**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.usda.gov>**

**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 (CO~2~) 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 CO~2~ (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 CO~2~ 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 (C~3~) grasses in these
rangeland ecosystems, shifting them to a plant community dominated by
the warm-season (C~4~) 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 C~3~ species western wheatgrass
(*Pascopyrum smithii*) (Derner et al., unpublished data). Reducing the
C~3~ 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 C~3~ herbaceous species. Similar changes
in soil organic C associated with a plant community shift from C~3~ to
C~4~ 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 C~4~ 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 CO~2~
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 C~3~ 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 CO~2~ 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 C~4~-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 \<240 mm displayed soil organic C pools that were only
27 to 57% of those found at the two sites mentioned previously.

Nine years following removal of all vegetation in a semiarid
shrubland/woodland in Spain, soil organic C in the top 20 cm of the soil
profile was 31% lower compared to non-disturbed areas (Martinez-Mena et
al. 2002). The authors estimated losses of 0.02 MT C/ha/yr from erosion
and 4.3 MT C/ha/yr from mineralization due to higher soil temperatures.

The C~3~ shrub *Cornus drummondii* has expanded in the tallgrass prairie
of Kansas with the removal of fire. Soil organic C to a depth of 5 cm
did not differ between native, C~4~-grass dominated prairie and the \>15
year old stands of *C. drummondii* (McCarron et al. 2003). However, C
cycling was negatively impacted by the woody plant invasion as soil
CO~2~ 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 CO~2~. 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 CO~2~ 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\_\_\_]{.underline}

[Grazing]{.underline}

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]{.underline}

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]{.underline}

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

Table 1 continued

Management Strategy/

[Ecosystem\_\_\_\_\_\_\_\_\_ Soil C Sequestration Location
Citation\_\_\_]{.underline}

[Restoration of Degraded]{.underline}

[Rangelands]{.underline}

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]{.underline}

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