Investigating the significance of dissolved organic contaminants in aquatic environments: Coupling passive sampling with in vitro bioassays

Article (PDF Available)inChemosphere 90(2) · July 2012with101 Reads
DOI: 10.1016/j.chemosphere.2012.06.041 · Source: PubMed
Abstract
We investigated the feasibility of coupling passive sampling and in vitro bioassay techniques for both chemical and ecotoxicological assessment of complex mixtures of organic contaminants in water. Silicone rubber passive sampling devices (SR-PSDs) were deployed for 8-9weeks in four streams and an estuary of an agricultural catchment in North East (NE) Scotland. Extracts from the SR-PSDs were analysed for freely dissolved hydrophobic organic contaminants (HOCs) and screened for wide range of pesticides. The total concentrations of dissolved PAHs (∑PAH(40), parent and branched) in the water column of the catchment varied from 38 to 69ngL(-1), whilst PCBs (∑PCB(32)) ranged 0.02-0.06ngL(-1). A number and level of pesticides and acid/urea herbicides of varying hydrophobicity (logK(OW)s ∼2.25 to ∼5.31) were also detected in the SR extracts, indicating their occurrence in the catchment. The acute toxicity and EROD induction potentials of SR extracts from the study sites were evaluated with rainbow trout liver (Oncorhynchus mykiss; RTL-W1) cell line. Acute cytotoxicity was not observed in cells following 48h exposure to the SR extracts using neutral red uptake assay as endpoint. But, on a sublethal level, for every site, statistically significant EROD activity was observed to some degree following 72h exposure to extracts, indicating the presence of compounds with dioxin-like effect that are bioavailable to aquatic organisms in the water bodies of the catchment. Importantly, only a small fraction of the EROD induction could be attributed to the PAHs and PCBs that were determined. This preliminary study demonstrates that the coupling of silicone rubber passive sampling techniques with in vitro bioassays is feasible and offers a cost effective early warning signal on water quality deterioration.
Investigating the significance of dissolved organic contaminants in aquatic
environments: Coupling passive sampling with in vitro bioassays
Emmanuel S. Emelogu
a,b,
, Pat Pollard
b
, Craig D. Robinson
a
, Foppe Smedes
c,d
, Lynda Webster
a
,
Ian W. Oliver
e
, Craig McKenzie
b
, T.B. Seiler
f
, Henner Hollert
f
, Colin F. Moffat
a,b
a
Marine Scotland Science (Marine Laboratory), 375 Victoria Road, Aberdeen AB11 9DB, UK
b
Institute for Innovation, Design and Sustainability in Research (IDEAS), Robert Gordon University, Aberdeen AB10 1FR, UK
c
Masaryk University RECETOX, Kamenice 126/3, 625 00 Brno, Czech Republic
d
Deltares, Utrecht, The Netherlands
e
Scottish Environment Protection Agency (SEPA), Avenue North, Heriot-Watt Research Park, Edinburgh EH14 4AP, UK
f
Department of Ecosystem Analysis, Institute for Environmental Research (Biology V), RWTH University, Aachen, Germany
highlights
"
PAHs and PCBs were determined in a rural catchment using passive sampling.
"
Semi-polar pesticides and herbicides were detected using silicone rubber samplers.
"
Coupling passive sampling with in vitro tests enabled evaluation of mixture effects.
"
No cytotoxicity was observed in RTL-W1 cells exposed to passive sampler extracts.
"
EROD activity was measured; only 12–21% of this was attributed to PAHs and PCBs.
article info
Article history:
Received 13 March 2012
Received in revised form 27 June 2012
Accepted 28 June 2012
Available online xxxx
Keywords:
Passive sampling
Silicone rubber
Hydrophobic organic contaminants
Water
Monitoring
Toxic equivalency (TEQ)
abstract
We investigated the feasibility of coupling passive sampling and in vitro bioassay techniques for both
chemical and ecotoxicological assessment of complex mixtures of organic contaminants in water. Sili-
cone rubber passive sampling devices (SR-PSDs) were deployed for 8–9 weeks in four streams and an
estuary of an agricultural catchment in North East (NE) Scotland. Extracts from the SR-PSDs were ana-
lysed for freely dissolved hydrophobic organic contaminants (HOCs) and screened for wide range of pes-
ticides. The total concentrations of dissolved PAHs (
P
PAH
40
, parent and branched) in the water column
of the catchment varied from 38 to 69 ng L
1
, whilst PCBs (
P
PCB
32
) ranged 0.02–0.06 ng L
1
. A number
and level of pesticides and acid/urea herbicides of varying hydrophobicity (logK
OW
s 2.25 to 5.31) were
also detected in the SR extracts, indicating their occurrence in the catchment. The acute toxicity and
EROD induction potentials of SR extracts from the study sites were evaluated with rainbow trout liver
(Oncorhynchus mykiss; RTL-W1) cell line. Acute cytotoxicity was not observed in cells following 48 h
exposure to the SR extracts using neutral red uptake assay as endpoint. But, on a sublethal level, for every
site, statistically significant EROD activity was observed to some degree following 72 h exposure to
extracts, indicating the presence of compounds with dioxin-like effect that are bioavailable to aquatic
organisms in the water bodies of the catchment. Importantly, only a small fraction of the EROD induction
could be attributed to the PAHs and PCBs that were determined. This preliminary study demonstrates
that the coupling of silicone rubber passive sampling techniques with in vitro bioassays is feasible and
offers a cost effective early warning signal on water quality deterioration.
Crown Copyright Ó 2012 Published by Elsevier Ltd. All rights reserved.
1. Introduction
Hydrophobic organic contaminants (HOCs), alongside other
environmental stressors, are major pressures on the ecological
and chemical status of many European freshwater and marine
water bodies. Numerous HOCs are persistent, bioaccumulative,
and have the potential to induce both acute and chronic toxicolog-
ical effects on aquatic organisms and humans (Warren et al., 2003).
0045-6535/$ - see front matter Crown Copyright Ó 2012 Published by Elsevier Ltd. All rights reserved.
http://dx.doi.org/10.1016/j.chemosphere.2012.06.041
Corresponding author at: Marine Scotland Science (Marine Laboratory), 375
Victoria Road, Aberdeen AB11 9DB, UK. Tel.: +44 01224 876 544; fax: +44 01224
295 511.
E-mail address: Emmanuel.Emelogu@scotland.gsi.gov.uk (E.S. Emelogu).
Chemosphere xxx (2012) xxx–xxx
Contents lists available at SciVerse ScienceDirect
Chemosphere
journal homepage: www.elsevier.com/locate/chemosphere
Please cite this article in press as: Emelogu, E.S., et al. Investigating the significance of diss olved organic contaminants in aquatic environments: Coupling
passive sampling with in vitro bioassays. Chemosphere (2012), http://dx.doi.org/10.1016/j.chemosphere.2012.06.041
Monitoring of HOCs such as polycyclic aromatic hydrocarbons
(PAHs), polychlorinated biphenyls (PCBs) and pesticides of varying
chemical composition is a major activity under a number of oblig-
atory monitoring programmes including the EU Water Framework
Directive (WFD; EC, 2000); Marine Strategy Framework Directive
(MSFD Directive 2008/56; EC, 2008) and the Convention for the
Protection of the Marine Environment of the North-East Atlantic
(OSPAR Commission, 2011).
The limitations of traditional water monitoring techniques
have been widely recognised (e.g. Booij et al., 2003; Vrana
et al., 2005) and many statutory monitoring programmes are
now embracing or investigating other methods of assessing HOCs
in the aquatic environment, including biota monitoring and the
use of passive sampler devices (Covaci et al., 2005; Greenwood
et al., 2009; MacGregor et al., 2010). Much work is still required
on the integration and refinement of these alternative methods of
monitoring HOCs. Work is also urgently required to widen the
range of substances that can be assessed, as under current mon-
itoring regimes evaluations of impacts are frequently limited to
substances for which agreed environmental quality standards
(EQSs) are in place (e.g. for WFD priority substances, Decision
No. 2455; EC, 2001).
The combination of accurate chemical data with biological
effects measurement can improve risk assessment for aqueous
organic contaminants; this is particularly true where complex mix-
tures of widely varying compounds occur and where interactions
amongst the components are possible. Further, understanding
the biological availability and interaction of complex mixtures of
HOCs in the environment and in biological systems is crucial in
predicting their toxicological impacts in those systems.
As an alternative to traditional bottle sampling and monitoring
techniques, passive sampling techniques can provide time-
weighted average concentrations (C
TWA
) of freely dissolved aque-
ous contaminants (i.e. bioavailable). Essentially, passive sampling
involves the free flow of contaminants from water, to the receiving
medium, e.g. the passive sampler, which is driven by differences
between the two media in terms of chemical activities. The ex-
change of the contaminants continues until equilibrium is reached
in the system or the sampling is discontinued. Compared to spot
sampling, much lower limits of detection (LOD) are attained with
passive sampling techniques through sampling a large volume of
water over the extended deployment period i.e. days to months.
A variety of passive sampling devices (PSDs) exist, including the
semi-permeable membrane device (SPMD; Huckins, 2006), polar
organic chemical integrative sampler (POCIS; Alvarez et al.,
2004), low density polyethylene (LDPE; Adams et al., 2007) and
silicone rubber (Smedes, 2007). Silicone rubber (SR) passive sam-
pling devices (PSDs), have been shown to be effective in monitor-
ing organic contaminants across a wide range of polarity, i.e.
octanol–water partition coefficient (log K
OW
) over the range 3–8
(Smedes, 2007) and the extraction and clean-up steps are straight-
forward compared to bi-phasic PSDs, e.g. SPMDs. These properties,
in addition to their low cost and relative ease of handling and
analysis, makes them ideal for application in an integrated chemi-
cal–biological effect analysis of environmental samples.
Polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/
Fs) as well as some planar PCBs and PAHs (referred to as diox-
in-like compounds, DLCs) exert their toxic effects on aquatic
organisms by the same mechanism of action; mainly through
the initial binding to the soluble receptor protein known as the
aryl hydrocarbon receptor (AhR). This initiates several biochemi-
cal effects, including the induction of cytochrome P450 1A
(CYP1A) (Stegeman et al., 2001). Evaluation of CYP1A induction
in aquatic organisms has proven to be a sensitive biomarker of
organic contaminants in the aquatic environment and can be rou-
tinely assessed in various ways, such as immunoblotting or mea-
suring the activity of 7-ethoxyresorufin-O-deethylase (EROD) in
various organisms and test systems (Hahn et al., 1996; Hallare
et al., 2011). DLCs usually occur in the environment as complex
mixtures with several other potentially toxic compounds. The di-
oxin-like toxic potencies of complex mixtures can be expressed
in terms of toxic equivalency (TEQ) to the reference compound;
2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). The TEQ can be esti-
mated directly from bioassays but is more routinely calculated
from the individual compounds in the complex mixtures and
their toxic equivalent factors (TEFs) or relative potency (REP)
(Van den Berg et al., 2006; Olsman et al., 2007). When using
the TEQ approach for risk assessment the process is still re-
stricted to the suite of chemicals analysed. Using bioassays with
EROD as the endpoint may determine the joint effects of all DLCs
present in complex environmental samples, although the pres-
ence of some compounds may inhibit the induction of EROD
activity. Importantly, bioassay is often a cheaper and more rapid
estimate of contaminant exposure than chemical analysis, and
can be used to detect the presence of DLCs at concentrations be-
low the LODs of chemical methods (Thain et al., 2008).
Possibly, the quantified contaminant concentrations in PSDs de-
ployed in water can be linked to biological effects determined via
concurrent toxicity assays and/or measurement of CYP1A induc-
tion in vitro. This is supported by a previous work, as Bauer
(2008) applied SR-PSD extracts in umuC (DNA damage) bioassays
and assessed the genotoxicity of the water in which the sampler
had been deployed. Unlike the SPMDs and POCIS (e.g. Muller
et al., 2006; Rastall et al., 2006; Alvarez et al., 2008), to date, there
is still a paucity of studies that have integrated SR-PSD with in vitro
toxicity testing of environmental samples as part of water quality
assessment.
The specific objectives of this preliminary study were: (1) to
investigate the feasibility of integrating SR-PSD and in vitro bioas-
says for chemical and biological effect analysis of aqueous organic
contaminants; (2) to quantify the dissolved concentrations in
water (C
W
) of PAHs and PCBs and to investigate if selected pesti-
cides and acid/urea herbicides could be detected in four streams
and an estuary draining an agricultural catchment; (3) to evaluate
the toxicological effects of complex mixtures extracted from de-
ployed SR-PSDs on a fish cell line using neutral red uptake (NR)
and EROD assays as endpoints; (4) to estimate the TEQ values of
the mixtures.
2. Materials and methods
2.1. Chemicals and materials
HPLC grade solvents (acetone, methanol, dichloromethane,
ethyl acetate, iso-hexane, toluene and acetonitrile) were purchased
from Rathburn Chemicals Ltd., Scotland, United Kingdom (UK).
Certified custom made solutions of PAHs (including deuterated
PAHs) and PCBs were obtained from QMX Laboratories, Essex,
UK. All chemicals and biological reagents used for the neutral red
and EROD induction assays were obtained from Sigma–Aldrich,
Deisenhofen, Germany unless stated otherwise. AlteSil
Ò
translu-
cent food grade SR sheet with a thickness of 0.5 mm and a dimen-
sion of 30 30 or 60 60 cm was purchased from Altec Products,
Ltd., Cornwall, UK. The SR sheets were cut to a dimension of
6 9 cm and pre-extracted in hot ethyl acetate for >100 h using
a Soxhlet apparatus (Laboratory Glass Specialists BV, Ubenna,
Netherlands). This removed low molecular weight silicone SR
oligomers that might affect instrumental analysis and bioassays.
Glass solid-phase extraction (SPE) C8 columns were supplied by
Mallinckrodt Baker, London, UK. Ultra-pure water (18.2 M
X
cm)
was used throughout the experiment.
2 E.S. Emelogu et al. / Chemosphere xxx (2012) xxx–xxx
Please cite this article in press as: Emelogu, E.S., et al. Investigating the significance of dissolved organic contaminants in aquatic environments: Coupling
passive sampling with in vitro bioassays. Chemosphere (2012), http://dx.doi.org/10.1016/j.chemosphere.2012.06.041
2.2. Preparation of silicone rubber passive samplers
Pre-extracted SR sheets were split into two batches, one for
chemical analysis and the other for biological effect assessments
(bioassays). The SR sheets for bioassays were thoroughly rinsed
with ultra-pure water to remove any trace of chemical solvent.
The SR sheets for chemical analysis were spiked with a mixture
of performance reference compounds (PRCs) including deuterated
PAHs (D12-chrysene, D12-benzo[e]pyrene, D10-fluorene and D10-
fluoranthene) and chlorinated biphenyl congeners (CBs 10, 14, 21,
30, 50, 55, 78, 104, 155, and 204) by equilibrating in a methanol/
water spiking solution (Booij et al., 2002). PRCs are a group of
non-environmentally occurring compounds and their release dur-
ing deployment enables determination of in situ sampling rates
(R
S
). The SR sheets for chemical and bioassay analyses were kept
in amber coloured jars with lids lined with aluminium foil and
stored at 20 °C until required.
2.3. Sampling locations and passive sampling
2.3.1. Sampling locations and site descriptions
The study was conducted in the Ythan catchment in north east
(NE) Scotland, which has an area of 675 km
2
. Approximately 85%
of the catchment is utilised for agriculture (mixed farming). The
area is sparsely populated and there are no known major industrial
facilities in the vicinity. Four stream locations and an estuary site
were targeted for study (Fig. 1). Site 1 was at the headwaters of
the River Ythan (the main river within the catchment); sites 2
and 3 were on a small tributary and were approximately 3.3 km
apart; site 4 was on the River Ythan just above the tidal limit;
and site 5 was located in the estuary of the river. The catchment
was selected for this study on the basis of preliminary ecology
and chemical assessments at sites 2 and 3, which indicated a de-
gree of pesticide influence. Site 4 was close to the largest town in
the catchment and all of the other study sites were in the vicinity
of moderate, rural vehicular road usage. At sites 2 and 3, the min-
imum and maximum temperature and pH values during the
sampling period where 1.31 and 7.10 °C and 6.70 and 7.01; at
the estuary, the mean salinity was 15 PSU, mean temperature
was 4.1 °C, and mean pH was 7.5.
2.3.2. Passive sampling
The procedures for the preparation, deployment, retrieval and
extraction of SR-PSDs in this study followed Smedes (2007).ASR
sampler consisted of six sheets weighing 20 g in total and had a
surface area of 600 cm
2
. One PRC-spiked and one un-spiked sam-
pler were deployed simultaneously at the five sites from November
2010 to January 2011. Samplers were deployed for 65 d at sites 1,
2, 3 and 4, and for 58 d at site 5. Upon retrieval, sampler surfaces
were gently and rapidly wiped using solvent-free household clean-
ing pads and water from the study sites in order to remove any bio-
fouling. Sets of PRC-spiked SR samplers served as field and
production control blanks and for time zero determination of PRC
loss. The field blanks were similarly taken to the study sites during
deployment and retrieval; but were only exposed to air (i.e. not
submerged in water) and the production blanks were kept in the
laboratory. A separate set of SR samplers for bioassay controls were
kept in ultra-pure water in the laboratory during the entire sam-
pling period. This served as control blank for the bioassays. Once
retrieved, SR samplers for chemical analysis and bioassays were
kept separately in amber coloured jars with lids lined with alumin-
ium foil and stored at 20 °C until required.
2.4. Extraction of silicone rubber passive sampling devices (SR-PSDs)
2.4.1. Extraction of SR-PSDs for chemical analysis
Extractions of SR samplers were performed with Soxhlet appa-
ratus for 24 ± 4 h in hot mixture of acetonitrile (ACN):methanol
(MeOH; 2:1 v/v). The design of this apparatus ensures that the
sheets are continuously submerged in sub-boiling solvent through-
out the extraction period. Prior to extraction, known amounts of
deuterated internal standards (D8-naphthalene, D10-biphenyl,
D8-dibenzothiophene, D10-anthracene, D10-pyrene, D12-ben-
zo[a]pyrene and D14-dibenzo [a,h]anthracene), a PCB recovery
Fig. 1. River Ythan catchment (showing River Ythan and main tributaries) in NE Scotland. Sample sites 1–5 are indicated. Inset shows map of Scotland and location of the
catchment.
E.S. Emelogu et al. / Chemosphere xxx (2012) xxx–xxx
3
Please cite this article in press as: Emelogu, E.S., et al. Investigating the significance of diss olved organic contaminants in aquatic environments: Coupling
passive sampling with in vitro bioassays. Chemosphere (2012), http://dx.doi.org/10.1016/j.chemosphere.2012.06.041
standard (CB112) and pesticide internal standards (azobenzene
and diphenamid) were added to each Soxhlet apparatus containing
each set of the SR samplers. The choice of azobenzene and diphen-
amid as internal standards for pesticides/herbicides was because
they are not prevalent in the environment and they are stable
and similar to other components in the suite of analysis. Inclusion
of low molecular weight PAH internal standard (i.e. D8-naphtha-
lene) to the Soxhlet extraction system was to correct for possible
loss of low molecular weight HOCs e.g. naphthalene. After extrac-
tion, samples were reduced to 2 mL via Kuderna–Danish
evaporation apparatus (Laboratory Glass Specialists BV, Ubenna,
Netherlands). Subsequently, extracts were added to glass solid-
phase extraction (SPE) C8 columns and eluted with ACN to remove
any co-extracted SR oligomers. The samples in ACN were further
concentrated to 2 mL and were solvent exchanged into iso-hex-
ane. The extracts were aliquoted into three equal fractions for
determination of (1) PAHs, (2) PCBs, and (3) selected pesticides
and acid/urea herbicides.
2.4.2. Extractions of SR-PSDs samplers for bioassays
Two sheets from each set of six deployed at each site for bioas-
say use, together with the laboratory process blanks, were ex-
tracted using the hot Soxhlet apparatus as explained in
Section 2.4.1 but without any added internal or recovery standards.
The other four sheets from each set were preserved for future tox-
icological assessment. Following extraction, extracts were purified
using the glass solid-phase extraction (SPE) C8 column, solvent
exchanged into MeOH and concentrated to 1 mL with Kuderna–
Danish evaporation apparatus and activated carbon purified nitro-
gen blow down. Bioassay results are expressed relative to the
equivalent mass of SR per mL extract (mg SREQ mL
1
). The extracts
were stored at 20 °C until needed for bioassays. Extracts from
blank samplers were utilised to confirm that the extraction proce-
dure and extraction solvents were not inherently toxic to cell
systems during bioassays.
2.5. Chemical analysis of silicone rubber extracts
Extracts of SR samplers were analysed for 40 PAHs (parent and
branched), 32 ortho and mono-ortho PCBs and several selected pes-
ticides and acid/urea herbicides using a combination of GC and li-
quid chromatography (LC). The detectors applied include mass
spectrometry (MS), electron capture detector (ECD), and MS–MS.
Detailed procedures for the analysis, including the complete list
of pesticides and acid/urea herbicides selected for this study are
provided in Supporting information.
2.5.1. Calculation of sampling rate and freely dissolved concentrations
of HOCs in water
To calculate the dissolved concentrations (C
W
;ngL
1
) of the
analytes from the amounts absorbed by the samplers during
deployment, the sampler water partition coefficients (K
SW
) and
sampling rates (R
s
,Ld
1
) are required. The K
SW
values for most
of the compounds used in the current study were obtained from
Smedes et al. (2009). The uptake of organic compounds and release
of PRCs by passive samplers is principally controlled by the resis-
tance to transport in the water boundary layer (WBL) and the sam-
pler material (Huckins, 2006). However, compared to the WBL, the
resistance to transport of the sampler material was found negligi-
ble on SR samplers for compounds with log K
OW
>3(Rusina et al.,
2007). As the molecular weight of the sampled compounds
increases, the diffusion and transport through the WBL decreases
and consequently the R
S
decreases. Hence, in this situation, where
uptake is entirely dependent on the thickness of the WBL, Rusina
et al. (2010) proposed a model which expresses R
S
as a function
of molecular mass (M) of the compounds:
R
s
¼ FAM
0:47
ð1Þ
where Aisthe surface area (m
2
) of the sampler and F is the flow pro-
portionality constant, which includes the flow dependence sam-
pling rate and factors to fit the units. This flow proportionality (F)
constant is derived from the release of PRCs loaded to the samplers
prior to deployment. The fraction (f) of PRCs retained in the SR sam-
plers after deployment is related to the sampling rate (R
S
) through:
f ¼
N
t
N
0
¼ exp
R
st
mK
SW

¼ exp
FAM
0:47
t
mK
SW
"#
ð2Þ
where N
0
is the initial amount (ng) of PRC in the sampler, N
t
the fi-
nal amount (ng) remaining in the deployed SR sampler, t is the
exposure time (d), and m is the mass (kg) of the SR sampler. By
applying non-linear least-squares (NLSs) regression with f as func-
tion of FAM
0.47
, the modelled f values can be fitted with the exper-
imental values using the proportionality constant (F) as the
adjustable variable as detailed in Booij and Smedes (2010).
Using the spreadsheet supplied by the authors, an estimate of the
R
S
and the standard error of the target compounds were obtained by
applying the deviations of the experimental value from the model.
The PRC derived R
S
values for an average compound of mass 300
in L d
1
were 47 ± 5.7 for site 1, 14 ± 2.0 for site 2, 29 ± 4.2 for site
3; 32 ± 14 for site 4, and 36 ± 5.2 for site 5. Graphs of the obtained
fits can be found in Supporting information (Fig. SI 1).
Conventionally, the C
W
can be calculated by applying the uptake
model that is valid for equilibrium and linear uptake situations
(Huckins, 2006):
C
W
¼
N
t
mK
SW
1
1 exp
R
s
t
mK
SW

ð3Þ
where N
t
is amount (ng) of target compound absorbed by the SR
sampler during exposure; by combining Eqs. (2) and (3):
C
W
¼
N
t
mK
SW
1 exp
FAM
0:47
t
mK
SW
hi
ð4Þ
In this study, the uptake of PAHs were sufficiently high, so
production and field blanks were insignificant (<10%), hence, no
corrections were needed. However, the absorbed amounts of PCBs
were very low and consequently closer to amounts in the produc-
tion and field blanks. Dealing with blanks is not a straight for-
ward subtraction in passive sampling, for example a production
blank will not influence the amount on the sampler for com-
pounds that attained equilibrium, but needs to be fully subtracted
for compounds that stay in the linear uptake phase during the
whole deployment (Booij et al., 2007). Further, production blanks
dissipate like PRCs during sampling, hence, from Eq. (4) a correc-
tion for production blank can be included (Smedes and Booij,
2012):
C
W
¼
N
t
N
0
exp
FAt
mK
SW
M
0:47

mK
SW
1 exp
FAt
mK
SW
M
0:47

ð5Þ
where N
0
is the initial amount (ng) of target compound in the pro-
duction blank. Calculated concentrations (ng g
1
SR) that were less
than twice the production blank were considered below the detec-
tion limit and marked accordingly. The whole data were included
when summing the compounds.
2.6. Biological effects analysis of silicone rubber extracts
2.6.1. Cell culture
The fibroblast-like permanent fish cell line, RTL-W1 cells was
used in bioassays. The choice of RTL-W1 cell line was due to its
4 E.S. Emelogu et al. / Chemosphere xxx (2012) xxx–xxx
Please cite this article in press as: Emelogu, E.S., et al. Investigating the significance of dissolved organic contaminants in aquatic environments: Coupling
passive sampling with in vitro bioassays. Chemosphere (2012), http://dx.doi.org/10.1016/j.chemosphere.2012.06.041
high potential to express CYP1A-based EROD activity on exposure
to dioxin-like compounds (Lee et al., 1993; Hallare et al., 2011).
RTL-W1 cells (Drs. N.C. Bols and L. Lee; University of Waterloo,
Canada) were cultured in 75 cm
2
plastic culture flasks (TPP,
Trasadingen, Switzerland) without additional gassing at 20 °Cin
Leibowitz’s L15 medium supplemented with 9% foetal bovine ser-
um (FBS) and 1% penicillin/streptomycin solution (10,000 U/
10,000
l
gmL
1
) in 0.9% sodium chloride (NaCl).
2.6.2. Neutral red uptake assay
Acute cytotoxicity of SR (deployed and controls) extracts and
the vehicle control on RTL-W1 cells were assessed with neutral
red uptake assay as detailed in Borenfreund et al. (1988) and Seiler
et al. (2006). In this study, the procedure was adapted to a 24-well
microtitre plate (TPP, Trasadingen, Switzerland). Each SR extract
(1.04–66.7 mg SREQ per mL) and vehicle control (methanol; high-
est concentration = 1% v/v) were tested on individual plates. Each
sample and dilution was tested in duplicate. The positive control
used for each test was 3,5-dichlorophenol (DCP; highest concentra-
tion = 40 mg L
1
of medium) in each test plate. DCP was also tested
separately on an individual plate at concentrations in the range
0.63–40 mg L
1
of medium. Duplicate negative control wells (with
cells but without test extracts/vehicle control) were located both
near the samples with the highest concentrations and those with
the lowest concentrations. Four further wells were left blank with
neither solvent nor cells. Neutral red uptake (cell viability) was
determined photometrically at 540 nm with a reference wave-
length of 690 nm using Infinite M200 multiwell plate reader
(Tecan, Crailsheim, Germany).
2.6.3. EROD induction assay
The CYP1A induction potentials of SR extracts from each of the
study sites were assessed using the EROD assay. The details of the
procedure have been previously described in Gustavsson et al.
(2004) and Wölz et al. (2011); in this study, the procedure was
modified and optimised to a 24 well microtitre plate. Two plates
were used for each test, one plate for the samples and positive con-
trol (2,3,7,8-tetrachlorodibenzo-p-dioxin, TCDD; Promochem, We-
sel, Germany) and the other plate for the measurement of resorufin
and protein calibration standard curves. Each test plate had five
concentration levels of SR extracts (4.17–66.7 mg SREQ per mL)
or vehicle control (methanol; highest concentration 1% v/v) and
TCDD (3.13–50 pM). Wells with cells but without solvent or ex-
tracts were used as negative controls (NC) and another two blank
wells contained neither cells nor test compound. All tests and dilu-
tions were conducted in duplicate. EROD activity was measured
fluorimetrically at an excitation and emission wavelength of 544
and 590 nm using an Infinite M200 plate reader (Tecan, Crailsheim,
Germany). The protein concentrations were determined fluorimet-
rically in parallel using the fluorescamine method at excitation and
emission wavelengths of 360 and 465 nm (Lorenzen and Kennedy,
1993), according to the protocol detailed in Hollert et al. (2002).No
cytotoxicity was observed in response to methanol (vehicle con-
trol; maximum 1% v/v).
2.7. Data analysis of biological effect assessment and calculation of
Chem-TEQ and Bio-TEQ values
2.7.1. Data analysis of neutral red and EROD assays
In both the neutral red uptake and EROD assays, the average
readings for blank wells were subtracted from the values obtained
for the test wells. With the neutral red uptake assay, each test was
considered valid if the two sets of negative controls did not differ
by more than 20% from each other. Statistical analyses were per-
formed using a one-way analyses of variance (ANOVA) followed
by Dunnett’s and Tukey’s multiple comparison tests. Extracts were
considered cytotoxic if the ANOVA and the multiple comparison
tests with mean values for two or more consecutive concentrations
were significantly (p < 0.05) higher than the two negative controls
and the lowest concentration imposed. Where possible, Boltzmann
sigmoidal concentration–response curves (with variable slopes)
were fitted to the mean (±SD) viability of four replicates at each
exposure concentration using Graphpad Prism 5.0 (GraphPad, San
Diego, USA). Viability of the exposed cells was expressed relative
to the NC and the cytotoxic potential of the test samples were cal-
culated as EC
50
values.
For the SR extracts and vehicle control to be evaluated as capa-
ble of inducing EROD activity, procedures described in Bols et al.
(1999) were followed. Concentration–response curves for the
EROD induction response to each sample were computed by non-
linear regression using the Boltzmann sigmoid curve as a model
equation and the concentration of each sample causing 25% of
the TCDD-induced maximum EROD activity (defined as extract
EC
25
TCDD) was calculated (Seiler et al., 2006).
2.7.2. Calculation of Toxicity Equivalent concentrations (TEQ)
Chemically derived toxic equivalent (Chem-TEQ) concentra-
tions for each SR extract were calculated by the sum of the prod-
ucts of the measured PAH and PCB concentrations with their
corresponding toxic equivalent factors (TEF) values as shown in
Eq. (6) while assuming additive effect (Eadon et al., 1986). TEF val-
ues for some PAHs including benzo[a]anthracene, chrysene,
benzo[b]fluoranthene, benzo[a]pyrene, dibenzo[a,h]anthracene,
and indeno[1,2,3-cd]pyrene and some mono ortho PCBs including
CBs 105, 118 and 156 have been derived from EROD assay using
RTL-W1 cells (Clemons et al., 1998; Bols et al., 1999). Chem-TEQs
were calculated as concentrations in picogram TCDD per gram
SREQ (pg TCDD g
-1
SREQ).
Chem TEQ ¼
X
½PAH
i
TEF
i
þ
X
½PCB
i
TEF
i
n ð6Þ
where TEF
i
is the TEF for the individual PAH or PCB congener and n
is the number of compounds in each extract.
Subsequently, the EC
25
TCDD values calculated for each SR ex-
tract were used for calculations of the bioassay TCDD-equivalents
(expressed as Bio-TEQ; pg TCDD g
1
SR) in each SR extract as
shown in Eq. (7):
Bio-TEQ ¼
TCDDEC
25
extract EC
25
TCDD
ð7Þ
where TCDD EC
25
(pg mL
1
) is the concentration of the TCDD posi-
tive control in each SR extract causing 25% of EROD induction and
extract EC
25
TCDD (g mL
1
) is the concentration of the SR extract
equivalent causing 25% of EROD induction (Engwall et al., 1999).
EC
25
TCDD was considered more appropriate than EC
50
in this
study because in several cases the EC
50
was not well defined by
the dose response curve. In addition, interactions are less likely
to occur at lower extract concentrations, therefore the lower con-
centration portion of the curves were considered more appropriate
to calculate the EROD inducing potencies of the extracts (Hollert
et al., 2002).
3. Results and discussion
3.1. Chemical analysis
3.1.1. PAHs
The concentrations of the environmental analyte PAHs were
above the limit of detection (LOD) in most of the samples, except
for the field blank in which most compounds were below the
LOD. The freely dissolved concentrations of the individual PAH
compounds are provided in Supporting information (Table SI 1).
E.S. Emelogu et al. / Chemosphere xxx (2012) xxx–xxx
5
Please cite this article in press as: Emelogu, E.S., et al. Investigating the significance of diss olved organic contaminants in aquatic environments: Coupling
passive sampling with in vitro bioassays. Chemosphere (2012), http://dx.doi.org/10.1016/j.chemosphere.2012.06.041
The total concentrations (sum of 40 parent and branched) of freely
dissolved PAHs in water at each of the study locations varied from
38 to 69 ng L
1
, being highest at sites 1 and 5 and lowest at site 2
(Table 1). Similar composition profiles where observed for all study
locations, with the low to medium molecular weight PAHs (2- to 3-
rings), particularly naphthalene and phenanthrene, dominating the
overall PAH compositions and accounting for more than 70% of the
total PAHs measured (Fig. 2). Naphthalene accounted for more
than 46% and 22% of the total PAH concentrations at sites 1 and
2 respectively. Using SPMDs, concentrations in the range of 79–
540 ng L
1
were reported for individual PAHs including fluoranth-
ene, fluorene and pyrene in surface waters in an area with inten-
sive agricultural activities in the United States (Alvarez et al.,
2008). Schafer et al. (2010) used silicone passive samplers to
estimate a total concentration of 0.1–10 ng L
1
for 16 PAHs in 9
streams in Victoria, Australia after a wildfire.
A high proportion of the heavier parent PAHs (5-and 6-rings)
implies predominately pyrolytic source, whilst a high proportion
of alkylated 2- and 3-ring PAHs suggests a petrogenic origin, hence
concentration ratios can identify possible sources (Witt, 2002; Neff
et al., 2005). At all of the study sites 5- and 6-ring PAHs comprised
<5% of the total (Fig. 2). The ratio of methylphenanthrene/phenan-
threne sequestered by SR-PSDs was 1.2 at site 1, 0.9 at site 2, 0.8 at
site 3, 1.7 at site 4 and 1.1 at site 5, suggesting a higher petrogenic
input at site 4 than at the other sites. This may reflect the proxim-
ity of this site to the largest town in the catchment. The freely dis-
solved concentrations of individual PAHs in the streams and
estuary of Ythan catchment were 3–4 orders of magnitude lower
than their annual average environmental quality standards
(AA-EQS) for surface waters under the WFD (EC, 2006).
3.1.2. PCBs
Freely-dissolved concentrations of thirty-two ortho and mono
ortho PCB congeners, including the Indicator-PCBs (CB28, 52, 101,
118, 153, 138, and 180) are listed in Supporting Information
(Table SI 2). The PCB extract for site 4 was lost during the clean-
up process. Sum-concentrations of the seven Indicator-PCBs
(
P
PCB
7
) followed the same pattern as the sum of all measured
PCBs (
P
PCB
32
), with about 3 times higher concentrations at site
2 than at the other sites (Table 1). Unlike for PAHs, the highest total
concentrations of the PCBs were found at sites 2 and 3. Total
concentrations of PCBs (
P
PCB
32
) measured in this study ranged
0.02–0.06 ng L
1
(Table 1) and were consistent with a previous
survey in Scottish waters which concluded that values of PCBs in
most surface waters were less than 1 ng L
1
for the sum of all
the congeners (SOAEFD, 1996). In other studies, a sum concentra-
tion of 0.12–1.47 ng L
1
of 20 PCBs were measured in spot water
samples from eight major riverine runoff outlets of the Pearl River
Delta (PRD), South China (Guan et al., 2009). Similarly, a study
using SPMDs to monitor 12 dioxin-like PCBs in Port Jackson (Syd-
ney Harbour), Australia estimated sum concentrations ranged from
0.021 to 0.54 ng L
1
(Roach et al., 2009).
Most of the PCB congeners were present at each of the four suc-
cessfully analysed locations and a predominance of the moderately
to relatively highly chlorinated PCBs, i.e. tetra, penta, and hexa
PCBs, was apparent (Fig. 3). The overall data obtained for PCBs in
the present study did not indicate any specific point source inputs
in the catchment.
3.1.3. Chem-TEQ from chemical analysis
The Chem-TEQ (based on concentrations in silicone rubber) val-
ues measured in this study varied considerably among the study
locations and ranged from 6 to 11 pg TCDD g
1
SREQ (Table 1).
The lowest and highest concentrations were measured for sites 1
and 2 respectively, and reflected the dominant contribution of
PAHs to the Chem-TEQ values (i.e. PAHs contributed >80% of
Chem-TEQ). Chrysene, benzofluoranthenes and benz[a]anthracene
were the predominant PAH compounds, while CBs 105 and 118
Table 1
Sum of dissolved concentrations of PAHs and of PCBs and the concentrations of
pesticides and acid/urea herbicides absorbed by silicone rubber passive sampling
devices (SR-PSDs) deployed in water at the sampling locations.
Site 12345
PAHs and PCBs (ng L
1
)
P
PAHs
40
69 38 41 59 69
P
PCBs
32
0.03 0.06 0.04 N.A 0.02
P
Indicator-PCBs (
P
PCB
7
) 0.01 0.03 0.01 N.A. 0.01
Chem-TEQ [pgTCDD g
1
SR sampler] 6 11 7 6 9
Pesticides (ng g
1
SR sampler)
Chlorpyrifos ethyl 8 34 287 9 11
Cyprodinil 8 3 19 9 11
Diazinon 20 3 8 31 43
Epoxiconazole 26 205 201 37 50
Fenpropimorph 1200 93 150 280 44
Flusilazole 27 460 301 54 71
Metazachlor 36 330 230 35 50
Pendimethalin 31 76 740 250 180
Hexaconazole N.D. 7 N.D. 12 15
Propiconazole 18 34 44 32 23
Diflufenican 27 106 440 190 160
Tebuconazole 11 120 120 47 23
Others
a
15 27 31 27 27
Acid/urea herbicides (ng g
1
SR sampler)
2,4-D
c
4.0 8.7 3.8 1.40 4.0
Mecoprop (MCPP) 6 30 15 10 12
Diuron N.D N.D 290 N.D 290
Isoproturon 50 80 275 340 220
Linuron 7600 3200 1500 900 1500
Chlorotoluron 180 90 4000 3000 2100
Others
b
3451411
N.A.-Not available.
N.D.-Not detected.
SR-PSDs were deployed for 65 d at sites 1, 2, 3 and 4 and for 58 d at site 5
(November 2010–January 2011).
a
Others = atrazine, chlorfenvinphos, disulfoton, iprodione, malathion, penta-
chlorobenzen, pirimicarb, pirimiphos methyl, terbuthylazine, terbutryn, triadime-
nol, trifloxystrobin, fenpropidin and trifluralin.
b
Others = 4-(2,4-dichlorophenoxy) butyric (2,4-DB), dichlorprop, bromoxynil,
loxynil, 4-(4-chloro-2-methylphenoxy) butyric acid (MCPB), triclopyr, monolinu-
ron, and fenuron.
c
2,4-dichlorophenoxy acetic acid (2,4-D).
Fig. 2. Percentage composition of PAH groups in water at the various sampling
locations. Term description: 2-ring =
P
naphthalenes (parent and C1-C4); 3-
ring =
P
acenaphthene; acenaphthylene; fluorene; phenanthrene and anthracene
(parent and C1–C3); DBTs =
P
dibenzothiophenes (parent and C1–C3);
4-ring =
P
fluoranthene and pyrene (parent and C1–C3); benzo[c]phenanthrene;
benz[a]anthracene; benz[b]anthracene and chrysene (parent and C1–C2); 5-
ring =
P
benzofluoranthene, dibenz[ah]anthracene, benzo[a]pyrene, benzo[e]pyr-
ene and perylene (parent and C1–C2). 6-ring =
P
indenopyrene, benzoperylene
(parent and C1 C2).
6 E.S. Emelogu et al. / Chemosphere xxx (2012) xxx–xxx
Please cite this article in press as: Emelogu, E.S., et al. Investigating the significance of dissolved organic contaminants in aquatic environments: Coupling
passive sampling with in vitro bioassays. Chemosphere (2012), http://dx.doi.org/10.1016/j.chemosphere.2012.06.041
were the highest PCB contributors. The measured Chem-TEQ val-
ues in each of the study locations would enable comparison to
Bio-TEQ values, so as to evaluate the contributions of the analysed
freely dissolved PAHs and PCBs to the overall biological effects
measured with the EROD assays. It has been suggested that the
presence of PAHs in environmental complex mixtures can domi-
nate the contributions of other DLCs such as PCBs, PCDF and PCDD
in the estimation of TEQ (Eljarrat and Barcelo, 2003).
3.1.4. Pesticides and acid/urea herbicides
Unfortunately, SR-PSDs K
SW
and diffusion coefficients (D
P
) are
not currently available for these compounds and consequently
aqueous concentrations of pesticides and acid/urea herbicides
could not be calculated. However, SR-PSDs samplings could be
used to assess their occurrence and absorbed amounts allowed a
relative comparison.
There was a downward trend in the absorbed amounts of pesti-
cides towards the estuary, indicating the possible influence of agri-
cultural activities in the upper parts of the catchment and
increasing water dilutions in the lower parts (Table 1). The distri-
bution pattern of pesticides in the catchment was variable, for
example, fenpropimorph, a cereal fungicide, was dominant at sites
1 and 4, accounting for more than 84% and 27% of the sum detected
components respectively, while pendimethalin, a selective herbi-
cide used for the control of broadleaf and grassy weeds, was dom-
inant at sites 3 and 5, accounting for over 28% and 24% of the sum
components in the two sites, respectively; flusilazole, a systematic
fungicide for broad spectrum disease control, was dominant (over
31%) at site 2. The occurrence and dimension of the detected pes-
ticides in the catchment may be influenced by a number of factors
including agricultural runoff potential and the physico-chemical
properties of the individual components. Using POCIS, Alvarez
et al. (2008) reported up to 3400 ng POCIS
1
of individual pesti-
cides (atrazine) in surface waters of an area with intensive agricul-
tural activities in United States.
A number of acid and urea herbicides were also detected in the
SR extracts at all sites, with linuron, isoproturon, chlortoluron and
Mecoprop (MCPP) being predominant (Table 1). Linuron accounted
for more than 90% of the sum amount of acid/urea herbicides ab-
sorbed to the samplers at sites 1 and 2, while chlorotoluron dom-
inated the profile of the herbicides at study sites 3, 4 and 5,
accounting for >65% of the components in sites 3 and 4, and
>50% in site 5. Linuron is widely used in vegetable production,
while chlorotoluron has applications in barley and wheat produc-
tion. Non-agricultural uses of acid/urea herbicides also exist, with
their application to control weed growth on hard surfaces, partic-
ularly roads, railways, airport runways, golf courses and public
parks, being an important example (Lapworth and Gooddy,
2006). Although their environmental fate in the aquatic environ-
ment is yet to be fully defined, most herbicides are transformed
by both biotic and abiotic processes and can be biodegraded to
their metabolites which in some instances may be more toxic than
the parent compounds (Virkutyte et al., 2010).
Relatively polar pesticides and acid/urea herbicides, e.g. met-
azachlor (logK
OW
= 2.49) and diuron (logK
OW
= 2.68), were ade-
quately sequestered by the SR-PSDs alongside non-polar
compounds, i.e. PAHs and PCBs with log K
OW
3–8. This demon-
strates the great utility of SR-PSDs, as this level of sensitivity in
detection would have been extremely challenging using conven-
tional sampling and analytical techniques (e.g. Kuster et al.,
2009). This study has demonstrated that SR-PSDs can be employed
to monitor occurrence and distribution of these pesticides in catch-
ment waters and also provides insights into future prospects of
using SR-PSDs quantitatively (for measurement, source identifica-
tion and establishing environmental fate) once the required D
P
and
K
SW
data become available.
3.2. Biological effects analysis
3.2.1. Cytotoxic effects of SR passive sampler extracts
Following 48 h exposure to SR extracts and MeOH (vehicle), no
statistically significant cytotoxicity was measured in RTL-W1 cells
due to any of the extracts from the study sites (e.g. Fig. 4A). In con-
trast, cytotoxicity was observed with the positive control (DCP;
Fig. 4B) indicating that the cells were indeed responsive and any
Fig. 3. Percentage composition of PCB congeners in water at the various sampling
locations. Term descriptions; Tri =
P
CBs 28 and 31; Tetra =
P
CBs 44, 49, 52, 70,
and 74; Penta =
P
CBs 97, 99, 101, 105, 110, 114, 118, and 123; Hexa =
P
CBs
149,132,153,137,138,158,128,167,156, and 157; Hepta =
P
CBs 170, 180,183, 187,
and 189; Octa and Deca =
P
CBs 194 and 209. Data for site 4 not available.
A
B
Fig. 4. Concentration–response curves following 48 h exposure of (A) blank SR
sampler extracts and (B) 3,5 DCP to RTL-W1 cells using neutral red retention assay.
Results of extracts from deployed SR-PSDs were analogous to that of the blank. Cell
viability expressed as percentage of unexposed controls (negative control, NC). Data
points are mean with ± standard deviation (SD) of four replicates at each exposure
concentration. NC = negative control; PC = positive control (1.6 mg L
1
3,5 DCP). PC
and NC are not plotted in the indicated units.
E.S. Emelogu et al. / Chemosphere xxx (2012) xxx–xxx
7
Please cite this article in press as: Emelogu, E.S., et al. Investigating the significance of diss olved organic contaminants in aquatic environments: Coupling
passive sampling with in vitro bioassays. Chemosphere (2012), http://dx.doi.org/10.1016/j.chemosphere.2012.06.041
toxicity present would have been measured. Considering that cyto-
toxicity was not measured with the blank SR extracts, it shows that
the extraction process and solvents used were not inherently toxic
to the cells in this study. The measured concentrations of contam-
inants sequestered by the SR-PSDs were relatively low, hence did
not elicit measureable cytotoxic effects in the neutral red cell via-
bility assays. This result agrees with a previous study in which no
toxicity was observed using Microtox assay in extracts from POCIS
(herbicides) and SPMDs deployed in surface water that was af-
fected by large pesticide inputs and with estimated PAH concentra-
tions three orders of magnitudes higher than observed in this study
(Alvarez et al., 2008). Reduction of cell numbers due to exposure to
xenobiotic compounds is often compensated by a parallel lyso-
somal proliferation which could, to some degree, mask cytotoxic
effects (Hollert et al., 2000).
3.2.2. EROD assays
In vitro bioassays were applied in this study as an alternative
chemical detector, with EROD activity used to determine the over-
all dioxin-like activity of organic compounds in water at each of
the study sites. EROD activity was not observed with either the
blank SR extract (Fig. 5B), or with the vehicle control (methanol,
maximum 1% v/v; not shown). Contrastingly, SR extracts from
the five study sites of the catchment induced statistically signifi-
cant EROD activity (e.g. Fig. 5C and D). Complete EROD activity
concentration–response curves were obtained for most of the SR
extracts and for the TCDD positive control (Fig. 5A). The potencies,
i.e. the maximal level of EROD activity (EC
25
TCDD), were different
in each sample and generally lower than the TCDD positive control
(maximum 50 pM) in each assay (e.g. Fig. 5B–D). At site 2 the high-
est extract concentration (dose equivalent to 66.67 mg SREQ mL
1
)
reduced the EROD activity compared to the peak induction (data
not shown). This should not be attributed to cytotoxicity consider-
ing the NR results, but was a result of sublethal inhibitory effects.
Extracts from site 2 displayed high EROD induction, while sites 1
and 4 were significantly lower. It is curious to note that site 3
showed low EROD induction in comparison to site 2, despite their
near proximity. At site 3, higher concentrations of pesticides and
acid/urea herbicides were detected in the SR-PSD and could be
causing EROD inhibition as has been postulated by other studies
(Babín and Tarazona, 2005; Han et al., 2007).
3.2.3. Bio-TEQ from EROD bioassay
The EC
25
values derived for the positive control TCDD run with
each extract (including the vehicle control) were approximately
the same for each bioassay and were used to calculate Bio-TEQ val-
ues for each site. The Bio-TEQ values were then compared to the
Chem-TEQ values (Fig. 6). Chem-TEQ values were significantly
lower than Bio-TEQ values in SR extracts from all the study sites,
indicating a higher sensitivity (detection) by the bioassay method,
and/or the presence of other dioxin-like compounds not captured
by the PAH and PCB TEQs (e.g. pesticides or other chemicals). Pre-
vious studies have shown similar discrepancies between
chemically calculated TEQ values and bioassay induction values
A
C
B
D
Fig. 5. Concentration–response curves for EROD induction in the RTL-W1 cells by (A) TCDD, (B) SR sampler extracts-blank (C) SR sampler extract-site 2 and (D) SR sampler
extract -site 4. Data points represent the mean ± standard deviation (SD) of four replicates at each exposure concentration. NC = negative control; within plot, B-D the TCDD
(50 pM) is marked. TCDD and NC are not plotted in the indicated units.
Fig. 6. Comparison of Bio-TEQ and Chem-TEQ values obtained from in vitro RTL-W1
assay and concentrations of PAHs and PCBs measured with silicone rubber (SR)
passive sampling technique in water of the study locations.
8 E.S. Emelogu et al. / Chemosphere xxx (2012) xxx–xxx
Please cite this article in press as: Emelogu, E.S., et al. Investigating the significance of dissolved organic contaminants in aquatic environments: Coupling
passive sampling with in vitro bioassays. Chemosphere (2012), http://dx.doi.org/10.1016/j.chemosphere.2012.06.041
(Willett et al., 1997; Brack et al., 2007). Keiter et al. (2008) illus-
trated that combinations of chemical analysis, fractionation tech-
niques and various in vitro assays do not necessarily explain
inductions, even when the concentrations of priority PAHs were
very high. It is possible that other environmental contaminants
including polybrominated diphenyl ethers (PBDEs), PCDD/Fs and
polychlorinated naphthalenes (PCNs) that were not measured in
this study might have contributed to the Bio-TEQ values. Applying
chemical and effects directed fractionation techniques to SR sam-
pler extracts prior to and after chemical analysis and bioassays
could help to identify the compounds present in the complex mix-
tures responsible for the observed EROD responses and help to
bridge the gap between the Chem-TEQ and Bio-TEQ values. Fur-
ther, HOCs may volatilise and adsorb to the walls of microplate,
leading to reduced sensitivity during bioassays. This may contrib-
ute, positively or negatively, to disparity between Chem-TEQ and
Bio-TEQ values depending upon whether EROD inducers or inhib-
itors are preferentially lost from the exposure system. The applica-
tion of partition controlled delivery for bioassays (passive dosing;
e.g. Smith et al., 2010) should provide more stable exposure
conditions and eliminate the use of carrier solvents during aqueous
toxicity assays.
4. Conclusions
The study demonstrates that extracts of SR-PSDs deployed in sur-
face water can be applied with minimal preparation to in vitro cell
line bioassays. These can be used to rapidly and economically
measure the potential impact of complex mixtures of organic con-
taminants and also to detect the presence of toxic compounds not
routinely analysed for. The concentration data of organic
contaminants presented in this study are significant from the
ecotoxicological perspective since SR-PSDs samples reflect the con-
taminant level aquatic organisms are exposed to. SR samplers ab-
sorbed relatively polar pesticides/herbicides, as well as non-polar
compounds extending the potential application of the SR-PSD tech-
nique in regulatory monitoring programmes, particularly in relation
to the challenging LODs set for some compounds under the WFD.
Acknowledgements
We acknowledge the financial support ESE received from the
Institute for Innovation, Design and Sustainability Research
(IDEAS), Robert Gordon University, Aberdeen UK and the German
Academic Exchange Service (DAAD). The authors are thankful to
Fiona Napier, Sue Bowers, Lucy Stevens and other members of
SEPA staff for their various supports in this study. The assistance
provided by Val Johnston (Oceanlab, University of Aberdeen) dur-
ing sampling at site 5 is also appreciated. Finally, we would like
to thank the team at Ecosystem Analysis, Institute for Environmen-
tal Research (Biology V) RWTH University Aachen, for their assis-
tance with the bioassay, particularly Sylvester Heger for helping
to design and implement the 24 well assay templates.
Appendix A. Supplementary material
Supplementary data associated with this article can be found, in
the online version, at http://dx.doi.org/10.1016/j.chemosphere.
2012.06.041.
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Please cite this article in press as: Emelogu, E.S., et al. Investigating the significance of dissolved organic contaminants in aquatic environments: Coupling
passive sampling with in vitro bioassays. Chemosphere (2012), http://dx.doi.org/10.1016/j.chemosphere.2012.06.041
    • "As these SPMDs had been previously deployed in the aquatic environment, the test organisms were thus exposed to mixtures directly collected in the field. In a number of these studies chemical analysis was performed on the passive sampler extracts in which mostly polycyclic aromatic hydrocarbons (PAHs) (Petty et al., 2000; Rastall et al., 2004; Bopp et al., 2007; Ke et al., 2007; Hillwalker et al., 2010; Emelogu et al., 2013), polychlorinated biphenyls (PCBs) (Petty et al., 2000; Hillwalker et al., 2010; Emelogu et al., 2013) and pesticides (Petty et al., 2000; Shaw et al., 2009; Hillwalker et al., 2010 Pesce et al., 2011; Morin et al., 2012) were the target substances. While in some of these studies a correlation between contaminants and the observed effects was found (e.g. "
    [Show abstract] [Hide abstract] ABSTRACT: This study presents a new approach in aquatic toxicity testing combining passive sampling and passive dosing. Polydimethylsiloxane sheets were used to sample contaminant mixtures in the marine environment. These sheets were subsequently transferred to ecotoxicological test medium in which the sampled contaminant mixtures were released through passive dosing. 4 out of 17 of these mixtures caused severe effects in a growth inhibition assay with a marine diatom. These effects could not be explained by the presence of compounds detected in the sampling area and were most likely attributable to unmeasured compounds absorbed to the passive samplers during field deployment. The findings of this study indicate that linking passive sampling in the field to passive dosing in laboratory ecotoxicity tests provides a practical and complimentary approach for assessing the toxicity of hydrophobic contaminant mixtures that mimics realistic environmental exposures. Limitations and opportunities for future improvements are presented. Copyright © 2015 Elsevier Ltd. All rights reserved.
    Article · Mar 2015
    • "The observed low CYP1A activity and lack of acute cytotoxicity corresponds to the low concentrations of freely dissolved PAHs and PCBs measured in the catchment. The total concentrations of freely dissolved PAHs (∑PAHs 40 ; parent and branched) and PCBs (∑PCBs 32 ; ortho and mono ortho-PCBs) measured in water at the five sites during the first sampling campaign varied from 38 to 69 ng l −1 and 0.02 to 0.06 ng l −1 (Emelogu et al. 2013a). In the second sampling campaign, the freely dissolved concentrations of ∑US-EPA 16 PAHs and ∑7 Indicator-PCBs detected in water at the three sites range from 16 to 28 ng l −1 and 0.003 to 0.008 ng l −1 , respectively (Emelogu et al. 2013b). "
    [Show abstract] [Hide abstract] ABSTRACT: A wide variety of organic contaminants including pesticides, polycyclic aromatic hydrocarbons (PAHs) and polychlorinated biphenyls (PCBs) have previously been detected in surface waters in the river Ythan catchment, North East Scotland UK. While the concentrations detected were below Water Framework Directive Environmental Quality Standards (WFD-EQSs) environmental exposures to the diverse mixtures of contaminants, known and unknown, may pose chronic and/or sublethal effects to non target organisms. The present study assessed the embryo and algal toxicity potential of freely dissolved organic contaminants from the Ythan catchment using silicone rubber passive sampling devices (SR-PSDs) and miniaturised bioassay techniques. Zebrafish (Danio rerio) embryos and marine phytoplankton species (Diacronema lutheri) were exposed to extracts from SR-PSDs deployed at different locations along the river Ythan and an undeployed procedural blank. Statistically significant developmental and algal toxicities were measured in all tests of extracts from deployed samples compared with the procedural blanks. This indicates environmental exposure to, and the combined toxicity potential of, freely dissolved organic contaminants in the catchment. The present and previous studies in the Ythan catchment, coupling SR-PSDs and bioassay techniques, have both helped to understand the interactions and combined effects of dissolved organic contaminants in the catchment. They have further revealed the need for improvement in the techniques currently used to assess environmental impact.
    Full-text · Article · Jan 2014
    • "Scott et al. (2012) showed levels of PAHs in semi permeable membrane devices (SPMDs) from five Irish lakes ranged from 125 to 577 pg L À1 which are lower than the PAHs estimated in this study. PAHs reported from PDMS PSDs deployed in four remote streams in Northeast Scotland ranged from 38 to 68 ng L À1 , which are similar to results reported in this study (Emelogu et al., 2013). Overall, low molecular weight PAHs dominate in water samples, while higher molecular weight compounds dominate in lake sediments , thus supporting water/particulate partitioning theory and the low levels as reported by Scott et al. (2012) Santillo et al. (2005) in a composite eel sample from the Netherlands and a mean of 1545 ng g À1 w.w. "
    [Show abstract] [Hide abstract] ABSTRACT: Homologue and congener profiles of PCDD/Fs in eels, passive sampler and sediment extracts from the Burrishoole, a rural upland catchment on the western Irish seaboard were compared with potential PCDD sources. ΣPCDD/F levels in eels ranged from 2.9 to 25.9 pg g(-1) wet weight, which are elevated compared to other Irish locations. The OCDD congener dominated the pattern of ΣPCDD/Fs in all matrices from Burrishoole. Passive samplers were successfully deployed to identify for the first time the presence in the water column of PCDD/Fs and dimethoxylated octachlorodiphenyl ether (diMeOoctaCDE), impurities found in pentachlorophenol (PCP) production. Principal component analysis (PCA) identified similarities between PCDD/F profiles in technical PCP mixtures and environmental samples from the Burrishoole region. Results strongly suggest residual PCDD contamination associated with historic local use of a dioxin contaminated product in the catchment area, with pentachlorophenol a strong candidate.
    Article · Oct 2013
    P WhiteP WhiteB McHughB McHughR PooleR Poole+1more author...[...]
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