BGBiogeosciencesBGBiogeosciences1726-4189Copernicus GmbHGöttingen, Germany10.5194/bg-12-4361-2015High methane emissions dominated annual greenhouse gas balances 30
years after bog rewettingVanselow-AlganM.marion.vanselow-algan@uni-hamburg.deSchmidtS. R.GrevenM.FienckeC.KutzbachL.https://orcid.org/0000-0003-2631-2742PfeifferE.-M.University of Hamburg, Center for Earth System Research
and Sustainability, Institute of Soil Science, Hamburg,
GermanyUniversity of Hamburg, Biocenter Klein Flottbek, Applied
Plant Ecology, Hamburg, GermanyM. Vanselow-Algan (marion.vanselow-algan@uni-hamburg.de)28July201512144361437116December201410February20151July2015This work is licensed under a Creative Commons Attribution 3.0 Unported License. To view a copy of this license, visit http://creativecommons.org/licenses/by/3.0/This article is available from https://bg.copernicus.org/articles/12/4361/2015/bg-12-4361-2015.htmlThe full text article is available as a PDF file from https://bg.copernicus.org/articles/12/4361/2015/bg-12-4361-2015.pdf
Natural peatlands are important carbon sinks and sources of methane
(CH4). In contrast, drained peatlands turn from a carbon sink to a
carbon source and potentially emit nitrous oxide (N2O). Rewetting of
peatlands thus potentially implies climate change mitigation. However, data
about the time span that is needed for the re-establishment of the carbon
sink function by restoration are scarce. We therefore investigated the annual
greenhouse gas (GHG) balances of three differently vegetated sites of a bog
ecosystem 30 years after rewetting. All three
vegetation communities turned out to be sources of carbon dioxide (CO2)
ranging between 0.6 ± 1.43 t CO2 ha-2 yr-1
(Sphagnum-dominated vegetation) and 3.09 ± 3.86 t CO2 ha-2 yr-1 (vegetation dominated by heath).
While accounting for the different global warming potential (GWP) of
CO2, CH4 and N2O, the annual GHG balance was calculated.
Emissions ranged between 25 and 53 t CO2-eq ha-1 yr-1 and were dominated by large
emissions of CH4 (22–51 t CO2-eq ha-1 yr-1),
with highest rates found at purple moor grass (Molinia caerulea) stands. These are to our
knowledge the highest CH4 emissions so far reported for bog ecosystems
in temperate Europe. As the restored area was subject to large fluctuations
in the water table, we assume that the high CH4 emission rates were caused
by a combination of both the temporal inundation of the easily decomposable
plant litter of purple moor grass and the plant-mediated transport through
its tissues. In addition, as a result of the land use history, mixed soil
material due to peat extraction and refilling can serve as an explanation.
With regards to the long time span passed since rewetting, we note that the
initial increase in CH4 emissions due to rewetting as described in the
literature is not inevitably limited to a short-term period.
Introduction
Covering only 3 % of the Earth's land surface, peatlands
store as much carbon as all terrestrial biomass and twice as much as all
global forest biomass (Parish et al., 2008). Today, merely 1 % of the
former peatlands areas in western Europe still has living mire vegetation
and accumulates peat (Koster and Favier, 2005). An area of 80 million ha of
peatlands have been destroyed worldwide, mainly due to drainage for
agriculture and forestry, and due to peat mining for fuel and horticulture
(Joosten and Clarke, 2002). While growing mires have a cooling effect on the
climate by acting as a carbon sink, degraded peatlands are a major and
growing source of the greenhouse gases (GHG) carbon dioxide (CO2) and
nitrous oxide (N2O; Joosten et al., 2012).
CO2 emissions from degraded peatlands are estimated to be equivalent to
more than 10 % of global fossil fuel emissions (Parish et al., 2008).
Even in an industrialized country, such as Germany, does the estimated peatland GHG
exchange account for 2.3–4.5 % of the anthropogenic emissions
(Drösler et al., 2008). Therefore, restoration of peatlands includes not
only the recovery of ecosystem functions and biodiversity, but also climate
change mitigation (Drösler et al., 2009). Rewetting of drained peat
soils as climate change mitigation measure presents a new challenge (Erwin,
2009; Couwenberg, 2009): in Germany, a potential reduction of 35 million
tons CO2 equivalent per year is possible through peatland restoration,
which is a cost-effective mitigation strategy (Joosten, 2006; Drösler et
al., 2009; Freibauer et al., 2009). However, rewetting of
peatlands leads to increased methane emissions by both reduced methane
oxidation and increased methane production (Komulainen et al., 1998;
Tuittila et al., 2000; Waddington and Day, 2007; Wilson et al., 2009; Cooper
et al., 2014). Flooding of peat soils for restoration should be avoided as
inundation leads to huge CH4 emissions especially if fresh plant litter
is available (Augustin and Joosten, 2007; Drösler et al., 2008;
Hahn-Schöfl et al., 2011). Although these high CH4 emissions are
generally assumed to be a transient phenomenon of limited duration (Morison,
2012; Joosten et al., 2012; Artz et al., 2013; Cooper et al., 2014), no
information exists on the length of this time span after rewetting (Augustin and Joosten, 2007). Nevertheless, the term “short-term”
is often used in grey literature (Trepel, 2008; Morison, 2012). Moreover,
there is high uncertainty about the magnitude of CH4 fluxes in
peatlands (Joabsson et al., 1999). However, it is generally agreed that
vascular plants stimulate CH4 emissions by allowing the gas to
bypass the oxygenated upper soil layer moving through the plant tissues
(Joabsson et al., 1999; Kutzbach et al., 2004; Lai, 2009). This
plant-mediated transport can represent between 30 and almost 100 % of
the total methane flux (Bhullar et al., 2013). Plants that provide such a
shortcut between the root zone and the atmosphere are referred to as
“shunt” species (Joosten et al., 2012). Many studies linked the increase of
CH4 emissions post-restoration to the colonization of the rewetted area
by Eriophorum species (Tuittila et al., 2000; Waddington and Day, 2007; Green and
Baird, 2012; Cooper et al., 2014).
In the future, restoration and management of peatlands will be more complex
and challenging due to climate change (Erwin, 2009; Grand-Clement et al.,
2013). Thus, extensive knowledge of peat soil processes will help in the
adaptation of management practices and in turn in stabilization of ecosystem
functions of peatlands against new environmental threats. However, data
about the C balance of restored and abandoned peatlands is scarce on
national and global scales and is urgently needed (Yli-Petäys et al.,
2007; Artz et al., 2013). As most studies focus on the boreal region
or are conducted mainly at recently rewetted sites (Beyer and Höper,
2015), research on the mid- and long-term effect of rewetting is essential,
especially in temperate peatlands.
We studied the CO2 and GHG balances of three different plant
communities at a regenerating mined area and of the active peat extraction
site within the same ombrotrophic peatland in northern Germany. Our main
goal was to investigate if restoration succeeded in creating a CO2 sink and to estimate the GHG mitigation potential of restoring the
industrial peat extraction site after abandonment. We hypothesized that the
restored site acted as a CO2 sink with varying magnitude depending on
plant community. Likewise, we expected that higher CH4 emissions were
related to high abundance of vascular plant species and thus varied between
the three plant communities. On the other hand, no significant N2O
fluxes were expected on the vegetated sites. Testing theses hypotheses could
provide recommendations for site-specific management actions, because peat
mining of the study area will be ceased in a few years.
Material and methodsStudy area Himmelmoor
The Himmelmoor (Quickborn, Germany; 53∘44′20′′ N,
9∘50′58′′ E) is located approximately 20 km northwest of
Hamburg in Schleswig-Holstein. The mean annual precipitation measured at the
climate station in Quickborn is 838 mm, and the mean air
temperature is 9.0 ∘C (long-term average from 1981 to 2010, data
source: Deutscher Wetterdienst). Climatically, there was
821 mm of precipitation and an average temperature of 9.6 ∘C during study year 2011, very close to the long-term average. We
therefore consider our measurements as representative for the present
climatic conditions.
With an extent of about 6 km2, the Himmelmoor is one of the
largest raised bogs in Schleswig-Holstein. Peat formation by
terrestrialization started after the last ice age 10.020 ± 100 years
before present and the total peat thickness reached a maximum of 10 m
(Pfeiffer, 1998; Grube et al., 2010). The original ombrotrophic peat bog was
altered by peat drainage and cutting began in the 18th century. Commercial
peat mining started in 1871 (Averdieck, 1957) and persists to this day.
According to the peat mining company, a volume of 38 000 m3 of peat
was harvested in the study year 2011 over an area of 70 ha of the former
130 ha extraction site (K. Czerwonka, personal communication, 2013). Peat
mining will cease in 2016 due to the exhaustion of usable peat resources,
and restoration will take place.
Measurement sites
Measurements were done on the bare and deeply drained active peat extraction
site and on a restored and vegetated part in the northwestern Himmelmoor.
All soils are classified as Fibric Ombric Histosol, while for the peat extraction site, the suffix
qualifier Drainic applies in addition (IUSS, 2006). The restored site was formerly used
by hand block-cutting for fuel. Before cutting the drained black peat, the
upper rooted layer called Bunkerde (Poschlod, 1988) was removed and used for filling
up cut-over areas as described in Koster and Favier (2005). In the 1960s the site
was abandoned and vegetation development started, mainly from whole plants or
generative and vegetative propagules in the Bunkerde. Restoration began in the 1980s
with drainage blocking and the repeated cutting of birches (last event: 2008) in
order to raise the water table. Today, the strips between former drainage
ditches (distance: 30–45 m) show differences in plant species composition.
Three strips with differing vegetation, typical for rewetted peatlands, were
chosen and named after the most prominent plant species or groups, later
referred to as “heath”, “Sphagnum” or “purple moor grass” site. According to
the decimal scale of Londo (1976), plotwise analyzes of vascular plant
species were done in the study year 2011 during summertime, providing
vegetation coverage and abundance to confirm the visual classification of
these sites. The moss cover was determined using pointintercept sampling
as described in Jonasson (1988), with the following categories: Sphagnum mosses,
other mosses, liverworts and bare peat.
GHG fluxes were measured at each of the four sites in four replicates
(plots). Therefore, PVC frames (60 cm × 60 cm) with a soil insertion depth
of about 50 cm were permanently installed blockwise. Positions were selected
based on vegetation and microtopography with the intention of representing the
whole site. All plots of the vegetated site were equipped with wooden
boardwalks to minimize soil gas disturbances during flux measurements.
Boardwalks were oriented northwards of the plots to avoid shading. Several
micrometeorological variables were continuously monitored at a central point
of the restored site, including air temperature and air pressure,
photosynthetically active radiation (PAR), wind speed and wind direction
(all at 2 m height) and precipitation. Water table depth and soil
temperature (10 cm depth) were measured at each site except for the peat
extraction site.
Chamber design and flux measurement procedure
GHG flux measurements were done using closed chambers. Their design conforms
to the latest recommendations for chamber design made by Pihlatie et al. (2013). CO2 fluxes were measured with a transparent, climate-controlled
chamber system connected to an infrared gas analyzer (IRGA, LI-840, LI-COR
inc.) as described by Schneider et al. (Schneider et al., 2011).
Additionally, the chamber was equipped with a PAR sensor inside.
Measurements were done for 3 min, recording CO2 concentration,
PAR and chamber air temperature every second. To gain a wide spectrum of
different light conditions for modeling, the transparent chamber was shaded
in two intensities with black gauze (Elsgaard et al., 2012; Görres et
al., 2014). After the first measurement with the transparent chamber, a
second measurement was performed while shading the chamber with one layer of
gauze (PAR approx. 50 %) and a third measurement was done with two layers
(PAR approx 30 %). Subsequently, the chamber was darkened in a fourth
measurement with an opaque cover (PAR = 0) to estimate ecosystem
respiration (Reco). Between each measurement the chamber was removed and
ventilated to obtain ambient CO2 concentrations within the chamber. If
vegetation exceeded chamber height, a transparent polycarbonate elongation
of 60 cm height was used, which was shaded and darkened correspondingly.
Measurements were generally conducted between 10:00 and 14:00 local time when PAR
reached the maximum, and the measuring order of the sites was randomized.
CO2 flux measurements were performed year-round from August
2010 until January 2012. Measurement intervals depended on vegetation growth
with higher frequency in summer than in winter (at least twice a month up to
twice a week). To compare day and night respiratory fluxes, nighttime
measurements were taken three times during the night of 30–31 August 2011
(around 21:00, 00:00 and 05:00). CH4 and N2O flux
measurements were performed over a 1-year period from April 2011 until
March 2012. Measurements were carried out every 2 weeks (except December
2011: once per month) for CH4 and monthly for N2O as first results
showed no significant N2O fluxes. CH4 was measured at all four
replicate plots, N2O only at three. Nighttime measurements were
performed during the night of 30 to 31 August 2011 around midnight. CH4
and N2O flux measurements were done using aluminum chambers (60 cm × 60 cm × 32 cm; an elongation of 60 cm was used if needed) which were
equipped with a fan, a pressure vent, a temperature sensor and a sampling
port. Two circular openings (4 cm diameter) at the front side were open
while placing the chamber on the collar and closed afterwards to reduce
initial pressure shocks (Schneider et al., 2009). During the closure time of
20 min, six samples were taken from the chamber headspace with 60 mL
plastic syringes connected to the sampling port via three-way stopcocks.
CH4 samples were analyzed subsequently in the lab using a gas
chromatograph (GC) equipped with a flame ionization detector (HP 5890
Packard Series II). Syringes were directly connected via a loop. Analyses
were done within 4 days after sampling, and each sample was analyzed
twice. Two standard gases were used for calibration (1.7 and 200 ppm
CH4), injected in triplicates before and after samples of three
plots.
N2O was measured at a GC provided with an electron capture detector
(Agilent Technologies 7890A). In the field, a sample volume of 20 mL was
injected into an air-filled septum vial, from where it was then taken with a
microliter syringe and injected into the GC. The GC was calibrated daily
with three standard gases (0.3, 0.9 and 1.5 ppm N2O) being
injected in triplicates before measurement. Since the samples were diluted
by injecting them into the air-filled vials, standards were treated similarly
to have the appropriate concentration. As this procedure might cause a
decrease in accuracy, it was tested with a standard gas. There was no
decrease in reproducibility in comparison to a standard gas that was
injected directly into the GC: the coefficient of variation was 0.01 in both
cases (N= 10).
Flux calculation
The GHG flux rates were calculated from the change in gas concentration as a
function of time during chamber closure. Gross primary production (GPP) was
calculated as the difference between the directly following measurements
with transparent and dark chambers, respectively. According to the
micrometeorological sign convention, positive values represent fluxes to the
atmosphere and negative values uptakes by the ecosystem.
Flux calculation of each single CO2 chamber measurement was done with
an updated version of the MATLAB® routine of Kutzbach et al. (2007) using a power series expansion of a nonlinear regression as
described by Görres et al. (2014). The first and the last 10 of the
180 s of each measurement were discarded and the flux rate was
calculated at t= 10 s from the remaining 160 CO2 concentration data
points applying a water vapor dilution correction. Each single flux curve
was reviewed for abnormalities such as abrupt changes in slope due to, e.g.,
changes in PAR derived from cloud movement. If possible, the flux was
recalculated by using only a part of the 160 s interval with constant
conditions (minimum 40 s). Flux calculations with outlying residuals were
checked for mistakes, e.g., in data preparation, and were discarded from the
data set if the mistake could not be eliminated. The standard deviation of the
residuals of most of the data (98 %) was lower than 0.55 ppm and had a
mean of 0.42 ± 0.06 ppm, which is remarkably low as the noise of the
IRGA is specified to be < 1 ppm. Upward concave flux curves, which
are not explainable by diffusion theory were calculated using linear
regression as executed by Schneider et al. (2011). It was shown that this
procedure achieves more robust and less biased flux estimates (Schäfer,
2012; Görres et al., 2014).
Flux calculation of each CH4 and N2O chamber measurement was done
with an updated version of the MATLAB® routine of Forbrich et al. (2010). Each single flux curve was reviewed for abnormalities such as
ebullition and discarded from data set if necessary. Linear or nonlinear
regression was used depending on the model performances according to the
Akaike information criterion. As only six concentration measurements were
available for flux calculation, we used the Akaike information criterion
with small sample correction (AICc), as proposed by Forbrich et al. (2010).
According to AICc, the majority of CH4 and N2O fluxes curves was
better explained by linear than by exponential regression (74 and 98 %,
respectively).
Flux modeling
Net ecosystem exchange (NEE), Reco and GPP were modeled over a complete year
(2011). Modeling of GPP and Reco was based on ambient PAR and air
temperature, respectively, which were measured half-hourly at the
meteorological station. Each single plot (N= 16) was modeled separately to
analyze the differences between and within the sites. Values of the four plots
per site were later averaged, thus standard deviations shown here display
the spatial variability of flux estimates. As day and night fluxes of
CO2, CH4 and N2O were not statistically different, modeling
was only based on daytime data.
GPP was modeled with a rectangular hyperbolic light response curve as
described in Elsgaard et al. (2012) using PAR values from inside the chamber
(GPP 1 model, Table 1). Applying this curve obtains light saturation points
(Pmax) for a certain period of the year (1 week up to 1 month depending
on measurement interval and season). Gap filling between these intervals was
done using linear interpolation. If GPP 1 could not explain the data
appropriately, a linear model was used instead (GPP 2, Table 1). In
wintertime when GPP ranged around zero, neither of the two models could be
fitted and mean values were used instead.
Two model approaches for gross primary production (GPP) as a
function of photosynthetically active radiance (PAR) and two temperature-driven ecosystem respiration (Reco) models, where t= air or soil
temperature, respectively, and a, b and c are fitting parameters.
Model titleModel formulaRemarksGPP 1GPP=PmaxαPARPmax+αPARRectangular hyperbolic function(Schäfer, 2012; Elsgaard et al., 2012).Pmax= maximum potential photosynthetic rate,α= initial light response efficiency.GPP 2GPP=a+bPARLinear model, a and b are fitting parameters.Reco 1Reco=a1+be-ktSimple logistic function (Richards, 1959;Rodeghiero and Cescatti, 2005; Schäfer, 2012).Reco 2Reco=aebtTwo-parameter exponential function(Schneider et al., 2009; Schäfer, 2012)
To estimate annual Reco fluxes, we used air and soil temperature as
explanatory variables and tested two different models (Table 1), using all
respiratory flux data of the study year for each plot. Comparing the
qualifying parameter Radjusted2 of the model results showed that
the respiratory fluxes were better explained by air than by soil temperature
and that Reco 1 model achieved better results than Reco 2 model. The annual
Reco fluxes of the industrial extraction site were estimated using linear
regression, as both models resulted in low Radjusted2 values
ranging between 0.03 and 0.33.
Methane fluxes did not show a soil temperature or water table depth
dependency. Annual fluxes were thus not modeled with the classical model of
Saarnio et al. (1997) as intended, but calculated with linear interpolation,
as done for the annual N2O flux rates.
Calculation of the GHG budget
For the calculation of the GHG budget, the fluxes of CH4 and N2O
were converted in CO2 equivalent according to their global warming
potentials on a 100-year timescale including climate–carbon feedbacks:
CH4= 34 and N2O = 298 (IPCC, 2013). The C loss due to peat
mining was also estimated in CO2 equivalent. The calculation was based
on the amount of peat harvested in the study year relative to bulk density
and C content of the peat, measured within the upper three soil horizons.
ResultsWater table
The four sites displayed differences in water table depth throughout the
year (Fig. 1). The Sphagnum site had the highest water table followed by the heath
and the purple moor grass site with an annual mean for 2011 of 1.2,
-0.8 and -2.7 cm, respectively. The lowest water table was recorded at
the drained active industrial extraction site with a minimum of 55 cm under
the soil surface and a mean of -23 cm for a 12-month period from April 2011 to March 2012.
The industrial extraction site faces a great range of water
table amplitude. In contrast, the variation in the water table depth of the
three vegetated sites was much less, but soil surfaces were often inundated
especially in the winter half of the year.
Mean daily water table and daily precipitation at four
different bog sites over the whole measurement period from July 2010 until
April 2012. No data available for the industrial extraction site before
April 2011. Negative values indicate water levels below soil surface and
positive values indicate inundation.
Vegetation
The vegetation analysis showed clear differences in species coverage between
the three defined sites (Table 2). The heath site was dominated by
ericaceous shrubs (49 % cumulative coverage of Erica tetralix, Calluna vulgaris, Vaccinium oxycoccos and
Andromeda polifolia). The Sphagnum site had 99 % coverage of Sphagnum mosses (mainly S. cuspidatum and S. fimbriatum) and the purple
moor grass site was dominated by the perennial deciduous grass Molinia caerulea (67 %
coverage). Sites additionally differed in the number of vascular plant
species decreasing in the following order: heath > Sphagnum > purple moor grass with in total nine, eight and six species, respectively.
Average coverage of vascular plant species according to Londo (1976) at the three rewetted bog sites in 2011. Coverage of the moss layer
and bare peat areas were estimated using pointintercept sampling (Jonasson,
1988). Values refer to means of 7–8 replicates ±SD.
Numbers in bold print indicate eponymous species for site denotation.
Sphagnum species were mainly S. cuspidatum and S. fimbriatum. A dash indicates that the species was not found.
The results of the modeled Reco (Fig. 2) and the modeled GPP were summed up
resulting in modeled net ecosystem exchange (NEE; Fig. 3). The temporal
dynamics between the sites were quite similar for the heath and the
Sphagnum site. However, the Sphagnum site displayed some more pronounced peaks in
photosynthesis. The purple moor grass site had a steep increase in
photosynthesis in May and was very productive in summertime, while
photosynthesis at the two other vegetated sites began earlier and with a
slower increase. Before the steep increase in photosynthesis at the purple
moor grass site, respiratory fluxes predominated and increased from March
until May. The industrial extraction site was characterized by comparatively
low respiratory fluxes.
Measured ecosystem respiration (Reco) as a function of air
temperature. Points indicate fluxes ±standard error of the flux
calculation. The replicate plots of each site are shown in different colors.
Lines represent the fits to raw data using Reco 1 model. For the industrial
extraction site a linear regression was used. The goodness-of-fit is
indicated (Radjusted2).
CO2 budget
The annual net ecosystem exchange (NEE) in 2011 was positive at all
vegetated sites as well as at the industrial extraction site (Fig. 4),
meaning a net CO2 release from the ecosystem. The highest values of NEE
were determined at the active industrial extraction site where ecosystem
respiration predominates (730 ± 67 g m2 yr-1), followed by
heath, purple moor grass and Sphagnum with 308 ± 386,
247.36 ± 330.29 and 59.56 ± 142.58 g m2 yr-1, respectively. These differences in NEE between
the sites were closely approximating significance (ANOVA, p= 0.055).
Model parameters for gross primary production (GPP 1 model) in
2011: (a) maximum photosynthetic activity (Pmax) and (b) values for the initial
light response efficiency (α). Points indicate means of four
replicate plots ±SD. (c) Mean modeled net ecosystem
exchange (NEE) in 2011. Positive values represent emissions and negative
values uptakes.
Annual CO2 budgets of four different bog sites in 2011. GPP:
gross primary production, Reco: ecosystem respiration and NEE: net ecosystem
exchange. Values represent means of the model results of four replicates
(plots) at each site ±SD. Positive values represent
emissions and negative values uptakes.
CH4 and N2O fluxes
CH4 emissions showed no seasonal trend and were not dependent on water
level or soil temperature (data not shown). However, CH4 fluxes
differed significantly among sites (ANOVA, p < 0.00): CH4
fluxes of the industrial extraction site ranged around zero, while the
vegetated sites acted as CH4 sources. CH4 emissions increased in
the following order (Table 3): heath < Sphagnum < purple moor grass.
N2O fluxes were significantly different between the industrial
extraction site and the vegetated sites (ANOVA, p= 0.002). Significant
emission of N2O were only measured at the industrial extraction site at
summertime, resulting in mean annual emissions of 0.0165 ± 0.0086 µg m-2 s-1. N2O fluxes of the rewetted
sites were very low, ranging around zero and varying between small uptakes
and emissions throughout the whole year.
Mean annual CH4 and N2O fluxes ±SD
at four different bog sites. Values refer to means of four replicates ±SD.
CH4 and N2O fluxes were calculated into CO2 equivalent
(CO2-eq) and summed with the CO2 fluxes as a GHG budget (Table 4). We found that all sites were significant sources of greenhouse gases.
The GHG balance of the rewetted sites was dominated by CH4 emissions, accounting for 88–98 % of the total GHG CO2-eq
emissions. The portion of N2O was negligible there, while it played
with about 21 % of the GHG balance a significant role at the industrial
extraction site (1.55 ± 0.81 t CO2-eq ha-1 yr-1). The uncertainty of the
GHG budget was high due to high spatial variability between the replicates.
Nevertheless, differences between the sites were statistically different
(p < 0.05, Kruskal–Wallis one way ANOVA). The highest GHG emissions
were detected at the purple moor grass site (53.05 ± 35.72 t CO2-eq ha-1 yr-1), whereas the lowest GHG
emissions were present at the industrial extraction site (8.9 ± 1.1 t CO2-eq ha-1 yr-1). However, the GHG balance
of the industrial extraction site is not complete without considering the
amount of mined peat. Including these C losses and assuming their release
as CO2, the GHG balance of the industrial extraction site was
considerably higher than those of the vegetated sites (123 ± 7 t CO2-eq ha-1 yr-1).
Discussion
The visual differences in plant communities between the three sites were
confirmed by analyses of species composition and coverage. Additionally,
different life forms dominated, emphasizing diverse functions and
adaptations to the ecosystem: evergreen, ericaceous shrubs (heath site),
Sphagnum mosses with sedges (Sphagnum site) and nearly monospecific stands of the perennial
deciduous grass M. caerulea (purple moor grass site). We assume the differences in
vegetation composition to be a result of the different water table levels.
As the lowest water table position was detected at the purple moor grass
site, the combination of both the relatively drier surface and the
appearance of the non-typical bog plant resulted in significant differences
in GHG fluxes compared to the other two vegetated sites. Here, the highest
CH4 fluxes were measured, as well as the highest GPP and highest
maximum photosynthetic activity. The fast growth of M. caerulea leaves and their autumn
senescence can explain the annual dynamics of the modeled NEE at the purple
moor grass site. This grass performs no photosynthesis in wintertime
except at green basal internodes (Jefferies, 1915). Thus, photosynthesis
steeply increases with springtime growth and abruptly stops with the
dieback of its leaves in fall. By contrast, photosynthesis was detected in
wintertime at the Sphagnum and the heath site as they were dominated by evergreen
plants.
Greenhouse gas (GHG) budgets of four bog sites. The GHG budget of
the active industrial extraction site is composed of measured fluxes on site
and the C loss due to peat mining in 2011, assuming its emission as
CO2. Values represent means of four (CO2 and CH4) or three
(N2O) replicates ±SD.
All three vegetation communities established at the restored study site
turned out to be sources of CO2. Thus, restoration failed to
re-establish the CO2 sink function after 30 years, in contrast to what
was hypothesized. However, the Sphagnum site had the lowest CO2 emissions
ranges and, considering the uncertainty related to the spatial variability,
close to a CO2 neutral status. These results are concordant with the
findings of Samaritani et al. (2010) who showed that a Sphagnum-dominated European
mid-latitude cut-over bog, was a CO2 source 29 years after rewetting,
but older regenerating sites (42 and 51 years) were taking up
CO2. The magnitude of the CO2 balances is consistent with the
analyses of 53 studies evaluated by Couwenberg et al. (2008), showing that
temperate European peatlands with similar water table positions emit less
than 3 t CO2 ha-1 yr-1 (utilization and restoration
status not mentioned). The net CO2 emissions measured at the industrial
extraction site were higher compared to the restored sites, as the soil is
deeply drained and only negligible photosynthesis due to algae growth takes
place. The annual CO2 emission of 7.3 t CO2 ha-1 yr-1 is in the range of 11 reported studies from European drained peat
cut sites with a median of 2 and a maximum of
13 t CO2 ha-1 yr-1 (Drösler et al., 2008). Rewetting of
the industrial extraction site provides a CO2 mitigation potential
between 422 and 671 g CO2 m-2 yr-1 (peat extraction not
included) calculated from the CO2 budgets of the three rewetted sites.
This is in total 295 to 470 t CO2 yr-1 for the whole site
(70 ha), depending on which of the three vegetation communities establishes.
However, restoring a cut-over peatland requires special techniques as most
Sphagnum species cannot re-establish spontaneously and re-vegetation is generally
slow due to no viable propagules and unfavorable conditions for plant growth
and seed germination (Quinty and Rochefort, 2003; Triisberg et al., 2011;
D'Astous et al., 2013).
The GHG budget of the restored site is dominated by CH4 fluxes (up to
98 %). While the magnitude of the modeled CO2 balances is similar to
data reported in other studies as described above, the magnitude of the
CH4 fluxes is very high. To our knowledge, the maximum rates reported
for European bogs are: approximately
17 t CO2-eq ha-1 yr-1, which was measured in the
Bavarian Alpine foreland on a natural bog site with
Sphagnum–Scheuchzeria palustris hollows (Drösler, 2005) and
24 t CO2-eq ha-1 yr-1 measured in
E.vaginatum-colonized infilled ditches of a blanked bog in Wales (Cooper et al., 2014).
In the present study, we found annual CH4 fluxes of 22, 34 and
51 t CO2-eq ha-1 yr-1. We hypothesize that the large
fluctuations in water level, which are typical for degraded peatlands
(Schouwenaars, 1993; Tuittila et al., 1999), might explain these high
CH4 emissions as fresh belowground or aboveground plant litter is
inundated episodically and undergoes anoxic fermentation (Augustin and
Joosten, 2007; Paul and Alewell, 2013). This applies especially for the
purple moor grass site although it features the lowest water level of the
three vegetated sites: here, CH4 fluxes were higher than at the
Sphagnum and the heath site (2.3 and 1.5 times, respectively). We suggest that this
is due to the large amounts of litter produced by M. caerulea, which is easily
decomposable in comparison to other plants present at the study site (van
Breemen, 1998). Like we found here, Bohdalkova et al. (2013) observed the
highest CH4 emissions at the site with the lowest water level and
suggested this to be the result of the invasion of easily degradable
vascular plants. Likewise, Hahn-Schöfl et al. (2011) explained extremely
high CH4 emissions in a German fen with the inundation of reed canary
grass (Phalaris arundinacea). Another likely explanation for the huge CH4 fluxes from the
purple moor grass site is the presence of aerenchyma (Jaiswal et al., 2000),
through which M. caerulea can act as gas conduit. It hence allows CH4 to bypass the
oxidized surface soil, and consequently reduces CH4 oxidation. Although
the flux of CH4 through the plant tissues is lower in M. caerulea compared to
E. angustifolium (Bhullar et al., 2013), the chimney effect might be increased by the high
cover of this grass species at the purple moor grass site. Thus, it was
shown that in some plots CH4 emissions coincided with high coverage of
aerenchymatous plant species. The composition of plant communities is
therefore the most important driver for CH4 fluxes at the study sites,
as reported by other authors (Samaritani et al., 2010; Couwenberg et al.,
2011; Bohdalkova et al., 2013).
Irrespective of the abundance of shunt species, CH4 emissions from the
rewetted study site are generally very high. We therefore assume that the
site's land use history determines the fundamental conditions for methane production.
As mentioned above, the upper vegetated soil layer (Bunkerde) was used for filling
up the mined area. We found differently decomposed soil material in layers
or lens-shaped zones in the soil profiles of all rewetted sites (data not
shown). We hypothesize that the peat material, which was exposed to
oxygen, enhances the methanogenic degradation of the organic matter by the
initiated decay process (priming process). Additionally, labile organic
matter in the form of aboveground plant biomass of the primary vegetation
was buried and thus might serve as substrate for CH4 production. A
similar effect was found by Cooper et al. (2014), who identified former
drainage ditches blocked with heather bales and colonized by aerenchymatous
species as hotspots for CH4 emissions. However, it must be kept in
mind that the heather bales in Cooper et al. (2014) were recently
incorporated, while in the present study the incorporation of fresh plant
material into the soil is several decades past and can therefore only play a
minor role in explaining the high methane emissions found here.
These findings are important for a large area of exploited peatlands,
because the peat extraction method applied here was long the traditional way of
peat cutting (Koster and Favier, 2005). It is therefore essential that more attention is paid
to monitoring and management of areas with a hand-cutting history. In terms
of CH4 mitigation and climate protection, an active reduction of
M. caerulea stands, by for example sod cutting or grazing, could help to establish target
species communities instead (Jacquemart et al., 2003; Reid et al., 2009;
Keddy, 2010; Meuser, 2012), and hence reduce CH4 emissions at the study
site. Avoiding water table fluctuations is additionally important to impede
inundation of fresh plant litter as a substrate for methanogenesis
(Hahn-Schöfl et al., 2011). The installation of overflows as management
measure could stabilize the water table level (Maitland and Morgan, 1997;
Quinty and Rochefort, 2003).
Conclusions
We found large differences in GHG fluxes between the three vegetation
communities. Differentiation between plant communities for estimation and
upscaling of GHG budgets is therefore imperative. The particular land use
history of this degenerated peat site, in combination with inundation due to
water tables fluctuations and a high cover of aerenchymatous plants seem to
form favorable conditions for both, high methanogenesis and efficient soil
atmosphere CH4 transport. This results in extremely high CH4
emissions and high GHG budgets in comparison to other bog ecosystems in
Europe. The annual GHG emissions are at least twice as high as the rates of
the industrial extraction site, although 30 years passed since rewetting. We
therefore state that the initial increase in CH4 emissions due to
rewetting was not limited to a short-term period at our study site, as
described in literature. We raise here the question if the emergence of
high methane fluxes is, on the contrary, permanent in severely damaged
rewetted peatlands, as it is not possible to fully reestablish ecosystem
functions.
Acknowledgements
We wish to acknowledge the support of the peat company manager in the
Himmelmoor, Klaus Czerwonka, and the committed field and lab assistance of
Birgit Grabellus, Mathias Schwarzer, Stephanie Langer and Kira Kalinski. We
appreciate the technical support from Christian Wille and Peter Schreiber
during the planning and preparation stages of this study. This work was
carried out within the research project KLIMZUG-Nord, which was funded by
the German Federal Ministry of Education and Research (grant no.
01LR0805D). L. Kutzbach was supported through the Cluster of Excellence
“CliSAP” (EXC177) at the University of Hamburg funded through the German
Research Foundation.
Edited by: X. Wang
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