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Toxicopathological Effects of the Sunscreen UV Filter, Oxybenzone (Benzophenone-3), on Coral Planulae and Cultured Primary Cells and Its Environmental Contamination in Hawaii and the U.S. Virgin Islands

  • Haereticus Environmental Laboratory

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Benzophenone-3 (BP-3; oxybenzone) is an ingredient in sunscreen lotions and personal-care products that protects against the damaging effects of ultraviolet light. Oxybenzone is an emerging contaminant of concern in marine environments-produced by swimmers and municipal, residential, and boat/ship wastewater discharges. We examined the effects of oxybenzone on the larval form (planula) of the coral Stylophora pistillata, as well as its toxicity in vitro to coral cells from this and six other coral species. Oxybenzone is a photo-toxicant; adverse effects are exacerbated in the light. Whether in darkness or light, oxybenzone transformed planulae from a motile state to a deformed, sessile condition. Planulae exhibited an increasing rate of coral bleaching in response to increasing concentrations of oxybenzone. Oxybenzone is a genotoxicant to corals, exhibiting a positive relationship between DNA-AP lesions and increasing oxybenzone concentrations. Oxybenzone is a skeletal endocrine disruptor; it induced ossification of the planula, encasing the entire planula in its own skeleton. The LC50 of planulae exposed to oxybenzone in the light for an 8- and 24-h exposure was 3.1 mg/L and 139 µg/L, respectively. The LC50s for oxybenzone in darkness for the same time points were 16.8 mg/L and 779 µg/L. Deformity EC20 levels (24 h) of planulae exposed to oxybenzone were 6.5 µg/L in the light and 10 µg/L in darkness. Coral cell LC50s (4 h, in the light) for 7 different coral species ranges from 8 to 340 µg/L, whereas LC20s (4 h, in the light) for the same species ranges from 0.062 to 8 µg/L. Coral reef contamination of oxybenzone in the U.S. Virgin Islands ranged from 75 µg/L to 1.4 mg/L, whereas Hawaiian sites were contaminated between 0.8 and 19.2 µg/L. Oxybenzone poses a hazard to coral reef conservation and threatens the resiliency of coral reefs to climate change.
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Toxicopathological Effects of the Sunscreen UV Filter,
Oxybenzone (Benzophenone-3), on Coral Planulae and Cultured
Primary Cells and Its Environmental Contamination in Hawaii
and the U.S. Virgin Islands
C. A. Downs
Esti Kramarsky-Winter
Roee Segal
John Fauth
Sean Knutson
Omri Bronstein
Frederic R. Ciner
Rina Jeger
Yona Lichtenfeld
Cheryl M. Woodley
Paul Pennington
Kelli Cadenas
Ariel Kushmaro
Yossi Loya
Received: 17 July 2015 / Accepted: 13 September 2015
Springer Science+Business Media New York 2015
Abstract Benzophenone-3 (BP-3; oxybenzone) is an
ingredient in sunscreen lotions and personal-care products
that protects against the damaging effects of ultraviolet
light. Oxybenzone is an emerging contaminant of concern
in marine environments—produced by swimmers and
municipal, residential, and boat/ship wastewater dis-
charges. We examined the effects of oxybenzone on the
larval form (planula) of the coral Stylophora pistillata, as
well as its toxicity in vitro to coral cells from this and six
other coral species. Oxybenzone is a photo-toxicant;
adverse effects are exacerbated in the light. Whether in
darkness or light, oxybenzone transformed planulae from a
motile state to a deformed, sessile condition. Planulae
exhibited an increasing rate of coral bleaching in response
to increasing concentrations of oxybenzone. Oxybenzone is
a genotoxicant to corals, exhibiting a positive relationship
between DNA-AP lesions and increasing oxybenzone
concentrations. Oxybenzone is a skeletal endocrine dis-
ruptor; it induced ossification of the planula, encasing the
entire planula in its own skeleton. The LC
of planulae
exposed to oxybenzone in the light for an 8- and 24-h
exposure was 3.1 mg/L and 139 lg/L, respectively. The
s for oxybenzone in darkness for the same time points
were 16.8 mg/L and 779 lg/L. Deformity EC
(24 h) of planulae exposed to oxybenzone were 6.5 lg/L in
the light and 10 lg/L in darkness. Coral cell LC
the light) for 7 different coral species ranges from 8 to
340 lg/L, whereas LC
s (4 h, in the light) for the same
species ranges from 0.062 to 8 lg/L. Coral reef
Electronic supplementary material The online version of this
article (doi:10.1007/s00244-015-0227-7) contains supplementary
material, which is available to authorized users.
&C. A. Downs
Haereticus Environmental Laboratory, P.O. Box 92, Clifford,
VA 24533, USA
Department of Zoology, George S. Wise Faculty of Life
Sciences, Tel Aviv University, 69978 Tel Aviv, Israel
Avram and Stella Goldstein-Goren Department of
Biotechnology Engineering and the National Institute for
Biotechnology in the Negev, Ben-Gurion University of the
Negev, 84105 Beer Sheva, Israel
Department of Biology, University of Central Florida, 4000
Central Florida Boulevard, Orlando, FL 32816-2368, USA
Pacific Biosciences Research Center, University of Hawaii,
Honolulu, HI 96822, USA
Department of Life Sciences, Ben-Gurion University of the
Negev, Beer Sheva 84105, Israel
Hollings Marine Laboratory, U.S. National Oceanic &
Atmospheric Administration, 331 Ft. Johnson Rd.,
Charleston, SC 29412, USA
Center for Coastal Environmental Health and Biomolecular
Research, U.S. National Oceanic & Atmospheric
Administration, 219 Ft. Johnson Rd., Charleston, SC 29412,
National Aquarium, 501 East Pratt Street, Baltimore,
MD 21202, USA
Arch Environ Contam Toxicol
DOI 10.1007/s00244-015-0227-7
contamination of oxybenzone in the U.S. Virgin Islands
ranged from 75 lg/L to 1.4 mg/L, whereas Hawaiian sites
were contaminated between 0.8 and 19.2 lg/L. Oxyben-
zone poses a hazard to coral reef conservation and threat-
ens the resiliency of coral reefs to climate change.
Oxybenzone (BP-3; benzophenone-3; 2-hydroxy-4-
methoxphenyl phenylmethanone; CAS No. 131-57-7) often
is used as an active ingredient in sunscreen lotions and
personal-care products, such as body fragrances, hair-sty-
ling products, shampoos and conditioners, anti-aging
creams, lip balms, mascaras, insect repellants, as well as
dishwasher soaps, dish soaps, hand soaps, and bath oils/
salts (CIR 2005;
184390-oxybenzone). BP-3 and other benzophenone
derivatives often are found as contaminants in boating,
residential, and municipal wastewater effluents and are
considered ‘‘emerging environmental contaminants of
concern’’ by the U.S. Environmental Protection Agency
(Eichenseher 2006; Richardson 2006,2007; Blitz and
Norton 2008; Gago-Ferrero et al. 2011; Kameda et al.
2011; Rodil et al. 2012; Aquero et al. 2013).
Between 6000and 14,000 tons of sunscreen lotion, manyof
which contain between 1 and 10 % BP-3, are estimated to be
released intocoral reef areas each year, putting at least 10 % of
the global reefs at risk of exposure, and approximately 40 % of
coral reefs located along coastal areas at risk of exposure
(Shaath and Shaath 2005;UNWTO2007; Danovaro et al.
2008; Wilkinson 2008). In Okinawa, BP-3 levels on coral
reefs that were 300–600 m away from public swimming
beaches ranged from 0.4 to 3.8 pptrillion (Tashiro and
Kameda 2013); in South America, sediments near coral
communities/reefs contained BP-3 concentrations between 54
and 578 pptrillion (Baron et al. 2013). Schlenk et al. (2005)
discovered through a Toxicity Identification Evaluation that
BP-3 was unequivocally identified as the source of estrogenic
activity in marine sediments near wastewater outfalls.
Although the half-life in seawater is several months, BP-3 can
act as a pseudo-persistent pollutant; its contamination of a site
may be constantly renewed, resulting in ecological receptors
experiencing persistent exposure (Vione et al. 2013). Con-
cerns regarding the adverse impacts of exposure to BP-3 on
coral reefs and other marine/aquatic ecosystems have led to
either banning oxybenzone-containing products in marine-
managed areas (e.g. Mexico’s marine ecoparks; Xcaret 2007;
´2007) or public relations campaigns by management
agencies to encourage reduction of environmental contami-
nation of sunscreen lotions by swimmers (e.g. ‘‘Protect
Yourself, Protect the Reef’’ Bulletin U.S. NPS 2012).
BP-3 exhibits a number of toxicological behaviors
ranging from the molecular level to multi-organ system
pathologies (Gilbert et al. 2012). Benzophenones,
including BP-3, are documented mutagens that increase the
rate of damage to DNA, especially when exposed to sun-
light (Popkin and Prival 1985; Zeiger et al. 1987; Know-
land et al. 1993; NTP 2006). BP-3 produced a positive
mutagenic response by inducing the umu operon (geno-
toxicity assay Nakajima et al. 2006). Benzophenones, and
especially BP-3, either can act directly as genotoxicants or
become genotoxicants by bioactivation via cytochrome
P450 enzymes (Takemoto et al. 2002; Zhao et al. 2013).
The types of damage to genetic material by benzophenones
include oxidative damage to DNA, formation of cyclobu-
tane pyrimidinic dimers, single-strand DNA breaks, cross-
linking of DNA to proteins, and an increase in the forma-
tion of DNA abasic sites (Cuquerella et al. 2012). Ben-
zophenones also exhibit pro-carcinogenic activities
(Kerdivel et al. 2013). BP-3 can generate reactive oxygen
species, which are potential mutagens, when applied topi-
cally to the skin followed by UV light exposure (Hanson
et al. 2006).
BP-3 is a reproductive toxicant whose mechanisms of
action and its pathological effects are poorly characterized
in various model species. In mice studies, BP-3 exposure
significantly affected fecundity, as well as inducing unex-
plained mortality in lactating mothers (Gulati and Mounce
1997). Studies in both mice and rats demonstrated that
generational exposure to BP-3 reduced body weight,
increased liver ([50 %) and kidney weights, induced a
30 % increase in prostate weight, a reduction in immuno-
competence, and significantly increased uterine weight in
juveniles (Gulati and Mounce 1997; French 1992; Sch-
lumpf et al. 2008; Rachon et al. 2006). In mammals, BP-3
is renowned for having estrogenic and anti-androgenic
activities, causing activation of estrogen receptor proteins
and inhibition of androgen receptors (Morohoshi et al.
2005; Suzuki et al. 2005; Kunz et al. 2006; Molina–Molina
et al. 2008; Nashez et al. 2010). Topical application of BP-
3 to the skin has been shown to be absorbed and transferred
to breast milk, creating risk to breast-fed neonates (Hany
and Nagel 1995). In addition, an association between
exposure to benzophenones and an increased occurrence of
endometriosis in women was recently found by Kunisue
et al. (2012).
In fish, BP-3 actions are similar to those in mammals,
causing an endocrine disruption by modulating estrogen
receptor signaling pathways, inducing reproductive
pathologies, and reducing reproductive fitness (Kunz et al.
2006; Coronado et al. 2008; Cosnefroy et al. 2011;
Bluthgen et al. 2012). Chronic exposure to BP-3 in fish
resulted in reduced egg production, induction of vitel-
logenin protein in males, and a significant reduction in egg
hatchings (Nimrod and Benson 1998; Coronado et al.
2008). These findings raise the possibility of ‘‘gender
shifts’’ in fish exposed to BP-3 during the entirety of their
Arch Environ Contam Toxicol
life history or during ‘‘windows of sensitivity’’ (Coronado
et al. 2008).
A few studies exist that have evaluated the effects of
BP-3 exposure in invertebrates. In insects, BP-3 inhibited
expression of the usp gene (ultraspiracle protein)—a pro-
tein that combines with the EcR protein to form the
ecdysone receptor, which controls aspects of develop-
mental and reproductive processes (Oza
´ez et al. 2013). Gao
et al. (2013) found that BP-3 exposure resulted in oxidative
injuries, reduced glutathione, and adversely affected cell
viability in the protozoan ciliate, Tetrahymena
Since the 1970s, coral reefs have been devastated on a
global scale. Regional weather and climate events often are
responsible for acute events of mass-mortality of coral
reefs (Carpenter et al. 2008). However, the long-term
causative processes of sustained demise often are locality
specific (Edinger et al. 1998; Rees et al. 1999; Golbuu et al.
2008; Smith et al. 2008; Downs et al. 2011,2012; Omori
2011). Records of coral recruitment in many areas of the
Caribbean, Persian Gulf, Red Sea, Hawaiian Islands, and
elsewhere have exhibited precipitous declines (Richmond
1993,1997; Hughes and Tanner 2000; Rogers and Miller
2006; Williams et al. 2008). This is most apparent in the
deterioration of juvenile coral recruitment and survival
rates along coastal areas (Dustan 1977; Miller et al. 2000;
Abelson et al. 2005; Williams et al. 2008). As with other
invertebrate species, coral larvae (i.e., planula) and newly
settled coral (i.e., recruits) are much more sensitive to the
toxicological effects of pollution compared with adults
(Kushmaro et al. 1997). Hence, even small impacts to
larval development and survival can have significant
effects on coral demographics and community structure
(Richmond 1993,1997). To manage BP-3 pollution and
mitigate its effect on the ecological resilience of coral
reefs, the toxicological effects of BP-3 on larval survival
and development need to be characterized (Fent et al. 2010;
US EPA 2012; NRC 2013).
In this study, we examined the toxicological effects of
exposures to varying concentrations of BP-3 on the larval
form (planula) of the scleractinian coral Stylophora pis-
tillata, the most abundant coral species in the northern
Gulf of Aqaba, Red Sea (Loya 1972). Many chemical
pollutants affect organisms differently when exposed to
light, a process known as chemical-associated phototox-
icity (Yu 2002; Platt et al. 2008). Because reef-building
corals are photosynthetic symbiotic organisms, and many
coral species have planulae that are photosynthetically
symbiotic (e.g., S. pistillata), we examined the effects of
BP-3 exposure in planulae subjected to either darkness or
to environmentally-relevant light conditions. Histopathol-
ogy and cellular pathology, planula morphology, coral
bleaching, DNA damage as the formation of DNA abasic
lesions, and planula mortality were measured in response to
BP-3 exposure. Median lethal concentration (LC
), effect
concentration (EC
), and no observable effect concentra-
tions (NOEC) were determined for coral planulae exposed
to BP-3 in both darkness and in light. Coral planulae are a
relatively difficult resource to procure for toxicological
studies. Therefore, primary coral cell cultures were used in
in vitro toxicological tests of BP-3 to examine their validity
as a surrogate model for coral planulae in generating an
effect characterization as part of an Ecological Risk
Assessment. The confidence in this model was examined
by comparisons of the LC
results of BP-3-exposed
planulae to the BP-3 LC
of coral cells (calicoblasts) from
adult S. pistillata colonies. Coral-cell toxicity testing was
conducted on six other species that originate from either
the Indo-Pacific or Caribbean Sea/Atlantic Ocean basins to
provide in vitro data on the species’ sensitivity distribution
of BP-3. To determine the environmentally relevant con-
centration of BP-3 in seawater on coral reefs, we measured
BP-3 concentrations at various locations in the U.S. Virgin
Islands and the U.S. Hawaiian Islands.
Materials and Methods
Planula Collection and Toxicity Exposures
Planula collection and planula-toxicity exposures were
conducted at the Inter-University Institute of Marine Sci-
ences (IUI) in Eilat, Israel. Stylophora pistillata (Esper
1797) planulae were collected from the wild within the IUI
designated research area by placing positively buoyant
planula traps over Stylophora colonies measuring more
than 25 cm in diameter. Permit for collection was given to
Y. Loya by the Israel National Park Authority. Traps were
set between 17:00 and 18:00 h, and then retrieved at
06:00 h the next morning. Planulae were inspected and
sorted by 07:15 h, and toxicity exposure experiments
began at 08:00 h.
Experimental design and culture conditions were based
on modified (for coral) guidelines set forth in OECD
(2013) and described in Downs et al. (2014). This experi-
ment for BP-3 was conducted concurrently with the study
conducted in Downs et al. (2014).
All seawater (ASW) was made artificially using Fisher
Scientific Environmental-Grade water (cat#W11-4) and
Sigma-Aldrich sea salts (cat#S9883) to a salinity of 38
parts per thousand at 22 C. Benzophenone-3 (BP-3;
2-Hydroxy-4-methoxyphenyl-phenylmethanone; Aldrich
cat#T16403) was solubilized in dimethyl sulfoxide
(DMSO) and then diluted with ASW to generate stock
solutions and exposure solutions. Solutions of BP-3 for
toxicity exposures each contained 5 microliters of DMSO
Arch Environ Contam Toxicol
per one liter and were of the following concentrations:
1 mM BP-3 (228 parts per million), 0.1 mM BP-3
(22.8 mg/L; parts per million), 0.01 mM BP-3 (2.28 mg/L;
parts per million), 0.001 mM BP-3 (228 lg/L; parts per
billion), 0.0001 mM (22.8 lg/L; parts per billion), and
0.00001 mM (2.28 lg/L; parts per billion). For every
exposure time-period, there were two control treatments
with four replicates each: (a) planulae in ASW, and
(b) planulae in ASW with 5 microliters of DMSO per 1 L.
There was no statistical difference between the two con-
trols for any of the assays.
Planulae were exposed to different BP-3 concentrations
during four different time-period scenarios: (a) 8 h in the
light, (b) 8 h in the dark, (c) a full diurnal cycle of 24 h,
beginning at 08:00 in daylight and darkness from 18:00 in
the evening until 08:00 h the next day, and (d) a full 24 h
in darkness. For the 24-h exposure, planulae from all
treatments were transferred to new 24-well microplates
with fresh ASW/BP-3 media at the end of the 8-h daylight
exposure before the beginning of the 16 h dark exposure.
At the end of the 8 and 24-h time points, chlorophyll
fluorescence, morphology, planula ciliary movement, and
mortality were measured, while at least one planula from
each replicate of each treatment was chemically preserved,
and the remaining living planulae were flash frozen in
liquid nitrogen for the DNA apyrimidinic (AP) site assay.
Chlorophyll Fluorescence as an Estimate
of Bleaching
Chlorophyll fluorescence was measured using a Molecular
Dynamics microplate fluorometer with an excitation
wavelength of 445 nm and an emission wavelength of
685 nm. Fluorescence measurements were taken at the end
of the 8-h light and dark periods of BP-3 exposure. All ten
planulae in each replicate well were measured in aggregate.
Each well was measured independently of the other wells.
Justification and caveats for this assay are described in
Downs et al. (2014).
DNA Abasic Lesions
DNA abasic or apurinic/apyrimidinic lesions (DNA AP
sites) were quantified using the Dojindo DNA Damage
Quantification Kit-AP Site Counting (DK-02-10; Dojindo
Molecular Technologies, Inc.) and conducted as described
in Downs et al. (2014). Four individual planulae (one from
each well) from each treatment were individually assayed.
Only planulae that were relatively intact were assayed,
even if scored as dead. Planulae from 228 ppm BP-3 at 8 h
in the light were not collected, because there were no
coherent planulae.
Transmission Electron Microscopy
Transmission electron microscopy was used for tissue and
cellular pathomorphology assessment on three planulae
from each treatment. Methodology for this technique was
described in Downs et al. (2014). At least three planula
from each treatment were collected and fixed for analysis.
Coral Cell Toxicity Assay
Cultured colonies of S. pistillata (Esper 1797) were
obtained from Exotic Reef Imports (www.exoticreefim and did not need a permit for possession. Cul-
tured colonies of Pocillopora damicornis (Linnaeus 1758)
was provided by the National Aquarium and did not need a
permit for possession. Montastrea annularis,Montastrea
cavernosa (Linnaeus 1766), and Porites astreoides (La-
marck 1816) were obtained from the Florida Keys National
Marine Sanctuary under permit# FKNMS-2011-139. Cul-
tured colonies of Acropora cervicornis (Lamarck 1816)
and Porites divaricata (Lesueur 1821) were provided by
Dr. Cheryl Woodley of the U.S. National Oceanic and
Atmospheric Administration and did not need a permit for
possession. Corals were maintained in glass and Teflon-
plumbed aquaria in 36 ppt salinity artificial seawater (Type
1 water using a Barnstead E-Pure filter system that inclu-
ded activated carbon filters) at a temperature of 24 C.
Corals were grown under custom LED lighting with a peak
radiance of 288 photosynthetic photon flux density lmol/
/s. Light Spectra ranged from 380 to 740 nm. Light was
measured using a Licor 250A light meter and planar inci-
dence sensor. Description of coral cell isolation from each
species is described in Downs et al. (2010,2014).
Exposure experiments of cells were conducted in PTFE-
Telfon microplates. Cells of all species except Acropora
cervicornis were exposed to BP-3 concentrations in cell
culture media of 570 parts per trillion to 228 parts per
million for 4 h in the light, whereas Stylophora cells also
were exposed for 4 h in the dark. Acropora cervicornis
cells were exposed to BP-3 concentrations in cell culture
media of 570 ng/L (parts per trillion) to 228 mg/L (parts
per million) for 4 h in the light. Lighting was from custom
LED fixtures that had wavelength emissions from 390 to
720 nm with a light intensity of 295 lmol/m
/s of photon
flux density.
Viability was confirmed using the trypan blue exclusion
assay. There were four replicate wells with cells per
treatment. Duplicate aliquots of cells from each replicated
wells were collected into a microcentrifuge tube, cen-
trifuged at 3009gfor 5 min, and the supernatant aspirated.
Cells were gently resuspended in culture media that con-
tained 0.5–1.5 % (w/v) of filtered trypan blue (Sigma-
Arch Environ Contam Toxicol
Aldrich, cat#T6146), and incubated for 5 min. Viable
versus dead cells were counted using a modified Neubauer
hemocytometer (Hausser-Levy Counting Chamber).
Sampling and analysis of benzophenones in seawater
samples via gas chromatography-mass spectrometry (GC–
MS) and liquid chromatography-mass spectrometry (LC–
MS). Dichloromethane, methanol, acetone are pesticide-
grade solvents (Fisher Scientific). Analytical standards were
purchased from Sigma Aldrich and included: Benzophenone
(cat# B9300), Benzhydrol (cat#B4856), 4-hydroxyben-
zophenone (cat#H20202), 2-hydroxy-4-methoxy benzophe-
none (cat#H36206), 2,4-dihydroxy benzophenone (cat#
126217), 2-20-dihydroxy-4-methoxy benzophenone (cat#
323578), 2,3,4-trihydroxy benzophenone (cat# 260576),
2,20,4,40-tetrahydroxy benzophenone (cat#T16403). Internal
standard solutions (phenanthrene-d10 and chrysene-d12)
were purchased from AccuStandard Inc. (New Haven, CT).
Field personnel collecting samples were subject to an
Alconox Liqui-Nox detergent decontamination immedi-
ately before entering the sampling site and did not apply
any sunscreen lotion or nonorganic personal-care products
to their body for at least 21 days before sampling. Between
100 and 500 mL of seawater were collected approximately
35 cm below the surface of the water into EPA-certified
clean, amber jars. In the field, water samples were
extracted using Phenomenex C18 solid phase extractions
columns that were first activated with methanol. All col-
umns were capped and then shipped and stored frozen at
-80 C or colder.
Extraction of analytes from seawater samples collected
in the U.S. Virgin Islands (under a U.S. National Park
Service permit, STT-045-08) followed the methodology
described in Jeon et al. (2006). Seawater samples were
collected using precleaned 1-L amber glass bottles with
Teflon lined lids (I-Chem, 300 series, VWR). Seawater
samples were extracted using C-18E cartridges (500 mg,
6 mL Phenomenex Inc.) on a vacuum manifold (Phenom-
enex Inc.). Cartridges were conditioned with 5 mL of
methanol and then 5 mL of water, after which the seawater
samples were then added to the column. Following
extraction, the cartridges were dried for 10 min, capped,
and frozen until processed. The cartridges were eluted with
2 mL of acetone followed by 2 95 mL dichloromethane.
The extracts were evaporated to dryness under a gentle
stream of nitrogen. Then, 50 lL of MSTFA (N-Methyl-N-
(trimethylsilyl) trifluoroacetamide, Sigma-Aldrich) was
added, capped, vortexed for 30 s, and heated at 80 C for
30 min. Extracts were transferred to gas chromatography
vials with a rinse step to a final volume of 1 mL and the
internal standard was added. Percentage recovery for all 8
target analytes using this method with seawater was
[95 %.
Seawater samples from Hawaii were collected using
precleaned one liter amber glass bottles with Teflon lined lids
(I-Chem, 300 series, VWR). Samples were extracted using
C-18E cartridges (500 mg, 6 mL Phenomenex Inc.) on a
vacuum manifold (Phenomenex Inc.). Cartridges were con-
ditioned as indicated in the previous paragraph and eluted
with 5 mL of methanol. For LC–MS analysis, samples were
run on an AB_SCIEX 5500 QTRAP Triple Quadrupole
Hybrid Linear Ion Trap Mass Spectrometer with a Spark
Holland Symbiosis HPLC for analytical separation. The
analytes were measured with MRM (multiple reaction
monitoring) followed by switching to ion trap functionality
(Q3- LIT) to confirm the fragmentation pattern of the MRMs.
The source was set at 700 C and the gasses were set to 60
arbitrary units of nitrogen. The curtain gas was set at 45
arbitrary units, and all MRMs were optimized using infusion
based introduction of analytical standards. Analytical sepa-
ration was performed using a Phenomenex Hydro RP
4.6 950 2.6 lm particle size stationary phase, with the
mobile phase composed of methanol and water with the
addition of 0.1 % formic acid and 5 mM of ammonium
acetate in both phases. The flow rate was set at 0.9 mL per
min, and a ballistic gradient and re-equilibration was run
over 5 min. Percentage recovery for target analytes was
[85 %, Limit of Detection was 100 pptrillion, and Quanti-
tative Limit of Measurement was 5 ppbillion (lg/L).
Statistical Methods
OECD (2006) was used as a guidance document for our
approach in the statistical analysis of the data. To address
different philosophies and regulatory criteria, Effect Con-
centration response (EC
and EC
) and median Lethal
Concentration response (LC
) were determined using
three initial methods: PROBIT analysis (Finney 1947),
linear or quadratic regression (Draper and Smith 1966), and
spline fitting (Scholze et al. 2001). Data were analyzed
using linear or quadratic regression and PROBIT methods
individually for each experiment, based on model residuals
being random, normally distributed, and independent of
dosing concentrations (Crawley 1993, Fig. 5.1), as well as
having good fit, statistically significant, and biologically
interpretable regressors (Agresti 2002; Newman 2013).
Spline fitting did not meet these criteria. In several analy-
ses, BP-3 concentrations as log
(x?1) were transformed
to conform to model assumptions.
Data were tested for normality (Shapiro–Wilk test) and
equal variance. When data did not meet the assumption of
normality and homogeneity, the no-observed-effect con-
centration (NOEC) was determined using Kruskal–Wallis
one-way analysis of variance, using Dunnett’s Procedure
(Zar 1996) to identify concentrations whose means differed
Arch Environ Contam Toxicol
significantly from the control (Newman 2013). When
variances among treatments were heterogeneous, we veri-
fied these results using a Welch ANOVA. In cases where
responses were homogeneous within the control treatment
(i.e., all planulae survived) or another concentration (i.e.,
all planulae died or were deformed), the Steel Method
(Steel 1959) was substituted, which is the nonparametric
counterpart to Dunnett’s Procedure (Newman 2013). Four
replicates of each experimental concentration provided
good statistical power for parametric analyses, but it is
cautioned that the relatively small sample size for the
nonparametric Steel Method (Steel 1959) made results of
this test less powerful. To facilitate comparisons among
other treatment means, figure legends include results of
Newman–Keuls Method post hoc test, which compares
each concentration to all others.
Parametric (Pearson’s r) or nonparametric (Spearman’s q)
regression analyses were used to determine the relationship
between mortality of coral planulae and coral cells. Coral
planulae are available only immediately after spawning and a
strong association between these two responses would allow
mortality of coral cells to serve as a surrogate for this repro-
ductive response. JMP version 9.0 or 10.0 (SAS Institute, Inc.,
Cary, NC), SAS version 9.3 and SigmaPlot 12.5 (Systat
Software, Inc., San Jose, CA) were used for analyses.
Planulae under control conditions have an elongated,
‘cucumber-like’’ morphology with organized rows of
zooxanthellae-containing gastrodermal cells running from
the aboral pole to the oral pole (Fig. 1a; ‘‘brown dots’’ in
the rows are individual zooxanthella cells). Normal plan-
ulae are in near-constant motion, being propelled by cilia
that cover the elongated body. Within the first 4 h of
exposure of planulae to BP-3 in both light and darkness,
planulae showed a significant reduction in ciliary move-
ment and the morphology had significantly changed from
the elongated form to a deformed ‘‘dewdrop’’ (Fig. 1b). At
228 lg/L BP-3, planulae contain noticeably less zooxan-
thellae (brown spots) indicative of ‘‘bleaching’’ (Fig. 1c).
The mouth of the planula at the oral pole began to increase
three- to fivefold in diameter at the end of the 8-h exposure
Fig. 1 Stylophora pistillata planulae exposed to various treatments
of benzophenone-3 (BP-3). aControl planula exposed for 8 h in light.
bPlanula exposed to 22.8 parts per billion (lg/L) BP-3 for 8 h in the
light. cPlanula exposed to 228 parts per billion (lg/L) BP-3 for 8 h in
the light. dPlanula exposed to 2.28 parts per million (mg/L) BP3 for
8 h in the light. ePlanula exposed to 28.8 parts per million (mg/L)
BP3 for 8 h in the light. Scale bar is 0.5 mm
Arch Environ Contam Toxicol
(Fig. 1d). By the end of the 8 h of exposure for all BP-3
concentrations, the oral pole was recessed into the body in
deformed planulae (Fig. 1b) and the epidermis of all the
deformed planulae took on a white opaque hue. For plan-
ulae exposed to the higher concentrations of BP-3, it was
apparent that the epidermal layer had lost its typical
transparency and become opaque (Fig. 1, bracket indicates
opaqueness of epidermal layer).
At the end of the 8-h exposure, all planulae exposed to
all of the concentrations of BP-3 became sessile. Addi-
tionally, there was a positive relationship between exposure
to increasing concentrations of BP-3 and planula bleaching
(Figs. 1a–e, 2). Bleaching is the loss of symbiotic
dinoflagellate zooxanthellae, photosynthetic pigments, or
both. Chlorophyll fluorescence as an indicator of the con-
centration of chlorophyll apigment corroborated these
visual observations; exposure to BP-3, whether in light or
darkness, caused planulae to bleach (Fig. 2). The Lowest
Observable Effect Concentration for inducing chlorophyll-
defined bleaching is 2.28 lg/L in the light (P\0.001,
Dunnett’s Method) and 22.8 lg/L in the dark (P\0.01,
Dunnett’s Method).
Normal planulae have four layers of organization. At the
surface of the planula is the epidermis (Fig. 3a–c). The
outer aspect of the epidermis has densely packed ciliated
cells (Fig. 3a), spirocysts and nematocysts/blasts (Fig. 3b),
and cells containing chromogenic organelles. Between the
epidermis and the gastrodermal tissue layers is the meso-
glea (Fig. 3c–d). Within the gastrodermal tissue are cells
that contain symbiotic dinoflagellate zooxanthellae within
an intracellular vacuole (Fig. 3e). Figure 3e depicts a
healthy morphology, with the presence of starch granules,
coherent chloroplasts, and the presence of a pyrenoid body
that interfaces with chloroplasts. Figure 3f illustrates the
integrity of chloroplasts (cp) within the dinoflagellate,
especially the structure of the tri-partite rows of the
Fig. 2 Relative chlorophyll
fluorescence emission at
685 nm with excitation at
445 nm of planulae of
Stylophora pistillata exposed to
various treatments of
benzophenone-3 (BP-3). Bars
show treatment means with
whiskers representing ±1
standard error of the mean.
N=4 replicates per treatment.
aPlanulae exposed to various
BP-3 concentrations for 8 h in
the light. Treatment means with
different letters differed
significantly from the control at
a=0.05, based on Kruskal–
Wallis one-way analysis of
variance on ranks followed by a
Dunnett’s Method post hoc test
against a control. bPlanulae
exposed to various BP-3
concentrations for 8 h in the
dark. Treatment means with
different superscript letters
differed significantly from the
control at a=0.05, based on
one-way analysis of variance
followed by a Dunnett’s Method
post hoc test against control
Arch Environ Contam Toxicol
thylakoid (t) membranes. Dinoflagellates from control
planulae contained an abundance of starch granules (S), as
well as the absence of vacuolated space between the
dinoflagellate’s thecal plate and the host’s symbiophagic
membrane (indicated by ‘‘{’’; Fig. 3f).
Transmission electron microscopy of planula exposed to
288 parts per billion BP-3 for 8 h in the light (Fig. 4)
showed that the planulae experienced catastrophic tissue
lysis and cellular degradation in both the epidermis and
gastrodermis, as well as partial collapse of the mesoglea
(Figs. 3vs. 4). At the surface of the epidermis, there was a
complete loss of ciliated cells (Fig. 4a). The development
and extent of cell death and tissue deterioration was
greatest at the surface of the epidermis and became less
pronounced at the center of the planula. In the middle area
of the epidermal tissue, between the outer surface of the
epidermis and its boundary with the mesoglea, the
incidence of autophagic cell death became more pro-
nounced (Fig. 4b; Tsujimoto and Shimizy 2005; Samara
et al. 2008). Individual cells were dense with autophagic
bodies, and many of the nuclei exhibited delamination of
the nuclear bilayer membrane and vacuolization of the
inner nuclear membrane containing chromatin (Fig. 4c;
‘}’’ indicates vacuolization; Eskelinin et al. 2011). None of
the nuclei observed in the micrographs exhibited any signs
of apoptosis, such as condensation of chromatin (Kerr et al.
1972; White and Cinti 2004; Taatjes et al. 2008). Spe-
cialized cells, such as spirocysts, also exhibited deteriora-
tion (Fig. 4d). The mesoglea exhibited structural
deterioration; this vascular space contained an abundance
of debris, including detached cells (Fig. 4e). The gastro-
dermis also exhibited extensive trauma (Fig. 4e–g). Many
gastrodermal cells exhibited considerable dense autophagic
bodies (Fig. 4f), although there were a few instances of
Fig. 3 Transmission electron microscopy of Stylophora pistillata
planula control treatment. aEpidermal surface, indicating the
presence of functional cilia (c) and tightly adjoined epidermal cells;
bar indicates 2000 nm. bEpidermal surface indicates intact nema-
tocysts (n) and nuclei (nuc); bar indicates 5000 nm. cMesoglea
(m) demarks the epidermal tissue (epi) from the gastrodermal tissue
(gd); bar indicates 5000 nm. dMicrograph indicates the interface of
the gastroderm (g), mesoglea (m), and epidermis (epi); bar indicates
5000 nm. eZooxanthella in the gastrodermal tissue of planula,
indicating the presence of intact chloroplasts (cp) and pyrenoid body
(p). Notice the absence of a vacuolar space between the coral vacuolar
membrane and the thecal plates/membrane of the zooxanthella; bar
indicates 2000 nm. fClose-up of cytosolic structure of zooxanthella.
Chloroplasts (cp) exhibit intact chloroplastic membrane and coherent,
parallel rows of thylakoid membranes. Bracket (]) indicates the
absence of vacuolar space between the coral vacuolar membrane and
the zooxanthella’s thecal plate/membrane; bar indicates 500 nm
Arch Environ Contam Toxicol
nuclear autophagy. Gastrodermal cells containing symbi-
otic zooxanthella exhibited the early stages of symbio-
phagy, with vacuolization occurring around the
zooxanthella (Fig. 4e–g). None of the zooxanthellae
showed ‘‘normal’’ morphologies. They instead displayed
extensive internal vacuolization, homogenization of chro-
matin density, and chloroplast degradation, especially of
the thylakoid membranes (Fig. 4g–h).
Transmission electron microscopy of planulae exposed
to 228 lg/L BP-3 for 8 h in darkness (Fig. 5) exhibited a
similar gradient of cell death and tissue deterioration from
the surface of the planula to its center as seen in planulae
exposed to BP-3 in the light, although the progression of
cellular deterioration was not as severe (Fig. 5a–h). Along
the surface of the epidermal tissue layer, ciliated cells were
undergoing cellular degradation (Fig. 5a). The cell layer
immediately below the ciliated cells was degraded, char-
acterized by an abundance of vacuolated bodies and loss of
the plasma membrane (Fig. 5b, c). Many of the nuclei
exhibited partial delamination of the bilayer nuclear
membrane, but unlike the nuclei observed in planulae
exposed to BP-3 in the light, vacuolization was not com-
plete and the bilayer was still partially anchored by nuclear
pores (Fig. 5b, c). Deeper into the epidermal layer, along
the boundary with the mesoglea, cellular degradation per-
sisted, especially of the spirocysts (Fig. 5d). There is an
extracellular matrix that acts as a barrier between the epi-
dermal tissue and mesoglea, and again between the gas-
trodermal tissue and mesoglea. Under these conditions, the
integrity of the boundary layer between the epidermis and
mesoglea had severely deteriorated, whereas the boundary
layer between the gastrodermis and mesoglea remained
intact (Fig. 5e). Within the gastrodermis, a vast majority of
the cells were alive, but exhibiting signs of massive
autophagy (Fig. 5f; Klionsky et al. 2012). It should be
noted that there were almost no instances of delamination
of the nuclear membrane in the gastrodermal cells; nuclei
looked healthy (Fig. 5f). Many of the cells were dense with
autophagosomic bodies, and most of the zooxanthellae
were undergoing symbiophagy, as indicated by the vac-
uolization around the dinoflagellate cell (Fig. 5f; Downs
et al. 2009). In zooxanthellae that were not significantly
degraded (Fig. 5f vs. h), thylakoids exhibited a pathomor-
phology similar to that found in zooxanthellae of corals
exposed to heat stress (32 C) in darkness; thylakoid
lamellae were diffuse (Fig. 5g; Downs et al. 2013), sug-
gesting that the zooxanthellae were directly affected by the
BP-3 exposure. In contrast to the findings of Danovaro
Fig. 4 Transmission electron microscopy of Stylophora pistillata
planula exposed to 228 parts per billion (lg/L) benzophenone-3 for
8 h in the light. aSurface of the epidermal layer; indicating a lack of
cilia and cells dying either via necrosis or autophagic cell death; bar
indicates 5000 nm. bEpidermal tissue where cells exhibit an
abundance of vacuolated bodies, especially the presence of vacuo-
lated nuclei (nuc); bar indicates 5000 nm. cMagnification of
vacuolated nuclei (nuc) that completely lacks nuclear blebbing (a
sign of apoptosis). ‘‘}’’ indicates vacuolization of delaminated nuclear
double membrane; bar indicates 1000 nm. dEpidermal layer with
vacuolated ciliated cells, spirocysts (sp) and nematocysts; bar
indicates 5000 nm. eMicrograph depicts intersection of mesoglea
(m) and gastrodermal tissue containing both zooxanthella (zx)
gastrodermal cells and yolk (y); bar indicates 5000 nm. fEpidermal
tissue adjacent to yolk exhibits extensive autophagic vacuolization;
bar indicates 5000 nm. GGastrodermal cells containing symbio-
phagic zooxanthellae. Zooxanthellae have undergone extensive
internal vacuolization; bar indicates 5000 nm. hIncreased magnifi-
cation focused on vacuolated zooxanthella, (v) indicates symbio-
phagic vacuole; bar indicates 2000 nm
Arch Environ Contam Toxicol
et al. (2008), viral inclusion bodies were not observed in
our electron microscopy examination.
During the initial examination of the planulae using
transmission electron microscopy, scratches in the micro-
sections under observation were readily apparent (Figs. 5a–c
and 6). Scratches to the microsection can arise as a result of
hardened particles from the sample that scrape between the
diamond blade and micro-sectioned sample (Carson 1997;
Crang and Klomparens 1988). This is a common occurrence
in biological samples that contain CaCO
skeleton (coral or
vertebrates). These scratches are preventable if the samples
are first decalcified before embedding in a resin and sec-
tioned (Crang and Klomparens 1988). Coral planula samples
do not normally need to be decalcified, because they should
contain no aragonite skeletal matrix. An Alizarin red stain
confirmed the presence of a CaCO
crystal matrix on the
surface of the planula (data not shown; Barnes 1972).
Decalcifying the fixed coral planulae with EDTA before
embedding the sample in resin alleviated the ‘‘scratch’
artifact and the remaining samples that were processed using
a decalcification step were devoid of scratches.
Increasing concentrations of BP-3 induced significantly
higher levels of DNA AP lesions in planulae exposed to the
light compared to the controls (Fig. 7a, b), as well as
planulae exposed to BP-3 in the dark (Fig. 7c, d).
No-Observed-Effect Concentration
Estimating Lowest-observed-effect Concentration
(NOECs) for planulae exposed to BP-3 for 8 h was prob-
lematic because responses in the control treatment were
homogeneous (Shapiro–Wilk; P\0.05); all planulae sur-
vived and were not deformed, so analyses defaulted to the
less powerful, nonparametric method (Steel 1959). The
NOEC for both the proportion of live coral planulae and
nondeformed planulae exposed to BP-3 for 8 h in either the
light or the dark was 228 ppmillion (mg/L) (Steel Method
(Steel 1959), all Z [2.32, P\0.0809; Fig.8a, c). In
contrast to the Steel Method, the NOEC for planulae in the
light determined by a Kruskal–Wallis One-Way Analysis
of Variance on Ranks was 228 lg/L (H Statistic =21.903;
PB0.001; Dunnett’s Procedure). The NOEC for planulae
Fig. 5 Transmission electron microscopy of Stylophora pistillata
planula exposed to 228 parts per billion (lg/L) benzophenone-3 for
8 h in the dark. aSurface of the epidermal layer; ciliated cells are
present, but undergoing early stages of autophagic cell death. Cells
beneath the cilia layer exhibiting late stage autophagic cell death and
necrosis. Note scratches in the micrograph; bar indicates 2000 nm.
bEpidermal tissue area between cilia and nematocyst layer showing
extensive vacuolization. Early stages of nuclear vacuolization (nuc).
Note scratches in the micrograph; bar indicates 2000 nm. cEpidermal
tissue in area exhibiting advanced stages of cell death; nucleus
vacuolization (nuc). Note scratches in the micrograph; bar indicates
2000 nm. dExtensive vacuolization of cells surrounding
nematocysts. Note scratches in the micrograph; bar indicates
5000 nm. eMesoglea (m), gastrodermal and epidermal tissues.
Symbiophagy occurring to zooxanthella (zx) surrounded by extensive
vacuolization in neighboring cells; bar indicates 2000 nm. fGastro-
dermal tissue and yolk (y). All cells exhibiting extensive vacuoliza-
tion (v), especially within the gastrodermal cell surrounding the
zooxanthella. Coral cells showing increased level of autophagosome
content but no signs of autophagic cell death or necrosis; bar indicates
5000 nm. gZooxanthella chloroplast with thylakoid dispersion-
pathomorphologies. Chloroplast (cp); bar indicates 1000 nm. hZoox-
anthella exhibiting extensive pyknosis; symbiophagic vacuole (v); bar
indicates 1000 nm
Arch Environ Contam Toxicol
in the dark determined by a Kruskal–Wallis One-Way
Analysis of Variance on Ranks was 228 lg/L (H Statistic =
22.402; PB0.001; Dunnett’s Procedure).
Estimates for NOECs for planulae exposed to BP-3 for
24 h in light or darkness also were problematic because
responses in the control and at all concentrations greater
than 22.8 lg/L (in certain cases, C2.28 lg/L) were
homogeneous (Fig. 8b, d); all planulae survived and were
not deformed in the control but died at the higher con-
centrations (Laskowski 1995). Using the nonparametric
Steel Method, we determined the NOEC as 2.28 lg/L for
the proportion of coral planulae alive after 24 h of expo-
sure to BP-3 in the light and 22.8 lg/L in the dark (both
Z=2.48, P=0.0543). The corresponding NOECs for
non-deformed planulae were identical to these values
(Fig. 9a, c). In contrast, the NOEC for planulae exposed for
24 h in the light, determined by a Kruskal–Wallis One-
Way Analysis of Variance on Ranks, was 228 lg/L
(Fig. 9b; H Statistic =22.084; PB0.001; Dunnett’s Pro-
cedure). The NOEC for planulae exposed for 24 h in
darkness, determined by a Kruskal–Wallis One-Way
Analysis of Variance on Ranks, was 228 lg/L (Fig. 9d;
HStatistic =22.112; PB0.001; Dunnett’s Method).
The NOEC for DNA abasic sites in planulae met
ANOVA assumptions and was determined as 22.8 lg/L
(100 nM; one-way ANOVA F
=73.1, P\0.0001,
=0.95; Dunnett’s Method for this comparison,
P\0.0001) when exposed in the light, and 22.8 lg/L
(100 nM) when exposed in the dark (Welch ANOVA
=142.1, P\0.0001; Dunnett’s Method for this
comparison, P\0.0001). The NOEC for mortality of S.
pistillata calicoblast cells was below the 570 ng/L concen-
tration for cells exposed to the dark for 4 h (Fig. 10a, b). The
NOEC for mortality of S. pistillata calicoblast cells was
570 ng/L for cells exposed to the light for 4 h (Fig. 10c, d).
, and EC
Regression models used to estimate median LC
centration expected to cause death in 50 % of the popula-
tion), EC
and median EC
(effective concentrations,
which adversely affect 20 and 50 % of the population,
respectively) after 8 h of exposure to BP-3 had coefficients
of determination (R
) between (0.91 and 0.97). Using
regression models, the median LC
for the proportion of
live coral planulae exposed in the light was 3.1 mg/L,
whereas for planulae exposed in the dark, the LC
was 5.4
times higher: 16.8 mg/L (Table 1; Supplemental Fig. 1a,
c). PROBIT analysis for LC
in the light was 2.876 mg/L
(mg/L), whereas LC
in the dark was 12.811 mg/L
(Table 1; Supplemental Fig. 2a, c).
Models used to estimate LC
and EC
, of coral plan-
ulae after 24 h of exposure to BP-3 continued to explain
the substantial variation (0.86 \R
B0.997). The 24 h-
for the proportion of live coral planulae, after expo-
sure in the light, was just 103.8 lg/L (ppbillion) compared
with 873.4 lg/L in the dark exposure (Table 1; Supple-
mental Fig. 1b, d). PROBIT analysis for 24-h LC
in the
light was 139 lg/L, whereas LC
in the dark was 799 lg/L
(Table 1; Supplemental Fig. 2b, d).
Fig. 6 ‘‘Scratch’’ artifacts in transmission electron microscopy
micrographs of Stylophora pistillata planula exposed to 288 parts
per billion (lg/L) benzophenone-3. When microsectioning planula
embedded in a plastic resin without first decalcifying the sample,
scratches can manifest on the mounted ultrathin sections. The
scratches form as a result of the diamond blade fracturing the
aragonite skeleton and pieces of the skeleton adhering to the edge of
the diamond blade. As the contaminated blade cuts through the
sample block, it scratches the ultrathin sections of the sample. These
scratches can be alleviated by cleaning the diamond blade and
removing aragonite skeleton in the sample through decalcification
before embedding the sample in a resin. aScratches apparent in
ultrathin section of epidermal section of a planula; bar indicates
2000 nm. bScratches apparent in ultrathin section of gastrodermal
section of a planula; bar indicated 5000 nm
Arch Environ Contam Toxicol
The 8-h EC
for nondeformed planulae exposed to BP-
3 in the light and dark were much lower: 107 and 436 lg/
L, respectively using regression modeling (Table 1; Sup-
plemental Fig. 3a, c). PROBIT analysis for 8-h EC
in the
light was 133 ppbillion (lg/L), whereas EC
in the dark
was 737 lg/L (Table 1; Supplemental Fig. 4a, c). PROBIT
analysis for 8-h EC
in the light was 6.3 lg/L, whereas
in the dark was 15.5 lg/L (Table 1; Supplemental
Fig. 4a, c). The 24-h EC
for nondeformed planulae
exposed in the light and dark were much lower: 17 ppbil-
lion and 105 lg/L, respectively using regression modeling
(Table 1; Supplemental Fig. 3b, d). PROBIT analysis for
24-h EC
in the light was 49 lg/L, whereas LC
in the
dark was 137 lg/L (Table 1; Supplemental Fig. 4a, d).
PROBIT analysis for 24-h EC
in the light was 6.5 lg/L,
whereas EC
in the dark was 10.4 lg/L (Table 1; Sup-
plemental Fig. 4b, d).
The number of DNA abasic sites increased approxi-
mately tenfold across the BP-3 concentration gradient in
the light, but nearly 20-fold in the dark (Fig. 7b, d). Sim-
ilarly, the percentage of dead coral cells increased dra-
matically with increasing concentrations of BP-3, but the
was much lower in the light at 39 lg/L than in the
dark at 842 lg/L. PROBIT analysis for 4-h LC
coral cells
in the light was 42 ppbillion, whereas LC
in the dark it
was 679 lg/L (Table 2; Supplemental Fig. 5a, b).
Species Sensitivity Distribution Using Coral Cell
Toxicity Assay
To provide a perspective of the differences in sensitivities
of various species of Indo-Pacific and Caribbean coral
reefs, the LC
s and LC
s with their corresponding upper
Fig. 7 Number of DNA apyrimidinic lesions in planulae of Sty-
lophora pistillata exposed to various concentrations of benzophe-
none-3 (BP-3). Bars show treatment means of four replicates with
whiskers representing ±1 standard error of the mean. Treatment
means with different letters differed significantly at a=0.05, based
on Kruskal–Wallis one-way analysis of variance on ranks followed by
a Student–Newman–Keuls Method post hoc test. aPlanulae exposed
for 8 h in the light. bLog-linear regression between DNA AP lesions
of coral planulae of Stylophora pistillata exposed to concentrations of
BP-3 for 8 h in the light. Quadratic regression line (solid) and 95 %
confidence intervals (dashed lines) are shown. cPlanulae exposed for
8 h in the dark. dLog-linear regression between DNA AP lesions of
coral planulae of Stylophora pistillata exposed to concentrations of
BP-3 for 8 h in the dark
Arch Environ Contam Toxicol
and lower 95 % confidence intervals for the two Indo-
Pacific and five Caribbean species are provided in Table 1.
Correction Factor Between Mortality of Coral
Planulae and Coral Cells
Coral cells were much more sensitive than coral planulae
across a wide range of BP-3 concentrations, which makes
cell mortality a potential indicator of reproductive and
recruitment failures. To estimate the correction factor
needed to translate coral cell mortality into potential
mortality of coral planulae, one option is the use of a
quadratic regression model to estimate these relationships:
In the light (F
=43.8, P\0.0001, R
=0.81) %
mortality of planulae =2.26 -0.28 (% mortality of
cells) ?0.0107 (% mortality of cells)
In the dark
=84.5, P\0.0001, R
=0.89) % mortality of
planulae =0.86 -0.0007 (% mortality of cells) ?0.0078
(% mortality of cells)
Environmental Chemistry Analysis
The purpose of the chemical analysis was to conduct a
cursory survey of BP-3 concentrations on coral reefs.
Seawater samples were collected from bays in St. John
Island, U.S. Virgin Islands: Caneel Bay, Hawksnest Bay,
and Trunk Bay in April 2007 (Fig. 11a, b). Caneel Beach is
managed by the resort, Caneel Bay. Samples were col-
lected at approximately 16:30 h near the dive platform that
adjoins the Caneel Beach and along a large coral com-
munity that spans from the edge of Caneel Beach to the
edge of Honeymoon Beach. There were 17 swimmers in
Caneel Bay in the 48-h period before sampling. Swimmers
were monitored from the shore of the resort from dawn to
dusk. No benzophenones could be detected in either of the
samples collected in Caneel Bay.
Hawksnest Bay is a densely visited beach within the
U.S. National Park system on St. John Island. In general,
more than 1000 visitors per day can enter into this bay. On
the day of sampling, more than 230 people entered the
Fig. 8 Percent mortality of planula of Stylophora pistillata exposed
to various concentrations of benzophenone-3. Bars show treatment
means with whiskers representing ±1 standard error of the mean.
Treatment means with different letters differed significantly at
a=0.05, based on Kruskal–Wallis one-way analysis of variance on
ranks followed by a Student–Newman–Keuls Method post hoc test.
aPlanulae exposed for 8 h in the light. bPlanulae exposed for 8 h in
the light and then 16 h of darkness. cPlanulae exposed for 8 h in the
dark. dPlanulae exposed for 24 h in the dark
Arch Environ Contam Toxicol
water and swam within 20 m of the three large Acropora
palmata spurs (coral reefs) indicated in Fig. 11c; the
majority swam in the sandy grooves that lie between the
coral-reef spurs. These spurs are very shallow (1–3 m
deep), with live coral often protruding above the surface
of the water during low tide. The concentration of BP-3 in
the western groove was 75 ppbillion (lg/L), whereas the
larger, eastern groove had a BP-3 level of 95 ppbillion
(lg/L). Samples were collected between 17:00 and
17:40 h.
Trunk Bay is an iconic landscape and a highly managed
natural resource area. Before 2009, there could be more
than 3000 visitors on the beach and in the water at Trunk
Bay. After 2009, National Park Service policy reduced the
number to 2000 visitors per day (personal communication,
Rafe Boulon, retired, USVI NP Chief, Resource Manage-
ment). A coral community surrounds the island in Trunk
Bay, as well as an abundance of gorgonians to the west of
the island, and there was once a very extensive stand of A.
palmata corals to the east of the island. At a site near the
edge of the Trunk Island coral community, BP-3 levels
were 1.395 ppmillion (mg/L) (Fig. 11d). A sampling site
93 m east of the first sampling site contained 580 ppbillion
(lg/L) BP-3 (Fig. 11d). Samples were collected at
11:00–11:24 h with more than *180 swimmers in the
water and *130 sunbathers on the beach within 100 m of
the two sampling sites.
Seawater samples were collected at five sites in Mau-
nalua Bay, Oahu Island, Hawai’i on May 30, 2011 between
11:00 and 15:00 h (Fig. 12a, b). ASW samples were col-
lected in public swimming areas in waters that were 1.3 m
in depth and 35 cm from the surface of the water. Sites 1–4
had detectable levels of BP-3 ([100 pptrillion; ng/L) but
were below the quantitative range of measurement (5
ppbillion (lg/L); Fig. 12b). Site 5 contained measurable
levels of BP-3—19.2 ppbillion (lg/L) (Supplemental
Fig. 6).
Samples were collected at two sites on June 3, 2011,
along the northwest coast of Maui Island, Hawai’i
(Fig. 12c). Kapalua Bay is a protected cove and has a
Fig. 9 Percentage of deformed planulae of Stylophora pistillata
exposed to various concentrations of benzophenone-3. Bars show
treatment means with whiskers representing ±1 standard error of the
mean. Treatment means with different letters differed significantly at
a=0.05, based on Kruskal–Wallis one-way analysis of variance on
ranks followed by a Student–Newman–Keuls Method post hoc test.
aPlanulae exposed for 8 h in the light. bPlanulae exposed for 8 h in
the light, then 16 h of darkness. cPlanulae exposed for 8 h in the
dark. dPlanulae exposed for 24 h in the dark
Arch Environ Contam Toxicol
public beach that can often see [500 swimmers/day in the
peak tourism season (personal communication, Kapalua
Dive Co.; Fig. 12d). A seawater sample was collected 40 m
from shore near the center of the bay, immediately above
remnants of a coral reef at 09:30 h. The Kapalua sample
had detectable levels of BP-3 but was below the quantita-
tive range of measurement (5 ppbillion, 5 lg/L). From
06:30 to 09:30 h on the day of sampling, 14 swimmers had
entered Kapaula waters. A seawater sample also was col-
lected at Kahekili Beach Park, Maui Island, Hawai’i
(Fig. 12e). Kahekili Beach is a public beach that also
serves visitors from a number of nearby hotels and resorts.
The sample was collected 30 m from shore, immediately
above a coral reef. Unlike Kapalua, Kahekili is an exposed
shoreline not protected within a bay, and retention time of
contaminants is thought to be minimal because of the
prevailing currents. The Kahekili sample had detectable levels
of BP-3 but was below the quantitative range of mea-
surement (5 ppbillion). Kahekili is a heavily visited beach
and had 71 swimmers within 200 m of the sampling site at
the time of sampling (11:45 h).
Benzophenone-3 is a phototoxicant and induces different
toxicities depending on whether the planulae are exposed
to the chemical in light or in darkness. Corals will usually
release brooded planulae at night or spawn gametes at night
(Gleason and Hofmann 2011). Planulae of broadcasting
species (those that spawn eggs and sperm that are fertilized
in the water column) are positively buoyant and planktonic,
residing at or near the surface of the ocean for 2–4 days
before they are able to settle (Fadlallah 1983; Shlesinger
and Loya 1985; Harii et al. 2007; Baird et al. 2009). Light
levels on a clear sunny day in tropic latitudes can be as
high as or higher than 2000 lmol/m
/s of photosyntheti-
cally active radiation—five times more than what the corals
experienced in this study, suggesting that actual environ-
mental conditions may aggravate the phototoxicity. Whe-
ther the BP-3 pollution comes from swimmers, or from
point and nonpoint wastewater sources, planulae will be at
Table 1 Regression and PROBIT determination of LC
for planulae
mortality when exposed to BP-3 in the light and dark, and the EC
for planulae deformity when exposed to BP-3 in the light and the dark
Planulae mortality LC
Regression to estimate LC
8-h light 3.1 mg/L
PROBIT to estimate LC
8-h light 2.9 mg/L
Regression to estimate LC
8-h dark 16.8 mg/L
PROBIT to estimate LC
8-h dark 12.8 mg/L
Regression to estimate LC
24-h light 103.8 lg/L
PROBIT to estimate LC
24-h light 1.39 lg/L
Regression to estimate LC
24-h dark 873.4 lg/L
PROBIT to estimate LC
24-h dark 799 lg/L
Planulae deformation EC
Regression to estimate EC
8-h light 107 mg/L
PROBIT to estimate EC
8-h light 133 mg/L
Regression to estimate EC
8-h dark 436 mg/L
PROBIT to estimate EC
8-h dark 737 mg/L
Regression to estimate EC
24-h light 17 lg/L
PROBIT to estimate EC
24-h light 49 lg/L
Regression to estimate EC
24-h dark 105 lg/L
PROBIT to estimate EC
24-h dark 137 lg/L
Planulae deformation EC
PROBIT to estimate EC
8-h light 6.3
PROBIT to estimate EC
8-h dark 15.5
PROBIT to estimate EC
24-h light 6.5
PROBIT to estimate EC
24-h dark 10.4
PROBIT determination of EC
for planulae deformity when exposed
to BP-3 in the light and the dark
Table 2 Differences in
sensitivities of various species
of Indo-Pacific and Caribbean
coral reefs, the LC
s and LC
of calicoblast cells exposed
in vitro to benzophenone-3 with
their corresponding upper and
lower 95 % confidence intervals
for the two Indo-Pacific and five
Caribbean species. (lg/L) =to
parts per billion. (ng/L) =parts
per trillion
Coral species LC
(lg/L) 95 % CI LC
95 % CI
Indo-Pacific species
Stylophora pistillata (light) 42 28; 60 2 lg/L 1.14; 3.61
Stylophora pistillata (dark) 671 447; 984 14 lg/L 7; 26
Pocillopora damicornis 8 4.96; 12.15 62 ng/L 24; 136
Caribbean-Atlantic species
Acropora cervicornis 9 5.4; 14.5 63 ng/L 22; 150
Montastrea annularis 74 40; 126 562 ng/L 166; 1459
Montastrea cavernosa 52 36; 72 502 ng/L 247; 921
Porites astreoides 340 208; 534 8 lg/L 3; 16
Porites divaricata 36 21; 57 175 ng/L 60; 420
Arch Environ Contam Toxicol
risk from both forms of toxicities (Brooks et al. 2009;
Futch et al. 2010; Pitarch et al. 2010).
As with our previous paper examining benzophenone-2
(Downs et al. 2014), the data in this paper are consistent
with the observation by Danovaro et al. (2008) that ‘‘sun-
screens compounds’’ cause coral bleaching. In the light,
BP-3 caused injury directly to the zooxanthellae, inde-
pendent of any host-regulated degradation mechanism.
Based on the pathomorphology of the thylakoids within the
chloroplasts, the most probable interpretation is that BP-3
induces photo-oxidative stress to the molecular structures
that form the thylakoid membranes (Downs et al. 2013). In
darkness, bleaching resulted from the symbiophagy of the
symbiotic zooxanthellae; a process whereby the coral
gastrodermal cell ‘‘digests’’ the zooxanthella (Downs et al.
2009). Nesa et al. (2012) demonstrated that following
exposure to light, the algal symbionts of corals increased
the DNA damage to coral cells in coral planulae.
Consistent with the Oxidative Theory of Coral Bleaching
(Downs et al. 2002), Nesa et al. hypothesized that the
sources of this damage was the production of oxygen
radicals. If this is the case, then darkness-associated, BP-3-
induced bleaching may reduce the exacerbated morbidity
experienced by ‘‘bleached’’ planulae that would occur
during the periods of daylight. Regardless of the toxico-
logical mechanism, managing exposure of corals to BP-3
corals will be critical for managing coral reef resilience in
the face of climate-change pressures associated with coral
bleaching (West and Salm 2003).
Autophagy was the dominant cellular response to BP-3
exposure (Figs. 4a–f, 5b–d; Yla-Antilla et al. 2009). Micro-
autophagosomes were abundant in all cell types and larger
vacuolated bodies of specific organelles were readily
observed. None of the nuclei in any coral cell-types
exhibited any of the classic signs of apoptosis, such as
pyknosis or karyorrhexis of the nucleus (Krysko et al.
Fig. 10 Percentage mortality of calicoblast cells of Stylophora
pistillata exposed to various concentrations of benzophenone-3. Bars
show treatment means (n=4) with whiskers representing ±1
standard error of the mean. Treatment means with different letters
differed significantly at a=0.05, based on one-way analysis of
variance followed by a Tukey’s Honestly Significant Difference Test.
aCalicoblast cells exposed for 4 h in the light. bLog-linear
regression between coral cell mortality and concentrations of BP-3 for
4 h in the light. Quadratic regression line (solid) and 95 % confidence
intervals (dashed lines) are shown. Larger symbols represent multiple
coincident data points, with symbol area proportional to the number
of replicates with the same value. cCalicoblast cells exposed for 4 h
in the dark. dLog-linear regression between coral cell mortality and
concentration of BP-3 for 4 h in the dark
Arch Environ Contam Toxicol
2008). The most fascinating aspect of these autophagic
events were the delamination of the nuclear bilayer mem-
brane (Figs. 4b, c, 5b, c), a classic hallmark of autophagic
cell death and further evidence arguing against apoptosis as
a regulated mechanism of cnidarian cell death (Tasdemir
et al. 2008; Yla-Antilla et al. 2009; Klionsky et al. 2012).
In both the light and the dark, there was a gradation of
vitiated cells beginning at the surface of the epidermis to
‘non-morbid’’ cells in the gastrodermis that surrounded the
yolk. In Figs. 4a and 5a, the cells are severely degraded; it
is difficult to distinguish any mechanism of cell death, and
the cells could easily be labeled as necrotic. Going
20,000 nm into the planula from the surface, cells exhib-
ited the hallmarks of autophagic cell death. This tissue
transect of the gradation of cell death is evidence for a
model of cell death, first demonstrated in C. elegans, that
requires autophagic degradation of cells for the manifes-
tation of necrosis (Samara et al. 2008; Eskelinin et al.
BP-3 is a genotoxicant to corals, and its genotoxicity is
exacerbated by light. Based on the current literature, this
was not unexpected, but our data do underscore the threat
that BP-3 may pose to not only corals but also to other
coral-reef organisms (Hanson et al. 2006; Cuquerella et al.
2012). DNA AP lesions can be produced in response to
oxidative interaction or alkylation events (Fortini et al.
1996; Drablos et al. 2004). Accumulation of DNA damage
in the larval state has implications not only for the success
of coral recruitment and juvenile survival, but also for
reproductive effort and success as a whole (Anderson and
Wild 1994; Depledge and Billinghurst 1999). Surviving
planulae exposed to BP-3 may settle, metamorphose, and
Fig. 11 Seawater analysis of benzophenone-3 (BP-3) in coral reef
areas in St. John Island, U.S. Virgin Islands. aAerial view of St. John
indicating the five sampling sites, indicated by a yellow dot.No
benzophenones were detected in samples from Red Point or at Tektite
Reef. All samples were taken between 12:00 and 14:00 h. Scale bar is
1.5 km. bAerial view of the three northwestern sites within St. John
National Park: Trunk Bay, Hawksnest Bay, and Caneel Bay. The two
sampling sites at Caneel Bay are indicated by yellow dots.No
benzophenones were detected in samples from Caneel Bay. Scale bar
is 500 m. cAerial view of the two sampling sites in Hawksnest Bay,
St. John Island. Yellow arrows indicate three coral reef spurs that are
dominated by the U.S. Threatened Species, Acropora palmata.Yellow
arrows pointing at red dots indicate the sample site. Values indicate
the concentration of BP-3 in the water column. Scale bar is 245 m.
dElevated view of Trunk Bay, St. John Island. Yellow arrows
pointing to red ‘X’’ indicate the sample site. The values indicate the
concentration of BP-3 in the water column at those two sites
Arch Environ Contam Toxicol
develop into colonial adults, but they may be unfit to meet
the challenges of other stressor events, such as increased
sea-surface temperatures. Cnidarians are rather unusual in
the animal kingdom in that the germline is not sequestered
away from the somatic tissue in early stages of develop-
ment; the somatic tissue gives rise directly to the germline
during seasonal reproductive cycles. Damage to the geno-
mic integrity of coral planulae therefore may have far-
reaching and adverse impacts on the fitness of both the
gametes in adults.
The ossification of the planulae from exposure to BP-3 is
one of the strangest cases of developmental endocrine dis-
ruption to wildlife, although skeletal endocrine disruption in
vertebrates is only now being recognized (Colburn et al.
1993; Depledge and Billinghurst 1999; Golub et al. 2004;
Lind et al. 2004; Doherty et al. 2004;Agasetal.2013). In
mammals, estrogen and estrogenic compounds may influence
different estrogen and thyroid hormone receptors, which
affect bone growth and composition (Rickard et al. 1999;
Lindberg et al. 2001; Golub et al. 2004). In classic vertebrate
physiology, estrogen plays a complex role in ossification and
skeletal maintenance, affecting both bone anabolism and
catabolism (Simmons 1966;Va
¨nen and Ha
¨nen 1996).
In vertebrates, exposure to high levels of estrogen can result
in skeletal hyperossification (Pfeiffer et al. 1940;Rickard
et al. 1999). For ‘‘classic’’ endocrine disruptors, such as tri-
butyltin and dioxin, ossification is inhibited, not induced
(Birnbaum 1995; Jamsa et al. 2001; Tsukamoto et al. 2004;
Finnila et al. 2010;Agasetal.2013). Osteo-endocrine dis-
ruption is both complex and complicated; different com-
pounds affect different cell types within the skeletal tissue
differently (Hagiwara et al. 2008a,b;Agasetal.2013).
Benzophenones as endocrine disruptors are no exception;
BP-3 and BP-2 showed contradictory effects on estrogen and
aryl hydrocarbon receptors, and both compounds induced
‘‘ a kind of endocrine disruption that is not assessed by
‘classical’ estrogenic markers’’ (Schlecht et al. 2004;Sei-
´-Wuttke et al. 2004; Ziolkowaska et al. 2006).
The ossification-induced opacity of the epidermal tissue
layer of planulae was readily observed at the three highest
Fig. 12 Seawater analysis of benzophenone-3 (BP-3) in coral reef
areas in Oahu and Maui islands, Hawai’i. Yellow dots indicate the
sampling location in the panels. aAerial view of Oahu indicating the
five sampling sites. Scale bar is 5 km. bAerial view of the five
sampling site along the coast of Maunalua Bay, Oahu. Sites 1–4 had
levels of BP-3 that were detectable, but below the quantitative range.
Scale bar is 1.5 km. cAerial view of the two sampling sites in Maui,
Hawai’i. Scale bar is 6 km. dElevated view of Kapalua Bay, Maui.
Scale bar is 100 m. eElevated view of Kahekili Beach, Maui. Scale
bar is 100 m
Arch Environ Contam Toxicol
concentrations of BP-3 exposure but was not visually
obvious at the lower concentrations, although we know
from the electron microscopy sample processing that
ossification was present to a lesser extent in the lower BP-3
exposures. Many endocrine disruptors do not exhibit a
‘classic’’ monotonic exposure–response curve, but instead
exhibit nonmonotonic behaviors (vom Saal et al. 1995;
Conolly and Lutz 2004;
tonic.html). Ossification of planulae can be assayed by a
variety of methods, including alizarin staining and calcein
fluorescence. This study was not designed to be an
exhaustive characterization of exposure–response behavior
(i.e., regulatory toxicology); hence lower BP-3 exposure
concentrations were not attempted. More comprehensive
studies that examine the ossification response of both acute
and chronic exposure of BP-3 in the lower pptrillion and
ppquadrillion need to be conducted to determine accurately
this endocrine behavioral response.
Ecotoxicology and Species Sensitivity
To conduct a relevant and accurate ecological risk or threat
assessment, it is imperative that the species chosen reflects
the structure of the specific coral-reef ecosystem being
affected (Suter 2007). Stylophora pistillata used in this
study, is indigenous to specific regions in the Indo-Pacific
basins, and hence may not be a valid representative for
coral-reef communities in Hawaii or the Atlantic/Car-
ibbean basins. The use of coral planulae in research studies
is a relatively difficult resource to obtain. It requires access
to healthy coral colonies that are reproductively viable,
spawning in specific dates and specific moon phases, and in
addition, obtaining the necessary collection and transport
permits. We therefore applied an in vitro primary cell
toxicity methodology using a specific coral cell type that
has been proposed as a surrogate for either planula or
colonial polyp studies (Downs 2010). Comparison of LC
of coral cells in the light (42 ppbillion; lg/L) and coral
planula in the light for 8 and 24 h [2.876 ppmillion (mg/L)
and 139 ppbillion (lg/L), respectively] exhibits a similar
response. The increased sensitivity of in vitro cell models
versus whole organism models is a common phenomenon
and accepted principle (Blaauboer 2008; Gura 2008).
Diffusion of BP-3 across the epidermal boundary layer and
reaching concentrations that are toxic in the interior of the
planula (e.g., gastroderm) versus direct exposure by cul-
tured cells could likely be the major factor influencing the
variation in LC
rate. Although there are obvious caveats
to using in vitro models, this may be the only way to
conduct ecotoxicological research and ecological risk
assessments on coral species that are currently endangered
with extinction, such as the species on the IUCN’s Red List
or species proposed/listed for protection under the U.S.
Endangered Species Act.
When an environmental stressor impacts a community
of organisms, different species may respond (tolerate)
dissimilarly to one another; some species may tolerate the
stressor at a particular level, whereas other species may
succumb (Johnston and Roberts 2009; Maloney et al.
2011). This species sensitivity distribution is a crucial
concept for ecological risk assessments and a predictor of
the species composition of a community (community
phase-shift) in reacting to a pollution stressor, as well as
defining the probability of success for community/ecolog-
ical restoration (Posthuma et al. 2002; van Woesik et al.
2012). This concept readily applies to corals and coral
reefs. Coral bleaching in response to heat stress or fresh-
water input is an excellent example of this community
behavior; some species have high tolerance to stress-in-
duced bleaching, whereas others are highly susceptible,
resulting in species-specific extinctions in localized areas
(Goreau 1990; Loya et al. 2001; Jimenez and Cortes 2003).
Species sensitivity distribution in response to pollutants in
corals is also well documented, including synergisms
between pollutants and heat stress (Loya 1975; Brown
2000; Fabricius 2005). For the Caribbean, the species
sensitivity to BP-3 toxicity is consistent with the model for
coral tolerance to general stress as set forth by Gates and
Edmunds (1999): corals with slower growth rates, such as
massive or boulder coral species, are inherently more tol-
erant than coral species with higher growth rates (e.g.,
branched species such as A. cervicornis and P. divaricata).
In fringing reefs that have been impacted by anthropogenic
stressors, especially fringing reefs near tourist beaches,
Acropora species are the first to experience localized
extinction. Species that tentatively endure a decade or
longer of sustained stress, but are intermediate in their
persistence, are the large boulder corals found in the genus
Montastrea (synomym Orbicella). Coral cell toxicity data
indicated that P. astreoides was at least 4.59more tolerant
to BP-3 toxicity than the second more tolerant coral species
and at least 389more tolerant than the most sensitive
species. This is consistent with observations that P.
astreoides is usually the last to become extinct in a pol-
luted-impacted locality and one of the first to recruit once
water quality parameters reach a minimum level of habit-
ability (Peters 1984; Lirman et al. 2003; Alcolado-Prieto
et al. 2012). From a management perspective, these data
can be used to predict the changes in coral-reef community
structure when challenged with BP-3, regarding which
species will become extinct, as well as the species that will
persist in areas that are adjacent to tourist beaches, popular
mooring sites, or near sewage discharges. These data also
can be integrated directly into reef resilience management
Arch Environ Contam Toxicol
plans against climate change, acting as both a measurable
endpoint for management effectiveness and as a target
(concentration of BP-3 in seawater on a reef) for estab-
lishing action values for reef management.
Management of BP-3 Pollution for Coral Reef
Conservation and Restoration
What do these pathological toxicities induced by BP-3
mean demographically and ecologically for corals and
coral reefs? Trunk Bay in St. John Island, the U.S Virgin
Islands, may represent an example of this effect. Ecologi-
cally, this area has been severely degraded in the past
25 years, despite the limited input from human activities in
the watershed and from marine sources. The most obvious
input is recreational swimming at Trunk (Downs et al.
2011). During our monitoring of this site from 2005 to
2010, settlement of planulae and recruitment/survival of
juvenile coral was almost 0 %. Established coral colonies
in this area were assayed for regeneration of tissue over
experimentally induced lesions (laceration-regeneration
assay, a single diagnostic test for the general health of a
coral; Fisher et al. 2007); not a single colony exhibited any
regeneration of any of the lesions during the 5-year
investigation (Downs et al. 2011). This was in contrast with
Caneel Bay, which had undetectable levels of BP-3
resulting from a much lower density/rate of swimmers and
has a flourishing coral community on its southern bank
with an abundance of recruitment. These demographic-
level pathologies are consistent with the pathologies that
manifest from BP-3 exposure. The pathologies exhibited at
this site can be seen at other coral reef swimming areas the
world over: Eilat, Israel (degraded with an abundance of
sunscreen lotion users) versus Aqaba, Jordan (thriving
coral reefs with swimmers that do not use sunscreen lotion;
Fuad Al-horani, personal communication), Honolua Bay in
Maui, Hawaii, Hanauma Bay Beach in Oahu, Hawaii,
Seven Mile Beach in Grand Cayman, Bathway Beach in
Grenada, Playa Langosta, and Playa Tortugas Beaches in
Cancun, Mexico. At Okinawa, Tashiro and Kameda (2013)
demonstrated that BP-3 contamination from beaches can
travel over 0.6 km in distance from the pollution source.
The threat of BP-3 to corals and coral reefs from swimmers
and point and non-point sources of waste-water could thus
be far more extensive than just a few meters surrounding
the swimming area.
Acknowledgments The study in Israel was partially funded by the
Israel Science Foundation (ISF) No. 1169/07 to Yossi Loya. No other
organization or government provided Grant-in-aid funding for this
project. The authors thank Dr. Jon Martinez and Dr. Katherine
Schaefer for assistance with water sampling in Oahu, Hawai’i, Ms.
Maya Vizel for her assistance with the planula exposure challenges,
Dr. Gideon Winters for assistance with Molecular Dynamics
microplate fluorimeter, and Dr. Fuad Al-Horani for his assistance with
toxicological exposures. We sincerely thank Dr. Sylvia Galloway and
Mr. James H. Nicholson for their work on formatting the figures for
publication. We also wish to thank the U.S. National Park Service of
the U.S. Virgin Islands National Park for their assistance. We wish to
thank the two anonymous reviewers for their comments in improving
the manuscript. C.A. Downs thanks the unidentified Virgin Islander in
Cruz Bay who gave him insight into the hypothetical cause of the
ecological collapse occurring at Trunk Bay; hypothesizing that the
visible ‘‘sheen’’ on the surface of the water produced from swimmers’
sunscreen lotions was somehow impacting coral reef health.
Compliance with Ethical Standards
Conflict of Interest The authors can identify no potential conflicts
of interest, neither financial nor ethically, involved in the writing or
publication of this manuscript.
Disclaimer The intent of this article is purely for dissemination of
scientific knowledge, and is neither an endorsement nor condemna-
tion of the activities of any government, corporation, their employees
or subsidiaries, nor to imply liability on their part. This publication
does not constitute an endorsement of any commercial product or
intend to be an opinion beyond scientific or other results obtained by
the U.S. National Oceanic and Atmospheric Administration (NOAA).
No reference shall be made to U.S. NOAA, or this publication fur-
nished by U.S. NOAA, to any advertising or sales promotion which
would indicate or imply that U.S. NOAA recommends or endorses
any proprietary product mentioned herein, or which has as its purpose
an interest to cause the advertised product to be used or purchased
because of this publication.
Abelson A, Ronen O, Gaines S (2005) Coral recruitment to the reefs of
Eilat, Red Sea: temporal and spatial variation, and possible effects
of anthropogenic disturbances. Mar Pollut Bull 50:576–582
Agas D, Sabbieti MG, Marchetti L (2013) Endocrine disruptors and
bone metabolism. Arch Toxicol 87:735–751
Agresti A (2002) Categorical data analysis, 2nd edn. Wiley, New
Alcolado-Prieto P, Aragon HC, Alcolado PM, Castillo AL (2012)
Stony coral recruitment in coral reefs at different distances from
pollution sources in Habana, Cuba. Rev Biol Trop 60:981–994
Anderson SL, Wild GC (1994) Linking genotoxic responses and
reproductive success in ecotoxicology. Environ Health Perspect
Aquera A, Martinez Bueno MJ, Fernandez-Alba AR (2013) New
trends in the analytical determination of emerging contaminants
and their transformation products in environmental waters.
Environ Sci Pollut Res Int 20:3496–3515
Baird AH, Guest JR, Willis BL (2009) Systematic and biogeograph-
ical patterns in the reproductive biology of scleractinian corals.
Annu Rev Ecol Evol Syst 40:551–571
Barnes DJ (1972) The structure and formation of growth-ridges in
scleractinian coral skeletons. Proc R Soc Lond B 182:331–350
Baron E, Gago-Ferrero P, Gorga M et al (2013) Occurrence of
hydrophobic organic pollutants (BFRs and UV-filters) in sedi-
ments from South America. Chemosphere 92:309–316
Birnbaum LS (1995) Developmental effects of dioxins. Environ
Health Perspect 103:89–94
Blaauboer BJ (2008) The contribution of in vitro toxicity data in
hazard and risk assessment: current limitations and future
perspectives. Toxicol Lett 180:81–84
Arch Environ Contam Toxicol
Blitz JB, Norton SA (2008) Possible environmental effects of
sunscreen run-off. J Am Acad Dermatol. doi:10.1016/j.jaad.
Bluthgen N, Zucchi S, Fent K (2012) Effects of the UV filter
benzophenone-3 (oxybenzone) at low concentrations in zebrafish
(Danio rerio). Toxicol Appl Pharmacol 263:184–194
Brooks AC, Gaskell PN, Maltby LL (2009) Importance of prey and
predator feeding behaviors for trophic transfer and secondary
poisoning. Environ Sci Technol 43:7916–7923
Brown BE (2000) The significance of pollution in eliciting the
‘bleaching’ response in symbiotic cnidarians. Int J Environ
Pollut 13:392–415
Carpenter KE, Abrar M, Aeby G et al (2008) One-third of reef-
building corals face elevated extinction risk from climate change
and local impacts. Science 321:560–563
Carson FL (1997) Histotechnology: a self-instructional text, 2nd edn.
American Society of Clinical Pathologists, Chicago
CIR (Cosmetic Ingredient Review) (2005) Annual review of cosmetic
ingredient safety assessments: 2003/2003. Int J Toxicol
Colborn T, vom Saal FS, Soto AM (1993) Developmental effects of
endocrine-disrupting chemicals in wildlife and humans. Environ
Health Perspect 101:378–384
Conolly RB, Lutz WK (2004) Nonmonotonic dose-response relation-
ships: mechanistic basis, kinetic modeling, and implications for
risk assessment. Toxicol Sci 77:151–157
Coronado M, De Haro H, Deng X et al (2008) Estrogenic activity and
reproductive effects of the UV-filter oxybenzone (2-hydroxy-4-
methoxyphenyl-methanone) in fish. Aquat Toxicol 90:182–187
Cosnefroy A, Brion F, Maillot-Marechal E et al (2011) Selective
activation of zebrafish estrogen receptor subtypes by chemicals
by using stable reporter gene assay developed in a zebrafish liver
cell line. Toxicol Sci 125:439–449
Crang RFE, Klomparens KL (1988) Artifacts in biological electron
microscopy. Plenum Press, New York
Crawley MJ (1993) GLIM for ecologists. Blackwell, London
Cuquerella MC, Lhiaubet-Vallet V, Cadet J, Miranda MA (2012)
Benzophenone photosensitized DNA damage. Acc Chem Res
Danovaro R, Bongiorni L, Corinaldesi C et al (2008) Sunscreens
cause coral bleaching by promoting viral infections. Environ
Health Persp 116:441–447
Depledge MH, Billinghurst Z (1999) Ecological significance of
endocrine disruption in marine invertebrates. Mar Pollut Bull
Doherty TM, Fitzpatrick LA, Inoue D et al (2004) Molecular,
endocrine, and genetic mechanisms of arterial calcification.
Endocr Rev 25:629–672
Downs CA, Fauth JE, Halas JC, Dustan P, Bemiss J, Woodley CM
(2002) Oxidative stress and seasonal coral bleaching. Free Radic
Biol Med 32:533–543
Downs CA, Kramarsky-Winter E, Martinez J et al (2009) Symbio-
phagy as a mechanism for coral bleaching. Autophagy
Downs CA, Fauth JF, Downs VD, Ostrander GK (2010) In vitro cell-
toxicity screening as an alternative animal model for coral
toxicology: effects of heat stress, sulfide, rotenone, cyanide, and
cuprous oxide on cell viability and mitochondrial function.
Ecotoxicology 19:171–184
Downs CA, Woodley CM, Fauth JE et al (2011) A survey of
environmental pollutants and cellular-stress biomarkers of
Porites astreoides at six sites in St. John, U.S. Virgin Islands.
Ecotoxicology 20:1914–1931
Downs CA, Ostrander GK, Rougee L et al (2012) The use of cellular
diagnostics for identifying sub-lethal stress in reef corals.
Ecotoxicology 21:768–782
Downs CA, McDougall KE, Woodley CM et al (2013) Heat stress and
light stress induce different cellular pathologies in the symbiotic
dinoflagellate during coral bleaching. PLoS One 8(12):e77173.
Downs CA, Kramarsky-Winter E, Fauth JE et al (2014) Toxicological
effects of the sunscreen UV filter, benzophenone-2, on planula
and in vitro cells of the coral, Stylophora pistillata. Ecotoxicol-
ogy 23:175–191
Drablos F, Feyzi E, Aas PA et al (2004) Alkylation damage in DNA
and RNA: repair mechanisms and medical significance. DNA
Repair 3:1389–1407
Draper NR, Smith H (1966) Applied regression analysis. Wiley, New
Dustan P (1977) Vitality of reef coral populations off Key Largo,
Florida: recruitment and mortality. Environ Geol 2:51–58
Edinger EN, Jompa J, Limmon GV, Widjatmoko W, Risk MJ (1998)
Reef degradation and coral biodiversity in Indonesia: effects of
land-based pollution, destructive fishing practices and changes
over time. Mar Pollut Bull 36:617–630
Eichenseher T (2006) The cloudy side of sunscreens. Environ Sci
Technol 40:1377–1378
Eskelinin EL, Reggiori F, Baba M, Kovacs AL, Seglen PO (2011)
Seeing is believing: the impact of electron microscopy on
autophagy research. Autophagy 7:935–956
Fabricius KE (2005) Effects of terrestrial runoff on the ecology of
corals and coral reefs: review and synthesis. Mar Pollut Bull
Fadlallah YH (1983) Sexual reproduction, development and larval
biology in scleractinian corals: a review. Coral Reefs 2:129–150
Fent K, Kunz PY, Zenker A, Rapp M (2010) A tentative environmental
risk assessment of the UV-filters 3-(4-methylbenzylidene-camphor),
2-ethyl-hexyl-4-trimethoxycinnamate, benzophenone-3, benzophe-
none-4 and 3-benzylidene camphor. Mar Environ Res 69:S4–S6
Finney DJ (1947) Probit analysis, a statistical treatment of the
sigmoid response curve. Cambridge University Press,
Finnila MA, Zioupos P, Herlin M, Miettinen HM, Simanainen U,
Hakansson H, Tuukkanen J, Viluksela M, Jamsa T (2010)
Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin exposure on bone
material properties. J Biomech 43:1097–1103
Fisher EM, Fauth JE, Hallock-Muller P, Woodley CM (2007) Lesion
regeneration rates in reef-building corals Monstrastrea Spp. as
indicators of colony condition. Mar Ecol Prog Ser 339:61–71
Fortini P, Raspaglio G, Falchi M, Dogliotti E (1996) Analysis of
DNA alkylation damage and repair in mammalian cells by the
COMET assay. Mutagen 11:169–175
French JE (1992) NTP technical report on the toxicity studies of
2-hydroxy-4-methoxybenzophenone (CAS No. 131-57-7)
administered topically and in dosed feed to F344/N Rats and
B6C3F1 mice. Toxic Rep Ser 21:1–14
Futch JC, Griffin DW, Lipp EK (2010) Human enteric viruses in
groundwater indicate offshore transport of human sewage to
coral reefs of the upper Florida keys. Environ Microbiol
Gago-Ferrero P, Diaz-Cruz MS, Barcelo D (2011) Occurrence of
multiclass UV filters in treated sewage sludge from wastewater
treatment plants. Chemosphere 84:1158–1165
Gao L, Yuan T, Zhou C, Cheng P, Bai Q et al (2013) Effects of four
commonly used UV filters on the growth, cell viability and
oxidative stress responses of the Tetrahymena thermophila.
Chemosphere 93:2507–2513
Gates RD, Edmunds PJ (1999) The physiological mechanisms of
acclimatization in tropical reef corals. Am Zool 39:30–43
Gilbert E, Pirot F, Bertholle V, Roussel L, Falson F, Padois K (2012)
Commonly used UV filter toxicity on biological functions:
review of last decade studies. Int J Cosmet Sci 35:208–219
Arch Environ Contam Toxicol
Gleason DF, Hofmann DK (2011) Coral larvae: from gametes to
recruits. J Exp Mar Biol Ecol 408:42–57
Golbuu Y, Fabricius K, Victor S, Richmond R (2008) Gradients in
coral reef communities exposed to muddy river discharges in
Pohnpei, Micronesia. Estuar Coast Shelf S 76:14–20
Golub MS, Hogrefe CE, Germann SL, Jerome CP (2004) Endocrine
disruption in adolescence: immunologic, hematologic, and bone
effects in monkeys. Toxicol Sci 82:598–607
Goreau TJ (1990) Coral bleaching in Jamaica. Nature 343:417
Gulati D, Mounce R (1997) NTP reproductive assessment by
continuous breeding study for 2-hydroxy-4-methoxybenzophe-
none in Swiss CD-1 mice. NTIS# PB91158477. Environ Health
Perspect 105(Suppl 1):313–314
Gura T (2008) Toxicity testing moves from the legislature to the Petri
dish—and back. Cell 134:557–559
Hagiwara H, Sugizaki T, Tsukamoto Y, Senoh E, Goto T, Ishihara Y
(2008a) Effects of alkylphenols on bone metabolism in vivo and
in vitro. Toxicol Lett 181:13–18
Hagiwara H, Suizaki T, Tsukamoto Y (2008b) Effects of alkylphenols
on bone metabolism in vivo and in vitro. Toxicol Lett 181:13–18
Hanson KM, Gratton E, Bardeen CJ (2006) Sunscreen enhancement
of UV-induced reactive oxygen species in the skin. Free Radic
Biol Med 41:1205–1212
Hany J, Nagel R (1995) Detection of sunscreen agents in human
breast milk. Dtsch Lebensm Rundsch 91:341–345
Harii S, Nadaoka K, Yamamoto M, Iwao K (2007) Temporal changes
in settlement, lipid content, and lipid composition of larvae of
the spawning hermatypic coral Acropora tenuis. Mar Ecol Prog
Ser 346:89–96
Hughes TP, Tanner JE (2000) Recruitment failure, life histories, and
long-term decline of Caribbean corals. Ecology 81:2250–2263
Jamsa T, Viluksela M, Tuomisto JT, Tuomisto J, Tuukkanen J (2001)
Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on bone in two rat
strains with different aryl hydrocarbon receptor structures.
J Bone Miner Res 16:1812–1820
Jeon HK, Chung Y, Ryu JC (2006) Simultaneous determination of
benzophenone-type UV filters in water and soil by gas
chromatography-mass spectrometry. J Chromatogr A 1131:
Jimenez CE, Cortes J (2003) Coral cover change associated to El
Nino, Eastern Pacific, Costa Rica, 1992–2001, PSZNI. Mar Ecol
Johnston EL, Roberts DA (2009) Contaminants reduce the richness
and evenness of marine communities: a review and meta-
analysis. Environ Pollut 157:1745–1752
Kameda Y, Kimura K, Miyazaki M (2011) Occurrence and profiles
of organic sun-blocking agents in surface waters and sedi-
ments in Japanese rivers and lakes. Environ Pollut 159:
Kerdivel G, Le Guevel R, Habauzit D, Brion F, Ait-Aissa S, Pakdel F
(2013) Estrogenic potency of benzophenone UV filters in breast
cancer cells: proliferative and transcriptional activity substanti-
ated by docking analysis. PLoS One 8:e60567. doi:10.1371/
Kerr JFR, Wullie AH, Currie AR (1972) Apoptosis: a basic biological
phenomenon with wide-ranging implication in tissue kinetics. Br
J Cancer 26:239–257
Klionsky DJ, Abdalalla FC, Abeliovich H et al (2012) Guidelines for
the use and interpretation of assays for monitoring autophagy.
Autophagy 8:445–544
Knowland J, McKenzie EA, McHugh PJ, Cridland NA (1993)
Sunlight-induced mutagenicity of a common sunscreen ingredi-
ent. FEBS Lett 324:309–313
Krysko DV, Berghe TV, D’Herde K, Vandenabeele P (2008)
Apoptosis and necrosis: detection, discrimination and phagocy-
tosis. Methods 44:205–221
Kunisue T, Chen Z, Buck Louis GM et al (2012) Urinary concentrations
of benzopheone-type UV filters in U.S. womenand their association
with endometriosis. Environ Sci Technol 46:4624–4632
Kunz PY, Galicia HF, Fent K (2006) Comparison of in vitro and
in vivo estrogenic activity of UV filters in fish. Toxicol Sci
Kushmaro A, Henning G, Hofmann DK, Benayahu Y (1997)
Metamorphosis of Heteroxenia fuscescens planulae (Cnidaria:
Octocorallia) is inhibited by crude oil: a novel short-term
toxicity bioassay. Mar Environ Res 43:295–302
Laskowski R (1995) Some good reasons to ban the use of NOEC,
LOEC, and related concepts in ecotoxicology. Oikos 73:140–144
Lind PM, Milnes MR, Lundberg R et al (2004) Abnormal bone
composition in female juvenile American alligators from a
pesticide-polluted lake. Environ Health Perspect 112:359–362
Lindberg MK, Erlandsson M, Alatalo SL et al (2001) Estrogen
receptor alpha, but not estrogen receptor beta, is involved in the
regulation of the OPG/RANKL (osteoprotegerin/receptor acti-
vator of NF-kappa B ligand) ratio and serum interleukin-6 in
male mice. J Endocrinol 171:425–433
Lirman D, Orlando B, Macia S, Manzello D, Kaufman L et al (2003)
Coral communities of Biscayne Bay, Florida and adjacent
offshore areas: diversity abundance, distribution, and environ-
mental correlates. Aquat Conserv 13:121–135
Loya Y (1975) Possible effects of water pollution on the community
structure of Red Sea corals. Mar Biol 29:177–185
Loya Y, Sakai K, Yamazato K, Nakano Y, Sambali H et al (2001)
Coral bleaching: the winners and the losers. Ecol Lett 4:122–131
Maloney KO, Munguia P, Mitchell RM (2011) Anthropogenic
disturbance and landscape patterns affect diversity patterns of
aquatic benthic macroinvertebrates. J N Am Benthol Soc
Miller MW, Weil E, Szmant AM (2000) Coral recruitment and
juvenile mortality as structuring factors for reef benthic
communities in Biscayne National Park, USA. Coral Reefs
Molina-Molina J-M, Escande A, Pillon A et al (2008) Profiling of
benzophenone derivatives using fish and human estrogen
receptor-specific in vitro bioassays. Toxicol Appl Pharmacol
Morohoshi K, Yamamoto H, Kamata R, Shiraishi F, Koda T, Morita
M (2005) Estrogenic activity of 37 components of commercial
sunscreen lotions evaluated by in vitro assays. Toxicol In Vitro
Nakajima D, Asada S, Kageyama et al (2006) Activity related to the
carcinogenicity of plastic additives in the benzophenone group.
J UOEH 28:143–156
Nashez LG, Schuster D, Laggner C et al (2010) The UV-filter
benzophenone-1 inhibits 17 beta-hydrozysteroid dehydrogenase
type 3: virtual screening as a strategy to identify potential
endocrine disrupting chemicals. Biochem Pharmacol 79:
Nesa B, Baird AH, Harii S, Yakovleva I, Hidaka M (2012) Algal
symbionts increase DNA damage in coral planulae exposed to
sunlight. Zool Stud 51:12–17
Newman MC (2013) Quantitative ecotoxicology. CRC Press, Boca
Nimrod AC, Benson WH (1998) Reproduction and development of
Japanese medaka following an early life stage exposure to
xenoestrogens. Aquat Toxicol 44:141–156
NRC (National Research Council) (2013) Assessing risks to endan-
gered and threatened species from pesticides. National Academy
of Sciences. ISBN 978-0-309-28583-4
NTP (National Toxicology Program) (2006) NTP technical report on
the toxicology and carcinogenesis of benzophenone in F344/N
rats and B6C3F1 mice. NIH Publication # 06-4469
Arch Environ Contam Toxicol
OECD (2006) Current approaches in the statistical analysis of
ecotoxicity data: a guidance to application. OECD Environment
Health & Safety Publications Series on Testing and Assessment.
No. 54. Organization for Economic Cooperation and Develop-
ment, Paris
OECD (2013) OECD Guidelines for the testing of chemicals: fish
embryo acute toxicity test. Organization for Economic Cooper-
ation and Development, Paris
Omori M (2011) Degradation and restoration of coral reefs:
experience in Okinawa, Japan. Mar Biol Res 7:3–12
´ez I, Martinez-Guitarte JL, Morcillo G (2013) Effects of in vivo
exposure to UV filters (4-MBC, OMC, BP-3, 4-HB, OC, OD-
PABA) on endocrine signaling genes in the insect Chironomus
riparius. Sci Total Environ 456–457:120–126
Peters EC (1984) A survey of cellular reactions to environmental
stress and disease in Caribbean scleractinian corals. Helgol
Meeresunters 37:113–137
Pfeiffer CA, Kirschbaum A, Gardner WU (1940) Relation of estrogen
to ossification and the levels of serum calcium and lipoid in the
English Sparrow, Passer domesticus. Yale J Biol Med
Pitarch E, Portole
´s T, Marı
´n JM et al (2010) Analytical strategy based
on the use of liquid chromatography and gas chromatography
with triple-quadrupole and time-of-flight MS analyzers for
investigating organic contaminants in wastewater. Anal Bioanal
Chem 397:2763–2776
Platt KL, Aderhold S, Kulpe K, Fickler M (2008) Unexpected DNA
damage caused by polycyclic aromatic hydrocarbons under
standard laboratory conditions. Mutat Res 650:96–103
Popkin DJ, Prival MJ (1985) Effects of pH on weak and positive
control mutagens in the AMES Salmonella plate assay. Mutat
Res 142:109–113
Posthuma L, Suter GW II, Traas TP (2002) Species sensitivity
distributions in ecotoxicology. Lewis Publishers, Boca Raton,
p 587 pp
Rachon D, Rimoldi G, Wuttke G (2006) In vitro effects of
benzophenone-3 and octyl-methoxycinnamate on the production
of interferon-cand interleukin-10 by murine splenocytes.
Immunopharmacol Immunotoxicol 28:501–510
Rees JG, Setiapermana D, Sharp VA, Weeks JM, Williams TM
(1999) Evaluation of the impacts of land-based contaminants on
the benthic faunas of Jakarta Bay, Indonesia. Oceanol Acta
Richardson SD (2006) Environmental mass spectrometry: emerging
contaminants and current issues. Anal Chem 78:4021–4046
Richardson SD (2007) Water analysis: emerging contaminants and
current issues. Anal Chem 79:4295–4324
Richmond R (1993) Coral reefs: present problems and future concerns
resulting from anthropogenic disturbance. Am Zool 33:524–536
Richmond R (1997) Reproduction and recruitment in corals: critical
links in the persistence of reefs. Life and death of coral reefs.
Chapman and Hall, New York, pp 175–197
Rickard DJ, Subramaniam M, Spelsberg TC (1999) Molecular and
cellular mechanism of estrogen action on the skeleton. J Cell
Biochem 75:123–132
Rodil R, Quintana JB, Concha-Grana E, Lopex-Mahia P, Muniatequi-
Lorenzo S, Prada-Rodriguez D (2012) Emerging pollutants in
sewage, surface and drinking water in Galicia (NW Spain).
Chemosphere 86:1040–1049
Rogers CS, Miller J (2006) Permanent ‘phase shifts’ or reversible
declines in coral cover? Lack of recovery of two coral reefs in St.
John, US Virgin Islands. Mar Ecol Prog Ser 306:103–114
Samara P, Syntichaki P, Tavernarakis N (2008) Autophagy is required
for necrotic cell death in Caenorhabditis elegans. Cell Death
Differ 15:105–112
Schlecht C, Klammer H, Wuttke W (2004) Effects of estradiol,
benzophenone-2 and benzophenone-3 on the expression pattern
of the estrogen receptors (ER) alpha and beta, the estrogen
receptor-related receptor 1 (ERR1) and the aryl hydrocarbon
receptor (Ahr) in adult ovariectomized rats. Toxicology
Schlenk D, Sapozhnikova Y, Irwin MA et al (2005) In vivo bioassay-
guided fractionation of marine sediment extracts from the
southern California bight, USA, for estrogenic activity. Environ
Toxicol Chem 24:2820–2826
Schlumpf M, Durrer S, Fass O et al (2008) Developmental toxicity of
UV filters and environmental exposure: a review. Int J Androl
Scholze M, Boedeker W, Faust M, Backhaus T, Altenburger R,
Grimme LH (2001) A general best-fit method for concentration-
response curves and the estimation of low-effect concentrations.
Environ Toxicol Chem 20:448–457
´-Wuttke D, Jarry H, Wuttke W (2004) Pures estrogenic effect
of benzophenone-2 (BP-2) but not of bisphenol A (BPA) and
dibutylphtalate (DBP) in uterus, vagina and bone. Toxicology
Shaath NA, Shaath M (2005) Recent sunscreen market trends. In:
Shaath NA (ed) Sunscreens, regulations and commercial devel-
opment, 3rd edn. Taylor & Francis, Boca Raton, pp 929–940
Shlesinger Y, Loya Y (1985) Coral community reproductive patterns:
red sea versus the great barrier reef. Science 228:1333–1335
Simmons DJ (1966) Collagen formation and endochondral ossifica-
tion in estrogen treated mice. Proc Soc Exp Biol Med
Smith TB, Nemeth RS, Blondeau J, Calnan JM, Kadison E, Herzlieb
S (2008) Assessing coral reef health across onshore to offshore
stress gradients in the US Virgin Islands. Mar Pollut Bull
Steel RGD (1959) A multiple comparison rank sum test: treatments
versus control. Biometrics 15:560–572
Suter GW II (2007) Ecological risk assessment. CRC Press, Boca
Suzuki T, Kitamura S, Khota R, Sugihara K, Fujimoto N, Ohta S
(2005) Estrogenic and anti-androgenic activities of 17 ben-
zophenone derivatives used as UV stabilizers and sunscreens.
Toxicol Appl Pharmacol 203:9–17
Taatjes DJ, Sobel BE, Budd RC (2008) Morphological and cyto-
chemical determination of cell death by apoptosis. Histochem
Cell Biol 129:33–43
Takemoto K, Yamazaki H, Nakajima M, Yokoi T (2002) Genotoxic
activation of benzophenone and its two metabolites by human
cytochrome P450s in SOS/umu assay. Mutat Res 519:199–204
Tasdemir E, Galluzzi L, Majuri MN et al (2008) Methods for
assessing autophagy and autophagic cell death. Methods Mol
Biol 445:29–76
Tashiro Y, Kameda Y (2013) Concentration of organic sun-blocking
agents in seawater of beaches and coral reefs of Okinawa Island,
Japan. Mar Pollut Bull 77:333–340
Tsujimoto Y, Shimizu S (2005) Another way to die: autophagic
programmed cell death. Cell Death Differ 15:1528–1534
Tsukamoto Y, Ishihara Y, Miyagawa-Tomita S, Hagiwara H (2004)
Inhibition of ossification in vivo and differentiation of
osteoblasts in vitro by tributyltin. Biochem Pharmacol
UNWTO (United Nations World Tourism Organization) website
(2007) Accessed 30 Jun 2007
US EPA (2012) Sunscreen use.
models/SunscreenUse.html. Accessed 28 July 2014
U.S. National Park Service (2012)
Arch Environ Contam Toxicol
¨nen HK, Ha
¨nen PL (1996) Estrogen and bone metabolism.
Maturitas 23:S65–S69
van Woesik R, Franklin EC, O’Leary J, McClanahan TR, Klaus JS
et al (2012) Hosts of the Plio-Pleistocene past reflect modern-day
coral vulnerability. Proc R Soc Lond B Biol Sci 279:2448–2456
Vione D, Caringella R, De Laurentiis E, Pazzi M, Minero C (2013)
Phtotransformation of the sunlight filter benzophenone-3 (2-
hydroxy-4-methoxybenzophenone) under conditions relevant to
surface waters. Sci Total Environ 463–464:243–251
Vom Saal F, Nagel S, Palanza P et al (1995) Estrogenic pesticides:
binding relative to estradiol in MCF-7 cells and effects of
exposure during fetal life on subsequent territorial behavior in
male mice. Toxicol Lett 77:343–350
West JM, Salm RV (2003) Resistance and resilience to coral
bleaching: implications for coral reef conservation and manage-
ment. Conserv Biol 17:956–967
White MK, Cinti C (2004) A morpholic approach to detect apoptosis
based on electron microscopy. Methods Mol Biol 285:105–111
Wilkinson C (2008) Status of coral reefs of the world. Global Coral
Reef Monitoring Network and Reef and Rainforest Research
Centre, Townsville
Williams DE, Miller MW, Kramer KL (2008) Recruitment failure in
Florida Keys Acropora palmata, a threatened Caribbean coral.
Coral Reefs 27:697–705
Xcaret Ecopark (2007) Home page.
faqs.php. Accessed 2 Dec 2013
´Ecopark (2007) Home page. Accessed
2 Dec 2013
Yla-Antilla P, Vihinen H, Jokitalo E, Eskelinin EL (2009) Monitoring
autophagy by electron microscopy in mammalian cells. Methods
Enzymol 452:143–164
Yu H (2002) Environmental carcinogenic polycyclic aromatic
hydrocarbons: photochemistry and phototoxicity. J Environ Sci
Health, Part C 20:149–183
Zar JH (1996) Biostatistical analysis, 3rd edn. Prentice Hall, New
Zeiger E, Anderson B, Haworth S, Lawlow T, Mortlemans K, Speck
W (1987) Salmonella mutagenicity Tests: 3. Results from the
testing of 255 chemicals. Environ Mutagen 9:1–110
Zhao H, Wei D, Li M, Du Y (2013) Substituent contribution to the
genotoxicity of benzophenone-type UV filters. Ecotoxicol Envi-
ron Saf 95:241–246
Ziolkowaska A, Rucinski M, Pucker A et al (2006) Expression of
osteoblast marker genes in rat calvarial osteoblast-like cells, and
effects of the endocrine disruptors diphenylolpropane, ben-
zophenone-3, resveratrol and silymarin. Chem-Biol Interact
Arch Environ Contam Toxicol
... As a leachable compound from plastics, phthalates are one of the most frequently detected persistent organic pollutants in the environment (Gao and Wen, 2016). In addition to plastics, oxybenzone is becoming an increasingly significant environmental concern for coral reefs because it has been detected in coastal surface waters and sediments, and has been shown to influence larval development (Downs et al., 2016;Mitchelmore et al., 2019). Steroid sex hormones, phthalates, and oxybenzone are all recognized as EDCs and all have been detected in sewage effluent and coastal marine environments where anthropogenic activity is prevalent Thibaut and Porte, 2004;Singh et al., 2010;Bargar et al., 2013;Tsui et al., 2014;French et al., 2015;Thorn et al., 2015;Al-Jandal et al., 2018;LaPlante et al., 2018;Matouskova et al., 2020). ...
... It has been proposed that lipophilic compounds will easily diffuse into cnidarian tissues and bioaccumulate (Peters et al., 1997;Imbs, 2013;Ko et al., 2014). Using a variety of mass-spectrometry assays, steroid hormones (and/or their conjugates), phthalates, and oxybenzone have been detected in coral tissues (from <10 ng up to 650 ng g −1 dry weight) and/or the environment (from <10 ng L −1 to >100 μg L −1 ), (Solbakken et al., 1985;Tarrant et al., 2003;Twan et al., 2003;Twan et al., 2006;Downs et al., 2016;Mitchelmore et al., 2019). Cnidarians have demonstrated the ability to uptake estrogens in the water column at concentrations as low as 300 pg/L (Tarrant et al., 2001). ...
... Acetone has previously been used as a wetting agent (Morgan et al., 2001;Morgan et al., 2012). The 20 ppb concentration was chosen to induce responses and does not necessarily reflect environmental conditions even though BP-3 has been detected at concentrations >20 ppb (Downs et al., 2016) and cholesterol at 11ppb (French et al., 2015). All treatments were nominal concentrations for 4 h in 1L artificial seawater under ambient laboratory lighting during the early spring season. ...
Full-text available
Endocrine disruption is suspected in cnidarians, but questions remain how occurs. Steroid sex hormones are detected in corals and sea anemones even though these animals do not have estrogen receptors and their repertoire of steroidogenic enzymes appears to be incomplete. Pathways associated with sex hormone biosynthesis and sterol signaling are an understudied area in cnidarian biology. The objective of this study was to identify a suite of genes that can be linked to exposure of endocrine disruptors. Exaiptasia diaphana were exposed to nominal 20ppb concentrations of estradiol (E2), testosterone (T), cholesterol, oxybenzone (BP-3), or benzyl butyl phthalate (BBP) for 4 h. Eleven genes of interest (GOIs) were chosen from a previously generated EST library. The GOIs are 17β-hydroxysteroid dehydrogenases type 14 (17β HSD14) and type 12 (17β HSD12), Niemann-Pick C type 2 (NPC2), Equistatin (EI), Complement component C3 (C3), Cathepsin L (CTSL), Patched domain-containing protein 3 (PTCH3), Smoothened (SMO), Desert Hedgehog (DHH), Zinc finger protein GLI2 (GLI2), and Vitellogenin (VTG). These GOIs were selected because of functional associations with steroid hormone biosynthesis; cholesterol binding/transport; immunity; phagocytosis; or Hedgehog signaling. Quantitative Real-Time PCR quantified expression of GOIs. In silico modelling utilized protein structures from Protein Data Bank as well as creating protein structures with SWISS-MODEL. Results show transcription of steroidogenic enzymes, and cholesterol binding/transport proteins have similar transcription profiles for E2, T, and cholesterol treatments, but different profiles when BP-3 or BBP is present. C3 expression can differentiate between exposures to BP-3 versus BBP as well as exposure to cholesterol versus sex hormones. In silico modelling revealed all ligands (E2, T, cholesterol, BBP, and BP-3) have favorable binding affinities with 17β HSD14, 17β HSD12, NPC2, SMO, and PTCH proteins. VTG expression was down-regulated in the sterol treatments but up-regulated in BP-3 and BBP treatments. In summary, these eleven GOIs collectively generate unique transcriptional profiles capable of discriminating between the five chemical exposures used in this investigation. This suite of GOIs are candidate biomarkers for detecting transcriptional changes in steroidogenesis, gametogenesis, sterol transport, and Hedgehog signaling. Detection of disruptions in these pathways offers new insight into endocrine disruption in cnidarians.
... Commonly used organic UV filters such as BP3, 4-MBC, OD-PABA and EHMC have been Overall, BP3 is the most studied compound with concentrations varying greatly among biota (Downs et al. 2016;Giokas, Sakkas, and Albanis 2004;Tovar-Sánchez et al. 2013;Tsui et al. 2014). ...
... Interestingly, a discrepancy in the published data was noted as concentrations at the same location varied from the nanogram to the milligram range (Downs et al. 2016). Chemical properties of UV filters make their accurate quantification a challenging task. ...
... Lastly, similar trends were observed for Firmicutes that are often found in the gut of marine invertebrates . Taken together, these results suggest that the studied UV filters might affect the growth of key players within coastal microbial communities where UV filters can reach high concentrations (Downs et al. 2016;Kim and Choi 2014;Tsui et al. 2014). ...
UV filters are the active components of sun protection products and are used as stabilizing agents in cosmetics and industrial products. Several studies reported the worldwide occurrence of UV filters in coastal waters and demonstrated their bioaccumulation and toxic effects on aquatic organisms. Marine bacteria dominate the marine biomass and are crucial for biogeochemical cycles. Notwithstanding their essential roles, the toxicity of UV filters has never been assessed on microorganisms. This thesis aims to explore the effects of the most occurrent organic UV filters, taking into account the physiological, morphological, and molecular response, on environmentally relevant heterotrophic bacteria, isolated from the bay of Banyuls (NW Mediterranean Sea, France). This project demonstrated the species and compound-dependent bactericidal effects of common organic UV filters on marine bacteria and contributed to a better understanding of the cellular response induced by UV filter exposure. We showed, via morphological investigation and quantitative proteomics, that bacteria exhibited a variety of stress-related molecular responses, emphasizing that bacteria are suitable for ecotoxicological studies. Future meta-omic research on natural and artificial communities will enable us to assess the impact of anthropogenic inputs on the functions of microbial communities.
... mineral UV filters and organic UV filters. The use of certain organic filters, such as octocrylene, oxybenzone, and 4-methylbenzylidene camphor, is subject to much controversy, as these filters can represent a danger to the environment [16,27]. It has been demonstrated that these products can cause hormonal effects that affect the fertility and reproduction of fish [8,9]. ...
... Concerning the effects of TiO2 on hard corals, it has been determined that this mineral filter may interfere with the symbiotic relationship between corals and zooxanthellae by reducing algal populations within symbioses, without leading to coral death, at a concentration of 10 mg/L over an exposure period of 17 days[36]. Some xycinnamate, have impacts on corals.Avobenzone can induce a significant decrease in photosynthetic efficiency at a concentration of 1 mg/L[31], and benzophenone 2 and oxybenzone are toxic to corals, causing damage to zooxanthellae and DNA damage to coral cells[27, 37]. Ethylhexylmethoxycinnamate and 4-methylbenzylidene camphor induce rapid bleaching at a concentration of 33 µg/L[11]. ...
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Every second, 0.8 litres of sunscreen enters ocean waters, which corresponds to the release of 25.000 tons per year. UV filters may present substantial threats to marine fauna and flora and have an impact similar to that of other contaminants. Coral reefs play a major role in marine biodiversity, and some publications suggest that they are threatened by the release of sunscreen into the environment, which should cause bleaching. The aim of this study was to evaluate the potential impact of sunscreen products on hard corals. Laboratory experiments in which Seriatopora hystrix coral fragments were exposed to 9 sunscreens at concentrations up to 100 mg/L for 96 hours were conducted, and the biological responses of the fragments were assessed. The examined parameters were coral bleaching and polyp retraction. The results obtained revealed that the 9 tested sunscreens had no effects on S. hystrix, with a recorded NOEC (No Observed Effect Concentration) of 100 mg/L for both tested parameters. This concentration is much higher than those of chemicals in the natural environment, Journal of Environmental Science and Public Health 16 which are on the order of µg/L or ng/L. Under the conditions in this experiment, the absence of toxic effects from the tested sunscreens allows us to argue the absence of potential danger on corals.
... The effects of oxybenzone in vivo are broad: it is toxic to cyanobacteria [14,15], green algae [14,16] and coral [17]. In rodents, oxybenzone alters the development of the mammary gland, alters the weight of the liver, kidney, and reproductive organs, decreases the number of spermatocytes in males [18][19][20][21] and induces DNA damage in the mammary epithelium [22]. ...
... First synthesized in the mid-twentieth century, oxybenzone quickly gained dominance among sunscreen products and equally rapidly began polluting coastal waters when washed off the skin of beachgoers. The levels of oxybenzone measured in coastal waters harboring coral reefs are no longer sufficiently diluted to avoid harm to these species [17,43]. The demand for sunburn-free midday beachgoing, combined with climatic and oceanographic conditions produce the present situation in which oxybenzone threatens the survival of aqueous species and ecosystems [44]. ...
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Background Technological advancements make lives safer and more convenient. Unfortunately, many of these advances come with costs to susceptible individuals and public health, the environment, and other species and ecosystems. Synthetic chemicals in consumer products represent a quintessential example of the complexity of both the benefits and burdens of modern living. How we navigate this complexity is a matter of a society’s values and corresponding principles. Objectives We aimed to develop a series of ethical principles to guide decision-making within the landscape of environmental health, and then apply these principles to a specific environmental chemical, oxybenzone. Oxybenzone is a widely used ultraviolet (UV) filter added to personal care products and other consumer goods to prevent UV damage, but potentially poses harm to humans, wildlife, and ecosystems. It provides an excellent example of a chemical that is widely used for the alleged purpose of protecting human health and product safety, but with costs to human health and the environment that are often ignored by stakeholders. Discussion We propose six ethical principles to guide environmental health decision-making: principles of sustainability, beneficence, non-maleficence, justice, community, and precautionary substitution. We apply these principles to the case of oxybenzone to demonstrate the complex but imperative decision-making required if we are to address the limits of the biosphere’s regenerative rates. We conclude that both ethical and practical considerations should be included in decisions about the commercial, pervasive application of synthetic compounds and that the current flawed practice of cost-benefit analysis be recognized for what it is: a technocratic approach to support corporate interests.
... Although less information is available regarding CECs, some groups have shown adverse effects on marine biota and/or potential to bioaccumulate (e.g. Downs et al., 2016;Gago-Ferrero et al., 2013;He et al., 2019;Langford et al., 2015;Picot-Groz et al., 2018). ...
... Thus, their presence in dolphin tissue from a remote island suggests influence of direct contamination sources related to tourism and diving activities in Fernando de Noronha Archipelago. This is especially concerning since potential adverse effects, including increasing coral bleaching rates and potential to bioaccumulate in marine mammals, have been associated with some UV filters (Gago-Ferrero et al., 2013;Langford et al., 2015;Downs et al., 2016;Picot-Groz et al., 2018). Moreover, the presence of contaminants represents an additional stressor to these marine mammals, for which a number of negative impacts are thought to result from tourism and diving activities (Curtin and Garrod, 2008). ...
The presence and distribution of persistent organic pollutants (POPs) and contaminants of emerging concern (CECs) were evaluated in spinner dolphins (Stenella longirostris) from the Fernando de Noronha Archipelago (Western Atlantic Ocean). Blubber samples (n = 37) were Soxhlet extracted and analyzed using gas chromatography tandem mass spectrometry. The levels of POPs reported in this study are far below those previously reported in spinner dolphins from the Indian and Pacific Oceans. Despite relatively low levels of contaminants, the presence of chemicals represents an additional stressor to these marine mammals, which are subject to increasing anthropogenic pressures, especially regarding tourism activities, in Fernando de Noronha.
... Within the genus Porites in particular, it is difficult to determine a genetic difference in species with clearly distinct morphological differences (Forsman et al. 2015). 2015). A correlation between colony size and brooding has been documented in the Atlantic, with brooding coral colony size being smaller than spawning corals occupying the same watershed (Szmant 1986 that are located in cleaner water. ...
... Furthermore, promiscuous larval settlement preferences may explain the broad population distribution of this species in the area.Materials and Methods:Colonies of Leptastrea purpurea maintained in an open seawater system at the Kewalo Marine Laboratory (Honolulu, Hawaiʻi) released brooded planula larvae on a continuous basis. These colonies collected from Honolulu Harbor were monitored in an open seawater system over four years(2013)(2014)(2015)(2016) and were found to release small quantities of larvae averaging 1-2 per day throughout the year. However, distinct brooding peaks were observed annually from September-October, in which an individual colony increased its brooding output to several hundred planula larvae per day (data not shown). ...
Coral reefs are in decline globally as anthropogenic induced climate change effects ravage our oceans. It is estimated that at least 50% of coral reefs have disappeared over the last 40 years and are declining at an alarming rate. While much scientific research focuses on healthy coral reef ecosystems, data suggests that degraded watersheds with high levels of selective pressure may harbor coral species that are well adapted to stress. A thriving population of corals exists in Honolulu Harbor, a highly degraded ocean habitat exposed to multiple anthropogenic stressors. Following the massive molasses spill in 2013, two species of corals have shown remarkable resilience to multiple stressors. Both species were observed to be brooders, with Leptastrea purpurea demonstrating a larval peak in the late summer. L. purpurea planula larvae are induced by a settlement cue originating from other coral colonies. When a coral scent is present, settlement rates are as high as 90-100% on biofilm and other substrates, including plain untreated glass. Field surveys reveal that L. purpurea colonies are found on average 18mm in distance from their nearest neighbor, and modeling suggests a non-random distribution of colonies at our survey sites. A second species, Harbor Porites, is genetically distinct from Porites lobata, though genetics show a similarity in origin. Harbor Porites larvae will settle in the presence of a biofilm cue, and both larvae and recruits show remarkable resilience to multiple chemical and physical stressors, as well as the ability to undergo reversible metamorphosis. Both coral species are in high abundance inside Honolulu harbor, and coral surveys reveal that the two species are found within an average 16mm of each other. To elucidate the high level of survival in these two species, a thermal tolerance exposure was performed to induce a bleaching response in both species. Molecular biomarkers were used to quantify relative stress levels. Molecular expression analyses could give us insights into how these corals are responding to stress, and if the basis for their resilience is tied to up-regulated molecular processes. While corals continue to face stress as a result of climate change, these two harbor coral species serve as excellent models for studying the resilience of corals to stress. Their persistence in a stressful environment makes them candidate species for coral reef restoration.
... [13,14] BP-3 also interfered with embryonic development and the reproduction of aquatic invertebrates, such as Chironomus riparius, [15,16] being harmful to corals by increasing bleaching rates. [17,18] Hematological parameters have been considered good indicators of the physiological changes and health status in fish, [19][20][21] being used as biomarkers for identifying effects of substances in aquatic environments. The occurrence of micronuclei and other nuclear abnormalities of erythrocytes have also been used as effective parameter to measure chromosomal damage in fish after exposure to contaminants. ...
... [13,14] BP-3 also interfered with embryonic development and the reproduction of aquatic invertebrates, such as Chironomus riparius, [15,16] being harmful to corals by increasing bleaching rates. [17,18] Hematological parameters have been considered good indicators of the physiological changes and health status in fish, [19][20][21] being used as biomarkers for identifying effects of substances in aquatic environments. The occurrence of micronuclei and other nuclear abnormalities of erythrocytes have also been used as effective parameter to measure chromosomal damage in fish after exposure to contaminants. ...
Benzophenone-3 (BP-3) is a common component of organic sunscreen widely used that can affect especially aquatic ecosystems health, including fish. To verify the biological effects of low concentrations of BP-3 on blood cells, one hundred and forty zebrafish (D. rerio) were used and then randomly divided into five groups: control group (water), solvent group (alcoholic water), and BP-3 group (BP-3 at 7 mg L À1 , BP-3 at 70 mg L À1 , and BP-3 at 700 mg L À1). The blood slices were stained with Panoptic stain and with Giemsa solution for the hematological analysis. During the exposure to BP-3, no behavioral changes were observed. Although no significant difference in total leuko-cytes occurred, an increase in neutrophils and a reduction of lymphocytes at the highest concentration on both 7th and 14th days were detected. The total and cytoplasmic area of erythrocytes on the 7th day at the highest concentration were reduced. In addition, alterations on the erythro-cyte nuclear morphology in fish exposed to BP-3 were usually visualized, mainly when considered the occurrence of blebbed nucleus and micronucleus, indicating that BP-3 exhibits cytotoxic and mutagenic effects. The results indicate that BP-3 can interfere with the morphophysiology of aquatic organisms. ARTICLE HISTORY
... Currently, several sunscreen formulations are on the market, with active ingredients ranging from inorganic UV filters such as titanium dioxide (TiO 2 ) and zinc oxide (ZnO) to organic UV filters, examples of which include avobenzone, oxybenzone, homosalate, and octocrylene, among many others [3,6]. Despite the many sunscreen formulations currently available, certain setbacks such as the potential toxicity to humans and the environment, as well as photo-instability, have resulted in the banning of some UV filters [7][8][9][10]. To this effect, sunscreen scientists have continued to search for safe and efficient UV filters, seeking inspiration from natural sources, including plants and microorganisms [11]. ...
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Plants, as with humans, require photoprotection against the potentially damaging effects of overexposure to ultraviolet (UV) radiation. Previously, sinapoyl malate (SM) was identified as the photoprotective agent in thale cress. Here, we seek to identify the photoprotective agent in a similar plant, garden cress, which is currently used in the skincare product Detoxophane nc. To achieve this, we explore the photodynamics of both the garden cress sprout extract and Detoxophane nc with femtosecond transient electronic absorption spectroscopy. With the assistance of liquid chromatography-mass spectrometry, we determine that the main UV-absorbing compound in garden cress sprout extract is SM. Importantly, our studies reveal that the photoprotection properties of the SM in the garden cress sprout extract present in Detoxophane nc are not compromised by the formulation environment. The result suggests that Detoxophane nc containing the garden cress sprout extract may offer additional photoprotection to the end user in the form of a UV filter booster.
The coastal marine ecosystems of Vietnam are one of the global biodiversity hotspots, but the biodiversity of marine fungi is not well known. To fill this major gap of knowledge, we assessed the genetic diversity (ITS sequence) of 75 fungal strains isolated from 11 surface coastal marine and deeper waters in Nha Trang Bay and Van Phong Bay using a culture-dependent approach and 5 OTUs (Operational Taxonomic Units) of fungi in three representative sampling sites using next-generation sequencing. The results from both approaches shared similar fungal taxonomy to the most abundant phylum (Ascomycota), genera (Candida and Aspergillus) and species (Candida blankii) but were different at less common taxa. Culturable fungal strains in this study belong to 3 phyla, 5 subdivisions, 7 classes, 12 orders, 17 families, 22 genera and at least 40 species, of which 29 species have been identified and several species are likely novel. Among identified species, 12 and 28 are new records in global and Vietnamese marine areas, respectively. The analysis of enzyme activity and the checklist of trophic mode and guild assignment provided valuable additional biological information and suggested the ecological function of planktonic fungi in the marine food web. This is the largest dataset of marine fungal biodiversity on morphology, phylogeny and enzyme activity in the tropical coastal ecosystems of Vietnam and Southeast Asia. Biogeographic aspects, ecological factors and human impact may structure mycoplankton communities in such aquatic habitats.