Carbon Sequestration by Rangelands: Management Effects …



Carbon Sequestration by Rangelands: Management Effects and Potential

Gerald E. Schuman Justin D. Derner

USDA-ARS USDA-ARS

High Plains Grasslands Res. Stn. High Plains Grasslands Res. Stn.

8408 Hildreth Road 8408 Hildreth Road

Cheyenne, Wyoming 82009 Cheyenne, Wyoming 82009

307-772-2433 ext. 107 307-772-2433 ext. 113

gschuman@lamar.colostate.edu Justin.Derner@ars.

Abstract

Lands grazed by wild and domesticated animals comprise 336 million hectares in the United States. Rangelands account for about 48% of that land area and more than one-third of the world’s terrestrial carbon reserves. Because of the large land area they have the potential to sequester a significant amount of additional carbon from the atmosphere. Grazing lands are estimated to contain 10-30% of the world’s soil organic carbon. Management practices, such as grazing, nitrogen inputs, and improved plant species have been shown to increase soil organic carbon storage in rangelands. Properly managed rangelands of the United States are estimated to have the capacity to store 19 million metric tonnes of C per year. Therefore rangelands can have a major impact in mitigating the effects of elevated atmospheric carbon dioxide levels on global climate change.

Introduction

Rangelands (including grasslands, shrublands, deserts, and tundra) occupy about half of the world’s land area and contain more than 33% of the above- and below-ground carbon (C) reserves (Allen-Diaz 1996). Rangelands account for 48% of the 336 million hectares (Mha) of U.S. grazing lands. Changes in soil C on rangelands can occur in response to a wide range of management and environmental factors. Although the magnitude of these changes per unit of land area are small relative to those reported for croplands and improved pastures, increases in terrestrial C resulting from management or inputs account for a significant amount of C sequesteration and reduction in atmospheric carbon dioxide (CO2) given the size of this land resource. For example, Schuman et al. (2001) estimated that improved management of 113 Mha of poorly managed U.S. rangelands could sequester 11 million metric tonnes (MMT) C annually whereas 13 Mha of Conservation Reserve Program (CRP) lands could sequester 8 MMT C per year. They also estimated potential avoided losses of C to be nearly 43 MMT C/yr by ensuring that (a) well-managed rangelands continue to be grazed and managed properly, (b) rangelands are not broken out and cultivated, and (c) that CRP lands be maintained in perennial grasses and not re-cultivated. These estimates should be viewed as conceptual illustrations that demonstrate the potential of rangelands as a vast terrestrial C pool and should not be ignored when assessing the impact of agricultural management on atmospheric CO2 (Schuman et al. 2001). Follett et al. (2001a) estimated that the net C sequestration by U.S. grazing lands to be 17.5 to 90.5 MMT C/yr with a mean of 54 MMT C/yr. Furthermore, these authors state that most of the potential for soil to sequester C is not being managed for and could be significantly increased by adoption of more intensive management practices. For comparison, recent estimates of the potential soil C sequestration on U.S. cropland soils are 60-70 MMT C/yr assuming producers widely adopt management practices that sequester C (Sperow et al. 2003).

Management Effects on Carbon Sequestration

Rangeland C sequestration research over the past 10-15 years has focused on assessing the effects of management practices on soil C dynamics. Soil organic C reserves in a given rangeland ecosystem will eventually approach a steady state; therefore, a shift in management, environment or inputs would be required to increase the potential for additional soil C sequestration (Schuman et al. 2001). Management practices such as grazing, nitrogen inputs via fertilization and interseeding of N-fixing legumes into rangelands, burning, woody plant encroachment, and restoration of degraded rangelands have been shown to influence soil C sequestration. Below we present the current state of knowledge pertaining to the effects of these management practices on C sequestration which is summarized in Table 1.

Grazing

Grazing of the shortgrass steppe in northeastern Colorado at a moderate stocking rate resulted in increased soil organic C compared to adjacent ungrazed exclosures (Derner et al. 1997). These authors found 19.8 MT C/ hectare (ha) in the surface 15 cm of the soil under grazing and only 13.2 MT C/ha in the ungrazed exclosure, but found no differences in soil C in the 15-30 cm soil depth between grazed and ungrazed areas. Since these pastures had been grazed season-long for 55 years, we estimate that the rate of C sequestration attributable to moderate grazing during this period would be 0.12 MT C/ha/yr. Another study on the shortgrass steppe demonstrated that soil organic C storage increased with heavy (16.9 MT C/ha) grazing, but not with light (14.3 MT C/ha) grazing compared to a non-grazed (13.0 MT C/ha) exclosure (Reeder and Schuman 2002). These pastures and exclosures had been under the existing grazing management practices for 56 years when these assessments were made; thus we estimate the C sequestration rate associated with heavy grazing in the shortgrass steppe would be about 0.07 MT C/ha/yr.

Twelve years of light or heavy grazing in a northern mixed-grass prairie, just 50 km north of the shortgrass steppe site, increased soil organic C in the surface 30 cm of the soil compared to non-grazed exclosures (Schuman et al. 1999). These authors estimated the C sequestration rate to be about 0.30 MT C/ha/yr compared to the ungrazed exclosures. Similar C sequestration rates on northern mixed-grass prairie using CO2 flux data were reported by Frank (2004). He estimated a C sequestration rate of 0.29 MT C/ha/yr for the 6 years of his study. Soil organic C on the northern mixed-grass prairie did not differ between a short-duration rotational grazing system, a rotationally-deferred grazing systems, and continuous season-long system when heavily grazed (Manley 1995). Severe drought and heavy grazing can result in significant losses of soil organic C that was previously stored during normal to above-normal production years in northern mixed-grass prairie (Ingram et al., unpublished data, Morgan et al. 2004). The heavy stocking rate appears to be detrimental in drought years compared to lighter moderate stocking rates. Therefore, additional information is needed to assess the long-term interactions of climate and grazing on C sequestration across a variety of rangeland ecosystems.

The season-long, moderate and heavy stocking rates at the shortgrass steppe and northern mixed-grass prairie resulted in a shift in plant community composition. Season-long grazing at these stocking rates greatly reduced the proportion of cool-season (C3) grasses in these rangeland ecosystems, shifting them to a plant community dominated by the warm-season (C4) species, blue grama (Bouteloua gracilis.) (Schuman et al. 1999, Derner et al, in review). Heavy stocking rates increased vegetative basal cover, canopy cover, amount of bare ground and density of blue grama, but also substantially reduced levels of litter and density of the dominant C3 species western wheatgrass (Pascopyrum smithii) (Derner et al., unpublished data). Reducing the C3 component of the plant community greatly lowers the production potential of these rangelands. Heavy, season-long grazing on the northern mixed-grass prairie site in southeastern Wyoming decreased production by over 36% in just 12 years (Schuman et al. 1999). Therefore, a portion of the increase in soil organic C storage is attributed to this shift in plant community composition. Coupland and Van Dyne (1979) found that blue grama dominated grasslands transfer more of the C to belowground plant parts. Blue grama also has a (a) greater root to shoot ratio, (b) greater proportion of its root system in the surface few centimeters of the soil, and (c) a root system that is much more fibrous than many of the C3 herbaceous species. Similar changes in soil organic C associated with a plant community shift from C3 to C4 species with grazing in a northern mixed-grass prairie in North Dakota and Canada have been reported by Frank et al. (1995) and Dormaar and Willms (1990).

Grazing by sheep in an alpine meadow in the Medicine Bow National Forest in Wyoming increased soil organic C (Povirk 1999). She found that soil organic C levels in ungrazed exclosures averaged 6.3% compared to 11% in the grazed allotment. These mountain meadows are generally only grazed for 1-3 months by domestic livestock, but may potentially be grazed by wild herbivores for a much longer time period.

Ten years of moderate and heavy grazing on southern mixed-grass prairie in Oklahoma reduced concentrations of soil organic C compared to ungrazed areas (Fuhlendorf et al. 2002). However, all of these pastures were grazed at moderately heavy to heavy stocking rates prior to implementation of the treatments.

Extremely light grazing in the Russian steppe for 100 years did not change the soil organic C pool (Torn et al. 2002). Carbon concentrations and pools to 140 cm in the soil profile were remarkably similar between the 1895 to 1903 sampling and the 1997 sampling.

Model estimates of the C dynamics of a semidesert community (mean annual precipitation of 200 mm) dominated by C4 grasses and overgrazed from 1942 to 2000 predicted lower soil organic C content, but a woodland savanna with mean annual precipitation of 358 mm that was overgrazed for the same time period showed a slight increase in soil organic C (Olsson and Ardö 2002). In addition, they determined that a savanna community with a mean annual precipitation of 269 mm would exhibit a stable soil organic C pool with light grazing intensity. They also noted that minor oscillations would be expected with climate fluctuations.

Nine native grassland sites on the southern Canadian prairies were assessed for soil organic C dynamics in grazed and ungrazed treatments (Henderson 2000). He found that organic C tended to be higher in the grazed vs ungrazed treatments, though the effect was significant at only two of the sites in the 0-10 cm soil surface. When evaluated for the entire soil profile (0-105 cm), C storage was dependent upon moisture regime. Soil organic C in semi-arid sites (mean annual precipitation of 328-390 mm) was higher under grazing compared to non-grazed treatments. At sub-humid sites (mean annual precipitation of 476 mm) the response was reversed. Consistent with Henderson’s findings, Derner et al. (in review) determined that moderate grazing, compared to non-grazed exlosures, increased soil organic C in the top 30 cm of the soil profile in a semiarid shortgrass steppe (mean annual precipitation of 321 mm), but reduced soil organic carbon by 7-8% in two more mesic rangeland sites, a southern mixed-grass prairie in west-central Kansas (mean annual precipitation of 588 mm) and a tallgrass prairie in eastern Kansas (mean annual precipitation of 835 mm). Derner et al. (in review) speculated that the changes in soil organic C may have been attributable to a 2-fold greater root mass in the semiarid site compared to the two more mesic sites. This greater root mass increased the ratio of root C to soil organic C (0.20-0.27). Therefore, grazing-induced increases in root mass may have a larger and more immediate effect on soil organic C pools in the semiarid shortgrass steppe; however, shifts in plant community composition due to grazing generally resulted in decreased aboveground plant production (Schuman et al. 1999).

Short-rotation grazing increased soil organic C to a depth of 50 cm by 22% compared to extensive grazing or haying (48.3 vs. 39.5 MT C/ha) in orchardgrass (Dactylis glomerata)-dominated pastures in Virginia (Conant et al. 2003). Soil sequestration rates were calculated to be 0.41 MT C/ha/yr with the short-rotation grazing. Grazing (5-19 yrs) also increased soil organic C compared to haying or ungrazed pastures of bermudagrass (Cynodon dactylon) in the Southern Piedmont of the U.S. (Franzluebber et al. 2000; Franzluebbers and Stuedemann 2001).

Researchers and land managers follow the long standing belief that well-managed grazing stimulates the growth of herbaceous species and improves nutrient cycling in grassland ecosystems. For example, early season (April-June) photosynthesis (as measured by chamber CO2 exchange rates) on grazed northern mixed-grass prairie was greater compared to ungrazed exclosures (LeCain et al. 2000). Researchers have also reported that grazing stimulates aboveground production (Mutz and Drawe 1983; Dodd and Hopkins 1985; Frank and McNaughton 1993; McNaughton et al. 1996) and increases tillering and rhizome production (Floate 1981; Schuman et al. 1990). Grazing has also been noted to stimulate root respiration and exudation rates (Dyer and Bokhari 1976). Nutrient cycling and distribution in the soil profile can be significantly altered by defecation and urination in grazing systems. These factors all likely contribute to the observed increases in soil organic C storage in the upper soil horizons/depths. Grazing processes also impact the rate of turnover/decomposition of the aboveground components (litter and standing dead plant residues) of the plant community. Shoot turnover was estimated to be 36 and 39% under light and heavy grazing compared to 28% in ungrazed exclosures in a northern mixed-grass prairie (Schuman et al. 1999). Animal traffic enhances the physical breakdown, soil incorporation and rate of decomposition of the residual plant material (Naeth et al. 1991; Sharif et al. 1994). Immobilization of C in aboveground plant residue in the ungrazed rangeland may contribute to the lower soil C often observed. Schuman et al. (1999) found that 72% of the aboveground biomass was standing-dead and litter in an ungrazed exclosure and hypothesized that this likely accounted for a portion of the lower soil organic C they observed in the 0-30 cm depth of the exclosure compared to the grazed treatments.

Nitrogen Inputs

Many rangelands are nitrogen (N) deficient and have been shown to exhibit increased production and water-use-efficiency in response to N addition. Addition of N fertilizer to the tallgrass prairie increased plant production and increased soil organic C by 1.6 MT C/ha after 10 years (Rice 2000). Reeder et al. (1998) reported increases in soil organic C of 0.41 and 1.16 MT C/ha/yr in the surface 7.5 cm and 10 cm, respectively, after 4 years of annual applications of 34 kg N/ha on two different CRP sites seeded to a mixture of native C3 grasses. Soil organic C increases of 5.4 to 9.3 MT C/ha occurred on a grassland in north-central Saskatchewan when both N and sulfur fertilizers were applied (Nyborg et al. 1994), and greater increases were observed in Alberta when N was annually applied (Malhi et al. 1991). Schnabel et al. (2001) also found increased soil organic C to a depth of 30 cm when high rates of inorganic fertilizers (336-37-139 kg NPK/ha/yr) were applied to bermudagrass pastures in the Southern Piedmont of Georgia for 15 years compared to low application rates (134-15-56 kg NPK/ha/yr). Application of other nutrients, where they are deficient, can also enhance soil organic C storage (Nyborg et al. 1998, 1999; Conant et al. 2001). However, the benefits of increased soil C sequestration must be compared to the “C costs” of fertilizer production to assess whether there is any net beneficial effects on the atmosphere (Schlesinger 1999).

Introduction of legumes into native rangelands of the Northern Great Plains has been the subject of research for years (Berdahl et al. 1989; Heinrichs 1975; Kruger and Vigil 1979; Tesar and Jacobs 1972; Waddington 1992). Kruger and Vigil (1979) estimated that 70% of the native rangelands in North America were in fair to poor condition and that 10 Mha of Canadian rangelands could benefit from the introduction of a legume into the system. Interseeding of yellow-flowered alfalfa (Medicago sativa ssp. falcata) into northern mixed-grass prairie increased organic C storage by 4, 8, and 17% in a 1998, 1987, and 1965 interseeding, respectively (Mortenson et al. 2004). This resulted in C sequestration rates of 1.56, 0.65, and 0.33 MT C/ha/yr, for the 1998, 1987, and 1965 interseeding (after 3, 14, and 36 years), respectively. These data demonstrate that C sequestration rates will be greater immediately after initiation of a new management practice because of the lower inherent soil organic C levels. Nitrogen fixation by the yellow-flowered alfalfa significantly increased soil total N and aboveground production (Mortenson 2003). This increase in production accounts for the enhanced soil organic C storage and does not represent any “C costs” in the production of the N, nor does it appear that the greater soil N increases nitrous oxide emission from soils that would offset the benefits to the atmosphere resulting from the increased soil C sequestration (Schuman et al. 2004).

Fire

Fire is a management tool and is a natural-occurring ecological process that influences the structure and function of rangelands. Historically, fire controlled the spread of woody species on rangelands. The lack of fire and/or changes in fire return intervals has appreciably altered rangeland plant communities. For example, reductions in fire frequency and increases in return intervals in the last several decades have facilitated the encroachment and spread of undesired woody species in many rangelands throughout the world.

Annual burning and moderate grazing of the tallgrass prairie in Kansas increased soil organic C storage by 2.2 MT C/ha after 10 years, which represents a C sequestration rate of 0.22 MT C/ha/yr (Rice 2000), whereas annual burning increased root growth by 25% compared to unburned areas and heavy grazing decreased root growth by 30% (in the same tallgrass prairie in Kansas) (Johnson and Matchett 2001). Additionally, these authors demonstrated that root quality decreased with annual burning as C:N ratio increased to 60, but grazing only increased these ratios to 40. Therefore, it appears that the combination of annual burning and moderate grazing may offset one another with net results of an equilibrium with regards to soil organic C. Burning can produce charcoal, a form of C very resistant to decomposition, which can account for a significant portion of the stored soil organic C in rangelands (Skjemstad et al. 1996). Fire management can also influence the amount of C stored in biomass by altering the density or encroachment of woody species (Sampson and Scholes 2000).

Restoration of degraded rangelands

Degraded rangelands, whether degraded by overgrazing, mineral extraction, or fire, have a significant potential for loss of soil organic C through erosion, enhanced decomposition of soil organic matter, and dilution of soil organic C through mixing of surface soil horizons.

The effects of restoring degraded southern mixed-grass prairie and grazing intensity (none, moderate and heavy) on soil organic C were evaluated by comparing soils from cultivated areas reseeded 30 to 50 years earlier to native grasses and native prairies that had not been previously cultivated (Fuhlendorf et al. 2002). Concentrations of soil organic C in the top 10 cm of the soil profile did not differ between reseeded and native prairies when grazed at a moderate intensity; however, heavy grazing reduced soil organic C by 65% in the reseeded compared to the native prairie. Olsson and Ardö (2002), using modeling, estimated that conversion of marginal agricultural lands to rangeland in Sudan would restore soil organic C levels to 80% of those found in native savannas in 100 years.

Soil organic C to a depth of 20 cm was 4.7 times greater on a restored site (70 MT C/ha) compared to a degraded semi-arid savanna in the western Chaco of Argentina (15 MT C/ha) (Abril and Bucher 2001). The authors estimated that the C sequestration rate for the highly restored site was 0.28 MT C/ha/yr.

To prevent potentially confounding effects of surface litter on soil organic C measurements, soil organic C pools of the 5-10 cm soil depth were determined for CRP fields in Nebraska that were either recently planted with grass mixtures or planted ten years previously (Baer et al. 2000). Although soil organic C did not differ between the two treatments, both plantings had 24-30% less soil organic C than the native prairie. However, Follett et al. (2001b) and Gebhart et al. (1994) both showed significant C sequestration within 5-10 years after CRP establishment. Gebhart et al. (1994) found soil C sequestration in CRP lands to range from 0.8 to 1.1 MT C/ha/yr in the 0-40 and 0-300 cm soil depth in Texas, Kansas and Nebraska. Follett et al. (2001b) reported C sequestration in the 0-20 cm depth of CRP land from Texas to North Dakota to average 0.9 MT C/ha/yr. Using these C sequestration rates he estimated that the CRP land area in the Great Plains could account for about 20% of agriculture CO2 emissions.

Soil organic C in the top 10 cm of the soil profile in shortgrass steppe was lowest in soils cultivated for > 50 years, intermediate for soils which had been cultivated but abandoned for >50 years and highest for native shortgrass steppe (Burke et al. 1995). Soil organic C was 20% greater in soils that had successional vegetation following abandonment of cultivation, yet these soils contained only 67% of the soil organic C found in the native shortgrass steppe, suggesting that there is considerable reservoir capacity to sequester additional C in these regenerating soils.

Similarly, Potter et al. (1999) investigated soil organic C storage in a tallgrass prairie in Texas 6, 26, and 60 years after grass re-establishment. Concentrations of soil organic C were intermediate in the restored grasslands compared to the native prairie (highest) and cultivated (lowest) soils. They estimated a C sequestration rate of 0.45 MT C/ha for the restored grasslands, and indicated that it would take nearly a century for these grasslands to reach soil C levels found in the native prairie.

Rangelands in the Great Plains disturbed by surface mining and then revegetated have a great potential for C sequestration because of the soil salvage process. Soil salvage results in the collection of at least the two surface soil horizons and in some cases stockpiling of this material for several years. This process results in the dilution of the soil organic C pool by mixing of the organic matter rich surface horizon with the lower soil horizons which are much lower in soil organic matter (Woods and Schuman 1986; Schuman 2002). This results in a soil material that is very similar to dryland cropland soils that have been under tillage for over 50 years (Haas et al. 1957, Tiessen et al. 1982, Burke et al. 1989, Bowman et al. 1990); therefore, these soils have a soil organic C sequestration potential similar to marginal, highly erodible croplands that have or are being put into CRP. Stockpiling soil also greatly diminishes the quality of the soil through enhanced organic matter degradation/decomposition and general loss of much of the microbial functions (Harris et al. 1989, 1993; Severson and Gough 1983). Use of alternative plant growth materials are also sometimes used in place of topsoil which can result in soil or mine spoil material with very limited soil organic C and also limited nutrient cycling potential (Schuman and Taylor 1978, Woods and Schuman 1986). Reclaimed mine soils (0-15 cm depth) in Wyoming have exhibited increases of about 400% over a 30 year period, increasing from 24.2 MT C/ha on soils reclaimed in 1999 to 82.5 MT C/ha on lands reclaimed in 1969, compared to native rangeland levels of 23.2 MT C/ha (Stahl et al. 2003). They hypothesized that the decomposition rates are slow in reclaimed mine soils due to the low microbial activity observed in the reclaimed mine soils which accounts for the large percentage increase in C.

.Woody plant encroachment

Soil organic C was examined along a rainfall and land-use gradient in southern Africa in which the relative abundance of trees decreased with lower rainfall (Feral et al. 2003). Concentrations of soil organic C were greater at the site with 460 mm of annual precipitation (0.39%) compared to those (0.30-0.32%) with lower mean annual precipitation (365 to 407 mm).

Honey mesquite (Prosopis glandulosa ) is a N-fixing shrub that has encroached into much of the southern mixed-grass prairie in Texas over the last 100 years. The effect of root-plowing on soil organic C was investigated by comparing undisturbed, 4, 11, 16 and 22 years old root-plowed sites (Teague et al. 1999). They observed no differences between native and root-plowed sites, and also no trend in soil organic C over the 22-year period suggesting that removal of the mesquite trees by root-plowing does not change soil C relative to the native rangeland. The C4-dominated grasslands/savannas of southern Texas have been transformed to thorn woodlands as tree and shrub abundance (primarily honey mesquite, P. glandulosa) has markedly increased in the past century (Boutton et al. 1998). Model predictions estimated a 2-fold increase in soil organic C in the top 10 cm of the soil profile beneath clusters of woody plants (23.5 MT C/ha) compared to open herbaceous vegetation (11.7 MT C/ha) (Hibbard et al. 2001). They estimated a soil C sequestration rate of 0.23 MT C/ha/yr for the woody plants. Field research addressing soil organic C storage following woody plant invasion in the Rio Grande Plains of southern Texas generally corroborates the above model predictions as soil organic C to a depth of 30 cm was >50 MT C/ha for woody plants and about 20 MT C/ha in grasslands (Boutton and Liao 2004). Their calculated C sequestration rate for the woody invaded areas was 0.14 MT C/ha/yr.

Soil organic C to a depth of 40 cm was 2-fold greater on a perennial grass (Stipa tenacissima) steppe (40.3 MT C/ha) compared to a steppe dominated by the perennial shrub Artemisia barrelieri (18.8 MT C/ha) in Spain (Gauquelin et al. 1998). Both of these steppe sites receive approximately the same mean annual precipitation (370-400 mm). Replacement of the perennial grass by the perennial shrub is usually associated with land degradation. Soil organic C pools decreased as precipitation decreased. Sites characterized by mean annual precipitation of 15 year old stands of C. drummondii (McCarron et al. 2003). However, C cycling was negatively impacted by the woody plant invasion as soil CO2 flux was reduced by 16% compared to the native prairie.

Jackson et al. (2002) investigated the effect of woody plant invasion into grasslands along a precipitation gradient (200 to 1,000 mm) in the southwestern U.S. They found that drier sites gained and wetter sites lost soil organic C with woody plant invasion. Landi et al. (2003) evaluated soil organic C storage along a grassland to forest gradient in Saskatchewan encompassing a mean annual precipitation range of 330 (shortgrass rangeland) to 450 mm (forest community). They found that soil organic C to a depth of 120 cm increased from shortgrass (91 MT C/ha) to tallgrass (150 MT C/ha) rangeland, but that forest-dominated vegetation (96 MT C/ha) was similar to the shortgrass rangeland and 36% lower than the tallgrass community despite similar mean annual precipitation amounts (420-450 mm). The authors estimated C sequestration rates of 0.006 MT C/ha/yr for the shortgrass community, 0.012 MT C/ha/yr for the tallgrass community and 0.008 MT C/ha/yr for the forest community, using soil formation times of 11,500-17,000 years.

Assessing Soil Carbon in Rangelands

To assess soil C sequestration in rangelands one must deal with the variability in soils and vegetation at multiple spatial scales ranging from plant community interspaces (Hook et al. 1991, Vinton and Burke 1995, Derner et al. 1997) to the landscape (Burke et al. 1999). Sampling strategies need to recognize spatial variability resulting from topography, grazing, microsites and plant species. Patchy vegetation patterns characterize many arid and semiarid rangelands and these patches are often associated with large differences in soil organic C (Hook et al. 1991, Vinton and Burke 1995, Derner et al. 1997). It should be noted, however, that differences in soil organic C associated with patch-scale variability are often inconsequential relative to landscape-level differences (i.e., topography, Burke et al. 1999) and may not be important in assessing soil C over large land areas.

Soil erosion and deposition can also play an important role in spatial distribution of soil organic C. Over long time periods, soil erosion and deposition are responsible for many of the landscape-level differences in C sequestration potential. Much of the soil organic organic C in rangelands is concentrated near the soil surface (Weaver et al. 1935, Gill et al. 1999), where it is more susceptible to loss or redistribution by wind and water. Therefore, sampling points should be spatially distributed based on the relative proportion of erosional and depositional surfaces.

Data collection at representative sites is necessary to monitor changes in soil organic C pools, and sensitive techniques are required that are capable of detecting small increments of C change over relatively short time frames (Batjes 2002). Yet, use of a paired-site approach for verifying C sequestration requires substantially large sample sizes (40 to 65 paired sites) to achieve an 80% confidence level (Kucharik et al. 2003). They also noted that it is essential to use statistical power analysis to ensure a high level of confidence in soil C sequestration. Field researchers, however, will have to balance the statistical requirements necessary to achieve this level of confidence with costs of obtaining a large number of samples. Attention to details regarding selection of appropriate and comparable sampling sites and re-sampling of these same sites over time are essential to economically obtain field data estimates of C sequestration.

Conclusions

In terms of vegetative community and soils, rangeland ecosystems are very complex, making it difficult to characterize soil C storage. Changes due to management, environment or the interaction of these factors are difficult to assess considering the complexity of these ecosystems; however, in recent years research has been aimed at making these assessments. Rangelands are a large repository of soil C because of their high C density and the vast land resource area they represent. Improved range management strategies have been shown to significantly increase soil C storage while concurrently providing other benefits such as improved water infiltration, increased water storage capacity, and greater nutrient reserves. Because productivity of rangelands are inherently low with traditional low-input management systems, suggested strategies for improving production, and concurrently soil C sequestration, include: 1) using appropriate plant species, 2) enhancing water-use efficiency, 3) controlling erosion and restoring degraded soil, and 4) managing and enhancing soil fertility (Izaurralde et al. 2001).

The estimates of soil C storage and rates of C sequestration for rangelands are being used by scientists and policymakers to estimate the potential of rangelands to help mitigate the elevated atmospheric levels of CO2. Considerable interest is being generated in terrestrial C storage and marketing of stored C is being initiated to be used by industry that is emitting CO2 into the atmosphere. Continued research, data synthesis and modeling will help to further refine estimates of terrestrial C storage in rangelands.

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1245.

Table 1. Management effects on soil carbon sequestration rates of rangelands across ecosystems.

Management Strategy/

Ecosystem_________ Soil C Sequestration Location Citation___

Grazing

Shortgrass prairie 0.12 MT C/ha/yr NE Colorado Derner et al.

1997

Shortgrass prairie 0.07 MT C/ha/yr NE Colorado Reeder & Schuman 2002

Northern mixed-grass 0.30 MT C/ha/yr SE Wyoming Schuman et al.

1999

Northern mixed-grass 0.29 MT C/ha/yr North Dakota Frank 2004

Alpine meadow ~80% increase Wyoming Provik 1999

Southern mixed-grass no change in C Oklahoma Fuhlendorf et al.

2002

Canadian prairie higher soil organic Canada Henderson 2000

C in grazed pastures

Planted pasture 0.41 MT C/ha/yr Virginia Conant et al.

2003

Planted pasture increase C sequest. Southern Piedmont Franzluebbers &

Stuedemann

2001

Nitrogen Inputs

N-fertilization, tallgrass 1.6 MT C/ha/yr Kansas Rice 2000

N-fertilization, CRP 0.41-1.16 MT C/ha/yr Wyoming Reeder et al.

1998

N & S fertilization increases of 5.4-9.3 Saskatchewan Nyborg et al.

MT C/ha 1994

Legume interseeded 0.33-1.56 MT C/ha/yr South Dakota Mortenson et al.

into native rangeland 2004

Fire

Tallgrass prairie 0.22 MT C/ha/yr Kansas Rice 2000

Table 1 continued

Management Strategy/

Ecosystem_________ Soil C Sequestration Location Citation___

Restoration of Degraded

Rangelands

Southern mixed-grass no difference in top Oklahoma Fuhlendorf et

10 cm (mod. graze) al. 2002

65% decrease (heavy

grazing)

Marginal croplands to restored soil C to 80% Sudan Olsson &

rangelands of native rangeland in Ardö 2002

100 yr

Restored semiarid 466% increase Argentina Abril &

savanna Bucher 2001

Tallgrass prairie ave. 447 kg C/ha, Texas Potter et al.

estimated 100 yr 1999

to achieve native

rangeland level

CRP-Great Plains 0.8-1.1 MT C/ha/yr Texas, Kansas, Gebhart et al.

Nebraska 1994

CRP-Great Plains ave. 0.9 MT C/ha/yr North Dakota to Follett et al.

Texas 2001

Mined land reclamation 400% increase over Wyoming Stahl et al.

30 years 2003

Woody Plant Encroachment

Southern mixed-grass removal of Prosopis Texas Teague et al.

glandulosa had no affect 1999

on soil C compared to

native rangeland

Grassland savanna-

thorn woodlands model predictions- Texas Hibbard et al.

0.23 MT C/ha/yr under 2001

woody plants

Rio Grande Plains 0.14 MT C/ha/yr Texas Boutton &

Liao 2004

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