Dynamic Article LinksC<Journal ofEnvironmentalMonitoringCite this: DOI: 10.1039/c1em10591d

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An examination of the toxic properties of water extracts in the vicinity of anoil sand extraction site

F. Gagn�e,*a C. Andr�e,a M. Douville,a A. Talbot,ab J. Parrott,b M. McMasterb and M. Hewittb

Received 22nd July 2011, Accepted 9th August 2011

DOI: 10.1039/c1em10591d

The industrial extraction of oil sands (OS) in northern Alberta, Canada, has raised concerns about the

quality of the Athabasca River. The purpose of this study was to examine the toxic properties of various

water extracts on Oncorhynchus mykiss trout hepatocytes. The water samples were fractionated on

a reverse-phase C18 cartridge and the levels of light-, medium- and heavy-weight polycyclic aromatic

hydrocarbons (PAHs) were determined by fluorescence spectroscopy. Primary cultures of trout

hepatocytes were exposed for 48 h at 15 �C to increasing concentrations of the C18 extract

corresponding to 0.02, 0.1, 0.5 and 2.5X concentrations from upstream/downstream sites in the

Athabasca River, lake and groundwater samples, OS tailings and interceptor well-water samples.

Changes in cell viability, phase I and phase II biotransformation enzymes (cytochrome P4501A and

glutathione S-transferase activities), oxidative damage (lipid peroxidation LPO) and genotoxicity

(single and double DNA strand breaks) were monitored in post-exposure cells. The water samples

decreased cell viability and increased all the above endpoints at thresholds of between 0.02 and 0.1X the

water concentration. The most responsive biomarker was DNA damage but it also offered the least

discrimination among sites. LPO was higher at sites downstream of the industrial operations compared

to upstream sites. A decision tree analysis was performed to formulate a set of rules by which to identify

the distinctive properties of each type of water samples. The analysis revealed that OS tailings and

interceptor waters were characterized by an increased concentration in light PAHs (>42 mg L�1) and

this fraction represented more than 85% of the total PAHs. These samples also inhibited GST activity,

which could compromise the elimination of genotoxic PAHs present in the system. An analysis of

groundwater samples revealed a contamination pattern similar to that for OS tailings. There is

a need for more research into specific biomarkers of toxicity from OS tailings compounds such as

naphthenic acids, light PAHs among others, which are a characteristic fingerprint of OS extraction

activities.

aFluvial Ecosystem Research, Aquatic Ecosystem Protection Division,Water Science and Technology, Environment Canada, 105 McGill Street,Montr�eal, Quebec, Canada H2Y 2E7. E-mail: francois.gagne@ec.gc.cabAquatic Ecosystem Protection Research Division, Water Science andTechnology, Environment Canada, 867 Lakeshore Rd., Burlington,Ontario, Canada L7R 4A6

Environment impact

This work examines the toxic effects of surface water extracts from a

understanding on the release and toxic properties of various wa

extraction area is mandatory to understand the impact of mining

Rainbow trout hepatocytes were used as an alternative to whole fish

of industrial and municipal effluents and single substances and to

pollutants.

This journal is ª The Royal Society of Chemistry 2011

Introduction

Oil sands (OS) are a sandy matrix rich in bitumen and containing

heavy lipophilic aliphatic and aromatic hydrocarbons. These

hydrocarbons make up most of the crude oil and are recovered

by either surface mining or in situ steam injection. The industrial

production of crude oil from bitumen is estimated at more than

1.3 million barrels per day in an area covering 530 km2 wherein

n oil sand rich area supporting intense extraction activities. The

ter samples collected upstream and downstream the oil sand

activities in the Athabasca River system (Alberta, Canada).

following a protocol that is currently used to monitor the quality

provide insights on the toxic mode of action of environmental

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tailings ponds take up approximately 20% of the area.1,2 It is

estimated that the production of crude oil from OS will reach

a volume on the order of 2.9 million barrels per day by 2020.3

While the increasing activity of OS extraction by industry is of

great economic interest for Canadians, it also raises some

concerns about potential harmful impacts on the environment,

given the tremendous scale of the development area.

Crude oil is obtained by an industrial process involving alka-

line hot water extractions wherein the crude oil forms an emul-

sion at the top of the aqueous alkaline water phase. This aqueous

phase is made up of fine particles and alkali-labile organic

compounds from bitumen and represents the bulk material

contained in tailing ponds. Indeed, the tailings are composed of

fine suspended particles, metals, anionic aromatic and aliphatic

hydrocarbons. For metals, concentrations of a number of

elements such as cadmium, copper, lead, mercury, nickel, silver,

and zinc are found at higher concentrations near the develop-

ment area than upstream of it.4 Moreover, a significant

proportion of metals are bound to the organic matrix involving

protoporphyrin-metal complexes.5,6 Naphthenic acids (NAs) are

a family of compounds composed mostly of carboxylated cyclic

aliphatic hydrocarbons. This important organic family has been

demonstrated to account for the main toxic properties of tail-

ings.7–9 NAs represent the major class of alkali-extractable

organics in OS tailings, reaching levels as high as 50 mg/L.10 NAs

are a complex class of compounds governed by the general

formula CnH2n+zOx whose specific measurements represent huge

analytical challenges because of their homogeneous physical and

chemical properties. A recent study used 2-dimensionnal gas-

chromatography and time of flight mass spectrometry to detect

bi and tri pentacyclic acids fitting the formula for NAs.11 Some

NAs have aromatic rings permitting detection by synchronous

fluorescence spectroscopy.12 The presence of aromatic

compounds is associated with both commercial mixtures of NAs

and OS tailing water samples. The above study also revealed that

1–3 aromatic ring polycyclic aromatic hydrocarbons (i.e. light

PAHs) were found at much higher concentrations in tailing

ponds than at upstream sites in the river. These upstream river

sites contained more of the heavy (and highly genotoxic) PAHs

(four or more aromatic rings) from the natural leaching of OS in

the river. Dissolved PAHs concentrations in the tributaries of the

Athabasca River increased from 0.009 mg L�1 upstream to 0.2 mg

L�1 downstream of the OS mining sites in the summer.8 This

study corroborated the contention that water samples from these

sites are contaminated by the high density of naturally occurring

OS, with industrial extraction activities contributing to the

increased mobilization of total PAHs in the area. A previous

study also showed that snow is an important vector of PAH-

contaminated dust in this ecosystem, with snow samples con-

taining up to 4.8 mg L�1 of PAHs at the end of winter and could

represents another source of contamination to the river systems.8

The release of light PAHs and other compounds into an

ecosystem already contaminated by naturally occurring and/or

industrially disturbed OS heavy PAHs might represent a risk on

aquatic ecosystems. The discrimination of ecotoxicological

effects between natural and anthropogenic release of OS-derived

contaminants is a challenge. In a previous study, adult Yellow

Perch Perca flavescens were exposed to aged OS tailings water

contained in experimental ponds for up to 10 months in an

J. Environ. Monit.

attempt to determine the potential toxic effects of OS tailings

contamination.13 Following the exposure period, severe gill

erosion and viral-induced tumours were observed. The frequen-

cies of these observations were correlated to concentrations of

OS-related compounds such as the major ions and total naph-

thenate concentrations. Medium (3–4 aromatic rings) and heavy

PAHs such as pyrene and benzo(a)pyrene are well-known

immunosuppressive, genotoxic and carcinogenic substances to

aquatic animals.14–16 The manifestation of genotoxicity requires

phase 1 biotransformation by cytochrome P450 1A1 to produce

DNA-reactive intermediates.17 The conjugating activity of

glutathione S-transferase (GST) activity is also involved in the

elimination of hydroxylated PAH intermediates by the bile.16

The activity of cytochrome P4501A1 (7-ethoxyresorufin O-dee-

thylase, EROD) was induced by heavy PAHs in the Nile tilapia

but not by light PAHs.18 Interestingly, GST activity was induc-

ible by smaller PAHs such as phenanthrene (three aromatic

rings), fluoranthrene (three aromatic rings) and chrysene (four

aromatic rings). No information exists on the effects of light

PAHs and NAs on both EROD and GST activities. The toxicity

of PAHs also involves oxidative stress during biotransformation,

which could lead to major cell damage in the form of lipid per-

oxidation (LPO) or genotoxicity.19 In fathead minnow Pime-

phales promelas larvae exposed to OS, larval mortality was

directly correlated to the cytochrome P4501A1 activity that fol-

lowed exposure to oil sands.20 This suggests that three or more

aromatic ring PAHs were responsible, at least in part, for the

observed toxicity. In mature goldfish Carassius auratus, the basal

levels of testosterone were reduced in both males and female

exposed to an OS-contaminated pond compared to a control

pond, suggesting compromised steroidogenic capacity in the

gonad.21 In addition, fathead minnows held in an aged OS water

pond (>15 years but still containing 10 mg L�1 NAs with

a conductivity of 2000 mS*cm�1) completely inhibited spawning

and diminished male secondary sexual characteristics.22

The purpose of this study was to examine the sub-lethal effects

of water extracts from an intensive OS-extraction area on

primary cultures of rainbow trout hepatocytes. The sub-lethal

biomarkers examined were biotransformation activity (cyto-

chrome P4501A1 and GST activity), oxidative stress LPO and

genotoxicity (DNA strand breaks). The water samples comprised

a selection of OS tailings, OS tailing pond interceptor wells,

a reference lake, groundwater and Athabasca River sites

following an upstream-downstream gradient. The relative levels

of light, medium and heavy PAHs in the water extracts were also

examined by scanning fluorescence spectroscopy to highlight

signatures with respect to the levels and proportion of light and

heavy PAHs. An attempt was made to formulate a set of rules

based on biomarker responses and PAH concentrations by which

to characterize the various types of water samples (river, lake,

groundwater, OS tailing ponds) in this OS-rich area.

Methods

Water sample collection and preparation

Grab-water samples (120 L) were collected at five sites in the

Athabasca River, two oil sands tailing ponds (OS1 and OS2) and

interceptor well samples, one surface-water sample fromGregoire

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Lake and groundwater samples. The reference sites were one site

upstream of Fort McMurray on the Athabasca River (near the

intakeof adrinkingwater treatment plant) and surfacewater from

Gregoire Lake. This lake (N 56 26.237; W 111 05.179) is located

approximately 15 km south of Fort McMurray, Alberta. The

upstream Athabasca River site is located about 2–3 km from the

city centre of FortMcMurray and the Northland sawmill site was

located just downstream the city of Fort McMurray below the

municipal sewage treatment plant discharge. The exact

geographic locations and basic physical and chemical properties

of the water samples are shown in Table 1. The samples were

transported to the laboratory, kept at 4 �C in the dark, and

immediately sent to various laboratories over the following days.

Upon reception in our laboratory, the water samples were frac-

tionated on C18 solid-phase mini-columns (Sep-Pak C18, Waters

Associates, Inc)whichwere activatedwith 2mLof ethanol (100%)

and with 10 mL bidistilled water. Before the extraction step, the

water samples were filtered on a 0.4-mm-pore polycarbonate

membrane filter to remove suspended solids andmicroorganisms.

A volume of 500 mL was passed through a reverse-phase C18

cartridge (360 mg) under vacuum (5 psi), washed with 10 mL of

bidistilled water and the material eluted with 1 mL of analytical-

grade ethanol (Absolute; Sigma-Aldrich Chemical Co., Ontario,

Canada). The ethanol fraction was kept at �20 �C until analysis.

Fixed wavelength fluorescence polycyclic aromatic hydrocarbon

(PAH) analysis

The levels of PAH were determined by fixed-wavelength fluores-

cence spectroscopy.23,24Avolumeof 50mLof each of the ethanolic

C18 water extracts was mixed with 150 mL of ethanol in a dark

microplate and analyzed by fluorescence using a dual mono-

chromator-basedmicroplate reader (Biotek Inc.,USA).The light-

weight PAH group includes PAHs containing 2–3 rings (e.g.

naphthalene andphenanthrene) andwas determined by excitation

at 290 nm and emission at 340 nm. Standard additions of naph-

thalene were used for calibration. Medium-weight PAHs (3–4

rings) include fluoranthene and chrysene; theywere determinedby

excitation at 325 nm and emission at 370 nm. Calibration was

achieved with standard additions of pyrene. The heavy-weight

PAH group includes PAHs with more than four rings, such as

benzo(a)pyrene and benzo(k)fluoranthene; they were determined

at 385 nm for excitation and at 440 nm for emission. Standard

Table 1 Site location and general physico-chemical characteristics of tailing

Sites Geographical location Date of co

Upstream N56 43 28.20; W111 24 06.90 Sept. 2009Northland N56 52 11.2; W111 26 26.5 Sept. 2009OS1 downstream N57 03 45.9; W111 31 11.8 Sept. 2009Muskegg N57 07 45.4; W111 36 22.4 Sept. 2009Ells N57 18 45.69; W111 39 98.6 Sept. 2009OS 1 interceptor well N 56 58 56.11; W 111 26 46.17 June 2010OS 2 interceptor well N 57 5 48.19; W 111 37 22.80 June 2010OS1 tailing pond N 56 53 56.19; W 111 23 3.23 Sept. 2009OS2 tailing pond N 57 4 46.54; W 111�3806.17 Sept. 2009Surface Water Gregoire Lake N 56 28.59 17; W 111 10 51 68 June 2010Groundwater OS2 Mildred Lake N56 99 16.3; W111 04 4.7 June 2010

a Not determined.

This journal is ª The Royal Society of Chemistry 2011

additions of benzo(a)pyrene were used for calibration at the cor-

responding wavelengths. The above wavelengths were selected to

avoid spectral overlap where the analytical signal from each PAH

group was measured independently.23 The data were expressed as

mg L�1 naphthalene, pyrene and benzo(a)pyrene equivalents for

the light, medium and heavy PAHs, respectively.

Preparation and exposure of rainbow trout hepatocytes

Primary cultures of rainbow trout (Oncorhynchus mykiss) hepa-

tocytes were freshly prepared using the double perfusion method

developed by Klauning et al.25 with some modifications.26

Young-of-the-year trout (15–20 cm fork length) were humanely

euthanized in 100 mg L�1 tricaine methanesulphonate (MS-222)

buffered with 100 mg L�1 NaHCO3, pH 7, for 5 min at 15 �C in

accordance with the recommendations of the Canadian animal

care committee. The liver of three fish were dissected, immedi-

ately transferred to a Petri dish and perfused with 20 mL of ice-

cold phosphate buffered saline containing 10 mM EDTA at

a flow rate of 2 ml/min. The perfusion buffer was then replaced

with phosphate buffered saline containing 1.5 mM CaCl2 and

100 units/ml collagenase, pH 7.4 and the livers perfused with

25 ml at a 5 ml min�1 flow rate followed by a 5–10 min incubation

at room temperature. The livers were then transferred in ice-cold

Leibovitz L-15 Medium (L15) supplemented with antibiotics

(100 mg mL�1 streptomycin and 100 units/mL penicillin), an

antimycotic (0.025 mg mL�1 amphotericin B), 10% Fetal Bovine

Serum (FBS) and 10 mM Hepes-NaOH, pH 7.4 and shaken

gently on ice for 30 min. The resulting suspension was dissociated

and filtered through a 100-mm stainless-steel mesh (Tissue

Dissociation Kit, Sigma-Aldrich Chemical Co., Ontario, Can-

ada) and cells were collected by centrifugation at 72� g for 2 min

at 4 �C, the medium removed and the cells washed with Dul-

becco’s phosphate buffered saline at pH 7.4. The cells were then

resuspended in 10 mM Tricine, 0.85% NaCl, pH 7.4 and purified

by iodixanol density barrier (OptiPrep�, Axis-Shield, Norway).

The hepatocytes were finally washed in Dulbecco’s phosphate

buffered saline, centrifuged (72 � g for 2 min at 4 �C) and the

pellet resuspended in L-15 medium without serum. The cell yield

was assessed by the use of an hemocytometer and the viability

was determined in the presence of 0.2% Trypan blue under

microscopic examination (dead cells retain the blue-coloured

dye). The initial cell viability was >96%. The cells were plated in

ponds, surface and groundwatera

llection Ammonia mg L�1 pH Conductivity uSx cm-1 DOC mg L�1

< 0.05 8.3 191 —< 0.05 7.7 200 —< 0.05 8.3 198 —< 0.05 8.3 190 —< 0.05 8.1 182 —— 7.6 1420 45— 7.6 3190 673 8.77 1920 —6 8.23 2150 —< 0.05 6.9 130 10.5— 7.3 2640 43

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24-well microplates at a density of 1 � 106 cells per mL of L-15

medium and incubated overnight at 15 �C in a dark and humid

atmosphere prior to replacement of half the initial volume of

culture medium by medium containing the test extracts or

controls. The cells were exposed to increasing concentrations of

ethanol extracts (0.004, 0.02, 0.1 and 0.5%) using ethanol as the

solvent control. The exposure period was 48 h at 15 �C. Cellswere also exposed to positive controls such as a mixture of 0.1 mg

mL�1 b-naphtoflavone to control for between batch of hepato-

cyte preparations. After the exposure period, the plates were

centrifuged briefly at 100 � g (5 min) at 4 �C and the exposure

media removed by aspiration. The cells were washed with 1ml of

D-PBS and resuspended in 0.5 ml of D-PBS.

Cytotoxicity and genotoxicity assessments

Cell viability was determined according to the fluorescein

retention test using carboxylfluorescein diacetate as the fluores-

cent probe.25 The activity of cytochrome P4501A1 was deter-

mined using the 7-ethoxyresorufin assay in hepatocyte

aggregates.27 An aliquot of hepatocyte suspension (5 � 104 cells)

was mixed with 20 mM of 7-ethoxyresorufin and 100 mM of

reduced NADPH. The incubation mixture was allowed to incu-

bate at 30 �C for 0, 10, 20 and 30 min and fluorescence

measurements were taken at 520 nm excitation and 590 nm

emission wavelengths (Biotek multiplate reader, USA). Standard

solutions of 7-hydroxyresorufin were used for calibration.

Fluorescence data were normalized with cell density or total

proteins using the protein dye binding assay.28 DNA damage

(single and double stranded DNA breaks) was assessed in

exposed hepatocytes by the alkaline precipitation assay modified

for the fluorescent quantification of DNA strand breaks in the

presence of trace amounts of detergent.29,30 The assay principle is

based on K-assisted SDS precipitation of genomic DNA (asso-

ciated with proteins), which leaves protein-free DNA breaks

(single and double stranded) in the supernatant. Salmon sperm

DNA standards were used for DNA calibration using the SYTO

dye methodology. The results were expressed as mg of superna-

tant DNA/cell density. LPO was determined by a spectrofluoro-

metric assay, which uses thiobarbituric acid for the

determination of malonaldehyde.31 Briefly, 20 mL of hepatocyte

Table 2 Fluorescence PAH analysisac

SitesLight PAHs(mg L�1)

Me(mg

Control < dL < dUpstream 12.8 � 0.8 24Northland 32 � 0.9 36Downstream OS1 14.8 � 0.9 28Muskeg River confluence 10.8 � 0.8 22Ells River confluence 24 � 1 12Gregoire Lake surface water 31.4 � 6 15.Mildred Lake groundwater 372 � 40 42OS1 tailings 1 970 � 87 60OS1 Interceptor well 972 � 63 136OS2 tailings 5 000 � 660 232OS2 Interceptor well 11 000 � 380 188

a Fluorescence analysis was performed on C18 SPE ethanol extracts. Light, mepyrene equivalents (mg L�1) respectively.b % light PAHs/(# light + medium +shown in bold.

J. Environ. Monit.

suspension was mixed with 130 mL of bi-distilled water and

mixed with 25 mL of 10% trichloroacetic acid solution containing

1 mM FeSO4 and 50 mL of 0.67% thiobarbituric acid. The

mixture then stood in a hot-water bath (70–80 �C) for 10 min,

cooled at room temperature for 15 min and fluorescence

measured at 540 nm excitation and 600 nm emission (Biotek

microplate reader, USA). Blanks and standards of tetrame-

thoxypropane (stabilized form of malonaldehyde) were prepared

in the presence of the PBS solution. Because the reagent could

react with other aldehydes, results were expressed as mg of thi-

obarbituric acid reactants (TBARS)/cell density. Glutathione

S-transferase (GST) was determined a spectrophotometric

methodology using 2,4-dichloronitrobenzene as the co-substrate

and reduced GSH.32 The data were expressed as the increase in

absorbance at 340 nm min�1/total proteins.

Data analysis

Rainbow trout hepatocytes were exposed to the surface-water

extracts in quadriplicate (n ¼ 4) and the exposure experiments

were repeated twice using a different water sample. The data

were normalized and reported as fold changes in the solvent (i.e.

ethanol) control group: value of treatment/mean value of

controls. The difference between treatments was determined

using ANOVA and critical differences between treatment groups

were appraised using the Mann-Whitney U test. Correlation

analyses were performed using the Pearson-moment procedure.

Significance was set at p < 0.05 but marginal changes (0.05 <

p < 0.1) were reported. A discriminant function analysis and

a decision tree analysis were performed for site classification

purposes and to identify which of the water properties (i.e., PAH

measurements and measured biomarkers) were associated with

the type of water sample.

Results

Fluorescence spectroscopic analyses

The levels of PAH in the water extracts were determined by

scanning fluorescence spectroscopy (Table 2). The data revealed

that all water extracts contained detectable amounts of light

dium PAHsL�1)

Heavy PAHs(mg L�1) % light PAHsb

L <dL —� 0.9 14 � 0.2 25� 0.9 14 � 0.2 39� 0.4 18 � 0.2 24� 0.25 22 � 0.2 20� 0.25 11.2 � 0.2 516 � 1 4 � 0.1 63� 5 6 � 0.3 89� 2 18 � 0.2 95� 14 10 � 0.7 87� 5 20 � 0.7 69� 9 8 � 0.5 98

dium and heavy PAHs are expressed as naphthalene, pyrene and benzo(a)heavy PAHs).c Water extracts from industrial extraction activities are

This journal is ª The Royal Society of Chemistry 2011

Table

3Cytotoxicityandsublethaleffectsofriver

surface

waterandoilsandtailingsa

bc

Sites

Upstream

Northland

OS1downstream

Muskeg

River

confluence

EllsRiver

confluence

OS1/O

S2

0.004

0.02

(0.1)

0.5%

0.004

0.02

0.1

0.5%

0.004

0.02

0.1

0.5%

0.004

0.02

0.1

0.5%

0.004

0.02

0.1

0.5%

0.004

0.02

0.1

0.5%

Viability

10.96

0.98

1.1

0.91

0.99

1.0

0.89

0.96

0.94

0.93

0.90

0.95

1.04

1.05

10.9

0.92

0.9

11

0.95

0.9

0.8*

(0.15)

(0.2)

(0.1)

(0.2)

(0.1)

(0.2)

(0.1)

(0.15)

(0.1)

(0.2)

(0.1)

(0.2)

(0.1)

(0.1)

(0.2)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.2)

(0.2)

0.95

11

0.85*

(0.1)

(0.1)

(0.1)

(0.05)

DNA

damage

1.6*

1.5*

1.5*

1.3*

1.3

1.2

1.9*

0.8*

1.1

1.4*

1.3*

1.6*

1.6*

2*

1.8*

2.3*

11.2

1.4*

1.7*

1.6*

1.3

0.8*

0.6*

(0.2)

(0.2)

(0.2)

(0.1)

(0.1)

(0.1)

(0.2)

(0.1)

(0.1)

(0.1)

(0.1)

(0.2)

(0.2)

(0.2)

(0.2)

(0.2)

(0.1)

(0.1)

(0.1)

(0.2)

(0.1)

(0.1)

(0.1)

(0.1)

1.1

1.2

1.5*

2.3*

(0.1)

(0.1)

(0.2)

(0.2)

LPO

1.2

0.95

1.1

0.92

1.8*

0.91

1.0

0.8*

1.9*

0.8

0.74*

1.1

2.6*

1.6*

2.1*

1.5*

1.1

0.9

1.4*

1.9*

2.2*

1.2

0.91

0.68*

(0.1)

(0.1)

(0.1)

(0.1)

(0.2)

(0.1)

(0.1)

(0.05)

(0.2)

(0.1)

(0.1)

(0.1)

(0.25)

(0.2)

(0.25)

(0.2)

(0.1)

(0.1)

(0.2)

(0.2)

(0.2)

(0.1)

(0.1)

(0.1)

1.4*

0.8

0.86

0.78*

(0.1)

(0.1)

(0.1)

(0.07)

GST

0.85

0.85

0.86

0.91

0.83

0.81

0.81

0.81

0.85

0.74*

0.83*

0.83*

0.9

0.95

10.95

0.7*

0.7*

0.7*

0.9

0.81

0.51*

0.49*

0.58*

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.2)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.15)

(0.15)

0.80

0.67*

0.52*

0.5*

(0.1)

(0.1)

(0.1)

(0.1)

EROD

0.9

0.82

0.93

0.9

11.4*

1.3*

0.42*

0.96

1.1

1.2

1.3*

0.87

1.1

1.05

0.94

1.1

1.0

0.97

1.2

1.4*

3*

2*

1(0.2)

(0.1)

(0.05)

(0.1)

(0.1)

(0.2)

(0.15)

(0.2)

(0.1)

(0.05)

(0.1)

(0.1)

(0.1)

(0.1)

(0.15)

(0.05)

(0.05)

(0.1)

(0.1)

(0.05)

(0.1)

(0.25)

(0.25)

(0.1)

10.95

0.75*

0.45*

(0.05)

(0.1)

(0.1)

(0.15)

aThedata

are

expressed

asmean(�

standard

deviation)oftheresponse

factor,i.e.,activityofthetreatm

ent/meanvalueofsolventcontrols.bSignificantdifferencesare

highlightedin

bold

andwiththe*

symbol.cTheexposure

concentrationsare

expressed

in%

v/v

oftheethanolicC18extract.

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.

Table

4Cytotoxicityandsublethaleffectsoftailingpondinterceptor,groundwaterandlakewatersamplesa

bc

Sites

GregoireLakesurface

water

SyncrudeMildredLakegroundwater

OS2interceptor

OS1interceptor

0.004

0.02

0.1

0.5%

0.004

0.02

0.1

0.5%

0.004

0.02

0.1

0.5%

0.004

0.02

0.1

0.5%

Viability

1.1

1.1

1.2

1.2

11

10.99

0.97

1.1

0.8

0.9

11.1

0.9

0.8

*

(0.1)

(0.1)

(0.2)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.2)

1.1

11

0.9

1.2

1.1

1.1

1.1

0.95

1.1

11.1

0.8

1.1

0.98

1(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.05)

(0.1)

(0.1)

(0.1)

(0.1)

(0.2)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

DNA

damage

1.4

*1.8

*2.6

*2.6

*1.5

*1.9

*2.1

*2*

1.6

*1.8

*2.2

*2.4

*1.2

*1.8

*1.7

*2.1

*

(0.2)

(0.2)

(0.2)

(0.3)

(0.2)

(0.2)

(0.3)

(0.2)

(0.1)

(0.1)

(0.2)

(0.1)

(0.1)

(0.1)

(0.2)

(0.2)

2.1

*2.3

*2.3

*2.3

*1.7

*2.1

*2*

1.8

*2*

1.7

*1.5

*1.3

*1.6

*1.7

*1.7

*1.3

*

(0.2)

(0.2)

(0.3)

(0.2)

(0.2)

(0.2)

(0.2)

(0.2)

(0.2)

(0.2)

(0.2)

(0.1)

(0.2)

(0.2)

(0.2)

(0.2)

LPO

1.4

*1.2

1.3

*1.2

1.9

*1.7

*1.6

*1.6

*1.7

*1.2

1.1

0.95

1.9

*1.3

*1.2

1.4

*

(0.1)

(0.1)

(0.2)

(0.1)

(0.2)

(0.2)

(0.2)

(0.2)

(0.1)

(0.1)

(0.2)

(0.1)

(0.2)

(0.1)

(0.1)

(0.1)

1.4

*1.2

1.3

*1.2

1.6

*1.4

*1.2

10.7

*1.6

*–

1.7

*–

(0.2)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.15)

(0.2)

(0.2)

GST

1.1

1.1

1.2

*1.1

11.1

11

0.95

1.1

0.98

10.9

0.98

0.97

0.99

(0.1)

(0.1)

(0.05)

(0.1)

(0.1)

(0.1)

(0.1)

(0.05)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.2)

(0.2)

1.1

1.1

1.2

*1.1

0.99

11

11.2

1.1

0.99

0.8

0.8

*1

0.98

0.92

(0.05)

(0.05)

(0.1)

(0.1)

(0.05)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.05)

EROD

0.85

1.2

1.3

*1.5

*0.8

11.2

1.4

*0.6

1.1

1.2

1.2

0.81

1.1

1.3

*0.9

(0.1)

(0.1)

(0.1)

(0.2)

(0.1)

(0.025)

(0.1)

(0.2)

(0.05)

(0.1)

(0.2)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

0.9

1.1

1.2

1.4

*1.1

1.2

1.3

*1.3

*0.8

1.2

1.3

*0.9

0.91

0.89

1.4

*0.5

*

(0.2)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.15)

(0.1)

(0.1)

(0.1)

(0.1)

(0.1)

(0.2)

(0.2)

aThedata

are

expressed

asmean(�

standard

deviation)oftheresponse

factor,i.e.,activityofthetreatm

ent/meanvalueofsolventcontrols.bSignificantdifferencesare

highlightedin

bold

andwiththe*

symbol.cTheexposure

concentrationsare

expressed

in%

v/v

oftheethanolC18extract.

J. Environ. Monit. This journal is ª The Royal Society of Chemistry 2011

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(2–3 aromatic rings), medium (4 aromatic rings) and heavy (5

aromatic rings) PAHs. In general, the extracts of surface water

from the Athabasca River contain relatively low levels of PAHs:

in descending order, heavy PAHs > medium PAHs > light

PAHs. However, water extracts taken from OS tailing ponds

and interceptor waters contained significantly higher levels of

light and medium PAHs: in descending order, light PAHs [

medium PAHs > heavy PAHs. This suggests that water samples

from an OS-rich area bring about the release of heavy PAHs to

the water column while the wastewaters from the industrial

extraction process bring about the release of light and medium

PAHs to the tailing ponds. Indeed, the proportion of light PAHs

from tailing ponds OS1 and OS2 represented 70–98% of total

PAHs, in contrast to the surface-water samples, in which light

PAHs represented on the order of 32% of PAHs. The proportion

of light PAHs in the surface-water extracts from Gregoire Lake

was 63%of totalPAHs, indicatingadifferent contaminationpattern

from the surface water in this area. The groundwater samples from

Mildred Lake near OS2 contained 89% of light PAHs, suggesting

a similar contamination to that of OS tailings water.

Cytotoxicity and sublethal effects of river water and OS tailings

samples

The cytotoxic properties of surface water and OS tailings water

extracts are reported in Tables 3 and 4. First, exposure to b-NF

at 0.1 mg mL�1 significantly reduced cell viability by 25% in cells

and increased EROD activity and DNA strand breaks levels in

cells. Cell viability was significantly reduced in OS tailings at

a threshold concentration of 0.2%. The concentration of 0.2% of

the ethanol extracts corresponds to the undiluted sample water

i.e., 1X. The surface water and OS extracts displayed strong

genotoxic responses. In the OS ponds, a significant increase in

DNA strand-break levels was observed for OS1 at the lowest

concentration (0.004% or 0.02X) in trout hepatocytes, which was

followed by a significant drop in DNA strand breaks indicative

of reduced DNA repair activity. In OS2, an increase in DNA

strand breaks was observed at a threshold concentration of 0.02

to 0.1X. A correlation analysis revealed that the cell viability

index was positively related to DNA strand breaks for both OS

tailings. In surface waters, DNA damage was also apparent at all

sites, with no significant upstream or downstream effect (Mann-

Whitney U test with upstream and downstream sites p > 0.1). A

correlation analysis revealed that cell viability and DNA damage

were significantly correlated at the Northland, OS1 downstream

and Muskeg sites; the strongest correlation was observed at the

second closest downstream site from OS1 (i.e. at the Muskeg

confluence site in the Athabasca River).

Oxidative stress in rainbow trout hepatocytes exposed to the

various types of water samples were also examined by measuring

LPO (Tables 3 and 4). OS tailings samples readily increased LPO

at the lowest exposure concentration (0.02X). OS1 was more

potent than OS2 in increasing LPO in exposed hepatocytes, with

a 2.2-fold increase in LPO for OS1 at the lowest exposure

concentration (0.02X). A correlation analysis revealed that LPO

was marginally correlated with cell viability (r ¼ 0.31; p < 0.1);

however, it was highly correlated with DNA strand breaks (r ¼0.74; p < 0.001). For surface waters, LPO levels were significantly

affected by almost all sites except the upstream site. The levels of

This journal is ª The Royal Society of Chemistry 2011

LPO were significantly higher at the downstream sites at the

confluence of the Ells and Muskeg rivers compared with the

upstream sites (upstream, Northland, OS1 downstream) (Mann-

Whitney U test; p < 0.05). A correlation analysis of the surface-

water extracts revealed that LPO was significantly correlated

with DNA damage at the upstream, Muskeg and Ells sites.

The activity of the phase-2 xenobiotic conjugating enzyme

GST was also examined (Tables 3 and 4). In OS tailings, GST

activity was significantly inhibited at 0.02% (0.1X) extract

concentrations. A correlation analysis revealed that GST activity

was marginally correlated with LPO (r ¼ 0.31; p < 0.1) and with

cell viability (r ¼ 0.31; p < 0.1) in OS2. In surface waters, GST

activity was apparently lower but not at the significance level

(p < 0.05) for most sites. At the farthest downstream site (Ells),

GST activity was significantly lower at the lowest concentration

tested (<0.04% v/v or < 0.2X). No significant upstream/down-

stream effects were observed but the downstream Ells site

exhibited significant decreases in GST activity. Nevertheless,

GST activity at the surface-water sites was systematically higher

than at OS1 and OS2. GST activity at the surface-water sites was

significantly correlated with cell membrane permeability at the

Northland, downstream OS1, Muskeg and Ells sites. GST

activity was significantly correlated with DNA damage (r ¼�0.33; p ¼ 0.05) at the downstream OS1 site. However at the

next downstream site (Muskeg), GST activity was no longer

negatively correlated with DNA damage (r ¼ 0.31; p < 0.1).

The activity of EROD was measured to detect the effects of

medium and heavy PAHs (Tables 3 and 4). In the OS samples,

EROD activity was significantly induced at OS1 only. The

activity of EROD at OS1 was negatively correlated with DNA

strand breaks (r ¼ �0.36; p ¼ 0.05) and positively so with GST

activity (r ¼ 0.42; p < 0.05). For the surface-water samples,

EROD activity was induced at the Northland and OS1 down-

stream sites, indicating the presence of heavy PAHs. Thus,

activity was significantly and positively correlated with DNA

strand breaks at all sites except the upstream site, where the

correlation was not significantly negative (r¼�0.31; p < 0.1) but

positively correlated with GST activity (r¼ 0.35; p < 0.1); EROD

and GST activities were also correlated at all river locations.

Cytotoxicity and sublethal effects of lake water, groundwater

and OS pond well water

Similar effects were also measured for lake surface water, ground-

water and interceptor well water samples (Table 4). DNA damage

was readily induced in all water samples with no clear difference

among them. The levels of LPO were also influenced by these

samples but less so than the other biomarkers. LPO was readily

induced by the groundwater samples drawn fromMildred Lake.

GST activity was not significantly affected in the surface-water

samples. However, a positive correlation between GST and

DNA strand breaks was apparent in all water samples, the

strongest being observed in the sample fromGregoire Lake; GST

activity was also significantly correlated with LPO at this site.

EROD activity was significantly induced at 0.1 and 0.5X

extract concentrations in the surface water and groundwater

samples, respectively. We found only a passing increase in

EROD activity at 0.1X for OS2 interceptor well water samples.

EROD activity was significantly and positively correlated with

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GST activity at the surface-water site, with DNA damage in the

groundwater samples and LPO in the OS1 interceptor water

samples.

Site discrimination and properties by multivariate analysis

In an attempt to provide an overall picture of the observed

properties of the surface-, groundwater and tailings water

extracts in an OS-rich area supporting intensive industrial

extraction activity, discriminant function and decision tree

analyses were performed on the relative PAH distribution (light,

medium and heavy) and on the five cellular toxicity endpoints.

First, a general correlation analysis revealed the following trends

(all sites combined). Cell viability was positively correlated with

DNA strand breaks (r ¼ 0.3; p < 0.001) and GST activity (r ¼0.34; p < 0.001). Cell viability was negatively correlated with

medium (r ¼ �0.33; p < 0.001) and heavy (r ¼ �0.19; p < 0.05)

PAHs. DNA strand breaks were significantly correlated with

LPO (r ¼ 0.27; p < 0.001) and GST activity (r ¼ 0.53; p < 0.001)

but not with the PAHs found in the extracts. LPO levels were

correlated with GST (r ¼ 0.23; p < 0.01), light PAHs (r ¼ �0.37;

p < 0.001), medium PAHs (r¼�0.22; p < 0.01) and heavy PAHs

(r ¼ 0.15; p ¼ 0.07 marginal). GST activity was significantly

correlated with medium PAHs (r ¼ �0.26; p < 0.01) and heavy

PAHs (r¼�0.47; p < 0.001). EROD activity was correlated with

medium PAHs (r ¼ 0.2; p < 0.05). The levels of light PAHs

followed the levels of medium PAHs (r¼ 0.65; p < 0.001) and the

levels of medium PAHs, in turn, followed the levels of heavy

PAHs (r¼ 0.28; p¼ 0.001). A discriminant function analysis was

used to examine the capacity of the above noted markers to

classify sites and determine which sites show similarities (Fig. 1).

The analysis revealed that most sites were correctly classified

with $75% efficiency. Only two sites were less well classified

using this procedure. Both the Northland and OS1 interceptor

Fig. 1 Discriminant function analysis of the cytotoxic responses and PAH

differences or similarities between the water sample sites based on the biomar

performance of each site is shown in parentheses. The dotted ellipse represen

J. Environ. Monit.

sites were correctly classified only 50% of the time. The North-

land site was sometimes mistaken for the upstream site, while the

OS1 interceptor site was misclassified evenly between Gregoire

Lake surface water and Mildred Lake groundwater. The most

important biomarkers on the X-axis were the heavy PAHs,

medium PAHs, light PAHs, EROD and GST activity. On the Y-

axis, the principal biomarkers were medium PAHs, EROD

activity, GST activity, DNA damage and light PAHs. The close

proximity between the OS1 tailings and the OS1 downstream

sites was noted and suggested some mutual influence. There is

also close proximity between the OS2 interceptor water and the

Mildred Lake groundwater samples located near OS2, suggesting

the influence of OS2 tailings water. It was noteworthy that the

interceptor waters from OS1 and OS2 were different from the

OS1 and OS2 tailing ponds water.

A decision tree analysis was also performed in an attempt to

highlight the most important water properties by establishing

a set of rules that enable site identification (Fig. 2A and 2B). The

analysis revealed that each site was properly identified by a given

set of rules that mainly involved PAH profiles and, occasionally,

the cellular toxic responses of the extract samples. The light-PAH

profile was the most important characteristic, with a perfor-

mance rate of 100%, followed by medium PAHs, with a perfor-

mance rate of about 92%. The four sublethal cytotoxic effects

measurements had performance criteria of > 52%, suggesting

that the endpoints used were less discriminatory among the water

properties. EROD activity was the most performing biomarker,

with a performance of 62%. The rules underpinning the site

characterization are summarized in Table 5. The control solvent

was characterized mainly by low levels of light, medium, and

heavy PAHs and by the lack of toxic effects. The upstream site in

the Athabasca River was close to the solvent control group but

showed increased GST and EROD activities (above 0.9-fold of

the controls) and contained low amounts of light (<42 mg L�1),

profiles. Discriminant function analysis was performed to highlight the

ker responses with the rainbow trout hepatocyte tests. The classification

ts the 90% confidence interval of the mean discriminant function value.

This journal is ª The Royal Society of Chemistry 2011

Fig. 2 Decision tree analysis of PAH profiles and in vitro toxicity data. Decision tree analysis was provided to generate rules that could discriminate the

sites from each other and determine which effects endpoints are the major contributors for site classification. Decision tree analysis A) and the

contribution of each biomarker or PAH B) is shown.

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medium (<8.6 mg L�1) and heavy (< 10 mg L�1) PAHs. At the

Northland site, a slight increase in medium PAHs was observed

with no changes in light or heavy PAHs or in cytotoxic response.

At the downstream OS1 site, a decrease in GST activity was

observed, with a slight increase in medium (>15 mg L�1) and

heavy (>10 mg L�1) PAHs. At the downstream sites near the

confluence of both the Ells and Muskeg rivers, no changes in the

patterns of PAHs were observed, indicating that the relative

proportion of light, medium and heavy PAHs did not change

significantly, although a slight decrease in GST activity was

observed at the Ells site. In the OS tailings from industries 1 and

2, a characteristic drastic rise in light PAHs (>797 mg L�1) and in

medium PAHs (<73 mg L�1) was found. The surface water from

Gregoire Lake was seemingly close to the solvent control and to

the upstream site, but the light PAHs ranged from 5 to 42 mg L�1.

This journal is ª The Royal Society of Chemistry 2011

The groundwater sample from Mildred Lake near OS2 was

characterized by a PAH distribution of 13 to 76 mg L�1 for

medium PAHs and between 42 and 797 mg L�1 for light PAHs.

This site showed evidence of contamination by light and medium

PAHs that is characteristic of OS tailings. Finally, the interceptor

waters from OS1 and OS2 showed significant amounts of light

PAHs (above 379 mg L�1) that are characteristic of OS tailings

water.

Discussion

The Athabasca River is located in an OS-rich area where the

release of PAHs to the water column occurs. The industrial

extraction process, which is intended to remove heavy weight

petroleum products (crude oil), contributes to raising the levels

J. Environ. Monit.

Table 5 Decision tree rule analysis of sample sites

Sites Decision tree rules Water properties

Solvent control A water sample that shows light PAHs < 5 mg L�1,heavy PAHs < 4 mg L�1 and 4 PAHs < 17 mg L�1 isconsidered equivalent to the solvent control.

The solvent corresponds to the baseline datawhere PAH levels are very low.

Upstream A water sample that shows normal GST activity,EROD activity > 0.9, heavy PAHs < 10 mg L�1,medium PAHs < 9 mg L�1 and light PAHs < 42 mgL�1 is considered an upstream site.

The increased proportion of PAHs compared tothe solvent control is a natural occurrence. ERODand GST activities are also naturally occurring.

Northland sawmill A water sample that shows medium PAHs > 17 mgL�1 but light PAHs < 42 mg L�1 is considered theNorthland site downstream of the city of FortMcMurray.

The site near the Northland sawmill locateddownstream of the city of Fort McMurrayreleases more medium PAHs compared to theupstream site.

OS1 downstream A water sample that shows GST activity < 0.85,EROD activity > 0.9, medium PAHs > 15 mg L�1,heavy PAHs > 10 mg L�1 and light PAHs < 42 mgL�1 is considered downstream OS1.

The proximity of OS1 brings about less GSTactivity than the upstream sites.

Muskeg A water sample showing medium PAHs < 9,heavy PAHs > 10, and light PAHs < 42 mg L�1 ischaracteristic of river water samples fromMuskegRiver.

The release of heavy PAHs is a characteristic ofthe confluence of the Muskeg and Athabascarivers.

Ells A water sample that shows medium PAHs < 17 mgL�1, heavy PAHs > 3.6 mg L�1 and light PAHs <42 mg L�1 is characteristic of the Ells Riverconfluence site

The release of heavy PAHs is a characteristic ofthe confluence of Ells and Athabasca rivers.

OS1 tailings A water sample that shows 17 < medium PAHs <73 mg L�1 and light PAHs > 800 mg L�1 isconsidered a tailing pond.

The industrial extraction process brings aboutmajor releases of light PAHs (2–3 rings) in thetailing water samples.

OS2 tailings A water sample that has medium PAHs >73 mgL�1 and light PAHs > 800 mg L�1 is considereda tailing pond at OS2.

Same as OS1 but contains more medium PAHs(4 rings).

Gregoire Lake surface water A water sample that has between 5 < light PAHs< 42, medium PAHs < 17 and medium PAHs < 4mg L�1 is considered a lake water sample.

Similar to the solvent control and upstream site.

OS1 Interceptor A water sample containing significant amounts oflight PAHs at concentrations > 380 mg/L.

Diluted OS tailings but still contains significantamounts of light PAHs.

OS2 interceptor A water sample containing 42 < light PAHs < 800mg L�1 is considered the Mildred Lake interceptorof OS2.

This interceptor sample shows dilution but stillexhibits elevated light PAH levels.

Mildred Lake groundwater (OS2) A water sample showing 13 < medium PAHs < 76and 42 < light PAHs < 800 mg L�1 isa groundwater sample.

Similar to OS2 interceptor water. Thegroundwater samples show evidence ofcontamination with OS2 because the high levels oflight and medium PAHs.

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and the proportions of dissolved of more polar organic

compounds in the tailing ponds. This was corroborated by the

significant increase in light PAHs in the OS tailing water extracts.

The prevalence of light PAHs in OS tailings agrees with previous

findings using synchronous fluorescence spectrometry, in

contrast to river water, which contains proportionally heavier

PAHs.12 Thus, where there is contamination by OS tailings water

in the surrounding environment; we might also expect a gradual

increase in the proportion and concentration of light PAHs

among other alkali-extractable compounds in the tailings. In this

context, the cumulative toxic effects of a water sample would be

the result of the natural release of heavy PAHs, with the possible

leaching of alkali-extractible organics including the light PAHs

fraction from tailing ponds. The search for specific biomarkers to

discriminate between natural and industrial (OS tailings) inputs

is important, especially for the purpose of relating the toxic

impacts observed in the field with industrial OS extraction

activities. Interestingly the release of the heavy fraction of PAHs

to the OS tailings ponds was not significantly different from the

upstream river water samples, suggesting that the tailings would

not significantly contribute to the ambient levels of heavy PAHs

in our hands. This generalisation needs further validation since

J. Environ. Monit.

the water was collected during one season. In a previous study,

the increased presence of heavy PAHs was reported at sites

downstream of OS extraction operations.8 These authors also

found that snow (entrapping fine dust particles) is a significant

vector for the entry of PAHs to the Athabasca River system. The

release of dust particles into the atmosphere could also be

a pathway of entry.

Clear differences were observed with the concentration and

relative proportion of light PAHs (2 to 3 aromatic rings). The

increased proportion of light PAHs was also found in the Mil-

dred Lake groundwater and Ells River samples, which could

indicate contamination from OS tailings. However, a decision

tree analysis integrating the toxic responses with the levels and

distribution of PAH data revealed inputs from the natural release

of heavier PAHs rather than from OS tailings. Indeed, the

proportion of light PAHs was below 42 mg L�1-slightly elevated,

but still characteristic of river water in this system. The

proportion of light PAHs cannot, on its own, be a definitive

criterion for tracking the contamination by OS tailings; the

actual concentration in the water is also of importance. The

analysis of the surface water from Gregoire Lake led us to the

same conclusion: although the proportion of light PAHs was

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high, the actual levels of 2–3 aromatic ring PAHs (i.e. < 42 mg

L�1) revealed a closer similarity to the Athabasca upstream site.

The Mildred Lake groundwater is close to the OS2 interceptor

well water, suggesting a positive input of OS tailing water to the

groundwater.

It is noteworthy that DNA damage (genotoxicity) was the

effects endpoint that exhibited the strongest responses in the

various water samples. However, this endpoint was the least

discriminatory between either the types of waters or upstream/

downstream trends indicating a more generalized response in this

area. It is well known that PAHs are potent genotoxic if not

carcinogenic compounds, especially those belonging to the

medium and heavy PAH groups, such as benzo(a)pyrene.33 The

genotoxic potential of heavy PAHs has to undergo biotransfor-

mation by cytochrome P4501A (EROD) to yield DNA-reactive

metabolites. This is in keeping with the positive correlation

between DNA strand breaks and EROD activity for the

upstream and downstream river site samples, which contain

proportionally more of the heavy PAHs (Table 3). Interestingly,

no significant correlations between DNA damage and EROD

were found in OS tailings water suggesting that other compounds

than EROD inducers were at play. In a study of the polar cod

Boreogadus saida exposed for four weeks to crude oil (i.e. con-

taining more of the heavy PAHs), DNA damage was significant

at PAH exposure concentrations < 15 mg L�1 and positively

correlated with the medium-weight (pyrene) and heavy-weight

(benzo[a]pyrene) PAHs.34 The effects of light PAHs such as

naphthalene on metabolic biotransformation, oxidative stress

and DNA damage is not well understood in aquatic organisms at

the present time. In Anguilla anguilla L. eels, both b-naphtho-

flavone (a medium weight PAH) and naphthalene (a light PAH)

were strong inducers of cytochrome P4501A1 activity, with the

latter requiring a longer exposure time to induce enzymatic

activity.35,36 Interestingly, no evidence of DNA damage was

observed in the naphthalene exposure group as determined by

erythrocytic nuclear abnormalities. In a follow-up study of eels,

an early naphthalene-induced genotoxicity was observed at <8 h

post-exposure but the DNA damage returned to control values at

higher exposure times, demonstrating naphthalene-induced

DNA repair capacity in fish.36 Studies of the genotoxicity of

NAs, an important contaminant family found in OS tailings, are

virtually nonexistent in aquatic organisms. NA-rich processed

light oil was shown to produce DNA adducts as determined

by32P post-labelling of nucleotides in treated mouse skin and the

light oil was positive with a bacterial mutagenicity index of 1.3–

4.3.37 Closer examination of the radiograms in the above study

revealed some differences between mouse skin exposed to either

bitumen or the untreated heavy paraffinic distillate aromatic

extract, suggesting that NAs could have adducted to DNA in

mice. A recent study revealed that NAs decreased the SOSDNA

repair activity of Escherichia coli using a microbial genome-wide

live cell reporter array system.38 The decrease in the SOS

response was observed at environmentally realistic concentra-

tions of NAs and after only a few hours of exposure. Although

not related to OS tailings, the water-soluble fraction of gasoline

was genotoxic to the bivalve Corbicula fluminea.39 Significantly

higher levels of micronuclei in hemocytes were detected after only

short-term (6–96 h) exposure to the gasoline water-soluble

fraction.

This journal is ª The Royal Society of Chemistry 2011

It was observed that GST activity was readily inhibited in OS

tailing wastewaters. The explanation for this inhibition remains

elusive at the present time, although characteristic of OS tailings.

GST activity was positively correlated with EROD activity,

suggesting that conjugation of PAH hydroxylated metabolites

was still at play in the treated hepatocytes. However, GST

activity was strongly correlated with DNA damage in the various

water samples, indicating a bioactivation mechanism. The co-

substrate GSH of GST could be depleted by sustained oxidative

stress but this was not apparent since we did not observe

a negative trend with LPO although LPOwas significantly higher

at the downstream sites from the OS extraction industries. There

is evidence that some organometallic compounds are potent

inhibitors of GST activity. Organic complexes of elements from

the IVA family of the periodic table of elements such as germa-

nium, lead and tin were potent inhibitors of hepatic GST activity

in the rat.40 This inhibition was potent enough to block the

biliary excretion of GSH conjugate of sulfobromophthalein.

Metal protoprophyrin complexes were found in bitumen

involving nickel and iron.6 If this is the case in OS tailings, the

presence of organometallic compounds could compromise the

natural elimination processes of many xenobiotics by blocking

GST activity, a major xenobiotic-conjugating enzyme family

involved in the (biliary) excretion of foreign compounds.

In conclusion, water extracts from this OS-rich area were able

to induce cytochrome P4501A and GST, which are involved in

the biotransformation of PAHs. The increased detoxifying

activity did not prevent the manifestation of important DNA

strand breaks and LPO levels in cells. All the types of water

sampled (river, surface water, groundwater, OS tailings) were

able to induce all the observed effects, with the exception of GST

activity, which was significantly inhibited by OS tailings and

some OS interceptor water. Moreover, DNA damage was not

related to cytochrome P4501A1 activity in OS tailings water

samples. A PAH profile analysis by scanning fluorescence

revealed that OS tailings could potentially be a significant source

of light PAH inputs to this river system. A decision tree analysis

revealed that concentrations of light PAHs (>42 mg L�1) on the

order of 87% of total PAHs are a clear sign of OS tailings

contamination. This study provides some insight into the

potential toxic effects of naturally occurring PAHs and the

contribution of the OS industry to the toxic potential of water.

Acknowledgements

The authors thank the following people for sample collection:

Gerald Tetreault, Jim Bennett, Adrienne Bartlett, for field

collections of water samples. The authors acknowledge the

technical help provided by Sarah-Ann Quesnel for performing

the biomarker analyses. The English editing of the manuscript

was done by Patricia Potvin from Environment Canada. This

research was supported by the Aquatic Ecosystem Protection

Research Division of Environment Canada.

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