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
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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,
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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
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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-,
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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
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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
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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
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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
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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.
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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
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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
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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
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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):
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033
01515
]ln[]ln[ −− −−
=NONONN
Nδδ
ε (3) 233
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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
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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).
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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
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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).
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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
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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
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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.
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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).
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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
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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.
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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).
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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).
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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 ∆
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δ459 15Nsurf, respectively)
and the subsurface isotopic enrichment factor for ∆δ460
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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.
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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
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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.
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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
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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
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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
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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