Nitrous Oxide Emissions from a Large, Impounded River: The Ohio River J . J . B E A U L I E U , * W . D . S H U S T E R , A N D J . A . R E B H O L Z National Risk Management Research Laboratory, Office of Research and Development, U.S. Environmental Protection Agency, 26 West Martin Luther King Drive, Cincinnati, Ohio 45268 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 measure- ments 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 the pool during two successive summer surveys. We quanti- fied several sources of N2O to the river including wastewater treatment plant (WWTP) effluent and microbial N2O pro- ...

1. Nitrous Oxide Emissions from a
Large, Impounded River: The Ohio
River
J . J . B E A U L I E U , * W . D . S H U S T E R , A N D
J . A . R E B H O L Z
National Risk Management Research Laboratory, Office of
Research and Development, U.S. Environmental Protection
Agency, 26 West Martin Luther King Drive, Cincinnati,
Ohio 45268
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

2. 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 measure-
ments 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
the pool during two successive summer surveys. We quanti-
fied several sources of N2O to the river including wastewater
treatment plant (WWTP) effluent and microbial N2O pro-
duction (e.g., nitrification and denitrification) in the water
column and sediments.
Experimental Section

3. 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 down-
stream 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

4. 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 pre-
evacuated 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*
Corresponding author e-mail: [email protected]
Environ. Sci. Technol. 2010, 44, 7527–7533
10.1021/es1016735 Not subject to U.S. Copyright. Publ. 2010
Am. Chem. Soc. VOL. 44, NO. 19, 2010 / ENVIRONMENTAL
SCIENCE & TECHNOLOGY 9 7527
Published on Web 08/30/2010
by incubating cores as described above and amending the

5. 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 Infor-
mation.
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 de-
scription 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 NO3-
and 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

6. through the fall to near equilibrium values in winter, and
rose during spring returning to high levels of saturation during
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).
FIGURE 2. A) Discharge, B) nitrate (NO3-), and C) ammonium
(NH4+) at km 23 during the study.
7528 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY /
VOL. 44, NO. 19, 2010
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

7. 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
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

8. 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).
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.
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.
VOL. 44, NO. 19, 2010 / ENVIRONMENTAL SCIENCE &
TECHNOLOGY 9 7529
We found that the water column was a net source of N2O

9. 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)
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 con-
tradictory 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 inde-
pendently (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

10. 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 NO3-
concentrations 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 NO3-
was high and temperature was low. This latter pattern
compares favorably with Beaulieu et al. (12) where increased
NO3- availability during the winter stimulated N2O produc-
tion, despite low temperatures. These examples illustrate
the importance of recognizing that covariance among NO3-
and 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

11. 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 deni-
trification) 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
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)a ambientb +nitratec +ammoniumd +acetatee
23 3.6 ( 1.6 70.0 ( 27.9f 6.1 ( 1.4 2.1 ( 2.7

12. 64 6.1 ( 3.9 190.4 ( 110.0 7.9 ( 3.3 4.3 ( 2.2
94 3.7 ( 2.0 113.4 ( 40.3f 3.7 ( 1.4 2.5 ( 1.3
124 15.8 ( 8.8 152.5 ( 61.7f 4.3 ( 1.6 3.4 ( 1.4
144 7.4 ( 1.8 274.4 ( 74.6f 13.1 ( 5.8 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.
7530 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY /
VOL. 44, NO. 19, 2010
(-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 trans-
formation rates. For example, Risgaard-Peterson et al. (22)
found that in situ wave forces can transform unconsolidated

13. 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 pro-
duction 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 (0.17-
20.59, <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 NO3-
partially 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, posi-
tively related to sediment organic matter content, possibly
because sediment organic matter was used as an energy

14. 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 denitrifi-
cation is the most likely source of the water column N2O
production. Nitrification, an aerobic microbial transforma-
tion 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.

15. 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 under-
appreciated 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

16. 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
VOL. 44, NO. 19, 2010 / ENVIRONMENTAL SCIENCE &
TECHNOLOGY 9 7531
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.

17. 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

18. source of N2O in aquatic systems (42). Unfortunately, no
data on groundwater flow rates or dissolved N2O concentra-
tion 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 under-
appreciated 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
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

19. differences.
Acknowledgments
We thank Chris Nietch (US EPA), Susan Crookall (USDA-
ARS), 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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ES1016735
VOL. 44, NO. 19, 2010 / ENVIRONMENTAL SCIENCE &
TECHNOLOGY 9 7533
cis363wk6iLab.jpg
e.jpg

27. library.js
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XYear: The 4-digit year value of the given date
XDay: December 31 in the year of the given date
DayCount: The number of days between New Year's Eve and
the given date
*/
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var XYear=CheckDay.getFullYear();
var XDay=new Date("December, 31, 2014");
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var DayCount=(XDay-CheckDay)/(1000*60*60*24);
DayCount=Math.round(DayCount);

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Anaerobic Transformation Kinetics and Mechanism of Steroid
Estrogenic Hormones in Dairy Lagoon Water
Wei Zheng,*,† Xiaolin Li,† Scott R. Yates,‡ and Scott A.
Bradford‡
†Illinois Sustainable Technology Center, Prairie Research
Institute, University of Illinois at Urbana−Champaign, 1
Hazelwood Drive,
Champaign, Illinois 61820, United States
‡Contaminant Fate and Transport Unit, Salinity Laboratory,
USDA-ARS, Riverside, 450 West Big Springs Road, California
92507,
United States
*S Supporting Information
ABSTRACT: Wastewater from concentrated animal feeding
operations
(CAFOs) frequently contains high concentrations of steroid
estrogenic
hormones. Release of these hormones into the environment may

31. occur when
CAFO wastewater is applied to agricultural lands as a nutrient
and water
source for crop production. To assess the potential risk of
hormone
contaminants derived from animal wastewater, we investigated
the trans-
formation kinetics and mechanisms of three natural estrogenic
hormones
(17α-estradiol, 17β-estradiol, and estrone) in aqueous solutions
blended with
dairy lagoon water under anaerobic conditions. Initial
transformations of the
three hormones in the dairy lagoon water were dominated by
biodegradation
and the degradation rates were temperature-dependent. The total
amounts of
hormones (initial concentration at 5 mg L−1) remaining in the
solution after
52 days at 35 °C accounted for approximately 85%, 78%, and
77% of the initial
amounts of 17α-estradiol, 17β-estradiol, and estrone,
respectively. This observation suggests that these hormones are
relatively
stable over time and may accumulate in anaerobic or anoxic
environments and anaerobic CAFO lagoons. A racemization
reaction
between 17α-estradiol and 17β-estradiol via estrone was
observed in aqueous solutions in the presence of CAFO
wastewater
under anaerobic conditions. The initial hormone concentrations
did not affect this degradation mechanism. A reversible reaction
kinetic model was applied to fit the observed transformation
dynamics. The degradation and regeneration of the parent
hormone
and its metabolites were successfully simulated by this model.

32. The information in this study is useful for assessing the
environmental risk of steroid hormones released from CAFO
wastewater and to better understand why these hormone
contaminants persist in many aquatic environments.
■ INTRODUCTION
As the scarcity of water supplies grows, wastewater reclamation
offers an essential and viable water management solution.1 It is
commonplace to reuse treated municipal wastewater for
landscape irrigation. The reuse of concentrated animal feeding
operation (CAFO) wastewater on agricultural fields is another
feasible and economic water recycling strategy,2 which can
provide nutrients and organic matter for plant growth and offer
an alternative water source to reduce the demand for high
quality water. However, water derived from CAFOs usually
retains many contaminants such as excess amounts of
nutrients,3,4 salinity,5−7 pathogens,7,8 heavy metals,9,10 and
organics (e.g., animal hormones and veterinary pharmaceut-
icals),11−15 posing a potential risk both to the receiving
ecosystems and drinking water resources.
Current management and regulations for CAFO wastewater
reuse on agricultural lands are primarily focused on nutrients.2
Emerging contaminants such as veterinary antibiotics and
hormones in CAFO wastewater are currently unregulated and
their potential adverse impacts on environmental resources and
public health are poorly understood.2,7 Unlike municipal
wastewater, which is treated to remove most contaminants,
CAFO wastewater receives no additional treatment before land
application. Therefore, CAFOs are attracting extensive
attention as an important source for the release of these
emerging contaminants into the environment.13−16
Dairy farms are one of the most important CAFOs. It has

33. been estimated that dairy cattle contributes 45 t of endogenous
hormones annually, accounting for approximately 90% of
estrogen excretion by livestock in the United States.17
Hormones excreted from livestock species occur as free
steroids, or sulfate or glucuronide conjugates. Free steroid
estrogenic hormones are classified as highly potent endocrine-
disrupting chemicals (EDCs), whereas most conjugated forms
of these hormones are biologically inactive. In vitro studies
have
shown that low concentrations of steroid estrogenic hormones,
even at levels as low as ng L−1, can adversely affect the
reproductive biology of aquatic wildlife.18−20 High concen-
Received: April 19, 2012
Accepted: April 20, 2012
Published: April 21, 2012
Article
pubs.acs.org/est
© 2012 American Chemical Society 5471
dx.doi.org/10.1021/es301551h | Environ. Sci. Technol. 2012,
46, 5471−5478
pubs.acs.org/est
trations of three natural estrogenic hormones, 17α-estradiol,
17β-estradiol, and estrone, are often detected in dairy manure
and manure-containing wastewater.2,12,14−16 Although all
naturally occurring hormones are readily degradable in waste
disposal systems,15 significant quantities of undegraded
hormones and their metabolites are introduced to soils from
land application of livestock wastes21 or from cattle feedlots by

34. runoff.22
Sorption and degradation are two primary processes affecting
the fate and transport of steroid hormones in the environment
once they enter the soil. The sorption of hormones has been
extensively studied in a variety of soils and sediments.23−28 It
has been reported that the sorption potential of steroid
hormones are moderate to high depending on the fraction of
organic matter content.24,28 Batch equilibration experiments
have shown that equilibrium times range from several hours to
several days.27,29 Rapid biodegradation of hormones in soils
affects their sorption processes, which may account for the
highly variable equilibrium times. Most hormones exhibit
relatively short half-lives in a variety of environmental media
including agricultural soils,24,30,31 sewage sludge,32,33 and
sediments.34 For example, the half-lives of 17α-estradiol, 17β-
estradiol, and estrone in activated sludge and agricultural soils
are often less than one day.24,31,33 The degradation rates are
also related to the hormone compounds. In a silt loam soil,
estrone was found to be relatively recalcitrant to degradation
compared to the two estradiol isomers.30 In addition, the
degradation rates of hormones are much higher under aerobic
conditions compared to anaerobic incubation.33−35 Previous
studies have shown that both estradiol hormones can be readily
oxidized to estrone under aerobic conditions and that estrone
can be further degraded in the environment.24,31,32 Thus,
aerobic biodegradation is the major removal mechanism for
these steroid hormones in the environment.
The release of steroid hormones into agricultural soils is
usually associated with manure, manure-contaminated waste-
water, or other biosolids. Addition of organic materials into soil
has a significant impact on the soil biota, for example,
increasing the soil organic carbon content, altering the soil
microbial community structure, and improving microbial
activity. Accordingly, it may also affect the transformation and

35. transport processes of hormones in agricultural soils. Previous
studies reported that animal manure amendment increased the
sorption and biodegradation of 17β-estradiol and estrone in
soils.36,37 Given the rapid degradation and high affinity in
manure amended soils, these estrogenic hormones were not
expected to be released to surrounding water bodies. However,
these hormones are frequently detected in tile drainage and
ditch water from the agricultural fields receiving livestock
wastes,21 which indicates that their persistence in the
environment may be underestimated. The objective of this
study is to investigate the mechanisms and kinetics of anaerobic
transformation for three steroid estrogenic hormones in dairy
lagoon water. The study provides a better understanding of the
persistence of hormones in the environment and will help to
address some of the knowledge gaps about the degradation of
estrogenic hormones in CAFO wastewater.
■ MATERIALS AND METHODS
Chemicals and Dairy Lagoon Water. The hormones
17α-estradiol, 17β-estradiol, and estrone were purchased from
the Sigma-Aldrich Chemicals (St. Louis, MO) at the highest
possible purity (>98%). Stock solutions of these estrogenic
hormones (1.0 mg mL−1) were prepared in methanol.
Deionized water was supplied by a Barnstead E-pure
purification system (Dubuque, IA). Other reagent chemicals
were obtained from Fisher Scientific (Fair Lawn, NJ). All
chemical reagents were used as received. All glassware was
autoclaved prior to use.
The dairy lagoon water was collected from a dairy farm
located in San Jacinto, California (CA). The farm has three
large anaerobic lagoons, which are sequentially connected to
store manure-contaminated water from stormwater runoff and
wastewater derived from the milking parlor. The lagoon water

36. used in this study was collected from the outlet of the tertiary
lagoon at a depth of about 15 cm below the surface using a
stainless steel bucket. The collected water samples were stored
in 4 L solvent bottles and immediately transported to the
laboratory in an ice cooler. Detailed information concerning the
lagoon water can be found in our previous research.15 The
oxidation−reduction potential of dairy lagoon water is reported
to be −277 ± 24 mV.2 Laboratory analysis revealed that the
total concentration of steroid hormones in this tertiary lagoon
water was less than 5 ng L−1.15 The lagoon water to be used
for
anaerobic incubations was passed through a 2.0 μm filter to
remove visible particles, flushed with nitrogen gas for 1 h, and
sealed tightly. The dairy lagoon water was stored at 4 °C
overnight, then thawed gradually to room temperature and
used within 24 h.
Experimental Systems. To investigate the transformation
processes of hormones in CAFO lagoon water, kinetic
experiments were conducted in amber glass bottles with
Teflon-lined screw caps under anaerobic conditions. Generally,
lagoon water needs to be diluted using surface water and
groundwater prior to its irrigation.2 Similar to lagoon water
application practice, the experimental solutions were prepared
by thoroughly mixing the dairy lagoon water and deionized
water (1:1 by volume) within an anaerobic glovebag. The
aqueous solutions blended with lagoon water were purged with
nitrogen gas for 30 min prior to use.
All solutions were preconditioned at a selected temperature
for 1 day and were then spiked with stock hormone solutions,
yielding an initial hormone concentration of 5 mg L−1. Similar
hormone concentration has also been chosen in previous
studies.28,38,39 All solution bottles were vigorously shaken
and
then incubated at 15, 25, 35, and 45 °C in the dark. At regular

37. time intervals, aliquots of incubation solutions were withdrawn
from each bottle and immediately transferred into a centrifuge
tube containing an equal volume of methanol within the
anaerobic glovebag. The samples were vortexed for 5 min at
room temperature for extraction. The tubes were centrifuged at
4000 rpm for 10 min and then filtered through a 0.45 μm
membrane (Iso-Disc, PTFE, Supelco, Bellefonte, PA) using a
syringe. All filtrate samples were stored in a freezer (−21 °C)
until analysis. Preliminary experiments revealed that the
addition of methanol could immediately quench hormone
transformation and effectively extract hormones that were
sorbed to suspended manure in the lagoon water. The
recoveries of the three estrogenic hormones ranged from 95
to 105% in aqueous solutions blended with 50% dairy lagoon
water.
A control experiment was concurrently performed using a
sterile lagoon water sample to determine abiotic degradation
and thereby deduce the effect of biodegradation in nonsterile
samples. Briefly, the solutions blended with lagoon water
(lagoon water: deionized water, 1:1 by volume) were
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46, 5471−54785472
autoclaved twice at 120 °C, each for 60 min within a 24 h
interval. The same procedures described above were performed
including solution preparation, incubation, sampling, extraction,
and analysis of steroid hormones. All experiments were carried
out in triplicate.
Analytical Methods. To determine the transformation

38. kinetics of three estrogenic hormones, filtrate samples were
analyzed using an Agilent 1100 series high performance liquid
chromatography (HPLC) with a diode array detector (DAD)
(Agilent Technologies, Palo Alto, CA). Separation for HPLC/
DAD analysis was performed using an Eclipse C18 column
(250 × 4.6 mm i.d., particle size 5 μm). The mobile phase
consisted of acetonitrile/water (50:50, v/v), the flow rate was
1.0 mL min−1, and the detector wavelength was 205 nm. Under
these conditions, the retention times for 17β-estradiol, 17α-
estradiol, and estrone were 6.5, 7.7, and 9.2 min, respectively.
The detection limits of the method were 0.05, 0.05, and 0.08
mg L−1 for 17β-estradiol, 17α-estradiol, and estrone,
respectively.
The biodegradation products of hormones were analyzed
using an Agilent 1100 series HPLC/DAD in tandem with a
mass spectrometer (MS) equipped with an electrospray
ionization (ESI) source. The biodegradation products of
hormones were identified using the HPLC/DAD method
described above by matching retention times to their
corresponding reference standards, and then quantified by
external calibration. The identities of biodegradation products
were further confirmed by LC/MS. LC/MS total ion current
(TIC) chromatograms were recorded between m/z 100 and
500 at a rate of 2 scans per second. The negative polarity
ionization mode was operated to obtain mass spectra for the
identification of transformation products of estrogenic
hormones. The electrospray source parameters were optimized
by infusion of hormone standard solutions. The operating
conditions for ESI were capillary voltage 4000 V for positive
mode, drying gas (nitrogen) flow rate 10 L min−1 at 300 °C,
and nebulizer gas pressure 45 psi.
Transformation of Hormones at Low Concentration.
Additional experiments were conducted to validate the
transformation of hormones at environmentally−relevant

39. concentrations. This study employed dairy lagoon water from
Champaign, Illinois (IL) and incubation solutions at an initial
hormone concentration of 5 μg L−1. Details for these
experiments are described in the Supporting Information
(SI). Specific chemical properties of this dairy wastewater are
provided in the Table S1 of the SI. New analytic procedures
had to be developed to determine the low hormone
concentrations in this study. Details of this analytical method
are summarized in Table S2 and Figure S1 of the SI.
■ RESULTS AND DISCUSSION
Initial Degradation of Three Estrogenic Hormones.
The first transformation study focused on the initial
degradation of the three steroid hormones (i.e., within 24 h
for 17α-estradiol and 17β-estradiol, and 72 h for estrone) in
aqueous solutions blended with dairy lagoon water. Time
courses for degradation of each estrogenic hormone (initial
concentration at 5 mg L−1) at different temperatures are shown
as semilogarithmic plots in Figure 1. No discernible hormone
degradation occurred in control experiments conducted in
sterile solutions mixed with the lagoon water over comparable
time periods (Figure 1). The result indicates that the initial
transformations of the three hormones in blended CAFO
wastewater were dominated by biological degradation and that
abiotic degradation such as hydrolysis was negligible.
For the initial degradation of three hormones at different
temperatures, the biodegradation rate can be represented by a
pseudo-first-order kinetic model:
= −
t
k

40. d[C]
d
[C]i (1)
Upon rearrangement and integration, eq 1 becomes
Figure 1. Time courses for initial degradation of three
estrogenic
hormones in aqueous solutions blended with dairy lagoon water
at
different temperatures under anaerobic conditions. The initial
concentration of each investigated steroid hormone was 5 mg
L−1.
Error bars represent standard deviations of triplicate samples.
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46, 5471−54785473
= − +k tln([C]) ln([C] )i 0 (2)
where ki is the initial biodegradation rate constant of the
investigated hormone under a certain temperature, [C] is the
concentration of the hormone, and [C]0 is the initial
concentration of the hormone in incubation solutions. Values
of ki were calculated from the slope of semilogarithmic plots of
hormone concentration versus time. These plots fit a log−linear
model well for the degradation of all three hormones during the
initial experimental time period (Figure 1). Details pertaining
to the initial biodegradation rate constants of three hormones
under different incubation temperatures are summarized in
Table 1.
The effect of temperature on the initial biodegradation rates

41. of the three hormones in blended CAFO wastewater is shown
in Figure 2. For all hormones, the biodegradation rate increased
with increasing temperature from 15 to 35 °C, but then
decreased when the incubation temperature was elevated to 45
°C. It appears that the most suitable temperature for hormone
biodegradation in this study is around 35 °C, which is close to
the well-known optimum temperature (∼37 °C) for the growth
of active microorganisms in various environmental media. Yang
et al.35 reported that the testosterone degradation at 37 °C was
much faster than that at room temperature (i.e., 22 °C),
because many fecal-derived enzymes have optimal activity at
physiological temperature. A similar effect of temperature was
observed for the biodegradation kinetics of ceftiofur in CAFO
wastewater.40
Statistically significant differences in ki were observed for the
three hormones investigated for each incubation temperature.
The initial biodegradation rate of the three hormones decreases
in the following order: 17β-estradiol >17α-estradiol > estrone
(Figure 2 and Table 1). This observation is consistent with a
previous study, in which the degradation of these estrogenic
hormones was conducted in soil.31 For instance, the
degradation rates of 17β-estradiol and 17α-estradiol were
approximately 35 and 12-fold higher than estrone in aqueous
solutions blended with the dairy lagoon water and incubated at
25 °C, respectively. The significant difference in the
biodegradation rates of these hormones is attributed to subtle
variations in their molecular structure.
Transformation Mechanism and Pathway at 35 °C
under Anaerobic Conditions. Following the initial kinetic
study, longer-term transformation experiments were conducted
for 52 days at 35 °C for the three hormones with an initial

42. concentration at 5 mg L−1. The concentration of each
estrogenic hormone is plotted as a function of time in Figure
3. The same transformations were observed for the three
hormones in studies conducted at their low initial concen-
trations under anaerobic conditions at 35 °C (SI, Figure S2).
Relevant information is reported in the SI for studies conducted
at an initial hormone concentration of 5 μg L−1. This section
primarily focuses on discussing the transformation studies
conducted at an initial hormone concentration of 5 mg L−1.
To elucidate the anaerobic degradation pathways for the
three hormones (initial concentration at 5 mg L−1) in blended
CAFO wastewater, solution aliquots were periodically with-
drawn and analyzed by HPLC/DAD/MS. A representative
HPLC/DAD chromatogram exemplifying the hormone trans-
formation process is shown in Figure S3 of the SI. Peak
identification was concurrently performed by LC/MS. To
better identify the transformation products, three hormone
standards were run to verify chromatographic separation and
mass spectra with desired fragmentations.
The initial loss of 17α-estradiol was accompanied by an
accumulation of a degradation product over the first two days
(Figure 3a). This product was characterized as estrone
according to the analysis of its mass spectrum (SI, Figure S4-
a) and retention time. This observation is consistent with
previous reports that estrone is the major degradation product
of 17α-estradiol.24,31−33 Interestingly, another product identi-
fied as 17β-estradiol was detected in the degradation process of
17α-estradiol after two days. 17α-Estradiol and 17β-estradiol
were observed with the same mass spectra (SI, Figure S4-b) but
at different retention times on the chromatogram (SI, Figure
S3). This provides an explanation for why both estradiol
isomers have been detected in dairy wastes,15 although cow
species only excrete 17α-estradiol.12 A recent study
investigated

43. the degradation of 17α-estradiol in the feedlot surficial soil
under simulated rainfall. It also showed that a concentration
decrease of 17α-estradiol was accompanied by an equivalent
increase in estrone and 17β-estradiol.22
Table 1. Initial Pseudo-First-Order Biodegradation Rate
Constants (h−1) and Correlation Coefficient (r) of Three
Hormones in
Aqueous
Solution
s Blended with Dairy Lagoon Wastewater at Different
Temperaturesa
17α-estradiol 17β-estradiol estrone
temperature ki r ki r ki r
15 °C 0.60 (±0.04) × 10−2 0.991 1.43 (±0.05) × 10−2 0.997
0.68 (±0.08) × 10−3 0.963
25 °C 1.28 (±0.10) × 10−2 0.988 3.68 (±0.19) × 10−2 0.994
1.06 (±0.07) × 10−3 0.990
35 °C 1.72 (±0.05) × 10−2 0.998 6.75 (±0.22) × 10−2 0.997
2.49 (±0.34) × 10−3 0.950
45 °C 0.73 (±0.12) × 10−2 0.938 4.30 (±0.15) × 10−2 0.997
1.71 (±0.20) × 10−3 0.962

44. aThe initial concentration of each investigated steroid hormone
was 5 mg L−1.
Figure 2. Effect of incubation temperature on the initial
biodegradation rates of three hormones in aqueous solutions
blended
with dairy lagoon water under anaerobic conditions.
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Similarly, estrone was immediately detected following the
rapid degradation of 17β-estradiol in aqueous solutions
containing dairy lagoon water (Figure 3b). The rate of estrone
formation equaled the rate of 17β-estradiol consumption within
the first two days indicating that it is the primary trans-
formation product. The formation of estrone achieved a
maximum when the lowest concentration of 17β-estradiol was
detected in the solution after two days. The concentration of

45. estrone then decreased with a concurrent formation of a new
-estradiol, inferring that estrone was subse-
quently transformed to 17α-estradiol with further incubation.
These two products were further confirmed using their
authentic standards. Simultaneously, the concentrations of
17β-estradiol were surprisingly observed to transiently increase
for 4 days, decrease over the subsequent 10 days, and then
accumulate again (Figure 3b). By contrast, the time course of
the major product estrone in the transformation process had a
reversed change pattern compared to that of 17β-estradiol
(Figure 3b). Similar degradation and formation patterns were
also observed for 17α-estradiol and its degradation products
(Figure 3a). These results indicate that 17α-estradiol and 17β-
estradiol could be readily oxidized to estrone in aqueous
solutions containing dairy wastewater, and the latter could be
reduced back to both estradiol isomers under anaerobic
conditions.
The transformation study of estrone in solution blended with
dairy lagoon water further confirmed that both estradiol
isomers were formed in its degradation process (Figure 3c).
A reversible transformation pathway related to the three
hormones is illustrated in Figure 4. The racemization reaction
between 17α-estradiol and 17β-estradiol via estrone has been

46. suggested in previous reports.15,34 The reversible trans-
formation among the three estrogenic hormones under
anaerobic conditions suggests that these compounds may
persist in an anoxic aquatic setting, such as anaerobic CAFO
lagoons and underlying sediments. For the studies conducted at
an initial hormone concentration of 5 μg L−1, the same
reversible transformation pathway was observed under
anaerobic conditions. Relevant information is provided in the
SI.
The total hormone contents with time in blended CAFO
wastewater under anaerobic conditions are shown in Figure 3
and Figure S2 of the SI. Throughout the test period, a small
reduction in total hormone content was observed for both
hormone degradation experiments conducted at 5 mg L−1 and
5 μg L−1. For instance, the total hormones (initial
concentration at 5 mg L−1) remaining in the solutions after
52 days at 35 °C were 3.85, 3.90, and 4.26 mg/L, which
corresponded to about 85%, 78%, and 77% of initial spiked
amounts of 17α-estradiol, 17β-estradiol, and estrone, respec-
tively. Also, the control experiments conducted in the sterile
solutions revealed that the abiotic degradation of three
hormones was minimal (Figure 3 and Figure S2 of the SI).
These results suggest there should be other degradation
pathways in addition to the reversible transformation and/or

47. estrone could be further degraded. Previous studies revealed
Figure 3. Anaerobic transformation of investigated hormones
and
formation of their degradation products in aqueous solutions
mixed
with dairy lagoon water at 35 °C: (a) 17α-estradiol; (b) 17β-
estradiol;
and (c) estrone. The initial concentration of each investigated
steroid
hormone was 5 mg L−1. Standard deviation of triplicate samples
is
shown as error bars.
Figure 4. Reversible transformation pathway among three
hormones
in aqueous solutions blended with dairy lagoon water under
anaerobic
conditions.
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46, 5471−54785475

48. that estriol was one of main metabolites during estrone
degradation,14,31 but not shown in this study. Additionally,
this
study did not detect any other likely metabolites, presumably
because estrone can be quickly degraded through a ring
cleavage and mineralized to carbon dioxide by a tricarboxylic
acid cycle (Figure 4).32,41 Thus, further work is needed to
elucidate all degradation pathways and mechanisms, although
they are minor compared to the reversible transformation of
the three hormones that occurred in this experimental system.
Transformation Kinetics of the Three Estrogenic
Hormones at 35 °C under Anaerobic Conditions. As
shown in Figure 4, the anaerobic transformation of the three
estrogenic hormones in aqueous solutions blended with dairy
lagoon water involved two consecutive reversible first-order
reactions:
←→ ←→
− −
k

49. k
k
k
A B C
1
1
2
2 (3)
The differential rate equations are
= − + −
C
t
k C k C
d
d

50. A
1 A 1 B (4)
= − + +− −
C
t
k C k k C k C
d
d
( )B 1 A 1 2 B 2 C (5)
= − −
C
t
k C k C
d
d
C
2 B 2 C (6)

51. where CA is the concentration of the investigated hormone
(17α-estradiol or 17β-estradiol), CB is the concentration of
estrone, CC is the concentration of transformation product
(17β-estradiol or 17α-estradiol), k1 and k2 are the trans-
formation rate constants denoting the forward steps, and k−1
and k−2 are the reversible transformation rate constants
denoting the backward steps.
These differential equations with constant coefficients can be
solved using Laplace transform theory.42 After taking the
Laplace transform of eqs 4−6, rearranging and obtaining the
inverse Laplace transform, the appropriate solutions to eqs 4−6
are given as follows, respectively
α α
α α β
β β
β α β
=
αβ
+

52. − + −
−
−
− + −
−
β
− − −α
− − −
− − − −
⎡
⎣ ⎢
⎤
⎦ ⎥
C t
C k k
e

53. C k k C k C m
e
C k k C k C m
( )
[ ( )]
( )
[ ( )]
( )
t
A
o 1 2 t
o 1 2 Bo 1 Ao
o 1 2 Bo 1 Ao
(7)

54. α α α
α α β
β β β
β α β
=
αβ
+
− + − + −
−
−
− + − + −
−
α
β
−

55. − − −
− − −
⎡
⎣ ⎢
⎤
⎦ ⎥
C t
C k k
e
C k k C k k C k
e
C k k C k k C k
( )
( ) [ ( ) ( )]
( )
( ) [ ( ) ( )]

56. ( )
t
t
B
o 1 2
o 2 1 Ao 2 1 Bo 1
o 2 1 Ao 2 1 Bo 1
(8)
α β β β
α α β
α β α α
β α β
=
αβ
+

57. − − + − + + −
−
−
+ − − − + + −
−
α
β
−
− − − −
−
− − − −
⎡
⎣ ⎢
⎤

58. ⎦ ⎥
C t
C k k
e
C k C k k C k k k
e
C k k k C k C k k
( )
[ ( ) ( )] ( )
( )
( ) [ ( ) ( )]
( )
o
t
t

59. Ao
C
1 2
Ao 2 Bo 2 2 o 2 2 2
o 2 2 2 2 Bo 2 2
(9)
where
α β+ = + + +− −k k k k1 1 2 2 (10)
αβ = + +− − −k k k k k k1 2 1 2 1 2 (11)
= + +− −m k k k1 2 2 (12)
and CAo, CBo, and CCo, respectively, are the initial concen-
trations of [CA], [CB], and [CC]. The mass balance relationship
for three differential equations in the process of hormone
transformation is CT = CA + CB + CC, where CT is the total
hormone concentration in solution.
The rate constants of the transformation reactions for the

60. three hormones exhibited in eq 3 were obtained by optimizing
the fit of the analytic solution to the corresponding
concentration−time data using nonlinear regression. The
reversible first-order kinetic model produced a good fit to the
data (r > 0.97). Table 2 lists the transformation rate constants
at 35 °C for each spiked hormone (initial concentration at 5 mg
L−1) in blended dairy lagoon water under anaerobic conditions.
The transformation rate constants for 17α-estradiol or 17β-
estradiol from the reversible kinetic model (k1) were close to
their initial degradation rate constants (Table 1). Similarly, the
initial biodegradation rate constant of estrone nearly equals the
sum of its transformation rate constant to both estradiol
isomers. This result indicates that the transformation kinetics of
the three hormones can be well-described by eqs 7−9.
Additionally, the rate constant for the transformation from
estrone to 17β-estradiol was approximately 4.5-fold higher than
that of the conversion from estrone to 17α-estradiol. This result
is in agreement with the formation of both estradiol isomers
during the degradation of estrone, in which the formation
concentration of 17β-estradiol was higher than 17α-estradiol
(Figure 3c).
Similarly, simulations based on the reversible first-order
reactions provided a good description of the data for studies

61. Table 2. Anaerobic Transformation Rate Constants and
Correlation Coefficient (r) of Hormones at 35 °C Calculated
from their
Consecutive Reversible First-Order Reactions Shown in eq 3
investigated hormone k1 (d
−1) k−1 (d
−1) k2 (d
−1) k−2 (d
−1) r2
17α-estradiol 0.414 ± 0.085 0.422 ± 0.111 0.064 ± 0.026 0.077
± 0.042 0.974
17β-estradiol 1.518 ± 0.294 0.455 ± 0.135 0.010 ± 0.008 0.021
± 0.048 0.943
estronea 0.076 ± 0.077 0.013 ± 0.001 0.058 ± 0.021 0.215 ±
0.096 0.992
aWhen estrone served as the source hormone, k1 and k−1
denoted transformation rate constants between estrone and 17α-
estradiol; k2 and k−2
represented transformation rate constants between estrone and

62. 17β-estradiol. The initial concentration of each investigated
hormone was 5 mg L−1.
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46, 5471−54785476
conducted at an initial hormone concentration of 5 μg L−1 (r >
0.97) (SI, Table S2). The effect of initial hormone
concentrations on the transformation rates is provided in the
SI.
Environmental Significance. In this study, a reversible
transformation process among 17α-estradiol, 17β-estradiol, and
estrone is clearly illustrated in the blended dairy lagoon water.
The racemization of the two estradiol isomers readily occurred
under anaerobic conditions due to the reduction of the ketone
group of estrone to the hydroxyl group of 17α-estradiol or 17β-
estradiol (Figure 4). 17α-Estradiol and 17β-estradiol have been
suggested as principal indicators to distinguish different
livestock sources of estradiol in environmental samples, since
cattle primarily excretes 17α-estradiol while 17β-estradiol

63. mainly occurs in the excreta of swine or poultry.12 However,
the stereotransformation of the two estradiol isomers through
estrone may result in inaccurate source apportionment analyses
if they are used to identify the livestock species contributing to
waterway contamination.14 More importantly, this trans-
formation mechanism may alter the endocrine-disrupting
activity of dairy lagoon water because 17β-estradiol and estrone
have much higher estrogenic potency compared to 17α-
estradiol. For example, a toxicological study on aquatic species
revealed that 17β-estradiol is generally several orders of
magnitude more potent than 17α-estradiol.43 Therefore, the
alteration of the biological activity resulting from the reversible
transformation of the three estrogenic hormones in CAFO
lagoons should be considered when addressing their potential
adverse impact on the environment.
Results from this study demonstrate that the total amount of
hormones slowly decreased in the aqueous solutions containing
animal wastewater under anaerobic conditions (Figure 3),
although the three estrogenic hormones could be readily
transformed during their initial degradation processes (Table
1). It can be surmised that these estrogenic hormones will
persist and possibly accumulate in anaerobic or anoxic
environments. For instance, aquatic sediments and ground-
water have the potential to be a reservoir for these endocrine

64. chemicals, especially when they are anoxic.34 This finding is
noteworthy because it may help explain why estrogenic
hormones are frequently detected in many natural aquatic
systems. By contrast, aeration is very effective at eliminating
hormone contaminants because the biodegradation of these
endocrine chemicals under aerobic conditions is very
rapid.44,45
Unlike aerobic ponds, CAFO lagoons typically function as
anaerobic reactors. Moreover, CAFO wastewater does not
require additional treatment as long as it does not discharge
directly into surface waters. This suggests that steroid
hormones frequently detected in aquatic systems may be
primarily attributable to anaerobic animal wastewater. Further
research is needed to better understand the contamination
potential related to releases of steroid hormones when reusing
CAFO lagoon water on agricultural lands.
■ ASSOCIATED CONTENT
*S Supporting Information
Detailed description of transformation of hormones at a low
concentration (5 μg L−1) and a high concentration (5 mg L−1),
main compositions and hormone concentrations of the dairy
wastewater, MRM setting and chromatograms of hormones and
their degradation products, anaerobic transformation rate

65. constants of hormones at the initial concentration of 5 μg
L−1, plots of degradation of hormones and formation of their
metabolites, an HPLC/DAD chromatogram showing three
estrogenic hormones, and electrospray LC/MS mass spectra.
This material is available free of charge via the Internet at
http://pubs.acs.org.
■ AUTHOR INFORMATION
Corresponding Author
*Phone: (217)-333-7276; fax: (217)-333-8944; e-mail:wz
[email protected]
Notes
The authors declare no competing financial interest.
■ ACKNOWLEDGMENTS
This study was supported by the Agriculture and Food
Research Initiate Competitive Grant no. 2010-65102-20403
from the USDA National Institute of Food and Agriculture. We
gratefully thank Christie Teausant and Brian Meschewski for
their assistance with some of the sample analysis. We thank
three anonymous reviewers for their valuable comments and
suggestions for improving this paper.
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