Chemosphere 89 (2012) 1015–1025
Contents lists available at SciVerse ScienceDirect
Chemosphere
journal homepage: www.elsevier.com/locate/chemosphere
Review
Review of measured concentrations of triphenyltin compounds in marine
ecosystems and meta-analysis of their risks to humans and the environment
Andy Xianliang Yi a,b, Kenneth M.Y. Leung a,b,c,d,⇑, Michael H.W. Lam c,d,e, Jae-Seong Lee c,f,
John P. Giesy a,b,c,d,e,g,h
a
The Swire Institute of Marine Science, The University of Hong Kong, Pokfulam, Hong Kong, China
School of Biological Sciences, The University of Hong Kong, Pokfulam, Hong Kong, China
Area of Excellence Centre for Marine Environmental Research and Innovative Technology (MERIT), The University of Hong Kong, Pokfulam, Hong Kong, China
d
State Key Laboratory in Marine Pollution, City University of Hong Kong, Tat Chee Avenue, Kowloon, Hong Kong, China
e
Department of Biology and Chemistry, City University of Hong Kong, Tat Chee Avenue, Kowloon, Hong Kong, China
f
Department of Chemistry and the National Research Lab of Marine Molecular and Environmental Bioscience, College of Natural Sciences, Hanyang University, Seoul 133-791,
South Korea
g
Department of Veterinary Biomedical Sciences, University of Saskatchewan, 44 Campus Drive, Saskatoon, SK, Canada S7N 5B3
h
Department of Zoology, and Center for Integrative Toxicology, Michigan State University, East Lansing, MI, USA
b
c
h i g h l i g h t s
" We reviewed the fate and toxicities of TPT compounds in marine environments.
" Health hazard of TPT to human via consuming contaminated seafood was highlighted.
" The current and future ecological risks of TPT to marine ecosystems were analyzed.
" Knowledge gaps and future research areas were identified and prioritized.
a r t i c l e
i n f o
Article history:
Received 19 March 2012
Received in revised form 10 May 2012
Accepted 17 May 2012
Available online 15 June 2012
Keywords:
TPT
Antifouling agents
Endocrine disruption chemicals
Bioaccumulation
Organotin
Marine pollution
a b s t r a c t
The state of scientific knowledge regarding analytical methods, environmental fate, ecotoxicity and ecological risk of triphenyltin (TPT) compounds in marine ecosystems as well as their exposure and health
hazard to humans was reviewed. Since the 1960s, TPT compounds have been commonly applied as biocides for diverse industrial and agricultural purposes. For instance, they are used as active ingredients in
antifouling systems on marine vessels and mariculture facilities, and as fungicides in agriculture. Due to
their intensive use, contamination of coastal waters by TPT and its products of transformation has
become a worldwide problem. The proportion of quantified TPT to total phenyltin compounds in the marine environment provides evidence that TPT is photodegradable in water and sediment but resistant to
biotransformation. Concentrations of TPT in marine biota are consistently greater than concentrations
in water and sediment, which implies potential of TPT to bioaccumulate. TPT is toxic to both marine
plants and animals. The predicted no effect concentration (PNEC) for TPT, as determined by use of the
species sensitivity distribution approach, is 0.64 ng L 1. In some parts of the world, concentrations of
TPT in seawater exceed the PNEC, indicating that TPT can pose risks to marine life. Although there is negligible risk of TPT to average human consumers, TPT has been detected in blood of Finnish people and the
concentration was greater in fishermen who ate more seafood. It is, therefore, advocated to initiate regular monitoring of TPT in blood and breast milk of populations that consume greater amounts of seafood.
Ó 2012 Elsevier Ltd. All rights reserved.
Contents
1.
2.
Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1016
Analytical methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1017
⇑ Corresponding author at: School of Biological Sciences, The University of Hong
Kong, Pokfulam, Hong Kong, China. Tel.: +852 22990607; fax: +852 25176082.
E-mail address: kmyleung@hku.hk (K.M.Y. Leung).
0045-6535/$ - see front matter Ó 2012 Elsevier Ltd. All rights reserved.
http://dx.doi.org/10.1016/j.chemosphere.2012.05.080
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A.X. Yi et al. / Chemosphere 89 (2012) 1015–1025
3.
Distribution and bioaccumulation of PTs in the marine ecosystem . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.1.
Distribution of TPT in the marine environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.1.1.
Concentrations of PTs in environmental compartments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.1.2.
Factors affecting distribution of PTs in the environment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.1.3.
Temporal variation in concentrations of PTs: implications on effectiveness of restrictions on the use of OTs . . . . . . . . . . . . .
3.2.
Proportions of TPT to PTs: relations with degradation and rates of biotransformation of TPT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.3.
Bioaccumulation of TPT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.3.1.
Factors affecting bioaccumulation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.3.2.
Bioconcentration of TPT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.3.3.
Biomagnification of TPT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
3.4.
Summary of environmental fate of TPT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Adverse effect of TPT to marine ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.1.
Endocrine disrupting effects on marine organisms . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.2.
Species sensitivity distribution for aquatic biota exposed to TPT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.2.1.
Toxicity test of TPT on marine organisms. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.2.2.
Predicted no effect concentration (PNEC) and evaluation of ecological risks of TPT . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Exposures and effects of TPT compounds to humans . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Acknowledgements . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Appendix A. Supplementary material . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
4.
5.
6.
1. Introduction
Organotin (OT) molecules are comprised of a tin (Sn) atom
which is covalently bound to one or more organic substituents,
such as methyl, ethyl, butyl or phenyl groups. Triphenyltin
(TPT) compounds are a group of OTs conforming to a general formula (C6H5)3Sn-X, where X is an anion or anionic group, such as
chloride, hydroxide and acetate. OTs are hydrophobic due to the
presence of hydrocarbon substituents and thus their solubility in
water is relatively small (Rüdel, 2003). In the pH range of 6–8,
the solubilities of TPT chloride (TPTCl) and TPT oxide (TPTO) in
water are approximately 1 mg L 1 at 25 °C (Inaba et al., 1995).
TPT compounds such as TPT acetate (TPTA) and TPTCl can be
hydrolyzed to TPTOH in the marine environment, and neutral
hydroxyl-complexes accounted for more than 93% of TPT
compounds in well-buffered seawater (Arnold et al., 1997; Veltman et al., 2006). The log KOW values of TPTs, such as TPTOH
and TPTCl are between 3 and 4 (Tsuda et al., 1990a; WHO,
1999), which suggests that these compounds are potentially
bioaccumulative.
TPTs are major ingredients of antifouling products and fungicides. OT-based antifouling products used on hulls of vessels and
submerged mariculture facilities are the main sources of TPT in
the marine environment. TPT compounds have also been widely
used in agriculture, as fungicides on potatoes, sugar beets, hops
and rice against fungal diseases, algae and molluscs (WHO,
1999). For example, in Minnesota, 10.73% of potato crops were
treated with TPTOH (Subramanyam, 1993). In Taiwan, 27% of surveyed farmers admitted illegal use of TPTA, even after its complete
prohibition in that country (Hu et al., 2009). Thus, agricultural runoff could also contribute to TPT in the marine environment. The
amount of total TPT produced annually in China is estimated to
be 200 tonnes (Hu et al., 2009), while in Japan, 140–160 tonnes
of TPT were produced for export annually during the period
1994–1996 (WHO, 1999).
As a result of their common industrial or agricultural applications, TPT compounds have entered various ecosystems, where
they interact with organisms. Studies of the distribution and the
effects of OTs in the marine environment have been conducted
since the late 1980s (Gibbs et al., 1987; Horiguchi et al., 1994;
Gomez-Ariza et al., 1998; Leung et al., 2001). OTs, especially trisubstituted molecules, are toxic to aquatic organisms (Fent,
1996a). Due to its effects on marine gastropods and oysters, appli-
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cation of OT-based antifoulants on boats of less than 25 m length
has been banned since the 1980s in countries including France,
the United Kingdom, the USA and Japan (Champ, 2000; Cao et al.,
2009). As initiated by the International Maritime Organization
(IMO), the Convention on the Control of Harmful Antifouling Systems on Ships (AFS Convention), which was adopted in 2001 and
has entered into force since September 2008, bans the use of OTs
as biocides in antifouling systems (IMO, 2001; IMO website:
http://www.imo.org). In order to facilitate ratification of the AFS
Convention by its member states, the European Community (EC)
published a specific regulation (EC No 782/2003) to ban the use
of harmful OTs in antifouling paints on ships in 2003 (EC, 2003).
However, it is likely that OT-based antifouling paints are still being
produced and used in some developing countries that are not
members of the IMO. Additionally, the continuing agricultural
use of TPT compounds as fungicides in Asia will remain a longterm concern.
To date, most information on the distribution and toxicity of
OTs is about tributyltin (TBT) and its products of transformation,
while little information on the occurrence of TPT compounds and
their derivatives. Thus, in comparison to TBT, there is a paucity of
information on the distribution, toxicity and accumulation of TPT
(Fent and Hunn, 1991; Tolosa et al., 1992; Shim et al., 2005a).
The only review of TPT, which was published by WHO, described
the toxicity of TPT to mammals and their potential toxicity to humans but did not adequately address the fate and ecological effects of TPT in the marine environment (WHO, 1999). This
review, which considered information published through September 2011, is the most comprehensive review on the state of scientific knowledge on TPT compounds and is provided as a basis
to assess current status and trends and potential risks posed by
the chemicals in the marine environment. In particular, the review is focused on measured concentrations of TPT in environmental compartments, such as water, sediment, and biota, their
fate in the marine environment, their toxicities and ecological
risks towards marine organisms, and their potential hazards to
human health via consumption of contaminated seafood. Secondary data on concentrations of TPTs and toxicities, which were
reported in the peer-reviewed literature, were used in a metaanalysis to determine current status and potential future trends
in exposures and possible effects. Finally, areas or issues which
require further research attention are prioritized and highlighted
in this review.
A.X. Yi et al. / Chemosphere 89 (2012) 1015–1025
2. Analytical methods
A variety of methods have been developed for identification and
quantification of OTs, and most of these methods share key steps in
the analytical procedures for chemical speciation of different compartments of the marine environment including water, sediments
and biota. These steps include sample preparation, extraction of
OTs, derivatisation, purification and application of appropriate
analytical techniques for identification and quantification (de Carvalho Oliveira and Santelli, 2010). The International Organization for
Standardization (ISO) has set up the standards for the determination of selected OTs including phenyltins (PTs) in water (ISO
17353, 2004) and soil samples (ISO/SIS 23161-2, 2007). Considering there have been comprehensive reviews of the analytical methods of OTs which summarized the analytical procedures in detail
(Abalos et al., 1997; de Carvalho Oliveira and Santelli, 2010), in this
section, only a brief introduction on the more widely adopted
methods is provided.
Collection and preparation of samples are important steps for
any chemical analysis. The location and time of sampling are
important since there can be spatial and temporal variability in
distributions of PTs. For example, the distance to potential sources
and seasons of sampling can result in variation in distribution of
PTs (see Section 3.1.2). After collection, to guarantee stability of
PTs before chemical analysis, samples should be stored in the dark
at 20 °C (Abalos et al., 1997). For both biotic and abiotic matrices,
the use of organic solvents of lesser to moderate polarity in combination with hydrochloric acid or acetic acid accounts for more than
50% of the extraction procedures (Abalos et al., 1997). Extraction
with tetramethylammonium hydroxide (Guðmundsdóttir et al.,
2011) or potassium hydroxide (Nagase et al., 1995), have also been
applied to decompose the biological matrices as an alternative to
acid leaching. However, there is no agreement on the most appropriate solvent to use for extraction. Extraction methods have been
developed to isolate and concentrate analytes from both biotic and
abiotic matrix including solid phase extraction (SPE), supercritical
fluid extraction (SFE) and solid-phase microextraction (SPME).
After extraction, derivatisation is used to produce volatile OTs that
can be separated by use of gas chromatography (GC). Derivatisation methods are based on alkylation or hydridization. Sodium tetraethylborate (NaBEt4) is widely applied for ethylation (Ashby and
Craig, 1991), and sodium borohydride (NaBH4) is adopted for hydridization (Abd-Allah, 1995). A clean-up step is crucial for removal
of matrix components, such as lipids, proteins, sulfur and high boiling point compounds that might interfere with the recovery, identification or quantification of OTs (de Carvalho Oliveira and
Santelli, 2010). This step is usually conducted after derivatisation
(Stewart and Thompson, 1994), and most commonly used adsorbents are silica, alumina, and florisil.
For identification and quantification of PTs, the most commonly
used technique is gas chromatography (GC) because this method of
separation can be coupled with an element-specific detection
method, such as mass spectrometry (GC/MS) (Ishizaka et al.,
1989; Ashby and Craig, 1991; Arnold et al., 1998), atomic absorption spectrometry (GC/AAS) (Tas et al., 1991; Ceulemans et al.,
1998), flame photometric detection (GC/FPD) (Carlier-Pinasseau
et al., 1997), microwave-induced or inductively-coupled plasma
atomic emission spectrometry (GC/MIP-AES or GC/ICP-AES) (Tutschku et al., 1994; Becker et al., 1997). Separation by liquid chromatography (LC) has also been employed as the separation step
in analyses of OTs. Derivatisation is not required when LC is used,
which minimizes both time and cost. Compared to GC, nevertheless, LC is less sensitive because of its poorer power to separate
chemicals (Rosenberg et al., 2000), thus limited numbers of chemical can be identified in a single analysis (Garcia-Alonso et al.,
1993). Because it is more sensitive than other commonly used
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techniques, ICP-MS is being used more as the detector for identification and quantification of PTs (Bianchi et al., 2006).
Apparently, more concerted efforts are needed to harmonize
and standardize the analytical methods for quantifying various
PTs in environmental samples. To advance this process, more inter-laboratory calibration studies (e.g., the European QUASIMEME
laboratory performance studies; http://www.quasimeme.org/)
should be fostered and conducted in different geographical
regions.
3. Distribution and bioaccumulation of PTs in the marine
ecosystem
There is little information on concentrations of PTs in marine
ecosystems. Concentrations of PTs in marine waters, sediments,
and biota have been assembled in the Supplementary materials
(i.e., Appendix A: Tables S1–S5). Whenever concentrations were
less than the limit of detection (DL), a surrogate value of half of
the DL was reported (USEPA, 1998; Helsel, 2006). Concentrations
were then compared among regions and locations by use of
standard statistical software, PASW 18 (SPSS Inc., Chicago, IL). To
facilitate comparisons among reported concentrations all values
were expressed dry weight (dw). Concentrations expressed wet
weight (ww) were converted to dw by use of the following dw/
ww proportions: 15% for invertebrate (Phillips, 1985), 20% for fish
(CRESP, 2006) and 33.3% for livers in bird/mammal (Scanlon, 1982;
Yang and Miyazaki, 2003). Concentrations of TPTs and their chemodynamics are discussed simultaneously.
3.1. Distribution of TPT in the marine environment
3.1.1. Concentrations of PTs in environmental compartments
Mean concentrations of PTs in various environmental compartments at different locations are illustrated (Fig. 1). Concentrations
of PTs in seawater are often near to, or less than the DL of currently
available instrumental analytical methods (Table S1). PTs can only
be detected at more contaminated locations, such as harbors or
marinas (Nemanič et al., 2002). Concentrations of TPT in sediments
are greater than in seawater, with the greatest concentration of
40 ng Sn g 1 dw occurring in Japan (Harino et al., 1998). TPT, which
has a relatively great sediment–water partition coefficient, with reported values ranging from 2.1 104 to 1.1 105 L kg 1 (Fent,
1996a), has a relatively strong affinity for sediments and suspended particulates. Therefore, TPTs become deposited into benthic sediments, which are the primary sink for TPT in the marine
environment.
PTs have been detected in invertebrates, fishes, birds and mammals from various locations (Tables S2–S4). The greatest concentration of TPT, 5.9 lg g 1 dw, was measured in the blue mussel
(Mytilus edulis) (Higashiyama et al., 1991). Concentrations of TPT
generally increase in the order: water < sediment < invertebrate
and fish (in Japan and south China), whereas in Japan and Western
Europe concentrations of TPT in birds and mammals were less than
those in fishes or invertebrates (Fig. 1). This might be due to variable patterns of bioaccumulation of TPT (Section 3.3) among organisms in marine food webs. Concentrations of TPT were compared
among invertebrates, fishes, birds and mammals, by use of the
non-parametric Kruskal–Wallis test. However, no significant differences among these three groups was observed (v2 = 2.314,
p > 0.05). This is due to the relatively large variance within groups
due to the wide range of locations and times included in the analysis. Thus, it is difficult to stratify and/or aggregate the available
data in a way to make meaningful comparisons among these
groups.
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A.X. Yi et al. / Chemosphere 89 (2012) 1015–1025
Fig. 1. Distribution of TPT in compartments of the marine environment. To allow for better visualization of a range of concentrations, mean TPT concentrations in each area
are expressed as Log10[mean conc.] + 2.
3.1.2. Factors affecting distribution of PTs in the environment
Distributions of OTs such as BTs and PTs in marine environments depend on several factors, among which distance from potential sources is of the primary importance (Minchin et al.,
1995; Leung et al., 2006; Guðmundsdóttir et al., 2011). Shipping
activities are the most intense in coastal areas such as harbors,
estuaries or marinas, thus these areas have been regarded as the
most problematic among sources of OTs. There is a general trend
of decreasing concentrations of OTs in organisms with distance
from sources (Evans, 1999), and their concentrations in pelagic
species are greater than those of demersal species (Rumengan
et al., 2008). Concentrations of OTs in deep-sea areas were poorly
studied because it had been deemed unlikely for these pollutants
to be transported to such depths. However, Borghi and Porte
(2002) conducted a survey around the northwest Mediterranean
Sea and for the first time, reported the maximum concentration
of TPT, as great as 1.4 lg Sn g 1 ww, in deep-sea fish. This observation of TPT in deep-sea fish indicates that TPT persists in the marine environment and can be transported to greater depths in the
water column.
Concentrations of PTs in the marine environment can vary
among seasons (Lee et al., 2006). In contrast to TBT, concentrations
of which were reported to be greater in winter and lesser in summer (Rivaro et al., 1997), concentrations of TPT in organisms have
been found to be greater in summer (Hung et al., 1998, 2001; Lee
et al., 2005). Because antifouling paints are not the only source of
PTs, this seasonal trend of PTs might be due to their use in aquaculture nets and application of PT-containing biocides in both mariculture and agriculture during the summer (Hung et al., 1998;
Meng et al., 2009).
PTs are accumulated differentially among tissues and organs of
marine organisms. In fish, liver is prone to accumulate greater concentrations of TPT than are other organs such as gills or digestive
glands (Morcillo et al., 1997; Harino et al., 2000). A similar trend
has been observed in marine mammals, where the greatest con-
centration of total OTs was found in liver (Berge et al., 2004). Because of the greater affinities of OTs to sulfhydryl groups such as
glutathione, the pattern of distribution among tissues or organs
is dependent on contents of protein, rather than lipids (Iwata
et al., 1997; Strand and Jacobsen, 2005; Hu et al., 2009). In addition
to the factors mentioned above, species-specific characters might
also influence the distribution of the PTs and thus contribute to
more variation among species, even at the same trophic level. For
instance, concentrations of PTs and BTs are lesser in marine mammals that are covered with fur, because OTs tend to be bound to
sulfhydryl groups in hair, which in turn can be shed (Tanabe,
1999; Berge et al., 2004). Maternal transfer of PTs is another possible mechanism to reduce the tissue burden of PTs in the mother,
which might also contribute to different patterns of distributions
among tissues. This phenomenon has been reported to occur in
Dall’s porpoises (Phocoenoides dalli) (Yang et al., 2007), the Japanese medaka (Oryzias lapipes) (Zhang et al., 2008) and Chinese sturgeons (Acipenser sinensis) (Hu et al., 2009).
3.1.3. Temporal variation in concentrations of PTs: implications on
effectiveness of restrictions on the use of OTs
In order to evaluate the effectiveness of restrictions of use, a
long-term monitoring of OTs concentrations was conducted in various developed countries. As expected, after the restriction of their
use in antifouling products, concentrations of OTs have been
decreasing in estuarine areas of Japan (Port of Osaka) (Harino
et al., 1999), the United Kingdom (Irish Sea) (Oliveira et al.,
2009), South Korea (East Sea and Yellow Sea) (Choi et al., 2009),
Iceland (Guðmundsdóttir et al., 2011), Northern Europe (North
Sea and Baltic Sea) (Rüdel et al., 2003) and Southern Europe (Sousa
et al., 2007). Thus, restrictions in use have been at least partially
success in reducing concentrations of OTs the coastal marine environment. Nonetheless, exceptions to this general decrease in concentrations of OTs have been reported for other locations. For
example, concentrations of OTs, in particular PTs have not been
A.X. Yi et al. / Chemosphere 89 (2012) 1015–1025
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et al., 2009). Based on the available results, it can be concluded that
the rate of biotransformation of TPT is relatively slow.
3.3. Bioaccumulation of TPT
Fig. 2. Ratios of triphenyltin (TPT) to total phenyltins (PT) (%) in compartments of
the marine environment. DPT: diphenyltin. MPT: monophenyltin.
decreasing in Hong Kong (Nakayama et al., 2009; Kevin Ho, unpublished data). These exceptions seem to occur in busy shipping areas
with fresh inputs of OTs, probably from the continued illegal uses
of OTs, and from uses of OT-based antifouling systems on ships and
mariculture facilities in countries which did not impose the IMO’s
regulation on their uses.
3.2. Proportions of TPT to PTs: relations with degradation and rates of
biotransformation of TPT
The ratio between concentrations of TPT and total PT can be
used as an approximate measure of transformation of TPT by
chemical or biological means. A greater proportion of TPT relative
to that of total PTs indicates a lesser rate of transformation of TPT
in the environment or a lesser rate of biotransformation of TPT.
Thus, ratios of concentrations of TPT, Di-PT (DPT) and Mono-PT
(MPT) to concentrations of total PTs were calculated for different
compartments among various marine environments (Fig. 2). TPT
constitutes only 15.7% of total PTs in seawater, while percentages
of DPT and MPT are 31.4% and 52.9%, respectively. These ratios
are consistent with measured rates of photolysis of TPT, whereby
DPT and MPT and some water-soluble, non-extractable OT polymers are formed (Soderquist and Crosby, 1980). Similar distributions were observed in sediments, in which proportions of TPT,
DPT and MPT were 12.5%, 28.1% and 59.4%, respectively. Degradation of TPT occurs relatively slowly in sediments, and photodegradation, rather than microbial degradation, is the primary pathway
of degradation of TPT in sediments (Fent et al., 1991a,b; Kannan
and Lee, 1996). The smaller proportion of TPT, relative to that of total PTs in sediments might also be attributable to a decreasing input of TPT in areas that were studied (Midorikawa et al., 2004).
Unlike abiotic compartments, TPT is consistently the dominant
compound in marine invertebrates and fish, with TPT contributing
63.4% and 65.8% of total PTs in invertebrates and fishes, respectively (Fig. 2). This is consistent with TPT being poorly biotransformed. For instance, during a 96 h exposure of larvae of the
minnow Phoxinus phoxinus biotransformation of TPT was not observed (Fent et al., 1991a,b). Only 2.5% of TPT was biotransformed
during 14 d in the goldfish Carassius auratus (Tsuda et al., 1988).
The half-life of TPT in the guppy Poecilia reticulata was at least
48 d (Tas et al., 1991). The rate of biotransformation of TPT is less
than that of TBT in some organisms, such as the mussels M. edulis
and M. coruscus and oyster Crassostrea gigas (Higashiyama et al.,
1991; Shim et al., 2005b).
TPT is also the predominant PT in birds and marine mammals,
with a slightly lesser ratio of 49.0%. Although there are no reports
on the rate of in vivo biotransformation of TPT in marine birds or
mammals, a recent study that used hepatic microsomes from Dall’s
porpoises found that little (<3%) TPT was biotransformed (Yang
3.3.1. Factors affecting bioaccumulation
Bioaccumulation and biomagnification of persistent organic pollutants (POPs), including organochlorine pesticides, such as DDT and
its transformation products, methylmercury and polychlorinated
biphenyls (PCBs) and OTs, can affect the resultant toxic effects of
these chemicals to aquatic organisms (van der Oost et al., 2003). In
general, xenobiotics that can be biomagnified have the following
properties: (1) a large octanol–water partition coefficient (KOW);
(2) persistence and (3) a slow rate of elimination (Clarkson, 1995).
Among these properties, the log KOW is of primary importance in
determining the bioaccumulation of a chemical where a greater log KOW value generally indicates a greater accumulative potential. The
log KOW values vary among different TPTs because of various anions/
anionic groups that are combined with the Sn atom. Generally
speaking, the log KOW values of TPTs range from 3.0 to 5.0, and
TPTOH, which is the most abundant PT species in marine environment, has a log KOW value of 3.5 (Arnold et al., 1997).
Bioaccumulation of chemicals depends not only on hydrophobicity, but also physicochemical parameters, such as salinity, presence
of humic acid and pH that can alter speciation, thereby affecting bioavailability of chemicals. For instance, bioconcentration factors
(BCFs) of both TBT and TPT for the freshwater guppy (Lebistes reticulatus) are approximately twice as great as those for marine fishes
(Tsuda et al., 1990a), which is indicative of a negative effect of
salinity on accumulation of TBT and TPT. Bioaccumulation of TBT
by marine mussels is inversely proportional to the concentration
of humic materials in the water (Laughlin et al., 1986). In contrast,
accumulation of OTs by aquatic organisms is directly proportional
to pH. Bioaccumulation of both TBT and TPT was greater for both
invertebrates and fishes at pH 7.8 than at pH 6.8 or 6.0 (Tsuda
et al., 1990b). Apart from these abiotic factors, biotic factors, such
as life stage of the organism and species-specific characteristics,
can potentially affect the bioaccumulation of TPTs (Fent, 1996a).
3.3.2. Bioconcentration of TPT
Accumulation of TPT from water/sediments has been studied in
algae, invertebrates and fishes. A BCF of 1.1 105 was reported for
the alga Scenedesmus obliquus after 96 h exposure to TPT (Huang
et al., 1993). Marine invertebrates can accumulate OTs quickly
and the BCFs for accumulation of TPT by marine mussels Mytilus
graynus and M. edulis were 4.3 104 and 3.6 104, respectively
(Suzuki et al., 1998). Bioconcentration of TPT has been determined
for several fishes, including the guppy (Poecilia reticulate), rainbow
trout (Salmo gairdneri), common carp (Cyprinus carpio), red sea
bream (Pagrus major), mullet (Mugil cephalus), filefish (Rudarius ercodes) and minnow (Phoxinus phoxinu); among which the greatest
BCF was 4.1 103, which was observed for R. ercodes (Fent et al.,
1991a,b; Tsuda et al., 1987; Tas et al., 1989; Yamada and Takayanagi, 1992).
3.3.3. Biomagnification of TPT
Little is known about biomagnification of TPT. Studies of trophic
transfer of TPT have rarely been conducted under controlled laboratory conditions. Most studies of trophic transfer of TPT have been
conducted in the field and because it is impossible to know the exact nature of exposure or trophic relationships, only apparent biomagnification factors (BMFs) can be estimated. There has been
only one documented study of biomagnification conducted in a controlled laboratory exposure of marine organisms to TPT. The BMF of
TPT was estimated to be 0.57, which is approximately twice as great
as that of TBT (Yamada et al., 1994). Since the BMF of TPT is less than
1020
A.X. Yi et al. / Chemosphere 89 (2012) 1015–1025
Fig. 3. Environmental fate of triphenyltin (TPT) in the marine ecosystem. DPT: diphenyltin. MPT: monophenyltin.
1.0, it is assumed that TPT would not be biomagnified. Based on the
available information collected during field investigations, however, it appears that TPT is biomagnified in marine food webs
(Stäb et al., 1996; Kono et al., 2008). The BMF for TPT was estimated
to be 3.7 in Bohai Bay, China (Hu et al., 2006) and from 2.2 to 5.4 in
coastal areas of Japan (Murai et al., 2008), while BMF values of TBT
in both regions were less than 1.0. Neither of these studies determined concentrations of TPT in organisms at the highest trophic level such as marine mammals or fish-eating birds. There is only one
study of accumulation of TPT along a more comprehensive food
web. In that study conducted in Denmark, species from seaweed
to marine mammals were investigated. Concentrations of TPT in
fish-eating birds and marine mammals were less than in the fishes
consumed by these predators. The authors of that report suggested
that the fish-eating birds and marine mammals could likely biotransform and excrete TPT. Thus, TPT is unlikely to be accumulated
to these higher trophic levels (Strand and Jacobsen, 2005).
3.4. Summary of environmental fate of TPT
Primary sources of TPT in the marine environment are releases
from antifouling products used in boats or mariculture facilities,
and agricultural runoff, although other potential sources include
treated or untreated sewage effluent discharges, and surface runoff
via storm drainage discharges (Fig. 3; Fent, 1996b; Arnold et al.,
1998). TPT can be enriched in sediments while it is photolized in
both surface seawater and sediments in shallow water. Marine
organisms can accumulate TPT directly from seawater or sediments, as well as through their diet. It has been confirmed that
TPT can be biomagnified from phytoplankton and zooplankton to
fish. But more information will be needed to confirm the details
of biomagnification to higher trophic levels, such as marine mammals and fish-eating birds.
4. Adverse effect of TPT to marine ecosystems
4.1. Endocrine disrupting effects on marine organisms
OTs have been implicated as endocrine disruptors primarily
through studies of aquatic organisms, among which imposex of
marine gastropods is the most well-defined endocrinal effect
caused by OTs. Therefore, monitoring for incidence of imposex in
neogastropod species has become a common method to monitor
for effects of OTs in coastal marine environment worldwide. Imposex is the development of vas deferens and penis in females (Gibbs
et al., 1997), which can lead to reproductive failure, and subsequent decreases in populations where imposex occurs (Gibbs
et al., 1987, 1988). Imposex was first associated with TBT in the
early 1980s, and since then it has been reported to occur in at least
195 species of gastropods (Lima et al., 2011), such as Nucella lapillus, Nassarius reticulates, Thais bronni and Thais clavigera (Gibbs and
Bryan, 1986; Bryan et al., 1988; Horiguchi et al., 1995; Barreiro
et al., 2001). Although TPT occurred with TBT in coastal areas
and TBT was known to cause imposex, it was not until the mid1990 when Japanese scientists confirmed that TPT could also cause
imposex in T. clavigera (Horiguchi et al., 1994, 1996). More recently
the occurrence of imposex in marine gastropods collected from
coastal areas of different countries has been associated with TPT
(Japan: Horiguchi et al., 1994; Spain: Solé et al., 1998; Korea: Shim
et al., 2000). Positive relationships between concentrations of TPT
and rates of imposex have been reported since the mid-1990. Results of a study conducted in the laboratory where TPT was injected into T. clavigera demonstrated that TPT does cause
imposex (Horiguchi et al., 1996). A similar result was subsequently
observed for Bolinus brandaris (Santos et al., 2006). However, due to
species-specific properties, this phenomenon is not necessarily observed in all gastropods. For example, when TPT was injected into
N. lapillus, it did not cause imposex, even at concentrations as great
as 1 lg TPTCl g 1 body weight (Bryan et al., 1988). Similarly, TBT
was also reported to not cause imposex in the dove snail, Columbella rustica (Gibbs et al., 1997).
Possible mechanisms by which OTs cause imposex in gastropods have been described by Horiguchi (2009). There are four previously proposed pathways, including: (1) increasing androgen
content through inhibition of cytochrome P450-dependent aromatase (Bettin et al., 1996); (2) inhibition of testosterone excretion
(Ronis and Mason, 1996); (3) interference with the neuroendocrine
system (Féral and LeGall, 1983), and (4) abnormal release of the
neuropeptide, APGWamide (Oberdörster and McClellan-Green,
2000). However, none of these hypotheses adequately explains
A.X. Yi et al. / Chemosphere 89 (2012) 1015–1025
onset of imposex, and thus the mechanism of OT-induced imposex
is still not completely elucidated (Horiguchi, 2009). Recently, a
hypothesis regarding the mechanisms by which OT induces imposex has been proposed by Nishikawa et al. (2004). Those authors
demonstrated that OT could directly bind to a retinoid X receptor
(RXR) and act as agonists, thereby inducing imposex in gastropods.
RXR has been shown to be involved in induction of male-type genitalia in female gastropods (Horiguchi et al., 2007; Horiguchi,
2009).
TPT-induced effects on the endocrine system have also been reported in fish. For instance, TPT can affect reproduction by suppressing spawning frequency and reducing the number of eggs
produced by female medaka, Oryzias latipes (Zhang et al., 2008),
or by inhibiting testicular development in male rockfish Sebastiscus
marmoratus (Sun et al., 2011); TPT’s binding to RXR can also cause
deformities of the eye in the Chinese sturgeon Acipenser sinenses
(Hu et al., 2009). In general, effects of OTs on the endocrine system
of fish are initiated by inhibition of enzymes like cytochrome P450
and production of the egg yolk phospholipoprotein vitellogenin
(Hinfray et al., 2006; Miller et al., 2007). Apart from gastropods
and fish, TPT have been reported to induce malformation in embryos of amphibians such as the African clawed frog (Xenopus tropicalis) (Yuan et al., 2011). Crustaceans form a ubiquitous group of
invertebrates, and many of them have been chosen as model species in aquatic toxicity tests, including the water flea (Daphnia
magna), the mysid Siriella armata, and the copepod Tigriopus japonicus. TPT and other OT compounds can have adverse effects on
molting, growth and reproduction of crustaceans (Rodríguez
et al., 2007). Effects of OTs on sexual differentiation process of
crustaceans have been neglected, which may deserve further research to fill the knowledge gaps.
4.2. Species sensitivity distribution for aquatic biota exposed to TPT
4.2.1. Toxicity test of TPT on marine organisms
OTs, in particular the tri-substituted compounds are more toxic
to aquatic organisms than are inorganic Sn compounds (Fent,
1996a; Hoch, 2001). The toxic effects of OTs, primarily TBTs, have
been of concern and have been widely studied during the past
few decades. Nevertheless, much of the information is for TBTs
(Hall and Pinkney, 1985). Information on toxicity of TPTs is available for only a few species of marine algae, invertebrate and fishes
(Table S5). For algae, inhibition of growth has been the most studied measurement endpoint. Median effect concentrations, EC50 values, for algae growth test ranged between 0.59 and 51.5 lg L 1
(Table S5). Besides inhibition of growth, other indices including
chlorophyll or phycocyanin content, photosynthesis inhibition
and nitrate reductase have also been applied to study the toxic effect of TPT on algal species (Mooney and Patching, 1995; Huang
et al., 1996). Among these indices, nitrate reductase was the most
sensitive to the effects of TPT (Huang et al., 2002). Sensitivity of
invertebrates to TPT varied among species’ median lethal concentrations (LC50) or EC50, ranging from 0.8 to more than
560 000 lg L 1. Among species of invertebrates tested, the copepod
Acanthocyclops venustus was the most sensitive to TPT while the bivalve Mercenaria mercenaria was the most tolerant (Table S5). Most
data on the toxicity of TPT to fishes have been obtained for larvae
or juveniles (Fent and Meier, 1992, 1994). LC50 values for fishes
ranges from 7.1 to 62 lg L 1. In addition, TPT compounds can affect
feeding and swimming behavior of tadpoles of the European frog,
(Rana esculenta; synonym of Pelophylax kl. esculentus) (Semlitsch
et al., 1995).
To determine the risks of chemicals it is useful to determine the
mechanism of action prior to conducting testing whole individuals
(Fent, 1996a). Uptake of OTs by microorganisms is primarily
through adsorption of organic moieties of OTs onto the cell surface
1021
(Avery et al., 1993). Triorganotin compounds such as TBT and TPT
exert their toxicities through interaction with membrane lipids
(Mooney and Patching, 1995), thereby disrupting respiration and
photosynthesis of algae. Toxicity of OTs is related to total surface
area of microorganisms such as unicellular algae, and the lipid solubility of OT chemicals (Cooney and Wuertz, 1989).
Toxicity of TPT in vitro has been determined. Toxicity of TPT to
PLHC-1 fish hepatoma cells was correlated with in vivo toxicity of
mortality in fish (Brüschweiler et al., 1995). TPTs can interact with
different intracellular enzymes. For example, cytochrome P450 and
glutathione-S-transferase are both involved in biotransformation
of xenobiotics, but TPTs can inhibit their expression, thus leading
to dysfunction of the detoxicification system (George and Buchanan, 1990; Fent and Bucheli, 1994). Also, TPT compounds increase
intracellular calcium storage by slowing mobilization of Ca2+,
which in turn causes hyperpolarization of cell membranes (Miura
and Matsui, 1991).
TBT and TPT are equally potencies to most species that have
been tested. Relatively few toxicological studies, especially full
life-cycle chronic toxicity tests have been conducted for TPT and
little is known about the mechanism of action. Toxicities of OTs
are thought to be additive (WHO, 1999), but there is virtually no
relevant, specifically designed study to investigate and elucidate
the interactions among different OT compounds.
4.2.2. Predicted no effect concentration (PNEC) and evaluation of
ecological risks of TPT
Acute toxicity of TPT to 15 marine species of different trophic
levels (six algal species, two arthropod species, four mollusc species and three fish species) were obtained from peer-reviewed literature or the ECOTOX database (USEPA, 2011a), with species
geometric means calculated if multiple studies had been conducted for the same species. The average toxicity values for the
15 species were ranked and used to construct a species sensitivity
distribution (SSD) according to the method described in Wheeler
et al. (2002). From the SSD analysis (Fig. 4) which was performed
on log-transformed data by use of an SSD generator (downloaded
from USEPA; USEPA, 2011b), an HC5 (hazardous concentration to
5% of species or protective of 95% of species) value for TPT was
determined to be 0.117 lg Sn L 1 (95% CI: 0.064–0.216 lg Sn L 1).
By applying an acute to chronic ratio of 10 and an assessment factor of 10 to the lower 95% CI of HC5, which is 0.064 lg Sn L 1
(Chapman et al., 1998), a chronic predicted no effect concentration
(PNEC) was estimated to be 0.64 ng Sn L 1. Larvae of the oyster
(Crassostrea virginica) were the most sensitive to TPTOH with an
LC50 value of 0.32 lg L 1 (or 0.10 lg Sn L 1). The Hazard Quotient
(HQ) approach was applied to estimate the ecological risks posed
by TPT. Based on values of HQs, which are defined as the quotient
of a measured concentration in the environment (MEC) divided by
the PNEC, TPT might pose risks to marine organisms in coastal
environments of Japan, Korea, and the Mediterranean Sea (Table 1).
Nevertheless, it is worth noting that most of these data are not recent, thus it is not reliable to predict the current ecological risks
caused by TPT using these data. Therefore, more current information for TPT distributions is needed to evaluate its current environmental risks.
5. Exposures and effects of TPT compounds to humans
Humans are exposed to multiple chemicals in the environment
and their diet, including OTs and thus the potential health risks of
OTs to humans should be considered (Golub and Doherty, 2004).
These sources of OTs are various and widespread. For instance,
food containers made of PVC polymers (Kannan et al., 1992), seafood from markets (Guérin et al., 2007) and even tap water that
1022
A.X. Yi et al. / Chemosphere 89 (2012) 1015–1025
Proportion of species affected
1.0
Central tendency
95% CI
0.8
R2 = 0.962
0.6
0.4
Crassostrea virginica
0.2
0.0
0.01
HC5 = 0.117 µg L-1
(95% CI: 0.064 – 0.216 µg L-1)
0.1
1
10
100
TPT conc. (µg/L as Sn)
Fig. 4. Species sensitivity distribution (SSD) of TPT to marine species.
Table 1
Hazard quotient (HQ) in different countries or areas. MEC: measured environmental
concentration.
Area
Year
MEC
(ng Sn L
Japan
1996
5.25
Korea
1997–
1998
2006
1.2
1988
19.5
China
France and Spain
(Mediterranean Sea)
0.09
Reference
1
HQ
)
Harino et al.
(1998)
Shim et al.
(2005a,b)
Wang et al.
(2008)
Alzieu et al.
(1991)
8.2
1.9
0.14
30.5
passes through PVC pipes (Sadiki et al., 1996) are sources of OTs.
Food, especially seafood and fishery products are considered the
primary sources of OTs (Guérin et al., 2007) to humans. In addition
to oral uptake from contaminated foodstuffs, cutaneous absorption
via the respiratory tract is likely to happen and should not be neglected (Colosio et al., 1991). There have been two reports related
to the adverse effects of TPTs on humans (Manzo et al., 1981; Colosio et al., 1991) via accidental exposure to TPT-based pesticides by
farmers. These patients exhibited similar symptoms of TPT poisoning, including dizziness and nausea. TPT may also lead to possible
impairment of the central nervous system and liver damage.
Since no quantitative toxicological data on humans is available,
potential toxic effects on humans can only be extrapolated from results of tests on other mammals like rats, rabbits or pigs and the
estimated human intake of these chemicals. A diet of 50 mg TPTOH
kg 1 of diet did not have adverse effect on rats even after 276 d of
the continuous dietary exposure (Kimbrough, 1976). The guineapig has been reported to be the most sensitive species and its
growth inhibition was observed at concentration as little as 1 mg
TPTA kg 1 of diet (Stoner, 1966). TPTs are slowly eliminated from
rats and guinea pigs (Stoner, 1966; Verschuuren et al., 1966). Similar as in marine organisms, aromatase may be a toxicological target of TPT in mammals as well. For example, in rats, TPT exposure
can affect brain and gonadal aromatase activity in a sex-dependent
fashion (Hobler et al., 2010). Immunotoxicity is considered to be
the most sensitive, or critical endpoint for exposure of mammals
to TPT (Boyer, 1989). Based on immunological responses of experimental animals, the acceptable daily intake (ADI) of TPT compounds for humans has been set at 0.5 lg kg 1 bw d 1 (Lu, 1994;
WHO, 1999). Considering that TBT, dibutyltin (DBT), dioctyltin
(DOT) and TPT have similar modes of immunotoxic effects to
organisms, it is more reasonable to derive an ADI for this whole
group of chemicals (Guérin et al., 2007). Based on an assumption
that toxic effects of these compounds are additive, the European
Food Safety Authority (EFSA) established the tolerable daily intake
(TDI) value of this group of OTs to be 0.25 lg kg 1 bw d 1 (EFSA,
2004).
There have been reports on concentrations of OTs in food and its
implications for public health risks. The results of these risk assessments published before 2000 have been summarized (Belfroid
et al., 1999). Over the last decade, studies of concentrations of
TPT in seafood have been conducted in China (Cao et al., 2009), Finland (Rantakokko et al., 2006), France (Guérin et al., 2007), Japan
(Ueno et al., 1999) and Portugal (Santos et al., 2009), respectively.
In some of these recent studies, TPTs were detected in seafood at
considerable concentrations (Ueno et al., 1999; Guérin et al.,
2007). Most of these surveys on OTs in foodstuffs found that the
daily uptake from food were less than the ADI or TDI values, and
therefore suggested negligible risks to average consumers. However, there were limitations of these surveys. For example, OTs
can also originate from other market products other than seafood,
such as potatoes, fruits and vegetables (Rantakokko et al., 2006).
Furthermore, dietary intakes of OTs among children of less than
average bodyweights have not been estimated. Daily intake can
also be significantly greater for those who consume large amounts
of contaminated food. OTs have been detected in human blood and
liver. Butyltins (BTs) were first found in human blood in late 1990s
(Kannan et al., 1999) and their residues in human livers were detected in samples collected from the Japanese and Polish people
(Kannan and Falandysz, 1997; Takahashi et al., 1999). Average concentrations of BTs in livers of Polish people (2.4–11 ng g 1 wet wt)
were less than those in Japanese (59–96 ng g 1 wet wt). TPT has
been detected in blood of Finnish people. Concentrations of TPT
in blood of humans were greater in people who consumed greater
amounts of seafood (Rantakokko et al., 2008). Therefore, from the
viewpoint of public health, it is rational to devote effort to further
reducing the OTs in the environment. It is also advocated to initiate
monitoring of TPT and its degradants in human blood and breast
milk samples, especially in the population with greater seafood
consumption such as fishermen and people living nearby the coast.
Such monitoring schemes can be easily integrated with the existing programmes for monitoring POPs.
6. Summary
Due to wide-spread application in antifouling products and
agriculture, TPT compounds can be detected in various compartments of the marine environment. TPT can undergo photolysis in
water and sediment but is not easily biotransformed by marine
organisms. TPT can be bioaccumulated and potentially biomagnified through the marine food web of lower trophic levels, such as
from phytoplankton and zooplankton to fishes. TPT is toxic to marine organisms, with the predicted no effect concentration (PNEC)
for chronic, continuous exposure of 0.64 ng Sn L 1. The PNEC was
derived by use of the species sensitivity distribution approach.
Based on this PNEC and concentrations measured in the marine
environment, TPT was determined to pose risks to marine organisms in some regions. Current concentrations of TPT pose a di minimus risk to humans, but there are significant uncertainties in the
assessment of risks to humans due to limitations in both estimates
of exposure and predicted thresholds for effects this assessment.
Acknowledgements
This study was substantially supported by the Area of Excellence Scheme under the University Grants Committee of Hong
A.X. Yi et al. / Chemosphere 89 (2012) 1015–1025
Kong SAR China (Project No. AoE/P-04/2004) and by the Research
Grants Council through a General Research Fund (Project No.
HKU 7034/07P). The research was also supported, in part, by a Discovery Grant from the Natural Science and Engineering Research
Council of Canada (Project # 326415-07). Prof. Giesy was supported by the Canada Research Chair program, an at large Chair
Professorship at the Department of Biology and Chemistry and
State Key Laboratory for Marine Pollution, City University of Hong
Kong, The Einstein Professor Program of the Chinese Academy of
Sciences and the Visiting Professor Program of King Saud
University.
Appendix A. Supplementary material
Supplementary data associated with this article can be found, in
the online version, at http://dx.doi.org/10.1016/j.chemosphere.
2012.05.080.
References
Abalos, M., Bayona, J.-M., Compañó, R., Granados, M., Leal, C., Prat, M.-D., 1997.
Analytical procedures for the determination of organotin compounds in
sediment and biota: a critical review. J. Chromatogr., A 788, 1–49.
Abd-Allah, A.M.A., 1995. Occurrence of organotin compounds in water and biota
from Alexandria harbors. Chemosphere 30, 707–715.
Alzieu, C., Michel, P., Tolosa, I., Bacci, E., Mee, L.D., Readman, J.W., 1991. Organotin
compounds in the Mediterranean: a continuing case for concern. Mar. Environ.
Res. 32, 261–270.
Arnold, C.G., Weidenhaupt, A., David, M.M., Müller, S.R., Haderlein, S.B.,
Schwarzenbach, R.P., 1997. Aqueous speciation and l-octanol-water
partitioning of tributyl- and triphenyltin: effect of pH and ion composition.
Environ. Sci. Technol. 31, 2596–2602.
Arnold, C.G., Berg, M., Müller, S.R., Dommann, U., Schwarzenbach, R.P., 1998.
Determination of organotin compounds in water, sediments and sewage sludge
using perdeuterated internal standards, accelerated solvent extraction and
large-volume-injection GC/MS. Anal. Chem. 70, 3094–3101.
Ashby, J.R., Craig, P.J., 1991. Speciation for analysis of organotin compounds by GC
AA and GC MS after ethylation by sodium tetraethylborate. Appl. Organomet.
Chem. 5, 173–181.
Avery, S.V., Codd, G.A., Gadd, G.M., 1993. Biosorption of tributyltin and other
organotin compounds by cyanobacteria and microalgae. Appl. Microbiol.
Biotechnol. 39, 812–817.
Barreiro, R., González, R., Quintela, M., Ruiz, J.M., 2001. Imposex, organotin
bioaccumulation and sterility of female Nassarius reticulatus in polluted areas
of NW Spain. Mar. Ecol. Prog. Ser. 218, 203–212.
Becker, G., Janak, K., Colmsjö, A., Östman, C., 1997. Speciation of organotin
compounds released from poly(vinyl chloride) at increased temperatures by
gas chromatography with atomic emission detection. J. Chromatogr., A 775,
295–306.
Belfroid, A.C., Purperhart, M., Ariese, F., 1999. Organotin levels in seafood. Mar.
Pollut. Bull. 40, 226–232.
Berge, J.A., Brevik, E.M., Bjørge, A., Følsvik, N., Gabrielsen, G.W., Wolkers, H., 2004.
Organotins in marine mammals and seabirds from Norwegian territory. J.
Environ. Monit. 6, 108–112.
Bettin, C., Oehlmann, J., Stroben, E., 1996. TBT-induced imposex in marine
neogastropods is mediated by an increasing androgen level. Helgoland Mar.
Res. 50, 299–317.
Bianchi, F., Careri, M., Maffini, M., Mangia, A., Mucchino, C., 2006. Optimization of
the solid phase microextraction procedure for the ultra-trace determination of
organotin compounds by gas chromatography–inductively coupled plasmamass spectrometry. J. Anal. At. Sepctrom. 21, 970–973.
Borghi, V., Porte, C., 2002. Organotin pollution in deep-sea from the northwestern
Mediterranean. Environ. Sci. Technol. 36, 4224–4228.
Boyer, I.J., 1989. Toxicity of dibutyltin, tributyltin and other organotin compounds
to humans and to experimental animals. Toxicology 55, 253–298.
Brüschweiler, B.J., Würgler, F.E., Fent, K., 1995. Cytotoxicity in vitro of organotin
compounds to fish hepatoma cells PLHC-1 (Poeciliopsis lucida). Aquat. Toxicol.
32, 143–160.
Bryan, G.W., Gibbs, P.E., Burt, G.R., 1988. A comparison of the effectiveness of tri-Nbutyltin chloride and five other organotin compounds in promoting the
development of imposex in the dog-whelk Nucella lapillus. J. Mar. Biol. Assoc.
UK 68, 733–744.
Cao, D., Jiang, G., Zhou, Q., Yang, R., 2009. Organotin pollution in China: an overview
of the current state and the potential health risk. J. Environ. Manage., S16–S24.
Carlier-Pinasseau, C., Lespes, G., Astruc, M., 1997. Determination of butyl- and
phenyltin compounds in sediments by GC–FPD after NaBEt4 ethylation. Talanta
44, 1163–1171.
Ceulemans, M., Slaets, S., Adams, F., 1998. Speciation of organotin in environmental
sediment samples. Talanta 46, 395–405.
1023
Champ, M.A., 2000. A review of organotin regulatory strategies, pending actions,
related costs and benefits. Sci. Total Environ. 258, 21–71.
Chapman, P.M., Fairbrother, A., Brown, D., 1998. A critical evaluation of safety
(uncertainty) factors for ecological risk assessment. Environ. Toxicol. Chem. 17,
99–108.
Choi, M., Choi, H.-G., Moon, H.-B., Kim, G.-Y., 2009. Spatial and temporal distribution
of tributyltin (TBT) in seawater, sediments and bivalves from coastal areas of
Korea during 2001–2005. Environ. Monit. Assess. 151, 301–310.
Clarkson, T.W., 1995. Environmental contaminants in the food chain. Am. J. Clin.
Nutr. 61, 682S–686S.
Colosio, C., Tomasini, M., Cairoli, S., Foà, V., Minoia, C., Marinovich, M., Galli, C.L.,
1991. Occupational triphenyltin acetate poisoning: a case report. Br. J. Ind. Med.
48, 136–139.
Cooney, J.J., Wuertz, S., 1989. Toxic effects of tin compounds on microorganisms. J.
Ind. Microbiol. Biotechnol. 4, 375–402.
Consortium for Risk Evaluation with Stakeholder Participation, 2006. Note on: Wet
to Dry Weight Conversions for Biota from Amchitka and Kiska. <http://
www.cresp.org/Amchitka/Final_WW_DW_3_13_06.pdf>.
de Carvalho Oliveira, R., Santelli, R.E., 2010. Occurrence and chemical speciation
analysis of organotin compounds in the environment: a review. Talanta 82, 9–
24.
The European Community, 2003. Regulation (EC) No. 782/2003 of the European
Parliament and of the Council of 14 April 2003 on the Prohibition of Organotin
Compounds on Ships. Off. J. Eur. Union: L 115/1. <http://europa.eu/
index_en.htm> (accessed 24.04.12, 09.05.03.
(EFSA) European Food Safety Authority, 2004. Opinion of the scientific panel on
contaminants in the food chain on a request from the commission to assess the
health risks to consumers associated with exposure to organotins in foodstuffs.
EFSA J. 102, 1–119.
Evans, S.M., 1999. Tributyltin pollution: the catastrophe that never happened. Mar.
Pollut. Bull. 38, 629–636.
Fent, K., 1996a. Ecotoxicology of organotin compounds. Crit. Rev. Toxicol. 26, 3–117.
Fent, K., 1996b. Organotin compounds in municipal wastewater and sewage sludge:
contamination, fate in treatment process and ecotoxicological consequences.
Sci. Total Environ. 185, 151–159.
Fent, K., Bucheli, T.D., 1994. Inhibition of hepatic microsomal monooxygenase
system by organotins in vitro in freshwater fish. Aquat. Toxicol. 28, 107–126.
Fent, K., Hunn, J., 1991. Phenyltins in water, sediment, and biota of freshwater
marinas. Environ. Sci. Technol. 25, 956–963.
Fent, K., Meier, W., 1992. Tributyltin-induced effects on early life stages of minnows
Phoxinus phoxinus. Arch. Environ. Contam. Toxicol. 22, 428–438.
Fent, K., Meier, W., 1994. Effects of triphenyltin on fish early life stages. Arch.
Environ. Contam. Toxicol. 27, 224–231.
Fent, K., Hunn, J., Sturm, M., 1991a. Organotins in lake sediment.
Naturwissenschaften 78, 219–221.
Fent, K., Lovas, R., Hunn, J., 1991b. Bioaccumulation, elimination and metabolism of
triphenyltin chloride by early life stages of minnows Phoxinus phoxinus.
Naturwissenschaften 78, 125–127.
Féral, C., LeGall, S., 1983. The influence of a pollutant factor (tributyltin) on the
neuroendocrine mechanism responsible for the occurrence of a penis in the
females of Ocenebra erinacea. In: Lever, J., Boer, H. (Eds.), Molluscan Neuroendocrinology. North Holland Publishing Co., Amsterdam, pp. 173–175.
Garcia-Alonso, J.I., Sanz-Medel, A., Ebdon, L., 1993. Determination of butyltin ion
species by ion-exchange chromatography with inductively coupled plasma
mass spectrometric and spectrofluorimetric detection. Anal. Chim. Acta 283,
261–271.
George, S.G., Buchanan, G., 1990. Isolation, properties and induction of plaice liver
cytosolic gluthathione-S-transferases. Fish Physiol. Biochem. 8, 437–449.
Gibbs, P.E., Bryan, G.W., 1986. Reproductive failure in populations of the dog-whelk,
Nucella lapillus, caused by imposex induced by tributyltin from antifouling
paints. J. Mar. Biol. Assoc. UK 66, 767–777.
Gibbs, P.E., Bryan, G.W., Pascoe, P.L., Burt, G.R., 1987. The use of dog-whelk, Nucella
lapillus, as an indicator of tributyltin (TBT) contamination. J. Mar. Biol. Assoc. UK
67, 507–523.
Gibbs, P.E., Pascoe, P.L., Burt, G.R., 1988. Sex change in the female dog-whelk,
Nucella lapillus, induced by tributyltin from antifouling paints. J. Mar. Biol.
Assoc. UK 68, 715–731.
Gibbs, P.E., Bebianno, M.J., Coelho, M.R., 1997. Evidence of the differential sensitivity
of neogastropods to tributyltin (TBT) pollution, with notes on a species
(Columbella rustica) lacking the imposex response. Environ. Technol. 18,
1219–1224.
Golub, M.S., Doherty, J.D., 2004. Triphenyltin as a potential human endocrine
disruptor. J. Toxicol. Environ. Health, Part B 7, 281–295.
Gomez-Ariza, J.L., Morales, E., Giraldez, I., 1998. Spatial distribution of butyltin and
phenyltin compounds on the Huelva coast (southwest Spain). Chemosphere 37,
937–950.
Guðmundsdóttir, L.Ó., Ho, K.K.Y., Lam, J.C.W., Svavarsson, J., Leung, K.M.Y., 2011.
Long-term temporal trends (1992–2008) of imposex status associated with
organotin contamination in the dogwhelk Nucella lapillus along the Icelandic
coast. Mar. Pollut. Bull. 63, 500–507.
Guérin, T., Sirot, V., Volatier, J.-L., Leblanc, J.-C., 2007. Organotin levels in seafood
and its implications for health risk in high-seafood consumers. Sci. Total
Environ. 388, 66–77.
Hall, L.W., Pinkney, A.E., 1985. Acute and sublethal effects of organotin compounds
on aquatic biota: an interpretative literature evaluation. Crit. Rev. Toxicol. 14,
159–209.
1024
A.X. Yi et al. / Chemosphere 89 (2012) 1015–1025
Harino, H., Fukushima, M., Yamamoto, Y., Kawai, S., Miyazaki, N., 1998.
Contamination of butyltin and phenyltin compounds in the marine
environment of Otsuchi Bay, Japan. Environ. Pollut. 101, 209–214.
Harino, H., Fukushima, M., Kawai, S., 1999. Temporal trends of organotin
compounds in the aquatic environment of the port of Osaka, Japan. Environ.
Pollut. 105, 1–7.
Harino, H., Fukushima, M., Kawai, S., 2000. Accumulation of butyltin and phenyltin
compounds in various fish species. Arch. Environ. Contam. Toxicol. 39 (1), 13–
19.
Helsel, D.R., 2006. Fabricating data: how substituting values for nondetects can ruin
results, and what can be done about it. Chemosphere 65, 2434–2439.
Higashiyama, T., Shiraishi, H., Otsuki, A., Hashimoto, S., 1991. Concentrations of
organotin compounds in blue mussels from the wharves of Tokyo Bay. Mar.
Pollut. Bull. 22, 585–587.
Hinfray, N., Porcher, J.M., Brion, F., 2006. Inhibition of rainbow trout (Oncorhynchus
mykiss) P450 aromatase activities in brain and ovarian microsomes by various
environmental substances. Comp. Biochem. Physiol. Part C: Toxicol. Pharmacol.
144, 252–262.
Hobler, C., Andrade, A.J.M., Grande, S.W., Gericke, C., Talsness, C.E., Appel, K.E.,
Chahoud, I., Grote, K., 2010. Sex-dependent aromatase activity in rat offspring
after pre- and postnatal exposure to triphenyltin chloride. Toxicology 276, 198–
205.
Hoch, M., 2001. Organotin compounds in the environment – an overview. Appl.
Geochem. 16, 719–743.
Horiguchi, T., 2009. Mechanism of imposex induced by organotin in gastropods. In:
Arai, T., Harino, H., Ohji, M., Langston, W.J. (Eds.), Ecotoxicology of Antifouling
Biocides. Springer, Ibaraki, Japan, pp. 111–124.
Horiguchi, T., Shiraishi, H., Shimizu, M., Morita, M., 1994. Imposex and organotin
compounds in Thais clavigera and T. Bronni in Japan. J. Mar. Biol. Assoc. UK 74,
651–669.
Horiguchi, T., Shiraishi, H., Shimizu, M., Yamazaki, S., Morita, M., 1995. Imposex in
Japanese gastropods (neogastropoda and mesogastropoda): effects of
tributyltin and triphenyltin from antifouling paints. Mar. Pollut. Bull. 31, 402–
405.
Horiguchi, T., Shiraishi, H., Shimizu, M., Morita, M., 1996. Effects of triphenyltin
chloride and five other organotin compounds on the development of imposex in
the rock shell, Thais clavigera. Environ. Pollut. 95, 85–91.
Horiguchi, T., Nishikawa, T., Ohta, Y., Shiraishi, H., Morita, M., 2007. Retinoid X
receptor gene expression and protein content in tissues of the rock shell Thais
clavigera. Aquat. Toxicol. 84, 379–388.
Hu, J., Zhen, H., Wan, Y., Gao, J., An, W., An, L., Jin, F., Jin, X., 2006. Trophic
magnification of triphenyltin in a marine food web of Bohai Bay, north China:
comparison to tributyltin. Environ. Sci. Technol. 40, 3142–3147.
Hu, J., Zhang, Z., Wei, Q., Zhen, H., Zhao, Y., Peng, H., Wan, Y., Giesy, J.P., Li, L., Zhang,
B., 2009. Malformations of the endangered Chinese sturgeon, Acipenser sinensis,
and its causal agent. Proc. Natl. Acad. Sci. USA 106, 9339–9344.
Huang, G., Bai, Z., Dai, S., Xie, Q., 1993. Accumulation and toxic effect of
organometallic compounds on algae. Appl. Organomet. Chem. 7, 373–380.
Huang, G., Dai, S., Sun, H., 1996. Toxic effects of organotin species on algae. Appl.
Organomet. Chem. 10, 37–387.
Huang, G., Song, Z., Liu, G., Zhang, W., 2002. Toxic effect of triphenyltin chloride on
the alga Spirulina subsalsa. Appl. Organomet. Chem. 16, 117–122.
Hung, T.C., Lee, T.Y., Liao, T.F., 1998. Determination of butyltins and phenyltins in
oysters and fishes from Taiwan coastal waters. Environ. Pollut. 102, 197–203.
Hung, T.C., Hsu, W.K., Mang, P.J., Chuang, A., 2001. Organotins and imposex in the
rock shell, Thais clavigera, from oyster mariculture areas in Taiwan. Environ.
Pollut. 112, 145–152.
International Marine Organization, 2001. International Convention on the Control of
Harmful Antifouling Systems on Ships. Retrieved from IMO website: <http://www.
imo.org/about/conventions/listofconventions/pages/international-convention-onthe-control-of-harmful-anti-fouling-systems-on-ships-(afs).aspx>.
Inaba, K., Shiraishi, H., Soma, Y., 1995. Effects of salinity, pH and temperature on
aqueous solubility of four organotin compounds. Water Res. 29,
1415–1417.
Ishizaka, T., Nemoto, S., Sasaki, K., Suzuki, T., Saito, Y., 1989. Simultaneous
determination of tri-n-butyltin, di-n-butyltin, and triphenyltin compounds in
marine products. J. Agric. Food Chem. 37, 152–1527.
ISO 17353, 2004. Water Quality – Determination of Selected organotin Compounds
– Gas Chromatographic Method.
ISO/DIS 23161-2, 2007. Soil Quality – Determination of Selected Organotin
Compounds – Gas Chromatographic Method.
Iwata, H., Tanabe, S., Mizuno, T., Tatsukawa, R., 1997. Bioaccumulation of butyltin
compounds in marine mammals: the specific tissue distribution and
composition. Appl. Organomet. Chem. 11, 257–264.
Kannan, K., Falandysz, J., 1997. Butyltins in sediment, fish, fish-eating birds, harbour
porpoise and human tissues from the Polish coast of the Baltic Sea. Mar. Pollut.
Bull. 34, 203–207.
Kannan, K., Lee, R.F., 1996. Triphenyltin and its degradation products in foliage and
soils from sprayed pecan orchards and in fish from adjacent ponds. Environ.
Toxicol. Chem. 15, 1492–1499.
Kannan, K., Tanabe, S., Tatsukawa, R., 1992. Occurrence of butyltin residues in
certain foodstuffs. Bull. Environ. Contam. Toxicol. 55, 510–516.
Kannan, K., Senthilkumar, K., Giesy, J.P., 1999. Occurrence of butyltin compounds in
human blood. Environ. Sci. Technol. 33, 1776–1779.
Kimbrough, R.D., 1976. Toxicity and health effects of selected organotin
compounds: a review. Environ. Health Perspect. 14, 51–56.
Kono, K., Minami, T., Yamada, H., Tanaka, H., Koyama, J., 2008. Bioaccumulation of
tributyltin and triphenyltin compounds through the food web in deep offshore
water. Coastal Mar. Sci. 32, 102–107.
Laughlin, R.B., French, W., Guard, H.E., 1986. Accumulation of bis(tributyltin)
oxide by the marine mussel Mytilus edulis. Environ. Sci. Technol. 20, 884–
890.
Lee, C.-C., Wang, T., Hsieh, C.-Y., Tien, C.-J., 2005. Organotin contamination in fishes
with different living patterns and its implications for human health risk in
Taiwan. Environ. Pollut. 137, 198–208.
Lee, C.-C., Hsieh, C.-Y., Tien, C.-J., 2006. Factors influencing organotin distribution in
different marine environmental compartments, and their potential health risk.
Chemosphere 65, 547–559.
Leung, K.M.Y., Wheeler, J.R., Morritt, D., Crane, M., 2001. Endocrine disruption in
fishes and invertebrates: issues for saltwater ecological risk assessment. In:
Newman, M.C., Roberts, M.H., Jr., Hale, R.C. (Eds.), Coastal and Estuarine Risk
Assessment. Lewis Publishers, Boca Raton, pp. 189–216.
Leung, K.M.Y., Kwong, R.P.Y., Ng, W.C., Horiguchi, T., Qiu, J.W., Yang, R.Q., Song, M.Y.,
Jiang, G.B., Zheng, G.J., Lam, P.K.S., 2006. Ecological risk assessments of
endocrine disrupting organotin compounds using marine neogastropods in
Hong Kong. Chemosphere 65, 922–938.
Lima, D., Reis-Henriques, M.A., Silva, R., Santos, A.I., Castro, L.F.C., Santos, M.M.,
2011. Tributyltin-induced imposex in marine gastropods involves tissuespecific modulation of the retinoid X receptor. Aquat. Toxicol. 101, 221–227.
Lu, F.C., 1994. A review of acceptable daily intake of pesticides assessed by WHO.
Regul. Toxicol. Pharmacol. 21, 352–364.
Manzo, L., Richelmi, P., Sabbioni, E., Pietra, R., Bono, F., Guardia, L., 1981. Poisoning
by triphenyltin acetate. Report of two cases and determination of tin in blood
and urine by neutron activation analysis. Clin. Toxicol. 18, 1343–1353.
Meng, P.-J., Lin, J., Liu, L.-L., 2009. Aquatic organotin pollution in Taiwan. J. Environ.
Manage. 90, S8–S15.
Midorikawa, S., Arai, T., Harino, H., Ohji, M., Cu, N.D., Miyazaki, N., 2004.
Concentrations of organotin compounds in sediment and clams collected
from coastal areas in Vietnam. Environ. Pollut. 131, 401–408.
Miller, D.H., Jensen, K.M., Villeneuve, D.L., Kahl, M.D., Makynen, E.A., Durhan, E.J.,
Ankley, G.T., 2007. Linkage of biochemical responses to population-level effects:
a case study with vitellogenin in the fathead minnow. Environ. Toxicol. Chem.
26, 521–527.
Minchin, D., Oehlmann, J., Duggan, C.B., Stroben, E., Keatinge, M., 1995. Marine TBT
antifouling contamination in Ireland, following legislation in 1987. Mar. Pollut.
Bull. 30, 633–639.
Miura, Y., Matsui, H., 1991. Inhibitory effects of phenyltin compounds on stimulusinduced changes in cytosolic free calcium and plasma membrane potential of
human neutrophils. Arch. Toxicol. 65, 562–569.
Mooney, H.M., Patching, J.W., 1995. Triphenyltin inhibits photosynthesis and
respiration in marine microalgae. J. Ind. Microbiol. Biotechnol. 14, 265–270.
Morcillo, Y., Borghi, V., Porte, C., 1997. Survey of organotin compounds in the
western Mediterranean using molluscs and fish as sentinel organisms. Arch.
Environ. Contam. Toxicol. 32, 198–203.
Murai, R., Sugimoto, A., Tanabe, S., Takeuchi, I., 2008. Biomagnification profiles of
tributyltin (TBT) and triphenyltin (TPT) in Japanese coastal food webs
elucidated by stable nitrogen isotope ratios. Chemosphere 73, 1749–1756.
Nagase, M., Kondo, H., Hasebe, K., 1995. Determination of tributyltin and
triphneyltin compounds in hair and fish using a hydrolysis technique and
gas chromatography with flame photometric detection. Analyst 120, 1923–
1926.
Nakayama, K., Matsudaira, C., Tajima, Y., Yamada, T.K., Yoshioka, M., Isobe, T.,
Takahashi, S., Tanabe, S., 2009. Temporal and spatial trends of organotin
contamination in the livers of finless porpoises (Neophocaena phocaenoides) and
their association with parasitic infection status. Sci. Total Environ. 407, 6173–
6178.
Nemanič, T.M., Leskovšek, H., Horvat, M., Vrišer, B., Bolje, A., 2002. Organotin
compounds in the marine environment of the Bay of Piran, Northern Adriatic
Sea. J. Environ. Monit. 4, 426–430.
Nishikawa, J., Mamiya, S., Kanayama, T., Nishikawa, T., Shiraishi, F., Horiguchi, T.,
2004. Involvement of the retinoid X receptor in the development of imposex
caused by organotins in gastropods. Environ. Sci. Technol. 38, 6271–6276.
Oberdörster, E., McClellan-Green, P., 2000. Mechanisms of imposex induction in the
mud snail, Ilyanassa obsoleta: TBT as a neurotoxin and aromatase inhibitor. Mar.
Environ. Res. 54, 715–718.
Oliveira, I.B., Richardson, C.A., Sousa, A.C., Takahashi, S., Tanabe, S., Barroso, C.M.,
2009. Spatial and temporal evolution of imposex in dogwhelk Nucella lapillus
(L.) populations from North Wales, UK. J. Environ. Monit. 11, 1462–1468.
Phillips, D.J.H., 1985. Organochlorines and trace metals in green-lipped mussels
Perna viridis from Hong Kong waters: a test of indicator ability. Mar. Ecol: Prog.
Ser. 21, 251–258.
Rantakokko, P., Kuningas, T., Saastamoinen, K., Vartiainen, T., 2006. Dietary intake of
organotin compounds in Finland: a market-basket study. Food Addit. Contam.
23, 749–756.
Rantakokko, P., Turunen, A., Verkasalo, P.K., Kiviranta, H., Männistö, S., Vartiainen,
T., 2008. Blood levels of organotin compounds and their relation to fish
consumption in Finland. Sci. Total Environ. 399, 90–95.
Rivaro, P., Frache, R., Leardi, R., 1997. Seasonal variations in levels of butyltin
compounds in mussel tissues sampled in an oil port. Chemosphere 34, 99–106.
Rodríguez, E.M., Medesani, D.A., Fingerman, M., 2007. Endocrine disruption in
crustaceans due to pollutants: a review. Comp. Biochem. Physiol. Part A: Mol.
Integr. Physiol. 146, 661–671.
A.X. Yi et al. / Chemosphere 89 (2012) 1015–1025
Ronis, M.J.J., Mason, A.Z., 1996. The metabolism of testosterone by the periwinkle
(Littorina littorea) in vitro and in vivo: effects of tributyltin. Mar. Environ. Res.
42, 161–166.
Rosenberg, E., Kmetov, V., Grasserbauer, M., 2000. Investigating the potential of
high-performance liquid chromatography with atmospheric pressure chemical
ionization–mass spectrometry as an alternative method for the speciation
analysis of organotin compounds. Fresenius. J. Anal. Chem. 366, 400–407.
Rüdel, H., 2003. Case study: bioavailability of tin and tin compounds. Ecotoxicol.
Environ. Saf. 56, 180–189.
Rüdel, H., Lepper, P., Steinhanses, J., 2003. Retrospective monitoring of organotin
compounds in marine biota from 1985 to 1999: results from the German
Environmental Specimen Bank. Environ. Sci. Technol. 37, 1731–1738.
Rumengan, I., Ohji, M., Arai, T., Harino, H., Arifin, Z., Miyazaki, N., 2008.
Contamination status of butyltin compounds in Indonesian coastal waters.
Coastal Mar. Sci. 32, 116–126.
Sadiki, A., Williams, D.T., Carrier, R., Thomas, B., 1996. Pilot study on the
contamination of drinking water by organotin compounds from PVC
materials. Chemosphere 32, 2389–2398.
Santos, M.M., Reis-Henriques, M.A., Vieira, M.N., Solé, M., 2006. Triphenyltin and
tributyltin, single and in combination, promote imposex in the gastropod
Bolinus brandaris. Ecotoxicol. Environ. Saf. 64, 155–162.
Santos, M.M., Enes, P., Reis-Henriques, M.A., Kuballa, J., Castro, L.F.C., Vieira, M.N.,
2009. Organotin levels in seafood from Portuguese markets and the risk for
consumers. Chemosphere 75, 661–666.
Scanlon, P.F., 1982. Wet and dry weight relationships of mallard (Anas
platyrhynchos) tissues. Bull. Environ. Contam. Toxicol. 29, 615–617.
Semlitsch, R.D., Foglia, M., Mueller, A., Steiner, I., Fioramonti, E., Fent, K., 1995.
Short-term exposure to triphenyltin affects the swimming and feeding
behaviour of tadpoles. Environ. Toxicol. Chem. 14, 1419–1423.
Shim, W.J., Kahng, S.H., Hong, S.H., Kim, N.S., Kim, S.K., Shim, J.H., 2000. Imposex in
the rock shell, Thais clavigera, as evidence of organotin contamination in the
marine environment of Korea. Mar. Environ. Res. 49, 435–451.
Shim, W.J., Hong, S.H., Kim, N.S., Yim, U.H., Li, D., Oh, J.R., 2005a. Assessment of
butyl- and phenyltin pollution in the coastal environment of Korea using
mussels and oysters. Mar. Pollut. Bull. 51, 922–931.
Shim, W.J., Yim, U.H., Kim, N.S., Hong, S.H., Oh, J.R., Jeon, J.K., Okamura, H., 2005b.
Accumulation of butyl- and phenyltin compounds in starfish and bivalves from
the coastal environment of Korea. Environ. Pollut. 133, 489–499.
Soderquist, C.J., Crosby, D.G., 1980. Degradation of triphenyltin hydroxide in water.
J. Agric. Food Chem. 28, 111–117.
Solé, M., Morcillo, Y., Porte, C., 1998. Imposex in the commercial snail Bolinus
brandaris in the northwestern Mediterranean. Environ. Pollut. 99, 241–246.
Sousa, A., Matsudaira, C., Takahashi, S., Tanabe, S., Barroso, C., 2007. Integrative
assessment of organotin contamination in a southern European estuarine
system (Ria de Aveiro, NW Portugal): tracking temporal trends in order to
evaluate the effectiveness of the EU ban. Mar. Pollut. Bull. 54, 1645–1653.
Stäb, J.A., Traas, T.P., Stroomberg, G., van Kesteren, J., Leonards, P., van Hattum, B.,
Brinkman, U.A.T., Cotino, W.P., 1996. Determination of organotin compounds in
the food web of a shallow freshwater lake in the Netherlands. Arch. Environ.
Contam. Toxicol. 31, 319–328.
Stewart, C., Thompson, J.A.J., 1994. Extensive butyltin contamination in
southwestern coastal British Columbia, Canada. Mar. Pollut. Bull. 28, 601–606.
Stoner, H.B., 1966. Toxicity of triphenyltin. Br. J. Ind. Med. 23, 222–229.
Strand, J., Jacobsen, J.A., 2005. Accumulation and trophic transfer of organotins in a
marine food web from the Danish coastal waters. Sci. Total Environ. 350, 72–85.
Subramanyam, B., 1993. Chemicals used on Four Crops in Minnesota During 1991.
University
of
Minnesota,
Station
Bulletin
603-1993.
<http://
www.extension.umn.edu/distribution/cropsystems/DC6235.html>.
Sun, L., Zhang, J., Zuo, Z., Chen, Y., Wang, X., Huang, X., Wang, C., 2011. Influence of
triphenyltin exposure on the hypothalamus–pituitary–gonad axis in male
Sebastiscus marmoratus. Aquat. Toxicol. 104, 263–269.
Suzuki, T., Yamamoto, I., Yamada, H., Kaniwa, N., Kondo, K., Murayama, M., 1998.
Accumulation, metabolism, and depuration of organotin compounds in the
marine mussels Mytilus graynus and Mytilus edulis under natural conditions. J.
Agric. Food Chem. 46, 303–313.
Takahashi, S., Mukai, H., Tanabe, S., Sakayama, K., Miyazaki, T., Masuno, H., 1999.
Butyltin residues in livers of humans and wild terrestrial mammals and in
plastic products. Environ. Pollut. 106, 213–218.
Tanabe, S., 1999. Butyltin contamination in marine mammals – a review. Mar.
Pollut. Bull. 39, 62–72.
1025
Tas, J.W., Hermens, J.L.M., van den Berg, M., Seinen, W., 1989. Bioconcentration and
elimination of triphenyltin hydroxide in fish. Mar. Environ. Res. 28, 215–218.
Tas, J.W., Seinen, W., Opperhuizen, A., 1991. Lethal body burden of triphenyltin
chloride in fish: preliminary results. Comp. Biochem. Physiol. Part C: Toxicol.
Pharmacol. 100, 59–60.
Tolosa, I., Bayona, J.M., Albaiges, J., Merlini, L., de Bertrand, N., 1992. Occurrence and
fate of tributyl- and triphenyltin compounds in western Mediterranean coastal
enclosures. Environ. Toxicol. Chem. 11, 145–155.
Tsuda, T., Nakanishi, H., Aoki, S., Takebayashi, J., 1987. Bioconcentration and
metabolism of phenyltin chlorides in carp. Water Res. 21, 949–953.
Tsuda, T., Wada, M., Aoki, S., Matsui, Y., 1988. Bioconcentration, excretion and
metabolism of bis (tri-n-butyltin) oxide and triphenyltin chloride by goldfish.
Toxicol. Environ. Chem. 18, 11–20.
Tsuda, T., Aoki, S., Kojima, M., Harada, H., 1990a. Differences between freshwater
and seawater-acclimated guppies in the accumulation and excretion of tri-nbutyltin chloride and triphenyltin chloride. Water Res. 24, 1373–1376.
Tsuda, T., Aoki, S., Kojima, M., Harada, H., 1990b. The influence of pH on the
accumulation of tri-n-butyltin chloride and triphenyltin chloride in carp. Comp.
Biochem. Physiol. Part C: Toxicol. Pharmacol. 95, 151–153.
Tutschku, S., Mothes, S., Dittrich, K., 1994. Determination and speciation of
organotin compounds by gas chromatography–microwave induced plasma
atomic emission spectrometry. J. Chromatogr., A 683, 269–276.
Ueno, S., Susa, N., Furukawa, Y., Komatsu, Y., Koyama, S., Suzuki, T., 1999. Butyltin
and phenyltin compounds in some marine fishery products on the Japanese
market. Arch. Environ. Health 54, 20–25.
(USEPA) US Environmental Protection Agency, Office of Research and Development
(1998). Guidance for Data Quality Assessment: Practical Methods for Data
Analysis (EPA/600/R-96/084). Washington, DC.
U.S. Environmental Protection Agency, 2011a. Ecotox Database (Version 4.0).
<http://www.epa.gov/ecotox>.
U.S. Environmental Protection Agency, 2011b. Species Sensitivity Distribution
Generator. <http://www.epa.gov/caddis/da_software_ssdmacro.html>.
van der Oost, R., Beyer, J., Vermeulen, N.P.E., 2003. Fish bioaccumulation and
biomarkers in environmental risk assessment: a review. Environ. Toxicol.
Pharmacol. 13, 57–149.
Veltman, K., Huijbregts, M.A.J., van den Heuvel-Greve, M.J., Vethaak, A.D., Hendriks,
A.J., 2006. Organotin accumulation in an estuarine food chain: comparing field
measurements with model estimations. Mar. Environ. Res. 61, 511–530.
Verschuuren, H.G., Kroes, R., Vink, H.H., van Esch, G.J., 1966. Short-term toxicity
studies with triphenyltin compounds in rats and guinea-pigs. Food Cosmet.
Toxicol. 4, 35–45.
Wang, X., Hong, H., Zhao, D., Hong, L., 2008. Environmental behaviour of organotin
compounds in the coastal environment of Xiamen, China. Mar. Pollut. Bull. 57,
419–424.
Wheeler, J.R., Grist, E.P.M., Leung, K.M.Y., Morritt, D., Crane, M., 2002. Species
sensitivity distribution: data and model choice. Mar. Pollut. Bull. 45, 192–202.
World Health Organization, 1999. Triphenyltin Compounds. Geneva Switzerland:
WHO. <http://www.inchem.org/documents/cicads/cicads/cicad13.htm>.
Yamada, H., Takayanagi, K., 1992. Bioconcentration and elimination of
bis(tributyltin)oxide (TBTO) and triphenyltin chloride (TPTC) in several
marine fish species. Water Res. 26, 1589–1595.
Yamada, H., Takayanagi, K., Tateishi, M., 1994. Bioaccumulation of organotin
compounds in the red sea bream (Pagrus major) by two uptake pathways:
dietary uptake and direct uptake from water. Environ. Toxicol. Chem. 13, 1415–
1422.
Yang, J., Miyazaki, N., 2003. Moisture content in Dall’s porpoise (Phocoenoides dalli)
tissues: a reference base for conversion factors between dry and wet weight
trace element concentrations in cetaceans. Environ. Pollut. 121, 345–347.
Yang, J., Harino, H., Miyazaki, N., 2007. Transplacental transfer of phenyltins from a
pregnant Dall’s porpoise (Phocoenoides dalli) to her foetus. Chemosphere 67,
244–249.
Yang, J., Oshima, Y., Sei, I., Miyazaki, N., 2009. Metabolism of tributyltin and
triphenyltin by Dall’s porpoise hepatic microsomes. Chemosphere 76, 1013–
1015.
Yuan, J., Zhang, X., Yu, L., Sun, Z., Zhu, P., Wang, X., Shi, H., 2011. Stage-specific
malformations and phenotypic changes induced in embryos of amphibian
(Xenopus tropicalis) by triphenyltin. Ecotoxicol. Environ. Saf. 74, 1960–1966.
Zhang, Z., Hu, J., Zhen, H., Wu, X., Huang, C., 2008. Reproductive inhibition and
transgenerational toxicity of triphenyltin on medaka (Oryzias latipes) at
environmentally relevant levels. Environ. Sci. Technol. 42, 8133–8139.