Changes in wetland forest structure, basal growth, and composition across a tidal gradient.
Because of sea level rise, many of our coastal ecosystems are under threat of increased inundation (Williams et al., 1999; Morris et al., 2002). Freshwater tidal systems may be most susceptible because of the change in salinity regime that is expected to accompany sea level rise (Craft et al., 2009). Freshwater tidal forests often occur at the upriver extent of tidal influence and tend to be most prevalent in larger river systems with substantial flow and gradual elevation gradients (Doyle et al., 2007). There has recently been increased attention focused on these ecosystems (Conner et al., 2007; Courtwright and Findlay, 2011; Martin et al., 2012) because these forests normally represent the freshwater extent of estuarine systems and there is concern about their resilience and long term response to rising sea levels. Hydrology is a critical forcing factor influencing the composition and functional role of all riverine wetlands (Williams, 1998). Wetland hydrology can shift dramatically as rivers become tidally influenced in the coastal zone. Related to freshwater tidal hydrology is the potential for periodic saltwater exposure in these zones, particularly during droughts with decreased freshwater flow, and after tropical storm events (increased tidal surge). Not surprisingly, riparian wetlands along this transition zone may undergo significant changes in species composition related to tidal influence (Wharton, 1982; Conner et al., 2007). Forest structure may also shift along this transitional zone and represents a variety of conditions. The size, distribution, and density of trees within a wetland have a substantial influence on several wetland functional attributes such as canopy cover, hydroperiod, habitat, and nutrient cycling (Sharitz and Mitsch, 1993; Brinson and Rheinhardt, 2000).
Several reports in the Southeastern United States have shown that forested wetlands in the freshwater tidal zone are characterized by lower basal areas and/or greater tree density compared to other floodplain forests (Brinson et al., 1985; Krauss et al., 2009). These conditions may contribute to other structural differences between tidal and upstream, nontidal wetland forests. The occurrence of tree saplings, shrubs, and herbaceous vegetation may shift considerably with less canopy cover and greater light penetration (Rheinhardt, 2007). Forest basal area and stem wood production reflects the current stock, productivity, and mortality of trees and could vary considerably along a tidal gradient. Similarly, changes in forest structure may influence the amount of coarse woody debris (CWD) on forest floors. Accumulation of woody debris is the balance between in situ production, transport into and out of the wetland, and litter decomposition and destruction (Facelli and Pickett, 1991). This debris is important as wildlife habitat and serves as an immobilization/-mineralization repository for nutrients (Harmon et al., 1986). Currently there is very little information about how much CWD biomass is present in a tidal freshwater swamp.
Historically, freshwater tidal swamps have been under-studied although recent work has expanded our understanding of these wetlands, particularly in the Southeastern United States (Conner et al., 2007; Krauss et al., 2009). Although less is known about tidal wetlands, nontidal floodplain forests in the Southeastern United States have been investigated and reviewed extensively (e.g., Conner and Buford, 2000; Sharitz and Mitsch, 1993; Wharton, 1982). Several comprehensive studies have examined forested wetlands and the management issues specific to the Apalachicola River, Florida (Elder, 1988; Light et al., 2006). These studies, however, only examined nontidal wetlands and stopped near the boundary of tidal influence. Consequently there are very few direct comparisons that have been made between tidal and nontidal sections of a lower river system.
Using upriver, nontidal floodplain forests for comparison, we examined how riparian forest structure and composition changes as hydrologic conditions (and related potential for saltwater exposure) shift from fluvial to tidal. Forest structure and composition are important measures for determining the capacity of wetlands to provide important functions such as nutrient cycling and wildlife habitat. Concurrent research has demonstrated that as forests along the Apalachicola River become tidal there are notable changes in foliar nutrient dynamics (Anderson and Lockaby, 2011a), hydrology (Anderson and Lockaby, 2011b), and tree communities (Anderson and Lockaby, 2011c). The objective of our study was to examine structural changes in forested wetlands along a tidal gradient on the Apalachicola River, Florida, U.S.A.
Our study was conducted along the lower 84-kin reach of the Apalachicola River (29[degrees]49'44.47"N, 85[degrees]01'26.80"W) (Fig. 1). Study plots were established in forested wetland stands along the river and major tributary/distributary edges. Ten wetlands along this reach were fitted with water level recorders (5 PSIG, Model Standard P, In-Situ, Inc., 2000) to monitor hydroperiod for >2 y. Based on these data and proximity to other stands (Anderson and Lockaby, 2011b), all wetlands were designated being either tidal (n = 12) or nontidal (n = 8). Tidal wetlands were dominated by swamp tupelo [Nyssa sylvatica var. biflora (Walt.) Sarg.], bald cypress [Taxodium distichum (L.) Rich.], cabbage palm [Sabal palmetto (Walt.) Lodd.ex J.S. Shult. and Shult.], and mixed hardwoods such as sweet bay (Magnolia virginiana L.), sweet gum (Liquidambar styraciflua L.), and pumpkin ash [Fraxinus profunda (Bush) Bush]. Nontidal forests were dominated by water tupelo (Nyssa aquatica L.), Carolina ash (Fraxinus caroliniana Mill.), Ogeechee tupelo (Nyssa ogeche Bartr. Ex Marsh.), and overcup oak (Quercus lyrata Walt.). We could not locate firm accounts of past logging specific to this reach of the Apalachicola River but all indications suggested that most of the area has been left uncut since the early 20th century, when wetland logging in the southern U.S. was at its peak (Lockaby, 2009). In general wetland soils along this reach consist of Sulfihemists, Sulfaquents, and Fluvaquents that tend to have organic C concentrations between 7.5 and 17.9% (Coultas, 1984). Soils series mapped in this area included Brickyard, Chowan, and Kenner series with Maurepas mucks reported in two tidal stands closer to the river mouth (Sasser et al., 1994).
Forest stands and data collection.--We examined forest structure at tidal and nontidal forested wetlands using data collected between Dec. 2006 and Mar. 2010. Twenty stands along the study reach were sampled, although only a subset was used to examine most of the structural components described below.
Forest canopy size class.--In each stand, we established a transect with either one or three plots (Fig. 1). Transects were aligned perpendicular to the creek/river edge, and we established study plots at 20 m, 60 m, and 140 m intervals. Stands with one plot were established at 60 m and were installed to increase the number of stands. Each plot was a 500 [m.sup.2] circle where we identified all tree species (>2.5 cm DBH) and measured diameter at breast height (DBH) (i.e., 1.3 m). For trees with multiple trunks, we measured each stem meeting our size criteria. For tress with a butt swell at breast height, we measured tree diameter immediately above the swell. All trees were measured at the end of the 2009 growing season (late Nov. to early Mar.) and the number and percentage of canopy trees were determined for the following DBH size classes: 2.5-5.0 cm, 5.0-10.0 cm, 10.0-15.0 cm, 15.0-20.0 cm, and >20.0 cm. We calculated the average size class distribution (stems [ha.sup.-1] and percentage) for each plot and averaged for stands with multiple plots per transects.
Forest basal growth.--To evaluate basal area growth between forested wetland treatments, we monitored eight stands (three tidal and five nontidal, all with three plots per transect) annually for 4 y. All trees in these plots were tagged, identified, and measured for DBH at the end of the growing season starting in Nov. 2006. Changes in basal area per plot reflected tree stem growth, recruitment of new trees (those >2.5 cm DBH), and loss of trees due to mortality. We summed the basal area of all living trees and converted total basal area per plot to basal area per hectare. To compare stand level productivity, we examined year to year change in basal area by calculating annual basal area increment (BAI) ([m.sup.2] [ha.sup.-1] and %) for each plot using the following equations:
BAI ([m.sup.2] [ha.sup.-1]) = [Ai.sub.Year X] - [Ai.sub.Year X - 1]
BAI(%) = [([AI.sub.Year X] - [Ai.sub.Year X-1])]/[Ai.sub.Year X - 1] * 100
Where [Ai.sub.Year X] = forest basal area per ha ([m.sup.2] [ha.sup.-1]) in year X
Sapling, shrub and herbaceous vegetation.--Using the same plots established for tree measurements, we also identified and counted sapling and shrub (<2.5 cm DBH) species at the end of the 2009 growing season. We counted all the stems in the plot including shrubs with multiple stems. Stem counts for each plot were converted to stems [ha.sup.-1] and averaged for stands with multiple plots per transect. We examined herbaceous cover and species occurrence at 10 forest stands (four tidal and six nontidal) in August 2007, all with three plots per transect. For each plot, three 1-[m.sup.2] quadrats were randomly distributed. We identified all understory vegetation to species (or lowest taxa possible) and its percent cover estimated to the nearest 5% (species <2.5% were rounded to 2.5%). Average total herbaceous cover and the number of species were determined per plot. We calculated species importance values (IV) for each stand based on relative frequency and dominance of each species.
Downed woody debris biomass.--For eight stands (three tidal, five nontidal), we used the line-intercept method (Brown, 1974) to estimate coarse woody debris (CWD) in forests. In each stand, we measured CWD along four parallel transects established 40 m apart. Each transect consisted of five subplots 20 m apart for 20 subplots per stand. At each subplot, a 20 m line was randomly extended. All woody debris that intersected that line was counted, measured for diameter, and put into a diameter size class: small (<7.62 cm) or large (>7.62 cm). Piece counts and size classes were then used to estimate standing CWD mass (Mg [ha.sup.-1]) per Brown (1974).
Statistical analyses.--We tested differences in mean stem density (total and per size class), basal area, BAI, herbaceous cover, and CWD (small and large) between tidal and nontidal forest stands using a hierarchal approach (stands nested in hydrologic groups). A nested- or partly nested-ANOVA using a general linear model (GLM) procedure was used to evaluate differences between our hydrologic groups (tidal v. nontidal). In all models, hydrologic group was designated as a fixed factor and forest stand as a random factor. For BAI data, plot-level data was crossed with year and nested within stand and hydrologic group. We tested data for normality and homogeneity of variance using the Ryan-Joiner test and Levene's test, respectively, and transformed data when needed using either square root or Log10. We examined differences in sapling/shrub stem density, sapling shrub richness, and herbaceous species richness between tidal and nontidal stands using the Kruskal-Wallis test because of nonnormality of data. Significant differences were reported at P < 0.05 and all statistical analyses were conducted using MiniTab 14 (MiniTab Inc., 2003).
RESULTS AND DISCUSSION
Forest canopy size class and basal growth.--We found that there were differences in tree size class, basal area, and BAI between treatments. Tidal wetlands along the Apalachicola River had a greater number and proportion of stems in smaller size classes than in the nontidal wetlands (Table 1). Overall tree density was higher in tidal wetlands compared to non-tidal (1446 [+ or -] 159 and 962 [+ or -] 100 stems [ha.sup.-1], respectively) and consistent with an earlier estimate made using a larger pool of forest stands along the study reach (Anderson and Lockaby, 2011c). The greatest number and proportion of stems for both treatments were in the smallest size class (<5.0 cm) with a decreasing number of larger size stem classes. The number of larger stems was much lower in tidal wetlands as only 10.7 stems [ha.sup.-1](or < 1% of the total trees) were >20.0 cm DBH compared to 68.9 stems [ha.sup.-1] (or 8.3% of the total trees) in nontidal wetlands. Tidal wetlands had nearly twice the number of small stemmed (2.5-5.0 cm DBH) trees compared to non-tidal wetlands (699 [+ or -] 112 vs. 359 [+ or -] 54 stems [ha.sup.-1]).
Our observation of smaller trees occurring in tidal forests was consistent with other studies (Brinson et al., 1985; Conner et al., 2007; Effier et al., 2007) and suggested slower growing and/or younger trees in a stressfill environment. Wetlands in the tidal reaches often experience year-round tidal flooding, periodic saltwater flooding, and greater nutrient limitation compared to nontidal swamps (Brinson et al., 1985; Doyle et al., 2007). In Sep. 2009, a survey of soil conditions along the lower Apalachicola River found that mean electrical conductivity was 2.79 [+ or -] 0.35 mmhos [cm.sup.-1] in tidal wetlands and 0.24 [+ or -] 0.02 mmhos [cm.sup.-1] in nontidal wetlands (Anderson and Lockaby, 2011c). The difference was likely related to recent low river flow conditions that increased upriver saltwater intrusion (Anderson and Lockaby, 2011b). Although episodes of saltwater intrusion into tidal freshwater wetlands are infrequent, it occurs periodically, but it is uncertain how long elevated soil salinity may persist. Flooding and soil saturation were prolonged throughout the year in tidal wetlands compared to nontidal wetlands (Anderson and Lockaby, 2011b). Because of the long hydroperiod, trees are often restricted to elevated hummocks that are a common feature in tidal freshwater forests (Rheinhardt, 2007; Duberstein and Conner, 2009). Along the Apalachicola River, hummocks were generally shallow compared to other reported tidal swamps (Anderson and Lockaby, 2011c) and flooding stress may be even more pronounced.
As expected average forest basal area was significantly higher for nontidal wetlands (64.4 [+ or -] 2.8 [m.sup.2] [ha.sup.-1]) than for tidal wetlands (35.6 [+ or -] 3.3 [m.sup.2] [ha.sup.-1]) (P = 0.002, F = 28.02, df = 1). Our estimates were consistent with an earlier estimate made using a larger pool of forest stands along the study reach (Anderson and Lockaby, 2011c). Both wetland types had basal areas that were within the range commonly reported for deepwater swamps in the southeastern U.S. (e.g., Conner and Buford, 1998) but tidal forest basal areas were at the low end of that range. The average BAI of tidal wetlands was also significantly higher in nontidal forests than tidal forests (P = 0.028, F = 8.41, df = 1) and substantial year-to-year changes in BAI were noted for both forest types between 2007-2010 (Table 2). BAI for tidal forests ranged from -0.07 to 0.30 [m.sup.2] [ha.sup.-1] [yr.sup.-1] (or -0.21 to 0.85% of standing basal area) and 0.07 to 0.55 (or 0.11 to 0.85%) for nontidal wetlands. Both forest types experienced substantial declines in BAI during the 2008 growing season but for different reasons. In nontidal forests, reduction in BAI was primarily the result of several larger trees that fell in or around plots as a result of a wind storm in winter 2008. The loss of basal area by fallen or crushed trees offset basal production of remaining trees for that year. Tree damage varied considerably between nontidal stands (some showed no effects from the storm) and common fallen trees included Ogeechee tupleo, sweet bay, and oak. Mortality rates for healthy wetland forests are normally low (~2%; Conner et al., 2002) but windstorm events can change canopy structure and composition by increasing tree mortality rates, particularly for large trees (Loope et al., 1994; Conner et al., 2002). It appeared that certain nontidal stands were more susceptible than others as evident by tree mortality from this event and the number of trees that were already partially tipped or leaning. Interestingly, there was less evidence of wind fallen trees in tidal wetlands although winds did knock down branches and break snags. Tree susceptibility to wind may be species specific and our observations were consistent with Conner et al., (2002) who found more trees had fallen from wind in periodically flooded swamps of Louisiana and South Carolina [common species there included sweet gum, American elm (Ulmus americana L.), oaks, and other hardwoods] than in swamps that are flooded most of the year (common species included bald cypress and water tupelo). Blowdown accounted for much of the tree mortality observed in these wetlands during a 12 y monitoring period (Conner et al., 2002).
For tidal wetlands, low BAI in 2008-2009 was the combined result of tree mortality and reduced growth of living trees. Mortality in tidal forests appeared to be caused by drought induced saltwater intrusion prior to and during the early portion of the study period. During regional drought years, saltwater can move farther upriver as the river flow decreases (Doyle et al., 2007). Most mortality in tidal swamps was detected in the 2008-2009 growing seasons, although the drought actually began in 2006 and was most severe in 2007 (Maxwell and Soule, 2009; Wang et al., 2010). In some cases, it was difficult to determine when tree mortality had occurred and we revisited several tidal plots during the growing season to confirm tree mortality. Two tidal stands (15 and WS1) showed indications of widespread tree mortality in the past and experienced the greatest mortality and least growth during our study (data not shown). These forests appear to be transitioning from forest to marsh, a condition that may occur in other tidal sites as sea-level rise continues.
Others have found drought induced salinity stress to be a major factor in reducing tree growth (Brinson et al., 1985; Krauss et al., 2009). Ozalp et al., (2007) attributed much of the variability in stemwood production of tidal swamps along the Pee Dee River in South Carolina to drought. In our study, stemwood production was most affected in the 2008 growing season, a year following peak drought conditions (Anderson and Lockaby, 2011b). The tinting of tree annual basal growth and exposure to salt stress may partially explain this pattern. Wetland floodplain stem growth in the southeastern U.S. is known to start early in the growing season (mid- to late-May; Keeland and Conner, 1997). Drought induced low river flows (and potential saltwater intrusion) may be most severe in late summer or early autumn (Anderson and Lockaby, 2011b).
Sapling, shrub, and herbaceous vegetation.--On averagetidal forests had more than twice the density and richness (549 [+ or -] 136 [ha.sup.-1], 6.9 [+ or -] 1.0; respectively) of saplings/shrubs than nontidal wetlands (210 [+ or -] 72 [ha.sup.-1], 3.0 [+ or -] 0.6, respectively), although these differences were not significant. This was likely because of the wide range of densities and species richness observed in tidal forests and the small number of stands sampled. Forest types also varied in sapling/shrub composition (Table 3). Tidal wetlands were dominated by Morella cerifera (176 stems ha l) and Fraxinus sp. (117 stems [ha.sup.-1]), which represented 53% of all stems in these forests. Species identification of Fraxinus saplings was uncertain but most of them were likely pumpkin ash based on their prevalence in the canopy. Both wax myrtle and pumpkin ash have been noted as common to other tidal freshwater wetlands throughout the southeastern U.S. as well (Conner et al., 2007; Light et al., 2007). To a lesser extent, dominant canopy species [bald cypress, Nyssa sp., and red maple (Acer rubrum L.)] were also common in the sapling/shrub layer (Table 3).
In the nontidal swamps, the sapling/shrub layer was dominated by Fraxinus sp. Unlike tidal forests, saplings in these forests were most likely Carolina ash (Fraxinus caroliniana), a common canopy species [although green ash (Fraxinus pennsylvanica Marsh.) was also observed]. We found little recent recruitment of water tupelo or Ogeechee tupelo, which are two of the dominant canopy species in these nontidal stands (Anderson and Lockaby, 2011c). The lack of tupelo and the dominance of ash in the sapling/shrub stratum may reflect a shift in species composition although the cause is uncertain. Some canopy species shifts have been noted further up the Apalachicola River associated with the lower water levels and decreased flows (Darst and Light, 2008), but we cannot attribute our observations to these factors. The current dominance of canopy-sized tupelo may reflect earlier selection logging of cypress, the capacity of tupelo to resprout from stumps, or both (Penfound, 1952; Sharitz and Mitsch, 1993). Cypress was certainly more dominant in the past based on our frequent sightings of large, decayed cypress stumps throughout the study area.
Tidal wetland forests had significantly greater herbaceous vegetation cover than nontidal wetland forests (72 [+ or -] 4 v. 19 [+ or -] 4%; P = 0.005, F = 13.89, df = 1). Plant community cover and species richness has been shown to increase overall habitat diversity while also influencing primary productivity and nutrient cycling (Harris and Gosselink, 1990). Herbaceous cover in tidal swamps was generally dominated by Panicum gymnocarpon which had an average importance value of 20.8. Herbaceous cover in nontidal swamps was often sparse and dominated by tree seedlings [planer elm (Planera aquatica Walt. ex J.F. Gmel.), ash, bald cypress] (Table 3). Species richness per [m.sup.2] was significantly higher in tidal wetlands (6.27 [+ or -] 0.95) than in nontidal wetlands (3.84 [+ or -] 0.58) (P = 0.033, H = 4.55, df = 1). Trends related to herbaceous vegetation cover and richness were consistent somewhat with other reports of freshwater tidal swamps in the southeastern U.S. (Baldwin, 2007; Hackney et al., 2007). Cover and species richness varied considerably among tidal forest stands. Many nontidal swamps were sparsely vegetated in the understory, which likely reflects greater shading and deeper flooding compared to their tidal counterparts. In many non-tidal swamps, the only understory vegetation observed were tree seedlings as reflected by the high importance values reported for tree species (Table 3). While herbaceous cover was not measured again, nearly all these seedlings were gone the following year probably because of subsequent flooding.
The amount of year to year change that occurs in herbaceous vegetation in these tidal forests is unknown. Species turnover in freshwater tidal swamps appears to be common and annual variation in salinity may play an important role (e.g., Hackney et al., 2007). We documented numerous herbaceous and sapling/shrub species in the tidal swamps that are considered somewhat salt tolerant based on their occurrence in brackish habitats [e.g., alligator weed (Alternanthera philoxereoides Greisb.), saw grass (Cladium jamaicense Crantz), groundsel bush (Bacharris halimifolia L.), wax myrtle, cabbage palm] (Tiner, 1993). Although none of these species were among the most prominent, we would expect these species to become more common in drought years or as sea levels continue to rise.
Coarse woody debris biomass.--Downed CWD biomass for the size class >7.62 cm ranged from 1.61 to 9.16 Mg [ha.sup.-1] across our forest plots and was significantly lower on tidal plots than on nontidal forests (P = 0.028, F = 4.87, df = 1; Fig. 2). This is at least partially because of the larger trees in the nontidal swamp (Table 1), but these swamps also tended to occur along the main channel of the river and along larger tributary creeks where there may be washed in woody debris. Although this material may breakdown slowly, it is an important component of forest food webs and nutrient cycles (Harmon et al., 1986). Downed woody debris biomass for the size class <7.62 cm ranged between 0.51 and 3.27 Mg [ha.sup.-1] across forests and we found no significant difference between forest types. Although smaller in biomass, this material is also important to wetland nutrient dynamics as it will release nutrients more rapidly from decomposition than larger woody debris (Stevens, 1997). Other work related to this study found a significant relationship between the amount of smaller woody debris and soil carbon, nitrogen, and phosphorous (Click, 2011).
Our estimates of downed CWD biomass fell at or below the low end of the range for mature deciduous temperate forests (4 to 40 Mg [ha.sup.-1]) (Tritton, 1980; Macmillan, 1981; Gore and Patterson, 1986; Muller and Liu, 1991) or other floodplain studies (3.8 to 22.5 Mg [ha.sup.-1]) (Ellis et al., 1999). It is unclear if our estimate of CWD is attributable to rapid decomposition in flooded environments, periodic transport from floods, or both. The percentage of total above-ground storage of P and N in CWD can range from 1 to 16.5% and 2 to 21%, respectively (Duvigneand and Denaeyer-DeSmet, 1970; Greenland and Kowal, 1960; Sollins et al., 1980). Our data suggested that that nontidal forests store more nutrients in CWD than tidal forests. These nutrients could be effectively stored in the wetland for several years because nutrients are slowly released from larger CWD over time.
It is well established that many wetland functions are closely tied to forest structural attributes. For instance tree size, tree density, species composition, and the range of vegetation strata provide various potential wildlife habitats (Hunter, 1990; Wakeley and Roberts, 1996). Similarly, wetland elemental cycling is dependent on forest structure measures such as basal area, tree mortality, species composition, understory vegetation, and CWD (Day, 1979; Sharitz and Mitsch, 1993). Along the lower reach of the Apalachicola River, floodplain forested wetlands transition from nontidal to tidal conditions. Along this transition, we found significant shifts in wetland forest structure that coincided with tidal influence. Consequently, we expect that the different attributes of tidal and nontidal wetlands combine to enhance the functional diversity of the lower Apalachicola River. As natural resource managers seek to better understand and manage coastal wetlands, a better understanding of the structure and functions provided by these wetlands will be necessary.
Acknowledgments.--Numerous individuals assisted with the collection of data for this project including: Ana Cerro, Robin Governo, Rachel Jolly, Tug Kesler, Lauren Levi, Helen Light, Jennifer Mitchell, Jennifer Morse, Wayde Morse, John Timpone, Jennifer Trusty, and Jennifer Wanat. The Apalachicola National Estuarine Research Reserve provided boat access to sites and housing to visiting researchers. This study was funded by the Auburn University Center for Forest Sustainability.
SUBMITTED 19 MARCH 2012
ACCEPTED 5 SEPTEMBER 2012
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CHRISTOPHER J. ANDERSON, B. GRAEME LOCKABY AND NATHAN CLICK (1)
School of Forestry and Wildlife Sciences, Auburn University, Auburn, Alabama 36849
(1) Present address: Kentucky Transportation Cabinet, 200 Mero Street, Frankfort, KY 40601
TABLE 1.--Mean ([+ or -] SE) density and relative frequency of tree size classes at tidal and non-tidal forest stands along the lower Apalachicola River. Significant differences between wetland hydrologic groups detected at P < 0.05 No. of stems [ha.sup.-1] Tree DBH class (cm) Tidal Non-tidal P (F, df) 2.5-5.0 699 [+ or -] 112 359 [+ or -] 54 0.013 (14.17,1) 5.0-10.0 483 [+ or -] 72 233 [+ or -] 39 0.004 (22.67,1) 10.0-15.0 201 [+ or -] 20 165 [+ or -] 29 NS 15.0-20.0 52 [+ or -] 11 136 [+ or -] 15 0.007 (18.16,1) >20.0 11 [+ or -] 6 69 [+ or -] 7 0.003 (24.22,1) TOTAL 1446 [+ or -] 159 962 [+ or -] 100 0.008 (17.41,1) % Tree DBH class (cm) Tidal Non-tidal P (F, df) 2.5-5.0 46.0 [+ or -] 4.2 36.9 [+ or -] 3.8 NS 5.0-10.0 33.3 [+ or -] 2.7 23.9 [+ or -] 1.8 0.011 (14.78, 1) 10.0-15.0 15.6 [+ or -] 2.0 17.0 [+ or -] 2.6 NS 15.0-20.0 4.2 [+ or -] 0.8 14.2 [+ or -] 1.1 0.001 (42.47, 1) >20.0 0.9 [+ or -] 0.5 8.3 [+ or -] 1.4 0.025 (8.99, 1) TOTAL TABLE 2.--Mean ([+ or -]SE) annual basal area increment (BAI) of tidal and non-tidal forests (2007-2010) BAI ([m.sup.2] [ha.sup.-1] [yr.sup.-1]) 2007-2008 2008-2009 2009-2010 3-yr mean Tidal 0.30 (0.05) -0.07 (0.26) 0.03 (0.17) 0.09 (0.16) Non-tidal 0.52 (0.08) 0.07 (0.27) 0.55 (0.22) 0.38 (0.19) BAI % 2007-2008 2008-2009 2009-2010 3-yr mean Tidal 0.85 -0.21 0.09 0.24 Non-tidal 0.82 0.11 0.85 0.59 TABLE 3.--Mean ([+ or -] SE) sapling/shrub stem count, sapling/shrub species richness (SR), herbaceous cover, herbaceous SR, and dominant species importance value (IV) in tidal and non-tidal forest stands Saplings and shrubs Stems (no. [ha.sup.-1]) SR Tidal 549 ([+ or -] 136) 6.9 ([+ or -] 1.0) Non-tidal 210 ([+ or -] 72) 3.0 ([+ or -] 0.6) Saplings and shrubs Dominant species (stems [ha.sup.-1]) Tidal Morella cenfera (176) Fraxinus sp. (117) Taxodium distichum (47) Cephalanthus occidentalis (32) Nyssa sp. (29) Magnolia virginiana (25) Non-tidal Fraxinus sp. (109) Planera aquatica (28) Taxodium distichum (25) Quercus sp. (17) Cephalanthus occidentalis (9) Acer rubrum (4) Herbaceous vegetation Cover (%) SR Tidal 72 ([+ or -] 4) 6.27 ([+ or -] 0.95) Non-tidal 19 ([+ or -] 7) 3.84 ([+ or -] 0.58) Herbaceous vegetation Dominant species (IV) Tidal Panicum gymnocarpon (20.8) Crinum americanum (7.8) Sabal palm (5.1) Leersia sp. (5.0) Saururus cernuus (4.4) Justicia ovata (3.1) Non-tidal Nyssa sp. (18.9) Planera aquatica (14.9) Fraxinus sp. (13.8) Taxodium distichum (10.6) Saururus cernuus (5.3) Panicum gymnocarpon (5.3)
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|Author:||Anderson, Christopher J.; Lockaby, B. Graeme; Click, Nathan|
|Publication:||The American Midland Naturalist|
|Date:||Jul 1, 2013|
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