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PAPER
An overview of dissolved organic carbon in groundwater
and implications for drinking water safety
S. Regan
1
&P. Hynds
2
&R. Flynn
3
Received: 16 September 2016 / Accepted: 23 March 2017 /Published online: 22 April 2017
#Springer-Verlag Berlin Heidelberg 2017
Abstract Dissolved organic carbon (DOC) is composed of a
diverse array of compounds, predominantly humic sub-
stances, and is a near ubiquitous component of natural ground-
water, notwithstanding climatic extremes such as arid and
hyper-arid settings. Despite being a frequently measured pa-
rameter of groundwater quality, the complexity of DOC com-
position and reaction behaviour means that links between con-
centration and human health risk are difficult to quantify and
few examples are reported in the literature. Measured concen-
trations from natural/unpolluted groundwater are typically be-
low 4 mg C/l, whilst concentrations above these levels gener-
ally indicate anthropogenic influences and/or contamination
issues and can potentially compromise water safety. Treatment
processes are effective at reducing DOC concentrations, but
refractory humic substance reaction with chlorine during the
disinfection process produces suspected carcinogenic disin-
fectant by-products (DBPs). However, despite engineered ar-
tificial recharge systems being commonly used to remove
DOC from recycled treated wastewaters, little research has
been conducted on the presence of DBPs in potable ground-
water systems. In recent years, the capacity to measure the
influence of organic matter on colloidal contaminants and its
influence on the mobility of pathogenic microorganisms has
aided understanding of transport processes in aquifers.
Additionally, advances in polymerase chain reaction tech-
niques used for the detection, identification, and quantification
of waterborne pathogens, provide a method to confidently
investigate the behaviour of DOC and its effect on contami-
nant transfer in aquifers. This paper provides a summary of
DOC occurrence in groundwater bodies and associated issues
capable of indirectly affecting human health.
Keywords Dissolved organic carbon .Health .Disinfectant
by-products .Microbial processes .Polymerase chain reaction
Introduction
Numerous studies have examined the occurrence, source, re-
activity, and transport of natural dissolved organic carbon
(DOC) in aquifers (Baker et al. 2000; Chapelle et al. 2009;
Longnecker and Kujawinski 2011;Shenetal.2015); howev-
er, to date, few have focused on the indirect associations be-
tween DOC and human health. This is in part due to the
diverse nature of organic compounds frequently encountered
in both polluted and unpolluted groundwater, in addition to
their variable levels of reactivity. While the precise composi-
tion of groundwater DOC is both temporally and spatially
variable (Clay et al. 1996; Goñi and Gardner 2003;Shen
et al. 2015), it typically comprises a relatively small propor-
tion of low molecular weight compounds (i.e. carbohydrates
and amino acids), and a larger proportion of more complex,
high molecular weight compounds (i.e. humic substances;
Artinger et al. 2000).
Organic matter of biological origin is the most frequently
oxidised substance in the aqueous environment and, due to the
reactive nature of a subset of biomolecular components
(Aiken 2002) groundwater DOC is indicative and/or causative
Published in the special issue BHydrogeology and Human Health^
*S. Regan
regans@tcd.ie
1
Department of Civil, Structural and Environmental Engineering,
University of Dublin Trinity College, Dublin 2, Ireland
2
Environmental Health and Sustainability Institute, Dublin Institute of
Technology, Graingegorman, Dublin 7, Ireland
3
School of the Natural and Built Environment, David Keir Building,
Queen’s University Belfast, Stranmillis Road, Belfast BT9 5AG, UK
Hydrogeol J (2017) 25:959–967
DOI 10.1007/s10040-017-1583-3
with respect to a number of subsurface microbial and geo-
chemical processes. Some of these processes are considered
beneficial, e.g. denitrification (Clay et al. 1996; Pabich et al.
2001; Thayalakumaran et al. 2008); however, DOC enrich-
ment directly affects microbial oxygen availability and, thus,
subsurface microbial survival and mobility (Aravenaa et al.
1995; Chapelle et al. 2013; Thayalakumaran et al. 2015).
Moreover, high levels of dissolved organic matter (DOM)
present in groundwater may decrease optical clarity (or in-
crease turbidity), which can significantly reduce the efficacy
of specific treatment processes (e.g. irradiation) and compli-
cations can arise with the formation of DOC by-products fol-
lowing disinfection. Similarly, the presence of high levels of
DOC typically concurs with high concentrations of humic and
fulvic acids (Artinger et al. 2000), potentially resulting in in-
hibition of molecular laboratory analyses, most notably poly-
merase chain reaction (PCR) and enhanced contaminant mo-
bility; though in some polluted waters, e.g. sewage effluent,
proteins and carbohydrates may dominate DOC composition
(Flynn et al. 2015).
This paper provides a summary of DOC occurrence in
groundwater, followed by a brief overview of some issues
capable of indirectly affecting human health. Finally, a sum-
mary of where future research efforts might be focused with
respect to groundwater DOC and human health is provided.
Dissolved organic carbon in groundwater
DOC represents the carbonic fraction of dissolved organic
matter (DOM), typically accounting for >90% of the total
organic carbon (TOC) budget found in natural groundwaters
(Batiot et al. 2003a), with particulate organic carbon (POC)
accounting for the remainder. Analytically, DOC is
characterised as the organic fraction capable of passing
through a 0.45-μm filter (Evans et al. 2005; Kolka et al.
2008). Measured DOC concentrations in groundwater have
been shown to cover a wide range, from below detection
limits to >50 mg/l (Thurman 1985; Evans et al. 2005). As
shown in Table 1, a DOC range of 0.1–4 mg C/l is frequently
encountered in unpolluted aquifers, however, significantly
lower levels of DOC may be associated with aquifers receiv-
ing recharge where an organic horizon in a soil profile is
absent (to less than detection limits, typically in the order of
0.006–0.05 mg/l) . Conversely, uncontaminated rivers are typ-
ically associated with DOC ranges of 1–10 mg C/l (Tao 1998),
although, as might be expected, elevated concentrations are
found in runoff from catchments characterised by extensive
areas of organic soils, e.g. peatlands (Kolka et al. 2008). Thus,
while DOC may be described as being relatively stable, in
both time and space, within many groundwater environments
(Pabich et al. 2001), this is not always the case, particularly
where significant surface-water/groundwater interaction
occurs, e.g. karst systems; accordingly, caution should be
employed in its interpretation.
Low DOC concentrations frequently associated with
groundwater contrast with those in pore water from overlying
soil horizons, with DOC ranges of 10–200 mg C/l typically
reported (Wassenaara et al. 1991;Goldscheideretal.2006). It
is in the unsaturated subsurface zone that DOC is leached
from decomposing organic materials, with the principal bio-
geochemical processes affecting DOC removal in the vadose
zone being microbial utilization, oxidation, adsorption, and
metal co-precipitation (Wassenaara et al. 1991; Shen et al.
2015). These processes influence the transport of soluble or-
ganic carbon to and below the water table, in parallel with
environmental, climatic and geological factors. Typically, ap-
proximately 90% of surface-derived DOC is removed prior to
reaching the saturated zone (Pabich et al. 2001;Shenetal.
2015). Several studies have reported decreasing DOC concen-
trations in concurrence with increasing subsoil thickness and
aquifer depth (Pabich et al. 2001; Goñi and Gardner 2003;
Datry et al. 2004;Goldscheideretal.2006), with DOC con-
centrations at or close to zero reported in deep (> 1 km) and
old groundwaters (Pabich et al. 2001). Consequently, shallow
groundwater DOC characteristics are largely determined by
processes operating in the soil organic horizon (Trumbore
et al. 1992; Kalbitz et al. 2000), while the transfer of organic
matter remaining in solution is primarily dependent on vadose
zone thickness (Pabich et al. 2001; Batiot et al. 2003b;Goñi
and Gardner 2003), and the level of hydraulic connectivity
between subsoil horizons and the water table (Kalbitz et al.
2000).
The hydrological connectivity between the surface and
subsurface is important for the transfer of DOC into shallow
groundwater ecosystems, which will also be influenced by soil
properties/structure and land management practice. Whilst el-
evated groundwater DOC (generally considered to be approx-
imately 4 mg C/l) can occur naturally due to recharge being
influenced by discharge originating from wetland environ-
ments and the presence of sedimentary organic deposits, such
as coal, at depth (e.g. values of up to 200 mg C/l reported;
Artinger et al. 2000), high DOC is often associated with an-
thropogenic activities (Chomycia et al. 2008;Table1).
Additional organic matter from the increased use of organic
fertilisers, such as manure, has been reported from dairy farm
operations resulting in DOC concentrations up to 55 mg C/l
(Chomycia et al. 2008) and irrigated sugar cane fields (up to
82 mg C/l), due to organic-rich sugar mill by-products applied
as fertilizer and/or sugarcane sap released during harvest
(Thayalakumaran et al. 2015). DOC may also be sourced from
the hyporheic zone—for example, previous studies suggest
that surface water containing untreated waste has driven a rise
in groundwater DOC levels up to 14 mg/l in the Bengal aqui-
fer system of Bangladesh due to groundwater pumping from
irrigated croplands during the dry season (Bhattacharya et al.
960 Hydrogeol J (2017) 25:959–967
2002;Harveyetal.2002). In extreme cases, aquifers polluted
by leachate from waste disposal such as a landfill, can contain
DOC concentrations as high as 300 mg/l (Christensen et al.
1999). Similarly, contamination of groundwater by leachate
from historical industrial activities, such as tannery facilities,
has been reported as the source of significant groundwater
plumes with DOC levels to as high as 400 mg/l (Davis et al.
1994).
Groundwater TOC and human health
The chemical composition of groundwater reflects the com-
position of water entering the subsurface, in addition to the
products of kinetically controlled reactions within the aquifer
matrix and overlying subsoil layers (Grützmacher et al. 2013).
The composition and bioavailability of DOM are key factors
affecting water quality, where observed increases in DOC
concentration may be indicative of changes in water quality.
For example, concentrations above background levels can
provide evidence of the occurrence of groundwater contami-
nation by organic compounds (Barcelona 1984;Harveyand
Barber 1992; Longnecker and Kujawinski 2011). Whilst there
are a number of issues that may be associated with TOC in
groundwater and human health, such as metal mobility
(Christensen et al. 1999), the following sections provide an
overview of three rarely examined issues with regard to
groundwater DOC measurements and their interpretation
within the context of human health.
Groundwater DOM, treatment approaches and efficacy
High levels ofDOM in water are known to cause aesthetic and
odour problems, while possibly promoting the growth and
proliferation of pathogenic bacteria (Pernthaler 2005;
Goldscheider et al. 2006; Gopal et al. 2007). Unpolluted
groundwater generally lacks any discernible colour and is
therefore considered optically clear (Chapelle et al. 2016),
though not necessarily in dynamic groundwater/surface-
water systems such as karst. However wastewater ingress to
groundwater recharge, due to an absence of natural subsurface
attenuation and/or localised preferential pathways (e.g. poorly
located or malfunctioning septic tanks), can increase ground-
water DOC concentrations and thus cause (dis)coloration via
elevated levels of humic materials (Westerhoff and Pinney
2000; Rittmann et al. 2002). The solubilization of heavy
metals by complexation with humic substances is known to
be of considerable importance in coloured natural waters
Tabl e 1 Selected groundwater
DOC concentrations from
unpolluted and polluted/enriched
aquifers
Author/s DOC (mg C/l) Aquifer Impact
Unpolluted
Artinger et al. 2000 0.3–1.5 Glacial sediment –
Artinger et al. 2000 0.1 to 0.5 Sandstone bedrock –
Baker et al. 2000 1–4 Fluvial sediment –
Barcelona 1984 2.7 to <4 Glacial sediment –
Batiot et al. 2003a 1.2 Karst bedrock –
Datry et al. 2004 0.5 Glacial sediment –
Wassenaara and Aravena 1991 1.4–2 Sandstone bedrock –
Wassenaara et al. 1991 1.5 to 3.2 Glacial sediment –
Polluted/enriched
Anawar et al. 2003 >1 to <6 Fluvial sediment Sedimentary
Aravenaa et al. 1995 1–18 Glacial sediment Sedimentary
Artinger et al. 2000 <1 to >200 Glacial sediment Sedimentary
Artinger et al. 2000 5–17 Fluvial sediment Agriculture
Barcelona 1984 <5 to >20 Glacial sediment Drilling fluids
Batiot et al. 2003b 1–5 Dolomite bedrock Irrigation water
Bhattacharya et al. 2002 <1 to >14 Fluvial sediment Sedimentary
Chomycia et al. 2008 4–55 Fluvial sediment Dairy farming
Clay et al. 1996 <1 to >10 Glacial sediment Agriculture
Davis et al. 1994 2 to >400 Fluvial sediment Tannery leachate
Goñi and Gardner 2003 6 to >120 Marine sediments Irrigation water
Pabich et al. 2001 <1–23 Glacial sediment Eutrophication
Thayalakumaran et al. 2015 <4 to >80 Fluvial and marine sediment Sugar-cane fields
Note: Sedimentary refers to naturally emplaced geological formations
Hydrogeol J (2017) 25:959–967 961
(Oliver et al. 1983). Also, the persistence of high DOC loading
has significant implications for the fate of other contaminants
such as pesticides, pathogens, and pharmaceuticals, as hypox-
ic conditions may restrict the degradation of many carbon-
based compounds (Chomycia et al. 2008), though in some
instances it may promote it.
There are numerous water treatment technologies applied
for DOC reduction such as flocculation and filtration engi-
neering techniques. Advanced treatment typically employs
ozonation, which cleaves the unsaturated bonds in the aromat-
ic molecular components of humic materials, thereby decreas-
ing colour, in addition to increasing the degradability of or-
ganic molecules (Rittmann et al. 2002). While ozonation has
been shown to reduce DOC, the magnitude of reduction is
limited (< 10%); accordingly, additional treatment processes,
typically engineered biofiltration approaches, are required to
provide further DOM reductions (Drewes and Jekel 1998).
However, when wastewater is to be released to groundwater,
further treatment can be achieved by soil aquifer treatment
(SAT), which utilises the attenuation properties of the subsoil
zone and concurrently recycles wastewater while augmenting
recharge (Drewes and Jekel 1998). SAT systems are typically
employed in areas characterised by limited conventional
freshwaterresources and overexploited local aquifers, e.g. arid
regions (Kim and Yu 2007), developing countries (Westerhoff
and Pinney 2000), and heavily urbanised environments
(Drewes and Jekel 1998).
Drinking-water treatment facilities, including the pre-
treatment of SAT recharge, typically involves disinfection via
chlorination in order to ensure satisfactory potable water sup-
plies and to prevent microbial growth during distribution
(Chomycia et al. 2008;Uyaketal.2008); however, this ap-
proach may be problematic as chlorine may react with DOC to
produce potentially harmful disinfection by-products (DBPs)
including trihalomethanes (THMs) and haloacetic acids
(HAAs), both of which are suspected carcinogens (WHO
2011). The concentrations of both are currently regulated
across the European Union by Council Directive 98/83/EC
and in the United States by the US Environmental Protection
Agency (Safe Water Drinking Act). The total concentration of
DBPs present and the distribution of individual compounds in
chlorinated water depend upon the initial raw water character-
istics, in addition to operational parameters during the treatment
process such as pH and temperature (Kim and Yu 2007;Uyak
et al. 2008). The complex polymeric properties of humic sub-
stances result in significant difficulties with respect to effective
environmental characterisation. Previous studies have shown
that humic substances are imperative precursors to the devel-
opment of DBPs and strongly influence the formation potential
of THMs, much more so than the non-humic faction of organic
matter (Westerhoff and Pinney 2000;KimandYu2007).
Investigations of DOC treatment processes consistently
confirm that biodegradation of DOC represents the dominant
removal mechanism during treatment, with preferential re-
moval of low molecular weight DOC (Drewes and Jekel
1998; Westerhoff and Pinney 2000;KimandYu2007).
DOC removal typically exhibits moderate (50%) to high
(90%) removal efficiencies (Quanrud et al. 2003a).
However, the bias towards refractory humic substance com-
ponents remaining in DOC solution complicates treatment
efficacy as DBP precursors with THM formation potential
(THMFP) are persistent factions of DOC and whose reactivity
may increase post-treatment due to removal on non-reactive
species (Westerhoff and Pinney 2000; Quanrud et al. 2003b).
Nevertheless, whilst associations have been found between
increased instance of specific cancers (e.g. rectal, kidney)
and consumption of chlorinated drinking water (Kuo et al.
2010;Liaoetal.2012), few studies have been carried out on
demonstrated cancer occurrences as a result of THM/DBP
presence in potable groundwater supplies. Subsequently, de-
spite engineered artificial recharge systems being commonly
used to remove DOC from recycled treated wastewaters, little
research has been conducted on the presence of DBPs in
groundwater systems receiving artificial recharge.
Organic matter and colloidal mobility
BColloidal contaminants^include some of the earliest known
pollutants recognised to impact water quality such as patho-
genic microorganisms (bacteria, viruses and protozoa; Macler
and Merkle 2000). More recent investigations have
recognised that natural colloidal materials can facilitate the
transport of low solubility substances (Kretzschmar et al.
1999), while, by products of recent technological develop-
ments, including engineered nanoparticles (Nuttall and Kale
1994), represent emerging contaminants of concern that may
also impact human health (Troester et al. 2016). As frequently
alluded to throughout the current Hydrogeology Journal spe-
cial issue Hydrogeology and Human Health, the importance
of aquifers as drinking water sources in many parts of the
world cannot be overstated, and thus, an increase in under-
standing of colloid mobility in the subsurface will likely prove
fundamental to the development of successful source protec-
tion strategies (Flynn et al. 2015; Hunt and Johnson 2016)as,
in contrast to solutes, pathogens can cause infection at exceed-
ingly low levels of exposure and are often associated with
rapid (<10 year) subsurface flow (Hunt and Johnson 2016).
In spite of considerable recent advances made in under-
standing fundamental processes responsible for attenuation
under controlled (laboratory) conditions, factors influencing
colloidal fate and transport in natural porous media remain
poorly defined. Complicating factors include the higher de-
grees of compositional and textural heterogeneity encountered
in natural deposits (Flynn et al. 2015), in addition to other
factors such as the co-occurrence of suites with other contam-
inants that may promote or inhibit colloid mobility.
962 Hydrogeol J (2017) 25:959–967
At present, the conceptual understanding of colloid mobil-
ity largely relies on approaches developed for engineered filter
beds, where removal capacity has been shown to depend on
the chemical and physical properties of the contaminants, col-
lector surfaces (i.e. grains and fissure walls), and the inorganic
hydrochemistry of the host solution, most notably ionic
strength and pH (Kretzschmar et al. 1999). The role of DOC
adds another layer of complexity to effective characterisation.
Despite its widespread occurrence in natural waters and the
frequently observed relationship between DOC and colloids
in certain polluted settings, e.g. sewage impacted groundwater
(Flynn et al. 2012), colloid-organic matter interactions have
received relatively little research attention until recently
(Bradbury et al. 2013;Yangetal.2015). Qualitative observa-
tions have indicated that diverse mechanisms may operate,
ranging from promotion to inhibition of mobility
(Kretzschmar et al. 1999). The contrasting responses are per-
haps unsurprising, given the range of organic compounds
present in the wider environment in both natural and polluted
settings (Harvey et al. 2010a,2010b).
The capacity of organic matter (OM) to stabilise colloids in
suspension has been recognised for some time (Kretzschmar
et al. 1997), while more recent studies have begun to shed
light on the capacity of OM to influence the attenuation ca-
pacity in saturated porous media—for example, Yang et al.
(2010) carried out multiple pulse dynamic column studies at
laboratory scale, coupled with random sequential adsorption
modelling (RSA) of tracer breakthrough curves to quantify the
influence of OM on colloid mobility. This approach provided
a basis for further investigations to highlight the complexity of
OM-colloidal contaminant interactions in the subsurface en-
vironment. Using RSA modelling, Yang et al. (2011)
employed engineered particles (fluorescent microspheres) to
show how changes in ionic strength and pH, with comparable
compositions to those observed in natural waters, dramatically
altered the capacity of a well characterised humic acid to in-
fluence the colloid attenuation capacity of a saturated sand;
such changes may contribute to increases in colloid contami-
nation in groundwaters following recharge events, as noted by
Hunt and Johnson (2016).
Subsequent studies further highlighted the complexity of
colloid-OM interactions. Flynn et al. (2012) employed a well-
characterised protein (Bovine Serum Albumin) in place of
humic acid, to reveal comparable blocking responses at low
concentrations, whereas application of higher concentrations
generated breakthrough curves reflecting Bripening^behav-
iour. Furthermore, Yang et al. (2012) have identified addition-
al processes of potential hydrogeological significance, includ-
ing the capacity of compounds to dissolve favourable deposi-
tion sites, such as iron oxide patches on grain surfaces, thus
promoting greater colloid mobility.
Despite progress in characterising colloid-OM interactions
in the laboratory, DOC’s influence on colloid migration in
groundwater systems remains largely uncharacterised.
Nonetheless, realistic conceptual models provide a reliable
basis for upscaling and identification of potentially relevant
settings—for example, while the impacts of changes in water
chemistry on colloid mobility are expected to be of greater
significance in surface water systems,similar changes can also
be anticipated in karst, where field experimentation, albeit in
the absence of organic matter, has demonstrated the capacity
of hydrochemical changes to effectively mobilise particles
deposited under conditions more conducive to detachment
(Flynn and Sinreich 2010). Similarly, as studies employing
DOC have demonstrated, contrasts in concentration, which
occur between the margins of a contaminant plume and its
centre may give rise to contrasting colloid attenuation capacity
(Harvey et al. 2011). This information, coupled with the find-
ings of Flynn et al. (2012) suggests that some DOC may be
capable of promoting or inhibiting colloid mobility at different
stages of a plume’sdevelopment. Furthermore, alterations to
the content/configuration of deposited OM, arising from pro-
cesses including OM degradation and/or changes in
hydrochemistry, may give rise to temporal changes in the ca-
pacity of aquifers to disinfect groundwater.
From the aforementioned, it is clear that characterising the
influence of DOC on colloidal contaminant occurrence in
aquifers is a complex topic; moreover, effective characterisa-
tion of colloid-OM interactions at the field scale continues to
present a significant challenge, largely due to inherently lower
levels of experimental control. Nevertheless, it is widely ac-
knowledged that findings from laboratory-based investiga-
tions require complementary in-situ testing in order to accu-
rately assess the relative importance of a potentially vast range
of mechanisms in which OM may influence colloid fate and
transport. Field-based artificial tracer testing, when employed
within a realistic hydrogeological framework, which con-
siders findings of studies completed under more controlled
(laboratory) conditions, represents a useful approach to better
understanding these processes (Flynn et al. 2015). Findings
from limited studies completed to date (e.g. Harvey et al.
2011; Metge et al. 2010;Pieperetal.1997) suggest that this
will prove a fruitful area of research for future investigations,
ultimately leading to more scientifically defensible groundwa-
ter protection strategies.
Inhibition of polymerase chain reaction (PCR)
Those who drink groundwater that has not been disinfected
are at increased risk of infection and disease from pathogenic
microorganisms (Macler and Merkle 2000; Borchardt et al.
2012; Hynds et al. 2014a). Polymerase chain reaction (PCR)
is increasingly being employed as the standard method for
detection and characterization of (pathogenic and non-
pathogenic) microorganisms and genetic markers in a variety
of water types including groundwater (Schrader et al.2012). A
Hydrogeol J (2017) 25:959–967 963
recent systematic review of groundwater contamination in the
United States and Canada found that almost three quarters of
relevant studies from 1990 to 2013 employed PCR as the
principal pathogen detection method (Hynds et al. 2014b).
The method employs a process whereby nucleic acid se-
quences (i.e. microbial Bbuilding blocks^) are enzymatically
amplified from non-detectable to detectable levels (Kong et al.
2002; De Man et al. 2014). During the process, strands of the
DNA double helix, initially present in very small numbers, are
physically separated via DNA thermal melting, followed by
highly selective amplification by DNA polymerase.
Amplification usually takes place over 35–45 cycles, resulting
in exponential multiplication of the number of DNA frag-
ments of interest originally present (Girones et al. 2010;
Ramírez-Castillo et al. 2015). Since its development in the
early 1980s, PCR has become a rapid, highly specific, low
cost method for microbial detection, and is now regularly used
in the field of environmental science (Abbaszadegan et al.
1999; Schrader et al. 2012). Moreover, due to increasing ac-
curacy, in parallel with decreasing costs, these techniques are
now frequently employed for quantitative risk assessment and
regulation of both drinking and recreational waters (Gibson
et al. 2012).
Despite the multiple advantages associated with PCR, the
process is prone to inhibition by naturally occurring sub-
stances (inhibitors) in groundwater; PCR inhibitors represent
a diverse group of organic and inorganic chemicals including
urea, phenol, ethanol, complex polysaccharides, bacterial de-
bris, metal ions, humic acids, and tannic acid, among others
(Radstrom et al. 2004;Schraderetal.2012). The presence of
varying concentrations of inhibitory compounds in water sam-
ples will disrupt amplification of target nucleic acids via a
number of biochemical processes (Shain and Clemens 2008;
Radstrom et al. 2008). This results in partial or total PCR
inhibition, leading to decreased sensitivity or false-negative
results, respectively. Accordingly, this potential under-
estimation represents the primary source of concern with re-
spect to identifying poor water quality using PCR
methodology.
As previously outlined, humic acids and other humic sub-
stances are typically present in groundwater with high levels
of dissolved organic carbon, (Artinger et al. 2000;Rocketal.
2010). This in turn will interact with template DNA and the
polymerase, even at low concentrations, thus preventing the
enzymatic reaction (Sutlovic et al. 2008). It is important to
note that levels of inhibition will primarily depend on the
relative concentration of the inhibitor and not its source. A
recent PCR inhibition study undertaken by Gibson et al.
(2012) comprised 2187 groundwater samples taken from do-
mestic and municipal boreholes associated with relatively
deep sandstone aquifers in Wisconsin. Overall, 4% of ground-
water samples were associated with false-negative results for
viral pathogens due to the effects of inhibitory compounds,
with inhibitory concentrations in groundwater samples found
to span four orders of magnitude (Gibson et al. 2012). Results
suggest that numerous factors including sample volume, site
location and seasonality affected inhibition levels; it was not
possible to spatially assess or predict inhibition potential, and
thus it is recommended that all groundwater samples be
analysed for inhibition prior to PCR, and particularly when
sampling is associated with a human health event and/or
(expected) high groundwater DOC concentrations, i.e.
groundwater from karstic areas and peatlands (Borchardt
et al. 2003; Kolka et al. 2008;Gibsonetal.2012).
Summary
The current paper presents an overview of the occurrence of
dissolved organic carbon (DOC) in groundwater, and its im-
plications for human health linked to interactions with other
substances. Treatment processes for DOM effectively reduce
DOC concentrations; however, humic substances are difficult
to remove from solution and react with chlorine in the disin-
fection process to produce DBPs, most notably THMs, many
of which are suspected carcinogens. Whilst connections be-
tween human cancer occurrences and ingestion have been
made for treated surface-water supplies, there are no such
studies from supplies using publicly supplied groundwater.
Considering many domestic and public drinking water wells
are often located in close proximity to arable and agricultural
activities that require treatment, this may be a significant and
persistent, yet poorly understood, health risk, both in develop-
ing and developed countries. Moreover, the use of reclaimed
and treated wastewater for engineered aquifer recharge is an
increasingly common practice in many regions of the world,
which indicates that aquifers utilised by human populations,
both in urban and rural areas, may already have undesirable
DBP concentrations.
Until recently, relatively little attention had been given to
mechanisms behind the interactions between OM and colloi-
dal contaminants, and their influence on the mobility of path-
ogenic microorganisms in the subsurface. The capacity to
quantify these processes now provides a means to more con-
fidently investigate the behaviour of OM in saturated porous
media and for colloidal contaminant transport in aquifers.
Further development of this topic, necessary to protect
groundwater supplies, will require investigations spanning a
range of investigative scales and disciplines. Although there
remains a need for further laboratory-based investigations to
identify fundamental processes, a need for further field-scale
investigations is outstanding if the importance of these pro-
cesses in natural systems is to be defined. The role of the
hydrogeologist is anticipated to prove fundamental in bridging
this divide.
964 Hydrogeol J (2017) 25:959–967
PCR techniques are now frequently used for the detec-
tion, identification, and quantification of waterborne path-
ogens—for example, Sezen et al. (2014) recently reported
on a large waterborne multi-pathogen gastroenteritis out-
break in Turkey, which was found to be caused by con-
sumption of contaminated groundwater from historic
neighbourhood fountains. This investigation comprised
the use of real-time multiplex PCR, with results proving
the source of the outbreak, in addition to the causative
pathogens (Shigella sonnei, astrovirus, and norovirus).
Accordingly, the study authors were able to provide fast,
reliable evidence-based recommendations pertaining to
both avoidance of new cases and treatment of existing
cases, thereby significantly contributing to public health
in the affected region. However, PCR inhibition by humic
substances is capable of resulting in significant under-
estimation or false-positives (Abbaszadegan et al. 1999;
Radstrom et al. 2008;Schraderetal.2012), thus poten-
tially resulting in a failure to correctly detect the source of
an ongoing human health event associated with ground-
water supplies. While groundwater derived from specific
settings may be expected to comprise high levels of PCR
inhibitors due to high groundwater DOC concentrations,
i.e. karstic areas and peatlands (Borchardt et al. 2003;
Kolka et al. 2008), and should therefore be systematically
analysed for inhibition prior to the use of PCR, a recent
studybyGibsonetal.(2012) found that 4% of ground-
water samples from deep non-karstic aquifers in the US
(n= 2187) produced false-negative results. Moreover, re-
sults from the same study showed that sample volume,
site location and seasonality affected inhibition levels;
thus, it is recommended that, in the short-term, all ground-
water samples be systematically analysed for levels of
inhibition due to the presence of humic acids prior to
PCR, and in the medium- to long-term, regionally or
hydrogeologically specific studies seek to quantify tempo-
ral variation associated with PCR inhibitors in groundwa-
ter, particularly in vulnerable areas, and those associated
with previous waterborne outbreaks.
This paper has identified three key issues associated
with DOC in groundwater, namely the presence of
DBPs, colloidal transport mechanisms and usage of PCR
techniques, in treated and/or untreated groundwater used
in municipal water supply. Future research efforts on
these issues should be directed with respect to groundwa-
ter DOC and human health, as currently, these issues are
under-represented in published literature. Other issues of
concern not covered here include optical clarity and the
presence of chromophoric dissolved organic matter, which
effects ultraviolet/visible light absorbance, and this also
warrants increased investigation in coloured municipal
groundwater supplies, as a means to better understand
the factors and processes leading to DOC enrichment.
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