Introduction
Greenhouse gas (GHG) emissions to the atmosphere have increased significantly
since pre-industrial times as a direct result of human activities, such as
fossil fuel burning, cement production and land use changes (IPCC, 2013). The
Intergovernmental Panel on Climate Change (IPCC) have estimated in their
Fifth Assessment Report (AR5) that around one third of all anthropogenic
emissions of carbon dioxide (CO2) for the period 1750–2011, were caused
by land use changes (IPCC, 2013). From 2000 to 2009, the Agriculture,
Forestry and Other Land Use (AFOLU) sector accounted for 24 % of all
global GHG emissions (around 10 Gt CO2-eq yr-1), with emissions
from peatland drainage and burning alone estimated at around
0.9 Gt CO2-eq yr-1.
Natural (i.e. undrained) peatlands function as long-term carbon (C) stores as
the sequestration of CO2 over time is greater than the amount of C that
is emitted from the peatland as methane (CH4) and leached in waterborne
exports (Roulet et al., 2007; Nilsson et al., 2008; Koehler et al., 2011;
Gažovič et al., 2013). Key to this role is the position of the water
table, which largely dictates the rate of decomposition within the peatland.
When the water table is positioned close to the peat surface, the breakdown
and degradation of organic matter typically proceeds very slowly in the
absence of oxygen. As a consequence, there is an accumulation of peat (and C
within; Dise, 2009).
In the Republic of Ireland (ROI) and the United Kingdom (UK), peat has been
extracted for energy use for many centuries (Chapman et al., 2003; Renou et
al., 2006). Traditionally, this involved the manual removal of the peat, i.e.
hand cutting; however, this has been largely superseded by highly mechanised
methods to extract the peat for both energy and horticulture requirements. In
the ROI, over 4 million tonnes of peat per annum are industrially extracted
from approximately 50 000 ha to provide ca. 5.5 % of primary energy
requirements (Howley et al., 2012) and for use in horticulture. A further 0.4
million tonnes per year is likely burned for domestic heating (Duffy et al.,
2014) and may impact as much as 600 000 ha of peatlands (Wilson et al.,
2013b). Although peat extraction areas in the UK have generally declined over
the last few decades, approximately 0.8 million tonnes of peat is still
extracted each year in England and Scotland (Webb et al., 2014), although it
is UK Government policy to phase out peat extraction in England by 2030
(Department of Environment Food and Rural Affairs, 2011). Peat extraction
areas in Wales are small (482 ha) and have remained unchanged in the
1991–2010 period (Webb et al., 2014). In Northern Ireland, the area of
peatland utilised for fuel (mechanical and hand cutting) has declined
considerably in the 1990–2008 period, although a slight increase in the
areas used for horticulture have been recorded (Tomlinson, 2010).
In industrial peatlands, the extraction of peat is facilitated by the
installation of drainage ditches at regular (typically 15–30 m) intervals
across the peatland. For peat used for horticultural purposes, the more
fibrous upper layers (e.g. Sphagnum peat) are extracted and
utilised. If the peat is to be used for energy production the more highly
decomposed peat is milled, dried in the production fields and removed for
immediate use or stockpiled for later requirements. Peat extraction ceases
for energy production when either the sub-peat mineral soil is reached, large
quantities of fossilised timber are encountered or drainage is no longer
practical (Farrell and Doyle, 2003). For peatlands used for the provision of
domestic heating, the peat is either removed by a digger from the margins of
peatlands, placed in a tractor-mounted hopper and extruded onto the surface
of the peatland, or the peat is extruded onto the surface of the peatland
from openings made in the peat by a chain cutter. Over a period of weeks the
peat is dried in situ and removed from the site. The effect of peat
extraction on the hydrological functioning is marked by a large fall in the
water level either throughout the peatland (industrial) or at the margins of
the peatland (domestic). In the latter, significant water level drawdown is
also experienced further inward towards the centre of the peatland (Schouten,
2002).
The impact of drainage on C cycling in peatlands has been widely documented.
In general, a lowering of the water table leads to increased CO2
emissions (Silvola et al., 1996; Salm et al., 2012; Haddaway et al., 2014) as
the aerobic layer is deepened and mineralisation rates are accentuated.
Concurrently, CH4 emissions (with the exception of emissions from ditches) may
decrease or cease (Salm et al., 2012; Turetsky et al., 2014), waterborne C
exports may increase (Strack et al., 2008; Evans et al., 2015) and there may
be a heightened risk of C loss through fire (Turetsky et al., 2015). In the
case of peat extraction, C cycling may be further altered by the removal of
vegetation (Waddington and Price, 2000), and losses of windblown particulate
organic carbon (POC) may be exacerbated from the bare peat surfaces (Lindsay,
2010).
Under the United Nations Framework Convention on Climate Change (UNFCCC) and
the Kyoto Protocol, “Annex 1” countries (i.e. countries that have committed
to targets that limit or reduce emissions) are obligated to prepare annual
National Inventory Reports (NIR) and up-to-date annual inventories, detailing
GHG emissions and removals from six different sectors. Emissions associated
with off-site peat combustion are reported under the Energy sector and are
not considered further here. The recent IPCC Wetlands Supplement (IPCC, 2014)
to the 2006 Good Practice Guidance (GPG; IPCC, 2006) derived new Tier 1
emission factors (EFs) for drained organic soils that differentiated between
on-site emissions (e.g. CO2-Con-site, fire) and off-site
losses (e.g. leaching of waterborne C). In the case of peatlands managed for
extraction in the temperate climate zone, the CO2-Con-site
values have increased from 0.2 (nutrient-poor, bogs) and 1.1 (nutrient-rich,
fens) t CO2-C ha-1 yr-1 in the 2006 GPG to a single higher
EF of 2.8 t CO2-C ha-1 yr-1 (covering the entire boreal
and temperate regions) in the Wetlands Supplement. On-site burning directly
consumes aboveground C stocks (prescribed and wildfire burning) and the
underlying peat C store (wildfire burning) and rapidly releases both gases
(e.g. CO2, CH4) and particulates (e.g. black carbon) to the
atmosphere. In the Wetlands Supplement, an EF for GHG emissions from
prescribed fire on drained peatlands is not provided due to a paucity of
published data at present. However, emissions from wildfires are addressed
and EFs of 362, 9 and 207 g kg-1 dry fuel burned are provided for
CO2-C, CH4 and CO respectively with a proviso that they were
derived from a very small data set.
Given the relatively large areas under peat extraction in both the ROI and
the UK, a move from Tier 1 to higher reporting levels is desirable,
particularly as (a) a wide range in uncertainty is associated with the IPCC
Tier 1 values (1.1–4.2 t CO2-C ha-1 yr-1), which reflects
the disparity in emissions from drained peatlands from different climate
zones and nutrient composition, (b) the most recently published annual
CO2 flux estimates (not included in the derivation of IPCC Tier 1
values) also display a very wide amplitude (cf. Järveoja et al., 2012;
Mander et al., 2012; Salm et al., 2012; Strack et al., 2014), (c) no data
from ROI or UK peatlands were included in the IPCC derivation, which might
mean that the Tier 1 value may not be appropriate for these countries, and
(d) no distinction is made between industrial or domestic extraction sites,
despite large differences in their drainage, vegetation cover and management
characteristics. In addition, previous studies of peatland fire EFs have
focused on the boreal peatlands of Alaska (Yokelson et al., 1997) and Canada
(Stockwell et al., 2014) and the temperate peatlands of Minnesota (Yokelson
et al., 1997) and North Carolina (Stockwell et al., 2014). These studies
found that the smouldering combustion of peats associated with low combustion
efficiency leads to relatively lower CO2 emissions (compared with other
ecosystems) and much higher carbon monoxide (CO), CH4, and other
non-CH4 hydrocarbon emissions. Therefore, it is important to quantify
emissions of these gases as they include strong GHGs (e.g. CH4) and
reactive gases responsible for tropospheric ozone formation and poor air
quality (e.g. CO, ammonia (NH3), hydrogen cyanide (HCN)).
The objectives of the study are (1) to provide estimates of the annual
CO2-C exchange (i.e. CO2-Con-site) for nine peat
extraction sites in the ROI and the UK, (2) to derive regional-specific
CO2-C EFs for drained peat extraction areas that would permit the ROI
and the UK to progress to the Tier 2 reporting level, (3) analyse the factors
that influence CO2-C dynamics in this region (i.e. land use, climate,
etc.), and (4) to report GHG emissions associated with the burning of Irish
Sphagnum moss peat in the first laboratory study to investigate fire
emissions from European temperate peats.
Materials and methods
Study sites
The study sites were located at nine peat extraction areas in the ROI and the
UK with a history of either industrial peat (IP) or domestic peat (DP)
extraction (Table 1). Boora (IP1), Blackwater (IP2), Bellacorick (IP3),
Turraun (IP4), Middlemuir Moss (IP5) and Little Woolden Hall Moss (IP6) are
industrial cutaway peatlands where significant areas of bare peat (i.e.
unvegetated microsites) have remained following the cessation of milled peat
extraction. At IP6, milled peat is currently extracted from areas close
(< 150 m) to the study site. The IP sites are former raised bogs
with the exception of IP3, which is a former Atlantic blanket bog. At all
sites, the drainage ditches have remained functional. Here we define
“drained” as a mean annual water table position deeper than -20 cm
(Couwenberg and Fritze, 2012; Strack et al., 2014). Physico-chemical
characteristics of all the sites are detailed in Table 1.
At Clara (DP1), Glenlahan (DP2) and Moyarwood (DP3) the peat has been
extracted from the margins of the sites for use in domestic heating. In the
case of Clara, peat extraction was an ongoing activity at the time of our
study despite the designation of the site as a Special Area of Conservation
(SAC). DP1 and DP3 are raised bogs and DP2 is a mountain blanket bog. The
vegetation component at all the sites is species poor and is composed mainly
of ling heather (Calluna vulgaris), cross-leaved heather
(Erica tetralix) and lichens (Cladonia spp.) A continuous
water table level was not observed at DP2, as the relatively shallow peat
deposit (∼ 40 cm) over bedrock at that site was prone to drying out at
various times throughout the study.
Site characteristics. Mean annual air temperature (∘C) and
mean annual rainfall (mm yr-1) are long-term values (1981–2010; Met
Éireann, http://www.met.ie/, and Met Office UK,
http://www.metoffice.gov.uk).
Site name
Boora
Blackwater
Bellacorick
Turraun
Middlemuir
Little
Clara
Glenlahan
Moyarwood
Moss
Woolden
Site code
IP1
IP2
IP3
IP4
IP5
IP6
DP1
DP2
DP3
Time since last
> 20 years
> 25 years
> 10 years
> 30 years
> 10 years
ca. 1
0
> 20 years
> 20 years
extraction*
Study period
1/9/2007–
1/5/2011–
1/1/2012–
1/1/2002–
1/11/2003–
1/1/2013–
1/4/2006–
1/4/2006–
1/4/2013–
30/8/2009
30/4/2014
31/12/2013
31/12/2003
31/10/2004
31/12/2014
31/3/2007
31/3/2007
31/3/2014
Latitude
53.203
53.297
54.128
53.260
57.60
53.451
53.316
53.103
53.346
Longitude
-7.726
-7.965
-9.556
-7.720
-2.15
-2.468
-7.647
-7.538
-8.514
Subregion
Irish
Irish
North-west
Irish
North-east
Northern
Irish
Irish
Western
Midlands
Midlands
Ireland
Midlands
Scotland
England
Midlands
Midlands
Ireland
Mean annual air
9.3
9.8
10.3
9.3
8.0
10.2
9.3
9.3
10.0
temperature (∘C)
Mean annual
970
907
1245
807
851
867
970
804
1193
rainfall (mm yr-1)
Vegetation
Calluna vulgaris, Erica tetralix, Cladonia sp.
Peat type
Phragmites
Phragmites
Cyperaceous
Phragmites
Sphagnum/
Sphagnum/
Sphagnum
Ericaceous
Sphagnum
cyperaceous
cyperaceous
Von Post scale
H7
H7
H5 to 6
H7
H8
H6 to 7
H6
H6
H6
Parent material
Limestone
Limestone
Shale
Limestone
Granite drifts
Triassic
Limestone
Old Red Sandstone
Limestone
and rocks
sandstone
Peat depth (m)
1.0
1.5
0.5
0.5–1.8
0.7–3.1
0.5–1.75
4
0.4
4.4
pH
4.3
4.9
3.8
6.3
3.6–4.1
2.9
4.0
3.8
4.4
C (%)
50
52.4
56
52
52
49.1
49.8
29.1
51.5
N (%)
1.09
2.14
0.97
2.1
1.4
1.34
1.46
0.69
1.32
C : N
45.9
24.5
57.7
24.8
37
36.6
34.1
42.2
39
* Time between cessation of peat extraction and
the study period.
Climatic conditions
All the sites are located within the temperate zone as defined by IPCC (2006)
and are characterised by an oceanic climate with prevailing south-west winds,
mild mean annual air temperatures (8 to 10.3 ∘C) and moderate to
high annual rainfall (804 to 1245 mm; Table 1).
Environmental monitoring
At each site, three to nine aluminium square collars (60 × 60 cm)
were inserted to a depth of 30 cm into the peat. At IP6, smaller circular
plastic collars were used (15 cm diameter) to facilitate the use of the
CPY-4 chamber (PP Systems, UK) at that site. Soil loggers (μ logger,
Zeta-tec, UK; HOBO external data loggers, Onset Computer Corporation, MA,
USA; Comark N2012 Diligence loggers, Norwich, UK) were established at all the
IP sites and recorded soil temperatures (∘C) at hourly intervals.
Weather stations were installed at all the DP sites and recorded
photosynthetic photon flux density (PPFD;
µmol m-2 s-1) and soil temperatures (5 and 10 cm
depths) at 10 min intervals. At DP3, soil volumetric moisture content (VMC,
%) was also recorded (at 10 min intervals) by the weather station at that
site. At sites IP5 and IP6, soil temperature was only measured manually
during CO2 flux measurements. In order to estimate soil temperature at
times where data was lacking at these two sites, a regression-based approach
between manually recorded T5 cm and air temperature recorded at
15 min intervals by a logger at the site was used to gap-fill the data
(r2= 88.7 %). The water table level (WT) was manually measured from
dipwells (internal diameter 2 cm) inserted adjacent to each collar. Wooden
boardwalks were established at each site (except at IP6).
Leaf area index (LAI)
At the IP sites, the vegetation had been removed prior to the commencement of
peat extraction and virtually no natural recolonisation has taken place
following the cessation of peat extraction. However, at the DP sites a
vegetation component was present and in order to incorporate the seasonal
dynamics of the plants into CO2-C exchange models, the leaf area index
(LAI) was estimated for each of the collars. This involved accounting for the
green photosynthetic area of all vascular plants (leaves and stems) within
the collar at monthly intervals. In short, the number of leaves and stems
were counted from five subplots (8 × 8 cm) within each collar. The
size (length, width) of the leaves was measured from sample plants outside
the collars. The LAI was then calculated by multiplying the estimated number
of leaves by an area estimate of the leaf. Moss and lichen percentage cover
was estimated at the same time. Species-specific model curves were applied to
describe the phenological dynamics of the vegetation of each collar, and the
models (vascular plants and moss) were summed to produce a plot-specific LAI.
For a detailed description of the method, see Wilson et al. (2007). At site
DP1 only, the vegetation was removed by regular clipping from one third of
the collars, in order to provide an estimate of the heterotrophic
contribution (RH) to ecosystem respiration (Reco).
On-site carbon dioxide flux estimation
Field measurements
At sites IP1–5 and DP1–3, Reco was measured with a static
polycarbonate chamber (60 × 60 × 33 cm) equipped with two
internal fans to ensure mixing of the air within the chamber and a cooling
system (submerged ice packs and water pumped to a radiator located within the
chamber) to maintain the temperature within the chamber close to the ambient
air temperature (for a more detailed description, see Alm et al., 2007b). At
IP6, Reco was measured with a CPY-4 (PP Systems, UK) clear
acrylic chamber (14.6 cm diameter, 14.5 cm height). The CPY-4 chamber was
equipped with an internal fan, PPFD sensor and thermistor. Sampling was
carried out at fortnightly or monthly (winter) intervals (two to four
measurements per collar per measurement day). For each Reco flux
measurement, the chamber was placed in a water-filled channel at the top of
the collar or connected with a rubber gasket (IP5), covered with an opaque
cover, and the CO2 concentration (ppmv) in the chamber headspace was
measured at 15 s (5 s at IP6) intervals over a period of 60–180 s using a
portable CO2 analyser (EGM-4, PP Systems, UK). Concurrently, air
temperature ( ∘C) within the chamber and soil temperatures at 5, 10
and 20 cm depths were recorded at each collar (soil temperature probe, ELE
International, UK). The WT position relative to the soil surface was manually
measured with a water level probe (Eijkelkamp Agrisearch Equipment, The
Netherlands). At the DP sites, net ecosystem exchange (NEE) was measured with
the same polycarbonate chambers described above under a range of ambient
light levels (PPFD; µmol m-2 s-1) prior to
Reco measurements. NEE measurements were carried out between
08:00 and 18:00 GMT in the summer and between 09:00 and
15:00 GMT in the winter
(three to eight measurements per collar per measurement day) to ensure that
the maximum PPFD was reached at each measurement date. Artificial shading was
used in the early morning to obtain low PPFD levels
(< 100 µmol m-2 s-1). PPFD was recorded from a
sensor (PAR-1, PP Systems) located within the chamber. The portable CO2
analysers were regularly calibrated with a CO2 standard gas.
Flux calculations
Flux rates (mg CO2-C m-2 h-1) were calculated as the linear
slope of the CO2 concentration in the chamber headspace over time, with
respect to the chamber volume, collar area and air temperature. A flux was
accepted if the coefficient of determination (r2) was at least 0.90. An
exception was made in cases where the flux was close to 0 (mainly in
wintertime when soil processes are typically slower) and the r2 is
always low (Alm et al., 2007b). In these cases the flux data were examined
graphically, and fluxes with obvious non-linearity (due to chamber leakage,
fan malfunction, etc.) were discarded. The remainder were accepted provided
that some of the environmental variables measured at the same time (e.g. soil
temperature) were sufficiently low to account for the low flux values (Wilson
et al., 2013a). In this study, we followed the sign convention whereby
positive values indicated a CO2-C flux from the peatland to the
atmosphere (source) and negative values indicated a flux from the atmosphere
to the peatland (sink). Gross primary production (GPP) was calculated as NEE
minus Reco (Alm et al., 2007b), and the closest Reco
flux value in time to a NEE flux value was used.
Modelling
Statistical and physiological response models (Alm et al., 2007b) were
constructed and parameterised for each study site. Model evaluation was based
on the following criteria; (a) statistically significant model parameters
(p < 0.05), (b) lowest possible standard error of the model
parameters and (c) highest possible coefficient of determination (adjusted
r2; see Laine et al., 2009). The basic Reco models, based
upon the Arrhenius equation (Lloyd and Taylor, 1994), are non-linear models
related to soil temperature. GPP was related to PPFD using the
Michaelis–Menten-type relationship that describes the saturating response of
photosynthesis to light (Tuittila et al., 1999). GPP model coefficients and
associated standard errors were estimated using the Levenberg–Marquardt
multiple non-linear regression technique (IBM SPSS Statistics for Windows,
Version 21.0. Armonk, NY, USA). During model construction, the relationship
between Reco or GPP and a range of independent environmental
variables (recorded in conjunction with flux measurements) was tested. Only
variables that increased the explanatory power of the model (i.e. improved
r2 values) were included. The models were accepted if the residuals were
evenly scattered around 0.
Annual CO2-C balance
The response functions estimated for Reco and GPP were used for
the reconstruction of the annual CO2-C balance. Recofluxes were
reconstructed for each collar in combination with an hourly time series of
(1) T5 cm, (2) VMC (at DP3) recorded by the data loggers or (3) WT
depths linearly interpolated from weekly measurements. The annual
CO2-C balance (g C m-2 yr-1) was calculated for each sample
plot by integrating the hourly Reco values over each 12-month
period. (Note: integration periods vary between study sites; see Table 1.) At
the DP sites, GPP was reconstructed in combination with (1) PPFD values
recorded by the weather station, (2) plot-specific modelled LAI and (3) an
hourly time series of T5 cm (DP1only). At the DP sites, annual NEE
was calculated as annual GPP plus annual Reco.
Statistical analysis
The CO2-C flux data (Reco for the IP sites and
Reco and GPP for the DP sites) had a non-normal distribution, so
the non-parametric Kruskal–Wallis (p= 0.05) and Mann–Whitney tests were
used to test for differences between sites. Uncertainty in reconstructed
annual Reco and GPP was calculated by summing up the maximum and
minimum standard errors associated with each of the model parameters (e.g.
Drösler, 2005; Elsgaard et al., 2012; Renou-Wilson et al., 2014).
Uncertainty in the annual Reco or NEE estimate was calculated
following the law of error propagation as the square root of the sum of the
squared standard errors of GPP and Reco (IPCC, 2006).
Peat fire emissions
Around 5 kg (dry mass) of loose Irish Sphagnum moss peat (H2–H3 on
the von Post decomposition scale) was used for measuring fire EFs. Subsamples
of the peat were taken and placed in a 22 × 12 × 10 cm
open-topped insulated chamber. The chamber was constructed from lightweight
Celcon insulation blocks and was used to replicate natural surface combustion
conditions, leaving only one surface of the peat exposed to open air and
thereby reducing heat loss and oxygen exchange from the other surfaces, in
accordance with the suggested peat combustion methodology of Rein et
al. (2009). Each sample was dried in an oven overnight at 60 ∘C. In
order to produce comparable replicates, the samples for the burning
experiment had to be dried to an absolute dry base to increase ignition
probability (Frandsen, 1997) and encourage pyrolysis (Rein et al., 2009).
Following drying, the chamber and sample were placed in a fume cupboard under
controlled airflow conditions and the peat was ignited using a coiled
nichrome wire heated to ∼ 600 ∘C and placed in contact with
the surface of the peat. This also best represents natural ignition
conditions (e.g. from a surface shrub fire), also in accordance with the
methodology of Rein et al. (2009). Once ignited, each 1 kg sample proceeded
to burn for ∼ 90 min. The resulting smoke was continuously sampled
using a pump and a 90 cm sample line with a funnel held ∼ 12 cm above
the smouldering peat. The smoke was sampled into an 8.5 L infrared White
(multipass) cell (Infrared Analysis, Inc.) where infrared spectra were
collected using a Fourier transform infrared (FTIR) spectrometer. Analysis of
the FTIR spectra was performed using the Multi-Atmospheric Layer Transmission
(MALT) software (Griffith, 1996), yielding trace gas mole fractions inside
the White cell, from which emissions factors may be calculated. A full
description of how EFs may be calculated from FTIR measurements of gas mole
fractions is given in Paton-Walsh et al. (2014) and Smith et al. (2014). Here
we use the C mass balance approach to calculate EFs for CO2 and CO
(Eq. (1) in Paton-Walsh et al., 2014). The C content of the peat (required
for calculating EFs via the C mass balance approach) is assumed to be
53.3 %, as measured in Scottish Sphagnum moss peat (Cancellieri
et al., 2012). For all other gas species considered in the study (CH4,
ethylene (C2H4) ethane (C2H6), methanol (CH3OH),
HCN, NH3), we use their respective emission ratios to CO and the EF for
CO to calculate EFs (via Eq. (5) in Paton-Walsh et al., 2014).
Combustion efficiency is a measure of the amount of fuel carbon released as
CO2 and may be approximated using the modified combustion efficiency
(MCE) formula, which requires only a measurement of CO and CO2 rather
than all the C-containing gases (Yokelson et al., 2008):
MCE=ΔCO2ΔCO2+ΔCO
where ΔCO2 and ΔCO represent the elevated mixing ratios
of these gases (the difference between mixing ratios measured in biomass burn
emissions and those in the ambient air). MCE is often expressed as a
percentage. Generally, an MCE lower than 0.9 (90 %) is considered a
low-combustion-efficiency burn (Lobert et al., 1991; Yokelson et al., 1996).
Discussion
There is a very wide range in reported CO2 emissions from both active
and abandoned peat extraction areas in the literature (Fig. 7). Much of
this variation can be attributed to differences in climate, drainage level,
peat type, peat extraction methods and the end use of the peat and, as such,
provides a useful framework to examine the variations in this study.
Mean modified combustion efficiency (MCE) and emission factors
(g kg-1 dry fuel burned) reported by this study and those for the same
trace gases reported by previous studies of temperate or boreal peat
(Yokelson et al. 1997; Stockwell et al. 2014). The mean and standard
deviation of the emission factor is calculated from individual sample burns;
nr: not reported.
Emission factor (g kg-1 dry fuel burned)
Trace gas
Irish Sphagnum
Canadian
North Carolina
Alaska and Minnesota
moss peat
boreal peat
temperate peat
peat
(this study)
Stockwell et al. (2014)
Stockwell et al. (2014)
Yokelson et al. (1997)
MCE
0.837 ± 0.019
0.805 ± 0.009
0.726 ± 0.009
0.809 ± 0.033
CO2
1346 ± 31
1274 ± 19
1066 ± 287
1395 ± 52
CO
218 ± 22
197 ± 9
276 ± 139
209 ± 68
CH4
8.35 ± 1.3
6.25 ± 2.17
10.9 ± 5.3
6.85 ± 5.66
C2H4
1.74 ± 0.23
0.81 ± 0.29
1.27 ± 0.51
1.37 ± 0.51
C2H6
1.53 ± 0.17
nr
nr
nr
CH3OH
0.60 ± 0.87
0.75 ± 0.35
2.83 ± 2.87
4.04 ± 3.43
HCN
2.21 ± 0.35
1.77 ± 0.55
4.45 ± 3.02
5.09 ± 5.64
NH3
0.73 ± 0.50
2.21 ± 0.24
1.87 ± 0.37
8.76 ± 13.76
Relationship between (a) ecosystem respiration (Reco:
t CO2-C ha-1 yr-1) and mean soil temperature (∘C)
at 5 cm depth at the IP sites and (b) net ecosystem exchange (NEE:
t CO2-C ha-1 yr-1) and leaf area index (LAI: m2 m-2). Circles indicate an annual value.
Effects of climate
While the study sites in this paper are all located within the temperate
zone, considerable variation in CO2-C emissions was evident. Given that
all the sites are drained to a similar depth (Fig. 1), it is not surprising
that the variation in emissions appeared to be controlled largely by
differences in soil temperatures between the sites (Fig. 6). The coldest site
in terms of mean soil temperatures and lowest in terms of annual emissions
was Muirhead Moss (IP5) in north-eastern Scotland. Although rainfall and site
water table levels were similar to the other sites, soil temperatures at this
site remained below 0 ∘C for a high proportion (∼ 14 %) of
the year and are likely to have resulted in a slowdown of extracellular
enzymatic diffusion (Davidson and Janssens, 2006), reduced microbial activity
(Fenner et al., 2005) and consequently lower rates of CO2 production
(Basiliko et al., 2007). Indeed, it is likely that our value of
0.93 t CO2-C ha-1 yr-1 at this site may be an
overestimation given that it was calculated from monthly mean values that
were measured during daytime hours (highest daily temperatures). As much of
the peatlands in Scotland fall within the same temperature regime (Chapman
and Thurlow, 1998), CO2-C emissions data from a wider range of peat
extraction sites in this region might significantly refine our EF derivation.
At the other end of the spectrum, the highest emissions and soil temperatures
were observed at Turraun (IP4) in the Irish Midlands. Data from this site had
been previously reported by Wilson et al. (2007). In this study, we only
utilised CO2-C flux data from plots where the mean annual water table
position was deeper than -20 cm. This resulted in a higher mean value
(taken over 2 years) in this current study. Three of the IP sites in the ROI
are located in the Midlands where more “extremes” in climate are generally
experienced (lower winter temperatures, higher summer temperatures) than
along the western coast (IP3). However, during this study, winter
temperatures at all the ROI sites seldom decreased below 0 ∘C
(Fig. 3) and the proportion of hourly temperatures higher than 20 ∘C
was somewhat similar between the sites. Although, Little Woolden Moss (IP6)
received the lowest annual rainfall of all sites in year 1 of the study at
that site (Fig. 1), mean annual soil temperatures were in the mid-range of
the nine study sites, hourly T5 cm values were normally
distributed (Fig. 3) and CO2-Con site emissions were close
to the derived EF value of 1.70 t CO2-C ha-1 yr-1
(Table 2).
Carbon dioxide emissions (t CO2-C ha-1 yr-1) from
peatlands managed for extraction in Canada, the ROI and the UK (this study)
and Fennoscandia. The 10th and 90th percentile are indicated by the bars, the
25th and 75th percentiles with the top and bottom of the box and the median
value by the centre line. (Data for Canada and Fennoscandia taken from the
following studies; Tuittila and Komulainen, 1995; Sundh et al., 2000;
Waddington et al., 2002; Glatzel et al., 2003; McNeil and Waddington, 2003;
Tuittila et al., 2004; Cleary et al., 2005; Alm et al., 2007a; Shurpali et
al., 2008; Waddington et al., 2010; Järveoja et al., 2012; Mander et al.,
2012; Salm et al., 2012; Strack et al., 2014.) Where studies reported
seasonal fluxes (typically May to October), these were converted to annual
fluxes by assuming that 15 % of the flux occurs in the non-growing season
(Saarnio et al., 2007).
The DP sites are all located in the ROI and within a 35 km radius, but
considerable variation in annual rainfall was apparent during this study
(Fig. 1), with DP3 (the furthest west) receiving the highest rainfall of all
sites in the study (on average 34 % more rainfall than at the other DP
sites). The east–west rainfall gradient in the ROI is well documented and
coincides with a change in peatland types (i.e. raised bogs to Atlantic
blanket bogs). This climatic variation is reflected in the annual
Reco values, which were similar at DP1 and DP2 but much lower at
DP3 (Fig. 5). There is an established relationship between rainfall amount
and the moisture content of peat (Price and Schlotzhauer, 1999; Strack and
Price, 2009). For the sites located in high-rainfall areas, such as DP3,
there may be a suppression of aerobic microbial activity within the peat
matrix, and as a consequence Reco values may be lower than would
be expected for a drained peat soil. Indeed, at some of these sites, occult
precipitation (e.g. dew and fog droplets) may also contribute significantly
to higher levels of soil moisture (Lindsay et al.,
2014. During the
growing season, the transpiration process is also likely to play a role in
determining the moisture content of the peat within the rooting zone
(∼ 20 cm depth) at these vegetated sites. Moisture losses are likely
to be accentuated on sunny days when air and soil temperatures are high, when
LAI values are highest (midsummer) and when vapour pressure deficit is not a
limiting factor. As CO2 emissions were closely correlated to soil
temperature at 5 cm depth, reduced moisture content in this zone is likely
to stimulate aerobic microbial activity. Annual GPP showed a similar trend to
annual Reco at these vegetated DP sites. GPP is strongly
controlled by the amount of light received by the plants (i.e. PPFD levels
and LAI) and the efficiency with which the plants use it. PPFD values (data
not shown) and the vegetation composition were broadly similar during the
sampling periods, which would seem to indicate that LAI is the driver of both
productivity and therefore NEE at these sites (Fig. 6). However, variations
in LAI are likely to be the result of subtle differences in a number of other
variables (e.g. nutrient status, site management) that were not captured in
our measurements.
Effects of drainage level
While a close relationship between WT position and CO2-C emissions has
been established in some peatland studies (Silvola et al., 1996; Blodau and
Moore, 2003; Blodau et al., 2004), soil temperature proved to be the
strongest determinant of CO2-Con-site emissions at our sites
and this relationship has also been observed by other studies in peat
extraction areas (e.g. Shurpali et al., 2008; Mander et al., 2012; Salm et
al., 2012). While the addition of WT or VMC improved the performance of the
Reco models at some of the sites, the improvement was only slight
and this is likely due to the fairly narrow range of WT and VMC values
recorded over the course of the 12-month study (e.g. the range in VMC values
at DP3 was between 56 and 64 %). Therefore, optimum WT and VMC
levels for respiration may not have been encountered. The Reco
models used here are only valid for the data that were measured over the
course of the study at each site and cannot be readily extrapolated beyond
the range of that data. For those sites where the water table did not appear
to influence Reco dynamics it may be that fluctuations in WT
level were missed with the interpolation approach and CO2-C flux
measurement regimes that we employed here, although these methodologies have
been widely used elsewhere (Riutta et al., 2007; Soini et al., 2010;
Renou-Wilson et al., 2014). Instead, it is probable that our results reflect
the complexity of the relationship between Reco and WT in very
dry soils as outlined by Lafleur et al. (2005), where factors such as a
stable, low surface soil moisture content, and decreased porosity (i.e.
limited oxygen availability) at the depths at which the WT is mainly located
ensure that when CO2-C fluxes are measured, the WT is deeper than the
zone where it has a discernible impact on Reco (Juszczak et al.,
2013). As such, the soil temperature regime at these sites may act as a
“proxy” for drainage level (i.e. higher soil temperatures are likely to
occur in conjunction with deeper water table levels and vice versa;
Mäkiranta et al., 2009).
Peat characteristics
Industrial peat extraction involves the removal of surface vegetation and
results in the exposure of decomposed peat at the surface. The level of
decomposition in the peat is related to depth and as extraction proceeds, the
more highly decomposed peat is exposed. The peat at industrial extraction
sites tends to have a lower aerobic CO2 production potential than at
natural sites, for example, due to differences in substrate and nutrient
availability, a more extreme physical environment (Glatzel et al., 2004) and
reduced labile organic matter supply in the absence of plant communities
(i.e. priming). In our study, the C content (with the exception of DP2) was
similar across all sites (Table 1). Although, Glatzel et al. (2004) noted
that CO2 production was negatively correlated with the von Post scale of
decomposition, no correlation with annual CO2-C emissions was evident in
our study (p > 0.05). Similarly, despite obvious differences
in nitrogen content and pH values between IP sites, no relationships with
CO2 fluxes were discerned. However, the residual peat at IP4 is strongly
influenced by the close proximity of limestone parent material, as evidenced
by high pH values and the lowest C : N ratio (Table 1), and is highly
minerotrophic. Given the high CO2-C emissions associated with this site,
consideration should be given to disaggregation by nutrient type should more
data become available in the future.
Organic matter quality has been closely linked to the soil respiration rate,
with lower emission rates associated with the poorer-quality organic matter
found at depth in drained peatlands (Leifeld et al., 2012). The lowest
emissions at our sites occurred where the residual peat was either of
Cyperaceous (IP3) or Sphagnum/cyperaceous (IP5) origin. However,
while the slow decomposition rate of Sphagnum litter in comparison
to other plant litter has been well documented (Verhoeven and Toth, 1995;
Bragazza et al., 2007), there is insufficient data from our study sites to
determine whether the limited relationship observed here between peat type
and CO2-C emissions at our study sites is coincidental rather than
causal.
Effects of peat extraction methods and peat end use
For peat utilised for horticulture, the more fibrous peat layers nearer the
surface are extracted. This may result in the oxidation of more labile
organic matter and may account for the very high emissions associated with
Canadian peatlands, for example (Fig. 7), in comparison to countries where
the deeper peat layers are extracted (Mander et al., 2012). However, the IP
sites in this study are highly decomposed peat and have been abandoned for 30
years or more in some cases (e.g. IP4) and have remained unvegetated. It is
possible that CO2-C emissions from active extraction areas may be higher
than those derived in this study given that over the summer period the
surface of the peat is regularly scarified and aerated. However, Salm et
al. (2012) reported higher emissions from abandoned areas in comparison to
active areas, although colonisation by vegetation in the former may have
accentuated respiration losses. High annual CO2-C emissions following
abandonment and recolonisation have also been reported by Strack and
Zuback (2013) and are in close agreement with the Reco values
reported here for the DP sites (Fig. 5).
We have estimated the contribution of heterotrophic respiration
(RH) to Reco at 49 %. Although this is based on
measurements at a single site (DP1), it is within the range reported by other
studies (Frolking et al., 2002; Moore et al., 2002; Shurpali et al., 2008).
The RH values measured at DP1 (Fig. 5) and estimated at DP2 are
higher than the Reco values at the IP sites, which would indicate
that decomposition of the belowground biomass (following clipping) and
subsequent “priming” effects may contribute significantly to CO2-C
dynamics at vegetated extraction sites. Furthermore, the methods employed to
extract the peat at some of the DP sites (the peat is extruded onto the
surface of the peatland from narrow openings made in the peat by a chain
cutter) has led to the formation of deep fissures (ca. 4 cm wide and
> 2 m deep) within the peat that may enhance oxidation
throughout the peat profile. Nonetheless, fissures (ca. 10 cm wide and
> 1 m deep) that formed in the peat during climatically dry
years and that were partially filled in during wetter and windier years were
also observed at IP5, where the lowest annual emissions were observed.
Fire emission factors
The mean MCE reported here (0.837) is typical of smouldering combustion (e.g.
Yokelson et al., 1996; Bertschi et al., 2003) and comparable with the
reported range of MCE in other studies of high-latitude peats (Yokelson et
al., 1997; Stockwell et al., 2014). Emission factors for CO2 and CO are
also typical of smouldering combustion and similar to those from other peat
studies, particularly Yokelson et al. (1997). As found in other studies of
peat fire emissions, our measurements confirm that the CH4 EF for Irish
peat is particularly high (8.35 g kg-1 dry fuel burned) when compared
with other forms of biomass burning. Given the high global warming potential,
where each gram of emitted CH4 is equivalent to 34 g of CO2
(100-year time horizon, IPCC, 2013), the CH4 emissions from Irish peat
fires may account for over 12 % of the CO2-equivalent emissions.
This result emphasises the importance of understanding the full suite of
trace gas emissions from biomass burning rather than focussing solely on
CO2 and CH4 emissions. In general, the other EFs reported here lie
within the range of variability observed by other peat burning studies, with
the exception of NH3, which is particularly low, possibly as a result of
the nitrogen-poor soils that are typical of Irish and UK blanket bogs. Here,
we also report the first C2H6 EF for peat
(1.53 ± 0.17 g kg-1 dry fuel burned), similar in magnitude to
C2H6 emissions from boreal forests (1.77 g kg-1 dry fuel
burned), according to Akagi et al. (2011). The use of prescribed
fire in the UK to burn off old heather growth to encourage new growth (e.g.
the muirburn practice) may not impact the underlying peat to any great
extent, given that the practice is restricted to the October–April period
when soil moisture conditions are highest. Emissions result from the burning
of the woody aboveground biomass, and the underlying peat is generally
unaffected. In contrast, wildfires typically occur during the summer months
when temperatures are highest and moisture levels are low, resulting in the
burning of both the vegetation and the peat itself. Indeed, recent work by
Kettridge et al. (2015) has highlighted the vulnerability of drained
peatlands, even at high latitudes, to increased risk of wildfire and
subsequent vegetation changes.
Implications for National Inventory reporting
The ROI currently employs the 2006 GPG default value of
0.2 t CO2-C ha-1 (nutrient-poor) in reporting of all peat
extraction areas, and estimated emissions for 2012 (the most recent
assessment year) were 9312 t CO2-C yr-1 (Table 4). In contrast,
the approach in the UK has been to differentiate between peat extracted for
fuel and horticulture and then applying the default EFs for nutrient-rich
(1.1 t CO2-C ha-1) and nutrient-poor peat
(0.2 t CO2-C ha-1) respectively. For 2012, CO2-C emissions
from UK extraction peatlands were estimated at 2118 t CO2-C yr-1
(Table 4).
Reported annual emissions are likely to increase considerably if the Tier 1
values in the IPCC Wetlands Supplement are adopted by inventory compilers. We
estimate that emissions from peatlands managed for extraction will be
approximately 16 and 10 times higher for the ROI and UK respectively
(Table 4). The EFs derived in this study for CO2-Con site
for both industrial and domestic peatlands (Table 2) are considerably lower
than the Tier 1 value of 2.8 t CO2-C ha-1 yr-1 provided in
the IPCC Wetlands Supplement (2014). Although the EFs derived in this study
fall within the lower confidence margin of the Tier 1 range, our new EFs have
a marked reduction in associated uncertainty. As the Tier 1 is a generic
value based on published literature rather than a targeted measurement
programme, it is naturally subject to a certain level of bias, which results
when the underlying studies are not representative of management practices,
climatic zones, or soil types in a particular region (Ogle et al., 2004) and
may lead to either an over- or underestimation of CO2-C emissions. Given
that no significant difference exists between the EFs derived for the IP and
DP sites in this study, we propose a single EF for
CO2-Con-site of 1.68 t CO2-C ha-1 yr-1 to
be applied to peatlands managed for extraction in the ROI and UK regardless
of peat type. This EF value could be further disaggregated by regional
climate, domestic peat extraction intensity (based on extraction rates) or by
the end use of the peat (horticulture or energy) if more data become
available. For the latter, it would be highly useful to determine
quantitatively whether CO2-Con-site emissions vary between the
less decomposed residual peat utilised for horticulture and the more
decomposed residual peat used for energy production. As the EFs derived in
this study have come from sites located within the same “climatic” region,
we feel that they are more appropriate for the ROI and the UK inventory
purposes than either the 2006 GPG or the 2013 Wetlands Supplement. If the
CO2-Con site EFs derived from this study are used in annual
NIRs, we estimate that annual emissions would be 9.5 and 6 times higher for
the ROI and UK respectively, in comparison to the emissions calculated with
the 2006 GPG Tier 1 value, and 40 % lower than emissions calculated with
the Wetlands Supplement EF.
Annual CO2-C emissions (in tonnes of CO2-C yr-1)
from peatlands managed for extraction in the ROI and UK calculated using the
IPCC 2006 Good Practice Guidance (Tier 1 value: 0.2 and
1.1 t CO2-C ha-1 yr-1 for nutrient-poor and nutrient-rich
peatlands respectively), the IPCC 2013 Wetlands Supplement (Tier 1 value: 2.8
t CO2-C ha-1 yr-1) and the Emission Factors derived in this
study (Table 2). Areas (ha) and CO2-C emissions using the IPCC 2006 Good
Practice Guidance values are taken from the 2014 National Inventory Reports
(NIR) for the ROI (Duffy et al., 2014) and the UK (Webb et al., 2014).
Country
Area (ha)
Emissions (tonnes CO2-C yr-1)
IPCC (2006)
IPCC (2013)
This study
ROI
52 422
9312
146 782
88 069
England
4790
960
13 412
8047
Scotland
1610
545
4508
2705
Wales
482
95
1350
810
N. Ireland
1030
518
2884
1730
UK
7912
2118
22 154
13 292
As reported CO2-Con-site emissions are henceforth likely to be
much higher for any country that moves from the 2006 GPG to the 2013 Wetlands
Supplement, some consideration of potential mitigation measures is required.
Wetland Drainage and Rewetting is a new elective activity under Article 3.4
of the Kyoto Protocol (second commitment period) and applies to all lands
that have been drained since 1990 and to all lands that have been rewetted
since 1990. Countries that elect to report under this activity will also be
able to claim C benefits from the rewetting of drained peatlands. In theory,
this should provide an impetus for the rewetting of highly emitting land use
categories such as peatlands managed for extraction, particularly as these
areas will remain persistent long-term emission hotspots in the absence of
rewetting actions (Waddington et al., 2002).
Information gaps
Greenhouse gas emissions from peatlands used for extraction are composed of
(a) on-site emissions (i.e. from peat extraction areas, ditches and
stockpiles) and (b) off-site emissions associated with waterborne losses and
the use of the peat for energy or horticulture. In this paper, we have
focused solely on the on-site CO2-C emissions from the peat extraction
areas and GHG emissions from fire. However, C losses from other pathways may
also be substantial. Research has shown that GHG emissions from on-site peat
stockpiles and ditches are considerable (Alm et al., 2007a, and references
therein). Currently, emissions data from stockpiles in the temperate zone are
not available and the IPCC Wetlands Supplement does not provide a Tier 1
value and instead encourages countries to move to higher tiers in terms of
reporting (IPCC, 2014). However, countries such as Finland have developed a
Tier 2 approach in which EFs (including CH4 and N2O) depend on regional
weather and in which emissions from ditches and stockpiles are taken into
account (Alm et al., 2007a; Lapveteläinen et al., 2007). The IPCC
Wetlands Supplement provides Tier 1 EFs for CH4 emissions from both peat
extraction areas and from ditches. The value for the latter is particularly
high (542 kg CH4 ha-1 yr-1 expressed per unit area of ditch
surface) and indicates the importance of this pathway in the full GHG balance
(Evans et al., 2015). Similarly, N2O emissions have been shown to be
significant from drained peatlands (Regina et al., 1996); yet despite this,
there are only a small number of published studies and more research is
critical in order to provide regionally specific EFs. While CH4 and
N2O fluxes have been quantified at some of the sites in this study, the data are
currently being processed with a view to publication in the future. In terms
of the fire study, N2O is a difficult gas to measure using the FTIR
set-up employed in this study, as it can only be determined from spectra with
very large enhancements of trace gases. This is because the N2O
absorption occurs in a similar wave number region to both the CO2 and CO
absorption bands (Paton-Walsh et al., 2014). Paton-Walsh et al. (2014) could
only determine N2O from two of their five open fires, whilst Smith et
al. (2014), who used a similar set-up, failed to determine N2O from any
of their 21 fires studied. In our study, we found that excess mole fractions
of N2O could not be correlated to either CO2 or CO for the
determination of emission ratios, precluding the calculation of EFs. One
explanation for this is that N2O is predominantly a product of flaming
combustion and is strongly correlated to CO2 (Paton-Walsh et al., 2014).
The lack of flaming combustion in our peat burns probably explains our
inability to detect significant excess N2O mole fractions.
Other pathways may be of equal importance. For example, the loss of POC from
bare peat surfaces may be considerable where the surface is exposed and
subject to wind or water erosion (Evans et al., 2006; Lindsay, 2010). While
some of the windborne POC is likely to be deposited within the extraction
field itself, a proportion undoubtedly leaves the peatland, although there
are currently few data available to quantify losses from either wind or water
erosion or the extent to which POC is converted to CO2 (IPCC, 2014). In
addition, high losses of DOC from drained peatlands have been reported (Evans
et al., 2015, and references therein). Although a Tier 1 EF value for DOC is
provided in the IPCC Wetlands Supplement, disaggregated by climate zone, with
the assumption that 90 % of the exported DOC is converted to CO2,
there is an obvious need to quantify these losses on a regional basis given
the high precipitation loads experienced by the ROI and the UK and the
associated differences in peat type (Evans et al., 2015). Emissions from
burning are not currently reported in either the ROI or UK inventory reports.
The EF provided in the IPCC Wetlands Supplement for CO2 emissions
associated with wildfire burning is similar to our value here (Table 3).
Furthermore, given the high CH4 emissions associated with the burning of
the peat that we have reported here (Table 3), and taking cognisance of the
strong global warming potential of
CH4, more research is urgently required to quantify this emission
pathway, particularly under field conditions.
The provision of activity data for inventory reporting varies between the ROI
and the UK, with the peat extraction industry the source of data in the
former (Duffy et al., 2014), and a multi-source approach (Directory of Mines
and Quarries, point locations with Google Earth imagery, scientific
reports/papers) used in the latter (Webb et al., 2014). However, CO2
emissions from domestic peat extraction in the ROI are not currently reported
due to a lack of activity data and could potentially be very high (Wilson et
al., 2013b). In the UK, areas under domestic extraction are included in the
Grassland category but may be moved as the UK considers changes post-Wetlands
Supplement. Determining to what degree peatlands have been affected by
domestic peat extraction and how far those impacts extend into the main
peatland area are obvious challenges facing future research. The use of
remote-sensing platforms could provide high-resolution data that will be able
to differentiate between domestic peat extraction and other types of
disturbed peatlands. In particular, the use of unmanned aerial vehicles (i.e.
drones), which have been used to map individual peatlands at a very high
resolution (e.g. Knoth et al., 2013), offer considerable potential for more
detailed mapping of domestic peatlands on the national scale.