-
Nitrate dynamics within the Pajaro River,
a nutrient-rich, losing stream
C. Ruehl1, A.T. Fisher1,2, M. Los Huertos3,4, S. Wankel5, C.G.
Wheat6, C. Kendall5, C.
Hatch1, and C. Shennan3,4
1 Earth Sciences Department, University of California, Santa
Cruz, CA, 95064, USA
2 Institute for Geophysics and Planetary Physics, University of
California, Santa Cruz, CA, 95064, USA
3 Environmental Studies Department, University of California,
Santa Cruz, CA, 95064, USA
4 Center for Agroecology and Sustainable Food Systems,
University of California, Santa Cruz, CA, 95064,
USA
5 Stable Isotope Laboratory, U. S. Geological Survey, Menlo
Park, CA 94301, USA
6 Global Undersea Research Unit, University of Alaska,
Fairbanks, AK, 99701, USA
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 1
-
Abstract 1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
The major ion chemistry of water from an 11.42 km reach of the
Pajaro River, a
losing stream in central coastal California, shows a consistent
pattern of increasing
concentrations during the second (dry) half of the water year,
along with conservation of
most species. Nitrate concentration ([NO3-]) decreases
consistently along this reach by
~30% at a time when the there is both a significant loss of
channel discharge and
extensive surface-subsurface exchange. The corresponding average
net NO3- uptake
length is 37 ± 13 km, or 42 ± 12 km when normalized to the
conservative solute Cl-. The
similarity of these values indicates that dilution by
discharging groundwater does not
significantly contribute to the decrease in [NO3-]. Given these
uptake lengths and typical
values of channel [NO3-], discharge, and width late in the water
year, the areal NO3-
uptake rate is 0.5 μmol m-2 s-1 The observed reduction in [NO3-]
and channel discharge
along the experimental reach represents an absolute NO3- sink of
~50%, comprising a net
removal rate of 200 - 400 kg day-1 nitrate-N. High-resolution
(temporal and spatial)
sampling indicates that most of the NO3- loss occurs along the
lower part of the reach,
which is also where most seepage loss and hydrologic exchange of
water occurs. Stable
isotopes of NO3-, phosphorus concentrations, and streambed
chemical profiles suggest
that denitrification is the most significant nitrate sink along
the reach. Denitrification
efficiency, as expressed through downstream enrichment in
15N-NO3- (assuming
denitrification is the dominant nitrate sink), varies
considerably during the water year:
when discharge is greater (typically earlier in the water year),
denitrification is least
efficient and downstream enrichment in 15N is greatest. When
discharge is lower,
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 1
-
23
24
25
26
27
28
29
30
31
32
33
34
35
36
37
38
39
40
41
42
43
44
45
denitrification in the streambed appears to occur with greater
efficiency, resulting in
lower downstream enrichment in 15N.
Keywords: Rivers and streams; Solution transport; Nitrates;
N-15/N-14; O-18/O-16
1. Introduction
The global rate of nitrogen fixation doubled during the
twentieth century due to
numerous human activities, the most important being increased
application of fertilizer
(Galloway et al., 1995; Vitousek et al., 1997). Because nitrate
(NO3-) is more stable and
mobile than other common fixed nitrogen compounds, increased
loading of nitrogen is
often expressed as elevated [NO3-] in streams (Duff and Triska,
2000), i.e., [NO3-] greater
than ~70 μM (e.g., Bohlke et al., 2004; Holloway et al., 1998;
Schiff et al., 2002).
Adverse effects of elevated [NO3-] in streams are well
documented, and include
eutrophication (Neal and Jarvie, 2005; Turner and Rabalais,
1994) and contamination of
groundwater (e.g., Bohlke and Denver, 1995; Bohlke et al., 2002;
Nolan, 2001). Human
health effects of NO3- in drinking water are widely known, and
have prompted the U.S.
EPA to set a standard for maximum [NO3-] in drinking water of
714 μM (10 mg N/L)
(e.g., Kendall, 1998; Nolan et al., 1997).
Nitrogen export via streams is often lower than watershed
inputs, implying that
nitrogen sinks, along with accumulation in biomass and export to
aquifers, are important
in many catchments (e.g., Alexander et al., 2000; Bernhardt et
al., 2003; Sjodin et al.,
1997). Relatively low [NO3-] in stream water, despite relatively
high inputs, can be
explained in some systems by NO3- removal in riparian buffer
zones between discharging
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 2
-
46
47
48
49
50
51
52
53
54
55
56
57
58
59
60
61
62
63
64
65
66
67
68
groundwater and stream channels (e.g., Cirmo and McDonnell,
1997; McMahon and
Bohlke, 1996; Peterjohn and Correll, 1984; Sebilo et al., 2003).
However, the
anthropogenic influence on global fixation of nitrogen is
relatively recent, and it is not
known how buffer zones may help to mitigate high [NO3-] in
groundwater on longer time
scales (Bohlke, 2002; Galloway et al., 1995); thus, NO3-
contamination in both aquifers
and streams may become a more significant problem in coming
decades. It is therefore
important to develop a better understanding of factors that
control the spatial and
temporal extent of NO3- sinks in catchments, particularly
in-stream sinks where [NO3-] is
high, to quantify, control and mitigate current and future
impacts on the quality of both
surface and ground water.
Many studies have documented uptake of nutrients during
transport in streams.
Uptake rates have been related to many environmental variables,
including solar flux
incident on the stream (Hill et al., 2001; Mulholland and Hill,
1997), dissolved organic
carbon concentrations (Bernhardt and Likens, 2002), dissolved
oxygen concentrations
(Christensen et al., 1990; Laursen and Seitzinger, 2004), types
of vegetation (Schade,
2001), and the effects of logging and other human activities
(Sabater et al., 2000). Due to
the importance of processes occurring in and on stream
sediments, the magnitude of
surface-subsurface exchange is an important control over the
potential nutrient uptake
rate for a given reach (e.g., Duff and Triska, 1990; Duff and
Triska, 2000; Valett et al.,
1996; Wondzell and Swanson, 1996). Many streambed processes can
remove solutes
from downstream transport, including weathering reactions
(Gooseff et al., 2002),
sorption to streambed sediments (McKnight et al., 2002), and
retention of colloids and
sorbed materials (Ren and Packman, 2004). But removal of
nutrients, particularly NO3-,
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 3
-
69
70
71
72
73
74
75
76
77
78
79
80
81
82
83
84
85
86
87
88
89
90
91
is commonly attributed to microbiological activity within the
streambed (e.g., Butturini
and Sabater, 1999; Cirmo and McDonnell, 1997; Grimaldi and
Chaplot, 2000; Hall et al.,
2002; Hinkle et al., 2001; Mulholland and Hill, 1997; Triska et
al., 1989). Unlike
assimilative uptake, transformation of NO3- to N2 gas via
denitrification removes fixed
nitrogen from the stream system. Streambed seepage and its
influence on denitrification
is of particular interest within losing streams that contribute
water to underlying aquifers,
and relatively few studies have been completed within losing
streams having [NO3-] near
or above drinking water standards (e.g., Grimaldi and Chaplot,
2000; Sjodin et al., 1997).
In this paper, we present geochemical results and quantify
relations between
streambed seepage and NO3- removal over a range of discharge and
seepage rates within
an experimental reach of a single stream. We use the term
"streambed seepage" to refer to
the movement of water across the streambed, both entering and
leaving the stream
channel. Results of differential discharge gauging and tracer
experiments in the same
reach are presented in a separate paper, including independent
estimates of hydrologic
exchange rates (Ruehl et al., in press). We identify parts of
the experimental reach where
and when there are significant NO3- sinks, determine the
dominant mechanism of NO3-
removal, and place quantitative constraints on rates of NO3-
cycling in this nutrient-rich,
losing stream.
2. Field Setting and Experimental Design
We instrumented and sampled an 11.42 km reach of the Pajaro
River in the Pajaro
Valley (Fig. 1); see also Ruehl et al. (in press) for more
information concerning local
geology, climate, and hydrology. The upper end of the
experimental reach is defined by a
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 4
-
92
93
94
95
96
97
98
99
100
101
102
103
104
105
106
107
108
109
110
111
112
113
114
gauging station developed and maintained by the U. S. Geological
Survey (Station
#11159000, Chittenden). Historical discharge records extend from
1939 to the present
whereas water quality data are available from 1952 to the
present
(http://waterdata.usgs.gov/nwis/uv/?site_no=11159000). Mean
daily discharge at this
station during the period of record varied from 0 to >600
m3/s, and peak discharge
exceeded >700 m3/s on several occasions. Annual precipitation
in the basin is generally
20-60 cm/yr. Most precipitation falls during winter and early
spring, whereas the late
spring to fall are generally dry. Temperatures rarely drop below
freezing, so most
precipitation in the basin falls as rain. Thus there are two
distinct hydrologic periods
apparent on stream flow hydrographs and chemographs during each
water year (1
October through 30 September of the following year): (1) winter
conditions,
characterized by high and highly variable discharge and
relatively low solute
concentrations, and (2) summer conditions, characterized by
lower flows (typically
-
115
116
117
118
119
120
121
122
123
124
125
126
127
128
129
130
131
132
133
134
135
136
137
extracted from shallow alluvial and underlying Aromas aquifers
(Muir, 1977). Of current
groundwater extraction in the Pajaro Valley, about 65% is
overdraft, resulting in seawater
intrusion near the coast and a loss of storage throughout the
basin (PVWMA, 2001). The
impacts of overdraft on surface water - ground water
interactions in the basin are not well
understood, but the experimental reach lost discharge via
streambed seepage at a rate of
0.2-0.4 m3/s during the second half of the 2002-04 water years.
It would be difficult to
quantify similar rates of streambed seepage during winter flows,
but assuming that the
documented loss extends throughout the water year, seepage along
the experimental
reach could comprise ~20-40% of current sustainable basin yield
(Ruehl et al., in press).
The historical influence of agricultural development in the
Pajaro Valley
hydrologic basin (Los Huertos et al., 2001) is readily apparent
in [NO3-] measured in
river water collected at the Chittenden gauging station.
Although there is considerable
variability in water quality within and between years, peak
[NO3-] has risen considerably
over the last 50 years. [NO3-] was generally below 0.1 mM in the
early-mid 1950's, but
now commonly exceeds the drinking water standard of 0.714 mM
(Fig. 1C).
The focus of the present study is on changes in water chemistry
that occur
downstream of the Chittenden station, the reference point for
all stream distances, along
an 11.42 km reach (Fig. 1B). We established an additional
gauging stations at Km 8.06
(2003 water year) and Km 11.42 (2002-04 water years); additional
stream discharge
measurements were made periodically at locations throughout the
experimental reach
both to calibrate rating curves for the gauging stations, and to
quantify changes in
discharge in the channel. During the summer and early fall,
discharge generally decreases
downstream along this reach, with most of the loss occurring in
the lower ~3 km of the
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 6
-
138
139
140
141
142
143
144
145
146
147
148
149
150
151
152
153
154
155
156
157
158
159
reach (Ruehl et al., in press). In this study of stream
chemistry we focused on the second
half of the water year. Mass balance of NO3- is easier when
there are no significant fluid
inflows or outflows along the experimental reach other than
channel discharge and
streambed seepage, and quantifying changes in stream discharge
(required for
quantifying NO3- fluxes) is more difficult and less accurate in
an absolute sense when
discharge is >5 m3/s. In addition, we knew that [NO3-] would
be elevated during much of
the measurement period, simplifying estimation of removal
rates.
3. Methods
3.1. Water Sampling
We conducted synoptic sampling of the experimental reach on 47
separate days
throughout the 2002-04 water years, in which stream water from
the top, the bottom, and
1-9 intermediate sites was obtained. Samples were collected in
pre-cleaned HDPE bottles
after the bottles were rinsed 3 times with stream water. We
immediately placed samples
on ice, and filtered them in the lab through 0.45 μm glass fiber
filters within 12 hours of
collection. Samples were either analyzed within 48 hours of
collection, or were frozen
(anions and nutrients) or chilled (cations) after filtration
until immediately before
analysis. In addition to synoptic sampling, we conducted diel
sampling at a frequency of
2 hours for 48-hr periods at the top and bottom of stretches
(subsections of the
experimental reach) associated with tracer tests. These samples
were collected with an
automated sampler and recovered from the field within 12 hours
of the last sample,
returned to the lab and immediately filtered and stored as
described above. Finally, a
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 7
-
small subset of samples for isotopic analysis of NO3- were
filtered through 0.22 μm filters
immediately after collection, transported back to the lab on
ice, and frozen until analysis.
160
161
162
163
164
165
166
167
168
169
170
171
172
173
174
175
176
177
178
179
180
181
We obtained samples from the streambed using passive dialysis
samplers
(“peepers”) and piezometers. Peepers consist of a series of 5 mL
chambers carved into a
rigid sheet of polycarbonate at 2 cm intervals. Chambers are
filled with deionized water
and a 0.4 μm semi-permeable polycarbonate membrane is attached,
covering the
openings to the chambers. Peepers are submerged in a container
filled with deionized
water and N2 is bubbled through the water around the peepers for
≥10 days to
deoxygenate the water in the chambers. Deoxygenated peepers are
inserted into the
streambed and left deployed for ≥2 weeks, ensuring ample time
for solutes to diffuse into
the chambers and equilibrate with adjacent pore waters. Although
peepers are not
intended to document transient processes, the chemistry of
fluids trapped in the chambers
is most influenced by the last 12-24 hours of diffusive
exchange. Peeper samples were
extracted from the chambers, filtered, and transported on ice to
the laboratory for
analysis. We also obtained subsurface samples for chemical
analyses from drive-point
piezometers, installed at depths of 0.5-1.0 m into the
streambed. Piezometer casings were
purged several times before each sample was collected. Finally,
a small number of
groundwater samples were taken from agricultural production
wells in the vicinity of the
experimental reach. Piezometer and well samples were filtered,
preserved and analyzed
using the same methods as surface water samples.
3.2. Analytical methods
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 8
-
182
183
184
185
186
187
188
189
190
191
192
193
194
195
196
197 198
199
200
201
202
203
204
We measured dissolved oxygen, temperature, and pH of surface
water in the field
with a Hydrolab multiprobe system. Ammonium, soluble reactive
phosphorus (SRP), and
nitrate + nitrite were measured by colorimetric flow-injection
analysis (QuickChem 8000,
Lachat Instruments, Loveland, CO). Chloride, nitrite, bromide,
nitrate, phosphate, and
sulfate were measured with ion chromatography (DX-100, Dionex,
Sunnyvale, CA).
Major cations, iron, manganese and total dissolved phosphorus
(TDP) were measured by
emission spectrometry (Optima 4300 ICP-OES, PerkinElmer,
Wellesley, MA). Samples
for iron and manganese were acidified to pH
-
LNO
NONOL
mean
bottomtopNO ×
−= −
−−
][][][
3
33]3[ (1) 205
206
207
208
209
210
211
212
213
214
215
216
217
218
219
220
221
where [NO3-]top and [NO3-]bottom are the nitrate concentrations
at the top and bottom of the
reach, [NO3-]mean is the average of [NO3-]top and [NO3-]bottom,
and L is the reach length.
Use of (1) to calculate L[NO3] assumes that NO3- uptake is
zeroth-order with respect to
[NO3-] (i.e., that the downstream decrease in [NO3-] is linear).
Although uptake lengths
are commonly calculated assuming a first order (i.e.,
exponential) decrease in [NO3-]
(e.g., Stream Solute Workshop, 1990), we use (1) because we
compare L[NO3] to other
hydrologic exchange rates which are independent of [NO3-], and
because results obtained
assuming a linear and an exponential decrease in [NO3-] were
virtually identical (see
Discussion). When [NO3-] data from more than two locations along
the reach were
available, L[NO3] is equal to the negative inverse of the slope
obtained by regression of
[NO3-] on downstream distance (Stream Solute Workshop,
1990).
We also calculated L[NO3]:[Cl], a dilution-corrected (e.g.,
Haggard et al., 2005)
NO3- uptake length based on the conservative solute Cl-. The
equation used was identical
to (1), except that all [NO3-] were divided by the [Cl-] of the
same samples. In addition,
we calculated areal uptake rates corresponding to uptake lengths
(Stream Solute
Workshop, 1990):
wL
NOQU
NONO ×
×=
−
]3[
33
][ (2) 222
223
224
225
226
where Q is stream discharge and w is stream width. Finally, we
note that although other
techniques such as 15N- NO3- additions are able to distinguish
between net and gross
nitrate removal (e.g., Bohlke et al., 2004; Mulholland et al.,
2004), these would be
expensive in the Pajaro River due to its relatively high N- NO3-
flux of ~500 kg/day.
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 10
-
Stable isotopes of NO3- are useful for assessing sources and for
distinguishing
between NO3- sinks (e.g., Burns and Kendall, 2002).
Biologically-mediated
denitrification enriches residual NO3- in both 15N and 18O,
whereas biological NO3-
uptake generally results in little or no enrichment. The
magnitude of δ15N-NO3
enrichment associated with NO3- removal is quantified through an
enrichment factor, εN,
defined by a form of the Rayleigh equation (Sebilo et al.,
2003):
227
228
229
230
231
232
033
01515
]ln[]ln[ −− −−
=NONONN
Nδδ
ε (3) 233
234
235
236
237
238
239 240 241
242
243
244
245
246
247
248
249
where δ15N = stable isotope ratio for N- NO3-, and δ15No and
[NO3-]0 are the nitrogen
stable isotope ratio and concentration, respectively, of
original (reference) NO3-. There is
a similar enrichment factor for oxygen, εO. Enrichment factors
for δ15N-NO3-and δ18O-
NO3- in surface waters were determined by regression of δ15N (or
δ18O) on ln[NO3-].
4. Results 4.1 Synoptic sampling
At individual sites along the reach, concentrations of major
cations and anions in
Pajaro River water increased during the second half of the water
year (Fig. 2), as
expected when stream discharge decreases. Concentrations of none
of the major ions
consistently changed from the upper to the lower end of the
experimental reach during
the last half of the water year, with one notable exception:
[NO3-] decreased by ~30% late
in the water year (Fig. 2C, Table 1). Uptake lengths were
calculated for 14 late-year days
in which synoptic sampling from at least 3 locations along the
reach occurred, and the
average values of L[NO3] and L[NO3]:[Cl] were 37 ± 13 and 42 ±
12 km, respectively (Table
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 11
-
1). Given typical late-year values of discharge, [NO3-], and
stream width, these uptake
lengths correspond to an areal uptake rate of ~0.5 μmol m-2 s-1.
At times, total dissolved
phosphorus (TDP) concentrations along the reach also decreased
(Fig. 2D), although the
magnitude of TDP reduction (~4 μM when observed) was much
smaller than that for
[NO3-] (~400 μM). These decreases in TDP were seen at the bottom
of the experimental
reach (Km 11.42) late in the water year, when discharge and
velocities at that location
were relatively low (
-
8.44 to 9.84 (Fig. 3C). L[NO3] was usually greater along the
upper portion of the reach, or
could not be calculated at all when there was no significant
change in [NO3-] (Table 2).
273
274
275
276
277
278
279
280
281
282
283
284
285
286
287
288
289
290
291
292
293
294
295
4.3 Pore water sampling
Pore water profiles obtained from peeper samples demonstrate
that NO3- was
removed, at times rapidly, in the streambed (Figs 4 and 5). Most
peepers were installed
where it was determined that water moved into or out of the
streambed, and conservative
solutes such as Cl- generally varied only slightly with depth.
In contrast, solutes likely to
be involved in mineralogical or biogeochemical reactions often
varied considerably with
depth below the stream bottom. Nitrogen species concentrations
were notably variable,
with [NO3-] generally decreasing to a local minimum at 5-10 cm
below the stream
bottom. Multiple local maxima and minima in [NO3-] are apparent
in some peeper
profiles (e.g., Figs 4A, 5A-C), with low [NO3-] values often
accompanied by increases in
[Mn] and [Fe], and decreases in [SO42-]. Small increases in
[NO2-] and [NH4+] were
sometimes collocated with decreases in [NO3-] (Figs 4A, 5A).
4.4 Stable isotopes of NO3-
There is a broad trend of correlated enrichment in 15N and 18O
of all NO3-
samples, with a relative fractionation (εN:εO) of ~2:1 (Fig. 6).
This is consistent with
biologically-mediated denitrification (e.g., Cey et al., 1999;
Lehmann et al., 2003). When
[NO3-] in the subsurface was lower than in the channel, the
associated εN was ~ -20‰
(Figs 4 and 5), except when conservative solutes such as Cl-
also varied with depth (Fig.
4B). Both δ18O-NO3- and δ15N-NO3- in surface water tend to
increase with distance along
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 13
-
the experimental reach as [NO3-] decreases. Collectively, these
data define apparent
Rayleigh enrichment factors of εN,surf = -6.0‰ to -20‰ and
εO,surf = -1.6‰ to -20‰ (Fig.
7). Two distinct regimes of δ15N-NO3- enrichment in surface
water were observed.
Samples collected on 8/26/04 and 5/21/04 experienced relatively
high downstream
enrichment in δ15N-NO3- (εN,surf = -20‰ and -17‰, respectively).
In contrast, samples
collected on five other days had εN,surf between -6.0 and -9.0‰
(Fig. 7A). Greater
enrichment was observed when discharge was relatively high
(>~0.3 m3/s), and apparent
NO3- uptake lengths were longer (L[NO3]>100 km) (Fig. 8A).
Downstream enrichment in
δ18O-NO3- was much more variable (Fig. 7B), but the ratio of
εN,surf to εO,surf consistently
decreased towards the end of the water year, from ~2 to ~0.5
(Fig. 8B).
296
297
298
299
300
301
302
303
304
305
306
307
308
309
310
311
312
313
314
315
316
317
318
5. Discussion
5.1. Rates of NO3- removal
NO3- was quantitatively removed from the experimental reach
during the second
half of the water year by one or more internal processes. This
interpretation is based on
quantitative reductions in [NO3-] relative to more conservative
solutes, an observation
that precludes dilution (either from lateral inflow of ground or
surface water) as a
possible explanation. Reductions in discharge and [NO3-]
correspond to removal of 200 -
400 kg N/day (Fig. 2E), with the missing NO3- either recharging
underlying aquifers, or
being lost to temporary or permanent sinks (e.g., assimilation
into biomass or
denitrification, respectively). If metrics of NO3- uptake are
based on this change in flux,
the reach-averaged NO3- uptake length would be 9.5 km,
equivalent to an areal uptake
rate (UNO3) of 1.4 μmol m-2 s-1. These values, however, include
any NO3- that enters and
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 14
-
319
320
321
322
323
324
325
326
327
328
329
330
331
332
333
334
335
336
337
338
339
340
341
remains in the subsurface but is not assimilated or denitrified,
and should therefore be
considered an lower (upper) limit for L[NO3] (UNO3). If the
discharge lost from the channel
is neglected, calculations are based on the change in [NO3-]
instead of the NO3- flux, and
L[NO3] and UNO3 are equal to 37 ± 13 km and 0.5 μmol m-2 s-1,
respectively. Uptake
lengths calculated assuming an exponential downstream decrease
in [NO3-] (e.g., Stream
Solute Workshop, 1990) were 37 ± 14 km, or virtually identical
to those assuming a
linear decrease. We therefore concluded that the assumption of a
linear [NO3-] decrease
inherent in (1) is appropriate in this system.
We found that dilution by inflow of groundwater does not
contribute significantly
to the downstream decrease in [NO3-]. Even in a strongly-losing
stream reach such as
this, it is possible that [NO3-] could be reduced via
groundwater dilution, but three
observations suggest that this process is negligible in this
system. First, average uptake
lengths based on late-year synoptic sampling were not
significantly increased when
corrected for the conservative solute chloride (L[NO3] = 37 ± 13
and L[NO3]:[Cl] = 42 ± 12
km). Also, other solute concentrations did not change
significantly from the top to the
bottom of the reach (Fig. 2). Finally, stable isotope ratios of
water (δ18O and δD) were
constant along the reach, despite the fact that δ18O and δD of
groundwater samples near
the river were distinct and would therefore have likely altered
the isotopic signature of
surface water if groundwater inflow was significant (Ruehl et
al., in press). We conclude
that L[NO3] is an appropriate metric of NO3- uptake in this
system.
L[NO3] along the lower part of the reach was of the same
magnitude as inflow
length (LI, where LI is the stream length required for inflow of
tracer-free water to equal
channel discharge), but one to two orders of magnitude larger
than transient storage
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 15
-
exchange lengths (L[NO3]>>LS, where LS is the average
distance traveled by a water
molecule before entering an adjacent storage zone) (Fig. 9). The
consistency of L[NO3] and
LI may appear at first to conflict with the earlier assertion
that dilution can not explain the
removal of NO3- during transport, but these interpretations are
entirely consistent. If
inflow of tracer-free water is primarily due to stream water
which enters the subsurface
and follows a spatially or temporally long path before
re-entering the main channel, then
the diluting water will have the same major ion chemistry as the
stream except for
nonconservative solutes like NO3- that change in the subsurface.
Ground water inflow
would also lack injected tracer, but has a distinct chemistry
and would be readily
identified on this basis. NO3- removal in the upper part of the
reach was observed during
a single tracer experiment (5/19/04); there was no significant
net change in [NO3-] during
other tracer experiments on this part of the experimental reach.
Uptake and inflow lengths
for the upper part of the reach calculated on the basis of the
5/19/04 tracer experiment are
about the same, but both values are 2-5 times greater than
equivalent lengths determined
for the lower part of the reach, implying less vigorous exchange
and nutrient cycling
processes in the upper part of the reach.
342
343
344
345
346
347
348
349
350
351
352
353
354
355
356
357
358
359
360
361
362
363
364
The uptake lengths reported above tend to be longer than those
reported for more
pristine systems, largely due to the much higher [NO3-] in the
Pajaro River. For example,
NO3- uptake lengths of ~0.1-1.0 km were reported in first-order
New Mexico streams
(Valett et al., 1996), and uptake lengths of 0.004-3.4 km were
reported for a higher-order
Arizona stream (Marti et al., 1997), both of which have [NO3-]
< 12 μM. Uptake lengths
of 0.1-0.4 km were reported in a Kansas stream with [NO3-] as
great as ~100 μM (Dodds
et al., 2002). These authors also reported areal uptake rates
(UNO3) of 0.1 - 1.2 μmol m-2 s-
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 16
-
1, similar to values determined for the Pajaro River and other
nitrogen-rich streams in
Denmark (Christensen et al., 1990), Ontario (Hill, 1979), and
Colorado (Sjodin et al.,
1997).
365
366
367
368
369
370
371
372
373
374
375
376
377
378
379
380
381
382
383
384
385
386
Peeper data are also useful for estimating uptake rates. Several
profiles were
collected from the lower part of the experimental reach, where
strong downward seepage
occurs (Fig. 5). Reductions in [NO3-] in the upper 20-40 cm of
the streambed approached
90% in many locations. Several peeper profiles included two
regions of low [NO3-], with
a typical spacing of [NO3-] variations of ~20 cm. Smaller
variations in dissolved [Mn]
and [NO2-] often accompanied variations in [NO3-], suggesting
that they may indicate
spatial variations in subsurface redox state. One explanation
for vertical variations in
streambed chemistry in this area is that stream water with high
[NO3-] was swept
downward into the streambed, feeding facultative denitrifiers
who consumed NO3- at
different rates throughout the day. Assuming that these microbes
became more active
when dissolved oxygen levels decreased at night, downward
seepage rates of ~20 cm/day
(~2 x 10-6 m/s) are implied by the spatial distribution of
[NO3-] variations in the stream
bed. This corresponds to length-averaged channel loss at ~3 x
10-5 m2/s, consistent with
differential discharge measurements and observed head gradients
(Ruehl et al., in press).
This seepage rate and the magnitude of [NO3-] variations
correspond to U[NO3] ~ 2 μmol
m-2 s-1 at the peeper sampling sites, consistent with the
stretch-specific calculations of
UNO3 described above.
5.2. Mechanisms of NO3- removal
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 17
-
NO3- removal in the Pajaro River may occur via both assimilative
and
dissimilative pathways. Assimilative uptake (production of new
biomass) comprises a
temporary change in the nature of the aquatic nitrogen
reservoir, in that nitrogen can
return rapidly to the stream through degradation and
mineralization. In contrast,
dissimilative removal through denitrification can lead to N2 gas
as an end product and
export from the system. Assimilative uptake is the dominant NO3-
sink in many pristine
stream systems (Duff and Triska, 2000), and is likely to be
active within the experimental
reach of the Pajaro River as well, but denitrification appears
to be the dominant NO3- sink
in this system.
387
388
389
390
391
392
393
394
395
396
397
398
399
400
401
402
403
404
405
406
407
408
409
Production of new biomass requires both nitrogen and phosphorus
in roughly a
30:1 atomic ratio for freshwater systems (Hecky et al., 1993).
Observed N:P removal
ratios in the experimental reach were consistently above 200,
and no significant decreases
in SRP accompanied large decreases in [NO3-] in high-resolution
records (Fig. 3). This
suggests that biomass production in the Pajaro River stream
ecosystem is more strongly
phosphorus-limited than it is nitrogen-limited, although the
extent to which organisms
may utilize P from sources other than the water column (e.g.,
mineralized phosphorous or
that sorbed onto sediments) is unknown. The mineralization of
organic matter would
release both dissolved nitrogen and phosphorus into the river
and thus would not result in
the observed net downstream decrease in NO3- relative to
phosphorus. Thus although
assimilative uptake may occur in this system, it is
insignificant relative to dissimilative
uptake (e.g., denitrification) and/or balanced by mineralization
of organic matter
followed by nitrification, and therefore not responsible for the
observed downstream
removal of NO3-.
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 18
-
NO3- isotopic data provide some of the strongest evidence that
denitrification is
the primary mechanism of NO3- removal in the experimental reach,
and variations in
observed isotopic enrichment are likely linked to the dynamics
of system hydrology.
Numerous studies have explored relations between NO3-
transformations and δ18O-NO3-
and δ15N-NO3- in aquatic settings. Enrichments observed in
laboratory cultures of
denitrifiers and marine settings and aquifers have often been
greater in magnitude (εN as
low as -40‰) than enrichment seen in streams and coastal
sedimentary settings
(Lehmann et al., 2003). εN values closer to zero in many field
settings may result from
extremely rapid denitrification, elevated temperatures, and
diffusion limitation of
denitrification (Mariotti et al., 1988).
410
411
412
413
414
415
416
417
418
419
420
421
422
423
424
425
426
427
428
429
430
431
432
Benthic dentrification may be diffusion limited in the absence
of significant
advective exchange between sediments and surface water (Brandes
and Devol, 1997;
Lehmann et al., 2004; Sebilo et al., 2003; Sigman et al., 2003).
In one set of lab
experiments, diffusion-limited denitrification indicated εN =
-3.7‰, whereas
denitrification with more extensive water-sediment interaction
resulted in εN = -18‰
(Sebilo et al., 2003). If diffusion limitation is sufficiently
strong, enrichment can be
driven close to zero (Brandes and Devol, 1997). Enrichment may
also be driven close to
zero if essentially all available NO3- is denitrified because
microorganisms are compelled
to consume the heavy isotopes as well as the light.
In Pajaro River surface water samples, εN values are often about
twice εO values
(Fig. 6), consistent with results from many field studies (e.g.,
Cey et al., 1999; Lehmann
et al., 2003). Downstream enrichment in δ15N-NO3- in the
experimental reach occurs in
two distinct regimes: a high discharge (high-Q) regime in which
[NO3-] reduction is
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 19
-
modest, and a low discharge (low-Q) regime in which [NO3-]
reduction is more intense.
During high-Q periods, εN,surf associated with the downstream
reduction in [NO3-] is -
17‰ to -20‰ (Fig. 8A), as observed in other river and shallow
marine systems (e.g.,
Brandes and Devol, 1997; Dhondt et al., 2003; Kellman and
Hillaire-Marcel, 1998;
Sebilo et al., 2003). In contrast, during low-Q periods, εN,surf
is between -6‰ and -9‰
(Fig. 8A).
433
434
435
436
437
438
439
440
441
442
443
444
445
446
447
448
449
450
451
452
453
454
Although diffusion-limited denitrification may occur in this
system, it is not a
satisfactory explanation for the bimodal enrichment pattern of
15N. The local influence of
diffusion limitation on isotopic fractionation during
denitrification is apparent in results
from one set of peeper samples recovered at Km 2.72 (Fig. 4B).
Whereas most other
streambed chemical profiles are consistent with rapid vertical
fluid advection, strong
gradients in conservative solutes such as Cl- from this location
indicate more diffusive
conditions, and enrichment in NO3- isotopes is closer to zero
(εN ~ -5‰ and εO ~ -3‰).
Diffusion-limited denitrification such as this, however, is
unlikely to contribute
significantly to overall NO3- removal along the reach. This can
be shown by calculating
the rate at which diffusion can remove NO3- along the entire
reach, based on the diffusion
coefficient for NO3- (Li and Gregory, 1974) and the geometry of
the stream. Even if
[NO3-] changed from 1 to 0 mM over an average length of 2 cm
along the entire stream-
streambed interface of the reach, this would result in a daily
NO3- removal rate of
-
Instead, we believe that the bimodal pattern of enrichment in
15N is caused by
variation in the efficiency of denitrification, which can be
quantified by considering an
idealized two-box model representing surface and subsurface
storage regions (Fig 10A).
For steady-state conditions and no net change in surface
discharge, we use observed
changes in surface [NO3] and ∆δ15N-NO3 values (∆NO3,surf and
∆
455
456
457
458
δ459 15Nsurf, respectively)
and the subsurface isotopic enrichment factor for ∆δ460
461
462
463
464
465
466
467
468
469
470
471
472
473
474
475
476
477
15N-NO3 (εN,sub ~ -20‰) to calculate
the resulting apparent enrichment in surface water (εN,surf). If
subsurface denitrification is
extremely efficient, there will be little fractionation in
surface water (Fig. 10B), even if
there is a large reduction in [NO3], because no NO3- will return
from the subsurface
(εN,surf 0). At the other extreme, if only a small fraction of
the NO3- that is exchanged
with the subsurface is denitrified, the enrichment apparent in
surface water will approach
that in the subsurface (εN,surf εN,sub). During high-Q
conditions on the Pajaro River, low
subsurface denitrification efficiency is consistent with
observed surface enrichment of -
17 to -20‰. During low-Q conditions, subsurface denitrification
efficiency appears to be
relatively high. This increased efficiency results in a much
smaller enrichment of ∆δ15N-
NO3 in surface water (Fig. 10B). Peeper profiles from the lower
portion of the reach tend
to support this interpretation: [NO3-] removal in the subsurface
was more complete when
discharge was ~ 0.2 m3/s and surface fractionation was
relatively weak (Figs 5A, 5B)
than when discharge was ~ 0.4 m3/s and εN,surf ~ -20‰ (Fig.
5C).
Based on this model and geochemical observations shown earlier,
we estimate
that during high-Q conditions, 25-45% of NO3- in the main
channel exchanges with the
subsurface, where it is inefficiently denitrified. This lowers
the [NO3-] of surface water
by 5-10%, and shifts δ15N-NO3 values such that εN,surf = -17‰ to
-20‰ (Fig. 10B). In
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 21
-
contrast, during low-Q conditions, 35-45% of NO3- in the channel
exchanges with the
subsurface and is efficiently denitrified, lowering the [NO3-]
of surface water by ~30%,
and shifting surface δ15N-NO3 values such that εN,surf = -6‰ to
-9‰ (Fig. 10B). If stream
water lost to underlying aquifers is subject to similar
biogeochemical processes as stream
water that enters the subsurface but later flows back into the
main channel, this implies
that the fraction of NO3- removed from aquifer recharge is also
higher during low-Q
conditions. This scenario contrasts with that interpreted for
many more pristine stream
systems, where the presence of carbon or nutrients appears to
provide the primary
limitation on denitrification. In the experimental reach of the
Pajaro River, the extent of
surface - subsurface exchange relative to discharge appears to
be the most important
control on denitrification, particularly when apparent NO3-
removal rates are highest.
478
479
480
481
482
483
484
485
486
487
488
489
490
491
492
493
494
495
496
497
498
499
500
The downstream enrichment in 18O-NO3 was not as consistent as
that for 15N-
NO3-, varying over a large range and not falling into distinct
regimes. However, the ratio
of εN,surf to εO,surf decreases consistently with time,from ~2
(consistent with
denitrification) to ~0.5 (Fig. 8B). One explanation for this
trend is that nitrification
becomes increasingly important as the water year progresses; a
subset of diel
observations suggest that there may be an internal source for
NO3- late in the water year
along parts of the experimental reach (Fig. 3C). The coupling of
denitrification to
nitrification could result in the incorportation into residual
NO3- of isotopically-heavy
atmospheric oxygen, which has δ18O = 23.5‰ (Keeling, 1995),
and/or “light” nitrogen (if
an NH4+ source were depleted in 15N). Either or both of these
effects could result in a
decrease in (εN/εO)surf. If nitrification is important along the
experimental reach, then
gross denitrification rates in this system may be considerably
greater than our estimates
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 22
-
of the net rate, because nitrification would act as an internal
NO3- source. Additional
work along the experimental reach will be required to assess the
importance of
nitrification relative to denitrification, and to determine the
cause for high variability in
εO,surf.
501
502
503
504
505
506
507
508
509
510
511
512
513
514
515
516
517
518
519
520
521
522
523
6. Conclusions
During the second (dry) half of the water year, [NO3-] decreased
consistently
along an 11.42 km experimental reach of the Pajaro River by
~30%, at a time when there
was both a significant loss of channel discharge and extensive
surface-subsurface
exchange. The observed reductions in [NO3-] and channel
discharge along this reach
represent an absolute NO3- sink of ~50%, comprising a net
removal rate of 200 - 400
kg/day N-NO3. The associated NO3- uptake length (L[NO3]) and
areal uptake rate (U[NO3])
were 37 ± 13 km and 0.5 μmol m-2 s-1, respectively. These values
did not change
significantly when [NO3-] was normalized to the conservative
solute [Cl-], nor when an
exponential downstream decrease in [NO3-], as opposed to a
linear decrease, was
assumed. These results suggest that the contribution to the
decrease in [NO3-] via dilution
by groundwater is negligible in this system. Furthermore, in
NO3--rich streams with
significant denitrification such as the Pajaro River, the
assumption of a linear
downstream decrease in [NO3-] may be as appropriate as the
assumption of an
exponential decrease.
High-resolution (temporal and spatial) sampling shows that most
of the NO3- loss
occurs along the lower part of the reach, which is also the
stretch along which seepage
loss and hydrologic exchange is most rapid. Pore water profiles
chemical profiles from
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 23
-
524
525
526
527
528
529
530
531
532
533
534
535
536
537
538
539
540
541
542
543
544
545
546
the lower part of the reach suggest that denitrification within
the streambed can occur at
rates consistent with rates derived from downstream changes in
[NO3-] and channel
geometry. Downstream enrichment in 15N- and 18O-NO3- suggests
denitrification is the
primary NO3- sink in the reach during the times studied. We used
an idealized box-model
which assumed stream δ15N-NO3- was controlled by changes in
denitrification efficiency
in the streambed (or anywhere isolated from, but exchanging
water with, the main
channel). Differences in the fraction of NO3- entering the
streambed that is denitrified
could explain variations in εN,surf during the water year: when
discharge is greater,
denitrification is least efficient and isotopic fractionation is
greatest. When discharge is
lower, denitrification appears to be more efficient, resulting
in lower isotopic
fractionation. If NO3- lost via net channel loss of water is
similarly denitrified, this
approach also allows estimates of the fraction of NO3-
potentially recharging aquifers that
is removed. Contemporaneous nitrification, suggested by
enrichment in 18O-NO3- relative
to 15N-NO3-, may lead to an underestimate of gross
denitrification rates within this
system.
7. Acknowledgements
This work was supported by the Committee on Research (UCSC), the
STEPS Institute
(UCSC), Center for Agroecology and Sustainable Food Systems
(UCSC), the United
States Department of Agriculture (projects # 2003-35102-13531
and 2002-34424-11762),
and the National Science Foundation Graduate Fellowship program.
Jonathan Lear and
colleagues with the Pajaro Valley Water Management Agency
assisted with field
logistics and sampling of ground water wells, and numerous land
owners and tenants
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 24
-
547
548
549
550
551
552
553
kindly provided access to monitoring and experimental sites
along the Pajaro River. In
addition, the authors gratefully acknowledge field assistance
provided by Gerhardt Epke,
Remy Nelson, Emily Underwood, Laura Roll, Randy Goetz, Nicole
Alkov, Mike Hutnak,
Claire Phillips, Ari Hollingsworth, Jean Waldbieser, Kena
Fox-Dobbs, Carissa Carter,
Pete Adams, Ty Kennedy-Bowdoin, Andy Shriver, Bowin
Jenkins-Warrick, Robert
Sigler, Heather McCarren, Iris DeSerio, Sora Kim, Heather
McCarren, Greg Stemler, and
Kevin McCoy.
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 25
-
References Alexander, R.B., Smith, R.A. and Schwarz, G.E., 2000.
Effect of stream channel size on
the delivery of nitrogen to the Gulf of Mexico. Nature,
403(6771): 758-761. Bernhardt, E.S. and Likens, G.E., 2002.
Dissolved organic carbon enrichment alters
nitrogen dynamics in a forest stream. Ecology, 83(6): 1689-1700.
Bernhardt, E.S., Likens, G.E., Buso, D.C., and Driscoll, C.T.,
2003. In-stream uptake
dampens effects of major forest disturbance on watershed
nitrogen export. Proceedings of the National Academy of Sciences,
100(18): 10304-10308.
Bohlke, J.-K., 2002. Groundwater recharge and agricultural
contamination. Hydrogeology Journal, 10(1): 153-179.
Bohlke, J.-K. and Denver, J.M., 1995. Combined use of
groundwater dating, chemical, and isotopic analyses to resolve
history and fate of nitrate contamination in two agricultural
watersheds, Atlantic coastal plain, Maryland. Water Resources
Research, 31(9): 2319-2339.
Bohlke, J.-K., Harvey, J.W. and Voytek, M.A., 2004. Reach-scale
isotope tracer experiment to quantify denitrification and related
processes in a nitrate-rich stream, midcontinent United States.
Limnology and Oceanography, 49(3): 821-838.
Bohlke, J.-K., Wanty, R., Tuttle, M., Delin, G. and Landon, M.,
2002. Denitrification in the recharge area and discharge area of a
transient agricultural nitrate plume in a glacial outwash sand
aquifer, Minnesota. Water Resources Research, 38(7).
Brandes, J.A. and Devol, A.H., 1997. Isotopic fractionation of
oxygen and nitrogen in coastal marine sediments. Geochimica Et
Cosmochimica Acta, 61(9): 1793-1801.
Burns, D.A. and Kendall, C., 2002. Analysis of δ15N and δ18O to
differentiate NO3 sources in runoff at two watersheds in the
Catskill Mountains of New York. Water Resources Research, 38(5):
10.1029/2001WR000292.
Butturini, A. and Sabater, F., 1999. Importance of transient
storage zones for ammonium and phosphate retention in a
sandy-bottom Mediterranean stream. Freshwater Biology, 41(3):
593-603.
Cey, E.E., Rudolph, D.L., Aravena, R. and Parkin, G., 1999. Role
of the riparian zone in controlling the distribution and fate of
agricultural nitrogen near a small stream in southern Ontario.
Journal of Contaminant Hydrology, 37: 45-67.
Christensen, P.B., Nielsen, L.P., Sorensen, J. and Revsbech,
N.P., 1990. Denitrification in Nitrate-Rich Streams - Diurnal and
Seasonal-Variation Related to Benthic Oxygen-Metabolism. Limnology
and Oceanography, 35(3): 640-651.
Cirmo, C.P. and McDonnell, J.J., 1997. Linking the hydrologic
and biogeochemical controls of nitrogen transport in near-stream
zones of temerate-forested catchments: a review. Journal of
Hydrology, 199: 88-120.
Dhondt, K., Boeckx, P., Cleemput, O.V. and Hofman, G., 2003.
Quantifying nitrate retention processes in a riparian buffer zone
using the natural abundance of 15N in NO3-. Rapid Communications in
Mass Spectrometry, 17(23): 2597-2604.
Dodds, W.K. et al., 2002. N uptake as a function of
concentration in streams. Journal of the North American
Benthological Society, 21(2): 206-220.
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 1
-
Duff, J.H. and Triska, F.J., 1990. Denitrification in sediments
from the hyporheic zone adjacent to a small forested stream.
Canadian Journal of Fisheries and Aquatic Sciences, 47(6):
1140-1147.
Duff, J.H. and Triska, F.J., 2000. Nitrogen biogeochemistry and
surface-subsurface exchange in streams, Streams and Ground Waters,
pp. 197-220.
Galloway, J.N., Schlesinger, W.H., II, Levy, H., Michaels, A.
and Schnoor, J.L., 1995. Nitrogen fixation: Anthropogenic
enhancement-environmental response. Global Biogeochemical Cycles,
9(2): 235-252.
Gooseff, M.N., McKnight, D.M., Lyons, W.B. and Blum, A.E., 2002.
Weathering reactions and hyporheic exchange controls on stream
water chemistry in a glacial meltwater stream in the McMurdo Dry
Valleys. Water Resources Research, 38(12), 1279,
doi:10.1029/2001WR000834.
Grimaldi, C. and Chaplot, V., 2000. Nitrate depletion during
within-stream transport: Effects of exchange processes between
streamwater, the hyporheic and riparian zones. Water Air and Soil
Pollution, 124(1-2): 95-112.
Haggard, B.E., Stanley, E.H., and Sotrm, D.E., 2005. Nutrient
retention in a point-source-enriched stream. Journal of the North
American Benthological Society, 24(1): 29-47.
Hall, R.O., Bernhardt, E.S. and Likens, G.E., 2002. Relating
nutrient uptake with transient storage in forested mountain
streams. Limnology and Oceanography, 47(1): 255-265.
Hanson, R.T., 2003. Geohydrologic framework of recharge and
seawater intrusion in the Pajaro Valley, Santa Cruz and Monterey
Counties, California. U.S.G.S. Water-Resources Investigations
Report 03-4096.
Hecky, R.E., Campbell, P. and Hendzel, L.L., 1993. The
Stoichiometry of Carbon, Nitrogen, and Phosphorus in Particulate
Matter of Lakes and Oceans. Limnology and Oceanography, 38(4):
709-724.
Hill, A.R., 1979. Denitrification in the nitrogen budget of a
river ecosystem. Nature, 281: 291-292.
Hill, W.R., Mulholland, P.J. and Marzolf, E.R., 2001. Stream
ecosystem responses to forest leaf emergence in spring. Ecology,
82(8): 2306-2319.
Hinkle, S.R. et al., 2001. Linking hyporheic flow and nitrogen
cycling near the Willamette River - a large river in Oregon, USA.
Journal of Hydrology, 244: 157-180.
Holloway, J.M., Dahlgren, R.A., Hansen, B. and Casey, W.H.,
1998. Contribution of bedrock nitrogen to high nitrate
concentrations in stream water. Nature, 395 (6704): 785-788.
Hunt, J.W. et al., 1999. Patterns of aquatic toxicity in an
agriculturally dominated coastal watershed in California.
Agriculture, Ecosystems and Environment, 75: 75-91.
Keeling, R.F., 1995. The Atmospheric Oxygen Cycle - the Oxygen
Isotopes of Atmospheric CO2 and O-2 and the O-2/N-2 Ratio. Reviews
of Geophysics, 33: 1253-1262.
Kellman, L. and Hillaire-Marcel, C., 1998. Nitrate cycling in
streams: using natural abundances of NO3--delta N-15 to measure
in-situ denitrification. Biogeochemistry, 43(3): 273-292.
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 2
-
Kendall, C., 1998. Tracing nitrogen sources and cycling in
catchments. In: C. Kendall and J. McDonnell (Editors), Isotope
tracers in catchment hydrology. Elsevier, Amsterdam, Netherlands,
pp. 519-576.
Laursen, A.E. and Seitzinger, S.P., 2004. Diurnal patterns of
denitrification, oxygen consumption and nitrous oxide production in
rivers measured at the whole-reach scale. Freshwater Biology,
49(11): 1448-1458.
Lehmann, M.F., Reichert, P., S.M. Bernasconi, S.M., Barbieri, A.
and McKenzie, J.A., 2003. Modelling nitrogen and oxygen isotope
fractionation during denitrification in a lacustrine
redox-transition zone. Geochimica et Cosmochimica Acta, 67(14):
2529-2542.
Lehmann, M.F., Sigman, D.M. and Berelson, W.M., 2004. Coupling
the 15N/14N and 18O/16O of nitrate as a constraint on benthic
nitrogen cycling. Marine Chemistry, 88: 1-20
(doi:10.1016/j.marchem.2004.02.001).
Li, Y.-H. and Gregory, S., 1974. Diffusion of ions in sea water
and in deep-sea sediments. Geochimica et Cosmochimica Acta, 38(5):
703-714.
Los Huertos, M., Gentry, L.E. and Shennan, C., 2001. Land use
and stream nitrogen concentrations in agricultural watersheds along
the central coast of California. The Scientific World, 1: 1-8.
Mariotti, A., Landreau, A. and Simon, B., 1988. 15N isotope
biogeochemistry and natural denitrification process in groundwater:
application to the chalk aquifer of northern France. Geochimica et
Cosmochimica Acta, 52(7): 1869-1878.
Marti, E., Grimm, N.B. and Fisher, S.G., 1997. Pre- and
post-flood retention efficiency of nitrogen in a Sonoran Desert
stream. Journal of the North American Benthological Society, 16(4):
805-819.
McKnight, D.M., Hornberger, G.M., Bencala, K.E. and Boyer, E.W.,
2002. In-stream sorption of fulvic acid in an acidic stream: A
stream-scale transport experiment. Water Resources Research, 38(1):
1-12.
McMahon, P.B. and Bohlke, J.K., 1996. Denitrification and mixing
in a stream-aquifer system: effects on NO3- loading to surface
water. Journal of Hydrology, 186: 105-128.
Muir, K.S., 1977. Initial Assessment of the Ground-Water
Resources in the Monterey Bay Region, California, USGS Water
Resources Investigation 77-46.
Mulholland, P.J. and Hill, W.R., 1997. Seasonal patterns in
streamwater nutrient and dissolved organic carbon concentrations:
Separating catchment flow path and in-stream effects. Water
Resources Research, 33(6): 1297-1306.
Mulholland, P.J. et al., 2004. Stream denitrification and total
nitrate uptake rates measured using a field N-15 tracer addition
approach. Limnology and Oceanography, 49(3): 809-820.
Neal, C., and Jarvie, H.P., 2005. Agriculture, community, river
eutrophication and the Water Framework Directive. Hydrological
Processes, 19(9): 1895-1901.
Newbold, J.D., Elwood, J.W., O'Neill, R.V., and VanWinkle, W.,
1981. Measuring nutrient spiralling in streams. Canadian Journal of
Fisheries and Aquatic Sciences, 38: 860-863.
Nolan, B.T., 2001. Relating nitrogen sources and aquifer
susceptibility to nitrate in shallow ground waters of the United
States. Ground Water, 39(2): 290-299.
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 3
-
Nolan, B.T., Ruddy, B., C., Hitt, K.J. and Helsel, D.R., 1997.
Risk of nitrate in groundwaters of the United States - a national
perspective. Environmental Science and Technology, 31(8):
2229-2236.
Peterjohn, W.T. and Correll, D.L., 1984. Nutrient dynamics in an
agricultural watershed: observations on the role of a riparian
forest. Ecology, 65(5): 1466-1475.
PVWMA, 2001. Pajaro Valley Water Management Agency State of the
Basin Report. Ren, J.H. and Packman, A.I., 2004. Modeling of
simultaneous exchange of colloids and
sorbing contaminants between streams and streambeds.
Environmental Science & Technology, 38(10): 2901-2911.
Sabater, F. et al., 2000. Effects of riparian vegetation removal
on nutrient retention in a Mediterranean stream. Journal of the
North American Benthological Society, 19(4): 609-620.
Schade, J.D.e.a., 2001. The Influence of a Riparian Shrub on
Nitrogen Cycling in a Sonoran Desert Stream. Ecology, 82(12):
3363-3376.
Schiff, S.L. et al., 2002. Two adjacent forested catchments:
Dramatically different NO3- export. Water Resources Research,
38(12): 1-13.
Sebilo, M., Billen, G., Grably, M. and Mariotti, A., 2003.
Isotopic composition of nitrate-nitrogen as a marker of riparian
and benthic denitrification at the scale of the whole Seine River
system. Biogeochemistry, 63(1): 35-51.
Sigman, D.M. et al., 2001. A bacterial method for the nitrogen
isotopic analysis of nitrate in seawater and freshwater. Analytical
Chemistry, 73(17): 4145-4153.
Sigman, D.M. et al., 2003. Distinguishing between water column
and sedimentary denitrification in the Santa Barbara Basin using
the stable isotopes of nitrate. Geochemistry Geophysics Geosystems,
4(5), 1040, doi:10.1029/2002GC000384.
Sjodin, A.L., Jr., W.M.L. and III, J.F.S., 1997. Denitrification
as a component of the nitrogen budget for a large plains river.
Biogeochemistry, 39(3): 327-342.
Stream Solute Workshop, 1990. Concepts and methods for assessing
solute dynamics in stream ecosystems. Journal of the North American
Benthological Society, 9(2), 95-119.
Triska, F.J., Kennedy, V.C., Avanzino, R.J., Zellweger, G.W. and
Bencala, K.E., 1989. Retention and transport of nutrients in a
third-order stream in northwestern California: hyporheic processes.
Ecology, 70(6): 1893-1905.
Turner, R.E. and Rabalais, N.N., 1994. Coastal eutrophication
near the Mississippi river delta. Nature, 368(6472): 619-621.
Valett, H.M., Morrice, J.A., Dahm, C.N. and Campana, M.E., 1996.
Parent lithology, surface-groundwater exchange, and nitrate
retention in headwater streams. Limnology and Oceanography, 41(2):
333-345.
Vitousek, P.M. et al., 1997. Human alteration of the global
nitrogen cycle: Sources and consequences. Ecological Applications,
7(3): 737-750.
Wondzell, S.M. and Swanson, F.J., 1996. Seasonal and storm
dynamics of the hyporheic zone of a 4th-order mountain stream .2.
Nitrogen cycling. Journal of the North American Benthological
Society, 15(1): 20-34.
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 4
-
Figure Captions Figure 1. Experimental locations and historical
[NO3-] from the Pajaro River, central coastal California. (A) Map
of the Pajaro River watershed. (B) The 11.42 km experimental reach
of the Pajaro River. All sampling and measurement locations
discussed in this study are identified on the basis of distance
downstream from the top of the reach (boldface numbers, in km). (C)
[NO3-] at Km 0.00 determined since 1952. Figure 2. Chemistry of the
Pajaro River during the end of the 2002-03 water years. Filled
(open) symbols indicate water at the top (bottom) of the reach. (A)
Major cation concentrations: sodium (diamonds), magnesium
(squares), and calcium (circles). (B) Major anions concentrations:
chloride (circles) and sulfate (diamonds). (C) [NO3-] at Km 0.00
(closed circles), Km 8.06 (open diamonds), and Km 11.42 (open
circles). (D) Total dissolved phosphorus (TDP) concentrations at Km
0.00 (closed circles), Km 8.06 (open diamonds), and Km 11.42 (open
circles). (E) NO3- -N flux (in kg/day) at Km 0.00 (closed circles),
Km 8.06 (open diamonds), and Km 11.42 (open circles). Figure 3.
High-resolution (spatial, temporal) records of [NO3-] and soluble
reactive phosphoris ([SRP]) at the top and bottom of stretches. (A)
Concentrations immediately before the 6/15/04 tracer experiment.
(B) Concentrations soon after the 6/17/04 tracer experiment. (C)
Concentrations immediately before the 7/20/04 tracer experiment. No
[SRP] data were available from this period. Figure 4. Pore water
profiles at Km 2.72 on 7/16/03, obtained from peeper (streambed)
samplers. Stable isotope ratios for NO3--N are indicated when
applicable. (A) Peeper from the left side of the channel. The
apparent enrichment in δ15N (εN) is -22‰. (B) Peeper from the
channel center. εN is -5‰. Figure 5. Pore water profiles near Km 8,
obtained from peeper (streambed) samplers. Stable isotope ratios
for NO3--N are indicated when applicable. (A) Peeper from Km 8.06
on 7/9/03, when channel discharge was ~ 0.2 m3/s. (B) Peeper from
Km 8.08 on 7/9/03. The apparent enrichment in δ15N (εN) is -16‰.
(C) Peeper from Km 8.06 on 9/1/04. εN is -16‰. Figure 6. Stable
isotope ratios of NO3-. (A) δ18O vs. δ15N for all samples analyzed.
(B) Magnification of Fig. 6A, showing surface water samples along
with isotopically similar piezometer and peeper samples. Figure 7.
Apparent fractionation of stable isotopes of NO3- in surface water
samples. (A) δ15N vs. ln[NO3-(μm)] for samples collected on seven
days from 7/13/03-8/26/04. The slopes of the best-fit lines are the
apparent enrichments in δ15N-NO3- (εN,surf, in ‰) (B) Same as 10A,
but for δ18O-NO3- and εO,surf. Figure 8. Trends in apparent
enrichment of NO3- stable isotopes in surface water. (A) Enrichment
in 15N (εN,surf) vs. NO3- uptake length (L[NO3]). (B) Ratio of
εN,surf to εO,surf vs. day of the year, for 2003 and 2004
samples.
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 5
-
Figure 9. NO3- uptake lengths (L[NO3]) plotted as a function of
channel discharge. Included are observed trends in inflow lengths
(LI, the stream length required for inflow of tracer-free water to
equal channel discharge), storage exchange lengths (LS, the average
distance traveled by a water molecule before entering an adjacent
storage zone), and channel loss lengths (LL, the distance at which
discharge would equal zero given observed seepage loss). (a)
Injections from Km 0.00. (b) Injections from Km 7.67. Figure 10.
Box model and calculations of surface - subsurface exchange and
associated denitrification. (A) Cartoon showing conceptual
configuration of the box model. Water passes along the main channel
with constant discharge. Denitrification occurs mainly in
subsurface storage zones, with a resulting Rayleigh isotopic
enrichment (εN,sub) of -20‰. This causes a decrease in surface
[NO3-] (∆[NO3-]surf), with an apparent fractionation of 15N-NO3-
(εN,surf). (B) Downstream surface enrichment in 15N-NO3- (curves of
equal εN,surf), assuming a fraction of NO3- entering the reach is
subject to denitrification in the subsurface. Regions corresponding
to low discharge (low-Q) high discharge (high-Q) conditions are
shown. See text for discussion.
Ruehl et al., Nitrate Dynamics in a Losing Stream Page 6
-
Table 1. NO3- uptake lengths, with correlation coefficients,
from late-year synoptic sampling. Date n L[NO3] r2[NO3] L[NO3]:[Cl]
r2[NO3]:[Cl] (km) (km) 6/17/02 3 44 0.99 50 0.99 6/24/02 3 35 0.98
38 0.99 6/30/02 3 31 0.97 32 0.97 7/8/02 3 37 0.99 34 0.99 7/15/02
4 31 0.75 27 0.89 7/22/02 4 39 0.94 37 0.98 7/30/02 4 38 0.99 36
0.97 8/5/02 5 23 0.53 39 0.87 6/29/03 6 34 0.83 38 0.86 7/6/03 6 79
0.28 67 0.51 7/13/03 8 30 0.87 26 0.89 7/24/03 8 36 0.91 55 0.86
8/17/03 11 34 0.45 53 0.25 7/23/04 5 28 0.77 54 0.75
-
Table 2. Summary of stretch-specific nitrate uptake, with
hydrologic length scales (see text and Ruehl et al, in press) when
available. Date injection length Qin Qout LNO3 LIa LSb LLc
(river Km) (m) (m3/s) (m3/s) (km) (km) (km) (km) 8/26/03 7.67
530 0.099 0.058 11 6.6 4.6 1.0 8/28/03 5.79 930 0.22 0.19 34 8.5
0.11 8.4 950 0.19 0.12 N/Cd 2.0 0.28 2.0 9/2/03 2.72 3070 0.19 0.16
330 15.3 1.1 19 9/4/03 0.00 1530 0.2 0.2 N/C 7.7 0.78 N/Ae 1190 0.2
0.19 N/C 23 5/17/04 5.79 1880 0.66 0.58 N/C 3730 0.58 0.49 107
5/19/04 0.00 2720 0.67 0.66 N/C 27 1.93 N/A 3070 0.66 0.61 N/C
6/15/04 7.67 770 0.26 0.26 19 9.6 1.4 N/A 1400 25 6/17/04 0.00 2720
0.32 0.29 N/C 28 7/20/04 7.67 650 0.13 0.11 -10 7.2 1.3 3.9 1520 22
7/22/04 0.00 1530 0.21 0.21 N/C 8.1 0.65 N/A 1190 0.21 0.21 N/C
8/31/04 7.67 770 0.28 0.27 N/C 7.0 0.34 21 9/2/04 0.00 1530 0.37
0.38 N/C 27 1.0 N/A 1190 0.38 0.33 N/C a Lateral inflow length, the
stream length required for inflow of tracer-free water to equal
channel discharge b Storage exchange length, the average distance
traveled by channel water before entering a storage zone. c Channel
loss length, the distance at which discharge would reach zero given
observed seepage loss. d Uptake lengths were not calculated when
downstream changes in [NO3-] were within the precision of the
analytical instrument (
-
0 .5 1 km
8.06
Paja
ro R
.
N
11.420.00
9.84
8.208.44
400' 400'
1.53
5.79
2.72
7.67
San Andreas
Fault Zone
1200'
400'
400'
400'
400'
55'
0.5
1.0
1.5
[NO
3-N
] (m
M)
1960 1970 1980 1990 2000Date
Drinking water standard
B
36°5
4' N
800'
00 25 mi
40 km
121º30' W 121º00'
Maparea
watershed boundaryN
inset
SalinasMonterey
San Andreas Fault Zone
37º0
0'
Watsonville
Gilroy
36º3
0' N
A
MontereyBay
Hollister
Pacific Ocean
37'121°39' W
C
Figure 1
-
2
3
4
5
6
6/02 8/02 6/03 8/03
200
400
600
NO
3- fl
ux (k
g N
/day
)[T
DP
] (µM
)
4
2
6
8
0.6
0.8
1.0
1.2
[NO
3-]
(mM
)[C
l- ], [
SO
42- ]
(mM
)2
4
6
8
10
[Na+
], [C
a2+ ],
[Mg2
+ ] (m
M) A
B
C
D
E
Na+
Ca2+
Mg2+
Cl-
SO42-
closed = km 0.00open = km 11.42
closed = km 0.00open = km 11.42
Km 8.06Km 0.00
Km 11.42
Km 0.00
Km 8.06
Km 11.42
Km 0.00
Km 8.06
Km 11.42
Drinking water standard
Figure 2
-
1.0
1.1
5
10
15 [SRP
] (µM)
6/12/04 6/13/046/11/04
Km 7.67Km 8.44
Km 9.84
Km 9.84
Km 8.44 Km 7.67
Km 0
Km 2.72
Km 1.54
Km 0Km 1.54
Km 2.72
6/25/04 6/26/046/24/04
7/19/04 7/20/047/18/04
Km 9.84
Km 8.44
Km 7.67
A
B
C
4
6
81.1
1.2
1.3
0.5
0.6
[SR
P-] ( µM
)
[NO
3-] (
mM
)
Date
[NO
3-] (
mM
)[N
O3-] (
mM
)
Figure 3
-
-20
-15
-10
-5
0
0 1 2 3 4
-16
-12
-8
-4
0
4El
evat
ion
(cm
)El
evat
ion
(cm
)
streambed
streambed
Concentration (mM, except [Mn] in 10-5 M)
[Cl-][SO4
2-][NO3
-][NO2
-][NH4
+][Mn]
14‰
24‰
δ15N
A
B
41‰
16‰
24‰
Figure 4
-
-40
-30
-20
-10
0
-40
-30
-20
-10
0
0 1 2 3 4
-30
-20
-10
0
streambed
streambed
streambed
Elev
atio
n (c
m)
Elev
atio
n (c
m)
Elev
atio
n (c
m)
A
B
C
15‰
20‰
14‰
28‰
20‰
Figure 5
Concentration (mM, except [Mn] in 10-5 M)
[Cl-][SO4
2-][NO3
-][NO2
-][NH4
+][Mn]δ15N24‰
-
Pajaro R.PiezometerPeeperWell
δ18 O
(‰)
A
10
20
10 20 30 40
10
12
12 14 16
B
δ18 O
(‰)
δ15N (‰)
δ15N (‰)
B
denitrif
ication
denitrific
ation
Figure 6
-
12
14
16
6.2 6.6 7ln[NO3
- (µm)]
19.9‰
δ15 N
(‰)
-8.2‰ -9.0‰
-7.6‰
-6.0‰-7.7‰
-17.3‰
7/13/0310/5/035/21/046/4/04
6/20/047/21/048/26/04
-20.0‰
-16.3‰ -9.7‰ -1.6‰
-3.5‰-4.0‰
-6.4‰
δ18 O
(‰)
8
10
12
A
B
Figure 7
-
-20
-10
-5
10 100L[NO3] (km)
ε N,s
urf (
‰)
-15
June/1 Aug/1 Oct/1Day of the year
20032004
(εN /
ε O) su
rf
A
B
denitrification
20032004
0.1
1
10
low Q
high Q
Figure 8
-
Leng
th s
cale
(km
)
LS
LI
Qin (m3 s-1)
LI
LS
A
B
1
10
0.2 0.3 0.4 0.5 0.6
1
10
0.1 0.15 0.2 0.25
Leng
th s
cale
(km
)
mor
e se
epag
em
ore
seep
age
LLln L[NO3] = 1.8 +
4.6Qin
L[NO3]LL
Figure 9
-
εsurf= ?
10%
20%
30%
40%
NO3- entering subsurface
[NO
3-] r
educ
tion
alon
g re
ach
εN,surf = 0‰-6‰
-10‰
-14‰
-16‰
-18‰
-19‰
10% 20% 30% 40%
low Q
high Q
Main channel
Subsurface NO3- flux in
εN,surf= ?
NO3- flux out
A
B
Qin[NO3
-]inδ15Nin
Qout[NO3
-]outδ15Nout
∆[NO3-]surf
∆δ15Nsurf
∆[NO3-]sub
∆δ15NsubεN,sub= -20‰
Figure 10