Environ. Sci. Technol. 2010, 44, 7527–7533
Nitrous Oxide Emissions from a
Large, Impounded River: The Ohio
River
the pool during two successive summer surveys. We quantified several sources of N2O to the river including wastewater
treatment plant (WWTP) effluent and microbial N2O production (e.g., nitrification and denitrification) in the water
column and sediments.
J. J. BEAULIEU,* W. D. SHUSTER, AND
J. A. REBHOLZ
National Risk Management Research Laboratory, Office of
Research and Development, U.S. Environmental Protection
Agency, 26 West Martin Luther King Drive, Cincinnati,
Ohio 45268
Experimental Section
Received May 17, 2010. Revised manuscript received
August 9, 2010. Accepted August 10, 2010.
Models suggest that microbial activity in streams and rivers
is a globally significant source of anthropogenic nitrous oxide
(N2O), a potent greenhouse gas, and the leading cause of
stratospheric ozone destruction. However, model estimates of
N2O emissions are poorly constrained due to a lack of
direct measurements of microbial N2O production and consequent
emissions, particularly from large rivers. We report the first
N2O budget for a large, nitrogen enriched river, based on direct
measurements of N2O emissions from the water surface and
N2O production in the sediments and water column. Maximum
N2O emissions occurred downstream from Cincinnati, Ohio,
a major urban center on the river, due to direct inputs of N2O
from wastewater treatment plant effluent and higher rates
of in situ production. Microbial activity in the water column
and sediments was a source of N2O, and water column production
rates were nearly double those of the sediments. Emissions
exhibited strong seasonality with the highest rates observed
during the summer and lowest during the winter. Our results
indicate N2O dynamics in large temperate rivers may be
characterized by strong seasonal cycles and production in the
pelagic zone.
Introduction
Atmospheric concentrations of nitrous oxide (N2O), a potent
greenhouse gas with a global warming potential nearly 300
times that of carbon dioxide (1) and the leading cause of
stratospheric ozone destruction (2), are rising by 0.26% per
year (1). The primary anthropogenic source of N2O is the
biological conversion of nitrogen (N) to N2O in terrestrial
and aquatic ecosystems (3). Nitrous oxide production in
agricultural soils has been well studied with over 1000
published measurements and is a relatively well constrained
component of the global N2O budget (4). Anthropogenic N2O
production in rivers which receive anthropogenic N in runoff
and sewage inputs may be as large as 1.7 Tg N y-1 or 25%
of the global N2O budget (1, 5). However, this estimate is
uncertain, partially due to a lack of N2O emission measurements from large rivers made over annual temporal scales.
In this study we measured the production and emission
of N2O from the Markland Pool of the Ohio River, which is
ranked by annual discharge as the third largest river in North
America. Nitrous oxide emission rates were measured
biweekly for 13 months at one site and along a transect of
* Corresponding author e-mail: [email protected].
10.1021/es1016735
Not subject to U.S. Copyright. Publ. 2010 Am. Chem. Soc.
Published on Web 08/30/2010
The Ohio River is formed by the confluence of the Allegheny
and Monongahela Rivers in Pittsburgh, Pennsylvania and
flows 1579 km to its confluence with the Mississippi River.
The river drains 508,202 km2, 48% of which is developed for
agricultural or urban land uses (6), and is divided into 21
pools by 20 dams designed to maintain a minimum water
depth of 4 m to facilitate barge traffic. This research was
conducted on the 153 km long Markland Pool which is bound
by the Markland Lock and Dam on the downstream end and
the Meldahl Lock and Dam on the upstream end (Figure 1).
At baseflow, the pool averages 0.4 km wide and depth ranges
from 6 m at the Meldahl Lock and Dam to 14 m at the
Markland Lock and Dam. Mean discharge during the study
period was 2371 m3 s-1. The city of Cincinnati is located near
the middle of the pool, and major tributaries draining into
the pool include the Little Miami River, Licking River, Mill
Creek, and Great Miami River.
This study consisted of four main sampling efforts, which
included 29 sampling sites defined by their distance downstream from the Meldhal Lock and Dam (Figure 1). The first
sampling effort consisted of measuring N2O emission rates,
dissolved N2O concentrations in the top 5 cm of the water
column, dissolved oxygen, and water chemistry biweekly from
August 2008 through September 2009 at the 23 km site. The
second sampling effort was conducted in August of 2008 and
included six sites forming a transect spanning the length of
the pool. In addition to the parameters described above, we
measured sediment N2O production rates and experimentally
determined which nutrients (e.g., ammonium (NH4+), nitrate
(NO3-), carbon) limited sediment N2O production at each
site. The third sampling effort entailed water chemistry
sampling at 18 sites distributed between km 42 and 104 in
mid August 2009. The final sampling campaign included
measurements of N2O production in the water column and
sediment, emissions of N2O from the water surface, and water
chemistry at 11 sites distributed between kilometers 5 and
79 (Figure 1). Table S1 contains a summary of the sampling
sites and measurements included in the four sampling
campaigns.
We measured N2O emission rates using 20 L floating acrylic
chambers tethered to a drifting boat. The headspace gas in
the chamber was sampled every 2-5 min for 12-30 min
though a flexible plastic tube. Gas samples were collected in
a plastic syringe and immediately transferred to preevacuated glass storage vials. Nitrous oxide emission rates
were computed from the linear regression of headspace N2O
partial pressure against time, after accounting for the
headspace volume and surface area of the river enclosed by
the chamber. During the biweekly and summer 2008 sampling
campaigns 4 chambers were deployed simultaneously,
whereas during summer 2009 two chambers were deployed
for each of two successive deployments. In total, we made
184 emission rate measurements. A full description of the
sampling procedure is included as Supporting Information.
We estimated sediment N2O production rates from the
change in dissolved N2O concentrations in the water overlying
sediment cores during a 6-h laboratory incubation. Sediment
N2O production rate nutrient limitation status was measured
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FIGURE 1. Markland Pool of the Ohio River showing major tributaries, urban centers, and sampling sites. Sampling sites are
identified by distance downstream from the Meldahl Lock and Dam. Black circles and diamonds represent sites sampled for N2O
during the summers of 2008 and 2009. White triangles represent sites sampled for nutrients (e.g., nitrate, ammonium) during the
summer of 2009. The 23 km site was sampled biweekly for 13 months (white square).
by incubating cores as described above and amending the
overlying water in five cores from each site with NO3-, NH4+,
or acetate. Dissolved N2O was sampled using the headspace
equilibration technique described in ref 7. A full description
of the sampling procedure is included as Supporting Information.
We measured water column N2O production rates by
incubating river water in 4 L flexible containers (Cubitaners
series 300; I-Chem; Rockwood, TN) in the laboratory at
ambient river water temperature for 48 h. Water samples
were collected at the beginning and end of the incubation
and analyzed for dissolved N2O concentration. A full description of the sampling and analytical procedure is included
as Supporting Information.
Results
Discharge at the biweekly sampling site (e.g., km 23) ranged
from 195 to 8076 m3 s-1 with the highest values during the
winter-spring (Figure 2A). Nutrient concentrations exhibited
little seasonal variation (Figure 2B, C) with average NO3and NH4+ concentrations of 0.82 ( 0.05 (SE) mg N L-1 and
50 ( 6 µg N L-1, respectively. These levels of N availability
in the river should provide ample substrate for the microbial
production of N2O. Dissolved organic carbon concentration
averaged 2.71 mg L-1, which is lower than has been reported
for lower portions of the Ohio River (see the Supporting
Information) (6, 8). Dissolved reactive phosphorus (DRP)
concentration was consistently low (mean )15 µg P L-1 (
1.8) (see the Supporting Information), and the high ratio of
inorganic N to DRP is indicative of the phosphorus limited
status of much of the Ohio River (9).
Unlike nutrient concentrations, dissolved N2O saturation
levels exhibited a strong seasonal pattern (Figure 3A). Nitrous
oxide saturation was high during the summer of 2008, fell
through the fall to near equilibrium values in winter, and
rose during spring returning to high levels of saturation during
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FIGURE 2. A) Discharge, B) nitrate (NO3-), and C) ammonium
(NH4+) at km 23 during the study.
FIGURE 3. A) Mean ((SE, n ) 3 per data point) degree of
nitrous oxide (N2O) saturation at km 23 expressed as the ratio
of the measured N2O concentration in the river (obs) and that
which is expected (exp) if the river were in equilibrium with
the atmosphere. Values above 1 indicate supersaturation and
values below 1 indicate undersaturation. B) Mean ((SE, n )
3-4 per data point) N2O emission rates at km 23.
the summer of 2009. Overall, 70% of the seasonal variation
in N2O saturation was explained by water temperature alone
(p < 0.001). Emission rates were positively related to N2O
saturation levels (p < 0.001, r2 ) 0.36) and exhibited a similar
seasonal pattern but tended to be more variable (Figure 3B)
which we attribute to differences in the air-water gas
exchange rate among sampling dates. Emission rates were
also positively related to water temperature (p < 0.001, r2 )
0.36).
The pool-wide survey conducted during the summer of
2008 showed the entire pool was supersaturated with N2O
(Figure 4A). Saturation levels (expressed as the ratio of
measured to equilibrium N2O concentrations) upstream of
Cincinnati were moderate (mean )1.6) but increased to 7.4
below the city’s WWTP and rapidly declined to ∼3.3 at sites
further downstream. Emission rates showed a similar pattern
with low rates upstream of the city (mean ) 16.3 µg N2O-N
m-2 h-1) and peak rates downstream of the WWTP (623 µg
N2O-N m-2 h-1) which rapidly declined at sites further
downstream (Figure 4B). The longitudinal N2O survey
conducted in 2009 was designed to provide greater spatial
resolution of N2O dynamics near Cincinnati. As we observed
in the 2008 survey, N2O emission rates were low upstream
of the city’s WWTP (mean ) 12.2 µg N2O-N m-2 h-1) and
peaked downstream of the WWTP (max ) 84.5 µg N2O-N
m-2 h-1, Figure 4B). However, the maximum emission rate
was a factor of 7 lower than in 2008 suggesting that N2O
emissions may exhibit significant intra-annual variation.
River-water NH4+ concentration reached 402 µg N L-1
below the WWTP outfall and was consistently higher than
upstream of the outfall (Figure 4C). Nitrate concentration
increased downstream from the WWTP in all three surveys
(Figure 4D).
We sampled the WWTP effluent at the top and bottom of
the 0.87 km long underground sewer pipe that conveyed the
effluent to river km 59 (8/24/2009 sampling date). The effluent
flow rate was 3.15 m3 s-1, and average NH4+ and
NO3-concentrations were 17.2 and 2.7 mg N L-1. The N2O
saturation ratio decreased from 135.6 ( 1.8 to 51.1 ( 1.3
FIGURE 4. River physicochemical parameters measured during
the three longitudinal surveys. The vertical dashed line
indicates the location of the WWTP outfall. Sampling sites are
identified by distance downstream from the top of the pool. A)
Mean ((SE, n ) 3 per data point) degree of nitrous oxide (N2O)
saturation expressed as the ratio of the measured N2O
concentration in the river (obs) and that which is expected
(exp) if the stream were in equilibrium with the atmosphere.
Values above 1 indicate supersaturation and values below 1
indicate undersaturation. B) Mean (((SE, n ) 3-4 per data
point) N2O emission rates. C) Ammonium (NH4+) concentration.
D) Nitrate (NO3-) concentration.
during transit through the pipe. Despite losing over half of
the dissolved N2O load during transport, the effluent
transported 110 g N2O-N h-1 to the river.
Results of the sediment core incubations ranged from a
slight net uptake of N2O to significant N2O production, but
on average all sites showed positive N2O production (Figure
5A). Sediment N2O production rates ranged from 0.2-15.8
µg N2O-N m-2 h-1 and were positively related to sediment
organic matter (SOM) content (p ) 0.004, r2 ) 0.20) which
ranged from 2.4-7.7% (Table S4). The nutrient limitation
assays demonstrated that sediment N2O production was
NO3- limited at all but one site (p < 0.05, Table 1) and not
limited by C or NH4+ at any site (Table 1).
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FIGURE 5. Mean ((SE) nitrous oxide (N2O) production rates
measured in (A) the sediments (n ) 5 per data point) and (B)
water column (n ) 3 per data point) of the river. Sediment N2O
production was measured in 2008 and 2009, while water
column production was only measured in 2009. Sampling sites
are identified by distance downstream from the top of the pool,
and the vertical dashed line indicates the location of the
WWTP outfall.
TABLE 1. Sediment Nitrous Oxide (N2O) Production Rates
(Mean ± SE) in the Ohio River under Ambient Levels of
Nutrient Availability and When Amended with Nitrate,
Ammonium, or Acetate
sediment N2O production rates (µg N2O-N m-2 h-1)
site (km)
23
64
94
124
144
ambientb
+nitratec
3.6 ( 1.6
6.1 ( 3.9
3.7 ( 2.0
15.8 ( 8.8
7.4 ( 1.8
70.0 ( 27.9f
190.4 ( 110.0
113.4 ( 40.3f
152.5 ( 61.7f
274.4 ( 74.6f
a
+ammoniumd +acetatee
6.1 ( 1.4
7.9 ( 3.3
3.7 ( 1.4
4.3 ( 1.6
13.1 ( 5.8
2.1 ( 2.7
4.3 ( 2.2
2.5 ( 1.3
3.4 ( 1.4
9.1 ( 3.2
a
Sites are identified by distance downstream from the
Meldahl Lock and Dam. b Sediment cores were incubated
with ambient river water. c Sediment cores were incubated
with river water amended with nitrate (potassium nitrate to
10 mg N L-1 above background). d Sediment cores were
incubated with river water amended with ammonium
(ammonium chloride to 1 mg N L-1 above background).
e
Sediment cores were incubated with river water amended
with acetate (to 25 mg C L-1 above background). f Ambient
and amended means significantly different at P < 0.05.
We found that the water column was a net source of N2O
at 9 of the 11 sites sampled (Figure 5B). Production rates
expressed per unit volume of water ranged from -0.0004 0.021 µg N2O-N L-1 h-1 and from -5 - 244 µg N2O-N m-2
h-1 on an areal basis. The three highest rates co-occurred
with high NH4+ concentration (Figure 4C) and N2O emission
rates (Figure 4B) downstream of the WWTP outfall. Nitrous
oxide production in the WWTP effluent (1.01 µg N2O-N L-1)
exceeded the highest rate observed in the river by a factor
of 49.
Discussion
The biweekly sampling showed that the river is a seasonally
variable net source of N2O to the atmosphere (Figure 3A, B)
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with the highest dissolved N2O saturation levels and emission
rate occurring during the summers and lowest during the
winter. Overall, 70 and 36% of the variation in N2O saturation
and emission rates could be explained by temperature alone.
The most likely explanation for this pattern is that microbial
N2O production rates in the river are partially controlled by
water temperature. However, our results contrast with Stow
et al. (10) who reported that N2O emission rates from 11
streams in North Carolina showed no seasonal pattern, while
Beaulieu et al. (11) reported the highest emission rates from
12 small streams in Michigan occurred during the winter
and the lowest during the summer. These seemingly contradictory results may be resolved by considering the relative
roles of temperature and nutrient availability in controlling
N2O production rates. Laboratory experiments have shown
that sediment denitrification, an important source of N2O
(3, 12), responds positively to temperature (up to 30 °C) and
NO3- availability when these factors are manipulated independently (12-14). In the field, however, these factors may
covary, obscuring the effects of either variable on N2O
production. In the current work NH4+ and NO3- remained
relatively constant throughout the study (Figure 2B, C), while
temperature followed a predictable seasonal cycle, which
suggests that seasonal variation in N2O emissions resulted
from changes in water temperature. This pattern contrasts
with the small stream work in Michigan (12) where NO3concentrations ranged across 4 orders of magnitude and
reached maximum levels during the winter when water
temperature was at a minimum, a common pattern in tile
drained agricultural landscapes (15). In this case, seasonal
variability in NO3- availability overwhelmed temperature
controls on microbial metabolism with the net result that
the highest emission rates occurred during the winter. Cole
and Caraco (16) reported that N2O emission rates from the
Hudson River exhibited a temporal pattern intermediate to
the two cases just discussed. Maximum emission rates were
observed during late summer when NO3- concentrations
were low and temperature was high, suggesting that high
temperature may have been important in stimulating N2O
production in late summer. However, a smaller peak in
emission rates was observed during midwinter when NO3was high and temperature was low. This latter pattern
compares favorably with Beaulieu et al. (12) where increased
NO3- availability during the winter stimulated N2O production, despite low temperatures. These examples illustrate
the importance of recognizing that covariance among NO3and temperature can result in complex temporal patterns in
N2O emissions.
The Markland Pool was a source of N2O throughout its
entire length during the summers of 2008 and 2009 (Figure
4A, B), though emission rates were low upstream of Cincinnati
and increased by factors of 8-200 immediately below the
city. The emission rates fall within the range of values reported
for several other nitrogen rich lotic systems including the
Platte River (17), Assabet River (18), and several small
agricultural rivers in the Midwestern USA [ (19), see ref 11
for a summary of published N2O emission rates]. Emissions
in this study were likely supported by several sources of N2O
including microbial metabolism (e.g., nitrification and denitrification) in the sediments or water column and possibly
the direct injection of N2O into the river from WWTP effluent
or tributaries. Sediment collected from the Markland Pool
was a net source of N2O at all sampling sites (Figure 5A).
Nitrous oxide production rates ranged from 0.2-15.8 µg
N2O-N m-2 h-1 which compares well to reports from a small
agricultural river in Indiana (4.2 µg N2O-N m-2 h-1 (20)) and
two sites in the NO3- rich Potomac River (15.4 and 140 µg
N2O-N m-2 h-1 (21)). Higher and more variable rates were
reported for small agricultural streams in Michigan (0.35-3236
µg N2O-N m-2 h-1 (12)) and the Wiske River in England
(-175-11,000 µg N2O-N m-2 h-1 (13)), but these sites
spanned a much broader range of NO3- concentrations
(0.003-27.4 and 2.3-31.9 mg N L-1, respectively) than
encountered in this work. Overall, sediment N2O production
rates in the Markland pool were too low to support the
observed emission of N2O from the water surface. Sediment
N2O production rates averaged only 14% of N2O emissions
at the sites sampled during the two longitudinal surveys,
which is surprisingly low and may reflect an artifact of the
incubation method we used to measure N2O production rates.
Enclosure-based methods have been shown to reduce the
delivery of water column solutes (e.g., NH4+, NO3-) to
sediments resulting in artificially low biogeochemical transformation rates. For example, Risgaard-Peterson et al. (22)
found that in situ wave forces can transform unconsolidated
sediments into a semifluid state resulting in accelerated solute
transport and greatly enhanced denitrification rates. While
our static core incubations may have supported lower rates
of advective solute transport than occurred in situ, we
nonetheless observed greatly enhanced sediment N2O production rates when the water overlying the cores was
amended with NO3-. This demonstrates solute transport had
occurred but does not verify that this process occurred at in
situ rates. To partially address this concern, we stirred the
water overlying the sediment for the 2009 incubations. We
subsequently observed higher sediment N2O production rates
in 2009 than 2008 (mean of 7.3 vs 2.8 µg N2O-N m-2 h-1),
though this difference was not statistically significantly (p )
0.08). Without an in situ sediment N2O production rate
measurement, it is impossible to determine if our incubation
method resulted in artificially low production rates, but the
weight of the evidence discussed above suggests that this is
not the case.
We found that sediment N2O production rates were not
related to NO3- concentration at the sampling site which is
inconsistent with reports from small streams (12), lakes (23),
and estuaries (24, 25). This lack of a relationship between
NO3- and sediment N2O production is likely due to the narrow
range of NO3- concentrations encountered in this work
(0.48-0.84 mg N L-1), relative to the other cited studies (0.1720.59, <0.005-0.63, and 0.14-0.80 mg N L-1, respectively);
however, sediment N2O production rates increased by a factor
of 25 when amended with NO3-. This suggests that NO3partially controls sediment N2O production in the Ohio River,
but the range in NO3- concentration among sites was not
great enough to explain the observed variation in production
rates. Sediment N2O production rates were, however, positively related to sediment organic matter content, possibly
because sediment organic matter was used as an energy
source by heterotrophic N2O producing bacteria (14) which
are ubiquitous in aquatic and terrestrial systems.
If we assume that our sediment cores are representative
of sediments throughout the river, then sediment N2O
production is estimated to account for 11% of the observed
N2O emissions from the pool during the late summer.
However, all of our sediment cores were collected from within
50 m of the shoreline. Outside this limited sampling area the
benthos consist of hard packed clays and gravels. We suspect
these unsampled sediments, which compose ∼75% of the
benthos in the pool, support low rates of N2O production
due to a combination of low hydraulic permeability (e.g.,
low rates of solute transport into sediments) and low organic
matter content. Overall, the scaling approach used in this
budget may actually lead us to overestimate the importance
of sediment N2O production. While additional research is
required to refine our estimate of sediment N2O production
rates at the scale of the Markland Pool, it is clear that sediment
N2O production alone cannot support the observed emission
of N2O from the river.
We found that the water column was a net source of N2O
at most of the sites sampled (Figure 5B). Since dissolved
oxygen was near saturation (mean ) 99% saturation, Table
S1) during the sampling, nitrification rather than denitrification is the most likely source of the water column N2O
production. Nitrification, an aerobic microbial transformation in which NH4+ is oxidized to NO3- and N2O, has been
shown to occur in sediments (26) and the water column (27)
(e.g., pelagic nitrification) of rivers. The relative importance
of pelagic versus sediment nitrification is determined by a
number of factors including 1) water residence time, 2) the
concentration of suspended sediments in the water column
which provide substrate for nitrifying bacteria to adhere to,
and 3) the ratio of benthic surface area to water volume.
Sediment nitrification predominates in small streams which
are characterized by short water residence times, low
suspended solids concentration, and high benthic surface to
water volume ratios. By contrast, pelagic nitrification can
exceed sediment-based nitrification in estuaries where high
suspended particle concentration and long residence times
associated with turbidity maximum zones promote high rates
of nitrification (28, 29). While several studies have shown
that pelagic nitrification can be the predominate source of
N2O production in estuaries (29, 30), few studies have
investigated the importance of pelagic N2O production in
rivers. Investigations of the Wiske river in England (31), the
Assabet River in Massachusetts (18), and the Tama River in
Tokyo have shown that pelagic N2O production rates are
undetectable or low in these systems. However, the discharge
of these rivers is less than 1% that of the Ohio River, and
pelagic N2O production would not be expected to be
important in these relatively small systems. By contrast, the
Seine River in France has a mean annual flow ∼20% that of
the Ohio River and supports substantial pelagic nitrification
(14 µg N L-1 h-1) (27, 32) which may be an important source
of N2O in the river (33). According to our findings pelagic
N2O production is the source of at least 26% of the N2O
emissions from the Ohio River during the summer. These
results suggest that pelagic N2O production may be underappreciated in large rivers and merits further research.
The WWTP serving the city of Cincinnati played an
important role in the N2O budget of the pool. The NH4+
concentration in the WWTP effluent was extremely high (14.4
mg N L-1) and likely accounted for the consistently elevated
NH4+ concentration in the river reach below the outfall (Figure
4C). The elevated NH4+ in this reach may have stimulated
pelagic nitrification accounting for the high rates of pelagic
N2O production observed below the outfall. It is well-known
that NH4+ availability can limit sediment nitrification rates
(34) and reports from the Upper Mississippi River have shown
that sediment nitrification rates are correlated with sediment
NH4+ concentration (26). While little work has been done to
identify the controls on pelagic nitrification, it is reasonable
to assume that NH4+ limitation could be an important factor
and that poorly nitrified WWTP effluent could stimulate
pelagic N2O production in large rivers.
It is possible that nitrifying bacteria discharged to the
river in the WWTP effluent contributed to the high rates of
pelagic N2O production observed below the outfall. We found
the effluent supported rates of N2O production 49 times
greater than the highest pelagic N2O production rate observed
in the river, demonstrating that N2O producing bacteria are
present in the effluent. This assertion is supported by work
in the Seine River in France where it was shown that WWTP
effluent seeded the river with nitrifying bacteria, which
subsequently contributed to high rates of pelagic nitrification
(27, 35). In a laboratory study on the same topic Bonnet et
al. (36) demonstrated that nitrifying bacteria in WWTP
effluent can survive the transition from effluent to a riverine
environment and Montuellee et al. (37) reported a larger and
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more active population of sediment-associated nitrifiers
downstream than upstream of a WWTP. While we did not
measure the size of the nitrifying community in this study,
it stands to reason that bacteria derived from the WWTP
may have played an important role in producing N2O in the
water column of the Ohio River.
We consistently observed the highest N2O emission rates
in the 12 km reach downstream from the WWTP. The elevated
emissions in this reach were derived not only from greater
in situ N2O production (see above) but also through subsidies
of dissolved N2O imported to the river in the WWTP effluent
which was supersaturated by a factor of 51 at the outfall.
This equates to an N2O flux to the river of 110 g N2O-N h-1,
equivalent to ∼5% of the N2O emissions from the pool during
summer. This is not the first study to report that WWTP
effluent can directly subsidize riverine N2O emissions.
Hemond and Duran (18) found the largest source of N2O to
a 370 m reach of a small river was WWTP effluent, and Toyoda
et al. (38) used isotopomers to demonstrate that a fraction
of the N2O dissolved in the Tama River was derived from
WWTP effluent. It is interesting to note that although the
WWTP directly subsidized the N2O load in the Ohio River,
this direct effect was outweighed by the indirect effect of
stimulating in situ production (see above).
Sediment N2O production, pelagic N2O production, and
direct input from WWTP effluent account for 11, 26, and 5%
of the summertime N2O emissions (50 kg N2O-N d-1),
respectively. Of these terms, pelagic and sediment N2O
production have the greatest uncertainty due to the high
degree of spatial variability. Since no pelagic N2O production
measurements were made downstream of km 79, we assumed
the water column was not a source of N2O below this site,
and our estimate of pelagic N2O production is therefore
conservative. Other potential sources of N2O to the river
include tributaries and groundwater. It is unlikely that
tributaries contributed a large amount of N2O to the river
since their combined flow was only ∼10% that of the Ohio
River, and rapid air-water gas exchange in small rivers
prevents dissolved N2O from building up to high levels (39).
Groundwater contaminated with nitrogen can contain high
levels of dissolved N2O (40, 41) and can be an important
source of N2O in aquatic systems (42). Unfortunately, no
data on groundwater flow rates or dissolved N2O concentration exist for the Markland Pool of the Ohio River.
This research has shown that the urbanized Markland
Pool of the Ohio River is a source of N2O throughout its
entire length and that emission rates exhibit strong seasonal
variation. Seasonal patterns in N2O emissions have been
examined in smaller lotic systems (10, 11, 16), but this is the
first study to examine a very large river (e.g., flow >400 m3
s-1) where the effect of water temperature on microbial
metabolism is more likely to outweigh that of nitrogen
availability (see above). Our finding suggests that N2O
emission models and inventories that do not account for
seasonally variable N2O production in large temperate rivers
may yield biased results.
We found that the water column is an important site of
N2O production. Pelagic N2O production may be an underappreciated process in large rivers where long water residence
times, high turbidity, and low ratios of benthic surface area
to water volume set the stage for pelagic N2O production.
This N2O production can be further stimulated by bacterial
and NH4+ subsidies derived from WWTP effluents which are
a common feature along large rivers across the globe. This
scenario contrasts with smaller lotic systems where the N
transformations occur almost exclusively in the benthos. This
suggests N2O production in large rivers may not be well
predicted by models that are based on mechanisms and
parameter sets drawn from smaller lotic systems. However,
few data on N cycling in large rivers are available, and
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modelers are forced to use data from small streams (43-45).
Our findings demonstrate that N2O production in large rivers
is controlled by different factors than in smaller lotic systems,
and N2O emission models should account for these
differences.
Acknowledgments
We thank Chris Nietch (US EPA), Susan Crookall (USDAARS), and Pegasus Technical Services Inc. for analytical
support. We also thank Chris Lorentz (Thomas More College,
Biology Department) for field support, the Ohio River Valley
Water Sanitation Commission for helpful discussions and
spatial data layers, the US CORP for hydrology data, and the
Municipal Sewer District of Cincinnati for access to the
wastewater treatment facility.
Supporting Information Available
Additional information on sample collection, analytical
methods, statistics, and ancillary data are available. This
material is available free of charge via the Internet at http://
pubs.acs.org.
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