0XX) inserted below the soil organic horizon in the first 1–2 cm of the mineral soil, and a photosynthetically active radiation (PAR) smart-sensor probe (model #S-LIA-M003). The PAR smart sensors were situated on a horizontal brace ≈75 cm above the forest floor, and the air temperature and relative humidity probe was situated ≈50 cm above the forest floor inside a solar radiation shield (model #RS3). The data loggers collected light and microclimate observations at 1-minute intervals; by using a short sampling interval, we sought to capture data on the understory light environment, which often exhibits rapid fluctuations due to sunflecks (Neufeld and Young 2003). Based on the microclimate and light data collected, daily mean values and standard deviations were calculated for each data logger for daytime (6:00 to 18:00 EST) air temperature, relative humidity, soil temperature, and PAR levels (μmol m-2 s-1 in the 400–700 nm wavelength range). We calculated the percent (%) difference in these mean values for each 12-hour daytime sampling period for each pair of adjacent plots as the mean Black Birch plot value minus the mean Hemlock plot value, divided by the Black Birch mean, multiplied by 100 (positive scores represent higher mean values in the young Black Birch plot). To provide a measure of microclimate variability in the two habitats, we also compared daily standard deviation (SD) values for each metric for each pair of plots, and calculated the % difference in SD between adjacent Black Birch and Hemlock plots. For PAR, air temperature, and soil temperature, we calculated mean % differences and standard deviations based on the differences observed in the 2 pairs of plots (i.e., n = 2 data logger pairs); for relative humidity, a sensor in 1 station malfunctioned, so we present data from a single pair of stations (i.e., n = 1 data logger pair). To evaluate the significance of microclimatic differences between the two forest types, we conducted one-sample t-tests to test the null hypothesis that the mean of the 30 daily mean % difference values for each microclimate factor was zero (i.e., environmental conditions did not systematically vary by forest type). Similarly, we used one-sample t-tests to test the null hypothesis that the variability in microclimate, estimated by % differences in standard deviation, did not differ systematically between the two habitat types. We conducted our analyses using the mosaic package in R (version 2.15.2). 67

2014 Vol. 13, Special Issue 6 Southeastern Naturalist J.M. Zukswert, J. Bellemare, A.L. Rhodes, T. Sweezy, M. Gallogly, S. Acevedo, and R.S. Taylor

Litterfall characterization To characterize and compare leaf-litter inputs to the forest floor of the mature Hemlock and young Black Birch plots, we collected leaf litterfall from July 2012 through June 2013 in 4 of the plots (n = 2 Hemlock, n = 2 Black Birch; Appendix A). We collected litter in rectangular laundry baskets (0.55 × 0.39 m) lined with nylon tulle mesh and held in place with landscape staples. Starting on 5 July 2012, we randomly placed 5 collection baskets in each of the 4 plots (n = 20 collectors at start); collection continued until 25 June 2013. We lost samples from December 2012 due to heavy accumulations of snow and ice that tore the mesh in many of the baskets. In addition, animals overturned 2 baskets in the Hemlock plots in Fall 2012; therefore, we excluded them from the analysis, resulting in a total of 8 Hemlock baskets and 10 young Black Birch baskets for analysis. In total, the collections represent 11 of 12 months for 2012–2013. We retrieved leaf litter from the baskets several times during the year, dried it at 70 °C for >48 h in a laboratory oven prior to sorting it by genus and type of litter (e.g., leaf, twig, seed), and weighed it to the nearest 0.01 g. We summed raw leaflitter data (i.e., dry mass per species per basket) over the study period, converted it to g m-2, and analyzed it with two-sample t-tests contrasting the two forest types; we also calculated % composition by tree species. Organic horizon characterization and N cycling In June 2013, we estimated organic horizon depth for each plot by extracting 10 randomly placed soil cores per plot and measuring the combined depth of the Oe and Oa portions of the O horizon. Additionally, on 20 June and 15 July 2013, we randomly collected ten ≈5 g samples from the Oe and Oa in a subset of 3 young Black Birch and 6 adjacent Hemlock plots (n = 90 samples total) and estimated % moisture content by measuring field-moist and oven-dry mass after >48 h at 75 °C. We analyzed data on organic-layer depth and moisture content with nested ANOVA in R, testing for significant forest-type effects. We investigated net nitrogen mineralization (hereafter net Nmin) and nitrification rates in one pair of adjacent mature Hemlock and young Black Birch plots and in the nearby mature deciduous forest plot using an incubated-core method modified from Orwig et al. (2008) and Robertson et al. (1999). We subdivided each plot into seven 2 m × 2 m subplots; during each sampling period, we obtained two 5-cmdiameter × 20–30-cm-long soil cores from each subplot using a soil core sampler with a sliding hammer (AMS, Inc., American Falls, ID). We immediately brought back one soil core (initial core) to the laboratory in an ice-chilled cooler to be analyzed for baseline N levels, and we placed a second core (incubated core), obtained adjacent to the first core (within 0.5 m), back into the ground in a PVC sleeve. The incubated core was capped at the base to prevent net Nmin products from leaching out with downward flow of soil water, and the PVC sleeve was loosely capped at the top to maintain airflow but prevent rainfall from entering the core (Robertson et al. 1999). We separated the organic (O + A) and mineral horizons (B) of each of the 7 soil cores, sieved the material to 2 mm, and then composited and homogenized 68

2014 Vol. 13, Special Issue 6 Southeastern Naturalist J.M. Zukswert, J. Bellemare, A.L. Rhodes, T. Sweezy, M. Gallogly, S. Acevedo, and R.S. Taylor

the samples to generate 1 organic-soil sample and 1 mineral-soil sample per plot per sampling period. We performed this process for the initial soil cores and then followed the same protocol 3 or more weeks later for the incubated soil cores in order to estimate daily rates of net Nmin over the incubation period. We performed soil incubations during May 2011–July 2013 with occasional breaks, e.g., during inactive periods in winter (Appendix B). We determined the amount of exchangeable NO3- and NH4+ in organic and mineral soil horizons after preparing soil extracts using 0.02M strontium chloride (SrCl2), which has been shown by Li et al. (2006) to be as effective as the more commonly used 1M KCl and 0.01M CaCl2 extractant solutions for both acidic and calcareous soils. The less concentrated 0.02M SrCl2 solutions enabled analysis of NO3- and NH4+ by ion chromatography (Dionex ICS Model 3000, Dionex Corporation, Sunnyvale, CA). Using the ion chromatograph, we measured NO3- with an AS19 column with an isocratic hydroxide eluent, and by using both conductivity and ultraviolet (UV) detection, which eliminates peak interference between NO3- and Cl-. The limit of quantification for NO3- was 0.09 mg L-1 for conductivity detection and 0.12 mg L-1 for UV detection, and the mean standard deviation was ± 0.096 mg L-1. We measured concentrations of NH4+ using a CS12 cation column with an isocratic methane sulfonic acid eluent and a conductivity detector. The limit of quantification for NH4+ was 0.03 mg L-1, and the mean standard deviation was ± 0.103 mg L-1. We prepared 3 SrCl2 replicates for each composite sample by mixing ≈10 g of field-moist soil with 100 mL of SrCl2, and concentrations of NO3- and NH4+ were corrected for mass of the soil sample, soil water, and bulk density of the