Altered ecosystem nitrogen dynamics as a consequence of land cover change in tallgrass prairie.
In the last several decades, the ecological impacts of land cover and land-use change have received increasing attention. Land cover and land-use changes can have a wide variety of ecological effects (Meyer and Turner, 1992; Vitousek et al., 1997), including alteration of ecosystem functioning (e.g., biogeochemical fluxes) (Houghton et al., 1999; Tilman et al., 2000) and the distribution and diversity of species (Dobson et al., 1997; Chapin et al., 1998). An important land cover change occurring in grasslands worldwide is an increase in the abundance and cover of woody plant species (Archer, 1995; McPherson, 1997; Gill and Burke, 1999; Briggs et al., 2005). Hypothesized drivers for this vegetation shift include climate change (Archer, 1993), increasing atmospheric C[O.sub.2] concentration (Polley et al., 1994; Bond and Midgely, 2000), and changes in land use including fire suppression and increased grazing intensity (McPherson et al., 1993; Scholes and Archer, 1997; Brown and Archer, 1999; Van Auken, 2000; Briggs et al., 2002). Some of the ecological consequences of woody vegetation expansion into grasslands include losses of grassland biodiversity and productivity (Hoch and Briggs, 1999; Van Auken, 2000), increased soil nutrient availability and cycling (Hibbard et al., 2001), decreased soil C[O.sub.2] flux (Smith and Johnson, 2004) and altered soil C storage (Jackson et al., 2002) and increased nutrient (especially C) storage in woody biomass (Gill and Burke, 1999; Houghton et al., 1999; Tilman et al., 2000; Norris et al., 2001b).
In the productive tallgrass prairies of central North America, there have been two predominant patterns of increased woody vegetative cover (Briggs et al., 2005). First, riparian gallery forests have expanded farther up slope in stream valleys, increasing in cover in some areas by 69% from 1939 to 2002 (Knight et al., 1994; Briggs et al., 2005). Second, reductions in fire frequency have facilitated the establishment and spread of woody vegetation including several species of shrubs and eastern redcedar (Juniperus virginiana L., redcedar hereafter) into drier upland sites that were historically grass dominated (Bragg and Hulbert, 1976; Towne and Owensby, 1984; Engle et al., 1988). In these grasslands, local plant community dynamics are controlled largely by disturbance regimes including grazing and burning (Hartnett and Fay, 1998; Bakker et al., 2003). Frequent fires in these grasslands hinder woody plant establishment and growth, enhance productivity of the dominant C4 grasses, and affect species diversity (Bragg and Hulbert, 1976; Knapp and Seastedt, 1986; Gibson and Hulbert, 1987; Collins, 1992; Briggs and Knapp, 1995).
Redcedar is characterized by rapid growth and high reproductive output, which enables the development of monospecific, closed-canopy stands in as little as 40 y (Blan, 1970; Hoch and Briggs, 1999; Briggs et al., 2002). This has contributed to substantial increases in abundance and cover of redcedar, both locally and regionally throughout the eastern Central Plains. Schmidt and Leatherberry (1995) noted the rapid expansion of redcedar in four states of the lower Midwest and estimated that it occurred in nearly 2 m ha of forestland in 1993, a 113% increase since the previous inventory just two decades earlier. In an 8000 ha study region of eastern Kansas, Hoch (2000) projected that closed-canopy redcedar forests could increase 34% in cover from 1997 to 2011 based on rates of expansion in the recent past. The expansion of redcedar cover may result in additional loss of native tallgrass prairie and valuable grazing lands, with concomitant changes in ecosystem processes including soil nutrient cycling and availability, which have potential feedbacks to plant community composition and primary productivity.
This research focused on documenting some of the biogeochemical consequences of the conversion of tallgrass prairie to eastern redcedar forest. Our specific objectives were to: (1) quantify key aspects of ecosystem N pools and fluxes such as plant biomass N, forest floor N, total and available soil N and ecosystem-level nitrogen use efficiency (NUE) in recently established, closed-canopy redcedar forests and (2) compare these measurements with similar data collected in paired adjacent grassland sites (soil N pools and fluxes) or with data from similar grasslands at the nearby Konza Prairie Biological Station (ecosystem-level N budgets and NUE).
This research was conducted in the Flint Hills of eastern Kansas, where the landscape consists of stream-dissected hills eroded in a complex dendritic pattern. Vegetation in the Flint Hills reflects a complex dynamic affected by geology, climate, fire, grazing and other human activities. The native vegetation of the region is primarily tallgrass prairie, dominated by warm-season ([C.sub.4]) grasses (e.g., Andropogon gerardii, Sorghastrun nutans, Schizachyrium scopanus, Panicum virgatum), which co-occur with a diverse assemblage of cool season ([C.sub.3]) graminoids and forbs (Freeman, 1998). Bands of deciduous forest border streams in the valley bottoms, and shrub thickets occur at the upper reaches of draws and on slopes around limestone outcrops (Freeman, 1998). Soil properties of the region are most influenced by prairie vegetation and organic matter accumulation. In general, soils are characterized as cherty silty clay loams (Ransom et al., 1998). The climate is characterized as temperate mid-continental. Mean annual temperature is 13 C. Annual precipitation is 835 mm of which about 75% occurs during the growing season. This amount of precipitation is sufficient to support the growth of woody vegetation and, as a result, land-use practices play a major role in affecting land-cover (Briggs et al., 2005).
Three paired forest-grassland study sites in the Flint Hills were selected within a 40 km radius of Manhattan, Kansas, and have been previously described by Norris et al. (200la, b). Each study site included a closed-canopy redcedar forest (Table 1) paired with an adjacent grassland site. The forest stands at these sites ranged in age from approximately 35 to 80 y old, as determined by both historical aerial photographs and tree ring analysis (Hoch, 2000). Based on historical photographs and [sup.13]C analysis of soil organic carbon pools (Smith and Johnson, 2003), these sites were grass-dominated prior to redcedar encroachment, which presumably occurred as a result of changes in land use (particularly reductions in fire frequency). These forests, representing the current endpoint (most advanced stage) of redcedar encroachment into tallgrass prairie, are now dominated by eastern redcedar with very little understory vegetation. Herbaceous biomass in the forests averaged only 0.18 g/[m.sup.2] (Briggs et al., 2002).
The individual grassland sites paired with each forest site were used primarily for soil N measurements in this study. These sites were located adjacent to a forest site, usually separated by only a fence. Each was grazed by cattle and frequently burned (approximately once every 3 y), representing the predominant land-use for grasslands in this region. Plant communities were characteristic of the region, dominated by C4 grasses co-occurring with a variety of C3 grasses and forbs (Freeman, 1998). By choosing paired forest and grassland sites that were adjacent to one another and on similar soils and topographic positions (slope, elevation, aspect), and with historically similar vegetative cover, we assumed that any differences in soil N pools or processes between forest and grassland represented a change that occurred as a consequence of land-use change (e.g., reductions in fire frequency) and the associated change in land cover, from grasslands to redcedar forests. Because the paired grassland sites were all grazed, which precluded estimating ANPP and potential biomass accumulation, we used data from ungrazed upland tallgrass prairie at the nearby Konza Prairie Biological Station for comparisons of potential ANPP and standing stocks of plant and litter N in redcedar forests and grasslands. Detailed descriptions of the Konza Prairie site are available in Knapp et al. (1998).
SOIL N FLUXES
At each forested site, soil sampling was done in six plots (2 x 2 m) spaced evenly along a 200 m transect bisecting the interior portion of the forest stand. Transects of equal length were established in each adjacent grassland site, in areas with similar soil type and topographic position relative to the paired forest transect. Extractable inorganic soil N and in situ net N mineralization rates were measured in each of the six sample plots (n = 4 cores per plot, 24 per transect) year round from May 1998 through Dec. 1999 using an intact soil core incubation method (Blair, 1997 modified from Raison et al., 1987). Soil cores (5 cm diameter, 10 cm deep) were collected for measurement of initial inorganic N concentrations, and adjacent cores, contained in polyvinyl chloride (PVC) sleeves, were incubated in the field for approximately 30-d intervals during the growing season (May through Oct.) and longer periods (2-3 mo) during the remainder of the year. A soil depth of 10 cm was used because this was the maximal depth that could be consistently sampled at all sites due to the occurrence of shallow limestone layers. The upper 10 cm in these grasslands typically includes 44% of grassland belowground plant biomass (Rice et al., 1998) and the greatest concentrations of soil organic matter and microbial biomass (Dodds et al., 1996). Before collecting initial soil samples or installing in situ cores, identifiable surface litter (Oi horizon) in the forest sites was excluded from both the initial samples and the incubated soil cores to facilitate comparisons with mineralization rates in grassland sites, where surface litter was generally absent due to frequent fires and grazing. Comparisons of forest cores with and without surface litter horizons indicated no significant differences in net N mineralization rates (McKinley, 2006). Field-moist soil from both the initial and incubated cores was sieved (4 mm screen) and subsamples extracted with 2 mol/L KC1. Concentrations of inorganic N (NH4-N and NO2/NO3-N) in the extracts were measured colorimetrically using an Alpkem FlowSolution analyzer. Additional subsamples of soil were dried at 60 C to determine gravimetric soil water content and extractable soil N concentrations were converted to p.g N/g dry soil. Daily net N mineralization rates for specific incubation periods were calculated as ([N.sub.F] - [N.sub.I])/days, where NF is the final concentration of total extractable N (N[H.sub.4]-N + N[O.sub.3]-N) in post-incubation PVC cores and NI is the initial concentration of total extractable N in the adjacent core taken at the start of the incubation period. Cumulative N mineralized over the growing season was calculated by summing net N mineralization rates from specific incubation periods, and extrapolating values for the short periods between incubations. Extractable N concentrations and net N mineralization rates were converted to a kg/ha basis using site-specific bulk density measurements (Smith and Johnson, 2003).
Standing stocks of N in plant biomass, litter and soil were quantified in the forest sites and compared either to data collected from the adjacent grassland sites as part of this study (soil N), or to published data from the nearby Konza Prairie Biological Station (plant biomass and litter N). Data from Konza Prairie Biological Station provided an index for potential plant productivity and accumulation of biomass N in the absence of grazing. Allometric equations relating diameter at breast height (dbh) to aboveground biomass N (Norris et al., 2001b) were applied to three 0.02 ha plots equally spaced along the sample transects in each forest site to determine standing stocks of aboveground tree biomass N. We did not measure belowground biomass directly in the forest sites, but estimated redcedar belowground biomass assuming a root:shoot ratio of 0.25 (Cairns et al., 1997) and applying a mean root N concentration of 0.50% (Norris et al., 200la). Estimates of forest floor ([O.sub.i] horizon) N mass were obtained by collecting twenty 0.1 [m.sup.2] litter samples consisting of identifiable litter along each sample transect (corresponding to the litter removed from soil cores for inorganic N analysis). Total soil N to a depth of 10 cm was quantified in both forest and grassland sites using soil samples from the initial N mineralization cores collected on two dates (May 1998 and Oct. 1999). Litter and soil samples were dried for 48 h at 60 C, weighed, ground and analyzed for N content on a Carlo Erba NA 1500 Analyzer (Carlo Erba, Milano, Italy).
NITROGEN USE EFFICIENCY
In order to compare forests and grasslands developed under the same climate and on similar soils, we calculated nitrogen use efficiency (NUE) for each forest site using data collected in this study and compared their NUE to that of a typical grassland site in this region using data from the Konza Prairie Biological Station. Ecosystem-level NUE was calculated two ways, using ANPP:litterfall N, an index of the amount of aboveground biomass produced per unit of N lost (Vitousek, 1982; Reich et al., 1997) and as ANPP:soil net N mineralized, an index of aboveground biomass produced per unit soil N available (Reich et al., 1997). In the forest sites, ANPP was calculated using allometric regression equations (Norris et al., 2001b) and litterfall N determined through the use of litter collectors and analysis of subsamples for N content (Norris et al., 200la). Grassland NUE was calculated using average ANPP (obtained by harvesting end-of-season aboveground biomass), litterfall (collected using 1 m long troughs on the soil surface) and standing dead N values from an annually burned upland site at Konza Prairie Biological Station (Seastedt, 1988; Knapp et al., 1998; Blair et al., 1998). Both surface litterfall N and standing dead N were used since both represent sources of annual N loss via litter production in grasslands.
Soil N measurements (total N, inorganic N, and mineralization rates) within a particular site were characterized by high spatial variability. For these measurements, we used median values from each site for our analyses to diminish the influence of extreme values (Zar, 1996). A repeated measures ANOVA was used for the inorganic N and mineralization data to determine land cover effects over the course of the entire study. A mixed model was used to incorporate the random effect of blocking on location with the fixed effect of land cover type (Proc mixed, SAS Institute, Cary, NC). The analysis included the determination of an appropriate covariance structure while accounting for unequal spacing between sample dates (Littell et al., 1996). Differences between forests and grasslands at each date were evaluated by t-tests using a Bonferroni
correction. Soil moisture was considered as a possible covariate; however, including it did not aid in the analysis. Results are reported in the text as significant where P < 0.05.
SOIL NITROGEN AVAILABILITY
Mean measurements of soil N pools and fluxes in both the grassland and forest sites were characterized by a large amount of spatial heterogeneity. Across all dates, the effect of land cover type on inorganic N concentrations was only marginally significant (P = 0.0889); however, there was a fairly consistent trend for inorganic N to be slightly higher in grassland sites on most sample dates (Fig. 1A). The greatest margin of difference between net N mineralization rates in forest and grassland sites occurred in May of both years when the grasslands exhibited greater mineralization rates than forested sites (Fig. 1B); however, this was only significant on one date (May 1999, P = 0.0237) and there was no overall effect of land cover type on mineralization rates across all dates. Cumulative net N mineralized was calculated at the end of the first year (May 1998-Apr. 1999) and again after the second growing season (May-Oct. 1999) (Fig. 1C) and although not significantly different, forest sites mineralized on average 15% less N (14.3 kg N x [ha.sup.-1] x [yr.sup.-1]) than adjacent grassland sites (16.9 kg N x [ha.sup.-1] x [yr.sup.-1]). Values for both the forest and grassland sites are within the range previously reported for tallgrass prairie sites at the nearby Konza Prairie Biological Station (Blair, 1997; Turner et al., 1997).
Redcedar forests stored considerable quantities of N in aboveground biomass and surface litter compared to the grassland sites they replaced. The forest sites had 20-27% of total ecosystem N in plant biomass (depending on stand age), most of which was aboveground (Fig. 2). This was substantially more than the proportion of ecosystem N sequestered in plant biomass in ungrazed grassland sites (approximately 3.6%). In forest sites, we estimated mean aboveground biomass N to be 617 kg N/ha (Norris et al., 2001b), which was approximately 15 x greater than typical shoot N in local tallgrass prairie (40 kg N/ha, Blair et al., 1998). Additionally, litter accumulations on the forest floor ([O.sub.i] horizon) contained, on average, 253 kg N/ha, which was substantially more than the litter pool in grasslands, which is subject to frequent removal by burning (Norris et al., 200la). Averaged across all three sites, total soil N content (0-10 cm) was not significantly different between forest and grassland sites although the mean proportion of total ecosystem N residing in the soil was higher in grasslands (96.4%) than in forests (77.6%).
NITROGEN USE EFFICIENCY
Perhaps the greatest difference between the grassland and forest sites was their nitrogen use efficiency (NUE). Estimates of ecosystem-level NUE, calculated as ANPP (kg x [ha.sup.-1] x [yr.sup.-1]) per unit N returned in litterfall (kg N x [ha.sup.-1] x [yr.sup.-1]), were 223 ([+ or -]21) for the redcedar forest sites and 93 for grassland sites (no standard error is included for grasslands sites, since we used the long-term mean of ANPP at upland, burned grassland sites and a mean annual litterfall value for this calculation). Estimates of NUE calculated as ANPP:soil N mineralized (kg N x [ha.sup.-1] x [yr.sup.-1]) were 1151 ([+ or -]459) for forest and 256 ([+ or -]62) for grasslands. Differences in NUE between forest and grassland were due to the much greater average ANPP in redcedar forests (9162 kg x [ha.sup.-1] x [yr.sup.-1], Norris et al., 2001b) relative to comparable upland grassland sites (3700 kg x [ha.sup.-1] x [yr.sup.-1], Knapp et al., 1998) coupled with the relatively minor differences in litterfall N and soil N mineralized.
Despite the extreme land cover change associated with the coniferous afforestation of tallgrass prairie, changes in soil nitrogen availability and net N mineralization in surface soils of recently established stands ([less than or equal to] 80 y) of closed canopy redcedar forest were relatively small compared to adjacent grasslands sites. Although there were only a few statistically significant differences between forest and grassland sites for either measure of soil N availability (extractable inorganic N or net N mineralization) across all sites, the data did suggest a modest reduction in N availability as a consequence of conversion of grasslands to forests, driven primarily by early season differences between land cover types (Fig. 1). It is important to note that grassland management may play a role in affecting these comparisons. The paired grasslands used in this study were all subject to long-term grazing, which is a dominant land-use for native prairie in the Flint Hills region. Grazing has been shown to significantly increase N availability in these grasslands, relative to non-grazed tallgrass prairie (Johnson and Matchett, 2001). Thus, the relative comparisons of soil N availability were likely affected by both differences in land-use (i.e., fire and grazing) as well as the changes in land-cover type that occur once fire and grazers are excluded from these grasslands.
[FIGURE 1 OMITTED]
[FIGURE 2 OMITTED]
The relatively small changes in soil N availability observed in this study contrasts with other studies of woody encroachment into North American grasslands, which often indicate significant increases in rates of nutrient cycling, availability, and accumulation (e.g., Schlesinger et al., 1990; Wilson and Kleb, 1996; Hibbard et al., 2001); however, there are limited studies in temperate grasslands that involve coniferous species and an evaluation of soil nutrient consequences. In southern Texas (Brown and Archer, 1989; Archer, 1995; Hibbard et al., 2001) and in Arizona (MacPherson et al., 1993), land cover change has been well documented following the invasion of honey mesquite (Prosopis glandulosa Torr.). This leguminous species obtains at least 50% of total N requirements from symbiotic [N.sub.2]-fixation (Rundle et al., 1982; Boutton et al., 1992) and has a foliar N concentration of 5.2% (Barnes and Archer, 1996), more than 4 x the live foliage nitrogen content of redcedar (1.25%). Despite increases in foliar inputs and litter quality, a modeling exercise showed that landscape changes in soil carbon associated with the mesquite-dominated uplands did not vary significantly from heavily grazed soils until well after 30 y of woody plant development (Hibbard et al., 2003).
The redcedar stands in our study were relatively young and may continue to alter soil N dynamics as they mature, especially considering the relatively low quality and slow decomposition of redcedar litter (Norris et al., 2001a). The oldest forest site at Tuttle Cr. Dam, double the age of the younger sites, exhibited greater soil N concentrations and mineralization rates relative to the younger forests and usually the grassland sites as well. For instance, average annual net N mineralized at the Tuttle Cr. Dam forest was 27.8 kg N x [ha.sup.-1] x [yr.sup.-1] compared to 8.5 and 6.7 at the other forest sites and 16.9 kg N x [ha.sup.-1] x [yr.sup.- 1] in the grasslands. These data suggest a potential time lag on the scale of decades of soil nutrient responses associated with this land cover change and that these sites have yet to reach equilibrium. Clearly, in order to explain the patterns of change in ecosystem processes associated with redcedar forest expansion, we need a better understanding of plant-soil interactions and the temporal dynamics of this afforestation.
In contrast to the relatively minor changes in soil N pools and processes, there were substantial increases in aboveground biomass and surface litter N pools associated with the conversion of grassland to forest. Likewise, large increases in ANPP (Norris et al., 2001b) led to dramatic increases in NUE in grassland sites converted to redcedar forests, providing an interesting metric for comparisons of these two land cover types. Previous studies of NUE have often found that NUE increases with diminishing N availability (Gosz, 1981; Lennon et al., 1985; Escudero et al., 1992; Reich et al., 1997). However, in this case, differences between forest and grassland do not appear to be linked to differences in soil N availability, but rather to intrinsic characteristics of the species themselves, providing an opportunity to evaluate how changes in plant life form can alter ecosystem-level patterns of NUE on sites of similar fertility and nutrient availability.
The comparatively low NUE in grassland, relative to redcedar forest, is particularly interesting given the efficiency of the [C.sub.4] photosynthetic pathway and ability of the dominant grasses to operate under conditions of low N availability (Knapp and Seastedt, 1986), typically the nutrient most limiting to grassland productivity (Risser and Parton, 1982; Seastedt et al., 1991; Blair et al., 1998) due to a strong interaction between frequent disturbances (e.g., fire, grazing) and N (Ojima et al., 1994; Blair, 1997). Conifers also possess traits that result in a high NUE, including conservative N cycling due to poor litter quality (e.g., low leaf N, high lignin), slow decomposition rates, and high internal redistribution of N (Gosz, 1981; Reich et al., 1992). Leaf longevity is a simple factor related to each of these mechanisms (Chapin, 1980; Reich et al., 1995; Aerts and Chapin, 2000) and it is probable that differences in the leaf life-span of redcedar stands (3-6 y) (Norris, 2000) and tallgrass prairie (5-6 mo) contribute to observed patterns of N cycling and NUE.
It is also interesting to note the changes in total ecosystem nitrogen associated with the conversion of grassland sites to forest. If we examine the major pools and fluxes of N in these mature redcedar forests and in frequently burned grassland (Fig. 2), there was no significant change in total soil N pools, but a substantial increase in N stored in aboveground biomass and the litter layer. We recognize that the changes in aboveground biomass might be offset or augmented by changes in belowground biomass which we did not measure directly. However, our estimates of plant biomass N belowground magnify the differences between N sequestered in plant total biomass in forest and grassland, as redcedar root N mass was estimated to be 150 kg N/ha compared to about 80 kg N/ha for root mass N in native tallgrass prairie. These estimates of root biomass in redcedar forest stands are first approximations that will require verification, but they suggest the importance of both above- and below-ground biomass in N accumulation in new redcedar forests.
Since the establishment of these forests, these sites appear to have accumulated 910 kg N/ha in total ecosystem N (15 to 23 kg N x [ha.sup.-1] x [yr.sup.-1]) relative to grassland. It is not uncommon for aggrading forests to accumulate nitrogen as a result of conservative nutrient cycling (Bormann and Likens, 1979; Pearson et al., 1987). In the case of conversion of tallgrass prairie to redcedar forest, the elimination of fire and associated N losses may account for at least a portion of the accumulated N. Two additional potential explanations for the high rate of N accumulation include enhanced N inputs via [N.sub.2]-fixation (Bormann et al., 1977, 1993) or more efficient capture of atmospheric deposition by coniferous trees (Lovett and Lindberg, 1984, 1993). While we cannot discount these two mechanisms for additional N accumulation, they are believed to be minor relative to other inputs (e.g., bulk precipitation) (Vance et al., 1983; Grant and Binkley, 1987; Heath et al., 1988). The apparent increase in N storage may also reflect a redistribution of existing N pools following the conversion of grasslands to redcedar stands as soil N from deeper depths is depleted and redistributed into plant biomass and litter. This seems unlikely, however, as these forests occurred on relatively shallow soils and we observed no significant difference in total N stored in the upper 10 cm of soil in grassland and forest sites.
This study highlights some important changes in ecosystem-level NUE and nitrogen storage as a result of land-cover change from native tallgrass prairie to redcedar forest. However, understanding the mechanisms underlying these changes in N cycling will require better knowledge of several processes. In particular, root production and turnover have not been quantified directly in comparable grassland and redcedar sites, but are likely responsible for the majority of total N turnover (Cox et al., 1978). In addition, a more detailed assessment of the temporal dynamics of change may provide insights into the apparent differences in N dynamics between forest sites of different age, as suggested by a possible stand age factor in net N mineralization rates. Lastly, identifying the mechanisms by which these forests have accumulated a substantial amount of N relative to the grasslands they replaced may shed light on the controls of biomass and C sequestration in these newly established forests. In summary, the establishment of redcedar forests in areas formerly dominated by native grasslands results not only in decreased biodiversity (Hoch and Briggs, 1999), but also significant changes in productivity (Norris et al., 2001b) and ecosystem N storage. The increases in productivity as well as N and C storage with redcedar encroachment are consistent with the reported effects of other types of woody plant expansion in the Midwestern U.S. (Tilman et al., 2000; Hughes et al., 2006), indicating that this land cover change may have important impacts on regional biogeochemistry (Pacala et al., 2001; Jackson et al., 2002). However, the long-term dynamics of these newly established ecosystems, with respect to disturbance (e.g., fire) and climate (e.g., drought and climate change), are still unknown and deserve further attention, as the current rates of biomass and nutrient accumulation will likely not be sustainable.
Acknowledgments.--This research was supported by a grant from NASA's Land Cover Land Use Change program. This research would not have been possible without the cooperation of property owners granting permission to pursue this research on their land. We thank Cris Oppert and Eric Nun for assistance in the field and lab and Dr. Jeff Pontius for statistical consultation. John Briggs, Peter Reich, Kathy Hibbard, and one anonymous reviewer provided insightful comments on earlier versions of this manuscript. Some grassland data used for comparisons with forest sites was extracted from the NSF-funded Konza Prairie LTER dataset. This is contribution #07-180-J of the Kansas Agricultural Experiment Station.
SUBMITTED 20 SEPTEMBER 2006
ACCEPTED 20 FEBRUARY 2007
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MARK D. NORRIS, (1) JOHN M. BLAIR AND LORETFA C. JOHNSON
Division of Biology, 232 Ackert Hall, Kansas State University, Manhattan 66506
(1) Corresponding author present address: Department of Environmental Science & Biology, SUNY Brockport, 350 New Campus Drive, Brockport, New York 14420; e-mail: firstname.lastname@example.org; Telephone: (585)395-5743; FAX: (585)395-5969
TABLE 1.--Descriptions of eastern redcedar forest study sites, including tree age, density, above-ground biomass, litterfall rates and ANPP (data from Norris et al., 2001a, 2001b) Approx. Tree density Aboveground total Site Location age (yr) (trees/ha) biomass (kg/ha) 1 Scenic Dr. 35 1733 114120 39[degrees]17'N, 96[degrees]34'W 2 Tuttle Cr. Dam 80 860 210952 39[degrees]10'N, 96[degrees]39'W 3 Randolph 40 1900 120738 39[degrees]27'N, 96[degrees]45'W Litterfall production ANPP Site (kg/ha/yr) (kg/ha/yr) 1 5193 9796 2 4503 7247 3 5159 10442
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|Author:||Norris, Mark D.; Blair, John M.; Johnson, Loretfa C.|
|Publication:||The American Midland Naturalist|
|Date:||Oct 1, 2007|
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