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Exposure of frogs and tadpoles to chiral herbicide fenoxaprop-ethyl
Accepted Manuscript Exposure of frogs and tadpoles to chiral herbicide fenoxaprop-ethyl Xu Jing, Guojun Yao, Donghui Liu, Chang Liu, Fang Wang, Peng Wang, Zhiqiang Zhou PII:
S0045-6535(17)31183-9
DOI:
10.1016/j.chemosphere.2017.07.132
Reference:
CHEM 19665
To appear in:
ECSN
Received Date: 25 April 2017 Revised Date:
25 June 2017
Accepted Date: 25 July 2017
Please cite this article as: Jing, X., Yao, G., Liu, D., Liu, C., Wang, F., Wang, P., Zhou, Z., Exposure of frogs and tadpoles to chiral herbicide fenoxaprop-ethyl, Chemosphere (2017), doi: 10.1016/ j.chemosphere.2017.07.132. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
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Graphical abstract
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Exposure of frogs and tadpoles to chiral herbicide
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fenoxaprop-ethyl Xu Jing, Guojun Yao, Donghui Liu, Chang Liu, Fang Wang, Peng Wang, Zhiqiang
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Zhou*
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Beijing Advanced Innovation Center for Food Nutrition and Human Health, Department of Applied
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Chemistry, China Agricultural University, Beijing, 100193, P.R. China.
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*Corresponding author:
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Zhiqiang Zhou, Department of Applied Chemistry, China Agricultural University, No.2
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Yuanmingyuan West Road, Beijing 100193, P.R. China; Tel: +8610-62733547; Fax:
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+8610-62733547; E-mail: [email protected]
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ABSTRACT
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Pesticides have long been considered to a risk factor of amphibian population declines.
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The bioaccumulation and elimination of fenoxaprop-ethyl (FE) in frogs and tadpoles
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were
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6-chloro-2,3-dihydrobenzoxazol-2-one (CDHB) were monitored. The acute toxicity
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and genotoxicity of the enantiomers to tadpoles was also studied. After both oral
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administration and aqueous solution exposure, FE was not found in frogs, while FA
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was formed and accumulated in liver, kidney, brain, eggs, skin, thigh muscle and
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blood with preferential accumulation of R-FA. The presence of FA in frog eggs
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suggested maternal transfer in females and potential impacts to offsprings. The
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elimination of FA in frog tissues was also enantioselective with a preferential
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metabolism of R-FA (kidney) or S-FA (liver, eggs, skin, muscle and whole blood). FE
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and FA were hardly detectable in tadpoles after aqueous solution exposure, while
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CDHB was accumulated and eliminated as first-order kinetics with half-life of 37.1 h.
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Mortality of tadpoles and micronucleus rate in peripheral blood erythrocytes of
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tadpoles were used to evaluate the enantioselective acute toxicity and genotoxicity.
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Only CDHB induced significant acute toxicity to tadpole with 96-h LC50 value of 30.4
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ߤg/mL, and rac-FA, S-FA and CDHB showed genotoxicity.
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the
main
metabolites
fenoxaprop
(FA)
and
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and
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studied
KEYWORDS: Fenoxaprop-ethyl; Frog; Tadpole; Chiral; Toxicity
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1. Introduction
Pesticides have been widely used in agriculture, in which about 30% are chiral
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(Williams, 1996). Although the enantiomers of a chiral pesticide almost have the same
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physical and chemical properties, they behave enantioselectively in ecology with
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different bioactivity, toxicity and environmental behavior. These pesticides pose
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unpredictable enantioselective environmental threats, affecting the food chain and
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entire ecosystems (Ye et al., 2015). It is urgent to understand the toxicity and
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metabolism of enantiomers of pesticides in target and non-target organisms on
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enantiomeric level. There are 19 major categories of herbicides on the market in
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which 14 categories contain chiral structures (Zhou and Liu, 2011). Aryloxyphenoxy
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propionate herbicides are widely used and the weed control is achieved by the active
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R-enantiomer.
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(FE,
Fig.
1) is
a post-emergent
herbicide of the
aryloxyphenoxy propionate group (Zawahir et al., 2009). It inhibits fatty acid
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synthesis in grasses by inhibition of acetyl-CoA carboxylase. The key step of
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degradation is ester hydrolysis to its related acid fenoxaprop (FA, Fig. 1) (Dong et al.,
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2015), which is 100 times more active than FE in inhibiting acetyl-CoA carboxylase
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(Yaacoby et al., 1991). FA may undergo further degradation, forming the metabolite
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6-chloro-2,3-dihydrobenzoxazol-2-one (CDHB, Fig. 1) (Hoagland and Zablotowicz,
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1998). Metabolite CDHB may also come directly from the breakdown of the
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benzoxazolyl-oxyphenyl ether linkage of FE (Lin et al., 2007; Jing et al., 2016). The
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degradation of FE have been widely studied in the environment, such as in soil
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ACCEPTED MANUSCRIPT (Zhang et al., 2010), water (Lin et al., 2007), microorganism (Dong et al., 2015),
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wheat (Singh et al., 2013), rice (Lucini and Pietro Molinari, 2010), oat (Xie et al.,
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1996), barley (Romano et al., 1993), crabgrass (Tal et al., 1993), rat (Moody and
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Ritter, 1992) and rabbit (Zhang et al., 2011). Influencing factors like pH (Zablotowicz
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et al., 2000; Lin et al., 2007), temperature and light, and so on (Xie et al., 1996; Lin et
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al., 2008) have been investigated. Both FE and FA are chiral and exist as two
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enantiomers (Fig. 1). According to the literature, S-FE decreased faster than R-FE in
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soils, and the inversion of S-enantiomer to R-enantiomer occurred (Zhang et al.,
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2010). In rabbits, S-FA decreased faster in plasma, heart, lung, liver, kidney, and bile
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than its antipode with similar inversion (Zhang et al., 2011). The conversion between
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the enantiomers was not found in Scenedesmus obliquus suspension (Zhang et al.,
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2008). To accurately evaluate the impacts of FE on the environment, the
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enantioselective behavior of FE and FA should be taken into account.
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The toxicity of FE and its metabolites to some non-target organisms has been
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studied. The 96-h LC50 values for FE to rainbow trout were 0.57 mg/L (Song et al.,
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2005). The 48-h EC50 values for FE, FA and CDHB to Daphnia magna were 5.29,
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14.6 and 8.4 mg/L, respectively (Lin et al., 2008). The 48-h EC50 values for rac-FA
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and R-FA to freshwater alga Scenedesmus obliquus were 8.42 and 8.03 mg/L,
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respectively (Zhang et al., 2008). Clinical toxicity data was achieved by patients with
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acute exposure to commercial products of FE. The main clinical features were
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epigastric burning sensation, vomiting and consciousness recession (Zawahir et al.,
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2009).
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terrestrial animals, while amphibians attracted little attention. Amphibians may be
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good environmental pollutant bioindicator due to the unique characteristics, such as
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aquatic and terrestrial habitats, complex life cycles, rapid larval growth rates, trophic
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position, poikilothermy and permeable skin (Burkhart et al., 2000; Fontenot et al.,
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2000). Worldwide amphibian decline has triggered massive research efforts to explore
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biotic and abiotic factors. Amphibians are sensitive to environmental chemicals,
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which have been implicated as factors contributing to large-scale losses of amphibian
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populations (Hayes et al., 2006). The adverse effects of pesticides on amphibian
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growth, development, reproduction and behavior have long been a problem, and the
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mixture may enhance the effects (Renner, 2003). It has been widely expected that the
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impacts of xenobiotics depend on their uptake, tissue distribution and metabolism. In
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recent years, the bioavailability and tissue distribution of heavy metals, PAHs, PCBs
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and PBDEs in amphibian was examined (Vogiatzis and Loumbourdis, 1997; Fontenot
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et al., 2000; Stabenau et al., 2006; Wu et al., 2009; Li et al., 2014). The
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bioaccumulation, transformation and elimination information about pesticides is
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relatively limited (Tilak et al., 2003; Reynaud et al., 2012). Besides, breeding and
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larval development of amphibians occur in spring and summer when there is frequent
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application of pesticides (Mann et al., 2009). Corresponding environmental behaviors
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and toxicity studies are vital to predict chemical fate in organisms, develop
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environmental quality criteria and assess the ecological risks. Further effort is needed
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to study the bioaccumulation, elimination and toxicity of pesticides in amphibians.
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behavior and toxicity studies of FE have not been conducted in amphibians, the
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present study was conducted to evaluate the bioaccumulation, elimination, acute
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toxicity and genotoxicity of FE and its metabolites FA and CDHB in frogs and
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tadpoles. Frogs were administered by a single oral dose of 1.0 ߤg/g of FE or exposed
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to 0.5 ߤg/mL aqueous solution of FE. Tadpoles were exposed to 0.5 ߤg/mL aqueous
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solution of FE. The concentrations of FE, FA and CDHB in frogs (liver, kidney, brain,
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eggs, skin, thigh muscle and whole blood) and tadpoles were determined on an
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enantiomeric level by high-performance liquid chromatography-tandem mass
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spectrometry (HPLC-MS/MS). Individual enantiomers of FE and FA were prepared to
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assess the acute toxicity and genotoxicity to tadpoles. The goal is to expand on the
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limited data on the bioaccumulation, elimination and toxicity of chiral pesticides in
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amphibians.
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2. Materials and methods
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2.1Chemicals
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Racemic FE (98.0%) was got from Institute for Control of Agrichemicals,
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Ministry of Agriculture of China (Beijing, China). The synthesis of racemic FA and
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the single enantiomer of FE and FA (98.0%) were conducted according to literatures
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(Amabilino et al., 1998; Kato et al., 2003; Moon et al., 2007). CDHB (98.0%) was
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obtained from TCI Development Company Limited (Shanghai, China). Dimethyl
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sulfoxide and corn oil was obtained from Sigma-Aldrich (St. Louis, MO, USA).
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Methanol and acetonitrile were obtained from Fisher Scientific (Fair Lawn, NJ, USA).
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Water was purified by a Milli-Q system.
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2.2 Animals
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Adult female bullfrogs (Rana catesbeiana) and premetamorphic tadpoles (Rana
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catesbeiana) were obtained from local aquafarm. The average weight of frogs and
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tadpoles was 500 ± 25 and 0.6 ± 0.03 g, respectively. The frogs and tadpoles were
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respectively placed in glass aquariums and glass beakers for 1 week at 25 °C with a
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photoperiod of 12 h: 12 h, for the purpose of acclimatization.
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2.3 Exposure to FE
Frogs were exposed individually in glass aquariums (60 cm × 45 cm × 30 cm)
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containing 5 L of water. Two exposure routes were applied. Oral gavage
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administration of FE: racemic FE was dissolved in corn oil containing 1% dimethyl
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sulfoxide and administered orally once at a dose of 1.0 ߤg/g by gavage. Tissue
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samples were collected at intervals of 1, 3, 5, 8, 12, 24, 48 and 72 hours after the oral
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administration. Aqueous exposure: racemic FE was spiked into the surrounding water
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at 0.5 ߤg/mL, and the water in the glass aquarium was replaced daily to maintain a
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constant concentration. Tissue samples were collected at exposure times of 3, 7, 14,
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21 and 28 days. Frogs were pithed and the liver, kidney, brain, eggs, skin and thigh
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muscle were excised and collected. Whole blood was sampled directly from the heart.
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Ten tadpoles were exposed in 4 L of water with 0.5 ߤg/mL racemic FE in a 5-L
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exposure times of 3, 7, 10, 12, 14 and 17 days. For the elimination, tadpoles in
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contaminated water were transferred to clean water after 17 days and recollected at
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day 0.125, 0.25, 0.5, 1, 2, 3, 4 and 7. The samples were immediately frozen and kept
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at -20 °C until further analysis.
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2.4 Extraction
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Frog tissues or tadpoles were homogenized by Mixer Mill MM400 (Retsch, Haan,
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Germany) and 0.5 g was placed into a 3-mL plastic centrifuge tube. The homogenized
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samples were extracted by the mixer mill with oscillation frequency of 30 r/sec for 3
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min with 1 mL of ethyl acetate and centrifuged at 10000 rpm for 3 min. Repeat the
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extraction and combine the upper solution. The combined solvent was evaporated to
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dryness at 35 °C, reconstituted in 1 mL of acetonitrile and then washed with n-hexane
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(2 mL) twice to remove the adipose. The acetonitrile layer was passed through a 0.22
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µm filter film for analysis.
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2.5 HPLC-MS/MS Analysis Analyses of CDHB and the individual enantiomers of FE and FA were performed
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on HPLC-MS/MS (TSQ Quantum Access Max, Thermo Scientific, Shanghai, China).
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A chiralpak IC chiral column (250×4.6 mm, Daicel Investment Company Limited,
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Shanghai, China) was used at 20 °C, and methanol/water/formic acid (75:25:0.1, v/v/v)
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with the flow rate of 500 ߤL/min was used as the mobile phase. The retention times of
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ACCEPTED MANUSCRIPT CDHB, R-FA, S-FA, R-FE and S-FE were 11.4, 16.8, 18.3, 77.3 and 85.3 min,
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respectively (Fig. S1). FE was analyzed in positive ionization mode ESI-MS/MS,
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while FA and CDHB were analyzed in negative ionization mode ESI-MS/MS. Setup
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for mass spectrometer were as follows: spray voltage of positive ionization mode
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3500 V; spray voltage of negative ionization mode 3000 V; vaporizer temperature
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300 °C; sheath gas pressure 40 Arb; axu gas pressure 10 Arb; ion sweep gas pressure
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0 Arb; collision gas pressure 1.5 mTorr. The fragmentation information of CDHB, FA
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and FE was 168.0/132.1 and 168.0/76.2, 332.0/260.0 and 332.0/152.0, 362.1/288.1
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and 362.1/244.0, respectively.
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2.6 Assay Validation
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Frog tissues and tadpoles were analyzed to ensure the absence of FE, FA and
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CDHB. No target compounds were found. Recoveries, calculated from spiked
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matrixes at 3 levels (racemic FE: 0.25, 0.5 and 1 ߤg/g; racemic FA and CDHB: 0.01,
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0.1 and 1 ߤg/g), were between 82.4% and 106.1% based on an RSD below 20%.
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Limits of detection (LODs), the concentrations that produced a signal-to-noise (S/N)
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ratio of 3, ranged between 0.0001 and 0.0370 ߤg/g (Table S1).
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2.7 Acute toxicity
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A tadpole was exposed to 400 mL of aqueous solution with a series of
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concentration of the chemical (rac-FE, R-FE, S-FE, rac-FA, R-FA, S-FA and CDHB)
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for 96 h. Controls were set with the cosolvent and without the chemicals. The aqueous
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preliminary tests, the concentration of rac-, R- and S-FE and FA was set to 20, 40, 60,
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80, 100 ߤg/mL and the concentration of CDHB was set as 22, 26, 30, 34, 38 ߤg/mL,
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respectively. After the 96-h exposure, the mortality was counted and analyzed using
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the SPSS Version 18.0.
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2.8 Genotoxicity
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The micronucleus test is extensively used to evaluate the genotoxicity of
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chemicals. Tadpoles were exposed to 400 mL of aqueous solution with three
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concentrations (0.1, 1 and 10 ߤg/mL) of the chemicals (rac-FE, R-FE, S-FE, rac-FA,
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R-FA, S-FA and CDHB) for 96 h. Controls were set with the cosolvent and without
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the chemicals. The aqueous solution was replaced daily. After the 96-h exposure,
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blood from tails was smeared onto microscope slides, fixed in methanol solution for
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15 min, stained in 10% Giemsa dye for 15 min, rinsed and dried. Micronucleus cells
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were scored for 1000 cells under a microscope. The micronucleus rate was calculated
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as permillage of micronucleus cells for three replicates.
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3. Results and discussion
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3.1 Bioaccumulation and elimination to frogs
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3.1.1 Oral administration
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Tissue distribution was quantified in the liver, kidney, brain, eggs, skin, muscle
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and whole blood, and the pharmacokinetics of oral FE was studied. Because of the
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concentrations of FE were found below LODs in all the tissues. The distribution and
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elimination of the main metabolite FA in each tissue were presented in Fig. 2 and the
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pharmacokinetic parameters were shown in Table S2. The results indicated that FA
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was widely distributed and reached maximum concentration in all the tissues within 8
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h after treatment. FA was found in brain, indicating that FA could permeate the blood
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brain barrier. But it was accumulated more slowly in the brain (8 h) than in other
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tissues (Table S2), probably because of the protective function of the blood brain
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barrier against the xenobiotics. Because both kidney and liver were the main
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metabolic tissues, a dominant tissue distribution of FA was observed in liver and
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kidney with higher Cmax (Table S2). In the elimination stage, FA was eliminated most
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rapidly in the whole blood with half-lives less than 3.7 h. In contrast, the half-lives
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were above 13.4 h in the other tissues.
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Enantiomer fraction (EF), defined as the concentration ratio of R-FA to the sum
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of S-FA and R-FA, was used as a measurement for the enantioselectivity of FA. In
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general, R-FA showed higher bioavailability (expressed as AUC in the whole blood,
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Table S2) and concentrations in all the tissues through the entire time, and the
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calculated EF values were greater than 0.5 (Fig. 2 and Fig. S2). It was reported there
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was chiral inversion of S-FA to R-FA in rabbits but no inversion of R-FA to S-FA
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(Zhang et al., 2011). Similar chiral inversion from S-enantiomer to R-enantiomer
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might be a reason for the enantioenrichment of R-FA in frogs. R-FA was found to be
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less enriched in brain compared with other tissues, with EF value about 0.63 in brain,
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concentration). In the elimination stage of FA, S-FA was preferentially eliminated in
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liver, eggs, skin, muscle and whole blood, while R-FA decreased more rapidly than
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S-FA in kidney. For example, the half-life of R-FA (35.0 h) was longer than S-FA
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(20.5 h) in liver. In contrast, the half-life of R-FA in kidney was 27.0 h which was
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slightly shorter than that of S-FA (29.5 h) (Table S2). The order of R-FA half-lives in
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the tissues was muscle (42.8 h) > skin (35.1 h) > liver (35.0 h) > kidney (27.0 h) >
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eggs (22.2 h) > brain (16.7 h) > whole blood (3.7 h). The order of S-FA half-lives was
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kidney (29.5 h) > muscle (27.1 h) > liver (20.5 h) > skin (15.7 h) > eggs (13.4 h) >
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whole blood (3.1 h). The concentration of S-FA in brain was too low to describe the
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elimination process and calculate the half-life (Table S2).
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After oral administration, the metabolite CDHB was only found in liver and
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kidney (Fig. 2), and the half-lives were less than 2 h. The concentration of CDHB did
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not increase with the metabolism of FA, indicating that FA was not mainly
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decomposed to CDHB in liver and kidney. The formation of CDHB may be from the
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breakdown of the benzoxazolyl-oxyphenyl ether linkage of FE (Lin et al., 2007; Jing
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et al., 2016).
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3.1.2 Aqueous exposure
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The concentrations of FE were below LODs in all frog tissues during the 28-day
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exposure. While the concentration of FA gradually increased, indicating that FA could
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be well accumulated in all the tissues by aqueous exposure of FE. The level of FA at
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ACCEPTED MANUSCRIPT day 28 was in the order of kidney > liver > skin > muscle > eggs > whole blood >
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brain (Fig. 3). Like oral administration, the concentrations were much higher in
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kidney and liver. The concentration-time curves of FA in liver and kidney were shown
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in Fig. S3. Because the skin directly contacted with contaminated water, there was a
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stronger accumulation tendency by skin compared with oral administration. The
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concentration of FA in skin was significantly higher than that in muscle, eggs, whole
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blood and brain. It was reported that pesticides could diffuse through frog skin
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because of the lack of a hydrophobic barrier in amphibian skin and high porosity to
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water molecules (Quaranta et al., 2009; Van Meter et al., 2014). Skin absorption was a
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dangerous route of exposure, because pollutants were carried directly through the
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blood circulation to organs without liver detoxification. According to UN Commodity
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Trade Statistics Database, the per year international trade in frog legs as food is 40
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million dollars (Turnipseed et al., 2012), thus the consumption of the contaminated
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thigh muscle may pose a potential threat to human health. The presence of FA in frog
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eggs should raise a concern about the potential reproductive risk of FE for frogs. FE
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has influence on the viability and maturation of porcine oocytes in vitro, indicating
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that FE may be a cause of reproductive dysfunction in animals (Casas et al., 2010).
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Predilection of R-FA in the enantioselective distribution in all tissues was found,
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which was consistent with that in oral administration exposed group. EF values after
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4-week exposure were in the order of whole blood (0.96) > muscle (0.95) > kidney
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(0.93) > liver (0.92) > eggs (0.87) > skin (0.85). The concentrations of S-FA in brain
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were below LOD, therefore the EF value of FA in brain was not given. Low amount
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of CDHB was detected only in the liver and kidney (Fig. S3).
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3.2 Bioaccumulation, elimination and toxicity to tadpoles
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3.2.1 Bioaccumulation and elimination
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During the accumulation period, the concentrations of FE and FA were below
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LODs. Thus the stereoselective bioaccumulation and elimination was not involved in
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tadpoles. Unlike in frogs, FA was not accumulated in tadpoles probably due to the
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existence of gill. For the elimination of chemicals in fishes, gill elimination rather
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than metabolism may dominate for chemicals with lower log KOW. Gill elimination
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rates tend to drop with increasing log KOW (Mackintosh et al., 2004). The log KOW of
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FA was 1.04, thus gill of tadpoles might play a similar role in the elimination of FA.
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In contrast, the log KOW of CDHB was 1.59. The concentration of CDHB gradually
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increased and reached an equilibrium level of 0.06 ߤg/g at day 14 (Fig. 4). The
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formation of CDHB may be from the breakdown of the benzoxazolyl-oxyphenyl ether
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linkage of FE or FA (Lin et al., 2007; Jing et al., 2016). During the elimination period,
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no FE or FA was detected and CDHB gradually eliminated as first-order kinetics with
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half-life of 37.1 h.
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3.2.2 Acute toxicity
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The 96-h acute toxicity of rac-FE, R-FE, S-FE, rac-FA, R-FA, S-FA and CDHB to
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tadpoles was studied by aqueous solution exposure. There were no deaths in solvent
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control group. Individual enantiomers and racemates of FE and FA exhibited low
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ߤg/mL and mortality was not observed even at 100 ߤg/mL showing the LC50 values of
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FE and FA were greater than 100 ߤg/mL. But CDHB exhibited higher acute toxicity
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with no tadpoles survived at a concentration of 38 ߤg/mL. The calculated 96-h LC50
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value of CDHB for tadpoles was 30.4 ߤg/mL (with 95% confidence interval of
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28.6-32.2 ߤg/mL).
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3.2.3 Genotoxicity
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Micronucleus test results were shown in Fig. 5. Compared with control, only
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rac-FA, S-FA and CDHB could induce statistically significant increase in the
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micronucleus rate in the highest dose group (10 ߤg/mL). The micronucleus rates were
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2.8, 2.4 and 1.7 times than that of control after 96-h exposure, respectively.
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Metabolites FA and CDHB had greater genetic toxicity than the parent compound FE.
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And enantioselective genetic toxicity of FA was found with R-FA less toxic than S-FA
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and rac-FA.
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4. Conclusion
In this study, the bioaccumulation and elimination of the chiral herbicide
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fenoxaprop-ethyl in frogs and tadpoles was determined by HPLC-MS/MS equipped
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with chiralpak IC chiral column. Fenoxaprop-ethyl was hardly accumulated because it
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was rapidly hydrolyzed to its related acid fenoxaprop. A dominant tissue distribution
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of fenoxaprop and 6-chloro-2,3-dihydrobenzoxazol-2-one was observed in frog
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chiral data indicated that preferential accumulation of R-fenoxaprop occurred in liver,
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kidney, brain, eggs, skin, thigh muscle and blood. A 28-day aqueous solution exposure
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experiment was also conducted, and it was found fenoxaprop in tissues continued to
335
increase during the whole exposure period indicating potential impact of
336
fenoxaprop-ethyl
337
6-chloro-2,3-dihydrobenzoxazol-2-one was accumulated in tadpoles after the 17-day
338
aqueous solution exposure of fenoxaprop-ethyl, and the elimination half-life was 37.1
339
h. Acute toxic and genotoxic potential to tadpoles was evaluated in this study. The
340
96-h LC50 value of CDHB for tadpoles was 30.4 ߤg/mL, while no death was induced
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by individual enantiomers and racemates of fenoxaprop-ethyl or fenoxaprop.
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Statistically significant increase of micronucleus rate in tadpole blood cells suggested
343
that genetic damage could be induced by rac-FA, S-FA and CDHB in the tested
344
concentrations. R-enantiomer are more likely to be bioaccumulated in tissues and
345
bring greater toxic effect. The results suggested the application of optically pure
346
fenoxaprop-ethyl (R-enantiomer) might pose less threat to amphibian.
environmental
water
to
frogs.
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Acknowledgements
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This work was supported by the National Natural Science Foundation of China
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(Contract Grants: 21337005, 21677175), supported by the New-Star of Science and
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Technology and by Beijing Nova program YETP0323.
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Romano, M.L., Stephenson, G.R., Tal, A., Hall, J.C., 1993. The Effect of Monooxygenase and
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Singh, S.B., Das, T.K., Kulshrestha, G., 2013. Persistence of herbicide fenoxaprop ethyl and its acid and Health, Part B 48, 324-330.
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Stabenau, E.K., Giczewski, D.T., Maillacheruvu, K.Y., 2006. Uptake and Elimination of Naphthalene from Liver, Lung, and Muscle Tissue in the Leopard frog (Rana pipiens). Journal of Environmental Science and Health, Part A 41, 1449-1461.
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Tal, A., Romano, M.L., Stephenson, G.R., Schwan, A.L., Hall, J.C., 1993. Glutathione Conjugation: A Detoxification Pathway for Fenoxaprop-ethyl in Barley, Crabgrass, Oat, and Wheat. Pestic. Biochem. Physiol. 46, 190-199.
Tilak, K.S., Veeraiah, K., Sastry, L.V., 2003. Bioaccumulation of fenvalerate technical grade in different organs of the frog Haplobatrachus tigerinus (Daudin). J. Environ. Biol. 24, 261-264.
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Turnipseed, S.B., Clark, S.B., Storey, J.M., Carr, J.R., 2012. Analysis of Veterinary Drug Residues in Frog Legs and Other Aquacultured Species Using Liquid Chromatography Quadrupole Time-of-Flight Mass Spectrometry. J. Agric. Food Chem. 60, 4430-4439. Van Meter, R.J., Glinski, D.A., Hong, T., Cyterski, M., Henderson, W.M., Purucker, S.T., 2014.
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Estimating terrestrial amphibian pesticide body burden through dermal exposure. Environ. Pollut. 193, 262-268.
Vogiatzis, K.A., Loumbourdis, S.N., 1997. Uptake, Tissue Distribution, and Depuration of Cadmium (Cd) in the Frog Rana ridibunda. Bul . Environ. Contam. Toxicol. 59, 770-776. Williams, A., 1996. Opportunities for chiral agrochemicals. Pesticide Science 46, 3-9. Wu, J.-P., Luo, X.-J., Zhang, Y., Chen, S.-J., Mai, B.-X., Guan, Y.-T., Yang, Z.-Y., 2009. Residues of Polybrominated Diphenyl Ethers in Frogs (Rana limnocharis) from a Contaminated Site, South China: Tissue Distribution, Biomagnification, and Maternal Transfer. Environ. Sci. Technol. 43, 5212-5217. Xie, H.S., Hsiao, A.I., QulCk, W.A., Hume, J.A., 1996. Influence of water stress on absorption, translocation and phytotoxicity of fenoxaprop-ethyl and imazamethabenz-methyl in Avena fatua. Weed Research 36, 65-71. Yaacoby, T., Hall, J.C., Stephenson, G.R., 1991. Influence of fenchlorazole-ethyl on the metabolism of
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ACCEPTED MANUSCRIPT fenoxaprop-ethyl in wheat, barley, and crabgrass. Pestic. Biochem. Physiol. 41, 296-304. Ye, J., Zhao, M., Niu, L., Liu, W., 2015. Enantioselective Environmental Toxicology of Chiral Pesticides. Chem. Res. Toxicol. 28, 325-338. Zablotowicz, R.M., Hoagland, R.E., Staddon, W.J., Locke, M.A., 2000. Effects of pH on Chemical Stability and De-esterification of Fenoxaprop-ethyl by Purified Enzymes, Bacterial Extracts, and Soils. J. Agric. Food Chem. 48, 4711-4716. Zawahir, S., Roberts, D.M., Palangasinghe, C., Mohamed, F., Eddleston, M., Dawson, A.H., Buckley,
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N.A., Ren, L., Medley, G.A., Gawarammana, I., 2009. Acute intentional self-poisoning with a herbicide product containing fenoxaprop-P-ethyl, ethoxysulfuron, and isoxadifen ethyl: a prospective observational study. Clin. Toxicol. 47, 792-797.
Zhang, A., Xu, C., Liu, W., 2008. Influence of toxicity and dissipation of racemic fenoxaprop and its
R-enantiomer in Scenedesmus obliquus suspension by cyclodextrins. Journal of Environmental Science
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and Health, Part B 43, 231-236.
Zhang, Y., Li, X., Shen, Z., Xu, X., Zhang, P., Wang, P., Zhou, Z., 2011. Stereoselective metabolism of fenoxaprop-ethyl and its chiral metabolite fenoxaprop in rabbits. Chirality 23, 897-903. Zhang, Y., Liu, D., Diao, J., He, Z., Zhou, Z., Wang, P., Li, X., 2010. Enantioselective Environmental Behavior of the Chiral Herbicide Fenoxaprop-ethyl and Its Chiral Metabolite Fenoxaprop in Soil. J. Agric. Food Chem. 58, 12878-12884.
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Zhou, Q., Liu, W., 2011. Phytotoxicity and Environmental Fate of Chiral Herbicides. Chiral Pesticides: Stereoselectivity and Its Consequences. American Chemical Society, pp. 135-150.
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Fig. 1. Chemical structures of FE and its metabolites FA and CDHB.
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* denotes chiral center
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R-FA in liver S-FA in liver
R-FA in kidney S-FA in kidney
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1.2 0.8
12 8 4
0.4
0
0.0 0
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36
48
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Time (h)
0.3
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24
36
0.3
24
36
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60
0.1
0.0
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0
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0.8
Time (h)
R-FA in skin S-FA in skin 0.2
0.1
36
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Time (h)
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R-FA in muscle S-FA in muscle
0.6
Conc. (µg/g)
Conc. (µg/g)
0.2
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Conc. (µg/g)
Conc. (µg/g)
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0
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R-FA in eggs S-FA in eggs
0.2
0.3
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Time (h)
R-FA in brain S-FA in brain
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Conc. (µg/g)
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0.4
0.0
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60
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Time (h)
0.6
Conc. (µg/g)
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36
48
60
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Time (h) CDHB in liver CDHB in kindey
0.20 0.15 0.10
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0.00 0
Time (h)
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0.0
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0.25
R-FA in blood S-FA in blood
0.4
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24
36
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60
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Time (h)
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Fig. 2. Concentration-time curves of R-FA, S-FA and CDHB in frog tissues following
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a single oral dose of 1.0 ߤg/g.
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R-FA S-FA
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kidney liver
skin muscle eggs blood brain
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Fig. 3. Concentration of R-FA and S-FA in frog tissues exposed to 0.5 ߤg/mL aqueous
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solution of FE at day 28
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Conc. (µg/g)
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0.00 0
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Time (d)
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Fig. 4. Concentration-time curves of CDHB in tadpoles exposed to 0.5 ߤg/mL
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aqueous solution of FE
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40 30 20 10 0
40 30 20 10 0
CK
0.1
1
10
Concentration (µg/mL)
50
CK rac-FA R-FA S-FA
* *
Micronucleus rate (‰)
50
CK rac-FE R-FE S-FE
CK CDHB
40
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Micronucleus rate (‰)
531
50
Micronucleus rates (‰)
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30
*
20 10 0
CK
0.1
1
Concentration (µg/mL)
10
CK
0.1
1
10
Concentration (µg/mL)
Fig. 5. Blood cells micronucleus rate in tadpoles exposed to 0, 0.1, 1, 10 ߤg/mL
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aqueous solution of FE, FA and CDHB
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Fig. 1. Chemical structures of FE and its metabolites FA and CDHB.
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* denotes chiral center
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Fig. 2. Concentration-time curves of R-FA, S-FA and CDHB in frog tissues following
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a single oral dose of 1.0 ߤg/g.
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Fig. 3. Concentration of R-FA and S-FA in frog tissues exposed to 0.5 ߤg/mL aqueous
551
solution of FE at day 28
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Fig. 4. Concentration-time curves of CDHB in tadpoles exposed to 0.5 ߤg/mL
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aqueous solution of FE
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Fig. 5. Blood cells micronucleus rate in tadpoles exposed to 0, 0.1, 1, 10 ߤg/mL
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aqueous solution of FE, FA and CDHB
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Highlights 1. Enantioselective bioaccumulation in frogs and tadpoles was investigated. 2. Metabolite fenoxaprop distributed in brain and eggs of frogs. 3. Enantioselective acute toxicity and genotoxicity in tadpoles was conducted. 4. Metabolites were more toxic than the parent compound fenoxaprop-ethyl to tadpoles.
S0045-6535(17)31183-9
DOI:
10.1016/j.chemosphere.2017.07.132
Reference:
CHEM 19665
To appear in:
ECSN
Received Date: 25 April 2017 Revised Date:
25 June 2017
Accepted Date: 25 July 2017
Please cite this article as: Jing, X., Yao, G., Liu, D., Liu, C., Wang, F., Wang, P., Zhou, Z., Exposure of frogs and tadpoles to chiral herbicide fenoxaprop-ethyl, Chemosphere (2017), doi: 10.1016/ j.chemosphere.2017.07.132. This is a PDF file of an unedited manuscript that has been accepted for publication. As a service to our customers we are providing this early version of the manuscript. The manuscript will undergo copyediting, typesetting, and review of the resulting proof before it is published in its final form. Please note that during the production process errors may be discovered which could affect the content, and all legal disclaimers that apply to the journal pertain.
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Graphical abstract
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Exposure of frogs and tadpoles to chiral herbicide
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fenoxaprop-ethyl Xu Jing, Guojun Yao, Donghui Liu, Chang Liu, Fang Wang, Peng Wang, Zhiqiang
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Zhou*
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Beijing Advanced Innovation Center for Food Nutrition and Human Health, Department of Applied
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Chemistry, China Agricultural University, Beijing, 100193, P.R. China.
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*Corresponding author:
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Zhiqiang Zhou, Department of Applied Chemistry, China Agricultural University, No.2
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Yuanmingyuan West Road, Beijing 100193, P.R. China; Tel: +8610-62733547; Fax:
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+8610-62733547; E-mail: [email protected]
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ABSTRACT
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Pesticides have long been considered to a risk factor of amphibian population declines.
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The bioaccumulation and elimination of fenoxaprop-ethyl (FE) in frogs and tadpoles
26
were
27
6-chloro-2,3-dihydrobenzoxazol-2-one (CDHB) were monitored. The acute toxicity
28
and genotoxicity of the enantiomers to tadpoles was also studied. After both oral
29
administration and aqueous solution exposure, FE was not found in frogs, while FA
30
was formed and accumulated in liver, kidney, brain, eggs, skin, thigh muscle and
31
blood with preferential accumulation of R-FA. The presence of FA in frog eggs
32
suggested maternal transfer in females and potential impacts to offsprings. The
33
elimination of FA in frog tissues was also enantioselective with a preferential
34
metabolism of R-FA (kidney) or S-FA (liver, eggs, skin, muscle and whole blood). FE
35
and FA were hardly detectable in tadpoles after aqueous solution exposure, while
36
CDHB was accumulated and eliminated as first-order kinetics with half-life of 37.1 h.
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Mortality of tadpoles and micronucleus rate in peripheral blood erythrocytes of
38
tadpoles were used to evaluate the enantioselective acute toxicity and genotoxicity.
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Only CDHB induced significant acute toxicity to tadpole with 96-h LC50 value of 30.4
40
ߤg/mL, and rac-FA, S-FA and CDHB showed genotoxicity.
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the
main
metabolites
fenoxaprop
(FA)
and
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studied
KEYWORDS: Fenoxaprop-ethyl; Frog; Tadpole; Chiral; Toxicity
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1. Introduction
Pesticides have been widely used in agriculture, in which about 30% are chiral
47
(Williams, 1996). Although the enantiomers of a chiral pesticide almost have the same
48
physical and chemical properties, they behave enantioselectively in ecology with
49
different bioactivity, toxicity and environmental behavior. These pesticides pose
50
unpredictable enantioselective environmental threats, affecting the food chain and
51
entire ecosystems (Ye et al., 2015). It is urgent to understand the toxicity and
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metabolism of enantiomers of pesticides in target and non-target organisms on
53
enantiomeric level. There are 19 major categories of herbicides on the market in
54
which 14 categories contain chiral structures (Zhou and Liu, 2011). Aryloxyphenoxy
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propionate herbicides are widely used and the weed control is achieved by the active
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R-enantiomer.
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(FE,
Fig.
1) is
a post-emergent
herbicide of the
aryloxyphenoxy propionate group (Zawahir et al., 2009). It inhibits fatty acid
59
synthesis in grasses by inhibition of acetyl-CoA carboxylase. The key step of
60
degradation is ester hydrolysis to its related acid fenoxaprop (FA, Fig. 1) (Dong et al.,
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2015), which is 100 times more active than FE in inhibiting acetyl-CoA carboxylase
62
(Yaacoby et al., 1991). FA may undergo further degradation, forming the metabolite
63
6-chloro-2,3-dihydrobenzoxazol-2-one (CDHB, Fig. 1) (Hoagland and Zablotowicz,
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1998). Metabolite CDHB may also come directly from the breakdown of the
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benzoxazolyl-oxyphenyl ether linkage of FE (Lin et al., 2007; Jing et al., 2016). The
66
degradation of FE have been widely studied in the environment, such as in soil
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ACCEPTED MANUSCRIPT (Zhang et al., 2010), water (Lin et al., 2007), microorganism (Dong et al., 2015),
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wheat (Singh et al., 2013), rice (Lucini and Pietro Molinari, 2010), oat (Xie et al.,
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1996), barley (Romano et al., 1993), crabgrass (Tal et al., 1993), rat (Moody and
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Ritter, 1992) and rabbit (Zhang et al., 2011). Influencing factors like pH (Zablotowicz
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et al., 2000; Lin et al., 2007), temperature and light, and so on (Xie et al., 1996; Lin et
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al., 2008) have been investigated. Both FE and FA are chiral and exist as two
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enantiomers (Fig. 1). According to the literature, S-FE decreased faster than R-FE in
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soils, and the inversion of S-enantiomer to R-enantiomer occurred (Zhang et al.,
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2010). In rabbits, S-FA decreased faster in plasma, heart, lung, liver, kidney, and bile
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than its antipode with similar inversion (Zhang et al., 2011). The conversion between
77
the enantiomers was not found in Scenedesmus obliquus suspension (Zhang et al.,
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2008). To accurately evaluate the impacts of FE on the environment, the
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enantioselective behavior of FE and FA should be taken into account.
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The toxicity of FE and its metabolites to some non-target organisms has been
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studied. The 96-h LC50 values for FE to rainbow trout were 0.57 mg/L (Song et al.,
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2005). The 48-h EC50 values for FE, FA and CDHB to Daphnia magna were 5.29,
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14.6 and 8.4 mg/L, respectively (Lin et al., 2008). The 48-h EC50 values for rac-FA
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and R-FA to freshwater alga Scenedesmus obliquus were 8.42 and 8.03 mg/L,
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respectively (Zhang et al., 2008). Clinical toxicity data was achieved by patients with
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acute exposure to commercial products of FE. The main clinical features were
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epigastric burning sensation, vomiting and consciousness recession (Zawahir et al.,
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2009).
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ACCEPTED MANUSCRIPT Currently, most animal studies concerning pollution were achieved by aquatic or
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terrestrial animals, while amphibians attracted little attention. Amphibians may be
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good environmental pollutant bioindicator due to the unique characteristics, such as
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aquatic and terrestrial habitats, complex life cycles, rapid larval growth rates, trophic
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position, poikilothermy and permeable skin (Burkhart et al., 2000; Fontenot et al.,
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2000). Worldwide amphibian decline has triggered massive research efforts to explore
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biotic and abiotic factors. Amphibians are sensitive to environmental chemicals,
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which have been implicated as factors contributing to large-scale losses of amphibian
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populations (Hayes et al., 2006). The adverse effects of pesticides on amphibian
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growth, development, reproduction and behavior have long been a problem, and the
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mixture may enhance the effects (Renner, 2003). It has been widely expected that the
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impacts of xenobiotics depend on their uptake, tissue distribution and metabolism. In
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recent years, the bioavailability and tissue distribution of heavy metals, PAHs, PCBs
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and PBDEs in amphibian was examined (Vogiatzis and Loumbourdis, 1997; Fontenot
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et al., 2000; Stabenau et al., 2006; Wu et al., 2009; Li et al., 2014). The
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bioaccumulation, transformation and elimination information about pesticides is
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relatively limited (Tilak et al., 2003; Reynaud et al., 2012). Besides, breeding and
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larval development of amphibians occur in spring and summer when there is frequent
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application of pesticides (Mann et al., 2009). Corresponding environmental behaviors
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and toxicity studies are vital to predict chemical fate in organisms, develop
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environmental quality criteria and assess the ecological risks. Further effort is needed
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to study the bioaccumulation, elimination and toxicity of pesticides in amphibians.
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behavior and toxicity studies of FE have not been conducted in amphibians, the
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present study was conducted to evaluate the bioaccumulation, elimination, acute
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toxicity and genotoxicity of FE and its metabolites FA and CDHB in frogs and
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tadpoles. Frogs were administered by a single oral dose of 1.0 ߤg/g of FE or exposed
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to 0.5 ߤg/mL aqueous solution of FE. Tadpoles were exposed to 0.5 ߤg/mL aqueous
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solution of FE. The concentrations of FE, FA and CDHB in frogs (liver, kidney, brain,
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eggs, skin, thigh muscle and whole blood) and tadpoles were determined on an
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enantiomeric level by high-performance liquid chromatography-tandem mass
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spectrometry (HPLC-MS/MS). Individual enantiomers of FE and FA were prepared to
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assess the acute toxicity and genotoxicity to tadpoles. The goal is to expand on the
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limited data on the bioaccumulation, elimination and toxicity of chiral pesticides in
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amphibians.
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2. Materials and methods
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2.1Chemicals
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Racemic FE (98.0%) was got from Institute for Control of Agrichemicals,
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Ministry of Agriculture of China (Beijing, China). The synthesis of racemic FA and
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the single enantiomer of FE and FA (98.0%) were conducted according to literatures
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(Amabilino et al., 1998; Kato et al., 2003; Moon et al., 2007). CDHB (98.0%) was
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obtained from TCI Development Company Limited (Shanghai, China). Dimethyl
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sulfoxide and corn oil was obtained from Sigma-Aldrich (St. Louis, MO, USA).
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Methanol and acetonitrile were obtained from Fisher Scientific (Fair Lawn, NJ, USA).
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Water was purified by a Milli-Q system.
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2.2 Animals
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Adult female bullfrogs (Rana catesbeiana) and premetamorphic tadpoles (Rana
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catesbeiana) were obtained from local aquafarm. The average weight of frogs and
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tadpoles was 500 ± 25 and 0.6 ± 0.03 g, respectively. The frogs and tadpoles were
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respectively placed in glass aquariums and glass beakers for 1 week at 25 °C with a
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photoperiod of 12 h: 12 h, for the purpose of acclimatization.
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2.3 Exposure to FE
Frogs were exposed individually in glass aquariums (60 cm × 45 cm × 30 cm)
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containing 5 L of water. Two exposure routes were applied. Oral gavage
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administration of FE: racemic FE was dissolved in corn oil containing 1% dimethyl
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sulfoxide and administered orally once at a dose of 1.0 ߤg/g by gavage. Tissue
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samples were collected at intervals of 1, 3, 5, 8, 12, 24, 48 and 72 hours after the oral
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administration. Aqueous exposure: racemic FE was spiked into the surrounding water
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at 0.5 ߤg/mL, and the water in the glass aquarium was replaced daily to maintain a
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constant concentration. Tissue samples were collected at exposure times of 3, 7, 14,
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21 and 28 days. Frogs were pithed and the liver, kidney, brain, eggs, skin and thigh
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muscle were excised and collected. Whole blood was sampled directly from the heart.
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Ten tadpoles were exposed in 4 L of water with 0.5 ߤg/mL racemic FE in a 5-L
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exposure times of 3, 7, 10, 12, 14 and 17 days. For the elimination, tadpoles in
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contaminated water were transferred to clean water after 17 days and recollected at
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day 0.125, 0.25, 0.5, 1, 2, 3, 4 and 7. The samples were immediately frozen and kept
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at -20 °C until further analysis.
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2.4 Extraction
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Frog tissues or tadpoles were homogenized by Mixer Mill MM400 (Retsch, Haan,
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Germany) and 0.5 g was placed into a 3-mL plastic centrifuge tube. The homogenized
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samples were extracted by the mixer mill with oscillation frequency of 30 r/sec for 3
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min with 1 mL of ethyl acetate and centrifuged at 10000 rpm for 3 min. Repeat the
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extraction and combine the upper solution. The combined solvent was evaporated to
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dryness at 35 °C, reconstituted in 1 mL of acetonitrile and then washed with n-hexane
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(2 mL) twice to remove the adipose. The acetonitrile layer was passed through a 0.22
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µm filter film for analysis.
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2.5 HPLC-MS/MS Analysis Analyses of CDHB and the individual enantiomers of FE and FA were performed
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on HPLC-MS/MS (TSQ Quantum Access Max, Thermo Scientific, Shanghai, China).
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A chiralpak IC chiral column (250×4.6 mm, Daicel Investment Company Limited,
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Shanghai, China) was used at 20 °C, and methanol/water/formic acid (75:25:0.1, v/v/v)
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with the flow rate of 500 ߤL/min was used as the mobile phase. The retention times of
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respectively (Fig. S1). FE was analyzed in positive ionization mode ESI-MS/MS,
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while FA and CDHB were analyzed in negative ionization mode ESI-MS/MS. Setup
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for mass spectrometer were as follows: spray voltage of positive ionization mode
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3500 V; spray voltage of negative ionization mode 3000 V; vaporizer temperature
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300 °C; sheath gas pressure 40 Arb; axu gas pressure 10 Arb; ion sweep gas pressure
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0 Arb; collision gas pressure 1.5 mTorr. The fragmentation information of CDHB, FA
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and FE was 168.0/132.1 and 168.0/76.2, 332.0/260.0 and 332.0/152.0, 362.1/288.1
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and 362.1/244.0, respectively.
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2.6 Assay Validation
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Frog tissues and tadpoles were analyzed to ensure the absence of FE, FA and
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CDHB. No target compounds were found. Recoveries, calculated from spiked
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matrixes at 3 levels (racemic FE: 0.25, 0.5 and 1 ߤg/g; racemic FA and CDHB: 0.01,
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0.1 and 1 ߤg/g), were between 82.4% and 106.1% based on an RSD below 20%.
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Limits of detection (LODs), the concentrations that produced a signal-to-noise (S/N)
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ratio of 3, ranged between 0.0001 and 0.0370 ߤg/g (Table S1).
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2.7 Acute toxicity
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A tadpole was exposed to 400 mL of aqueous solution with a series of
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concentration of the chemical (rac-FE, R-FE, S-FE, rac-FA, R-FA, S-FA and CDHB)
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for 96 h. Controls were set with the cosolvent and without the chemicals. The aqueous
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preliminary tests, the concentration of rac-, R- and S-FE and FA was set to 20, 40, 60,
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80, 100 ߤg/mL and the concentration of CDHB was set as 22, 26, 30, 34, 38 ߤg/mL,
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respectively. After the 96-h exposure, the mortality was counted and analyzed using
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the SPSS Version 18.0.
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2.8 Genotoxicity
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The micronucleus test is extensively used to evaluate the genotoxicity of
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chemicals. Tadpoles were exposed to 400 mL of aqueous solution with three
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concentrations (0.1, 1 and 10 ߤg/mL) of the chemicals (rac-FE, R-FE, S-FE, rac-FA,
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R-FA, S-FA and CDHB) for 96 h. Controls were set with the cosolvent and without
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the chemicals. The aqueous solution was replaced daily. After the 96-h exposure,
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blood from tails was smeared onto microscope slides, fixed in methanol solution for
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15 min, stained in 10% Giemsa dye for 15 min, rinsed and dried. Micronucleus cells
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were scored for 1000 cells under a microscope. The micronucleus rate was calculated
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as permillage of micronucleus cells for three replicates.
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3. Results and discussion
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3.1 Bioaccumulation and elimination to frogs
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3.1.1 Oral administration
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Tissue distribution was quantified in the liver, kidney, brain, eggs, skin, muscle
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and whole blood, and the pharmacokinetics of oral FE was studied. Because of the
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concentrations of FE were found below LODs in all the tissues. The distribution and
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elimination of the main metabolite FA in each tissue were presented in Fig. 2 and the
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pharmacokinetic parameters were shown in Table S2. The results indicated that FA
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was widely distributed and reached maximum concentration in all the tissues within 8
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h after treatment. FA was found in brain, indicating that FA could permeate the blood
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brain barrier. But it was accumulated more slowly in the brain (8 h) than in other
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tissues (Table S2), probably because of the protective function of the blood brain
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barrier against the xenobiotics. Because both kidney and liver were the main
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metabolic tissues, a dominant tissue distribution of FA was observed in liver and
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kidney with higher Cmax (Table S2). In the elimination stage, FA was eliminated most
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rapidly in the whole blood with half-lives less than 3.7 h. In contrast, the half-lives
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were above 13.4 h in the other tissues.
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Enantiomer fraction (EF), defined as the concentration ratio of R-FA to the sum
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of S-FA and R-FA, was used as a measurement for the enantioselectivity of FA. In
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general, R-FA showed higher bioavailability (expressed as AUC in the whole blood,
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Table S2) and concentrations in all the tissues through the entire time, and the
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calculated EF values were greater than 0.5 (Fig. 2 and Fig. S2). It was reported there
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was chiral inversion of S-FA to R-FA in rabbits but no inversion of R-FA to S-FA
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(Zhang et al., 2011). Similar chiral inversion from S-enantiomer to R-enantiomer
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might be a reason for the enantioenrichment of R-FA in frogs. R-FA was found to be
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less enriched in brain compared with other tissues, with EF value about 0.63 in brain,
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concentration). In the elimination stage of FA, S-FA was preferentially eliminated in
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liver, eggs, skin, muscle and whole blood, while R-FA decreased more rapidly than
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S-FA in kidney. For example, the half-life of R-FA (35.0 h) was longer than S-FA
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(20.5 h) in liver. In contrast, the half-life of R-FA in kidney was 27.0 h which was
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slightly shorter than that of S-FA (29.5 h) (Table S2). The order of R-FA half-lives in
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the tissues was muscle (42.8 h) > skin (35.1 h) > liver (35.0 h) > kidney (27.0 h) >
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eggs (22.2 h) > brain (16.7 h) > whole blood (3.7 h). The order of S-FA half-lives was
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kidney (29.5 h) > muscle (27.1 h) > liver (20.5 h) > skin (15.7 h) > eggs (13.4 h) >
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whole blood (3.1 h). The concentration of S-FA in brain was too low to describe the
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elimination process and calculate the half-life (Table S2).
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After oral administration, the metabolite CDHB was only found in liver and
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kidney (Fig. 2), and the half-lives were less than 2 h. The concentration of CDHB did
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not increase with the metabolism of FA, indicating that FA was not mainly
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decomposed to CDHB in liver and kidney. The formation of CDHB may be from the
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breakdown of the benzoxazolyl-oxyphenyl ether linkage of FE (Lin et al., 2007; Jing
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et al., 2016).
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3.1.2 Aqueous exposure
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The concentrations of FE were below LODs in all frog tissues during the 28-day
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exposure. While the concentration of FA gradually increased, indicating that FA could
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be well accumulated in all the tissues by aqueous exposure of FE. The level of FA at
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brain (Fig. 3). Like oral administration, the concentrations were much higher in
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kidney and liver. The concentration-time curves of FA in liver and kidney were shown
268
in Fig. S3. Because the skin directly contacted with contaminated water, there was a
269
stronger accumulation tendency by skin compared with oral administration. The
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concentration of FA in skin was significantly higher than that in muscle, eggs, whole
271
blood and brain. It was reported that pesticides could diffuse through frog skin
272
because of the lack of a hydrophobic barrier in amphibian skin and high porosity to
273
water molecules (Quaranta et al., 2009; Van Meter et al., 2014). Skin absorption was a
274
dangerous route of exposure, because pollutants were carried directly through the
275
blood circulation to organs without liver detoxification. According to UN Commodity
276
Trade Statistics Database, the per year international trade in frog legs as food is 40
277
million dollars (Turnipseed et al., 2012), thus the consumption of the contaminated
278
thigh muscle may pose a potential threat to human health. The presence of FA in frog
279
eggs should raise a concern about the potential reproductive risk of FE for frogs. FE
280
has influence on the viability and maturation of porcine oocytes in vitro, indicating
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that FE may be a cause of reproductive dysfunction in animals (Casas et al., 2010).
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Predilection of R-FA in the enantioselective distribution in all tissues was found,
283
which was consistent with that in oral administration exposed group. EF values after
284
4-week exposure were in the order of whole blood (0.96) > muscle (0.95) > kidney
285
(0.93) > liver (0.92) > eggs (0.87) > skin (0.85). The concentrations of S-FA in brain
286
were below LOD, therefore the EF value of FA in brain was not given. Low amount
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of CDHB was detected only in the liver and kidney (Fig. S3).
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3.2 Bioaccumulation, elimination and toxicity to tadpoles
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3.2.1 Bioaccumulation and elimination
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During the accumulation period, the concentrations of FE and FA were below
292
LODs. Thus the stereoselective bioaccumulation and elimination was not involved in
293
tadpoles. Unlike in frogs, FA was not accumulated in tadpoles probably due to the
294
existence of gill. For the elimination of chemicals in fishes, gill elimination rather
295
than metabolism may dominate for chemicals with lower log KOW. Gill elimination
296
rates tend to drop with increasing log KOW (Mackintosh et al., 2004). The log KOW of
297
FA was 1.04, thus gill of tadpoles might play a similar role in the elimination of FA.
298
In contrast, the log KOW of CDHB was 1.59. The concentration of CDHB gradually
299
increased and reached an equilibrium level of 0.06 ߤg/g at day 14 (Fig. 4). The
300
formation of CDHB may be from the breakdown of the benzoxazolyl-oxyphenyl ether
301
linkage of FE or FA (Lin et al., 2007; Jing et al., 2016). During the elimination period,
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no FE or FA was detected and CDHB gradually eliminated as first-order kinetics with
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half-life of 37.1 h.
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3.2.2 Acute toxicity
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The 96-h acute toxicity of rac-FE, R-FE, S-FE, rac-FA, R-FA, S-FA and CDHB to
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tadpoles was studied by aqueous solution exposure. There were no deaths in solvent
308
control group. Individual enantiomers and racemates of FE and FA exhibited low
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310
ߤg/mL and mortality was not observed even at 100 ߤg/mL showing the LC50 values of
311
FE and FA were greater than 100 ߤg/mL. But CDHB exhibited higher acute toxicity
312
with no tadpoles survived at a concentration of 38 ߤg/mL. The calculated 96-h LC50
313
value of CDHB for tadpoles was 30.4 ߤg/mL (with 95% confidence interval of
314
28.6-32.2 ߤg/mL).
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3.2.3 Genotoxicity
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Micronucleus test results were shown in Fig. 5. Compared with control, only
318
rac-FA, S-FA and CDHB could induce statistically significant increase in the
319
micronucleus rate in the highest dose group (10 ߤg/mL). The micronucleus rates were
320
2.8, 2.4 and 1.7 times than that of control after 96-h exposure, respectively.
321
Metabolites FA and CDHB had greater genetic toxicity than the parent compound FE.
322
And enantioselective genetic toxicity of FA was found with R-FA less toxic than S-FA
323
and rac-FA.
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4. Conclusion
In this study, the bioaccumulation and elimination of the chiral herbicide
327
fenoxaprop-ethyl in frogs and tadpoles was determined by HPLC-MS/MS equipped
328
with chiralpak IC chiral column. Fenoxaprop-ethyl was hardly accumulated because it
329
was rapidly hydrolyzed to its related acid fenoxaprop. A dominant tissue distribution
330
of fenoxaprop and 6-chloro-2,3-dihydrobenzoxazol-2-one was observed in frog
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ACCEPTED MANUSCRIPT kidney and liver. Fenoxaprop could permeate the blood brain barrier into brain. The
332
chiral data indicated that preferential accumulation of R-fenoxaprop occurred in liver,
333
kidney, brain, eggs, skin, thigh muscle and blood. A 28-day aqueous solution exposure
334
experiment was also conducted, and it was found fenoxaprop in tissues continued to
335
increase during the whole exposure period indicating potential impact of
336
fenoxaprop-ethyl
337
6-chloro-2,3-dihydrobenzoxazol-2-one was accumulated in tadpoles after the 17-day
338
aqueous solution exposure of fenoxaprop-ethyl, and the elimination half-life was 37.1
339
h. Acute toxic and genotoxic potential to tadpoles was evaluated in this study. The
340
96-h LC50 value of CDHB for tadpoles was 30.4 ߤg/mL, while no death was induced
341
by individual enantiomers and racemates of fenoxaprop-ethyl or fenoxaprop.
342
Statistically significant increase of micronucleus rate in tadpole blood cells suggested
343
that genetic damage could be induced by rac-FA, S-FA and CDHB in the tested
344
concentrations. R-enantiomer are more likely to be bioaccumulated in tissues and
345
bring greater toxic effect. The results suggested the application of optically pure
346
fenoxaprop-ethyl (R-enantiomer) might pose less threat to amphibian.
environmental
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frogs.
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Acknowledgements
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This work was supported by the National Natural Science Foundation of China
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(Contract Grants: 21337005, 21677175), supported by the New-Star of Science and
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Technology and by Beijing Nova program YETP0323.
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N.A., Ren, L., Medley, G.A., Gawarammana, I., 2009. Acute intentional self-poisoning with a herbicide product containing fenoxaprop-P-ethyl, ethoxysulfuron, and isoxadifen ethyl: a prospective observational study. Clin. Toxicol. 47, 792-797.
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Zhou, Q., Liu, W., 2011. Phytotoxicity and Environmental Fate of Chiral Herbicides. Chiral Pesticides: Stereoselectivity and Its Consequences. American Chemical Society, pp. 135-150.
AC C
EP
TE D
461
19
ACCEPTED MANUSCRIPT 462 463 464
RI PT
465 466
M AN U
SC
467
468
Fig. 1. Chemical structures of FE and its metabolites FA and CDHB.
470
* denotes chiral center
473 474 475 476
EP
472
AC C
471
TE D
469
477 478 479
20
ACCEPTED MANUSCRIPT 480 2.0
20
R-FA in liver S-FA in liver
R-FA in kidney S-FA in kidney
16
1.2 0.8
12 8 4
0.4
0
0.0 0
12
24
36
48
60
0
72
Time (h)
0.3
12
24
36
0.3
24
36
48
60
0.1
0.0
72
0
12
24
0.8
Time (h)
R-FA in skin S-FA in skin 0.2
0.1
36
48
Time (h)
60
72
R-FA in muscle S-FA in muscle
0.6
Conc. (µg/g)
Conc. (µg/g)
0.2
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12
72
SC
Conc. (µg/g)
Conc. (µg/g)
0.1
0
60
R-FA in eggs S-FA in eggs
0.2
0.3
48
Time (h)
R-FA in brain S-FA in brain
0.0
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Conc. (µg/g)
Conc. (µg/g)
1.6
0.4
0.0
TE D
0.2
0.0
0
12
24
36
48
60
72
Time (h)
0.6
Conc. (µg/g)
EP
AC C
Conc. (µg/g)
0.2
12
24
36
482
36
48
60
72
Time (h) CDHB in liver CDHB in kindey
0.20 0.15 0.10
48
60
72
0.00 0
Time (h)
481
24
0.05
0.0
0
12
0.25
R-FA in blood S-FA in blood
0.4
0
12
24
36
48
60
72
Time (h)
483
Fig. 2. Concentration-time curves of R-FA, S-FA and CDHB in frog tissues following
484
a single oral dose of 1.0 ߤg/g.
485
21
ACCEPTED MANUSCRIPT 486 487 488
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489 490 491
SC
492
R-FA S-FA
2
1
0
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Conc. (µg/g)
18 17
kidney liver
skin muscle eggs blood brain
TE D
493
Fig. 3. Concentration of R-FA and S-FA in frog tissues exposed to 0.5 ߤg/mL aqueous
495
solution of FE at day 28
497 498 499 500
AC C
496
EP
494
501 502 503
22
ACCEPTED MANUSCRIPT 504 505 506
RI PT
507 508
510
0.08
Conc. (µg/g)
0.04
0.02
0.00 0
M AN U
CDHB in tadpoles 0.06
SC
509
7
14
21
28
Time (d)
511
Fig. 4. Concentration-time curves of CDHB in tadpoles exposed to 0.5 ߤg/mL
513
aqueous solution of FE
516 517 518
EP
515
AC C
514
TE D
512
519 520 521 522 23
ACCEPTED MANUSCRIPT 523 524 525
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526 527 528
533
40 30 20 10 0
40 30 20 10 0
CK
0.1
1
10
Concentration (µg/mL)
50
CK rac-FA R-FA S-FA
* *
Micronucleus rate (‰)
50
CK rac-FE R-FE S-FE
CK CDHB
40
M AN U
532
Micronucleus rate (‰)
531
50
Micronucleus rates (‰)
530
SC
529
30
*
20 10 0
CK
0.1
1
Concentration (µg/mL)
10
CK
0.1
1
10
Concentration (µg/mL)
Fig. 5. Blood cells micronucleus rate in tadpoles exposed to 0, 0.1, 1, 10 ߤg/mL
535
aqueous solution of FE, FA and CDHB
538 539 540 541
EP
537
AC C
536
TE D
534
542 543 544
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ACCEPTED MANUSCRIPT Figure Captions
546
Fig. 1. Chemical structures of FE and its metabolites FA and CDHB.
547
* denotes chiral center
548
Fig. 2. Concentration-time curves of R-FA, S-FA and CDHB in frog tissues following
549
a single oral dose of 1.0 ߤg/g.
550
Fig. 3. Concentration of R-FA and S-FA in frog tissues exposed to 0.5 ߤg/mL aqueous
551
solution of FE at day 28
552
Fig. 4. Concentration-time curves of CDHB in tadpoles exposed to 0.5 ߤg/mL
553
aqueous solution of FE
554
Fig. 5. Blood cells micronucleus rate in tadpoles exposed to 0, 0.1, 1, 10 ߤg/mL
555
aqueous solution of FE, FA and CDHB
SC
M AN U
EP AC C
557
TE D
556
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545
25
ACCEPTED MANUSCRIPT
AC C
EP
TE D
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Highlights 1. Enantioselective bioaccumulation in frogs and tadpoles was investigated. 2. Metabolite fenoxaprop distributed in brain and eggs of frogs. 3. Enantioselective acute toxicity and genotoxicity in tadpoles was conducted. 4. Metabolites were more toxic than the parent compound fenoxaprop-ethyl to tadpoles.