Science of the Total Environment 817 (2022) 152606
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Science of the Total Environment
j ourna l homepage: www.e lsev ie r .com/ locate /sc i totenvIsotopic evidence for nitrate sources and controls on denitrification in
groundwater beneath an irrigated agricultural districtStephen J. Harris a,b, Dioni I. Cendón a,b,⁎, Stuart I. Hankin b, Mark A. Peterson b, Shuang Xiao a,b, Bryce F.J. Kelly a
a School of Biological, Earth and Environmental Sciences, UNSW Sydney, NSW 2052, Australia
b Australian Nuclear Science and Technology Organisation, Lucas Heights, NSW, 2234, AustraliaH I G H L I G H T S G R A P H I C A L A B S T R A C T• Groundwater NO−3 in an irrigated agricul-
tural district was studied using isotope
tracers.
• Groundwater NO−3 is linked tomodern irri-
gation practices commencing in the 1970s.
• NO−3 is derived from N fertilisers which
was previously retained in the SON pool.
• Denitrification is a key feature of the
groundwater system.
• A conceptual model for district-scale NO−3
production and attenuation is discussed.⁎ Corresponding author at: School of Biological, Earth and
E-mail address: dce@ansto.gov.au (D.I. Cendón).
http://dx.doi.org/10.1016/j.scitotenv.2021.152606
0048-9697/Crown Copyright © 2022 Published by El
nd/4.0/).A B S T R A C TA R T I C L E I N F OArticle history:
Received 23 September 2021
Received in revised form 16 December 2021
Accepted 18 December 2021
Available online 8 January 2022
Editor: Christian HerreraThe application of N fertilisers to enhance crop yield is common throughout the world. Many crops have historically
been, or are still, fertilised with N in excess of the crop requirements. A portion of the excess N is transported into un-
derlying aquifers in the form of NO−3 , which is potentially discharged to surface waters. Denitrification can reduce the
severity of NO−3 export fromgroundwater.We sought to understand the occurrence and hydrogeochemical controls on
denitri cation in NO−fi 3 -rich aquifers beneath the Emerald Irrigation Area (EIA), Queensland, Australia, a region of ex-
tensive cotton and cereal production. Multiple stable isotope (in H2O, NO
−
3 , DIC, DOC and SO
2−
4 ) and radioactive iso-
tope (3H and 36Cl) tracers were used to develop a conceptual N process model. Fertiliser-derived N is likely
incorporated and retained in the soil organic N pool prior to its mineralisation, nitrification, and migration into aqui-
fers. This process, alongside the near absence of other anthropogenic N sources, results in a homogenised groundwater
NO−3 isotopic signature that allows for denitrification trends to be distinguished. Regional-scale denitrification mani-
fests as groundwater becomes increasingly anaerobic duringflow from an upgradient basalt aquifer to a downgradient
alluvial aquifer. Dilution and denitrification occurs in localised electron donor-rich suboxic hyporheic zones beneath
leaking irrigation channels. Using approximated isotope enrichment factors, estimates of regional-scale NO−3 removal
ranges from22 to 93% (average: 63%), and from 57 to 91% (average: 79%) beneath leaking irrigation channels. In the
predominantly oxic upgradient basalt aquifer, raised groundwater tables create pathways for NO−3 to be transported to
adjacent surface waters. In the alluvial aquifer, the transfer of NO−3 is limited both physically (through groundwater-
surface water disconnection) and chemically (through denitrification). These observations underscore the need to un-
derstand regional- and local-scale hydrogeological processes when assessing the impacts of groundwater NO−3 on ad-
jacent and end of system ecosystems.Keywords:
Groundwater nitrate
Isotope tracers
Denitrification
Agriculture
Emerald Irrigation AreaEnvironmental Sciences, UNSW Sydney, NSW 2052, Australia.
sevier B.V. This is an open access article under the CC BY-NC-ND license (http://creativecommons.org/licenses/by-nc-
S.J. Harris et al. Science of the Total Environment 817 (2022) 1526061. Introduction
Excess nitrate (NO−3 ) in groundwater, especially if subsequently
discharged to surface waters, is a global environmental problem that con-
tinues to threaten aquatic ecosystems (Adimalla et al., 2021; Burow et al.,
2010; Gu et al., 2013; Mateo-Sagasta et al., 2017; Rabalais, 2002;
Spalding and Exner, 1993; Vitousek et al., 1997). The recent enhancement
of NO−3 in groundwater is in large part due to diffuse pollution from inten-
sive agriculture (Foster, 2000; Kendall et al., 2007). In Australia, the issue
of N pollution is pertinently demonstrated within the World Heritage-
listed Great Barrier Reef (GBR), which continues to be affected by land-
based N runoff that manifests in the eutrophication and acidification of
many coral reef ecosystems (Bell, 1992; Brodie et al., 2012; Great Barrier
Reef Marine Park Authority, 2019). The Fitzroy Basin drains the largest
area (~142,000 km2) of the GBR catchments. Prior to the development of
agriculture, it has been estimated that only 1700 t/year total N (TN) was
exported from the Fitzroy Basin to the GBR (Kroon et al., 2012). It is now
estimated that approximately 15,000 t/year TN is exported from the Fitzroy
Basin. This is the highest proportional increase in TN loads of all GBR
catchments (Kroon et al., 2012). Currently, the Fitzroy Basin accounts
for approximately 20% of TN delivered to the GBR (Bartley et al.,
2017), making it a key target for implementing improved N manage-
ment practices.
The Emerald Irrigation Area (EIA) is of particular note for potential
groundwater-surface water NO−3 transfer in the Fitzroy Basin. Irrigated
cropping, particularly cotton farming, is widespread throughout the EIA.
These crops rely on the application of N fertiliser to enhance crop yield.
Fertiliser application rates for fully irrigated cotton in the Fitzroy Basin av-
erage 296.9 kg N/ha, which is lower than the industry average 335.9 kg N/
ha (Cotton Research and Development Corporation, 2018). However, these
rates are still higher than amounts recommended by research, which indi-
cates that application rates exceeding 200–240 kg N/ha do not improve
lint yield (Macdonald et al., 2018). Nitrogen management is made more
complex given the variable Central Queensland climate. N-based fertilisers,
such as urea and urea ammonium nitrate are the most widely used
fertilisers in the region (Australian Bureau of Statistics, 2018). The applica-
tion of manure to cotton crops in the EIA is not as widespread, with only ap-
proximately 50% of cotton farmers using manure (chicken, cow, feedlot,
and pig manure) for fertilisation in Central Queensland (Roth Rural,
2016). Application rates are lower and highly variable, ranging from 1 to
20 t/ha of manure which is roughly equivalent to 5–20 kg N/ha based on
manure type (Roth Rural, 2016). Historical data indicate that NO−3 occurs
in EIA aquifers at varying concentration, reaching as much as 29 mg/L N
(Queensland Government, 2020), suggesting that excess N from inorganic
fertiliser and/or manure is migrating to EIA aquifers. Despite these histori-
cal measurements, the relative contributions from inorganic fertilisers and
manure to groundwater NO−3 , as well as the aquifer-scale hydrogeological
controls on its distribution, have not been studied in the EIA. Elucidating
the sources and controls on these processes provides the knowledge re-
quired to improve land and groundwater management practices in the
EIA. This also advances our understanding of both catchment- and
basin-wide NO−3 sources and their potential impacts to downstream eco-
systems such as the GBR.
Denitrification attenuates NO−3 in groundwater (Böttcher et al., 1990;
Postma et al., 1991; Trudell et al., 1986), and therefore could play an im-
portant role in controlling the magnitude of NO−3 loads exported from the
EIA. The removal of NO−3 via denitrification in groundwater is predomi-
nantly observed under anaerobic conditions with sufficient electron donors
(Rivett et al., 2008). Many studies have shown strong hydrological controls
on groundwater denitrification, especially redox boundaries that occur
along hydraulic gradients (Clague et al., 2015; Hinkle et al., 2007;
Stenger et al., 2018; Tesoriero et al., 2000; Tesoriero et al., 2007;
Tesoriero and Puckett, 2011). Hyporheic zones tend to be sites for denitri-
fication because of the anaerobic conditions, presence of electron donors,
and occurrence of microbial communities that promote denitrification
(Böhlke et al., 2009; Harvey et al., 2013).2
Likemany agricultural districts situated in semi-arid climates, the EIA is
equippedwithwater infrastructure that enables intensive cultivation.Much
of the irrigated cropping in the area has been exclusively sustained bywater
from Fairbairn Dam, whose construction was completed in 1972 to form
LakeMaraboon. Irrigationwater is distributed via an irrigation channel net-
work, the majority of which was built by 1979 (Geddes et al., 2019). Seep-
age from these earthen channels, combined with deep drainage from crop
irrigation and land clearing, have enhanced groundwater infiltration into
EIA aquifers. This has resulted in rising groundwater levels and decreased
unsaturated zone thickness (Pearce and Hansen, 2007). Many studies
have shown that these processes can alter groundwater flow paths and pre-
vailing geochemical conditions (e.g., Böhlke, 2002; Pulido-Bosch et al.,
2018; Scanlon et al., 2008, 2009; Stewart and Aitchison-Earl, 2020). By ex-
tension, the alteration of hydrogeochemical conditions caused by irrigation
in the EIA could have a strong influence on groundwater NO−3 cycling. For
instance, groundwater mounding can enable evaporation near the ground-
water surface, which concentrates dissolved salts and NO−3 in remaining
groundwater (Pulido-Bosch et al., 2018). Waterlogging also enhances min-
eralisation rates in soil profiles, transforming organic forms of N into more
soluble and mobile forms of oxidised N, such as NO−3 (Scanlon et al., 2008,
2009). Leakage from irrigation infrastructure, such as channels and dams,
can further enhance groundwater recharge and distort groundwater flow
paths (Fernald and Guldan, 2006; Harvey and Sibray, 2001; Kendy and
Bredehoeft, 2006). Locally, irrigation channel leakage can change the geo-
chemical conditions of aquifers via their recharge to groundwater. This ren-
ders them conducive to biologically mediated processes such as
denitrification (Fernald and Guldan, 2006).
It is well-documented that the isotopes of NO− (δ15N-NO−3 3 and δ18O-
NO−3 ) can assist in identifying probable NO−3 sources, in addition to detect-
ing the occurrence of denitrification and/or dilution in groundwater
systems (Aravena and Robertson, 1998; Aravena et al., 1993; Carrey
et al., 2021; Kendall et al., 2007; Widory et al., 2004). Additional isoto-
pic tracers, such as δ13C in dissolved inorganic and organic carbon (DIC
and DOC) and δ34S and δ18O in sulfate, have been used to trace the co-
fractionation of the electron donors (organic C and reduced Fe) in deni-
trification processes (Otero et al., 2009; Puig et al., 2017; Valiente et al.,
2018; Vitòria et al., 2008). δ34S and δ18O in sulfate are additionally use-
ful as a proxy for detecting fertiliser signatures in groundwater, espe-
cially because the δ15N signatures of different NO−3 sources have
considerable overlap (Xu et al., 2016).
In addition to stable isotopes, radioactive tracers such as tritium (3H)
have been used to determine whether concentrations of NO−3 are related
to modern anthropogenic activities (Böhlke and Denver, 1995; Clague
et al., 2015; Erostate et al., 2018; Koh et al., 2010; McMahon and Böhlke,
2006; Pastén-Zapata et al., 2014; Stenger et al., 2018). These tracers have
also been used to estimate the timeframe over which denitrification persists
based on apparent groundwater residence times (Hinkle et al., 2007). The
utility of radioactive isotope tracers potentially extends beyond that of
being robust indicators of groundwater age. For example, the 36Cl/Cl
ratio is a useful tool to study salt origins and transport processes in younger
groundwater, such as halite dissolution and evapotranspiration (Cartwright
et al., 2006). To our knowledge, 36Cl/Cl ratios have not been explicitly used
to study N pollution but may have utility in resolving the mobilisation/
mixing of solutes from the unsaturated zone. For example, Cl− andNO−3 in-
creases in aquifers receiving enhanced recharge prompted by land use
change (e.g., Scanlon et al., 2008).
Considering the above, the aim of this study is to investigate NO−3 sources
and its transport and fate across the EIA groundwater system. To address
these issues, we combine geochemical (major ion), stable isotope (δ2H-H2O,
δ18O-H2O, δ15N-NO−, δ18O-NO−, δ133 3 C-DIC, δ13C -DOC, δ34S-SO2−4 and
δ18O-SO2−4 ) and radioactive isotope (3H and 36Cl) tracers to:
1. identify the source (s) of groundwater NO−3 ;
2. determine the hydrogeological controls on denitrification and how this
may have been altered by irrigation practices; and
3. evaluate the potential for NO−3 export from the aquifers of the EIA.
S.J. Harris et al. Science of the Total Environment 817 (2022) 1526062. Study area
2.1. Location and climate
The EIA (~150 km2) is located on both thewestern and eastern banks of
the Nogoa River, which intersects the township of Emerald, in the Lower
Nogoa sub-catchment of the Fitzroy Basin, Queensland, Australia (Fig. 1).
It is located in a hot semi-arid climate (BSh; Peel et al., 2007). The mean
January and July maximum temperatures at Emerald Airport (site number
035264) are 34.6 °C and 23.4 °C, respectively, with the average rainfall for
the samemonths being 83.0mm and 16.6mm, respectively (Australian Bu-
reau of Meteorology, 2021). Themean annual precipitation (1992–2021) is
544 mm/year whilst evaporation ranges between 2000 and 2400 mm/
year. Rainfall is strongly seasonal, with 74% falling in the summer months
between October to March.
Fig. 1 presents the location of 24 groundwater and 4 surface water sam-
pling sites used for this study. The groundwater sample locations are re-
ferred to as NogGW1 to NogGW24. These sites are located within and
adjacent to the EIA, covering an area of ~408 km2 (see Fig. 1). SurfaceFig. 1.Map of the study area, including groundwater sampling sites and irrigation chann
Natural Neighbour algorithm over the study area to highlight the hydraulic gradient. Tra
“S”) are also indicated.
3
water samples are referred to as sites NogSW1, NogSW2, NogSW3 and
NogSW5. Further details on sample collection are outlined in Section 3.
Groundwater samples were confined to the existing QLD Department of
Natural Resources, Mines and Energy groundwater monitoring network.
These monitoring wells provide both good spatial and depth sampling cov-
erage of the different hydrogeological units in the study area.
2.2. Hydrogeology
Within the EIA, the oldest outcropping formation is the Late Permian
Freitag Formation comprising interbedded clastic sedimentary rocks that
are exposed at Fairbairn Dam (lithic sandstone, siltstone, coal, and carbona-
ceous shale; Fielding andMcloughlin, 1992; Power, 1966). The Freitag For-
mation is underlain by the Early Permian Aldebaran Sandstone
(sandstones, conglomeratic sandstones, conglomerates, and coal; Dickins
and Malone, 1973), which is intersected by deeper groundwater monitor-
ing wells with screens below 30 m (NogGW16 and 17).
The Permian sedimentary sequences are overlain byfluviatile and lacus-
trine sediments of the Emerald Formation (unconsolidated to semi-els used for crop irrigation in the EIA. Piezometric head data is interpolated using the
nsect A is indicated in red. The locations of river sampling sites (designated with an
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606unconsolidated claystones, siltstones, sandstones with minor interbedded
basalt) deposited throughout the Paleocene and Oligocene (Day et al.,
1983). These sedimentary rocks have low permeability and poorly devel-
oped fracture networks, resulting in low groundwater extraction yields.
The sedimentary aquifer is therefore not exploited for irrigated agriculture
(Pearce and Hansen, 2007). Only one monitoring well screening this aqui-
fer was sampled in this study (NogGW24) due to a lack of intersecting
groundwater monitoring wells.
The preceding sedimentary sequences are overlain by basalts and tra-
chytes, most likely emplaced during the Oligocene based on dating of re-
gional volcanics (Jones et al., 2018). These volcanics form an unconfined
aquifer, with groundwater resources contained within weathered and frac-
tured zones that are hydraulically connected (Pearce and Hansen, 2007;
Webb and McDougall, 1967). Most monitoring wells sampled in this
study are screened within a localised basalt flow that is exposed on the
western bank of the Nogoa River. This exposure forms a topographic high
that has weathered to form dark cracking clay soils (McDonald and
Baker, 1986). The soils on the basalt are typically less than 1 m deep, mak-
ing the underlying aquifers vulnerable to anthropogenic pollution from ir-
rigated cropping practices.
From the Miocene to Late Pleistocene, a paleo-valley was carved
through these rocks and filled with unconsolidated colluvial and alluvial
sediments (coarse sands and gravels with varying amounts of clay). The dis-
tribution of these sediments follows the drainage system of the Nogoa River
and its tributaries, forming a floodplain that varies in width from less than
100 m to 6 km either side of the riverbanks (Tucker et al., 2003).
A prominent feature of the EIA is that groundwater is not pumped for
irrigation. Instead, irrigation waters are sourced exclusively from dam
water distributed via irrigation channels. Groundwater recharge occurs
via a combination of rainfall, crop irrigation and irrigation channel water
leakage (discussed further in Section 2.3), with no input from irrigation re-
turn flows because groundwater is not used for crop irrigation. Groundwa-
ter recharged in the upgradient basalt aquifer flows in an easterly direction
into the downgradient alluvial aquifer. Transect A traces south-easterly
groundwater flow over a portion of the study area (Fig. 1) and is used
later in the Discussion to demonstrate the behaviour of NO−3 as groundwa-
ter flows from the basalt aquifer into the alluvial aquifer.
2.3. The EIA irrigation channel network
An important hydraulic feature of the EIA is the surface irrigation chan-
nel network that distributes the water used for irrigated agriculture (see
Fig. 1). This network originates 19 km upstream of Emerald at Fairbairn
Dam (1.3 × 109 m3 capacity), which supplies water for irrigation by con-
trolling flows to the channel network. In the absence of any surface runoff
and precipitation, the flows of the channel networks and Nogoa River are
entirely controlled by water releases from Lake Maraboon through
Fairbairn Dam (Tucker et al., 2003). In total, the channels and pipelines
supplying water for irrigation in the EIA extends for over 126 km and can
distribute irrigation water to an area of approximately 15,000 ha
(Fairbairn Irrigation Network, 2019). The Selma Channel System supplies
water to the western bank of the Nogoa River and the Weemah Channel
System supplies water to the eastern bank of the Nogoa River. A surface
drainage system of over 144 km in length is in place to provide off-farm
drainage (Sunwater, 2012).
The construction of the Selma and Weemah Channel Systems elevated
groundwater levels in the EIA because many channels were not internally
lined. In the absence of regional groundwater abstraction, the water table
has risen between 1.7 and 21 m (6 m on average across sites sampled in
this study) on both banks of the Nogoa River. Analysis of groundwater
hydrographs indicates that groundwater levels have steadily risen since
the 1980–1990s (Queensland Government, 2020). On the western bank
of the Nogoa River, the potentiometric head data indicate that the increase
in recharge – the result of both enhanced areal recharge (crop irrigation)
and point recharge (irrigation channel leakage) – has increased the natural
hydraulic gradient. Recently, portions of the channel network have been4
internally lined to limit leakage and restore natural groundwater levels
(Sunwater, 2020). As a result, the geochemistry of the groundwater that
is affected by irrigation channel leakage may represent transient geochem-
ical signals. Given the strong influence that irrigation channels have had on
groundwater, the presence of leaking irrigation infrastructure may be an
important factor to consider when interpreting groundwater NO−3 cycling
in the EIA.
3. Methods
3.1. Sample collection
Twenty-two groundwater monitoring wells and two privately-owned
wells of varying depths were sampled in October and November 2018
(Fig. 1). Fifteen monitoring wells are screened in the Cenozoic basalt aqui-
fer at depths between 6 and 29 m. These monitoring wells are located on
the western bank of the Nogoa River where the most intensive irrigated
cropping occurs. Five monitoring wells and one private well are screened
in the alluvial aquifer at depths between 13 and 20 m. These monitoring
wells are locatedwithin the Nogoa River floodplain, to the east of the Ceno-
zoic basalt aquifer. The remaining two wells (one private and one monitor-
ing well) are screened in the weathered Cenozoic sedimentary rock aquifer
at depths ranging from approximately 30 m in the private well to 95 m in
the monitoring well. These are located on the periphery of the main irri-
gated cropping practices of the EIA. Monitoring wells with screens less
than 30 m in depth were sampled with a 12 V Proactive impeller pump,
with the pump intake placed approximately 1 m above the screen. Deeper
wells were sampled with a bladder pump using low-flow methods
(Iverach et al., 2017). During this procedure, the pump was placed approx-
imately 10 m below the groundwater level with a drop tube positioned
within the screened interval. Both privately-owned wells were sampled
from a pump outlet because these wells are sealed. Sample collection proce-
dures for each chemical analyte are provided in Supplementary Material 1.
During the groundwater sampling campaign, auxiliary surface water
samples were taken from Nogoa River (sites NogSW1, NogSW2, NogSW4
and NogSW5; Fig. 1). These samples were acquired using a peristaltic
pump and prepared and analysed using the same sampling techniques as
the groundwater samples. The chemistry of the surface water samples is
not the focus of this study but is used to infer groundwater-surface water
mixing processes in the study area. Measurements obtained from these sur-
face water samples are broadly consistent with those collected during a
two-year surface water monitoring campaign conducted by our research
group in the Lower Nogoa sub-catchment. These point-in-time samples
thus provide an adequate representation of surface water inputs into EIA
aquifers.
3.2. Chemical analyses
Total alkalinity concentrations were determined in the field by titration
using aHACHdigital titrator and external pHmeter. Fe2+ andHS− ion con-
centrations were determined using a HACH DR890 portable colorimeter.
NH3-N concentrations were measured on a HACHDR3900 Spectrophotom-
eter using the HACH Method 10,205. Samples for cations were analysed
using inductively coupled plasma atomic emission spectroscopy (ICP-
AES) at ANSTO. Samples for anions were analysed using ion chromatogra-
phy (IC) at ANSTO. The charge balance error for the measured major ion
concentrations did not exceed 5%. Total dissolved solids (TDS) were calcu-
lated as the sum of all major ions.
Samples for δ2H-H O and δ182 O-H2Owere analysed using an established
method on a cavity ring-down spectroscopy (CRDS) on a Picarro L2130-i
analyser at ANSTO. These values are reported as‰ deviations from the in-
ternational standard V-SMOW (Vienna Standard Mean Ocean Water) and
results have a precision of ±1‰ for δ2H and ± 0.15‰ for δ18O. Samples
for δ13C-DIC were also analysed at ANSTO using a Delta V Advantage
mass spectrometer, and a GasBench II peripheral. The results are reported
as ‰ deviations from International Atomic Energy Agency (IAEA)
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606secondary standards that have been certified relative to the international
standard V-PDB (Vienna Pee Dee Belemnite) for carbon with a precision
of±0.3‰. DOC sampleswere analysed at UC-Davis Stable Isotope Facility,
USA using a total organic carbon (TOC) analyser connected to a PDZ Eu-
ropa 20–20 IRMS using a GD-100 Gas Trap interface. Results were
corrected based on laboratory standards calibrated against National Insti-
tute of Standards and Technology (NIST) Standard Reference Materials.
Samples for δ34S-SO and δ184 O-SO4 were analysed using a Carlo Erba
1108 elemental analyser and TC-EA pyroliser (for the δ18O), both coupled
to a Thermo FinniganDelta Plus XP Spectrometer atUniversity of Barcelona
(CCiTUB). Values are reported as‰ deviations from the international stan-
dard Cañon Diablo Troilite (CDT), and the analytical error (2σ) is ±0.3‰
for δ34S-SO and δ184 O-SO4. Values obtained for the international standard
NBS-127 were δ34S: 20.3 ± 0.1‰, and δ18O: 9.3 ± 0.2‰. Samples for
δ15N-NO− and δ183 O-NO−3 analysed on an autosampler / PreCon /
GasBench II assembly coupled to a Finnigan Delta Plus Advantage IMRS,
using the bacterial denitrification method (Sigman et al., 2001) at IsoLab
at the University of Washington, USA. Long-term precision (1σ) for δ15N
and δ18O was ±0.3‰ and ± 0.5‰, respectively.
Samples for 3H were analysed using established liquid scintillation
methodologies at ANSTO as outlined in Cendón et al. (2014). The 3H con-
centrations are expressed in tritium units (TU) with an average combined
standard uncertainty of ±0.05 TU and quantification limit of 0.04 TU.
36Cl/Cl and 36Cl/37Cl ratios were measured by accelerator mass spectrom-
etry (AMS) using the ANSTO 6MV SIRIUS Tandem Accelerator as outlined
in Wilcken et al. (2017).
4. Results
Table S2-1 (Supplementary Material 2) presents the physicochemical,
major ion and isotope results for EIA groundwater. Auxiliary geochemical
data obtained for Lake Maraboon and the Nogoa River, collected at the
same time as the groundwater sampling campaign, are also provided in
Table S2-1.
To aid the description and discussion of results, the EIA groundwater sam-
ples are differentiated into three clusters and referred to as the following:
1. pre-irrigation groundwater samples, which are located on the periphery
of themain irrigated cropping practices of the EIA and have screened in-
tervals in the deeper portions of the aquifers in the study area;
2. mixed groundwater samples, which include most samples collected in
the main irrigation district of the EIA; and
3. leakage-dominated groundwater samples, which are in proximity
(< 1 km) to leaking irrigation channels.
The differentiation of these groundwater samples has been undertaken to
streamline the discussion of the results based on trends in the data, and is not
the focus of this study. To ensure that the categories are sufficiently capable of
describing NO−3 dynamics in the EIA, the statistical validity of these clusters
was assessed using agglomerative hierarchical clustering (see Supplementary
Material 3). The agglomerative hierarchical clusteringwas used to objectively
organise the data into groups based on their δ2H-H 182O, δ O-H O, δ132 C-DIC,
3H, 36Cl/Cl andNO−3 compositions (which form thebasis of our data interpre-
tation). Dissimilarities (and by analogy similarities) between groundwater
sample compositions were measured using squared Euclidean distances ac-
cording to Ward's method and classified using a dendogram (e.g, Lambrakis
et al., 2004; Swanson et al., 2001). There was good agreement between the
derived statistical clusters and those interpreted, indicating that these clusters
are statistically viable representations of processes affecting the cycling of
NO−3 in the EIA groundwater system.
4.1. Physicochemistry and major ions
Across all groundwater samples, temperature varied between 24 and
30 °C, pH was between 5.5 and 8.2, TDS between 395 and 3747 mg/L,
and dissolved oxygen (DO) ranged from 0.1 to 6.4 mg/L. Dissolved NO−35
concentrations ranged from<0.02 to 24.39 mg/L N. Nine samples had con-
centrations that exceeded the WHO maximum guideline value of
10 mg/L N NO−3 for drinking water (WHO, 2011). Seventeen groundwater
samples had concentrations that exceeded the Australian/New Zealand
long-term trigger value for NO−3 in irrigation water of 5 mg/L N NO−3
(ANZECC/ARMCANZ, 2000). Two samples had detectable NO−2 concentra-
tions (0.03 mg/L N NO−2 in both samples). Two samples had detectable NH3
concentrations (0.11 and 0.34mg/L NNH3). The surface water samples from
Lake Maraboon and Nogoa River did not have any detectable NO−3 .
The deep pre-irrigation groundwater samples had no detectable NO−3 but
had detectable NH3 (reaching a maximum 0.34 mg/L N NH3). These sam-
ples had low DO (0.12 to 0.20 mg/L) and variable TDS (1033 to
1554 mg/L). The mixed groundwater cluster consists of Na-HCO3 and Ca-
HCO3 type groundwater associated with the Cenozoic basalt aquifer on
the western portion of the study area. These samples had variable DO
(0.19 to 6.43 mg/L), TDS (521 to 1734 mg/L) and NO−3 (<0.02 to
24.39 mg/L N). Groundwater from the Cenozoic basalt aquifer flows east
to mix with the groundwater in the alluvial aquifer, which is classified as
Na-Cl type groundwater. This groundwater has higher TDS (>3000 mg/L)
and lower DO (<0.6 mg/L) and NO−3 (<0.02 to 9.46 mg/L N). The leak-
age-dominated groundwater samples had low TDS (395 to 649 mg/L) and
NO−3 (<0.02 to 1.83 mg/L N), and were mostly suboxic (<0.5 mg/L DO).
4.2. H2O stable isotopes
H2O stable isotope compositions across all EIA groundwater samples
ranged from −34.4 to +0.9‰ for δ2H-H2O and −5.2 to +1.2‰ for
δ18O-H2O (Fig. 2a). The most depleted groundwater samples were from
the deeper pre-irrigation groundwater systems on the periphery of the
main irrigated cropping practices of the EIA. Their δ2H and δ18O composi-
tions ranged from −34.4 to −28.6‰ and −5.2 to −4.6‰, respectively.
These compositions plot near the global meteoric water line (GMWL;
Craig, 1961). The mixed groundwater samples plot to the right of the
GMWL on an evaporation trendline extending from the pre-irrigation
groundwater to the leakage-dominated groundwater samples (notably
GW10 and 15; δ2H from −0.4 to +0.9‰ and δ18O from +0.8 to
+1.2‰). In Fig. 2a, the evaporation trendline was extended to include
those H2O stable isotopic compositions of Lake Maraboon and Nogoa River
(δ2H from +7.57 to +11.32‰ and δ18O from +2.20 to +2.86‰). The
slope of this evaporation trendline (~5.5) is consistent with those reported
elsewhere in eastern Australia (Hollins et al., 2018; Iverach et al., 2017).
4.3. 3H activities
3H activities ranged from 0.05 to 1.55 TU (average: 0.60 TU). The low-
est 3H activities (0.05 to 0.10 TU) were measured for the pre-irrigation
groundwater (Fig. 2b). In contrast, the highest 3H activities (0.85 to 1.55
TU) were recorded for leakage-dominated groundwater samples. These
higher activities approach those measured in Lake Maraboon and Nogoa
River (1.82 to 1.99 TU). The mixed groundwater samples had 3H activities
between these two extremes, ranging from 0.16 to 1.1 TU.
To determine tracer-based ages for EIA groundwater, 3H input from
modern groundwater recharge were estimated. To do this, a composite
3H rainfall record was calculated based on 3H records from Brisbane,
Queensland (extending from the 1960s to 2012), Charleville, Queensland
(2006 to 2017), other southern hemisphere records for rainfall prior to
1962 (Morgenstern and Taylor, 2009). The effect on 3H due to distance
from the coastline was also taken into account. A loess best-fit was used
to smooth the composite record. The resulting 3H input from rainfall in
the EIA was estimated to be 1.73 TU, which is within the range for 3H in
rainfall predicted by Tadros et al. (2014) for the study area (1.6 to 2.0
TU). This estimate is similar to 3H activities in Lake Maraboon and the
Nogoa River (1.82 to 1.99 TU).
Assuming the average input 3H for modern groundwater recharge is 1.7
TU, the mixed and leakage-dominated groundwater samples (which all have
quanti 3fiable H) have tracer-based ages of less than ~75 years (based on
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606
Fig. 2. a) δ2H-H O and δ182 O-H2O compositions of the different groundwater clusters in the EIA. The global meteoric water line (GMWL; Craig, 1961) and the Brisbane MWL
(Hollins et al., 2018) are indicated by the dashed and dotted black lines, respectively. All groundwater samples plot on an evaporation line (solid line) with the equation
δ2H = 5.47 × δ18O – 4.16‰ (r2 = 0.99, p < 0.001). b) 3H versus δ2H-H2O in the different groundwater groups in the EIA.radioactive decay). However, a further age constraint can be placed on
these samples because they show variable degrees of mixing with infiltrat-
ing irrigation channel water (Fig. 2a and b). This indicates that these
groundwater locations have received recharge from water originating
from the mid-1970s when the first irrigation channels were constructed
in the EIA, placing their tracer-based ages to less than ~40–50 years. This
timeframe is consistent with deep drainage infiltration rates estimated for
the EIA (0.7 to 7.7 m/year; Shaw and Yule, 1978), considering the depth
of screens for these monitoring wells (9.3 to 29.6 m below ground surface;
mbgs). In contrast, the deeper pre-irrigation groundwater samples with low
3H activities (<0.05 TU) have tracer-based ages of more than
~40–50 years, which pre-dates any irrigation practices in the EIA.
4.4. 36Cl/Cl ratios
36Cl/Cl (x10−15) ratios ranged from 72.8 to 170.0. Pre-irrigation ground-
water samples had low 36Cl/Cl ratios (72.8 to 90.8), while themixed ground-
water samples had variable ratios (72.8 to 146.5). Leakage-dominated
groundwater samples had elevated 36Cl/Cl ratios (142.2 to 170.0), which ap-
proach, and at times exceed, those measured for Lake Maraboon and Nogoa
River (109.6 to 159.2). Given that the measured 3H activities pertain to
young tracer ages in EIA groundwater, an estimation of groundwater “age”
using 36Cl/Cl ratios – which require considerably longer groundwater resi-
dence times due to the half-life of 36Cl (301,000 years) – is not justified. In-
stead, these ratios are used to trace the mobilisation of solutes in the young
EIA groundwater, as well as to trace the input of modern 36Cl/Cl-rich water.
4.5. δ15N-NO− and δ18O-NO−3 3
δ15N-NO− signatures ranged from +1.9 to +19.6‰ and δ18O-NO−3 3
ranged from+4.2 to+24.8‰. Neither the pre-irrigation groundwater sam-
ples, nor Lake Maraboon and Nogoa River, had detectable NO−3 concentra-
tions. Therefore, no isotopic signatures were attainable for these samples.
The mixed and leakage-dominated groundwater samples had similar varia-
tions in their isotopic composition.
4.6. δ34S-SO2− and δ184 O-SO2−4
δ34S-SO2−4 ranged from +3.9 to +15.4‰ and δ18O-SO2−4 ranged from
+9.6 to +19.0‰. Pre-irrigation groundwater samples had δ34S-SO2−46
compositions ranging from +9.0 to +9.5‰, and δ18O-SO2−4 ranged from
+10.5 to +11.6‰. Mixed groundwater samples had δ34S-SO2−4 composi-
tions ranging from +3.9 to +15.4‰, and δ18O-SO2−4 ranged from +9.6
to +17.6‰. The leakage-dominated groundwater samples had δ34S-SO2−4
compositions ranging from +4.7 to +8.8‰, and δ18O-SO2−4 ranged from
+15.4 to +19.0‰. These δ18O compositions are notably higher than the
vast majority of those in the mixed groundwater cluster (barring GW7).
No isotopic signatures were attained for Lake Maraboon and Nogoa River
during this sampling campaign.
4.7. HCO− and δ133 C-DIC
HCO−3 concentrations ranged from 31.5 to 925.1 mg/L and δ13C-DIC
compositions ranged from−15.2 to−8.6‰. In the lower pH groundwater
samples (pH < 6.4), HCO−3 concentrations were low (31.5 to 276.8 mg/L)
and δ13C-DIC signatures were depleted (−15.2 to −13.1‰). For the re-
maining circum-neutral groundwater, δ13C-DIC compositions extended
from c. -13‰ (mostly mixed groundwater) to −8‰ (mostly pre-irrigation
and leakage-dominated groundwater samples). Lake Maraboon and Nogoa
River HCO−3 concentrations ranged from 164.3 to 176.9 mg/L and δ13C-
DIC compositions ranged from −6.6 to−3.5‰.
4.8. DOC and δ13C-DOC
DOC concentrations ranged from 0.37 to 3.25 mg/L but were typi-
cally between 0.37 and 1.5 mg/L. δ13C-DOC compositions ranged
from −32.6 to −17.2‰, yet most samples had compositions between
−28 to −25‰. There were no clear distinctions in the DOC content
and δ13C-DOC signature across the three geochemical clusters. In Lake
Maraboon and Nogoa River, DOC concentrations ranged from 5.2 to
5.5 mg/L and δ13C-DOC compositions ranged from −24.9 to −24.1‰.
5. Discussion
5.1. Groundwater recharge sources and distribution of NO−3
The pre-irrigation groundwater in the study area is characterised by low
3H activity, 36Cl/Cl ratios, and non-evaporated water stable isotopes, repre-
senting natural recharge signatures (Fig. 2a and b). The lack of NO−3 in
these groundwater samples might be explained by the absence of NO−3
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606contamination, and/or complete removal of NO−3 via denitrification given
the suboxic conditions. Although the latter cannot be confirmed via the use
of NO−3 isotope data, it is likely that this groundwater cluster contains nat-
ural background NO−3 concentrations given their depth, lack of evidence
for evaporation and tracer-based ages (> 40–50 years). Detectable, yet
minor, NH3-N concentrations may be of natural origin.
The tendency of the mixed groundwater samples to trend towards the
water stable isotopic (Fig. 2a) and 3H (Fig. 2b) composition of surface
water samples suggests that they are a mixture of modern recharge water
sources (i.e., evaporated irrigation waters sourced from Lake Maraboon)
and pre-irrigation groundwater. The stable water isotopic enrichment in
themixed groundwater samples is not likely the result of in-situ evaporation
given the depth of the groundwater sampling. Instead, elevated 3H activity
and enrichment in 2H and 18O in the mixed groundwater samples is most
likely derived throughmixingwith infiltrating evaporated irrigationwaters
which have atmospheric 3H signals and enriched 2H and 18O. NO−3 concen-
trations in these samples were broadly dependent on whether they were in
the upgradient Na-HCO−3 type waters of the basalt aquifer (lower TDS, and
higher DO andNO−3 ), or the downgradient Na-Cl typewaters of the alluvial
aquifer (higher TDS, and lower DO and NO−3 ). This is reflected in theFig. 3. NO−3 (mg/L N) plotted against a) TDS, b) DO, c) 3H, and d) 36Cl/Cl ratios. In p
irrigation groundwater values are indicated by a range (in violet) because NO−3 concen
by a range (in light blue). Symbols are as in Fig. 2.
7
moderate negative correlation between NO− and TDS (r23 = 0.60,
p < 0.001; Fig. 3a) and moderate positive correlation between NO−3 and
DO (r2 = 0.50, p < 0.005; Fig. 3b). This suggests that the distribution of
NO−3 is in part controlled by processes that occur over the regional flow
paths implied by the hydraulic gradient.
All samples containing detectable NO−3 in the mixed cluster had 3H ac-
tivities above quantification limits (0.05 TU; Fig. 3c), suggesting NO−3 is de-
rived from modern anthropogenic sources. Interestingly, 36Cl/Cl ratios
were positively correlated with NO−3 concentration in this groundwater
cluster (r2 = 0.69, p < 0.001; Fig. 3d). One possible explanation for this
may be that the variation in 36Cl/Cl ratios reflect the incorporation of Cl
(and other salts) from the unsaturated zone into the saturated zone follow-
ing the introduction of irrigation in the EIA (Scanlon et al., 2008, 2009).
Prior to the mid-1970s, this Cl likely accumulated over millennia in the un-
saturated zone in the form of dispersed halite resulting in lower 36Cl/Cl ra-
tios. The dissolution and subsequentmobilisation of old chlorides following
cultivation may have resulted in a decline in the 36Cl/Cl ratio along the
groundwater flow path. After modern NO−3 -rich water with elevated
36Cl/Cl ratios was introduced to the system post-1970s, a simultaneous in-
crease in NO−3 and 36Cl/Cl was possibly established in more-recentlyanel d), one sample (GW6) was removed from the linear regression. Note that pre-
trations were below detection limit. Similarly, surface water samples are indicated
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606recharged groundwater. There does not appear to be simultaneous isotopic
enrichment in water stable isotopes as salinity increases (Fig. 4a), which is
consistent with the origin of salinity being from the dissolution of pre-
existing unsaturated zone salts rather than recent evapo-concentration.
Moreover, NO−3 concentrations do not appear to have been increased via
evapo-concentration (Fig. 4b).
The leakage-dominated groundwater cluster samples have H2O stable
isotope, 3H and 36Cl/Cl ratios that range from those of themixed groundwa-
ter towards those of Lake Maraboon and Nogoa River (Fig. 2a and b). The
elevated 3H and 36Cl/Cl ratios in these samples indicate that they have
been influenced more heavily at a local scale by modern irrigation supply
channel leakage than the mixed groundwater samples (Fig. 3c and d).
Plots of NO−3 vs TDS and DO (Fig. 3a and b) show the behaviour of NO−3
in the groundwater at these sample locations is distinct from the mixed
groundwater cluster: they have low NO−3 and TDS and are mostly suboxic
(<0.5 mg/L DO). This suggests that localised modern surface water input
via irrigation supply channel leakage dilutes and, potentially, attenuates
NO−3 at these locations. Indeed, the dilution effect of irrigation supply chan-
nel leakage on both TDS and NO−3 on this groundwater cluster can be seen
in Fig. 4. The possibility of denitrification is explored further in Section 5.3.
Overall, themixing trends shown in Fig. 2b indicate that there is a contin-
uum of mixing between recharge sources in the EIA (rainfall, crop irrigation
and irrigation channel leakage), and that this is a modern process occurring
within the past 40–50 years. The different groundwater clusters showvarying
degrees ofmixing across the different recharge sources, with the leakage-dom-
inated groundwater samples displaying the greatest degree of mixing with ir-
rigation surface waters. The evaporated irrigation supply channel waters,
which originate upstream from Lake Maraboon, have low NO−3 contents.
These irrigation waters are therefore unlikely to contribute to NO−3 loads
via leakage, but rather dilute underlying groundwater.
5.2. Identifying NO−3 sources
The potential anthropogenic sources of NO−3 include inorganic
fertilisers and manure, which have recent (post-1970s) and widespread ap-
plication across the EIA. Fig. 5a and b show the δ15N-NO−3 and δ18O-NO−3
values for both themixed and leakage-dominated groundwater samples in re-
lation to potential NO−3 isotopic source signatures (derived from Xu et al.,
2016 and refences therein). Although a wide range of δ18O-NO−3 signatures
have been reported for different NO−3 sources via nitrification in the litera-
ture (ranging from−10 to+15‰; Xu et al., 2016 and references therein),
the measurement of δ18O-H2O permits further refinement of signatures ofFig. 4. δ2H-H2O plotted against a) TDS and, b) NO−3 concentration in the groundwater
irrigation groundwater samples and surface water samples are plotted as a range
concentrations. Symbols are as in Fig. 2.
8
NO−3 derived from nitrification. During nitrification, one atom of oxygen
from each of H −2O and dissolved O2 is incorporated into the NO2 molecule
during the first step (Buchwald et al., 2012; DiSpirito and Hooper, 1986;
Hollocher, 1984), and one atom of oxygen from H2O is incorporated into
the NO−3 molecule in the second step (Kumar et al., 1983). It has been
noted, however, that various kinetic isotopic fractionation effects occur
during both steps (Buchwald and Casciotti, 2010; Casciotti, 2009;
Casciotti et al., 2010; Granger and Wankel, 2016). Therefore the δ18O sig-
nature of the product NO−3 do not always reflect the isotopic composition
of H2O and O2 in a 2:1 ratio (Snider et al., 2010). Nonetheless, in keeping
with most groundwater NO−3 studies (e.g., Minet et al., 2017; Osaka
et al., 2010; Paredes et al., 2020; Valiente et al., 2018), in this paper the ex-
pected δ18O-NO−3 compositions derived via the nitrification of NH+4
fertilisers and manure is approximated using Eq. (1), after Mayer et al.
(2001):
δ18O−NO − ¼ 2=3 ∙ δ183 estimate O−H2O 18sample þ 1=3 ∙ δ O−O2 atmosphere (1)
where δ18O-NO−3 estimate is the expected δ18O-NO−3 composition of the
groundwater sample, δ18O-H2O sample is the measured δ18O-H2O composi-
tion of the sample, and δ18O-O2 atmosphere is the δ18O composition of atmo-
spheric O2 (+23.5‰; Kroopnick and Craig, 1972).
The expected δ18O-NO−3 composition for nitrification derived using
Eq. (1) for the mixed and leakage-dominated groundwater samples ranged
from+4.3 to+7.3‰ and+7.0 to+8.6‰, respectively. Although these re-
sults are interpreted with caution, when compared to the individual mea-
sured δ18O-NO−3 signatures, nine mixed and three leakage-dominated
groundwater samples had δ18O-NO−3 compositions exceeding those derived
from Eq. (1) (Fig. 6). This suggests possible enrichment via denitrification.
Those samples containing nitrified NO−3 had δ15N-NO−3 signatures
ranging from +1.9 to +9.9‰, which plot in the fields for both nitrified
NH+4 fertiliser and soil orgnaic nitrogen (SON; Fig. 5a and b). The dominant
fertilisers applied to EIA crops are NH+4 -based, which explains the tendency
of data to plot towards this isotopic source field.
Due to the significant overlap in δ15N-NO−3 signatures of NH+4 fertiliser
and SON, the contribution from the nitrification of reduced SON is chal-
lenging to resolve. Prior to cultivation and irrigation commencing in the
EIA, SON was presumably derived from fixed N2 that was incorporated
into soil and plants. However, this natural source of soil NO−3 alone does
not account for the elevated groundwater NO−3 concentrations in EIA aqui-
fers. During microbial immobilisation, inorganic N is reincorporated into
the SON pool and N is retained in the soil profile as organic N (Wellsand surface water samples. Similar observations are made for δ18O-H2O. In b), pre-
for demonstrative purposes, as these samples did not contain detectable NO−3
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606
Fig. 5. Dual NO−3 isotope plots for the a)mixed and b) leakage-dominated groundwater samples. The range of δ18O composition of these sources (apart from NO−3 fertilisers) was
limited to those expected via nitrification assuming theNO−3 incorporatedO atoms fromH2O and atmospheric O2 in a 2:1 ratio as in Eq. (1) (Mayer et al., 2001). Note that thefield
for sewage-derived NO−3 is not provided, as the groundwater locations sampled in this study were not proximal to any possible sewage sources. Symbols are as in Fig. 2.et al., 2015). Several authors have postulated that NO−3 derived from
fertilisers is not directly leached to groundwater. Instead, it is retained in
the soil-plant system prior to its subsequent mineralisation and re-
oxidation during nitrification (Mengis et al., 2001; Somers and Savard,
2009; Stewart and Aitchison-Earl, 2020). This process is known as
mineralisation-immobilisation-turnover (Mengis et al., 2001), and has
been shown to alter and homogenise δ15N-NO−3 and δ18O-NO−3 composi-
tions in groundwater (Kloppmann et al., 2018). In addition, waterlogging
can enhance microbial activity in the soil zone causing increased rates of
mineralisation of SON (Scanlon et al., 2008). Waterlogging is well-
documented in the EIA (Department of Natural Resources Queensland,
1998; Silburn et al., 2013; Yule, 1997), meaning this could also play an im-
portant contributing factor during this turnover process. Thus, it is plausible
that prior to being leached to groundwater, fertiliser-derived NO−3 is
retained in the SON poolwhere its isotopic composition is homogenised, re-
sulting in the isotopic signatures shown in Fig. 5. It is only after its subse-
quent mineralisation and nitrification that it is leached to EIA aquifers.Fig. 6. δ18O-NO−3 vs. δ18O-H2O for EIA groundwater samples. Symbols are as in Fig. 2.
9
Itmust also be noted that volatilisationmay occur during the surface ap-
plication of NH+4 fertilisers. This process enriches the residual NH+4 , which
is then transferred to the NO−3 during nitrification (Vitòria et al., 2005).
Thereby, the measured δ15N signatures of NO−3 derived from NH+4 -based
fertilisers which have undergone volatilisation can verge towards those of
manure (Valiente et al., 2018). The occurrence of volatilisation in the EIA
is well-documented (Weier, 1994), and therefore could be responsible for
the apparent vergence of δ15N-NO−3 signatures towards more enriched
NO−3 sources such as manure.
δ34S-SO2−4 and δ18O-SO2−4 signatures in the mixed and leakage-domi-
nated groundwater samples lend further credibility to a fertiliser source of
NO−3 in the EIA. In Fig. 7a, the δ34S-SO2−4 and δ18O-SO2−4 compositions
plot in the same range typical for fertiliser-derived sulfate (Cravotta,
2002; Finlay and Kendall, 2007; Otero et al., 2007; Rees et al., 1978;
Vitòria et al., 2004, 2008). They also roughly plot on a mixing line between
fertiliser-derived and seawater SO2−4 , suggesting aminor contribution from
seawater SO2−4 via rainfall deposition. It is noteworthy, however, that the
pre-irrigation groundwater cluster also plots in a similar field, and therefore
the signatures cannot be adequately separated from natural background
signatures. Nonetheless, the observed signatures are significantly more
enriched than those expected for the oxidation of sulfides (Vitòria et al.,
2008), and the δ18O-SO2−4 signatures are more elevated than would be ex-
pected formanure (Cravotta, 2002; Otero et al., 2009) and soil organicmat-
ter (Finlay and Kendall, 2007).
Collectively, the δ15N-NO−3 , δ18O-NO−3 , δ34S-SO2−4 and δ18O-SO2−4 data
indicate the source of NO−3 in EIA groundwaters is most likely the nitrifica-
tion of fertilisers and less so from manure. The contribution of NO−3
fertilisers, as well as the occurrence of additional processes such as
volatilisation could not be ruled out by the trends apparent in Fig. 5.5.3. Evidence for denitrification using NO−3 isotopes
During denitrification, as the NO−3 concentration decreases, there is
a simultaneous increase in both δ15N and δ18O of the residual NO−3 .
Many groundwater studies have shown that denitrification results in
Δδ18O-NO−:Δδ153 N-NO−3 ratios ranging from 0.5 to 0.8 (e.g., Aravena
and Robertson, 1998; Baily et al., 2011; Böhlke et al., 2006; Bourke
et al., 2019; Cey et al., 1999; Minet et al., 2017; Singleton et al.,
2007), which deviate substantially from the ratio of ~1 expected for
bacterial denitrification (Sigman et al., 2005). Granger and Wankel
(2016) showed that Δδ18O-NO−3 :Δδ15N-NO−3 ratios <1 arise when
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606
Fig. 7. a) δ34S-SO4 versus δ18O-SO4. The range of values for the different SO2−4 sources indicated were taken from references in-text. b) δ34S-SO4 versus δ15N-NO−3 . The
expected trend for autotrophic denitrification is indicated by an arrow after Hosono et al. (2014). Symbols are as in Fig. 2.NO−2 derived from denitrification is re-oxidised to NO−3 by incorporat-
ing atoms of oxygen from ambient H2O with depleted δ18O-H2O. In con-
trast to freshwater systems, ratios equal to or greater than 1 have been
reported for marine systems (Buchwald et al., 2015). This is attributed
to elevated rates of NO−2 re-oxidation relative to NO−3 reduction and
the incorporation of O atoms with enriched δ18O-H2O values
(Casciotti et al., 2013).
In the mixed groundwater samples, there was a strong positive linear
correlation between δ15N-NO− and δ183 O-NO−3 (r2 = 0.90, p < 0.001;
Fig. 5a), indicating the occurrence of denitrification (Mariotti et al.,
1988). Δδ18O-NO−:Δδ153 N-NO−3 increased at a ratio of 0.71, which were
in a similar range to previously reported ratios for groundwater denitrifica-
tion. This suggests that denitrification may have been superimposed by
concurrent nitrification, with the producedNO−3 incorporating atoms of ox-
ygen from ambient H 182Owith depleted δ O-H2O (as shown byGranger and
Wankel, 2016). Plots of δ15N-NO−3 and δ18O-NO−3 vs ln(NO−3 ) (Fig. 8a and
b) showedmoderate toweak tomoderate linear correlations (r2=0.49 and
0.32, respectively, p < 0.05). Estimated isotope enrichment factors (ε) cal-
culated using a simplified Rayleigh equation (Mariotti et al., 1988) were
−5.2‰ for 15ε and −3.2‰ for 18ε.Fig. 8. δ15N-NO−3 and δ18O-NO−3 plotted against the natural logarithm of the NO−3 co
Associated enrichment factors (ε) are indicated next to each regression line. Symbols ar
10Similarly, leakage-dominated groundwater samples showed a strong pos-
itive linear correlation between δ15N-NO− 18 − 23 and δ O-NO3 (r = 0.97, p=
0.01; Fig. 5b), indicating the occurrence of denitrification. Δδ18O-NO−3 :
Δδ15N-NO−3 increased at a ratio of 1.34, which is greater than typically re-
ported values for groundwater denitrification. Plots of δ15N-NO−3 and δ18O-
NO−3 vs ln(NO3) (Fig. 8a and b) showed strong linear correlations (r2 =
0.98 and 0.97, respectively, p < 0.05), with isotope enrichment factors of
−6.1‰ for 15ε and − 8.2‰ for 18ε. Importantly, because these samples
plot along a straight line in Fig. 8, this suggests that the isotopic enrichment
of NO−3 in these samples is caused by denitrification, rather than dilution or
mixing of different NO−3 sources (Mariotti et al., 1988). The deviation of the
Δδ18O-NO−3 :Δδ15N-NO−3 ratio from 1 also suggests that the simultaneous
re-oxidation of NO−2 occurs concurrently to NO−3 reduction in this ground-
water cluster. However, in contrast to the mixed cluster, NO−2 re-oxidation
to NO−3 likely incorporates O atoms with enriched δ18O-H2O originating
from evaporated surface waters. There was indeed a strong positive linear
correlation between δ18O-NO− 183 and δ O-H2O for the leakage-dominated
groundwater (r2 = 0.98, p < 0.01), suggesting that the NO−3 has derived
part of its δ18O-NO− signature from isotopically enriched δ183 O-H2O (see
Fig. 6). Similar trends indicating partial reduction/re-oxidation processesncentration in mg/L N for the mixed and leakage-dominated groundwater clusters.
e as in Fig. 2.
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606were observed for δ18O-SO2−4 in this groundwater cluster (r2 = 0.95,
p < 0.05), lending further support for partial reduction/re-oxidation pro-
cess occurring in these locations. Collectively, these trends suggest that
suboxic hyporheic zones have been created beneath leaking irrigation sup-
ply channels. These locations likely represent zones of enhanced exchange
between surfacewater and groundwater, creating favourable conditions for
denitrification to occur (Fernald and Guldan, 2006).
In view of the multiple lines of evidence suggesting the presence of de-
nitrification in our study area, we used the estimated enrichment factors to
calculate the percentage of denitrification for both the mixed and leakage-
dominated groundwaters. This approach is outlined in Ostrom et al.
(2002) and Otero et al. (2009), wherein Eq. (2) is used to quantify the per-
centage of denitrification in the groundwater samples:


ð Þ ¼ ½

NO
DEN 1− 3

% residual½   100h NO3 initial i (2)
¼ 1− eðδresidual − δinitial=εÞ 100
To quantify the percentage of denitrification, an initial δ15N-NO−3 com-
positionmust be assumedwhen 15ε is used in the calculation (Ostrom et al.,
2002; Otero et al., 2009). For the mixed groundwaters, NogGW24 was se-
lected as this corresponds to the location that had the highest NO−3 and
DO concentrations, and lowest δ15N-NO−3 signature (+1.9‰), and thus
was the most likely to correspond to nitri −fication-derived NO3 . Similarly,
NogGW4 (δ15N-NO−3 = +5.1‰) was selected for the leakage-dominated
groundwater cluster. Excluding NogGW24 and NogGW4, the extent of de-
nitrification ranged from 22 to 93% (average: 63%; median: 64%) in the
mixed groundwaters, and from 57 to 91% (average: 79%; median: 89%)
in the leakage-dominated groundwaters (Table S2-1). Refinement of esti-
mated enrichment factors and, by extension, the amount denitrified,
would be possible with more rigorous temporal studies and/or from the in-
stallation of multi-piezometer networks in the EIA.
5.4. What are the hydrogeochemical controls on denitrification?
5.4.1. Dissolved oxygen
δ15N-NO−3 compositions in the study area increase with decreasing DO
regardless of groundwater cluster (Fig. 9a), suggesting that DO is an over-
arching control on denitrification in the EIA groundwater system. A similar
trend is evident in the δ18O-NO−3 data (data not shown). NogGW19, which
clearly contained denitrified NO− (δ153 N-NO−3 = +10.2‰; δ18O-NO−3 =
+10.1‰), had a DO concentration of 2.6 mg/L, which represents an ap-
proximate limit below which denitrification occurs. The most denitrifiedFig. 9. a) δ15N-NO−3 plotted against DO. Suboxic conditions (<0.5 mg/L DO) are indica
11samples were those with suboxic DO concentrations (<0.5 mg/L), reinforc-
ing the strong DO control on denitrification.
5.4.2. Groundwater flow path
Analogous to the strong DO control on denitrification in the mixed
groundwaters, there was a strong positive linear correlation between
δ15N-NO−3 and TDS (r2= 0.83, p < 0.001, NogGW6 excluded as an outlier;
Fig. 9b). This indicates that denitrification tended to follow the flow path as
the groundwater incorporated solutes and became progressively anaerobic.
The groundwater flow path control on denitrification within the mixed
groundwaters was confirmed by examining the geochemistry of groundwa-
ters downgradient along Transect A as it flows from the upgradient basalt
aquifer to the downgradient alluvial aquifer (shown in Fig. 1). Along this
transect, DO and NO−3 concentrations progressively decrease alongside
the hydraulic gradient, while δ15N-NO−, δ183 O-NO−3 , % denitrified and
TDS increase (Fig. 10), highlighting the strong influence of groundwater
flow duration on denitrification.
In contrast, there was no relationship between TDS and δ15N-NO−3 in
the leakage-dominated groundwater cluster, and the groundwater flow
path had little effect on denitrification (Fig. 9b). This suggests that denitri-
fication in the leakage-dominated groundwaters were primarily controlled
by suboxic hyporheic zone processes occurring near leaking irrigation
channels, as opposed to over the regional hydraulic gradient.
5.4.3. Electron donors (C and S)
The groundwater samples with circum-neutral pH showed a mixing
trend between DIC with near-atmospheric δ13C composition (a first end-
member; c. -8‰; Clark and Fritz, 1997) and DIC with δ13C compositions
c. -12‰ (a second end-member; Fig. 11a). The first end-member with
higher δ13C values likely reflects modern dissolved CO2 within infiltrating
surface waters. The mixing of this surface water with groundwater is
most pronounced in the leakage-dominated groundwater samples
(Fig. 11a). The second δ13C end-member of c. -12‰ may reflect
naturally-occurring processes, such as the weathering of silicate minerals
with soil CO2, or leakage-dominated overprinting from the weathering of
pedogenic carbonates. Thus, δ13C compositions of −12‰ likely reflect
pre-irrigation groundwater signatures, with the subsequent enrichment in
δ13C-DIC reflecting modern input of DIC from irrigation waters.
There was no clear decreasing trend in δ13C-DIC when plotted against
δ15N-NO−3 in themixed and leakage-dominated clusters (Fig. 11b), possibly be-
cause the natural weathering and mixing with modern waters containing ele-
vated proportions of atmospheric CO2 have overprinted any evidence of
heterotrophic denitrification. In the mixed groundwater samples, there was
only a weak negative correlation between NO−3 and HCO−3 (r2 = 0.37;ted by the dashed line. b) δ15N-NO−3 plotted against TDS. Symbols are as in Fig. 2.
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606
Fig. 10. Geochemical evolution ofmixed groundwater samples taken along Transect A as it flows from the oxic basalt aquifer into the suboxic alluvial aquifer. Denitrification
progresses along the flow path. Note: NogGW21 was sampled from the pump outlet, and therefore piezometric head could not be determined.p < 0.05; Fig. S4-1a in Supplementary Material 4). Similarly, there were only
weak positive correlations between δ15N-NO−3 and HCO− (r23 = 0.37,
p < 0.05; Fig. S4-1b in Supplementary Material 4), further confirming the
lack of insight that inorganic C can provide into the denitrification process
in the mixed groundwater cluster. Interestingly, there was a strong positive
correlation between NO−3 and HCO−3 (r2 = 0.95, p < 0.05; Fig. S4-2a in Sup-
plementaryMaterial 4) and a strong negative relationship between δ15N-NO−3
and HCO− 23 (r = 0.96, p < 0.05; Fig. S4-2b Supplementary Material 4) in the
leakage-dominated cluster. These relationships are seemingly at odds with het-
erotrophic denitri cation (a decrease in NO−fi 3 and increase in δ15N-NO−3
should coincide with increasing HCO−3 ). One possible explanation for this12relationship is that the extent of denitrification is directly related to the mag-
nitude of irrigation channel leakage at these locations. That is, denitrification
occurs most at locations receiving high amounts of irrigation channel waters
that have low HCO−3 content. Further research is required to determine the
spatial effect of irrigation leakage on denitrification in the EIA groundwater
system.
There was no relationship between DOC concentration and δ15N-NO−3 ,
nor between δ13C-DOC and δ15N-NO−3 in the mixed and leakage-dominated
groundwater clusters (Fig. S4-3 Supplementary Material 4). This suggests
that the dissolved C fraction was not consumed during denitrification.
The lack of a clear link between DOC and denitrification, however, is
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606
Fig. 11. a and b) δ13C-DIC plotted against δ15N-NO− and HCO−. Note that δ15N-NO−3 3 3 data was unattainable for the pre-irrigation groundwater samples and for the modern
irrigation waters (Lake Maraboon and Nogoa River). Note that the y-axis scale is changed in b). Symbols are as in Fig. 2.consistent with the findings of Hinkle et al. (2007), who found dissolved C
fractions were less significant in denitrification compared to solid phase OC
contained within the aquifer matrix. Nonetheless, three groundwater sam-
ples (GW9, 10 and 15), which all belong to the leakage-dominated cluster,
had DOC concentrations exceeding 2 mg/L. This could indicate that DOC
was contributed via the infiltration of irrigation channel waters, which
have higher DOC contents ranging from 5.2 to 5.5 mg/L. The NO−3 in
GW10 and GW15 were also highly denitrified. Although this DOC may not
be directly consumed during the denitrification process, the elevated DOC
content suggests that the hyporheic zones created beneath the leaking irriga-
tion channel may act as a pathway for labile OC into the groundwater system
where it can act an electron donor (Harvey et al., 2013; Hinkle et al., 2001).
There was no significant relationship between δ15N-NO−3 and δ34S-SO4
(Fig. 7b), suggesting autotrophic denitrification did not occur in the study
area (Hosono et al., 2014; Otero et al., 2009; Pauwels et al., 2010; Vitòria
et al., 2008). Furthermore, the positive δ34S-SO4 and δ18O-SO4 composi-
tions of all EIA groundwater samples greatly exceed those expected for sul-
fate derived from sulfide oxidation during autotrophic denitrification
(Vitòria et al., 2008).
Considered collectively, therewas no conclusive evidence of an electron
donor control on denitrification in the EIA groundwater system.
Overprinting from increased recharge render it difficult to rule out the oc-
currence of heterotrophic denitrification in the absence of other geochem-
ical data. However, there is some evidence to suggest that hyporheic
zones near leaking irrigation channels provide the necessary C for hetero-
trophic denitrification.5.5. Conceptual model and management implications
A conceptual model is shown in Fig. 12 that captures the origin and fate of
NO−3 across the EIA groundwater system. Fertiliser-derivedN is likely retained
in the SON pool prior to its mineralisation and subsequent nitrification, after
which it is leached to underlying groundwater. Within the groundwater sys-
tem, denitrification is strongly related to DO availability. At a regional-scale,
thismanifests in denitrification along the groundwaterflowpath as it becomes
progressively more oxygen-depleted and incorporates solutes. At a local-scale,
denitrification occurs in suboxic hyporheic zones beneath leaking irrigation
supply channels, independent of the regional groundwater flow.
What is the potential for NO−3 export from the aquifers of the EIA?
Waterlogging from rising shallow groundwater tables has previously been re-
ported in the extensively irrigatedfields overlying the basaltic soils in the EIA.
This is managed using tile and surface drainage to nearby tributaries13(Department of Natural Resources Queensland, 1998; Sunwater, 2012;
Yule, 1997). Irrigation farm drainage water in the EIA has been found to con-
sistently exceed Australian environmental guidelines for TN concentrations
(Department of Natural Resources Queensland, 1998). The elevated ground-
water NO−3 concentrations measured in this study indicate that the shallow
groundwater therefore remains an important transport mechanism for NO−3
export to surface waters, and more so under high rainfall or flood conditions.
Although denitrification was detected in hyporheic zones beneath leaking ir-
rigation channels, it is currently unknown to what spatial extent this may
occur throughout the EIA. Thus, denitrification should not be relied upon to
attenuate groundwater NO−3 at a regional-scale in this aquifer.
In contrast, standing water levels in the monitoring wells in the alluvial
aquifers underlying the Nogoa River remain below 10 m, despite water
levels having also risen ~5m since the 1980s. Aside from the EIA, ground-
water levels in the Lower Nogoa sub-catchment are mostly disconnected
from overlying fluvial systems (typically between 10 and 20 mbgs; Pearce
and Hansen, 2007; Queensland Government, 2020), making regional-
scale groundwater-surface water discharge in the Nogoa River alluvial
floodplain under current groundwater levels unlikely. Nonetheless, if
groundwater levels continue to rise, the alluvial aquifers risk becoming hy-
draulically connected. This could become an important pathway for NO−3
discharge into the Nogoa River and surrounding tributaries. Indeed,
groundwater discharge from the alluvial aquifer has been documented in
localised instances further east (downstream) of the study area (SLR
Consulting Australia, 2020). However, such groundwater expressions
have not been documented on a region-scale. Thus, catchment-wide trans-
fer of groundwaterNO−3 from the alluvial aquifer can be considered less im-
portant than in the basalt aquifer, especially given that denitrification
persists as groundwater flows into this aquifer.
While groundwater N may reflect past practices, considered collec-
tively, groundwater management in the EIA should focus on:
1. improved irrigation water management, including the lowering of shallow
groundwater tables of the basalt aquifer. This may have the benefit of de-
creasing mineralisation rates in soil profiles and decrease the rate at
which solutes aremobilised to the groundwater system. Furthermore, low-
ering groundwater levels in this aquifer would decrease the rate at which
water levels rise in the adjacent alluvial aquifers, albeit with a degree of
time lag; and
2. improved N fertiliser use-efficiency, to reduce the amounts of anthropo-
genic N reaching these aquifers via deep drainage. Current attempts to
re-line leaking irrigation channels (Sunwater, 2020) will likely assist in
lowering shallow groundwater tables.
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606
Fig. 12. Conceptual diagram (not to scale) of the processes leading to regional-scale denitrification and hyporheic zone denitrification beneath leaking irrigation channels.
Processes relating to N are shown in red, and hydrological processes indicated in blue. b.d.l: below detection limit.6. Conclusion
We applied multiple stable and radioactive isotope tracers to elucidate
the sources and hydro-geochemical controls on groundwater NO−3 cycling
in the EIA. Our isotopic data are consistent with a fertiliser source of
groundwater NO−3 , which is likely retained in the SON pool (via microbial
immobilisation) prior to being mineralised and nitrified. This NO−3 is deliv-
ered to EIA aquifers following the beginning of modern irrigation practices
in the mid-1970s. NO−3 isotopes indicate that denitrification is a prominent
feature of the EIA groundwater system. Denitrification manifests progres-
sively as groundwater flows from the upgradient oxic basalt aquifer to the
downgradient suboxic alluvial aquifer. In locations where groundwater is
hydraulically connected to leaking irrigation channels, dilution and denitri-
fication occurs in electron donor-rich suboxic hyporheic zones. This process
acts independently of the regional groundwater flow. Denitrification is not
likely to be coupled to sulfide oxidation. However, evidence for organic
matter oxidation during denitrification could not be decoupled from the
strong isotopic influence of infiltrating irrigation waters and natural
weathering processes. Importantly, the application of both 3H and 36Cl to
identify zones of channel leakage, as well as regional mixing processes,
was critical for enabling the de-coupling of local- (hyporheic zones beneath
leaking irrigation channels) and regional-scale (groundwater flow path) de-
nitrification.
The ability to successfully characterise (with strong statistical signifi-
cance) denitrification and associated enrichment factors using NO−3 iso-
topes in the EIA is noteworthy. This is likely due to the blended isotopic
signature of groundwater NO−3 resulting from the immobilisation of NO−3
in the SON pool prior to migration to the groundwater system, in addition
to there being a single predominantN source (N fertiliser) to EIA groundwa-
ter. To further confirm the role of SON retention in soils, more detailed
studies into SON content and the isotopic signature of both SON and differ-
ent crop types are required in the EIA. Further studies are also needed to
confirm the potential role of organic C in denitrification processes in
these aquifers.
In the predominantly oxic upgradient basalt aquifer, raised groundwa-
ter tables have created pathways for NO−3 to be transported to adjacent riv-
ers and tributaries. In contrast, in the alluvial aquifers of the Nogoa River
floodplain, the transfer of NO−3 is limited both physically (through14groundwater-surface water disconnection) and chemically (through deni-
trification along the groundwater flow path) in the study area. The lower-
ing of groundwater levels in the basalt aquifers and limiting future
groundwater table elevation in the alluvial aquifers, are thus important
steps in minimising the export of NO−3 from EIA groundwaters. A unified
approach focused on improving water use and fertiliser use efficiency in
the EIA is therefore needed tominimise any risk of fertiliser use on adjacent
and end of system ecosystems.
CRediT authorship contribution statement
Experimental conceptualisation and design was carried out by DIC,
BFJK and SJH. Fieldwork and in-field geochemical analyses were con-
ducted by all authors. Data analysis was primarily undertaken by SJH,
with guidance from DIC and BFJK. The manuscript was written by SJH
with input from all authors. Funding for the research was obtained by
DIC and BFJK.Declaration of competing interest
The authors declare that they have no conflict of interest.
Acknowledgements
This researchwas funded by the Cotton Research andDevelopment Cor-
poration (CRDC) via project ANSTO1801 “Quantifying the nitrogen cycle:
from farm gate to catchments, groundwater and atmosphere”. Stephen Harris
is supported by PhD scholarships from the Australian Government, the
Australian Institute of Nuclear Science and Engineering (AINSE),
Australian Nuclear Science and Technology Organisation (ANSTO) and
CRDC. ANSTO support and analytical staff are thanked for their continuous
efforts (Chris Dimovski, Klaus Wilcken, Henri Wong, Robert Chisari,
Jennifer van Holst). We thank Peter Voltz and Shane Answer from the
QLD Queensland Department of Natural Resources, Mines and Energy
groundwater monitoring network for facilitating access to monitoring
wells. Shane inspected sites with us, shared his local knowledge and ar-
ranged for key monitoring wells to be maintained/fixed ahead of our sam-
pling. Landholders who provided access to monitoring wells on their land
S.J. Harris et al. Science of the Total Environment 817 (2022) 152606are also thanked. Dr. Javier García Veigas from the CCiTUB Scientific and
Technological Center, Universitat de Barcelona is thanked for his expertise
and care taken in the analysis of sulfate isotopes samples.
Appendix A. Supplementary data
Supplementary data to this article can be found online at https://doi.
org/10.1016/j.scitotenv.2021.152606.References
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