What can ecological science tell us about opportunities for carbon sequestration on arid rangelands in the United States?

Abstract

Scientific interest in carbon sequestration on rangelands is largely driven by their extent, while the interest of ranchers in the United States centers on opportunities to enhance revenue streams. Rangelands cover approximately 30% of the earth's ice-free land surface and hold an equivalent amount of the world's terrestrial carbon. Rangelands are grasslands, shrublands, and savannas and cover 312 million hectares in the United States. On the arid and semi-arid sites typical of rangelands annual fluxes are small and unpredictable over time and space, varying primarily with precipitation, but also with soils and vegetation. There is broad scientific consensus that non-equilibrium ecological models better explain the dynamics of such rangelands than equilibrium models, yet current and proposed carbon sequestration policies and associated grazing management recommendations in the United States often do not incorporate this developing scientific understanding of rangeland dynamics. Carbon uptake on arid and semi-arid rangelands is most often controlled by abiotic factors not easily changed by management of grazing or vegetation. Additionality may be impossible to achieve consistently through management on rangelands near the more xeric end of a rangeland climatic gradient. This point is illustrated by a preliminary examination of efforts to develop voluntary cap and trade markets for carbon credits in the United States, and options including payment for ecosystem services or avoided conversion, and carbon taxation. A preliminary analysis focusing on cap and trade and payment for avoided conversion or ecosystem services illustrates the misalignment between policies targeting vegetation management for enhanced carbon uptake and non-equilibrium carbon dynamics on arid United States rangelands. It is possible that current proposed carbon policy as exemplified by carbon credit exchange or offsets will result in a net increase in emissions, as well as investment in failed management. Rather than focusing on annual fluxes, policy and management initiatives should seek long-term protection of rangelands and rangeland soils to conserve carbon, and a broader range of environmental and social benefits.

Full text

What can ecological science tell us about opportunities for carbon sequestration
on arid rangelands in the United States?
Kayje Booker
a
, Lynn Huntsinger
a,
*, James W. Bartolome
a
, Nathan F. Sayre
b
, William Stewart
a
a
Department of Environmental Science, Policy, and Management, 133 Mulford Hall MC 3110, University of California, Berkeley, CA 94720-3110, United States
b
Department of Geography, 507 McCone Hall MC 4740, University of California, Berkeley, CA 94720-4740, United States
1. Introduction
Rangelands are one of the most widely distributed landscapes
in the world. Found at the more arid end of the earth’s climates,
approximately 30% of the ice-free global land surface can be
considered rangeland (FAO, 2009), although estimates vary widely
depending on the particular definition used (Lund, 2007). In turn,
rangelands are thought to have as much as 30% of terrestrial carbon
stocks (Schuman et al., 2002; FAO, 2009). Debates about the
impacts of livestock grazing, climate change, and cultivation on
rangelands now include concerns about their effects on carbon
cycling. Interest in increasing carbon flux from the atmosphere into
the soils and vegetation of rangelands in the United States has led
to a number of national policies and market-based projects
designed to encourage management that enhances this flux
(McCarl and Sands, 2007). It is now commonplace to use the
rationale of increasing carbon sequestration to argue for changes in
grazing management. Focusing on the U.S., we argue that, given
recent developments in the scientific understanding of rangeland
ecological dynamics, grazing management strategies and associ-
ated management practices cannot lead to reliably increased
capture of carbon on many arid rangelands. For this reason, policies
for such rangelands that are based on additionality are unlikely to
be effective, and may even lead to increased emissions.
Proposals for managing rangelands for climate change mitiga-
tion are gaining attention at state and federal levels in the United
States. Primarily because they are so extensive, the 312 million ha
of U.S. rangelands (USFS, 1989), defined here as grassland,
Global Environmental Change 23 (2013) 240–251
ARTICLE INFO
Article history:
Received 30 July 2011
Received in revised form 12 August 2012
Accepted 5 October 2012
Available online 1 November 2012
Keywords:
Non-equilibrium dynamics
Arid lands
Soil carbon
Cap and trade
Additionality
Rangeland management
ABSTRACT
Scientific interest in carbon sequestration on rangelands is largely driven by their extent, while the
interest of ranchers in the United States centers on opportunities to enhance revenue streams.
Rangelands cover approximately 30% of the earth’s ice-free land surface and hold an equivalent amount
of the world’s terrestrial carbon. Rangelands are grasslands, shrublands, and savannas and cover 312
million hectares in the United States. On the arid and semi-arid sites typical of rangelands annual fluxes
are small and unpredictable over time and space, varying primarily with precipitation, but also with soils
and vegetation. There is broad scientific consensus that non-equilibrium ecological models better
explain the dynamics of such rangelands than equilibrium models, yet current and proposed carbon
sequestration policies and associated grazing management recommendations in the United States often
do not incorporate this developing scientific understanding of rangeland dynamics. Carbon uptake on
arid and semi-arid rangelands is most often controlled by abiotic factors not easily changed by
management of grazing or vegetation. Additionality may be impossible to achieve consistently through
management on rangelands near the more xeric end of a rangeland climatic gradient. This point is
illustrated by a preliminary examination of efforts to dev elop voluntary cap and trade markets for carbon
credits in the United States, and options including payment for ecosystem services or avoided
conversion, and carbon taxation. A preliminary analysis focusing on cap and trade and payment for
avoided conversion or ecosystem services illustrates the misalignment between policies targeting
vegetation management for enhanced carbon uptake and non-equilibrium carbon dynamics on arid
United States rangelands. It is possible that current proposed carbon policy as exemplified by carbon
credit exchange or offsets will result in a net increase in emissions, as well as investment in failed
management. Rather than focusing on annual fluxes, policy and management initiatives should seek
long-term protection of rangelands and rangeland soils to conserve carbon, and a broader range of
environmental and social benefits.
ß2012 Elsevier Ltd. All rights reserved.
* Corresponding author. Tel.: +1 510 685 1884; fax: +1 510 643 5098.
E-mail addresses: kayje@berkeley.edu (K. Booker), huntsinger@berkeley.edu
(L. Huntsinger), jwbart@berkeley.edu (J.W. Bartolome), nsayre@berkeley.edu
(N.F. Sayre), billstewart@berkeley.edu (W. Stewart).
Contents lists available at SciVerse ScienceDirect
Global Environmental Change
journal homepage: www.elsevier.com/locate/gloenvcha
0959-3780/$ – see front matter ß2012 Elsevier Ltd. All rights reserved.
http://dx.doi.org/10.1016/j.gloenvcha.2012.10.001

shrubland, and savanna, contain significant carbon stocks.
Traditional land use, largely grazing, does not involve tillage,
potentially resulting in less soil carbon loss than that connected to
cultivation (Uri and Bloodworth, 2000). It has been estimated that
grazing lands contribute about 15% of U.S. soil carbon sequestra-
tion potential (Lal et al., 2003). U.S. rangeland livestock producers,
generally operating with low and variable financial returns,
continue to express considerable interest in diversifying income
streams to include payments related to carbon sequestration (Diaz
et al., 2009). Land management and conservation organizations
also seek to promote management for increased carbon seques-
tration on private and public rangelands (Audubon California,
2012). As the U.S. failed to ratify the Kyoto treaty, the voluntary
markets for trading carbon credits have thus far been the main
thrust of initiatives for incentivizing management for carbon
sequestration domestically.
While the specific applications are still contested, there is broad
scientific consensus that non-equilibrium models better explain
the ecological dynamics of arid rangelands, in the U.S. and
throughout the world, than equilibrium models (Briske et al., 2005;
Vetter, 2005). The ecological behavior of rangeland systems has
been much debated and researched in the last twenty years, but it
is not clear that what has been learned through investigation,
experimentation, and theoretical development has been integrated
into carbon sequestration initiatives and management recom-
mendations. Further, a lack of information has led to over-
generalized applications of scientific and traditional ecological
knowledge despite the fact that such knowledge is linked to locales
of specific environmental characteristics within rangeland sys-
tems. Just as different definitions of the term ‘‘rangeland’’ can lead
to vastly different estimates of how much rangeland there is, over-
generalization of ecological knowledge to areas of differing
environmental parameters can lead to incorrect assumptions
about potential management outcomes. Site specificity is impor-
tant because rangelands are so widespread, temporally and
spatially diverse, and diverse in structure and function.
Because synthesis of information about rangelands has suffered
from poorly defined terms and variable usage, this paper begins
with a definition of rangeland and a review of the development of
explanatory rangeland vegetation change models and their linkage
to ecological sites. Next, the interaction of rangeland ecological
dynamics and management for carbon sequestration is analyzed.
Finally, the implications of this science for carbon sequestration
management and policy initiatives are presented and discussed,
and we offer recommendations for rangeland carbon policies that
accommodate recent developments in rangeland ecological
science.
2. Rangelands and rangeland ecosystem dynamics
Rangelands have been defined as a type of vegetation, a land
use, or what is left when other types are excluded. Definitions of
rangeland that include specific uses, usually livestock grazing
(NRCS, 1997; Holechek et al., 2010), are not a good basis for stable
descriptions of extent or processes. Defining rangelands as ‘‘land
not permanently ice and snow, urban, cropland, or forest’’
(Stoddard et al., 1975) does not identify what rangelands actually
are. Defining rangelands as grasslands, shrublands, and savanna
(Heady and Child, 1994) incorporates a wide range of communi-
ties from arid to semi-arid and can be distinguished from other
more productive systems like woodlands, forests, wetlands, and
croplands. These distinctions are essential for predicting and
measuring carbon at the landscape scale. Included within this
definition are what have been defined as ‘‘grazing lands’’ (NRCS,
1997; Follett and Reed, 2010) to emphasize the importance of
large herbivore grazing, and intensively managed lands used for
grazing that have been termed ‘‘pasturelands’’ (Holechek et al.,
2010). Rangelands can be temporally transient, especially at the
margins with forest, wetlands, and croplands (Heady and Child,
1994). Rangelands with sufficient rainfall, or suitable for
irrigation, may be temporarily or permanently converted to
cropland or forest. U.S. arid and semi-rangelands generally fall to
the west of the 100th Meridian.
For nearly a century, the management of U.S. rangelands has
been informed by predictive models for vegetation change linked
to geographic areas known first as ‘‘range sites’’ and now as
‘‘ecological sites’’ (Brown, 2010). Early in the twentieth century,
Sampson (1917) adapted the then new concepts of Clementsian
plant succession into a model relating grazing pressure to
vegetation change away from and towards an equilibrial ‘‘climax’’
of ideal plant species composition. This linear, deterministic model
was used in developing a general framework for evaluating
progress in sustainable livestock grazing and rehabilitation of
deteriorated rangeland. The utility of this approach was greatly
enhanced by the development of what was called the ‘‘quantitative
range condition’’ model (Dyksterhuis, 1949), which measured
range condition as the difference between the current species
composition and productivity and the ideal climax state. What
operationalized this approach was combining Sampson’s ideas
about species composition with newer theories of an edaphic
climax to identify what were termed range sites, defined as
rangeland areas with a similar potential climax state (SCS, 1976).
This formed the basis for evaluating the ‘‘health’’ of rangelands and
for informing grazing management.
The term ecological site replaced range site by the early 1990s
(NRCS, 1994). This was more than just an alteration in terminology,
as the change reflected significant advances since the 1980s in
models describing succession. It has been found that non-
equilibrium models better explain ecological dynamics then do
equilibrium-based models, particularly when rangeland is at the
arid end of a gradient from dry to mesic conditions (Briske et al.,
2005; Vetter, 2005). Non-equilibrium or disequilibrium models
posit that abiotic factors such as weather, soil structure, erosion,
and water table depth are the dominant drivers of rangeland
productivity and species composition (Ho, 2001), and that the
relationship with livestock grazing is often non-linear (Westoby
et al., 1989; Ellis and Swift, 1988). On arid rangelands spatial and
temporal variation in water and forage resources is high, annual
production is as unpredictable as rainfall and temperature
patterns, and extremes of precipitation or temperature are not
uncommon. Non-equilibrium models also posit the existence of
multiple stable (within a management timeframe) vegetation
states maintained largely by abiotic factors, rather than a single
endpoint climax or stable equilibrium state (Westoby et al., 1989;
Stringham et al., 2003) created mostly by biotic interactions,
including grazing pressure. As a result, an ecological site is
described more by climate, topography, and soils, than reference to
a climax vegetation (Brown, 2010). Assessments of range condition
have been largely decoupled from the use of linear distance to
climax.
Westoby et al. (1989) provided an alternative approach to
describing the dynamics of managed ecological sites using state
and transition models which accommodate non-linear, non-
equilibrium ecology and varied management objectives. Current
use of ecological site by federal agencies emphasizes concepts of
stable states and thresholds and utilizes recent advances in
available soil information and Geographic Information System
(GIS) technology (Brown, 2010). The more traditional goals of
sustainable grazing management and enhanced forage production
have been joined by the need to evaluate and anticipate response
of rangelands to global change and the potential for carbon
sequestration.
K. Booker et al. / Global Environmental Change 23 (2013) 240–251
241

3. Rangeland ecosystem dynamics and carbon sequestration
Effective climate change mitigation strategies for rangelands
must be based on knowledge of how management and land use
interact with carbon cycling. Carbon sequestration is the
movement of carbon stock, or stored carbon, from the atmospheric
pool to other carbon pools, including soils and vegetation. In
general, the movement of carbon between pools, usually measured
and reported on an annual basis, is described as carbon flux, and
management can influence both flux and stocks of carbon.
Incentivizing management that increases the net flux of carbon
from the atmosphere into soils and vegetation was a goal of the
carbon credit trading market of the Chicago Carbon Exchange, and
developing grazing management programs to create additional
carbon sequestration is a common subject of research and
speculation. Less attention has been given to the protection of
carbon stocks already in soils.
Much of the political and scientific discussion on terrestrial
carbon sequestration has centered on forests, where annual net
flux can be significantly increased through management of tree
growth and is easily measured. Yet, in terms of long-term carbon
storage, rangelands can be superior to forests because relatively
more of the total site carbon is stored in the soil (White et al., 2000;
Paruelo et al., 2010) where it is usually better protected from
atmospheric release than carbon stored in vegetation. However,
carbon inputs in rangelands, i.e. net carbon flows from the
atmosphere, are comparatively small, and soil carbon is more
difficult to measure than carbon in trees.
Soil carbon stocks are dynamic and sequestration may be
described as a function of inputs, losses, and storage time (Silver,
pers. com). For example, plant residues and organic amendments
add carbon, while carbon is lost through organic matter
decomposition (Subak, 2000). Especially on more arid ecological
sites, net primary productivity on rangelands is relatively low,
limiting carbon input. Due to the rapid turnover of grasses and
forbs, most carbon flux from the atmosphere into vegetation
through photosynthesis is quickly lost to the atmosphere through
decay. A small part of the carbon in the vegetation goes into the soil
and is stored longer as soil organic matter. Increased exposure of
soil to rainfall and the atmosphere by disturbance or tillage can
enhance flux to the atmosphere under some conditions.
Although the relationship between soil carbon and plant
production is not universal, low production, and, thus, low carbon
inputs to the soil, necessarily limit annual carbon sequestration.
Vegetation production in rangelands varies widely, in part because
rangeland as a land classification masks major differences in
climate and vegetation. In the United States rangelands include
montane meadows, blue oak woodlands, valley grasslands, basin
sagebrush, bluebunch wheatgrass prairie, bluestem and grama
grass prairie, and warm desert. The wettest of these rangelands are
capable of producing more than 3500 kg/ha per year, while the
driest yield on average only 200–300 kg/ha (Huntsinger and Starrs,
2006). These variations are linked to the climate and geomorphol-
ogy that define ecological sites (Parton et al., 1994). As might be
expected, given this variability, estimates of the average carbon
uptake on temperate rangelands vary widely.
4. Implications for management
Because rangeland vegetation mediates and constrains the
carbon flux from the atmosphere into soils and plants, three major
non-exclusive carbon management principles can be identified
when rangeland ecological dynamics are considered. First, in
rangeland ecosystems carbon flux into plants and soils is low,
highly spatially and temporally variable, strongly influenced by
stochastic events like weather, and largely outside the control of
management. Second, in some rangeland environments, because of
limited and slow plant growth, and significant storage of carbon in
mineral form close to the surface, management that causes soil loss
can significantly increase carbon flux to the atmosphere. Finally,
carbon flows and pool sizes may be less variable and more
amenable to enhancement through management at the less arid
end of the rangeland climate gradient. These principles largely
determine the outcome of carbon sequestration strategies in
rangelands, and must be considered in assessing the ability to
mitigate climate change through rangeland management.
In the United States, as in many parts of the world, the most
productive rangeland ecological sites, such as tallgrass prairie,
have largely been converted to crops; extant rangelands tend to be
at the less productive and more arid end of the spectrum (Fig. 1). In
these rangelands, the annual carbon flux into the soil is low, and
even, at times, negative, especially during drought (Zhang et al.,
[(Fig._1)TD$FIG]
Fig. 1. Characteristics of ecological sites pertaining to carbon sequestration as influenced by aridity.
K. Booker et al. / Global Environmental Change 23 (2013) 240–251
242

2010; Svejcar et al., 2008; Wylie et al., 2007)(Fig. 1). Because the
impacts of biotic interactions such as plant competition are
overshadowed by abiotic influences such as the timing and amount
of rainfall, temperature variations, and soil type (Westoby et al.,
1989; Parton et al., 1994; Briske et al., 2003), the most important
factors influencing carbon sequestration are not amenable to
management (Westoby et al., 1989; Briske et al., 2003, 2005)
Building management strategies for such rangelands based on
equilibrium type community models may exaggerate the role of
grazing and potential management impacts (Briske et al., 2005;
Huntsinger and Bartolome, 2004; Bartolome et al., 2008).
In contrast, more mesic and productive rangeland systems have
a greater likelihood of responding to grazing and other types of
management (Vetter, 2005). In rangeland ecosystems where
environmental conditions support plant growth sufficient for
plant competition and other biotic interactions to play a major role
in vegetation development, grazing management that leads to
increased soil carbon storage by plants, and increased woody and
perennial vegetation with extensive root systems, can positively
influence carbon sequestration, in scenarios similar to those of
other mesic ecosystems. In fact, most information documenting
carbon response to grazing is from less arid rangelands (Gilmanov
et al., 2010; Conant and Paustian, 2002), and the highest estimates
of potential rangeland carbon sequestration (Conant and Paustian,
2002; Ogle et al., 2004; Morgan et al., 2010) include management
activities such as fertilization and sowing of legumes that have
little to no effect where productive potential is limited by aridity
(Table 1).
Over-generalizing results from one rangeland type to another
can lead to false expectations. Much of the estimated potential for
increasing carbon uptake is based on the goal of shifting rangeland
communities either among or within vegetation states (Westoby
et al., 1989) or to other land uses like irrigated pasture, cropland, or
forestland. Yet such opportunities are severely constrained by
rangeland ecological dynamics specific to each ecological site.
Opportunities to increase carbon sequestration on rangelands are
highly variable and best predicted at a coarse scale by the position
of the ecological site along an aridity gradient. The magnitude of
carbon sequestration and management influences diminishes with
decreasing rainfall (Fig. 1). At a broad regional scale, precipitation
can be used as an initial filter to tailor interventions to the
possibilities of the ecological site.
5. What can rangeland management accomplish?
Management actions should be linked to ecological site and aim
to conserve the existing stocks of carbon in soils, as well as increase
flux from the atmosphere and length of storage time while
minimizing loss (FAO, 2009). Manipulation of rangeland vegeta-
tion and above-ground carbon stocks must consider both long and
short term effects on wildfire, erosion, wildlife, and water
dynamics. Below are discussed the major ways that management
might either prevent release of carbon or gradually increase carbon
stocks, especially in previously degraded areas (Table 1). However,
when considering the effects of these management actions on
carbon stocks, it must be remembered that abiotic factors will
constrain and at times overwhelm the results of any management
action. As the primary economic use of rangelands throughout the
world is livestock grazing, special attention is directed to grazing
management.
5.1. Grazing management
Although grazing management is the focus of many proposals
for and studies of rangeland carbon sequestration, grazing by wild
and domestic herbivores has been shown to have mixed and
unpredictable effects on carbon cycling in rangelands (Derner and
Schuman, 2007; Ingram et al., 2008; Diaz et al., 2009; Pin
˜eiro et al.,
2010; Briske et al., 2011). Processes on arid rangelands are
dominated more by abiotic factors than biotic interactions like
grazing (Bartolome et al., 2008). Much of the variability in results
and interpretations of grazing effects stem from inadequate
Table 1
Estimates of potential annual carbon sequestration on US rangelands.
Author Tonnes
C/ha
Tonnes
CO2e/ha
Million TCO2e US
Range-lands
%US CO2e emissions
d
(%) Management scenario
Morgan et al. (2010) (low)
a
0.07 0.26 80 1.37 ‘‘Best management practices,’’ citing Schuman
et al. (2002) and Derner and Schuman (2007)
Morgan et al. (2010) (high)
a
0.3 1.10 343 5.88
IPCC (2007)
a
0.03 0.11 34 0.59 ‘‘Cool dry’’ and ‘‘warm dry’’ types, using grazing,
fire, fertilization
Schuman et al. (2002)
b
0.34 1.243 388 6.64 Considers both mitigation gains and avoided losses using
conversion from croplands to grasslands, proper grazing,
and keeping lands in CRP
Lal et al. (2007) (High)
c
n.a. n.a. 257 4.40 Includes pasture and semi-arid rangelands using fertilization,
manure, and planting of improved species
Lal et al. (2007) (Low)
c
n.a. n.a 48 0.82
Conant et al. (2001) (average)
a
0.54 1.98 618 10.59 ‘‘Most common types of improvement were fertilization,
improved grazing management, and conversion to pasture
from native and cultivated lands, accounting for >90%
of all studies’’
Ogle et al. (2004) (low)
a
0.4 1.47 458 7.84 Assumes ‘‘most managed grasslands are in a nominal condition’’
and a single management change including sowing legumes,
planting more productive varieties, irrigating, and applying
fertilizer
Ogle et al. (2004) (high)
a
0.9 3.30 1031 17.65 Assumes 67% of rangelands currently degraded and multiple
of the above listed interventions
a
Reported sequestration potential in tonnes of carbon per hectare per year. These were converted to tonnes in US rangelands assuming 312 million hectares of rangelands
in US based on USFS (1989)
b
Reported sequestration potential as total MMT carbon, assuming 183 Mha rangelands in U.S. This was converted to tonnes of carbon per hectare then multiplied by 312
million so as to be consistent with other cited sources.
c
Reported sequestration for grazing and pasture lands entire US but did not give either a per hectare number nor a total number of grazing and pasture hectares in US based
on USFS (1989).
d
Based on US emissions of 5839.3 MMT CO
2
in EIA (2008).
K. Booker et al. / Global Environmental Change 23 (2013) 240–251
243

attention to differences among ecological sites. Arid rangelands are
underrepresented in grazing studies, yet within the bounds of
grazing intensities that can sustain livestock production, manipu-
lating grazing systems, for example using rotational, delayed, or
season-long grazing systems, is unlikely to make a significant
difference in carbon stocks or flows in the abiotically-controlled
non-equilibrium arid systems of most extant rangelands in the US
today (Briske et al., 2008).
The major potential impacts of grazing management on carbon
stocks are in driving or preventing vegetation state transitions,
such as increasing or decreasing woody plants, and in preventing
soil disturbance and exposure. On croplands, tillage increases
carbon loss by mixing and exposing soil and organic matter to the
atmosphere and weather, increasing decomposition and both
organic and inorganic carbon loss (Uri and Bloodworth, 2000),
while leaving more durable residue leads to a higher level of
carbon retention in the soil (Subak, 2000). Minimizing soil
disturbance, and leaving adequate residue, is also likely to prove
useful in retaining carbon stored in rangelands.
5.2. Grazing and woody vegetation
If grazing on a particular ecological site influences the
abundance of woody plants, the effect on carbon dynamics can
be large. Growth of shrubs sequesters carbon in woody material
that persists longer than herbaceous matter. The effect of grazing
and its relation to the growth of shrubs depends on the grazing
system itself (number and type of animals, duration, frequency,
and spatial distribution) (Huntsinger, 1996) and on the ecological
site. In drier areas of the southwestern US and much of Mexico,
heavy grazing can contribute to a transition from grasses and forbs
to woody shrub cover (Laycock, 1991). Conversely, in wetter areas,
such as coastal California, decreased grazing can lead to woody
shrub invasion in which grasses are choked out by dense shrubs
(Russell and McBride, 2003). Grazing of cheatgrass in Great Basin
sagebrush vegetation may reduce the chance of wildfire that
eliminates native woody shrubs (Daubenmire, 1970).
Research examining the carbon effects of these vegetation
changes shows that results are also linked closely to ecological site
(Jackson et al., 2005). Recent work suggests that drier sites are
more likely to retain organic carbon and wetter sites more likely to
lose it through increased oxidation rates with increases in woody
vegetation (Jackson et al., 2002). Any gains and losses in soil carbon
must then be compared with concomitant gains and losses in
aboveground carbon to assess the net carbon gain or loss. In desert
grasslands, woody cover increases, but as grasses decline due to
shading and root competition, there is more exposed soil. While it
seems likely that such a situation would increase carbon
sequestration over time due to the increase in woody plants
(Silver et al., 2010), the bare ground and lack of grasses may mean a
loss in soil carbon stocks from exposure and decay (Subak, 2000).
Dense shrub stands also leave little room for grasses that slow
runoff and increase percolation, potentially reducing soil moisture
and net primary production. Soil carbon may be sequestered for a
longer period of time and be less vulnerable to fire and drought
than aboveground woody biomass, meaning that continued grass
dominance may prove to be more valuable for climate change
mitigation in the long term. In California coastal grasslands, shrub
increases almost certainly increase carbon stocks. However, the
shrubs may increase fire potential and the possibility that much of
the stored carbon will be released to the atmosphere within a few
decades (Russell and McBride, 2003).
In sum, manipulating the amount of woody vegetation through
grazing, where feasible, remains an intervention opportunity that
is manageable, tractable, and likely has a significant effect on
carbon stocks. However, more needs to be known about the effects
of these vegetation state changes on carbon, especially soil carbon,
in different ecological sites, and how to balance increases in above
ground carbon stores with possibly higher fire probabilities. The
consequences of altered disturbance regimes also must be
evaluated, or short-term gains may result in long term loss.
Finally, other considerations, such as wildlife habitat, viewsheds,
and water cycling, must be taken into account before recommend-
ing actions to cause or prevent major vegetation change.
5.3. Reforestation and afforestation
Afforestation or reforestation are assumed to increase carbon
stocks on rangelands (Morgan et al., 2010; Jackson et al., 2005), but
again outcomes are linked to ecological site. Afforestation is the
planting of trees in areas that are not currently forests, while
reforestation is the replanting of trees that have been removed
through fire or cutting. On some sites, protection of trees from
browsing or other influences that suppress them is another
approach to increasing trees. Although many arid rangeland sites
are too dry for trees, some more mesic rangelands were once, and
in some cases are still, home to a variety of broadleaf and
coniferous trees. Unlike increases in woody shrubs, the introduc-
tion, or re-introduction, of broadleaf trees appears to have clear,
positive effects on rangeland carbon sequestration (O’Halloran
et al., 2009; Baldocchi et al., 2010; Morgan et al., 2010; Jackson
et al., 2005). Besides the carbon stored in the aboveground woody
material, broadleaf trees typically have large root systems, and, in
some rangeland systems, actually increase production of grasses
and forbs beneath the canopy (Frost and McDougald, 1989).
Reforestation of previously forested rangelands, or increasing
the tree cover on sparsely forested rangelands, can have positive
carbon impacts, but environmental and social tradeoffs arise and
can also be linked to ecological site. Trees are heavy water users
and affect the hydrological cycle both in their immediate area and
downstream (Jackson et al., 2005, 2007), reducing water availabil-
ity for other vegetation and other uses where water is limiting. On
arid lands, water is often the major limitation on tree growth, and
there is a tradeoff between water use and carbon fixation, with
trees transpiring as much as 500 kg of water for each kg of carbon
fixed on an annual basis. In addition, roughly 60% of this carbon has
been found to return to the atmosphere by respiration. In short,
plant loss of water vapor through stomata is far more than 1000
times the net carbon gain (Sabate
´and Gracia, 2011), which may
make for an unfortunate tradeoff in arid areas. In Mediterranean
oak (Quercus ilex) forests, respiration of up to 90% of annual rainfall
has been recorded (Gracia et al., 2011). Additionally, while planting
trees at lower densities can actually increase understory grass and
forb production on some ecological sites, in other cases thick tree
cover may choke out undergrowth and may not be compatible
with grazing use or habitat for grassland and shrubland wildlife.
Concerns have been raised that carbon policies could promote
monocultural tree plantations on rangelands that may be more
susceptible to epidemics of disease and pests. Management that
balances promotion of broader biodiversity goals with carbon
sequestration is often better aligned with visual and cultural
preferences (Caparro
´s et al., 2010).
As with woody vegetation in general, drought- and fire-prone
ecological sites raise additional concerns when it comes to
afforestation and reforestation. Plantings of trees or other woody
vegetation on sites where they are vulnerable to drought or likely
to burn may result in a net carbon loss. Attention to the density of
woody vegetation, the characteristics of surrounding vegetation,
and the resulting fuel structure is essential. Some widespread fast
growing plantation trees, such as eucalyptus, suppress understory
vegetation through allelopathy, shading, and heavy duff, and are
highly vulnerable to fire.
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5.4. Erosion
Soil erosion, which may result not only from poor grazing
practices but also from cultivation, road construction, or other
management actions and natural processes, may cause a net
decrease of soil carbon stocks and increased carbon flux to the
atmosphere. Lal et al. (1998) estimate that 15 MMT of soil carbon is
emitted to the atmosphere each year due to water erosion of US
soils. Eroded soils may have reduced production potential, and
reduced carbon storage capacity (Rhoton and Tyler, 1990).
However, the net carbon effect of erosion is not entirely clear as
recent work focusing primarily on croplands suggests that erosion
can function as a carbon sink in some circumstances (Van Oost
et al., 2007; Harden et al., 2008).
The degree of soil erosion that takes place depends primarily on
vegetation cover, vegetation residue on the soil (Holechek et al.,
1998), soil type, slope, and rainfall amount and seasonal
distribution. On rangelands, vegetation cover and the accumula-
tion of vegetation residue, also known as litter, mulch or residual
dry matter, can be successfully manipulated by managing grazing
intensity (Bartolome et al., 2006). A number of natural resource
protection programs currently exist that use residual dry matter as
an indicator for the appropriate amount of livestock impact. On
croplands, residue has been identified as a key component in
retaining soil carbon (Subak, 2000).
5.5. Restoration
Restoring woody vegetation to sites that were cleared, restoring
croplands by reconverting them to rangelands, restoring vegeta-
tion to bare soils, and restoring soil stability can all increase carbon
sequestration as well as storage. Plants are the primary vehicle for
adding carbon to terrestrial environments and planting grass may
help to increase soil carbon storage (Diaz et al., 2009). However
most of the research investigating revegetation by planting grass
and vegetation manipulation has taken place in the context of
converting marginal agricultural lands to grasslands, rather than
restoring lands that were already classified as rangeland (Derner
et al., 2005). Studies examining the restoration of agricultural fields
to grasslands have shown that doing so increases carbon storage by
replenishing some of the carbon lost from years of soil disturbance
through tillage and erosion (Diaz et al., 2009). Converting
rangelands that have been cultivated back to rangeland can
change the soil from a source to a sink and restore carbon storage
levels to that of native range (Uri and Bloodworth, 2000).
In some cases removal of invasive species may enhance soil and
plant storage by reducing fire hazard. For example, cheatgrass
(Bromus tectorum), an invasive grass, is associated with increased
fire frequency (Bradley et al., 2006). Organic amendments, such as
manure and compost, have shown promise for increasing soil
carbon storage (Conant et al., 2001; Follett et al., 2000; Moffet et al.,
2005; Paschke et al., 2005) and are being investigated for their
potential to increase sequestration (Marin Carbon Project, 2012). It
is possible that organic amendments do not actually enhance
carbon sequestration following their application, but increase soil
carbon storage due to remnant carbon from the organic materials
themselves (Conant et al., 2001). These efforts have not been
studied at the more arid end of the spectrum, and treatment of
extensive areas may not be practical, especially if transport results
in an increase in emissions.
5.6. Fire
While rangeland fires result in an immediate pulse of carbon to
the atmosphere, their effect on carbon stocks over the longer term
is not entirely clear, perhaps because research results may not be
linked to specific ecological sites, fire characteristics vary, carbon
stores are largely below ground, and weather is highly variable. In
the tallgrass prairie, for example, Bremer and Ham (2010) found
moderate soil carbon loss from annual burning, whereas Fornara
and Tilman (2008) found no decrease in soil carbon accumulation.
Fire may maintain a specific vegetation state, as when frequent
burning prevents shrub encroachment into grassland, or revitalize
a chaparral stand. Conversely, frequent fire may also lead to a
removal of woody vegetation and a transition to grassland, for
example when sagebrush is eliminated and replaced by cheatgrass
in Great Basin sagebrush. Grazing is sometimes used to reduce the
risk of fire through elimination of fine fuels, preventing shrub
encroachment, or suppressing shrub regrowth after a fire or
clearing (Huntsinger et al., 2007). It is important to consider
ecological site and the fuel conditions that will result from
management actions.
6. Implications for policies for carbon sequestration
Carbon sequestered as a result of policy implementation is
considered additional, meaning that it would not have been
sequestered, or would have been released to the atmosphere, in the
absence of the policy. Carbon policies that exchange sequestration
and emission credits should ensure that participants sequester
additional carbon as a result of a policy, yet the non-equilibrium
dynamics at the arid end of the climate spectrum makes reliably
attaining additionality through management difficult if not
impossible. How can policies to sequester additional rangeland
carbon accommodate ecological findings that the net flux of carbon
on arid rangelands is low, highly variable, and largely controlled by
abiotic factors?
6.1. Flux-based policies for carbon sequestration
Flux-based policies for carbon sequestration offer incentives for
increasing average flux from the atmosphere to soils, usually on an
annual basis, through changes in management. These policies
include most proposals for cap-and-trade as well as payments for
environmental services. They face a variety of obstacles in arid
rangeland systems including the size and variability of rangeland
flows, transaction costs associated with measurement, and,
particularly, the assumption that changes in management can
reliably increase terrestrial carbon sequestration.
The small and variable net carbon flux on arid rangelands
means that annual payments based on average flux measurements
will not be large on an areal basis, and in some cases may even be
negative. It has been argued that, though small, these flows and the
rewards based upon them may still ‘‘add up.’’ That is to say that
some rangeland landowners, including the U.S. government, own
such large acreages that small rewards per unit area would still be
meaningful once aggregated, as long as the net flux is positive. A
small increase in income could be meaningful for an enterprise
where returns from rangeland grazing itself are also small and
variable, and vulnerable to unpredictable drought.
On the flip side, small carbon losses driven by drought could
also add up to a meaningful loss for which the landowners would
be responsible. Aggregation may work for some landowners under
some types of policies for carbon sequestration. However, under
any flow-based policy that attempts to account accurately for
carbon sequestered, transaction costs posed by measurement and
monitoring are likely to diminish any yearly reward paid to
landowners to the point that few would find it worthwhile to
participate (Subak, 2000).
Table 2 provides examples of the potential range of annual
carbon payments landowners might expect before transaction
costs are accounted for, showing average, best, and worst-case
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245

scenarios using carbon flux data from Svejcar et al. (2008) in 2010
carbon offset prices from both the regulated (European Emissions
Trading Scheme, Clean Development Mechanism) and voluntary
(the now-defunct Chicago Climate Exchange) markets (Table 2).
While neither of these markets currently makes payments for U.S.
rangeland carbon, they offer evidence of the range of carbon prices
that might be expected in certified and voluntary carbon trading.
Data on sequestration were derived from the northern mixed
prairie (Svejcar et al., 2008) and chosen for analysis because they
were from a site that most consistently sequestered (rather than
emitted) carbon, with sequestration occurring six of seven years.
Notably, every site from the rangeland data set was a carbon
emitter at least one year out of seven. The worst case scenario
occurs when a site is a net emitter, a case that is dealt with
differently depending on the carbon policy. It is likely that the
possibilities for income augmentation increase along the gradient
from arid to mesic sites (Fig. 1). While even small income
augmentations may be motivating for ranchers, pinpointing where
along this gradient net flux shifts to consistently positive over the
long term is a challenge in itself, and will vary by ecological site.
Those benefiting from payments based on additionality are
expected to also take responsibility for negative fluxes.
The spatial and temporal heterogeneity of rangelands raises
transaction costs, which are likely to heavily undercut potential
payments. The transaction costs of quantification of flows depend
to some extent on the degree of accuracy that policymakers and
the public demand, or inaccuracy that they are willing to accept
(Conant and Paustian, 2002). Current methods of directly
measuring soil carbon can be accurate but, given high inter- and
possibly intra- site variability, if used for monitoring individual
projects, would need to be replicated in such large numbers as to
be cost prohibitive (Laca et al., 2010; Conant and Paustian, 2002).
Although there are ongoing attempts to create models to substitute
for direct measurement, current techniques do not provide
acceptable levels of certainty on arid and semiarid rangelands
(Brown et al., 2010), and data for arid shrublands are almost
entirely lacking (Morgan et al., 2010). In more productive
rangelands such as tallgrass prairie, Bremer and Ham (2010)
report that carbon inputs and losses in the system are both so large
that small errors in measurement or modeling may affect results to
such an extent that it becomes difficult to determine whether
ecosystems are sources or sinks, and multiple authors cite the low
ratio of flows to stock as a complicating factor in detecting a signal
(Morgan et al., 2010; Conant and Paustian, 2002). The relatively
high levels of below-ground storage complicate the use of more
efficient and less costly measurement methods such as remote
sensing (Gonzalez, 2008).
If landowners representing a large area of rangelands were to
participate in some type of coordinated carbon scheme, then it
might be possible to estimate sequestration on the aggregated
rangelands with reasonable reliability, although the total amount
is still likely to be quite low. The Chicago Climate Exchange, for
example, used averages ranging from 0.12 to 0.32 tCO2 per acre per
year (0.3–0.8 tCO2 per hectare per year). Perhaps the average could
be used as a conservative benchmark, and landowners could
request higher rates on the condition that they pay for their own
measurement and monitoring to demonstrate their achievements.
However, at current values, research shows that such broad-based
rangeland landowner participation is not likely unless prices rise
significantly and other forms of support are offered to help
landowners change their practices (Gosnell et al., 2011).
While landowner incentives and transaction costs raise
questions regarding the implementation of flux-based carbon
sequestration policies, the more fundamental issue is whether
such policies will result in additional rangeland carbon sequestra-
tion. These policies inherently assume that actions taken by
rangeland managers can reliably and predictably drive additional
carbon sequestration. However, as discussed above, rangeland
science suggests that in arid rangelands, carbon flows, and
potential additionality, are largely beyond the control of rangeland
managers (De Steiguer et al., 2008; Brown et al., 2010; Westoby
et al., 1989). This means that most additional carbon sequestered
after a policy is implemented is likely not the result of
management changes, and would have been sequestered even
in the absence of the policy. Even if the problems previously
mentioned – high variability and irregular losses – can be solved
through broad-based averaging schemes, the carbon sequestered
is still not truly additional unless it increases sequestration over
previous levels due to changes made as a result of the policy.
Such management-based additionality is undermined where
management actions lack predictable, consistent, and measurable
effects on carbon sequestration, as they do in all but the most mesic
rangeland ecological sites. For those management interventions
that do have an effect, the response follows a precipitation
gradient, with larger and more predictable effects in more mesic
rangelands and smaller, less predictable effects as aridity
increases. This gradient of effects must be kept in mind when
considering carbon sequestration policies. Policies that depend on
carbon sequestration as a result of management changes may be
appropriate for more mesic rangeland sites that are responsive to
management interventions (Fig. 1), but they are inconsistent with
the findings of ecological science on arid rangelands, where
management effects are largely unpredictable and heavily con-
strained by abiotic factors.
If participants are not sequestering more carbon than non-
participants, or increasing sequestration rates above those prior to
program implementation, then such policies have little use in
climate change mitigation. Moreover, offering incentives based on
the idea that increased uptake compensates for emissions
elsewhere, when uptake cannot be verified to be increasing, is
fraudulent and potentially harmful. If the incentives are financed
by additional fossil fuel based emissions, as with private carbon
offset markets, then any gap between promised and actual
achievement will involve a real addition of greenhouse gases to
the atmosphere. Therefore, there is a high likelihood that policies
that allow others to emit carbon using credits or offsets based on
assumed but difficult to measure increased rangeland carbon
sequestration will result in an overall increase in emissions.
6.2. Policy principles for carbon sequestration in arid U.S. rangelands
Based on the discussion above, we posit four principles for
policies aimed at carbon sequestration in arid U.S. rangelands to
make them consistent with modern rangeland ecology:
1.
Policies should not require short-term carbon accounting. Based
on the difficulties of measuring and monitoring soil carbon in
heterogeneous rangelands as well as the low and variable flows
of carbon, this type of accounting currently incurs transaction
Table 2
Average, best, and worst-case scenarios from annual carbon payments per hectare
for northern mixed prairie, assuming zero transaction costs, based on carbon flux
data from Svejcar et al. (2008). Prices were taken from the Ecosystem Marketplace
website in August of 2010.
Type of carbon
payment program
Average
$/ha/yr
Best
$/ha/yr
Worst
$/ha/yr
Certified ($10.24/tonne)
a
$19.86 $44.64 $10.14
Voluntary ($3.03/tonne)
b
$5.88 $13.20 $3.00
a
Clean development mechanism.
b
Chicago climate exchange.
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246

costs that divert much of the revenue away from landowners.
Moreover, short-term variations in carbon are unlikely to be the
result of management changes and are thus not additional.
2.
Policies should not assume that changes in management always
function as the primary basis for additional carbon storage. As
shown above, especially at the more xeric end of the spectrum,
which includes most extant U.S. rangelands, management
actions are constrained and often overwhelmed by abiotic
factors to such an extent that carbon sequestration or release
cannot be attributed to management.
3.
Credits from rangeland carbon sequestration based on manage-
ment should not be considered to offset emissions. Because the
amount of carbon sequestered on rangelands resulting from
changes in management is usually very small, and is effectively
impossible to determine relative to the role of abiotic factors, the
carbon that can be considered truly additional is both negligible
and undefined. Supporting the many other co-benefits of
ranching is not sufficient justification for carbon sequestration
payments or credits—only sequestration that is additional
should be used to justify payments or credits that offset
emissions. Until and unless arid rangeland sequestration can
meet the additionality standard reliably, policies should not
allow emissions elsewhere to be offset by such sequestration, as
they would thereby risk an overall increase in emissions.
4.
Policies should seek to conserve rangelands and encourage
restoration through conversion of marginal or degraded
agricultural lands back to rangelands. Although the carbon
sequestered on a short-term basis is low, variable, and not
reliably enhanced through management, over the long-term
most rangelands, even on the arid end of the spectrum, function
as significant sinks, whereas more intensive land uses are likely
to be a source. Rangelands that have been previously used for
cropland have an especially high capacity for sequestration
upon conversion back to rangeland because they are removed as
a high soil carbon emission source. If avoidance of rangeland
conversion or reconversion to rangeland results directly from
policy, the carbon sequestered over time can be considered
additional. However, to be consistent with Principles 1 and 3,
conservation or reconversion of these rangelands should not
require explicit short-term carbon accounting or be used to
offset other emissions.
7. Applying policy principles based on ecological science to cap
and trade, payment, and taxation options for creating
sequestration incentives
The above policy principles can be used as a basis to assess
whether or to what extent policies are consistent with the
ecological dynamics of rangelands. Two general types of policies
for carbon sequestration are highlighted here: (1) those that are
flux-based, as described above, including cap and trade and
payment for ecosystem services and (2) those that focus on long-
term rangeland conservation and restoration through payments
for avoided conversion or restoration. In addition, we consider the
potential impacts of another type of carbon policy, a carbon tax,
which, if implemented, could have an indirect effect on rangelands.
These policies represent the most widely proposed and considered
current options for rangelands in the United States. Examining
them in the light of rangeland ecological dynamics enables a
general assessment of their potential to influence carbon cycling.
The global policy situation for carbon sequestration is quite fluid,
as across the globe organizations and states explore ways of
overcoming social and ecological challenges to successfully
encouraging carbon storage and sequestration. At the current
time the U.S. policy environment is generally hostile to non-market
regulatory mechanisms, and policy options are further limited by
poor links between economic decision-making criteria and
ecological science (Norgaard, 2009).
7.1. Flux-based policies for carbon sequestration
These types of incentives reward additionality.
7.1.1. Cap and trade
Cap and trade schemes for rangelands, which functioned
through the Chicago Climate Exchange (CCX) until December 2010,
have gotten the most attention from U.S. policy makers and
analysts and are still under consideration for future domestic
climate change mitigation efforts. However, many supporters of
cap and trade systems for rangeland carbon have expressed doubts
about their potential to actually increase carbon sequestration
(Laca et al., 2010). Instead, some proponents have focused on the
co-benefits supported by carbon credit income that can contribute
to rangeland conservation. Supporters have cited biodiversity,
preservation of ranching culture and economies, improved
management and adoption of new practices, and prevention of
sprawl as some of the benefits to be expected from cap-and-trade
(Kroeger et al., 2009).
In analyzing cap-and-trade programs in general and the results
of the CCX in particular, it is clear that they are inconsistent with
the policy principles for rangeland carbon sequestration stated
earlier. One of the critical flaws of the CCX offset program, from a
climate change perspective, was that it assumed changes in
management would lead to additionality. Yet there is no assurance
that carbon sequestered under the CCX would not have been
sequestered in the absence of such a scheme. This is both because
many participants did not actually change management practices
(Diaz et al., 2009; Gosnell, 2009) and because of the predominance
of abiotic drivers in these systems. Carbon cannot be reliably
increased on arid rangelands purely through changes in manage-
ment.
The lack of management changes may have been peculiar to the
design and enforcement of the CCX, which, as a voluntary trading
scheme, was not subject to some of the same concerns and
oversight as some other certified programs. But the lack of
additionality from management changes is a fundamental
contradiction between any cap-and-trade or flux-based policies
and the ecological science of rangelands. If the target of the policy
is reducing net emissions, then achieving additionality is essential.
In a cap and trade system, the magnitude and additionality of the
sequestration generating the offset credit must be certain to
compensate for the emissions of the buyer of the credit, or net
emissions will not decrease.
To better assure additionality by screening out background
abiotic factors, trading schemes could require baseline flux
measurements and only pay managers for the excess annual
carbon sequestered above the baseline. However, additional
carbon resulting from management action would likely be
negligible, and flows are inherently low. Once the transaction
costs of measuring and monitoring are taken to account, it seems
unlikely that such a system would result in net income for
landowners. Moreover, although baseline flux measurements
provide some information on typical carbon flows independent
of management changes, determining how much of the carbon flux
to attribute to management (i.e. the portion of carbon that is truly
additional), rather than due to changes in precipitation or other
abiotic factors, would be impossible.
Another way around the additionality problem is for rangelands
to be considered under the cap instead of outside the system as
offsets. In such a scheme, yearly flows would be measured and
landowners would be allotted a credit (either purchased or free,
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depending on the specifics of the policy) based on the historical
emissions of their entire operation. As the cap was lowered,
landowners would have to decrease their emissions commensu-
rate with the lowering cap or purchase credits from others in order
to maintain their level of emissions. If they were able to lower their
emissions below their credit allotment, they could sell the extra
credits. This is the basic cap and trade scheme as it applies to all
covered industries. However, to cover rangelands under such a
system requires the expense of measuring, monitoring, and
regulating emissions on all rangelands in the United States with
some accuracy, and political opposition would likely be ferocious
as landowners would be lumped with true polluters such as the
coal industry. Moreover, as most landowners would probably be
unable to decrease annual emissions, over time the system would
penalize landowners for something out of their control, possibly
creating perverse incentives for rangeland conversion.
Thus, either as offsets or under a cap, most extant U.S.
rangelands are not suited to a cap-and-trade system based on
additionality of net flux. The co-benefits of payments to land-
owners from a carbon trading scheme must not obscure the fact
that carbon emissions reduction is the policy goal, and such
reductions are unlikely to result reliably from changes in
management.
7.1.2. Payments for ecosystem services
Direct payments for ecosystem services (PES) can include
payments for increased carbon sequestration. However, direct
payments, whether from state or private entities, cannot overcome
the problems of achieving additionality on arid rangelands through
management. In arid environments, payments for management
that protects the soil, already part of responsible grazing
management, might contribute to protecting carbon stocks, but
this response will be constrained by abiotic factors.
To the extent that PES relies on changes in management to
increase carbon sequestration, PES suffers from the same lack of
additionality issues as other flux and management-based policies
such as cap-and trade. The PES framework, however, is preferable
to cap-and-trade in that, in a PES framework, any carbon that is
sequestered cannot be used to offset increased emissions
elsewhere. In this way payments for ecosystem services are more
consistent with our policy principles because net emissions are not
increased as a result of implementation. Instead, potential
emissions are prevented.
Management that protects the soil is also valuable for
protecting watershed, wildlife, recreational, and scenic values,
as well as the long term productivity of the land for grazing. The
ecosystem service of preventing carbon loss can be seen as bundled
within a diversity of potential ecosystem services that together
might justify a payment that would alter the management and long
term use of the land. If, together, these payments resulted in
preventing the conversion of rangelands into intensive agricultural
or other uses, the policy itself, though not the changes in
management, could prevent additional carbon emissions and
would be consistent with our suggested policy principles. The issue
of avoided conversion is more fully discussed below.
7.2. Long-term conservation: payments for avoided conversion or
restoration
Payments for avoided conversion or loss of forests have
attracted growing international interest and might be considered
for U.S. rangelands (Verified Carbon Standard, 2012). If correctly
designed, payments for avoided rangeland conversion or restoring
marginal croplands to rangelands could yield additional rangeland
carbon storage and are consistent with our suggested policy
principles. Although the process is very slow, rangelands do
accumulate carbon over time, and land uses associated with them
are much less carbon-intensive than the common alternative of
industrial agriculture. In addition to the various greenhouse gas
emissions associated with crop inputs such as fertilizer, croplands
have been found to be net carbon emitters when consumption and
emissions of crops is considered, while rangelands are generally
either carbon neutral or small sinks (Perez-Quezada et al., 2010;
Uri and Bloodworth, 2000). Moreover, conversion from grassland
to annual crops can lead to a 60% loss of soil carbon stocks and a
95% loss of above ground carbon (FAO, 2009). Therefore, keeping
lands as rangelands, with associated land uses such as extensive
grazing, can at least prevent increases in emissions, and restoring
croplands to rangelands can yield increased sequestration.
To a certain extent, such a system already exists and has been
successful on a carbon basis–the Conservation Reserve Program of
the U.S. Department of Agriculture (Uri and Bloodworth, 2000). By
paying farmers to hold land in reserve, the CRP caused farmers to
restore some of their land to rangelands. Studies of recovering
rangelands generally and CRP land in particular show that they
quickly accumulate and store soil carbon (IPCC, 2007; Gebhart
et al., 1994; Derner and Schuman, 2007). Similar policies might
avoid the transaction costs of cap-and-trade as long as the
payment system is not tied to annual flows of carbon. It also does
not create problems of additionality or environmental robustness
as long as the carbon sequestered does not count as a credit against
other emissions.
A policy not based on payment for actual carbon flux may be
hard to justify politically since payments and sequestration
services would not be directly linked. There would be no
mechanism to determine how much carbon a single landowner
was storing, or how much flux to the atmosphere was prevented, in
return for the payment. One option might be payments based on an
algorithm that takes into account risk of conversion and loss of
future flows at a time-discounted rate, again using modeling or
proxies for estimation and verification. Another option might be a
voluntary, market-based scheme, along the lines of conservation
easements, in which individuals or organizations would pay
landholders either not to convert extant rangelands or to restore
degraded croplands to rangelands. Although the carbon seques-
tration benefits themselves might not generate the necessary
interest, they could usefully complement other conservation
values that are currently driving such programs.
Preserving rangelands over the long run or converting degraded
croplands to rangeland is consistent with our policy principles and
would result in more carbon sequestered by playing to the
strengths of rangelands as carbon sinks–they are nationally
extensive and associated with low intensity land uses. However,
the great extensiveness of rangelands may also prove to be a
problem if this policy is widely pursued. Given that there are so
many acres of rangelands in the United States, the size of
payments, if paid to all owners of rangelands, may have to be
small per unit area. If too small, the size of payment could limit the
effectiveness of the policy.
7.3. Carbon tax
The final policy type to be considered is a national carbon tax.
Although it is not generally seen to affect rangelands directly
because rangelands per se would not be taxed, a strong carbon tax
can indirectly affect rangelands through impacts to intensive
agriculture, although the impacts of such a policy are difficult to
predict.
Currently ranchers relying on extensive rangeland grazing are
of less economic importance relative to intensive beef producers
whose cattle gain weight more quickly and predictably. The
unpredictability of forage production on arid rangelands is one of
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the factors that make returns low and variable. In 2009, less than
3% of the beef market in the United States came from production
systems other than conventional grain-finished beef (Mathews
and Johnson, 2010). A carbon tax could tip the balance of beef
production more towards extensive grazing by increasing the cost
of corn and other feeds, thereby increasing total costs of feedlot
beef production.
Rangeland beef production is based on extensive grazing
requiring few energy, crop, and labor inputs. The production
system is not without greenhouse gas impacts: ranchers use
gasoline for their vehicles, grass-fed cattle have been shown to
release more methane per pound of gain than feedlot cattle
(Harper et al., 1999; Peters et al., 2010), and many ranchers fertilize
and irrigate some small pasture areas to produce supplemental
feed for their herd or move their herds to pastures in different
biomes seasonally in an effort to buffer highly variable rangeland
production. Yet, range cattle feed primarily on native or natural-
ized grass grown without fertilization or irrigation, waste is
returned to the soil to fertilize future grass production and
contribute to soil organic carbon, and density of animals is low.
Feedlots, in contrast, rely primarily on industrially produced corn,
which may be the most carbon intensive conventional crop in the
US, as well as other feedstuffs produced with inputs that are large
sources of GHG emissions, and they support a high density of
livestock with consequently high emissions (Harrington and Lu,
2002; West and Marland, 2002).
From an overall climate change standpoint, grass-fed beef is
likely preferable because of the broader long-term context of
production. For example, if corn prices were to increase enough to
make rangeland beef cost-competitive, causing the industry to
shift back to more range-fed beef, it would diminish the largest
market for commodity corn and reduce the number of cattle
produced. This, in turn, might encourage more Midwestern farms
to convert from corn to other crops that are less energy intensive,
perhaps, on more marginal lands, even to grazing. On the more
mesic former rangelands where corn predominates, carbon
sequestration from such a shift would be both larger and more
certain to occur than on drier rangeland types. It is also likely that
fewer cattle could be produced overall, potentially reducing total
methane emissions.
Indirectly, then, a carbon tax might encourage conversion from
cultivation back to rangelands, a move that would vastly increase
the sequestration of carbon, both through avoided emissions and
the high soil sequestration rates seen on restored agricultural lands
(Derner and Schuman, 2007; Morgan et al., 2010).
However, predicting the effects of a national carbon tax is
complicated by the global nature of beef production. If the result of
a carbon tax is higher beef prices, as predicted, then there could be
unfavorable land use effects on a global, or even national scale, as
the higher beef prices entice more people to raise beef and either
raise grain to feed those cattle or expand grazing area at the
expense of carbon rich forests. These indirect international land
use effects and equilibrium price dynamics, which have been
hypothesized for biofuel production (Searchinger et al., 2008),
must be taken into account when considering whether a carbon tax
would be the best policy to promote carbon sequestration on
rangelands.
8. Conclusions
Attention to ecological site and the link to ‘‘best fit’’ ecological
models is essential to assessing carbon sequestration potential on
rangelands as well as other types of natural landscapes (Norgaard,
2009). Reviewing what is known of rangeland ecological dynamics
and the distribution and extent of ecological sites suggests that
management cannot reliably increase carbon uptake on most
rangelands. Carbon uptake on the relatively small remaining areas
of more mesic and productive rangelands may be more responsive
to management, but care must be taken to differentiate these
rangelands and their potential from typical arid and semiarid
rangelands. On the other hand, over the great extent of rangeland,
protection of carbon stocks present in soils or conversion to
rangelands from more intensive uses would make a significant
contribution to global carbon balance. Management strategies to
accomplish this should be developed and implemented.
We have described four policy principles that will ensure that
policies for rangeland carbon sequestration are consistent with the
ecological science of rangelands: policies should (1) not require
short-term explicit carbon accounting (2) not assume that changes
in management can create additional carbon sequestration (3) not
use arid rangeland sequestration that is not consistent and verified
to offset emissions; and (4) should focus on conserving rangelands
or reverting degraded agricultural lands to rangeland.
Applying these principles to proposed rangeland carbon
sequestration policies shows that cap and trade is problematic in
terms of environmental integrity, transaction costs, and addition-
ality. Particularly troubling is a common assertion that absent the
ability to measure or assure that additional carbon is sequestered as
a result of managementpractices, carbon paymentssupport an array
of co-benefits from rangeland ecosystems. While this is no doubt
true, under cap and trade, credits allowing increased carbon
emissions are granted to others under the assumption that they
are compensated for by increased uptake in the rangeland system.
Payment for ecosystem services may be consistent with the
ecological dynamics of rangelands when they reward carbon
storage and management practices that protect the carbon in soils,
rather than increased sequestration. Payments for avoided
conversion (‘‘avoided de-rangification’’) or reconversion of mar-
ginal agricultural lands to rangelands are also consistent with the
science of carbon dynamics on rangelands, but low prices may
limit the effectiveness of direct payments. Perhaps through
bundling them with payments for other co-benefits of extensive
production and land conservation, payments could reach a level
allowing them be a viable strategy. This type of policy could
support enhancement of soil storage capacity, by conserving
rangelands so that they can continue to store carbon over the long-
term. It would also preserve the contribution of rangelands to the
global food supply. A national carbon tax, if high enough and
properly designed, could have a strong, though indirect, impact on
rangeland carbon sequestration if it encouraged conversion from
cropping to rangeland grazing.
Carbon sequestration incentives will be a disappointing source
of income for ranchers on arid lands unless they are part of a
realistic policy program that can provide consistent, even if small,
income augmentation. Ranchers are receptive to diversifying
income streams as long as such income is consistent with their
personal and stewardship goals, is voluntary, and avoids an
intrusive government role (Ma and Coppock, 2012; Cheatum et al.,
2011; Gosnell et al., 2011). Research has also shown that ranchers
in general are not particularly familiar with the details of carbon
sequestration (Cheatum et al., 2011; Ma and Coppock, 2012) and
that their response to the likely financial benefits alone as an
incentive is ‘‘tepid’’ (Ma and Coppock, 2012). Ranchers reported
that increasing forage production, improving soil quality, increas-
ing water retention and infiltration and enhancing drought
resistance were more compelling reasons to change management
practices (Ma and Coppock, 2012). Bundling of carbon sequestra-
tion with an array of other co-benefits (Cheatum et al., 2011)as
part of a payment for ecosystem services or avoided conversion
program may be the best received type of approach for the
ranching community, as well as being more suited to the ecology of
arid ecosystems.
K. Booker et al. / Global Environmental Change 23 (2013) 240–251
249

At the international level, a diverse array of programs have
attempted to cope with the ecological constraints of rangelands.
However, many of the dynamics underlying rangeland carbon
fluxes and management are not well understood and warrant
further study. In particular, modeling and remote sensing methods
to reduce transaction costs of soil carbon measurement and
monitoring are promising. In general, more information about
carbon dynamics along the gradient from arid to more mesic
rangeland ecosystems is needed. Research that increases the data
on the ecological dynamics of rangeland ecological sites, and how
they pertain to carbon sequestration, is also needed. Other topics in
need of further inquiry include a full systems-level carbon
comparisons of feedlot and grass-fed beef.
Acknowledgements
We thank participants and presenters in the Berkeley Institute
of the Environment roundtable on ‘‘Rangelands and Climate
Change’’. Funding for the roundtable was provided by the Berkeley
Institute of the Environment at UC Berkeley. We thank Joel Brown
for his advice about soils, and our reviewers for their valuable
suggestions.
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