Ecological Engineering 36 (2010) 1567–1574
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Ecological Engineering
journal homepage: www.elsevier.com/locate/ecoleng
Denitrification activity, wood loss, and N2 O emissions over 9 years from a wood
chip bioreactor
Thomas B. Moorman ∗ , Timothy B. Parkin, Thomas C. Kaspar, Dan B. Jaynes
USDA, Agricultural Research Service, National Laboratory for Agriculture and the Environment, 2110 University Boulevard, Ames, IA 50011, United States
a r t i c l e
i n f o
Article history:
Received 18 November 2009
Received in revised form 4 March 2010
Accepted 7 March 2010
Keywords:
Nitrate leaching
Denitrification wall
Microorganisms
Methane
Nitrate removal
Denitrifying microorganisms
Nitrous oxide emissions
Drainage
a b s t r a c t
Loss of nitrate in subsurface drainage water from agricultural fields is an important problem in the
Midwestern United States and elsewhere. One possible strategy for reducing nitrate export is the use
of denitrification bioreactors. A variety of experimental bioreactor designs have been shown to reduce
nitrate losses in drainage water for periods up to several years. This research reports on the denitrification
activity of a wood chip-based bioreactor operating in the field for over 9 years. Potential denitrification
activity was sustained over the 9-year period, which was consistent with nitrate removal from drainage
water in the field. Denitrification potentials ranged from 8.2 to 34 mg N kg−1 wood during the last 5 years
of bioreactor operation. Populations of denitrifying bacteria were greater in the wood chips than in adjacent subsoil. Loss of wood through decomposition reached 75% at the 90–100 cm depth with a wood
half-life of 4.6 years. However, wood loss was less than 20% at 155–170 cm depth and the half-life of this
wood was 36.6 years. The differential wood loss at these two depths appears to result from sustained
anaerobic conditions below the tile drainage line at 120 cm depth. Pore space concentrations of oxygen
and methane support this conjecture. Nitrous oxide exported in tile water from the wood chip bioreactor
plots was not significantly higher than N2 O exports in tile water from the untreated control plots, and
loss of N2 O from tile water exiting the bioreactor accounted for 0.0062 kg N2 O-N kg−1 NO3 -N.
Published by Elsevier B.V.
1. Introduction
Loss of nitrate from rain-fed agricultural fields in subsurface
drainage water contributes to the degradation of water quality in
streams, lakes and coastal waters of the United States, Europe and
elsewhere (Blann et al., 2009). Subsurface drainage from agricultural land contributes a substantial portion of base flow in rivers in
tile-drained areas of North America. Concentrations of NO3 -N exiting subsurface drains frequently exceed 15 mg L−1 in spring and
early summer (Baker et al., 1975; Gast et al., 1978; Patni et al., 1996;
Jaynes et al., 1999; Kladivko et al., 2004; Tomer et al., 2008). This
nitrogen (N) export from Midwestern tile-drained watersheds is a
contributing factor to the hypoxia problem in the Gulf of Mexico
(Rabalais et al., 1996). Research has shown that changing fertilizer
application rates and timing will reduce NO3 -N concentrations in
drainage water, but substantial losses of nitrate still occur and additional measures may be necessary to meet the nitrate-related goals
for drinking water and hypoxia reduction (Dinnes et al., 2002).
In situ bioreactors or denitrification walls are designed to intercept tile drainage or groundwater where they promote nitrate
∗ Corresponding author. Tel.: +1 515 294 2308; fax: +1 515 294 8125.
E-mail address: [email protected] (T.B. Moorman).
0925-8574/$ – see front matter Published by Elsevier B.V.
doi:10.1016/j.ecoleng.2010.03.012
removal by stimulating denitrification. Indigenous soil denitrification capacity generally decreases deeper in the soil due to the
decreased carbon substrate or smaller populations of denitrifiers
(Parkin and Meisinger, 1989; Yeomans et al., 1992; Sotomayor
and Rice, 1996; Richards and Webster, 1999). Thus, nitrate that
leaches beneath the surface soil is prone to entry into subsurface
drains. Bioreactors and denitrification walls contain carbon substrates, such as wood chips, that support microbial metabolism and
denitrification as the drainage water passes through the wall or
bioreactor.
Bioreactors have successfully decreased nitrate concentrations
in drainage water and shallow ground water at a number of locations throughout the world (Blowes et al., 1994; Robertson and
Cherry, 1995; Volokita et al., 1996; Schipper and Vojvodic-Vukovic,
1998; Robertson et al., 2000). Bioreactors constructed by Blowes
et al. (1994) decreased the 3–6 mg NO3 -N L−1 in the agricultural
drainage water to <0.2 mg L−1 using tree bark, wood chips and
leaf compost as carbon (C) sources. Schipper and Vojvodic-Vukovic
(1998, 2000) obtained NO3 -N removal from agricultural ground
water passing through a trench filled with a mixture of soil and
sawdust (a denitrification wall). Removal was attributed to denitrification based on enhanced denitrification enzyme activity within
the wall compared to unamended soil outside the wall. Jaynes et
al. (2008) used wood chip-based denitrification walls placed on
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T.B. Moorman et al. / Ecological Engineering 36 (2010) 1567–1574
either side of a subsurface drainage line to remove nitrate from
corn–soybean rotation drainage water. Over 5 years of operation,
flow-weighted nitrate concentrations in subsurface drainage water
averaged 8.8 mg NO3 -N L−1 in the treated drainage water compared
to 22 mg NO3 -N L−1 in untreated drainage water. Two small wood
chip bioreactors in Ontario reduced nitrate over 4 years by 32% and
53% in drainage from a corn field and a golf course, respectively
(van Driel et al., 2006).
These prior results demonstrate that long-term nitrate removal
can be achieved, but several important issues remain unresolved
concerning bioreactor design and performance. The duration of
effectiveness is determined by the duration of carbon supply to the
denitrifying microorganisms. Denitrification consumes some of the
substrate, but other microbial processes such as aerobic respiration,
sulfate reduction, and fermentation also consume the substrate.
The dynamics of wood substrate decay and associated denitrifier
activity and populations have not been extensively investigated. In
addition, there are concerns over increased emissions of the greenhouse gas N2 O, which is a product of the denitrification reaction. In
short-term laboratory experiments, N2 O production accounted for
less than 0.03% of the N denitrified (Greenan et al., 2009). However,
it is unclear if bioreactors will be as efficient in fully converting
nitrate to N2 in the field where weather, hydrology, and nitrate
fluxes all fluctuate.
This report describes measurements conducted on the bioreactors described by Jaynes et al. (2008) and on a similar adjacent
bioreactor. Our objectives were to determine the potential denitrification rates in samples of wood and soil and to determine the
changes of wood mass over time from this system. Additionally, we
report on an in situ estimation of denitrification based on losses of
15 NO -N introduced into the wall, and on N O-N produced within
3
2
the wall and exported in tile drainage water.
2. Materials and methods
2.1. Experimental site and bioreactor design
The research was conducted on samples of wood, soil, gas and
water collected from denitrification walls described previously
(Jaynes et al., 2008) and from an additional wall established concurrently. The site is located in Boone County 8.0 km northwest
of Ames, IA (42.05◦ N, 93.71◦ W). Average annual precipitation is
83.7 cm and drainage typically occurs from late March till midJuly. The soils are mapped as the Canisteo and Nicollet series
with the surface soils averaging 27.6% clay, 54.6% silt and 17.8%
sand. Soil organic carbon content was 4.5% in the 0–7.5 cm depth,
3.2% in the 7.5–15 cm depth, 2.1% in the 15–30 cm depth and
0.7–0.1% at depths from 30 to 120 cm. These soils require subsurface drainage due to the underlying glacial tills that restrict vertical
water flow (Eidem et al., 1999). Replicated plots (n = 4), 30.5 m wide
and 42.7 m long, were installed with a perforated plastic drainage
pipe (7.62 cm diameter, 1.2 m depth) oriented lengthwise in center of the plot. The denitrification walls consisted of two trenches
0.6 m wide and 1.83 m deep, and offset 3.05 m on either side of
the drainage pipe. These also ran lengthwise through the plots
parallel to the drainage pipe. The bioreactors consisted of these
trenches filled with wood chips and a small amount of soil and
covered with 30 cm of soil at the surface. The bioreactors were
installed in September of 1999 and the plots were cropped with
a corn–soybean rotation starting with a corn crop in 2000. The
effect of the bioreactors on nitrate concentration and load was
compared to plots with subsurface drainage, but without the wood
chip trenches (controls). These replicated plots were used for the
quantification of drainage volume and nitrate losses as described
by Jaynes et al. (2008), for the measurement of dissolved gasses in
drainage water, and the in situ measurement of nitrate removal by
push–pull methods (described subsequently).
In addition to these replicated plots, a similar trench filled with
wood chips was also constructed approximately 8 m from the edge
of the plots at the same time to provide a wall that could be disturbed to obtain samples over time for measurement of wood
loss and denitrification potential. Weighed quantities (70–90 g dry
weight) of the same wood chips used in the main plots were
enclosed in fiberglass mesh bags, tagged and buried in this ancillary trench at depths of 60 and 150 cm. Nylon cord extended from
each bag to the soil surface to facilitate recovery of the bags. Corn
and soybeans were cropped over this area in a manner similar to
that in the main plots. Pore space gas samplers were installed in
the test wood chip trench at depths of 60 cm, 90 cm, 120 cm and
150 cm below the soil surface. Samplers consisting of 3 m lengths
of porous silicone tubing (3.18 mm inner diameter, 0.79 mm wall
thickness (Cole Parmer Inst. Co., Vernon Hills, IL) were buried at
each depth. Non-porous plastic tubing was connected to the ends
of the silicone tubing and extended to the soil surface where they
were capped with rubber septa to facilitate gas sample removal
with a needle and syringe.
2.2. Sampling
On June 19, 2002, July 15, 2003, August 19, 2004, August, 7,
2008, and August 18, 2009, we excavated wood chip bags from
different depths of the ancillary trench. There is some variation
in the depths reported for recovery of wood and the denitrification potential measurements and this is likely due to settling of
the wood over time. At the same time additional samples of the
wood chips adjacent to the bags and soil samples from the exposed
sidewall of the trench at the same depth (±10 cm) were also taken.
Four replicate sets of samples were obtained at each depth. Surface
soil was sampled at the 0–15 cm depth. Samples were transported
to the laboratory and stored at 4 ◦ C until analysis. Wood chips in
the mesh bags were vigorously washed with water to remove any
soil, dried at 65 ◦ C, and weighed. Wood chips were also ground and
analyzed for total C and N (Greenan et al., 2006). On June 10, 2009
we used a 5-cm diameter hydraulic soil sampler to obtain continuous cores of soil and wood from the main plots in the drainage
study. Four replicate plots were sampled and wood from 90 ± 15 cm
and 180 ± 15 cm was obtained. Soil from outside the denitrification
walls was also obtained from these depths.
Pore space gasses were collected on 3 occasions during April
and May, 2000 and on 3 occasions during April and May, 2001 for
determination of O2 , CH4 and N2 O. Gas samples were collected by
inserting a 20 cc syringe equipped with a stopcock valve through
the rubber septa of the gas sampling tubing. The plunger on the
syringe was withdrawn to the full capacity of the syringe and the
stopcock valve was closed. The syringe plunger was then set to 10 cc
and the stopcock briefly opened to release the overpressure in the
syringe. The 10 cc gas sample remaining in the syringe was injected
into an evacuated glass vial sealed with a butyl rubber stopper.
The vials were returned to the laboratory for gas chromatographic
analyses (see below for details).
Five times during the period of March–May, 2001 tile water
from the main wood chip and control plots was sampled to determine dissolved N2 O concentrations. Samples were collected by
withdrawing a 10 ml volume of water flowing from the tile lines
with a syringe. The water was immediately injected into evacuated 26.5 ml test tubes sealed with butyl rubber stoppers, and the
test tubes were placed on ice. In the laboratory the headspace pressure in the test tubes was adjusted to atmospheric pressure with
He, the tubes were shaken to equilibrate the dissolved N2 O with
T.B. Moorman et al. / Ecological Engineering 36 (2010) 1567–1574
the headspace, and N2 O concentrations of the headspace of each
vial was determined with a gas chromatograph. The Bunsen coefficient for N2 O was used to determine dissolved N2 O concentrations
in the tile water (Tiedje, 1994).
Nitrate in the tile water from each plot was measured on 5 occasions during March through June, 2004 using the sampling and
analysis procedures described by Jaynes et al. (2008).
2.3. Gas analyses
A Tracor (Model 450) gas chromatograph equipped with a flame
ionization detector (FID) and a thermal conductivity (TC) detector was used to determine CH4 and O2 concentrations. For CH4 , a
stainless steel column (1.8 m long × 0.318 cm diam.) packed with
Porapak Q (60/80 mesh) was used with He carrier at 30 ml min−1
and oven temperature of 45o C and an FID temperature of 200 ◦ C.
Oxygen was determined with a 1.8 m long × 0.318 cm diameter
stainless steel column packed with molecular sieve. Helium carrier gas (20 ml min−1 ) and an oven temperature of 45 ◦ C with the
TC detector temperature of 175 ◦ C were used. Nitrous oxide concentrations in samples were determined with Shimadzu Mini-2
gas chromatograph (Kyoto, Japan) equipped with a 63 Ni electron
capture detector and a stainless steel column (0.318 cm diameter × 1.8 m long) with Porapak Q (80/100 mesh), with N2 carrier
(20 ml min−1 ) and column and detector temperatures of 70 ◦ C and
325 ◦ C, respectively. Certified standard gases (Scott Specialty Gases,
Troy, MI) were used to construct standard curves for each gas. The
detection limits for N2 O, CH4 , and O2 were 0.1 ppm, 0.5 ppm, and
0.1%, respectively.
2.4. Denitrification potential assays
Denitrification potential assays were performed using the
acetylene block method described in general by Tiedje (1994). This
technique supplies non-limiting amounts of glucose and nitrate to
denitrifying microorganisms. This, in conjunction with chloramphenicol to inhibit further protein synthesis, results in production
of N2 O in proportion to the quantity of active denitrifying microorganisms present in the original sample. Fifty g of field-moist wood
or soil was immersed in 100 mL of solution containing 1 mM KNO3 ,
250 mg L−1 of chloramphenicol, and 1 mM glucose in 500 ml glass
bottles. Bottles were sealed and the air was vacuum exchanged
with He three times. Acetylene at 8–10% of the headspace volume
was added and mixed using a gas tight syringe. Pressure in each jar
was adjusted to one atmosphere and the bottles were incubated
for up to 48 h and sampled 4–5 times in that period. Nitrous oxide
was measured by gas chromatography and rates of N2 O production
were calculated and expressed per unit dry weight of wood or soil.
The longer incubation times were used to ensure that measurable
N2 O concentrations were obtained in the subsurface soils. Rates of
N2 O production from wood or surface soils were typically obtained
from measurements made in the first 12 h or less. Beginning in 2003
and subsequent years, the denitrification assay was also conducted
using solutions without glucose.
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were made from these extracts and 5 replicate tubes were inoculated at appropriate dilutions, and then incubated for 14 d at 25 ◦ C.
The presence of nitrate and nitrite was evaluated with diphenylamine, but the primary endpoint was production of N2 O which
was detected by gas chromatography, as described above. The mpn
was calculated from a 5-place mpn table. Uninoculated media produced no nitrous oxide and tested positive for nitrate and nitrite.
In general, production of N2 O was observed in tubes at greater
dilutions, even when nitrate and nitrite were not completely
consumed.
2.6. In situ denitrification
In mid-July of 2004 we performed an in situ push-pull denitrification test by injecting 57 L of 25 mg NO3 -N L−1 enriched with 15 N
at 20 atom% and Br (628 mg L−1 ) into a small diameter PVC well
installed to a depth of 1.2 m in the center of the wood trench. Three
additional wells were installed at 50 cm spacings in either direction in the center of the wood chip trench. The nitrate solution also
contained 9.7 mg NO2 -N L−1 and 11.1 mg Cl L−1 . Well water samples taken 3 days prior to injection contained no nitrate or nitrite,
9.9–11.1 mg Cl L−1 , and 0–0.9 mg Br L−1 as determined by anion
chromatography (Dionex ICS-2000). Rates of in situ denitrification
were estimated from the recovery of nitrate and bromide using
methods described by Haggerty et al. (1998). To simplify the analysis, sample concentrations of nitrate and nitrite were summed,
expressed as concentrations relative to the initial concentrations
(C/C0 ), and the rate constant describing the disappearance of these
species was determined by non-linear regression.
2.7. Statistical analysis
We evaluated trends in nitrate losses from tile drainage by performing linear regression of the mean loss of NO3 -N against time
(years after installation of the bioreactor). A similar linear regression analysis was used to evaluate the change over time of the
denitrification potentials of the wood chips. The glucose amended
denitrification potential data for the 90–127 cm depths were evaluated in one regression and 155–180 cm depth were evaluated in
a second regression. Differences in the denitrification potentials of
wood and soil, including the effect of glucose on potential denitrification, were determined by analysis of variance. The data for
the surface soil were omitted from this analysis because there is
no corresponding wood chip layer. Because different depths were
analyzed in different years, the analysis of variance was performed
separately for each year. The analysis of variance was performed as
a completely randomized factorial design with the matrix (wood or
soil), depth, and glucose (amended or unamended) as main effects.
The analysis was performed using SAS (SAS Institute, 1985) using
the general linear models procedure (Proc GLM). When significance
was indicated by the analysis of variance, mean separation tests
were performed.
3. Results and discussion
2.5. Enumeration of denitrifying bacteria
3.1. Nitrate removal from drainage water
Bacteria capable of denitrification were enumerated using a
most-probable-number (mpn) technique (Tiedje, 1994). The mpn
medium contained 8 g L−1 of nutrient broth and 3 g L−1 of KNO3 .
Tubes were sealed with butyl rubber septa, evacuated and flushed
with He three times, then filled with He and sterilized. Acetylene
(10% of headspace) was added after autoclaving. Twenty grams of
wood or soil were placed into 0.0125 M phosphate buffer (pH 7.2)
and shaken on a reciprocating shaker for 20 min. Serial dilutions
Fig. 1 shows NO3 -N loss in tile drainage data from Jaynes et al.
(2008) with additional data obtained from 2006 to 2008. From 2001
to 2008 annual nitrate loss in plots with conventional drainage
averaged 54.5 kg NO3 -N ha−1 compared to 24.5 kg NO3 -N ha−1 in
plots with the denitrification walls. Slopes from linear regression
analysis (NO3 mass loss vs. time) were not significantly different
from zero, showing that both the control treatment and the wood
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T.B. Moorman et al. / Ecological Engineering 36 (2010) 1567–1574
Fig. 1. Annual losses of NO3 -N in subsurface drainage for a conventional drainage
system (control) and drains with wood chip denitrification walls.
trench bioreactors did not establish a trend over time in their NO3 N losses in drainage water.
3.2. Denitrification potential activity
Analysis of variance showed a significant (P ≤ 0.01) effect of
the matrix (wood or soil) and a significant effect of depth for
years 2002–2008 in denitrification potential activity at the P ≤ 0.05
level. Depth was also significant at the P ≤ 0.10 level in 2009
(Tables 1 and 2). Depth was also significant for the 2009 data, but at
the P ≤ 0.10 level. In each year there was a matrix × depth interaction that was significant at the P ≤ 0.05 level, except for 2009, which
was significant at the P ≤ 0.10 level. The mean potential denitrification rates for the wood and soil at each depth averaged across the
glucose amended and unamended treatments are shown in Table 2.
In the subsoil, denitrification potentials are smaller than those of
the surface soils, but there were no significant differences among
the subsoil denitrification potentials. The denitrification potentials
of the wood chips were greater than the potentials of the subsoils,
but there were no consistent effect of depth within the wood chip
trench. Averaged over all depths and glucose treatments (amended
and unamended) wood increased denitrification potentials by a
range of 31-fold in 2003 to 4000-fold in 2004 over the denitrification potential in subsurface soil.
The effect of glucose amendment and the glucose × depth interaction were significant only at the P ≤ 0.10 level in 2004 (Table 1).
In all other years these effects were not significant. The mean denitrification potential rates for the wood and soil at each depth
averaged across the glucose amended and unamended treatments
are shown in Table 2. Glucose amendment increased the denitrification potential of surface soil in 2003 and 2008, by factors of 11
and 38 respectively. Our results are in contrast to reports by Murray
et al. (2004) and McCarty and Bremner (1992) showing that subsoil denitrification was enhanced by glucose additions. However,
Richards and Webster (1999) reported only marginal stimulation of
subsurface denitrification potential in subsoils by glucose. In wood
chips, the glucose amendment had no consistent effect on denitrification potential suggesting that the microbial community in wood
was not carbon-limited or was non-responsive to glucose. Schipper
et al. (2005) also reported no effect of glucose on denitrification in
a wood-based denitrification wall.
The trend in glucose-amended denitrification potential over
time was evaluated by linear regression. After grouping the data
from 90 to 127 cm depths, the regression analysis (r2 = 0.15)
resulted in slope of 1738 ± 1850, which was not significantly different from zero. Similarly, for the data from the 155 to 180 cm
Fig. 2. Median populations of denitrifying microorganisms in soil at different depths
and in adjacent samples of wood collected in 2009 from the denitrification wall, as
determined by a most-probable-number technique.
depths the regression (r2 = 0.04) line slope was not different from
zero (490 ± 1350). Schipper and Vojvodic-Vukovic (2001) reported
that after 5 years denitrification potential in a sawdust based denitrification wall declined to about 10% of the activity measured when
the wall was initially installed.
The denitrification potential measurement provides an unambiguous measure of denitrification (nitrous oxide formation in
the presence of acetylene under anaerobic conditions), and these
results also support our earlier laboratory study (Greenan et al.,
2009) indicating that denitrification is the mechanism of nitrate
removal in wood chip bioreactors. Greenan et al. (2009) reported
that under different water flows nitrate removal rates for small
wood chip bioreactors ranged from 11 to 15 mg of N kg−1 wood d−1 .
The potential denitrification rates for wood in this study ranged
from 8.2 to 34.4 mg N kg−1 wood d−1 in 2004, 2008 and 2009. These
rates of N removal are also similar to those reported for various aged
woods by Robertson (2010).
3.3. Populations of denitrifying microorganisms
Populations of denitrifying microorganisms exceeded 108 g−1
wood, compared to populations above 107 g−1 in surface soil and
populations near 106 g−1 in the subsurface soils (Fig. 2). These populations were correlated (r2 = 0.92) to the potential denitrification
potential rates obtained from the same wood and soil materials
sampled in 2009 (Table 2). The populations of denitrifiers in surface soils are slightly greater than populations reported in other
studies which range between 103 and 106 cells g−1 soil (Martin
et al., 1988; Murray et al., 1995) or 1–5 × 105 nosZ copies g−1 soil
(Ma et al., 2008). Parkin and Meisinger (1989) reported declines in
denitrifier populations with increasing soil depth, reaching nondetectable levels at 160 cm depth. However, Sotomayor and Rice
(1996) reported populations between 105 and 106 cells g−1 soil to
depth of 10 m in a Kansas soil. Clearly, the wood provides a habitat
that supports a larger population of denitrifiers than the adjacent
subsoil.
3.4. In situ nitrate removal
The in situ push-pull test also confirmed rapid nitrate removal
(Fig. 3). Immediately after injection, recovery of Br was only 39%
indicating rapid mixing and dispersal of the nitrate + bromide solution. Isotope ratio analysis indicated that the dilution of injected
T.B. Moorman et al. / Ecological Engineering 36 (2010) 1567–1574
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Table 1
Summary of analysis of variance results for denitrification potentials in subsoil and bioreactor wood chips.
ANOVA component
Matrix (wood, soil)
Depth
Glucose (+/−)
Matrix × depth
Matrix × glucose
Depth × glucose
Matrix × depth × glucose
a
b
c
Year
2002
2003
2004
2008
2009
P ≤ 0.01a
P ≤ 0.01
NDb
P ≤ 0.01
ND
ND
ND
P ≤ 0.01
P ≤ 0.01
NSc
P ≤ 0.01
NS
P ≤ 0.01
P ≤ 0.01
P ≤ 0.01
P ≤ 0.05
P ≤ 0.10
P ≤ 0.05
P ≤ 0.10
NS
NS
P ≤ 0.01
P ≤ 0.05
NS
P ≤ 0.05
NS
NS
NS
P ≤ 0.01
P ≤ 0.10
NS
P ≤ 0.10
NS
NS
NS
P: level of significance.
ND: not determined.
NS: not significant.
Table 2
Denitrification potential of bioreactor wood and soil collected in 2002 through 2009.
Depth
Denitrification potential (␮g N kg−1 d−1 )a
Soil
Unamended
Wood
b
Glucose
Mean
2002
15
60
90
122
155
NDc
ND
ND
ND
ND
10,100
11
13
17
158
±
±
±
±
±
5000
4
7
2
95
2003
15
91
127
160
2770 ± 1390
12 ± 8
36 ± 25
570 ± 321
30,700
40
71
849
±
±
±
±
7110
33
47
110
2004
100
170
1.4 ± 0.8
1.0 ± 0.6
2008
15
100
170
338 ± 393
211 ± 105
72 ± 51
2009
15
90
180
3760 ± 2740
46 ± 46
19 ± 13
2.5 ± 1.4
7.8 ± 8.7
12,800 ± 8810
174 ± 103
21 ± 35
4620 ± 4130
90 ± 107
22 ± 17
Meanb
Unamended
Glucose
ND
ND
ND
ND
ND
ND
168 ± 71
236 ± 32
16,600 ± 7300
20,100 ± 4170
25 a
53 a
709 a
ND
4220 ± 1600
12,800 ± 1130
7070 ± 924
ND
2450 ± 258
18,000 ± 2390
4070 ± 485
3330 b
15,400 d
5570 c
1.9 a
4.3 a
13,500 ± 4270
8208 ± 1480
17,065± 2136
12,100± 4110
15,300 c
10,200 b
192 a
46 a
ND
34,400 ± 13,700
14,100 ± 8130
ND
32,600 ± 16,400
18,800 ± 12,300
33,500 c
16,500 b
68 a
20 a
ND
7830 ± 5360
11,400 ± 4580
ND
7810 ± 5510
14,300 ± 7460
7820 b
12,900 c
a
Values shown are means of four replicates and standard deviations.
Mean rates for glucose-amended and unamended assays for soil and wood, respectively. Differences in means for wood and soil by depth within the same year are
indicated by the letters following the mean. Means followed by the same letter are not significantly different at the P ≤ 0.05 level.
c
ND: not determined.
b
15 NO -N with unlabeled NO -N was negligible. The first-order non3
3
linear regression of the ratio of nitrate (C/C0 ) to bromide (C/C0 )
results in a rate constant of 1.08 ± 0.23 d−1 , which correspond to
a nitrate half-life of 0.64 d. Assuming 50% porosity of a bioreactor, a change in concentration of 20 to 10 mg NO3 -N over 0.64 d
and a wood density of 0.66 g cm−3 (Greenan et al., 2009), a linearized rate of nitrate disappearance was estimated at 23.6 mg N kg
wood−1 d−1 , which is similar to the rates estimated for wood using
the potential denitrification methods. Similar results were seen in
the observation well 50 cm to the south of the injection well (Fig. 3).
Analysis of the NO3 /Br ratio resulted in a NO3 disappearance rate
constant of 0.86 ± 0.19, which corresponds to a half-life of NO3 of
0.81 days in this well. Similar results were obtained in the two other
observation wells (data not shown).
3.5. Tile water N2 O
It is calculated, based on a N2 O Bunsen absorption coefficient of
0.882, that water at 10 ◦ C (the average soil temperature at the 1.5 m
depth in 2004 was 10.1 ◦ C) in equilibrium with N2 O at ambient
atmospheric concentrations should contain N2 O at approximately
0.36 ␮g N2 O-N L−1 . Dissolved N2 O concentrations from both treatments, at all sampling times were higher than expected background
concentrations and ranged from 2.6 to 73.2 N2 O-N L−1 (Table 3). We
observed that N2 O concentrations tended to be higher in tile water
Table 3
Average tile water flow rates and dissolved N2 O concentrations in tile water on 5
dates in 2004.
Date
March 11
March 23
May 20
June 2
June 16
a
Tile water flow (L d−1 )a
Dissolved N2 O (␮g-N L−1 )a
Control
Control
Woodchip
5.68 (1.54)
4.33 (4.02)
2.64 (0.72)
45.2 (58.1)
41.3 (14.9)
23.3 (20.9)
13.5 (8.49)
15.9 (12.2)
32.2 (61.1)
73.2 (47.8)
1681 (605)
2150 (423)
1266 (611)
2942 (301)
1624 (1450)
Woodchip
2010 (755)
2359 (834)
1275 (359)
3478 (1500)
1600 (857)
Values shown are means of 4 replicate plots and associated standard deviations.
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T.B. Moorman et al. / Ecological Engineering 36 (2010) 1567–1574
Fig. 3. The ratio of relative concentrations (C/C0 ) of nitrate (NO3 -N + NO2 -N) to bromide in the wood chip trench groundwater recovered over time from the injection
well (IW) and an observation well (OW) 50 cm from the injection well. Solid lines
show the pseudo-first order regressions for the nitrate/bromide ratio.
Fig. 4. Cumulative export of dissolved N2 O in tile water of control and woodchip
treatments over period of March 11 to June 16, 2004. Shown are means and associated standard errors.
flowing from the woodchip treatment than the control; however,
these differences were not significant (P > 0.10) on any sampling
date. Export of N2 O in tile water was calculated from the tile water
N2 O concentrations along with water flow rates (Fig. 4). Over the
3 month period from March 11, 2004 to June 16, 2004 cumulative N2 O export in tile drainage water averaged 15.1 g N ha−1 for
the woodchip treatment and 9.5 g N ha−1 for the control, although
there was no significant difference between treatments (P = 0.47).
Nitrate export in tile drainage waters over this same time period is
shown in Fig. 5. Cumulative NO3 export from the control plots was
significantly higher (P < 0.05) than the woodchip treatment (2.45 kg
NO3 -N ha−1 vs. 1.15 kg NO3 -N ha).
Removal of NO3 by the wood chip bioreactor has obvious
beneficial water quality effects; however, a potential negative
impact of NO3 removal by denitrification is the concomitant production of the greenhouse gas, N2 O. Little information exists on
the export of N2 O leached from the soil profile and exported
in tile drainage water. The Intergovernmental Panel on Climate
Change designated an emission factor (EF5g ) of 0.0025 kg N2 ON kg−1 NO3 -N for calculating indirect N2 O emissions based on the
mass of NO3 leached in ground and surface waters from agricultural lands (IPCC, 2006), and an additional 0.005 kg N2 O-N kg−1
NO3 -N transported from drainage water to rivers and estuaries.
In this study, from the control plots we observed an average NO3
Fig. 5. Cumulative export of NO3 in tile water of control and woodchip treatments
over period of March 11 to June 16, 2004. Shown are means and associated standard
errors.
leaching loss of 2.45 kg NO3 -N ha−1 over the period of March
11, 2004, to June 16, 2004. Over this same period average N2 O
from the control treatment was 0.0095 kg N2 O-N ha−1 yielding
a calculated EF5g emission factor of 0.0039 kg N2 O-N kg−1 NO3 N leached. Export of N2 O in the tile water of the wood chip
treatment averaged 0.015 kg N2 O-N ha−1 yielding an EF5g emission factor of 0.0062 kg N2 O-N kg−1 NO3 -N. However, this higher
observed EF5g in the wood chip treatment does not necessarily
mean higher overall indirect emissions, because lower amounts
of NO3 were exported in the tile water of the woodchip treatment. Thus, less NO3 was available for subsequent denitrification
downstream. Application of the IPCC default emission factors associated with rivers and estuaries (EF5r,e = 0.005 kg N2 O-N kg−1 NO3 )
to the NO3 exported in the tile drainage waters of the control and
wood chip treatments yield estimates of 0.012 kg N2 O-N ha−1 and
0.0057 kg N2 O-N ha−1 , respectively. Because lower amounts of NO3
were exported in the tile water of the woodchip treatment, corresponding estimates of indirect N2 O emissions are also lower.
By combining the measured tile water N2 O values with the IPCC
estimates of N2 O emissions from rivers and estuaries we determine that the overall indirect emissions associated with NO3 loss
of our treatments of our study to be 0.0088 kg N2 O-N kg−1 NO3 N for the control plots and 0.0085 kg N2 O-N kg−1 NO3 -N for the
wood chip plots. Thus, the wood chip treatment, while facilitating
greater loss of drainage NO3 through denitrification did not significantly increase overall indirect N2 O emissions compared to the
control.
3.6. Pore space gasses
Differences in aeration of the wood chip material at different
depths in the test wood chip trench were observed (Fig. 6). In the
spring of 2000 O2 pore space concentrations exceeded 18% at the
30, 60, 90, and 120 cm depths, but at 150 cm O2 concentrations
were 1.3% (Fig. 6A). A similar trend was observed in 2001, except
that O2 concentration at the 120 cm depth was 1.5% and was <1.0%
at 150 cm. The anaerobic nature of the woodchip trench is further
indicated by the elevated pore space CH4 concentrations at the
greater depths (Fig. 6B). Methane concentrations exceeding ambient levels (approx. 1.7 ␮L L−1 ) were detected at 90 cm, 120 cm, and
150 cm in 2001 and at 120 cm and 150 cm in 2000. Nitrous oxide
concentrations were only slightly greater than ambient concentrations at 60 cm, 90 cm, and 120 cm in 2000, but in 2001 pore
space N2 O concentrations exceeded 2 ␮L L−1 at the 90 cm depth
and exceeded 5 ␮L L−1 at 120 cm (Fig. 6C). Nitrous oxide concen-
T.B. Moorman et al. / Ecological Engineering 36 (2010) 1567–1574
Fig. 6. Pore space concentrations of oxygen, methane and nitrous oxide in the test
woodchip trench. Samples were collected on three dates in 2000 and on three dates
in 2001. Shown are means for each year and corresponding standard errors.
1573
tions of the year, the 90–100 cm depth would become aerobic as
the water table drops to the level of the drainage pipe (120 cm),
which would accelerate the wood decay. In contrast, the wood
at the deeper depths, 100–190 cm, there was less loss with time,
which is likely due to longer periods where the wood is saturated
and decomposes more slowly. The pore space gas measurements
(Fig. 6) demonstrate that the lower depths of the bioreactor are
more anaerobic. There were generally lower denitrification potentials measured at the 160–190 cm depths in 2003, 2004, and 2008,
but not in 2009 (Table 2). Schipper and Vojvodic-Vukovic (2001)
reported no loss of carbon over 5 years in that denitrification wall.
Changes in wood C and N content also occurred over the first
7.9 years of operation. The C content of the wood at the 90–100 cm
depth decreased from 49.4 ± 0.1% initially to 43 ± 2.7% in 2004, but
was 48.7 ± 0.7% in 2008. Total N content increased from 0.11% to
0.26 ± 0.03% in 2004 and remained nearly constant at 0.25 ± 0.03%
in 2008. The wood carbon content at the deeper depth was nearly
constant, changing from 49.4 ± 0.1% to 46.8% over the 7.9 years
period. The increased wood N content at the 90–100 cm depth
could be due to immobilization of NO3 -N by microorganisms or to
microbial retention of the wood-derived N during decomposition.
Previously, our studies utilizing 15 NO3 -N have shown 1.0–6.9% of
added NO3 -N to be immobilized in relatively short term studies
with these same wood chip substrates (Greenan et al., 2006, 2009).
The longevity of the wood chips is an important factor in bioreactor design and expected performance. While the denitrification
potentials expressed on a mass basis show sustained activity over
time, the loss of wood at the 90–100 cm depth should be taken into
account in the interpretation of those data. While 75% of the wood
was degraded over 8.9 years at the 90–100 cm depth, the denitrification capacity of the remaining material supports nitrate removal
in the field. The recovery of the wood chip bags at depths 30 cm
lower than where they were originally placed indicates that the
decayed wood has been partly replaced by wood from above the
sample bags through subsidence.
4. Summary and conclusions
trations at the 150 cm depth in both years were near background
levels.
The sustained potential denitrification activity in wood chips
samples over the 9-year period is consistent with the performance of the denitrification wall in the field. The wood chip matrix
within the bioreactor supports enhanced populations of denitrifying microorganisms in the wood chip bioreactors compared to
adjacent subsoils. The in situ test results also suggest that hydraulic
residence times of 24 h should be sufficient to reduce initial nitrate
concentrations of 20–25 mg NO3 -N L−1 entering the bioreactor to
10 mg NO3 -N L−1 or below. Wood mass loss (75%) within the wall
was greatest in the more shallow part of the bioreactor, where aerobic decomposition would be more prevalent. Indirect losses of
N2 O due to denitrification in this bioreactor are equivalent to N2 O
emissions from tile drainage without the bioreactor.
3.7. Loss of wood over time
Acknowledgements
The loss of wood at the 90–100 cm depth averaged 50% for years
2003 and 2004 (3.8 and 4.9 years, respectively), then increased to
75% loss by 2009 (8.9 years) (Fig. 7). Less than 13% of the wood was
decomposed at the 155–170 cm depth. A first-order decay curve
for wood loss at the 90–100 cm depth produced a significant fit
(P ≤ 0.01, r2 = 0.94) and predicted half-life of 4.6 years. For the wood
at the 155–170 cm depth, the first-order regression was also significant (P < 0.05, r2 = 0.91) and resulted in a 36.6 years half-life. The
differential loss of wood at these two depths is most likely due to
the greater degree of water saturation at the lower depth. For por-
This project was funded in part by grant 98-35102-6953 from
CSREES National Research Initiative NRI-CGP and grant 59-3625604 from the American Farm Bureau Foundation for Agriculture.
We thank Kent Heikens, Ben Knutson, Colin Greenan, Amy Morrow,
Beth Douglass and Otis Smith for their help in this research.
Fig. 7. Loss of wood chip mass from a denitrification wall in the field. Wood chips
were placed in mesh bags and recovered from the indicated depths over time.
Loss was determined by weight difference. Solid lines show the first-order nonlinear least squares regression. Points indicate means of four samples and associated
standard errors.
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Dr. Thomas Moorman serves as Microbiologist in the USDA-ARS National Laboratory for Agriculture and the Environment where he conducts research on fate of
agricultural chemicals in soils and water.
Dr. Timothy Parkin is a Microbiologist in the USDA-ARS National Laboratory for
Agriculture and the Environment where he investigates nitrogen transformations
in soil and greenhouse gas emissions from agricultural systems.
Dr. Thomas Kaspar is a Plant Physiologist in the USDA-ARS National Laboratory for
Agriculture and the Environment where he researches the management of cover
crops and nitrogen fertilizer in corn and soybean production systems.
Dr. Dan Jaynes is a Soil Scientist in the USDA-ARS National Laboratory for Agriculture and the Environment where he conducts research on agricultural drainage
management and water quality.