Leaching of nitrate (NO3-) from animal waste or fertilisers at
agricultural operations can result in NO3- contamination of
groundwater, lakes, and streams. Understanding the sources and fate of
nitrate in groundwater systems in glacial sediments, which underlie many
agricultural operations, is critical for managing impacts of human food
production on the environment. Elevated NO3- concentrations in
groundwater can be naturally attenuated through mixing or denitrification.
Here we use isotopic enrichment of the stable isotope values of
NO3- to quantify the amount of denitrification in groundwater at
two confined feeding operations overlying glacial sediments in Alberta,
Canada. Uncertainty in δ15NNO3 and
δ18ONO3 values of the NO3- source and
denitrification enrichment factors are accounted for using a Monte Carlo
approach. When denitrification could be quantified, we used these values to
constrain a mixing model based on NO3- and Cl-
concentrations. Using this novel approach we were able to reconstruct the
initial NO3-N concentration and NO3-N/Cl- ratio at the
point of entry to the groundwater system. Manure filtrate had
total nitrogen (TN) of up to 1820 mg L-1, which was predominantly
organic N and NH3. Groundwater had up to 85 mg L-1 TN, which
was predominantly NO3-. The addition of NO3- to the
local groundwater system from temporary manure piles and pens equalled or
exceeded NO3- additions from earthen manure storages at these
sites. On-farm management of manure waste should therefore increasingly focus
on limiting manure piles in direct contact with the soil and encourage
storage in lined lagoons. Nitrate attenuation at both sites is attributed to
a spatially variable combination of mixing and denitrification, but is
dominated by denitrification. Where identified, denitrification reduced
agriculturally derived NO3- concentrations by at least half and,
in some wells, completely. Infiltration to groundwater systems in glacial
sediments where NO3- can be naturally attenuated is likely
preferable to off-farm export via runoff or drainage networks, especially if
local groundwater is not used for potable water supply.
Introduction
The contamination of soil and groundwater with nitrate from agricultural
operations is a global water quality issue that has been extensively
documented (Power and Schepers, 1989; Spalding and Exner, 1993; Rodvang and
Simpkins, 2001; Galloway et al., 2008; Zirkle et al., 2016; Arauzo, 2017;
Ascott et al., 2017). Leaching of nitrate (NO3-) from animal waste
or fertilisers can result in groundwater NO3- concentrations
that exceed drinking water guidelines and pose human health risks (Fan and
Steinberg, 1996; Gulis et al., 2002; Yang et al., 2007). The discharge of
high-NO3- groundwater, runoff, or drainage can contaminate
streams and lakes, resulting in eutrophication and ecosystem decline
(Deutsch et al., 2006; Kaushal et al., 2011). In saturated groundwater
systems with low oxygen concentrations, elevated NO3- can be
naturally attenuated by microbial denitrification (Wassenaar, 1995;
Robertson et al., 1996; Smith et al., 1996; Tesoriero et al., 2000;
Singleton et al., 2007). Concentrations of NO3- will also decrease
along groundwater flow paths due to attenuation via dilution by hydrodynamic
dispersion (referred to hereafter as mixing). Because of these natural
attenuation mechanisms, infiltration to groundwater may be preferable to
off-site drainage and runoff of nitrate-rich waters. Many agricultural
operations are undertaken on fertile soils associated with glacial sediments
(Spalding and Exner, 1993; Ernstsen et al., 2015; Zirkle et al., 2016).
Understanding the sources and fate of agriculturally derived nitrate in
groundwater systems in glacial sediments is therefore critical for managing
impacts of human food production on the environment.
Identification of the sources and fate of NO3- at agricultural
operations can be challenging because of spatial and temporal variations in
sources (e.g. earthen manure storage, temporary manure piles, or fertiliser)
and heterogeneity in hydrogeologic systems (Spalding and Exner, 1993;
Rodvang et al., 2004; Showers et al., 2008; Kohn et al., 2016). These
spatial and temporal variations can result in complex subsurface solute
distributions that are difficult to interpret using classical transect
studies or numerical groundwater models (Green et al., 2010; Baily et al., 2011).
Groundwater containing significant agriculturally derived NO3-
also typically has elevated chloride (Cl-) concentrations (Saffigna and
Keeney, 1977; Rodvang et al., 2004; Menció et al., 2016). Decreasing
NO3-N/Cl- (or NO3-/Cl-) ratios have been used to
define denitrification based on the assumption that NO3- is
reactive while Cl- is non-reactive (conservative), such that
denitrification results in a decrease in the NO3-N/Cl- ratio
(Kimble et al., 1972; Weil et al., 1990; Liu et al., 2006; McCallum et al.,
2008). However, NO3N/Cl- ratios can also change in response to
mixing of groundwater with different NO3-N/Cl- ratios or when
groundwater sampling traverses hydraulically disconnected formations (Bourke
et al., 2015b). If NO3-N/Cl- ratios vary among potential
sources and the NO3-N/Cl- ratio at the point of entry to the
groundwater system can be reconstructed, this information could be used to
show that anthropogenic NO3- at different locations within an
aquifer is derived from the same or different sources.
The stable isotopes of NO3- (δ15NNO3 and
δ18ONO3) provide an alternative approach to characterising the
source and fate of NO3- in groundwater systems. In agricultural
areas, multiple sources of NO3- are common and could include
precipitation, soil NO3-, inorganic fertiliser, manure, and septic
waste (Komor and Anderson, 1993; Liu et al., 2006; Pastén-Zapata et al.,
2014; Clague et al., 2015; Xu et al., 2015). While source identification is
theoretically possible using δ15NNO3 and
δ18ONO3 (particularly with a dual-isotope approach), in practice
this can be difficult due to geologic heterogeneity, overlapping source
values, and the complexity of biologically mediated reactions (Aravena et
al., 1993; Wassenaar, 1995; Mengis et al., 2001; Choi et al., 2003; Granger
et al., 2008; Vavilin and Rytov, 2015; Xu et al., 2015).
NO3- attenuation by denitrification in groundwater systems can be
identified based on the characteristic enrichment of δ15NNO3
and δ18ONO3. Numerous studies have made
qualitative assessments that identified denitrification in groundwater using
the stable isotope approach (Böttcher et al., 1990; Wassenaar, 1995;
Singleton et al., 2007; Baily et al., 2011; Clague et al., 2015; Xu et al.,
2015). Recently published papers have also used stable isotopic values of
NO3- and water as the basis for mixing models in agricultural
settings (Ji et al., 2017; Lentz and Lehersch, 2019). Isotopic fractionation
effects can also allow for quantitative assessment of the proportion of
substrate that has undergone a given reaction, if enrichment factors and
source values are known; as in the case of evaporative loss of water, for
example (Dogramaci et al., 2012). To date, there have been very few attempts
to quantify denitrification using dual-isotope enrichment, largely due to
uncertainty in source values and enrichment factors (Böttcher et al.,
1990, Xue et al., 2009).
The only published calculations of the fraction of NO3- remaining
after denitrification the that we are aware of assumed a constant enrichment
factor and the same isotopic source values across the field site (Otero et
al., 2009). However, the enrichment factor will vary across a field site in
response to reaction rates (Kendall and Aravena, 2000), and isotopic values
of even the same type of source (e.g. manure) can vary substantially (Xue et al., 2009).
If the variation in source values and enrichment factors can be characterised
from measured data then these uncertainties can be accounted for using a
Monte Carlo approach (Joerin et al., 2002; Bourke et al., 2015a; Ji et al.,
2017), thereby extending the application of the dual-isotope technique to
allow for a robust quantitative assessment of denitrification in
agricultural settings.
A synthesised analysis of stable isotopes of NO3- with additional
ionic tracers can further improve the assessment of NO3-
attenuation mechanisms and sources of NO3-in agricultural
settings (Showers et al., 2008; Vitòria et al., 2008; Xue et al., 2009;
Xu et al., 2015; Ji et al., 2017). We hypothesise that if the amount of
denitrification can be quantified based on δ15NNO3 and
δ18ONO3, then this estimate of the fraction of NO3-N
removed through denitrification can be used to constrain a mixing model
based on NO3-N and Cl- concentrations. This novel approach allows
for the ratio of NO3-N/Cl- at the point of entry to the
groundwater system to be reconstructed from measured NO3- and
Cl- concentrations (see Sect. 2.4). Where the NO3-N/Cl-
ratio varies between sources, this ratio can then be used to assess the
source of the NO3- in groundwater (e.g. temporary manure
piles or feeding pens). These data can also then be used to estimate the
initial concentrations of NO3- and Cl- at the point of entry
to the groundwater system and quantify attenuation by mixing.
Map of study sites CFO1 and CFO4, showing locations of groundwater
monitoring wells, core collection, earthen manure storages (EMSs), dairy and
feedlot pens, manure piles, and irrigated land. Blue rectangle indicates extent
of CFO1 inset.
In this study, we present the application of this approach at two confined
feeding operations (CFOs) in Alberta, Canada, with differing lithologies and
durations of operation (Fig. 1). Concentrations of Cl- and nitrogen
species (N species) and the stable isotopes of NO3- were measured
in groundwater samples collected from monitoring wells and continuous soil
cores, as well as manure filtrate at both sites. These data were interpreted
to (1) assess the extent of agriculturally derived NO3- in
groundwater, (2) identify sources and initial concentrations of
NO3- at the point of entry to the groundwater system, and
(3) assess mixing and denitrification as attenuation mechanisms at these sites.
Materials and methodsExperimental sites
This study was conducted using data from two of the five sites investigated
by Alberta Agriculture and Forestry during an assessment of the impacts of
livestock manure on groundwater quality (Lorenz et al., 2014). To the best
of our knowledge (including discussions with farm operators) fertilisers
have not been applied at either of these sites. As such, manure waste from
livestock is assumed to be the sole source of agricultural nitrogen (N) and
elevated NO3- concentrations in groundwater at these sites.
The first study site (CFO1) is located 25 km northeast of Lethbridge,
Alberta (Fig. 1). Agricultural operations at this site were initiated with
the construction of a dairy in 1928, which has the capacity for 150 dairy
cattle. A feedlot for beef cattle was added in 1960s along with an earthen
manure storage (EMS) facility for storing liquid dairy manure (approx. 4 m
deep) and a catch basin that receives surface water runoff. This feedlot was
expanded in the 1980s to the 2000-head capacity it was at the time of this
study. There is also a dugout (or slough, a shallow wetland) on-site that
receives local runoff and an irrigation drainage canal at the southern
boundary of the property.
The second study site (CFO4) is located approximately 30 km north of Red
Deer, Alberta, and 300 km north of CFO1. This dairy and associated EMS
(approx. 6 m deep) were constructed in 1995 and the facility had 350 head of
dairy cattle at the time of the study. Runoff will drain either to the small
dugout in the northwest of the site, or the natural drainage features
(ephemeral ponds or a creek approx. 1.5 km east).
Sampling and instrumentationGroundwater monitoring wells
Groundwater samples were collected from water table wells and piezometers
(hereafter both are referred to as wells) installed at both sites (Table 1).
At CFO1, groundwater samples were collected from six individual water table
wells (DMW1, DMW2, DMW3, DMW4, DMW5, DMW6) and eight sets of nested wells
with one well screened at the water table and one well screened 20 m below
ground (BG) (DP10-2 and DP10-1, DMW10 and DP11-10b, DMW11 and DP11-11b,
DMW12 and DP11-12b, DMW13 and DP11-13b, DMW14 and DP11-14b, DMW15 and
DP11-15b, and DMW16 and DP11-16b). Wells DP10-2 and DP10-1 were located
directly adjacent to the EMS on the hydraulically down-gradient side. At
CFO4, groundwater samples were collected from eight water table wells (BC1,
BC2, BC3, BC4, BC5, BMW1, BMW3, BMW7) and four sets of nested wells, with
wells screened across the water table and at 15 m BG. Two of these nests
were located adjacent to the EMS (BMW2 and BP10-15e, BMW4 and BP10-15w) and
two were hydraulically down-gradient of the EMS (BMW5 and BP5-15, BMW6 and BP6-15).
Details of groundwater monitoring wells and continuous core collection
at CFO1 and CFO4 (all screens installed at bottom of the well).
SiteWell/coreTypeaLateralGroundTotalScreenLithology ofK (m s-1)hole IDdistanceelevationdepth (mlengthscreened intervalfrom(m a.s.l.)below(m)EMSb (m)ground)CFO1DMW1WTW60869.75.04.0SandDMW2WTW10867.26.04.0Sand1.2×10-7DMW3WTW2867.53.72.0SandDMW4WTW1604.24Sand1.3×10-6DMW5WTW270866.46.84.0Clayey sand1.7×10-5DMW6WTW3106.74DP10-1Piezo2867.818.60.5Clay1.6×10-9DP10-2Piezo2867.98.01.5Sand3.6×10-5DMW10WTW340868.07.23.0Clay3.0×10-7DP11-10bPiezo340868.0200.5Clay2.2×10-8DMW11WTW470864.87.03.0Sand and clay4.2×10-5DP11-11bPiezo470200.5Clay6.3×10-9DMW12WTW50867.67.03.0Sand and clay7.4×10-6DP11-12bPiezo50867.620.11.0Clay1.1×10-8DMW13WTW35867.17.03.0Sand8.9×10-6DP11-13bPiezo + core35867.120.00.5ClayDMW14WTW105865.77.03.0Clay5.7×10-6DP11-14bPiezo105865.720.00.5Sand1.1×10-6DMW15WTW1857.03Clay2.4×10-8DP11-15bPiezo18520.00.5Clay1.4×10-7DMW16WTW320866.06.03.0Sand and clay–DP11-16bPiezo32020.00.5Clay3.2×10-9DC15-20Core7615DC15-21Core4510.5DC15-22Core2212DC15-23Core915CFO4BC1WTW110857.06.93.1Clay and sandstoneBC2WTW365859.47.03.1Clay and sandstone2.2×10-7BC3WTW145858.66.83.1Clay and sandstone1.3×10-6BC4WTW95858.85.93.0Clay and sandstone3.4×10-6BC5WTW105859.57.54.5Clay and sandstoneBMW1WTW4858.67.13.1Clay and sandstone4.3×10-6BMW2WTW3857.97.54.5Clay and sandstone8.5×10-7BMW3WTW8858.66.03.0Clay and sandstoneBMW4WTW14858.07.54.8Clay and sandstone1.0×10-5BMW5WTW60858.07.54.5Clay and sandstoneBP5-15Piezo60858.115.31.5Sandstone1.0×10-7BMW6WTW150856.97.54.5Clay and sandstone4.0×10-6BP6-15Piezo150856.815.21.5Sandstone3.0×10-6BMW7WTW140856.77.54.5Clay and sandstone1.0×10-6BP10-15ePiezo4858.214.91.5Sandstone2.9×10-5BP10-15wPiezo10858.015.01.5Sandstone1.0×10-5
a WTW: water table well, Piezo: piezometer,
Core: continuous core. b EMS: earthen manure storage.
Groundwater samples were collected for ion analysis (Cl- and N species)
quarterly between April 2010 and August 2015. All water samples were
collected using a bailer after purging (1–3 casing volumes) and stored at
≤4∘C prior to analysis. Samples for
δ15NNO3 and δ18ONO3 were collected from wells at
CFO1 on 1 January and 1 May 2013. Samples for δ15NNO3
and δ18ONO3 at CFO4 were collected on 27 October 2014.
Wells were purged prior to sample collection (1–3 casing volumes), and
samples filtered into high-density polyethylene (HDPE) bottles in the field
and frozen until analysis.
Hydraulic heads in monitoring wells were determined using manual
measurements (approximately monthly, 2010–2015). Hydraulic head response
tests were conducted on the majority of the wells at the sites to determine
hydraulic conductivity (K) of the formation media surrounding the intake
zone. These tests were either a slug test (water level decline after water
addition) or bail test (water level recovery after water removal) depending
on the location of the water level within the well at the time of testing.
K was determined from the hydraulic head responses using the method of Hvorslev (1951).
Continuous core
A continuous core was collected at CFO1 immediately adjacent to well DP11-13b
on 1 May 2013 (Fig. 1). Additional core samples were collected from 1 to
5 June 2015 along a transect hydraulically down-gradient of the southeastern
side of the EMS at CFO1, where hydrochemistry data suggested leakage from the
EMS (see Sect. 3). During this 2015 drilling campaign, core samples were
collected at four locations (DC15-20, DC15-21, DC15-22, DC15–23) to
depths of up to 15 m below surface and distances of up to 100 m from the EMS
between wells DMW3 and DP11–14.
Continuous core samples were retrieved using a hollow stem auger (1.5 m core
lengths) with 0.3 m sub-samples collected at approximately 1 m intervals
ensuring that visually consistent lithology could be sampled. Core samples
for Cl- were stored in ZiplocTM bags and kept cool until analysis.
Core samples for N-species analysis were stored in Ziploc bags filled
with an atmosphere of argon (99.9 % Ar) to minimise oxidation and kept
cool until analysis. Subsamples of each core (250–300 g) were placed under
50 MPa pressure in a Carver Auto Series NE mechanical press with a 0.5 µm
filter placed at the base of the squeezing chamber, which was placed within
an Ar atmosphere to minimise oxidation. A syringe was attached to the base
of the apparatus and 15 mL of filtered pore water were collected for
analyses within 3.5 to 6.0 h (Hendry et al., 2013).
Liquid manure storages
Samples of liquid manure slurry were collected directly from the EMS at both
sites and the catch basin (containing local runoff from the feedlot) at CFO1
using a pipe and plunger apparatus to sample from approximately 0.5 m below
the surface. The slurry collected was subsequently filtered (0.45 µm)
to separate the liquid and solid components. The water filtered from samples
collected from the EMS or catch basin is hereafter referred to as manure filtrate.
Laboratory analysis
Groundwater samples from wells were analysed by Alberta Agriculture and
Forestry (Lethbridge, Alberta). Concentrations of Cl- were determined
using potentiometric titration of H2O, with a detection limit of
5.0 mg L-1 and accuracy of 5 % (APHA 4500-Cl-D). Concentrations of
NH3 as N (NH3-N), NO3- as N (NO3-N),
and NO2- as N (NO2-N) were measured by air-segmented
continuous-flow analysis (APHA 4500-NH3 G, APHA 4500-NO3 F). Total nitrogen (TN)
was determined by high temperature catalytic combustion and
chemiluminescence detection using a Shimadzu TOC-V with attached TN unit
(ASTM D8083-16). Total organic nitrogen (TON) was calculated by subtracting
NH3-N, NO3-N, and NO2-N from TN. Bicarbonate
(HCO3-) was analysed by titration (APHA 2320 B). Dissolved organic
carbon (DOC) was analysed by a combustion infrared method (APHA 5310 B)
using a Shimadzu TOC-V system. Manure filtrate was analysed by ALS
(Saskatoon, Saskatchewan) using similar methods for Cl- (APHA 4110 B),
TN (RMMA A3769 3.3), NO3+NO2 as N (APHA 4500-NO3-F),
NH3-N (APHA 4500-NH3 D), HCO3- (APHA 2320), and DOC (APHA 5310 B).
Pore-water samples squeezed from the continuous core were analysed at the
University of Saskatchewan (Saskatoon, Canada) for Cl-, NO3-N, and
NO2-N using a Dionex IC25 ion chromatograph (IC) coupled to a Dionex
As50 autosampler (EPA Method 300.1, accuracy and precision of 5.0 %)
(Hautman and Munch, 1997). Ammonia as N (NH3-N) was measured by Exova
laboratories using the automated phenate method (APHA Standard 4500-NH3 G,
detection limit of 0.025 mg L-1, accuracy of 2 % of the measured
concentration, and a precision of 5 % of the measured concentration).
δ15NNO3 and δ18ONO3 in groundwater samples
(from wells and pore water from the continuous core) and manure filtrate were
measured at the University of Calgary (Calgary, Alberta) using the
denitrifier method (Sigman et al., 2001) with an accuracy and precision of
0.3 ‰ for δ15NNO3 and
0.7 ‰ for δ18ONO3. Groundwater samples
collected for NO3- isotope analysis in January 2013 were also
analysed for NO3-N by the University of Calgary (denitrifier
technique, Delta + XL).
Modelling approachQuantification of denitrification based on δ15NNO3 and δ18ONO3
Nitrate in groundwater that has undergone denitrification is commonly
reported as being identified by enrichment of δ15NNO3 and
δ18ONO3 with a slope of about 0.5 on a cross-plot (Clark
and Fritz, 1997). However, published studies of denitrification in
groundwater report slopes of up to 0.77 (Mengis et al., 1999; Fukada et al.,
2003; Singleton et al., 2007). The relationship between isotopic enrichment
of 15NNO3 and 18ONO3 and the fraction of
NO3-N remaining during denitrification can be described by a Rayleigh equation:
R=R0fd1β-1,
where R0 is the initial isotope ratio (relative to the standard) of the
NO3- (δ18ONO3 or δ15NNO3), R is
the isotopic ratio when fraction fd of NO3- remains, and
β is the kinetic fractionation factor (>1) (Böttcher
et al., 1990; Clark and Fritz, 1997; Otero et al., 2009; Xue et al., 2009).
Kinetic fraction effects are commonly also expressed as the enrichment
factor, ε=11000(β-1) . In the
case of a constant enrichment factor, fd can be calculated from measured
δ15NNO3 (or δ18ONO3), if the initial
δ15NNO3 (δ15N0) is known;
fd=expδ15NNO3-δ15N0ε.
The fraction of NO3-N removed from groundwater through denitrification
is then given by (1-fd). The concentration of NO3-N that would have
been measured if mixing was the only attenuation mechanism
(NO3-Nmix) can also be calculated by dividing the measured
concentration by fd.
A subset of 20 samples with isotopic values of NO3- indicative of
denitrification were identified, and for each of these samples fd (mean
and standard deviation) was calculated from Eq. (2) using a Monte Carlo
approach with 500 realizations. The distribution of ε values
was defined based on measured data. If the initial δ15NNO3
is known, ε for δ15NNO3 (ε15N)
can be determined from the slope of the linear regression line on
a plot of ln(fd) vs. δ15NNO3 (Böttcher et al.,
1990). If the initial δ15NNO3 and fd are not known, as
is the case here, ε15N can be determined from the slope of
the regression line on a plot of ln(NO3-N) vs. δ15NNO3,
which will be the same as on a plot of ln(fd) vs. δ15NNO3.
In situ variations in temperature and reaction
rates may affect the enrichment factor (Kendall and Aravena, 2000) and this
was accounted for by allowing for variation in ε15N within
the Monte Carlo analysis. The enrichment factor for δ18ONO3
(ε18O) was calculated by multiplying the
δ15NNO3 by a linear coefficient of proportionality
determined for each CFO from the slope of the denitrification trend on an
isotope cross-plot (see Sect. 3.2).
For each realisation, initial isotopic values (δ15N0 and
δ18O0) were determined by Excel Solver such that the
difference between fd calculated from δ15NNO3 and
δ18ONO3 was minimised (<1 % difference). The
ranges of δ15N0 and δ18O0 were limited
based on measured data and literature values (see Sect. 3.2). This approach
neglects the effect of mixing of groundwater with differing isotopic values
and is valid if the concentration of NO3- in the source is much
greater than background concentrations such that the isotopic composition of
NO3- is dominated by the agriculturally derived end-member.
Quantification of mixing and initial concentrations of Cl- and NO3-N
A binary mixing model that also accounts for decreasing NO3-N
concentrations in response to denitrification was used to quantify
NO3- attenuation by mixing and estimate the initial concentrations
of Cl- and NO3-N. The measured concentration of Cl- was
assumed to be a function of two end-members mixing, described by
Cl=fmCli+1-fmClb,
where Cl is the measured concentration of Cl- in the groundwater
sample, Cli is the concentration of Cl- at the initial point of
entry of the agriculturally derived NO3- to the groundwater
system, Clb is the concentration of Cl- in the background ambient
groundwater, and fm is the fraction of water in the sample from the
source of agriculturally derived Cl- (and NO3-)
remaining in the mixture.
The concentration of NO3-N was also assumed to be a function of two
end-members mixing but with an additional coefficient, fd (the fraction
of NO3-N remaining after denitrification), applied to account for
denitrification. The measured NO3-N concentration was thus described by
NO3-N=fdfmNO3-Ni+1-fmNO3-Nb,
where NO3-N is the concentration of NO3-N measured in the groundwater
sample, NO3-Ni is the concentration of NO3-N in the source of
agriculturally derived NO3- at the initial point of entry to the
groundwater system, and NO3-Nb is the concentration of NO3-N in
the background ambient groundwater. This mixing calculation was only
conducted on samples for which NO3- dominated total-N
(NH3-N< 10 % of NO3-N) so that nitrification of
NH3 could be neglected.
If Cli is much greater than Clb and NO3-Ni is much
greater than NO3-Nb, then fm is insensitive to background
concentrations and these terms can be neglected (see Sect. 4.2 for further
discussion of this assumption). In this case, Eqs. (3) and (4) reduce to
Cl=fmCli,NO3-N=fdfmNO3-Ni.
Solving Eq. (6) for fm and substituting into Eq. (5) yields
NO3-NiCli=1fdNO3-NCl.
Thus, for each groundwater sample, the ratio of NO3-N/Cl- at
the initial point of entry of the agriculturally derived NO3- to
the groundwater system NO3-NiCli can be simply
calculated using measured concentrations, and fd estimated from
NO3- isotope data. This provides a relatively simple method to
identify agriculturally derived NO3- from different sources (e.g.
EMS vs. manure piles) if they have different NO3-N/Cl- ratios.
Estimated Cli and NO3-Ni are reported as the mid-range value with
uncertainty described by the minimum and maximum values. These initial
concentrations are at the water table for top-down inputs, or at the
saturated point of contact between the EMS and the aquifer for leakage from
the EMS. This analysis assumes that a sampled water parcel consists of water
with agriculturally derived NO3- that entered the aquifer from one
source at one point in time and space and has since mixed with natural
ambient groundwater. Any NO3- produced during nitrification after
the anthropogenic source water enters the aquifer is implicitly included in
NO3-Ni. The error in
NO3-NiCli- was assumed to be
dominated by error in the estimated fd, with the measurement error in
NO3-N and Cl- considered negligible.
The initial concentrations of the agriculturally derived NO3-
source (NO3-Ni and Cli) were estimated by simultaneously solving
Eqs. (5) and (6) using Excel Solver (GRG nonlinear). The absolute minimum
values of NO3-Ni and Cli were defined by measured
concentrations (e.g. if Cli=Cl, fm=1). Maximum
values of NO3-Ni and Cli were defined based on measured
concentrations of NO3-N and Cl- in groundwater and manure filtrate
(NO3-N≤ 150 mg L-1 and Cl-≤ 1300 mg L-1; see
Sect. 3.2). These maximum values of NO3-Ni and Cli correspond
to the minimum fm. The value of fd was assumed to be the
mean fd estimated from NO3- isotopes using Eq. (2), and
NO3-NiCli was required to be
within 1 standard deviation of the estimate from Eq. (7).
The resulting estimates of fm are reported as the mid-range, with
uncertainty described by the minimum and maximum values. Larger values
of fm indicate less mixing (a shorter path for advection–dispersion) and
suggest a source close to the well. Smaller values of fm indicate
extensive mixing (a longer path for advection–dispersion) and suggest a
source further away from the well. The relative contributions of mixing and
denitrification to NO3- attenuation at each site were evaluated by
comparing fm and fd for each sample. This analysis was conducted
using isotope values from the samples collected on 1 May 2013 at CFO1, which
were combined with the Cl- and NO3-N data from 6 June 2013. At
CFO4, results from stable isotopes collected on 27 October 2014 were
combined with Cl- and NO3-N data collected on 7 October 2014.
ResultsSite hydrogeologyCFO1
The geology at CFO1 consists of clay and clay–till interspersed with sand
layers of varying thickness to the maximum depth of investigation (20 m BG,
bedrock not encountered). Hydraulic conductivities (K) calculated from slug
tests on wells ranged from 1.2×10-7 to 4.2×10-5 m s-1
(n=10) for sand, 1.1×10-8 to
2.8×10-8 m s-1 (n=2) for clay–till, and
1.6×10-9 to 3.0×10-7 m s-1 (n=8) for
clay. Depth to the water table throughout the study site ranged from 0.5 m
at DMW14 to 3.8 m at DMW11. Seasonal water table variations were about 0.5 m
with no obvious change in the annual average during the 6-year measurement
period. Water table elevation was highest at DMW10 and DMW1 on the west side
of the site and lowest at DMW11 on the northeast side of the site (see
Supplement). Measured heads indicate groundwater flow from the
vicinity of the EMS to the northeast and southeast. Mean horizontal
hydraulic gradients at the water table ranged from 4.4×10-3
to 1.4×10-2 m m-1. Vertical gradients were predominantly
downward in the upper 20 m of the profile (mean gradients ranging from
1.8×10-3 to 0.18 m m-1), with the
exception of DMW11 where the vertical gradient was upward (mean gradient
-2.8×10-2 m m-1). Using the geometric mean K for the sand
(5.0×10-6 m s-1) and a lateral head gradient of
1.4×10-2 m m-1 yields a specific discharge (Darcy flux, q)
of 2.2 m yr-1. Assuming an effective porosity of 0.3 (Rodvang et al.,
1998), the average linear velocity (v‾) is 7.4 m yr-1. This
suggests that, in the absence of attenuation by mixing or denitrification,
agriculturally derived NO3- could have been transported through
the groundwater system by advection about 400 m since 1960 and
630 m since 1930.
CFO4
The geology at CFO4 consists of about 5 m of clay (with minor till)
underlain by sandstone, to the maximum depth investigated (20 m BG).
Hydraulic conductivities measured using slug tests on wells were
1.0×10-8 to 1.0×10-5 m s-1 (n=12) for
the clay and sandstone (many shallow wells were screened across the
clay–till and into the sandstone) and 1.0×10-5 to
2.9×10-5 m s-1 (n=4) for the sandstone. The depth to
water table ranged from 1.0 to 3.4 m, increasing from west to east across
the study site. Seasonal water table variations were on the order of 1.5 m
with water table declines on the order of 0.3 m yr-1. The horizontal
hydraulic gradient was consistently from west to east, with a mean gradient
at the water table of 3.9×10-3 m m-1 between BC2 and
BMW2 and 4.3×10-3 m m-1 between BMW2 and BMW7. Vertical
hydraulic gradients were 4.2×10-2 to 4.6×10-2 m m-1
downward. Using the geometric mean K for the site (2.9×10-5 m s-1)
and a lateral head gradient of 4.3×10-3 m m-1 yields
a q of 0.4 m yr-1. Assuming an effective
porosity of 0.3 yields a v‾ of 1.3 m yr-1. These values suggest
that, in the absence of attenuation by mixing or denitrification,
anthropogenic NO3- could have been transported through the
groundwater systems about 10 m by advection between 1995 and the time of sampling.
Values and evolution of stable isotopes of nitrate
The range of isotopic values of NO3- in groundwater was similar
at both sites (Fig. 2). At CFO1, δ18ONO3 ranged from
-5.9 to 20.1 ‰ and δ15NNO3 from -5.2 to
61.0 ‰. At CFO4, δ18ONO3 ranged from
-1.9 to 31.6 ‰ and δ15NNO3 from -1.3 to
70.5 ‰. The isotopic values of δ18ONO3
in groundwater are commonly assumed to be derived from a mix of a
one-third atmospheric-derived oxygen (+23.5 ‰) and
two-thirds water-derived oxygen (Xue et al., 2009). Given the average
δ18OH2O for both sites (-16 ‰; see
Supplement), a one-third atmospheric two-thirds groundwater mix would result
in a δ18ONO3 of -3.7 ‰. Manure
filtrate from the EMS at CFO1 had δ15NNO3 ranging from
0.4 to 5.0 ‰ and δ18ONO3 ranging from
7.1 to 19.0 ‰. A curve showing the co-evolution of
δ18ONO3 (mixing of atmospheric δ18O with
groundwater-derived δ18O) and δ15NNO3
(Rayleigh distillation, β=1.005) during nitrification is shown
in Fig. 2. Isotopic values in DMW3, where direct leakage from the EMS was
evident, are consistent with partial nitrification following this trend of
isotopic evolution (δ18ONO3 of -1.2 ‰ and
δ15NNO3 of 7.8 ‰).
(a) Cross-plot of stable isotopes of nitrate at CFO1 and CFO4
showing hypothetical nitrification trend, boundary of manure-sourced
NO3- values and linear enrichment trends associated with
denitrification. (b) Enrichment of δ15NNO3 during
denitrification (only samples within source region and with evidence of
denitrification are shown); dashed lines represent ±1 SD of enrichment
factor (ε=-10).
At both sites, co-enrichment of δ18ONO3 and
δ15NNO3 characteristic of denitrification was evident in some
samples (slopes of 0.42 and 0.72 in Fig. 2a). At CFO1, this includes samples
from DP10-2, DMW5, DMW11, DMW12, DP11-12b, and DMW13 (and associated core)
and some pore water from cores DC15-22 and DC15-23. These samples had
NO3-N concentrations of 0.6 to 23.7 mg L-1,
δ18ONO3 ranging from 4.8 to 20.6 ‰, and
δ15NNO3 ranging from 22.9 to 61.3 ‰.
At CFO4, samples exhibiting evidence of denitrification were from BMW2,
BMW5, BMW6, BMW7, and BC4. These samples had NO3-N concentrations
ranging from 0.4 to 35.1 mg L-1, δ18ONO3 ranging
from 1.6 to 22.1 ‰, and δ15NNO3
ranging from 20.9 to 70.1 ‰. Although the isotopic
values of DMW5 suggest enrichment by denitrification, the data plot away
from the rest of the CFO1 data and close to the denitrification trend at
CFO4 (Fig. 2), suggesting these samples were affected by some other process
(possibly mixing or nitrification); therefore, fd was not calculated. Also, well DMW3, which clearly
receives leakage from the EMS, did not contain substantial NO3-N and so
fd was not calculated.
In the Monte Carlo analysis the potential range of original isotopic values
of the NO3- source prior to denitrification (δ15N0
and δ18O0) varied from 5 to
27 ‰ for δ15NNO3 and from -2 to
7 ‰ for δ18ONO3 based on isotopic
values measured during this study (Fig. 2a). These values are consistent
with literature values for manure-sourced NO3-, which report
δ15NNO3 ranging from 5 to 25 ‰ and
δ18ONO3 ranging from -5 to 5 ‰ (Wassenaar,
1995; Wassenaar et al., 2006; Singleton et al., 2007; McCallum
et al., 2008; Baily et al., 2011). ε15N was defined by a
normal distribution with a mean of -10 ‰ and standard
deviation of 2.5 ‰ (Fig. 2b). At CFO1, the coefficient of
proportionality between the enrichment factor of δ15NNO3
and δ18ONO3 was described by a normal distribution
with mean of 0.72 and standard deviation of 0.05. At CFO4, the coefficient
of proportionality was also described by a normal distribution with a mean
of 0.42 and standard deviation of 0.035 (see Fig. 2a). These enrichment
factors are consistent with values from denitrification studies that report
ε15N ranging from -4.0 to
-30.0 ‰ and ε18O ranging from -1.9 to
-8.9 ‰ (Vogel et al., 1981; Mariotti et al., 1988;
Böttcher et al., 1990; Spalding and Parrott, 1994; Mengis et al., 1999;
Pauwels et al., 2000; Otero et al., 2009).
Distribution and sources of agricultural nitrate in groundwater
At both sites TN concentrations in filtrate from the EMS and catch basin
were generally an order of magnitude larger than concentrations in
groundwater (Table 2). The one exception is well DMW3 at CFO1, which
intercepted direct leakage from the EMS (see Sect. 3.3.1 for further discussion of
this well). The dominant form of N differed between manure filtrate and
groundwater. In the EMS filtrate, N was predominately organic N (TON up to
71 %) or NH3-N (up to 90 %), with NOx-N< 0.1 % of
TN. In the catch basin at CFO1 TON was >99 % of TN. In
groundwater TN concentrations ranged from <0.25 to 84.6 mg L-1,
and this N was predominantly NO3- (again, with the
exception of DMW3).
Range of measured concentrations of TN, NH3-N, NOx-N
(NO2-N+NO3-N), and TON at each study site. At CFO1 results from
monitoring well DMW3 are presented separately because values in this well
differed substantially from all other wells.
SiteN poolTNNH3-NNOx-NTON(mg L-1)(mg L-1)(mg L-1)(mg L-1)CFO1EMS550–1820275–747<0.1–0.473–1301Catch basin200–14402.5–7.3<0.1196–1437DMW3278–548219–479<0.1–50a31.3–73.9Other monitoring wells<0.25–33.4<0.05–2.9<0.1–31.4b<0.2–3.7CF04EMSc1000–1240724–7470.25–0.29275–492Monitoring wells<0.25–84.6<0.05–0.23<0.1–80.4<0.2–13.9
aNOx-N of 50 mg L-1 in DMW3
consisted of 12.6 mg L-1 as NO3-N and 37.4 mg L-1 as
NO2-N. bNOx-N max in groundwater
was measured in core (NO3-N= 66.4 mg L-1,
NOx-N= 67.8 mg L-1). c Range across
three replicates was measured on 25 August 2011.
CFO1
Agriculturally derived NO3- was generally restricted to the upper
20 m (or less) at CFO1 (NO3-N≤ 0.2 mg L-1 and
Cl-≤ 57 mg L-1 in seven wells screened at 20 m). The one exception
was DP11-12b, which had up to 4.1 mg L-1 of NO3-N. The southeast
portion of the site also does not appear to have been significantly
contaminated by agriculturally derived NO3-, with NO3-N
concentrations < 1 mg L-1 in five water table wells (DMW4,
DMW6, DMW14, DMW15, DMW16). In DMW6, Cl- and TN concentrations were
elevated (see Supplement) but NO3-N concentrations were
<2 mg L-1. Collectively, these data suggest the catch basin is
not a significant source of NO3- to the groundwater at this site.
Temporal variations in (a)NO3-N,
(b)Cl-, and (c)NO3-N/Cl- at CFO1.
Only wells with NO3-N> 10 mg L-1 are shown.
Leakage of manure slurry from the EMS at CFO1 is clearly indicated by the
data from DMW3, which feature the highest concentrations of TN in
groundwater (up to 548 mg L-1) and elevated Cl-, HCO3-,
and DOC in concentrations similar to EMS manure filtrate (see Supplement).
Nevertheless, NO3-N concentrations in this well were
consistently low (1.1±2.7 mg L-1, n=22). The potential for
nitrification in the vicinity of this well is indicated by NO2-N
production (2.7±8.3 mg L-1, n=22). However, the data
demonstrate that only a small proportion of the NH3-N in DMW3
(373.4±79.4 mg L-1, n=22) could have been converted to
NO3- within the subsurface (NO3-N in groundwater ≤ 66 mg L-1).
Further work is required to assess the importance of cation exchange as an
attenuation mechanism for direct leakage from the EMS at this site.
(a) Estimated NO3-Ni/Cli ratios
(mean and SD) in water table wells with evidence of denitrification at CFO1,
plotted with distance from earthen manure storage (EMS), where dashed lines are
the upper and lower bounds of DP10-2 (EMS source) and labelled values are maximum measured
NO3-N (mg L-1). (b) Estimated concentrations of
NO3-Ni and Cli at CFO1 (mid-range, error
bars are max and min values).
Contamination by agricultural NO3- that exceeds the drinking water
guidelines (NO3-N> 10 mg L-1) was observed in four
wells (DMW1, DMW11, DMW13, and DP10-2) and in the continuous core (DC15-23) (Fig. 3). DMW2
and DMW12 also had NO3-N concentrations that were elevated but did not
exceed the drinking water guideline (≤3.7 mg L-1). Given the
evidence of partial nitrification in DMW3 (and low NO3-N
concentrations), the NO3-N/Cl- ratio of contamination from the EMS
was assumed to be best represented by DP10-2, which is located directly
down-gradient of the EMS. Data for this well indicate values of
NO3-N/Cl- predominantly ranging from 0.1 to 0.3 with
NO3-Ni/ Cli estimated at 0.3±0.13 (Fig. 4).
The maximum NO3-N concentration in groundwater at CFO1 (66.4 mg L-1)
was measured in core sample DC15–23 (clay at 2 m b.g.l., 7 m
hydraulically down-gradient of DMW3). Pore water extracted from the
unsaturated zone (sand) at the top of this core profile contained 865 mg L-1
of NO3-N and had a NO3-N/Cl- ratio of 1.04,
consistent with the ratio of 0.95 in the core sample. Given this
consistency, and that NO3-N concentrations in the well immediately
up-gradient were low (DMW3), the NO3-N in this core sample was most
likely introduced into the groundwater system by vertical infiltration or
diffusion from above. In contrast, elevated NO3-N (up to 21.1 mg L-1)
within the sand between 6 and 12 m depth in this core had
NO3-N/Cl- ratios consistent with an EMS source (0.07 to 0.31).
Stable isotope values in pore water from this sand layer do not indicate
substantial denitrification (δ18O≤ 5.9 ‰,
δ15N≤ 16.7 ‰), suggesting these ratios will be similar to the
initial ratios at the point of entry to the groundwater system.
In DMW13 (33 m down-gradient from DP10-2) the ratio of
NO3-Ni/ Cli was 0.75±0.29, similar to the
NO3-N/Cl- ratio in DC15-23 at 2 m (0.95), which is interpreted as
reflecting a top-down source. The NO3- in DMW13 is therefore
unlikely to be sourced solely from leakage from the EMS, and could be
sourced from the adjacent dairy pens or a temporary manure pile that was
observed adjacent to this well during core collection in 2015 (or a
combination of EMS and top-down sources).
In DMW12 the NO3-Ni/ Cli ratio was not inconsistent with an EMS
source, but the hydraulic gradient between DMW2 and DMW12 is negligible,
indicating a lack of driving force for advective transport from the EMS
towards DMW12. This is also the case for well DMW1, which is up-gradient of
the EMS but had elevated NO3-N concentrations (6.5±3.6,
n=18). The source of nitrate in these wells is therefore unlikely to be
related to leakage from the EMS, but alternative sources (i.e. nearby
temporary manure piles) are not known.
Well DMW11, 470 m from the EMS, had consistently low NO3-N/Cl-
ratios (<0.05), similar to DP10-2, but estimates of Cli were
3 times higher than Cli for DP10-2 (Fig. 4b). NO3-Ni and
Cli estimated for DMW11 were consistent with measured values in that
well, indicating a local top-down source. Well DMW11 is located
hydraulically down-gradient of feedlot pens and adjacent to a solid manure
storage area, in a local topographic low. Elevated NO3-N in this well
is therefore interpreted to be from surface runoff and top-down
infiltration, rather than lateral advection from the EMS.
CFO4
At CFO4, measured data indicate that effects from agricultural operations on
NO3- concentrations in groundwater are restricted to the upper
15 m of the subsurface. NO3-N concentrations in wells screened at 15 m
depth were <0.5 mg L-1, with the exception of one sample from
BP10-15w (May 2012) with 4.3 mg L-1 of NO3-N. Water table wells in
the west and north of the study site (BC1, BC2, and BC3) also indicate
negligible impacts of agricultural operations, with Cl-< 10 mg L-1
and NO3-N< 0.1 mg L-1.
Temporal variations in (a)NO3-N,
(b)Cl-, and (c)NO3-N/Cl- at CFO4.
Only wells with NO3-N> 10 mg L-1 are shown.
Concentrations of NO3-N> 10 mg L-1 were measured in
three water table wells (BMW2, BMW3, BMW4) adjacent to the EMS, indicating
that they have been impacted by the EMS (Fig. 5). Of these, BMW2 had much
higher Cl- concentrations (502±97 mg L-1, n=22 in BMW2
compared to 182±81 mg L-1 in BMW3 and 188±74 mg L-1
in BMW4), and therefore lower NO3-N/Cl- ratios (<0.05).
Cl- concentrations in BMW2 were consistent with concentrations
in the EMS suggesting direct leakage, while stable isotopes of
NO3- and initial concentrations (NO3-Ni≥127 mg L-1)
indicate substantial denitrification (Table 3, Fig. 6). The
NO3-Ni/ Cli ratio in BMW2 is consistent with measured
NO3-N/Cl- in BMW4, which therefore likely reflects leakage from
the EMS without denitrification (consistent with stable isotope of values of NO3-).
(a) Estimated NO3-Ni/Cli ratios
(mean and SD) in water table wells with evidence of denitrification at CFO4,
plotted with distance from earthen manure storage (EMS), where dashed lines are
upper and lower bounds of BMW2 (EMS source) and values are maximum measured
NO3-N (mg L-1). (b) Estimated concentrations of
NO3-Ni and Cli at CFO4 (mid-range, error
bars are max and min values).
Given that the estimated subsurface travel distance during operations at
this site is 10 m, agriculturally derived NO3- in other wells not
immediately adjacent to the EMS is unlikely to be related to leakage from
the EMS. Wells BMW5 and BMW7 are 60 and 140 m hydraulically down-gradient
from the EMS, respectively. NO3-Ni/ Cli ratios in these wells were
not inconsistent with BMW2 (i.e. the range of values overlap), but given
the distance from the EMS the source of NO3-N in these wells is most
likely the adjacent dairy pens. Concentrations of NO3-N> 10 mg L-1
were also measured in BC4, which is located 95 m hydraulically
up-gradient of the EMS. The ratio of NO3-Ni/ Cli at BC4 was the
highest at CFO4 (0.6) and did not overlap with BMW2. The NO3- in
this well is interpreted to have been sourced from an adjacent manure pile,
which was observed during the study.
Mechanisms of attenuation of agriculturally derived NO3-
Attenuation of agriculturally derived NO3- in groundwater is
dominated by denitrification at both CFO1 and CFO4, with estimates of fm
consistently higher than estimates of fd (Tables 3 and S10, Fig. 7).
Calculated fd values indicate that where denitrification was
identified, at least half of the NO3-N present at the initial point of
entry to the groundwater system has been removed by this attenuation
mechanism. Comparison of NO3-Nmix (the concentration of NO3-N
that would be measured if mixing was the only attenuation mechanism) with
measured concentrations (which reflect attenuation by both mixing and
denitrification) suggests that the sample from 20 m depth (DP11-12b) is the
only sample that would be below the drinking water guideline if mixing was
the only attenuation mechanism (Fig. 8).
Calculated fd and fm based on measured
Cl- and NO3-N concentrations and stable isotope values of NO3-.
a Central depth of core samples, x, indicated as
SampleID_xm. b Maximum fm is 1 for all samples, which
implies no mixing.
Relative contributions to NO3- attenuation by mixing and
denitrification, as indicated by estimated fm and fd at
(a) CFO1 and (b) CFO4, for groundwater samples with
denitrification indicated by stable isotope values of NO3-.
At both sites, the stable isotope values of NO3- indicate that
denitrification proceeds within metres of the source. At CFO1, calculated
fd in well DP10-2 (2 m from the EMS) is 0.52±0.22; at CFO4,
fd in well BMW2 (3 m from the EMS) is 0.13±0.06. Denitrification
also substantially attenuated NO3-N concentrations in wells where the
source is not the EMS but instead is adjacent solid manure piles (e.g.
DMW11 at CFO1, BC4 at CFO4). In BMW6 at CFO4, denitrification completely
attenuated the agriculturally derived NO3-. This well had
negligible NO3-N (0.4±0.2 mg L-1, n=8) and the
lowest fd of 0.01. Measured DOC in this well was consistent with other wells at
both sites (6.9±1.7 mg L-1, n=3), suggesting DOC depletion
does not limit denitrification at these CFOs.
Measured concentrations of NO3-N (blue circles – attenuation
by mixing and denitrification) and NO3-Nmix (red triangles – attenuation
by mixing only) vs. mid-range estimate of NO3-Ni at
(a) CFO1 and (b) CFO4. Dashed lines are drinking water
guideline (10 mg L-1 of NO3-N).
DiscussionImplications for on-farm waste management
Agriculturally derived NO3- at these two sites with varying
lithology was generally restricted to depths < 20 m, consistent with
previous studies at CFOs (Robertson et al., 1996; Rodvang and Simpkins,
2001; Rodvang et al., 2004; Kohn et al., 2016). Attenuation of
agriculturally derived NO3- in groundwater was a spatially
varying combination of mixing and denitrification, with denitrification
playing a greater role than mixing at both sites. In the samples for which
fd could be determined, denitrification reduced NO3-
concentrations by at least half and, in some cases, back to background
concentrations. Given that the range of source isotopic composition was
allowed to vary to its maximum justifiable extent, these quantitative
estimates of denitrification based on stable isotopes of NO3- are
likely to be conservative. Redox conditions within the groundwater system
were not able to be determined in this study due to the sampling method used
to collect groundwater from wells screened across low-K formations (well
bailed dry then sample collected after water level recovery). However,
denitrification appears to proceed within metres of the NO3-
source, suggesting relatively short subsurface residence times are required
and that redox conditions close to the water table are conducive to
denitrification reactions (Critchley et al., 2014; Clague et al., 2015).
The substantial role of denitrification within the saturated glacial
sediments at these study sites indicates the potential for significant
attenuation of agriculturally derived NO3- by denitrification in
similar groundwater systems across the North American interior and Europe
(Ernstsen et al., 2015; Zirkle et al., 2016). Denitrification in the
unsaturated zone is limited by low water contents and oxic conditions,
resulting in substantial stores of NO3- in vadose zones
(Turkeltaub et al., 2016; Ascott et al., 2017). NO3- in water that
is removed rapidly from the site is also unlikely to be substantially attenuated
by denitrification due to oxic conditions and rapid transit times (Ernstsen
et al., 2015). Therefore, water management focussed on reducing the effects
of NO3- contamination in similar hydrogeological settings to this
study should aim to maximise infiltration into the saturated zone where
NO3- concentrations can be naturally attenuated, provided that
local groundwater is not used for potable water supply.
At both sites there is evidence of elevated NO3- due to leakage
from the EMS, but the impact appears to be limited to within metres of the
EMS. This suggests that saturation within the clay lining of the EMS has
limited the development of extensive secondary porosity that would allow
rapid water percolation (Baram et al., 2012). Infiltration of NO3--rich water that has passed through temporary solid manure piles and dairy
pens has resulted in groundwater NO3-N concentrations as high as those
associated with leakage from the EMS (e.g. DMW11, BC4). At CFO4, this is in
spite of the presence of clay at the surface, reflecting secondary porosity in
the upper part of the profile that has led to hydraulic conductivities
comparable to sand. This is consistent with the findings of Showers et al. (2008),
who investigated sources of NO3- at an urbanised dairy
farm in North Carolina, USA. Construction of EMS facilities in Alberta has
been regulated under the Agriculture Operation Practices Act since 2002,
which requires them to be lined with clay to minimise leakage (Lorenz et
al., 2014). On-farm waste management should increasingly focus on minimising
temporary manure piles that are in direct contact with the soil to reduce
NO3- contamination associated with dairy farms and feedlots.
Critique of this approach and applicability at other sites
At both sites, leakage from the EMS had NO3-Ni/ Cli of between 0.1
and 0.4, but this alone was not diagnostic of the source. The sources of
manure-derived NO3- (manure piles vs. EMS) are distinguishable
based on NO3-Ni/ Cli ratios, provided there is also an understanding
of the history of each site, local hydrogeology, and potential sources.
Calculated fd and fm generally decreased with increasing subsurface
residence time and distance from source, providing additional evidence for
source attribution. For example, at CFO4, well BMW2, which is adjacent to
the EMS, had the highest fm (0.92), indicating the least attenuation of
NO3 by mixing and consistent with the EMS being the source of
NO3- to this well. Temporal variability in NO3-Ni/ Cli for each source could not be determined
based on the snapshot isotope sampling conducted, but this could be
investigated by measuring NO3- isotopes in conjunction with
NO3-N and Cl- at multiple times.
Effect of neglecting background concentrations (Clb or
NO3-Nb) in the mixing model on calculated fm over the
range of values in this study.
Calculation of NO3-Ni/ Cli assumed that background concentrations
could be neglected in the mixing model. At these study sites, background
concentrations are likely to be <20 mg L-1 for Cl- and
<1 mg L-1 for NO3-N. Estimated NO3-Ni values
were at least 20 times background NO3-N concentrations, and over
100 times background concentrations in some wells. The estimated Cli values
were at least 3 times as high as the background concentrations at CFO1 and at least
10 times as high as the background concentrations at CFO4. The error introduced by neglecting
background concentrations was assessed by comparing fm calculated with
and without background concentrations included, using the full range of
values in this study (Fig. 9). Neglecting background concentrations results
in overestimation of fm (i.e. underestimation of the amount of
attenuation mixing) with the largest errors occurring when measured concentrations are
close to background concentrations. For Cl- the maximum difference
of 0.13 is in the mid-range of fm values. For NO3-N, the
difference is consistently <0.1 with the largest errors at the
lowest values of fm. The uncertainty in fm is primarily related to
uncertainty in the initial concentrations (Cli and NO3-Ni), which
depends on measured Cl- and NO3-N. The largest uncertainties in
NO3-Ni and Cli correspond to the lowest measured concentrations
(i.e. furthest from the upper limit), with less uncertainty at higher
measured concentrations as they approach the maximum values.
Although applicable at these sites, this approach may not be valid at other
sites if additional sources of NO3 in groundwater (e.g. fertiliser or
nitrification) are significant, or if NO3 concentrations in groundwater
are naturally elevated (Hendry et al., 1984). The combination of the approach
outlined here with measurement of groundwater age indicators would allow for
better constraints on groundwater flow velocities and determination of
denitrification rates (Böhlke and Denver, 1995; Katz et al., 2004;
McMahon et al., 2004; Clague et al., 2015).
Comparison with isotopic values of NO3- in previous studies
Nitrate isotope values in groundwater at the two CFOs studied were generally
consistent with previous studies reporting denitrification of manure-derived
NO3- at dairy farms (Wassenaar, 1995; Wassenaar et al., 2006;
Singleton et al., 2007; McCallum et al., 2008; Baily et al., 2011). However,
the isotopic values of NO3- in the manure filtrate from the EMS
at CFO1 were not consistent with values for manure-sourced NO3-
reported in other groundwater studies (Wassenaar, 1995; Wassenaar et al.,
2006; Singleton et al., 2007; McCallum et al., 2008; Baily et al.,
2011). This is likely to be because nitrification within the EMS was
negligible (NO3-N< 0.7 mg L-1), such that the isotopic
values of NO3-N in the manure filtrate reflect volatilisation of
NH3 and partial nitrification within the EMS. δ18ONO3
values may also have been affected by evaporative enrichment of the
δ18OH2O being incorporated into NO3- (Showers et al., 2008).
A number of groundwater samples collected during this study had relatively
enriched δ18ONO3 (>15 ‰) with
depleted δ15NNO3 (<15 ‰).
Some of these isotopic values are within the range previously reported
for NO3- derived from inorganic fertiliser
(δ15NNO3 from -3 to 3 ‰ and
δ18ONO3 from -5 to 25 ‰), with the
δ18ONO3 depending on whether the NO3- is from
NH4+ or NO3- in the fertiliser (Mengis et al., 2001;
Wassenaar et al., 2006; Xue et al., 2009). To the best of our knowledge,
however, no inorganic fertilisers have been applied at these study sites.
Another potential source is NO3- derived from soil organic N, but
this should have δ15NNO3 values of 0 to
10 ‰ and δ18ONO3 values of -10 to
15 ‰ (Durka et al., 1994; Mayer et al., 2001; Mengis et
al., 2001; Xue et al., 2009; Baily et al., 2011). Incomplete nitrification
of NH4+ can result in δ15NNO3 lower than the
manure source (Choi et al., 2003), but as there was no measurable NH3-N
in these samples this is also unlikely. These isotope values may reflect the
influence of NO3- from precipitation, which usually has values
ranging from -5 to 5 ‰ for δ15NNO3 and
40 to 60 ‰ for δ18ONO3 and has been
reported to dominate NO3- isotope values of groundwater under
forested landscapes (Durka et al., 1994). Alternatively, they may be
affected by microbial immobilisation and subsequent mineralisation and
nitrification, which can mask the source δ18ONO3 in
aquifers with long residence times (Mengis et al., 2001; Rivett et al., 2008).
Conclusions
A mixing model constrained by quantitative estimates of denitrification from
isotopes substantially improved our understanding of nitrate contamination
at these sites. This novel approach has the potential to be widely applied
as a tool for monitoring and assessment of groundwater in complex
agricultural settings. NO3-N concentrations in excess of the drinking
water guideline were measured at both sites, with sources including manure
piles, pens, and the EMS. Even though these sites are dominated by clay-rich
glacial sediments, the input of NO3- to groundwater from
temporary manure piles and pens resulted in
NO3-N concentrations comparable to (or greater than) leakage from the EMS. This is attributed to
the development of secondary porosity within unsaturated clays.
Nitrate attenuation at both sites is dominated by denitrification, which is
evident even in wells directly adjacent to the NO3- source. In the
wells for which denitrification was identified, concentrations of
agriculturally derived NO3- had been reduced by at least half and,
in some wells, completely. In the absence of denitrification all but one of
these wells would have had NO3-N concentrations above the drinking water guideline.
These results indicate that infiltration to groundwater systems in glacial
sediments where NO3- can be naturally attenuated is likely to be
preferable to off-farm export via runoff or drainage networks, provided
that local groundwater is not a potable water source. On-farm management of
manure waste at similar operations should increasingly focus on limiting
manure piles that are in direct contact with the soil to limit
NO3- contamination of groundwater.
Data availability
Alberta Agriculture and Forestry are the custodians of the
data used in this paper.
The supplement related to this article is available online at: https://doi.org/10.5194/hess-23-1355-2019-supplement.
Author contributions
Investigation was carried out by SB, MI and JM with
assistance from staff at AAF, NRCB and USask. SB developed the data analysis
methodology and prepared the paper. All co-authors contributed to
supervision, conceptualization, review and editing.
Competing interests
The authors declare that they have no conflict of interest.
Acknowledgements
This research was supported by Alberta Agriculture and Forestry (AAF) and
the Natural Resources Conservation Board (NRCB), who provided assistance
with field work and laboratory analysis. Funding was also provided by a
Natural Sciences and Engineering Research Council of Canada (NSERC)
Industrial Research Chair (IRC) (184573) awarded to M. Jim Hendry. The authors
thank Barry Olson at AAF for reviewing the paper. Our thanks also to the
local producers, whose cooperation made this research possible. And finally,
our thanks to Sebastien Lamontagne, Huaiwei Sun and anonymous reviewers for
their valuable comments and suggestions during the review process.
Edited by: Bill X. Hu
Reviewed by: Huaiwei Sun and three anonymous referees
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