Download Influence of Climate and Land Use Change on Carbon in Agriculture

4
Influence of Climate and Land Use
Change on Carbon in Agriculture,
Forest, and Peatland Ecosystems
across Canada
J.S. Bhatti
Natural Resources Canada, Canadian Forest Service, Northern Forestry Centre,
Edmonton, AB, Canada
C. Tarnocai
Agriculture and Agri-Food Canada, Research Branch (ECORC), Ottawa, ON, Canada
Carbon is exchanged between terrestrial ecosystems and the
atmosphere through photosynthesis, respiration, decomposition, and combustion, which could be net sources or sinks. This source or sink status is not a static
ecosystem characteristic, however, but can shift over time as a result of changes
in the physical, chemical, or biological processes within these systems (Kauppi
et al., 2001). Consequently, changes in climate will result in modifications to the
C fluxes of both natural and managed ecosystems throughout Canada. Therefore, there is a strong need to quantify these fluxes in relation to various changes
resulting from alternate management options. Quantifying these fluxes can be
challenging because of difficulties inherent in detecting very small changes relative to the large total C pools. When studying C stocks and fluxes in agriculture,
forest, and peatland ecosystems, it is often helpful to consider the C stored in
the basic components of biomass and soil (Fig. 4–1). Following this structured
approach, it is then important to understand the factors able to affect changes in C
fluxes from different pools, and to identify the key factors that determine whether
these pools function as sinks or sources of C.
The objective of this chapter is to illustrate the influence of climate change on
the C pools in managed ecosystems across Canada. Specifically, the ecosystems
considered are agricultural environments, forests, and peatlands. Much of the
focus of discussion is on the interaction between different C pools, under changing climatic conditions, with specific reference to C fluxes in the three managed
ecosystems. The need for the intense management of Canadian peatlands–wetlands, forests, and agricultural ecosystems is increasing to fulfill the requirements
of food, livestock feed, fiber, and fuel production.
Soil Carbon Sequestration and the Greenhouse Effect, 2nd edition. SSSA Special Publication 57.
Copyright © 2009. ASA‒CSSA‒SSSA, 677 S. Segoe Rd., Madison, WI 53711, USA.
47
48
Bhatti & Tarnocai
Fig. 4–1. Components of the different managed ecosystems’ carbon stocks. (Modified from Lal, 2005.)
Carbon Stocks in Relation to Canadian Terrestrial Ecosystems
Agricultural Ecosystems
Due to climatic and soil limitations, only 7% of Canada’s landmass is under
agriculture (Natural Resources Canada, 2004) and contains about 9.36 Pg of soil
organic carbon (SOC). This is estimated to represent 6% of the total organic
C pool in all agricultural soils worldwide (Tarnocai and Lacelle, 1996). Large
amounts of organic matter occur in Chernozemic soils, formed in grassland ecosystems, and in Luvisolic soils, formed in clay-rich regions, both of which are
common throughout the prairies. The total soil C content is estimated at 4.6 to
16.4 kg m−2 for Chernozems, and between 5.3 and 7.8 kg m−2 for Luvisols (Anderson and Coleman, 1985). Tarnocai (1998) reported the average soil C content of
12.4 kg m−2 for Chernozemic soils and 9.3 kg m−2 for Luvisols (Table 4–1). Soils
common to eastern Canada, such as Gleysols (soils formed under conditions of
poor drainage) and Podzols (acidic soils formed under coniferous forests), also
have relatively large amounts of organic matter but do not make up a large proportion of agricultural land in Canada. Organic soils (those found in bogs, fens,
and some swamps and marshes) contain the greatest amount of organic matter
but occupy only a small area of Canada’s agricultural region.
Forest Ecosystems
Approximately half of the total Canadian landmass (410 Mha) is covered
by forest (Natural Resources Canada, 2004), with boreal forests as the dominant
forest type, spanning the entire width of the country. Distribution of different
ecozones in Canada is presented in Plate 4–1 (see color image section). The C pool
Influence of Climate and Land Use Change on Carbon...
49
Table 4–1. Amount of soil organic carbon in various soil orders of Canada. (Data modified from Tarnocai, 1998.)
Soil classification
U.S.
Canadian
Inceptisol
Boroll
Gelisol
Aqua suborders
Boralf, Udalf
Histosol
Spodosol
Entisol
Mollisol, Alfisol
Total
Brunisol
Chernozem
Cryosol
Gleysol
Luvisol
Organic
Podzol
Regosol
Solonetz
Soil C content
Surface†
Total‡
—— kg m−2 ——
5.2
9.3
7.4
12.4
11.3
40.6
11.7
20
4.9
9.3
18.7
133.7
9.9
19.3
5.6
11.8
5.8
11.5
Soil C pool
Surface†
Total‡
—— Pg ——
6.1
10.9
3.2
5.4
28.7
102.7
2.7
4.6
3
5.6
14.9
106.3
12.5
24.4
0.8
1.7
0.3
0.6
72.2
262.3
† 0–30 cm depth.
‡ 0–100 cm depth for mineral soils; total depth for organic soils (peat deposits).
in Canadian forests (excluding peat deposits) was estimated at 86.6 Pg C in 1989,
of which 71.7 Pg occurred in the dead organic matter in litter and soils, and 14.9
Pg in the living biomass (Apps et al., 1999). Carbon density (kilograms C per hectare) in Canadian forest ecosystems varies greatly among regions, reflecting the
differences in growing conditions and species. The Forest Ecosystem C Database
(FECD) compiled by Shaw et al. (2005) estimated the mean total ecosystem C for
the different terrestrial ecozones in Canada (Table 4–2). Estimates of average total
ecosystem C in Canadian ecozones range from 11.3 kg m−2 in the Western Taiga
Shield to 48.7 kg m−2 in the Pacific Maritime. In general, at the terrestrial ecozone
level, mineral soils account for 40 to 80% of total ecosystem C, while tree biomass
and organic soil horizons account for 15 to 40% and 2 to 25% of total ecosystem C,
Table 4–2. Distribution of total ecosystem carbon among tree biomass, soil organic
horizon, and mineral horizon in Canadian terrestrial ecozones. (Data compiled from
Shaw et al., 2005.)
Terrestrial ecozone
Western Taiga Shield
Western Boreal Shield
Hudson Plains
Boreal Plains
Taiga Plains
Montane Cordillera
Eastern Boreal Shield
Mixwood Plains
Atlantic Maritime
Pacific Maritime
C content
Organic
Mineral
Tree biomass
Total
horizon
horizon
——————————— kg m−2 ———————————
4.21
2.26
4.83
11.30
3.50
5.54
13.74
22.75
3.58
2.55
17.20
23.33
10.17
4.23
9.35
23.74
8.83
6.31
12.24
27.40
9.48
2.85
17.53
29.87
6.03
3.74
21.28
31.04
9.85
1.10
28.20
39.20
7.85
3.76
30.60
42.20
7.05
3.87
38.70
48.70
50
Bhatti & Tarnocai
respectively. Live tree biomass ranges from 3.5 kg m−2 in the Western Boreal Shield
to 10.2 kg m−2 in the Boreal Plains. Carbon estimates for mineral soil horizons range
from 4.8 kg m−2 in the Western Taiga Shield to 38.7 kg m−2 in the Pacific Maritime
ecozones (Table 4–2).
Soils
Since about 75% of Canada’s landscape is under tundra and boreal ecosystems, a large proportion of the global soil C pool occurs in Canadian soils
(Tarnocai and Lacelle, 1996). The Organic soil order has the highest average C
content followed by Cryosols, with the remaining soil orders containing much
lower amounts (Table 4–1). It is estimated that 84.3% (205 Pg) of the organic C pool
in Canadian soils occurs in mid to high latitudes, where two major soil orders
are Cryosols (103 Pg) and Organic soils (106 Pg). The other soil orders with their
respective C pools are Podzols (17 Pg), Brunisols (10 Pg), Luvisols (5 Pg), Gleysols
(4 Pg), Regosols (1 Pg), Chernozems (1 Pg), and Solonetz (<1 Pg) (Tarnocai, 1998).
Wetland–Peatland Ecosystems
Canada contains the world’s second (after Russia) largest area of peatlands
(114 Mha), which cover approximately 13% of the Canadian land area and 16% of
the soil area (Environment Canada, 1986). The majority of wetlands (96%) occur
in the Boreal (72 Mha) and Subarctic (37 Mha) regions. The dominant peatland
types are bogs (67%) and fens (32%), while swamps and marshes together account
for less than 1% of Canadian wetlands. The western boreal forests of Alberta, Saskatchewan, and Manitoba, along with British Columbia, contain approximately
40% of Canada’s peatland area, while eastern Canada (Ontario eastward) contains about 37%, and northern Canada (the three territories) approximately 23%
(National Wetlands Working Group, 1988).
Across Canada, peatlands are estimated to store 103 to 184 Pg C (Apps et
al., 1993). Tarnocai et al. (2005) estimate that the organic C pool of Canadian
peatlands is 147 Pg, of which 67% occurs in the Boreal and 30% in the Subarctic
regions. Together, these two peatland regions contain 97% of the organic C pool
of all Canadian peatlands. Soils in the permafrost region of Canada contain about
75% (197 Pg) of the total SOC (Table 4–3) in North America (Tarnocai et al., 2007).
Although mineral soils in the permafrost region occupy a larger area than the
peatlands, the organic soils contain approximately 57% (112 Pg) of the region’s
total soil organic soil pool. In continental western Canada (Alberta, Saskatchewan,
Table 4–3. Organic carbon mass in peatlands (organic soils) and mineral soils in various
permafrost zones in Canada. (Data compiled from Tarnocai et al., 2007.)
Permafrost zones
Continuous
Discontinuous
Sporadic
Isolated patches
Total
Carbon pool†
Organic soils
Mineral soils
Total
——————————— Pg ———————————
21.82
51.1
72.92
26.54
10.33
36.86
30.66
9.15
39.81
32.95
13.59
46.54
111.97
84.17
196.14
† Calculated to the total depth of the peat deposit.
Influence of Climate and Land Use Change on Carbon...
51
and Manitoba), peatlands contain 48 Pg C, with 42 Pg C stored as peat, and 6 Pg C
as living, aboveground biomass (Vitt et al., 2000).
Carbon Exchange
A terrestrial ecosystem can be either a sink or a source of atmospheric carbon dioxide (CO2). The ecosystem is a sink when photosynthesis results in a net
increase in the vegetation C pool, which may subsequently enrich the soil C pool
via input of vegetation-derived biomass. The ecosystem is a source when processes, such as decomposition and fire, release C to the atmosphere in excess of
that fixed by photosynthesizing vegetation. The net C balance of the ecosystem is
estimated as the net change of ecosystem C pool over time:
Net Carbon Balance = dCecosys dt−1
where Cecosys is the sum of C pool in vegetation, detritus, and soil. Ignoring, for the
moment, any export of organic C from the ecosystem, the net C balance is identical to the net ecosystem productivity (NEP):
dCecosys dt−1 = NEP = GPP − R
where GPP (gross primary production) is the rate of CO2 uptake through photosynthesis, and R is the total ecosystem respiration flux comprising autotrophic
(plant) respiration Ra and heterotrophic respiration R h (decomposition) of the
accumulated detritus and soil pools.
R = Ra + R h
The net accumulation of C in the ecosystem is, thus, a summation over time of
the difference between a large incoming CO2 flux (GPP) and a nearly equal outgoing flux (R). Fluxes are controlled by different processes whose rates change
over time and space and vary both with environmental conditions and the type
of ecosystem. These processes include those regulating the internal redistribution of organic C within the ecosystem (e.g., allocation of photosynthate within
the plants and breakdown of fresh litter into less decomposable forms of soil
organic matter) and disturbances (e.g., land use changes, soil degradation, erosion, windthrow, insect predation, harvest, or fire).
Agricultural Ecosystems
In agricultural soils, increasing C sequestration is an important strategy for
reducing the net emission of CO2 into the atmosphere while restoring degraded
soils, improving soil and water quality, and increasing agronomic productivity
and farm income. The amount of C stored in agricultural soils can be increased in
two ways: (i) by increasing the quantity of plant C added to the soil or (ii) by suppressing the rate at which soil organic matter decomposes. Soil C sequestration
can be achieved through adoption of conservation tillage based on crop residue
mulch and use of cover crops, increasing use efficiency of nitrogenous fertilizers and organic amendments, erosion management, and restoration of degraded
soils. These practices increase the rate of biomass C input and decrease losses
from erosion and mineralization (Janzen, 1998). Several important mechanisms
protect SOC and include physical, chemical, and biological processes (Laird et
al., 2001; Six et al., 2002). The combined effect of no-tillage on C sequestration
has been estimated by Liang et al. (2005), who reported that, under continuous
52
Bhatti & Tarnocai
cropping, the soil C pool increased by 0.22 Mg C ha−1 yr−1 for Brown, 0.32 Mg C
ha−1 yr−1 for Dark Brown, 0.39 Mg C ha−1 yr−1 for Black, and 0.30 Mg C ha−1 yr−1 for
Dark Gray Chernozemic soils with no-tillage.
Forest Ecosystems
Changes in the climate regime also affect current forest C pools, although the
direction and magnitude of these changes is still uncertain and difficult to predict because of the complex interactions between component species in different
forest ecosystems. Over a period of time, from years to decades, the stimulation
of GPP through longer growing seasons may result in increased forest vegetation biomass, an effect that may already be apparent in the global atmospheric
CO2 record (Keeling and Whorf, 2005). Although GPP may increase with increasing mean annual temperature, so may the heterotrophic decomposition rate,
which approximately doubles with every 10°C increase in soil temperature. The
very large size of the C pool in forest litter and soils is the cause of concerns
that increased heterotrophic respiration may generate a positive feedback mechanism by releasing additional quantities of CO2 to the atmosphere. However, in
some ecosystems, increased heterotrophic respiration may be largely offset by
increased detrital production from trees (Lavigne et al., 2003). Thus, the detrital
and soil C pool would be relatively unchanged as long as the forest composition
remains unchanged.
On the basis of published data for the boreal forests of the United States,
Canada, Finland, Sweden, and China, Gower et al. (1997) calculated an average
net primary production (NPPA) of 3.6 and 1.4 Mg C ha−1 yr−1 for deciduous and
coniferous boreal forests, respectively. These values are relatively consistent with
field measurements, as well as average belowground NPPA values of 1.1 and 1.0
Mg C ha−1 yr−1 calculated for deciduous and coniferous boreal forests, respectively
(Gower et al., 1997). Total aboveground and belowground net primary production (NPPT) has also been estimated from field measurements in the United States,
Canada, Finland, Sweden, and China to be 4.7 Mg C ha−1 yr−1 for deciduous, broadleaved species, and 2.7 Mg C ha−1 yr−1 for coniferous species (Gower et al., 1997).
These estimated NPPT values are comparable to the total NPP simulated using
the C budget model of the Canadian Forest Sector (CBM-CFS2). Simulated CBM
NPP values for western Canada vary between 1.5 and 3.9 Mg C ha−1 yr−1 depending on the ecozone, forest type, stand age, and site productivity (Li et al., 2003).
The majority of boreal forest sites investigated sequester C (measured as
NEP) at annual rates of up to 2.5 Mg C ha−1 yr−1 (Black et al., 1996; Jarvis et al.,
1997; Chen et al., 1999). The values depend primarily on latitude, soil type, forest
type, and successional stage. The NEP measurements made over periods of up to
5 yr in the northern Canada BOREAS experiment (Sellers et al., 1997) show that a
few old-growth coniferous stands may be C neutral (Goulden et al., 1998) and in
warm and cloudy years can be a C source with a loss rate of up to 1.0 Mg C ha−1
yr−1 (Lindroth et al., 1998).
Wetland–Peatland Ecosystems
Carbon accumulation in peatlands represents the balance between aboveand belowground net primary production and decomposition in both the upper,
aerobic (acrotelm) and the underlying, anaerobic (catotelm) peat layers. In general,
peatlands have much lower productivity than other natural ecosystems (Vitt, 2006),
Influence of Climate and Land Use Change on Carbon...
53
and peat accumulation is controlled by cool, wet conditions that limit decomposition (Moore et al., 1998). Regional rates of C accumulation in Canadian peatlands
range from 0.14 to 0.28 Mg C ha−1 yr−1 (Gorham, 1991; Mäkilä, 1997; Vitt et al., 2000)
and vary with changes in soil moisture, soil temperature, reduction–oxidation
conditions (Reader and Stewart, 1972), acidity and alkalinity (Thormann et al.,
1999), species composition (Johnson and Damman, 1991), and litter quality (Updegraff et al., 1995; Yavitt et al., 1997). Campbell et al. (2000b) compiled the NPP data
for fens and bogs of continental western Canada (Table 4–4) and reported that the
Table 4–4. Biomass and net primary productivity of different vegetation in continental
western Canadian wetlands. (Compiled from Campbell et al., 2000b.)
Trees
Shrubs
Herbs
Moss
Total
Biomass
———————————— g m−2 ————————————
N/A†
1511 ± 1767
1411
–‡
–
1471 ± 1458
N/A
253 ± 191
316
558
–
358 ± 299
N/A
45
78
128
365 ± 458
142 ± 118
N/A
N/A
N/A
N/A
N/A
N/A
N/A
1768 ± 1408
1860
372 ± 245
295
1198 ± 1562
5456 ± 3049
–
–
5456 ± 3049
351 ± 765
1023
–
800 ± 523
77
492 ± 683
N/A
650 ± 523
N/A
N/A
N/A
N/A
654 ± 197
2483 ± 916
843 ± 380
2291 ± 2330
Wetland type
Peatlands
Permafrost bogs
Bogs
Forested fen
Nonforested fens
Open fen
Fens and bogs
Non-peataccumulating
Treed swamps
No-treed swamps
Marsh
Swamps and
marshes
Net Primary Productivity
——————————— g m−2 yr−1 ———————————
Peatlands
Permafrost bogs
Bogs
77
106 ± 192
–
247 ± 104
–
13
44
–
–
88 ± 68
108
63
–
255 ± 296
34
125
365 ± 458
166 ± 298
542 ± 279
–
–
542 ± 279
31 ± 29
481 ± 260
–
255 ± 296
62
727 ± 667
999 ± 529
820 ± 592
Forested fen
Nonforested fens
Open fen
Fens and Bogs
Non-peataccumulating
Treed swamps
No-treed swamps
Marsh
Swamps and
marshes
† N/A, not available.
‡ –, not in this wetland type.
24
190 ±
157
81
118
163
139 ±
298
–
–
–
–
176
449 ± 215
358
263
268 ± 142
337 ± 142
654 ± 197
1232 ± 405
1034 ± 156
924 ± 463
54
Bhatti & Tarnocai
total a NPP is 5.1 Mg ha−1 yr−1, comprising 3.4 Mg ha−1 yr−1 for aboveground and 1.7
Mg ha−1 yr−1 for belowground components. In comparison, marshes and swamps
have higher total NPP (Table 4–4). Joosten and Clarke (2002) reported that pristine fens presently remove 2.5 Mg ha−1 yr−1 (as CO2) and release 2.97 Mg ha−1 yr−1
(as CH4), while bogs currently remove 3.1 Mg ha−1 yr−1 and release 0.53 Mg ha−1 yr−1.
While these trends indicate that pristine fens act as a C source and bogs as a C
sink, total sequestration of 5.6 Mg ha−1 yr−1 and release of 3.5 Mg ha−1 yr−1 indicates
that peatlands are a net C sink.
A little information is available about the C fluxes in both unfrozen and frozen
mineral soils in the permafrost regions. Trumbore and Harden (1997) reported a
C sequestration rate of 60 to 100 g C m−2 yr−1 for an unfrozen, upland, mineral soil
in northern Manitoba, due to slow rate of decomposition. Robinson et al. (2003)
observed that, in the sporadic discontinuous permafrost zone, mean C sequestration rates were 88.6 and 78.5 g C m−2 yr−1 for unfrozen bog and frost mound soils,
respectively. However, the rate of C sequestration in frozen peat plateau soils was
13.31 g C m−2 yr−1. Tarnocai et al. (2007) estimated that the peatlands in the permafrost region of Canada sequester 18 Tg C yr−1. In permafrost soils, cryogenic
processes are responsible for moving the organic matter from the surface layers
into the deeper soil layers, resulting in long-term storage of C in Cryosols.
Understanding the ability of terrestrial ecosystems to adapt to environmental change requires fundamental knowledge of numerous ecological processes.
The changes considered here include not only climatic change, but also land use
changes and alterations in disturbance patterns that may, or may not, be brought
about by the changing climate. Ecological responses to climate change are complex and nonlinear, with the variables involved being strongly interactive and
nonindependent. Climate change affects both the distribution and character of
the landscape through changes in temperature, precipitation, and natural disturbance patterns. These impacts are not entirely separable from the effects of
other anthropogenic changes such as land use change, soil degradation, erosion, and
drainage, all of which may be exacerbated by climate change. The following sections
deal with the impacts of climate change and other disturbances on ecosystems and
the effects that these agents of change may have on the ecosystem C storage.
Impacts of Climate Change
Under present conditions, climate has a major influence on the year-to-year
variations in agricultural productivity (Gitay et al., 2001; Lal, 2006). Increasing
atmospheric concentration of CO2 may enhance plant growth in some crops, but
the increase may be limited by the lack of water and essential elements (e.g., N, P,
S, and some micronutrients) (Oren et al., 2001). The positive impacts of warmer
temperature and enhanced CO2 on the rates of crop maturation and production are expected to mitigate the impact of moisture limitation, so that increased
growth rates in grasslands and pastures are generally expected (Campbell et al.,
2000a). However, agronomic and biomass productivity may decline under climate
change because of increases in intensity and frequency of drought, reduction in
nutrient use efficiency, and increase in the incidence of pests and diseases. With
a doubling of CO2 concentrations, an average increase of about 17% in grassland
productivity is anticipated, with relatively more increases in the northern than
southern regions (Campbell et al., 2002). However, some studies suggest that
Influence of Climate and Land Use Change on Carbon...
55
under projected climate change scenarios, particularly with extreme weather
events, the invasion of alien species into grasslands could reduce the nutritional
quality of the grass (White et al., 2001). Farm-level adaptation, including new crop
strains, can generally offset the detrimental effects of climate change, but poor
soil conditions may be a major factor limiting the northward expansion of agricultural crops (Campbell et al., 2005).
Northward expansion of agricultural land use, due to climate warming,
may depend primarily on the availability of suitable soils and climatic conditions. The most favorable soil conditions for agriculture occur in the Yukon
Territory and the Mackenzie River Valley of the Northwest Territories. Tarnocai et al. (1988) evaluated the effect of climate warming on agriculture in three
areas of the Yukon Territory (Watson Lake, Whitehorse, and Dawson City) and
observed that a significant moisture deficit would develop as a result of higher
rates of evapotranspiration, leading to the replacement of the present thermal
limitation by a moisture limitation of equal severity. Irrigation would remove this
limitation, resulting in much more favorable growing conditions in these areas,
especially in the vicinity of Dawson City, where the long daylight hours would
further improve the growing conditions. Brklacich and Tarnocai (1991) reported
that similar limitations of moisture availability may also apply to the Mackenzie
River Valley. They concluded that, in the Central and Upper Mackenzie River
Valley, there would be sufficient warming, but the moisture deficit would impose
considerable restrictions on agriculture unless irrigation is available. However, in
the Lower Mackenzie Valley (Mackenzie Delta), climate warming would not be
sufficient to support commercial agriculture.
The projected changes in climate are expected to have some adverse impacts
on Canadian soils. These impacts may be described in relation to their effects
on the components of ecosystem C stocks and are outlined in Fig. 4–1. Increases
in soil temperatures may enhance the rate of decomposition of soil organic matter. Consequently, the soil organic matter pool may decline with adverse impacts
on soil structure, plant-available water retention capacity, and nutrient cycling.
More specifically, and with all other factors remaining the same, a lowering
of soil organic matter content may have severe adverse impacts on soil quality,
leading to (i) increase in susceptibility to crusting, compaction, and erosion; (ii)
decrease in available water-holding capacity; (iii) reduction in activity and species diversity of soil fauna; (iv) decline in use efficiency of water and nutrients, (v)
reduction in agronomic productivity; and (vi) decrease in probability of achieving sustainable use of soil and water resources (Lal, 2006). However, the impact
of projected climate change on soil quality may differ among regions, with more
adverse impacts on soils of higher than lower latitudes.
Climate also affects the distribution, health, and productivity of the forest
and has a strong influence on the disturbance regime. Realization of potential
increases in plant productivity because of climate change depends on a range of
factors including species changes and competitive interactions, water and nutrient availability, and the effect of temperature increase on photosynthesis and
respiration (Bauer et al., 2006). Already notable changes in forest growth have
been attributed to climate-related drivers such as increased CO2 concentration,
higher temperatures, greater water stress, changing nutrient loadings, and permafrost thaw (Apps et al., 2006). The scientific knowledge on the effects of these
drivers is limited. However, Hogg et al. (2005) observed that increased drought
56
Bhatti & Tarnocai
stress may result in the boreal region, becoming a long-term source of atmospheric CO2. Other studies suggest that warmer and drier conditions projected
for midcontinental Canada may result in greater losses of boreal regions in the
south (due to grassland and agricultural encroachment) than gains in the north
(due to forest migration into present tundra), which would lead to a net reduction
in the C pool (Apps et al., 1993).
Projected climate change scenarios for the boreal forest generally predict
warmer and somewhat drier conditions, with the disturbance patterns also
expected to change (Amiro et al., 2001). In general, Canadian temperatures have
been increasing steadily over the last 50 yr, with winter temperatures being above
normal between 1985 and 2005. Concurrently, winter precipitation has shown
a general decline across Canada. The greatest warming has occurred in western Canada, with up to 6°C increase in the mean daily minimum temperature
(Environment Canada, 2006). In addition, the frequency of days with extreme
temperatures, both high and low, is expected to increase, snow and ice cover to
decrease, and heavy precipitation events to increase (Folland et al., 2001). Higher
temperatures may result in higher evaporation rates; hence, soils in some boreal
regions are expected to be drier during the summer (Amiro et al., 2001).
In addition to the direct influence of climate change, other variables, such as
anthropogenic and natural disturbances, also have profound influences on forest
distribution and productivity. The future forest C balance may largely depend
on the type and frequency of disturbances, changes in species composition, and
alterations in the nutrient and moisture regimes under changing climate conditions (Apps et al., 2006). In the shorter term (the next 100 yr), it is likely that
climate change may be accompanied by an increase in natural disturbances (fire,
insects, disease, windthrow) that can reduce the forest C pool by releasing large
quantities of CO2 to the atmosphere (Fig. 4–2). Since approximately 1970, Canada’s
forests have been subjected to increases in natural and anthropogenic disturbances (Kurz et al., 1998). Changes in the disturbance regime have resulted in
a change in Canada’s forest ecosystems; the net C sink (about 225 Tg C yr−1) for
the period 1920 to 1970 became a small net C source (about 75 Tg C yr−1) by 1989
(Kurz and Apps, 1999). Similarly, Bond-Lamberty et al. (2007) reported that the
C balance of Canadian boreal forest was driven by changes in fire disturbance
between 1948 and 2005. Recent analysis performed for Canadian managed forests
between 1990 and 2005 shows that the managed forest in Canada was an overall
sink except during 5 yr when it was a source mainly due to emission resulting
from extensive forest fires (Fig. 4–3) (Environment Canada, 2007). The mountain
pine beetle has killed trees over 10 million ha in central British Columbia between
1999 and 2005, resulting in significant increase (W.A. Kurz and Canadian Forest
Service–Carbon Accounting Team, Victoria, personal communications, 2007) in
emissions over the last few years (Fig. 4–2). Climate change affected the variability, but not the mean, of the landscape C balance, with precipitation exerting
a more significant effect than temperature. While regeneration of the disturbed
forest may sequester some of this C, it is not known if the increased C uptake of
the younger forest (the spatial structure and growth characteristics of which are
determined by the altered climatic and environmental conditions) may equal the
C losses of the forest it replaced (Apps et al., 1993).
Slow rates of decomposition in peatlands result in the rapid build up of
organic matter. However, with climate change and the predicted increase in fire
Influence of Climate and Land Use Change on Carbon...
57
Fig. 4–2. Canadian managed forest area disturbed by fire, insect, and harvest between
1990 and 2005. (Data compiled by Kurz and CFS-CAT team, personal communication.)
Fig. 4–3. Carbon source–sink for Canadian managed forest between 1990 and 2005.
58
Bhatti & Tarnocai
activities in peatlands, most of the surface C pool may be depleted (Turetsky et al.,
2007). The average fire return frequency for western Canadian forested peatlands
is between 400 and 1700 yr (Kuhry, 1994). Published rates of organic matter combustion (vegetation and soil) in peatlands average 3.2 ± 0.4 kg C m−2 per fire event
(Turetsky and Wieder, 2001; Turetsky et al., 2002). Using this average combustion
rate, Turetsky et al. (2004) estimated that up to 5.9 Tg C is released annually to the
atmosphere as a result of peatland burning in continental western Canada.
Using the peatland sensitivity model developed by Kettles and Tarnocai
(1999), Tarnocai (2006) determined the effect of climate warming on the various
peatland regions. He found that the greatest effect is likely to occur in the Subarctic Region, where approximately 78% of the organic C mass may be severely
or extremely affected by climate change. The second-largest effect is expected to
occur in the Boreal Region, where approximately 41% of the organic C mass may
be severely or extremely affected by climate change. Further, approximately 97%
of the SOC pool of Canadian peatlands is contained in these two regions.
Permafrost is an important component of many northern forest–peatland
ecosystems and has profound influences on hydrology, vegetation, and C storage.
Warming of temperatures has resulted in increased permafrost melt in northern
regions of Canada. For example, air temperatures have warmed significantly in
the Mackenzie Valley (ranging from 60 to 64 oN) over the past five decades (Robinson et al., 2001). The projected warming and associated changes in precipitation
influence both NPP and decomposition in peatlands. Recent thawing in southern portions of the discontinuous permafrost zone has also been noted at boreal
sites in Manitoba (Thie, 1974) and other prairie provinces (Beilman et al., 2001). In
boreal forests, permafrost affects soil moisture by inhibiting permeability and
providing a continuous source of water to the ecosystem as the permafrost surface thaws during the growing season. Soils with shallow permafrost may be
particularly sensitive to climate change because of the limiting role that high soil
moisture plays in these ecosystems. For example, shifts toward wetter or cooler
conditions may suppress fires and enhance ecosystem C storage. Shifts toward
drier summers may favor fire activity and enhance fire-induced emissions. While
these climatic trends apply to all ecosystems, forest ecosystems with permafrost
soils are at much greater risk. In years that favor deep thawing, fire can tap into
these vast quantities of C-rich fuels (Harden et al., 2000). Peatland fires result
in an initial ecosystem response of decreased NPP and elevated postfire decomposition rates, but little is known about the longer-term recovery of peatland C
balance after a fire event (Zoltai et al., 1998).
Land Use Change
Rapid expansion of agriculture along its southern border has been a recognized risk to the boreal forest for more than 50 yr (Davidson, 1998). The
conversion of native upland and lowland into agriculture and urban lands has
escalated, resulting in the contemporary patchwork of ecosystem types (Houghton, 2000). In the prairie provinces of Canada alone, it has been estimated that
there was a net deforestation of 12.5 million ha (Mha) between 1869 and 1992
(Ramankutty and Foley, 1999). Using the Canadian Land Inventory Database to
estimate changes between 1966 and 1994, Hobbs and Theobald (2001) estimated
that forests of the southern boreal plains of Saskatchewan declined from 1.8 Mha
Influence of Climate and Land Use Change on Carbon...
59
to 1.35 Mha, an overall conversion of 24% of the boreal transition zone to agriculture over the 28-yr period. Other studies have shown that forest land is being
converted into agriculture, industrial, and urban development at the rate of 1215
ha yr−1 along the southern boreal zone of Canada (Fitzsimmons, 2002). This rate
is approximately three times the world average: the loss of boreal forests and
wetlands is equal to, and in some regions greater than, that occurring in tropical
rainforests. These estimates suggest that with the current rate of conversion, all
the wetland and forested areas in the boreal transitional zone will be lost by 2050
unless purposeful action is taken to reverse the present trend.
Conversion from forest ecosystem to agriculture and urban land uses results
in losses of C through both the initial depletion, associated with the removal of
natural vegetation, and the subsequent losses from soil, through mineralization,
erosion, and leaching (Lal, 2003a). Most agricultural soils in North America have
lost 30 to 50% (30–40 Mg C ha−1) of the antecedent C pool following conversion
from natural to agricultural ecosystems (Lal, 2006). In addition to removal of natural vegetation cover, agricultural activities also deplete the soil C pool through
reduction of biomass inputs and cause changes in soil temperature and moisture
regimes, which further accelerate decomposition. Soil drainage, aimed at managing the water table, and soil cultivation, intended to control weeds and prepare
seed beds, also accelerate soil erosion and mineralization of the SOC pool.
The above discussion has focused on CO2, but similar conclusions can be drawn
for other greenhouse gases such as CH4 and N2O. For example, N2O emissions are
influenced by the rate, type, and timing of fertilizer applications. Changes in land
use also alter the uptake of CH4 by soils, and different agricultural practices differ
in their CH4 emission profiles (Moss et al., 2005). Increases in livestock populations
have also contributed to the increase in atmospheric CH4. Enteric fermentation, the
digestion process in ruminant animals such as cattle, sheep, and goats, adds an estimated 100 Pg of CH4 yr −1 to the atmosphere (Boadi et al., 2004).
Land Degradation
Land degradation can occur through either degradation of the vegetation
cover or the underlying soil but, ultimately, results in reduced C storage of both
ecosystem components. Degradation of soil occurs as a result of excessive and
inappropriate utilization, environmental changes, and/or careless management of
agricultural, pasture, or forest lands. Soil degradation may be physical, chemical,
or biological (Fig. 4–4) and can range in severity from vegetation cover reduction
to drastic soil erosion. These degradation processes adversely affect NPP, both
directly and indirectly, which, in turn, reduces the amount of biomass input into
the soil. Consequently, the C input to the soil system is lower than output from
the system, resulting in depletion of the pool which becomes an atmospheric
source. Biological (as opposed to physical or chemical) soil degradation is directly
related to depletion of the SOC pool, which also leads to a reduction in soil biodiversity. The process of soil degradation leads to a positive feedback because
of the interactions among different processes involved, so that, once triggered,
the degradation may accelerate over time in both magnitude and rate (Lal, 2006).
Consequently, most degraded soils are severely depleted of their SOC pool (Lal et
al., 2003a,b). The historic loss of the SOC pool in degraded soils and ecosystems,
however, has created soil C sink capacity. Thus, restoration of degraded soils and
60
Bhatti & Tarnocai
Fig. 4–4. Soil degradation effects on the soil carbon pool. SOM, soil organic matter.
(Modified from Lal, 2006.)
ecosystems provides an opportunity to sequester some of the atmospheric CO2
into the depleted SOC pool.
Land use change associated with a loss of vegetation typically has an initial rapid loss of C and nutrients that is generally ascribed directly to the land
use change itself. However, land use change may result in a long-term ecosystem degradation that generates additional depletion of the SOC pool. The loss of
belowground C is especially significant because this loss can be much more rapid
than the rate of its formation or replacement (Lal, 2006); these pools are usually
regarded as long-term storage. Since almost 40% of the world land area is designated as drylands, and 70% of this area is being subjected to land use change,
the magnitude of degradation processes associated with land use change is very
important (Lal, 2003b). Soil degradation is also an important factor in the C balance in the Tundra Region. Given the recent trends of increase in atmospheric CO2
concentration and the attendant climate warming, a number of questions become
relevant when considering these degraded lands. Is there a CO2 fertilization effect
on degraded lands? If antidesertification or land management measures are taken,
will the C pool attain the antecedent values, or will changed atmospheric and climate conditions result in higher or lower equilibrium ecosystem C pools?
Influence of Climate and Land Use Change on Carbon...
61
Soil degradation may also occur on forested lands due to various operations
and activities including harvesting, wildfire, fire control, pest and disease outbreaks, and conversion to nonforest uses, particularly agriculture and pasture
land. These disturbances often cause forests to become sources of CO2 (NEP <
0) because the rate of NPP is exceeded by total respiration or oxidation of plants,
soil, and dead organic matter (International Geosphere–Biosphere Programme,
1998). Intermediate between uplands and wetlands lie the riparian zones, which
are commonly degraded through cultivation or overgrazing. Degradation of the
sensitive riparian zones not only has the direct impact of reducing the amount of
vegetated habitat available to sequester C, but also has a negative impact on the
adjacent wetland through nutrient loading. Healthy riparian vegetation functions
to slow runoff from adjacent agricultural lands, trapping sediments, associated
pesticides, and fertilizers in the riparian zone, thus preventing nutrient loading
and contamination of adjacent wetlands or aquatic systems.
Soil Erosion
Among major land uses in Canada, soil erosion is most extensive and severe in
agricultural regions (Bhatti et al., 2006). Soil erosion reduces agricultural productivity and sustainability as well as having adverse effects on air and water quality
(Lal, 2006). Wind and water erosion may increase significantly in agricultural soils
due to increases in extreme weather conditions such as heavy precipitation and
prolonged droughts (Lal, 2003a). Warmer winters may decrease snow cover, and
the consequential reduced soil moisture content further increases the wind erosion
risks during spring. Water erosion hazard occurs in all provinces, wind erosion in
the prairie provinces (Alberta, Saskatchewan, and Manitoba), and tillage erosion
on most of the cropland under conventional management practices. Soil erosion
depletes the SOC pool through preferential removal of soil organic material comprising the light soil fraction in the surface horizon. Consequently, the SOC pool
of eroded soils is severely depleted, often by as much as 30 to 45 Mg C ha−1 (Lal,
2003b). The fate of C displaced by erosion is an obviously important, but highly
debated, topic. Lal (2003b) estimated that on the global scale, water erosion translocates about 4.0 to 6.0 Pg C yr−1. Of this, 2.8 to 4.2 Pg C yr−1 is redistributed over
the landscape and transferred to local depressional areas, 0.4 to 0.6 Pg C yr−1 is
transported into the ocean by world rivers, and 0.8 to 1.2 Pg C yr−1 is emitted to the
atmosphere as CO2. Thus, an adoption of effective conservation measures can drastically reduce the erosion-induced emission of soil C into the atmosphere.
A recent analysis of Canadian soil erosion risk assessments by Van Vliet et al.
(2003) indicates that changes in cropping systems and tillage practices have significantly reduced the risk of erosion between 1981 and 1996. The combination of
reduced tillage, less-intensive cropping, decreased summer fallow, and removal
of marginal land from production have resulted in lower erosion rates. Techniques used to reduce the rate of erosion also have a large potential to be used
in the sequestration of C, which may then offset a fraction of the anthropogenic
emissions. Suggested cropping practices that may restore some of the depleted
SOC pool in eroded agricultural soils include reduction in tillage, growing of
perennial forage cover, and application of organic amendments to the soil. An
almost 70% increase in the total C pool has been observed in the Canadian Prairies over a 5-yr period, with continuous legume and cereal production, reduced
62
Bhatti & Tarnocai
tillage, and nutrient additions via fertilization or composted manure applications
(Duchemin et al., 1995).
Drainage and Flooding
In the Canadian prairie and parkland regions, wetland drainage is still a
current practice, although the major peat deposits lie further north in the boreal
forest. Approximately 17% of farmers whose lands supported wetlands drained
one or more of these between 1990 and 1992 (Canadian Wetlands Conservation
Task Force, 1993). Drainage of wetlands has serious consequences because it fundamentally changes the habitat by lowering the water table and altering both the
processes of photosynthesis (C uptake) and decomposition (C release). In Canada,
agriculture alone has accounted for an estimated loss of 20 Mha of the presettlement wetlands (Lal, 2003b). The emission of CO2 increases when northern
peatlands are drained or degraded (Alm et al., 1999). It has been suggested that
draining an additional 5% of Canadian peatlands would be sufficient to offset the
putative existing peatland C sink of the country (Waddington et al., 2002). Draining of peatlands also significantly increases N2O emissions (Regina et al., 1998). In
western Canada, most drained peatlands are only marginally productive under
crop management (Environment Canada, 1986). The loss of SOC from such wetlands
when converted to agricultural usage may be as much as 50% (Schlesinger, 1997).
Conversely, the flooding of forest and wetland areas in the boreal zone for
hydroelectric reservoirs generates massive fluxes of dissolved organic C into the
water, accelerates peat decomposition, and increases CH4 and CO2 fluxes to the
atmosphere (Munn and Maarouf, 1997). For example, Kelly et al. (1997) experimentally flooded a boreal wetland in Ontario, causing the C dynamics of the site
to change from a sink of 6.6 g C m−2 yr−1 to a source of 130 g C m−2 yr−1. Turetsky et al.
(2002) estimated that 0.8 ± 0.2 Gt C yr−1 is released from peatlands occupying approximately 780 km2 within hydroelectric reservoirs across western boreal Canada.
Mitigation Options for Different Ecosystems
As discussed previously, mitigation strategies promoting the preservation
and maintenance of healthy terrestrial ecosystem functioning may be as valuable
as land management strategies that aim to enhance the net uptake, and decrease
the releases of CO2 from terrestrial ecosystems. Logically, the amount of soil C
stored in agriculture ecosystems can be increased in two ways: first, by increasing the input of biomass to the soil and, second, by reducing soil respiration rates.
Adoption of many practices increases the soil C pool in agricultural ecosystems.
These practices include reducing the tillage intensity, maximizing the return of
crop residues, using improved agronomic practices to enhance crop yield, using
forage crops in rotation, encouraging vegetation cover on cultivated land, discouraging soil degradation practices, and improving water management practices
(Janzen, 2005). However, this gain in the soil C pool can only continue for a short
period of time before the system reaches a new equilibrium state (West and Post,
2002). The management practices to enhance SOC sequestration in agricultural
soils include increasing the cycle time of C in plant materials and soil organic matter by reducing tillage, taking full advantage of the growing season to produce
more shoot and root biomass by including perennial forages in the crop rotation,
Influence of Climate and Land Use Change on Carbon...
63
increasing the use of fertilizer to enhance biomass production, and growing forage varieties selected for yield and root mass production (Anderson and Coleman,
1985). In addition, Canada has more than 7 Mha of surplus agricultural land, a
significant portion of which can be brought into production for establishing biomass crops. Such crops are typically fast-growing perennials and can produce
lignocellulosic biomass at two to four times the average C sequestration rate in
soils of Canada’s current food crops. Assuming 7 Mha producing lignocellulosic
biomass at a conservative 3 Mg C ha−1 yr−1, there would be an annual biomass C
stream of more than 20 Tg C (Layzell and Stephen, 2006).
Much of the focus on C sequestration in forest ecosystems has been on
enhancement of the aboveground biomass. Recently, a shift to more comprehensive ecosystem management appears to be taking place, together with renewed
interest in rehabilitating degraded lands, mitigating the effects of deforestation,
and enhancing numerous ecosystem services such as wildlife, water quality, elemental cycling, and so on. In the forest ecosystem, Binkley et al. (1998) outlined
12 strategies to increase the C pool in managed forests such as those of Canada.
These strategies involve actions such as reducing the loss to fire, insects, and
similar disturbances (a short-term benefit only), increasing forest productivity
(e.g., through silvicultural practices such as fertilizing), improving regeneration
immediately following disturbance or harvesting, and choosing the appropriate mix of species to plant (Table 4–5). By planting forests where none previously
Table 4–5. Classes of management activities, subjective assessment of short- (<10 yr) and
long-term (>10 yr) benefits. (Compiled from Binkley et al., 1998, and Apps et al., 2006.)
Activities
Increase forest area
Afforestation and reforestation
Establish and manage reserves
Multiple use (e.g., agroforestry,
shelterbelts)
Restoration of degraded lands
Urban forestry
Increase carbon density
Longer rotation length
Enhance tree productivity
Control stand density (thinning)
Enhance nutrient availability
Control water table
Selected species and genotypes
Protect from natural disturbance,
reduce risk
Reduced impact logging
Reduce regeneration delay
Manage on site logging residues
Vulnerable to climate
change/human
activity‡
Short-term†
(>25 yr)
Longerterm†
++
+
+
−−
–
+
++
+
−
−
Va
Vc
++
+
−?
++
+
+
+++
−
Va, Vc
Va, Vc
Vc
Va
Vc
Vc
Va, Vc
?
+
+
−
−
−−
−−
Vc, Va
Vc, Va
+
Va
+
Vc, Va
† Number of plus signs (minus signs) indicates expected magnitude of C benefit (decrement).
‡ Vc, expected to be vulnerable to climate change; Va, potentially vulnerable to changes in human activity.
64
Bhatti & Tarnocai
existed (afforestation) or where previous land management practices had led to
degraded cover (reforestation), the total C pool is generally increased. Nutrient
fertilization has long been used to enhance stand productivity, and it can result in
increased C pools in trees and soils (Nohrstedt, 2001). Since the success of nutrient fertilization depends on site conditions, it is, therefore, potentially susceptible
to rapid climate changes that may drastically alter these conditions. For example, on more fertile sites the effect of fertilization is reduced by limited supply of
nutrients (Saarsalmi and Malkonen, 2001). Planting fast-growing species, such
as hybrid poplar, can produce high rates of C accumulation over short periods
of time. However, for long-term C sequestration, planting species adapted to the
local climate may be more effective (Schroeder and Kort, 2001).
Since approximately two-thirds of the ecosystem C, especially in the boreal
and subarctic forests, is stored in the soil (Tarnocai, 1998) forest management
practices must be based on minimal soil disturbance. Soil disturbances, such as
by erosion initiated by logging and road construction, as well as the thawing of
ice-rich permafrost, create favorable conditions for decomposition of soil organic
matter and the attendant release of CO2 to the atmosphere (Tarnocai et al., 2007).
Peatlands, and their ability to sequester C, are very sensitive to natural
and anthropogenic disturbances, including wildfire, road construction, drainage, peat harvesting, and overburden removal from any form of open pit mining,
including oil-sands (Turetsky and St. Louis, 2006). Consequences of peatland disturbance may be direct, where the disturbance itself removes C from the peatland,
or indirect, through reduced photosynthesis or increased decomposition. Decomposition, through the process of microbial respiration, converts previously stored
C to CO2, CH4, and/or N2O, which are then released in to the atmosphere. In general, the peatland disturbances described above all cause current C sequestration
to be either reduced or eliminated (Turetsky and St. Louis, 2006). Protection of C
pools from intensifying and recurring disturbance events, solely as a mitigation
strategy, is likely neither efficient nor effective as a long-term measure (Vitt and
Wieder, 2006).
Conclusions
This chapter focuses on a key question: Will the present C source–sink
relationships of several different ecosystems, namely agriculture, forest, and wetland–peatland, be maintained in Canada? More specifically, will the presently
observed C pools contained within these ecosystems decrease over time, or can
they be maintained or possibly increased over the next 50 to 100 years?
To answer these questions, a reliable projection of the C budget for 50 to 100
yr into the future is needed. However, such predictions are difficult to make with
certainty, given the present state of knowledge. With current knowledge and data,
it is possible to predict the likely trends in C balances for these ecosystems (i.e.,
whether there will be increase or decrease in the relative size of different components of the terrestrial ecosystems) so that a general C budget trend can be projected.
Based on the above discussions, the following trends appear to be likely:
• Carbon loss from land use or land cover change from forest and wetland–
peatland systems to agriculture is likely to increase, given the sustained
increase in food and fiber demand over the next 50 yr.
Influence of Climate and Land Use Change on Carbon...
65
• The emission of CO2 from the soils (agricultural and forest) as climate warms
will become an increasingly important source through the 21st century.
• Risks of soil erosion and other degradation processes, with attendant emission of CO2 and other greenhouse gases, are likely to increase with increase
in population pressure and warming climate.
·· In some locales, the greenhouse gases emissions may be offset by increased
productivity due to improved growing conditions. Methane emissions from
wetlands and peatlands are expected to decrease in southern regions, but
these greenhouse gas emission reductions may to some extent be offset by
an increase in emission of CO2. Moreover, there may be an increase in CH4
emissions in northern regions because of increase in temperature, lengthening of growing seasons, and thawing of permafrost.
Since forests and agricultural ecosystems continue to provide both the goods
(e.g., food and fiber) and ecosystem services (e.g., recreation, spiritual, and social)
of interest to humans, there is an urgent need to assess the impact of anthropogenic activities on the C pool of these ecosystems. Management activities that
enhance or protect C pools in forest ecosystems include reducing the regeneration
delay through seeding and planting, enhancing forest productivity, changing the
harvest rotation length, using forest products judiciously, and protecting forests
through control and suppression of disturbance by fire, pests, and disease. At the
same time, the flow of material goods and services from a thriving forest products sector not only reduces the dependence on more energy-intensive products
(e.g., cement), but also provides economic benefits that can help pay for such forest-enhancing activities.
It is possible that the current C source–sink relationships for terrestrial ecosystems in Canada may remain approximately in balance, especially if the current
sinks can be increased. However, a more likely outcome for the foreseeable future
is that the biosphere as a whole may become a net C source. Such a drastic change
may have important implications on any strategies designed to stabilize the concentration of greenhouse gases in the atmosphere.
Acknowledgments
The authors thankfully acknowledge the thoughtful comments by two anonymous
reviewers. The inputs and comments from Ruth Errington on an earlier draft, which
improved the manuscript, are greatly appreciated.
References
Alm, J., L. Schulman, J. Silvola, J. Walden, H. Nykänen, and P.J. Martikainen. 1999. Carbon balance of a
boreal bog during a year with an exceptionally dry summer. Ecology 80:161–174.
Amiro, B.D., B.J. Stocks, M.E. Alexander, M.D. Flannigan, and B.M. Wotton. 2001. Fire, climate change,
carbon, and fuel management in the Canadian boreal forest. Int. J. Wildland Fire 10:405–413.
Anderson, D.W., and D.C. Coleman. 1985. The dynamics of organic matter in grassland soils. J. Soil Water
Conserv. 40:211–216.
Apps, M.J., P. Bernier, and J.S. Bhatti. 2006. Forests in the global carbon cycle: Implications of climate
change. p. 175–200. In J.S. Bhatti, R. Lal, M.J. Apps, and M.A. Price (ed.) Climate change and managed ecosystems. CRC Press, Taylor and Francis Group, Boca Raton, FL.
Apps, M.J., W.A. Kurz, S.J. Beukema, and J.S. Bhatti. 1999. Carbon budget of the Canadian forest product
sector. Environ. Sci. Policy 2:25–41.
Apps, M.J., W.A. Kurz, R.J. Luxmoor, L.O. Nilsson, R.A. Sedjo, R. Schmidt, L.G. Simson, and T.S. Vinson.
1993. Boreal forest and tundra. Water Air Soil Pollut. 70:39–53.
66
Bhatti & Tarnocai
Bauer, I.E., M.J. Apps, J.S. Bhatti, and R. Lal. 2006 Climate change and terrestrial ecosystem management: Knowledge gaps and research needs. p. 411–426. In J.S. Bhatti, R. Lal, M.J. Apps, and M.A.
Price (ed.) Climate change and managed ecosystems. CRC Press, Taylor and Francis Group, Boca
Raton, FL.
Beilman, D.W., D.H. Vitt, and L.A. Halsey. 2001. Localized permafrost peatlands in western Canada: Definitions, distributions, and degradation. Arct. Antarct. Alp. Res. 33:70–77.
Bhatti, J.S., M.J. Apps, and R. Lal. 2006. Anthropogenic changes and the global carbon cycle. p. 71–92.
In J.S. Bhatti, R. Lal, M.J. Apps, and M.A. Price (ed.) Climate change and managed ecosystems. CRC
Press, Taylor and Francis Group, Boca Raton, FL.
Binkley, C.S., M.J. Apps, R.K. Dixon, P. Kauppi, and L.-O. Nilsson. 1998. Sequestering carbon in natural
forests. Crit. Rev. Environ. Sci. Technol. 27:S23–S45.
Black, T.A., G. den Hartog, H.H. Neumann, P.D. Blanken, P.C. Yang, C. Russell, Z. Nesic, X. Lee, S.G. Chen,
R. Staebler, and M.D. Novak. 1996. Annual cycles of water vapour and carbon dioxide fluxes in and
above a boreal aspen forest. Global Change Biol. 2:219–230.
Boadi, D., C. Benchaar, J. Chiquette, and D. Massé. 2004. Mitigation strategies to reduce enteric methane emissions from dairy cows: Update review. Can. J. Anim. Sci. 84:319–332.
Bond-Lamberty, B., S.D. Peckham, D.E. Ahl, and S.T. Gower. 2007. The dominance of fire in determining
carbon balance of the central Canadian boreal forest. Nature 450:89–92.
Brklacich, M., and C. Tarnocai. 1991. Agricultural potential and climate change in the Mackenzie Valley.
p. 83–90. In Proceedings of the 4th Meeting on Northern Climate, Yellowknife, NWT, 17–18 Oct.
1990, G. Wall and L. Weber (ed.) LRRC Cont. No. 90–92. Canadian Climate Program, Environment
Canada, Toronto.
Campbell, B.D., D.M. Stafford Smith, and GCTE. 2000a. Pastures and Rangelands Network members. A
synthesis of recent global change research on pasture and rangeland production: Reduced uncertainties and their management implications. Agric. Ecosyst. Environ. 82:39–52.
Campbell, C.A., H.H. Janzen, K. Paustian, E.G. Gregorich, L. Sherrod, B.C. Liang, and R.P. Zentner. 2005.
Carbon storage in soils of North American Great Plains: Effect of cropping frequency. Agron. J.
97:349–363.
Campbell, C., D.H. Vitt, L.A. Halsey, I.D. Campbell, M.N. Thormann, and S. Bayley. 2000b. New primary production and standing biomass in northern continental wetlands. Northern Forestry Centre
Information Report NOR-X-369. Canadian Forest Service, Northern Forestry Centre, Edmonton, AL.
Campbell, C.A., R.P. Zentner, S. Gameda, B. Blomert, and D.D. Wall. 2002. Production of annual crops on
the Canadian prairies: Trends during 1976–1998. Can. J. Soil Sci. 82:45–57.
Canadian Wetlands Conservation Task Force. 1993. Wetlands: A celebration of life. Issue Paper No. 1931. North American Wetlands Conservation Council, Ottawa, ON.
Chen, W.J., T.A. Black, P.C. Yang, A.G. Barr, H.H. Neumann, Z. Nesic, M.D. Novak, J. Eley, and R. Cuenca.
1999. Effects of climate variability on the annual carbon sequestration by a boreal aspen forest.
Global Change Biol. 5:41–53.
Davidson, C. 1998. Issues in measuring landscape fragmentation. Wildl. Soc. Bull. 26:32.
Duchemin, E., M. Lucotte, R. Canuel, and A. Chamberland. 1995. Production of the greenhouse gases CH4
and CO2 by hydroelectric reservoirs in the boreal region. Global Biogeochem. Cycles 9:529–540.
Environment Canada. 1986. Wetlands in Canada: A valuable resource. Fact Sheet 86-4. Lands Directorate, Ottawa, ON.
Environment Canada. 2005. Terrestrial ecozones of Canada. Available at www.ec.gc.ca/soer-ree/English/vignettes/Terrestrial/terr.cfm (verified 21 July 2008).
Environment Canada. 2006. Canada’s greenhouse gas inventory 1990–2002. Submission to the UNFCCC
Secretariat, Ottawa, ON.
Environment Canada. 2007. National inventory report 1990–2005: Greenhouse gas sources and sinks in
Canada. Submission to the UNFCCC Secretariat, Ottawa, ON.
Fitzsimmons, M. 2002. Effects of deforestation and reforestation on landscape spatial structure in
boreal Saskatchewan, Canada. Can. J. For. Res. 32:843.
Folland, C.K., T.R. Karl, R. Christy, R.A. Clarke, G.V. Gruza, J. Jouzel, M.E. Mann, J. Oerlemans, M.J. Salinger, and S.W. Wang. 2001. Observed climate variability and change. p. 99–182. In J.T. Houghton, Y.
Ding, D.J. Griggs, M. Noguer, P.J. van der Linden, X. Dai, K. Maskell, and C.A. Johnson (ed.) Climate
change 2001: The scientific basis. Contribution of Working Group I to the 3rd assessment report
of the Intergovernmental Panel on Climate Change. Cambridge Univ. Press, New York.
Gitay, H., S. Brown, W.E. Easterling, B. Jallow, J. Antle, M.J. Apps, R. Beamish, T. Chapin, W. Cramer, J.
Franji, J. Laine, L. Erda, J.J. Magnuson, I. Noble, C. Price, T.D. Prowse, O. Sirotenko, T. Root, E.-D.
Schulze, B. Sohngen, and J.-F. Soussana. 2001. Ecosystems and their services. p. 235–342. In J.T.
Influence of Climate and Land Use Change on Carbon...
67
Houghton, Y. Ding, D.J. Griggs, M. Noguer, P.J. van der Linden, X. Dai, K. Maskell, and C.A. Johnson (ed.) Climate change 2001: The scientific basis. Contribution of Working Group I to the 3rd
assessment report of the Intergovernmental Panel on Climate Change. Cambridge Univ. Press,
New York.
Gorham, E. 1991. Northern peatlands: Role in the carbon cycle and probable responses to climatic
warming. Ecol. Applic. 1:182–195.
Goulden, M.L., S.C. Wofsy, J.W. Harden, S.E. Trumbore, P.M. Crill, S.T. Gower, T. Fries, B.C. Daube, S.M.
Fan, D.J. Sutton, A. Bazzaz, and J.W. Munger. 1998. Sensitivity of boreal forest carbon balance to
soil thaw. Science 279:214–217.
Gower, S.T., J.G. Vogel, J.M. Norman, C.J. Kucharik, S.J. Steele, and T.K. Stow. 1997. Carbon distribution
and aboveground net primary production in aspen, jack pine, and black spruce stands in Saskachewan and Manitoba, Canada. J. Geophys. Res. 102(D24):29029–29041.
Harden, J.W., S.E. Trumbore, B.J. Stocks, A. Hirsch, S.T. Gower, K.P. O’Neill, and E.S. Kasischke. 2000. The
role of fire in the boreal carbon budget. Global Change Biol. 6(Suppl. 1):174–184.
Hobbs, N.T., and D.M. Theobald. 2001. Effects of landscape change on wildlife habitat: Applying ecological principles and guidelines in the western United States. In V.H. Dale and R.A. Haeuber (ed.)
Applying ecological principles to land management. Springer Verlag, New York.
Hogg, E.H., J. Brandt, and B. Kochtubajda. 2005. Factors affecting interannual variation in growth of
western Canadian aspen forests during 1951–2000. Can. J. For. Res. 35:610–622.
Houghton, R.A. 2000. Interannual variability in the global carbon cycle. J. Geophys. Res.
105:20121–20126.
International Geosphere–Biosphere Programme (IGBP) (Terrestrial Carbon Working Group: W. Steffan,
I. Noble, P. Canadell, M.J. Apps, E.-D. Schulze, P.G. Jarvis, et al.) 1998. The terrestrial carbon cycle:
Implications for the Kyoto Protocol. Science 280:1393–1394.
Janzen, H.H. 1998. Management effects on soil C storage on the Canadian prairies. Soil Tillage Res.
47:181–195.
Janzen, H.H. 2005. Soil carbon: A measure of ecosystem response in a changing world? Can. J. Soil
Sci. 85:467–480.
Jarvis, P.G., J.M. Massheder, S.E. Hale, J.B. Moncrieff, M. Rayment, and S.L. Scott. 1997. Seasonal variation
of carbon dioxide, water vapour, and energy exchanges of a boreal black spruce forest. J. Geophys.
Res. 102(D24):28953–2896.
Johnson, L.C., and A.W.H. Damman. 1991. Species-controlled Sphagnum decay on a south Swedish
raised bog. Oikos 61:234–242.
Joosten, H., and D. Clarke. 2002. Wise use of mires and peatlands: Background principles including a
framework for decision-making. International Mire Conservation Group and International Peat
Society, Saarijärvi, Finland.
Kauppi, P.E., R.A. Sedjo, M.J. Apps, C.C. Cerri, T. Fujimori, H. Janzen, O.N. Krankina, W. Makundi, G. Marland, O. Masera, G.J. Nabuurs, W. Razali, and N.H. Ravindranath. 2001. Technical and economic
potential of options to enhance, maintain, and manage biological carbon reservoirs and geoengineering. p. 301–323. In Working Group III contribution to the 3rd assessment report of the
Intergovernmental Panel on Climate Change. Cambridge Univ. Press, Cambridge.
Keeling, C.D., and T.P. Whorf. 2005. Atmospheric CO2 records from sites in the SIO air sampling network.
In Trends: A compendium of data on global change. Carbon Dioxide Information Analysis Center,
Oak Ridge National Laboratory, U.S. Department of Energy, Oak Ridge, TN.
Kelly, C.A., J.W.M. Rudd, and R.A. Bodaly. 1997. Increases in fluxes of greenhouse gases and methyl mercury following flooding of an experimental reservoir. Environ. Sci. Technol. 31:1334–1345.
Kettles, I.M., and C. Tarnocai. 1999. Development of a model for estimating the sensitivity of Canadian
peatlands to climate warming. Geogr. Phys. Quat. 53:323–338.
Kuhry, P. 1994. The role of fire in the development of Sphagnum-dominated peatlands in western boreal
Canada. J. Ecol. 82:899–910.
Kurz, W.A., and M.J. Apps. 1999. A 70-year retrospective analysis of carbon fluxes in the Canadian forest sector. Ecol. Applic. 9:526–535.
Kurz, W.A., S.J. Beukema, and M.J. Apps. 1998. Carbon budget implications of the transition from natural to managed disturbance regimes in forest landscapes. Mitigation Adapt. Strategies Global
Change 2:405–421.
Laird, D.A., D.A. Martens, and W.L. Kingery. 2001. Nature of clay-humic complexes in an agricultural soil:
I. Chemical, Biochemical, and Spectroscopic Analyses. Soil Sci. Soc. Am. J. 65:1413–1418.
Lal, R. 2003a. Soil erosion and global carbon budget. Environ. Int. 29:437.
68
Bhatti & Tarnocai
Lal, R. 2003b. Off-setting global CO2 emissions by restoration of degraded soils and intensification of
world agriculture and forestry. Land Degrad. Devel. 14:309–345.
Lal, R. 2006. Carbon dynamics in agricultural soils. p. 127–148. In J.S. Bhatti, R. Lal, M.J. Apps, and M.A. Price
(ed.) Climate change and managed ecosystems. CRC Press, Taylor and Francis Group, Boca Raton, FL.
Lal, R. 2005. Soil erosion and carbon dynamics. Soil Tillage Res. 81:137–142.
Lal, R., T.M. Sobecki, T. Iivari, and J.M. Kimble. 2003. Soil degradation in the U.S. CRC/Lewis Publ., Boca
Raton, FL.
Lavigne, M.B., R. Boutin, R.J. Foster, G. Goodine, P.Y. Bernier, and G. Robitaille. 2003. Soil respiration
responses to temperature are controlled more by roots than by decomposition in balsam fir ecosystems. Can. J. For. Res. 33:1744–1753.
Layzell, D., and J. Stephen. 2006. Linking biomass energy to biosphere greenhouse gas management. p.
218–230. In J.S. Bhatti, R. Lal, M.J. Apps, and M.A. Price (ed.) Climate change and managed ecosystems. CRC Press, Taylor and Francis Group, Boca Raton, FL.
Li, Z., M.J. Apps, E. Banfield, and W.A. Kurz. 2003. Estimating net primary production of forests in the
Canadian Prairie Provinces using an inventory-based carbon budget model. Can. J. For. Res.
32:161–169.
Liang, B.C., C.A. Campbell, B.G. McConkey, G. Padbury, and P. Collas. 2005. An empirical model for estimating carbon sequestration on the Canadian prairies. Can. J. Soil Sci. 85:549–556.
Lindroth, A., A. Grelle, and A.S. Moren. 1998. Long-term measurements of boreal forest carbon balance
reveal large temperature sensitivity. Global Change Biol. 4:443–450.
Mäkilä, M. 1997. Holocene lateral expansion, peat growth, and carbon accumulation on Haukkasuo, a
raised bog in southeastern Finland. Boreas 26:1–14.
Moore, T.R., N.T. Roulet, and J.M. Waddingtot. 1998. Uncertainty in predicting the effect of climatic
change on the carbon cycling of Canadian peatlands. Clim. Change 40:229–245.
Moss, A.R., J.-P. Jouany, and J. Newbold. 2005. Methane production by ruminants: Its contribution to
global warming. Ann. Zootech. 49:231–240.
Munn, R.E., and A.R. Maarouf. 1997. Atmospheric issues in Canada. Sci. Total Environ. 203:1–14.
National Wetlands Working Group. 1988. Wetlands of Canada. Ecological Land Classification Series, No.
24. Sustainable Development Branch, Environment Canada, Ottawa, and Polyscience Publications,
Inc., Montreal.
Natural Resources Canada. 2004. The state of Canada’s forests 2003–2004. Natural Resources Canada,
Ottawa, ON.
Nohrstedt, H.O. 2001. Response of coniferous forest ecosystems on mineral soils to nutrient additions:
A review of Swedish experiences. Scand. J. For. Res. 16:555–573.
Oren, R., D.S. Ellsworth, K.H. Johnsen, N. Phillips, B.E. Ewers, C. Maier, K.V.R. Schafer, H. McCarthy, G.
Hendrey, and S.G. McNulty. 2001. Soil fertility limits carbon sequestration by forest ecosystems in
a CO2-enriched atmosphere. Nature 411:469–474.
Ramankutty, N., and J.A. Foley. 1999. Estimating historical changes in land cover: North American croplands from 1850 to 1992. Global Ecol. Biogeogr. 8:381–394.
Reader, R.J., and J.M. Stewart. 1972. The relationship between net primary production and accumulation for a peatland in southeastern Manitoba. Ecology 53:1024–1037.
Regina, K., H. Nykänen, J. Silvola, and P.J. Martikainen. 1998. Emissions of N2O and NO and net nitrogen
mineralization in a boreal forested peatland treated with different nitrogen compounds. Can. J.
For. Res. 28:132–145.
Robinson, S.D., J. Park, and I.M. Kettles. 2001. Northern peatlands to monitor recent and future warming. Final Report to the Climate Change Action Fund, Project A313. Natural Resources Canada,
Ottawa, ON.
Robinson, S.D., M.R. Turetsky, I.M. Kettles, and R.K. Wieder. 2003. Permafrost and peatland carbon
sink capacity with increasing latitude. p. 965–970. In M. Phillips, S.M. Springman, and L.U. Arenson (ed.) Proceedings of the 8th International Conference on Permafrost. A.A. Balkema, Lisse,
The Netherlands.
Saarsalmi, A., and E. Malkonen. 2001. Forest fertilization research in Finland: A literature review. Scand.
J. For. Res. 16:514–535.
Schlesinger, W.H. 1997. Biogeochemistry: An analysis of global change. 2nd ed. Academic Press, New York.
Schroeder, W., and J. Kort. 2001. Temperate agroforestry: Adaptive and mitigative roles in a changing
physical and socio-economic climate. Proceedings of the 7th Biennial Conference on Agroforestry in North America, 12–15 Aug. 2001, Regina, SK.
Influence of Climate and Land Use Change on Carbon...
69
Sellers, P.J., F.G. Hall, and R.D. Kelly. 1997. BOREAS in 1997: Experiment overview, scientific results, and
future directions. J. Geophys. Res. 102:28731–28770.
Shaw, C.H., J.S. Bhatti, and K. Sabourin. 2005. An ecosystem carbon database for Canadian forests.
Northern Forestry Centre Information Report NOR-X-403. Canadian Forest Service, Northern Forestry Centre, Edmonton, AL.
Six, J., R.T. Conant, E.A. Paul, and K. Paustian. 2002. Stabilization mechanisms of soil organic matter:
Implications for C-saturation of soils. Plant Soil 241:155–176.
Tarnocai, C. 1998. The amount of organic carbon in various soil orders and ecoprovinces in Canada. p.
81–92. In R. Lal, J. Kimble, R.F. Follett, and B.A. Stewart (ed.) Soil processes and the carbon cycle.
CRC Press, Boca Raton, FL.
Tarnocai, C. 2006. The effect of climate change on carbon in Canadian peatlands. Global Planet. Change
53:222–232.
Tarnocai, C., I.M. Kettles, and B. Lacelle. 2005. Peatlands of Canada database. Digital database. Research
Branch, Agriculture and Agri-Food Canada, Ottawa, ON.
Tarnocai, C., and B. Lacelle. 1996. Soil organic carbon of Canada database. Digital database. Eastern Cereal
and Oilseed Research Centre, Agriculture and Agri-Food Canada, Research Branch, Ottawa, ON.
Tarnocai, C., C.-L. Ping, and J. Kimble. 2007. Carbon cycles in the permafrost region of North America. p.
101–112. In A.W. King, L. Dilling, D.F. Fairman, R.A. Houghton, G. Marland, A.Z. Rose, T.J Wilbanks,
and G.P. Zimmerman (ed.) The first state of the carbon cycle report (SOCCR): The North American carbon budget and implications for the global carbon cycle. A report by the U.S. Climate
Change Science Program and the Subcommittee on Global Change Research. National Oceanic
and Atmospheric Administration, Climate Program Office, Silver Spring, MD.
Tarnocai, C., C.A.S. Smith, and D. Beckman. 1988. Agricultural potential and climate change in the Yukon.
p. 181–196. In Canadian climate program, proceedings of the 3rd Meeting on Northern Climate.
Atmospheric Environment Service, Department of Environment and Northern Affairs Program,
Department of Indian and Northern Affairs, Ottawa.
Thie, J. 1974. Distribution and thawing of permafrost in the southern part of the discontinuous permafrost zone in Manitoba. Arctic. 27:189–200.
Thormann, M.N., A.R. Szumigalski, and S.E. Bayley. 1999. Aboveground peat and carbon accumulation potentials along a bog-fen-marsh gradient in southern boreal Alberta, Canada. Wetlands
19:305–317.
Trumbore, S.E., and J. Harden. 1997. Accumulation and turnover of carbon in organic and mineral soils
of the BOREAS northern study area. J. Geophys. Res. 102:28817–28830.
Turetsky, M.R., B. Amiro, and J. S. Bhatti. 2004. Peatland burning and its relationship to fire weather in
western Canada. Global Biogeochem. Cycles 18:GB4014, doi:10.1029/2004 GB002222.
Turetsky, M.R., and V. St. Louis. 2006. Disturbance in boreal peatlands. In R.K. Wieder and D.H. Vitt (ed.)
Boreal peatland ecosystems. Ecological Study Series. Springer-Verlag, Heidelberg.
Turetsky, M.R., and R.K. Wieder. 2001. A direct approach to quantifying organic matter lost as a result
of peatland wildfire. Can. J. For. Res. 31:363–366.
Turetsky, M.R., R.K. Weider, L.A. Halsey, and D.H. Vitt. 2002. Current disturbance and the diminishing
peatland carbon sink. Geophys. Res. Lett. 29:1526.
Turetsky, M.R., R.K. Wieder, D.H. Vitt, R.J. Evans, and K.D. Scott. 2007. The disappearance of relict permafrost in boreal north America: Effects on peatland carbon storage and fluxes. Global Change
Biol. 13:1922–1934.
Updegraff, K., J. Pastor, S.D. Bridgham, and C.A. Johnston. 1995. Environmental and substrate controls
over carbon and nitrogen mineralization in northern wetlands. Ecol. Applic. 5:151–163.
Van Vliet, L.J.P., D.A. Lobb, and G.A. Padbury. 2003. Soil erosion risk indicators used in Canada. In Proceedings of OECD expert meeting on soil erosion and biodiversity indicators. Rome, Italy, March
2003. OECD, Paris.
Vitt, D.H. 2006 Peatlands: Canada’s past and future carbon legacy. p. 201–216. In J.S. Bhatti, R. Lal, M.J.
Apps, and M.A. Price (ed.) Climate change and managed ecosystems. CRC Press, Taylor and Francis Group, Boca Raton, FL.
Vitt, D.H., L.A. Halsey, and S.C. Zoltai. 2000. The changing landscape of Canada’s western boreal forest:
The current dynamics of permafrost. Can. J. For. Res. 30:283–287.
Vitt, D.H., and R.K. Wieder. 2006. Boreal peatland ecosystems: Our carbon heritage. p. 453–462. In
R.K. Wieder and D.H. Vitt (ed.) Boreal peatland ecosystems. Ecological Study Series. SpringerVerlag, Heidelberg.
Waddington, J.M., K.D. Warner, and G.W. Kennedy. 2002. Cut-over peatlands: A persistent source of
atmospheric CO2. Global Biogeochem. Cycles 16:10–26. doi:1029/2001GB001398.
70
Bhatti & Tarnocai
West, T.O., and W.M. Post. 2002. Soil carbon sequestration rates by tillage and crop rotation: A global
data analysis. Soil Sci. Soc. Am. J. 66:1930–1946.
White, T.A., B.D. Campbell, P.D. Kemp, and C.L. Hunt. 2001. Impacts of extreme climatic events on competition during grassland invasion. Global Change Biol. 7:1–14.
Yavitt, J.B., C.J. Williams, and R.K. Wieder. 1997. Production of methane and carbon dioxide in peatland
ecosystems across North America: Effects of temperature, aeration, and organic chemistry of
peat. Geomicrobiol. J. 14:299–316.
Zoltai, S.C., L.A. Morrissey, G.P. Livingston, and W.J. de Groot. 1998. Effects of fire on carbon cycling in
North American boreal peatlands. Environ. Rev. 6:13–24.