Analytical and toxicity characterization of halo-hydroxyl-benzoquinones as stable halobenzoquinone disinfection byproducts in treated water

Article abstract of DOI:10.1021/ac5007238

Exposure to chlorination disinfection byproducts (DBPs) is potentially associated with an increased risk of bladder cancer. Four halobenzoquinones (HBQs) have been detected in treated drinking water and have shown potency in producing reactive oxygen species and inducing damage to cellular DNA and proteins. These HBQs are unstable in drinking water. The fate and behavior of these HBQs in drinking water distribution systems is unclear. Here we report the high-resolution mass spectrometry identification of the transformation products of HBQs as halo-hydroxyl-benzoquinones (OH-HBQs) in water under realistic conditions. To further examine the kinetics of transformation, we developed a solid-phase extraction with ultrahigh-performance liquid chromatography tandem mass spectrometry (SPE-UHPLC-MS/MS) method to determine both the HBQs and OH-HBQs. The method provides reproducible retention times (SD < 0.05 min), limits of detection (LODs) at subnanogram per liter levels, and recoveries of 68%-96%. Using this method, we confirmed that decrease of HBQs correlated with increase of OH-HBQs in both the laboratory experiments and several distribution systems, supporting that OH-HBQs were more stable forms of HBQ DBPs. To understand the toxicological relevance of the OH-HBQs, we studied the in vitro toxicity with CHO-K1 cells and determined the IC₅₀ of HBQs and OH-HBQs ranging from 15.9 to 72.9 μM. While HBQs are 2-fold more toxic than OH-HBQs, both HBQs and OH-HBQs are substantially more toxic than the regulated DBPs.

Full text of DOI:10.1021/ac5007238

Article
pubs.acs.org/ac
Analytical and Toxicity Characterization of Halo-hydroxyl-
benzoquinones as Stable Halobenzoquinone Disinfection
Byproducts in Treated Water
Wei Wang, Yichao Qian, Jinhua Li, Birget Moe, Rongfu Huang, Hongquan Zhang, Steve E. Hrudey,
and Xing-Fang Li*
Division of Analytical and Environmental Toxicology, Department of Laboratory Medicine and Pathology, University of Alberta,
Edmonton, Alberta T6G 2G3, Canada
S
* Supporting Information
ABSTRACT: Exposure to chlorination disinfection byproducts (DBPs) is
potentially associated with an increased risk of bladder cancer. Four halobenzoqui-
nones (HBQs) have been detected in treated drinking water and have shown potency
in producing reactive oxygen species and inducing damage to cellular DNA and
proteins. These HBQs are unstable in drinking water. The fate and behavior of these
HBQs in drinking water distribution systems is unclear. Here we report the high-
resolution mass spectrometry identification of the transformation products of HBQs
as halo-hydroxyl-benzoquinones (OH-HBQs) in water under realistic conditions. To
further examine the kinetics of transformation, we developed a solid-phase extraction
with ultrahigh-performance liquid chromatography tandem mass spectrometry
(SPE−UHPLC−MS/MS) method to determine both the HBQs and OH-HBQs.
The method provides reproducible retention times (SD < 0.05 min), limits of
detection (LODs) at subnanogram per liter levels, and recoveries of 68%−96%.
Using this method, we confirmed that decrease of HBQs correlated with increase of OH-HBQs in both the laboratory
experiments and several distribution systems, supporting that OH-HBQs were more stable forms of HBQ DBPs. To understand
the toxicological relevance of the OH-HBQs, we studied the in vitro toxicity with CHO-K1 cells and determined the IC50 of
HBQs and OH-HBQs ranging from 15.9 to 72.9 μM. While HBQs are 2-fold more toxic than OH-HBQs, both HBQs and OH-
HBQs are substantially more toxic than the regulated DBPs.
isinfection of drinking water is essential to prevent
waterborne disease.1 However, disinfection byproducts
(WTP) and distribution system (WDS).11,12 The DBP species
and concentrations in the water arriving at customers’ taps may
be different from when the water originally left the WTP.13,14
The transformation process can largely affect human exposure
and health risk of these DBPs.15,16 Our previous studies have
shown that HBQs are not stable at neutral pH or after exposure
to UV irradiation.7,17 However, the transformation processes of
HBQs and their transformation products have not been
identified or shown to exist in WTPs and WDSs.
Benzoquinone derivatives (BQs) can undergo spontaneous
reactions both in the environment and within living
organisms.18−22 BQs in water can undergo redox, photo-
chemical, and nucleophilic reactions to produce products such
as semiquinones, benzene-1,2,4-triols, hydroquinones, and
hydroxyl-quinones.23−28 On the basis of the chemical proper-
ties of BQ, we hypothesize that HBQs can undergo oxidation
reactions to form halo-hydroxyl-benzoquinones (OH-HBQs) in
water and that these OH-HBQs are a stable form of HBQ
DBPs in drinking water. To confirm this hypothesis, we
D
(DBPs) are unintentionally produced from the reactions of
disinfectants (e.g., chlorine) with natural organic matter
(NOM) in water.2 Epidemiological studies have found a
potential association between exposure to chlorination DBPs
and adverse health effects, primarily as an increased risk of
bladder cancer.3 Animal toxicological studies indicate that the
current regulated DBPs may not account for the observed risk
of bladder cancer.4 It is likely that more toxicologically relevant
DBPs may exist in drinking water but have not yet been
identified.5
Halobenzoquinones (HBQs) are a class of DBPs that are of
toxicological relevance and likely to be carcinogenic.5 These
compounds are predicted to be 10 000 times more toxic than
the regulated DBPs such as chloroform based on their lowest
observed adverse effect levels (LOAEL).5 Our research group
discovered some HBQs as DBPs in drinking water and
swimming pool water.6−8 Cytotoxicity studies have shown
that the four commonly occurring HBQs are highly cytotoxic to
T24 bladder cancer cells and can damage DNA and
proteins.9,10
Received: February 3, 2014
Accepted: April 15, 2014
Published: April 15, 2014
When DBPs are formed, some will undergo spontaneous
transformation reactions in the drinking water treatment plant
© 2014 American Chemical Society
4982
dx.doi.org/10.1021/ac5007238 | Anal. Chem. 2014, 86, 4982−4988

Analytical Chemistry
Article
conducted studies (1) to identify the transformation products
of HBQs to elucidate the transformation pathways, (2) to
confirm whether OH-HBQs exist as DBPs in drinking water
samples, and (3) to demonstrate the toxicological relevance of
OH-HBQs.
voltage, −4500 V; gas I, 60 arbitrary units; gas II, 60 arbitrary
units; curtain gas, 25 arbitrary units; source temperature, 450
°C; declustering potential (DP), −90 V; collision energy (CE),
−40 V; accumulation time, 0.25 s; scan range, m/z 100−1000.
The information dependent acquisition (IDA) was utilized to
obtain MS/MS spectra. The MS scan range of IDA was m/z
100−700, and the collision energy spread (CES) was 10 V. The
accurate masses of HBQs were set in the inclusion list to track
the peaks of HBQs at all times.
Multiple-reaction monitoring (MRM) methods were per-
formed using a triple quadrupole ion-trap tandem mass
spectrometer (AB SCIEX QTRAP 5500, Concord, ON,
Canada) for the quantification of the four HBQs and their
transformation products. The optimized MS instrumental
parameters were as follows: ion-spray voltage, −4500 V; source
temperature, 450 °C; gas I, 50 arbitrary units; gas II, 60
arbitrary units; curtain gas, 30 arbitrary units; entrance
potential, −10 V; accumulation time for each ion pair, 100
ms. The MRM ion pairs and the optimized values of DP, CE,
and cell exit potential (CXP) are listed in Supporting
Information Table S3. Analyst and PeakView (AB SCIEX)
software were used for data analysis. The method confirmed the
identity of the peak by matching the relative ratio of two
specific parent−product ion pairs and quantified it by the peak
area of one ion pair of higher intensity.
Sample Collection and Solid-Phase Extraction. Water
samples were collected from defined locations of five WTPs
and WDSs, including source water, water plant effluent, and tap
water in the distribution systems of different water ages
(halfway, maximum distance). Some samples were also
collected from locations that showed high concentrations of
regulated DBPs. Water samples were stored in amber bottles
which were precleaned three times by water and methanol that
are HBQ-free. Formic acid (0.25%, v/v) was added to the
samples immediately after collection to quench chlorine
residual and stabilize HBQs.8,29 The samples were transported
back to our laboratory in coolers with ice packs and analyzed
immediately on arrival. The time between collection and
analysis was within 2 days.
The water samples were extracted for the HBQs and OH-
HBQs using Waters Oasis HLB cartridges (6 mL, 200 mg per
cartridge; Milford, MA). The solid-phase extraction (SPE)
method was improved upon the previous one for the four
HBQs.7 The details of the SPE procedures for HBQs and OH-
HBQs are described in the Supporting Information.
To identify transformation products, we used high-resolution
triple quadrupole time-of-flight (QTOF) mass spectrometry to
obtain the mass spectra, accurate mass measurements, and
tandem mass spectra of the products. To quantitatively examine
the transformation kinetics and determine the products in
laboratory reactions, we developed a solid-phase extraction with
ultrahigh-performance liquid chromatography tandem mass
spectrometry (SPE−UHPLC−MS/MS) method using triple
quadrupole ion-trap (QTRAP) mass spectrometry. We further
confirmed these products in the field samples. Finally, we
evaluated the in vitro toxicity of both the HBQs and the OH-
HBQs to elaborate the toxicological relevance of the trans-
formation products.
MATERIALS AND METHODS
■
Chemicals and Solvents. 2,6-Dibromo-(1,4)-benzoqui-
none (DBBQ) was purchased from Indofine Chemical
Company (Hillsborough, NJ). 3,5-Dichloro-2-methyl-(1,4)-
benzoquinone (DCMBQ) and 2,3,6-trichloro-(1,4)-benzoqui-
none (TriCBQ) were synthesized by Shanghai Acana
Pharmtech (Shanghai, China); 2,6-dichloro-(1,4)-benzoqui-
none (DCBQ) was purchased from Sigma-Aldrich (St. Louis,
MO). Chemical structures and molecular weights of these
HBQs are listed in Supporting Information Table S1. 3-
Hydroxyl-2,6-dichloro-(1,4)-benzoquinone (OH-DCBQ), 5-
hydroxyl-2,3,6-trichloro-(1,4)-benzoquinone (OH-DCMBQ),
5-hydroxyl-2,3,6-trichloro-(1,4)-benzoquinone (OH-TriCBQ),
and 3-hydroxyl-2,6-dibromo-(1,4)-benzoquinone (OH-DBBQ)
were synthesized in our laboratory by dissolving solid DCBQ,
DCMBQ, TriCBQ, and DBBQ in Optima water for 12 h at 4
°C, respectively. The purity and identity of the synthesized
compounds were assessed using UHPLC−MS analysis. Only
one peak was observed in each chromatogram, and isotope
patterns confirmed the peak as OH-HBQ. Water (Optima LC/
MS grade; the grade means that the solvent goes through 0.03
μm filtration, and the purity is confirmed by UHPLC-UV and
HPLC−MS detection), methanol (Optima LC/MS grade), and
hydrochloric acid (ACS grade) were purchased from Fisher
Scientific (Nepean, ON). Formic acid (HPLC grade, 50% in
water) was purchased from Fluka.
Quality Control and Quality Assurance. A travel-blank
sample (500 mL of Optima water, 0.25% FA) was included in
each sampling trip. Two SPE-blank samples (500 mL of
Optima water, 0.25% FA) were extracted along with other
water samples in each batch of SPE. Analysis-blank samples
(500 μL, 20% methanol, 80% water, 0.25% FA) were analyzed
between every five samples. These blank samples were analyzed
to examine whether contamination occurred during sampling,
pretreatment, or analysis. Triplicate extractions and triplicate
runs of each extract were performed for each water sample to
determine the average concentration and standard error.
Recoveries and matrix effects of individual analytes were
determined from the spiked water samples.
Liquid Chromatography−Mass Spectrometry Anal-
ysis. A liquid chromatography system (UHPLC, Agilent 1290
Infinity Quaternary LC series) was applied with a Luna C18(2)
column (100 mm × 2.0 mm i.d., 3 μm; Phenomenex, Torrance,
CA) at room temperature to separate the HBQs and their
transformation products. The mobile phase consisted of solvent
A (0.1% FA in water) and solvent B (0.1% FA in methanol)
with a flow rate of 0.17 mL/min. A gradient program was
performed as follows: linearly increased B from 20% to 90% in
20 min; kept B at 90% for 5 min; changed B to 20% for column
equilibration at 25.1−30 min. The sample injection volume was
20 μL.
Accurate mass measurements and isotopic patterns were
obtained with a quadrupole time-of-flight mass spectrometer
(AB SCIEX TripleTOF 5600, AB SCIEX, Concord, ON,
Canada) to identify transformation products of the four HBQs.
The conditions of the TripleTOF mass spectrometry experi-
ments were as follows: negative ionization mode; ion source
Cell Culture and Cytotoxicity Testing. The CHO-K1
(Chinese hamster ovary, CCL-61, ATCC, Manassas, VA) cell
line was chosen to evaluate the toxicity of HBQs and OH-
HBQs. This cell line is widely used in DBP toxicity studies, so
comparisons can be readily made. The cells were cultured in
4983
dx.doi.org/10.1021/ac5007238 | Anal. Chem. 2014, 86, 4982−4988

Analytical Chemistry
Article
(1:1) DMEM/F12 media (Gibco), 10% fetal bovine serum
(FBS) (Sigma-Aldrich, Oakville, ON, Canada), and 1%
penicillin−streptomycin (Invitrogen, Carlsbad, CA) and
maintained at 37 °C, 5% CO2, 90% humidity. The cytotoxicity
of HBQs and OH-HBQs was examined using the real-time cell
electronic sensing (RT-CES) system (ACEA Biosciences, San
Diego, CA). The RT-CES experimental details are described in
the Supporting Information.
RESULTS AND DISCUSSION
■
Figure 1 shows the typical total ion chromatograms (TIC) of
four HBQs in freshly prepared solution (red), and solutions
stored for 3 h (blue) and 12 h (purple), and the blank (black)
at 4 °C. The samples were solid standards dissolved in pure
water (Optima LC/MS grade). Before analysis, methanol was
added to adjust the ratio of methanol in the sample to 20%. A
new peak was clearly detected in the 3 h old HBQ solution,
suggesting a product of HBQ degradation in water. After 12 h,
HBQs were completely undetectable and only one peak (new)
corresponding to individual HBQs was detected. We carefully
examined the accurate mass (Supporting Information Table S2)
and isotope ratios of the new peaks obtained from IDA analysis
(Figure 2). A search using PeakView matched the new peak in
DCBQ solution with OH-DCBQ. We then used the same
procedures to examine DCMBQ, TriCBQ, and DBBQ. As
shown in Figure 2, the accurate mass and isotope patterns of
the new peaks in the individual solutions of DCMBQ, TriCBQ,
and DBBQ in water correspond to OH-DCMBQ, OH-
TriCBQ, and OH-DBBQ, respectively. The mass accuracy for
the most abundant isotope of the four OH-HBQs was 0.4, 0.8,
0, and 0.2 ppm for the four OH-HBQs, respectively (Figure 2).
Supporting Information Figure S1 shows the extracted ion
chromatograms (XIC) of OH-DCBQ, OH-DCMBQ, OH-
TriCBQ, and OH-DBBQ produced in the DCBQ, DCMBQ,
TriCBQ, and DBBQ solutions, respectively, over 12 h storage
time. The formation of OH-DCBQ, OH-DCMBQ, OH-
TriCBQ, and OH-DBBQ (Supporting Information Figure
S1A−D) increased as the solutions of DCBQ, DCMBQ,
TriCBQ, and DBBQ aged, respectively.
Figure 2 shows that the OH-HBQ compounds form [M −
H]−•, [M]−•, and [M + H]−• as the major ions with negative
electrospray ionization (ESI). [M − H]−• ion was the most
abundant for OH-DCBQ and OH-DBBQ, while [M + H]−•
was the most abundant for OH-DCMBQ and OH-TriCBQ
under the optimized conditions. The possible ionization
pathways are described in Supporting Information Figure S2.
The formation of [M + H]−• of OH-HBQs is similar to the ESI
pathways of the HBQs that was previously reported.29 The [M
+ H]−• ions can be explained by two possible processes: one is
direct addition of two electrons and one proton; the other is via
two steps: OH-HBQ first undergoes transformation to
hydroxyl-halodihydroquinone (OH-HDHQ), which then loses
one proton. OH-HBQs also form [M − H]−• ions, which are
rarely observed from the ionization of HBQs. This finding
could be explained by the ionization of the hydroxyl groups.
There were also minor [M]−• ions observed in the mass
spectra, which may be produced by direct ionization via
addition of an electron.
Figure 1. Total ion chromatograms of blank solution (only solvent,
black) and (A) DCBQ, (B) DCMBQ, (C) TriCBQ, and (D) DBBQ
in freshly prepared (red), 3 h old (blue), and 12 h old (purple)
solution. The mass scan range was m/z 100−1000.
measurements, and the MS instrumental parameters were
optimized. Supporting Information Table S3 describes the
optimized MRM conditions. The baseline separation of the
eight OH-HBQs and HBQs was achieved using a C18 reversed-
phase LC.
We also optimized the UHPLC and ionization conditions.
Formic acid in the mobile phase can stabilize HBQs, while the
addition of weak acid in the mobile phase may suppress the
signal of negative electrospray ionization.30 On the basis of the
signals of HBQs and OH-HBQs (Supporting Information
Having synthesized and confirmed the OH-HBQs, we aimed
to confirm the existence of these compounds in the field
samples. To achieve this, we developed a UHPLC−MS/MS
method to determine both the OH-HBQs and the parent
HBQs. Two pairs of transition ions were used in the MRM
4984
dx.doi.org/10.1021/ac5007238 | Anal. Chem. 2014, 86, 4982−4988

Analytical Chemistry
Article
used for quantification. The UHPLC−MS/MS method was
validated for analysis of the four HBQs and four OH-HBQs in
tap water. The detection limit of the UHPLC−MS/MS method
is 0.01−0.7 ng/mL (Supporting Information Table S4, LOD1).
The concentrations of HBQs in water have been previously
reported to be around several nanograms per liter to several
hundred nanograms per liter.7,8 Therefore, it is necessary to
concentrate the compounds from water samples prior to the
UPHLC−MS/MS analysis. No SPE method for OH-HBQs was
available; thus, we developed one to concentrate these
compounds as well as HBQs. The SPE procedures were
optimized, including conditioning of the cartridge, loading of
the sample, washing/cleaning, and elution of the analytes. The
retention of analytes is dependent on the washing solvent and
elution solvent.31,32 On the basis of our LC separation
(Supporting Information Figure S4), methanol is suitable to
elute the HBQs and OH-HBQs. The optimized elution
condition was 10 mL of methanol (0.25% FA, v/v). The
washing step was optimized to remove the interference
matrixes and retain the desired analytes. The optimized SPE
procedures and recoveries are presented in Supporting
Information Figure S5.
To validate the SPE−UHPLC−MS/MS conditions, we
examined the retention time, limit of detection (LOD), limit
of quantitation (LOQ), recovery, and matrix effect. Supporting
Information Table S4 presents the performance of the method:
repeatable retention time (SD < 0.05 min), subnanogram per
liter LOD (LOD2, 0.020.8 ng/L) and LOQ (0.072.8 ng/
L), and recovery (6896%). Even after SPE, the matrix effect
(7998%) persisted; therefore, the standard addition method
was used for quantification of these compounds in authentic
water samples.
Having established a SPE−UHPLC−MS/MS method for
both HBQs and OH-HBQs, we were able to quantitatively
study the transformation of HBQs to OH-HBQs. Figure 3
presents the time course of HBQs converting to OH-HBQs
over 24 h after fresh preparation of an HBQ solution. As the
concentrations of HBQs decrease, the concentrations of OH-
HBQs increase accordingly. After 12 h, the reaction reached
equilibrium. The mass balance (sum) of HBQs and OH-HBQs
was maintained around 80%−120%. OH-HBQs were stable for
60 h at the initial pH 7 and initial concentration 50 μg/mL
(Supporting Information Figure S6), indicating that OH-HBQs
are much more stable than HBQs in water.
We further investigated the presence of OH-HBQs in the
field water systems where the HBQs were determined. Using
the method for the four HBQs and four OH-HBQs, we
analyzed water samples from five WTPs and WDSs. The water
treatment processes of the five WTPs are summarized in
Supporting Information Table S5. Supporting Information
Figure S7 demonstrates that OH-HBQs are present in the
treated tap water but not in the source water. To further
confirm the presence of OH-HBQs, we investigated the
occurrence frequency and concentrations of OH-HBQs in
these samples as summarized in Table 1. OH-DCBQ was the
most commonly identified OH-HBQs of the four OH-HBQs
tested, which was consistent with DCBQ being the most
frequently detected HBQ. The concentrations of HBQs and
OH-HBQs in each water sample are summarized in Supporting
Information Table S6. The samples containing HBQs were
confirmed to also contain OH-HBQs, suggesting that the
transformation of HBQs to OH-HBQs may occur in the WDSs.
To further examine this, we investigated the distribution system
Figure 2. Isotope patterns of four OH-HBQs produced in the
solutions of HBQs. The red trace is the theoretical value and the black
trace is the measured value.
Figure S3), we used the formic acid concentration in the mobile
phase at 0.1% in the MRM quantification methods.
Supporting Information Figure S4 shows typical UHPLC−
MS/MS (MRM) chromatograms obtained from analysis of the
four HBQs and four OH-HBQs. The identification of the target
compounds was based on the following criteria: retention times
are identical for the two ion transitions of the specific
compound and relative intensity (ratio) of these two ion
transitions detected in the samples is consistent with that in the
standard solutions. When an HBQ was identified, the ion
transition with higher abundance (the first ion pair listed in
Supporting Information Table S3 for each compound) was
4985
dx.doi.org/10.1021/ac5007238 | Anal. Chem. 2014, 86, 4982−4988

Analytical Chemistry
Article
Figure 4. Concentrations of HBQs and OH-HBQs in Plant 1
(sampled on 2013.2.24, 6.24 and 9.25). nd: not detected; the
concentration is lower than the detection limit.
Figure 3. Time course of HBQs to OH-HBQs. Initial HBQ solution
was 50 μg/mL at pH 4.5 and was maintained at 4 °C.
established in our laboratory.33,34 Figure 5 illustrates the
temporal response and IC50 of CHO-K1 cells exposed to
DCBQ (as a representation of the tested eight compounds).
Each well was dosed with the tested compounds when the cell
index (CI) was 1 (after about 20 h of growth). The cells
displayed concentration-dependent response curves. At any
given time point, as the concentration of DCBQ increased, the
normalized CI decreased, demonstrating a concentration-
dependent cytotoxic effect on CHO-K1 cells. Cytotoxic
responses of OH-DCBQ, DCMBQ, OH-DCMBQ, TriCBQ,
OH-TriCBQ, DBBQ, and OH-DBBQ are similar to those in
Figure 5.
of Plant 1. A set of samples were collected from different
locations and analyzed for both HBQs and OH-HBQs. The
results (Figure 4) show that the concentrations of HBQs in
water samples decreased while OH-HBQs increased with the
increasing distance from the WTP. Repeated samplings show
the same trend in Plant 1 (Figure 4). Similar results were found
in Plant 2 WDS where DCBQ was detected (Supporting
Information Figure S8). Both laboratory experiments and the
field study supported that the decrease of HBQs were
correlated with the increase of OH-HBQs.
To assess the health relevance of OH-HBQs as DBPs, we
evaluated the toxicity of OH-HBQs compared with that of
HBQs. We used the CHO-K1 cells with the RT-CES method
(details in the Supporting Information) that has been
On the basis of the temporal cytotoxicity profile, we
calculated the IC50 values for each compound on CHO-K1
cells, as represented by the IC50 histogram for DCBQ in Figure
Table 1. Occurrence Frequency and Concentrations of the Four HBQs and OH-HBQs in Treated Water Collected from Five
a
Water Treatment Plants
compd
frequency
concn (ng/L)
compd
frequency
concn (ng/L)
DCBQ
34/37
11/37
10/37
6/37
nd−20
nd−4
nd−20
nd−10
OH-DCBQ
OH-DCMBQ
OH-TriCBQ
OH-DBBQ
34/37
12/37
6/37
nd−20
nd−7
nd−20
nd−10
DCMBQ
TriCBQ
DBBQ
6/37
a
Note: nd, not detected. The concentration is lower than the detection limit.
4986
dx.doi.org/10.1021/ac5007238 | Anal. Chem. 2014, 86, 4982−4988

Analytical Chemistry
Article
OH-DBBQ, i.e., the same order. This indicated that the basic
HBQ structure likely plays a key role in determining the
toxicity of the corresponding OH-HBQ.
ASSOCIATED CONTENT
* Supporting Information
Additional information as noted in text. This material is
available free of charge via the Internet at http://pubs.acs.org.
■
S
AUTHOR INFORMATION
Corresponding Author
*Phone: 780 492-5094. Fax: 780 492-7800. E-mail: xingfang.
li@ualberta.ca.
■
Notes
The authors declare no competing financial interest.
ACKNOWLEDGMENTS
■
The authors acknowledge funding support from the Natural
Sciences and Engineering Research Council of Canada, the
Canadian Water Network, Alberta Innovates−Energy and
Environment Solutions, and Alberta Health. W. Wang
acknowledges the University of Alberta FS Chia Doctoral
Scholarship, the Provost Doctoral Entrance Award, and the
Alberta Innovates−Technology Futures Graduate Scholarship.
The authors thank Katerina Carastathis for her editing of the
manuscript.
Figure 5. RT-CES profile (A) and IC50 histogram (B) of DCBQ on
CHO-K1 cells.
5(B). The 24, 48, and 72 h IC50 values for all eight compounds
are summarized in Table 2. Generally, the IC50’s HBQs and
Table 2. IC50 Values of the Four HBQs and Four OH-HBQs
on CHO-K1 Cells
a
REFERENCES
■
(1) United States Environmental Protection Agency. Fed. Regist.
2006, 71, 388−493.
IC50 (μM)
(2) Hrudey, S. E. Water Res. 2009, 43, 2057−2092.
(3) Villanueva, C. M.; Cantor, K. P.; Cordier, S.; Jaakkola, J. J. K.;
King, W. D.; Lynch, C. F.; Porru, S.; Kogevinas, M. Epidemiology 2004,
15, 357−367.
compd
DCBQ
24 h
48 h
72 h
27.3 1.0
61.0 3.0
11.4 0.5
20.4 0.6
45.5 2.5
64.4 3.7
19.8 1.5
42.8 6.2
35.5 1.0
69.5 5.6
13.7 0.5
21.5 0.8
63.7 2.1
67.1 4.7
29.2 1.8
51.8 3.4
41.5 1.3
90.6 33.6
15.9 0.9
24.0 1.1
72.9 3.6
69.8 6.8
35.5 0.7
50.4 7.6
OH-DCBQ
DCMBQ
(4) Richardson, S. D.; Plewa, M. J.; Wagner, E. D.; Schoeny, R.;
DeMarini, D. M. Mutat. Res. 2007, 636, 178−242.
(5) Bull, R. J.; Reckhow, D. A.; Li, X.-F.; Humpage, A. R.; Joll, C.;
Hrudey, S. E. Toxicology 2011, 286, 1−19.
OH-DCMBQ
TriCBQ
OH-TriCBQ
DBBQ
(6) Qin, F.; Zhao, Y.; Zhao, Y.; Boyd, J. M.; Zhou, W.; Li, X.-F.
Angew. Chem., Int. Ed. 2010, 49, 790−792.
OH-DBBQ
(7) Zhao, Y.; Qin, F.; Boyd, J. M.; Anichina, J.; Li, X.-F. Anal. Chem.
2010, 82, 4599−4605.
a
Note: IC50 is the concentration of the compound determined from a
regression analysis of the data (Figure 5), at which cell index is
reduced to 50% of the negative control.
(8) Wang, W.; Qian, Y.; Boyd, J. M.; Wu, M.; Hrudey, S. E.; Li, X.-F.
Environ. Sci. Technol. 2013, 47, 3275−3282.
(9) Du, H.; Li, J.; Moe, B.; McGuigan, C. F.; Shen, S.; Li, X.-F.
Environ. Sci. Technol. 2013, 46, 2823−2830.
OH-HBQs are at micromole per liter level. The IC50’s of
trihalomethanes (THMs) on CHO cells (72 h) were reported
at the range of 3.96−11.5 mM,35 and IC50’s of haloacetic acids
(HAAs) were at the range of 8.90 μM to 17.52 mM.36 (The
data are summarized in Supporting Information Table S7.) The
IC50’s of HBQs and OH-HBQs (16−91 μM) are much lower
(more toxic) than most of the regulated DBPs, indicating that
chronic cytotoxicity of HBQs and OH-HBQs are significantly
higher than the regulated DBPs. The IC50’s of the eight
compounds are in the order DCMBQ < DBBQ < OH-
DCMBQ < DCBQ < OH-DBBQ < TriCBQ < OH-DCBQ <
OH-TriCBQ. Comparing HBQ and OH-HBQ in pairs, the
IC50’s are DCBQ < OH-DCBQ, DCMBQ < OH-DCMBQ,
TriCBQ < OH-TriCBQ, and DBBQ < OH-DBBQ, indicating
that the addition of the hydroxyl group to HBQ decreases the
cytotoxicity and that the transformation process is partially a
detoxifying process. The IC50 values of the four HBQs are
DCMBQ < DBBQ < DCBQ < DBBQ, and those of the four
OH-HBQs are OH-DCMBQ < OH-DBBQ < OH-DCBQ <
(10) Anichina, J.; Zhao, Y.; Hrudey, S. E.; Schreiber, A.; Li, X.-F.
Anal. Chem. 2011, 83, 8145−8151.
(11) Glezer, V.; Harris, B.; Tai, N.; Lev, L. O. Water Res. 1999, 33,
1938−1948.
(12) Urbansky, E. T. Chem. Rev. 2001, 101, 3233−3243.
(13) Brisson, I. J.; Levallois, P.; Tremblay, H.; Serodes, J.; DeBlois,
C.; Charrois, J.; Taguchi, V.; Boyd, J. M.; Li, X.-F.; Rodriguez, M. J.
Environ. Monit. Assess. 2013, 185, 7693−7708.
(14) Carter, J. M.; Moran, M. J.; Zogorski, J. S.; Price, C. V. Environ.
Sci. Technol. 2012, 46, 8189−8197.
(15) Farre, M. L.; Perez, S.; Kantiani, L.; Barcelo, D. TrAC, Trends
Anal. Chem. 2008, 27, 991−1007.
(16) Escher, B. I.; Fenner, K. Environ. Sci. Technol. 2011, 45, 3835−
3847.
(17) Qian, Y.; Wang, W.; Boyd, J. M.; Wu, M.; Hrudey, S. E.; Li, X.-F.
Environ. Sci. Technol. 2013, 47, 4426−4433.
(18) Uchimiya, M.; Stone, A. T. Chemosphere 2009, 77, 451−458.
(19) Hubig, S. M.; Kochi, J. K. J. Am. Chem. Soc. 1999, 121, 1688−
1694.
(20) Arumugam, S.; Popik, V. V. J. Org. Chem. 2010, 75, 7338−7346.
4987
dx.doi.org/10.1021/ac5007238 | Anal. Chem. 2014, 86, 4982−4988

Analytical Chemistry
Article
(21) Monks, T. J.; Hanzlik, R. P.; Cohen, G. M.; Ross, D.; Graham,
D. G. Toxicol. Appl. Pharmacol. 1992, 112, 2−16.
(22) Bolton, J. L.; Trush, M. A.; Penning, T. M.; Dryhurst, G.;
Monks, T. J. Chem. Res. Toxicol. 2000, 13, 135−160.
(23) Brunmark, A.; Cadenas, E. Free Radicals Biol. Med. 1989, 7,
435−477.
(24) Gorner, H. J. Phys. Chem. A 2003, 107, 11587−11595.
(25) Kutyrev, A. A.; Moskv, V. V. Russ. Chem. Rev. 1991, 60, 134−
158.
(26) Sarr, D. H.; Kazunga, C.; Charles, M. J.; Pavlovich, J. G.; Aitken,
M. D. Environ. Sci. Technol. 1995, 29, 2735−1740.
(27) Zhu, B.-Z.; Kalyanaraman, B.; Jiang, G.-B. Proc. Natl. Acad. Sci.
U.S.A. 2007, 104, 17575−17578.
(28) Zhu, B.-Z.; Zhu, J.-G.; Mao, L.; Kalyanraman, B.; Shan, G.-Q.
Proc. Natl. Acad. Sci. U.S.A. 2010, 107, 20686−20690.
(29) Huang, R.; Wang, W.; Qian, Y.; Boyd, J. M.; Zhao, Y.; Li, X.-F.
Anal. Chem. 2013, 85, 4520−4529.
(30) Wu, Z.; Gao, W.; Phelps, M. A.; Wu, D.; Miller, D. D.; Dalton, J.
T. Anal. Chem. 2004, 76, 839−847.
(31) Poole, C. F. TrAC, Trends Anal. Chem. 2003, 22, 362−373.
(32) Kruve, A.; Auling, R.; Herodes, K.; Leito, I. Rapid Commun.
Mass Spectrom. 2011, 25, 3252−3258.
(33) Du, H.; Li, J.; Moe, B.; McGuigan, C. F.; Li, X.-F. Anal. Methods
2014, 6, 2053−2058.
(34) Boyd, J. M.; Huang, L.; Xie, L.; Moe, B.; Gabos, S.; Li, X.-F.
Anal. Chim. Acta 2008, 789, 83−90.
(35) Plewa, M. J.; Wagner, E. D. Mammalian Cell Cytotoxicity and
Genotoxicity of Disinfection By-products; Water Research Foundation:
Denver, CO, 2009; pp 107−108.
(36) Plewa, M. J.; Simmons, J. E.; Richardson, S. D.; Wagner, E. D.
Environ. Mol. Mutagen. 2010, 51, 871−878.
4988
dx.doi.org/10.1021/ac5007238 | Anal. Chem. 2014, 86, 4982−4988