Boreal peatlands represent a globally important store of
carbon, and disturbances such as wildfire can have a negative feedback to
the climate. Understanding how carbon exchange and greenhouse gas (GHG)
dynamics are impacted after a wildfire is important, especially as boreal
peatlands may be vulnerable to changes in wildfire regime under a rapidly
changing climate. However, given this vulnerability, there is very little in the
literature on the impact such fires have on methane (CH4) emissions.
This study investigated the effect of wildfire on CH4 emissions at a
boreal fen near Fort McMurray, Alberta, Canada, that was partially burned by the
Horse River Wildfire in 2016. We measured CH4 emissions and
environmental variables (2017–2018) and CH4 production potential (2018)
in two different microform types (hummocks and hollows) across a peat burn
severity gradient (unburned (UB), moderately burned (MB), and severely burned
(SB)). Results indicated a switch in the typical understanding of boreal
peatland CH4 emissions. For example, emissions were significantly lower
in the MB and SB hollows in both years compared to UB hollows.
Interestingly, across the burned sites, hummocks had higher fluxes in 2017
than hollows at the MB and SB sites. We found typically higher emissions at
the UB site where the water table was close to the surface. However, at the
burned sites, no relationship was found between CH4 emissions and water
table, even under similar hydrological conditions. There was also
significantly higher CH4 production potential from the UB site than the
burned sites. The reduction in CH4 emissions and production in the
hollows at burned sites highlights the sensitivity of hollows to fire,
removing labile organic material for potential methanogenesis. The
previously demonstrated resistance of hummocks to fire also results in
limited impact on CH4 emissions and likely faster recovery to pre-fire
rates. Given the potential initial net cooling effect resulting from a
reduction in CH4 emissions, it is important that the radiative effect
of all GHGs following wildfire across peatlands is taken into account.
Introduction
Northern peatlands are an important component of the global carbon (C)
cycle, acting as long-term sinks of atmospheric carbon dioxide (CO2).
They are also large sources of methane (CH4) (Bridgham
et al., 2013), with northern peatlands contributing approximately 40–155 Tg to global CH4 emissions (Neef et al.,
2010; Turetsky et al., 2014). CH4 dynamics in peatlands results from
a combination of various biogeochemical processes (Lai, 2009).
Controls on CH4 production, oxidation, and emissions include
microtopography (Cresto Aleina et al., 2016), water
table depth (Bubier et al., 1995; Granberg et al., 1997),
soil temperature (Granberg et al.,
1997; Saarnio et al., 1998), substrate quality and availability
(Granberg et al., 1997; Segers, 1998; Joabsson et
al., 1999), and vegetation cover (Ström et
al., 2005; Strack et al., 2017).
However, disturbances such as wildfire can have a significant impact on the
magnitude of C fluxes across peatlands (fire can release between 10 and 85 kg C m-2 through combustion and smouldering; Turetsky et al.,
2011), potentially causing a negative feedback to the climate
(Randerson et al., 2006).
Western boreal Canada is undergoing increasing pressure from wildfire, with
fire extent and frequency expected to double by the end of this century
(Benscoter et al., 2005;
Flannigan et al., 2008). Understanding how ecosystem C cycling and
greenhouse gas (GHG) dynamics are impacted after a wildfire is important,
especially as boreal peatlands may be vulnerable to changes in wildfire
regime under a rapidly changing climate (Flannigan et
al., 2008). Fire can remove surface vegetation, increasing net radiation at
the ground surface (Brown et
al., 2015), and can “re-set” vegetation communities back to the primary
succession stage (Johnstone, 2006;
Benscoter and Vitt, 2008). Fire can alter soil organic matter quality in
the soil column (Neff et
al., 2005; Olefeldt et al., 2013a) and reduce belowground C stores in
peatlands (Wilkinson et al., 2018). Overall, wildfire
can lead to a decrease in C accumulation rate through combustion loss,
reduction in vegetation productivity, and increased organic matter
decomposition post-fire (Ingram et
al., 2019; Robinson and Moore, 2000; Wieder et al., 2009). Furthermore,
increased ash deposition after wildfire can increase soil pH (Molina et al.,
2007; Davies et al., 2013) and change the physical characteristics of the
soil, including blocking of peat macropores and altering hydrology (Noble et
al., 2017). Reduction in vegetation cover and soil organic matter can also
lead to drier conditions across peatlands (Tarnocai, 2009; Thompson and
Waddington, 2013; Kettridge et al., 2015), with the drop in water table
level causing an increase in the aerobic zone
(Waddington et al., 2015). This
could lead to a reduction in CH4 emissions or even uptake of CH4
via oxidation (Strack et al., 2004; Turetsky et
al., 2008; Moore et al., 2011). Conversely, high water tables can occur
post-fire (Kettridge et al., 2015), although they are often associated
with low surface moisture contents due to hydrophobicity of the peat
(Doerr et al., 2000). Low
soil moisture rates can also occur under increased ash deposition after
fire, with increased closure of soil pores by ash causing reduced capacity
to hold water and increased runoff (Heydari et al., 2017).
Microtopography (microforms) across peatlands can be impacted through fire
by increasing the prominence of hollows (low-lying areas close to the water
table; Belyea and Clymo, 1998) on the landscape through altering elevation
(Benscoter et al., 2015), and often hollows will
have a higher severity of burn compared to other areas across the landscapes
(Mayner et al., 2018; Benscoter et al.,
2005). Conversely, hummocks (mounded microtopography, approximately 0.2 m or
higher above the water table; Belyea and Clymo, 1998) are generally
resistant to fire, namely due to moisture retention differences between the
different moss species present at both microform types, as Sphagnum spp. is much
more resilient to fire than feather moss (Kettridge et al.,
2015).
Despite the increasing pressures from wildfires across northern peatlands, a
knowledge gap still persists on CH4 emissions after wildfire,
especially in boreal regions. In a study on the impact of wildfire on
methanotrophic communities from an ombrotrophic peat bog,
Danilova et al. (2015) found a reduction in
the activity of the methanotrophs in burned sites 7 years post-fire. This
reduction following wildfire could therefore lead to a potential increase in
CH4 emissions from bog systems. Grau-andrés et
al. (2019) also showed an increase in CH4 emissions at an ombrotrophic
bog in the UK 1 year after a prescribed fire, most likely due to increased
graminoid coverage. Conversely, studies at other bog sites in the UK report
a decrease in emissions after fire (Ward et al.,
2007; Davies et al., 2013). In non-peatland ecosystems across boreal
regions, wildfire has been shown to cause an increase in CH4 uptake
(Burke et al., 1997; Song et al.,
2017, 2018). However, in permafrost zones, wildfires can often typically
lead to substantial permafrost thaw and increasing moisture levels across
the landscape (Gibson et al., 2018), potentially leading to an
increase in CH4 emissions (Kim and Tanaka, 2003;
Turetsky et al., 2008; Olefeldt et al., 2013b; Helbig et al., 2017).
However, Köster et al. (2017, 2018) found an increase in CH4 uptake
across continuous permafrost sites in both Canada and Russia after fire. To
date, we cannot find any reference on the impact of fire on CH4
emissions across fens, despite being the dominant peatland type in western
boreal Canada (Vitt et al., 2000).
Therefore, the objectives of this study are to (i) determine the impact of
wildfire on fen CH4 emissions across a peat burn severity gradient; (ii) evaluate the controls on CH4 emissions within each site; and (iii) examine CH4 production potential across a peat burn severity gradient.
We hypothesize that CH4 emissions and production will be lower at
burned sites due to lowering of the water table, changes in substrate
availability, and reduction in vegetation cover.
MethodsStudy site and collar locations
The study was undertaken in a treed moderate-rich fen (hereafter referred to
as Poplar Fen), 20 km north of Fort McMurray, Alberta, Canada (56∘56.330′ N, 111∘32.934′ W), which was partially burned by the Horse
River Wildfire in 2016. The mean annual temperature (1981–2010) is 1 ∘C and the mean annual precipitation is approximately 420 mm
(Environment Canada, 2017). This treed fen is dominated by Larix laricina (Du Roi)
K.Koch, Picea mariana (Mill.) Britton, Betula pumila (L.), Equisetum fluviatile (L.), Smilacina trifolia (L.) Sloboda, Carex spp., and Sphagnum fuscum (Schimp.) Klinggr and
brown mosses, largely Tomenthypnum nitens (Hedwig) Loeske. All vegetation was identified to the
species level, with nomenclature for vascular plants and mosses as per the Flora of
North America Editorial Committee (1993). Average peat depth ranges between
1 and 1.5 m. The landscape consists of approximately 47 % hummocks and
53 % hollows within the fen area (Gabrielli, 2016). This site was split
into three sections along a peat burn severity gradient as assessed by depth
of burn (DOB). The unburned (UB) site was situated within the fen interior
and was not affected by the wildfire. The moderately burned (MB) site had a
DOB of approximately 9–11 cm and the severely burned (SB) site had a DOB
of approximately 14.5–17 cm. Both burned sites were situated on the
peatland margin, closer to the adjacent upland at a slightly higher
elevation. The DOB was determined following the protocols used by
Lukenbach et al. (2015a). In summary, this
method assumes a pre-fire flat surface between multiple reference points
across the site, including adventitious roots in the burned sites and
unburned reference points. A string is attached between two reference points
and 10 measurements were taken along the length, from string to burned
ground surface, giving an estimate of the depth of the burn. At each site,
PVC collars (height 15 cm × diameter 20 cm) were placed in both hollows and
hummocks. A total of four replicate collars were placed at each microform
location at each site to a depth of approximately 15 cm in spring 2017,
totaling eight collars at the UB site, eight collars at the MB site, and
eight collars at the SB site. The height of the collar was measured from the
soil surface in order to have the correct chamber headspace volume for
CH4 emission calculations.
Environmental conditions
Water table (WT) depth (relative to the ground surface) was measured
adjacent to each pair of collars at all three sites. A PVC pipe (4 cm
(diameter) × 100 cm (length)) fully slotted along the full length and
covered in mesh was used. A soil temperature (ST) profile was collected at
each collar during each CH4 measurement at -30, -25, -20, -15, -10, -5, and
-2 cm from the ground surface. Percentage cover of plant functional type
(bryophyte and overstory graminoid and dwarf shrub) as well as bare
ground, burned material, and standing water were estimated from photographs
taken at peak growing season during 2018 to produce percentage cover
estimates of the flux collars.
Measurements of field CH4 emissions
Methane fluxes were measured using the closed chamber method, 8 times
between 7 May and 16 August 2017 and 14 times between
11 May and 16 August 2018. A cylindrical opaque chamber (20 cm × 50 cm) was placed on the collar, with water poured around the collar edge
to create a seal. A battery-powered fan was used to mix the chamber
headspace. A thermocouple located within the chamber, attached to a
thermometer, was used to measure temperature during sampling. A 20 mL syringe
was used to collect gas samples at intervals of 7, 15, 25, and 35 min
following chamber closure and injected into Exetainers (Labco, UK). A gas
chromatograph (GC; Shimadzu GC2014, Mandel Scientific, Canada) with a flame
ionization detector (250 ∘C), helium gas carrier, and standards of
5 and 50 ppm was used to determine CH4 concentration of the gas samples
collected during the field seasons. The emissions were determined from the
linear change in concentration over time, which includes corrections for
temperature and volume of the chamber. Any small negative or positive values
in which the change in concentration did not exceed 10 % (precision of
concentration analysis) were assigned a zero-emission value. Large negative
emissions (<-5 mg CH4 m2 d-1) were removed from the
analysis as it is unlikely this system would have consumption rates as large
as this and were likely caused by disturbance during chamber placement.
These procedures resulted in a total loss of 6 % of the data across both
years. In order to determine whether emissions measured from the UB site
were representative of emissions from Poplar Fen as a whole, we also
compared our fluxes to a previous study of CH4 emissions collected
between 2011 and 2014 at the fen.
Potential CH4 production
Peat samples were collected adjacent to each collar (2–3 m away to avoid
disturbing the collar) on 13 August 2018 and immediately shipped
to the laboratory and frozen until analyzed. Peat samples were obtained
using a tin can to a depth of 20 cm from the ground surface (two samples
were collected per sampling location: 0–10 and 10–20 cm).
Potential CH4 production was determined under anaerobic conditions
following a similar methodology to Strack et al. (2004).
Peat slurries of approximately 20 g of wet peat were made in 250 mL
incubation jars. Distilled water was added to the sample to saturate,
without allowing for standing water. Samples were flushed with N2 for
15 min and then sealed. Slurries were incubated at room temperature
(approximately 20 ∘C) and sampled at 0, 24, and 48 h and then twice
weekly between 16 October 2018 and 2 November 2018 (total incubation
length of 17 d). Samples were agitated by hand before sampling commenced
to mix the gases within the peat pore spaces and the jar headspace.
Samples (10 mL) were extracted from the jars and injected into a Fast
Methane Analyzer (FMA; Los Gatos, USA). A 10 mL sample of N2 was
replaced in each jar after the sample was extracted to maintain headspace
pressure. Potential methane production was determined from the linear
increase in CH4 concentration within the jars over the incubation
period after correcting for dilution by N2 (Strack et
al., 2017). Gravimetric soil moisture (GWC) was determined by weighing
subsamples of peat not used in the incubation, drying the samples at
60 ∘C for 2–3 d, and reweighing. Organic matter content was
determined by loss on ignition (LOI), burning samples at 550 ∘C
for 4 h.
Data analysis
All statistical analysis was undertaken in R (R Core Team, 2013) using the
package nlme (Pinheiro et al., 2018), and all output and models were inspected
for normality and homogeneity of residuals (Zuur et al., 2009). Data were
log transformed if required, with a value of 10 being added prior to
transformation to account for zeros and negative values in the dataset.
Seasonal mean values of CH4 flux and associated environmental variables
at each collar were used in all model levels (Treat et al., 2007;
Turetsky et al., 2014), with statistical significance considered at
α=0.05.
In order to evaluate the effect of burn severity on CH4 flux, a linear
mixed effects model (LMM) was used with burn severity, microform, the
two-way interactions between these, and year as fixed factors, with collar
ID as a random effect to take into account repeated measures (Pinheiro et
al., 2018). Another LMM was used to evaluate the environmental controls on
CH4 flux, with the effect of burn severity, WT, ST at 30 cm depth, and
the two-way interactions between each included as fixed effects. If
significant factors were found, Tukey pairwise comparisons were completed
using the lsmeans package (Lenth, 2016). Any insignificant factors were
removed from the model until the final model was found. An individual
insignificant factor was kept in the model if its interaction with another
factor was significant. The amount of variance explained by the model
(RGLMM2) was calculated using the method described by Nakagawa
and Schielzith (2013).
A two-way analysis of variance (ANOVA) was used to test for differences
between CH4 production rate, burn severity, and microform. No
significant difference between sampling depths was found, and thus depths were
combined in further analyses. Again, if significant differences were found,
Tukey pairwise comparisons were completed.
ResultsEnvironmental and vegetation variables
Water table depth was linked to microtopographic position, with hollows
having the highest WT position across all sites. The deepest WT depth (± standard deviation) during both 2017 and 2018 was found at the SB hummocks,
approximately -45±5.6 and -44±6.0 cm below the surface,
respectively (Table 1). The shallowest WT was found at the unburned hollows,
being -8.6±4.1 and -6.9±1.8 cm below the surface in 2017 and
2018, respectively (Table 1). Average ST at 30 cm depth (taken as spot
measurement during each gas flux measurement) was similar across the burn
severity gradient, at approximately 10–12 ∘C (Table 1).
Environmental characteristics (mean (± standard deviation)) for each microform type across the burn severity gradient.
* UB is unburned, MB is moderately burned and SB is severely burned.
The vegetation survey of the collars undertaken in 2018 indicated that
bryophytes dominated across all collars at all sites, regardless of
microtopographic position. UB hollows were dominated by the moss T. nitens, while the
hummocks were dominated by S. fuscum. The SB hummocks and hollows both had the
highest percentage of bare ground at ∼49 % and
∼55 % cover, respectively, indicating vegetation (mostly T. nitens)
was completely removed during the fire. It was noted that Polytrichum strictum Bridel, J. Bot
(Schrader) moss was beginning to colonize these bare areas. The MB hummocks
had the highest percentage of burned material (∼40 %
cover; predominantly singed S. fuscum) (Fig. 1).
Vegetation cover (%) for each dominant plant functional type
and ground cover for the flux collars in 2018 across the peat burn severity
gradient.
CH4 emissions
The average CH4 flux (± standard deviation) at unburned hollows
was 126.5±80.5 and 56.3±18.9 mg CH4 m-2 d-1 in 2017 and 2018, respectively (Fig. 2). CH4 emissions
were much lower in the MB and SB hollows in both years, with the average
flux being -0.38±1.6 and -0.46±0.9 mg CH4 m-2 d-1 in 2017 and 0.21±1.7 and 0.62±2.5 mg CH4 m-2 d-1 in 2018 (Fig. 2), respectively. Interestingly, across
the burned sites, hummocks had higher fluxes in 2017 than hollows, with the
average flux being 1.10±2.04 mg CH4 m-2 d-1 at the MB
site and 4.53±9.3 mg CH4 m-2 d-1 at the SB site
(Fig. 2). During 2018, fluxes were lower, with hummocks being a slight
sink of CH4 with average fluxes of -0.18±2.06 mg CH4 m-2 d-1 at the MB site and 0.43±1.6 mg CH4 m-2 d-1 at the SB site (Fig. 2). No significant difference
in emissions was found (t(22.5)=1.5, p=0.154) between the present
study and the study undertaken between 2011 and 2014 (Fig. S1 in the Supplement).
Methane (CH4) emissions at each microform type across the
peat burn severity gradient for 2017 and 2018. UB is unburned, MB is
moderately burned and SB is severely burned. Note that CH4 values were
log transformed +10; therefore, a value of 1 represents the measured value
zero.
Results of the LMM illustrate that there was a significant effect of burn on
CH4 flux (Table 2) but no significant effect of microform or year
(Table 2). The second LMM considering environmental controls explained
63 % of the variance in CH4 emissions and found a significant
interaction between burn severity and WT depth (Table 2; Fig. 3) but no
significant relationship between CH4 flux with WT depth, ST (Fig. S2)
or burn severity alone.
Seasonal mean methane (CH4) flux at each collar across the
peat burn severity gradient plotted against seasonal mean water table (WT)
depth. UB is unburned, MB is moderately burned and SB is severely burned.
Note that CH4 values were log transformed +10; therefore, a value of 1
represents the measured value zero.
Statistical results of the linear mixed effects modela.
Flux componentEffectFPRGLMM2bCH4cBurn severityF1,22=83.02<0.00010.52MicroformF2,17=2.40.14YearF1,22=1.20.3Burn severity × microformF2,17=5.50.014InterceptF1,22=2075.3<0.0001CH4cBurn severityF2,20=83.4<0.00010.63WT depthF1,17=0.50.5ST at 30 cm depthF1,17=3.00.1Burn severity × WT depthF2,17=3.30.046Burn severity × ST at 30 cm depthF2,17=3.40.06InterceptF1,20=2085.4<0.0001
a All models have a random factor of collar ID to take into account the repeated measures across both years. b We report the marginal RGLMM2 accounting for variance explained by fixed factors only. c The model was calculated using log10CH4 values.
Potential CH4 production
Measured potential CH4 production was highest in the unburned hollows
ranging from between 0.006 and 0.13 µg g-1 peat h-1
(Fig. 4); however, there was no significant effect of burn severity
(ANOVA, F=2.959, p=0.065). CH4 production was much lower across
the burned sites, ranging between 0.0001 and 0.004 µg g-1 peat h-1. The MB hummocks followed a similar pattern to the field
measurements of CH4 flux, having higher potential CH4 production
than the hollows. No significant difference in organic matter content or
gravimetric water content was found between sites or microform types (Table S1 in the Supplement).
Potential methane (CH4) production across the peat burn
severity gradient and microform type. Each microform represents four sample
replicates.
Discussion
Fire had a strong effect on CH4 emissions in this study, causing a
large decrease in CH4 flux in the MB and SB hollows in comparison to
the UB hollows. Conversely, this study also highlights the resistance of
hummocks to fire (Wieder et
al., 2009; Benscoter et al., 2015), with hummocks across the burned sites
maintaining higher CH4 emissions after the fire compared to hollows.
Methane production in the laboratory followed a similar trend to the field
study, with the highest production in the UB hollows and virtually no production
in the burned hollows, again highlighting this reversal of typical peatland
CH4 emissions. These results contrast with other studies looking at
CH4 emissions post-fire at peatland sites, with
Danilova et al. (2015)
indicating that fire across an ombrotrophic bog could decrease CH4
oxidation due to removal of the methanotrophic community, while
Grau-andrés et al. (2019) note a potential increase in CH4 emissions
due to increased graminoid cover. We did not specifically measure CH4
oxidation in this study.
We hypothesized that the lower CH4 emissions at the burned sites could
be due to the intensity of the burn, reducing substrate availability (labile
carbon) and minimizing the methanogenesis community, resulting in lower
emissions. An increase in fire frequency has the potential to reduce organic
matter quality and change vegetation communities in peatlands (Lukenbach
et al., 2015b). Associated with a change in vegetation communities is the
potential change in biogeochemical cycling and microbial processes (Ward et
al., 2007). For example, the slight recovery in CH4 emissions in 2018
(2 years post-fire) could be due to vegetation recovery (Ward
et al., 2007), providing more available substrate through root exudates for
CH4 production (Greenup
et al., 2000; Robroek et al., 2015). The presence of graminoids in the SB
hollows post-fire was also found, which could also lead to increasing
CH4 emissions in the future, as plant-mediated transport of CH4 is
well documented across peatland ecosystems (Ström et al., 2005).
Higher emissions at the UB site could result from overall shallower WT at
this location compared to the MB and SB sites (Table 1), which were located
at the fen margins. Poplar Fen has a highly variable connection to
groundwater (Elmes et al., 2018) and the hydrogeologic setting of Poplar Fen
likely contributed to the limited effect of the wildfire at this location,
but could also result in higher CH4 emissions than would have occurred
naturally at the burned sites prior to the fire. However, the comparison of
our results to emissions measured between 2011 and 2014 at another location
in Poplar Fen burned during the fire indicates there was no significant
difference in CH4 emissions. Interestingly, we see no relationship with
CH4 emissions and WT depth at the burned sites. This switch in the
typical understanding of the relationship between CH4 emissions and WT
further strengthens our argument on the overriding influence of fire. Even
under suitable hydrological conditions, there is a lack of CH4
production, as shown in the incubation study. Removal of vegetation and soil
organic matter can lead to drier conditions (Thompson and Waddington,
2013), with a lower water table creating a larger aerobic zone, potentially
leading to lower rates of CH4 production and potentially greater rates
of CH4 consumption (Strack et al., 2004; Moore et al., 2011). However,
fire can also cause a higher water table, which could potentially lead to
larger anaerobic zones and potentially higher CH4 emissions. However,
this is dependent on the severity of the burn, where a low-severity fire
which only removes vegetation and does not impact the microbial community
and organic matter content of the soil may still allow for CH4
production. Conversely, a high-severity burn which potentially has removed
these communities and organic matter may no longer allow for CH4
production, even with suitable hydrological conditions.
The higher CH4 production found at the MB hummocks is likely due to the
small methanogen community surviving the fire, due to the resistance of S. fuscum to fire (Benscoter et al., 2011). After
fire, there could be chemical changes in the soil substrate, such as an
increase in availability of terminal electron acceptors, that could
contribute to the reduction in CH4 production and emissions
(Wilson et al., 2017). Therefore, there is a
potential long-term impact on the biogeochemical processes of peatlands
(Danilova et al., 2015), and in order to fully understand the overall impact
of wildfire on CH4 emissions, additional studies at other sites
encompassing the full range of boreal peatland types would be key. This is
especially true given the conflicting results in the literature regarding
the overall impact of fire across a variety of peatland sites. Continuous
monitoring of the recovery of the ecosystem over time could help evaluate
the amount of time required for CH4 emissions to return to similar levels
to the undisturbed site.
Conclusion
This study investigated the impact of wildfire on CH4 emissions at a
treed, moderate-rich fen in northern Alberta. We believe this is the first
study to investigate the impact of wildfire on CH4 emissions at a
non-permafrost boreal fen. The results showed a significant impact of fire
on the magnitude of CH4 flux, with a significant reduction in flux
observed at the burned sites in comparison to the unburned (UB) site. No
relationship was found with water table at the burned sites, contrasting with the
significant relationship at the UB site, further illustrating that
methanogenesis was limited following fire. This was further supported by a
lower rate of CH4 production from peat collected at burned sites
compared to UB, likely linked to reduced methanogen population and/or
substrate availability due to the resistance of Sphagnum spp. hummocks reducing burn
severity. With the expected increase in wildfire frequency across western
boreal Canada, it is vital we fully understand the impact of fire on
CH4 dynamics. If fire is to reduce CH4 emissions and production
across these peatland ecosystems, there is a potential initial net cooling
effect; therefore, it is important to take into account the radiative effect
of all GHGs following wildfire.
Data availability
The data that support the findings of this study are available from the
corresponding author upon reasonable request.
The supplement related to this article is available online at: https://doi.org/10.5194/bg-16-2651-2019-supplement.
Author contributions
MS and RP secured the funding; SJD, CVB, RP, and MS designed the study; SJD
performed the research; SJD analyzed the data with input from MS; and SJD,
CVB, RP, and MS wrote the paper.
Competing interests
The authors declare that they have no conflict of interest.
Acknowledgements
Funding for this project was provided by a Natural Sciences and Engineering
Research Council of Canada (NSERC) Collaborative Research and Development
(CRD) grant to Maria Strack and Richard Petrone co-funded by Suncor Energy Inc., Imperial Oil
Resources Limited, Teck and Shell Canada Energy, and an NSERC Discovery Grant
awarded to Maria Strack. The authors would like to acknowledge Canada's Oil Sands
Innovation Alliance (COSIA) for its support of this project. We thank James Michael Waddington and Manuel Helbig for helpful comments on an earlier version of the
manuscript. Finally, we would like to thank Matthew Coulas, Mariah Smith, Dryden Miller, and Emily Prystupa for their help in the field.
Financial support
This research has been supported by the Natural Sciences and Engineering Research Council of Canada (grant nos. 418557 and 342020).
Review statement
This paper was edited by Paul Stoy and reviewed by two anonymous referees.
References
Belyea, L. R. and Clymo, R. S.: Do hollows control the rateof peat bog growth?, in: Patterned mires and mire pools, edited by: Standen, V., Tallis, J. H., and Meade, R., 55–65, British Ecological Society, London, 1998.
Benscoter, B. W. and Vitt, D. H.: Spatial patterns and temporal trajectories
of the bog ground layer along a post-fire chronosequence, Ecosystems, 11,
1054–1064, 2008.
Benscoter, B. W., Vitt, D. H., and Wieder, R. K.: Association of postfire peat
accumulation and microtopography in boreal bogs, Can. J. Forest
Res., 35, 2188–2193, 2005.
Benscoter, B. W., Thompson, D. K., Waddington, J. M., Flannigan, M. D., Wotton,
B. M., De Groot, W. J., and Turetsky, M. R.: Interactive effects of vegetation,
soil moisture and bulk density on depth of burning of thick organic soils,
Int. J. Wildland Fire, 20, 418–429, 2011.
Benscoter, B. W., Greenacre, D., and Turetsky, M. R.: Wildfire as a key
determinant of peatland microtopography, Can. J. Forest
Res., 45, 1132–1136, 2015.
Bridgham, S. D., Cadillo-Quiroz, H., Keller, J. K., and Zhuang, Q.: Methane
emissions from wetlands?: biogeochemical, microbial, and modeling
perspectives from local to global scales, Glob. Change Biol., 19,
1325–1346, 2013.
Brown, L. E., Palmer, S. M., Johnston, K., and Holden, J.: Vegetation
management with fire modifies peatland soil thermal regime, J.
Environ. Manage., 154, 166–176, 2015.
Bubier, J. L., Moore, T. R., Bellisario, L., Comer, N. T., and Crill, M.:
Ecological controls on methane emissions from a northern peatland complex in
the zone of discontinuous permafrost, Manitoba, Canada, Global
Biogeochem. Cy., 9, 455–470, 1995.
Burke, R. A., Zepp, R. G., Tarr, A., Miller, L., and Stocks, J.: Effect of fire
on soil-atmosphere exchange of methane and carbon dioxide in Canadian boreal
forest sites, J. Geophys. Res., 102, 289–300, 1997.Cresto Aleina, F., Runkle, B. R. K., Brücher, T., Kleinen, T., and Brovkin, V.: Upscaling methane emission hotspots in boreal peatlands, Geosci. Model Dev., 9, 915–926, 10.5194/gmd-9-915-2016, 2016.
Danilova, O. V., Belova, S. E., Kulichevskaya, I. S., and Dedysh, S. N.: Decline
of activity and shifts in the methanotrophic community structure of an
ombrotrophic peat bog after wildfire, Microbiology, 84, 624–629, 2015.
Davies, G. M., Gray, A., Rein, G., and Legg, C. J.: Forest Ecology and
Management, Peat consumption and carbon loss due to smouldering wildfire in
a temperate peatland, Forest Ecol. Manage., 308, 169–177, 2013.
Doerr, S. H., Shakesby, R. A., and Walsh, R. P. D.: Soil water repellency: Its
causes, characteristics and hydro-geomorphological significance, Earth
Sci. Rev., 51, 33–65, 2000.Elmes, M. C., Thompson, D. K., Sherwood, J. H., and Price, J. S.: Hydrometeorological conditions preceding wildfire, and the subsequent burning of a fen watershed in Fort McMurray, Alberta, Canada, Nat. Hazards Earth Syst. Sci., 18, 157–170, 10.5194/nhess-18-157-2018, 2018.Environment Canada: Canadian Climate Normals 1981–2010 Station Data,
Government of Canada, Ottawa, available at:
http://climate.weather.gc.ca/climate_normals (last access: 30 March 2019), 2017.
Flannigan, M. D., Stocks, B., Turetsky, M. R., and Wotton, M.: Impacts of
climate change on fire activity and fire management in the circumboreal
forest, Glob. Change Biol., 15, 1–12, 2008.Flora of North America Editorial Committee: Flora of North America
North of Mexico, 19+ Vols., New York, NY, Flora of North America, 1993.
Gabrielli, E. C.: Partitioning Evapotranspiration in Forested Peatlands
within the Western Boreal Plain, Fort McMurray, Alberta, Canada, MSc
thesis, Wilfrid Laurier University, 2016.Gibson, C. M., Chasmer, L. E., Thompson, D. K., Quinton, W. L., Flannigan, M. D.,
and Olefeldt, D.: Wildfire as a major driver of recent permafrost thaw in
boreal peatlands, Nat. Commun., 9, 3041, 10.1038/s41467-018-05457-1,
2018.
Granberg, G., Catharina, M., Ingvar, S., Svensson, B. H., and Mats, N.:
Sources of spatial variation in methane emission from mires in northern
Sweden: A mechanistic approach in statistical modelling, Global
Biogeochem. Cy., 11, 135–150, 1997.
Grau-andrés, R., Gray, A., Davies, G. M., Scott, E. M., and Waldron, S.:
Burning increases post-fire carbon emissions in a heathland and a raised
bog, but experimental manipulation of fire severity has no effect, J. Environ. Manage., 233, 321–328, 2019.Greenup, A. L., Bradford, M. A., Mcnamara, N. P., Ineson, P., and Lee, J. A.: The
role of Eriophorum vaginatum in CH4 flux from an ombrotrophic peatland,
Plant Soil, 227, 265–272, 2000.Heydari, M., Rostamy, A., Najafi, F., and Dey, D. C.: Effect of fire severity
on physical and biochemical soil properties in Zagros oak (Quercus brantii Lindl.) forests
in Iran, J. Forest Res., 28, 95–104, 2017.
Helbig, M., Chasmer, L. E., Kljun, N., and Quinton, W. L.: The positive net
radiative greenhouse gas forcing of increasing methane emissions from a
thawing boreal forest-wetland landscape, Glob. Change Biol., 23, 2413–2427,
2017.Ingram, R. C., Moore, P. A., Wilkinson, S., Petrone, R. M., and Waddington,
J. M.: Post-fire soil carbon accumulation does not recover boreal peatland
combustion loss in some hydrogeological settings, J. Geophys.
Res.-Biogeo., 124, 775–788, 10.1029/2018JG004716, 2019.
Joabsson, A., Christensen, T. R., and Wallen, B.: Vascular plant controls on
methane emissions from northern peatforming wetlands, Trends Ecol. Evol., 14, 385–388,
1999.
Johnstone, J. F.: Response of boreal plant communities to variations in
previous fire-free interval, Int. J. Wildland Fire, 15,
497–508, 2006.Kettridge, N., Turetsky, M. R., Sherwood, J. H., Thompson, D. K., Miller, C. A.,
Bensocter, B. W., Flannigan, M. D., Wotton, B. M., and Waddington, J. M.:
Moderate drop in water table increases peatland vulnerability to post-fire
regime shift, Sci. Rep.-UK, 5, 8063, 10.1038/srep08063, 2015.Kim, Y. and Tanaka, N.: Effect of forest fire on the fluxes of CO2,
CH4 and N2O in boreal forest soils, interior Alaska, J.
Geophys. Res., 108, 8154, 10.1029/2001JD000663, 2003.
Köster, E., Köster, K., Berninger, F., Aaltonen, H., Zhou, X., and
Pumpanen, J.: Carbon dioxide, methane and nitrous oxide fluxes from a fire
chronosequence in subarctic boreal forests of Canada, Sci. Total
Environ., 601–602, 895–905, 2017.
Köster, E., Köster, K., Berninger, F., Prokushkin, A., and Aaltonen,
H.: Changes in fluxes of carbon dioxide and methane caused by fire in
Siberian boreal forest with continuous permafrost, J. Environ.
Manage., 228, 405–415, 2018.
Lai, D. Y. F.: Methane Dynamics in Northern Peatlands?: A Review, Pedosphere,
19, 409–421, 2009.Lenth, R. V.: Least-Squares Means: The R Package lsmeans, J.
Stat. Softw., 69, 1–33, 10.18637/jss.v069.i01,
2016.
Lukenbach, M. C., Hokanson, K. J., Moore, P. A., Devito K. J., Kettridge, N., Thompson, D. K., Wotton, B. M., Petrone, R. M., and Waddington, J. M.: Hydrological controls on deep burning in a northern forested peatland, Hydrol. Process., 29, 4114–4124, 2015a.
Lukenbach, M. C., Devito, K. J., Kettridge, N., Petrone, R. M., and Waddington,
J. M.: Hydrogeological controls on post-fire moss recovery in peatlands,
J. Hydrol., 530, 405–418, 2015b.
Mayner, K. M., Moore, P. A., Wilkinson, S. L., Petrone, R. M., and Waddington,
J. M.: Delineating boreal plains bog margin ecotones across hydrogeological
settings for wildfire risk management, Wetl. Ecol. Manag., 26,
1037–1046, 2018.
Molina, M., Fuentes, R., and Calderón, R., Escudey M., Avendaño, K., Gutiérrez, M., and Chang, A.: Impact of forest fire ash on surface charge characteristics of Andisols, Soil Sci., 172, 820–834, 2007.
Moore, T. R., Young, A., Bubier, J. L., Humphreys, E. R., Lafleur, P. M., and
Roulet, N. T.: A Multi-Year Record of Methane Flux at the Mer Bleue Bog,
Southern Canada, Ecosystems, 14, 646–657, 2011.
Nakagawa, S. and Schielzeth, H.: A general and simple method for obtaining R2 from generalized linear mixed-effects models, Methods Ecol. Evol., 4, 133–142, 2013.
Neef, L., Van Weele, M., and Van Velthoven, P.: Optimal estimation of the
present-day global methane budget, Global Biogeochem. Cy., 24, 1–10,
2010.
Neff, J. C., Harden, J. W., and Gleixner, G.: Fire effects on soil organic
matter content , composition , and nutrients in boreal interior Alaska,
Can. J. Forest Res., 2187, 2178–2187, 2005.
Noble, A., Palmer, S. M., Glaves, D. J., Crowle, A., and Holden, J.: Impacts of
peat bulk density, ash deposition and rainwater chemistry on establishment
of peatland mosses, Plant Soil, 419, 41–52, 2017.Olefeldt, D., Devito, K. J., and Turetsky, M. R.: Sources and fate of terrestrial dissolved organic carbon in lakes of a Boreal Plains region recently affected by wildfire, Biogeosciences, 10, 6247–6265, 10.5194/bg-10-6247-2013, 2013a.
Olefeldt, D., Turetsky, M. R., Crill, P. M., and Mcguire, A. D.: Environmental
and physical controls on northern terrestrial methane emissions across
permafrost zones, Glob. Change Biol., 19, 589–603, 2013b.Pinheiro, J., Bates, D., DebRoy, S., Sarkar, D., and R Core Team: nlme:
Linear and Nonlinear Mixed Effects Models. R package version 3.1-137, available at:
https://CRAN.R-project.org/package=nlme (last access: 30 March 2019), 2018.
Randerson, J. T., Liu, H., Flanner, M. G., Chambers, S. D., Jin, Y., Hess,
P. G., Pfister, F., Mack, M. C., Treseder, K. K., Welp, L. R., Chapin, F. S.,
Harden, J. W., Goulden, M. L., Lyons, E., Neff, J. C., Schuur, E. A. G., and
Zender, C. S.: The Impact of Boreal Forest Fire on Climate Warming, Science,
314, 1130–1132, 2006.R Core Team: R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria, available at: http://www.R-project.org/ (last access: 5 July 2019), 2013.
Robinson, S. D. and Moore, T. R.: The Influence of Permafrost and Fire upon
Carbon Accumulation in High Boreal Peatlands, Northwest Territories, Canada,
Arct. Antarct. Alp. Res., 32, 155–166, 2000.
Robroek, B. J. M., Jassey, V. E. J., Kox, M. A. R., Berendsen, R. L., Mills,
R. T. E., Cécillon, L., Puissant, J., Meima-Franke, M., Bakker, P. A. H. M.,
and Bodelier, P. L. E.: Peatland vascular plant functional types affect
methane dynamics by altering microbial community structure, J.
Ecol., 103, 925–934, 2015.Saarnio, S., Alm, J., Martikainen, P. J., and Silvola, J.: Effects of raised
CO2 on potential CH4 production and oxidation in, and CH4
emission from, a boreal mire, J. Ecol., 86, 261–268, 1998.
Segers, R.: Methane production and methane consumption: a review of
processes underlying wetland methane fluxes, Biogeochemistry, 41 23–51,
1998.Song, X., Wang, G., Ran, F., Chang, R., Song, C., and Xiao, Y.: Effects of
topography and fire on soil CO2 and CH4 flux in boreal forest
underlain by permafrost in northeast China, Ecol. Eng., 106,
35–43, 2017.Song, X., Wang, G., Hu, Z., Ran, F., and Chen, X.: Boreal forest soil
CO2 and CH4 fluxes following fire and their responses to
experimental warming and drying, Sci. Total Environ., 644,
862–872, 2018.
Strack, M., Waddington, J. M., and Tuittila, E.: Effect of water table
drawdown on northern peatland methane dynamics?: Implications for climate
change, Global Biogeochem. Cy., 18, 1–7, 2004.
Strack, M., Mwakanyamale, K., Hassanpour Fard, G., Bird, M., Bérubé,
V., and Rochefort, L.: Effect of plant functional type on methane dynamics in
a restored minerotrophic peatland, Plant Soil, 410, 231–246, 2017.
Ström, L., Mastepanov, M., and Christensen, T. R.: Species-specific
effects of vascular plants on carbon turnover and methane emissions from
wetlands, Glob. Change Biol., 75, 65–82, 2005.
Tarnocai, C.: The Impact of Climate Change on Canadian Peatlands, Can.
Water Resour. J., 34, 453–466, 2009.
Thompson, D. K. and Waddington, J. M.: Peat properties and water retention in
boreal forested peatlands subject to wildfire, Water Resour. Res., 49,
3651–3658, 2013.Treat, C. C., Bubier, J. L., Varner, R. K., and Crill, P. M.: Timescale dependence
of environmental and plant-mediated controls of CH4 flux in a temperate fen,
J. Geophys. Res.-Biogeo., 112, 1–9, 2007.Turetsky, M. R., Treat, C. C., Waldrop, M. P., Waddington, J. M., Harden, J. W.,
and Mcguire, A. D.: Short-term response of methane fluxes and methanogen
activity to water table and soil warming manipulations in an Alaskan
peatland, J. Geophys. Res., 113, G00A10, 10.1029/2007JG000496,
2008.
Turetsky, M. R., Donahue, W. F., and Benscoter, B. W.: Experimental drying
intensifies burning and carbon losses in northern peatland, Nat.
Commun., 2, 514–519, 2011.
Turetsky, M. R., Kotowska, A., Bubier, J., and Dise, N. B.: A synthesis of
methane emissions from 71 northern, temperate, and subtropical wetlands,
Glob. Change Biol., 20, 2183–2197, 2014.
Vitt, D. H., Halsey, L. A., Bauer, I. E., and Campbell, C.: Spatial and temporal
trends in carbon storage of peatlands of continental western Canada through
the Holocene, Can. J. Earth Sci., 37, 683–693, 2000.
Waddington, J. M., Morris, P. J., Kettridge, N., Granath, G., Thompson, D. K.,
and Moore, P. A.: Hydrological feedbacks in northern peatlands, Ecohydrology,
8, 113–127, 2015.
Ward, S. E., Bardgett, R. D., McNamara, N. P., Adamson, J. K., and Ostle, N. J.:
Long-Term Consequences of Grazing and Burning on Northern Peatland Carbon
Dynamics, Ecosystems, 10, 1069–1083, 2007.Wieder, R. K., Scott, K. D., Kamminga, K., Vile, M. A., Vitt, D. H., Bone, T.,
Xu, B., Benscoter, B., and Bhatti, J. S.: Postfire carbon balance in boreal
bogs of Alberta, Canada, Glob. Change Biol., 15, 63–81, 2009.
Wilkinson, S. L., Moore, P. A., Flannigan, M. D., Wotton, B. M., and Waddington,
J. M.: Did enhanced afforestation cause high severity peat burn in the Fort
McMurray Horse River wildfire, Environ. Res. Lett., 13, 014018, 10.1088/1748-9326/aaa136,
2018.Wilson, R. M., Tfaily, M., Rich, V. I., Keller, J. K., Bridgham, S. D.,
Medvedeff Zalman, C., Meredith, L., Hanson, P. J., Hines, M.,
Pfeifer-Meister, L., Saleska, S. R., Crill, P., Cooper, W. T., Chanton, J. P., and
Kostka, J. E.: Hydrogenation of organic matter as a terminal electron sink
sustains high CO2:CH4 production ratios during anaerobic
decomposition, Org. Geochem., 112, 22–32, 2017.
Zuur, A. G., Ieno, E. N., Walker, N. J., Saveliev, A. A., and Smith, G. M.: Mixed Effect Models and Extensions in Ecology with R, Springer-Verlag, New York, 547 pp., 2009.