Introduction
Peatlands are widely distributed across the Northern Hemisphere covering
5–30 % of national land areas in northern Europe, North America and Russia
and play a key role in the global carbon (C) cycle (Gorham, 1991; Joosten
and Clarke, 2002; Vasander et al., 2003; Charman et al., 2013). Throughout
the Holocene, northern peatlands have accumulated ∼ 270–450 Gt C
as peat and presently store about a third of the global soil C pool
(Gorham, 1991; Turunen et al., 2002). They also provide a small but
persistent long-term C sink (between 20 and 30 g C m-2 yr-1)
(Gorham, 1991; Vitt et al., 2000; Roulet et al., 2007; Nilsson et al.,
2008). Carbon accumulation in peatland ecosystems occurs mainly due to the
slow decomposition rate under the anoxic conditions caused by high water
table levels (WTLs) (Clymo, 1983). Within the past century, however, a large fraction of
peatlands has been exploited for energy production and horticultural use.
Since commercial peat extraction requires initial vegetation removal and
drainage, harvested peatlands are turned into C sources by eliminating the
carbon dioxide (CO2) uptake during plant photosynthesis and increasing
CO2 emission due to enhanced aerobic decomposition of organic matter.
Thus, following the cessation of peat extraction activities, after-use
alternatives that mitigate the negative climate impacts of these degraded
and abandoned areas are required.
Among different after-use alternatives, re-establishment of peatland
vegetation, which is essential for returning the extracted peatlands back
into functional peat-accumulating ecosystems, has been shown to provide
climate benefits (Tuittila et al., 1999, 2000a; Graf and Rochefort, 2009;
Waddington et al., 2010; Strack and Zuback, 2013) as well as high ecological
value (Rochefort and Lode, 2006; Lamers et al., 2015). However, due to the
harsh environmental conditions of bare peat surfaces and the lack of a
propagule bank, spontaneous regeneration of self-sustaining ecosystems
rarely occurs and thus human intervention is necessary to initiate this
process. For instance, active re-introduction of natural peatland vegetation
communities (i.e., primarily fragments of Sphagnum mosses and companion species)
combined with rewetting has been shown to be an effective method to initiate
the recovery of Sphagnum-dominated ecosystems with resumed long-term peat
accumulation (Quinty and Rochefort, 2003).
Re-establishing peatland vegetation and raising the WTL both affect the ecosystem C balance and peat accumulation through their
impact on the production and decomposition of organic matter. Specifically,
vegetation development results in increased plant photosynthesis and
respiration (i.e., autotrophic respiration) as well as in greater substrate
supply for methanogenesis. In addition, restoring the hydrological regime
affects the CO2 uptake by vegetation and the microbial decomposition of
organic matter (i.e., heterotrophic respiration) by increasing water
availability and decreasing soil oxygen status of the upper peat layer.
Moreover, an increase in the WTL also reduces the depth of the aerobic peat
layer in which methane (CH4) oxidation may occur. As a consequence, a higher WTL following the filling or blocking of the drainage ditches commonly
results in decreased CO2 emissions (Tuittila et al., 1999; Waddington
and Warner, 2001) and increased emissions of CH4 (Tuittila et
al., 2000a; Waddington and Day, 2007; Vanselow-Algan et al., 2015) relative
to the abandoned bare peat area. The depth of the WTL is therefore, in
addition to the vegetation biomass recovery, a key controlling variable of
the ecosystem CO2 and CH4 exchanges following peatland
restoration.
Considering the strong effects of the WTL on plant succession and ecosystem
C exchanges, differences in the depth of the re-established WTL baseline
(i.e., the mean WTL) due to the varying effectiveness of initial restoration
activities (e.g., ditch blocking, surface peat stripping) may have
implications for the trajectories of vegetation development and recovery of
the C sink function following restoration. To date, only a few studies (e.g.,
Tuittila et al., 1999, 2004) have investigated the impact of contrasting
WTLs on the subsequent ecosystem C balance within the same restoration site.
Understanding the sensitivity of the C balance to differences in the
re-established WTL baseline is, however, imperative when evaluating the
potential of restoration for mitigating the negative climate impacts of
drained peatlands. Moreover, estimates of the C sink–source strength of
restored and unrestored peatlands have been limited to the growing season
period in most previous studies (Tuittila et al., 1999, 2000a, 2004;
Waddington et al., 2010; Samaritani et al., 2011; Strack et al., 2014). In
contrast, data on annual budgets, which are required to evaluate the full
climate benefits of peatland restoration relative to the abandoned peat
extraction area, are currently scarce and to our knowledge only reported in
a few studies (e.g., Yli-Petäys et al., 2007; Strack and Zuback, 2013).
Furthermore, the full ecosystem greenhouse gas (GHG) balance also includes
emissions of nitrous oxide (N2O), a greenhouse gas with an almost 300
times stronger warming effect relative to CO2 (IPCC, 2013). Highly
variable N2O emissions ranging from < 0.06 to 26 kg N ha-1 yr-1
have been previously reported for drained organic soils, with the
highest emissions occurring from mesic and nutrient-rich sites (Martikainen
et al., 1993; Regina et al., 1996; Maljanen et al., 2010). In contrast,
N2O emissions are generally low in natural peatlands because
environmental conditions (i.e., uptake of mineral N by the vegetation and
anaerobic conditions due to high WTL favoring the complete reduction of
N2O to dinitrogen) diminish the potential for N2O production
(Martikainen et al., 1993; Regina et al., 1996; Silvan et al., 2005;
Roobroeck et al., 2010). Thus, while the focus of most previous studies in
restored peatlands has been limited to the CO2 and CH4 exchanges,
accounting for N2O emissions might be imperative when assessing the
climate benefits of peatland restoration as an after-use option for
abandoned peat extraction areas. To our knowledge, however, N2O fluxes
in restored peatlands have not been quantified to date.
This study investigated the GHG fluxes (i.e., CO2, CH4 and
N2O) and their biotic and abiotic controls in a restored peat
extraction area with high (ResH) and low (ResL) WTLs and in an unrestored
bare peat (BP) site. The two main objectives were (i) to investigate the
impact of contrasting mean WTLs on the annual C and GHG balances of a restored
peatland and (ii) to assess the potential of peatland restoration for
mitigating the C and GHG emissions from abandoned peat extraction areas. Our
hypotheses were that
(i) the C and GHG balances are improved in ResH
relative to ResL since the increased net CO2 uptake, as a result of
reduced peat mineralization and greater water availability enhancing gross
primary production (GPP), outweighs the increase in CH4 emissions; and (ii) the C and GHG balances of the two restoration
treatments are ameliorated relative to BP due to the decreased CO2
emissions from peat mineralization and lower N2O emissions under more
anoxic conditions following rewetting of drained peatlands.
Material and methods
Experimental area
The study was conducted in the Tässi peat extraction area located in
central Estonia (58∘32′ 16′′ N, 25∘51′43′′ E). The
region has a temperate climate with long-term mean (1981–2010) annual
temperature and precipitation of 5.8 ∘C and 764 mm, respectively
(Estonian Weather Service). Peat extraction in the peatland started in
late 1960s and today peat continues to be harvested for horticultural
purposes using the milling technique on about 264 ha.
The current study was carried out on a 4.5 ha area which was set aside from
peat extraction in the early 1980s. The residual Sphagnum peat layer depth is about
2.5 m. A section approximately 0.24 ha in size within the abandoned site was
restored in April 2012. The restoration was done following a slightly
modified protocol of the moss layer transfer technique (Quinty and
Rochefort, 2003) aimed at restoring the growth of Sphagnum mosses and initiating the
development of a natural bog community. The first restoration steps included
stripping the uppermost oxidized peat layer (20 cm) and flattening the
freshly exposed surface. In addition, the peat along the borders of the
restoration area was compressed and the outflow drainage ditch was dammed
with peat material to reduce the lateral water outflow from the experimental
site.
To study the impact of WTL on restoration success in terms of
vegetation development and GHG fluxes, the restoration site was
divided into wetter and drier sections by lowering the peat surface by 10 cm
for approximately one-third of the area. This resulted in restoration
treatments with high and low WTLs (i.e., ResH and ResL). In
addition, an unrestored BP site was included in the study as a
reference. Two replicate plots (20 × 20 m) were established for each of the
three treatments.
To enhance vegetation succession, living plant fragments from
Sphagnum-dominated hummocks were collected from a nearby (10 km) donor site
(Soosaare bog) and spread out in the ratio of 1 : 10 (i.e., 1 m2 of
collected plant fragments were spread over 10 m2) in the ResH and ResL
treatments. As the last step, straw mulch was applied to protect plant
fragments from solar radiation and to improve moisture conditions. Further
details about the restoration procedure at this study site have been given
in Karofeld et al. (2015).
Three years following restoration, the bryophyte species found at the
restored site were dominated primarily by Sphagnum mosses (e.g., S. fuscum,
S. rubellum and S. magellanicum). The common
vascular plant species observed post-restoration included shrubs and trees
such as common heather (Calluna vulgaris L.), common cranberry
(Oxycoccus palustris Pers.), downy birch
(Betula pubescens Ehrh.), bog rosemary (Andromeda polifolia
L.) and
Scots pine (Pinus sylvestris L.), with a minor cover of
accompanying herbaceous sedge and forb species such as tussock cottongrass
(Eriophorum vaginatum L.) and round-leaved sundew (Drosera rotundifolia L.) (Karofeld et al., 2015).
Environmental measurements
A meteorological station to continuously monitor environmental variables was
set up on-site in June 2014. This included measurements of air temperature
(Ta; model CS 107, Campbell Scientific Inc., Logan, UT, USA),
photosynthetically active radiation (PAR; model LI-190SL, LI-COR Inc.,
Lincoln, NE, USA) and precipitation (PPT; tipping bucket model 52202, R. M.
Young Company, Traverse City, MI, USA) at 1.2 m height above the ground.
Soil temperature (Ts; depths of 5 and 30 cm) was measured with temperature
probes (model CS 107, Campbell Scientific Inc., Logan, UT, USA) and
soil volumetric water content (VWC; depth 5 cm) with water content
reflectometers (model CS615, Campbell Scientific Inc., Logan, UT, USA). All
automated abiotic data were collected in 1 min intervals and stored as
10 min averages on a data logger (CR1000, Campbell Scientific Inc., Logan, UT,
USA). In addition, continuous 30 min records of the WTL relative to the soil
surface were obtained with submerged water level loggers (HOBO, Onset
Computer Corporation, Bourne, MA, USA) placed inside perforated 1.0 m long
PVC pipes (∅ 5 cm; sealed in the lower end).
The on-site meteorological measurements were complemented by Estonian
Weather Service data to obtain complete time series of Ta, PPT and PAR over
the entire year. Hourly means of Ta and daily sums of PPT were obtained from
the closest (∼ 20 km away) Viljandi meteorological station.
Global radiation (hourly sums) data from the Tartu meteorological station
(∼ 40 km away) were converted to PAR based on a linear
correlation relationship to on-site PAR.
In addition, manual measurements of Ts (depths 10, 20, 30 and
40 cm) were recorded by a handheld temperature logger (Comet Systems Ltd.,
Rožnov pod Radhoštěm, Czech Republic) and VWC (depth 0–5 cm) using a handheld soil moisture sensor (model GS3,
Decagon Devices Inc., Pullman, WA, USA) during each sampling campaign.
Furthermore, groundwater temperature, pH, redox potential, dissolved oxygen
content, electrical conductivity as well as ammonium and
nitrate concentrations were measured in observation wells
(∅ 7.5 cm, 1.0 m long PVC pipes perforated and sealed in the lower end)
installed at each sampling location using YSI Professional Plus handheld
instruments (YSI Inc., Yellow Springs, OH, USA). In addition, soil samples
(depth 0–10 cm) in three replicates were taken from each of the treatments
and analyzed for pH as well as total C, total N, P, K, Ca and S contents at
the Tartu Laboratory of the Estonian Environmental Research Centre. Three
additional samples were taken from the same depth to determine bulk density
in each treatment. Mean values for these soil properties are summarized in
Table 1.
Soil properties in restoration treatments with high (ResH) and low
(ResL) water table level and bare peat (BP); numbers in parenthesis
indicate standard error.
Soil property
ResH
ResL
BP
pH
4.0 (0.07)
3.9 (0.07)
3.9 (0.06)
Bulk density (g cm-3)
0.08 (0.002)
0.09 (0.003)
0.13 (0.004)
C (%)
49 (0.6)
50 (0.3)
48 (0.6)
N (%)
0.61 (0.04)
0.76 (0.05)
0.85 (0.04)
C / N
80.3
65.8
56.5
P (mg g-1)
0.2 (0.03)
0.2 (0.02)
0.4 (0.03)
K (mg g-1)
0.2 (0.007)
0.2 (0.003)
0.1 (0.004)
Ca (mg g-1)
2.1 (0.07)
2.1 (0.07)
3.4 (0.23)
S (mg g-1)
0.9 (0.12)
1.0 (0.05)
1.4 (0.09)
Vegetation cover estimation
To assess the effect of vegetation development on GHG fluxes,
vegetation cover (%) and species composition were recorded inside each of
the flux measurement collars (see Sect. 2.4) in late spring. In each
collar, the cover was estimated visually for each species and rounded to the
nearest 1 %. Bryophyte, vascular plant and total vegetation cover were
computed as the sum of their respective individual species coverages.
Net ecosystem CO2 exchange (NEE), ecosystem respiration (Re), GPP and net
primary production (NPP) measurements
To evaluate the impact of WTL on NEE
in the restored ResH and ResL treatments, CO2 flux measurements were conducted
biweekly from May to December 2014 at three sampling locations within each
replicate plot (i.e., six locations per treatment) using the closed dynamic
chamber method. At each sampling location, a collar (∅ 50 cm) with a
water-filled ring for air-tight sealing was permanently installed to a soil
depth of 10 cm. NEE measurements were conducted in random plot order (to
avoid diurnal effects) using a clear Plexiglas chamber (95 % transparency;
h 50 cm, V 65 L) combined with a portable infrared gas analyzer (IRGA;
EGM-4, PP Systems, Hitchin, UK). The chamber was equipped with a sensor to
measure PAR and Ta (TRP-2, PP
Systems, Hitchin, UK) inside the chamber. Ambient Ta was also
recorded with an additional temperature sensor placed on the outside of the
chamber. Cooling packs placed inside the chamber were used to avoid a
temperature increase inside the chamber during measurements. The chamber was
also equipped with a low-speed fan to ensure constant air circulation. After
every NEE measurement, Re was determined from a
subsequent measurement during which the transparent chamber was covered with
an opaque and light reflective shroud. CO2 concentrations, PAR,
Ta, pressure and relative humidity were recorded by the IRGA system
every 4.8 s over a 4 or 3 min chamber deployment period for NEE and Re
measurements, respectively. Since the aim of this study was to assess the
atmospheric impact of restoration, all fluxes are expressed following the
atmospheric sign convention in which positive and negative fluxes represent
emission to and uptake from the atmosphere, respectively.
GPP was derived from the difference between NEE
and Re (i.e., GPP = NEE -Re). In addition, an estimate of NPP was derived from the difference between NEE and
heterotrophic respiration (Rh; see Sect. 2.5) (i.e., NPP = NEE - Rh).
Re estimates during the non-growing-season months of March to April 2014 and
January to February 2015 were determined by closed static chamber
measurements (described in Sect. 2.6). Air samples collected during these
measurements were analyzed for their CO2 concentrations on a Shimadzu
GC-2014 gas chromatograph with an electron capture detector. These Re
estimates also represented non-growing-season NEE for all treatments.
In the BP treatment, Re was determined by measurements using a separate
closed dynamic chamber setup as described below in Sect. 2.5. Due to the
absence of vegetation, GPP as well as NPP were assumed to be 0 and NEE
subsequently equaled Re in the BP treatment.
Heterotrophic and autotrophic respiration measurements
From May to December 2014, Rh was measured
simultaneously with NEE from separate PVC collars (∅ 17.5 cm) inserted to
a depth of 10 cm beside each NEE collar. The soil around the Rh collars was
cut with a sharp knife to a depth of 30 cm in April 2014 to exclude
respiration from the roots. The area inside the collars was cleared of
living moss and vascular plants and kept free of vegetation during the
remaining year. For Rh measurements, a second set of instrumentation was
used which included an opaque chamber (h 30 cm, V 0.065 L; equipped with a
low-speed fan) combined with an EGM-4 infrared gas analyzer. During each Rh
measurement, CO2 concentration and Ta inside the chamber
were recorded every 4.8 s over a period of 3 min. Autotrophic respiration
(Ra) was derived from the difference between the measured Re and Rh fluxes
(i.e., Ra = Re - Rh). Due to the absence of vegetation, Ra was not
determined in BP.
Methane and nitrous oxide flux measurements
To assess the impact of WTL on CH4 and N2O exchanges, flux
measurements were conducted with the closed static chamber method at a
biweekly to monthly interval from March 2014 to February 2015 at the same
locations (i.e., same collars) as were used for the NEE measurements
(described in Sect. 2.4). During each chamber deployment period, a series
of air samples were drawn from the chamber headspace (h 50 cm, V 65 L; white
opaque PVC chambers) into pre-evacuated (0.3 mbar) 50 mL glass bottles 0,
0.33, 0.66 and 1 h after closing the chamber. The air samples were analyzed
for CH4 and N2O concentrations with a flame ionization detector
and an electron capture detector, respectively, using a Shimadzu
GC-2014 gas chromatograph combined with a Loftfield automatic sample
injection system (Loftfield et al., 1997).
Flux calculation
Fluxes of CO2, CH4 and N2O were calculated from the linear
change in gas concentration in the chamber headspace over time, adjusted by
the ground area enclosed by the collar, volume of chamber headspace, air
density and molar mass of gas at measured chamber Ta. The
linear slope in case of the dynamic chamber measurements was calculated for
a window of 25 measurement points (i.e., 2 min) moving stepwise (with
one-point increments) over the entire measurement period after discarding
the first two measurement points (i.e., applying a 9.6 s “dead band”). The
slope of the window with the best coefficient of determination (R2) was
selected as the final slope for each measurement. In the static chamber
method, the linear slope was calculated over the four available
concentration values.
All dynamic chamber CO2 fluxes with a R2≥ 0.90 (p < 0.001)
were accepted as good fluxes. However, since small fluxes generally
result in a lower R2 (which is especially critical for NEE
measurements), dynamic chamber fluxes with an absolute slope within ±0.03 ppm s-1
were always accepted. The slope threshold was determined
based on a regression relationship between the slope and respective R2
values. For static chamber measurements, the R2 threshold for accepting
CO2, CH4 and N2O fluxes was 0.90 (p < 0.05), 0.80
(p < 0.1) and 0.80 (p < 0.1), respectively, except when the
maximum difference among the four concentration values was less than the
gas-specific GC detection limit (i.e., < 20 ppm for CO2,
< 20 ppb for CH4 and < 20 ppb for N2O), in which
case no filtering criterion was used. Based on these quality criteria 11 %
of NEE, 9 % of Re, 21 % of Rh, 33% of CH4 and 6 % of N2O
fluxes were discarded from subsequent data analysis.
Annual balances
To obtain estimates for the annual CO2 fluxes, nonlinear regression
models were developed based on the measured CO2 flux, PAR, WTL and Ta
data following Tuittila et al. (2004). As a first step, measured GPP fluxes
were fitted to PAR inside the chamber using a hyperbolic function adjusted
by a second term which accounted for additional WTL effects (Eq. 1):
GPP=α×Amax×PARα×PAR+Amax×exp-0.5×WTL-WTLoptWTLtol2,
where GPP is the gross primary production (mg C m-2 h-1), PAR is the
photosynthetically active radiation (µmol m-2 s-1), α is the light use efficiency of photosynthesis (i.e., the initial slope of
the light response curve; mg C µmol photon-1), Amax
is the
maximum photosynthesis at light saturation (mg C m-2 h-1), WTL is
the water table level (cm), WTLopt is the WTL at which maximum
photosynthetic activity occurs and WTLtol is the tolerance (i.e., the
width of the Gaussian response curve of GPP to WTL).
Secondly, Re fluxes were fitted to Ta using an exponential function (Eq. 2):
Re=R0×exp(b×Ta),
where Re is the ecosystem respiration (mg C m-2 h-1), Ta is the air
temperature (∘C), R0 is the respiration (mg C m-2 h-1)
at 0 ∘C and b is the sensitivity of respiration to Ta.
Both GPP and Re were modeled with hourly resolution using hourly PAR, WTL
and Ta as input variables. Growing season (1 May to October 31) GPP and
annual Re were then derived from the cumulative sums of these modeled
fluxes. The balance between growing season GPP and annual Re estimates
resulted in the annual NEE in ResH and ResL, whereas annual Re represented
annual NEE in BP. The GPP and Re model parameters for the different
treatments are summarized in Table 2.
Parameters for the gross primary production (GPP) and ecosystem
respiration (Re) models in restoration treatments with high (ResH) and low
(ResL) water table level (WTL) and bare peat (BP): α is the quantum use
efficiency of photosynthesis (mg C µmol photon-1); Amax is
the maximum rate of photosynthesis at light saturation (mg C m-2 h-1);
WTLopt is the WTL at which maximum photosynthetic activity
occurs; WTLtol is the tolerance, i.e., the width of the Gaussian response
curve of GPP to WTL; R0 is the respiration (mg C m-2 h-1)
at 0 ∘C; b is the sensitivity of respiration to air temperature;
numbers in parenthesis indicate standard error; Adj. R2 is the adjusted
R2.
Model parameter
ResH
ResL
BP
GPP model
α
-0.20 (0.07)
-0.23 (0.07)
n/a
Amax
-98.0 (39.9)
-121.9 (43.4)
n/a
WTLopt
-18.7 (8.4)
-24.9 (6.4)
n/a
WTLtol
16.4 (10.0)
21.0 (9.7)
n/a
Adj. R2
0.58
0.61
n/a
Re model
R0
13.0 (1.5)
13.4 (1.5)
18.6 (2.7)
b
0.056 (0.005)
0.064 (0.005)
0.055 (0.005)
Adj. R2
0.62
0.71
0.60
n/a is not applicable
Annual sums of CH4 and N2O fluxes were estimated by scaling their
hourly mean and median flux values, respectively, to annual sums. The median
flux was used for N2O to avoid a positive bias caused by episodic high
peak fluxes measured directly after rainfall events. The annual sums were
converted to CO2 equivalents (CO2 eq) using the global warming
potentials (over a 100-year time frame including carbon–climate
feedbacks) of 34 and 298 for CH4 and N2O, respectively (IPCC,
2013).
Statistical analysis
Collar flux data were averaged for each plot before conducting further
statistical analysis to avoid pseudoreplication. The non-parametric Friedman
one-way analysis of variance (ANOVA) by ranks test for dependent samples was
used to account for repeated measurements in time when testing for treatment
effects (i.e., ResH, ResL and BP) on the growing season or annual means of
the various component fluxes. This analysis was followed by a Bonferroni
post hoc comparison to determine significant differences among treatment
means. The Mann–Whitney U test was used when comparing only the restoration
treatments for significant effects (i.e., on GPP, NPP and Ra fluxes).
Pearson's correlations were used to investigate the effects of vegetation
cover on mean growing season fluxes. The significance level was P < 0.05
unless stated otherwise. All calculations and statistics were computed
using the Matlab software (Matlab Student version, 2013a, Mathworks, USA).
Results
Environmental conditions
The annual mean Ta and total PPT from March 2014 to February 2015 were
7.2 ∘C and 784 mm, respectively, which suggests warmer conditions
with normal wetness when compared to the long-term climate normal (5.8 ∘C
and 764 mm). PAR peaked in the first week of July while the
seasonal Ta curve peaked at around 23 ∘C in late July (Fig. 1a).
A prolonged warm and dry period occurred from early to late July with a
mean Ta of 20.0 ∘C and total rainfall of 43.3 mm.
(a) Daily means of air temperature (Ta) and photosynthetically
active radiation (PAR) and (b) daily sums of precipitation (PPT) and daily means of water table level (WTL) in restoration treatments
with high (ResH) and low (ResL) WTL, and bare peat (BP) from March 2014 to February 2015; Ta, PAR
and PPT data are taken from the Viljandi and Tartu meteorological stations (until 17 June)
and measured at the study site (from 18 June onward).
The WTL ranged from -2 to -52 and from -8 to -59 cm in the restored ResH
and ResL treatments, respectively, while remaining between -26 and -69 cm
in the unrestored BP site (Fig. 1b). The mean WTLs in ResH and ResL were
-24 and -31 cm, respectively, resulting in a mean annual difference of 7 cm
between the restored treatments. Throughout the year, the WTL in ResH was
always higher than in ResL with the difference varying between 3 and 10 cm.
The mean WTL in BP was -46 cm resulting in mean differences of -22 and -15 cm
compared to ResH and ResL, respectively.
Vegetation cover and composition
The total surface cover, i.e., the fraction of re-colonized surface area,
inside the flux measurement collars was higher in the wetter ResH (63 %)
than in the drier ResL (52 %) treatment. Bryophytes were more abundant in
ResH (62 %) than in ResL (44 %) (Table 3). The bryophyte cover
consisted primarily of Sphagnum species which contributed 98 and 96 % in ResH and
ResL, respectively. Vascular plants occurred more frequently in the drier
ResL (14 %) than in the wetter ResH (4 %) treatment and were dominated
by woody plants (i.e., shrubs and tree seedlings) (Table 3). The cover of
sedges was < 1 % in both restored treatments.
Vegetation cover (%) inside the collars for greenhouse gas flux
measurements in restoration treatments with high (ResH) and low (ResL)
water table level. Total surface cover represents the area of bare peat
surface re-colonized by vegetation; numbers in parenthesis indicate the
range among individual collars.
Species
ResH
ResL
Bryophytes
62 (32 to 93)
44 (15 to 74)
Sphagnum mosses
61 (31 to 91)
43 (12 to 70)
Vascular plants
4 (2 to 9)
14 (5 to 22)
Shrubs and tree seedlings
2 (0 to 7)
13 (5 to 22)
Sedges
< 1
< 1
Total surface cover
63 (35 to 95)
52 (20 to 85)
Carbon dioxide fluxes
Daytime NEE was positive indicating CO2 emissions during the
non-growing-season months (November to April) in all three treatments
(Fig. 2a). During the early (i.e., June) and late (i.e., mid-August to
September) summer, net CO2 uptake occurred in both ResH and ResL with
maximum rates of -42 and -41 mg C m-2 h-1, respectively. However,
during the warm and dry mid-summer period, CO2 emissions of up to 36
and 27 mg C m-2 h-1 were observed in ResH and ResL,
respectively. In contrast, NEE remained positive in BP throughout the
growing season and followed the seasonal pattern of Ta with maximum emission
rates of 104 mg C m-2 h-1 occurring in early August. The annual
mean midday NEEs in ResH and ResL were significantly lower than in BP but
not significantly different between the two restored treatments (Table 4).
(a) Net ecosystem CO2 exchange (NEE), (b) ecosystem
respiration (Re), (c) gross primary production (GPP), (d) net primary
production (NPP), (e) autotrophic respiration (Ra) and (f) heterotrophic
respiration (Rh) in restoration treatments with high (ResH) and low (ResL)
water table level and bare peat (BP); error bars indicate standard error;
the horizontal dotted line in (a) visualizes the zero line above and below
which CO2 emission and uptake occur, respectively.
Means of measured CO2 fluxes (mg C m-2 h-1)
including net ecosystem exchange (NEE), ecosystem respiration (Re), gross
primary production (GPP), net primary production (NPP), autotrophic
respiration (Ra) and heterotrophic respiration (Rh), as well as means of
measured methane (CH4; µg C m-2 h-1) and nitrous oxide
(N2O; µg N m-2 h-1) fluxes in restoration treatments
with high (ResH) and low (ResL) water table level and bare peat (BP);
negative and positive fluxes represent uptake and emission, respectively; numbers in parenthesis indicate standard error; different letters indicate
significant (P < 0.05) differences among treatments.
Component flux
ResH
ResL
BP
NEE
0.57 (4.9)c
-2.82 (4.9)c
44.9 (8.2)ab
Re
29.9 (5.1)c
35.1 (6.4)c
44.9 (8.2)ab
GPP*
-49.3 (7.4)a
-65.5 (7.3)b
n/a
NPP*
-41.5 (5.3)
-48.1 (4.2)
n/a
Ra*
7.9 (2.6)a
16.2 (3.4)b
n/a
Rh*
37.0 (5.1)c
38.5 (5.9)c
71.2 (8.4)ab
CH4
23.0 (10.7)
10.9 (6.1)
14.7 (3.7)
N2O
-0.12 (0.25)c
2.13 (1.29)c
27.1 (9.1)ab
* Growing season mean (1 May to October 31);
n/a is not applicable.
Midday Re was similar for all treatments during the non-growing-season
months (Fig. 2b). During the growing season, however, midday Re differed
among treatments with lowest and highest Re observed in ResH and BP,
respectively. Re in ResH and ResL reached maximum values
of 74 and 96 mg C m-2 h-1 during early July, respectively, whereas Re peaked at
104 mg C m-2 h-1 in early August in BP. The annual mean midday Re was
significantly lower in ResH and ResL than in BP (Table 4).
From early June to late August, both the daytime GPP and NPP were more
negative
(i.e., representing greater production) in the drier ResL than in the wetter
ResH treatment (Fig. 2c, d). Greatest GPP
occurred in late June and mid-August reaching -90 and -98 mg C m-2 h-1
in ResH and ResL, respectively. GPP temporarily decreased
(i.e.,
resulting in less negative values) to -14 and -41 mg C m-2 h-1
during the warm and dry mid-summer period in both ResH and ResL. The
seasonal patterns of NPP followed closely those of GPP, reaching -65 and -68 mg C m-2 h-1 in
ResH and ResL, respectively. The growing season
mean GPP in ResH (-49.3 mg C m-2 h-1) was significantly higher
than that in ResL (-65.5 mg C m-2 h-1) (Table 4). The difference
in the growing season means of NPP in ResH and ResL was not statistically
significant.
Midday Ra was more than 2 times greater in the drier ResL than in the
wetter ResH treatment for most of the growing season sampling dates (Fig. 2e).
The seasonal pattern of Ra coincided with that of GPP in both restored
treatments with greatest Ra occurring in late June and mid-August, reaching
maximum values of up to 27 and 36 mg C m-2 h-1 in ResH and ResL,
respectively. The growing season mean Ra was significantly higher (by about
2 times) in ResL than in ResH (Table 4). The ratio of Ra to Rh was on
average 0.21 and 0.42 in ResH and ResL, respectively.
Midday Rh was consistently lower in ResH and ResL than in BP throughout
the growing season (Fig. 2f). Maximum Rh of up to 61, 73 and 104 mg C m-2 h-1
in ResH, ResL and BP, respectively, were observed in
early July (restored treatments) and early August (unrestored BP). The
growing season mean Rh was significantly lower (by about 50 %) in ResH
and ResL than in BP (Table 4).
Methane fluxes
Throughout most of the year, CH4 fluxes were observed in the range of
-13 to 60 µg C m-2 h-1 in all three treatments (Fig. 3a).
However, occasional peak CH4 emission of up to 170 and 92 µg C m-2 h-1
occurred in ResH and ResL, respectively. During the non-growing-season months, CH4 exchange was variable, showing both small uptake as
well as large emission (-6 to 138 µg C m-2 h-1). The mean
annual CH4 exchange was about 2 times greater in the wetter ResH
than in the drier ResL treatment although the differences among the three
treatments were not statistically significant (Table 4).
Measured fluxes of (a) methane (CH4; µg C m-2 h-1)
and (b) nitrous oxide (N2O; µg N m-2 h-1) in
restoration treatments with high (ResH) and low (ResL) water table level
and bare peat (BP); error bars indicate standard error; the horizontal
dotted line in (a) visualizes the zero line above and below which CH4
emission and uptake occur, respectively.
Nitrous oxide fluxes
N2O fluxes in ResH and ResL remained within the range of -2.8
to 25 µg N m-2 h-1 for most of the year (Fig. 3b). In contrast,
high N2O emissions of 66 to 133 µg N m-2 h-1 occurred
during July and August in BP. The annual mean N2O exchanges of
-0.12 µg N m-2 h-1 in ResH and 2.13 µg N m-2 h-1
in ResL were not significantly different (Table 4). Meanwhile, the
mean N2O exchanges in the two restored treatments were significantly
lower (by 1–2 magnitudes) compared to the 27.1 µg N m-2 h-1
in BP (Table 4).
Biotic and abiotic controls of GHG fluxes
The differences in mean growing season NEE, GPP, NPP and Ra among individual
collars (i.e., the spatial variability) were significantly correlated to
bryophyte but not to vascular plant cover in ResH (Table 5). In contrast,
spatial variations in NEE, GPP, NPP and Ra were significantly correlated to
vascular plant but not to bryophyte cover in ResL. In addition, Re was
significantly correlated to vascular plant cover in ResL. Meanwhile, the
CH4 and N2O exchanges were not significantly correlated to
vegetation cover in either ResH or ResL.
Correlation coefficients of vegetation (bryophytes and vascular
plants) cover (%) with mean growing season CO2 fluxes including the
net ecosystem CO2 exchange (NEE), ecosystem respiration (Re), gross
primary production (GPP), net primary production (NPP) and autotrophic
respiration (Ra) and with mean growing season methane (CH4) and nitrous
oxide (N2O) fluxes in restoration treatments with high (ResH) and low
(ResL) water table level. Total vegetation represents the sum of bryophyte
and vascular plant cover; significant correlations are marked with asterisks.
ResH
ResL
Vegetation cover
NEE
Re
GPP
NPP
Ra
CH4
N2O
NEE
Re
GPP
NPP
Ra
CH4
N2O
Bryophytes
-0.95**
0.74
-0.95**
-0.84*
0.97**
-0.53
-0.56
-0.75
0.67
-0.81*
-0.70
0.78
-0.33
-0.34
Vascular plants
-0.70
0.49
-0.76
-0.68
0.60
-0.07
-0.05
-0.92**
0.93**
-0.97**
-0.93**
0.89*
0.13
0.22
Total vegetation
-0.95**
0.74
-0.95**
-0.84*
0.96**
-0.50
-0.53
-0.82*
0.72
-0.84*
-0.75
0.88*
-0.21
-0.19
* indicates P <0.05 and ** indicates P < 0.01.
Growing season (GS; 1 May to 31 October) and annual (A) sums of the
carbon balance components (g C m-2) including gross primary production
(GPP), ecosystem respiration (Re), net ecosystem exchange (NEE) of CO2
and methane (CH4) fluxes, as well as the greenhouse gas (GHG) balance
components (t CO2 eq ha-1) including NEE, CH4 and nitrous
oxide (N2O) exchanges (using global warming potentials of 34 and 298
for CH4 and N2O, respectively) in restoration treatments with high
(ResH) and low (ResL) water table level and bare peat (BP); negative and
positive fluxes represent uptake and emission, respectively.
ResH
ResL
BP
Component flux
GS
A
GS
A
GS
A
C balance components
GPP
-78.0
-78.0
-110.5
-110.5
n/a
n/a
Re
127.5
188.6
148.8
213.2
180.5
267.8
NEE
49.5
110.6
38.3
102.7
180.5a
267.8a
CH4
0.130
0.190
0.036
0.117
0.076
0.137
Total C balanceb
110.8
102.8
268.0
GHG balance components
NEE
1.81
4.05
1.40
3.76
6.62
9.82
CH4
0.059
0.086
0.016
0.053
0.035
0.062
N2O
0.002
0.004
0.010
0.020
0.167
0.332
Total GHG balancec
4.14
3.83
10.21
a GPP for BP was assumed to be 0 and NEE therefore equal to
Re.
b The total C balance (g C m-2 yr-1) is the sum of NEE and
CH4 fluxes.
c The total GHG balance (t CO2 eq ha-1 yr-1) is the sum
of NEE, CH4 and N2O fluxes.
n/a is not applicable.
Ts measured at 10 cm depth was the abiotic variable that best
explained variations in Re (R2= 0.79, 0.84 and 0.81 in ResH, ResL
and BP, respectively) in the form of an exponential relationship (Fig. 4) with
higher temperatures resulting in higher respiration rates. The basal
respiration and temperature sensitivity parameters were lowest in the wetter
ResH treatment and highest in BP.
Response of ecosystem respiration (Re; mg C m-2 h-1) to
changes in soil temperature (Ts) measured at 10 cm soil depth in restoration
treatments with high (ResH) and low (ResL) water table level and bare peat
(BP).
N2O fluxes correlated best with VWC measured at
0–5 cm soil depth in ResL (R2= 0.60) and in BP (R2= 0.39)
(Fig. 5). In contrast, N2O fluxes were not correlated to soil
VWC or any other abiotic variable in ResH. Similarly,
the CH4 exchange did not show any significant relationships with any
abiotic variable for any of the three treatments.
Response of nitrous oxide (N2O) fluxes (µg N m-2 h-1)
to changes in volumetric water content (VWC) measured at 0–5 cm
soil depth during the growing season in restoration treatments with high
(ResH) and low (ResL) water table level and bare peat (BP).
Annual carbon and GHG balances
In the restored ResH and ResL treatments, the modeled annual Re estimates
were 188.6 and 213.2 g C m-2 yr-1, respectively, whereas in
the unrestored BP treatment annual Re was 267.8 g C m-2 yr-1
(Table 6). The annual GPP was estimated at -78.0 and -110.5 g C m-2 yr-1
in ResH and ResL, respectively. This resulted in annual NEEs of 110.6, 102.7 and 267.8 g C m-2 yr-1 in the
wetter ResH, drier ResL and BP treatments, respectively. The growing
season net CO2 loss represented 45 and 37 % of the annual
NEE in ResH and ResL, respectively, while it accounted for
67 % in BP. The additional C losses via CH4 emission were 0.190,
0.117 and 0.137 g C m-2 yr-1 in ResH, ResL and BP,
respectively. In total, all treatments acted as C sources; however, the
annual C balance was lower in the restored ResH (110.8 g C m-2 yr-1)
and ResL (102.8 g C m-2 yr-1) treatments than in the
unrestored BP (268.0 g C m-2 yr-1) treatment. The total GHG
balance, including NEE as well as CH4 and
N2O emissions expressed as CO2 eq, was 4.14, 3.83 and 10.21 t CO2 eq ha-1 yr-1
in ResH, ResL and BP, respectively (Table 6). The GHG balance
was driven by NEE (96 to 98 %)
in all three treatments. The contribution of CH4 emission was highest
(2.1 %) in the wetter ResH treatment, while the contribution of N2O
emission was highest (3.9 %) in the unrestored BP treatment.
Discussion
GHG fluxes and their controls in restored and abandoned peat
extraction areas
Coupling of water table level and vegetation dynamics
Three years following restoration, contrasting vegetation communities in
ResH and ResL had developed as a result of a mean annual WTL difference of
7 cm. Specifically, a greater cover of bryophytes (63 %) (primarily
Sphagnum spp.), which rely on capillary forces for acquiring water and thus require
moist conditions (Rydin, 1985), was present in the wetter ResH treatment.
In contrast, the lower WTL in ResL resulted in a lower bryophyte cover
(44 %) but greater abundance of vascular plants, likely due to the
extended zone of aeration for plant roots. Apart from having roots to absorb
water and nutrients from the soil, vascular plants also differ from
bryophytes by having leaf stomata to regulate water transport and CO2
exchange (Turner et al., 1985; Schulze et al., 1994). Thus, the
establishment of contrasting vegetation communities as a result of different
WTL baselines has potential implications for the biogeochemical cycles and
GHG fluxes following peatland restoration (Weltzin et al., 2000).
Carbon dioxide fluxes
In this study, the significantly higher GPP in ResL was likely due to the
greater vascular plant cover compared to ResH, since vascular plants reach
higher photosynthesis rates at higher light levels compared to mosses
(Bubier et al., 2003; Riutta et al., 2007a). Similarly, Strack and Zuback (2013)
reported a strong correlation between vascular plant cover and GPP in
a restored peatland in Canada. In return, the greater GPP also explains the
higher Ra observed in ResL compared to ResH. This highlights the
implications of hydrological differences and the associated vegetation
development on plant-related CO2 fluxes. Furthermore, it has been
suggested that the presence of vascular plants can facilitate greater
survival and better growth of the re-introduced mosses as they can provide
shelter from the intense solar radiation and wind and thus create a more
favorable micro-climate (Ferland and Rochefort, 1997; Tuittila et al.,
2000b; McNeil and Waddington, 2003; Pouliot et al., 2012). Since Sphagnum mosses are
generally more sensitive to drought compared to vascular plants, restoration
strategies allowing the development of a diverse vegetation cover (i.e.,
bryophytes accompanied by vascular plants) could therefore be considered to
have greater potential for limiting the CO2 loss and regaining the C sink
function (Tuittila et al., 1999). Nevertheless, despite the significant
effects of the re-established WTL baseline on vegetation development and the
associated CO2 component fluxes (i.e., Re and GPP), the NEE of the two restored treatments was similar. Our study therefore
suggests that the greater GPP was partly counterbalanced by greater Ra in
ResL compared to ResH. However, while differences in the re-established
WTL baseline had no significant effect on the CO2 sink–source strength
3 years after restoration of the abandoned peat extraction area,
vegetation characteristics are likely to further diverge in the future which
might essentially result in contrasting net CO2 balances over longer
time spans (Weltzin et al., 2000; Yli-Petäys et al., 2007; Samaritani et
al., 2011; Vanselow-Algan et al., 2015).
Compared to the unrestored BP treatment, growing season Rh was considerably reduced in the
restored treatments which suggests that raising the WTL effectively
mitigated C losses from the ecosystem by reducing the potential for aerobic
peat decomposition (Silvola et al., 1996; Frolking et al., 2001; Whiting and
Chanton, 2001). Furthermore, the significantly lower Re
in ResH and ResL compared to BP demonstrates that the additional
Ra from the vegetation was negligible compared
to the large reduction in Rh. Likewise, Strack and Zuback (2013) found a
significantly lower Rh and Re in a restored compared to an unrestored site
in Canada 10 years following peatland restoration. Furthermore, the lower Re
in the restored treatments relative to BP might also result from the lower
temperature sensitivity of Rh
observed in this study which is likely due to greater oxygen limitation in
the restored treatments following the raising of the WTL. Thus, our findings
highlight the effectiveness of raising the WTL in reducing peat
decomposition and associated CO2 emissions from drained organic soils.
Methane fluxes
Both WTL and vegetation dynamics have been previously highlighted as major
controls of the CH4 exchange in natural, restored and drained peatlands
(Bubier, 1995; Frenzel and Karofeld, 2000; Tuittila et al., 2000a; Riutta et
al., 2007b; Waddington and Day, 2007; Lai, 2009; Strack et al., 2014).
Specifically, the WTL determines the depth of the lower anaerobic and the upper
aerobic peat layers and thus the potential for CH4 production and
consumption occurring in these respective layers (Bubier, 1995; Tuittila et
al., 2000a). The relatively low mean annual WTLs (i.e., -24, -31 and -46 cm
in ResH, ResL and BP, respectively) might therefore explain the generally
low CH4 emission rates observed in our study compared to those
previously reported in similar ecosystems (Tuittila et al., 2000a; Basiliko
et al., 2007; Waddington and Day, 2007; Lai, 2009; Vanselow-Algan et al.,
2015). Nevertheless, high autumn peak emissions were observed in all
treatments that might be caused by a rapid drop in the WTL during which
CH4 may have been released from the pore water and emitted to the
atmosphere as shown in previous studies (e.g., Windsor et al., 1992; Moore
and Dalva, 1993). These episodic emission peaks indicate a potential for
higher annual CH4 emissions following peatland restoration than those
estimated in this study.
Vegetation composition affects the CH4 production through substrate
supply (i.e., quality and quantity) (Saarnio et al., 2004; Ström et al.,
2005) and by offering a direct emission pathway for CH4 from the deeper
anaerobic layer to the atmosphere via the aerenchymatous cell tissue of deep
rooting sedge species such as Eriophorum vaginatum (Thomas et al., 1996; Frenzel and
Karofeld, 2000; Ström et al., 2005; Waddington and Day, 2007). Given the
considerable differences in vegetation composition, the lack of significant
effects on CH4 emissions among the restored and BP treatments in our
study was surprising. Most likely, similar CH4 emissions in ResH and
ResL were the result of opposing effects counterbalancing the production
and consumption of CH4. For instance, enhanced anaerobic CH4
production due to the higher WTL in ResH could have been partly compensated by
greater CH4 oxidation within or immediately below the more developed
moss layer (Frenzel and Karofeld, 2000; Basiliko et al., 2004; Larmola et
al., 2010). In ResL, however, greater vascular plant substrate
supply might have sustained substantial CH4 production despite a
reduction of the anaerobic zone (Tuittila et al., 2000a; Weltzin et al.,
2000). Also noteworthy is that, while very few aerenchymatous sedge species
were established at the time of this study, a future increase in the
sedge cover is likely to occur (Tuittila et al., 2000a; Weltzin et al.,
2000; Vanselow-Algan et al., 2015) which could considerably increase the
CH4 emission in the restored treatments over longer time spans.
Overall, the potential effects from enhanced anaerobic conditions due to the
raised WTL, CH4 oxidation in the moss layer or greater vascular plant
substrate supply on the net CH4 fluxes were small, considering that
CH4 emissions were not significantly different from those in BP which
was characterized by a considerably lower WTL and absence of vegetation.
Thus, our study suggests that in non-flooded conditions WTL changes
following peatland restoration have a limited effect on the CH4
emissions during the initial few years.
Nitrous oxide fluxes
Soil moisture and WTL effects on the soil oxygen status have been previously
identified as the main control of N2O emissions from pristine and
drained peatlands (Firestone and Davidson, 1989; Martikainen et al., 1993;
Klemedtsson et al., 2005). Highest N2O emissions commonly occur in
mesic soils with intermediate WTLs, which allows both aerobic
and anaerobic N2O production during nitrification and denitrification,
respectively, while avoiding the anaerobic reduction of N2O to N2
(Firestone and Davidson, 1989; Martikainen et al., 1993). In addition,
substrate supply (i.e., C and inorganic N) is a key prerequisite for N2O
production (Firestone and Davidson, 1989). In our study, similar N2O
fluxes in the two restored treatments therefore suggest that the differences
in WTL, soil moisture and substrate supply from mineralization of organic
matter were too small to affect the magnitudes of N2O emission 3
years following restoration with different WTL baselines. In contrast,
the enhanced anaerobic conditions due to a higher WTL as well as lower soil N
concentrations due to reduced mineralization and enhanced plant N uptake
might explain both the reduced N2O emissions and their lower
sensitivity to soil moisture in the restored ResH and ResL treatments
compared to BP. Thus, peatland restoration has the potential for reducing
the N2O emissions commonly occurring in drained, abandoned peatlands by
altering both soil hydrology and N substrate supply.
The carbon and GHG balances of restored and abandoned peat
extraction areas
Both restored treatments were C sources during the growing season, which
indicates that the CO2 uptake by the re-established vegetation was not
able to compensate for the C losses via respiration and CH4 emissions
3 years following restoration. Several studies have previously reported
estimates for the growing season C sink–source strength of restored
peatlands, with contrasting findings due to different restoration
techniques, environmental conditions during the study year and time passed
since the initiation of the restoration (Tuittila et al., 1999; Bortoluzzi
et al., 2006; Yli-Petäys et al., 2007; Waddington et al., 2010;
Samaritani et al., 2011; Strack et al., 2014). For instance, restored
peatlands in Finland (Tuittila et al., 1999) and Canada (Waddington et al.,
2010; Strack et al., 2014) were C sinks during the growing season 3 to
6 years after restoration. In contrast, other studies suggested that
several decades may be required before restored peatlands resume their
functioning as C sinks (Yli-Petäys et al., 2007; Samaritani et al.,
2011). However, while growing season studies can provide important
information on processes governing the fluxes, it is necessary to quantify
and compare full annual budgets to better evaluate the climate benefits of
peatland restoration relative to abandoned peatland areas (and other
after-use options, e.g., afforestation or energy crop cultivation).
In our study, the annual C source strength of the restored
and BP treatments was about 1.5 to 2.5 times greater than on the
growing season scale. This highlights the importance of accounting for the
considerable non-growing-season emissions when evaluating the C sink
potential of restored peatlands. In comparison, the annual C source strength
of the two restored treatments (111 and 103 g C m-2 yr-1) was
lower than the annual emissions of 148 g C m-2 yr-1 reported for a
restored cutaway peatland in Canada 10 years following restoration (Strack
and Zuback, 2013). Similarly, the C balance of BP (268 g C m-2 yr-1)
in our study was about half of the 547 g C m-2 yr-1
emitted at the Canadian unrestored site. However, high emissions in the
study of Strack and Zuback (2013) were partly attributed to the dry
conditions during the study year. This indicates that restored
peatlands are unlikely to provide an annual C sink during the first decade
following restoration of peat extraction sites. However, compared to
naturally re-vegetating peatlands, which may require 20–50 years to reach a
neutral or negative C balance (Bortoluzzi et al., 2006; Yli-Petäys et
al., 2007; Samaritani et al., 2011), initiating the restoration by rewetting
in combination with re-introduction of peatland vegetation might reduce the
time required for the ecosystem to return to being a C sink similar to that
of a natural peatland (Tuittila et al., 2004; Roulet et al., 2007; Nilsson
et al., 2008).
The similar GHG balances in the two restored treatments ResH and ResL
suggest that the differences in the mean WTL had a limited effect on the GHG
balance within the few years following restoration of the peat extraction
area. Moreover, the GHG balances in the restored treatments were driven
primarily by the NEE, while the contribution of CH4
and N2O exchanges remained minor in our study. In contrast, 30 years
after rewetting of a German bog, high CH4 emission were reported as the
main component of the GHG balance (Vanselow-Algan et al., 2015). The same
study also reported GHG balances ranging from 25 to 53 t CO2 eq ha-1 yr-1
which are considerably higher compared to our study. This
indicates that the GHG balances of restored peatlands may vary greatly over
longer time spans. Moreover, this also suggests the GHG balance of peatland
restoration with differing WTL baselines is likely to further diverge over
time due to contrasting trajectories in vegetation development and changes
in soil biogeochemistry (e.g., pH, nutrient contents and soil moisture
dynamics).
While the two restored treatments had similar GHG balances, the difference
between the GHG balances in restored and BP treatments was considerable.
Only 3 years following restoration, the GHG balance in the restored
treatments was reduced to about half of that in BP. This reduction was
mainly due to lower annual CO2 emissions (i.e., lower NEE) in the
restored treatments compared to BP likely as a result of increased WTL and
vegetation development. In addition, annual N2O emissions were also
significantly reduced in the restored treatments, although, compared to the
differences in the CO2 balance, the impact of the reduction in N2O
emissions on the GHG balance was relatively small. Overall, our study
suggests that peatland restoration may provide an effective method to
mitigate the negative climate impacts of abandoned peat extraction areas in
the short term. However, due to the lack of long-term observations and
recent reports of potential high CH4 emissions occurring several
decades after rewetting (Yli-Petäys et al., 2007; Vanselow-Algan et al.,
2015), it remains uncertain whether restoration of abandoned peat extraction
areas may also provide an after-use solution with climate mitigation
potential in the long term.