3/2/2015
Impacts of Endocrine Disrupting Chemicals to Human and the Environment
香港大學 理學院 副院長 梁美儀教授
Professor Kenneth Mei-Yee Leung
Associate Dean (Research & Graduate Studies)
Faculty of Science, The University of Hong Kong
講座內容大綱
1. 從科學探究例子學習歸納思維：學習如何分
辨海洋生物的性別及探討動物的性別由來
2. 以環境雌激素作例子來說明人工合成化學物
品如何能干擾生物賀爾蒙系統
3. 介紹兩個本地的相關個案
4. 教學經驗分享及總結
5. 問答及討論環節
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講座內容大綱
1. 從科學探究例子學習歸納思維：學習如何分
辨海洋生物的性別及探討動物的性別由來
2. 以環境雌激素作例子來說明人工合成化學物
品如何能干擾生物賀爾蒙系統
3. 介紹兩個本地的相關個案
4. 教學經驗分享及總結
5. 問答及討論環節
How can you sex them?
如何分辨牠們的性別?
Photo source: http://mfrost.typepad.com/cute_overload/farm_animals/index.html
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How can you sex crabs? 如何分辨牠們的性別?
How can you sex shrimps?
如何分辨牠們的性別?
Between the base of the 4th and 5th
pereiopods
Source: http://www.hk-fish.net/
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How can you sex
mantis shrimps?
如何分辨牠們的性別?
Source: http://www.hk-fish.net/
馬蹄蟹在遠
古時已出現
何解牠的
血是藍色
的? 有何用?
如何分辨牠們的性別?
Blood of horseshoe
crabs contains Limulus
Amoebocyte Lysate
(LAL) which
immediately binds and
clots around fungi,
viruses, and bacterial
endotoxins. LAL is now
used to test microbial
contamination of drugs
and surgical apparatus.
http://blog.nativefoods.com/nativefoods/2012/02/the‐
horseshoe‐crab‐harvest‐million‐dollar‐blue‐bloods.html
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Female
雌
Male
雄
Male
Female
How can you sex an octopus?
如何分辨牠們的性別?
Source: http://1.bp.blogspot.com/
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How can you sex an octopus?
如何分辨牠們的性別?
http://chestofbooks.com
http://www.wallawalla.edu/academics/departments/biology/rosario/inv
erts/Mollusca/Cephalopoda/Enteroctopus_dofleini.html
Photo source: http://mfrost.typepad.com/cute_overload/farm_animals/index.html
你知道動物性別的由來嗎?
What are the key factors controlling the sex of
an individual?
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1. Genetics 性別基因
(不同的性染色體)
Photo source: http://mfrost.typepad.com/cute_overload/farm_animals/index.html
1. Genetic defects 遺傳出錯
資料來源: http://goo.gl/N1g5GN
日本出現罕見雌雄同體龍蝦。（圖／翻攝自日本
《每日新聞》）
日本三重縣鳥羽市答志島地區一名旅館老闆日前
進貨500隻龍蝦，其中一隻雌雄同體，可能是受
精卵分裂成2個細胞時，決定了雌雄及體色基因
交錯導致。
http://www.ettoday.net/news/20140928/406630.htm
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1. Genetic Defects – ‘Intersex’
some examples
Genotype
Defects
Phenotype
XX
Congenital adrenal
hyperplasia (due to 21hydroxylase deficiency
Adrenals and ovaries cannot produce sehormones, inhibiting breast development
XX
Freemartinism due to
In cattle’ fused twins, male hormones can go
temporal fusion of fraternal into the female and masculinize her. The
twins
affected female is infertile (like castrated male)
XY
Androgen insensitivity
syndrome
XY
5-alpha-reductase
From infertility with male genitalia to male
deficiency; it is required for underdevelopment
male genitalia development
Other
Unusual sex chromosome
e.g. XXX, XXY, XYY, etc.
Other
Mosaicism
Some cells have the common XX or XY, while
some have one of the less usual chromosomal
contents.
Having a large clitoris or a small penis, with 1
or both testes undescended
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Some Youtube Videos about Intersex (For Adult Only)
Other related videos:
https://www.youtube.com/watch?v
=cJxZe4KAdqU
https://www.youtube.com/watch?v
=vFd-EIkaQ3k
https://www.youtube.com/watch?v
=cr96b9v1YB8
https://www.youtube.com/watch?v
=veM0h--ZM-4
2. Epigenetics 表觀基因體機制
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表觀基因體機制
Specific epigenetic processes include paramutation, bookmarking, imprinting, gene silencing,
X chromosome inactivation, position effect, reprogramming, transvection, maternal effects, the
progress of carcinogenesis, many effects of teratogens, regulation of histone modifications etc.
2. Epigenetic 表觀基因體機制
基因體印記 (Genomic imprinting) 是涉
及DNA甲基化和組蛋白修飾的一種附基
因調控，會造成單一等位基因表現
(monoallelic gene expression)
Lefebvre et al. 1998. Abnormal maternal
behaviour and growth retardation associated with
loss of the imprinted gene Mest. Nature Genetics
20: 163-169
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http://www.ncbi.nlm.nih.gov/books/NBK9989/
http://www.flickr.com/photos
3. Temperature-dependent sex determination
爬蟲類的性別是由溫度控制
http://www.public.iastate.edu/~jneuwald/Research.html
4. Genetic and Environmental Control – Natural sex
change in fishes 有些魚類能變性，例如石班魚
Hermaphroditism
De Mitcheson and Liu (2008), Fish and Fisheries 9:1-43
Liu and De Mitcheson (2009), Aquaculture 287: 191-202
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5. Food availability and maternal conditions
營養也可影響下一代的性別
>
The lesser black-backed
gull Larus fuscus
Nager RG et al. (1999) Proc Natl Acad Sci USA 96:570–573
Nager RG et al. (2000) Ecology 81:1339–1350
5. 營養也可影響下一代的性別
http://www.jaunted.com/files/34094/reindeer.jpg
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6. Hormones 賀爾蒙的影響
Apple Daily 23.8.2011
http://mfrost.typepad.com/cute_overload/kittens/index.html
7. 人工合成的化學物品能干擾生物賀爾蒙系統
[Endocrine Disruptors: 環境賀爾蒙]
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講座內容大綱
1. 從科學探究例子學習歸納思維：學習如何分
辨海洋生物的性別及探討動物的性別由來
2. 以環境雌激素作例子來說明人工合成化學物
品如何能干擾生物賀爾蒙系統
3. 介紹兩個本地的相關個案
4. 教學經驗分享及總結
5. 問答及討論環節
人體內的一些賀爾
蒙例 子 : 如皮質
醇(Cortisol)、雌激
素(Estradiol) 及雄
激素 (Testosterone)
雌及雄激素的化學
結構十分相似
Source: Withers 1992
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Source: Google
Source: Colborn et al.
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生物放大作用
Bio-concentration Bio-accumulation Bio-magnification
生物放大
隨食物鏈, 原先在
水中一個單位的
DDT可擴大至千
萬倍, 導致頂級捕
食者中毒及產生
亞致死反應。
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例子一
為甚麼農藥滴滴涕能使鳥類絕種?
http://www.eagles.org/photo.html
例子一
因滴滴
涕使蛋
殼變薄
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例子一
有機氯化合物導致海洋捕乳類動物的幼兒雌雄同體，影響繁殖
Pictures downloaded from BBC’s homepage
例子一
滴滴涕也引致女孩早熟: 八歲長乳房, 十歲有月經
BBC News: 16 May, 2001 Premature
puberty link to DDT
• In Belgium, immigrant children from countries (e.g. India
and Colombia) were 80 times more likely to start puberty
unusually young.
• They began developing breasts before 8 years old, and
started their periods before they were 10.
• 75% of these children had high levels of a chemical
derivative of DDT in their blood.
Source: BBC News
Time Magazine OCTOBER 30, 2000 VOL. 156 NO. 18
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例子一
HK Economic Daily 29/7/07 A21
例子二
塑化劑如 雙酚A 和 雙酚S 均是環境雌激素
雙酚S 是用來取替雙酚A,
但其內分泌干擾性及毒
性較雙酚A高。
http://db2.photoresearchers.com/feature/infocus97
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例子二
塑化劑 雙酚A 無處不在
http://db2.photoresearchers.com/feature/infocus97
前列腺癌病人血內雙酚A濃渡較高
例子二
Figure 1. Scatter plots of LnBPA.
Tarapore P, Ying J, Ouyang B, Burke B, et al. (2014) Exposure to Bisphenol A Correlates with Early-Onset Prostate Cancer and
Promotes Centrosome Amplification and Anchorage-Independent Growth In Vitro. PLoS ONE 9(3): e90332.
doi:10.1371/journal.pone.0090332
http://www.plosone.org/article/info:doi/10.1371/journal.pone.0090332
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上海9‐12歲女孩中，如她們尿液中BPA高於每升
兩毫克，她們超重的機會比正常者高出2.3倍。
June 2013 | Volume 8 | Issue 6 | e65399
其它塑化劑(DEHP, DINP, DBP)污染亦嚴重
例子二
http://www.videohealthy.com/yNBFTl25Vz0wP
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塑化劑無處不在
例子二
例子二
38%台產塑膠包裝食品含塑化劑
[2014年10月4日 大公報]
41件樣品中測出13件含有微量鄰苯二甲酸酯
類（即塑化劑）。
13件檢測出微量塑化劑的樣品中，7-11鹽烤
豬肉夾心飯糰、麥當勞芝士蛋堡、漢堡王華鱈魚
堡、佳德鳳梨酥、維格鳳梨酥、維力炸醬罐、牛
頭牌沙茶醬等10件樣品被驗出含有塑化劑DEHP；
而康師傅正宗紅燒牛肉麵、味味A排骨雞湯麵、
統一滿漢大餐燒豬肉麵與佳德鳳梨酥則驗出另一
種塑化劑DINP。佳德鳳梨酥同時含有DBP塑化
劑，污染問題較為嚴重。
攝入過多恐幹擾內分泌: 倘若食用過多含有塑化
劑的食品，則將會幹擾內分泌，女童容易早熟，
而男童也恐將出現女性化傾向。
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講座內容大綱
1. 從科學探究例子學習歸納思維：學習如何分
辨海洋生物的性別及探討動物的性別由來
2. 以環境雌激素作例子來說明人工合成化學物
品如何能干擾生物賀爾蒙系統
3. 介紹兩個本地的相關個案
4. 教學經驗分享及總結
5. 問答及討論環節
Brief Question & Answer
A Break for 10 Minutes
講座內容大綱
1. 從科學探究例子學習歸納思維：學習如何分
辨海洋生物的性別及探討動物的性別由來
2. 以環境雌激素作例子來說明人工合成化學物
品如何能干擾生物賀爾蒙系統
3. 介紹兩個本地的相關個案:
–
例子一：香港的污水
4. 教學經驗分享及總結
5. 問答及討論環節
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本地例子一
塑化劑 雙酚A 無處不在
http://db2.photoresearchers.com/feature/infocus97
本地例子一
Nonylphenol (壬基酚聚氧乙烯) 常見於污水
處理後排放的污水中 ，是表面活性劑 (如洗頭
水、洗手液) 的分解物，也是環境雌激素。
Nonylphenol
Nonylphenol 能使
雄性魚雌性化
Pictures downloaded from BBC’s homepage
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本地例子一
鶴咀海岸公園也受到污染的影響
3 km
Xu (2014) PhD Thesis, HKU
本地例子一
在鶴咀海岸公園鄰近的污水廠處理完的污水中測得12種內
分泌干擾物質，濃渡最高是Nonylpehnol、接著是雙酚A 及
Tricosan (二氯苯氧氯酚，用於抗菌皂及護膚品，但去年被
發現「不抗菌」) ， 三者均是環境雌激素。
BPA
Tricosan
BPA
Tricosan
BPA
Tricosan
Xu et al. (2014) Marine Pollution Bulletin
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Endocrine Disrupting Chemicals (EDCs)
plasticizers
HO
surfactants
More in coastal waters which are closer to shipping, and outfall discharge.
Very low concentration of EDCs can exhibit significant effects on the environment
and human health.
(Atkinson et al. 2003; Melnick et al. 2002)
Xu et al. (2014) Mar Pollut Bullet
Removal efficiencies of different STPs
• At Shek O, the removal efficiency was low. • Higher removal efficiency in summer > winter
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本地例子一
鶴咀海岸公園也受到污染的影響
大腸桿菌量
3 km
水流方向
Xu (2014) PhD Thesis, HKU
本地例子一
Nonylpehnol 和 雙酚A 在海洋生物體內的濃渡也
偏高， 威脅鶴咀海岸公園內的生物多樣性。
In biota samples, BPA and NP were found between 12.8-241.8 and 7.5739.4 ng/g dry weight, respectively
Risk Quotient > 1: 49% for NP and 33% for BPA (Alarming!!)
Anthocidaris crassspina
Risk distribution of NP based on
Monte Carlo simulation
Septifer virgatus
Monodonta labio
Holothuria spilonota
sediment
Grapsus albolineatus
Thais clavigera
Nerita costata
Bathybobius fuscus
Palaemon pacificus
NP
Barbatia virescens
BPA
Ulva spp.
1
10
100
Mean concenration (ng/g, dw)
1000
Xu et al. (2015) Marine Pollution Bulletin
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Biological Effects of the Diluted Effluents on Medaka
Different Dilutions
Outfall
1%
10%
Artificial SW
Seawater
Mortality
2 to 10 dpf
Hatchability
Growth
mRNA Expression
1st fry
Negative effects of diluted effluents on fish health
Sample size 4-11, harmonic mean size=5.66
Higher Mortalities: SO10% and SW10%
Lower Hatchability: SO10% and SW10%
Smaller Larvae: ST10% and SO1%
Xu et al. (2014) Mar Pollut Bullet
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mRNA Expression after Exposure to Diluted Effluents
Exposure
RNA extraction
RNA quality validation
Real‐time PCR
Example : Vitellogenin (VTG: egg yolk protein) gene expression at 1st fry stage of medaka
VTG1 and VTG2 in SO1% were increased by 2 and 5 times relative to control groups, respectively.
Xu et al. (2014) Mar Pollut Bullet
講座內容大綱
1. 從科學探究例子學習歸納思維：學習如何分
辨海洋生物的性別及探討動物的性別由來
2. 以環境雌激素作例子來說明人工合成化學物
品如何能干擾生物賀爾蒙系統
3. 介紹兩個本地的相關個案:
–
例子二：香港的三丁基錫 (TBT)及 三苯基锡
(TPT) 污染問題
4. 教學經驗分享及總結
5. 問答及討論環節
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三丁基錫 (TBT)及 三苯基锡 (TPT)
本地例子二
• 用於船底油漆以防正生物生長於船殼上
© Kevin Ho
© www.clr.pdx.edu/images/iansampling.jpg
• 常用農藥
© landscapeonline.com
© chinapost.com.tw
© en.wikipedia.org
三丁錫 (TBT)及 三丁基錫 (TPT)
本地例子二
Highest levels of mono‐ and dibutyltin compounds in corks ranged from 3.3 to 6.7 μg Sn/g
(Jiang et al. 2004. Environ Sci Technol, 38:4349‐52)
In China, the drinking water sources contained concentrations of BTs and PTs up to 1.1 μg Sn/L. (Gao et al., 2013)
Photo sources: BBC, ORTEPA, http://www.vectormap.com/tour/co1.htm
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本地例子二
Normal female without penis
Imposex female with penis
Penis
Vas
Deferens
• TBT及TPT 能導致蠔生長變異，
只長殼不長肉。
• 它們使雌性海螺類長出雄性性器
官，以致不育，群落滅絕。香港
情況非常嚴重。
Field survey in HK and Shenzhen
Hong Kong (2010)
Rock shell
Reishia clavigera
Shenzhen (2013)
High imposex levels with high tissue concentrations found close
to shipping facilities / mariculture zones
Ho et al. (in preparation); Ho and Leung (2014) Marine Pollution Bulletin
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Field survey in HK and Shenzhen
Hong Kong (2010)
Shenzhen (2013)
Positive correlation between imposex indices and tissue OTs concentration
TPT was the most abundant OT residue
Highest TPT
concentration
of 22,000 μg
kg-1 dry
weight found
in Shenzhen
Ho et al. (in preparation); Ho and Leung (2014) Marine Pollution Bulletin
Field survey in HK and Shenzhen
RQ for phenyltin compounds
• Over 17% of R. clavigera
population were at risk
due to exposure to
phenyltins in HK
2010
RQ for butyltin compounds
2006
2010
Ho et al. (in preparation); Ho and Leung (2014) Marine Pollution Bulletin
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Studied 11 species
採樣11種本地海產
本地例子二
Bivalves
Fishes
Meretrix lusoria
(Common oriental
calm)
Collichthys lucidus
(Lion head croaker)
Harpadon nehereus
(Bombay duck)
Johnius belangerii
(Belanger's croaker)
Nibea albiflora
Paraplagusia blochii Siganus canaliculatus
(White flower croaker) (Bloch's tonguesole)
(Rabbitfish)
Ruditapes
variegatus
(Variegate
venus)
Gastropods
Babylonia
areolata
(Babylon shell)
Bufonaria
rana
(Frog shell)
Hemifusus
tuba
(Spiny shell)
Photos © Kevin Ho, Yanny Mak, dictall.com, CNKI
本地例子二
TBT及TPT在海產內的濃渡
56-97% were TPT
a
b
c
**
i
a
ab
a
a
b
a
Error bar = +1 SEM
t test or Kruskal-Wallis test
followed by Dunn’s test
Ho & Leung (2014) Marine Pollution Bulletin
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Hazard Index (TBT及TPT 的風險)
Gastropod
例子三
HI = 1
Fish
HI = 1
本地底棲魚類如
龍利魚的有機錫
濃渡甚高， 毒
害風險甚高，
因此少吃為妙。
i
Ho & Leung (2014) Marine Pollution Bulletin
講座內容大綱
1. 從科學探究例子學習歸納思維：學習如何分
辨海洋生物的性別及探討動物的性別由來
2. 以環境雌激素作例子來說明人工合成化學物
品如何能干擾生物賀爾蒙系統
3. 介紹兩個本地的相關個案
4. 教學經驗分享及總結
5. 問答及討論環節
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如何減少環境賀爾蒙對我們和生態的影響?
1. 必須了解環境賀爾蒙的來源及其害處
2. 多認識我們常用的產品、 飲料及食物
3. 減少其攝取量
4. 支持及推動可持續(永續) 的經濟發展模式
塑化劑無處不在
• 少吃加工及包裝的飲料
和食物
• 不用或減少使用這類有
害化學物品
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多認識我們常用的產品
There is 1% of
Zinc Pyrithione
which is a
highly toxic
pesticide!!!
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有機產品真的安全嗎?
Organic shower gel is not really naturally organic!
環保的選擇
天然肥皂 = 水+油+鹼(NaOH)
https://www.youtube.com/watch?v=
8pnBZmb2WYQ
讓學生自製肥皂、
了解化學及實踐環保
https://www.youtube.com/watch?
v=BORkyqPzsYY
https://www.youtube.com/watch?v=br
WtxZ025aA
https://www.youtube.com/watch?v=13
kb8lm28Xo
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Lets do an experiment (齊來做實驗)
Is synthetic detergent better than simple soap?
化學鹼液真的比天然肥皂好嗎?
Natural Bar Soap
(fat and alkaline)
Synthetic Detergent
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Learning outcomes (學習結果)
• Reduce (amount/frequency) 願意減少用鹼液
• Use environmentally friendly alternatives (sacrificing)
願意付出適度的犧牲金錢和時間去採用天然肥皂
知
其
不
可
而
為
之
知
而
不
行
如
不
知
• If the knowledge cannot translate into appropriate actions, then one still does
not understand the knowledge.
• Even though it is an uphill battle with little chance of success, I still try my
very best to do the right thing.
總結
故事還未完結………
納米
化學物
40
3/2/2015
納米氧化鋅
在納米化學物對人及環境的影響仍
不知時 已被大量使用….. 後果是???
納米氧化鋅
例如港鐵已經大量使用納米銀/二氧化鈦塗
層噴塗來殺菌。
解鈴還須繫鈴人
41
3/2/2015
http://goo.gl/fQ0ptk
你可做甚麼?
•
•
•
•
多認識及了解環境賀爾蒙
不用或減少使用這類有害化學物品
少吃加工及包裝的飲料和食物
將來從事環保工作或相關科研
42
3/2/2015
講座內容大綱
1. 從科學探究例子學習歸納思維：學習如何分
辨海洋生物的性別及探討動物的性別由來
2. 以環境雌激素作例子來說明人工合成化學物
品如何能干擾生物賀爾蒙系統
3. 介紹兩個本地的相關個案
4. 教學經驗分享及總結
5. 問答及討論環節
Contact Email:
kmyleung@hku.hk
43
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Page 1 of 3
Total number of document(s): 1
1 .Ming Pao Daily News | 2015-03-01
Newspaper | S03 | 綠色生活 | By 李佩雯
「藍眼淚」背後的海洋污染
香港的大自然景色和生態有多美，近幾個月見過不少，先是大東山芒草、大棠紅葉、元旦鳳
凰山日出，今年一月，又有在大埔內海出現的自然奇景「藍眼淚」。
美麗的景色卻帶來生態災難，有人留下垃圾，採摘花草，更為拍攝相片而破壞生態。
早前出現的藍眼淚，有人在海上投擲石頭，刺激海藻發光，甚至點火增加相片效果，美麗背
後，除了顯示人們沒公德心，真相是因為海洋受污染了！
最近一周內出現四宗紅潮，也是拜這些藻類所賜。
即使沒採花擲石，原來我們用的洗手液、洗頭水、潤膚露等，當中所含的化學物質，也悄悄
地改變海洋生物的性別，同樣帶來生態災難！
保護珍貴的大自然，不只要有公德心，減用化學品、截污減排同樣重要。
文/ 李佩雯
圖/ 資料圖片、網上圖片
編輯/ 蔡曉彤
污水氮磷引發紅潮 截污減排最奏效
號稱「藍眼淚」的夜光藻，年初在大埔內海出現，吸引無數市民前往拍照。台灣也出現藍眼
淚，有人配合「藍眼淚」的現象，大力推行觀光旅遊，打 「一生不容錯過的生態饗宴」、
「一生必去的台灣自然景點」旗號，吸引市民來體驗藍眼淚生態。諷刺的是，我們眼前看到
的美麗景象，事實上，它在危害生態，而它的出現，與我們不無關係。
夜光藻常見於受農地或家居污水污染的水域出現，研究海洋污染、生態毒理學的香港大學生
物科學學院生態學教授梁美儀表示， 藍眼淚夜光藻（Noctiluca scintillans） ， 屬海藻的一
種，符合幾個條件才能形成，包括水體營養豐富、水體平靜及陽光充足等。
營養過剩海藻爆發
藻類和植物一樣，需要氮磷鉀三種生長要素，而我們排出的生活廢水，含豐富的氮，例如洗
碗水中，含食物殘渣中的蛋白質和氮。海水中氮的比例較多，假如水體富營養化，即氮磷等
植物營養物質含量過多引起水質污染，有條件令海藻爆發，引致紅潮出現。
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這種海藻日間呈啡紅色，晚上受刺激下變藍色，美麗背後，帶來種種害處，它們日間會浮上
水面進行光合作用，假如增生迅速，水體氧氣不足，海藻死亡，在細菌分解作用下，需用更
多水中的氧氣，加劇水底缺氧情 ，令珊瑚、貝類、固定底棲生物遭殃。
「吐露港以前好多珊瑚，因為水體富營養化，引致大量藻類增生，出現紅潮，令水底缺氧，
珊瑚死光，以前由元洲仔對出全部都是珊瑚。」要避免海藻出現，梁教授斬釘截鐵說，「就
是要截污和減排。」污水處理清不了氮磷日常生活中，洗碗、煮食，用了大量水資源，同時
排出大量廢水，污水經喉管排走，勿以為眼不見為淨，污水就此消失。雖然污水處理廠幫手
處理污水，但梁教授表示，污水處理廠只能去除了大部分有機物及沉澱物，但去除不了營
養，令氮和磷的含量依然偏高，最後流出大海。
昔日沙田污水處理廠處理完的水排出吐露港，令吐露港氮和磷的含量頗高，最後加設污水管
道，污水運去九龍灣，再排出維多利亞港，維港水流較急，有助 散。
糞便致病原龍尾灘不宜游泳
早前政府在眾多反對聲音下仍堅持進龍尾人工海灘計畫，梁教授認為， 「想辦龍尾海灘，第
一件事一定要截污。現在龍尾灘不宜游泳，因有由人類而來、與糞便有關的致病原。新界許
多村屋，包括龍尾大尾篤一帶，排污系統是將糞水排入化糞池，表層的水會透過管道滲落泥
土，泥土的細菌幫忙將污染物吸收。但隨 村屋日漸增多，系統已不勝負荷。要截污，龍尾
一帶的村屋不能再用化糞池，改而接駁污水渠，再由污水處理廠處理後才排放。」
香港大學生物科學學院生態學教授梁美儀
少用化學清潔劑
化學合成的清潔劑、洗手液、淋浴乳等，所含的界面活性劑，會被細菌分解成壬基酚，與雌
性荷爾蒙類同，流入大海後影響魚兒的性別。
洗髮沐浴液導致 雄魚變雌
海洋面對的另一生態災難，是環境荷爾蒙。我們常用的洗手液、洗髮水、洗衣液淋浴乳，甚
至潤手霜，普遍含壬基酚類界面活性劑，在自然界中分解後， 成為殘留物壬基酚
（Nonylphenol），含雌性荷爾蒙的功能，會造成雄性魚兒變雌性，精子數量減少而影響繁
殖，最後或會令整個群族滅絕。
梁教授曾做過相關研究，在香港三個污水處理廠排出的水中，檢驗出十二種會影響海洋生物
系統的「內分泌干擾化學物質」（Endocrine Disrupting Chemicals），及找到四種壬基酚及
雙酚A。亦有其他研究，顯示有抗生素殘餘，有機會使水體中的細菌發展出一些基因來抵抗
抗生素。長遠來說，將來或有很多細菌用藥都禁止不了，令人類生命受威脅。
液殘留難分解
此外，人工合成的 液、洗頭水、清潔劑等，因有一條長長的碳鏈（carbon chain），難以
分解而殘留在環境一段長時間。相反，傳統的肥皂以油及鹼性物質製造，細菌能將之分解。
但一般人愛用 液，因其方便又帶有強烈香味，梁教授建議市民可節省使用，「市民有一種
錯覺，以為要用好多 液才洗得乾淨，其實只要擠少許就可以，一來慳錢，二來對環境好。
或將 液、清潔劑、洗髮水等稀釋，由一支變三支，功能都一樣。」
環境荷爾蒙影響免疫力
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環境荷爾蒙又稱為或「內分泌干擾化學物」，具有類似於生物體內荷爾蒙作用， 一旦進入生
物或人類體內，將會產生類似荷爾蒙的影響或干擾原有內分泌系統的平衡及功能，進而對生
物成長、發育與生殖等產生不良影響。
食物安全中心指出，一些流行病學研究顯示，這些化學物可能會對人體健康造成不同的影
響，例如損害生殖能力、神經系統功能和免疫力，以及引致不同類型的癌症。在國際間比較
常出現的內分泌干擾化學物，包括有機氯類除害劑、二噁英和二噁英樣多氯聯苯、雙酚A、
苯乙烯、鄰苯二甲酸酯、有機錫及壬基酚等等。
夜光藻屬甲藻的一種，是微小的單細胞、浮游性生物， 可自由游動。在香港東面、東北面、
南面、東南面、吐露港水域出現。雖然沒有毒性，但與眾多藻類一樣，大量爆發後造成紅
潮，耗用水體氧分。
紅潮
海藻同魚爭氧氣
漁護署資料指出，當海水裏一些微型單細胞、浮游藻類暴發性地大量繁殖，藻類含有的色素
引致海水變色，這一種自然生態現象便稱為「藻華」。因應不同的藻類色素，藻華可令海水
變成粉紅色、紅色、褐色、褐紅色、深綠色或其他顏色，而每年大概有二、三十宗紅潮在香
港水域之內發生。
海藻的出現，除了耗盡水中的氧氣，亦間接令海洋生物死亡，有些藻類更會分泌一些黏液或
毒素，堵塞或破壞魚鰓組織，阻礙魚的呼吸，引致窒息死亡。例如一九八二年，裸甲藻在香
港造成十噸養殖海魚損失；一九九八年，溝凱倫藻在香港水域大規模爆發，因波及大部分養
魚區，超過一半的養殖戶受到影響。
紅潮
當藻類迅速繁殖，便會形成紅潮，消耗水體氧氣，危害海洋生物。
DOCUMENT ID: 201503010040156
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Marine Pollution Bulletin 85 (2014) 352–362
Contents lists available at ScienceDirect
Marine Pollution Bulletin
journal homepage: www.elsevier.com/locate/marpolbul
The occurrence and ecological risks of endocrine disrupting chemicals
in sewage effluents from three different sewage treatment plants,
and in natural seawater from a marine reserve of Hong Kong
Elvis G.B. Xu a, Shan Liu b, Guang-Guo Ying b, Gene J.S. Zheng c, Joseph H.W. Lee d, Kenneth M.Y. Leung a,⇑
a
The Swire Institute of Marine Science and School of Biological Sciences, The University of Hong Kong, Pokfulam, Hong Kong, China
State Key Laboratory of Organic Geochemistry, CAS Centre for Pearl River Delta Environmental Pollution and Control Research, Guangzhou Institute of Geochemistry,
Chinese Academy of Sciences, Guangzhou, China
c
Department of Chemistry, Hong Kong Baptist University, Kowloon, Hong Kong, China
d
Department of Civil and Environmental Engineering, Hong Kong University of Science and Technology, Clear Water Bay, Kowloon, Hong Kong, China
b
a r t i c l e
i n f o
Article history:
Available online 17 March 2014
Keywords:
Endocrine disrupting chemicals
Sewage treatment plant
Marine reserve
Marine protected areas
Environmental risk assessment
Ecotoxicology
a b s t r a c t
We determined the concentrations of 12 endocrine disrupting chemicals (EDCs) in sewage effluents
collected from three different sewage treatment plants (STPs) in Hong Kong, and found 4-nonylphenol
(NP) and bisphenol A (BPA) were the most abundant EDCs. Effluent concentrations of NP and BPA were
higher in dry season than in wet season, but opposite seasonal changes of NP were observed in receiving
waters, probably due to the surface runoff. The two secondary STPs showed higher removal efficiency for
these compounds than the preliminary STP, while having higher removal efficiency in wet season.
Therefore, it is necessary to upgrade the preliminary STP and improve the EDC removal efficiency in
dry season. Seawaters from the Cape D’ Aguilar Marine Reserve adjacent to these STPs also exhibited
elevated NP levels with a hazard quotient >1. Furthermore, diluted effluents from the STPs elicited
significant transcriptional responses of EDC-related genes in the marine medaka fish.
Ó 2014 Elsevier Ltd. All rights reserved.
1. Introduction
Endocrine disrupting chemicals (EDCs) are of global concern and
broadly defined as chemicals that interfere with the normal function of endocrine systems in wildlife and humans (Burkhardt-Holm,
2010). Numerous laboratory experiments indicate that EDCs can
cause negative health effects (e.g. growth, behaviour, reproduction
and immune function) in fishes through disrupting their endocrine
systems (Mills and Chichester, 2005). Estrogenic EDCs can
adversely affect male fishes through induction of vitellogenin and
inhibition of the development of secondary sexual characteristics
at very low exposure concentrations (Länge et al., 2001).
Most EDCs are man-made organic chemicals being introduced
to the marine environment through anthropogenic inputs such as
contaminated sewage effluents and surface runoff. Typical representatives of synthetic EDCs, 4-nonylphenol (NP) and bisphenol A
(BPA) are the major contributors to the endocrine-disrupting activities in aquatic environments (Auriol et al., 2006). NPs are the main
degradation products of alkylphenols polyethoxylates which have
⇑ Corresponding author. Tel.: +852 22990607; fax: +852 25176082.
E-mail address: kmyleung@hku.hk (K.M.Y. Leung).
http://dx.doi.org/10.1016/j.marpolbul.2014.02.029
0025-326X/Ó 2014 Elsevier Ltd. All rights reserved.
been widely used as surfactants in household, agriculture, and
industrial processes (White et al., 1994). At an exposure concentration as low as 10 ppb, NP can cause an increase of vitellogenin
mRNA and a decrease in the growth rate of testes in male rainbow
trout (Lech et al., 1996). BPA is an industrial raw material mainly
used in plastic, rubber, adhesive, and cable industries, and known
to cause a decrease in sperm production in mice (Von Saal et al.,
1998), and lead to a delay in hatching of eggs and a suppression
of growth in juvenile rainbow trout (Aluru et al., 2010). It has been
widely recognized that effluent discharges from sewage treatment
plants (STPs) are the major source of the EDCs to aquatic environments (Zhang and Zhou, 2008). A growing body of research has
indicated that sewage effluents and even their receiving waters
can introduce estrogen-like effects in fishes (Gibson et al., 2005).
There are limited documented studies examining the composition and concentrations of EDCs in sewage effluents and natural
seawaters of Hong Kong (Li et al., 2007; Kueh and Lam, 2008). Li
et al. (2007) discovered that concentrations of NP ranged from 29
to 2591 ng/L in surface water samples collected from Mai Po
Marshes Nature Reserve, northwest of Hong Kong and its levels
were higher in winter (dry season; November and December) than
in late summer (moderately wet season; September and October).
E.G.B. Xu et al. / Marine Pollution Bulletin 85 (2014) 352–362
Kueh and Lam (2008) surveyed the ambient occurrences of selected EDCs, such as nonylphenol and nonlyphenol ethoxylates
(NPEO), in coastal waters, rivers, sediments and biota, and their results suggested that sewage effluents acted as primary sources for
these chemical contaminants. However, little is known about (1)
the composition of EDCs in sewage effluents, and (2) the removal
efficacy of EDCs from raw sewage by different types of STPs in
Hong Kong. Since sewage effluents often comprise of a complex
mixture of EDCs, it is essential to examine the composition of EDCs
in local sewage effluents and identify the dominated chemical contaminants. Furthermore, the seasonal variability of EDC concentrations in STPs and receiving waters in sub-tropical Hong Kong are
still largely unknown. This knowledge is important to decide
appropriate measures for minimizing ecological risks from EDC
emissions to sensitive receivers such as marine reserves in the
marine environment.
EDCs can alter the expression of estrogenic-related (ER) genes,
such as cyp19a and cyp19b, which may result in developmental
and reproductive abnormalities in fishes (Kortner et al., 2009).
EDCs can also cause disruptive endocrine effects through aryl
hydrocarbon receptors (AHRs) and the peroxisome proliferatoractivated receptors (PPARs) in fishes. The AHR pathway mainly
regulates the activation of several genes that encode phase I and
phase II xenobiotic metabolism enzymes, while the PPAR pathway
intermediates receptors and genes involved in the regulation of energy homeostasis, cell proliferation, differentiation and survival
(Fang et al., 2012). In this study, we used embryos and larvae of
the marine medaka fish (Oryzias melastigma) to assess their transcriptional response to sewage effluents and receiving waters,
involving 13 genes in the ER, AHR and PPAR signalling pathways.
There were three main objectives in this study. First, an attempt
was made to quantify, for the first time, the concentrations of 33
common EDCs and identify the most abundant ones in sewage
effluents collected from the three STPs which are located at south
of Hong Kong Island and close to the Cape D’ Aguilar Marine Reserve (Fig. 1). The 33 EDCs include natural and synthetic estrogens,
androgens, progestagens and glucocorticoids. Second, since both
NP and BPA were identified as the most abundant EDCs in our
study, we further monitored the concentrations of NP and BPA in
influents and effluents, as well as in the receiving waters from
the Cape D’ Aguilar Marine Reserve during both dry and wet seasons. Based on the measured concentrations of NP and BPA in
the sewage effluents and receiving waters and their corresponding
predicted no effect concentrations, ecological risks of these two
compounds were assessed. Third, we investigated the effect of
diluted sewage effluents and natural seawaters from the marine
reserve on the mortality, hatching success, and gene expression
in embryos of the marine medaka O. melastigma through a laboratory study. The results would provide supplementary information
for evaluating the current ecological risk of NP and BPA to marine
organisms in the marine reserve.
2. Materials and methods
2.1. Sampling
The three Sewage Treatment Plants (i.e., Shek O, Stanley, and
SWIMS STP) are located in south of Hong Kong Island, Hong Kong,
within 3 km distance to the Cape D’ Aguilar Marine Reserve (Fig. 1).
Shek O STP is a preliminary treatment facility (i.e., screening plant)
designed for removal of suspended matters with a diameter larger
than 6 mm and this plant only treats about 864 m3/day of raw
sewage from Shek O Village District (Table 1; Fig. 2). Stanley STP
is a secondary treatment plant and daily treats about 8100 m3 of
raw sewage from Stanley and Tai Tam Districts. It has the largest
treatment capacity among the three STPs. Its primary treatment
353
consists of a screen and a grit chambers, and its secondary treatment includes an aerobic bioreactor followed by a secondary sedimentation and disinfection (Table 1; Fig. 2). The sludge is returned
from the secondary sedimentation chamber for biological treatment, while the surplus activated sludge is dried before being taken
to landfill. The SWIMS STP, located within the Cape D’ Aguilar Marine Reserve, is designed to mainly treat the wastewater generated
from the Swire Institute of Marine Science (SWIMS) within the marine reserve (Table 1; Fig. 2). As a trickling biofilter treatment plant,
SWIMS STP consists of septic tank, biofilter tank, sand leach tank,
disinfection tank and sedimentation chamber. The three STPs
represent different levels of treatment efficiency: a screening treatment plant (Shek O STP), a secondary biological treatment plant
(Stanley STP), and a trickling filter plant (SWIMS STP).
We sampled both influents and effluents from the three STPs,
and receiving seawaters within the marine reserve three times
during wet season (April, May, and June 2012) and three times during dry season (December 2012, January and February 2013).
Hence, three replicates were used for each STP in each season.
For both influents and effluents, 15-min wastewater samples were
composited at the Stanley STP to prepare 24-h flow-weighted samples, and the 8-h composite samples with 1-h interval were collected at the SWIMS STP. For Shek O STP, influent and effluent
samples were grabbed in triplicates between 11:00 and 13:00,
with a 0.5 h interval between each influent or effluent sample.
Natural seawater samples were collected at 0.5 m below sea surface in the Cape D’ Aguilar Marine Reserve. In investigation of
EDC composition in effluents and bioassay, we grabbed effluent
samples in duplicates from each STP with a pre-cleaned stainless-steel bucket in April 2012. After collection, samples were
transported on ice to laboratory and immediately filtered through
GF/C glass fiber filter papers, and then stored at 4 °C. All samples
were extracted for EDCs within a week. Blank water was taken to
field as control to monitor any contamination during the transport.
2.2. Chemical analysis
Thirty-three phenolic EDCs, including androgens, estrogens,
xenoestrogens, glucocorticoids, and progestagens, were analysed
with RRLC–MS/MS as described by Liu et al. (2011). For each of
the effluent samples, 1 L of water sample was filtered through a
glass fiber filter (Whatman GF/F, 0.7 lm, UK). 100 lL each of
1 mg/L of E1-d4, E2-d4, T-d3, S-d3, CRL-d2 and P-d9 were added
as the internal standards. Solid Phase Extraction (SPE) cartridges
(Oasis HLB, 6 mL and 500 mg each) were preconditioned with
methanol and HPLC water. The filtered water samples passed
through the SPE cartridges at 5–10 mL/min. The target compounds
were eluted using ethyl acetate. The extracts were dried and re-dissolved in 1 mL of methanol for clean-up. The glass cartridge was
filled with glass wool, silica gel and anhydrous sodium sulphate
from bottom to top. Each extract was added to the cartridge, which
was preconditioned with methanol, ethyl acetate/methanol (90:10,
v/v), and hexane. After the cartridge was rinsed with hexane, the
target compounds were eluted with ethyl acetate/methanol. The
eluate was then dried and reconstituted. The target compounds
were analysed by RRLC–MS/MS with EI. Liquid chromatography
was performed on an Agilent 1200 series RRLC system (Agilent
Technologies). The chromatographic separation was performed on
an Agilent Zorbax SB-C18 (100 mm 3 mm, 1.8 lm) column with
pre-column filter (2.1 mm, 0.2 lm). The column oven temperature
was 40 °C and the injection volume was 10 lL. Mass spectrometry
was performed using an Agilent 6460 Triple Quadrupole detector
with ESI in both negative and positive modes (Agilent Corporation,
USA). The quantitative analysis was performed in MRM mode.
The analytical procedure for NP and BPA was based on Zhao
et al. (2009). Briefly, 1 L of each of the influent, effluent or seawater
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E.G.B. Xu et al. / Marine Pollution Bulletin 85 (2014) 352–362
Fig. 1. A map showing the outfall locations of the three sewage treatment plants and the Cape D’ Aguilar Marine Reserve.
samples was filtered through a glass fiber filter (Whatman GF/F,
0.7 lm, UK). Methanol was used to elute non-filterable particles
on the filter and combined with the filtered sample. For solid phase
extraction, 100 lL of 1 mg/L of 4-n-NP were added to each sample
as internal standards. SPE cartridge (Oasis HLB, 6 mL, 500 mg) was
preconditioned with methanol/HPLC water (1:1, v/v). The filtered
water samples passed through the SPE cartridges at 10 mL/min.
The target compounds were eluted from the cartridges using
methanol and DCM. The extracts were dried and then re-dissolved
in methanol. For derivatization, 100 lL of an extract were removed
to a tube and dried. 2 mL of 1 M NaHCO3 solution and 1 mL of 1 M
NaOH solution were added to the tube. 2 mL of n-hexane, 50 lL of
10% pyridine in toluene and 50 lL of 2% PFBOCl in toluene were
added in sequence. After separation, the supernatant of n-hexane
phase was transferred and dried. The final extract was re-dissolved
in n-hexane, which was ready for GC–MS analysis. The MS was
operated in the selected-ion monitoring mode with electron-impact ionization (ionization voltage, 70 eV). The target compounds
were separated by gas chromatography (Agilent 6890N) with a
DB-5MS capillary column (length: 30 m; i.d.: 0.25 mm; coated film
thickness: 0.25 lm). A mass spectrometry (Agilent 5973) was used
as the detector. The oven temperature program was set as follows:
the initial temperature was 70 °C for 1 min, then increased to
170 °C at 20 °C/min, to 230 °C at 6 °C/min, to 280 °C at 12 °C/min
for 6 min, and held at 300 °C for 2 min. The injector was set at
280 °C. The GC–MS interface and the ion-source temperature were
at 280 °C and 250 °C, respectively. Helium was used as the carrier
gas at 1 mL/min. A 1 lL sample was injected in splitless mode with
solvent delay of 3 min. The characteristic ions and retention times
of the target compounds were obtained and identified with full
scan mass spectra from m/z 50 to 500. The compounds of interest
were identified in SIM mode.
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E.G.B. Xu et al. / Marine Pollution Bulletin 85 (2014) 352–362
Table 1
Description of the sewage treatment system and treatment capacity of the three STPs considered in this study (information from personal contact with Drainage Service
Department, HKSAR, in 2012).
STP
Population
Stanley
Shek O
SWIMS
Hydraulic retention
time (h)
BOD
(ton/year)
Suspended
solid
(ton/year)
Sources of the
catchment area
Treatment process
11,600 (design)
8100 (Average daily
flow at present)
21.9 (design)
13.6 at present
9
9
Main areas of Stanley
and Tai Tam Districts
Secondary treatment processes:
1,180 (design)
864 (Average daily
flow at Present)
Not applicable
3
0.5
Treatment capacity
About 27,000 at
present
About 3050 at
present
About 30
Flow (m3/day)
1.
2.
3.
4.
100
80
Main areas of Shek O
village District
Screening of coarse material;
Settlement of grit particles;
Biological treatment of sewage;
Disinfection
Preliminary treatment processes:
Screening suspended matter
with diameter >6 mm and grit
removal.
0.04
0.02
Area of Cape D’ Aguilar
Marine Reserve
Trickling filter treatment processes:
1.
2.
3.
4.
Analyses of blank controls showed no contamination during the
sampling, transportation and extraction processes. The recoveries
of surrogate standard 4-n-NP were more than 65% for all the samples. Recoveries determined were 44.2–127.0% for wastewater
samples and 71.7–113.0% for seawater samples. Ranges of the
limit of detection (LOD) and limit of quantitation (LOQ) for 4-n-NP
were 0.01–0.39 and 0.02–1.63 ng/L for wastewater samples and
0.01–0.24 and 0.03–0.8 ng/L for seawater samples, respectively.
2.3. Medaka embryo-larval bioassay and gene expression
Fertilized eggs of the marine medaka O. melastigma were
cultured and acclimatized in artificial seawater at a salinity of
30‰ and a temperature of 28 ± 1 °C with a 14 h-light/10 h-dark
photoperiod for 2 days. The embryos of 2 day post fertilization
(dpf) were exposed to 1% and 10% (v/v) of effluents from each
STP and surface natural seawaters from the Cape D’ Aguilar Marine
Reserve, and to artificial seawater as the control. The effluents of
different dilutions (i.e., 1:100 and 1:10 dilution) were selected to
represent environmentally relevant concentrations that animals
living close to the outfall of effluent might experience. Each
experimental group contained 50 embryos, which were randomly
distributed into petri dishes containing 30 mL of exposure solution.
The media were daily renewed. Three replicates were conducted
for each experimental group. The mortality rates of the embryos
from 2 to 10 dpf, the hatchability of eggs, and the size of juveniles
after 21-d exposure were recorded. For each replicate, 10 embryos
at 4 dpf, 10 embryos at 10 dpf, and two juveniles at the first fry
stage were collected for quantitative real time polymerase chain
reaction (qRT-PCR) analysis.
The primers of 13 endocrine disruption related genes, including
ERa, ERb, ERc, VTG1, VTG2, AHR, ARNT, cyp1a, cyp19a, cyp19b,
PPARa, and PPARb and PPARc are presented in Table S1 in Appendix A. The procedures of qRT-PCR followed the methods described
in Fang et al. (2012). The embryos or juveniles were randomly collected and homogenized in 1 mL RNA-Solv reagent (Omega) using
a glass homogenizer. The total RNA was extracted from the homogenates using Omega kits according to the manufacturer’s instructions. Equal amounts of RNA were applied to qRT-PCR using
SYBR Premix Ex TaqTM kit (TaKaRa) on a Bio Red CFX 96 Real-Time
System. The PCR thermal profile was as follows: an initial denaturation step at 95 °C for 30 s, followed by 40 cycles at 95 °C for 5 s,
and 60 °C for 34 s, and ending with a dissociation curve analysis.
Septic tank;
Bio-filter tank;
Sand leach tank;
Disinfection and sedimentation
Gene expression levels were normalized to the 18 s rRNA
expression levels. The fold change of the tested genes was analysed
by the 244Ct method (Livak and Schmittgen, 2001).
2.4. Data analysis
Two-way analysis of variance (ANOVA) followed by post hoc
Tukey test was used to examine statistical differences of removal
rates of NP or BPA from wastewater samples among the different
STPs (3 levels) and between different seasons (2 levels). Student’s
t test was used to examine the seasonal differences of concentrations of NP or BPA in natural seawater samples. For the expression
results of each target gene, one-way ANOVA followed by post hoc
Tukey test was used to examine statistical differences amongst different treatment and control groups. For all statistical tests,
p < 0.05 was considered significant.
3. Results
3.1. Composition of EDCs in sewage effluents
Twelve different EDCs were detected in effluent samples from
the three STPs with the mean concentrations ranging between
5.25 ng/L (progesterone) and 4510 ng/L (NP) (Fig. 3). In general,
xenoestrogens showed higher concentrations than other steroids,
and estrogens were detected more frequently than androgens.
Eight common EDCs including NP, BPA, 4-tert-octylphenol, triclosan, triclocarban, estrone (i.e., E1), 4-androstene-3,17-dione, and
progesterone were universally detected in all samples from the
three STPs. The effluent samples of Stanley STP only contained
these eight compounds. In effluent samples obtained from Shek
O STP, three more EDCs namely, 17a-boldenone, androsterone
and testosterone were found. Norgestrel was only detected in
effluent samples collected from the SWIMS STP, in addition to
the eight common EDCs. Glucocorticoid group was not detected
in any sample. Due to varying sources of influents and different
treatment processes of the three STPs, the composition profiles of
target EDCs from the three STPs were slightly different, while on
average over 90% of the total EDC concentration was contributed
by NP and BPA (Fig. 3). To further investigate the removal efficiency of the EDCs detected in STPs, we chose these two most
abundant compounds (i.e., NP and BPA) as the targeted model EDCs
for our subsequent investigations.
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Fig. 2. Schematic diagrams to illustrate various sewage treatment processes at each of the three different sewage treatment plants.
3.2. Removal efficiency of NP and BPA in the three STPs
Both NP and BPA were continually detected in all sewage influent and effluent samples during both wet and dry seasons (Table 2).
Concentrations of NP and BPA in the effluents and influents measured in the current study were within the ranges reported for
other STPs in China (Jin et al., 2008; Nie et al., 2012) and Hong Kong
(Kueh and Lam, 2008). Amongst the three STPs, mean influents
concentrations of NP ranged from 646.3 to 2235.3 ng/L and from
906.9 to 1467.5 ng/L in wet and dry season, respectively (Table 2).
Mean influent concentrations of BPA ranged from 159.3 to
617.2 ng/L and from 454.1 to 1141.2 ng/L in wet and dry season,
respectively (Table 2). Mean effluent concentrations of NP and
BPA were between 441.4–762.5 ng/L and 57.0–159.4 ng/L in wet
season, and between 520.3–1562.0 ng/L and 258.0–713.0 ng/L in
dry season, respectively (Table 2). Clearly, both NP and BPA
showed significantly higher effluent concentrations in dry season
than in wet season (Student’s t test: t P 2.03, df = 52, p < 0.001).
Nonetheless, such a seasonal trend was not observed in the
influent.
The removals of NP and BPA in the STPs were dependent on the
type of the sewage treatment facility and also dependent on the
season (Table 2). There was a significant interaction of STP type
and season (two-way ANOVA: F2, 48 = 5.52, p < 0.05). The mean removal rates for NP and BPA were 13.0% and 42.5% at Shek O STP,
69.2% and 73.0% at Stanley STP, and 61.4% and 53.7% at SWIMS STP,
respectively. Higher removal rates were observed at the two
biological treatment plants (i.e. Stanley and SWIMS STPs) in wet
season than in dry season (Table 2). We also found significantly
higher removal efficiencies of NP and BPA in the two biological
treatment plants than in the preliminary treatment plant at Shek
O (two-way ANOVA: F2,48 = 103.59, p < 0.01; Table 2).
3.3. NP and BPA in the receiving environment
NP and BPA were also detected within the Cape D’ Aguilar
Marine Reserve, which is relatively close to the three STPs
(Fig. 1; Table 2). The mean concentrations of NP and BPA were
392.5 and 64.5 ng/L in wet season, and 109.4 and 69.5 ng/L in
dry season, respectively. Interestingly, the mean concentration
of NP in the marine reserve was significantly higher in wet
season than that in dry season (Student’s t test: t = 7.70, df = 45,
p < 0.001) and this was opposite to the seasonal pattern observed
for sewage effluents. In contrast, the concentration of BPA varied
insignificantly between the two seasons.
3.4. Implications from the medaka embryo-larval bioassay
The results of mortality, hatchability and growth of the marine
medaka embryos after 21 days of exposure to effluents and
seawaters are summarized in Fig. 4. During exposure from 2 to
10 dpf, mortalities of the embryos were significantly higher in
the treatments of SO10% (10% effluent from Shek O STP) and
SW10% (10% effluent from SWIMS STP) than in the control groups,
with average mortality rates of 59% and 18%, respectively
(One-way ANOVA: F7,16 = 328.00, p < 0.05). A significantly lower
E.G.B. Xu et al. / Marine Pollution Bulletin 85 (2014) 352–362
hatchability of embryos was also observed in SO10% and SW10%
treatment groups, having 13% and 54% hatching success, respectively (F7,16 = 328.00, p < 0.05). The effects of the effluents on
growth rate were assessed by measuring the interorbital distance
(i.e., width) and the total length of larvae after 21-day exposure.
We observed no significant difference of interorbital distance
among all treatment groups, however, juveniles exposed to
ST10% (i.e., 10% effluent from Stanley STP) and SO1% (i.e., 1% effluent from Shek O STP) were significantly smaller than the control
fish in terms of their total length (F6,41 = 5.22, p < 0.05). Effects of
the natural seawater treatment on mortality, hatchability and
growth of medaka were not significantly different when compared
with those in the artificial seawater control group. Overall, the
results indicated that the 10 times diluted effluents from the STPs
which were contaminated with various EDCs negatively affected
the health of medaka embryos.
The genes related to endocrine disruption responded differently
to different exposures (i.e., effluents with different dilutions, seawater collected from the marine reserve, and artificial seawater;
see Figs. S1–S3 in Appendix A), yet some general trends were
observed as follows. Firstly, magnitudes of the mRNA expression
in the medaka fish responding to the different treatment
groups generally followed an order as: diluted effluents > natural
seawaters > artificial seawater. For example, exposure to natural
seawaters from the marine reserve caused no significant alteration
in expression of VTG1 in 4 dpf fish, but VTG1 was significantly upregulated by 6-fold in 4 dpf fish after exposure to ST1% treatment
(One-way ANOVA: F7,16 = 41.99, p < 0.05). The largest fold change
amongst all gene expressions was observed in 4 dpf fish after exposure to Shek O effluents, with 50-fold of up-regulation in cyp1a
(Fig. 5a). Shek O effluents also elicited significant up-regulations
of VTG1 and VTG2 at the first fry stage (F6,14 P 46.15, p < 0.05;
Fig. 5b and c). In contrast, the mRNA expression levels of ERb,
ARNT, cyp19b, PPARa, and PPARb were all significantly down-regulated (F6,14 P 13.20, p < 0.05), while no significant difference was
observed for ChgH, ChgL and PPARc (see Figs. S1–S3 in Appendix
A). At 10 dpf, a significant up-regulation was observed for cyp1a
(F6,14 = 129.95, p < 0.05) (Fig. 5d), while ERc, VTG1, VTG2, ChgH,
ChgL, ARNT, cyp19b, and subtypes of PPAR were significantly
down-regulated (F6,14 P 5.91, p < 0.05). At the first fry stage, ChgL,
ARNT, cyp1a, cyp19a, cyp19b, PPARa, and PPARb were all significantly down-regulated (F6,14 P 15.66, p < 0.05).
For fish exposed to the natural seawater collected from the
marine reserve, their expression profiles for certain genes were
significantly different from those of the control animals. The genes
357
of ERb, cyp19b, ARNT and PPARb were significantly downregulated for fish embryos at 4 dpf after exposure to the natural
seawater (Fig. S1). For fish embryos at 10 dpf, the genes ERa, ERb
and ChgH were significantly down-regulated whereas cyp19a gene
was significantly up-regulated (Fig. S2). The genes of ERb, cyp19a,
cyp19b, ChgH, ChgL, PPARa, PPARb were all significantly downregulated in the fish at the first fry stage (Fig. S3). The pattern of
down-regulation of various genes was somewhat similar to those
exposed to diluted sewage effluents.
4. Discussion
4.1. Composition of EDCs in sewage effluents
Our results agree with a previous study that NP and BPA are the
mostly detected phenolic EDCs in sewage effluents and their concentrations are consistently higher than the other EDC compounds
(Wang et al., 2012). The total unit loads for NP and BPA were calP
culated based on the following equation: ULt = 365 Cifi, where
ULt is total EDC unit load (g/year) of the three STPs, Ci is the mean
EDC concentration in effluent from each STP, and fi is the mean
flow of each STP (m3/day). Based on this rough estimation, about
3500 g of NP and 1300 g of BPA would be discharged into the marine environment from the three STPs every year. We found the
highest concentrations of most EDCs (7 out of 12 samples) from
Shek O effluents, suggesting that it is an important source of EDC
pollution in this area. In Shek O effluents, for example, the mean
concentrations of NP (3595.03 ng/L) and E1 (24.14 ng/L) were
around one order of magnitude higher than their corresponding
predicted no effect concentration (PNEC) at 330 ng/L (EC, 2001)
and 3 ng/L (UKEA, 2002), respectively. The occurrence of the EDCs
at such high concentrations might have disrupted the endocrine
system in marine organisms living nearby the sewage outfall of
Shek O STP.
4.2. Removal efficiency of NP and BPA in the three STPs
The main objective of wastewater treatment systems in Hong
Kong is to remove organic substances, phosphorus, nitrogen,
ammonia and E. coli from wastewater. Although EDCs can also be
reduced by STPs, incomplete removal of EDCs would be largely
attributable to the processes of the STPs (i.e., physical, chemical
and/or biological treatment) and operational conditions (e.g. retention time of the sewage) (Birkett and Lester, 2003). In the present
study, the three STPs were chosen to represent different types of
Fig. 3. Composition profiles of the target EDCs in effluent samples collected from each of the three sewage treatment plants.
258.0 (65.7–395.2)
64.5 (11.0–407.5)
Not applicable
69.5 (25.1–243.7)
Not applicable
57.0 (44.7–81.0)
267.5 (147.7–472.7)
159.3 (40.8–342.3)
64.2b
454.1 (343.2–576.7)
43.2b
159.4 (64.0–265.4)
713.0 (356.4–1035.3)
617.2 (262.3–915.6)
74.2b
1014.6 (243.2–2262.1)
73.6b
140.9 (61.7–373.0)
537.2 (249.4–930.1)
1562.0 (391.7–2916.1)
Wet season
Removal rate (%)
Dry season
Removal rate (%)
BPA (ng/L)
268.7 (35.2–1204.1)
47.6a
1141.2 (404.2–1836.7)
37.5a
520.3 (176.3–747.7)
441.4 (246.8–623.5)
2235.3 (448.5–4443.9)
80.3b
906.9 (401.9–1220.8)
42.6b
459.2 (185.8–1167.5)
1835.9 (1526.9–2042.6)
75.0b
1467.5 (772.8–1818.8)
63.4b
762.5 (541.9–1144.2)
Influent
646.3 (530.0–959.5)
18.0a
1446.9 (469.5–2465.0)
8.0a
Wet season
Removal rate (%)
Dry season
Removal rate (%)
NP (ng/L)
Influent
Stanley STP
Influent
Effluent
Shek O STP
Effluent
SWIMS STP
Effluent
Seawater
392.5 (139.1–496.8)
Not applicable
109.4 (60.8–327.9)
Not applicable
E.G.B. Xu et al. / Marine Pollution Bulletin 85 (2014) 352–362
Table 2
Summary of mean concentrations and their ranges (in brackets) of NP and BPA in influents and effluents from the three STPs and in seawaters from the Cape D’ Aguilar Marine Reserve. The mean removal rates of NP and BPA from each of
the STP are given as bold numbers. For removal rates at the same row, the rates with different letters are significantly different (ANOVA followed by post hoc Tukey test: p < 0.05).
358
treatment processes (Fig. 2). The mechanical treatment consisted
of two steps were used in all the investigated STPs. Screening
was to remove objects such as rags and pieces of over 6 mm in
diameter. The sewage then was passed through a sand trap where
main solid organic material with lipophilic compounds would settle. We also detected NP and BPA at micrograms per gram dry
weight levels in solid samples collected in the sand trap (data
not shown). Stanley STP represented the activated sludge treatment, which was the most common type of biological treatment
in Hong Kong. The sewage was pumped into large open-air basins
containing suspended bacteria, where degradation or transformation of EDCs occurred (Auriol et al., 2006). Mixing was carried
out by aeration, and the residence time for the treatment was
13.5 h. Trickling filters was used at the SWIMS STP. After settled
in the septic tank, sewage was passed through a system of meandering biofilter tank. Such tank provided varying depths to create
aerated and anoxic zones for biodegradation or transformation of
EDCs. The residence time for the SWIMS STP was only about
0.5 h. However, such a short residence time can be a negative influence on the EDC removal (DEPA, 2003).
The removal efficiency of BPA found in the present study (see
Table 2) was in agreement with literature values (Auriol et al.,
2006). The removal efficiency of NP in Shek O STP was much lower
than the reported literature values of up to 99% (Auriol et al., 2006).
This may be attributed to the fact that Shek O STP is only a preliminary treatment plant, and the degradation of NPEO in the sewer
may increase the concentration of NP in the final effluent. In contrast, the removal efficiency of NP at Stanley and SWIMS STPs
was high and consistent with the literature data (Auriol et al.,
2006). Besides, the SWIMS STP, which applies biological filter technology, was generally less efficient in removing NP and BPA when
compared with the activated sludge process adopted at Stanley
STP. This could be due to a short hydraulic retention time in the
biofilter STP (Clara et al., 2005). Svenson et al. (2003) also observed
that the activated sludge process showed higher estrogenic removal (81%) than trickling filters (28%). Therefore, it is necessary
to upgrade the treatment facilities at Shek O and SWIMS STPs to
better remove EDC residues from the raw sewage.
Both NP and BPA showed significantly higher effluent concentrations in dry season than in wet season. Such seasonal patterns
can be partially explained by the dilution effect associated with
high rainfall and elevated water utilization during the summer,
wet season in Hong Kong. The flow of sewage (m3/day) was
approximately 30% and 115% higher during the samplings in wet
season at Stanley and Shek O STP, respectively. Wang et al.
(2010) also reported two to fivefold lower concentrations of 7 selected EDCs from 3 STPs in wet season than in dry season, which
was mainly due to the dilution by rain water. Furthermore, when
temperature is high during the summer, wet season, both NP and
BPA may undergo a faster physicochemical and biological degradation, and perhaps STPs with biological treatment may have higher
removal efficiency for these compounds in summer due to higher
microbial metabolisms and activities (Manzano et al., 1999). Ko
et al. (2007) and Nie et al. (2012) also demonstrated that concentrations of NP and BPA in the effluents were higher in winter and
spring than in summer and autumn, which was closely related to
the microbial activity and concentrations of mixed liquor suspended solids.
Higher removal rates for NP and BPA were observed at Stanley
and SWIMS STPs in wet season than in dry season. As discussed
above, the faster degradation rates for NP and BPA in summer were
probably attributable to the presence of more active microorganisms (e.g. more arylsulfatase enzymes) at higher temperature
(Manzano et al., 1999). In the present study, the average temperature of sewage was 16 °C during the samplings in dry season and
that was 26 °C in wet season. Some researchers also reported
E.G.B. Xu et al. / Marine Pollution Bulletin 85 (2014) 352–362
Fig. 4. Average mortality, hatchability and growth of O. melastigma after exposure
to treatments with different proportions of effluents and to the natural seawater
collected from the marine reserve [ST: Stanley STP; SO: Shek O STP and SW: Swire
Institute of Marine Science STP]. Mean and SD are shown; ⁄ denotes a significant
difference between the treatment and control groups, p < 0.05.
varying concentrations and different removal rates of EDCs in different seasons, and concluded that temperature can affect the biodegradation process to a large degree (Jin et al., 2008). Lian et al.
(2009) investigated the fate of NPEO and their metabolites
(including NP) in 4 STPs. They found that NP was not only
biodegraded but was also produced from its parent compound
(NPEO) during wastewater treatment process, and higher removal
efficiencies of both NP and NPEO in summer were probably due to
high temperature.
4.3. NP and BPA in the receiving environment
The recent knowledge of NP and BPA occurrence and behaviour
in river water is well established, but such knowledge in marine
environments is still limited (Sharma et al., 2009). The measured
concentration range of NP in the present study was slightly higher
than that reported in marine waters of Hong Kong by Kueh and
Lam (2008), which ranged from less than 100 up to 270 ng/L. For
BPA, although no data is available for the comparison with our
study in the marine environment of Hong Kong, our reported
BPA concentrations are one order of magnitude lower than those
quantified in the Pearl River, China (Gong et al., 2012).
The higher concentrations of NP in summer, wet season in the
receiving waters could be caused by several factors. Firstly, this
could be due to high temperature and associated microbial activity,
leading to an enhanced degradation of NPEO in marine sediment
and hence an increased NP concentrations in water column during
summer (Li et al., 2004). Levels of NP in both natural surface water
and in suspended particles were found decreasing with decreasing
water temperature (Xu et al., 2006). Fu et al. (2007) also reported
that a higher concentration of NP in summer in coastal waters of
359
Qingdao, China, was mainly due to the higher degradation rate of
NPEO and re-suspended sediment under strong wind. Some phenomena might also occur in the Cape D’ Aguilar Marine Reserve,
though further investigation would be required.
It is also postulated that the other major reason for the elevated
NP concentrations in natural seawater during wet season may be
attributable to an increased input of NP from the increased surface
runoff (Zhao et al., 2009). In the present study, surface runoff and
storm water discharges may have also increased the input of NP
during wet season. Kueh and Lam (2008) found the storm water
in Hong Kong containing 80–12,000 ng/L of NP and 260–
29,200 ng/L of NPEO. To better understand its environmental
behaviour, the concentrations of NP should be further measured
in suspended particle, sediment, and surface runoff samples collected around and within the marine reserve. Investigations should
also involve its parent compounds (e.g. NPEO) and its major aerobic metabolites (e.g. NP1EC, NP2EC) and anaerobic metabolites
(e.g. NP1EO, NP2EO) (Ahel et al., 1994).
Seasonal variation of BPA in the seawaters of the marine reserve
was not observed in the present study. The level of BPA in seawater
might have reached an equilibrium condition of leaching from
chemical products (e.g. epoxy and polycarbonate plastics), solution
in seawater, sorption to suspended particles, incorporation into organic matters, aerobic degradation with hydroxyl radicals,
bioaccumulation by marine organisms, and mineralization by
bacteria (Cousins et al., 2002). These processes can be affected by
various factors, including seawater temperature, pH, inorganic ions
and phytoplankton in seawater, and reactive oxygen species (Sajiki
and Yonekubo, 2002). However, Patrolecco et al. (2006) found
different seasonal patterns of EDCs in different rivers in Italy, and
concluded that the levels of BPA in aquatic compartments were
affected by differences in hydrological conditions between different sampling campaigns, and that the process of re-suspension
and re-dissolution from sediment was an important source of
EDCs. Thus, the environmental fate of BPA and NP in the Cape D’
Aguilar Marine Reserve should be further investigated in more
detail.
The ecological risks from the NP and BPA in the marine reserve
were assessed using the risk quotient (RQ) approach, i.e., a ratio
between the measured environmental concentration (MEC) and
PNEC (RQ = MEC/PNEC). The RQ must remain below 1 to ensure
an acceptable risk to the environment (EU, 1994). Given that the
proposed PNECs of NP and BPA were 330 and 150 ng/L for marine
water (EC, 2001; EU, 2010), mean RQs for NP and BPA in the marine
reserve were calculated as 1.1 and 0.4 in wet season, and 0.4 and
0.5 in dry season, respectively. The risk of BPA was consistently
low, but that for NP was high with the RQ exceeding 1 during
wet season. Therefore, organisms in the marine reserve were likely
adversely influenced by the elevated level of NP during wet season.
4.4. Implications from the medaka bioassay
The overall changes in expression of various genes in
O. melastigama after exposure to the diluted sewage effluents are
summarized in Fig. 6. The mRNA expression profiles were dependent on the developmental stages of medaka embryos and effluent
concentrations in a gene-subtype-specific manner. For instance,
VTG1 and VTG2 were significantly up-regulated at 4 dpf but then
inhibited after 6 days of further exposure to the effluents (i.e.,
10 dpf). PPARa and PPARb exhibited similar expression profiles in
the fish with significant down-regulation upon the exposure
regardless of the developmental stage, but PPARc did not change
in its expression among all developmental stages and among the
treatments. Overall, more genes (11 out of 13) were up or down
regulated at the late embryonic stage (10 dpf) than those at the
early embryonic stage (4 dpf with 7 genes) and the 1st fry stage
360
E.G.B. Xu et al. / Marine Pollution Bulletin 85 (2014) 352–362
Fig. 5. Examples showing mean expression levels (i.e., relative fold changes) of selected genes at different developmental stages of O. melastigma: (a), cyp1a at 4 dpf; (b),
VTG1 at 1st fry stage; (c), VTG2 at 1st fry stage; and (d), cyp1a at 10 dpf after exposure to the control (artificial seawater), five different effluent samples (A: ST10%, B: ST1%, C:
SO10%; D: SO1%, F: SW10% and G: SW1%, respectively) and a natural seawater sample obtained from the marine reserve (E) [ST: Stanley STP; SO: Shek O STP and SW: Swire
Institute of Marine Science STP]. The data in triplicate are presented as the mean and SD, relative to the control; ⁄p < 0.05.
Fig. 6. A summary of the gene expression profile of O. melastigma at 4 dpf, 10 dpf and 1st fry stage, respectively, after exposure to diluted sewage effluents. The relative
expression levels of genes in the treatment vs. the control were indicated as follows: significant up-regulation, up arrow; significant down-regulation, down arrow; no effect,
horizontal line, based on the results from one-way ANOVAs at p < 0.05.
(with 7 genes). These results indicated that the expression of subtypes of ER or PPAR genes were dependent on the developmental
stage of the fish, which was consistent with results reported from
other studies (Seo et al., 2006; Cocci et al., 2013). However, cyp19b
mRNA expression levels in the medaka embryo and juvenile were
significantly reduced at all developmental stages. Similarly, NP also
exhibited potent inhibitory effects on cyp19 genes and
significantly reduced brain aromatase activity in Atlantic salmon
(Kortner et al., 2009). Reduced ovarian aromatase activity in red
mullet was suggested to be also caused by NP (Martin-Skilton
et al., 2006). On the contrary, NP was found to induce cyp19A2
gene in dose-dependent manner in zebrafish juveniles (Kazeto
et al., 2004), and cause significant induction of cyp19 isomers in
immature Atlantic salmon (Meucci and Arukwe, 2006). The differential abundance and expression of cyp19 genes in different fish
species after exposure to estrogenic compounds have been
reported previously (Trant et al., 2001; Cheshenko et al., 2008).
Cytochrome aromatase as well as estrogen receptor genes isotypes
showed differential organ-specific, NP and BPA concentration- and
time-dependent expression patterns after exposure to environmental relevant concentrations of NP and BPA (Lee et al., 2006).
5. Conclusion
In this study, we first screened 33 common EDCs and found that
there were twelve EDCs present in effluents from three STPs
located at south of Hong Kong Island, and NP and BPA were the
most abundant EDCs. Afterwards, this study comprehensively
investigated the occurrence, seasonal variation and biological effects of NP and BPA in influent and effluent samples collected from
the three STPs, and in seawaters collected from the Cape D’ Aguilar
Marine Reserve adjacent to these STPs. We discovered that concentrations of NP and BPA in influents were comparable to those in
effluents from the preliminary STP in Shek O, indicating its poor removal efficiency for these compounds. In contrast, concentrations
of the two compounds were significantly decreased at Stanley
and SWIMS STPs following more efficient biological treatments.
Effluent concentrations of NP and BPA were higher in dry season
than in wet season, but opposite seasonal changes of NP were
observed in receiving waters (i.e., the Cape D’ Aguilar Marine
Reserve), probably because of the increased input of NP from the
increased surface runoff during the wet season. Our results also
showed that Stanley STP using an activated sludge process was
E.G.B. Xu et al. / Marine Pollution Bulletin 85 (2014) 352–362
more effective to remove NP and BPA from wastewater than the
biological filter adopted at SWIMS STP. Lower removal rates were
observed at these two biological STPs in dry, winter season than
in wet, summer season, suggesting that the EDC removal process
is temperature dependent.
Natural seawater samples taken from the marine reserve also
exhibited elevated levels of NP with a risk quotient greater than
one in wet season, indicating potential hazards of this compound
to marine organisms. In addition, our laboratory experiment further confirmed that diluted effluents from the three STPs and natural seawaters from the marine reserve can elicit transcriptional
responses of genes related to endocrine disruption pathways in
the marine medaka fish. Overall, our results demonstrated that
sewage effluents can act as the major source for the continuous input of estrogenic compounds into the marine environment. The
existing sewage treatment facilities at Shek O and SWIMS STPs
should be upgraded as a means to reduce the discharge of EDCs
into the marine environment and hence lower their ecological risks
to marine organisms living in the receiving waters including those
inhabiting the Cape D’ Aguilar Marine Reserve.
Acknowledgements
This work is jointly supported by the Area of Excellence (AoE)
Scheme under the University Grants Committee of the Hong Kong
Special Administration Region (HKSAR), China (Project No. AoE/P04/2004), and by a research grant from the Swire Educational Trust.
The authors thank the Drainage Services Department of the HKSAR
Government for granting us a permission to collect the sewage
influent and effluent samples for this study. Elvis Xu would also like
to thank John Swire & Sons Limited and the Swire Educational Trust
for providing him a James Henry Scott (Hong Kong) PhD Scholarship. The authors also thank Andy Yi and Karen Villarta for their
valuable comments on early drafts of this manuscript, and staff
and postgraduate students at the Swire Institute of Marine Science
for assisting this project. The authors are grateful to the Agriculture
Fisheries and Conservation Department for granting a permit for
taking samples from the Cape D’ Aguilar Marine Reserve.
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.marpolbul.2014.
02.029.
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Marine Pollution Bulletin 91 (2015) 128–138
Contents lists available at ScienceDirect
Marine Pollution Bulletin
journal homepage: www.elsevier.com/locate/marpolbul
Environmental fate and ecological risks of nonylphenols and bisphenol A
in the Cape D’Aguilar Marine Reserve, Hong Kong
Elvis G.B. Xu a, Brian Morton a, Joseph H.W. Lee b, Kenneth M.Y. Leung a,⇑
a
b
The Swire Institute of Marine Science and School of Biological Sciences, The University of Hong Kong, Pokfulam, Hong Kong, China
Department of Civil and Environmental Engineering, Hong Kong University of Science and Technology, Clear Water Bay, Kowloon, Hong Kong, China
a r t i c l e
i n f o
Article history:
Available online 2 January 2015
Keywords:
Endocrine disrupting chemicals
Yeast estrogen screen assay
mRNA assay
Marine reserve
Ecological risk assessment
a b s t r a c t
Nonylphenols (NPs) and bisphenol A (BPA) are the most common endocrine disruptors detected in the
coastal waters of Hong Kong. The Cape D’Aguilar Marine Reserve (CAMR), the only marine reserve in
Hong Kong is close to urbanized areas, thus the resident marine organisms are inevitably influenced
by partially treated wastewater from adjacent sewage treatment plants (STPs). Elevated levels of NPs
and BPA were detected in all seawater, sediment and biota samples collected from the CAMR. Estrogenic
activities of seawater from the CAMR, and sludge and sewage from a nearby STP were assessed using
yeast estrogen screen assay. We found aromatase, estrogen receptor and vitellogenin genes in the marine
medaka fish Oryzias melastigma were significantly up-regulated after exposure to the reserve’s seawater.
According to a tissue-residue-based probabilistic risk assessment, the marine species living in the CAMR
are having 35% and 21% of chance to be at risk due to exposure to NPs and BPA, respectively.
Ó 2014 Elsevier Ltd. All rights reserved.
1. Introduction
Cape D’Aguilar is situated on the southeastern tip of Hong Kong
Island in the Hong Kong Special Administrative Region (SAR) of
China. In 1996, the Hong Kong SAR Government, in recognition
of the importance and urgency of protecting the geomorphological
and ecological environment of Cape D’Aguilar, designated this area
as the first and, to date, only marine reserve in Hong Kong (Fig. 1).
The marine reserve has a rich marine biodiversity, with numerous
habitats such as intertidal sea arches and subtidal caves (zawns),
hermatypic and ahermatypic coral reefs, exposed and sheltered
rocky beaches, cobble beaches, intertidal pools of varying dimensions and elevations and protected under-boulder landscapes,
resulting in an associated rich marine biodiversity (Morton and
Harper, 1995). Owning to its small size (0.2 km2) and location close
to sources of human activities, the marine reserve is considered
vulnerable to chemical contaminants from point sources such as
nearby sewage treatment plants (STPs) (Xu et al., 2014, 2015).
These discharges were identified as major vectors for the entry of
toxic and endocrine disrupting chemicals (EDCs) such as nonylphenols (NPs) and nonylphenol ethoxylates (NPEOs) into receiving
waters in Hong Kong (Kueh and Lam, 2008). Li et al. (2007)
⇑ Corresponding author. Tel.: +852 22990607; fax: +852 25176082.
E-mail address: kmyleung@hku.hk (K.M.Y. Leung).
http://dx.doi.org/10.1016/j.marpolbul.2014.12.017
0025-326X/Ó 2014 Elsevier Ltd. All rights reserved.
discovered that concentrations of NPs ranged from 29 to
2591 ng/L in surface estuarine water samples collected from the
Mai Po Marshes Nature Reserve in northwestern Hong Kong. There
are increasing concerns over the adverse effects that may result
from the exposure of wildlife to these EDCs. Numerous laboratory
experiments indicate that the EDCs may cause negative health
effects (e.g. growth, behaviour, reproduction and immune function) in fishes through disrupting their endocrine systems (e.g.
Länge et al., 2001; Marta and Charles, 2012).
Of particular concerns are NPs and bisphenol A (BPA), which
have been identified as major anthropogenic contributors to endocrine-disrupting activities in aquatic environments (Auriol et al.,
2006). NPs are the main degradation products of alkylphenol polyethoxylates, which have been used widely as surfactants in household detergents, agriculture and the dyeing industry (White et al.,
1994). NPs can cause an increase of vitellogenin mRNA and a
decrease in the growth rate of testes in male rainbow trout at a
concentration of 1 lg/L (Lech et al., 1996; Bonefeld-Jørgensen
et al., 2007). BPA is an industrial raw material mainly used in plastic, rubber, adhesive, and cable industries, and is known to cause a
delay in the hatching of eggs and a suppression of growth in juvenile rainbow trout (Aluru et al., 2010; Bonefeld-Jørgensen et al.,
2007). Due to concerns about the possible impacts of NPs and
BPA on aquatic organisms, the levels of these chemicals have been
quantified in different aquatic systems around the world with
E.G.B. Xu et al. / Marine Pollution Bulletin 91 (2015) 128–138
concentrations ranging from a few ng/L to hundreds lg/L (Isobe
et al., 2001; Mayer et al., 2007; Lee et al., 2013). In parallel with
chemical analysis, in vivo and in vitro bioassays have been applied
to assess the integral estrogenic activity in environmental samples
(Sonneveld et al., 2006). Estrogenic activities derived from bioassays have been reported upon in different environmental matrices
from coastal marine environments of various countries (Ra et al.,
2011; Cargouet et al., 2004; Vermeirssen et al., 2005).
In a previous study, we found that effluents from STPs in Hong
Kong contained significant concentrations of varying EDCs, with
NPs and BPA being the most abundant (Xu et al., 2014). Little is
known, however, about the occurrence and distribution of the
two compounds in the coastal waters of Hong Kong. Cape D’Aguilar
Marine Reserve (CAMR) is an important habitat and spawning
ground for a highly diverse assemblage of marine organisms, but
there are no data concerned with the bioaccumulation of EDCs in
the various species the reserve protects. Despite the previous
detection of NPs and BPA at hundreds of ng/L in waters collected
from the marine reserve, overall estrogenic activities in the waters
and sediments of the marine reserve remain unknown.
The objectives of this study were: (1) to investigate the occurrence of NPs and BPA in Hong Kong’s coastal waters; (2) to analyse
the concentrations of NPs and BPA in surface seawater, sediment,
and tissues of eleven marine organisms sampled from the CAMR
in both the wet (summer) and dry (winter) seasons and (3) to
assess the estrogenic activities of NPs, BPA and their admixture
at environmentally realistic concentrations, as well as different
environmental samples collected from the CAMR using in vitro
yeast estrogen screen (YES) assay and in vivo mRNA assays with
the marine medaka fish Oryzias melastigma. This study provides a
better understanding of the occurrence and ecological risks of EDCs
in the Hong Kong coastal waters and, in particular, the Cape
D’Aguilar Marine Reserve.
2. Materials and methods
2.1. Sampling
Surface seawater samples were collected from the east, south
and west coasts of Hong Kong in August and October 2012
(Fig. 1). Seawater, sediment, and biota samples (i.e., of the algae,
snails, mussels, sea urchins, sea cucumbers, shrimps, crabs and
fishes) were collected from the CAMR in January and August
2013 (see Fig. 1). Water samples were taken from 1.5 m below
the surface. About 50 ml of methanol was immediately added to
1 L water sample to suppress biological activities and the pH
adjusted to 3–4 using 4 M H2SO4 on the research boat SWIRE Asterina. Water samples were transported on ice to the laboratory and
stored at 4 °C in darkness. The samples were subjected to chemical
analysis within 24 h. Surface sediment samples (up to the top
20 cm) were collected using a stainless steel grab sampler, scooped
into glass jars and stored at 20 °C until analysis. Wastewater and
sludge samples were collected from the Marine Reserve STP in
August 2013. All sediment tissue and sludge samples were
freeze-dried and stored at 20°C until analysis.
2.2. Water quality measurement in the CAMR
A Water Quality Monitoring Buoy System (EMM68, Yellow
Springs, USA) equipped with sensors for measuring water
temperature, salinity, dissolved oxygen, conductivity, pH, turbidity, chlorophyll a and phycocyanin was deployed in the CAMR
(see Fig. 1). The results of the monitored water quality parameters
during the sampling time (January and August 2013) are given in
Table S1 (Appendix A).
129
2.3. Chemical analysis
The analytical procedure for the NPs and BPA was based on
Zhao et al. (2011). Briefly, 1 L each of the influent, effluent and seawater samples was filtered through a glass fibre filter (Whatman
GF/F, 0.7 lm, UK). Methanol was used to elute non-filterable particles on the filter and these were combined with the filtered sample. For solid phase extraction (SPE), 100 lL of 1 mg/L of 4-n-NP
was added to each sample and served as internal standards. SPE
cartridges (Oasis HLB, 6 mL, 500 mg) were preconditioned with
methanol/HPLC water (1:1, v/v). The filtered water samples were
passed through the SPE cartridges at 10 mL/min. The target compounds were eluted from the cartridges using methanol and
dichloromethane. The extracts were then dried and re-dissolved
in methanol. Target compounds in the sediment, sludge, or tissue
samples were extracted by ultrasonic-assisted solvent extraction.
Five grams of the prepared samples were mixed with ethyl acetate
(10 mL). The tube was ultrasonicated for 15 min and centrifuged at
1370g for 10 min, and the supernatant collected. The sediments
were extracted twice more with 10 mL and 5 mL of ethyl acetate,
respectively. The supernatants were combined and concentrated
to about 1 mL on a rotary evaporator. The extract was further purified with a glass column loaded with 1 g of silica gel. Elution was
carried out using ethyl acetate. The eluted sample was concentrated to near dryness under a gentle nitrogen stream and re-dissolved in 1 mL methanol for further treatment. For derivatization,
100 lL of an extract was removed to a tube and dried. Prior to
2 mL of 1 M NaHCO3 solution and 1 mL of 1 M NaOH solution being
added into the sample, 2 mL of n-hexane, 50 lL of 10% pyridine in
toluene and 50 lL of 2% PFBOCl in toluene were added in sequence.
After separation, the supernatant of the n-hexane phase was transferred to a vial and dried. The final extract was re-dissolved in hexane for gas chromatography-mass spectrometry (GC–MS) analysis
or in DMSO for yeast estrogen screen (YES) assay. The target compounds were separated by the GC (Agilent 6890N) with a DB-5MS
capillary column (length: 30 m; i.d.: 0.25 mm; coated film thickness: 0.25 mm). The MS (Agilent 5973) was used as the detector
and operated in the selected-ion monitoring mode with electronimpact ionization (ionization voltage, 70 eV). The oven temperature programme was as follows: the initial temperature was
70 °C for 1 min, then increased to 170 °C at 20 °C/min, to 230°C
at 6 °C/min, to 280 °C at 12 °C/min for 6 min, and held at 300 °C
for 2 min. The injector was set at 280 °C. The GC–MS interface
and the ion-source temperatures were at 280 °C and 250 °C,
respectively. Helium was used as the carrier gas at 1 mL/min. A
1 lL sample was injected in splitless mode with a solvent delay
of 3 min. The characteristic ions and retention times of the target
compounds were obtained and identified with full scan mass spectra from m/z 50 to 500, the compounds of interest being identified
in SIM mode.
2.4. Yeast estrogen screen (YES) assay
YES bioassay was conducted according to Ma et al. (2005). The
recombinant yeast (Saccharomyces cerevisiae) cells were constructed by Brunel University, U.K. (Routledge and Sumpter,
1996) and kindly provided by Dr. Mei Ma from the State Key Laboratory of Environmental Aquatic Chemistry, Chinese Academy of
Science, Beijing, China. The yeast strain was grown at 30 °C,
130 rpm overnight, in Erlenmeyer flasks using CuSO
4 supplemented (108 M) SC medium. In performing the assay, exponentially growing overnight cultures were diluted with the SC
medium to an OD at 600 nm of 0.25; the absorbance was measured
using a spectrophotometer (SpectroMax M2e, Molecular Devices,
USA). All environmental samples were assayed with a minimum
of three replicates. Each sample assayed included positive
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E.G.B. Xu et al. / Marine Pollution Bulletin 91 (2015) 128–138
Fig. 1. Locations of the sampling sites in Hong Kong’s coastal waters and in the Cape D’Aguilar Marine Reserve (inserted figure).
(estradiol) and negative controls (DMSO). The effects of estrogenic
compounds were standardized against estradiol. Five micro-litres
of DMSO dissolved samples were combined with 995 lL of medium containing 5 103 yeast cells/ml resulting in a test culture.
The test cultures (100 lL) were transferred into each well of a
96-well plate and incubated at 30 °C with vigorous orbital shaking
(130 rpm) on a titer plate shaker for 2 h. The cell density of the
culture was then measured at 600 nm wavelength. Fifty microlitres of test culture were transferred to a new 96-well plate and,
after the addition of 120 lL of Z-buffer and 20 lL chloroform,
the assays were mixed carefully (vortex 25 s) and pre-incubated
for 5 min at 30 °C. The enzyme reaction was started by adding
40 lL o-NPG (13.3 mM, dissolved in Z-buffer). The assays were
incubated at 30 °C on a titer plate shaker. The reactions were
E.G.B. Xu et al. / Marine Pollution Bulletin 91 (2015) 128–138
terminated by the addition of 100 lL Na2CO3 (1 M). After centrifugation at 12,000g for 15 min, 200 ml of the supernatant was transferred into a new 96-well plate and the OD at 420 nm determined
using the spectrophotometer (SpectroMax M2e, Molecular Devices,
USA). The total estrogenic activity in the environmental samples
was measured by comparing the activity of the natural estrogen
of estradiol (E2) and expressed as estradiol equivalent (EEQ)
(Wagner and Oehlmann, 2009).
2.5. Oryzias melastigma embryo bioassay and gene expression
Fertilized eggs of the marine medaka fish O. melastigma
(McClelland, 1839) (Pisces, Adrianichthyidae) were cultured and
acclimated in artificial seawater at a salinity of 30‰ and a temperature of 28 ± 1 °C with a 14 h-light/10 h-dark photoperiod for two
days. The embryos, 2 day post fertilization (dpf), were exposed to
three environmental concentrations of NPs (0, 500, 1000 ng/L),
two mixtures (medium conc. 500 ng/L NPs + 100 ng/L BPA; high
conc. 1000 ng/L NPs + 500 ng/L BPA), positive control E2 (1 ng/L),
surface natural seawater from the CAMR, and to artificial seawater
as a control. Each experimental group contained 50 embryos,
which were randomly distributed into petri dishes containing
30 mL of exposure solution. The media were renewed daily. Three
replicates were conducted for each experimental group. Mortality
rates of the embryos and egg hatchability from 2 to 10 dpf were
recorded. For each replicate, 10 embryos at 4 dpf, 10 embryos at
10 dpf, and two juveniles at the first fry stage were collected for
quantitative real time polymerase chain reaction (qRT-PCR)
analysis.
The primers of 14 endocrine disruption related genes, including
ERa, ERb, ERc, ChgH, ChgL, VTG1, VTG2, ARNT, cyp1a, cyp19a,
cyp19b, PPARa, and PPARband PPARc are presented in Table S2
in Appendix A. The procedures for qRT-PCR followed the methods
described in Fang et al. (2012). The embryos were collected
randomly and homogenized in 1 mL RNA-Solv reagent (Omega)
using a glass homogenizer. The total RNA was extracted from the
homogenates using Omega kits according to the manufacturer’s
instructions. Equal amounts of RNA were applied to qRT-PCR using
SYBR Premix Ex TaqTM kit (TaKaRa) on a Bio Red CFX 96 Real-Time
System. The PCR thermal profile was as follows: an initial denaturation step at 95 °C for 30 s, followed by 40 cycles at 95 °C for 5 s,
and 60 °C for 34 s, and ending with a dissociation curve analysis.
Gene expression levels were normalized to the 18 s rRNA expression levels. The fold change of the tested genes was analyzed
following the 2 44Ct method (Livak and Schmittgen, 2001).
2.6. Probabilistic ecological risk assessments
In order to assess the ecological risk of NPs and BPA on the marine organisms in the CAMR, a probabilistic risk assessment
approach based on tissue concentrations of the two compounds
was applied for this purpose. The risk can be described using the
following equation: Risk Quotient (RQ) = Measured tissue concentration (MTC)/Predicted no effect tissue concentration (PNETC)
(Leung et al., 2006). If RQ < 1, there is low risk to the marine organisms. If RQ P 1, the marine organisms are at risk. The MTC was the
tissue concentration of the two compounds obtained from 11
resident marine organisms in the CAMR (Sections 2.1 and 2.3).
The PNETC of a species was converted from its chronic toxicity
endpoint [i.e., no observed effect concentration (NOEC) or lowest
observed effect concentration (LOEC) in seawater by applying the
bio-concentration factor (BCF) (Leung et al., 2006)]; PNETC =
(NOEC or LOEC)/BCF, where the NOEC and LOEC were obtained
from the USEPA’s ECOTOX database (http://cfpub.epa.gov/ecotox/)
and peer-reviewed literature, and the corresponding BCF values
were also extracted from relevant peer-reviewed literature
131
(Tables S3 and S4 in Appendix A). If there were multiple toxicity
data for the same species, a geometric mean of the chronic endpoints would be taken to calculate the PNETC for the species. With
the distribution of MTC obtained from this study, Monte Carlo simulation with 10,000 iterations was conducted using @Risk software
(version 6.2 Industrial, Palizade Corporation) to compute the RQ
values by random sampling of MTC and PNETC values from the distributions and finally construct a distribution of the RQ values
(Leung et al., 2006). An average probability of the marine organisms in the CAMR, being at risk (i.e., RQ P 1) or at high risk (i.e.,
RQ P 10) was obtained by taking the mean value of five times of
the simulation.
2.7. Data analysis
The sigmoidal YES dose-response curves for E2, NPs, BPA and
the two mixtures of NPs and BPA were fitted to a logistic function
(based on the Hill’s equation) using Prism (GraphPad, San Diego,
USA). Two-way analysis of variance (ANOVA) was employed to
compare the temporal (sampling times: August vs. October in
2012) and spatial (sampling regions: eastern, southern and
western waters) differences in the waterborne concentration of
NPs or BPA across Hong Kong waters. For the expression results
of each target gene, one-way ANOVA followed by a post hoc Tukey
test was used to examine statistical differences amongst different
treatments and control groups. The O. melastigma stage-specific
transcriptional model was constructed and visualized using
GenMAPP 2.1 hhttp://www.genmapp.orgi. A Student’s t test was
used to examine for seasonal differences in the concentrations of
NPs or BPA in the seawater and sediment samples collected from
the CAMR. For all statistical tests, p < 0.05 was considered
significant for rejecting the null hypothesis.
3. Results
3.1. Occurrence of NPs and BPA in coastal waters of Hong Kong
NPs and BPA were found to be ubiquitous in seawater samples
collected around Hong Kong (Fig. 2). The concentrations of NPs
ranged from 110.3 to 721.4 ng/L in August 2012 and from 99.4 to
847.0 ng/L in October 2012. The concentrations of BPA were consistently lower than that of NPs, ranging from 34.8 to 608.8 ng/L
in August 2012 and from 41.7 to 250.2 ng/L in October 2012. Spatially, concentrations of NPs in eastern waters were significantly
higher than those measured in southern and western waters
(2-way ANOVA: F2, 42 = 7.860, p = 0.0013), while such spatial difference for BPA was not statistically significant (F2, 42 = 2.525,
p = 0.092) (Fig. 2). Across all sites, BPA showed significantly higher
mean concentrations in August 2012 than in October 2012
(F2,42 = 6.727, p = 0.013), while such a temporal difference for NPs
was not significant (F2,42 = 2.451, p = 0.125).
3.2. NPs and BPA in water, sediment and biota from the CAMR
The concentrations of NPs and BPA in seawater, sediment and
biota samples collected from the CAMR are presented in Table 1.
NPs and BPA in the CAMR water samples ranged from 91.7 to
154.9 ng/L and from 14.1 to 34.5 ng/L in January 2013, respectively
and varied from 285.1 to 473.9 ng/L and from 72.1 to 206.5 ng/L in
August 2013, respectively. Both NPs and BPA showed significantly
higher mean concentrations in August 2013 (summer) than in January 2013 (winter) in seawater (t test, NPs: t = 22.4, df = 46,
p < 0.01; BPA: t = 16.4, df = 46, p < 0.01) as well as in sediment samples (t test, NPs: t = 7.8, df = 10, p < 0.01; BPA: t = 20.2, df = 10,
p < 0.01). Highest concentrations of both NPs and BPA were
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Fig. 2. Mean concentrations of nonylphenols (NPs) and bisphenol A (BPA) in water samples collected from the southern, western and eastern zones (n = 4 for each zone)
during August 2012 (Left) and October 2012 (Right), respectively. The error bars represent standard deviations.
Table 1
Mean concentrations of nonylphenols (NPs) and bisphenol A (BPA) in surface seawater (sw, ng/L), sediment samples (sd, ng/g d.w.) and marine organisms (ng/g d.w.) collected
from the Cape D’Aguilar Marine Reserve. The standard deviation values of the means are given in parentheses.
Nonylphenols
sw1
sw2
sw3
sw4
sw5
sw6
sw7
sw8
sd1
sd2
Alga Ulva spp.
Bivalve Barbatia virescens
Shrimp Palaemon pacificus
Fish Bathygobius fuscus
Gastropod Nerita costata
Gastropod Thais clavigera
Crab Grapsus albolineatus
Sea cucumber Holothuria leucospilota
Gastropod Monodonta labio
Bivalve Septifer virgatus
Sea urchin Anthocidaris crassispina
Bisphenol A
n
January 2013
August 2013
January 2013
August 2013
91.7 (3.0)
154.9 (7.2)
144.2 (8.3)
137.1 (15.1)
112.0 (9.0)
107.3 (11.0)
134.4 (6.9)
124.6 (15.1)
527.3 (20.3)
602.1 (19.0)
7.5 (2.8)
35.5 (5.1)
36.7 (4.8)
45.1 (12.5)
58.5 (4.0)
73.2 (17.3)
93.7 (19.2)
358.1 (29.7)
388.2 (25.1)
538.6 (20.0)
739.4 (14.0)
285.1 (25.6)
422.8 (27.6)
398.2 (20.5)
405.0 (10.8)
473.6 (18.9)
473.9 (6.7)
402.1 (6.6)
405.8 (8.3)
724.0 (12.0)
800.0 (16.0)
50.4 (5.6)
231.2 (16.3)
63.9 (6.6)
333.2 (9.9)
301.7 (14.9)
369.9 (18.1)
13.7 (9.0)
194.7 (21.6)
318.2 (16.9)
364.7 (20.2)
788.0 (20.8)
14.1 (4.0)
34.5 (2.1)
31.3 (5.9)
21.6 (7.4)
17.1 (6.0)
20.0 (9.0)
31.6 (11.3)
24.1 (6.2)
60.8 (10.6)
70.1 (11.0)
12.8 (3.0)
25.9 (4.5)
32.6 (5.0)
75.2 (7.3)
158.1 (9.1)
40.1 (11.0)
159.2 (7.9)
241.9 (31.0)
79.5 (13.1)
207.1 (19.0)
198.6 (20.1)
72.1 (9.8)
206.5 (13.2)
186.8 (9.0)
159.3 (3.7)
171.0 (8.9)
166.8 (6.7)
153.4 (2.8)
154.5 (5.5)
238.0 (13.1)
265.9 (15.1)
5.0 (1.4)
27.2 (5.7)
4.8 (2.1)
127.2 (4.6)
231.3 (8.7)
90.3 (7.0)
168.1 (23.0)
51.2 (18.3)
111.5 (23.0)
78.8 (18.1)
306.1 (17.6)
obtained from Lobster Bay (Site sw2; Fig. 1, Table 1). NPs and BPA
were detected at higher concentrations in sediment samples than
in water samples, at average concentrations of 564.7 and 65.5 ng/g
d.w. in January 2013, respectively and 762.0 and 252.0 ng/g d.w.
in August 2013, respectively.
NPs and BPA detected in samples of biota were in the range of
7.5–739.4 and 12.8–241.9 ng/g d.w. in January 2013, respectively
and 50.4–788.0 and 5.0–231.3 ng/g d.w. in August 2013, respectively. Among the 11 marine species examined, the highest mean
concentrations of BPA were obtained from the black sea cucumber
Holothuria leucospilota (Brandt, 1835) (241.9 ng/g, d.w.), followed
by the black mussel Septifer virgatus (Wiegmann, 1837)
(207.1 ng/g, d.w.), while the highest mean concentration of NPs
was identified in the sea urchin Anthocidaris crassispina (Agassiz,
1863) (739.4 ng/g, d.w.), followed by S. virgatus (538.6 ng/g, d.w.)
3
3
3
3
3
3
3
3
3
3
5
5
9
3
5
10
3
3
10
5
5
(Table 1). The lowest mean concentrations of both NPs and BPA
were found in the alga Ulva spp. (7.5 and 5.0 ng/g, d.w., respectively) (Table 1).
3.3. Estrogenic activities assessed by YES
The concentration-response relationship obtained with each
EDC solution (E2 alone, NPs alone, BPA alone, and a mixture of
NPs and BPA) in DMSO is shown in Fig. S1 (Appendix A). The
calculated EC50 values of E2, NPs and BPA were 11.3 ng/L,
16.3 lg/L, and 165 lg/L, respectively. Additive effects of the
in vitro estrogenic activity were identified for the mixture. Fig. 3
illustrates the in vitro estrogenic activity (measured by YES assay)
in different environmental samples. The average EEQ values in
surface water and sediment samples from the CAMR were
E.G.B. Xu et al. / Marine Pollution Bulletin 91 (2015) 128–138
133
Fig. 3. YES-estrogenic activities of wastewater and sludge samples collected from SWIMS sewage treatment plant (STP), and seawater (sw) and sediment (sd) samples
collected from the Cape D’Aguilar Marine Reserve. The triplicate data are presented as the mean and SD; bars with different letters indicate significantly different means
(ANOVA followed by post hoc Tukey test: F12, 26 = 107.7, p < 0.01).
2.1 ng/L and 4.2 ng/g, respectively. EEQs in sludge, influent and
effluent samples from the sewage treatment plant at the Marine
Reserve were significantly higher than those found in the natural
seawater samples (one-way ANOVA, F12, 26 = 106.7, p < 0.01, Fig. 3).
3.4. Implications from the medaka bioassay
The results of mortality and hatchability of the O. melastigma
embryos after 21 days of exposure to E2, NPs, and mixtures of
NPs and BPA are summarized in Fig. S2 (Appendix A). In mixed
solutions, mortality and hatchability of the embryos ranged from
1.7% to 10.0%, and from 60% to 77%, respectively; in NPs solutions
mortality and hatchability was between 6.0% and 11.0%, and
between 31% and 82%, respectively. Nevertheless, no significant
difference in either mortality or hatchability was observed among
all treatment groups (one-way ANOVA, p > 0.05).
A total of 14 genes representing three key endocrine signaling
pathways in O. melastigma were examined in this study (Fig. 4).
The results showed that not all endocrine disruption related genes
were affected by the seawater collected from the marine reserve.
Of a total of 14 genes, 5, 4, and 8 genes were significantly altered
in O. melastigma at 4 dpf, 10 dpf and the 1st fry stage, respectively
(Fig. 4; Table S5 in Appendix A). Cyp19a (encoding aromatase
catalyzing the conversion of androgens to estrogens) and ER
genes (nuclear receptor activated by estrogen) were significantly
up-regulated at all three developmental stages (Fig. 4; Table S5).
The mRNA levels of vitellogenin genes (VTG1 and VTG2) and
CYP1A gene were also significantly higher than those in control
Fig. 4. The gene expression profile of Oryzias melastigma at 4 dpf, 10 dpf and 1st fry stage, respectively, after exposure to natural seawater collected from the Cape D’Aguilar
Marine Reserve. Each ‘‘striped gene box’’ contains two parts. The right part represents the fold change of mRNA expression, and the left part illustrates the statistically
significant fold change of each gene, based on the results from one-way ANOVAs at p < 0.05.
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E.G.B. Xu et al. / Marine Pollution Bulletin 91 (2015) 128–138
groups at the 1st fry and 4 dpf stage, respectively (Fig. 4). The largest fold change of all gene expressions was observed in ERa (7-fold
of up-regulation), cyp19a (8-fold of up-regulation) and PPARa (0.1fold of down-regulation) (Table S5). In addition, E2, NPs, and mixtures of NPs and BPA at environmentally realistic concentrations
resulted in similar responses in these three genes in O. melastigma
embryos. Fig. 5 shows the transcriptional responses of these three
genes in O. melastigma embryos after 2 weeks of exposure to NPs
and the mixtures of NPs and BPA as compared to the artificial seawater control. Both high concentrations of NPs (1000 ng/L) and the
‘high mixture’ (1000 ng/L NPs + 500 ng/L BPA) elicited significant
up-regulations of ERa and cyp19a, and down-regulation of PPARa
(Fig. 5). Cyp19a mRNA was significantly up-regulated at lower concentrations of NPs (500 ng/L) and the ‘medium mixture’ (500 ng/L
NPs + 100 ng/L BPA) (Fig. 5).
3.5. Probabilistic ecological risk assessment
The PNETCs of NPs contained seven marine species categorized
into different trophic levels (e.g., crustaceans, fish and molluscs),
covering endpoints such as mortality, and inhibition of development and reproduction (Table S3 in Appendix A). The most sensitive species were crustaceans including Americamysis bahia and
Tisbe battagliai. The PNETCs of BPA contained eight marine species,
with the most sensitive species being fishes, such as Psetta maxima
and Gadus morhua (Table S4 in Appendix A). From the computed
probability density function, the average probability of having RQ
for NPs equal to and greater than 1 and 10 was 35% and 8%,
respectively (Fig. 6). The average probability of RQ for BPA equal
to and greater than 1 and 10 was 21% and 5%, respectively (Fig. 6).
4. Discussion
In general, both NPs and BPA in seawater samples collected
from Hong Kong coastal waters showed higher mean concentrations at all sites in August 2012 than in October 2012. The concentrations of the two compounds in both water and sediment
samples collected from the marine reserve were also higher in
the wet season (August) than in the dry season (January) (Table 1).
This seasonal variation was in accordance with our previous findings (Xu et al., 2014). The higher concentrations of NPs and BPA
in the wet season (summer) in the CAMR could be caused by several factors. Firstly, the average seawater temperature in the CAMR
was 28.32 °C in August, but decreased by over 10 °C in January
(Table S1 in Appendix A). High temperature and associated microbial activity lead to enhanced degradation of nonylphenol ethoxylates (NPEO) in marine sediments and, hence, increased NPs
concentrations in the water column during summer (Li et al.,
2004). Levels of NPs in both natural surface seawater and in suspended particles were shown to decrease with decreasing water
temperature (Xu et al., 2006). Secondly, occasional tropical
cyclones in the wet season increased seawater turbidity in the
CAMR by almost 20 times in August over that in January
(Table S1 in Appendix A), and hence promoted re-suspension of
pollutants from sediments. Fu et al. (2007) also reported that
Fig. 5. Mean expression levels (i.e., relative fold changes) of ERa, cyp19a and PPARagenes at 1st fry stage of Oryzias melastigma. Right column: different concentrations of NPs;
left column: different concentrations of the mixture of NPs and BPA. E2 treatment was applied as a positive control. The data in triplicate are presented as the mean and SD,
relative to the control; ⁄p < 0.05.
E.G.B. Xu et al. / Marine Pollution Bulletin 91 (2015) 128–138
Fig. 6. The distribution of the computed risk quotients of NPs (Upper) and BPA
(Lower) generated from the Monte Carlo simulation.
higher concentrations of NPs in summer in the coastal waters of
Qingdao, China, were mainly due to the higher degradation rate
of NPEO and re-suspension of sediments under strong winds. The
third major reason for the elevated EDCs concentrations in natural
seawater during the wet season may be attributable to an
increased input of EDCs from non-point pollution sources, such
as increased surface runoff (Zhao et al., 2009; Ying et al., 2011).
Possibly, surface runoff and storm water discharges may have also
increased the input of NPs into CAMR during the wet season. Significant amounts of rainfall recorded in August 2013 led to large
amounts of surface runoff and storm water discharges (such heavy
rainfalls also reduced the salinity of the seawater in the CAMR from
>30‰ to <23‰; Table S1 in Appendix A). A previous study showed
that the concentrations of NPs in Hong Kong’s storm waters ranged
from 80 to 12,000 ng/L (Kueh and Lam, 2008), affirming storm
waters as a potential source of NPs.
The estrogenic activities in the CAMR (i.e., 2.1 ng/L and 4.2 ng/g
EEQ for seawater and sediment, respectively) fell into similar
ranges as those reported from the U.S.A. (1–90 ng/L EEQ, Schlenk
et al., 2005), Europe (Switzerland 0.3–7.0 ng/L EEQ, Vermeirssen
et al., 2005; France 0.3–4.52 ng/L EEQ, Cargouet et al., 2004; Portugal 0.1–1.7 ng/L EEQ, Quirós et al., 2005), and Asia (South Korea
0.38–6.27 ng/L EEQ, Ra et al., 2011; Japan 0.7–4.01 ng/L EEQ,
Hashimoto et al., 2005; China 0.08–2.4 ng/L EEQ, Jiang et al.,
2012). EEQs in the Rhine River in Germany and in the Pearl River
in southern China were higher than those in the above regions,
with a mean of 19.42 and 6.81 ng/L, respectively (Pawlowski
et al., 2004; Zhao et al., 2011). Our data showed that the sigmoidal
concentration-response curve of the mixture of NPs and BPA lies
between those of single NPs and single BPA, acting additively in
stimulating the estrogen receptor in the YES assay. From a pharmacological point of view, BPA can be thought of as behaving as a
dilutor of NPs. This observation is comparable with the results
135
reposted by Payne et al. (2000) and Kang et al. (2002) on four different estrogenic chemicals including NPs or/and BPA in the YES
assay, supporting that the concentration addition concept (Loewe
and Muischnek, 1926) is suitable for predicting and assessing mixture effects of xenoestrogens. According to our calculations, the
EC50 values of NPs and BPA in our YES assay were 16.3 lg/L and
165 lg/L, respectively, which are 3–4 orders higher than those
detected in the environmental samples collected from the CAMR.
These weak xenoestrogens alone, therefore, may compose relatively low risk to the CAMR. Considering additive combination
effects and the amount of unknown/undetectable EDCs, however,
the outcomes of single-chemical based risk assessment is likely
to be under-estimated (Rajapakse et al., 2002; Kortenkamp,
2008). In this paper, the integral environmental risks were initially
assessed according to the method of Zhao et al. (2011), by dividing
the predicated no-effect concentration (PNEC) of estradiol (E2) by
the measured E2 EEQ in seawater and sediment samples. The mean
risk quotients (RQ = measured EEQ / PNEC) for seawater and sediment in the CAMR were 1.3 and 2.6, respectively, indicating high
risks to marine organisms, especially the benthos. The results of
the tissue-residue-based probabilistic risk assessment further suggested that the marine species living in the CAMR were having 35%
and 21% of chance to be at risk due to exposure to NPs and BPA,
respectively. There were, however, some uncertainties existing in
this kind of probabilistic risk assessment that have been lengthily
discussed in previous studies (Suter, 1990; Skinner et al., 2014).
For instance, the major source of uncertainty is probably associated
with the potential error in the derivation of predicted no effect tissue concentrations (PNETCs). It is because the PNETCs are only
estimated from a limited amount of toxicity data and bioconcentration factors (BCFs) generated from few non-native marine species. To increase the accuracy and appropriateness of the risk
assessment, we need to generate more relevant information on
the toxicities and BCFs of NPs and BPA on native marine species
of different trophic levels.
In this study, the CYP19 genes, encoding aromatase for vertebrate reproduction and development, which catalyze the conversion of androgens to estrogens, were up-regulated in O.
melastigma regardless of their developmental stage. Some previous
studies investigating effects of environmental waters, which were
contaminated with NP, also identified such an elevated aromatase
mRNA expression in different fish species. For example, NP was
found to induce up-regulation of CYP19A genes in a dose-dependent manner in juveniles of zebra fish (Kazeto et al., 2004), and
cause a significant induction of CYP19 isomers in immature Atlantic salmon (Meucci and Arukwe, 2006). Likewise, Massari et al.
(2010) reported that river waters contaminated with environmental estrogens significantly increased the level of aromatase CYP19
mRNAs in frog and carp.
As shown in the present study, the CYP1A gene, which is
involved in phase I xenobiotic metabolism, was also up-regulated
in early development stages of O. melastigma after exposure to
the seawater collected from the CAMR, suggesting the presence
of xenobiotics in this marine reserve. Similarly, An et al. (2011)
identified higher levels of CYP1A mRNA expression in mullet fish
collected from a polluted coastal area compared to a reference site
in Bohai Bay, which were positively correlated with the concentrations of different xenobiotics in their tissues. Such induction of the
CYP1A gene may be a consequence of exposure to NPs and BPA or
other ligands such as polycyclic aromatic hydrocarbons (PAHs)
binding to the aryl hydrocarbon receptor (AhR) (Cousinou et al.,
2000; Rees and Li, 2004), resulting in an increase of estradiol
metabolism (Arcaro et al., 1999).
Estrogen receptor (ER) genes are the other key players in sexual
hormone synthesis and action in teleosts. Jin et al. (2009) reported
that NPs induced ER mRNA levels in both larval and adult zebra
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E.G.B. Xu et al. / Marine Pollution Bulletin 91 (2015) 128–138
fish. Recently, Palermo et al. (2012) demonstrated that ER mRNA
expression was increased in juvenile sole following a short exposure to environmental concentrations of NPs (106 M). Our data
are in accordance with several studies which show that NPs and
BPA are able to induce fish ER expression in vivo (Yadetie et al.,
1999; Yamaguchi et al., 2005) and in vitro (Flouriot et al., 1995;
Bonefeld-Jørgensen et al., 2007). ER genes are involved in the transcriptional regulation of VTG genes, but different subtypes of ER
show different binding affinities of estrogen in goldfish (Nelson
and Habibi, 2010). We also discovered that different subtypes of
ER gene resulted in different sensitivities to xenoestrogen and natural seawater exposures. Greytak et al. (2010) showed that levels
of VTG mRNA in killifish embryos did not differ between a highly
polluted site and an unpolluted site, and that, similarly, induction
of ER alpha by estradiol exposure in reproductively inactive male
killifish did not differ between the two sites. In the present study,
significant up-regulation of VTG mRNA was observed only at the
1st fry stage of O. melastigma through activation of ER beta
(Fig. 4). These findings suggest that fish development stage be considered when VTG is used as a biomarker of estrogenic effects since
responses to estrogens are mediated differently by ER subtypes.
Most recently, similar researches regarding EDC pollution in a marine reserve documented high levels of NPs and low levels of BPA in
sediment samples, and extremely high concentrations of NPs in the
bile of the mullet fish (up to 44,896 ng/g), and demonstrated that
the mullet fish from this marine reserve showed clear signs of
endocrine disruption indicated by their expression of VTG, ER,
CYP and retinoid X receptor (RXR) mRNAs (Puy-Azurmendi et al.,
2013).
This is the first study to report that NPs and BPA are present in
marine organisms resident in a marine reserve in Hong Kong,
South China. A large variation in the concentrations of NPs and
BPA in different marine species was recorded, which might be correlated with the feeding modes and EDC removal abilities of the
species. It is interesting to note that both the highest concentrations of NPs and BPA were obtained from echinoderms. This may
be explained by their unusually expanded external epithelia,
through which the uptake of dissolved substances occurs continuously, and their benthic habits which make them easy targets for
environmental contamination, particularly by micro-pollutants
stored in seabed sediments (Candia Carnevali, 2005). Sea cucumbers, for example, scavenge organic debris in the sediment. Hong
et al. (2011) also identified a higher bioaccumulation factor (BAF;
revealing integral uptake through partitioning and diet) and
biota-sediment accumulation factor (BSAF) of other persistent
organic pollutants (POPs) in sea cucumbers than those in mussels,
prawns, crabs, clams, polychaetes and fish.
Aquatic organisms acquire synthetic organic pollutants principally from the surrounding water through partitioning and from
dietary uptake. In benthic ecosystems, uptake of these pollutants
through the sediment is an additional route (Park and Lee, 1993).
Higher concentrations of both NPs and BPA were detected in sediment samples than in seawater samples both obtained from the
CAMR, suggesting that the former act as a sink for these hydrophobic EDCs (Ying et al., 2002). In this situation, the benthos may act as
a common entry route for EDCs into the trophic chain (David et al.,
2009). On the other hand, high accumulation of both NPs and BPA
in the mussel S. virgatus, occurring in dense intertidal clusters, and
with a wide distribution and ease of handling (Morton, 1995),
make them a good candidate for biomonitoring of EDCs pollution
in Hong Kong.
The tissue concentrations of EDCs in the biotic samples
obtained for this investigation were compared with those of other
domestic and overseas investigations (Table S6 in Appendix A). NPs
tissue concentrations noted in the present study are similar to
those in Bohai Bay (Hu et al., 2005) and Shenzhen (Liu et al.,
2009), China, and are significantly lower than those in Xiamen
(Zhang et al., 2011) and the northeastern coast of China (Wang
et al., 2010). The BPA tissue concentrations in this study are comparable to those in Xiamen (Zhang et al., 2011) and Shenzhen,
China (Liu et al., 2009). In comparison with other overseas investigations, the tissue concentrations of NPs in this study fell in the
middle of the world-wide range (Table S6 in Appendix A). The pollution levels of NPs in the CAMR are lower than those in Taiwan
(Cheng et al., 2006), Canada (Diehl et al., 2012), U.S.A. (Diehl
et al., 2012) and Japan (Isobe et al., 2007), but higher than those
of the investigated freshwater species in the U.S.A. (Keith et al.,
2001) and U.K. (Lye et al., 1999). At the current tissue concentrations, NPs and BPA may have affected the reproductive and developmental biology of the marine species inhabiting in the CAMR.
For instance, Roepke et al. (2005) reported unsuccessful development of sea urchin embryos after exposure to EDCs, such as TBT,
o, p-DDD, 4-octylphenol and BPA. NPs can adversely affect the
regenerative capabilities of echinoderms by inducing marked
anomalies in growth rate, cell proliferation and migration, cell/tissue rearrangement and turnover, and general cytological disorders
(Candia Carnevali, 2005). Overall, given the relatively high levels of
NPs and BPA accumulated in the marine organisms resident in the
CAMR and their high ecological risks, there is concern over their fitness and well being.
5. Conclusions
Our study investigated the distribution of NPs and BPA in the
coastal waters of Hong Kong and indicated their contaminations
even in relatively remote areas, such as marine protected areas.
The two chemicals were present in all seawater, sediment and
biota samples collected from the Cape D’Aguilar Marine Reserve,
the only marine reserve in Hong Kong, indicating that it has been
polluted by these two common endocrine disruptors. We also present evidence for the ecological risks posed by the two endocrine
disruptors as assessed by in vivo and in vitro bioassays. Monte Carlo
simulations result in the probability functions for the corresponding risk quotient (RQ = MTC/PNETC), which could be used to characterize the likelihood and severity of the adverse effects of the NPs
and BPA. The high RQs to the marine organisms identified in this
study call for more systematic, ecological and field studies on Hong
Kong’s marine life alongside the monitoring of EDC levels in Hong
Kong, especially. The ubiquity of NPs and BPA in Hong Kong’s
coastal environment supports the need for greater awareness and
control of EDC pollution, particularly for safeguarding the MPAs,
the only local sensitive shelters for precious marine creatures.
Acknowledgements
This work is jointly supported by the University Grants Committee of the Hong Kong Special Administration Region (HKSAR),
China via the Area of Excellence Scheme (Project No. AoE/P-04/
2004), and by a research grant from the Swire Educational Trust.
The sampling of surface seawater samples from the eastern, southern and western waters of Hong Kong was funded by the Research
Grants Council of HKSAR via a Collaborative Research Fund (Project
No. HKU5/CRF/12G). The authors thank the Agriculture, Fisheries
and Conservation Department of the HKSAR Government for granting us permission to collect environmental and biota samples at
the marine reserve for this scientific study. Elvis Xu would also like
to thank John Swire & Sons Limited and the Swire Educational Trust
for providing a James Henry Scott Hong Kong PhD Scholarship. The
authors thank Dr. Mei Ma of the State Key Laboratory of Environmental Aquatic Chemistry, Chinese Academy of Sciences for kindly
providing the recombinant yeast (Saccharomyces cerevisiae) cells
E.G.B. Xu et al. / Marine Pollution Bulletin 91 (2015) 128–138
which were prerequisite for the YES assay adopted in this study.
The authors are grateful to the staff and postgraduate students at
the Swire Institute of Marine Science for assisting in this project,
and thankful to Dr. June Leung for her helpful comments on an
earlier draft of this manuscript.
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.marpolbul.2014.
12.017.
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Marine Pollution Bulletin 85 (2014) 254–260
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papers which clearly identify the quality of the data will be considered for publication. Contributors to Baseline should refer to
‘Baseline—The New Format and Content’ (Mar. Pollut. Bull. 60, 1–2).
Spatio-temporal comparisons of imposex status and tissue organotin
concentration in the whelk Reishia clavigera collected along the coasts
of Dapeng Bay and Daya Bay, Shenzhen, China
Kevin K.Y. Ho a, Kenneth M.Y. Leung a,b,⇑
a
b
The Swire Institute of Marine Science and School of Biological Sciences, The University of Hong Kong, Pokfulam, Hong Kong, China
The State Key Laboratory in Marine Pollution, City University of Hong Kong, Tat Chee Avenue, Kowloon, Hong Kong, China
a r t i c l e
i n f o
Article history:
Available online 3 July 2014
Keywords:
Triphenyltin
Thais
Tissue concentration
Gastropod
Container terminal
Antifouling
a b s t r a c t
Organotin compounds (OTs) have caused widespread adverse effects on marine organisms. As no local
restrictions on OT-based antifouling paints have been implemented in China, high concentrations of
OTs still occur in coastal environments. In this study, we measured the imposex status and tissue
concentrations of OTs in the whelk Reishia clavigera collected along the coast of Dapeng Bay and Daya
Bay of Shenzhen, China in 2013. Our results showed that all female whelks suffered from the onset of
imposex. The highest concentration of total OTs was 27,756 lg kg 1 dry weight in the samples collected
from Shuitousha. Triphenyltin was the most predominant residue, accounting for more than 97.8% of
total OTs across all sites. Compared with the results from previous studies, the marine environment of
this region is still heavily contaminated with OTs. Therefore, a tightened control is necessary to regulate
the use and release of OTs in China.
Ó 2014 Elsevier Ltd. All rights reserved.
Organotin compounds (OTs), in particular tributyltin (TBT) and
triphenyltin (TPT), have been widely used as effective biocides on
ship hulls, fish cages and agriculture since 1960s (Hoch, 2001).
However, their toxic effects were not limited to target fouling
organisms. The two most well documented examples of OTs toxicity were shell thickening of oysters and imposex development on
more than 200 species of gastropods worldwide (Shi et al., 2005).
Due to their widespread toxic effects on marine ecosystems, the
International Maritime Organization (IMO) called for a complete
⇑ Corresponding author at: The Swire Institute of Marine Science, School of
Biological Sciences, The University of Hong Kong, Pokfulam, Hong Kong, China. Tel.:
+852 22990607; fax: +852 25176082.
E-mail
addresses:
kevinho2@hku.hk
(K.K.Y.
Ho),
kmyleung@hku.hk
(K.M.Y. Leung).
http://dx.doi.org/10.1016/j.marpolbul.2014.06.017
0025-326X/Ó 2014 Elsevier Ltd. All rights reserved.
ban of organotin-based antifouling paints through an enactment
of the International Convention on the Control of Harmful Antifouling Systems on Ships (i.e., AFS Convention) in September
2008 (IMO, 2008; Sonak et al., 2009). Hence, it is expected to see
a reduction in OT contamination and associated ecosystem recovery in the marine environment worldwide.
In China, however, no local regulations on OT-based antifouling paints have ever been made (Cao et al., 2009; Gao et al.,
2013). Therefore, high concentrations of OTs especially in coastal
areas were still observed (Deng et al., 2010). In the Pearl River
Delta, rapid industrialisation and economic growth in recent
years have made Yantian Port in Shenzhen (see Fig. 1) to become
an important hub in transferring cargos to other parts of China.
This is reflected by an ever-increasing container throughput since
its operation in 1994, reaching 10 million TEU in 2010 (TEU:
K.K.Y. Ho, K.M.Y. Leung / Marine Pollution Bulletin 85 (2014) 254–260
Fig. 1. Map showing the sampling locations in Dapeng Bay and Daya Bay, Shenzhen,
China. Sites C4–8 and C10 were previously visited by Chan et al. (2008) while D2–5
were previously visited by Deng et al. (2010). Names and coordinates of the
sampling sites, as well as the sampling dates are listed in Table 1.
twenty-foot equivalent unit) and serving about 40 shipping lines
from major ports around the world (YICT, 2012). Nationally,
China produces 200 tonnes of TPT every year, which is used as
antifoulants and fungicides. Given shipping facilities being a
major source of OTs to the marine environment together with
the terrestrial runoff of TPT from agricultural practices (Yi et al.,
2012), the level of OT contamination in coastal areas of China
could be inevitably high.
The intertidal whelk Reishia clavigera (= Thais clavigera, see
Claremont et al., 2013) was extensively used as a biomonitor for
OT contamination in the Asia–Pacific region (e.g., Horiguchi et al.,
1994; Shim et al., 2000; Leung et al., 2006; Qiu et al., 2011). The
species is very sensitive to OTs, which can develop imposex at concentrations as low as 1 ng L 1 (Horiguchi et al., 1995). This species
(Chan et al., 2008) and other gastropods (Deng et al., 2010) were
successfully used in previous studies to quantify the degrees of
OT contamination in this region (Fig. 1). These studies found a negative relationship between the distance to the container port in
Yantian and the OT contamination in the various molluscs including gastropods and bivalves.
Previous studies in this region only measured butyltin compounds (BTs) in tissues of R. clavigera (Chan et al., 2008; Deng
et al., 2010) and ignored phenyltin compounds (PTs). As demonstrated by Horiguchi et al. (1997), TPT can trigger imposex in R.
clavigera as TBT does. Therefore, this study aimed to determine
the contemporary imposex status and concentrations of both BTs
and PTs in tissues of R. clavigera collected along the coasts of Dapeng Bay and Daya Bay, Shenzhen, China. We examined the spatial
variation among the studied sites, and temporal variation for BT
compounds between 2006 (Chan et al., 2008) and Chan et al.,
2013 (this study).
About 30 individuals of adult R. clavigera (shell
length P 17 mm; see Tong (1986); actual shell length ranged from
20 to 31 mm in this study) were randomly collected from each of
the 10 rocky shores (ca. 0.5–1.5 m above chart datum) along the
coasts of Dapeng Bay and Daya Bay, Shenzhen, China during January 2013 (Fig. 1; Appendix A). The sites were previously visited by
either Chan et al. (2008) or Deng et al. (2010), and thus were correspondingly labelled as group ‘‘C’’ or group ‘‘D’’ on the map
255
(Fig. 1). These sites covered different proximities to the shipping
activities associated with Yantian Port. Samples of the whelks were
covered with ice and placed in a cool box, and transferred to The
University of Hong Kong (HKU) within eight hours of collection.
Samples were stored at 20 °C for further analysis.
Total shell length of R. clavigera was measured to the nearest
0.1 mm using vernier calipers. Fresh weight of the animals was
measured individually using an electronic balance (Libror EB430HU, Shimadzu, Japan) to the nearest 1 mg. The animals were
crushed using a bench vice; the soft tissues were pulled out carefully and weighed while the broken pieces of shell were removed.
Each individual was sexed under a stereomicroscope (Olympus
SZH10): the presence of a prostate gland indicated a male while
a sperm-ingesting gland represented a female (Blackmore, 2000).
The presence of a penis was recorded, and it was straightened
and the length was measured (to the nearest 0.1 mm) under the
stereomicroscope. Vas Deferens Sequence Index (VDSI) and relative Penis Size Index (RPSI) were used to determine the degree of
imposex and indirectly indicate the degree of OT contamination
across the study sites.
The seven-stage VDSI (0–6) indicates the progressive imposex
development, and it has been widely used for assessing the severity of imposex and OT contamination (e.g., Gibbs et al., 1987; Leung
et al., 2006; Abidli et al., 2012). Detailed descriptions of each stage
of VDSI were given by Cheung et al. (2010) and Horiguchi et al.
(2012). RPSI indicates the percentage of the mean bulk of the
female penis to that of the male penis (Gibbs et al., 1987). The larger the value of RPSI is, the more seriously the imposex developed
on the gastropod.
The soft-body tissues were pooled as four replicates; each having 5–8 individuals. They were freeze-dried for at least 48 h (VirTis
#6KBTES-55 freeze dryer, Gardiner, NY, U.S.A.) before homogenization using a blender (Philips HR2860, Holland). Chemical analysis
of OTs followed previously established method (Guðmundsdóttir
et al., 2011). Chemical standards were bought from Sigma–Aldrich
(St. Louis, MO, USA) and Chiron (Trondheim, Norway), while solvents were purchased from Tedia (Fairfield, OH, USA) in HPLC
Grade. Quantification of the six residues of OTs, i.e., monobutyltin
(MBT), dibutyltin (DBT), TBT, monophenyltin (MPT), diphenyltin
(DPT) and TPT, was conducted using a gas chromatograph (GC;
Bruker 450-GC, Bruker Inc., Billerica, MA, USA) equipped with a
mass-selective detector (MS; Bruker 320-MS, Bruker Inc., Billerica,
MA, USA). A VF-5MS fused silica capillary having 0.25 mm
i.d. 30 m 0.25 m film thickness (Bruker Inc., Billerica, MA,
USA) was used as the GC column. The method was previously validated by using the certified reference material ERM-CE477 (see
Guðmundsdóttir et al., 2011). Procedural blank was analysed
simultaneously for every batch of five samples. The detection limits were estimated as 0.2–1.5 lg kg 1 dry weight (dw). Concentrations of all OTs were reported without correction from the recovery
rates.
Spatial differences of tissue OT concentrations (i.e., total OTs,
TBT and TPT), imposex index (VDSI) and condition index (= fresh
tissue weight x 100/[fresh tissue weight + dry shell weight]; Lau
and Leung, 2004) among the 10 study sites were identified using
one-way Analysis of Variance (ANOVA) if data had homogeneous
variances (tested using Levene’s test); otherwise non-parametric
Kruskal–Wallis test was used. Parametric Tukey test or non-parametric Dunn’s test, were used respectively for multiple comparisons. Relationships among tissue OT concentrations, imposex
indices and condition index were tested using non-parametric
Spearman’s rank correlation analysis due to the small sample size.
VDSI, RPSI and total BTs concentrations in R. clavigera measured in
this study were compared with a study conducted in 2006 (Chan
et al., 2008) using Wilcoxon signed ranks test, in which both studies visited the six identical locations in Dapeng Bay (Fig. 1: sites
K.K.Y. Ho, K.M.Y. Leung / Marine Pollution Bulletin 85 (2014) 254–260
21.1
16.0
14.3
12.8
8.8
4.4
6.9
6.7
10.3
13.4
29.6 ± 2.79
26.6 ± 3.22
30.5 ± 3.76
24.6 ± 3.65
26.3 ± 3.58
23.0 ± 3.52
23.9 ± 3.67
24.4 ± 3.47
21.7 ± 3.02
26.4 ± 3.30
0.8
4.5
7.0
9.8
11.7
2.0
4.5
0.4
4.2
7.1
Shipping
activities
(km)
% Sterile F
% Imposex
100
100
100
100
100
100
100
100
100
100
4.00
4.00
3.00
4.00
4.00
5.00
4.00
4.00
3.00
3.00
VDSI
VDSI
4.43
3.76
2.69
4.07
3.87
4.90
4.17
3.82
3.00
3.56
11.89 ± 1.83
8.99 ± 1.54
7.00 ± 0.80
7.19 ± 2.24
6.81 ± 2.00
10.67 ± 1.46
9.23 ± 1.31
9.05 ± 2.10
5.36 ± 2.10
6.28 ± 1.52
Mean ± SD
Mean ± SD
1.31 ± 0.33
0.71 ± 0.29
0.51 ± 0.15
0.48 ± 0.13
0.38 ± 0.15
0.40 ± 0.14
0.56 ± 0.33
0.75 ± 0.35
0.39 ± 0.42
0.59 ± 0.21
30.50 ± 2.28
24.34 ± 3.62
20.05 ± 1.88
21.37 ± 2.46
20.41 ± 2.93
20.88 ± 2.73
23.43 ± 4.22
26.32 ± 4.14
21.01 ± 4.84
24.12 ± 2.29
Mean ± SD
Mean ± SD
13.26 ± 1.95
11.34 ± 1.96
10.00 ± 2.88
10.24 ± 2.24
9.29 ± 1.81
11.79 ± 2.08
10.90 ± 1.64
10.83 ± 2.93
9.51 ± 3.96
11.52 ± 2.31
1.18 ± 0.45
0.64 ± 0.23
0.54 ± 0.20
0.63 ± 0.26
0.52 ± 0.23
0.45 ± 0.14
0.49 ± 0.22
0.62 ± 0.27
0.39 ± 0.28
0.57 ± 0.22
Mean ± SD
Mean ± SD
29.01 ± 3.52
23.48 ± 3.29
20.38 ± 2.73
23.25 ± 3.41
22.92 ± 3.80
21.68 ± 2.61
22.41 ± 3.43
24.86 ± 3.41
20.89 ± 4.18
24.05 ± 3.62
46.7
56.7
53.3
50.0
50.0
33.3
40.0
37.9
40.0
60.0
Seafood Street
Dameisha
Meisha
Xichong
Guanhu
Shuitousha
Egong
Dongshan
Yangmeikeng
Luzui
C4
C5
C6
C7
C8
C10
D2
D3
D4
D5
30
30
30
30
30
30
30
29
25
30
Median
Mean
Penis length
(mm)
Tissue
weight (g)
Shell length
(mm)
Tissue
weight (g)
Shell length
(mm)
Penis length
(mm)
Female
Male
%F
n
Site
No.
Table 1
Morphological measurements and imposex status of the whelk Reishia clavigera collected from various sites in Dapeng Bay and Daya Bay during winter of 2013.
42.9
23.5
0.0
26.7
13.3
60.0
25.0
18.2
0.0
16.7
72.07
49.84
34.30
34.60
39.44
74.23
60.56
58.46
17.90
16.21
Mean ± SD
RPSI
Condition
Index
Distance to
Mariculture
activities
(km)
256
C4–C8 and C10) and used the same analytical method (i.e., GC–
MS). All statistical tests were performed using IBM SPSS Statistics
20.0 (SPSS Inc., Chicago, IL, USA, 2011) and Microsoft EXCEL 2003
(Microsoft Corporation, Redmond, WA, USA, 2003).
Incidence of imposex was 100% at all sites (Table 1). Highest
imposex level was found at Shuitousha (C10) where highest mean
VDSI (4.90), RPSI (74.23) and proportion of sterile female (i.e.,
VDSI > 4; 60.0%) were recorded. Median VDSI of the whelks from
this site reached 5.00. Comparably high imposex level could be
found in R. clavigera collected at Seafood Street (C4), where mean
VDSI, RPSI and proportion of sterile female were 4.43, 72.07 and
42.9%, respectively.
VDSI significantly differed among the 10 sites (Kruskal–Wallis
test: Chi square = 51.01, p < 0.001). Shuitousha (C10) had the highest VDSI (4.90) while Meisha (C6) had the lowest VDSI (2.69;
Fig. 2a). Only two sites (i.e., Meisha and Yangmeikeng) had mean
VDSI 6 3. The lowest imposex stage recorded in all samples was
2, which was observed in individuals collected from Meisha (C6),
Yangmeikeng (D4) and Luzui (D5). The percentages of females
having VDSI = 2 at these three sites were 37.5%, 20% and 16.7%,
respectively. Individuals having VDSI of 6 were only found at
Xichong (C7), Shuitousha (C10) and Luzui (D5), and their relative
percentages to the total number of females in each site were
13.3%, 30% and 11.1%, accordingly.
RPSI ranged from 16.21 (Luzui; D5) to 74.23 (Shuitousha; C10).
The whelks in Yangmeikeng (D4) and Luzui (D5) had RPSI < 20
whereas those in Seafood Street (C4), Shuitousha (C10) and Egong
(D2) had RPSI > 60 (Fig. 2b). Sterile females (i.e., with VDSI > 4)
were found at all sites (13.3–60.0% of females) except in Meisha
(C6) and Yangmeikeng (D4) (Table 1). The whelks at Seafood Street
(C4) and Meisha (C6) had higher condition indices than those in
other sites (One-way ANOVA: F = 19.18, p < 0.001 followed by
Tukey test; Fig. 2c).
Total OTs concentration in the tissues of R. clavigera ranged
from 5,063.6 to 23,463.8 lg kg 1 dw for all of the 10 sites (Table 2),
among which significant site differences in the tissue concentration of total OTs were observed (Kruskal–Wallis test: Chi
square = 27.72, p < 0.01). The whelks in Shuitousha (C10) had the
highest concentration of total OTs while those in Yangmeikeng
(D4) had the lowest concentration (Fig. 2d).
TPT accounted for 97.8–99.1% of total OTs in the whelk tissues.
The whelks in Shuitousha (C10), again, had the highest tissue TPT
concentration (23,232.2 lg kg 1 dw). The tissue concentration of
TBT, however, only ranged from 7.2–24.3 lg kg 1 dw (i.e., 6 0.4%
of total OTs) across all sites. In general, the tissue concentrations
of total PTs (sum of MPT, DPT and TPT) were of hundred folds
higher than those of total BTs (sum of MBT, DBT and TBT) at all
sites (Table 2).
Both imposex indices showed good agreement with each other
(Spearman’s rank correlation tests between median VDSI and RPSI:
p < 0.001; Fig. 3). The proportions of sterile females were also positively correlated to imposex indices (Spearman’s rank correlation
tests: p < 0.01 for median VDSI and RPSI). No correlation was found
between condition index and imposex indices (p > 0.05). The tissue
concentrations of total OTs in R. clavigera increased with VDSI
(p < 0.001), RPSI (p < 0.01) and proportions of sterile females
(p < 0.01; Fig. 3).
The distances to the areas with heavy shipping or mariculture
activities did not show significant linear relationships with imposex status (VDSI, RPSI and% sterility) or tissue concentration of
OTs (all p > 0.05). However, there were general decreasing trends
showing that higher imposex levels and tissue OT concentrations
were found in the whelks from sites closer to marina and mariculture zone.
Median VDSI was not statistically different between 2013 (this
study) and 2006 (Chan et al., 2008) across the six study sites
K.K.Y. Ho, K.M.Y. Leung / Marine Pollution Bulletin 85 (2014) 254–260
257
Fig. 2. (a) Vas Deferens Sequence Index, VDSI; (b) Relative Penis Size Index, RPSI; (c) Condition index and (d) Total organotins (OTs) and triphenyltin (TPT) concentrations in
the tissues of the whelk Reishia clavigera. Significant differences are represented by different letters. Names and coordinates of the sampling sites are referred to Appendix A.
(Fig. 4; Wilcoxon signed ranks test: Z = 1.363, p > 0.05); yet five
out of the six sites had elevated VDSI in 2013. RPSI of R. clavigera
in all the six sites had higher RPSI in 2013 than in 2006 (Fig. 4;
Z = 2.201, p < 0.05). Furthermore, significant declines of the concentration of total BTs were observed in 2013 samples when compared with those measured in 2006 (Fig. 4; Z = 2.201, p < 0.05).
This study applied a variety of parameters, including VDSI, RPSI,
percentage of sterile females, tissue concentration of various OT
compounds and condition index to assess the vulnerability of
OTs to the whelk populations along the coasts of Dapeng Bay and
Daya Bay in South China. Condition index was suggested to be a
good indicator of health condition in marine gastropods (Lau and
Leung, 2004) and the index was generally increased in eutrophic
or moderately polluted sites (Leung and Furness, 2001). In the field,
we also noted that R. clavigera living in areas with less OT contamination tended to be smaller and grow slower when compared with
those inhabiting in OT-polluted sites in Hong Kong (Ho, 2014).
However, the present study was unable to reveal any significant
258
K.K.Y. Ho, K.M.Y. Leung / Marine Pollution Bulletin 85 (2014) 254–260
Table 2
Tissue concentrations (in lg kg 1 dry weight) of six species of organotins (monobutyltin, MBT; dibutyltin, DBT; tributyltin, TBT; monophenyltin, MPT; diphenyltin, DPT; and
triphenyltin, TPT) in the whelk Reishia clavigera collected from various sites in Dapeng Bay and Daya Bay during winter of 2013.
MBT
DBT
TBT
MPT
DPT
TPT
Total OTs
C4
C5
C6
C7
C8
C10
D2
D3
D4
D5
3.3
19.6
13.0
103.7
64.1
17905.5
18109.3
12.8
33.7
24.3
38.8
22.5
5929.3
6061.4
5.3
10.9
11.4
36.2
10.4
7544.4
7618.5
3.8
12.9
8.9
46.1
37.6
10790.0
10899.3
5.6
16.3
15.8
44.4
16.1
9903.7
10001.8
9.3
19.3
14.7
153.2
35.1
23232.2
23463.8
1.5
11.4
8.7
50.8
48.9
13894.3
14015.6
4.6
10.1
13.0
26.7
15.1
6055.2
6124.7
2.8
5.8
7.2
26.6
2.7
5018.5
5063.6
3.6
11.7
8.6
27.8
17.5
7037.7
7107.1
Fig. 3. Correlations among Vas Deferens Sequence Index (VDSI), Relative Penis Size Index (RPSI), percentage of sterile female and tissue concentration of total organotins
(OTs) in Reishia clavigera.
relationship between the condition index and imposex status or
tissue concentrations of OTs. Apart from OT contamination, other
factors such as food availability, habitat condition and biological
community structure might play a more important role to influence the body condition of R. clavigera.
As demonstrated in this study, a combined use of conventional
indicators such as VDSI, RPSI and tissue concentrations of OTs in R.
clavigera was useful to effectively assess the impact and extent of
OT contamination. In terms of temporal comparison between
2006 and 2013, imposex status of R. clavigera collected in the six
study sites along Dapeng Bay and Daya Bay did not show any
improvement, though there was a significant decline in TBT levels.
During 2000–2002, a survey reported that high VDSI values ranging from 4.07 to 5.38 were found in R. clavigera at locations along
the southeastern coast of China (sites 31 and 32; see Shi et al.,
2005). Subsequently, Chan et al. (2008) showed that both VDSI
and RPSI in R. clavigera at this region remained high in 2006.
When comparing to the survey results of Chan et al. (2008),
increases in VDSI and RPSI were recorded at most sites in 2013
(this study). Our results, therefore, suggested a deterioration of
the OT contamination in this region as shown by the tissue concentration of total OTs (in particular TPT which accounted for more
than 97.8% of total OTs) in R. clavigera increased along with elevated imposex indices in 2013. In spite of the significant decline
of BT contamination over the past seven years, the high concentration of TPT detected in R. clavigera indirectly suggested a significant
rise of TPT contamination in the region over the same period. Also,
the high imposex incidence found in R. clavigera is highly possibly
due to the emerging contamination of PTs but not BTs.
Spatial variation was found in imposex status and tissue concentrations of OTs in R. clavigera collected along the Shenzhen
coast. In general, animals living in Dapeng Bay had higher levels
of both parameters than those inhabiting in Daya Bay. As both
the potential sources of OTs, the Yantian container terminal and
the mariculture zone near Shuitousha are located in Dapeng Bay,
and thus this embayment is expected to be more polluted than
Daya Bay. These results also agreed with previous study (Deng
et al., 2010) which showed that the marine water and gastropods
K.K.Y. Ho, K.M.Y. Leung / Marine Pollution Bulletin 85 (2014) 254–260
Fig. 4. Temporal trends of (a) Vas Deferens Sequence Index (VDSI); (b) Relative
Penis Size Index (RPSI) and (c) tissue concentration of total butyltins (BTs) in Reishia
clavigera collected along the coast of Dapeng Bay. The 2006 survey was conducted
by Chan et al. (2008) while the 2013 survey refers to the current study. Between
2006 and 2013, RPSI significantly increased while tissue concentration of total BTs
significantly decreased (Wilcoxon signed ranks test: p < 0.05). Error bars in (a) and
(c) represent standard deviation.
in Daya Bay were less contaminated with TBT than those in Dapeng
Bay. Although the present results did not show significant correlations between the distances to the areas with heavy shipping and
mariculture activities and the imposex status or tissue concentrations of OTs in R. clavigera (probably due to small sample size;
n = 10 only), the spatial profile of imposex indices and OT concentrations in R. clavigera is, by and large, coincided with the pattern of
the intensity of shipping and mariculture activities in this region.
This suggested that OT contamination is likely associated with
the leaching of PTs and BTs from OT-based antifouling paints.
However, it could also be attributed to the re-suspension of these
compounds from highly contaminated sediment, contaminated
sewage effluent discharges and surface runoff contaminated with
OT-based pesticide residues from agriculture uses, as these potential sources have been demonstrated to contain PTs and BTs (Hoch,
2001; Yi et al., 2012). The exact origin of OTs, therefore, deserves
further source appointment investigations.
In the Pearl River Delta, while low concentrations of TBT were
reported in sediments, levels of TBT in marine water were comparatively very high (Fu et al., 2003). Such elevated TBT levels were
more pronounced at coastal waters in Shenzhen and Hong Kong
where intensive shipping activities occurred (Cao et al., 2009).
Therefore, R. clavigera living in this region were highly susceptible
to bioaccumulation of OTs such as TBT (Leung et al., 2006; Chan
et al., 2008). However, very few studies incorporated measurements of tissue concentration of PTs in biota. In this study, we
259
recorded low levels of BTs in the tissue of R. clavigera but very high
levels of PTs. Similar finding was obtained in a study in Xiamen,
where TPT levels in R. clavigera were higher than those of TBT
(Xie et al., 2010). Our measured TPT concentrations (up to
23,232.2 lg kg 1 dw), however, were much higher than those
measured in the same species collected in Hong Kong (up to
11,108.0 lg kg 1 dw; Ho, 2014) and in Xiamen (12.7 lg kg 1 dw;
Xie et al., 2010). Overall, concentrations of total OTs in R. clavigera
collected from Shenzhen were among the top 10% of those collected from Hong Kong. This showed that the OT contamination,
in particular TPT, in the coastal areas of Shenzhen was much more
serious than that in Hong Kong and in Xiamen.
Apart from in southern China, TPT becomes a growing concern
in other parts of China as well. Hu et al. (2009) discovered that in
Yangtze River the Chinese sturgeon Acipenser sinensis had high concentrations of TPT in different tissues. A more recent study conducted in northern part of China also detected higher TPT levels,
as compared to TBT, in the muscles of the whelk Rapana venosa
(An et al., 2013). They also showed a recent input of PTs, as
reflected by the PT degradation index which were smaller than 1
across all sites. In addition, the degradation process of TPT in
sediments is long, making it more persistent in the marine
environment (Harino et al., 1997). All these studies further
confirmed that TPT is now an emerging concern in both freshwater
and marine environments in China.
After the partial ban of OT-based antifouling system in some
countries in 1990s, recoveries from OT contamination were widely
documented in many European countries (e.g., Galante-Oliveira
et al., 2006; Oliveira et al., 2009; Guðmundsdóttir et al., 2011).
Such recoveries, however, were not observed in many locations
in Asia (e.g., Leung et al., 2006; Choi et al., 2010). In China, although
it is a member state of IMO and has ratified the AFS Convention to
ban the use of OTs in antifouling systems on sea going vessels
(IMO, 2013), local restrictions have not been implemented (Cao
et al., 2009) and thus high concentrations of OTs are still observed
in many coastal areas (Jiang et al., 2001; Wang et al., 2008; this
study). Therefore, a tightened control on the use and release of
OTs is necessary; otherwise the environmental impacts of OTs such
as the OT-mediated imposex development and sterilisation of
females in the whelks could not be reduced.
Acknowledgements
The authors thank Alex Yeung, Samuel Wang and Yanny Mak
for their assistance of sampling in the field. This work was jointly
supported by the University Grants Committee via the Area of
Excellence Scheme (Project No.: AoE/P-04/2004) and by the
Research Grants Council via a General Research Fund (Project
No.: HKU 771212 M). We thank Helen Leung and Jessie Lai for their
technical assistance and Prof. Huahong Shi for his constructive
comments on this work. Kevin Ho thanks HKU for providing him
a Type A PhD studentship.
Appendix A. Supplementary material
Supplementary data associated with this article (i.e., Appendix
A) can be found, in the online version, at http://dx.doi.org/
10.1016/j.marpolbul.2014.06.017.
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Marine Pollution Bulletin 85 (2014) 634–640
Contents lists available at ScienceDirect
Marine Pollution Bulletin
journal homepage: www.elsevier.com/locate/marpolbul
Organotin contamination in seafood and its implication for human
health risk in Hong Kong
Kevin K.Y. Ho, Kenneth M.Y. Leung ⇑
The Swire Institute of Marine Science and School of Biological Sciences, The University of Hong Kong, Pokfulam, Hong Kong, China
a r t i c l e
i n f o
Article history:
Available online 21 January 2014
Keywords:
Triphenyltin
Tributyltin
Hazard quotient
Hazard index
Human health
Endocrine disruption
a b s t r a c t
Organotins (OTs) have caused widespread adverse effects on marine organisms, while they can also
induce health problems to humans via consumption of contaminated seafood. This study aimed to quantify the tissue concentrations of OTs in 11 seafood species in Hong Kong, and assess the human health risk
for consuming these species. The tongue sole Paraplagusia blochii had the highest concentration of total
OTs. Triphenyltin (TPT) accounted for 56–97% of total OTs. The highest hazard quotient (HQ) for TPT was
1.41 in P. blochii, while the HQs for butyltins were much less than 1. The results indicated that it is likely
to have certain health risks for consuming P. blochii due to its high TPT contamination. Therefore, TPT
should be a priority pollutant of concern. Appropriate management actions should be taken to control
its use and release in the region in order to safeguard the marine ecosystem and human health.
Ó 2014 Elsevier Ltd. All rights reserved.
1. Introduction
Organotin compounds (OTs), in particular tributyltin (TBT) and
triphenyltin (TPT), are well-known endocrine disruptors and have
contaminated our environments for more than 40 years (Yi et al.,
2012). Ever since their wide application as biocides in antifouling
systems, aquaculture facilities and agriculture starting from the
1960s, OTs have caused widespread adverse effects to many marine and freshwater organisms (Clarke and Smith, 2011). These
chemicals have been known to induce the abnormal development
of imposex, intersex and female masculinisation in over 260 species of gastropods, and inhibit the growth and development of oysters (Titley-O’Neal et al., 2011). The International Maritime
Organization (IMO) has therefore enacted a mandatory global
ban on the application of OT-based antifouling systems on all seagoing vessels since September 2008 (IMO, 2008). As a result, it is
logical to anticipate that there would be a reduction in OT contaminations in marine waters around the world.
Many marine organisms, however, are still suffering from OT
contaminations. As OTs can be easily accumulated in biota as well
as along the food chain, marine organisms at higher trophic levels
are more susceptible to OTs (Howell and Behrends, 2010). Marine
mammals, for example, being the top predator in the marine ecosystem, have higher concentrations of OTs than their prey (Kannan
and Falandysz, 1997).
⇑ Corresponding author at: School of Biological Sciences, The University of Hong
Kong, Pokfulam, Hong Kong, China. Tel.: +852 22990607; fax: +852 25176082.
E-mail addresses: kyho2@hku.hk (K.K.Y. Ho), kmyleung@hku.hk (K.M.Y. Leung).
0025-326X/$ - see front matter Ó 2014 Elsevier Ltd. All rights reserved.
http://dx.doi.org/10.1016/j.marpolbul.2013.12.039
Previous studies showed that OTs are probably able to induce
harmful health effects to humans including reproductive and
developmental abnormalities, immunosuppression and possible
carcinogenic activity (Antizar-Ladislao, 2008). For example, TBT
or TPT can inhibit enzyme activity in ovary cells at concentrations
as low as 2 ng mL1 (Saitoh et al., 2001) and promote development
of prostate cancer cells at 100 nM (Yamabe et al., 2000). Humans
are exposed to OTs mainly through three ways namely skin contact, inhalation and ingestion (WHO, 1980). Among them, dietary
consumption through contaminated seafood is regarded as the major pathway of OT intake (Yi et al., 2012). These compounds are
degradable via bacterial action (i.e., biodegradation) and light irradiation (i.e., photodegradation). However, they cannot be destroyed by cooking (Willemsen et al., 2004). In general, the
higher the trophic levels the organisms are, the more OTs are accumulated in their body tissues. Some predatory fishes, such as tuna,
salmon, mackerel and cod, are regularly consumed by humans and
contribute to nearly 38% of our total OT exposure (Guérin et al.,
2007).
Hong Kong has a long history of OT contamination (Ko et al.,
1995; Leung et al., 2006; Qiu et al., 2011) although the applications
of OT-based antifouling paints on small ships (<25 m in length) and
fish cages have been banned since 1992. In the freshwater and
marine environments of Hong Kong, these chemicals are still
detected in water, sediments and biota. For instance, TBT and its
degradation products were detected up to 23.2 lg L1 and
38.6 lg g1 dry weight (dw) in Hong Kong’s river waters and sewage sludge respectively (Kueh and Lam, 2008). TBT concentrations
in marine sediments could be as high as ca. 130,000 lg kg1 dw
635
K.K.Y. Ho, K.M.Y. Leung / Marine Pollution Bulletin 85 (2014) 634–640
(53,000 ng Sn g1; see Ko et al., 1995). In biota, we have
recently detected a very high TPT concentration (up to
11,279 lg kg1 dw) in the tissue of the rock shell Reishia clavigera
(Ho and Leung, unpubl. data). However, we also noted that TBT
concentrations in R. clavigera were not very high when compared
to that of TPT.
Like most coastal cities, Hong Kong has easy accessibility towards fisheries resources and maritime trade. Traditionally, Hong
Kong people consume a large amount of seafood in their daily life.
According to a food consumption survey performed during 2005–
2007, the average consumption rates of fish and molluscs for Hong
Kong people were 57.48 g d1 and 5.95 g d1 respectively (CFS,
2010). However, the figures increased to 87.9 g d1 (32.1 kg yr1)
for fish and 52.3 g d1 (19.1 kg yr1) for molluscs according to
the most up-to-date figures documented by Food and Agriculture
Organisation (FAO, 2012). These seafood consumption rates ranked
among the highest in the world (see Appendix 1). As intake of contaminated seafood is a major pathway of human exposure towards
OTs, the human health risk associated with OTs is expected to be
high to Hong Kong people.
In a recent report published by the Government of the Hong
Kong Special Administrative Region, OTs have been highlighted
as one of the top seven endocrine disrupting chemicals of concern
in food (CFS, 2012). However, only a few studies reported OT concentrations in seafood (e.g., Harino et al., 2000; Chen et al., 2008).
No surveys of OT contamination, including both butyltin (BTs) and
phenyltin compounds (PTs), have been conducted in the seafood
from Hong Kong and Pearl River Delta region of South China. The
associated human health risk concerning OTs from consuming seafood has never been evaluated in this region.
This study, therefore, aimed to (1) investigate, for the first time,
the OT tissue concentrations, including both butyltin and phenyltin
compounds, in selected seafood species in Hong Kong markets; (2)
identify the species or the group of seafood that contains high levels of OTs and (3) perform human health risk assessments to investigate the potential health risk under two estimated seafood
consumption rates of Hong Kong people.
2. Materials and methods
This study was designed to measure the tissue concentrations
of OTs (i.e., mono-BT, di-BT and TBT, mono-PT, di-PT and TPT) in
11 commonly available seafood species in Hong Kong including
three gastropod species (Babylonia areolata, Bufonaria rana and
Hemifusus tuba), two bivalve species (Meretrix lusoria and Ruditapes
variegatus) and six fish species (Collichthys lucidus, Harpadon nehereus, Johnius belangerii, Nibea albiflora, Paraplagusia blochii and Siganus canaliculatus). The results were then applied to the human
health risk assessment associated with OT-contaminated seafood.
2.1. Sample preparation
2.1.1. Gastropods and bivalves
About 20–40 individuals of each species were bought from two
wet markets, namely Tai Po Market (22°260 46.7200 , 114°090 59.1800 )
and Mongkok (22°190 4.400 , 114°100 0.5200 ), during the wet season
of 2012 in Hong Kong. All samples were of marketable size (see
Appendix 2). They were iced immediately after purchase and
transferred to laboratory. Identification of the animals followed
the descriptions and keys in Yang et al. (1992), Shao et al. (1996)
and Zhang (2008). For gastropods, we measured the shell length
of the gastropods by callipers (to the nearest 0.1 mm). For bivalves,
shell width and height were measured instead. Whole body wet
weight (with shell) and tissue wet weight were also measured by
an electronic balance (to the nearest 1 mg) for all animals. The soft
body tissues were taken out for the following analysis.
Imposex identification was conducted for the three gastropod
species following the generalised scheme described in Muenpo
et al. (2011). Features of their reproductive systems were observed
under a stereomicroscope (Olympus SZH10). Penis, if present, was
straightened and the length was measured using the scale in the
microscopic lens (to the nearest 0.1 mm). We used two imposex
indices, namely Vas Deferens Sequence Index (VDSI) and Relative
Penis Size Index (RPSI), to access the severity of imposex development. VDSI measures the progressive imposex development by seven stages (0–6). In brief, stage 0 means no imposex developed and
stage 6 indicates the most severe stage of imposex. Stages 5 and 6
indicate the infertility of the female due to the blockage of the oviduct opening (Muenpo et al., 2011). RPSI is the fraction between
the mean bulk of the female penis and that of the male penis
(Gibbs et al., 1987). The bulk of the penis can be expressed as the
cube of its length, thus:
RPSI ¼ ðMean length of female penisÞ
3
100=ðMean length of male penisÞ
3
For each species, five replicates were analysed for tissue concentrations of OTs and each replicate contained a pool of 4–8
individuals.
2.1.2. Fish
Fishes were obtained from two commercial shrimp trawlers
operated around western (22°120 14.400 , 113°550 58.800 ) and southern
Hong Kong waters (22°120 39.600 , 114°170 600 ) during the wet season
of 2012. Each trawler out-rigger was of 16.9 m. The trawler operated at a speed of 5–7 km h1 for 30 min at each site. Samples were
taken from nine replicate nets (beam size: 3.28 m; stretched mesh
size: 1.3–2.5 cm). Fish samples were immediately frozen after
landing and transferred to laboratory for identification following
the descriptions in Shen (1993) and AFCD (2012). Total length,
standard length (if applicable) and wet weight were measured
for every individual. Only the dorsal muscle was dissected and
used for chemical analysis. For large fishes, each individual was
treated as a replicate, while for small fishes, several individuals
were pooled as a replicate. Five replicates per species were analysed for OTs concentrations. Names of all studied seafood species
were checked against and followed the World Registry of Marine
SpeciesÒ (WoRMSÒ, 2013).
2.2. Chemical analysis
Analysis of BTs and PTs followed the protocol described by
Guðmundsdóttir et al. (2011) with slight modifications. Quantification of OTs was performed using a gas chromatograph (GC; Bruker
450-GC, Bruker Inc., Billerica, MA, USA) equipped with a massselective detector (Bruker 320-MS, Bruker Inc., Billerica, MA,
USA). A VF-5MS fused silica capillary having 0.25 mm i.d. 30 m
0.25 m film thickness (Bruker Inc., Billerica, MA, USA) was used
as the GC column. The certified reference material ERM-CE477
(mussel tissue) validated the method previously with recoveries
of 82–92%. Also, a surrogate standard (di-n-heptyltin dichloride)
was spiked into each sample to check the recovery. Procedural
blanks were analysed simultaneously with each batch of five samples to check for any interference or contamination during the
analysis. Limits of detection ranged from 0.2 to 1.5 lg kg1 dw of
the six compounds. No correction was made for the recoveries of
surrogate standard to the concentrations reported.
All standards were bought from Sigma–Aldrich (St. Louis, MO,
USA) and Chiron (Trondheim, Norway). All solvents were in HPLC
Grade bought from Tedia (Fairfield, OH, USA).
636
K.K.Y. Ho, K.M.Y. Leung / Marine Pollution Bulletin 85 (2014) 634–640
2.3. Statistical analysis on OT contamination
Data of tissue concentrations of MBT, DBT, TBT, total BTs (i.e.,
summation of MBT, DBT and TBT), MPT, DPT, TPT, total PTs (i.e.,
summation of MPT, DPT and TPT) and total OTs (i.e., summation
of all six OTs) in gastropods, bivalves and fish species were analysed, respectively. Data in each group of animals (gastropod, bivalve or fish) were tested for homogeneity of variances using
Levene’s test, followed by Student’s t-test (for two samples) or
One-Way Analysis of Variance (ANOVA; for more than two samples) in order to infer the mean differences between/among species. If ANOVA confirmed the difference among the mean values,
a post hoc Tukey’s multiple comparison test would be used to identify significantly different means. If the data failed the assumption
of homogeneity of variances, they were logarithmic-transformed
(base 10) and then proceeded to ANOVA, or used non-parametric
Kruskal Wallis test followed by Dunn’s test. Spearman rank correlation analyses were conducted among total OTs, total BTs, total
PTs, TBT and TPT. For all statistical tests, the significance level
was set as 0.05.
2.4. Human health risk assessment
2.4.1. Hazard quotient and hazard index
Non-cancer hazard quotient (HQ) was used to determine the
health risk of TBT and TPT towards humans. It was calculated as:
2.4.2. Tolerable average residue level
The tolerable average residue level (TARL) measures the level of
a particular contaminant in seafood that is tolerable for the average
consumer with an average weight of 60 kg (Belfroid et al., 2000). It
is calculated as:
TARL ¼ ðTDI 60 kg body weightÞ
=average daily seafood consumption
The tolerable daily intake (TDI) of TBT is 0.25 TBTO lg kg1 body weight (bw) d1 which was derived from a toxicity endpoint
based on the immune response of rats, and a safety factor of 100
has been applied to extrapolate the result from rats to humans
(Penninks, 1993). With reference to the two estimates of average
daily seafood consumption of Hong Kong people (Table 1), TARLs
of TBT for fish and molluscs were calculated, respectively (Appendix 4).
TARL of TPT was also computed using the acceptable daily intake (ADI) instead of TDI (Eguchi et al., 2010). ADI was established
as 0.5 TPT lg kg1 bw d1 by Joint Food and Agricultural Organization (FAO)/World Health Organization (WHO) Meeting on Pesticide
Residues (JMPR, 1992). TARLs of TPT for fish and molluscs were
also shown in Appendix 4.
3. Results
3.1. Imposex status in gastropods
HQ ¼ ½ðCF IR EF EDÞ=ðBW ATÞ=RfD
where CF is the measured chemical concentration (lg g1 wet
weight). The original dry weight data were converted to wet weight
according to different dry weight-wet weight ratios of different species (Appendix 3). IR is the ingestion rate of seafood (g d1). Two
sets of data, from CFS (2010) and FAO (2012) respectively, were
used in the calculation (see Table 1). EF is the exposure frequency
(365 d yr1). ED is the exposure duration (yr). Males have an average life expectancy at birth of 80.5 yr in Hong Kong while females
have 86.7 yr (CHP, 2012). Therefore, an average of 83.6 yr was used.
BW is the body weight (kg). Average body weight of an adult Asian
is normally assumed to be 60 kg (Belfroid et al., 2000). AT is the
average age time (d) = 365 d yr1 83.6 yr = 30,514 d. RfD is the
oral reference dose (lg kg1 body weight d1). RfD for TBT is
0.25 lg kg1 body weight d1 (USEPA, 1997) and 0.5 lg kg1 body
weight d1 for TPT (WHO, 1992). A combined RfD for TBT and
DBT of 0.25 lg kg1 body weight d1 is suggested by Penninks
(1993) and Belfroid et al. (2000).
Hazard index (HI) was determined for each selected seafood
species based on the overall non-carcinogenic effect of the chemicals. HI is equal to the sum of all individual HQs, i.e., summation of
HQTBT+DBT and HQTPT (Lee et al., 2005). If HQ or HI was less than 1, a
daily exposure of the seafood to the respective compound(s) is unlikely to cause any deleterious effects to human. Otherwise, it is
likely to pose adverse health risks. Average and maximum HQs
and HIs were calculated to provide more conservative estimates
on the potential health risks of OTs via seafood consumption.
Table 1
Per capita seafood consumption rate of Hong Kong people in terms of food fish supply.
Data are extracted from CFS (2010) and FAO (2012).
Total
Molluscs
Fish
a
b
c
CFS (2010)
FAO (2012)
70.78 g d1
5.95 g d1
57.48 g d1
65.9 kg yr1a (ca. 180.5 g d1)c
19.1 kg yr1 (ca. 52.3 g d1)c
32.1 kg yr1b (ca. 87.9 g d1) c
Total food fish excluding freshwater and diadromous fish.
Including only demersal and pelagic fish, and marine fish other.
Calculation based on the assumption of 365 days per year.
Among the three gastropod species, only B. areolata and H. tuba
showed imposex development, while no imposex incidence was
observed in B. rana. All females of B. areolata (i.e., 100%) showed
imposex, whereas 31.8% of female H. tuba displayed imposex.
Average VDSI for B. areolata and H. tuba were 4.17 and 1.00 respectively, while RPSI were recorded as 0.399 and 0.005 for the two
species. One-third of the female B. areolata was found to be infertile (i.e., having VDSI > 4), while the VDSI stages of all H. tuba were
below 4.
3.2. Tissue organotin concentration
An average recovery of the surrogate standard was found to be
71% (range: 45–102%). In agreement with imposex results, B. areolata also had the highest total OTs and TPT tissue concentrations
among the three gastropod species; the corresponding average
concentrations were 1751.4 lg kg1 dw and 1695.0 lg kg1 dw
respectively (Fig. 1). Bufonaria rana and H. tuba had significant
lower levels of total OTs and TPT (ANOVA using log10-transformed
data; total OTs: F0.05(1),1,12 = 9028.499, p < 0.001; TPT: F0.05(1),1,12 =
7867.380, p < 0.001). Bufonaria rana had a significantly lower level
of TBT compared to B. areolata and H. tuba (ANOVA using
log10-transformed data; F0.05(1),1,12 = 11.329, p < 0.01; Fig. 1).
For the bivalves, higher concentrations of total OTs, TPT and TBT
were found in R. variegatus than those in M. lusoria (Mann–Whitney
U test: All U = 25, p < 0.01; Fig. 1). Among the fishes, P. blochii had
the highest concentrations of total OTs and TPT at 2325.8 and
2237.7 lg kg1 dw, respectively (ANOVA; total OTs: F0.05(1),5,24 =
4.032, p < 0.01; TPT: F0.05(1),5,24 = 4.718, p < 0.01). Total OTs in other
five fish species ranged from 957.7 to 1346.5 lg kg1 dw and TPT
ranged from 905.7 to 1217.4 lg kg1 dw. Higher TBT concentrations
were in C. lucidus and H. nehereus (Kruskal–Wallis test; X 20:05;5 ¼
18:874, p < 0.01). TBT tissue concentrations of C. lucidus and
H. nehereus were 62.9 and 105.3 lg kg1 dw respectively, while
TBT concentrations in other fish species ranged from 18.3 to
43.0 lg kg1 dw (Fig. 1).
TPT was the most predominant compound among the six OT
residues, ranging from 75.5 (H. tuba) to 2237.7 (P. blochii) lg kg1
K.K.Y. Ho, K.M.Y. Leung / Marine Pollution Bulletin 85 (2014) 634–640
637
Fig. 1. Tissue concentrations (mean and SE for bar charts, median for box plots; n = 5) of organotins (OTs) including total OTs, triphenyltin (TPT) and tributyltin (TBT) in each
of the 11 selected common seafood species in Hong Kong. All concentrations were in lg kg1 dry weight.
dw and accounting for 55.5–96.8% of total OT concentrations
(Fig. 2). TBT was ranked the second, representing only 0.6–18.6%
of total OT concentrations (Fig. 2). Thus, the ratios between TPT
and TBT were consistently greater than 1 in all species (Appendix
5).
Tissue TBT concentrations were positively correlated to that of
total BTs (Spearman rank correlation: rs 0.05(2),9 = 0.936, p < 0.001).
Significant positive correlations were also found between TPT
and total PTs, TPT and total OTs, and total PTs and total OTs (all
rs 0.05(2),9 = 1.000, p < 0.001; Appendix 6). Nevertheless, tissue concentrations of TPT were not significantly correlated to TBT concentrations (rs 0.05(2),9 = 0.364, p > 0.05). No significant relationship was
found between total BTs and total PTs as well (rs 0.05(2),9 = 0.391,
p > 0.05).
Butyltin degradation indices (BDIs = ([MBT] + [DBT])/[TBT]) ranged from 0.16 to 3.29. Phenyltin degradation indices (PDIs =
([MPT] + [DPT])/[TPT]) of all species were below 0.27 (Appendix 5).
3.3. Human health risk assessment
The mean and maximum HQs for TBT and TBT + DBT were below 1 (all 60.11) for all species. However, the maximum HQ of
TPT exceeded 1 in P. blochii (1.30) under the estimate of seafood
consumption rate given by CFS (2010). For the FAO’s estimate of
seafood consumption rate, the maximum HQ for TPT exceeded 1
in P. blochii (2.00), N. albiflora (1.24) and B. areolata (1.05). Only
P. blochii had the mean HQ for TPT larger than 1 (i.e., 1.41; Fig. 3).
Similar patterns were observed for HIs. Under the estimate of
seafood consumption rate by the Centre for Food Safety of the
Hong Kong SAR Government, only P. blochii showed a maximum
HI > 1. While for FAO’s estimate, P. blochii, N. albiflora and B. areolata had maximum HIs > 1 and only P. blochii exhibited the mean
HI > 1 (Fig. 3).
Tissue TBT concentrations of all species were well below the
TARLs for TBT calculated from the two estimates of seafood consumption rates in Hong Kong (Fig. 4). Molluscs had tissue TBT concentrations <1.7% of TARLs from both estimates, while TBT
concentrations of fish were <6.3% of the TARLs. For TPT, all molluscs had tissue TPT concentrations lower than the TARLs while
in fish, only P. blochii had an average tissue TPT concentration
(480.3 lg kg1 wet weight; ww) significantly higher than the TARL
of TPT estimated from FAO’s seafood consumption rate (i.e.,
341.1 lg kg1; One-sample t test: t0.05(1),4 = 2.484, p < 0.05; Fig. 4).
4. Discussion
Fig. 2. Relative concentrations of different residues of organotins (OTs), including
tributyltin (TBT) and triphenyltin (TPT), in each of the 11 selected common seafood
species in Hong Kong.
Imposex is one of the most sensitive biomarkers to OT pollution.
However, not all gastropod species showed imposex upon exposure to OTs because they vary in sensitivities (Fent, 1996). The results of the present study agreed with previous finding that B. rana
showed no imposex development in Hong Kong because of its low
638
K.K.Y. Ho, K.M.Y. Leung / Marine Pollution Bulletin 85 (2014) 634–640
Fig. 3. Hazard quotients (HQs) for triphenyltin (TPT) and hazard indices (HIs) for the studied gastropods and fishes. The vertical dotted line represents the critical point at HQ
or HI equals to 1. Results for bivalves were not shown because their HQ or HI values were below 1.
Fig. 4. Tissue tributyltin (TBT) and triphenyltin (TPT) concentrations of molluscs (i.e., gastropods and bivalves; a and b) and fishes (c and d), with indications of tolerable
average residue levels (i.e., the two vertical dotted lines) calculated from the two estimates of seafood consumption rates of Hong Kong people from CFS (2010) and FAO
(2012). Asterisk () indicates that tissue concentration was significantly higher than the tolerable level at p < 0.05.
sensitivity to OTs and living in deeper habitats compared to intertidal species (Proud and Richardson, 1997). Nevertheless, the same
species has been recorded with imposex elsewhere (= Bursa rana,
see Titley-O’Neal et al., 2011).
Conversely, B. areolata was once believed to be less sensitive to
develop OT-mediated imposex (Swennen et al., 2009). In the present study, however, 100% imposex was found in B. areolata which
also had the highest OT concentrations among the three studied
gastropod species. The severity of OT contamination in the marine
environment of Hong Kong could have made this insensitive species much vulnerable to OTs.
Our results showed different OT profiles from other places in
the world. In the fishes from Japan, TBT was detected at a range
of 11–182 lg kg1 wet weight (ww) while the range of TPT was
<1–130 lg kg1 ww only (Harino et al., 2000). Our TBT concentrations in fishes detected in the present study were about 17 folds
lower than their results. For TPT, however, our results were at least
4 times higher than the fish in Japan measured by Harino et al.
(2000). In Korea, the predominant residues of OTs in shellfish were
from BTs. The highest TBT concentration in mussels was
23.0 lg kg1 ww (Oh, 2009), which was five times lower than the
maximum TBT concentration of the shellfish observed in the
current study. Our maximum TPT concentration in shellfish was
>120 folds higher than that in Korea. With the existing information
of OT profiles in the seafood among these places, it demonstrated
that the seafood in Korea and Japan was dominated by BTs whereas
the seafood in Hong Kong was dominated by PTs instead.
Similar patterns were observed in Taiwan as in Hong Kong,
where concentrations of PTs in fish were higher than those of BTs
(Lee et al., 2005). In Taiwan, the ratios of TPT/TBT and PTs/BTs were
above 1 for all the species sampled in summer, although a few
species sampled in winter had the ratios below 1. Lee et al. (2005)
suggested that TPT is more resistant to degradation due to its chemical structure and hence it has a higher bioaccumulative potential to
fishes compared to that of TBT. Nevertheless, TPT concentrations in
fishes in Hong Kong (range: 905.7–2237.7 lg kg1 dw; this study)
were significantly higher than those in Taiwan (range: 120.7–
655.2 lg kg1 dw; Lee et al., 2005). In particular, the highest tissue
TPT concentration found in P. blochii in Hong Kong was about 3.5
times higher than that in Pagrus major from Taiwan (Lee et al.,
2005) in which both of them are demersal species. The applications
of TPT as fungicide and pesticide from agricultural fields were suspected to contribute a great extent of TPT contamination in marine
waters in Taiwan. Hong Kong, however, with limited agricultural
K.K.Y. Ho, K.M.Y. Leung / Marine Pollution Bulletin 85 (2014) 634–640
activities nowadays, is not expected to have agriculture-based TPT
pollution as a major source.
In addition, PDIs for fishes in Hong Kong were all below 0.05
(Appendix 5), although in Taiwan only half of the studied fish
species had PDI < 1 (Lee et al., 2005). The present results clearly
indicated that TPT was more predominant than their mono- or
di-substituted counterparts in fishes and molluscs in Hong Kong
waters. The results also agreed with previous studies by Suzuki
et al. (1992) and Uhler et al. (1993). The high proportion of TPT
to total PTs could be due to recent inputs of TPT and/or slow degradation of TPT in the marine environment.
Interestingly, there were no correlations between the concentrations of TBT and TPT, nor the concentrations of total BTs and total PTs in this study. These did not agree with previous findings in
Taiwan (Lee et al., 2005) and in Japan (Harino et al., 1998), in which
they showed negative correlations between TBT and TPT in fishes,
and positive correlations between these two OT compounds in
mussels respectively. It is, therefore, speculated that there are different contamination sources for TBT and TPT in Hong Kong. Given
that there are high concentrations of TPT in Hong Kong’s seafood
species and little is known regarding the origin of OTs pollution,
source-appointment studies are therefore urgently needed in Hong
Kong.
Fishes living in different habitats varied in tissue OT concentrations. In this study, P. blochii, which is a demersal fish living on the
sediment, had the highest tissue OT concentrations. Lee et al.
(2005) also showed that demersal fishes had higher concentrations
of PTs compared to pelagic or cultured species. Organisms living
closer to the sediment showed higher TPT/TBT ratios as there were
more TPT in the sediment which could be taken up by the organisms (Stäb et al., 1996).
In this study, we only analysed the OT concentrations in muscles of the fish. However, many organs/tissues of fish could have
much higher OTs than that of muscles. For instance, liver was demonstrated as the sink for BTs and PTs in fish body (Stäb et al., 1996),
and a large amount of TBT and TPT could also be found in gills and
digestive systems of some deep sea fishes (Borghi and Porte, 2002).
Therefore, tissue-specific accumulation of OTs should be taken into
consideration for future risk assessment study. The potential risks
of consuming the seafood could be under-evaluated if tissue concentration is only measured in muscles and other tissue types
are also consumed by people.
The elevated seafood consumption of Hong Kong people led to
higher possibility to take in more OTs plus other chemical contaminants through contaminated seafood. It should be noted that in
the previous survey conducted by CFS (2010) the 97.5th percentile
of seawater fish consumption was 250.0 g d1, which was actually
higher than the estimation from FAO. Therefore, the risk of OTs
through fish intake by the high consumption group should not be
underestimated in Hong Kong.
With increasing seafood consumption by Hong Kong people as
well as growing concerns on food safety, the present study addressed the chronic sub-lethal health risk of OTs to humans. The
results of this study also filled the gap of OTs distribution among
marine seafood species in Hong Kong, with a hope that more species could be included in future studies thereby having a more
comprehensive understanding to safeguard human health.
5. Conclusions
We quantified the concentrations of six species of OTs in 11
selected market seafood species in Hong Kong. The tongue sole
P. blochii had the highest concentration of total OTs, in which over
90% were TPT. Under the current seafood consumption rate of
Hong Kong people, the mean hazard quotient for TPT for this
639
species was larger than 1, indicating potential health risk of eating
this species. To reduce the health risk of OTs, Hong Kong people
should avoid consuming too much of this species. Source appointment studies should be conducted to fully understand the sources
of OTs. Appropriate actions such as legislation and restriction of
the use of OTs, in particular TPT, in antifouling paints and biocide
products should be taken by the government with a view to controlling their use and release, and thus safeguarding the marine
ecosystem and human health.
Acknowledgements
The authors thank the student helpers who assisted in sampling
and species identification. This work was jointly supported by the
University Grants Committee via the Area of Excellence Scheme
(Project No.: AoE/P-04/2004) and by the Research Grants Council
via a General Research Fund (Project No.: HKU 771212M). The
authors also thank Helen Leung and Jessie Lai for their technical
assistance. Kevin Ho thanks the University of Hong Kong to provide
him a Type A Ph.D. studentship. This work was presented at the 7th
International Conference on Marine Pollution and Ecotoxicology by
Kevin Ho who was awarded The Marine Pollution Bulletin Young
Scientist Award for an Outstanding Oral Presentation.
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.marpolbul.2013.
12.039.
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