REVIEW

Mineral soil carbon fluxes in forests and implications for

carbon balance assessments

THOMAS BUCHHOLZ * , ANDREW J. FRIEDLAND * , CLAIRE E. HORNIG * ,

WILLIAM S. KEETON † , GIULIANA ZANCHI ‡ and JARED NUNERY§

*Environmental Studies Program, 6182 Steele Hall, Rm. 113, Dartmouth College, Hanover, NH 03755, USA, †Rubenstein

School of Environment and Natural Resources, 209 Hills Building, University of Vermont, 81 Carrigan Drive, Burlington, VT

05405, USA, ‡Physical geography and ecosystem science, Lund University, S€olvegatan 12, S-223 62, Lund, Sweden, §Vermont

Department of Forest, Parks and Recreation, 1229 Portland Street, St. Johnsbury, VT 05819, USA

Abstract

Forest carbon cycles play an important role in efforts to understand and mitigate climate change. Large amounts

of carbon (C) are stored in deep mineral forest soils, but are often not considered in accounting for global C

fluxes because mineral soil C is commonly thought to be relatively stable. We explore C fluxes associated with

forest management practices by examining existing data on forest C fluxes in the northeastern US. Our findings

demonstrate that mineral soil C can play an important role in C emissions, especially when considering intensive forest management practices. Such practices are known to cause a high aboveground C flux to the atmosphere, but there is evidence that they can also promote comparably high and long-term belowground C fluxes.

If these additional fluxes are widespread in forests, recommendations for increased reliance on forest biomass

may need to be reevaluated. Furthermore, existing protocols for the monitoring of forest C often ignore mineral

soil C due to lack of data. Forest C analyses will be incomplete until this problem is resolved.

Keywords: carbon accounting, deep soil mineral carbon, Forest carbon pool assessments, forest soil, stand level carbon dynamics

Received 16 October 2012 and accepted 8 November 2012

Introduction

Analysis of forest carbon (C) cycles is central to understanding and mitigating climate change (IPCC, 2007).

Globally, forests store an estimated 861 gigatons of C

(Lal, 2008), representing 25%–27% of the total terrestrial

C pool of ~3,300 gigatons C, and have a sink capacity of

around 2.4 gigatons C per year (Pan et al., 2011). Forests

also account for 16%–20% of total annual anthropogenic

Carbon dioxide (CO2) emissions (Lal, 2005, 2008), mainly

due to ongoing net deforestation.

Understanding forest C cycles requires an in-depth

analysis of the storage in and fluxes among different

forest C pools. These pools include aboveground live

and dead biomass, as well as the belowground organic

soil horizon, mineral soil horizon and roots. Accurate

accounting of these pools is a precondition for national

forest C statistics reported to the United Nations Framework Convention on Climate Change (UNFCCC, 2005),

quantifying the CO2 emissions associated with harvesting and processing forest products (Werner et al., 2010)

and forest management practices (Nunery & Keeton,

2010), bioenergy C accounting frameworks (European

Commission, 2010; EPA, 2011), calculating C emissions

from bioenergy systems (Zanchi et al., 2012), or forestbased C offset markets.

Forest soils are a critical part of any forest C accounting effort (Fig. 1). Forest soils are the largest active terrestrial C pool (2,500 gigatons to a 1 m depth, Lal, 2008)

and account for 34% of the global soil C pool (Pan et al.,

2011). Soil characteristics, climate, and land use change

affect the rate of biological and chemical processes that,

in turn, impact soil C content on timescales ranging

from hours to thousands of years (Fontaine et al., 2007;

Trumbore 2009). C input to the soil comes from roots,

dead trees, and litterfall, and is released through root

and heterotrophic respiration (Dixon 1994; Fahey et al.,

2005). In the case of soil organic C in the forest floor, the

relationships between forest harvest practices and soil C

responses are increasingly well understood (Lal, 2005):

the loss of aboveground biomass results in increased

solar radiation to the soil and decreased evapotranspiration from the soil (Lal, 2005; Mariani et al., 2006), and

soils are often compacted and may experience mechanical mixing, although this is typically confined to the

organic horizon (Yanai et al., 2003). These changes

impact decomposition rates and soil microbial communities, potentially increasing soil C respiration rates

(Diochon et al., 2009).

Correspondence: Thomas Buchholz, tel. +1 802 881 5590, fax +1 603

646 1682, e-mail: thomasbookwood@gmail.com

© 2013 Blackwell Publishing Ltd 1

GCB Bioenergy (2013), doi: 10.1111/gcbb.12044
The mineral component of forest soils stores more

than 50% of the C in forest soils (Jobbagy & Jackson,

2000; Fig. 1). However, due to limited understanding of

mineral soil C fluxes in response to forest harvesting (e.

g. Zummo & Friedland, 2011), and because mineral soil

C pools are commonly assumed to be stable (e.g. Smith

et al., 2006), mineral soil C fluxes are not considered in

empirical simulation models commonly used to project

forest C dynamics over time (e.g. the U.S. Forest Service’s Forest Vegetation Simulator (FVS), Hoover & Rebain, 2011). In the policy realm, the lack of sound

scientific data on mineral soil C in forests often leads to

exemptions for reporting mineral soil C storage capacity

and C stock changes. For example, under C market

accounting protocol soil C is often optional or excluded

in forestry projects (e.g. VCS, 2012).

In this article, we review data from the northern temperate forest in eastern North America as well as anecdotal evidence from a growing body of literature

around the globe to (i) review the current knowledge,

practice, and requirements for including and quantifying mineral soil C balances in forest C accounting systems, (ii) elucidate how recent insights into mineral soil

C fluxes challenge conventional wisdom in forest C

accounting, (iii) describe the current limitations to quantifying and tracking mineral soil C, and (iv) suggest

steps to incorporate mineral soil C fluxes in forest C

accounting, policy, and management.

Current views on soil C

Conventional wisdom

Contemporary practice-oriented forest C accounting

models and literature on soil C fluxes regularly exhibit

two assumptions: (i) only C fluxes to and from the upper

organic soil horizons (<40 cm) and root compartment

respond to forest management practices and (ii) a postharvest soil C equilibrium is reached in the short-term

within 20 years, even under intense harvest practices (e.

g. Johnson & Curtis, 2001 on a global scale; Jones et al.,

2011 for bioenergy applications in New Zealand). Deeper

soil C pools are considered stable. For instance, several

recent review papers on landscape C analysis provide an

in-depth overview of C accounting, but avoid any reference to mineral soil C (e.g. Ryan et al., 2010; McKinley

et al., 2011; or Fahey et al., 2010 for forestry, Conant et al.,

2011 for agriculture). Likewise, the US-wide lookup

tables for soil C fluxes in conjunction with clearcutting

regimes provided by Smith et al. (2006) assume that the

mineral soil C pool remains constant throughout the

125 years postharvest. Meanwhile, many studies analyzing soil C changes focus on examples such as converting

agricultural land to forest or vice versa (e.g. Cowie et al.,

2006; Searchinger et al., 2009) rather than soil C change

on land continuously categorized as forested.

As a result of these assumptions, forest C accounting

frameworks frequently consider upper soil horizon C

fluxes only (e.g. IPCC, 2006). Partly due to this exclusion of soil C, study results then find rapid net C-emission benefits from intensified forestry (e.g. Perez-Garcia

et al., 2005; Cowie et al., 2006) and support the substitution of forest-based products and fuels for energy-intensive products and fossil fuels, respectively.

Such outcomes reinforce the prevailing wisdom that

“research suggests that harvest operations have no

effect on soil carbon” (Perschel et al., 2007 pg 25, for the

Northeastern US), and have led some researchers to

exclude soil C from their analyses until further evidence

of harvest impacts on this C pool is found (e.g. Lippke

et al., 2011). In other cases, the change in mineral soil C

as a response to forest management is acknowledged,

but omitted due to data uncertainty (e.g. Holtsmark,

2012), or the change is declared marginal in comparison

to potential greenhouse gas mitigation gains (e.g. Cowie

et al., 2006). Sometimes it is discussed as a potential

additional and marketable C sink rather than a potential

source (e.g. Lorenz et al., 2011).

These approaches are not unreasonable given the

many articles that have indicated that mineral soil C in

managed forests is stable. However, additional evidence

points to changes in mineral soil C brought on by harvesting, and its potentially large impact makes it especially worthy of consideration. We present this evidence

in the following section.

Challenges to the conventional wisdom

Further research suggests that mineral soil C responses

can be highly variable depending on harvesting intensity, surface disturbance, and soil type (e.g. Nave et al.,

Foliage,

branches, twigs:

29.2 Mg ha–1

Bark and bole:

65.8 Mg ha–1

Standing dead:

6.6 Mg ha–1

Coarse woody

debris:

4.7 Mg ha–1

Roots:

25.1 Mg ha–1

Dead roots:

1.9 Mg ha–1

Soil (organic

horizons)

29.7 Mg ha–1

Soil (0–20 cm):

57.7 Mg ha–1

Soil (+20 cm):

70.0 Mg ha–1

Fig. 1 Major forest carbon (C) pools for a 100-year old hardwood forest in New Hampshire (Fahey et al., 2005). The

depicted size of the pools is in proportion to their relative C

content.

© 2013 Blackwell Publishing Ltd, GCB Bioenergy, doi: 10.1111/gcbb.12044

2 T. BUCHHOLZ ET AL.
2011). However, several recent studies suggest that certain forest harvesting practices might cause significant

and long-term C losses in the mineral soil. Pregitzer &

Euskirchen (2004) observed a decrease in mineral soil C

with decreasing age class in a study encompassing tropical, temperate and boreal forest biomes. Johnson (1995)

noticed that “the decline in mineral soil C was significant” over 8 years following whole tree harvest clearcuts in the temperate US northern hardwood forest (pg

1349). For the same region, Zummo & Friedland (2011)

found, in Spodosols, a significantly lower total C

amount in the 10–45 cm depth soil horizons of forests

55 years after clear-cuts. In a global meta-analysis

encompassing 432 sites, Nave et al. (2011) detected “a

significant decline in deep mineral soil Carbon” (pg

860), also for Spodosols, by 9% following harvest,

although no significant mineral soil C loss was detected

when other soil types were included in the analysis.

Diochon et al. (2009) observed a 50% loss in mineral soil

C 30 years following harvest in the boreal forest of

Nova Scotia. In temperate forests in the Great Lakes

region, Tang et al. (2009) discovered that “total soil C at

0–60 cm initially decreased after harvest, and increased

after stands established” (pg 153). There is no consensus

on what processes cause the decrease in mineral soil C

(Jandl et al., 2007; Diochon et al., 2009; Zummo & Friedland, 2011).

Including forest soil C fluxes in forest management

analysis could have significant impacts on modeling

results. For instance, Zanchi et al. (2012) included volatile soil C fluxes in their analysis of increasing fellings

in an Austrian forest from 60% to 80% of net annual

increment. Their model assumed only moderate forest

C losses with the loss rate peaking at around 0.03 Mg/

ha after ~60 years. In this work, we assessed that the

accumulated soil C losses constituted 12% of total forest

C losses or 3.6 Mg/ha (Fig. 2a) once a C payback period

was reached after 175 years compared to a coal substitution scenario (Fig. 2b). Excluding even these presumably minor changes in soil C would lead to a

significantly shorter (by about 25 years) C payback period. Including forest soil C fluxes is especially important

in bioenergy C accounting because residues are often

the primary fuel considered and residue decomposition

patterns have a significant impact on litter and soil C

stock changes (Repo et al., 2011).

Limitations to forest soil C analysis

Forest mineral soil datasets

Collecting robust data on mineral soil C has proven difficult due to the labor necessary to make the required

measurements. As a result, regional, national, and global

soil C datasets are often restricted to the upper soil

strata, utilize only short-term observations, and are incapable of associating forest management regimes to

belowground C fluxes. Examples include the periodic

nationwide U.S. Forest Inventory and Analysis National

Program, which samples several aboveground forest C

attributes, but has not been designed to provide sufficient data to analyze upper-strata soil (Fahey et al.,

2010) or deep soil characteristics (Harrison et al., 2010)

in response to different forest management regimes.

The soil-specific datasets that do exist are extensive,

but often are not robust enough to provide stand- or

treatment-specific analysis of mineral soil C. Examples

include the USDA Natural Resources Conservation Service STATSGO database or the Harmonized World Soils

Database, which contain estimates of mineral soil C up

to 1 m depth, but without any connection to aboveground C data. One of the largest available soil C databases, the ISRIC_WISE World Inventory of Soil

Emissions Potentials, has ~10,250 ISRIC_WISE plots, but

most of these include soil C measurements only to a

depth of 30 cm (Batjes, 2011). The spatially explicit

–3.5

–3.0

–2.5

–2.0

–1.5

–1.0

–0.5

0.0

0.5

0 50 100 150 200

Forest C stock change (Gg C)

Year

Tree C

LiƩer C

Soil C

0

2

4

6

8

10

12

0 50 100 150 200

Emissions (Gg CO2)

Year

Biomass

Biomass_no_soil

Fossil_fuel

(a)

(b)

Fig. 2 Forest Carbon (C) stock change (a) and effects of

excluding forest soil C (b) in a bioenergy substitution scenario

modeling C fluxes in a 90 ha Austrian forest when increasing

fellings from 60% to 80% of net annual increment (Zanchi et al.

in press).

© 2013 Blackwell Publishing Ltd, GCB Bioenergy, doi: 10.1111/gcbb.12044

MINERAL SOIL CARBON FLUXES IN FORESTS 3
design for large-scale analysis prohibits the use of these

datasets for analysis in smaller scales where “the use of

local data is preferable” (Smith et al., 2012 pg 2091).

Inclusion of mineral soil C in standard forest C

accounting tools

Although reliable soil C data are a precondition for full

forest C accounting, another requirement for successful

integration of soil C fluxes is the availability of C

accounting tools that are capable of accounting for all

above- and belowground C fluxes. A range of existing

forest C accounting models have this capacity. One

example is FORCARB (Birdsey, 2006). Another, the ecosystem process model Biome-BGC, is able to include

forest mineral soil C fluxes (e.g. Peckham & Gower,

2011 using mineral soil data from the STATSGO database). Models that use forest growth models for data

input, such as the GORCAM model, are among the

most flexible ones. It calculates C fluxes to and from the

atmosphere for different forest management strategies,

incorporates C pools in wood products and fuels (Schlamadinger & Marland, 1996), and allows independent

integration of soil C (Zanchi et al., 2012).

However, such models acknowledge limitations by

available data. For example, the fact that “empirical evidence is lacking for consistent changes in average organic

C stocks in the mineral soil following harvesting and

immediate regeneration” (pg 376) forced Heath et al.

(2002) to use soil C constants in FORCARB instead of a

dynamic soil C model for cases of continuous forest cover.

Furthermore, many forest and soil C modeling tools offer

no link to the mineral soil C pool. Examples include the

FVS and the Carbon OnLine Estimation tool (COLE). Mineral soil C fluxes can be modeled in standard soil C models such as CENTURY and Yasso07 based on land

management. However, both models are focusing on soil

organic C and the mineral soil C component is less established (Hilinski, 2001; Tuomi et al., 2011).

Soil C in regulatory forest C accounting frameworks

Periodic national and international forest C inventories

provide a snapshot in time of different forest C pools (e.

g. IPCC, 2006 or EIA, 2011 for an international or

national level, respectively). A range of voluntary as

well as mandatory frameworks also specify forest C

accounting standards. Many require tracking and identifying causal relationships between C fluxes and forest

management practices. Key questions are which forest

C pools to include and which downstream emissions

should be accounted for [e.g. C stored in wood products

or emissions avoided by substituting wood products for

other materials (Law & Harmon, 2011)].

In contrast to the inventory and monitoring frameworks mentioned above, frameworks or protocols

developed for marketing forest C storage and sequestration are based on modeling future trends. Most of these

C market protocols acknowledge the role of mineral soil

C in forest ecosystems, especially when land use change

occurs. However, the complex relationship between

below- and aboveground C pools makes the inclusion

of mineral soil C in these market protocols difficult and

therefore rare when no land use change is assumed

(Table 1, see also Fahey et al., 2010) – as is the case in

projects focusing on improving forest management or

reducing emissions from deforestation. The reasoning is

that such projects will increase aboveground C through

reduced harvest intensity such as extended rotation

periods, and therefore will not reduce soil C (e.g. Gershenson & Barsimantov, 2011). For instance, the Verified

Carbon Standard (VCS) has soil C monitoring only as

an option in its protocol. Even if soil C is included, VCS

limits measurements of soil C to the top 30 cm, therefore excluding part of the mineral soil C. Meanwhile,

the Climate Action Reserve (CAR) acknowledges the

potential of soil C fluxes triggered by harvest practices

and long recovery times (Gershenson & Barsimantov,

2011) and requires soil C inclusion in any registered

project but it also restricts soil monitoring to the top

30 cm (CAR, 2012). Given that the California Compliance Offset Program, a recently approved regulatory

protocol administered by the California Air Resources

Board, used a previous version of the CAR forestry protocol for its offset protocol (ARB, 2011) and the extensive changes made between the two versions, it is not

clear if this new version will be endorsed by ARB. Mineral soil C losses are therefore not completely considered in these programs. Similarly, the Regional

Greenhouse Gas Initiative (RGGI), a framework specific

to large electric utilities in the Northeastern US, does

stress the monitoring of soil C prior to project start but

fails to specify a sampling protocol.

An additional, nonmarket protocol, the IPCC C inventory standards for managed forests, takes a similar

approach to the market-based protocols. As the default

for Tier 1 applications, it assumes no soil C change if

land remains forest and applies a constant mineral soil

C factor and the assumption that SOC stabilizes within

20 years for land use change scenarios (Table 1). Tier 2

and 3 applications consider temporal dynamics in soil C

and allow the use of soil C model estimates and

national inventories (Ortiz et al., 2011), but are rarely

applied (e.g. Batjes, 2011) due to inherent uncertainties.

Although fluxes in forest mineral soil C are recognized,

forest management is not considered a potential highimpact factor on mineral soil C fluxes (IPCC, 2006,

chapter 3.2).

© 2013 Blackwell Publishing Ltd, GCB Bioenergy, doi: 10.1111/gcbb.12044

4 T. BUCHHOLZ ET AL.
Protocols of potential global significance are still

under development, including a potential UN-REDD

(Reducing Emissions from Deforestation and Forest

Degradation, Goetz et al., 2010) instrument as well as an

effort by the EU to account for C in international biomass trade (European Commission, 2010). Both documents acknowledge the potential impact of forest

management practices on and the significance of the

mineral soil C pool, but fall short of outlining methods

on how to account for mineral soil C fluxes. Without a

clear accounting procedure for mineral soil C fluxes, it

is not clear how future voluntary or regulatory frameworks will accurately account for total forest C flux.

Conclusions

Mitigating climate change through forest management

strategies is a major component of forest policy, but the

technical and organizational difficulties of long-term

and mineral soil C research have led current practice to

assume stable or replenished forest soil C pools 20 years

postharvest (e.g. IPCC, 2006). Recent forest mineral soil

C research (e.g. Diochon et al., 2009; Zummo & Friedland, 2011) and forest C accounting studies (e.g. Zanchi

et al., 2012) suggest that the exclusion of mineral soil C

in forest C flux analysis can result in not fully accounting for the C flux under specific forest management or

site conditions. These results also emphasize the importance of considering sufficiently long temporal scales in

forest C flux analysis (Marland, 2011). Forest management alternatives that result in a delayed or avoided

release of GHGs might be more effective in mitigating

climate change relative to a scenario characterized by

high, although short-term, CO2 emissions.

The potentially significant and long-term release of

mineral soil C by practices such as a clearcutting (Zummo & Friedland, 2011) might have profound impacts

on the forest use vs. preservation debate. Although

some studies stress the beneficial impact of intensified

forest management for biomass on atmospheric CO2

reduction targets by replacing fossil-fuel intensive products (Werner et al., 2010), other studies suggest that less

intensive forest management strategies might be more

beneficial from a climate change perspective (Nunery &

Keeton, 2010). While such contradictory results can be

largely explained by differences in boundary settings (e.

g. the inclusion or exclusion of product substitution),

our research indicates that a more nuanced debate

regarding forest management strategies is appropriate.

This debate must include mineral soil C. Including even

a moderate forest soil C loss can have significant

impacts on calculating a C payback time when assessing

bioenergy substitution scenarios for fossil fuel use comparing forest-based bioenergy with fossil fuel scenarios

(Fig. 2). Likewise, the timing of emissions is of considerable concern when aiming for climate change mitigating

strategies even if the CO2 is re-sequestered at a later

stage (Kendall et al., 2009; Cherubini et al., 2011; Sathre

& Gustavsson, 2011).

Regarding forest C credit markets, forestry projects

generally realize climate benefits through reducing

management impacts (e.g. through retention harvesting

systems and extended rotation periods). Avoiding

intensive harvesting practices might justify the exclusion of the mineral soil C in forest C accounting in this

case where a stable or enlarged forest C pool is the goal.

There is a need to bridge knowledge gaps in mineral

soil C fluxes and to accurately, cost-effectively and

transparently monitor all major forest C pools, including soil. Techniques for increasing the frequency and

efficiency of mineral soil sampling may be a major first

step toward better understanding of mineral soil C

Table 1 Carbon protocol requirements on monitoring soil

Carbon (C)

Protocol Soil C requirements Source

American Carbon

Registry (ACR)

Soil C not specified ACR, 2011;

Verified Carbon

Standard (VCS)

Soil C not included or

optional in improved

forest management

or reduced emissions

from deforestation

projects

VCS, 2012;

Assessments usually to

a depth of 30 cm only

Climate Action

Reserve (CAR)

Soil C included in all

forest projects

CAR, 2012;

Assessments usually to

a depth of 30 cm only.

Regional Greenhouse

Gas Initiative (RGGI)

Soil C inclusion

mandatory

RGGI, 2008;

No specification of

assessment and

monitoring methods

IPCC inventory

standard for

managed forests

Impact of forest

management regime

on mineral soil C

considered marginal

IPCC, 2006

Range of C assessment

methods offered:

Tier 1 assumes constant

mineral soil C if land

remains forest

Tier 2 and 3 allow the

use of soil C model

estimates and national

inventories if preferred

by contender

© 2013 Blackwell Publishing Ltd, GCB Bioenergy, doi: 10.1111/gcbb.12044

MINERAL SOIL CARBON FLUXES IN FORESTS 5
fluxes (Harrison et al., 2010). New technologies, such as

inelastic neutron scattering, laser-induced breakdown

spectroscopy or ground-penetrating radar, might allow

cost-effective in situ, nondestructive analysis of deeper

soil strata (Johnston et al., 2004). This increased production of data is essential to meaningfully couple land

cover management methods with soil structure which

in turn can be built into larger soil datasets. Only then

can regional datasets be justifiably applied in smaller

spatial scales as well (Smith et al., 2012) and soil C pools

can be widely included in forest C flux analysis.

As of now, ecosystem process models or forest product

C accounting models offer increasingly sophisticated analytical capacities. A combination of available models

could more fully account for above- and belowground C

fluxes within a given project than any single model existing today. Using several models or the extended use of

sensitivity analysis would provide a helpful step toward

improved C monitoring and the communication of C flux

uncertainties to researchers in other fields as well as practitioners. However, until the production of robust science

describing the relationship of forest management practices to mineral soil C matures, we recommend a precautionary approach by avoiding intensified forest

management practices such as an increased harvest frequency and intensity if the primary forest management

objective is to increase forest C storage (Jandl et al., 2007).

Acknowledgements

This work was supported by awards to A. Friedland from the

Northeastern States Research Cooperative and the Porter Fund.

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