Introduction
Peatlands worldwide contain 500–600 Pg carbon (C) (Gorham, 1991; Yu et al.,
2010; Page et al., 2011) that has been fixed from the atmosphere. Wet, anoxic
conditions constrain the decomposition of organic matter and thus enable the
accumulation of carbon as peat. Since wet conditions are a prerequisite for
peat accumulation, drying of peatlands through drainage or climate change
has been assumed to result in the release of sequestered carbon back to the
atmosphere.
The effect of drainage of forested peatlands on carbon stocks has been under
debate at least since the 1980s, when large carbon dioxide (CO2)
emissions were reported from drained peatlands in Finland (Silvola, 1986).
Studies from agricultural peat soils show that carbon stocks are usually
greatly diminished under efficient drainage (e.g. Oleszczuk et al., 2008;
Tiemeyer et al., 2016), with some exceptions (e.g. Merbold et al., 2009;
Fleischer et al., 2016). Similar C loss has often been assumed for all
drained peatlands, including those drained for forestry. However, in some
peatlands, soil has been reported to sequester carbon even after drainage
(e.g. Lohila et al., 2011; Turetsky et al., 2011). Minkkinen and
Laine (1998a) and Minkkinen et al. (1999) showed, based on peat C stock
measurements, that many nutrient-poor peatland sites remained C sinks after
drainage. Later, Ojanen et al. (2013) showed the same relation with site type
and soil C balance, nutrient-poor ones being sinks and fertile ones sources.
The continued C sequestration on relatively nutrient-poor sites has been
related to the increased litter production and changes in litter quality
(Laiho et al., 2003; Straková et al., 2012) vs. only moderately increased
decomposition of old peat (Minkkinen and Laine, 1998a). This view has,
however, still been challenged (e.g. Simola et al., 2012), and, for example
according to IPCC guidance, drained peatlands are assumed to be C sources
(Drösler et al., 2014).
Climate warming, in addition to drainage, has been predicted to increase C
loss from peatlands because of increased soil temperatures and droughts
(e.g. Moore, 2002). In warmer and drier conditions the decomposition of soil
organic matter (SOM) is expected to increase, although increased primary
production and a possible long-term shift towards more shrub and tree
dominated vegetation communities (Laiho et al., 2003; Tahvanainen, 2011;
Straková et al., 2010, 2012) may partly compensate for the increased
decomposition rates (Flanagan and Syed, 2011). The reported impacts of
droughts on ecosystem CO2 fluxes are, however, variable. Droughts have
been shown to decrease photosynthesis and increase ecosystem respiration,
especially on wet and nutrient-rich fens (Bubier et al., 2003; Adkinson et al., 2011), while on naturally drier bogs, the effects may be reversed
(Sulman et al., 2010). CO2 emissions from the decomposition of peat are
often shown to increase linearly with water level drawdown (e.g. Silvola et al., 1996; Jauhiainen 2012) but there are also indications of an optimum
water table depth in boreal peatlands below which soil respiration would not
further increase (Mäkiranta et al., 2009). Thus, in some cases
decomposition of SOM might even decrease during droughts (Sulman et al., 2010).
Our study site, the forestry-drained peatland Kalevansuo in southern Finland,
was earlier reported to be a strong C sink in terms of net ecosystem
CO2 exchange (NEE) during 2004–2005 (Lohila et al., 2011). The
magnitude of the C sink was remarkably higher than the estimated tree stand
C pool increment, which led us to the conclusion that the soil must also act
as a C sink. Whether this was just a single-year result or whether it holds
through several years with varying weather conditions will be investigated
in this paper.
The aims of this study were to estimate the full C balance of a drained
peatland forest ecosystem over 4 years, and to analyse the impact of
seasonal drought on the C fluxes. We measured the C pools in the ecosystem
(peat soil, vegetation above and below ground), CO2 fluxes between the
ecosystem and atmosphere, namely NEE, gross primary production (GPP),
ecosystem respiration (RECO) and forest floor respiration (RFF)
divided to component fluxes (peat, litter, roots, ground vegetation) and the
C flux in litter (L). We complemented the results with measurements of
methane (CH4) fluxes and peat subsidence.
Material and methods
Site
The measurements were carried out in a drained peatland forest, Kalevansuo,
in southern Finland (60∘38′49′′ N, 24∘21′23′′ E,
123 m.a.s.l.). The peatland was drained by digging open ditches in 1971.
Kalevansuo is a typical dwarf-shrub type peatland forest according
to the classification of Vasander and Laine (2008). The dominant tree species
is Scots pine (Pinus sylvestris L.), comprising 98 % of the
stand volume and 53 % of the stem number. Pubescent birch (Betula pubescens Roth) and Norway spruce (Picea abies L.) form the sparse
understorey.
The site has been naturally forested since long before drainage as evidenced by
very old scattered stumps of Scots pine found in all parts of the site. Tree
ages of the present Scots pine stand, as determined in 2005 from increment
cores of sample trees (n=7), varied from 67 to 179 years with an average of
120 years. In 2008, the stand stem volume was 130 m3 ha-1, basal
area 18 m2 ha-1, dominant height 16 m and stem number
1670 ha-1. Microtopographically the site is rather even (lawn level),
with small hummocks covering about 25 % of the area. A more detailed
description of the stand is given by Lohila et al. (2011).
Following drainage, the tree stand has grown bigger and the coverage of mire
species has decreased and forest species increased in the bottom and field
layers. However, many mire species are still present at the peatland. Forest
floor vegetation consists mainly of forest and mire dwarf shrubs
(Vaccinium myrtillus L., V. vitis idaea L., V. uliginosum L., Ledum palustre L.), with patches of cottongrass
(Eriophorum vaginatum L.) and cloudberry (Rubus chamaemorus
L.). The dominant moss species are Pleurozium schreberi (Brid.)
Mitt., covering 48 % of the study area, and Dicranum polysetum (37%), but Sphagnum mosses such as S. angustifolium
(Russ.) C. Jens., S. russowii Warnst., and S. magellanicum
Brid. are also abundant in moist patches (coverage 15 %; Badorek et al.,
2011). The ditches have not been cleaned since digging in 1971 and are
nowadays totally vegetated, mainly with Sphagnum riparium (and
S. russowii, S. angustifolium), some cottongrass (Eriophorum vaginatum) and sporadic dwarf shrubs (Ledum palustre).
Peat depth, measured from the 33 sample plots, varies from 1.3 to 3.0 m,
the average being 2.2 m. Mean peat bulk density is 94 kg m-3 in the
0–20 cm layer. The peat accumulated prior to drainage at the study area is
composed mainly of the remains of Sphagna (Sphagnum fuscum, S. magellanicum), Ericaceous shrubs and cottongrass (Eriophorum vaginatum) (Mathijssen et al., 2017). After drainage the remains of forest
mosses and woody roots have increased their share in surface peat. Drainage
has increased surface peat oxidation which is seen as a shallow layer of more
decomposed peat about 10–20 cm below the surface. Remains of several forest
fires are also present, especially in the surface layers 30–50 cm, where a
charcoal layer is clearly visible, but also in the deeper layers from 70 to
180 cm.
The mean air temperature in 2005–2008 was 15.3 ∘C during summer
months (June–August) and -3.8 ∘C during winter
(December–March). The annual mean air temperature was 5.1 ∘C and
temperature sum (> 5 ∘C) 1356 degree-days. Annual average
precipitation was 722 mm and maximum wintertime snow depth 20–60 cm.
Measurement setup
The site was set up for C flux measurements in June–August 2004. The
micrometeorological eddy covariance (EC) measurements were conducted from a mast, erected in
August 2004, in the centre of the peatland at a 200–250 m distance from an
upland forest in the north-west and from a small lake in the south-west. To the
north-east the homogenous fetch was longer, about 600 m. The EC footprint
was thus concentrated to the fairly homogenous peatland pine forest with at
least 200 m radius (Lohila et al., 2011).
The chamber measurements of CO2 and CH4 fluxes were
conducted at four plots, located 50–100 m from the mast. The measurement
collars were inserted and litter and plants removed from the treated collar
plots in June 2004. As every plot consisted of 16 measurement points
(collars), the whole setup contained 4 × 16, i.e. 64 measurement
points. In addition, CH4 fluxes from ditches were measured in 2011
at four points on two parallel ditches located on both sides of the mast.
The depth of the water table (WT) was manually measured from two perforated
plastic pipes at each plot, along with chamber measurements. The WT was also
continuously recorded close to the EC mast by a PDCR 830 (Druck Messtechnik
GmbH, in 2004–2006), and a Hobo U20-001-01 (Onset Computer Corporation, MA,
USA, in 2007–2009). Soil temperatures were recorded with temperature loggers
(i-Button DS1921G, Maxim Integrated Products) from the depths of 5 cm
(T5) and 30 cm (T30) below soil surface at intervals of 1–3 h.
In 2005–2006, T5 was recorded from every measurement point and T30
from 16 points (four per plot). In 2007–2008, T5 recordings were taken from
two points per plot and T30 recordings from two points in total.
The tree stand, ground vegetation and soil properties were measured on 33
plots located evenly along 8 radial transects extending 160 m from the
mast (the centre plot). Four transects with plots spaced at 20, 60, 100 and
140 m distances from the mast were alternated with four other transects with
plots spaced at 40, 80, 120 and 160 m distances from the mast. The area of
each plot was 200 m2.
Measurements
Ecosystem–atmosphere exchange of CO2
The turbulent fluxes of CO2, water vapour (H2O), sensible
heat and momentum were measured with the eddy covariance technique on top of
the 21.5 m telescopic mast (17.5 m from April 2005 to April 2006).
Supporting meteorological measurements included, for example, relative humidity (RH),
photosynthetic photon flux density (PPFD), air and soil temperatures, and soil
moisture (see Lohila et al., 2011, for closer description). Measurements were
carried out from August 2004 to March 2009. Here we report results for the
full years 2005–2008.
We used an SATI-3SX (Applied Technologies, Inc.) sonic anemometer/thermometer
from 2004 to November 2006, after which a METEK USA-1 (METEK GmbH, Elmshorn,
Germany) was used. The atmospheric concentrations of CO2 and
H2O were measured with an LI-7000 (LI-COR, Inc.) analyser. This
instrument was calibrated bimonthly to monthly with two known CO2
concentrations [CO2] (0 and 421 ppm). CO2-free synthetic
dry air was used as a reference gas. The heated inlet tube (3.1 mm Bevaline
IV) for the LI-7000 was 17 m long, and a flow rate of 6 L min-1 was used.
The signals were sampled at a frequency of 10 Hz, and the turbulent fluxes
were calculated on-line as 30 min averages, applying standard EC procedures. The effect of density
fluctuations related to the water vapour flux (Webb et al., 1980) was
included in the calculations, and the fluxes were corrected for systematic
losses using the transfer function method of Moore (1986), including the
losses due to autoregressive running mean filtering and the imperfect
high-frequency response of the measurement system. Details of the flux
calculation and correction procedures can be found in Pihlatie et al. (2010)
and Lohila et al. (2011).
To estimate the storage fluxes of CO2, the mean [CO2]
observed at a height of 4 m with a LI-820 CO2 analyser and the
[CO2] measured at the top of the mast were used. The storage term
was calculated with the central difference method from the mean concentration
during the subsequent and preceding 30 min periods and added to the measured
turbulent flux. Hereafter NEE refers to the sum of turbulent and storage
fluxes. In this paper, we use the convention that a positive value of NEE
indicates a flux from the ecosystem to the atmosphere.
Forest floor CO2 efflux
CO2 efflux from the forest floor was measured with an opaque closed
steady state chamber (diameter 31.5 cm, height 14.9 cm) attached to a
portable infrared gas analyser (EGM-4, PP-Systems, Hitchin, UK; NSF11 in
Pumpanen et al., 2004). Chamber closure time was 81 s. Measuring points were
delimited with permanent collars and had four different treatments including
the following respiration components: (A) peat soil (including cut roots),
(B) A + above-ground litter, (C) B + living roots and (D)
C + ground vegetation. In plots with vascular plants, extra collars of
5–10 cm height were used to fit the plants inside the chamber. The chamber
volume was corrected accordingly.
In order to exclude autotrophic respiration, treatments A and B involved
trenching with 30 cm deep collars and removing above ground parts of living
vegetation by repeated clippings every time before measurements if new plant
growth had emerged. From treatment A, the above-ground litter was also
removed every time before measurements. From treatment C, only the above-ground parts of plants were removed and treatment D was left intact. Collar
depth in treatments C and D was only 2–3 cm to minimise disturbance to
roots. Treatment D (RD) thus includes all respiration components
of forest floor respiration (RFF) and treatment A respiration
from peat soil only (RPEAT). Respiration from treatment B
(RB) equals heterotrophic respiration (RHET), and
autotrophic respiration (RAUT) is calculated as
RD-RB. Autotrophic respiration of above-ground
vegetation (RGV) is defined as RD-RC, root
respiration (RROOT) equals RC-RB and
RLITTER equals RB-RA.
CO2 fluxes from treatments A and D were measured during the whole
period 2005–2008, while treatments B and C were only measured from 2005 to 2006.
Forest floor and ditch CH4 fluxes
Soil CH4 fluxes from the strips between ditches were measured with
static chambers from the D points and reported by Lohila et al. (2011). To
complement the CH4 flux estimate for the whole area, fluxes from
ditches were measured with the same equipment and methods as earlier. Fluxes
were measured from four points on two parallel ditches on the both sides of
the mast, altogether 7 times between 28 June and 8 December 2011. The annual
flux was estimated as 365 × daily mean flux.
Organic carbon pools and fluxes
The carbon stock in peat, and biomasses and litter production of the tree
stand and ground vegetation, were measured to estimate organic carbon pools
and fluxes in the peatland. Peat C stock was estimated based on average peat
layer thickness on the tree stand transects (Lohila et al., 2011) and average
carbon density in peat (Mathijssen et al., 2017). Tree stand properties were
measured in spring 2005 and autumn 2008. In 2005, the sample trees were cored
to estimate diameter increment during the previous 5 years. Tree stand
biomasses and C pools for years 2000, 2005 and 2008 were then estimated from
these data using models of Repola (2008, 2009) and Laiho and Finér (1996)
for pine below-ground biomasses (root d > 1 cm), as described in
detail by Ojanen et al. (2012). In all biomass C stock and flux calculations,
C content of 50 % was assumed.
Above-ground biomass of ground vegetation vascular plants was sampled along
the tree stand transects (n plots = 39), from an area of
0.25 m2 per plot. Moss samples
(n=64) were collected from the same sites using corers with a diameter of
93 or 125 mm. In the lab, the dead part of the moss was cut and removed,
based on ocular assessment (colour change of the moss). The samples were
separated by species, and dry mass (105 ∘C) was determined for each
sample.
The biomass of roots (and rhizomes of shrubs) were determined by taking a
soil sample of 15×15×20 cm
(width × length × depth) along the tree stand transects,
adjacent to the mid-points of the tree sample plots (n=32). In the
laboratory all roots were carefully separated from peat, divided according to
species or functional groups (pine, spruce, birch, shrubs, grasses and herbs)
and diameter (below and over 2 mm), dried in 105 ∘C and weighed.
According to Bhuiyan et al. (2016), 15 % of the fine roots in Kalevansuo
are located deeper than 20 cm. The biomasses estimated here were corrected
accordingly.
C flux in above-ground litter was estimated with 14 litter traps
(20 × 20 cm) per chamber plot (i.e. altogether 56 traps). Litter
was collected 2–3 times per year, separated by species, dried in
105 ∘C and weighed. As moss litter is not captured by litter traps,
moss litter production was estimated by harvesting moss biomass production
over 2 and 5 years (Ojanen et al., 2012). As the whole moss biomass
eventually dies and forms litter on site, annual moss biomass growth equals
annual litter production.
Coarse root (> 2 mm) litter production was estimated as
biomass × turnover rate (0.12 for pine, 0.08 for shrub rhizomes;
Finer and Laine, 1998). Fine root litter production was estimated with
root-ingrowth cores by Bhuiyan et al. (2016). Sixty cores (diameter 3 cm,
length 50 cm) filled with Sphagnum peat were installed into soil in
October 2009, and 20 cores were collected every year for 3 years. The
fine root production rate was calculated as the average fine root mass
(live + dead) in the cores divided by incubation years (average for 2nd
and 3rd years).
Change in peat layer thickness
To survey the changes in peat layer thickness, caused by compaction and
decomposition of soil organic matter, litter production and moss height
growth, soil surface around the mast was levelled in 2004, 2011 and 2014. In
the beginning of measurements in 2004, a 20 mm thick steel rod was hammered
through the peat layer firmly to the subsoil, serving as a stable benchmark.
The soil surface at the undisturbed chamber measurement points (D collars),
was repeatedly levelled in relation to the benchmark. A manual levelling
instrument with a levelling rod was used and the readings were recorded with
the precision of ±0.5 cm.
Gas flux calculations
NEE
The NEE data obtained from the EC measurements were screened as described by
Lohila et al. (2011). In short, screening criteria were applied to remove
spikes in the 10 Hz anemometer data and to discard poor-quality 30 min data.
For the latter, the criteria were based on the expected range of the mean
[CO2] and air temperature (from the sonic anemometer), and of the
variances of [CO2], vertical wind speed and air temperature. In
addition, a cumulative flux footprint of 70 % was required, and a
threshold of 0.1 m s-1 was set to the friction velocity (Lohila et
al., 2011). The procedures of gap-filling of the EC flux data and
partitioning of NEE to the GPP and RECO components are described
in Appendix A. The estimation of uncertainties in annual NEE is described in
Appendix B.
Forest floor CO2 efflux
CO2 efflux from the forest floor is a result of heterotrophic and
autotrophic processes from different layers (vegetation and soil), which have
different temperature dynamics. Therefore an additive, layerwise model was
used, in which soil temperatures T5 and T30 predict fluxes from
different layers, with different temperature dynamics. An Arrhenius-type
function (Lloyd and Taylor, 1994) was fitted to the measured CO2
efflux (g CO2 m-2 h-1) from the forest floor:
CO2efflux=RREF5expE051TREF-T0-1T5-T0+RREF30expE0301TREF-T0-1T30-T0,
where RREF5 and RREF30 are respirations at reference
temperatures (TREF= 10 ∘C), and E05 and E030
describe temperature sensitivities of respiration in 5 and 30 cm peat
depths, respectively. T0=-46.02 ∘C is a constant.
Parameter values were estimated separately for different treatments (A–D)
representing different components of RFF, the four gas measurement
plots and two groups of years (2005–2006 and 2007–2008; Appendix C), as
the decomposability of soil organic matter changes in time at A collars. The WT
was also tested as an explanatory variable, but as it predicted the temporal
flux variation poorly, it was not included in the final models. The models
were used with measured soil temperature data to simulate the temporal
dynamics and annual fluxes of different flux components.
Modelling of the tree stand CO2 fluxes
To analyse the contribution of the tree stand (above ground) to the ecosystem
CO2 exchange, we used the GPP and shoot respiration (R) models in
the Stand Photosynthesis Program (SPP). SPP predicts canopy light interception,
photosynthesis and shoot respiration in half-hourly time steps
(Mäkelä et al., 2006). PPFD, air CO2 concentration, air
temperature and relative air humidity measured at the site were used as
inputs for SPP. The photosynthesis model used was OPAC (Mäkelä et
al., 2006). Tree stand was described as three size classes (Ojanen et al.,
2012), foliar masses for each class were estimated using the models of
Repola (2009), and these were converted to leaf area index with specific leaf
area of 11 m2 kg-1 (Luoma, 1997). Stem respiration was estimated
with the model of Zha et al. (2004).
Results
Meteorological conditions
Of the studied years, 2008 was the warmest, especially during the winter
months January–March, which were almost snowless. It was also the rainiest
year. The summer (June to August) of 2008 was significantly cooler, but
otherwise similar to the other summers. In contrast, the year 2006 was
exceptionally dry from January until the end of September, including a severe
drought during the growing season. In summer 2006, air temperature and PPFD
were higher than in other years, whereas relative humidity and the water table
were lower (Table 1, Figs. 1 and 2). The dry and warm growing season 2006 was
preceded by a cold winter, which is why soil surface temperatures (T5)
were much below average in the spring, and down at 30 cm stayed below
average until September, i.e. for almost the whole growing season (Fig. 2).
In September–October the deeper peat layers finally warmed up and stayed
warmer than average for the rest of the year.
Meteorological parameters for the full years and summer months,
June–August. T = mean air temperature, P = precipitation sum,
PPFD = mean daily sum of photosynthetic photon flux density,
RH = mean relative humidity, VPD = mean vapour pressure deficit in
the afternoon (12:00–16:00 LT).
Year
June–August
Year
T
P
T
P
PPFD
RH
VPD
(∘ C)
(mm)
(∘C)
(mm)
(mol m-2
(%)
(kPa)
d-1)
2005
4.7
725
15.1
285
35.0
76.0
0.89
2006
5.1
600
16.5
95
37.6
66.9
1.35
2007
4.9
724
15.1
234
35.1
75.4
0.93
2008
5.6
839
14.3
237
31.3
73.2
0.91
Daily weather data: precipitation (average from nearby weather
stations), air (Tair) and soil temperatures at 5 cm (T5;
d collars) and 30 cm (T30) depths, and average water table (WT) level
at Kalevansuo peatland.
The WT typically fluctuated between -30 and -50 cm in a year, being on
average 42 cm below ground surface during the snow-free season
(April–November) and only about 5 cm higher during the winters
(December–March) (Figs. 1 and 2). The WT varied also spatially (mean range
between water-wells 24 cm), being deeper in the hummocks (-49 cm)
compared to the lawns (-35 cm). During the drought in 2006, the WT started
dropping down in July, reached -79 cm in the end of September, and rose
again after heavy rainfalls in the beginning of October. The average WT in
2006 was 10 cm deeper than in other years.
Weather variables by year and month, measured at Kalevansuo, except
precipitation, which is an average from nearby weather stations. Mean daily
air temperature (∘C) and temperature sum (> 5 ∘C
d.d.) at 2 m height, soil temperatures (∘C) at 5 and 30 cm depths,
mean daily PPFD (mol m-2 d-1) at 21.5 m. height, monthly
precipitation sum (mm month-1), water table depth (cm) and vapour
pressure deficit (kPa) at 21.5 m height in the afternoon (12:00–16:00 LT).
Ecosystem CO2 exchange
According to the EC flux measurements, the site acted as a CO2
source typically during the winter months (October–March) and a sink during
the growing season (April–September) (Figs. 3, 4a). The variation in NEE
during winter was small, ranging from about -0.1 to
0.1 mg m-2 s-1 (Fig. 3). While there were occasional warm
days with net CO2 uptake during the winter, the actual spring
recovery of photosynthesis seemed to occur typically in the beginning of
April, the only exception being the spring after the warm winter of
2006–2007, when the recovery started already in March. In summer
(June–August), the highest night-time CO2 emission values,
representing RECO, were on average 0.35 mg m-2 s-1,
and the highest day-time CO2 uptake typically fluctuated around
-0.75 mg m-2 s-1. Only in summer 2006 the amplitude in the
diurnal dynamics was smaller.
The site was a sink of CO2 in all years, NEE varying between -520
and -990 g CO2 m-2 a-1 (Table 2). The average NEE
for the 4 years was -860 g CO2 (i.e.
-234 g C m-2 a-1). With the exception of the dry year of
2006, the annual NEE was surprisingly similar in other years, varying from
-950 to -990 g CO2 m-2 a-1. The estimated
uncertainty in the annual budget, including the random measurement and
gap-filling error and the uncertainty in the high-frequency loss correction
and gap-filling of the gaps longer than 2 days, varied from 35 to
114 g m-2 a-1, corresponding to 3.6 to 22 % of the
respective annual balance (Appendix B).
Quality-controlled half-hourly NEE measured with the eddy
covariance method at Kalevansuo peatland in 2005–2008.
The drought during the spring and the growing season of 2006 was clearly
reflected in the CO2 exchange. The (gap-filled) NEE and GPP were
markedly less negative in June and July 2006, indicating lower CO2
uptake by photosynthesis compared to the other years (Fig. 4b, c).
However, in July and August RECO was also clearly suppressed
(Fig. 4d), thus decreasing the net loss of CO2 from the peatland
(NEE). In October 2006, GPP had fully recovered to the level of other years,
but RECO stayed at a slightly higher level during the rest of the
year, leading to clearly higher NEE during the last months of the year.
After the first week of June until the end of July 2006, there were only a
few days with accepted NEE observations (Fig. 3), so the results shown for
these months (Fig. 4) largely depend on gap-filling. However, the main
parameters of respiration and photosynthesis (RREF= respiration
at 10 ∘C, GPMAX= photosynthesis in optimal
light conditions; see Appendix A) indicate that both the photosynthetic
capacity and ecosystem respiration were reduced in the summer of 2006
(Fig. 5). The data coverage was considerably better in August, making it
possible to reliably study the impact of drought on NEE. In August 2006, both
RREF and GPMAX had values that were
significantly different from the other years (Fig. 5). While typically
RREF reached its maximum in August and decreased thereafter
(GPMAX having similar but opposite dynamics), in 2006
the trend was reversed and both RREF and
GPMAX increased towards October. This suggests that the
ecosystem was affected by the drought in August and September 2006 and slowly
recovering in October. Thus, the distinct decrease in the annual net
CO2 uptake in 2006 (Table 2) was likely to be caused by the GPP
decrease during the summertime, although RECO decreased during
the drought as well. In addition to the summer depression in net
CO2 uptake, the higher RECO in autumn months after the
drought and heavy rains in October (Fig. 2) furthermore increased the
difference to other years: the cumulative NEE in October-December in 2006 was
as high as 320 g CO2 m-2, while in other years it varied
from 130 to 190 g CO2 m-2.
(a) Gap-filled and partitioned daily NEE, GPP and
RECO at Kalevansuo 2005–2008. Full days with missing data shown
with dark blue (NEE_gap). (b) Monthly NEE; (c) monthly
GPP; and (d) monthly RECO.
EC-measured (and gap-filled and partitioned) annual net ecosystem
exchange (NEE ± error; Appendix B), gross primary production (GPP) and
ecosystem respiration (RECO) of the Kalevansuo peatland, in
comparison with the simulated tree stand GPP (GPPTREES), tree
stand above-ground respiration (RTREES_AG) and forest floor
respiration (RFF). Unit: g CO2 m-2 a-1.
EC measurements + gap-filling
Model simulations
+ flux partitioning
Year
NEE
GPP
RECO
1GPPTREES
1RTREES_AG
RTREES_AG
+2RFF
2005
-991±37
-3816
2821
-2311
1033
3345
2006
-516±114
-3231
2725
-2590
1160
3468
2007
-952±35
-4149
3207
-2530
1010
3122
2008
-970±35
-4023
3089
-2463
1007
3063
Mean
-857
-3805
2961
-2473
1053
3250
as C
-234
-1038
807
-674
287
886
1 SPP model (Mäkelä et al., 2006).
2 Eq. (1) (App. 3, treatment D).
Forest floor CO2 flux
The measured instantaneous CO2 fluxes from the forest floor
(RFF) varied between -0.02 and
1.80 g CO2 m-2 h-1 (Fig. 6), following the dynamics
in soil temperature. For the treatments A, B, C and D, the mean ± SD respiration fluxes in non-winter seasons
(April–November) 2005–2006 were 0.23 ± 0.11, 0.31 ± 0.13,
0.38 ± 0.22 and
0.42 ± 0.24 g CO2 m-2 h-1, respectively. In
winter (i.e. over a snowpack or frozen ground between December and March),
the mean fluxes were almost the same in the different treatments (0.022,
0.019, 0.022 and 0.035 g CO2 m-2 h-1 from A to D,
respectively).
The regression models with T5 and T30 as explanatory variables
(Eq. 2) explained 70 % (46–90 %) of the variation in the fluxes of
the entire dataset (Appendix C). Respiration rates at 10 ∘C
(RREF5 and RREF30) increased from A to D collars,
i.e. as respiration components were added, and decreased at A collars with
time since the beginning of the study (05–06 to 07–08).
The modelled annual respiration ranged during the first 2 years from
1233 g CO2 m-2 a-1 in A collars (RPEAT)
to 2312 g CO2 m-2 a-1 in D collars (RFF,
Table 3). During 2007–2008, RPEAT clearly decreased from the
previous years, to ca. 830 g CO2 m-2 a-1, whereas
RFF varied little between the studied years. In A collars the
decomposability of organic matter is likely gradually decreased when the
labile components are decomposed and the recalcitrant ones are enriched.
Also, as we had to remove a newly grown moss layer from A collars in
spring 2007 (inevitably with some soil organic matter attached), this
procedure probably decreased the proportion of labile components on the soil
surface.
Based on the modelled fluxes of the first 2 years, RHET
contributed 75 % and RAUT 25 % to the mean annual
RFF (Table 3). RPEAT comprised 53 % of the flux,
RLITTER 22 %, RROOT 16 % and RGV
8 %. The 4-year mean of RFF was
2197 g CO2 m-2 a-1, i.e. ca.
600 g C m-2 a-1. Using this mean value with the proportions
from 2005–2006, we get an estimate for RHET of
450 g C m-2 a-1 and RAUT
150 g C m-2 a-1.
Modelled annual forest floor CO2 effluxes
(mean ± SEM; g CO2 m-2 a-1) in the four
treatments at Kalevansuo peatland. A=peat, B=peat+litter, C=peat+litter+roots and D=peat+litter+roots+groundvegetation. SEM
is the standard error between the four plot means.
Year
A
B
C
D
2005
1233 ± 48
1745 ± 121
2118 ± 67
2312 ± 170
2006
1233 ± 48
1741 ± 108
2117 ± 76
2308 ± 179
2007
822 ± 40
No data
No data
2112 ± 141
2008
835 ± 49
No data
No data
2056 ± 143
In 2006, the main part of summertime (15 June to 12 September) measurements
were lost due to instrument failure. Thus, we cannot reliably analyse the
impact of 2006 summer drought on forest floor respiration. The existing soil
CO2 efflux data from September 2006, when the WT was extremely low, do
show higher effluxes than those in early June 2006, although soil surface
temperatures (T5) were lower in September. However, at the same time
T30 was much higher (10.7 ∘C) than in June (5.4 ∘C),
explaining the increased efflux. Compared to the other years, soil
temperatures in September were at their highest in 2006 (Fig. 2), and the
temperature response models thus predicted higher fluxes for September 2006
than for the other years. Following the heavy rains in the beginning of
October, respiration decreased at the same time with the rise of the WT –
and the decrease in T5.
The impact of the WT on forest floor respiration was ambiguous. Correlations
between the WT and CO2 efflux were weak and variable by year and
treatment. The residuals of the model (Eq. 1) estimates vs. The WT indicated a
positive response especially in D collars (lower RFF with lower
WT). However, this effect was caused mostly by spatial variation, as
measurement points in hummocks generally had a lower WT and lower respiration
than the points in the lawn level. Since the models were used for predicting
temporal dynamics, the WT was not included in the models.
Simulated tree stand CO2 flux
The SPP model simulated the tree stand GPP and respiration well. For the year
2008 with the most complete NEE data, the RECO, derived from the
gap-filling and partitioning of the EC measurements, matched very well
(0.8 % difference) the model-derived sum of RFF and
above-ground tree respiration (RTREE, Fig. 7a, Table 2). Not
surprisingly, the model was not able to simulate the suppression of
respiration in 2006 (Fig. 7b), apparently since it does not have linkages to
soil moisture. The simulated 4-year average was 9 % higher than the
EC-derived RECO (Table 2).
The simulated 4-year average GPP of the tree stand was
2473 g CO2 m-2 a-1 (675 g C). The GPP for the
ground vegetation, measured by manual flux chambers in another campaign, was
1040 g CO2 m-2 a-1 (Badorek et al., 2011). Altogether
the tree stand and the ground vegetation GPP sum up to
3513 g CO2 m-2 a-1, which is relatively close
(92 %) to the ecosystem GPP obtained from the partitioning of the EC
fluxes (3805 g CO2 m-2 a-1). These independent
findings suggest that the tree stand contributes about 70 % and the
ground vegetation 30 % of the GPP at Kalevansuo.
Parameter values (a) GPMAXx and (b)
RREF ± 95 % confidence intervals (see Eqs. A2 and A3 in
Appendix A, respectively) for June–October in 2005–2008. For the
respiration model, a constant value of E0= 200 K was used, and the GPP
model was used here without the VPD term.
CH4 fluxes
CH4 flux from ditches was very variable, especially spatially but
also temporally. The instantaneous fluxes varied between -0.098 and
1.757 mg CH4 m-2 h-1. The wettest plot, with
cottongrass (Eriophorum vaginatum), emitted on average
0.936 mg CH4 m-2 h-1, significantly more
(p < 0.001) than the other three, slightly drier plots (mainly
Sphagnum riparium), with mean fluxes of 0.006, 0.056 and
-0.006 mg m-2 h-1. Temporal variation was high but no clear
seasonality was observed. At the wettest plot, fluxes had similar temporal
pattern with the WT, i.e. the highest flux took place in September during the
highest WT.
The average flux from ditch plots was
0.248 mg CH4 m-2 h-1, which calculated for the whole
drained area (ditches 2.5 % of the area) increased the estimated total
flux by 0.006 mg CH4 m-2 h-1. As the flux at the
strips was on average -0.015 mg m-2 h-1 (Lohila et al., 2011),
the site would therefore remain as a small sink for CH4. The annual
areally weighted flux was -0.06 g CH4 m-2 a-1 (i.e.
-0.12 g CH4 m-2 a-1×0.975
(strips) + 2.2 g CH4 m-2 a-1×0.025
(ditches)).
Change in peat layer thickness
The soil surface on the undisturbed D collars had subsided on average by
1.4 cm in 10 years from 2004 to 2014, i.e. 1.4 mm a-1 (Fig. 8).
There was considerable variability between points from an increase in
elevation by 2 cm to a subsidence of 5 cm, so that the change was not quite
statistically significant (p=0.067). Also, some back and forth variation in
peat thickness between years was observed: in August 2011 all but four points
had lower elevation than in 2014. This can be either a measurement error or
real shrink–swell behaviour (breathing) of the peatland.
Carbon balance
The biggest carbon pool at Kalevansuo (Fig. 9.) was the 2.2 m thick peat
layer making up 95.3 % of the total carbon pool. Tree stand (without fine
roots) comprised 4.3 % and ground vegetation only 0.4 %. Fine roots
comprised 0.2 %. The total C pool in vegetation in 2008 was
5.5 kg m-2, which corresponds to about 10 cm layer of peat.
Above-ground parts comprised 62 % of the total biomass. Of the moss
biomass, Sphagna comprised 20 % and forest mosses 80 %.
The tree stand volume increased from 90 m3 ha-1 in 2000 to
130 m3 ha-1 in 2008, i.e. on average by
5 m3 ha-1 a-1. The corresponding carbon pool was
4.6 kg m-2 in 2008 and 3.2 kg m-2 in 2000. The tree stand thus
sequestered ca. 170 g C m-2 a-1. This made 74 % of the
carbon accumulation at Kalevansuo, while the rest was attributed to peat soil
(Fig. 9).
Total litter production was estimated at 437 g C m-2 a-1. Of
this, mosses comprised 20 % and vascular plants 80 %. Of the litter
production by vascular plants, trees comprised 79 % (above ground) and
66 % (below ground). Fine root production was estimated at
120 g C m-2 a-1 (Bhuiyan et al., 2016), comprising 76 % of
the below-ground litter.
As the average of the 4 years, the Kalevansuo peatland ecosystem fixed ca.
1040 g C m-2 a-1 through photosynthesis, 70 % of which was
attributed to the tree stand. Simultaneously it lost
810 g C m-2 a-1 through RECO. Ca. 50 % of
RECO resulted from heterotrophic respiration and 50 % from
autotrophic respiration of trees and ground vegetation. RFF was
comprised mainly of heterotrophic respiration of peat and litter (75 %),
and less by autotrophic respiration of tree roots and ground vegetation
(25 %).
Some C may have been lost through leaching (not measured), but this is
considered a minor component due to ineffective ditches and high
transpiration. No C was lost as methane, as the site was a small
CH4 sink (-0.06 g CH4 m-2 a-1), which is
insignificant for the C balance.
Discussion
Ecosystem CO2 fluxes – the effects of drought
The Kalevansuo drained peatland forest was a strong CO2 sink in all
the 4 years studied (2005–2008). The annual sinks were similar, except
for the dry year 2006, when it was only about 50 % of that in other
years. Interestingly, this decrease in ecosystem CO2 sink was not
caused by increased RECO in drier conditions, as could be
expected. Both GPP and RECO were reduced in summer, and the
reduction in GPP was larger. In addition, the higher-than-normal soil
temperatures in September and October and the very high precipitation in
October resulted in higher RECO in autumn 2006, which partly
explained the much lower annual CO2 net uptake.
Despite the long gap in the NEE data in June and July 2006, we were able to
demonstrate with the data from August, one of the driest months, that drought
had a clear impact on the potential CO2 exchange. Based on the
direct responses between the night-time NEE (respiration) and temperature,
and between the day-time NEE and PPFD, the parameters describing the potential
ecosystem respiration and day-time net CO2 uptake were reduced in
August 2006 compared to the other years. However, GPP was not markedly
different from the other years due to the larger number of clear-sky days
with higher irradiation levels. On the other hand, the higher temperatures
were not able to compensate for the reduced respiration potential (i.e. the
parameter RREF), resulting in a reduced monthly RECO.
Drought has been shown to strongly affect NEE through decreased GPP in
pristine mires where vegetation is adapted to a high water table (Alm et al.,
1999; Bubier et al., 2003; Lafleur et al., 2003). Although Scots pine, the
main tree species in Kalevansuo peatland, is a drought-tolerant species,
summer droughts have been reported to decrease its radial growth in drained
peatlands (Huikari and Paarlahti, 1967). The water table in Kalevansuo is
usually rather high, which means that the roots of pines are located mainly
in the top 40 cm (Bhuiyan et al., 2016), i.e. in the oxic layer above the
average water table. During drought, when the water table may drop down to
80 cm for several weeks, even the pines will probably suffer from water
deficit, and close their stomata.
In contrast to GPP, RECO and soil respiration have often been
shown to increase in peatlands, when the water table is lowered and more peat
is exposed to oxidation (e.g. Silvola et al., 1996; Flanagan and Syed, 2011;
Ballantyne et al., 2014; Munir et al., 2014, 2017). However, many studies
have shown only a weak or no impact of the WT on RECO, whereas
soil temperature has been driving the respiration fluxes (Lafleur et al.,
2005; Nieveen et al., 2005; Juszczak et al., 2013; Olefeldt et al., 2017). In
Kalevansuo, the latter seems to be the case. RECO was slightly
lower during the drought in August 2006 compared to other years (Fig. 5).
RFF was strongly controlled by soil temperatures, whereas the WT
had only a weak and varying effect in different treatments and years.
Forest floor CO2 efflux from different treatment collars
in 2005–2008. A: peat, B: peat + litter, C:
peat + litter + roots and D:
peat + litter + roots + ground vegetation. Points mark individual
chamber measurements and lines modelled daily average fluxes (Eq. 1, App. 3).
Note that two individual fluxes (9.8.2005) with values 1.80 (D) and 1.84 (C)
are outside the graph range, and were excluded as outliers from the
regression models (Table 3).
The decrease in RECO may be caused by decrease of both
RAUT and RHET. As the drought decreases GPP, it will
also decrease photosynthetically driven autotrophic respiration (Olefeldt et
al., 2017), while heterotrophic respiration may well continue in deeper,
still moist but now more oxic, peat layers. However, a large part of
RHET is originated from the decomposition of the new organic
matter (Chimner and Cooper, 2003), i.e. above-ground and root litter,
deposited mainly in the very surface of the peat soil. In drained peatlands
the decomposition rate of this surface layer is hardly ever restricted by a too-high WT, but sometimes it can be restricted by too-low moisture content
(Mäkiranta et al., 2009).
If water levels were lowered for a longer period, e.g. through deeper
ditching, the effect might be different than that of drought: a more
efficient drainage would induce higher decomposition and heterotrophic
respiration through changes in microbial communities (Mäkiranta et al., 2009) but also probably increased root growth into the deeper layers.
(a) Ecosystem respiration based on measured and gap-filled
EC data (RECO) and simulated respiration effluxes of different
components at Kalevansuo in 2008. Since RECO is the total
ecosystem respiration, it should equal the sum of above-ground respiration of
trees (RTREE) and total forest floor respiration
(RFF). RTREE was simulated using the SPP model, while
RFF is based on measured flux data and statistical models for the
same site and year. (b) RECO compared with the sum of
RTREE+RFF during the whole measurement period
2005–2008.
Soil subsidence
Even though the flux and biomass data indicate a steady increase in soil C
stock, a small (insignificant) subsidence of the soil surface was measured
(0.14 cm year-1). The value is considerably smaller than that reported
for agricultural fields (0.3–3 cm year-1; Oleszczuk et al., 2008), or
for palm oil plantations on peat with high observed C losses
(4.2 cm year-1; Couwenberg and Hooijer, 2013). In peatlands drained
for forestry, subsidence is in the long term usually much smaller (Lukkala,
1949; Minkkinen and Laine, 1998a) because of shallower drainage and
continuous litterfall and humus formation on the soil surface. The only
published long-term study from drained peatland forest reports rates of
0.4–0.7 cm year-1 for a dwarf-shrub site in southern Finland (Ahti,
2002).
Subsidence of peat is caused by physical compaction and loss of organic
matter through oxidation. In physical compaction, solid matter is compacted
into a smaller space. The result is the increase in bulk density, which is
evident in all drained peatlands (e.g. Minkkinen and Laine, 1998b). We do not
have bulk density measurements from Kalevansuo peatland prior to drainage,
but compared to similar pristine sites (38 kg m-3 natural pine mires
(Minkkinen and Laine, 1998), bulk density of the surface 0–20 cm layer is
higher in Kalevansuo (94 kg m-3). It is therefore likely that bulk
density has increased in Kalevansuo after drainage. In oxidation, organic
matter is lost as CO2 from the peat to the atmosphere. In peat
soil, both processes take place at the same time, and in forested sites
especially, the C loss through oxidation is to varying extent compensated for
by litter production. Thus, given the estimated positive soil C balance
(i.e. accumulation of C in the soil) at Kalevansuo, we conclude that the
observed small subsidence is caused by compaction, not by loss of peat.
Carbon balance
Kalevansuo accumulated atmospheric C during every year of the study. Given that the average net carbon uptake of the site was
230 g m-2 a-1 and that 170 g m-2 a-1 was
sequestered to the growing tree stand, the remaining
60 g C m-2 a-1 must have been accumulated in the other parts of
the ecosystem. If the ground vegetation biomass is assumed to be constant,
the surplus must be in the peat soil. This assumption is based on ocular
assessment at the site. It is reasonable to assume that the ground vegetation
biomass is not increasing, since the tree stand is steadily growing bigger
and the correlation between tree stand and ground vegetation biomass is
negative (Reinikainen et al., 1984). Furthermore, an increase of
60 g C m-2 a-1 would equal the doubling of shrub biomass in 5
years, and that should be clearly visible. Thus the method should not be
overestimating soil C pool increase. However, as the C pool in ground
vegetation is 1/10 of that in the tree stand, the change in C pool would be
irrelevant, assuming the same relative growth rate.
(a) Elevation of soil surface in 2004, 2011 and 2014 in the
middle of the undisturbed D plots, relative to the fixed benchmark beside the
EC mast. (b) Change in elevation relative to 2004, mean and
2 ⋅ standard error of the mean. Only the points measured at every
occasion are included in the mean and SEM values.
Despite the small biomass pool compared to the tree stand, ground vegetation
was estimated to produce above-ground litter at a rate of
130 g C m-2 a-1, i.e. almost as much as the tree stand
(Fig. 9). The majority of this litter originates from mosses, the coverage of
which is almost 100 % in Kalevansuo. Another rapidly renewing biomass
pool was that of fine roots, which was composed almost totally of tree and
shrub roots. About half of this pool is renewed annually, producing root
litter at a rate of 120 g C m-2 a-1 (Bhuiyan et al., 2016).
When decomposed, part of the released C is translocated as a solute into
deeper peat layers (Domisch et al., 2000). Thus, despite being small C
pools, both ground vegetation and fine roots have a large impact on the soil
C balance..
In our estimation, the C in the below-ground parts of trees (stumps and
roots > 1 cm diameter) was considered as tree biomass, which
increases as the stand grows. When trees die, either naturally or as they are
harvested, the below-ground part of C becomes a part of the soil C pool.
Considering this below-ground biomass as a part of the soil C pool would
increase the soil C accumulation estimate to over
100 g C m-2 a-1. The biomass of smaller roots could of course
also change, but as the biomass pool of the 2–10 mm roots is only a small
fraction of that of the bigger ones (Fig. 9), and as the fine root turnover
is rapid (50 % a-1), this is not considered a major uncertainty.
Taking into account the leaching of C would have only a minor effect on the
NEE estimate. We do not have dissolved organic carbon (DOC) measurements from
Kalevansuo, but leaching of DOC, i.e. the output of dissolved C, from Finnish
drained peatlands ranges from 10 to 15 g C m-2 a-1 (Sallantaus
and Kaipainen, 1996; Kortelainen et al., 1997; Sarkkola et al., 2009;
Rantakari et al., 2010). This is 4–7 % of the estimated NEE and
17–25 % of the soil C balance at Kalevansuo. As the ditches in
Kalevansuo are ineffective and the transpiration of the tree stand and ground
vegetation is an important pathway for water output (Sarkkola et al., 2010),
leaching of DOC at Kalevansuo is likely at the lower end of the observed
range. Thus, taking leaching into account would not change the conclusion on
soil C sink.
Based on 4-year NEE and tree growth data, we estimated that the accumulation
of C in soil was on average 60 g C m-2 a-1 during the 4-year
period. Since the tree stand growth data is based on 5-year average, we
cannot say whether the soil C balance has been positive in all the studied
years. Neither can we say if the long-term soil C balance of the peatland
would stay similar in the future. In natural mires, where long-term peat C
accumulation can be reliably estimated from peat coring and radiocarbon
(14C) dating, the multi-year mean NEE derived from EC measurements
has typically been similar to the long-term accumulation rate (Aurela et al.,
2004; Roulet et al., 2007; Nilsson et al., 2008), although not in all cases
(Ratcliffe et al., 2017). It is well known that long-term average rates,
determined by peat coring and radiocarbon dating, are not necessarily the
same as the actual, current (or decadal average) rates (e.g. Clymo et al.,
1998; Frolking et al., 2014).
Kalevansuo has been cored extensively, and historical C accumulation has been
determined using radiocarbon dating (Mathijssen et al., 2017). However, the
drainage took place so recently (35 years before our study) that
post-drainage C accumulation cannot be reliably determined using
14C dating. Even if the surface peat could be dated accurately,
root growth into deeper layers would mess up the C accumulation estimate.
Several peat-coring methods have been tried to estimate post-drainage changes
in peat C stocks (e.g. Kruger et al., 2016; Minkkinen and Laine, 1998;
Minkkinen et al., 1999; Simola et al., 2012; Turetsky et al., 2004) but they
all have large uncertainties. However, as discussed above, we were not trying
to estimate long-term peat accumulation, only the current rate. Eddy
covariance combined with biomass growth measurements is the most accurate
method for this purpose.
Measured carbon pools (rounded boxes; g C m-2, and the
changes in pools in italics; g C m-2 a-1) and fluxes (arrows and
square boxes; g C m-2 a-1) in Kalevansuo drained peatland. Soil
C accumulation is calculated as NEE (230 g C m-2 a-1) – C
sequestration in tree stand biomass (170 g C m-2 a-1; above-
126 g C m-2 a-1 and below-ground 43 g C m-2 a-1).
Other fluxes and pools are based on measured and modelled values derived from
EC and chamber measurements, biomass measurements, and litterfall measurements in
Kalevansuo. The tree stand biomass is from the autumn 2008 measurement. (1)
Badorek et al., 2011; (2) fine root production of trees (Bhuiyan et al.,
2016); (3) fine root production of shrubs and herbs (Bhuiyan et al.,
2016).
Ojanen et al. (2012) evaluated different chamber-based methods for
calculating the soil C balance, and compared these to the EC-based method
described above. The “L–RHET method” (litter production minus
heterotrophic respiration) produced varying results depending on the variable
fine root turnover rates available from the literature. Using the recent results
of fine root production in Kalevansuo (Bhuiyan et al., 2016) we end up with L
of 437 g C m-2 a-1 – and RHET of 450 g, which
results in a loss of 13 g C m-2 a-1. Thus there is still a
difference of about 73 g C m-2 a-1 to the EC-based estimate.
This difference is probably caused by uncertainties in estimating
RHET (Ojanen et al., 2012). The cutting of roots causes an extra
litter input (e.g. Subke et al., 2006) and on the other hand prevents further
input. Roots may also reach under the 30 cm deep collars (Bhuyian et al.,
2016). Trenching also affects soil moisture that regulates respiration (Subke
et al., 2006).
In addition to the L–RHET method, soil C balance can be
estimated using data from transparent chamber and tree litter measurements,
as follows:
Soil C balance=GPPFF-LTREE+RFF-RAUTof tree roots,
where GPPFF is chamber-measured GPP of forest floor vegetation
and LTREE total litter from trees. Since the chamber-measured
RFF includes also tree stand root respiration this must be
subtracted from RFF.
We estimated GPPFF at -288 (Badorek et al., 2008),
RFF at 600, RAUT of tree roots at 89 and
LTREE at 253 g C m-2 a-1 (Fig. 9). This gives an
estimate for the soil C balance of -30 g C m-2 a-1 (sink),
which is relatively close to the EC-based estimate of
-60 g C m-2 a-1, and supports our finding of the soil C sink.
Can the carbon sink last?
Here we have shown that the Kalevansuo drained peatland ecosystem and even
the soil is currently a carbon sink despite the drainage. It would be
reasonable to assume that drainage would turn a peatland soil into a carbon
source, because the decomposition of peat is typically increased after
drainage. Drainage in Kalevansuo is, however, rather superficial, the average
water table being at 35–40 cm, i.e. only about 15–20 cm lower than in
natural dwarf-shrub pine bogs (Minkkinen et al., 1999). The site is
topographically rather even, as is typical for nutrient-poor pine bogs, so
draining with open ditches is not efficient. Thus the ditches are blocked by
vegetation and drainage is mainly mediated by the transpiration through the
tree stand (Sarkkola et al., 2010). The soil is also almost fully covered
with vegetation, including mire species like Sphagnum mosses. Such a
small change in vegetation structure is typical for drained dwarf-shrub pine
bogs (Minkkinen et al., 1999). It thus appears that this peatland has not
lost the ability to keep up the relatively high water table and surface
moisture supporting the continuous growth of mosses. Only very dry seasons,
like summer 2006, may disturb the hydrology so much that C dynamics are
seriously affected.
It is evident that most boreal and temperate peatland forest ecosystems,
where drainage has been successful, act as contemporary C sinks (Ojanen et
al., 2013; Meyer et al., 2013; Hommeltenberg et al., 2014), because the tree
stand C sequestration exceeds the loss of C from soil. In peatlands used for
forestry it is, however, the soil C storage that is important in the long
term, given that the tree stock will eventually be harvested and the C in
wood products will gradually be lost back to the atmosphere. Thus the most
relevant question is as follows: will sites like Kalevansuo remain C sinks in
the long term if they are managed for forestry? After the site is harvested,
as is typical, by clear-cutting, soil decomposition processes will go on,
whereas litter production from tree stand is ceased for several years.
Logging residues will decompose rather fast, and may enhance the
decomposition rate of the underlying peat soil (Mäkiranta et al., 2012,
Ojanen et al., 2017). This will create a loss of soil C through soil
respiration, the magnitude of which is dependent on soil quality (von Arnold
et al., 2005a, b; Minkkinen et al., 2007).
On the other hand, in the typical stem-harvesting method, tree stumps and roots
are left at the site, increasing the C stock in the soil significantly. This
C pool of coarse woody debris is not easily decomposed (Laiho and Prescott,
2004), especially when buried in peat soil, and its inclusion will compensate
for the soil C losses for several years. Also, after clear-cut, the water
table will rise because of the removal of the transpiring tree stand, likely
reducing peat decomposition rate (Mäkiranta et al., 2010). This reduction
is, however, probably quite small and the site is likely to be a strong C
source at least for the first 5 years, after which the growing vegetation
again starts to bind carbon to the system (Mäkiranta et al., 2010; Kolari
et al., 2004). However, no data of C dynamics of the young stand phase on
forested peatlands exist. To answer the question of the climatically best
option to manage different kinds of drained peatlands, simulations with
mechanistic models verified for peatland conditions (e.g. He et al., 2016)
are promising tools.