The effect of clear-fell harvesting on soil greenhouse gas
(GHG) fluxes of carbon dioxide (CO2), methane (CH4), and nitrous
oxide (N2O) was assessed in a Sitka spruce forest growing on a peaty
gley organo-mineral soil in northern England. Fluxes from the soil and
litter layer were measured monthly by the closed chamber method and gas
chromatography over 4 years in two mature stands, with one area harvested
after the first year. Concurrent measurements of soil temperature and
moisture helped to elucidate reasons for the changes in fluxes. In the 3 years after felling, there was a significant increase in the soil
temperature, particularly between June and November (3 to 5 ∘C higher),
and in soil moisture, which was 62 % higher in the felled area, and these
had pronounced effects on the GHG balance in addition to the removal of the
trees and their carbon input to the soil. Annual soil CO2 effluxes
reduced to almost a third in the first year after felling (a drop from 24.0
to 8.9 t CO2 ha-1 yr-1) and half in the second and third year
(mean 11.8 t CO2 ha-1 yr-1) compared to before felling, while
those from the unfelled area were little changed. Annual effluxes of
N2O more than doubled in the first two years (from 1.0 to 2.3 and 2.5 t CO2e ha-1 yr-1, respectively), although by the third year
they were only 20 % higher (1.2 t CO2e ha-1 yr-1). CH4
fluxes changed from a small net uptake of -0.03 t CO2e ha-1 yr-1 before felling to a small efflux increasing over the 3 years to
0.34 t CO2e ha-1 yr-1, presumably because of the wetter soil
after felling.
Soil CO2 effluxes dominated the annual net GHG emission when the three
gases were compared using their global warming potential (GWP), but N2O
contributed up to 20 % of this. This study showed fluxes of CO2,
CH4, and N2O responded differently to clear-felling due to the
significant changes in soil biotic and abiotic factors and showed large
variations between years. This demonstrates the need for multi-year
measurements of all GHGs to enable a robust estimate of the effect of the
clear-fell phase on the GHG balance of managed forests. This is one of very
few multi-year monitoring studies to assess the effect of clear-fell
harvesting on soil GHG fluxes.
Forests cover approximately 30 % (4.03 billion ha) of the Earth's land
surface and play a major role in the cycling of soil carbon and greenhouse gases
(GHGs)
(Le
Quéré et al., 2015). Afforestation and forest management can
contribute to GHG net emission mitigation aims by increasing the land-based
carbon sink (Grassi et al., 2017). In the UK, 3.2 million ha or 13.2 % of the land area (Forest Research, 2020) are covered by
forests, a substantial
proportion of which has been planted over the last 60 to 100 years on peat
and peaty gley upland soils (Smith et
al., 2018). The UK government has pledged to reach “net zero” emissions of
GHGs by 2050, and to contribute to this goal the Committee on Climate Change (2019) has recommended afforestation targets of more than 30 000 ha per
year. Forest harvesting is an important activity as part of normal forestry
practice but also for conversion to other land uses such as part of the peatland
restoration programmes. Approximately 20 000 ha of trees are felled annually
in Great Britain for timber and other harvested wood products (Forestry
Commission, 2016).
Forest clear-fell harvesting is the phase of the forest management cycle
that produces the most disturbance, and therefore it is important that it
does not impair long-term productivity and benefits. Clear-felling alters
(typically increases) many soil factors that influence GHG fluxes. The
physical factors include: soil water content and water table height owing to
the absence of evapotranspiration from trees (Zerva and Mencuccini, 2005;
Wu et al., 2011; Kulmala et al.,
2014; Sundqvist et al.,
2014; Korkiakoski
et al., 2019), soil temperature through reduced shading (Zerva and
Mencuccini, 2005; Wu et al., 2011;
Kulmala et al., 2014), and soil bulk
density due to soil disturbance and compaction caused particularly by
mechanized harvesting equipment (Yashiro et al., 2008;
Mojeremane et al., 2012). Chemical
factors include: soil pH (Smolander et al., 1998; Kim, 2008; Kulmala et al.,
2014; Sundqvist et al., 2014) and inputs of soil N and organic C
(Smolander et al., 2015;
Tate
et al., 2006; Hyvönen et al.,
2012; Kulmala et al., 2014; Sundqvist et al., 2014) due to nutrient release
from the decomposition of residual organic matter and root biomass.
Interactions between these factors, particularly with the loss of plant
litter input and root activity after felling, will strongly affect soil
biological processes responsible for production and consumption of GHGs.
They will particularly affect nitrous oxide (N2O) production from
aerobic nitrification and anaerobic denitrification, methane (CH4)
production by methanogenic organisms from anaerobic decomposition in
oxygen-poor environments or uptake by methanotrophs through oxidation in
aerated soils, and carbon dioxide (CO2) efflux during respiration and
decomposition and uptake during photosynthesis. The cessation of the
autotrophic respiration (Ra) component of the total soil respiration (Rt)
after felling should cause a large decline in CO2 efflux as a
meta-analysis of soil respiration partitioning studies reported that the
Ra/Rt ratio in temperate forest soils ranges from 20 % to 59 %
(Subke et al., 2006). In addition,
the death of tree roots after felling will inhibit the microbial
decomposition of root exudates, reducing the CO2 effluxes further.
Conversely, the increased soil temperatures following tree removal and
higher nutrient availability from decomposition of litter and other plant
materials such as brash may increase soil heterotrophic respiration (Rh)
(Yashiro et al., 2008).
There are a wide range of results published on the effects of forest
clear-felling on soil GHG fluxes. Some studies have shown CO2 effluxes
were reduced (Zerva and Mencuccini, 2005; Korkiakoski et al., 2019), increased
(Tate et al., 2006; Kim, 2008; Kulmala et al., 2014), or unchanged (Butnor
et al., 2006; Takakai et al., 2008; Yashiro et al., 2008). According to
Lavoie et al. (2013), the impact is often site specific, affected by the
severity of the disturbance or removal of surface organic matter or by the
length of time following harvest. Goutal et al. (2012) examined the duration
of physical, chemical, and biological disturbances in the soil following
mechanized harvesting of an oak forest in north-east France. Their measurements
showed that soil CO2 effluxes reduced, which they attributed to an
increase in the frequency and duration of anoxic conditions resulting from
poor soil gas diffusion after heavy forestry traffic. Similarly, Kulmala et
al. (2014) observed a slight decrease in soil CO2 efflux in the first
growing season after clear-cutting of a boreal Norway spruce stand, although
this was probably due to decreased tree root respiration. However, during
the following 2 years, CO2 efflux at their clear-cut site was
significantly higher than in their mature stand, which was attributed to
increased decomposition stimulated by higher soil moisture and temperature.
They observed no significant difference in CH4 uptake due to
clear-cutting. However, others have shown that after harvesting, forest soils
turned from a CH4 sink to a source (Zerva and Mencuccini, 2005;
Sundqvist et al., 2014; Korkiakoski et al., 2019), showed reduced CH4
uptake (Bradford et al., 2000; Wu et al., 2011; Yoshiro et al., 2008), or showed
increased CH4 uptake (Lavoie et al., 2013). For N2O, some studies
showed increased fluxes (Zerva and Mencuccini, 2005; Yoshiro et al., 2008;
Takakai et al., 2008; Ullah et al., 2009; Korkiakoski et al., 2019) or had no
clear change (Tate et al., 2006; Lavoie et al., 2013).
There is therefore a wide range of observed effects, making it difficult to
predict the changes in the GHG balance caused by clear-felling in any
particular situation, and very few multiple-year studies exist that characterize
the timescale and duration of changes post-felling. Therefore, there is an
urgent need to understand and quantify with long-term measurements the
effect of forest clear-felling on the GHG budget and to incorporate these
into analyses for the complete forest growth cycle for the main forest
systems relevant to a country or region (Skiba et al., 2012). The objective
of this study was to conduct a relatively long-term (4 year) assessment of
the effect of clear-fell harvesting on soil (including the litter layer) fluxes
of CO2, CH4, and N2O in a spruce plantation on organo-mineral
soil that is typical of many British upland forests.
Materials and methodsSite description and study layout
The study site was in Harwood Forest (55∘13.1′ N, 2∘1′ W), Northumberland, north-east England, and comprised a forest area of
approximately 4000 ha with an elevation of 200 to 400 m above sea level
(Fig. 1). The regional climate is temperate oceanic with a mean annual
rainfall of 1472 mm and mean air temperature of 7.5 ∘C (min. -7
and max. 26.4 ∘C), measured by an automatic weather station (AWS)
mounted above the tree stand during the period from April 2015 to April 2018. The tree cover consisted predominantly of even-aged Sitka spruce
(Picea sitchensis, Bong. Carr.) stands. Sitka spruce is the most common conifer in Great
Britain (GB) and represents 26 % of the total forested area in GB and
51 % of the conifer area (Forest Research, 2020). The main soil type is a
seasonally waterlogged organic-rich peaty gley classified as histogleysol
with a peat (O Horizon) thickness varying from 15 to 40 cm (Zerva and
Mencuccini, 2005), developed in clayey glacial till derived from
carboniferous sediments (Pyatt, 1970).
For the purpose of this study, two nearby areas (A and B, 2 km apart) of
mature stands within the forest with similar previous management, soil type,
and elevation (280 to 290 m) were chosen to carry out measurements over 4 years between 18 February 2014 and 16 April 2018. During this period, one
area (A) was left unfelled (A-year 1 to A-year 4) and the other (B) was
clear-felled after one year (B-year 1 before felling and B-year 2 to B-year 4 after
felling). In area A, the 40 ha stand was of second rotation, even-aged
mature Sitka spruce planted in 1973, with yield class of 18 m3 ha-1 and mean tree density of 1348 trees ha-1. In area B the
Sitka spruce stand was planted in 1958, with yield class of 16 m3 ha-1 and mean tree density of 1375 trees ha-1 prior to felling;
the felled area covered 42 ha. Felling operations were carried out
between late January and early March 2015 and followed standard practices of
the Forestry Commission (Murgatroyd and Saunders, 2005). Only timber larger
than 7 cm diameter was removed from site, leaving tree tops and branches on
site in rows, with some used as brash-mats to prevent compaction of soil by
the heavy harvesting machinery.
Gas flux measurements and analysis
Forest soil fluxes of CO2, CH4, and N2O were measured at
approximately monthly intervals over the 4-year study period from the two
areas of the forest. Flux measurements were made using a modified design of
the manual static chamber method described by Yamulki et al. (2013). Each
chamber was constructed of opaque PVC with dimensions of 40 cm × 40 cm × 25 cm height to provide a volume of 40 L and placed
temporarily on permanently installed frames. The frames were inserted
tightly into the ground to a depth of about 3 cm prior to the start of the
measurements. The bottom of the chamber had a neoprene rubber foam gasket to
ensure a gas-tight seal with the frame and the top of the chamber had a
pressure vent.
Within each area, eight chambers were positioned randomly in a transect within a
100 m2 area. During each gas flux measurement, the chambers were placed
on top of the frames for up to 60 min and duplicate gas samples of the
chamber headspace were taken immediately after closure and then at three
subsequent 20 min intervals. Gas samples were taken after the chamber was
closed by connecting a polypropylene syringe to a chamber sampling port
fitted with a three-way stopcock. The syringes were immediately used to fill
(under atmospheric pressure) pre-evacuated 20 mL vials fitted with
chlorobutyl rubber septa. Concentrations of CO2, CH4, and N2O
were determined within a week using a headspace sampler (TurboMatrix 110)
and gas chromatograph (GC; Clarus 500, PerkinElmer) equipped with an
electron capture detector (ECD) for N2O analysis, a flame ionization
detector (FID) for CH4 analysis, and a catalytic reactor (methanizer)
for CO2 analysis by reducing CO2 to CH4 before analysis by
the FID detector. The repeatability of the GC gas analysis (assessed as 3× the standard deviation of 20 repeated measurements of standard
CO2, CH4, and N2O concentrations at ambient levels) was better
than 4 % for all gases.
Gas fluxes were calculated based on linear increases of gas concentrations
inside the chambers with time. For CO2, however, if the concentration
increase was not linear then fluxes were determined using the R HMR package
(version 1.0.1) to plot a best-fit line to the data
(Pedersen, 2020) to correct
for the non-linearity. We did not apply the HMR model for CH4 and
N2O fluxes as the non-linear fitting plot is very sensitive to
variability and outliers in the measured GHG concentrations, particularly
for low fluxes and with only four data points per chamber (as also noted by
Pihlatie et al.,
2013; Brümmer et al., 2017;
Korkiakoski et al., 2017), which results in large apparent “spikes”, failure
to calculate the fluxes on many occasions, and likely overestimation of
calculated GHG fluxes (Pavelka et al., 2018). If CO2
concentration changes with time were not significant, fluxes for all gases
were rejected for that sample as this was judged to be indicative of gas
leakage within the chamber headspace. Overall 18 samples were rejected, nine of
which were during a snowfall period in January 2016.
Although most studies now measure soil CO2 fluxes with infrared gas analyser systems (IRGAs; as noted
by Yashiro et al., 2008), which is viewed as more accurate than gas sampling
and GC analysis, we needed to use the GC method to measure all three GHGs within
the logistical constraints of the experiment. Therefore, to give further
assurance in the gas flux calculations, CO2 effluxes were compared with
those from another 25 static chambers (20 cm diameter, 4.2 L volume,
LI-8100-103 Survey Chamber, Li-Cor Inc., Lincoln, Nebraska, USA) measured
in situ in both areas by a closed loop IRGA (LI8100A, Li-Cor Inc.), each over a 2 min duration after chamber closure
(Xenakis et al., 2021). These chambers were positioned over a much wider
area than the GC chambers to characterize spatial variation, and effluxes
were measured for 3 years but only after felling from February 2015. The
results (Fig. S1 in the Supplement) showed a mean flux difference over the 3-year period of
only 6.8 % higher by the IRGA method in the unfelled area and 19.5 %
higher in the felled area compared with the fluxes measured by GC. These are
relatively small differences between the two methods when considering the
higher site heterogeneity in the felled area, the inherent differences in
the analytical methodologies, and that the vegetation in the IRGA chambers was
not cut as it was in the GC chambers.
There was no understorey vegetation in the forest stands and no visual
evidence of vegetation growth within the chambers in the first 2 years of
the study; the chamber ground surfaces were covered by dense Sitka spruce
needle litter. However, in the third year there was growth of Juncus sp. in
one chamber in the felled area B between May and September 2016 (maximum
height was 40 cm before cutting but the total volume was <5 % of
the chamber volume), so the Juncus was cut continually thereafter. The effect
of the vegetation cut on the fluxes was assessed from the flux measurements
in all eight chambers before and directly after the first cutting. The
statistical analysis revealed no significant differences between mean
chamber fluxes before and after cutting for all gases, indicating that
variations between the chamber fluxes were greater than that due to cutting.
In the third and fourth years there was a proliferation of moss in two of
the chambers in area A. This could not be removed without substantial
disturbance of the soil surface; comparison of chambers with and without
moss showed no significant differences between mean fluxes. For CO2,
the effluxes measured will therefore be from aerobic and anaerobic
decomposition processes and respiration of soil organisms and roots.
Soil moisture and temperature measurements
During each flux sampling day, soil temperatures (∘C) at 2 and 10 cm
depths were recorded from one point around each chamber and volumetric
moisture content (m3 m-3) at 6 cm depth was recorded from three
points around each chamber. Soil temperature was recorded by a digital
temperature probe (Hanna model Checktemp 1) and the volumetric moisture
content by a moisture sensor (SM 200 attached to a handheld HH2 moisture
meter, Delta-T Devices Ltd, Cambridge, UK). The sensitivity of CO2 efflux
to temperature was determined with a Q10 function (Shi and Jin, 2016), which
is the proportional change in respiration resulting from a 10 ∘C
increase in temperature, derived from mean daily values of CO2 efflux
for each year using the equation
Q10=exp10⋅b,
where b is the slope of the exponential relationship between soil CO2
effluxes and soil temperature at 2 cm depth, estimated from a log-linear
fit. The apparent Q10 values were calculated from the temperature measured
near the soil surface (at 2 cm) as recommended by Pavelka et al. (2007),
because the strength of the CO2 efflux and temperature relationship
decreases with soil depth.
Soil parameters and root biomass measurements
To assess changes in soil characteristics due to the felling, additional
soil parameters that were likely to have an impact on GHG fluxes were
measured between 21 and 23 July 2015, approximately 5 months after
the end of felling. Three replicated soil samples were taken from 0–20 cm
depth below the litter layer around each chamber at each site for bulk
density, pH, total C content, and total N content. Soil pH was measured by mixing 5 g
soil samples with 25 mL H2O with the analysis performed using a pH meter probe
(Sentek). Soil C and N stocks were measured by the flash combustion method
in an NC soil analyser (Flash EA, series 1112, Thermo Scientific). Soil bulk
density was measured as the mass of oven-dried (at 105 ∘C until
constant weight) soil samples divided by the volume of the cores taken. Live
and dead fine root biomass, length, and diameter were measured from three
replicated samples taken from 0–15 cm depth below the litter layer around
each chamber in each site. As soil parameters were only measured on one
occasion of post-felling, the differences between areas could be due to the sites
rather than the felling.
Statistical analysis
Statistical analyses were made using statistical software R (version 3.5.2;
R Core Team, 2020). Data were analysed separately for pre-felling and
post-felling periods. As measurements were taken from the same eight
chambers through time, the analysis was conducted as a repeated measures
design, with the significance of felling and non-felling differences determined
against the individual chamber data (n=16), not against individual
observations. A linear and non-linear mixed-effects model (nlme; Pinheiro et al., 2018) was
used to structure and analyse the data for all flux data. Where necessary,
flux data were transformed to obtain a normal distribution. Management type
(i.e. felled or unfelled), temperature at 2 and 10 cm, soil moisture (plus
all two-way interactions), and date (plus interaction with management type)
were treated as fixed effects. Chambers were treated as a random effect (for the
repeated measures design). For post-felling data, residual analysis
indicated that the variance was larger for chambers in the felled area and
by individual chamber; therefore, weighted variance structures were
incorporated within the mixed effects models to account for within-type and
within-chamber heterogeneity.
A range of autoregressive moving average (corARMA; Pinheiro et al, 2018)
models were applied to each model to account for temporal autocorrelation
within chambers. Analysis of the (partial) autocorrelation function
indicated potential temporal autocorrelation up to one previous time point;
therefore, all combinations of corARMA structure (up to one previous time
point) were applied and the best-fit model determined using Akaike's
information criteria (AIC; R Core Team, 2020) applied to the maximum
likelihood fits across each of the gas fluxes separately. The significance
of the fixed effects were subsequently determined using analysis of deviance
(ANOVA chi-square tests; Fox and Weisberg, 2011) with non-significant
interactions and effects (p>0.05) removed from the final models.
Annual cumulative fluxes of CH4, N2O, and CO2 were estimated
to assess the inter-annual variations in each area. Within each year, median
flux values were calculated across the eight replicate chambers, along with
lower and upper quartiles and maximum and minimum values. These values were
then accumulated across the year by taking the mean of two consecutive flux
values and multiplying it by the number of days between the measurements and
summing over the monitoring period every 12 months (as indicated in Table S1 in the Supplement) and adjusting to 365 d. For year 1, where the felling interrupted
the measurement in area B, the same cumulative time period was taken for
both areas. Annual cumulative fluxes were converted into CO2 equivalent
(CO2e) using the global warning potential (GWP) for a 100-year time
horizon of 34 for CH4 and 298 for N2O (IPCC, 2013).
ResultsSoil temperature and moisture
Over the 4 years, soil temperature measured during gas sampling varied
between 0.1 and 26.0 ∘C (mean 9.7 ∘C) at 2 cm soil depth (Fig. 2a)
and 1.3 and 15.5 ∘C (mean 8.3 ∘C) at 10 cm depth (data not
shown), although soil temperatures were never above 15 ∘C under the
trees in area A. Before felling in year 1, there were only small differences
(p<0.001) in mean soil temperature between areas A and B
(7.1 and 8.5 ∘C, respectively) at 2 cm soil depth and at 10 cm soil depth
(7.1 and 7.9 ∘C) (Table 1), probably caused by sampling the area
B later in the day than A. Due to removal of the
tree cover, the mean soil temperature
increased in the felled area at 2 cm soil depth to a mean of 14.3 ∘C compared to 9.2 ∘C in the unfelled area (p<0.001) and at 10 cm depth to a mean of 10.2 ∘C compared to 8.4 ∘C. In the following two years, the soil
temperature remained higher in the felled area by up to 3.4 ∘C compared
to area A at 2 cm depth (1.6 ∘C at 10 cm) and was on average 2 ∘C higher than before felling (there was no change in the mean soil
temperature in A).
Annual mean of eight replicate measurements (± SE) of
soil temperatures at 2 and 10 cm depths and moisture at 6 cm depth
measured close to the soil GHG flux chambers in area A (mature spruce stand,
remaining) and area B (mature spruce stand, clear-felled after year 1) in
Harwood Forest. Significance p values are all <0.01.
Mean soil temperature (∘C) at 2 cm depth (a),
volumetric soil moisture content (cm3 cm-3) (b), and median soil
CO2(c), CH4(d), and N2O (e) fluxes measured from eight chambers per area approximately monthly throughout the experiment in Harwood
Forest. Black symbols and lines are for area A (mature spruce stand,
remaining) and red are for area B, (mature spruce stand, clear-felled after
one year) as indicated by the dotted vertical line. Error bars are standard
error of mean of eight replicate measurements of the soil temperatures and
moisture.
No significant differences in soil moisture content (by volume) were
observed between the two areas before felling with a mean of 0.36 in area A
and 0.39 m3 m-3 in B in year 1 (Fig. 2b and Table 1). However,
after felling the soil moisture content was significantly higher in the
felled area than the unfelled (mean 0.62 m3 m-3 compared with
0.38 m3 m-3; p<0.001), due to the reduced
evapotranspiration after tree removal (Xenakis et al., 2021). In both areas,
there was a pronounced seasonal variation in soil moisture in the year
before felling (2014) and in the first few months of 2015, with higher
moisture in winter months than in summer, but this pattern was not clear
thereafter because the rainfall was more evenly distributed during the last
2 years of this study (Fig. 3).
Monthly total rainfall measured after felling in each
area in Harwood Forest atop the 30 m tower in the mature spruce stand (area A; black) and on the ground in the clear-fell site (area B; red).
Measurements only started in area B in June 2015.
Soil parameters and root biomass measurements
Soil parameters for the 0–15 cm layer measured 5 months after felling in
year 2 showed no significant differences between the unfelled and felled
areas in mean soil pH (3.6 and 3.8, respectively) and soil total N content
(1.8 % and 1.95 %), but the mean soil total C content was about 17 % lower
(p<0.001) in the felled area (52.0 % and 43.3 %). The soil bulk
density was significantly (p<0.001) higher at 0.30 g cm-3 in
the felled area compared with 0.22 g cm-3 in the unfelled, but both
values are typical for peaty gley soils (Vanguelova et al., 2013). The
higher bulk density and lower C content could be an effect of felling,
although as no pre-felling measurements were taken it may be because of site
differences. Mean fine (<2 mm) live root mass was much lower in the
felled area, as expected (1.6 t ha-1 compared with 4.9 t ha-1), but the
difference was smaller for the mean fine dead root mass (0.35 and 0.53 t ha-1, respectively), probably due to partial or complete decomposition
during the 5 months after felling prior to the measurements.
GHG fluxes
Large variations were observed in the fluxes between the eight replicate
chambers after felling with some high outliers, particularly for CH4 (Fig. 2c–e). Therefore, to reduce bias in the annual budget
estimates of CO2, CH4, and N2O fluxes and enable a robust
comparison between annual fluxes before and after felling, the annual and
cumulative fluxes were based on the median of the replicate chambers as
described in the statistical analysis section.
CO2 effluxes
In the first year before felling, there were no significant differences in
soil CO2 effluxes between areas A and B (median 1.54 and 1.75 g
CO2-C m-2 d-1, respectively; Fig. 2c). In the
following 3 years after felling, CO2 effluxes became significantly
(p<0.001; Table 2) lower in the felled area (median 1.10, 0.90, and
0.92 g CO2-C m-2 d-1 in year 2, 3, and 4, respectively) than in
the unfelled area (2.44, 1.69, and 1.44 g CO2-C m-2 d-1).
There was a clear seasonal variation in the CO2 effluxes at both areas,
which as expected followed that of soil temperature with maximum effluxes
during June to September.
Results from the analysis of deviance (chi-square tests)
showing the probability (p) values for the effects of the explanatory
factors and variables: felling, soil temperature (at two depths 2 and 10 cm),
volumetric soil moisture measurement date, and their interactions (as fixed
effects) on soil CO2, CH4, and N2O fluxes in Harwood Forest.
Note: p<0.05 is deemed significant; pre-felling denotes the comparison
between area A (mature spruce stand remaining) and area B (before
clear-felling) in year 1; post-felling denotes comparisons between the areas
for year 2 to 4.
There was no significant correlation between CO2 effluxes and soil
moisture, in part because in the felled area the moisture was high most of
the time (Fig. 2b), although low CO2 effluxes were generally associated
with high soil moisture (and lower temperatures) in the winter and the
highest effluxes were observed during low soil moisture periods in the
warmer temperatures of the summer, particularly in the unfelled area.
CO2 effluxes increased (p≤0.002) with soil temperature at the 2
and 10 cm depths at both areas, which was best described by exponential
correlation relationships as shown in Fig. 4 for the periods before and
after felling. Before felling, the CO2 efflux response to soil
temperature was not significantly different between area A and B, with
comparable Q10 values (3.77 and 3.16, respectively; Table 3). However, in
the years after felling, the Q10 values became much lower in the felled area B (mean Q10= 2.7) than in the unfelled area A (mean Q10= 4.23) with the
lowest Q10 value of 2.1 in year 3.
Annual apparent Q10 values for soil CO2 effluxes in
area A (mature spruce stand, remaining) and area B (mature spruce stand,
clear-felled after year 1) in Harwood Forest during the study period (as
defined in Table S1 in the Supplement).
A (unfelled) B (felled after year 1) YearR2SlopeQ10R2SlopeQ1010.8800.1333.770.6050.1153.1620.8660.1494.440.4860.0942.5730.7000.1203.300.3190.0722.0640.9730.1604.940.6540.1243.46
Exponential relationship between soil CO2 effluxes
and soil temperature at 2 cm depth measured from area A (mature spruce
stand; black) and area B (clear-fell site; red) before (dashed lines; year 1)
and after (solid lines; years 2–4) felling. Equations and R2 for fitted
lines shown.
The annual cumulative CO2 effluxes from the felled and unfelled areas
were not significantly different before felling with large overlap between
the 95 % confidence intervals (Fig. 5a), and the annual CO2 effluxes
were 19.8 and 24.0 t CO2e ha-1 yr-1 in areas A and B,
respectively (Table 4). After felling, however, there was a clear divergence
in the effluxes between the areas, with the CO2 efflux dropping sharply
in the felled area (B). In the first year after felling, the annual CO2
efflux reduced to its minimum value of 8.9 compared with 23.0 t CO2e ha-1 yr-1 from the unfelled area A. In years 3 and 4 the annual
effluxes in the felled area increased gradually to 11.3 and 12.2 t CO2e ha-1 yr-1, respectively, but was still lower than the unfelled
area (20.3 and 18.4 t CO2e ha-1 yr-1).
Annual soil GHG fluxes expressed as CO2 equivalents
(t CO2e ha-1 yr-1) using the IPCC global warming potential
value for each gas in area A (mature spruce stand, remaining) and area B
(mature spruce stand, clear-felled after year 1) in Harwood Forest. Values in
brackets are the % contribution to total CO2 equivalent.
Cumulative soil fluxes of CO2(a), CH4(b), and
N2O (c) in Harwood Forest from area A (black; mature spruce stand,
remaining) and area B (red; mature spruce stand, clear-felled after year 1)
during each year, calculated from median fluxes of eight replicated chambers.
Ribbons are the estimated 95 % confidence intervals.
CH4 fluxes
Soil fluxes of CH4 in both areas of the forest were generally low
throughout the study period, particularly before felling with no significant
differences (Fig. 2d). Fluxes were predominantly negative (i.e. removal from
the atmosphere) in unfelled area A and before felling in area B with a
median flux of -0.33 and -0.21 mg CH4-C m-2 d-1, respectively. After felling, area B became a significant
(p<0.001) source of CH4 and fluxes increased rapidly in the
following 2 years to its maximum in year 4 (2.48 mg CH4-C m-2 d-1) compared to the unfelled area, which remained unchanged with a
small CH4 sink (-0.33 mg CH4-C m-2 d-1).
Fluxes of CH4 varied more between the flux chambers after felling
particularly in year 3 and 4 (Fig. 5b). Although both soil moisture and
CH4 fluxes increased in the felled area after felling, the increased
fluxes and the variation between chambers cannot directly be related to the
soil moisture as no significant overall correlation was observed. In
addition, including interactions between soil moisture and temperature in
the statistical model did not better explain the variation in CH4
fluxes and difference between areas (Table 2). However, the analysis showed
that CH4 fluxes after felling were best explained by the soil
temperature (negative association) at both the 2 cm and 10 cm depths.
Annual fluxes (expressed as CO2 equivalents using the global warming
potential) of CH4 in the first period before felling were -0.038 and
-0.026 t CO2e ha-1 yr-1 from areas A and B, respectively
(Table 4). In the following years, the felled area (B) became a consistent
source of CH4 with 0.018, 0.194, and 0.335 t CO2e ha-1 yr-1 in year 2, 3 and 4, respectively. In contrast, the unfelled area
(A) remained a small CH4 sink with a mean flux value of -0.050 t CO2e ha-1 yr-1.
N2O fluxes
There were no significant differences in N2O fluxes between the two
areas before felling (Fig. 2e), although the median flux was higher in B
(0.33 and 0.55 mg N2O-N m-2 d-1 for A and B, respectively). Maximum N2O fluxes of 1.30 and 2.12 mg N2O-N m-2 d-1 were measured from area A and B,
respectively, between May and June 2014. After felling, N2O fluxes in
the felled area B became significantly (p=0.01) higher in year 2, 3, and
4 (median 1.83, 1.45, and 0.72 mg N2O-N m-2 d-1) than in the
unfelled area A (0.63, 0.39, and 0.28 mg N2O-N m-2 d-1) with
a maximum flux of 6.01 mg N2O-N m-2 d-1 measured in the first
year after felling in August 2015.
There were no significant correlations between N2O fluxes and soil
temperature before felling. However, after felling, N2O fluxes showed a
seasonal pattern that followed (p<0.001) that of soil temperature
at both depths in both areas with maximum fluxes during periods from June to
October. Soil moisture and its interactions with soil temperature (at 10 cm)
were the main driver for N2O fluxes before felling (p=0.002 and p=0.007, respectively). After felling, however, the soil moisture remained
high (Fig. 2b) and no direct effect on N2O fluxes was observed. This
could be due to the significant (p<0.01) negative correlation
between soil temperature (at both depths) and moisture before felling,
compared with after felling where the seasonal pattern in soil moisture was
less clear (Figs. 2b and 3) and no significant correlation occurred, so
that the soil temperature became the main driver of N2O fluxes.
There was a large variation between chambers in N2O fluxes throughout
the study period in both areas of the forest (evident in the confidence
intervals shown in Fig. 5c). Before felling, N2O annual fluxes
(expressed as CO2 equivalents) were 0.62 and 1.03 t CO2e ha-1 yr-1 in areas A and B, respectively (Table 4), a much smaller
contribution to the total GWP than the CO2 effluxes. After
felling, the annual fluxes of N2O increased and the highest annual
fluxes were measured from the felled area in the two consecutive years after
felling with 2.25 and 2.52 t CO2e ha-1 yr-1 in year 2 and 3,
respectively. However, at the end of the monitoring period in year 4, the
annual flux of N2O returned to a rate similar to that before felling
(1.24 t CO2e ha-1 yr-1).
DiscussionEffect of felling on CO2 effluxes
Felling and removal of the trees reduced soil CO2 effluxes by 55 %,
comparing the mean over the 3 years post-felling with the pre-felling value, or
47 %, comparing the clear-felled with the mature stand (Table 4).
Presumably this was a consequence of the reduction in autotrophic root and
rhizosphere respiration (e.g. Boone et al., 1998; Takakai et al., 2008), and
measurements about 5 months after felling showed a reduction from 4.9 to 1.6 t ha-1 in the live fine root mass (i.e. diameter
<2 mm). Living fine roots and their associated mycorrhizae can
contribute up to 59 % of total respiration (Ewel et al., 1987; Irvine and
Law, 2002; Subke et al., 2006), and for similar spruce stands in Harwood
Forest about 40 % contribution has been estimated previously
(Zerva and Mencuccini,
2005). However, CO2 effluxes might also increase directly after felling
due to an increase in decomposition of fine roots and associated
ectomycorrhizal biomass and litter. Therefore, the net CO2 efflux will
be determined by the balance between the reduction in autotrophic root and
rhizosphere respiration and the increased decomposition (Köster et al.,
2014), which can be short-lived depending on environmental factors such as
soil temperature and moisture (Davidson et al., 1998; Skopp et al.,
1990).
Before felling and in the unfelled stand, CO2 effluxes showed a strong
seasonal pattern that followed that of soil temperature with higher effluxes
during summer, as expected, when fine root density and plant growth activity
are highest. The apparent Q10 maximum values of these CO2 effluxes
(Table 3) were at the higher end of those reported from temperate deciduous
forests in the UK (1.60 to 3.92; Yamulki and
Morison, 2017 and 2.2; Fenn et al., 2010). The response of CO2 efflux to
temperature became weaker after felling (Fig. 4), even though the soil
temperature was substantially higher (a decrease in the apparent Q10 values
by up to 36 % over the 3-year period; Table 3). This agrees with the study
of Zerva and Mencuccini, (2005) who also observed a weaker association of CO2 effluxes with soil
temperature over a 10-month period after felling at another site in this
forest. They noted that this was probably because of the increased water
content (also evident in our study; Fig. 2b), the death of fine roots, and
the disturbance of the soil caused by tree harvesting and suggested that
autotrophic root respiration was more responsive to temperature than
heterotrophic microbial respiration. However, this effect could also be
because the apparent Q10 of soil CO2 efflux determined from field
measurements like these with trees present is influenced by the seasonal
changes in radiation and photosynthesis and their positive association with
seasonal temperature, rather than an altered temperature sensitivity after
felling. A larger reduction of 64 % in the Q10 value of soil CO2
efflux, compared to that found in this study, was reported from a larch
forest in eastern Siberia comparing a disturbed clear-cut site with a forest
site (2.1 compared with 5.9; Takakai et al., 2008). At the end of our study period in year 4, the Q10
value for soil CO2 efflux in the felled area (3.46) was close to that
before felling (3.16 in B-year 1), which could be due to ground vegetation
growth near but outside the gas flux chambers. It may also be a result of a
drop in the rainfall and soil moisture in the period between May and June in
year 4 (Figs. 2b, 3), so the response to soil temperature became
stronger, as in the period before felling. There may also have been a
recovery from any compaction effects during harvesting, as Epron et al. (2016) showed that compaction by timber forwarding machinery after
harvesting a French oak forest on a mineral soil decreased the Q10 values of
soil CO2 by 16 %–22 %. In contrast,
Kulmala et al. (2014) found that
the temperature dependency of the CO2 efflux was not affected by
clear-cutting of a Norway spruce forest on organo-mineral soil in southern
Finland.
The response of CO2 efflux to soil temperature is likely to have been
affected by the substantial increase in soil moisture after felling (Fig. 2b) as noted by Zerva and Mencuccini (2005) and
Kulmala et al. (2014). This
increase in soil moisture (with values frequently >0.6 cm3 cm-3) could have affected soil respiration by limiting the diffusion of
substrates and O2 to microorganisms (Skopp et al., 1990). Yamulki and
Morison (2017) could not detect an effect of soil moisture alone on soil
respiration in an oak forest in south-east England, but the combined model of soil
temperature and moisture explained 73 % of the CO2 efflux variations.
It is also possible that some of the CO2 produced may have dissolved in
the soil water and gone undetected
(Zerva and Mencuccini,
2005), but this effect was probably negligible here because of the low
solubility of CO2 at the low soil pH (3.8).
Effect of felling on CH4 fluxes
Clear-felling changed the soil from a small annual sink of CH4 to a
small net source over the 3-year monitoring period after felling, while the
unfelled stand remained a sink in all years. The shift in CH4 fluxes
from net uptake to net emissions by clear-felling has been reported in
several other studies: Zerva and Mencuccini (2005) from another site within
this spruce forest, Castro et al. (2000) from two slash pine plantations in
Florida, Takakai et al. (2008) from a Siberian larch forest soil, Yashiro et
al. (2008) from a tropical rain forest in Malaysia, Sundqvist et al. (2014)
from a forest site in central Sweden, and more recently by Korkiakoski et al. (2019) from a Scots pine nutrient-rich peatland forest in southern Finland. As
soil CH4 production requires anaerobic conditions (Conrad, 2007), this
shift from sink to source was probably caused by the substantial increase in
soil moisture (62 % higher) and soil temperature (6 ∘C) in the first
year after felling. An increase in soil moisture can increase the anaerobic
conditions that favour CH4 production by methanogenic archaea (e.g.
Sundqvist et al., 2014) and therefore can change the direction of the
CH4 flux. Generally, soil temperature and particularly moisture are
considered to be good predictors for CH4 behaviour (Lavoie et al.,
2013), but disentangling the two influences is difficult in field conditions.
In this study, CH4 fluxes and the flux variations between chambers
increased significantly in the following years after felling. The increase
in fluxes was modest in the first year after felling but more substantial in
the second and third years (Fig. 5b), which may reflect a time lag in the
microbial community changing and the fungal decomposition (Glassman et al.,
2018). Although the increase in fluxes might be attributed to the
substantial increase in the soil moisture, there was no direct correlation
between CH4 fluxes and soil moisture. Some previous studies have also
attributed the increase in CH4 fluxes after felling to an increase in
soil moisture or rise in the water table (Sundqvist et al., 2014; Zerva and
Mencuccini, 2005; Korkiakoski et al., 2019; Epron et al., 2016), while some
reported no direct correlations between CH4 fluxes and soil moisture
after felling (Lavoie et al.,
2013; Takakai et al., 2008; Wu et al., 2011; Mäkiranta et al., 2012;
Sundqvist et al., 2014; Zerva and Mencuccini, 2005). The lack of direct
correlation between CH4 fluxes and soil moisture has previously been
attributed to: (i) insensitivity of methanotrophic activity to small
variations in soil moisture and temperature (Sjögersten and Wookey,
2002; Peichl et al., 2010; Wu et al., 2011; Mäkiranta et al., 2012)
particularly when fluxes are relatively low such as in this study, (ii) CH4 being produced at a depth greater than that of the measured soil
moisture (Zerva and Mencuccini, 2005), and (iii) to other overriding
biological and physical factors (Lavoie et al., 2013).
Soil temperature determines CH4 fluxes by influencing methanogenic and
methanotrophic activity differently (Luo et al., 2013; Aronson et al.,
2013). Generally, CH4 consumption by methanotrophs is less responsive
to temperature than CH4 production by methanogens as consumption is
mainly limited by atmospheric CH4 diffusion (Dunfield et al., 1993;
Kruse et al., 1996). This is in line with the results here, as the
statistical analysis showed that soil temperature better explained CH4
fluxes after felling when it became a source than before when the soil was a
CH4 sink. However, soil temperature can be positively related to
CH4 uptake (Maljanen et al., 2003; Wu et al., 2011;
Ullah and Moore, 2011;
Yamulki and Morison, 2017), emissions
(Zerva and Mencuccini, 2005;
Dunfield et al.,
1993; Ullah and Moore, 2011), or can have no
correlation (Takakai et al., 2008, Sjögersten and Wookey, 2009; Lavoie
et al., 2013) depending on soil moisture and other factors that affect
microbial CH4 production and consumption. We found no significant
interactive effect between temperature and moisture on CH4 fluxes here.
Clear-felling can also affect other factors that play a key role in microbial
CH4 production or oxidation (and CO2 and N2O fluxes) by
increasing the substrate availability and soil N (NH4+ and
NO3-) concentrations (e.g. Bradford et al., 2000; Wang and Ineson,
2003), for example, as a result of N release from litter or brash, or from
reduced soil pH (Dalal and Allen, 2008). The higher bulk density in the
felled area (0.30 g m-3 compared with 0.22 g m-3), which may have been caused
by compaction as a result of the machinery traffic, may have contributed to
reduced CH4 uptake, increased CH4 production, and/or CH4
release (Teepe et al., 2004; Frey et al., 2011). However, the change in soil
substrates (organic matter and microorganisms) after felling was not
measured in this study and the differences between the unfelled and felled
area for total soil N content (1.95 % and 1.83 %, respectively) and soil
pH (3.6 and 3.8, respectively) were small.
Effect of felling on N2O fluxes
N2O fluxes (Fig. 2e) increased after felling and there was an
association with soil temperature at 10 cm depth (Table 2). Soil moisture
was a significant driver for N2O fluxes before felling, but it became
less significant after felling. The effect of clear-felling in increasing
N2O fluxes by similar magnitudes has also been shown from forests on
both mineral and peat soils (e.g. Saari et al., 2009; Zerva and Mencuccini,
2005; Mäkiranta et al., 2012; Pearson et al., 2012), but an order of
magnitude higher N2O flux was measured after clear-felling a
nutrient-rich drained peatland forest in southern Finland (Korkiakoski et
al., 2019).
N2O fluxes in temperate forest soil are generally expected to be low
because of the high C : N ratios in the litter and topsoil (Butterbach-Bahl
and Kiese, 2005; Jarvis et al., 2009) but can have a high spatial
variability because of the variability in the controlling environmental
factors (Peichl et al., 2010; Fest et al., 2009). In this study, both areas
of the forest showed large between-chamber variation in N2O fluxes
throughout the study period. This could be due to variations in soil
moisture within different flux chambers, particularly after rainfall, and
the variability between chambers in soil characteristics, litter amount,
mineral N availability, and microbial biomass, which were not measured. As
the soil moisture was consistently higher after felling (Table 1), the lack
of relationship with soil N2O fluxes may indicate that the moisture
content was not limiting for soil N2O production by the main microbial
nitrification and denitrification processes and that other factors were
responsible for N2O flux variations.
As mentioned previously, the total soil N content measured some 5 months
after felling was very similar to that in the unfelled area. However,
microbial N2O production in the following years after felling was
probably influenced by a declining slow N release into the soil from the
decomposition of fresh tree harvest residues and roots (Zerva and
Mencuccini, 2005; Yashiro et al., 2008; Saari et al., 2009) but stimulated
by warmer soil temperatures in the summer (Kulmala et al., 2014). This is
consistent with the significant relationship between N2O fluxes and
soil temperature, particularly at the 10 cm depth, which might imply that
N2O production was more prominent at the deeper, more anaerobic soil
depth, probably caused by denitrification brought about by an increased
respiratory sink for O2. In the absence of plants, there is no
competition for this newly available N, thereby maximizing substrate
availability for microbial N2O production and release (Skiba et al.,
2012). Three-fold higher N2O emissions from Finnish drained peatland
pine forest plots with logging residues have been
reported (Mäkiranta et al., 2012). Such effects of increasing soil
temperature combined with microbial activities and microbial biomass N in
increasing N2O fluxes have also been reported by other studies (Ineson,
et al., 1991; Smolander et al.,1998; Zerva and Mencuccini, 2005; Papen and
Butterbach-Bahl, 1999; Smith et al., 2018).
It is pertinent to mention here the study of Liimatainen et al. (2018) from
a range of afforested northern peat soils in Finland, Sweden, and Iceland,
where they suggested that high N2O fluxes were linked to availability
of peat phosphorus (P) and copper (Cu), which could be released with other
nutrients by harvesting from soil disturbance and brash (Rodgers et al.,
2010), and that low P and Cu concentrations can limit N2O production
even with sufficient N availability.
Ratios of annual soil GHG fluxes in area B (mature stand,
felled after year 1) to that in area A (mature spruce stand, remaining) in
Harwood Forest. Note: CH4 fluxes are presented as absolute values. The dashed
line represents ratio = 1 (no difference between areas).
Clear-fell harvesting effect on the GHG balance
Figure 6 summarizes the GHG flux changes by showing the ratio of the annual
soil GHG fluxes in B to that in A, with the assumption that the year-to-year
variation in the unfelled area A is an indicator equally of what conditions
would have been for the mature stand in B if it had not been felled. The
figure shows a clear annual GHG flux response to felling where: CO2
effluxes reduced directly after felling, increasing gradually thereafter;
CH4 effluxes increased sharply in year 3 and 4 after felling; and
N2O effluxes increased in year 2 and 3 after felling but decreased
thereafter. In order to compare the contributions of each gas to the total
GHG soil flux and the effect on emissions and therefore radiative balance,
we expressed the fluxes in CO2 equivalents as normally done (using
their global warming potential values (Table 4). Before felling, the total
flux (sum of CO2, CH4, and N2O emissions in t CO2e ha-1) was 20.4 in the A area and 25.0 in the B area and CO2 effluxes
dominated the total GWP, contributing up to 97 %. The total flux dropped
in the first and second year after felling to approximately half (44 % and
56 % lower annual effluxes, respectively). The contribution of CO2 to
the total GHG flux expressed as CO2 equivalents in the unfelled area
remained constant at about 98 % throughout the study period but decreased
in the felled area to ca. 80 % in the first and second year after felling.
This was due to the doubling of N2O flux, which contributed up to 20 %
of the total flux in the two years after felling. Although the felled site
became a source of CH4, its contribution to the total flux was always
small (<2 %). For the same periods, the contributions of N2O
and CH4 to the total flux in the unfelled area were small and remained
constant at 2.9 % and -0.2 %, respectively, similar to their values in
year 1. In the last year, N2O annual efflux in the felled area halved
to 1.2 t CO2e ha-1, still 20 % higher than before felling, but the
efflux of CH4 continued to increase to 0.34 t CO2e ha-1. Over
the 3 years since felling, the total soil GHG flux was reduced by 45 %
(from 25.0 to 13.8 t CO2e ha-1) due to the much larger reduction
in soil CO2 efflux than the increases in N2O and CH4 fluxes.
In the unfelled area there was a reduction of approximately 8 %,
presumably due to changing weather conditions.
This study is one of very few longer-term assessments of the impacts of
clear-fell harvesting on the GHG balance of a Sitka spruce forest. However,
because of the limitations of the periodic, manual, closed-chamber measurement
technique used here, it does not take into account the daily temporal flux
variations or cover fluxes from ground occupied by brash or stumps.
Simultaneous eddy covariance (EC) measurements (Xenakis et al., 2021) over
the 3-year period after clear-felling at this study area showed approximately
3 times higher ecosystem respiration (32.8 t CO2 ha-1 yr-1)
than our estimated mean annual soil CO2 efflux (10.8 t CO2 ha-1 yr-1). As also suggested by Zerva
and Mencuccini (2005), the difference presumably indicates high CO2
effluxes coming from the brash mats and stumps, which could not be measured by the small chambers
used in this study, plus above- and belowground respiration of the colonizing
vegetation, which was not included in the chambers. The differences could
also be due to spatial heterogeneity over the site as the soil flux chambers
are only partially sampling ground that is representative of the EC
footprint. Moreover, it has been indicated that peat decomposition after
felling is stimulated by nutrient release from brash (Vanguelova et al.,
2010) and therefore higher N2O and CH4 fluxes may also be expected
as a result of increasing mineral N release and the presence of more labile
organic matter, respectively, from areas with brash. Mäkiranta et al. (2012) observed over 3-fold increase in seasonal average (over 3 years
between May–October) soil N2O flux and 2-fold in CO2 efflux from
plots with logging residues than without but observed no change in CH4
emissions. This indicates that brash removal post-harvesting (e.g. for
biofuel as suggested by Mäkiranta et al., 2012) might be a way of
limiting GHG effluxes from peat decomposition. More information on GHG
fluxes from brash and stumps and the underlying soil processes that might
be influenced by felling are priorities for future research.
Conclusions
In this upland Sitka spruce plantation on organic-rich peaty gley soil,
clear-fell harvesting affected soil GHG fluxes by increasing soil temperature
and moisture content and reducing fine root mass, which affected the soil
nutrient and organic C supply and associated microbial populations,
activities, and decomposition rates. Although soil moisture increased
significantly after felling, there were no direct correlations with the soil
GHG fluxes, probably because there was limited variation in the high soil
moisture after felling. By contrast, there was a good correlation between
GHG fluxes and the soil temperature, which exhibited much larger temporal
variation. This study does not take into account fluxes from brash
decomposition because of the small flux chamber areas; therefore, our
total measured soil GHG efflux after felling probably underestimates that of
the site as a whole. Over the 3-year measurement period after felling, soil
CO2 effluxes reduced substantially (55 %) due to cessation of root
respiration outweighing increased decomposition. For the same period,
CH4 fluxes changed from a small net sink to a net source, increasing
throughout, and N2O fluxes increased substantially in the 2 years after
felling. Mean soil CO2, CH4, and N2O fluxes over the 3-year
period after felling contributed 83 %, 1 %, and 16 %, respectively, to total
GHG flux on a CO2 equivalents basis with an overall reduction of
45 % due to much larger soil CO2 flux reduction than the combined
soil CH4 and N2O flux increases.
Code availability
Codes have been uploaded to Zenodo, available at: 10.5281/zenodo.5081968 (Yamulki et al., 2021a).
Data availability
Data have been uploaded to Zenodo, available at: 10.5281/zenodo.5082059 (Yamulki et al., 2021b).
The supplement related to this article is available online at: https://doi.org/10.5194/bg-18-4227-2021-supplement.
Author contributions
SY, JILM, GX, and MP designed the study. GX,
AA, and SY carried out the gas sampling and soil environmental measurements.
JB carried out the GC analyses and SY did all the data analyses. JF was
responsible for statistical analysis and SY prepared the manuscript with
contributions from JILM and GX.
Competing interests
The authors declare that they have no conflict of interest.
Disclaimer
Publisher’s note: Copernicus Publications remains neutral with regard to jurisdictional claims in published maps and institutional affiliations.
Acknowledgements
We would like to thank Forest Research
colleagues Elena Vanguelova, Ed Eaton, Sue Benham, and Vladimir Krivtsov for
soil and vegetation sampling and analysis. We are indebted to the Forestry
England staff who gave permission for the study in Harwood Forest and for
their huge support throughout, particularly the late Jonathan Farries. We
also thank Russell Anderson for his valuable comments on the paper.
Financial support
The project was funded by the Forestry Commission and partly by the UKRI Natural Environment Research Council GREENHOUSE project (grant no. NE/K002619/1).
Review statement
This paper was edited by Ben Bond-Lamberty and reviewed by two anonymous referees.
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