Introduction
There has been a renewed interest recently in developing new or enhancing
existing measurement approaches for improving our ability to constrain
dinitrogen (N2) fluxes due to denitrification in terrestrial ecosystems
(Kulkarni et al., 2014; Lewicka-Szczebak et al., 2013; Wang et al., 2011;
Yang et al., 2014). Denitrification, the reduction within soils of nitrogen
oxides (NO3- and NO2-) to NO, N2O, and ultimately
N2 gas, constitutes the most important mechanism for the removal of
reactive nitrogen (Nr) in
terrestrial ecosystems (Galloway et al., 2008; Groffman, 2012). Despite its
importance, denitrification is considered the most un-constrained process in
the global N cycle (Groffman, 2012; Kulkarni et al., 2008) due to
uncertainties in N2 flux estimations that are likely leading to
underestimations of denitrification rates at multiple scales (Butterbach-Bahl
et al., 2013). Considering contemporary atmospheric N deposition rates
globally including the UK (Dore et al., 2012; Galloway et al., 2008; Payne,
2014), the available Nr pool in soils may be greater than the capacity of
denitrification for its removal, with important consequences of chronic N
enrichment of natural terrestrial ecosystems (Galloway et al., 2008; Limpens
et al., 2003). Moreover, nitrous oxide (N2O), an obligate intermediate
of denitrification, is a potent greenhouse gas involved in the breakdown of
stratospheric ozone (Ravishankara et al., 2009). Therefore, a reliable
estimation of the relative magnitude of the major denitrification end
products (N2+ N2O) in soils is crucial in evaluating the role of
denitrification as an Nr sink (Kulkarni et al., 2008).
N2 comprises ∼ 78 % of the atmosphere and thus it is extremely
difficult to measure small N2 fluxes from soil against this high
background, particularly in natural terrestrial ecosystems (Groffman et al.,
2006). Available methods for measuring both N2 and N2O are limited
and can be categorized into the direct flux and 15N isotope tracer
methods (Kulkarni et al., 2014), whilst micrometeorological approaches (Eddy
covariance) are impossible in the N2-rich atmosphere (Felber et al.,
2012). The gas-flow soil core method (Burgin and Groffman, 2012;
Butterbach-Bahl et al., 2002; Scholefield et al., 1997; Wang et al., 2011)
allows the direct measurement of N2 flux (without the addition of any
substrate such as nitrate) from intact soil cores where the soil atmosphere
is replaced by a mixture of He / O2. However, despite the high
precision of the technique, cores still need to be extracted from the field
and conditioned over lengthy periods of time for the complete removal of
N2 from the soil atmosphere. This method is therefore time and resource
intensive, which precludes its application to intensive temporal and large
spatial scales (Kulkarni et al., 2014). Moreover, the gas-flow soil core
method cannot discriminate between sources of N2O, thus overestimating
the denitrification product ratio N2O / (N2 + N2O;
Butterbach-Bahl et al., 2013; Morse et al., 2015). The acetylene inhibition
technique (AIT) is also a direct flux method that exploits the ability of
acetylene (C2H2) at high concentrations (10 % v/v) to inhibit
the reduction of N2O to N2 (Tiedje et al., 1989); thus, total
denitrification (N2 + N2O) is measured in C2H2
amended soil cores in situ, whilst N2 flux is estimated indirectly by
difference from un-amended soil cores. Despite its simplicity and
cost-effectiveness, the AIT is becoming increasingly unpopular due its
several limitations (Groffman et al., 2006), of which the catalytic
decomposition of NO in the presence of C2H2 under oxic or suboxic
conditions in the field (Bollmann and Conrad, 1997; Nadeem et al., 2013), in
particular, precludes its use for reliable estimates of in situ
denitrification rates (Felber et al., 2012).
The 15N gas-flux method (Mosier and Klemedtsson, 1994) has the advantage
of providing in situ measurements of both N2 and N2O
simultaneously, thus allowing its application over large temporal and spatial
scales. It requires the addition of a 15N-labelled tracer in a soil
enclosure in the field which is subsequently covered by a chamber while the
chamber headspace is progressively enriched with 15N–N2 and
15N–N2O produced by denitrification (Stevens and Laughlin, 1998).
Assuming that both N2 and N2O originate from the same uniformly
labelled soil NO3- pool (Stevens and Laughlin, 2001), the true
denitrification product ratio can be more accurately estimated as opposed to
the direct flux approaches (Bergsma et al., 2001). Field applications of the
15N gas-flux method so far have been limited to fertilized
agro-ecosystems (Baily et al., 2012; Cuhel et al., 2010; Graham et al., 2013)
and more recently restored peatland soils (Tauchnitz et al., 2015) with high
15N tracer application rates (between 10 and 200 kg N ha-1),
with the exception of Kulkarni et al. (2014), who have measured
denitrification rates in northern hardwood forests of the USA by adding
tracer amounts of 15N-labelled nitrate, and Morse and Bernhardt (2013),
who applied the same technique in intact soil cores collected from mature and
restored forested wetlands in North Carolina, USA. These recent studies hold
much promise that the 15N gas-flux method can be applied to a range of
natural and semi-natural terrestrial ecosystems, allowing the quantification
of the relative magnitude of N2 and N2O fluxes due to
denitrification from these under-represented ecosystems.
Natural and semi-natural terrestrial ecosystems in the UK (i.e. peatlands,
heathlands, acid grasslands, deciduous and coniferous forests), where there
is no fertilizer use and the impact from grazing and commercial forestry is
minimal (Mills et al., 2013), along with improved and unimproved grasslands
(grazed and/or fertilized) constitute approximately 49 and 85 % of rural
land use cover in England and Wales respectively (Morton et al., 2011).
Unlike arable agriculture, these land use types have been poorly investigated
for their role in Nr loss through denitrification.
The major challenge in measuring 15N–N2 at near-natural abundance
levels is the possibility of interference at m/z 30 (30N2) due
to the reaction of oxygen in the ion source with N and the formation of
NO+ ions that also have m/z 30 (Stevens et al., 1993). Commonly, this
issue is addressed in a continuous flow isotope ratio mass spectrometer
(CF-IRMS) with the inclusion of a copper (Cu) oven for reducing O2 in
the gas sample (Russow et al., 1996). Recently, it has been suggested that
the interference at m/z 30 can be further reduced by including a molecular
sieve column in gas chromatograph IRMS (GC-IRMS) systems to not only separate
N2 and O2 in the gas sample, but also to quantitatively remove
O2 and other trace gases such as carbon monoxide (Lewicka-Szczebak et
al., 2013; Yang et al., 2014). We hypothesize that the precision for
m/z 30 determination can be greatly improved by using a custom-built
preparative unit for the removal of H2O, CO2, N2O, NO+
and CO, a device which also permits the micro-scale injection of volumes of
< 5 µL. These injection volumes are much smaller than have
previously been reported in the literature.
Studies that have compared the 15N gas-flux method with the AIT in the
field are rare and have exclusively focused on highly fertilized
agro-ecosystems with moderate to low soil moisture contents (Aulakh et al.,
1991; Mosier et al., 1986; Rolston et al., 1982). These studies have measured
comparable denitrification rates with both field techniques, although the
relatively low soil moisture contents have probably allowed greater diffusion
of C2H2 to the anaerobic microsites where denitrification occurs
(Malone et al., 1998), whilst the high nitrate application rates have
probably favoured nitrate reduction over N2O reduction (Dendooven and
Anderson, 1995), resulting in high denitrification rates from the AIT.
Conversely, laboratory studies have shown that the AIT significantly
underestimates total denitrification compared to the 15N tracer approach
(Yu et al., 2010) and the direct N2 flux approach (Qin et al., 2012) due
to the incomplete inhibition of N2O reduction to N2 by
C2H2 in wet soils (Yu et al., 2010) or in soils with low nitrate
content, where N2O reduction is more energetically favourable (Qin et
al., 2013, 2014). A comparison of the 15N gas-flux method with the AIT
under in situ conditions across a range of natural and semi-natural
terrestrial ecosystems has not been attempted before. It can provide valuable
insights in terms of the validity and applicability of the two field
techniques for measuring denitrification rates across broad spatial and
temporal scales.
Schematic of the 15N–N2 analysis system.
The objectives of the present study were (1) to determine the precision and
suitability of our preparative-IRMS instrumentation for measuring
15N–N2 and 15N–N2O at low enrichment levels, (2) to
adapt the 15N gas-flux method for application across natural and
semi-natural terrestrial ecosystems, and (3) to compare the validity and
applicability of the 15N gas-flux method with the AIT for measuring in
situ denitrification rates.
Materials and methods
IRMS system
For N2 gas isotopic analysis we used an Isoprime isotope ratio mass
spectrometer (Isoprime Ltd, UK, Wythenshawe) coupled to an in-house built
N2 preparative interface (Fig. 1). Headspace gas (4 µL) was
manually injected with a gas-tight syringe (SGE Analytical science) into the
preparative interface via an open split. Prior to its introduction into the
IRMS, the sample was treated as follows: (a) dried by passing through
Mg(ClO4)2 (Elemental Microanalysis Ltd, Devon, UK), (b) CO2
removed with 0.7–1.2 mm Carbosorb (Elemental Microanalysis Ltd, Devon, UK),
(c) N2O cryogenically trapped under liquid nitrogen, and (d) O2
removed over a copper-packed reduction furnace heated at 600 ∘C. The
N2 was then directed towards the triple collectors of the isotope ratio
mass spectrometer where m/z 28, m/z 29 and m/z 30 mass ions were
measured. Mass / charge ratios for the m/z 28, m/z 29, and m/z 30
nitrogen (28N2, 29N2, and 30N2) were recorded
for each sample at a trap current of 300 µA. Instrument stability
checks were performed prior to each analysis by running a series of 10
reference pulses of N2 (BOC special gases) until a standard deviation of
δ15N better than 0.05 ‰ was achieved. Additionally, 10
consecutive injections (4 µL) of atmospheric air were analysed
prior to the analysis of actual samples. The precision of the instrument was
better than δ15N 0.08 ‰ in all quality control tests.
Nitrous oxide was analysed using modified headspace methods described for the
analysis of nitrogen gas above. Headspace gas (ca. 4 mL) was injected into a
TraceGas™ Preconcentrator coupled to an
Isoprime™ IRMS (GV instruments Ltd, UK)
whereupon the sample was directed through a series of chemical traps designed
to remove H2O and CO2. The N2O was cryogenically trapped under
liquid nitrogen. The waste was flushed out of the instrument. The N2O
was further cryofocused in a second liquid nitrogen trap prior to being
introduced onto a 25 m × 0.32 mm Poraplot Q gas chromatography
column (Chrompack column, Varian, Surrey, UK). The column separated N2O
from any residual CO2, and both entered the IRMS via an open split. The
retention time between the first eluting CO2
(< 2 × 10-10 amplitude) and second eluting N2O
peak typically fell in the range between 60 and 70 s to avoid isobaric
interference of the CO2 with the calculated 15N. The N2O was
directed towards the triple collectors of the isotope ratio mass spectrometer
where m/z 44, m/z 45, and m/z 46 mass ions were measured and
recorded. Instrument stability checks were performed prior to each analysis
by running a series of 10 reference pulses of N2O (BOC special gases)
until a standard deviation of δ15N better than 0.05 ‰ was
achieved. Prior to each sample batch analysis, trace gas N2O
measurements were made on three 100 mL flasks containing atmospheric air
collected from outside the stable isotope laboratory. δ15N
precisions using the TraceGas Preconcentrator and Isoprime IRMS were better
than 0.3 ‰ respectively at the 600 µA trap current.
Field application of the 15N gas-flux and AIT methods
In situ measurements of N2 and N2O were made using static chambers
according to the 15N gas-flux method (Mosier and Klemedtsson, 1994).
Five plots were randomly established in June 2013 in each of four study sites
in the Ribble–Wyre River catchments (area 1145 km2; north-western
England; 53∘59′99′′ N, 2∘41′79′′ W). The study
sites were a heathland (R-HL), a deciduous woodland (R-DW), an unimproved
grassland (R-UG), and an improved grassland (R-IG). In August 2013, four more
study sites were tested in the Conwy River catchment (area 345 km2;
northern Wales; 52∘59′82′′ N, 3∘46′06′′ W)
following a similar sampling design. These sites were an acid grassland
(C-UG), an ombrotrophic peat bog (C-PB), a mixed deciduous and coniferous
woodland (C-MW), and an improved grassland (C-IG). Further details on the
location, land management status, and major soil properties for all study
sites can be found in Sgouridis and Ullah (2014).
Measured ratios of R29 and R30 for N2 in ambient air
(n= 10), ratios of R45 and R46 in standard N2O gas (0.5 ppm
concentration, n= 15), and 15N at % abundance calculated from the
respective ratios for both gases. SD: standard deviation; CV: coefficient of
variation.
R29 (N2)
R30 (N2)
R45 (N2O)
R46 (N2O)
15N at % (N2)
15N at % (N2O)
Mean
7.38 × 10-3
5.16 × 10-5
8.00 × 10-3
2.21 × 10-3
3.71 × 10-1
3.88 × 10-1
SD
2.77 × 10-7
2.26 × 10-7
1.25 × 10-5
1.04 × 10-5
2.09 × 10-5
1.01 × 10-3
CV (%)
0.00
0.44
0.16
0.47
0.01
0.26
In each plot a round PVC collar (basal area 0.05 m2; chamber volume
4 L) was inserted into the soil at ca. 10 cm depth (15 cm for the R-HL and
C-PB plots) 2–4 weeks before the measurement date. The collars were open at
the bottom to maintain natural drainage and root growth during the
measurements. The natural vegetation cover at the soil surface of each
installed collar remained unchanged. The PVC collars were fitted with a
circular groove of 25 mm depth to fit in an acrylic cylindrical cover
(chamber) providing a gas-tight seal when filled with water (Ullah and Moore,
2011). The gas leak rate from the chamber was determined in the laboratory by
placing the sealed collar and chamber over a tray of water, injecting
CH4 (10 ppm) and determining the change in CH4 concentration
within the chamber headspace over time (Yang et al., 2011). The CH4
concentration change within 24 h was negligible, with the relative standard
deviation (RSD) being < 5 %. We did not use a vent tube for
pressure equilibration, as suggested by Hutchinson and Mosier (1981), in our
chamber design, which could have diluted the chamber headspace with
atmospheric N2, as part of our effort to increase the probability of a
detectable 15N–N2 signal in the chamber headspace. Instead
chambers were covered with reflective foil for minimizing temperature
increase within the chamber headspace during the incubation period (Ullah and
Moore, 2011). Labelled K15NO3- (98 at % 15N,
Sigma-Aldrich) was applied in each plot via 10 injections of equal volume
through a grid (4 × 6 cm) using custom-made 10 cm long lumber
needles (15 cm for the R-HL and C-PB plots) attached to a plastic syringe
(Ruetting et al., 2011). The 15N tracer was delivered as the needle was
pushed into the soil from the surface up to 10 or 15 cm depth, aiming to
achieve as uniform as possible labelling of the soil volume enclosed by the
collar, as required by the 15N gas-flux method (Mosier and Klemedtsson,
1994). The volume and concentration of the labelled K15NO3-
tracer solution was determined from measurements of soil nitrate and moisture
content, as well as bulk density adjacent to each plot made during the
installation of the collars (Morse and Bernhardt, 2013). Lower application
rates (< 0.1 kg N ha-1) were administered to natural study
sites (e.g. peat bog, heathland) and higher rates
(< 1 kg N ha-1) administered to semi-natural ones (e.g.
unimproved and improved grasslands). The tracer solution (50–200 mL) was
adjusted between 3 and 5 % of the ambient volumetric water content (see
Supplement Table S1 for detailed data from each sampling plot). It should be
noted that no time was allowed for the equilibration of the added tracer
solution in the soil enclosure to avoid significant loss of the low amount of
added nitrate via plant uptake.
Following the 15N tracer application, the collars were covered with the
acrylic chamber fitted with a rubber septum for gas sampling. Two sets of gas
samples (20 mL each) were collected with a gas-tight syringe (SGE Analytical
Science) through the septum of the chamber cover at T= 1 h, T= 2 h,
and T≈ 20 h after the tracer injection, while a T= 0 h
sample was collected immediately after tracer injection above the plot
surface before fitting the chamber cover. The gas samples were transferred
into pre-evacuated (< 100 Pa) 12 mL borosilicate glass vials with
butyl rubber septa (Exetainer vial; Labco Ltd., High Wycombe, United Kingdom)
for storage under positive pressure and were analysed within 8 weeks from
collection without any significant change in the gas concentration (Laughlin
and Stevens, 2003).
Adjacent to each PVC collar in each plot, two intact soil cores (50 mm I.D.,
15 cm long) were extracted from 10 cm depth, leaving the top 5 cm void as
a headspace volume. The cores were capped on both ends with the top cap
fitted with a rubber septum for gas sampling. One set of cores was amended
with pure C2H2 with 5 mL injected through the septum directly in
the middle of the soil core before 10 % of the headspace was also
replaced with pure C2H2. The second set of cores was not amended
with C2H2 and both cores were placed back in the ground where they
came from. Gas samples (5 mL) were collected with a gas-tight syringe (SGE
Analytical Science) through the septa of the cores at T= 1 h and T= 2 h after amendment with acetylene. The gas samples were transferred
into pre-evacuated (< 100 Pa) 3 mL borosilicate glass vials with
butyl rubber septa (Exetainer vial; Labco Ltd., High Wycombe, UK) for storage
under positive pressure.
Flux calculations
The 15N content of the N2 in each 12 mL vial was determined using
the IRMS system described above and the ratios R29
(29N2 / 28N2) and R30
(30N2 / 28N2) were measured in both enriched
(T= 1, 2, and 20 h) and reference samples (T= 0 h). The inclusion of
air reference standards between every 10 samples indicated an upward drift
for R30 over time, potentially due to the formation of NO+ in the ion
source despite the inclusion of the Cu reduction step (Lewicka-Szczebak et
al., 2013). Subsequently, every sample batch was drift corrected by fitting a
linear regression through the air reference standards and calculating an
offset correction for both R29 and R30 (Yang et al., 2014). The minimum
detectable change (MDC) in R29 and R30 was defined with repeated manual
analyses of air reference standards (n= 10) and was calculated using the
following equation (Matson et al., 2009):
MDC=μpair diff+(2σpair diff),
where μ is the mean difference of all possible unique pairs of air
reference standards (n= 45) and σ is the standard deviation
between sample pairs. The MDC for R29 was 7.7 × 10-7 and for
R30 it was 6.1 × 10-7, and these values were used to determine
whether each time step sample was significantly different from ambient
reference samples (T= 0 h), and, if not, they were excluded from the flux
calculations.
For calculating the total N2 flux from a uniformly labelled soil nitrate
pool when both R29 and R30 are measured, the “non-equilibrium” equations
were applied as described by Mulvaney (1984) for estimating first the
15N fraction in the soil NO3- denitrifying pool
(15XN) as
15XN=2(ΔR30/ΔR29)/(1+2(ΔR30/ΔR29)),
where ΔR29 and ΔR30 are the difference between R29 and
R30 respectively between enriched (T= 1, 2, and 20 h) and reference
samples (T= 0 h). Subsequently, the 15XN allows the
quantification of the fraction of the N2 evolved from the
15N-labelled pool (d) using either ΔR30 or ΔR29:
d=ΔR3015XN2,d=ΔR29215XN1-15XN2.
Using d and the concentration of [N2] (µg N) in the chamber
headspace, the evolved N2 from the soil pool was calculated:
EvolvedN2=d[N2]/(1-d),
The N2 flux was then calculated using linear regression between the
maximum evolved N2 and the respective incubation time per plot surface
area, and was expressed in µg N m-2 h-1 representing
the total N2 flux from the mixture of the 15N-labelled tracer and
the soil N at natural abundance (Stevens and Laughlin, 1998).
The 15N content of the N2O in the same 12 mL vials as well as the
ratios R45 (45N2O / 44N2O) and R46
(46N2O / 44N2O) were measured in both enriched
(T= 1, 2, and 20 h) and reference samples (T= 0 h). The application
of the “non-equilibrium” equations to N2O is analogous to N2
after correcting for the naturally occurring oxygen isotopes (Bergsma et al.,
2001). Therefore, the ratios R45 and R46 were converted to ratios of R29 and
R30 respectively by applying the following equations:
R29=R45-R17,R30=R46-R29R17-R18,
where for R17 (17O / 16O) the value 0.000373 was used and for
R18 (18O / 16O) the value 0.0020052 was used (Bergsma et al.,
2001). There was no significant instrumental drift for the ratios R45 and R46
over time. The MDC was defined, for the converted R29 and R30, with repeated
automatic analyses of 0.5 ppm N2O standards (n= 15) as
3.4 × 10-5 and 2.9 × 10-5 respectively. The
second set of gas samples collected at the same time in the field was
analysed for total N2O on a GC-μECD (7890A GC Agilent Technologies
Ltd., Cheshire, UK) and the concentration of [N2O] (µg N) was
used in Eq. (5) to calculate the N2O flux due to denitrification of the
mixture of the 15N-labelled tracer and the soil N and expressed in
µg N–N2O m-2 h-1. Assuming that the N2O
originates from the same uniformly labelled pool as N2, the
15XN from N2O was used to estimate d for N2 using
either R30 (Eq. 3) or R29 (Eq. 4), thus lowering the limit of detection for
N2 (Stevens and Laughlin, 2001) and allowing measurement of N2 gas
flux from natural terrestrial ecosystems at low 15N-tracer application
rates.
Gas samples collected from the intact soil cores with or without acetylene
amendment were analysed for N2O on a GC-μECD (7890A GC Agilent
Technologies Ltd., Cheshire, UK) and for CO2 on a GC-FID (7890A GC
Agilent Technologies Ltd., Cheshire, UK) and flux rates were determined by
linear regression between 0 and 2 h. The instrument precision was determined
from repeated analyses of 6 ppm N2O and 200 ppm CO2 standards
respectively (n= 8), and the RSD was < 1%.
Statistical analysis
Using factor analysis on selected soil physico-chemical properties, the
samples from the eight field sites were grouped into three broad land use
types: organic soils (C-PB, C-UG, R-HL), forest soils (C-MW, R-DW), and
grassland soils (C-IG, R-UG, R-IG) according to Sgouridis and Ullah (2014).
All subsequent statistical analyses were performed on the broad land use
types rather than individual field sites. The data were analysed for
normality and homogeneity of variance with the Kolmogorov–Smirnov test and
the Levene statistic respectively and logarithmic transformations were
applied as necessary. One-way ANOVA combined with Hochberg's GT2 post hoc
test for unequal sample sizes or the Games–Howell post hoc test for
unequal variances was performed for comparing the variance of the means
between land use types for all gas fluxes. The non-parametric Kruskal–Wallis
test was used to compare mean flux rates between incubation time intervals.
Pearson correlation was used between log-transformed flux rates. Comparisons
between the 15N gas-flux and AIT techniques were made with an
independent samples t test. All statistical analyses were performed using
SPSS® 21.0 for Windows (IBM Corp., 2012,
Armonk, NY).
Results
IRMS system evaluation
The precision of the IRMS systems was evaluated using repeated analyses of
ambient air samples for N2 (n= 10) injected manually in one batch and
repeated analyses of N2O gas standard at natural abundance and 0.5 ppm
concentration (n= 15) using automated injections. The mean measured ratios
of R29 and R30 for N2 and of R45 and R46 for N2O are shown in
Table 1. Measurement precision was defined as the coefficient of variation
(%), and it was lower for R29 compared to R30 and lower for R45 compared
to R46, but still less than 0.5 % for all four measured ratios. We
estimated the 15N atom% abundance for both gases as per Yang et
al. (2014), and the precision was less than 0.01 % for N2 in air and
0.26 % for standard N2O at natural abundance. The mean measured R30
(5.16 × 10-5) was higher than the theoretical value of
1.35 × 10-5 for N2 in ambient air, suggesting some
interference at m/z 30 potentially due to the formation of NO+ ions
in the ion source of the mass spectrometer despite the inclusion of the Cu
reduction oven. The contribution of NO+ ions (R30 measured–R30
theoretical) was 3.81 × 10-5, whilst the ratio of R30
theoretical / R30 measured was 0.26. Correcting the R30 ratio for the
contribution of NO+ ions results in a lower “true” precision for the
R30 (CV = 1.67 %).
The ambient soil nitrate pool, the 15N tracer application rate,
the estimated enrichment of the total soil nitrate pool, the calculated
15XN value from N2O, and the slope of the 15XN
change with incubation time in the three land use types. Data are means with
standard errors in parentheses.
Land use type
Ambient
Tracer
Enrichment of
15XN
15XN
NO3-
application rate
total soil NO3- pool
(%)
slope
(kg N ha-1)
(kg15N ha-1)
(15N at %)
Organic soil (n= 3)
0.53 (0.44)
0.04 (0.02)
25 (11.8)
90 (1.5)
0.003 (0.0054)
Woodland (n= 2)
3.86 (2.42)
0.62 (0.41)
13 (0.7)
79 (8.3)
-0.007 (0.0025)
Grassland (n= 3)
1.81 (0.96)
0.51 (0.19)
24 (5.1)
81 (8.4)
0.000 (0.0037)
Field application of the 15N gas-flux method
The 15N tracer application rate was variable between land use types and
ranged between 0.03 and 1 kg 15N ha-1, while it was lower in the
case of the organic soils and higher for the woodland and grassland soils
(Table 2). Based on the soil nitrate content on the day of the tracer
amendments (Table 2), the estimated enrichment of the total soil nitrate pool
was on average between 13 and 25 15N at % (detailed data on the
15N tracer application per field site are shown in Table S2).
The 15N fraction in the denitrifying pool (15XN), as
calculated from the measured isotopic ratios of the N2O after 1 h of
incubation using Eq. (2), ranged between 65 and 93 15N at %. The
average change in the 15XN with incubation time, indicated by
the slope shown in Table 2, was not different from 0 in the case of the
organic (t test; t= 0.520, df = 18, p > 0.05) and
grassland soils (t test; t= 0.047, df = 28,
p > 0.05), whilst it was significantly below 0 for the
woodland soils (t test; t= 2.917, df = 18, p < 0.05).
Separating the woodland soils into C-MW and R-DW sites, only the former
displayed a significant negative slope of 15XN with incubation
time (t test; t= 3.306, df = 8, p < 0.05), suggesting
N2O production from a second nitrate pool, possibly nitrate produced
from the oxidation of NH4+ via nitrification, in the C-MW. In cases
where the 15XN could be calculated from the N2 isotope
ratio data (woodland and grassland soils; data shown in Table S3), this was
not significantly different from their respective 15XN
calculated from the N2O isotope ratio data (t test; t-WL= 0.929, df = 12, p > 0.05; t-GL= 1.511,
df = 20, p > 0.05).
Evolved (a) N2 and (b) N2O gas measured
between 1, 2, and 20 h incubation time intervals using the 15N gas-flux
method in the organic soil (OS), woodland (WL), and grassland (GL) land use
types. Data points are means and the error bars represent standard errors.
The mean evolved amount of N2 and N2O gases due to denitrification
in each land use type increased with increasing incubation time (Fig. 2). The
increase in the evolved N2 was statistically significant after 20 h
incubation in GL (ANOVA; F= 19.8, p < 0.01), whilst due to
the high variability among plots, shown by the large error bars at 20 h
incubation in Fig. 2a, it was not significant for the OS and WL soils. The
amount of N2O accumulated after 20 h (Fig. 2b) was significantly higher
than at the previous time points for all land use types (ANOVA;
FOS= 4.6, FWL= 5.1, FGL= 14.7,
p < 0.05). However, this pattern was not consistent in every
sampling plot (data presented in Tables S4 and S5); for example, in C-MW, the
highest N2 accumulations were observed after the first or second hour of
incubation, whilst in most cases the increase in N2 and N2O
concentrations was not linear throughout the incubation period (Tables S4
and S5). This suggested a complex temporal sequence of events, which was not
consistent between plots among the different land use types, probably as a
result of complex interactions between environmental controls of
denitrification and the length of the incubation period (details below).
Consequently, the N2 flux rate decreased with increasing incubation time
(Fig. 3a), and this decrease was significant between each time interval in
the OS (Kruskal–Wallis; χ2= 11.35, p= 0.003), between 1 and
20 h in the WL (Kruskal–Wallis; χ2= 10.78, p= 0.005), and
between 1 and 2 h in the GL (Kruskal–Wallis; χ2= 10.10,
p= 0.006). Conversely, the N2O flux rates increased between the first
and second hour of incubation (Fig. 3b), followed by a decrease after 20 h,
albeit the mean differences between time intervals were not statistically
significant in any land use type (Kruskal–Wallis;
χOS2= 3.58, χWL2= 3.47,
χGL2= 3.01, p > 0.05).
Comparison of mean flux rates and ratios between land use types for
the two field methods using one-way ANOVA. All variables are log-transformed.
F: F statistic; P: probability level.
15N gas-flux
F
P
Denitrification
19.4
< 0.001
N2O emission
31.1
< 0.001
N2O / (N2+ N2O)
7.4
< 0.01
Total bulk N2O
19.4
< 0.001
CO2 production
19.8
< 0.001
AIT
Denitrification
12.7
< 0.001
Total bulk N2O
9.4
< 0.01
N2O / (N2+ N2O)
0.3
> 0.05
CO2 production (un-amended cores)
11.2
< 0.001
CO2 production (C2H2 amended cores)
11.7
< 0.001
Mean rates of (a) N2 flux and (b) N2O
flux due to denitrification at the three incubation time intervals in the
three land use types (OS: organic soils; WL: woodland; and GL: grassland).
Same lower-case letters indicate no significant differences
(p > 0.05) between incubation time intervals according to the
non-parametric Kruskal–Wallis test. Error bars represent standard errors.
The N2 flux ranged between 2.4 and
416.6 µg N m-2 h-1 and was significantly different
among land use types based on 20 h incubation duration for comparison
purposes (Table 3). The grassland soils showed on average 3 and 14 times
higher denitrification rates than the woodland and organic soils respectively
(Fig. 4a). A similar pattern was observed for the N2O flux due to
denitrification (range: 0.003–20.8 µg N m-2 h-1),
with the grassland soils emitting on average 14 and 120 times more N2O
than the woodland and organic soils respectively (Fig. 4b), whilst the
N2O flux was on average 20 to 200 times lower than the N2 flux
among land use types. Consequently, the denitrification product ratio
N2O / (N2+ N2O) was low, ranging between 0.03 and
13 %, and was highest in the GL and similar between the WL and OS
(Fig. 4c). The change in the denitrification product ratio with incubation
time was evaluated in each sampling plot where both N2 and N2O
fluxes were available (data shown in Table S6). Generally, there was no
consistent pattern between individual sampling plots, with the exception of
the grassland soils, where the maximum product ratio was observed after 2 h
of incubation (ANOVA; F= 6.11, p < 0.05). This was an
indication of some reduction of the denitrification-derived N2O to
N2 during the extended closure period (up to 20 h) in the grassland
soils.
Mean rates of (a) N2 flux, (b) N2O
emission due to denitrification, and (c) the denitrification product
ratio N2O / (N2 + N2O) in the three land use types (OS:
organic soils; WL: woodland; and GL: grassland). Same lower-case letters
indicate no significant differences (p > 0.05) between land
use types according to one-way ANOVA and the Games–Howell post hoc test.
The sample size (n) is given in parentheses for each land use type on the
x axis. Error bars represent standard errors.
Comparison with the AIT
The total denitrification rate measured from the C2H2 amended
intact soil cores in the same land use types ranged between 0.5 and
325.2 µg N m-2 h-1 and correlated positively with the
total denitrification rate (N2 and N2O fluxes combined) measured
with the 15N gas-flux method (Pearson; r= 0.581, n= 25,
p < 0.01) following a similar trend among land use types, albeit
with only the OS being significantly lower than the grassland and woodland
soils (Table 3). The AIT denitrification rates were between 3 and 5 times
lower than the total denitrification from the 15N gas-flux (Fig. 5a),
with the difference being significant in woodland (t test; t= 3.914,
df = 18, p < 0.01) and grassland (t test; t= 3.521,
df = 25, p < 0.01) soils.
The total N2O flux measured from the un-amended intact soil cores ranged
between 0.15 and 86.6 µg N m-2 h-1 and was between 1
and 3 times lower than the total denitrification rate from the C2H2
amended cores. There were no significant differences between bulk N2O
fluxes measured with the static chambers and the un-amended intact soil cores
(Fig. 5b), which indicated that total N2O emissions were comparable
between the two field techniques. Consequently, estimating the
denitrification product ratio from the un-amended and C2H2 amended
intact soil cores resulted in significantly higher ratios compared to the
15N gas-flux approach (Fig. 5c), which were on average between 50 and
60 % and not significantly different among land use types (Table 3).
(a) Mean total denitrification measured with the 15N
gas-flux method and the AIT, (b) mean bulk N2O emission
measured in the static chambers of the 15N gas-flux method and in
un-amended intact soil cores, and (c) the denitrification product
ratio N2O / (N2+ N2O) with the 15N gas-flux method
and the AIT in the three land use types (OS: organic soils; WL: woodland; and
GL: grassland). Same lower-case letters indicate no significant differences
(p > 0.05) between measurement methods according to an
independent samples t test. The sample size (n) is given in parentheses
for each land use type and each method on the x axis. Error bars represent
standard errors.
The mean CO2 production rate was similar irrespective of whether it was
measured in static chambers, in C2H2 amended or un-amended intact
soil cores (Fig. 6), indicating that soil respiration (including both
microbial and plant respiration) was not affected by the measurement
technique.
Discussion
IRMS system evaluation
The precision of our trace gas isotope ratio mass spectrometer (TG-IRMS) for
manual analysis of 15N–N2 in gas samples was comparable for both
the R29 and R30 ratios to the recently developed gas chromatograph-IRMS
(GC-IRMS) systems that included a combination of a copper reduction oven and
a molecular sieve (Lewicka-Szczebak et al., 2013) or only a molecular sieve
(Yang et al., 2014) for the removal of O2 from the samples. This was
achieved while injecting a trace amount of headspace gas sample
(4 µL), which is less than half of what is used by Lewicka-Szczebak
et al. (2013) and 10 times less than the required sample volume by Yang et
al. (2014). Furthermore, the interference at m/z 30 by NO+ ions was
reduced by an order of magnitude (3.81 × 10-5) compared to the
value (1.6 × 10-4) reported by Lewicka-Szczebak et al. (2013).
Consequently, correcting the R30 ratio for the NO+ ion interference led
to a CV value of < 2 %, which was significantly lower than the
precision reported for natural abundance samples in previous studies
(Lewicka-Szczebak et al., 2013; Russow et al., 1996; Stevens et al., 1993),
thus constituting a significant improvement in m/z 30 determination in
N2 gas samples with low 15N enrichment. However, the correction of
the R30 ratio is only useful for estimating the “true” instrument precision
for m/z 30 and is not necessary for calculating N2 fluxes as shown by
Lewicka-Szczebak et al. (2013), unless using the mathematical formulations of
Spott and Stange (2007).
Mean CO2 production measured in the static chambers of the
15N gas-flux method, in un-amended and C2H2 amended intact
soil cores in the three land use types (OS: organic soils; WL: woodland; and
GL: grassland). Same lower-case letters indicate no significant differences
(p > 0.05) between measurement methods according to an
independent samples t test. The sample size (n) is given in parentheses
for each land use type on the x axis. Error bars represent standard
errors.
The TraceGas™ Preconcentrator IRMS system used
for 15N–N2O analysis displayed similar precision for the
determination of R45 and R46 in standard N2O gas at circa ambient
concentration to a similar system used by Bergsma et al. (2001) while
injecting only 4 mL of gas sample as opposed to 0.5 L used by Bergsma et
al. (2001). When expressed in delta values (δ15N), the precision of
our system was better than 0.05 ‰, which is significantly better
than the respective precisions reported in Lewicka-Szczebak et al. (2013) and
Yang et al. (2014), but comparable to Well et al. (1998). Therefore, the
analytical precision achieved for both the 15N–N2 and
15N–N2O analyses, using smaller gas sample volumes than previously
reported, allowed us to quantify in situ N2 and N2O fluxes with low
tracer addition under field conditions.
Field application of the 15N gas-flux method
The average 15N tracer application rate
(0.04–0.5 kg 15N ha-1 or 0.4–1.2 mg15N kg-1 dry
soil) across land use types was 1 to 2 orders of magnitude lower than
previous applications of the 15N gas-flux method in highly fertilized
agricultural systems (Baily et al., 2012; Bergsma et al., 2001; Cuhel et al.,
2010; Graham et al., 2013) and in restored peatland soils (Tauchnitz et al.,
2015). The estimated enrichment of the total soil NO3- pool was
variable (2–40 15N at %, Table S2) and this wide range was due to the
fact that the tracer concentration was calculated based on the previous
campaign's soil nitrate data, which in some cases did not reflect the soil
nitrate content on the day of the tracer application a month later. It should
be noted that the soil nitrate enrichment levels reported in this study
correspond to the high end of the average soil NO3- pool enrichment
(10–15 15N at %, Table S2) for the period April 2013 to October 2014,
which is presented in a separate publication (Sgouridis and Ullah, 2015). To
our knowledge, only Kulkarni et al. (2014) have applied the 15N gas-flux
method in the field with soil nitrate enrichment levels (5 15N at %)
lower than in our study, but this had as a consequence poorly detected
15N–N2 fluxes. Nevertheless, for the organic soils the average
tracer application rate corresponded to current estimates of daily
atmospheric N deposition (0.05 kg N ha-1 d-1) in the UK
(∼ 15–20 kg N ha-1 yr-1; Dore et al., 2012; Payne,
2014), whilst for the grassland soils the tracer application mimicked a daily
fertilizer application rate of 0.5 kg N ha-1 d-1. Due to the
inclusion of the NO3--rich C-MW site in the woodland soils, tracer
application rates were higher than the daily atmospheric N deposition rates,
but also reflected internal N cycling processes (e.g. nitrification) as an
additional source of nitrate in these well-drained forest soils. Therefore,
the application of the 15N tracer at these low rates should not be
expected to enrich the soil nitrate pool significantly, and potentially
enhance the denitrification activity, in excess of the amount of nitrogen
normally deposited via natural processes and common management practices.
The major assumptions of the 15N gas-flux method and the associated
“non-equilibrium equations” are that the denitrifying soil NO3-
pool is uniformly labelled with 15N and that the N2 and N2O
originate from the same denitrifying pool (Stevens and Laughlin, 1998). The
15N fraction in the denitrifying pool (15XN), calculated
non-destructively from the measured isotope ratios, ranged between 65 and
93 % and was well above the 10 % threshold for the correct
application of the “non-equilibrium equations” (Lewicka-Szczebak et al.,
2013). However, the calculated 15XN was higher than the
estimated total soil NO3- pool enrichment (range: 2–40 15N at
%), suggesting non-homogeneous mixing of the added tracer (98 15N at
%) with the ambient soil nitrate at natural abundance despite our effort
for uniform tracer application with multiple injections across the
investigated soil depth (Ruetting et al., 2011). Wu et al. (2011) have
optimized the number of injections and the volume of tracer needed to achieve
homogeneous labelling of a soil core (diameter 15 cm; height 20 cm) and
reported that 38 injections of 4 mL volume each were necessary. We have used
only 10 injections of 5–20 mL volume (depending on the soil water content
of each land use type) to minimize the disturbance of the soil matrix,
particularly in the highly porous media such as peatland soils, and this was
clearly sub-optimal for the homogenous labelling of the soil enclosure but
probably a necessary compromise for large-scale intensive measurements. We
were not able to sample the soil within the chamber collars to directly
estimate the 15NO3- content of the soil pool due to time and
budget constraints. However, in cases where destructive soil sampling was
used to measure the soil nitrate pool enrichment (Kulkarni et al., 2014), the
results were significantly different from the estimated enrichment due to
sampling bias of the volume of soil affected by the tracer application.
Non-uniform mixing of the 15N label may lead to overestimation of the
15XN and underestimation of the denitrification flux rates
(Boast et al., 1988). However, under field conditions, it is unlikely to
achieve complete mixing of the added tracer with the ambient nitrate pool,
and experimental studies (Mulvaney, 1988; Mulvaney and Van den Heuvel, 1988)
have shown that the associated error is well constrained and that accurate
measurements can be made even with a less-uniformly labelled denitrifying
pool.
The larger volume of tracer per injection (> 4 mL) in
combination with the lower number of injections compared to Wu et al. (2011)
may have created localized saturation effects (saturated soil cylinders
around the injection holes), even if the total soil moisture content of the
enclosure was not increased by more than 5 %, which would require several
hours to equilibrate with the ambient soil moisture. We did not allow time
for this soil moisture equilibration to occur following the tracer injection
to avoid significant loss of the added nitrate via plant uptake (measurements
occurring during the growth season). Therefore, it is likely that in plots
where denitrification activity may have been limited by soil moisture (e.g.
C-MW with mean WFPS 42 ± SE 0.76 %), the flux rates after 1 and
2 h of incubation may be overestimated due to moisture-induced
denitrification events.
Most studies using 15N tracers and static chambers in highly fertilized
systems typically deploy their chambers between 1 and 2 h (Baily et al.,
2012; Cuhel et al., 2010; Tauchnitz et al., 2015), but it has been shown that
longer incubation periods (up to 24 or 48 h) may be needed in case of low
15N enrichment applications in intact soil cores (Morse and Bernhardt,
2013) and laboratory incubations (Yang et al., 2014) for a more precise and
accurate detectable 15N–N2 signal. However, it should be noted
that in these cases the soil cores or slurries were incubated in fully
enclosed systems and were thus not affected by potential bias from diffusion
of evolved N2 and N2O to the subsoil (Clough et al., 2005). The
open-bottom, un-vented static chamber design used in this study in
combination with the extended incubation period up to 20 h may have
potentially allowed some loss of the evolved N2 and N2O through
downward subsoil diffusion and/or reduction of gas exchanges at the
soil–atmosphere interface due to decreasing concentration gradients (Healy
et al., 1996). This could partly explain the non-linear increase in the
evolved N2 and N2O in the chamber headspace (Fig. 2a and b) and
also the decrease in the N2 flux rate with increasing incubation time
(Fig. 3a). The N2O flux rate increased up to 2 h incubation followed by
a decrease after 20 h consistently across land use types (Fig. 3b),
indicating that the extended enclosure period lowered N2O fluxes due to
subsoil diffusion and enhanced N2O reduction to N2. However, due to
the high spatial heterogeneity within each land use type, the mean N2O
flux rate was not significantly different between the different incubation
intervals. In other words, the non-linearity of N2O evolution had less
effect on the flux rate estimation than the inherent spatial variability
within each land use type, which is in agreement with the findings of
Chadwick et al. (2014), who suggested that the spatial variability of
N2O fluxes far exceeds the bias due to assumed linearity of fluxes.
The lack of a consistent pattern of N2 flux rate change with incubation
time among the different land use types suggested a more complex temporal
variability of N2 fluxes that apart from the duration of incubation
could have also been affected by the distribution of the added nitrate
tracer. In the OS sites with the lowest average nitrate content (Table 2) and
the highest water filled pore space (mean WFPS:
C-PB = 70 ± SE 3.21 %; C-UG = 66 ± SE 1.58 %;
R-HL = 69 ± SE 2.00 %), non-homogeneous tracer distribution
(15XN= 90 15N at %) could have led to the creation of
hotspots of denitrification activity due to substrate availability resulting
in potentially overestimated flux rates in the first or even the second hour
of incubation. However, analytical uncertainty due to fluxes being close to
the limit of detection could not be ruled out. Conversely, in the soil
moisture-limited forest site (C-MW), the injection of even 50 mL of tracer
solution could have led to an increased moisture-induced denitrification
activity within the first 1–2 h of incubation, until the added water
started to equilibrate with the ambient soil moisture. Therefore the N2
flux rate in C-MW may be significantly overestimated after 1 h of
incubation. In the grassland sites and the R-DW forest site with intermediate
soil moisture (mean WFPS: R-DW = 65 ± SE 1.79 %;
R-UG = 64 ± SE 1.41 %; C-IG = 60 ± SE 1.45 %;
R-IG = 61 ± SE 2.46 %) and nitrate content, the tracer
injection is unlikely to have significantly affected the denitrification rate
when all the conditions (i.e. soil moisture and substrate availability) were
favourable, and therefore flux rates estimated after 1 h of incubation
should be more reliable as long as the bias from analytical uncertainty was
low. In these sites denitrification rates estimated after 1 or 20 h of
incubation were not significantly different (Fig. 3a), suggesting a
quasi-linear N2 evolution throughout the incubation period (at least in
37.5 % of the sampling plots; see Table S4). However, the N2 flux
rates were significantly lower after 2 h of incubation, whereas the N2O
flux rates were maximum at 2 h of incubation, consequently leading to an
increased product ratio N2O / (N2+ N2O) (Table S6). This
observation could potentially be explained by a delay in the de novo
synthesis of denitrification enzymes and the fact that the N2O reductase
is known to have a slower expression than the preceding reduction enzymes
(Knowles, 1982), leading to N2O accumulation and lower N2
production after 2 h of incubation. After 20 h incubation, the decrease in
the product ratio could be explained by a higher reduction rate of N2O
to N2 due to probably higher N2O reductase activity but also slower
soil–atmosphere exchange of N2O due to the decreasing concentration
gradient (Healy et al., 1996).
It has been shown that the N2 flux estimation with the 15N gas-flux
method is sensitive to the incubation time interval and the homogeneity of
the tracer distribution due to the combination of several antagonistic
effects such as decreasing gas diffusion gradients and soil moisture and
substrate availability effects due to the added tracer solution. The
uncertainty in the estimated in situ N2 fluxes can be significantly
reduced if additional effort is made to increase the homogeneity of the
tracer application by increasing the number of injections while reducing the
volume of the applied tracer (particularly in soils where denitrification is
limited by moisture). Moreover, allowing the equilibration of the added
tracer solution with the ambient soil water before gas sampling commences and
by closely monitoring the linear evolution of the produced gases with more
frequent gas sampling at shorter equal incubation intervals could help in
identifying the appropriate length of incubation, thus avoiding potential
overestimation of denitrification in nitrate and moisture-limited ecosystems
and potential underestimation due to subsoil diffusion of evolved gases. The
detailed uncertainty analysis of the N2 and N2O flux estimation
presented in this study complements the large-scale application of the
15N gas-flux method in the same land use types between April 2013 and
October 2014 for estimating annual rates of denitrification and N2O
emission, which is presented in Sgouridis and Ullah (2015).
The minimum detectable N2 and N2O fluxes depend on the precision of
the IRMS systems, the soil NO3- pool enrichment, and the incubation
parameters, such as the dimensions of the static chamber and the incubation
time (Bergsma et al., 2001; Stevens and Laughlin 2001). For our chamber
design, an incubation time of up to 20 h (which integrates the equilibration
of the added tracer solution within the soil enclosure), and using the
estimated MDC values (for both N2 and N2O) for calculating a
15XN value of 60 15N at %, the minimum detectable flux
rates were 4 µg N m-2 h-1 and
0.2 ng N m-2 h-1 for the N2 and N2O fluxes
respectively. These were significantly better than the minimum rates
(175–900 µg N2–N m-2 h-1 and
0.04–0.21 µg N2O–N m-2 h-1) reported by Bergsma
et al. (2001), Kulkarni et al. (2014), and Tauchnitz et al. (2015), using
similar field 15N tracer approaches, and comparable to the minimum rates
measured by a high-precision 15N gas flux approach in a laboratory soil
incubation (Yang et al., 2014) and the gas-flow soil core method
(8 µg N2–N m-2 h-1 and
< 1 µg N2O–N m-2 h-1) by Wang et
al. (2011). Our N2 fluxes from woodland soils compare well with the
rates reported in the literature for restored forested wetlands in North
America (Morse and Bernhardt, 2013) and with the rates from northern hardwood
forests in the USA (Kulkarni et al., 2014), using 15N tracers at
application rates similar to or lower than ours. Our results are also
comparable to the rates reported from central European forests, under similar
atmospheric N deposition rates, using the gas-flow soil core method
(Butterbach-Bahl et al., 2002). For the grassland soils, the N2 fluxes
measured in the present study were significantly lower than previous
applications of the 15N gas-flux method at high fertilizer application
rates (Baily et al., 2012; Cuhel et al., 2010; Graham et al., 2013), whilst
for the organic soils our rates were significantly lower than the ones
reported by Tauchnitz et al. (2015) since their 15N tracer application
rate (30 kg N ha-1) was 300 times higher than ours. The N2O
fluxes were up to 200 times lower than the N2 fluxes leading to low
denitrification product ratios in all land use types, a result which is in
line with the N2O yields reported from 15N tracer studies in forest
(Kulkarni et al., 2014; Morse and Bernhardt 2013) and grassland soils (Baily
et al., 2012; Bergsma et al., 2001). In the present study we have compared
the in situ denitrification rates between three major land use types using an
extended field incubation period to increase the probability of detecting a
reliable 15N–N2 signal, particularly under conditions of low
denitrifier activity due to seasonality of denitrification and/or inherent
capacity of soils (for example, organic and deciduous forest soils). However,
these rates should be considered conservative since confounding issues such
as subsoil diffusion and non-homogeneous labelling of the soil nitrate pool
may in some cases have led to underestimations of the in situ denitrification
rates.
Comparison with the AIT
The total denitrification rates measured with the C2H2 amended
intact soil cores followed the same trend as the total denitrification
(N2 and N2O fluxes combined) from the 15N gas-flux
measurements, while they were on average 168 times lower than the
denitrification potential measured in the same land use types in anaerobic
soil slurries amended with acetylene and nitrate in a previous study
(Sgouridis and Ullah, 2014), thus reflecting lower in situ rates. The AIT
denitrification rates were between 3 and 5 times lower than the 15N
gas-flux rates despite the fact that the AIT intact soil cores were capped at
the bottom, thus not allowing any subsoil diffusion of the evolved gases due
to denitrification. Therefore, the AIT rates should have been higher than the
15N gas-flux rates if serious underestimation was occurring due to
subsoil diffusion in the open-bottom static chambers, which was not the case.
Adding nitrate to the C2H2 amended cores would have been desirable
for directly evaluating the priming effect of the added substrate on
denitrification rates. The 15N tracer addition to the static chambers
corresponded to the amounts of N naturally deposited in these land use types
either via management practices and/or atmospheric deposition, thus avoiding
excessive N fertilization of the sampling plots. However, it cannot be
conclusively argued that the same amount of applied nitrate would not have
led to similar denitrification rates between the AIT and the 15N
gas-flux method. Previous comparisons between the AIT and the 15N tracer
method in field studies showed no significant difference between the two
methods in measuring in situ total denitrification rates when the tracer is
applied at high fertilization rates (50–200 kg N ha-1) and
relatively low soil moisture contents (WFPS: 40–60 %; Aulakh et al.,
1991; Mosier et al., 1986). Conversely, in laboratory incubations it was
shown that the AIT significantly underestimated total denitrification
compared to the 15N tracer approach (Yu et al., 2010) and the direct
N2 flux approach (Qin et al., 2012) due to the incomplete inhibition of
N2O reduction to N2 by C2H2 in wet soils (Yu et al.,
2010) or in soils with low nitrate content (Qin et al., 2013, 2014). In our
study, the soil WFPS ranged between 60 and 70 % in all land use types,
with the exception of the C-MW site (mean WFPS 42 %), whilst the
15N–NO3- tracer application rate was low
(< 1 kg N ha-1). Moreover, the disturbance of the soil
structure during the extraction of the soil cores and the effect of the
acetylene addition to microbial activity were not significant, as was
suggested by the similar CO2 production rates (Aulakh et al., 1991),
representing soil respiration (Felber et al., 2012), in the static chambers
and the C2H2 amended and un-amended intact soil cores. Therefore,
we could argue that it is possible that the AIT underestimated total
denitrification rates compared to the 15N gas-flux method due to the
likely incomplete inhibition of N2O reduction to N2 under
relatively high soil moisture contents, although the shorter incubation time
(2 h for the intact cores) may have limited the ability of C2H2 to
fully equilibrate within soil pore spaces. Other confounding factors such as
the catalytic decomposition of NO in the presence of C2H2 (Bollmann
and Conrad, 1997; Nadeem et al.,
2013) may have also contributed to the lower denitrification rates measured
by the AIT. This study has confirmed some of the drawbacks of the AIT as a
quantification method of in situ denitrification rates compared to the
15N gas-flux.
The estimation of the denitrification product ratio using the AIT method,
from the un-amended cores (N2O only) and the C2H2 amended
cores (N2 + N2O), is usually overestimated since the source of
N2O cannot be discriminated with the AIT, whilst the N2 flux is
underestimated due to the incomplete inhibition of N2O reduction
(Butterbach-Bahl et al., 2013). This was confirmed in the present study for
all the land use types, and even the maximum denitrification product ratio
after 2 h incubation in the case of the grassland soils (23 %) was still
significantly lower than the respective ratio from the AIT (50 %).
Therefore, the much lower denitrification product ratio estimated from the
15N gas-flux measurements is significantly more reliable, and the wider
application of this field technique across a range of land use types can have
important implications for evaluating the role of denitrification as a
reactive nitrogen sink and as a source of N2O emissions (Butterbach-Bahl
et al., 2013; Kulkarni et al., 2008).