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Tracing the Biotransformation of PCBs and PBDEs in Common Carp (Cyprinus carpio) Using Compound-specific and Enantiomer-specific Stable Carbon Isotope Analysis Bin Tang, Xiaojun Luo, Yan-Hong Zeng, and Bi Xian Mai Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b05130 • Publication Date (Web): 16 Feb 2017 Downloaded from http://pubs.acs.org on February 16, 2017
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Environmental Science & Technology
Tracing the Biotransformation of PCBs and PBDEs in Common Carp (Cyprinus carpio) Using Compound-specific and Enantiomer-specific Stable Carbon Isotope Analysis Bin Tang †, ‡, Xiao-Jun Luo †,*, Yan-Hong Zeng †, Bi-Xian Mai †
†
State Key Laboratory of Organic Geochemistry and Guangdong Key Laboratory of
Environmental Resources Utilization and Protection, Guangzhou Institute of Geochemistry, Chinese Academy of Sciences, Guangzhou, 510640, P. R. China ‡
University of Chinese Academy of Sciences, Beijing, 100049, P. R. China
* Corresponding author Phone: +86-20-85297622; Fax: 86-20-85290706; E-mail address: luoxiaoj@gig.ac.cn .
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ABSTRACT
2
Compound-specific and enantiomer-specific carbon isotope composition was
3
investigated in terms of biotransformation of polychlorinated biphenyls (PCBs) and
4
polybrominated diphenyl ethers (PBDEs) as well as atropisomers of chiral PCB
5
congeners in fish by exposing common carp (Cyprinus carpio) to certain PCB and
6
PBDE congeners. The calculated carbon isotope enrichment factors (εC) for PCBs 8,
7
18, and 45 were -1.99‰, -1.84‰, and -1.70‰, respectively, providing evidence for
8
the metabolism of these congeners in fish. The stable carbon isotopic compositions of
9
PBDE congeners clearly reflect the debromination of PBDEs in carp. Significant
10
isotopic fractionation was also observed during the debromination process of BDE
11
153 (εC = -0.86‰). Stereoselective elimination of the chiral PCB congeners 45, 91,
12
and 95 was observed, indicating a stereoselective biotransformation process. The
13
similar εC for E1- (-1.63‰) and E2-PCB 45 (-1.74‰) indicated that both atropisomers
14
were metabolized by the same reaction mechanisms and stereoselection did not occur
15
at carbon bond cleavage. However, the εC values of (+)-PCB 91 (-1.5‰) and (-)-PCB
16
95 (-0.77‰) were significantly different from those of (-)-PCB91 and (+) PCB95,
17
respectively. In the latter, no significant isotopic fractionations were observed,
18
indicating that the stereoselective elimination of PCBs 91 and 95 could be caused by a
19
different reaction mechanism in the two atropisomers.
20
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INTRODUCTION
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Polychlorinated biphenyls (PCBs) and polybrominated diphenyl ethers (PBDEs)
23
are two kinds of anthropogenic organic compounds, which are well known for their
24
persistence, bioaccumulation, long-range transport potential, and toxicity1, 2. Because
25
of these characteristics, PCBs and components of the penta-BDE and octa-BDE
26
technical mixtures have been regulated by the Stockholm Convention on persistent
27
organic pollutants (POPs)3, 4.
28
Despite their persistence in the environment, PCBs and PBDEs can undergo
29
biotransformation in wildlife and humans, and generate metabolites that may be more
30
toxic than their parent compounds, resulting in a serious threat to the biota5, 6. To date,
31
several studies have demonstrated the biotransformation of PCBs in fish7-9, although
32
fish are considered inefficient in the biotransformation of PCBs compared to birds and
33
mammals10,
34
(MeSO2-PCBs), two kinds of PCB metabolites, have been detected in fish7,
35
Additionally, biotransformation of PBDEs in fish was reported in previous studies13-15.
36
Fish have not demonstrated much ability to form hydroxylated PBDEs (OH-PBDEs);
37
however, they have reductively debrominated PBDEs both in vivo and in vitro, and
38
species-specific differences in metabolic rates and products were observed14, 15.
11
.
Hydroxylated
PCBs
(OH-PCBs)
and
methylsulfone
PCBs 12
.
39
During recent decades, compound-specific isotope analysis (CSIA) has
40
undergone rapid development, which has led to important applications for the
41
assessment of the origin and degradation of organic compounds in the environment16.
42
The concept of isotope fractionation relies on the observation of shifts in ratios of
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stable isotopes caused by the breakage of chemical bonds during bio-/chemical
44
transformations, which can lead to the enrichment of heavier isotopes in the residual
45
fractions and their depletion in the degradation products16-18. The extent of stable
46
isotope fractionation allows for the qualitative and quantitative assessment of
47
pollutant biodegradation, as well as the elucidation of the reaction mechanism19. The
48
use of CSIA for environmental investigation has been applied to study the
49
biotransformation and trophic dynamics of PCBs and PBDEs in fish in our previous
50
studies17, 20-22.
51
Of the 209 PCB congeners, a group of 19 congeners are axially chiral and form
52
stable atropisomers under ambient conditions 23. The stereoisomeric patterns of chiral
53
PCBs (described as enantiomeric fractions, EFs) is a useful tool for characterizing
54
biochemical processes23 because biological reactivity would lead to a preferential
55
biotransformation of individual atropisomers. Stereoselective biotransformation of
56
chiral PCBs has been observed in fish7, 9, 20. However, little information is available
57
regarding the mechanisms of stereoselective biotransformation. In recent years,
58
enantiomer-specific isotope analysis (ESIA) has become a promising new approach
59
that could provide insight into stereoselective fate and source apportionment of
60
environmental
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biochemical enzymatic reaction mechanisms during biotransformation19. The ESIA
62
approach has been applied in the study of stereoselective biodegradation of
63
α-hexachlorocyclohexane (α-HCH), polar herbicides (phenoxy acids), and galaxolide
64
and phenoxy alkanoic methyl herbicides24. Currently, no study has been conducted to
organic
contaminants24,
and
provide
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investigate the changes in stable isotope signatures of atropisomers of chiral PCB
66
congeners during the biotransformation processes.
67
In the present study, a dose of certain PCB and PBDE congeners was
68
administered to common carp (Cyprinus carpio) via their diet for 28 days. This was
69
followed by a depuration period of 84 days, during which the carp were fed
70
unfortified, non-spiked food. The primary objective of this study was to trace the
71
biotransformation of PCBs and PBDEs in fish using CSIA. We further investigated
72
the changes in atropisomeric composition of chiral PCB congeners and the carbon
73
stable isotope fractionation of individual atropisomers to verify the stereoselective
74
biotransformation and its mechanism using ESIA.
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EXPERIMENTAL SECTION
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Standards and Reagents. Nineteen PCB congeners ( PCBs 8, 18, 20, 31, 44, 45,
77
49, 52, 91, 95, 101, 132, 138, 149, 174, 177, 180, 183, and 187) and five PBDE
78
congeners (BDEs 85, 99, 100, 153, and 154) were obtained from AccuStandard (New
79
Haven, CT, USA). MeSO2-PCB, OH-PCB, OH-PBDE, methoxy PCB (MeO-PCB),
80
and methoxy PBDE (MeO-PBDE) standards (given in details in supporting
81
information, SI) were obtained from Wellington Laboratories (Guelph, ON, Canada).
82
Pesticide grade acetone (Ace), dichloromethane (DCM), and n-hexane (Hex) were
83
purchased from CNW Technologies GmbH (Dusseldorf, Germany). Guaranteed
84
reagent grade concentrated sulfuric acid (H2SO4) and anhydrous sodium sulfate were
85
acquired from Guangzhou Chemical Reagent Factory (Guangzhou City, China).
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Food Preparation. Commercial fish food (protein > 40% and crude fat > 4.5%)
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was obtained from Zhongshan President Enterprise Co., Ltd. (Guangdong, China).
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Approximately 400 – 450 µg of each PCB congener (PCBs 8, 18, 20, 31, 44, 45, 49,
89
52, 91, 95, 101, 132, 138, 149, 174, 177, 180, 183, and 187) and 1.8 mg of each
90
PBDE congener (BDEs 85, 99, 100, 153, and 154) were first dissolved in 10 g of cod
91
liver oil (Peter Moller, Norway), which was then mixed with 140 g of fish food pellets.
92
The initial concentrations of each PCB and PBDE congener in the food were 2.8–3.2
93
µg g-1 and 12.9 µg g-1 dry weight (dw), respectively. Non-spiked food, which was
94
used for the depuration phase and control group, was treated in an identical manner
95
but without addition of PCBs and PBDEs. Food was homogenized by mixing in a
96
shaking incubator (24 h, 20°C), and then air-dried for 24 h and stored in the dark at
97
-20°C in amber stopper-sealed jars throughout their use. Spiked food samples were
98
collected at the beginning and end of the feeding intervals to confirm the associated
99
PCB and PBDE concentrations.
100
Exposure and Sampling. Sixty-four common carp with average initial weights
101
and lengths of 17.1 ± 2.7 g (mean ± SD, similarly hereafter) and 10.3 ± 0.5 cm,
102
respectively, were purchased from a local aquarium market in Guangzhou, China. At
103
the beginning of the experiment, six fish were removed as background samples. The
104
remaining fish (n = 58) were randomly distributed between two rectangular glass
105
aquariums (150 cm × 45 cm × 100 cm). One was designated as control group (n = 18),
106
in which fish were fed non-spiked food throughout the experiment. The other
107
aquarium was the treated group (n = 40). Each tank was filled with filtered 7
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dechlorinated tap water, which was maintained at a temperature of 22 ± 1°C and
109
circulated using a submerged pump at a rate of 2.5 L min-1. The water in each
110
aquarium was gently aerated to maintain oxygen saturation under a 12-h light and
111
dark photoperiod cycle. Fish were acclimated to the non-spiked diet in the laboratory
112
for two weeks prior to exposure, and were fed food at a rate of 1% of their average
113
body weight per day.
114
After 28 days of exposure (uptake period), the fish were fed non-spiked food for
115
84 days (depuration period). Fish were sampled on days 0, 7, 21, and 28 of the uptake
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period, and on days 14, 28, 42, 56, 70, and 84 of the depuration period. On each
117
sampling day, four fish were randomly chosen from the exposed group to determine
118
their fork length and weight. Blood samples were obtained from the dorsal aorta using
119
syringes, transferred into 5-mL Teflon tubes, and centrifuged at 3000 rpm for 30 min
120
to obtain the serum. Then, the fish were dissected and separated into the gill, liver,
121
gonad, gastrointestinal tract (GI, including the stomach and intestines with undigested
122
food removed, spleen, kidney, heart and adipose fat associated with these organs), and
123
carcass (whole fish minus gill, gonad, liver and GI tract). The gill, liver, gonad, and
124
GI of fish sampled on the same day were weighed and respectively pooled to form
125
two composite samples, the sera were pooled into one sample, whereas carcass was
126
combined correspondingly to form two samples prior to extraction for CSIA. All
127
samples were freeze-dried, ground into powder, weighed, and stored at -20°C prior to
128
being analyzed. Similarly, two fish were randomly collected from the control group
129
on each sampling day, and treated in the same way as those in the exposure group. 8
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Sample Preparation and Extraction. The gill, liver, gonad, GI, and one-tenth
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of the carcass were used for PCB, PBDE, and MeSO2-PCB quantification analysis,
132
and the sera were analyzed for PCBs, PBDEs, OH-PCBs, and OH-PBDEs. The
133
remainder of the carcass material sampled on the same day was respectively
134
combined into two samples and used for compound-specific stable carbon isotope
135
analysis.
136
The extraction and quantification analysis procedures used for fish tissues (serum,
137
gill, liver, gonad, GI, and carcass) were similar to those described in previous
138
studies12, 25, with minor modifications, and given in details in the SI.
139
Approximately 10 g dry weight for each sample was used for stable carbon
140
isotope analysis. The method for purification of PCBs and PBDEs in fish for CSIA
141
was based on our previous studies20,
142
descriptions of the sample extraction and cleanup procedures are given in the SI. No
143
significant isotope fractionation of the target compounds was observed during the
144
purification process20, 21.
21
, with minor modifications. Detailed
145
Instrumental Analysis. Determination of PCBs was performed using an Agilent
146
7890A gas chromatography (GC) coupled with a 5975C mass selective detector (MS)
147
in an electron impact (EI) ion source mode. A DB-5 MS column (60 m × 0.25-mm i.d.
148
× 0.25-µm film thickness) was used for PCB separation. A Chirasil-Dex column (25
149
m × 0.25-mm i.d. × 0.25-µm film thickness) was used to separate PCB 91, 95, 132,
150
149, and 174 atropisomers. A BGB-172 column (30 m × 0.25-mm i.d. × 0.18-µm film
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thickness) was used to separate PCB 183 atropisomers, and a Cyclosil-B column (30
152
m × 0.25-mm i.d. × 0.25-µm film thickness) was used to separate PCB 45
153
atropisomers. Chiral PCB compositions were expressed as EFs, which were defined
154
as follows: EF =
155
A A+B
156
where A and B represent the areas of the (+)- and the (-)-atropisomer peaks in the
157
stereoselective chromatograph column, respectively, for PCBs 91, 95, 132, 149, and
158
for the first-eluting (E1) and second-eluting (E2) atropisomers, respectively, for PCBs
159
45, 174 and 183. The (-)-atropisomer elutes first for PCBs 95, 132, 136, and 149, and
160
the (+)-atropisomer elutes first for PCB 9126, 27, whereas the eluting orders for PCBs
161
45, 174 and 183 atropisomers were unknown. The oven temperature programs are
162
given in detail in the SI.
163
PBDEs, MeSO2-PCBs, OH-PCBs, and OH-PBDEs (OH-PCBs and OH-PBDEs
164
were derivatized to their methoxy analogues by diazomthane before instrumental
165
analysis, details in SI) were analyzed using an Agilent 6890N GC coupled with a
166
5975B MS in the electron capture negative ionization (ECNI) mode, and were
167
separated with a DB-XLB capillary column (30 m × 0.25-mm i.d. × 0.25-µm film
168
thickness). Details of the GC conditions and oven temperature programs are given in
169
the SI. Quantification was based on internal calibration curves constructed from
170
standard solutions at eight concentrations.
171
GC-C-Isotope Ratio Mass Spectrometry (IRMS) Analysis. The purities of the
172
extracts used for CSIA were first checked using an Agilent 7890A GC-5975B MS 10
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system with the EI ion source in the full scan mode. A secondary Aroclor mixture
174
1242/1248/1254/1260 (1:1:1:1) and penta-BDE mixture (DE-71) were used as
175
standards for the qualitative analysis of PCBs and PBDEs, respectively. The
176
individual congeners of the PCBs and PBDEs were identified by comparing the mass
177
spectrum and the retention time of the target compounds with the calibration
178
standards.
179
CSIA measurements of PCBs and PBDEs were performed using a method
180
similar to that used in our previous study22, with minor modifications. The same
181
columns were used for CSIA as those used for quantification analysis. Detailed
182
descriptions of the CSIA procedures are given in the SI. A co-injected standard,
183
2,4,6-trichlorobiphenyl (PCB 30), whose δ13C value was first determined offline with
184
a Flash 2000 EA-Delta V Plus IRMS (Thermo-Fisher Scientific, USA) and was
185
spiked into the extract before conducting the CSIA. The online-measured δ13C value
186
for PCB30 (-29.14‰ to -28.86‰) was close to the offline-measured value (-28.80‰),
187
indicating data reliability.
188
Bioaccumulation Parameters. All concentrations in fish tissues and serum were
189
lipid-based. The bioaccumulation parameters, including assimilation efficiencies (αa),
190
depuration rate constants (kd), half-lives (t1/2), and biomagnification factors (BMFs) of
191
PCB and PBDE congeners were calculated according to equations similar to those
192
described in a previous study8, and given in detail in the SI.
193
Carbon Stable Isotope Calculations. The carbon isotope ratios were reported in 11
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δ notation in parts per thousand (‰) relative to the international carbon isotope
195
standard Vienna Pee Dee Belemnite (V-PDB), according to the following equation
196
(Eqn. (1)):
δ13C =
197
R sample − R standard R standard
×1000
(1),
198
where Rsample and Rstandard represent the 13C/12C ratios of the sample and the V-PDB
199
standard for the carbon isotopic analysis, respectively.
200 201
202
The Rayleigh equation was used to quantify isotope fractionation upon biodegradation:
R C ln t = ( α-1) ln t R0 C0
(2), 13
C/12C) of the target
203
where Rt and R0 are the isotopic compositions (ratio of
204
compounds at time t and time 0 of the depuration period, α is the carbon isotope
205
fractionation factor, and Ct and C0 are the concentrations of the substrate at time t and
206
time 0 of the depuration period, respectively. The carbon isotope enrichment factor ε
207
was calculated according to Eqn. (3):
208
ε = ( α-1) × 1000
(3)
209
Statistical Analysis. All data are presented as means ± standard deviations
210
unless otherwise specified. Statistical analyses were performed using the SPSS 21
211
software for Windows (SPSS). The level of significance was set at p = 0.05
212
throughout the study. The statistical differences in the EFs of chiral PCBs and δ13C
213
values between different groups of samples were determined by one-way analysis of
214
variance (ANOVA) with Tukey’s post-hoc test.
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RESULTS AND DISCUSSIONS
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Background Levels and Quality Control. No natural mortality was observed
217
throughout the experiment period. BDEs 47, 100, 153, and 154 were detected in
218
background common carp at concentrations ranging from 5.8 ± 1.2 to 11.0 ± 2.3 ng
219
g-1 lipid weight (lw), 7.2 ± 1.5 to 23.2 ± 1.6 ng g-1 lw, 1.2 ± 0.52 to 15.4 ± 0.08 ng g-1
220
lw, and 2.9 ± 0.75 to 9.5 ± 0.97 ng g-1 lw in fish tissues, respectively. All 19 PCB
221
congeners
222
concentrations of PCB congeners in the carcass, liver, gonad, GI, gills, and serum
223
ranged from 1.3 ± 0.68 to 14.3 ± 2.8 ng g-1 lw, 2.3 ± 1.3 to 23.4 ± 4.5 ng g-1 lw, 1.4 ±
224
0.72 to 12.1 ± 2.4 ng g-1 lw, 1.3 ± 0.70 to 11.6 ± 2.3 ng g-1 lw, 1.2 ± 0.62 to 5.8 ± 1.1
225
ng g-1 lw, and 2.5 ± 0.78 to 8.1 ± 0.05 ng g-1 lw, respectively. No OH-PCBs,
226
OH-PBDEs, or MeSO2-PCBs were detected in the background or control samples at
227
the beginning or end of the experiment. The PCB and PBDE levels in the background
228
samples were two to three orders of magnitude lower than those in the exposed fish.
229
And spiking test using these background fish confirmed that the influence of
230
background PCBs and PBDEs on the isotopic composition of PCBs and PBDEs in the
231
exposure group was negligible (Table S1). The concentrations of PCB and PBDE
232
congeners in the spiked food pellet homogenate were ranged from 2.9 ± 0.06 to 3.3 ±
233
0.13 µg g-1 dw and 12.8 ± 0.20 to 12.9 ± 0.29 µg g-1 dw, respectively, which were
234
very close to the nominal concentrations. More details regarding quality assurance
235
and control are given in the SI.
236
were
detected
in
background
common
carp.
The
background
Bioaccumulation Parameters of PCB and PBDE. The uptake curves were 13
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similar for most PCB congeners. All congeners reached their highest concentrations at
238
the end of the uptake period (28 d) (Figure S1), and no steady state was observed
239
during the 28-d exposure. The elimination of PCB congeners in all tissues followed
240
first order depuration kinetics. The uptake and depuration kinetics constants,
241
including assimilation efficiencies (αa), depuration rate constants (kd), half-lives (t1/2),
242
and biomagnification factors (BMFs), of PCB congeners in the carcass, liver, gonad,
243
GI, gill, and serum were calculated (Table S2).
244
The calculated bioaccumulation parameters of PCBs in the present study were
245
quite similar to those reported by Fisk et al.8 in juvenile rainbow trout (Oncorhynchus
246
mykiss). Assimilation efficiencies of PCB congeners in liver (63–86%) and GI
247
(43–57%) were much higher than those in the carcass (28–41%), gonad (26–33%),
248
gill (15–23%), and serum (21–36%). The liver exhibited the highest assimilation
249
efficiencies (Table S2), which could be related to the fact that liver is the first organ in
250
which contaminants deposit after absorption from the GI28. The depuration rates were
251
fastest in the liver and slowest in the carcass. The rapid elimination rate in the liver
252
resulted because it is a rich blood-perfused organ and the main organ for the
253
metabolism of xenobiotic chemicals. The poor blood-perfusion in muscle could be
254
partly responsible for the low-elimination rate of chemicals in the carcass. The
255
depuration rates of PCBs 8, 18, and 45 were 2 or 3 times that of other PCB congeners
256
and the BMF of these three congeners was less than 1 (for other congeners >1). This
257
result indicated that there was another elimination pathway, such as metabolism, for
258
these chemicals (discussed below, this section). 14
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Exclusive of the five spiked PBDE congeners, five metabolic debromination
260
congeners (BDEs 28, 47, 49, 66, and 101) were also detected in fish tissues (Figure
261
S2). Previous studies have demonstrated that BDE 85, BDE 99, and BDE 153 can
262
debrominate to lower brominated congeners, such as BDE 101, BDE 47, BDE 49, and
263
BDE 6614, 15, and BDE 66 can be further debrominated to BDE 2829, whereas BDE
264
100 and BDE 154 were resistant to metabolic debromination in the common carp14, 15.
265
Our results further supported this finding. Although the five spiked PBDE congeners
266
had similar concentrations in the spiked food, the levels of BDE 100 and BDE 154
267
were more than two times those of BDE 153 and were 2 to 3 orders of magnitude
268
higher than those of BDE 85 and BDE 99. Conversely, BDE 47, the potential
269
debrominated product of BDEs 85, 99, and 15315, had the highest concentration in
270
carp tissues. After a 14-d depuration, BDE 85 could not be detected and BDE 99 was
271
at a very low concentration (< 32 ng g-1 lw) in carp.
272
The assimilation efficiencies and BMFs for BDEs 100, 153, and 154, and the
273
depuration rates and half-lives for all PBDE congeners, except for BDEs 85 and 99,
274
were calculated for all target tissues. The absence of BDEs 85 and 99 was caused by
275
their rapid metabolism in carp and they were generally lower than the detection limits
276
during the depuration phase. BDE 100 and BDE 154 showed similar assimilation
277
efficiencies and depuration rates in fish tissues. Thus, it was reasonable to assume that
278
all five BDE congeners would exhibit similar bioaccumulation behavior if no
279
debromination occurred for BDEs 85, 99 and 153. A ratio of the concentration of
280
(BDE 85 + BDE 99 + BDE 153 + all debrominated congeners) to that of (BDE 100 + 15
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BDE 154) would be predicted to be 1.5 based on the similar concentrations of the five
282
congeners in food. Indeed, the ratio was 1.5 after 7 d exposure. However, the ratio
283
decreased from 1.5 after day 7 of exposure to 0.9 at the end of experiment. Two
284
possible explanations can be proposed for the decrease. Firstly, hydroxylation of the
285
lower brominated BDE congeners could be possible because several OH-BDE
286
chemicals (details in Figure S3) were detected in the fish. Secondly, the excretive rate
287
of BDE 28, 47, and 66 may be faster than that of BDE 154 and BDE 100. The
288
assimilation efficiencies and BMFs of BDEs 100 and 154 were significantly higher,
289
whereas the depuration rates were significantly lower than those of BDE 153 (Table
290
S3), which could attributed to the metabolism of BDE 153. The depuration rates and
291
half-lives of the metabolic debromination congeners (i.e., BDEs 28, 47, 49, 66, and
292
101) were also calculated, but these values cannot be interpreted as the actual
293
depuration rates because congeners are formed continuously during the entire
294
experiment.
295
The half-lives (t1/2) and BMFs of the exposed PCB and PBDE congeners in fish
296
carcasses are plotted versus log Kow in Figure 1. The t1/2 and BMFs increased with
297
increasing Kow at log Kow < 7 and then decreased with increasing Kow. This trend was
298
also observed and explained in previous studies, in which rainbow trout was exposed
299
to organic compounds, including PCBs7, 8. However, the half-lives and BMFs of PCBs
300
18, 45, 91 and 95, and BDE 153 were generally lower than their expected values
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according to the log Kows in fish, indicating a metabolic process occurred for these
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chemicals7. 16
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Biotransformation
and
Compound-specific
Stable
Carbon
Isotope
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Signatures of PCB and PBDE. The bioaccumulation parameters provided clues to
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the possible biotransformation of chemicals. To elucidate the metabolism of PCBs and
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PBDEs in carp, compound-specific stable carbon isotope signatures of PCB and
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PBDE were measured. All 19 PCB congeners in the fish samples were available for
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isotope analysis, except PCBs 8 and 18 at the last two sampling points, and PCB 45 at
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the last sampling point (Figure S4). No significant isotopic fractionation was observed
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throughout the experiment for PCBs 20, 28, 44, 49, 52, 91, 95, 101, 132, 138, 149,
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174, 177, 180, 183, and 187 (F8, 53 = 0.608–1.959 , p = 0.074-0.766)20, 24, indicating
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that no biotransformation occurred or the biotransformation occurred with no
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detectable isotope fractionation.
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The stable carbon isotope compositions of PCBs 8, 18, and 45 were the same as
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those in the spiked food during the exposure period (Figure 2). However, a heavy
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isotope enrichment trend with depuration time was observed for these three congeners
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(Figure 2). The carbon isotope composition of PCBs 8, 18, and 45 increased from
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-29.5 ± 0.2‰, -32.4 ± 0.1‰, and 30.1 ± 0.2‰ at the beginning of depuration to -25.0
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± 0.1‰, -26.5 ± 0.2‰, and -24.43 ± 0.2‰ at the end of the experiment, respectively.
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These results affirmed that biotransformation occurred in the fish22,
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neither OH-PCBs nor MeSO2-PCBs were detected in the sera or fish tissues of the
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exposed carp throughout the experiment. This likely occurred because other kinds of
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metabolites or conjugates were generated, but they could not be detected by the
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methods used the present studies20, 31. 17
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. However,
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Stable carbon isotope ratios were precisely measured for four PBDE congeners
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(BDEs 47, 100, 153, and 154) (Figure 3). No obvious isotope fractionation was
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observed for BDEs 100 (F8, 53 = 1.75, p = 0.113) and 154 (F8, 53= 1.185, p = 0.329)
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(Figure 3), indicating that these two congeners were accumulated by common carp
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with or without undetectable biotransformation, which is in accordance with the
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results of laboratory exposure experiments14,
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increased from -29.4 ± 0.3‰ in the spiked food to -26.1 ± 0.1‰ in the carp at the end
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of the experiment (Figure 3). BDE 153 could undergo metabolic debromination to
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form BDE 47 and BDE 101 at similar rates14. The observed heavy isotope enrichment
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indicated isotope fractionation during the metabolic debromination of BDE 153 in
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common carp. The stable carbon isotopic fractionation of BDE 99 in tiger barbs
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(Puntigrus tetrazona) was similarly observed17. A previous study demonstrated that
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reductive dechlorination of PCBs during the microbial degradation process did not
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produce stable carbon isotopic fractionation32. No remarkable isotopic fractionations
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for long chain n-alkanes or polycyclic aromatic hydrocarbons were reported during
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microbial degradation33. It remains unknown whether stable carbon isotopic
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fractionation occurs during microbial degradation of PBDEs. The isotopic signature
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might allow for the differentiation of debromination by fish from debromination by
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microbial degradation in future studies.
21
. The δ13C values of BDE 153
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Of the five metabolic debromination congeners, only BDE 47 occurred in a
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sufficient amount for CSIA. The isotope ratio for BDE 47 ranged from -28.9 ± 0.3‰
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to -28.1 ± 0.1‰ (Figure 3). This was consistent with the hypothesis that BDE 47 18
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primarily originates from BDE 85, BDE 99, and BDE 153. Using the sample from 7 d
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exposure as an example, the concentrations of BDE 154 and BDE 100 in carp
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carcasses were 3,700 ng g-1 and 3,780 ng g-1 lw, respectively. The concentrations of
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BDEs 85, 99, and 153 would approximately equal to 3,700 ng g-1 lw if no
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debromination occurred. However, the concentration of BDEs 85, 99, and 153 were
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only 5 ng g-1, 50 ng g-1, and 1730 ng g-1, respectively. The concentration of BDE 101
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was 660 ng g-1 lw. As a rough estimate, we calculated that 42.5%, 42.5%, and 15% of
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BDE 47 in the carcasses was derived respectively from BDE 85, BDE 99, and BDE
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153. The calculated δ13C value was -28.49‰, which was close to the measured value
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(-28.33‰). In most cases, the measured δ13C value of BDE 47 was slightly higher
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than the calculated δ13C value, which occurred because a small fraction of BDE 47
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had been metabolized to form OH-PBDE. In fact, hydroxylated BDE 47 was
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frequently detected (Figure S3) in the sera of fish.
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The isotopic enrichment factors (εC) calculated according to the Rayleigh
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equation were -1.99‰, -1.84‰, -1.70‰, and -0.86‰ for PCBs 8, 18, and 45, and
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BDE 153, respectively (Figure 4). To the best of our knowledge, the only reported
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carbon enrichment factor for hydrodebromination of PBDEs was determined to be
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-2.11‰ during UV irradiation of BDE 4734, whereas no carbon isotope enrichment
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factors were available for the metabolic debromination of PBDEs and metabolism of
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PCBs in fish prior to this study.
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Enantiomer-specific Stable Carbon Isotope Analysis. Of the seven chiral PCB
19
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congeners (PCBs 45, 91, 95, 132, 149, 174, and 183), PCBs 132, 149, 174, and 183
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were found to be racemic (EF = 0.5) throughout the course of the experiment (Figure
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5), which implies that either no metabolism occurred or their biotransformation was
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achiral7, 20. PCBs 45, 91, and 95, on the other hand, were racemic during the uptake
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phase but non-racemic during the depuration period (Figure 5). No significant
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differences were found for the EF values of each chiral PCB congeners among
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different tissues (F5, 116 = 0.02–1.504, p = 0.149–0.99). The relative abundances of the
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E1 - atropisomer of PCB 45, (-)-atropisomer of PCB 91, and (+)-atropisomer of PCB
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95, increased with time, indicating the preferential metabolism of E2-PCB 45,
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(+)-PCB 91, and (-)-PCB 95 in carp (Figure 5), which was in agreement with our
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previous study20. Buckman et al.7 observed that (+)-PCB 91 (E1-PCB 91) and
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(+)-PCB 136 were preferentially biotransformed by rainbow trout, whereas PCB95
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was racemic. This was slightly different from our results. Species-specific metabolism
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of chiral PCBs has been suggested as a potential reason for this difference 20.
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Carbon isotope compositions were obtained for each atropisomer of PCBs 45, 91,
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95, 132, and 149 (Figure 6), whereas they were absent for atropisomers of PCBs 174
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and 183 because of their relatively poorly resolved chromatograph on chiral columns.
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The initial δ13C values were the same for the pair of atropisomers of each chiral PCB
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congener (Figure 6). For the congeners with no EF changes (PCBs 132 and 149),
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isotopic fractionation was also not observed in either atropisomer, which implied no
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biotransformation occurred for these congeners in common carp.
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Regarding the congeners with EF changes, different situations were exhibited for
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different congeners. A shift in carbon isotope composition of E1- and E2-PCB 45
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toward a more positive δ13C with similar extent (from -29.9 ± 0.3‰ to -24.4 ± 0.2‰
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for E1-PCB 45 and from -30.0 ± 0.3‰ to -23.4 ± 0.4‰ for E2-PCB 45) was observed
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during the depuration period (Figure 6). This indicated that both atropisomers were
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involved in metabolic processes, but the E2-atropisomer was metabolized faster than
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the E1-atropisomer because the EF values increased with depuration time. The εC of
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E1-PCB 45 (-1.63‰) was similar to that of E2-PCB 45 (-1.74‰) (Figure S6),
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indicating both atropisomers were metabolized by similar reaction mechanisms.
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However, because E2 was biotransformed faster than E1, other steps, such as
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substrate uptake into the cell and binding of the substrate to enzyme, than isotope
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sensitive carbon bond cleavage might be rate limiting for the metabolism of PCB 45,
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and those steps were stereoselective.
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A significant heavy isotope enrichment was observed for (+)-PCB 91 (from
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-31.65 ± 0.03‰ to -30.36 ± 0.17‰) during the depuration period (F8, 53 = 27.867, p