Northeast Natural History Conference 2011: Selected Papers 2012 Northeastern Naturalist 19(Special Issue 6):89–100 Land-Cover Change in Western New York: Implications for Soil Carbon Dynamics Mark D. Norris* Abstract - Woody plant expansion is a global phenomenon and has been demonstrated to have impacts on the global carbon (C) cycle as a substantial C sink. Land-cover change in western New York has followed a pattern common to the northeastern US as presettlement forests were extensively cleared for agriculture use. In the past several decades, a substantial portion of this agricultural land has been left to natural succession. This study investigates soil C dynamics across a chronosequence of habitats representing this landcover change including old fields, shrublands, and early successional forests. Despite substantial shifts in plant community composition and structure, neither soil respiration nor soil organic matter changed significantly with habitat type. While consequences of this land-cover change in western New York remain inconclusive, the scale of this change could result in substantial shifts in regional ecosystem C dynamics. Introduction Land-cover/land-use change is one of the major anthropogenic drivers of global ecological change (Foster et al. 2003, Ojima et al. 1994). A major aspect of landcover change involves shifts in the balance of woody and herbaceous vegetation, including a worldwide phenomenon of afforestation and woody plant expansion (Archer et al. 2001). For example, 239,000 km2 of croplands in the United States have been reverted to forests in the past century (Williams 1990). Given the worldwide extent of this woody plant expansion, the terrestrial carbon (C) sink of atmospheric CO2 has been greatly enhanced by this land-cover change in the temperate zone in the past half century (Myneni et al. 2001, Houghton 2003). Land-cover change in the northeastern United States is responsible for some of this terrestrial sink (Houghton 2003, Schimel et al. 2000). Although New England was historically (pre-European settlement) largely forested, forest cover declined to 20–40% in 1830–1890 when agriculture peaked following European settlement. As relatively poor farmland was abandoned, secondary forests returned and remain prevalent in these states, which are now approximately 65– 85% forested (Foster 1995). Aggrading forests in this region have accumulated large stores of ecosystem C, primarily in aboveground biomass and the forest floor, and will continue to be C sinks for at least 200 years following establishment (Hooker and Compton 2003). More elusive are consistent patterns of plant-soil interactions, including plant belowground dynamics and soil C pools and fluxes as a result of land-cover change. Carbon stored aboveground in plant biomass is potentially vulnerable to *Department of Environmental Science and Biology, The College at Brockport State University of New York, 350 New Campus Drive, Brockport, NY 14420; mnorris@ brockport.edu. 90 Northeastern Naturalist Vol. 19, Special Issue 6 loss through disturbances (e.g., fire); however, belowground C processes potentially represent longer-term C sequestration or loss. Increasing woody vegetation is generally associated with increased soil C and attributed to enhanced root production, a shift to more recalcitrant litter, and/or an accumulation of the forest floor (Foote and Grogan 2010, Hibbard et al. 2001, Kaye and Hart 1998, McKinley and Blair 2008, Morris et al. 2007); however, this pattern is not uniform, as several studies have found that soil C does not change significantly with this increased woody plant cover (Billings 2006, Hughes et al. 2006, Lett et al. 2004, McCarron et al. 2003, Scharenbroch et al. 2010). The effects of woody plant proliferation on soil C fluxes via soil respiration are more consistent, generally finding reduced soil respiration rates beneath woody vegetation and attributing this pattern to a shift in microclimate (e.g., cooler temperatures), coarser woody roots, reduced belowground net primary production, or reduced litter quality (e.g., increased % lignin and/or decreased % N) (Kaye and Hart 1998, Lett et al. 2004, McCarron et al. 2003, McKinley and Blair 2008, Smith and Johnson 2004). This reduced rate of soil C losses may correspond to enhanced soil C storage. Vegetative change in western New York is characteristic of the broader pattern of land-cover change in the northeastern United States. The presettlement vegetation of the region has been described as mostly late-successional forest dominated by Fagus grandifolia Ehrh. (American Beech) and Acer saccharum Marshall (Sugar Maple) (Seischab 1990, Wang 2007). Following European settlement, as much as 84% of the forest was cleared for agriculture on a per county basis (Smith et al. 1993, Wang et al. 2010). Wang et al. (2010) found that after agriculture peaked in the late 19th century, forest cover returned in some areas, resulting in a substantial increase in early and mid-successional tree species. This study investigates the consequences of this pattern of land-cover change in western New York as old fields succeed to shrublands and to early successional forests. I focus on soil C dynamics, linking these to dominant plant community habitats along a successional chronosequence. I hypothesize that 1) total soil C increases, and 2) soil respiration decreases over time as these systems shift from herbaceous to woody vegetation. Methods Site description The study took advantage of the habitat diversity at the Iroquois National Wildlife Refuge (43°6'35.2"N, 78°24'03.7"W) in western New York, located midway between Rochester and Buffalo in the rural towns of Shelby and Alabama in Orleans and Genesee counties. The refuge contains nearly 4400 ha, of which more than 700 ha are upland forest, 400 ha are shrubland, and 480 ha are grassland. Dominant herbaceous vegetation included Solidago spp. (goldenrod), various graminoid and Carex spp. (sedge) species, and Asclepias syriaca L. (Common Milkweed). Shrubs primarily included various Cornus spp. (dogwood) and Lonicera spp. (honeysuckle) species and Elaeagnus angustifolia L. (Russian Olive). Common tree species encountered in the late successional shrublands and forests included Fraxinus spp. (ash) and Salix spp. (willow) species, and lesser contributions by Acer spp. (maple), Quercus spp. (oak), Populus tremuloides 2012 M.D. Norris 91 Michx. (Quaking Aspen), and Juglans nigra L. (Black Walnut). The landscape is nearly level to gently sloping. Soils represent several series developed from glacial lakes deposits and are generally silt loams or very fine sandy loams and moderately well drained in study sites (Higgins et al. 1977, Wulforst et al. 1969). The climate is fairly humid continental with strong modification from the Great Lakes, with precipitation evenly distributed throughout the year. Mean annual temperature is 8.8 ºC, and total precipitation is 103 cm/yr. Experimental design Within the refuge, sample sites were selected based on representation of common grassland-shrubland-forest successional habitat types and accessibility. Sites were characterized as one of four habitats spanning the successional gradient from grassland to woodland: grassland/meadow (less than 10% shrub cover), early successional shrubland (40–60% shrub cover), late successional shrubland (>80% shrub cover), and early successional forest with a mature tree canopy, while lacking evidence of old-growth characteristics (e.g., shade-tolerant tree species, pit and mound topography) (Fig. 1). This approach assumes a space-for-time substitution that all sites would otherwise be similar (e.g., soils, water availability, potential vegetation). Site selection emphasizes plant functional types rather than the plant community composition of individual habitats, though it was determined that Figure 1. Stages of old-field succession investigated including A) herbaceous old field, B) early successional shrubland, with shrub islands in a herbaceous matrix, C) late successional shrubland, with limited herbaceous vegetation in a shrub matrix, and D) early successional forest. Photos taken summer 2008. 92 Northeastern Naturalist Vol. 19, Special Issue 6 composition was consistent across habitats of the same type. Most sites are not included in the refuge’s grassland management program and detailed land-use history is unavailable for specific sites. Several of the meadows or shrublands have been maintained with infrequent mowing or hydro-axing, effectively halting succession (Paul Hess, Iroquois National Wildlife Refuge, NY, pers. comm.). Each habitat was replicated three times, with the exception of the early successional shrubland habitat that had four replicate sites for a total of 13 study sites. These young shrubland sites also differed from the rest in that shrubs exist as shrub “islands” or distinct patches within an herbaceous matrix. Thus, to characterize this habitat as a whole, both dominant vegetative covers (shrub vs. herbaceous) were analyzed separately and then combined based on the percent shrub cover of that site for comparison with other habitats. Three sample plots (subplots) were utilized in each of the habitats (or sub-habitats in the early successional shrublands) for all data collection and then averaged prior to statistical analysis. This approach may mask variation due to shrub island size and age (Wheeler et al. 2007), but sites demonstrated relatively little heterogeneity in shrub patch structure. Data collection Soil respiration rates were measured six times, approximately every three weeks between mid-June and mid-October 2008. Rates were determined in situ using a Li- Cor 6400 infrared gas analyzer with soil CO2 flux chamber (Li-Cor, Lincoln, NE), with soil temperature measured simultaneously. Measurements were generally made in the morning and early afternoon. For each sampling date, the order in which sites were visited differed. Small soil cores (2.5 cm diameter, 20 cm depth) were removed to determine gravimetric soil moisture via oven drying. On 13 September 2008, numerous other variables were collected centered on or immediately adjacent to each of the three soil-respiration measurement points. Herbaceous aboveground biomass was harvested in 0.125-m2 plots, representing peak biomass. This biomass was returned to the lab, dried to a constant weight at 60 ºC, and weighed. Shrub stem density and basal area just above ground level were determined in 0.25-m2 plots. In the early successional forests, tree density and basal area (at breast height) were determined in 3-m-radius plots. Soil cores (5 cm diameter, 20 cm deep) were collected following removal of aboveground biomass for determination of bulk density, soil moisture, root biomass, and soil C. Cores were kept on ice until returned to the lab, and then were passed through a 4-mm sieve. Any rocks or aboveground plant tissue were discarded. Bulk roots and soil samples were oven-dried separately and weighed. Subsamples of dried soil were composited in equal quantities by habitat site (or sub-habitat for the early successional shrubland sites) and analyzed for total C content via an elemental analyzer (NC 2100 Soil Analyzer, ThermoQuest, Milan, Italy) as well as for soil organic matter (SOM, loss on ignition at 500 ºC for 2 hours). Because the two measures (SOM and total soil C) have been used interchangeably and were strongly correlated (P < 0.0001), only SOM is reported in the results. Analyses One time measurements and the means of repeated measurements were statistically analyzed for effects of successional stage (habitat) by ANOVA (n = 13) 2012 M.D. Norris 93 with post hoc Tukey’s tests to determine differences between habitats. The woody basal area data did not satisfy ANOVA assumptions, so the non-parametric alternative Kruskal Wallis was used. Paired t-tests were conducted to compare the same variables in herbaceous- versus shrub-dominated patches in the early successional shrubland sites. Soil respiration rates across the four habitats and within the two sub-habitats of the early successional shrubland were analyzed using a repeated measures analysis. Graphs of soil respiration and microclimate over the course of the season are not shown, as patterns were fairly consistent (i.e., there were no time x habitat interactions) and represented by the means over time. All analyses were completed with SPSS (SPSS, Inc.) Results Habitat and vegetation There was relatively little variation in plant species composition and shrub cover among sites of the same habitat type. This homogeneity may have been due to management to maintain that habitat type; it is expected that any such management would amplify effects of that habitat, essentially prohibiting succession from occurring and retaining the character of that successional stage. Through the successional sequence studied, herbaceous biomass decreased significantly (F = 17.82, P < 0.0001, r2 = 0.856; Fig. 2A), generally as woody biomass or canopy cover increased. Woody basal area varied significantly between habitats (H = 10.16, P = 0.017), but did not increase linearly, as the dense late successional shrublands had the greatest basal area compared to the early successional shrublands and forests, which were similar (Fig. 2B). Belowground, fine root biomass did not statistically differ between habitats (Fig. 2C). The comparison of the shrub-dominated and herbaceous-dominated communities in the early successional shrubland habitat were similar to that across all habitat types. Herbaceous biomass decreased with shrub establishment (t = 5.56, P = 0.012, Fig. 3A), while woody basal area increased (t = 18.46, P < 0.0001, Fig. 3B). Root biomass also increased, albeit marginally, with woody plant cover (t = 2.43, P = 0.093; Fig. 3C) Soil C dynamics and microclimate Soil respiration rates generally decreased with shrub development, then increased with forest establishment; however, these differences were not statistically significant when analyzed by repeated measures (F = 1.896, P = 0.201) or when averaged across dates (F = 1.90, P = 0.201, r2 = 0.387, Fig. 2D). Soil temperatures generally decreased with successional stage from meadow to forest, resulting in a significant effect of habitat when analyzed by repeated measures (F = 5.732, P = 0.018) or when averaged across dates (F = 5.73, P = 0.018, r2 = 0.657; Fig. 2E). Soil moisture generally increased during succession, but was characterized by greater variability, and thus there were no statistical differences due to habitat type when analyzed over time (F = 0.391, P = 0.763) or when averaged over time (F = 0.39, P = 0.763, r2 = 0.115; Fig. 2F). As before, patterns of soil respiration and microclimate in the two contrasting sub-habitats of the early successional shrubland largely mirrored those 94 Northeastern Naturalist Vol. 19, Special Issue 6 Figure 2. Ecosystem responses (mean + standard error) across 4 stages of succession from meadow to forest including A) herbaceous biomass, B) woody basal area, C) belowground biomass (0–20 cm), D) soil CO2 efflux (averaged across 6 sample dates), E) soil temperature (0–10 cm; averaged across 6 sample dates), F) soil moisture (0–20 cm; averaged across 6 sample dates), and G) soil organic matter (0–20 cm). Bars with different lowercase letters within each panel indicate statistical differences between habitats (α = 0.05). 2012 M.D. Norris 95 Figure 3. Ecosystem responses (mean + standard error) under contrasting vegetative cover in early successional shrublands including A) herbaceous biomass, B) woody basal area, C) belowground biomass (0–20 cm), D) soil CO2 efflux (averaged across 6 sample dates), E) soil temperature (0–10 cm; averaged across 6 sample dates), F) soil moisture (0–20 cm; averaged across 6 sample dates), and G) soil organic matter (0–20 cm). 96 Northeastern Naturalist Vol. 19, Special Issue 6 across the entire successional gradient. Soil respiration rates were highest in the herbaceous-dominated communities four of the six dates, but the overall pattern was insignificant (F = 3.171, P = 0.125). Soil temperature was typically reduced by shrub cover, such that there was a significant effect of plant community (F = 7.223, P = 0.036). Again, there were no effects of sub-habitat on soil moisture (F = 0.001, P = 0.976). SOM increased initially with shrub establishment in the early successional shrublands, but then decreased with increasing woody plant establishment (Fig. 2G), and there was no significant effect of habitat (F = 0.63, P = 0.614, r2 = 0.173). Within the early successional habitat, the increase in SOM from herbaceous cover to shrub cover was marginally significant (t = 2.71, P = 0.073; Fig. 3G). Discussion Despite substantial shifts in the plant community structure and composition along this chronosequence, soil C dynamics have not changed correspondingly. In short, changes in the plant community were largely predictable, but mechanisms driving soil C dynamics remain inconclusive. Given that there was relatively little variation in species composition and woody cover within each successional stage, it was expected then that the dominant functional group in the plant community would exert greater influence over ecosystem functioning, including soil C dynamics. It was anticipated that SOM would reflect patterns of aboveground biomass and productivity and increase over time. In contrast to this hypothesis, SOM increased from the meadows to the early successional shrublands, but then declined with further woody development, although with no statistical differences. It may be that the trends of SOM follow more closely that of aboveground litter quality rather than litter quantity. In other words, the litter C:N ratio or some other measure of litter recalcitrance may increase from herbaceous flora to shrubs (e.g., greater woody litter) followed by a subtle decrease as shrubs succeed to early successional forest (greater leaf litter relative to woody inputs). Previous studies have shown that during old-field succession or woody encroachment, decay rates generally decrease due to a reduction in litter quality (Cortez et al. 2007, Kazakou et al. 2006, Norris et al. 2001b); however, few studies have explicitly addressed litter chemistry dynamics and patterns of litter recalcitrance in the context of old-field succession. Patterns in soil respiration also conflicted with the hypothesis, largely due to increased soil C fluxes from the late successional shrublands to forests; however, there was a negative relationship between SOM and soil respiration (r = -0.534, P = 0.060). If aboveground litter recalcitrance peaks in shrubland habitats as suggested above, it may also explain patterns in soil C flux by retarding decay. Belowground productivity may also help explain patterns of both SOM and soil respiration (Hibbard et al. 2001). Studies in the tallgrass prairie found that as shrubs established in grassland, root production decreased (Lett et al. 2004) and was associated with coarser roots (McCarron et al. 2003), both concomitant with decreases in respiration rates. The one-time measurement of root biomass is insufficient to address this pattern here, but the increased root biomass in the 2012 M.D. Norris 97 forests likely contributed to elevated soil respiration rates. Soil C fluxes are often attributed to soil microclimate. As has been found in other woody encroachment or forest succession studies (McCarron et al. 2003, Smith and Johnson 2004, Tang et al. 2009), soil respiration was positively linked to soil temperature across all sites (P < 0.0001, r2 = 0.361). Generally, as woody basal area increases, soil temperature decreases, likely depressing soil microbial activity and soil respiration rates, a pattern also reflected in the data here (r = -0.484, P = 0.094). There are numerous alternative explanations for why SOM is not increasing in this successional chronosequence. First, increased C input to the soil associated with woody plants (i.e., litterfall, root exudates, fine root turnover) could stimulate microbial activity and actually decrease SOM as has been found in lownutrient soils (Fontaine et al. 2004); however, this seems unlikely given probable decreases in litter quality and because atmospheric N deposition in this region is relatively high (>4 kg N/ha/yr; National Atmospheric Deposition Program 2008). Second, Jackson et al. (2002) found that patterns of SOM during woody encroachment across a broad precipitation gradient depended on soil moisture in that more mesic sites lost enough SOM with woody establishment to offset C gains in plant biomass. They attributed this loss to reduced belowground production. In this case, such a scenario is unlikely given the general lack of substantial variability in soil moisture between habitats and because patterns of root biomass were not correlated with SOM or soil respiration. Third, there may not have been enough time for soil C to accumulate substantially in these early successional habitats. Despite the decadal time scale, litter quantity and quality may not vary enough across the chronosequence to drive significant soil C sequestration; however, Knapp et al. (2008) found that time since conversion was not important in differences in SOM between shrubs and prairie. Fourth, although extirpated from the region with glaciation, earthworms are present on site (M.D. Norris, pers. observ.) and may have dramatic influences on soil C dynamics. Earthworm impacts may include consumption of plant litter, soil mixing, altered soil microbial community, and increased nutrient cycling (including C) (Bohlen et al. 2004). Fifth, land-cover change may be impacting SOM at depths much deeper than 20 cm as measured here, as much more soil C is found below this (Batjes 1996). Jobbágy and Jackson (2000) found that the plant functional types studied here alter soil C distribution, and that as grassland succeeds to shrubland and forest, an increasing proportion of soil C is found in the surface layers. This change is due in part to an associated increase in aboveground litter but also to a shift in belowground biomass as rooting depth varies with plant functional type (Jackson et al. 1996). Despite the lack of apparent soil C sequestration in this chronosequence, woody basal area increased and likely contributes dramatically to ecosystem C. Several related studies have found that the vegetation represents the greatest change amongst the C pools during this land-cover transformation, primarily driven by a large increase in aboveground net primary productivity (Briggs et al. 2005, Hibbard et al. 2003, Hughes et al. 2006, Lett et al. 2004, McKinley and Blair 2008, Norris et al. 2001a, Scharenbroch et al. 2010). Another C pool of importance is the forest floor, which is likely to accumulate detritus and may also far outweigh increases of C in mineral soil (Hooker and Compton 2003, Kaye and Hart 1998); however, there was 98 Northeastern Naturalist Vol. 19, Special Issue 6 no apparent accumulation of the forest floor in our shrublands or forests. By metaanalysis of global afforestation, Paul et al. (2002) indicated that soil C depended on a variety of factors, but that soil C accumulation was small relative to biomass C. In a similar analysis of afforestation following agricultural abandonment, Laganière et al. (2010) found that soil C sequestration was greatest in croplands compared to pastures or grasslands. The refuge does have a history of farming, but detailed records for these study sites are unavailable. In conclusion, this land-cover change of meadow to shrubland to forest is predominant in this region. Much of this change can be attributed to the recent history of agricultural abandonment in western New York (Wang et al. 2010). My results suggest that this land-cover change does appear to influence patterns of ecosystem functioning, but not to the magnitude or sometimes direction expected. To better predict the ecosystem C implications of this change in the region, we need a better understanding of the complexity of the mechanisms behind soil C dynamics, including litter chemistry and belowground production and turnover. Additionally, we need to quantify biomass aboveground and on the forest floor, as this likely represents the greatest shift in ecosystem C across the chronosequence, then link these patterns to regional acreage representing these habitats. Acknowledgments This project would not have been possible without the cooperation of Thomas Roster and Paul Hess at the Iroquois National Wildlife Refuge. Justin Rogers, Emily Reilly, and Ryan Stotz provided help in sample collection. Erik Lindquist and Nate Grosse aided in sample processing. The Cornell Nutrient Analysis Laboratory performed soil nutrient analyses with funding from the College at Brockport Scholarly Incentive Award Program. Previous versions of this manuscript were improved substantially with comments and suggestions from Dan Potts and two anonymous reviewers. Literature Cited Archer, S., T.W. Boutton, and K.A. Hibbard. 2001. Trees in grasslands: Biogeochemical consequences of woody plant expansion. Pp. 115–138, In E.-D. 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