Nitrate dynamics within the Pajaro River,afisher/CVpubs/pubs/Ruehl... · m3/s) and higher solute...

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


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.


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


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


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


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-


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


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-.


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


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


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


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.


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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.


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.


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

Nitrate dynamics within the Pajaro River,afisher/CVpubs/pubs/Ruehl... · m3/s) and higher solute...

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Transcript of Nitrate dynamics within the Pajaro River,afisher/CVpubs/pubs/Ruehl... · m3/s) and higher solute...