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137
Journal of Hydrology (NZ) 61 (2): 137-150
© New Zealand Hydrological Society (2022)
Arsenic in groundwater in the Waitaha/Canterbury region
Andrew R. Pearson,1,2 Lisa Scott,1 Scott R. Wilson3 and
Michael S. Massey1
1 Environment Canterbury, 200 Tuam St., Christchurch 8011, Canterbury,
New Zealand.
2 Institute of Environmental Science and Research (ESR) 8041, Christchurch,
Canterbury, New Zealand.
3 Lincoln Agritech Ltd, PO Box 69-133, Lincoln 7640, Canterbury, New Zealand.
* Corresponding author: Andrew.Pearson@esr.cri.nz
Abstract
Arsenic contamination in groundwater
affects tens of millions of people worldwide
and is regarded as a major groundwater issue
in New Zealand. This study was conducted
to evaluate the connection between elevated
groundwater arsenic concentrations and
inferred redox state of aquifers across the
Canterbury region. In total, 2,428 samples
from 1,172 wells were analysed for arsenic
concentrations, and groundwater redox state
was inferred using geochemical indicators for
each well.
Most groundwater (~88% of wells;
N = 1,031) had arsenic concentrations below
laboratory detection limits, ~9.5% (N = 111
wells) had detectable arsenic concentrations
below the New Zealand drinking-water
Maximum Acceptable Value of 0.01 mg/L,
and ~2.6% (N = 30 wells) had concentrations
exceeding 0.01 mg/L.
Arsenic concentrations were strongly
associated with reducing groundwater;
47 (~43%) of 109 wells with reducing
groundwater had detectable concentrations
of arsenic, including 19 wells (~17%) with
arsenic greater than 0.01 mg/L. Conversely,
840 wells had water classified as oxic,
and only one of those wells exceeded the
0.01 mg/L Maximum Acceptable Value
for arsenic in drinking-water. Arsenic
concentrations were higher in coastal areas
north of Christchurch, where confined
aquifers are overlain and underlain by layers
of peat and low-permeability sediment.
Biogeochemical processes driving release
are the same as those that cause arsenic
contamination in groundwater elsewhere in
the world. However, differences in regional
geology and a relatively lesser availability of
labile organic carbon appear to constrain
the concentrations and spatial prevalence
of groundwater arsenic in Canterbury
compared to other arsenic-impacted regions
(e.g., Southeast Asia).
Keywords
arsenic, redox, groundwater, drinking water,
aquifer
Introduction
Arsenic (As) is a toxicant and carcinogen that
is found in soil, minerals, and groundwater
worldwide (Nordstrom, 2002; Podgorski and
Berg, 2020; Smedley and Kinniburgh, 2002).
Groundwater is the world’s largest source of
fresh water (McDonough et al., 2020), and
globally, consumption of naturally occurring
138
arsenic in groundwater affects 94 million
to 220 million people (Podgorski and Berg,
2020). The World Health Organization
(WHO) standard concentration for arsenic
in drinking-water is 0.01 mg/L (Gorchev and
Ozolins, 1984), which has been adopted by
New Zealand as the ‘Maximum Acceptable
Value’ (MAV) in drinking-water (Ministry of
Health, 2018).
Globally, most arsenic contamination
in groundwater is associated with arsenic
mobilisation due to natural processes
rather than human influence (Smedley
and Kinniburgh, 2002), although human
activities, such as mining, timber treatment
and pesticide use, have contributed to
arsenic pollution, including in New
Zealand (Kerr and Craw, 2021; Safa et
al., 2020; Robinson et al., 2004). Arsenic
(predominantly arsenate) is present in or
on aquifer minerals, including iron and
manganese oxide and hydroxide minerals,
which can undergo microbially-mediated
reductive dissolution when groundwater
is in a reducing redox state (Borch et al.,
2010; Postma et al., 2016). Thus, reductive
dissolution of arsenic-bearing minerals is
the primary trigger of arsenic release to
groundwater (McMahon and Chapelle,
2008; Smedley and Kinniburgh, 2002)
and is associated with elevated arsenic
concentrations in Argentina (Litter et al.,
2019), Bangladesh (Ying et al., 2017),
China (Zhang et al., 2020), India (Bindal
and Singh, 2019), Italy (Rotiroti et al.,
2015), the United States (Welch et al.,
2000) and Vietnam (Wallis et al., 2020).
Although reductive dissolution of arsenic-
bearing minerals is the primary driver of
arsenic release, groundwater redox state
can also influence the toxicity of arsenic;
under reducing conditions, arsenate can
be reduced to arsenite, which is more toxic
(Singh et al., 2011). After desorption from
aquifer minerals, arsenic concentrations are
dependent on hydrological factors, including
dilution and flushing rates (Fendorf et al.,
2010; Smedley and Kinniburgh, 2002).
Arsenic is widespread in New Zealand’s
aquatic and terrestrial environments
(Robinson et al., 2004), and arsenic
contamination is regarded as a major
groundwater issue (Daughney and Wall,
2007). Concentrations above the drinking-
water MAV have been observed in Waikato
(Piper and Kim, 2006), Otago (Levy et al.,
2021) and Canterbury (Scott et al., 2016).
In Canterbury, arsenic contamination has
been observed in reducing redox-state
groundwater in some coastal areas (Scott et
al., 2016), but this study represents the first
regional-scale analysis of the relationship
between groundwater redox state and arsenic
concentrations. To evaluate the relationship
between groundwater redox state and
arsenic contamination, we assessed arsenic
concentration data from 1,172 wells, and
used geochemical indicators to estimate
groundwater redox state for each well,
following the approach used by Close et al.
(2016) and Wilson et al. (2020).
The Canterbury region
The Canterbury (or Waitaha) region is
located on the east of Aotearoa New Zealand’s
South Island and is bounded by the Southern
Alps to the west and the Pacific Ocean to the
east (Fig. 1). Canterbury is New Zealand’s
largest geographic region (approximately
50,000 km2), occupying approximately 17%
of New Zealand’s land area (Painter, 2018).
In Canterbury, most groundwater is found
in alluvial sediments (predominantly gravels)
underlying the Canterbury Plains, which
are one of the world’s major alluvial plains
(Leckie, 2003). The groundwater contained
in Canterbury aquifers represents 73% of
New Zealand’s total groundwater by volume
(Stats NZ, 2017), is the primary source
of drinking-water to a population of over
600,000 (including the city of Christchurch)
and provides water for 70% of New Zealand’s
irrigated land (Painter, 2018).
139
Canterbury has a temperate maritime
climate (Macara, 2016). Groundwater
recharge is heavily influenced by interactions
between the Southern Alps and westerly
winds; orographic rainfall in the Alps is a
significant source of recharge for rivers and
groundwater. From their sources in the
Southern Alps, groundwater and braided
rivers flow dominantly eastward across the
Canterbury Plains before draining into
the South Pacific Ocean. The Southern
Alps are primarily composed of greywacke,
tectonically uplifted by the collision of the
Australian and Pacific plates (Browne and
Naish, 2003). In some areas of the Southern
Alps, arsenic concentrations are elevated
in greywacke and argillite (Becker et al.,
2000; Horton et al., 2001) and present in
other minerals (e.g., calcite, quartz, and
sulphides) (Horton et al., 2001; Craw et
al., 2002). For example, in the Wilberforce
valley, arsenic concentrations have been
measured at 200 mg/kg (compared to a
background concentration of 10 mg/kg)
in some greywackes and argillites (Becker
et al., 2000). Weathering of the Southern
Alps, followed by the erosion and eastward
transport of sediment toward the coast, led
to the formation of a thick assemblage of
overlapping alluvial fans, which formed the
Canterbury Plains (Ballance, 2017). This
process has been occurring for at least the
last five million years, with increased erosion
and deposition during glacial periods
(Ballance, 2017). As a result of this erosion
and transport of arsenic-bearing bedrock
material, geogenic sources of arsenic may
be widespread in the alluvial gravels, with
arsenic generally present in source rocks
and associated sedimentary deposits.
The mostly gravel aquifers of the upper
plains and river valleys of Canterbury are
dominated by oxic groundwater conditions
(Wilson et al., 2020).
In some coastal areas east of the plains
there are series of confined aquifers (Fig. 1),
which typically sit in glacio-fluvial gravels
interbedded between aquitards formed from
silts, clays, sands, and buried layers of peat
extending to a depth of over 200 m (Brown
et al., 1988). The silts, clays, and sands were
deposited in low-energy marine or estuarine
environments that occurred during marine
high stands during interglacial periods
through the Quaternary (Browne and Naish,
2003). Groundwater is also found in some
inland basins, which were formed by glacial
activity and tectonic movement and partially
infilled with glacio-fluvial sediments. In
some parts of south Canterbury, known as
the ‘downlands’, groundwater is overlain
by layers of loess (of varying thickness),
which were deposited during the Pleistocene
(Schmidt et al., 2005) (Fig. 1). In some
coastal areas, particularly to the north of
Christchurch and in some parts of southern
Canterbury, reducing redox state conditions
can be prevalent in groundwater (Close et al.,
2016; Wilson et al., 2020).
Methods
Arsenic data analysis
Our data set includes 2,185 groundwater
arsenic measurements, sampled from 1,172
wells (tested from 1 to 25 times, though most
wells were only sampled once) from 1989 to
2021. The wells used in this study cover all
the main aquifer types found in Canterbury
(Fig. 1), and well depths ranged from 1.1 m to
433 m below ground level (mean depth of 33
m; median depth of 20 m). For wells sampled
more than once, mean concentrations of
arsenic were calculated for each well.
Samples were collected for various
reasons, as part of long-term monitoring
programmes or for specific investigations.
Due to changing laboratory service providers
and improving analytical technology over the
course of 30+ years, analytical methods used
for arsenic measurements changed through
the period of sample collection. In summary,
140
arsenic concentrations were predominantly
measured by hydride generation atomic
absorption spectroscopy (AAS) from 1989
to 1995; graphite furnace atomic absorption
spectroscopy (GFAAS) from 1995 to 2000,
and inductively coupled plasma mass
spectrometry (ICPMS) from 2000 to 2021.
Laboratory measurements were undertaken
at IANZ accredited laboratories. Detection
limits ranged from 0.0005 to 0.01 mg/L,
though reported ‘non-detections’ in samples
with detection limits of 0.01 mg/L (i.e., the
Figure 1 – Main aquifer types of Canterbury region (inset figure shows New
Zealand) include coastal confined aquifers, plains aquifers dominated by gravels,
downlands aquifers overlain by loess, fluvial aquifers associated with river systems,
and inland basins. Broadly, groundwater flows from the Southern Alps toward the
east coast, before draining into the South Pacific Ocean.
drinking-water MAV) were excluded from
this study.
Groundwater samples were collected
by trained sampling technicians following
groundwater sampling procedures based
on international best practice. From 2006,
sample collection followed a national protocol
for state of the environment groundwater
sampling (Ministry for the Environment,
2006) later replaced by the National
Environmental Monitoring Standard for
discrete groundwater quality sampling
141
(NEMS, 2019). Over time, Environment
Canterbury’s approach to measuring
groundwater arsenic concentrations varied
between measurements of dissolved arsenic
and total arsenic in groundwater, depending
on whether the technician filtered the
samples before analysis. In New Zealand, the
MAV refers to arsenic concentration (i.e., it
does not distinguish between dissolved and
total arsenic). Measurements of total arsenic
may capture a more complete picture of the
actual amount of arsenic, so we have used
‘total’ arsenic concentrations if available
and dissolved arsenic concentrations
otherwise. This study aims to assess arsenic
concentrations relative to the drinking-water
MAV (Ministry of Health, 2018); therefore,
we have not distinguished ‘dissolved’ from
‘total’ arsenic in this study.
Redox indicators
Redox state for each well was estimated using
a total of 19,687 groundwater samples. After
Close et al. (2016) and Wilson et al. (2020)
we broadly classified groundwater redox state
into three categories: ‘oxic’ (or ‘oxidising’),
‘mixed’, and ‘reducing’. Redox state can be
estimated by measuring concentrations of
redox-sensitive parameters, such as dissolved
O2 (DO), nitrate (NO3−), manganese (Mn),
and iron (Fe) (Close et al., 2016). For example,
under reducing conditions DO and NO3−
will have low concentrations (i.e., < 1.0 mg/L
for DO, < 0.5 mg/L for nitrate-N), whilst Fe
and Mn might have elevated concentrations
(i.e., > 0.1 mg/L for Fe; > 0.05 mg/L for Mn)
due to reductive dissolution and release of
iron and manganese from aquifer sediments.
Following Close et al. (2016) groundwater
was classified as ‘mixed’ when nitrate-N,
Fe, Mn, and DO concentrations provided
an inconsistent indication of redox state.
Inconsistent indicators may be observed
because the sample was taken from a well that
samples multiple redox zones, the sample
was in disequilibrium, or because of the
thresholds applied to estimate redox state.
Our study used the same redox parameters
as Wilson et al. (2020), but rather than
mapping redox state, arsenic concentrations
in individual wells were compared against
redox state estimated from parameters in the
same well. This approach enabled a direct
comparison of groundwater redox state
against arsenic concentrations in individual
wells. We compared the distribution of the
redox state classes for wells with no arsenic
detections, wells with detectable arsenic
concentrations, and wells with arsenic
concentrations exceeding the drinking-
water MAV of 0.01 mg/L for arsenic. Due
to sporadic measurements of individual
geochemical indicators through time, the
redox estimates were based on the mean of
each redox indicator measurement as the
best estimate of long-term status of a given
well. Therefore, redox estimates were not
exclusively linked to the same samples used
for arsenic measurements. One confounding
factor resulting from the use of mean values
over time is that temporal variations in redox
state were not considered in this work.
Results
Arsenic and redox data
Of the 1,172 wells sampled for arsenic, most
wells (88%) had arsenic concentrations
below detection limits, 9.5% (N = 111) had
detectable values below the drinking-water
MAV (0.01 mg/L), and 2.6% (N = 30)
had arsenic values above the drinking-water
MAV (Table 1). Most wells contained oxic
groundwater (N = 840; 71.7%), groundwater
was mixed in 223 wells (19.0%), and
reducing groundwater was observed in 109
(9.3%) wells (Table 1).
Arsenic presence and exceedances of the
drinking-water MAV were strongly influenced
by groundwater redox state (Table 1).
In oxic groundwater, 94.9% of wells did not
have detectable concentrations of arsenic and
5% of wells had arsenic detected but below
142
the MAV; only one well exceeded the arsenic
drinking-water MAV (Table 1). In mixed
groundwater, 4.5% of wells had arsenic
concentrations above the drinking-water
MAV and 18.4% of wells had detectable
concentrations below the MAV. Still, in
mixed groundwater, most wells (77.1%) did
not have detectable concentrations of arsenic.
By contrast, in reducing groundwater,
Table 1 – Exceedances of the arsenic drinking-water Maximum Acceptable Value (MAV)
(0.01 mg/L), arsenic detections, and arsenic concentrations below the laboratory
detection limit in relation to predicted groundwater redox state.
Arsenic
concentration
category
Number of wells
(% of total wells)
Count of wells per redox classification
(% of total wells in each redox state class)
Oxic Mixed Reducing
≥ MAV 30 (2.6%) 1 (0.1%) 10 (4.5%) 19 (17.4%)
Detected below MAV 111 (9.5%) 42 (5.0%) 41 (18.4%) 28 (25.7%)
Not detected 1031 (87.9%) 797 (94.9%) 172 (77.1%) 62 (56.9%)
Total wells: 1172 840 223 109
17.4% of wells had concentrations of arsenic
exceeding 0.01 mg/L, whilst a further
25.7% had detectable concentrations of
arsenic below the MAV. Most wells with
reducing groundwater (56.9%) did not have
detectable concentrations of arsenic (Table 1),
but the differences between oxic and
reducing groundwater with respect to arsenic
concentrations were nonetheless evident.
Figure 2 – (a) Average groundwater arsenic concentrations in individual wells (N = 1,172) in the
Canterbury region, compared to the MAV (0.01 mg/L); (b) Redox state per individual well in the
Canterbury region. Redox state was estimated for each well using the same parameters (dissolved
oxygen, nitrate, iron, and manganese) and approach as Wilson et al. (2020). Canterbury’s main
aquifer types are also shown in each map.
143
Arsenic and redox state across Canterbury
In the inland Canterbury Plains, the
groundwater is dominated by oxic conditions
(Fig. 2b), with almost no exceedances of the
arsenic drinking-water MAV of 0.01 mg/L.
Although a geogenic arsenic source is present,
groundwater redox conditions do not enable
arsenic mobilisation to occur. Similarly, in
the inland basin aquifers, oxic conditions
are prevalent (Fig. 2b), with arsenic
concentrations generally below laboratory
detection limits (Fig. 2a).
By comparison, arsenic in groundwater
is more commonly found in low-lying
coastal areas (Fig. 2a), which in Canterbury
are typically underlain by confined aquifer
systems. Measured arsenic concentrations
in coastal aquifers ranged from values below
detection to a maximum concentration
of 0.92 mg/L (i.e., 92 times the MAV for
drinking-water). In the aquifers underlying
the South Canterbury downlands, reducing
conditions are more common, with some
exceedances of the drinking-water MAV for
arsenic, though arsenic was not detected in
most wells.
The spatial scale of elevated arsenic
concentrations (Figs. 2a and 3) indicate that
groundwater arsenic is diffusely distributed in
some coastal areas of Canterbury. However,
there is also spatial variability, particularly
within the reducing coastal aquifers. For
example, north of Christchurch there are
wells where arsenic has not been detected in
relative proximity (i.e., within hundreds of
metres) to wells where arsenic concentrations
exceeded the drinking-water MAV (Fig. 3).
Figure 3 – Arsenic concentrations relative to the MAV (0.01 mg/L) in an area
of North Canterbury (black rectangle in inset map shows location within
Canterbury). Groundwater arsenic concentrations are generally higher in
reducing aquifers, whilst inland oxic aquifers have lower concentrations and
fewer occurrences of arsenic. *Aquifer redox map reproduced from Wilson et al.
(2020) using data kindly supplied by the authors.
144
Discussion
Relatively higher risk of arsenic contamination
in coastal aquifers
Globally, reducing conditions tend to exist
within geologically young aquifers in low-
lying areas with low flushing rates (Smedley
and Kinniburgh, 2002), such as the coastal
confined aquifers of Canterbury. Elevated
arsenic concentrations are very strongly
associated with mixed or reducing redox
state groundwater in Canterbury (Table 1),
which is found primarily in the coastal
confined aquifer systems and the South
Canterbury downlands (Fig. 2b). In reducing
groundwater systems, the drinking-water
MAV for arsenic was exceeded in ~17% of
wells, with a further ~26% of wells having
detectable concentrations below the MAV.
Further, arsenic was also present in mixed
redox state groundwater, with 4.5% of wells
exceeding drinking-water MAV and a further
~18% of wells having detectable arsenic
concentrations below the MAV (Table 1).
These data show that arsenic is a potential
indicator of reducing conditions but reducing
conditions do not necessarily indicate the
presence of arsenic.
The prevalence of reducing conditions
in coastal areas is due to the presence of
in situ organic matter and confinement
of aquifers by low-permeability material
deposited by marine transgressions during
the Quaternary (Brown and Weeber, 1992).
Low-permeability material inhibits transport
of oxygen from the atmosphere, recharge,
and flow, potentially leading to a build-
up of arsenic through time (Fendorf et al.,
2010). Further, layers of subsurface peat
that overlie or underlie the coastal confined
aquifers (Brown et al., 1988) are a source of
labile organic matter, which is often a critical,
limiting component of reductive dissolution
of arsenic-bearing minerals in groundwater
(deLemos et al., 2006; Fendorf et al., 2010;
Islam et al., 2004). Organic matter release
is known to fuel microbially-mediated
reduction processes that release arsenic into
groundwater in countries where arsenic
contamination in groundwater is much more
common and has been extensively studied,
such as Bangladesh, Cambodia, and Vietnam
(Erban et al., 2013; Fendorf et al., 2010;
Harvey et al., 2002; Stuckey et al., 2016).
The presence of organic matter (which drives
microbial respiration) depletes dissolved
oxygen in groundwater, causing groundwater
to become suboxic or anoxic, which from
a chemical perspective is a ‘reducing’
environment. Under reducing conditions,
other favourable available electron acceptors
can be used in microbial respiration, such
as nitrate, manganese, iron, sulphate, and
carbon dioxide, and many trace metals and
metalloids, such as arsenic, chromium, and
uranium (Borch et al., 2010; McMahon and
Chapelle, 2008). Since arsenic commonly
adsorbs to the surface of iron (hydr)oxide
minerals or can be found as trace contaminants
within these oxides, desorption and release of
arsenic can occur when reduction of iron and
manganese oxides is favourable. Thus, the
occurrence of arsenic release in aquifers with
in situ sources of organic matter and reducing
groundwater aligns with expectations based
on the biogeochemical behaviour of arsenic
observed elsewhere in the world.
Local-scale variability is common in arsenic-
impacted groundwater (Jakobsen et al., 2018;
van Geen et al., 2003), and our sample data
also reveal complex patterns across small
spatial scales in some areas of Canterbury.
For example, in the reducing coastal confined
aquifers north of Christchurch there are wells
in which arsenic was not detected, less than
one hundred metres from wells where arsenic
concentrations exceeded the MAV. Although
redox state influences arsenic release,
variations in the source of water and the three-
dimensional patterns of groundwater flow
influence patterns of both redox state and
145
arsenic contamination, which are typically
complex in arsenic-impacted areas (Neumann
et al., 2010). For example, in some areas
of the coastal confined aquifers, mixed and
oxic groundwater is predicted (Wilson et al.,
2020), possibly due to preferential upwards
leakage though aquitards, whilst seepage
from the Waimakariri River (Fig. 3) supplies
groundwater flow towards Christchurch
(Stewart and van der Raaij, 2022), potentially
diluting arsenic and supplying oxygen-rich
water. Thus, in Canterbury’s coastal aquifers,
a combination of local-scale geochemical and
hydrological variability means that arsenic-
contaminated groundwater is not uniformly
observed.
Oxic groundwater and low arsenic
concentrations in the Canterbury Plains
In Canterbury’s oxic groundwater, 95% of
wells had arsenic values below detection
limits, with 5% of wells having arsenic
detections and one well (0.1% of oxic wells)
exceeding the drinking-water MAV of
0.01 mg/L (Table 1). Most of the groundwater
in the unconfined aquifers of the Canterbury
Plains and inland basins is oxic (Close et al.,
2016; Wilson et al., 2020) because the open-
framework gravels that form the unconfined
aquifers, and the thin soils that overlie
them, allow land-surface recharge (i.e., the
sum of irrigation and rainfall recharge) and
transport of oxygen from the atmosphere to
groundwater. Further, in some areas, recharge
is dominated by seepage from the gravel-bed
braided rivers, which transport oxygen-rich
water from the foothills and Southern Alps.
The lack of detectable concentrations of
arsenic in oxic zones of the Plains (despite
arsenic being present in the greywacke, rocks,
and minerals of the Alps that were eroded to
form them) implies that arsenic may be stored
in oxic aquifer sediments, or when arsenic is
released, it might be diluted to concentrations
below laboratory detection limits. Further,
significant parts of the Canterbury Plains
are dominated by agricultural activity
(Stewart and Aitchison-Earl, 2020), causing
unconfined gravel aquifers to have excessive
nitrate concentrations. Nitrate is a powerful
oxidant (Galloway et al., 2003), and the
diffuse addition of anthropogenic nitrate,
a more favourable electron acceptor than
manganese or iron (McMahon and Chapelle,
2008), might further limit reduction of
arsenic-bearing iron and manganese minerals
in the aquifers of the Canterbury Plains.
In the aquifers of the Canterbury Plains,
arsenic concentrations could be low because
of a higher rate of sorption rather than
desorption, which is influenced by redox
state, although dilution could also be a factor
due to high flow rates (Postma et al., 2017;
Sø et al., 2018). For example, the much
larger and fast-flowing Plains’ oxic aquifers
can dilute dissolved organic carbon and
competing adsorbates (e.g., phosphate) whilst
transporting higher concentrations (and total
loads) of oxygen and nitrate, moderating
arsenic release from aquifer minerals (Sø et
al., 2018). Thus, in aquifers on the Plains,
mass loads of arsenic could be high but very
difficult to spot in groundwater: rates of
sorption to aquifer minerals are likely to be
higher than the rate of arsenic release (owing
to the oxic redox state), whilst the relatively
high flow rates mean that any desorbed
arsenic may be diluted to concentrations
below detection limits. For example, Burbery
et al. (2020) installed a denitrifying bioreactor
in a fast-flowing (hydraulic conductivity of
1332 m/d) shallow unconfined gravel aquifer,
in which woodchips within the bioreactor
released labile organic carbon. The release
of labile organic carbon drove ‘pollution
swapping’, whereby denitrification took place
under reducing conditions, but manganese,
iron, and arsenic were released from aquifer
minerals (Burbery et al., 2022). Notably,
although concentrations of arsenic reached
a maximum of 0.022 mg/L (i.e., more than
double the drinking-water MAV), detectable
146
concentrations of arsenic travelled less
than 60 m down-gradient before dilution
or attenuation by mechanisms such as
sorption (Burbery et al., 2022). Thus, in
Canterbury, arsenic release appears to be
heavily influenced by redox state, yet on
the Plains, high concentrations of geogenic
arsenic in groundwater are likely only occur
on local scales, where there is a source of
organic carbon. In most of Canterbury, high
arsenic concentrations are likely to be diluted
or attenuated to concentrations below
laboratory detection limits within a relatively
short distance (e.g., Burbery et al., 2022).
Future research
More frequent monitoring would improve
understanding of arsenic behaviour in
Canterbury’s groundwater. For example,
most wells have only been sampled once,
yet regular sampling of arsenic-affected wells
would allow assessment of seasonal and
long-term (i.e., decadal) variability and any
potential trends. Analysis of arsenic bound
to aquifer sediments could shed light on
mechanisms of arsenic retention and release,
though these mechanisms are expected to be
similar to those observed elsewhere the world.
In the coastal confined aquifers, labile organic
carbon is likely to be an important influence
on the spatial variability observed in the
coastal confined aquifers, thus measurements
of dissolved organic carbon could be of value.
From a monitoring perspective, the spatial
variability of arsenic concentrations in some
coastal areas means that it is likely to be
difficult to reliably predict the concentration
of arsenic in a particular well based on the
results from neighbouring wells. Analysis of
arsenic speciation in groundwater could also
be valuable because, although the drinking-
water MAV does not discriminate between
arsenic species, arsenite is much more toxic
than arsenate (Singh et al., 2011).
Ongoing and future human-induced
changes could impact arsenic contamination
on local scales, even in the oxic aquifers of
the Plains. Although arsenic contamination
is largely of geogenic origin, there are
numerous anthropogenic arsenic sources
(e.g., timber treatment plants, landfills,
sheep dip sites) in Canterbury, whilst human
activities can also alter groundwater redox
state and flow behaviour. For example,
irrigation pumping and return flow can
alter recharge rates, mixing, and natural flow
patterns, which can influence geochemical
conditions. Further, the release of organic
matter (e.g., from manure spreading,
constructed ponds, septic tanks, or landfill
sites) can cause groundwater redox state to
become reducing, whilst phosphate (released
from fertiliser) can compete with arsenate for
binding sites on aquifer minerals (Fendorf
et al., 2010; Lin et al., 2016; Smedley and
Kinniburgh, 2002), though the impact of
these activities on arsenic release is not well
studied in Canterbury.
Conclusions
In Canterbury, arsenic in groundwater
originates from greywacke and fault-fluid
minerals in the Southern Alps, before being
fluvially transported coastward by way of
the Canterbury Plains. Region-wide, most
wells (88%) had groundwater with arsenic
concentrations below detection limits, even in
wells with reducing redox state groundwater
(56.9%). Arsenic contamination was very
rare in the mostly oxic groundwater of the
inland basins and the Canterbury Plains,
because of dilution and oxic conditions.
Conversely, in some coastal areas, there
are diffuse sources of arsenic and conditions
favourable to arsenic release to groundwater.
These processes contributed to exceedances
of the drinking-water MAV for arsenic in
2.6% of wells sampled across the region in
the period from 1989 to 2021. In reducing
groundwater, which is relatively common
in the coastal confined aquifers, arsenic
147
concentrations were highly variable at
relatively small spatial scales. The observed
variability is likely due to a combination of
the presence of arsenic-bearing iron oxides,
the availability of labile organic carbon,
differences in local-scale groundwater
residence times, and sluggish flow rates.
Thus, arsenic concentrations in Canterbury
groundwater are dependent on the presence
of arsenic sources, groundwater redox state,
and the influence of physical hydrological
factors such as mixing and dilution.
Acknowledgements
Andrew Pearson was affiliated with
Environment Canterbury at the time most
of this work was undertaken; however, this
manuscript was completed after he became
affiliated with the Institute of Environmental
Science and Research (ESR). The authors
wish to thank Environment Canterbury’s
Groundwater Science field team for sample
collection, and Philippa Aitchison-Earl
(Environment Canterbury) for providing
the map of aquifer types. Gratitude is also
due to Jennifer Tregurtha for advice with
map production and to Kurt Van Ness and
Ben Wilkins (Environment Canterbury)
for assistance preparing the dataset. The
authors also wish to thank Murray Close of
ESR for sharing the methods for estimating
groundwater redox state for individual wells.
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Manuscript received 16 February 2022; accepted for 19 April 2022