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Marine Pollution Bulletin 57 (2008) 493–502
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Marine Pollution Bulletin
journal homepage: www.elsevier .com/ locate /marpolbul
Ecotoxicological evaluation of polycyclic aromatic hydrocarbons using marineinvertebrate embryo–larval bioassays
Juan Bellas *, Liliana Saco-Álvarez, Óscar Nieto, Ricardo BeirasDepartamento de Ecoloxía e Bioloxía Animal, Universidade de Vigo, Estrada Colexio Universitario s/n, 36310 Vigo, Galicia, Spain
a r t i c l e i n f o
Keywords:
Polycyclic aromatic hydrocarbonsMarine invertebratesEmbryo–larval bioassaysPhototoxicityMixture toxicity0025-326X/$ - see front matter � 2008 Elsevier Ltd. Adoi:10.1016/j.marpolbul.2008.02.039
* Corresponding author. Tel.: +34 986 814087; fax:E-mail address: [email protected] (J. Bellas).
a b s t r a c t
The toxicity of polycyclic aromatic hydrocarbons (PAHs) was determined using mussel, sea-urchin andascidian embryo–larval bioassays. Fluorescent light exposure enhanced phenanthrene, fluoranthene, pyr-ene and hydroxypyrene toxicity in comparison with dark conditions, but not naphthalene and fluorenetoxicity. The toxicity of PAHs was inversely related to their KOW values following QSAR models derivedfor baseline toxicity of general narcotics, whereas the obtained regression using toxicity data from photo-activated PAHs significantly departed from the general narcosis model. Also, the mixture toxicity of fivePAHs to the larval growth of the sea-urchin was compared with predictions derived from the concentra-tion addition concept, indicating less than additive effects. Finally, we compared our toxicity data withworst-case environmental concentrations in order to provide a preliminary estimate of the risk to themarine environment. Naphthalene, fluorene and pyrene are not considered to pose a risk to sea-urchin,mussel or ascidian larvae, whilst phenanthrene and fluoranthene may pose a risk for mussel and sea-urchin. Moreover, a higher risk for those species is expected when we consider the photoactivation ofthe PAHs.
� 2008 Elsevier Ltd. All rights reserved.
1. Introduction
Polycyclic aromatic hydrocarbons (PAHs) are a large group ofwidespread organic compounds of high environmental concern.Even though PAHs occur naturally, the highest concentrations aremainly due to human activities that cause a continuous increasein PAH levels of estuarine and marine waters (Kennish, 1992;Walker et al., 2001). Direct discharges into the marine environ-ment from point sources such as wastewater treatment plantsrange from <1 lg/L to over 625 lg/L, whilst concentrations of PAHsin industrial effluents range from undetectable to 4.4 mg/L (Lati-mer and Zheng, 2003). Major sources of PAHs to the marine envi-ronment are combustion products and petroleum principallyfrom atmospheric deposition (5 � 104 tonnes/year) and oil spillage(1.7 � 105 tonnes/year) (Kennish, 1992; Meador, 2003). For in-stance, in November 2002 the oil tanker Prestige sank 130 milesoff the Galician coast (NW Spain), spilling more than 60,000 tonnesof number 2 fuel oil into the open sea (CEDRE, 2007). This is not anisolated disaster, since Galicia has received nine oil spills duringthe last 50 years (CEDRE, 2007), being one of the regions withthe highest number of oil spills in the world. Galician Rías arehighly productive estuaries well known because of its coastal fish-
ll rights reserved.
+34 986 812556.
ing and shellfish production. Among the shellfish production of theGalician Rías stands out the culture of the mussel (Mytilus gallopro-vincialis) as the main cultivated species, with an annual productionof more than 2.5 � 105 tonnes (1.3 � 108 euros) (Pérez-Camachoet al., 1995; Labarta and Corbacho, 2002). Therefore, it is necessaryto study the impact of those pollutants to marine organisms of eco-logical and commercial relevance. Furthermore, growing evidencesuggests that solar radiation may enhance the toxicity of certainPAHs. Some PAHs can absorb the radiation and become photoacti-vated, the photoactivated molecule may then transfer the radiationenergy to molecular oxygen thereby forming reactive superoxideanions capable of oxidative damage in the organisms (Landrumet al., 1986; Arfsten et al., 1996).
The present study focussed on the toxicity assessment of lowand intermediate molecular weight PAHs (2–4 rings), since theyconstitute the most toxic components of oil for the marine biota,using a battery of sensitive embryo–larval bioassays with marineorganisms of commercial and ecological relevance. The specificgoals of the present study were to investigate the toxicity and po-tential phototoxicity (under fluorescent light and in dark condi-tions) of selected PAHs individually and in mixtures, based oncurrent models of narcotic toxicity. Bioassays with marine organ-isms intended to be used in the assessment of marine pollutionmust meet some fundamental requirements: (i) ecological rele-vance, using ecologically relevant or commercial organisms; (ii)
494 J. Bellas et al. / Marine Pollution Bulletin 57 (2008) 493–502
feasibility, the bioassays must be easy to standardise, based on pre-cisely defined protocols, and using simple, rapid and cost-effectivebioassays; and (iii) sensitivity, using for instance, early life stagesof development, which are less tolerant to toxicants than adults(e.g. Connor, 1972; Marin et al., 1991; Ringwood, 1991; His et al.,1999), and sublethal biological responses. We conducted em-bryo-larval bioassays with the bivalve M. galloprovincialis, the echi-noid Paracentrotus lividus and the ascidian Ciona intestinalis. Thosespecies were chosen due to their commercial importance, theirabundance and their importance in the functioning of the marineecosystem (Dybern, 1965; Gosling, 1992; Boudouresque and Verl-aque, 2001).
2. Materials and methods
2.1. Biological material
Mature C. intestinalis and P. lividus were collected in pristinesites from local populations in the Ría de Vigo (Galicia, NW Spain).Mature M. galloprovincialis were purchased at the local market inVigo. Animals were transported to the laboratory in a portable ice-box and maintained in aquaria with running natural seawater untilthe experiments for a least one week. Handling conditions of theadult stock were 17.43 ± 0.54 �C temperature, 35.01 ± 1.52 pptsalinity, 7.24 ± 1.14 mg l�1 O2 and 7.88 ± 0.09 pH (mean ± std).
2.2. Experimental solutions
Stock solutions were made by dissolving analytical grade naph-thalene, phenanthrene, fluoranthene, fluorene, pyrene and hydrox-ypyrene (Sigma–Aldrich, Steinheim) in acetone, due to the lowsolubility of PAHs in seawater (Kennish, 1992). The experimentalconcentrations were obtained by diluting the stock solutions inartificial seawater (ASW) prepared as in Zaroogian et al. (1969).During this dilution, equal amounts of acetone (less than200 ll L�1), found not to be toxic in preliminary tests, were addedto each experimental beaker with PAHs solutions. All glasswarewas acid-washed (HNO3 10% vol.) and rinsed with acetone and dis-tilled water before the experiments.
Experimental concentrations were chosen on the basis of range-finding trials and on data from the literature. Tested concentrationsfor each compound were below their water-saturation levels (Nag-pal, 1993). Incubations were made in 25 ml glass vials with airtightTeflon-lined screw caps, to avoid losses of the tested compounds.All glassware was acid-washed (HNO3 10% vol.) and rinsed withacetone and distilled water before the experiments. Physico-chem-ical conditions of the experiments were 34.20 ± 0.15 ppt salinity,7.32 ± 0.70 mg l�1 O2 and 8.29 ± 0.11 pH (mean ± std, n = 15).
2.3. Chemical analysis
Test solutions intended for chemical analysis were collectedfrom the experimental vials at the beginning and end (48 h) ofthe tests. Chemical analyses were conducted with two (high andlow) concentrations for each compound. The test solutions werepoured into a separatory funnel and PAHs were extracted withdichloromethane (USEPA, 1980). After substitution of the solventby acetonitrile, the concentration of the PAHs was determined byHPLC with fluorimetric detection. Twenty microlitre of samplewere injected into the chromatographic column and a gradient elu-tion was performed by using water and methanol as eluents (Lópezet al., 1996; Viñas, 2002). The fluorimetric detection was carriedout by programming the specific excitation and emission wave-lengths of each PAH analysed. The recoveries in the extractionmethod were about 90% for the measured compounds.
2.4. Experimental procedure
The method used in the M. galloprovincialis test has been previ-ously described by Bellas et al. (2005). Mature M. galloprovincialiswere induced to spawn by thermal stimulation in separated beak-ers with 0.2 lm filtered seawater. Eggs from a single female weretransferred to a 100 ml measuring cylinder and their quality waschecked under microscope. Sperm solution was stored at 4 �C untiluse. A few microlitre of motile sperm were added to the egg sus-pension and carefully stirred to allow fertilization. Fertilized eggs(ca. 30 eggs/ml) were transferred to vials containing the experi-mental solutions that were incubated at 18 �C for 48 h, until thesecond larval stage (D-veliger), characterized by a straight dorsalhinge which gives the larva the shape of a capital letter D, wasattained.
P. lividus gametes were obtained by dissection from a single pairof adults according to the methods described by (Beiras and Saco-Álvarez, 2006). Approximately 400 fertilized eggs were deliveredinto vials with the experimental solutions. The vials were incu-bated at 18 �C until larvae reached the pluteus stage (approxi-mately 48 h after fertilization), which in this species developsfour arms covered by cilia and supported by calcareous skeletalrods.
To test the effects of the selected PAHs on the embryonic devel-opment of C. intestinalis, 250 fertilized eggs obtained by in vitro fer-tilization following the methods of Bellas et al. (2003), weredelivered into experimental vials containing 20 ml of the studiedcompounds. These vials were incubated in a culture chamber at18 �C until the tadpole larvae stage is reached (20 h after fertiliza-tion). The tadpole larvae consists of a trunk which contain the sen-sory organs (otolith, ocellus and adhesive papillae) and a tail whichcontains the notochord.
Concurrently with the experiments conducted in the dark, wecarried out bioassays under fluorescent light exposure with mus-sel, sea-urchin and ascidian larvae, in order to study the potentialphotoinduced toxicity of PAHs. One set of vials was incubated indarkness and the other set under fluorescent light with a 14:10 hlight:dark photoperiod. Cool daylight lamps (Osram L15W/765)were used in order to simulate natural irradiation (emission spec-trum range: 380–780 nm; Photosynthetically Active Radiation:70 lE m�2 s�1).
The combined toxicity of the five selected PAHs to P. lividus em-bryos was determined according to the concentration addition (CA)model (Loewe and Muischnek, 1926). CA assumes that all compo-nents in the mixture are similarly acting substances sharing iden-tical mechanism of action in the exposed organism. An equitoxicmixture of the studied compounds was used i.e. the ratio of theconcentrations of the individual mixture components was keptconstant while the total concentration of the mixture was varied.CA can be formulated as
Xn
i¼1
ci
ECxi
¼ 1;
where n is the number of mixture components, ci is the concentra-tion of the ith component of the mixture, ECxi is the concentrationof the mixture component that induce a x% effect when applied sin-gly. The quotient ci/ECxi represent the concentration of the mixturecomponent scaled for its relative toxicity and is usually termed theToxic Unit (TU) of that component. According to CA, each individualcompound in the equitoxic mixture contributes equally to the totaltoxicity. Due to solubility constraints, EC20 rather than EC50 valueswere used to test the mixtures.
After the incubation period mussel, sea-urchin and ascidian lar-vae were preserved by adding a few drops of 40% buffered forma-lin, and the percentage of D-veliger and normal tadpole larvae
J. Bellas et al. / Marine Pollution Bulletin 57 (2008) 493–502 495
(n = 100), and the mean larval length of the pluteus larvae (n = 35)were recorded. Four replicates per treatment, four ASW controls,and four acetone controls were assayed for each experiment. Con-trol embryogenesis success was always above 90% for sea-urchinand above 80% for mussel and ascidian.
2.5. Statistical analyses
Statistical analyses were conducted using the SPSS� version14.0 statistical software. Differences between treatments weretested for significance by means of one-way analysis of variance(ANOVA). When differences among groups were significant theDunnett’s test was employed to compare the control group andeach of the experimental groups for calculation of the Lowest Ob-served Effect Concentrations (LOEC). The EC10, EC20 and EC50 and
Table 1Nominal and measured concentrations (lg/L) of the tested compounds in theexperimental vessels in freshly made solutions (t0) and at 48 h (t48)
Nominal Measured
t0 t48
Naphthalene 12.8 11.0 9.96562 4523 4140
Phenanthrene 54 70.3 37.6214 243.2 118.9
Pyrene 16 12.6 8.665 90.7 41.3
Fluoranthene 32 41.6 13.1127 107.1 98.7
Fluorene 100 140.0 97.4400 325.4 267.9
Naphthalene [ M]
1 10 100
% D
-vel
iger
0
20
40
60
80
100
Pyrene [nM]
10 100 1000
% D
-vel
iger
0
20
40
60
80
100
μ
Fig. 1. Percentage of D-veliger larvae after 48 h exposure of M. galloprovincialis fertilizefluoranthene. Filled-in squares represent experiments carried out in the dark. Open cirepresent standard deviations.
their 95% confidence intervals (95CI) were calculated accordingto the Probit method after normalizing data to the mean control re-sponse using Abbot’s formula (Emmens, 1948). For analysis, datawere first arcsine-transformed to achieve normality (Hayes,1991). The results of the experiments conducted in the dark andunder fluorescent light were analyzed using two-way ANOVA. Also,significant differences between parameters for pairs of curves fromthe dark and light exposure experiments were tested using an ex-tra sum-of-squares F-test to determine if curves were statisticallyindistinguishable (Motulsky and Christopoulos, 2004).
3. Results
3.1. Chemical analysis
Results from the analytical chemistry applied for checking thenominal PAH concentrations in seawater are shown in Table 1.Measured concentrations were within 15–30% of the nominal con-centrations, except for the highest concentration of pyrene and thelowest concentration of fluorene analyzed, which were 40% abovethe nominal concentrations. In general, measured values averaged25% of the nominal ones. Concentrations decreased by 8–54% after48 h incubation in the experimental vials, except for the lowestconcentration of fluoranthene measured which decreased by 69%.In average, PAHs concentrations decreased by 33% after 48 h.
3.2. Mytilus galloprovincialis
Naphthalene inhibited the embryonic development of M. gallo-provincialis according to a sigmoidal toxicity curve (Fig. 1), withEC10 and EC50 values of 31.5 and 51.7 lM in dark conditions and
Phenanthrene [nM]
100 1000
% D
-vel
iger
0
20
40
60
80
100
Fluoranthene [nM]
100 1000
% D
-vel
iger
0
20
40
60
80
100
d eggs to different concentrations (nM) of naphthalene, phenanthrene, pyrene andrcles represent experiments carried out under fluorescent light (n = 4). Error bars
Tabl
e2
LOEC
,EC 1
0an
dEC
50
for
naph
thal
ene
(lM
),ph
enan
thre
ne,fl
uora
nthe
ne,p
yren
e,flu
oren
ean
dhy
drox
ypyr
ene
(nM
)
Mu
ssel
Sea-
urc
hin
Asc
idia
n
EC1
0EC
50
LOEC
EC1
0EC
50
LOEC
EC1
0EC
50
LOEC
Nap
hth
alen
eD
ark
31.5
(29.
4–33
.3)
51.7
(49.
8–53
.7)
16a
5.06
(2.9
7–8.
59)
37.3
(29.
6–47
.1)
7.41
4.76
(2.3
6–7.
09)
15.2
(11.
0–21
.0)
25.6
a
Ligh
t64
.3(6
1.5–
65.8
)77
.4(7
3.2–
91.8
)66
5.78
(4.0
7–8.
20)
34.0
(29.
4–39
.4)
7.41
23.6
(nc)
33.4
(nc)
102.
4Ph
enan
thre
ne
Dar
k16
5(n
c)80
9(4
88–1
262)
1200
2582
(103
0–64
30)
>240
071
80>2
400
>240
0>2
400
Ligh
t29
6(n
c)12
58(9
73–1
677)
1200
590
(320
–108
0)>2
400
1800
a14
72(1
034–
1725
)23
47(2
056–
2874
)24
00a
Pyre
ne
Dar
k46
5(3
29–6
58)
>640
640
342
(260
–450
)>6
4059
0>6
40>6
40>6
40Li
ght
40.8
(14.
7–69
.6)
657
(417
–149
7)32
0a11
6(9
0–15
0)42
7(3
80–4
80)
150a
>640
>640
>640
Flu
oran
then
eD
ark
>125
0>1
250
>125
013
2(9
0–20
0)>1
250
170
>125
0>1
250
>125
0Li
ght
166
(139
–188
)26
3(2
38–2
89)
312a
103
(80–
140)
239
(210
–270
)17
0a11
96(n
c)>1
250
1250
Flu
oren
eD
ark
nm
nm
nm
2960
(234
0–37
10)
>119
0059
60n
mn
mn
mLi
ght
nm
nm
nm
2960
(211
0–41
40)
>119
0059
60n
mn
mn
mH
ydro
xypy
ren
eD
ark
nm
nm
nm
nm
nm
nm
>800
>800
>400
Ligh
tn
mn
mn
mn
mn
mn
m29
3(2
26–3
38)
434
(386
–487
)40
0a
The
95%
con
fide
nce
inte
rval
s(9
5CI)
are
give
nin
brac
kets
.n
cn
otca
lcu
late
d.n
mn
otm
easu
red.
aD
enot
essi
gnifi
can
tdi
ffer
ence
sbe
twee
nda
rkan
dli
ght
trea
tmen
ts(s
eete
xtfo
rde
tail
s).
496 J. Bellas et al. / Marine Pollution Bulletin 57 (2008) 493–502
64.3 and 77.4 lM under fluorescent light, respectively (Table 2).Exposure to fluorescent light decreased the toxic effects of naph-thalene on embryos of the mussel M. galloprovincialis. Phenan-threne showed also toxicity to mussel embryos at theconcentrations tested here both in dark conditions (EC10 = 165,EC50 = 809 nM) and under fluorescent light (EC10 = 296,EC50 = 1258 nM). Pyrene resulted toxic to mussel embryos onlyat the highest tested concentration (640 nM) in dark conditions,reducing the percentage of D-veliger larvae by 20% with respectto controls and yielding an EC10 value of 465 nM. CalculatedEC10 and EC50 under fluorescent light were 40.8 and 657 nM. Onthe other hand, fluoranthene was not toxic at the tested concen-trations in dark conditions but toxicity was greatly increased un-der fluorescent light with EC10 and EC50 values of 166 and263 nM.
Significant differences between experiments carried out in thedark and under fluorescent light were observed for naphthalene,pyrene and fluoranthene. Naphthalene toxicity was significantlyhigher in the dark than under fluorescent light (p < 0.0001) asshown by the two-way ANOVA, although no significant differ-ences between the EC50s were detected with the F-test. Both pyr-ene and fluoranthene were significantly more toxic underfluorescent light (p < 0.01 and p < 0.0001), but significant differ-ences between EC50s were found only for fluoranthene. On thecontrary, no significant differences in phenanthrene toxicity weredetected in the dark and under fluorescent light exposure.
3.3. Paracentrotus lividus
For all tested PAHs P. lividus larval growth was reduced bothunder dark and fluorescent light exposures (Fig. 2). The EC10
and EC50 for naphthalene were 5.1 and 37 lM in darkness and5.8 and 34 lM under fluorescent light exposure (Table 2). Phen-anthrene significantly affected sea-urchin larval growth at thehighest tested concentration, although only a 10% decrease wasobserved in dark conditions and a 30% decrease under fluorescentlight. The obtained EC10 were 2582 and 590 nM, respectively. Pyr-ene exposure caused a 20% decrease of larval growth in darknessand a 65% decrease under fluorescent light at the highest concen-trations, with EC10 values of 342 and 116 nM, respectively, and anEC50 value of 427 nM under fluorescent light. Fluoranthene re-duced sea-urchin larval growth by 45% in darkness, yielding anEC10 value of 132 nM, whilst a larval growth reduction of 90%was observed under fluorescent light, with EC10 and EC50 valuesof 103 and 239 nM, respectively. Fluorene caused a 30% and40% reduction of larval growth in dark conditions and under fluo-rescent light, respectively, yielding similar EC10 values (2960 nM).
For sea-urchin, significant differences between experimentsconducted in the dark and under fluorescent light were foundfor phenanthrene (p < 0.0001), pyrene (p < 0.0001) and fluoran-thene (p < 0.0001), indicating a phototoxic response of those com-pounds. Also, the EC50 values of pyrene and fluoranthene weresignificantly higher in the dark than under fluorescent light(p < 0.0001), but not in the case of phenanthrene.
3.4. Ciona intestinalis
Naphthalene was the only tested PAH showing toxicity to C.intestinalis embryos both in dark conditions and under fluorescentlight exposure (Fig. 3). The EC10 and EC50 values were 4.8 and15 nM in darkness and 24 and 33 nM under fluorescent light (Ta-ble 2). Phenanthrene was not toxic to C. intestinalis embryos indarkness, but when incubations were made under fluorescentlight, a 50% decrease on the percentage of normal larvae was ob-served at the highest tested concentration, with EC10 and EC50
values of 1472 and 2347 nM. Pyrene did not show toxicity to C.
Naphthalene [ M]
1 10 100
% L
arva
l gro
wth
0
20
40
60
80
100
Phenanthrene [nM]
100 1000
% L
arva
l gro
wth
0
20
40
60
80
100
Pyrene [nM]
10 100 1000
% L
arva
l gro
wth
0
20
40
60
80
100
Fluoranthene [nM]
100 1000
% L
arva
l gro
wth
0
20
40
60
80
100
Fluorene [nM]
1000 10000
% L
arva
l gro
wth
0
20
40
60
80
100
μ
Fig. 2. Larval growth after 48 h exposure of P. lividus fertilized eggs to different concentrations (nM) of naphthalene, phenanthrene, pyrene, fluoranthene and fluorene. Filled-in squares represent experiments carried out in the dark. Open circles represent experiments carried out under fluorescent light (n = 4). Error bars represent standarddeviations.
J. Bellas et al. / Marine Pollution Bulletin 57 (2008) 493–502 497
intestinalis embryos in darkness or under fluorescent light. Fluoran-thene was not toxic either in dark conditions and only a 10% signif-icant reduction in the percentage of normal larvae was registeredunder light exposure, with an EC10 value of 1196 nM. An experi-ment was also carried out with hydroxypyrene. Although hydrox-ypyrene was not toxic to C. intestinalis embryos in dark conditions,a significant decrease on the percentage of normal larvae was ob-served under light exposure, yielding EC10 and EC50 values of 293and 434 nM.
In the experiments carried out with ascidian embryos, naph-thalene was significantly more toxic in the dark than underfluorescent light (p < 0.0001), but the EC50s were not significantlydifferent (p = 0.07). On the contrary, phenanthrene and hydrox-ypyrene resulted significantly more toxic under fluorescent
light than in the dark (p < 0.0001) showing photoenhancedtoxicity.
3.4.1. QSARsA baseline QSAR (quantitative structure–activity relationship)
model, based on the assumption that a target lipid is the site of ac-tion of narcotic chemicals in the organism (Di Toro et al., 2000),was used to predict the toxicity of the studied PAHs (in terms ofthe estimated EC20s) on the basis of their octanol–water partitioncoefficients (KOW). Despite using different species and endpoints,toxicity of single PAHs was inversely related to their KOW valueswith a regression slope of �1.09 ± 0.12 (Fig. 4). The estimatedregression slope is not significantly different (Student’s test,p < 0.001) from the universal narcosis slope (m = �0.95) reported
Hydroxypyrene [nM]
10 100 1000
% N
orm
al la
rvae
0
20
40
60
80
100
Naphthalene [ M]
0.1 1 10 100
% N
orm
al la
rvae
0
20
40
60
80
100
Phenanthrene [nM]
100 1000
% N
orm
al la
rvae
0
20
40
60
80
100
Pyrene [nM]
10 100 1000
% N
orm
al la
rvae
0
20
40
60
80
100
Fluoranthene [nM]
100 1000
% N
orm
al la
rvae
0
20
40
60
80
100
μ
Fig. 3. Percentage of normal larvae after 24 h exposure of C. intestinalis fertilized eggs to different concentrations (nM) of naphthalene, phenanthrene, pyrene, fluorantheneand hydroxypyrene. Filled-in squares represent experiments carried out in the dark. Open circles represent experiments carried out under fluorescent light (n = 4). Error barsrepresent standard deviations.
498 J. Bellas et al. / Marine Pollution Bulletin 57 (2008) 493–502
for a large database comprising 156 chemicals and 33 marine andfreshwater species (Di Toro et al., 2000); or from the regressionslope obtained by Neff et al. (2005) (m = �1.16) with more than300 toxicity values of 25 aromatic hydrocarbons (14 PAHs) forfreshwater and marine invertebrates and fish. The regression slopeobtained here with data from photoactivated PAHs was�1.47 ± 0.15 which significantly differs (Student’s test, p < 0.001)from the universal narcosis slope mentioned above and from theregression slope obtained by Neff et al. (2005).
3.4.2. Mixture toxicityResults of the mixture toxicity experiments are given in Fig. 5.
Five nominal equitoxic concentrations (0.2, 1.2, 2.4, 3.5 and 4.7TU after being recalculated), were used to test the mixture toxicity
of PAHs to the sea-urchin embryonic development. Although thetested compounds are considered as type I narcotic chemicals withthe same mode of action, the prediction according to the CA con-cept does not accurately predict the observed mixture toxicities.The estimated EC20 (95% CI) of the mixture, 1.47 (1.42–1.52) TU,was significantly higher than 1 TU (F-test, p < 0.0001), thus denot-ing slightly less than additive effects of PAHs mixtures.
4. Discussion
The present study reports acute effects of naphthalene, phenan-threne, fluoranthene and pyrene to early developmental stages ofM. galloprovincialis, P. lividus and C. intestinalis. In addition, fluoreneand hydroxypyrene were tested with P. lividus and C. intestinalis
Log K OW
3.0 3.5 4.0 4.5 5.0
Log
EC
20
1
2
3
4
5
Log EC20= 7.91 - 1.09 Log KOW
r2 = 0.790
Log EC20 = 9.48 - 1.47 Log KOW
r2 = 0.916
Fig. 4. Relationship between the toxicity of the tested PAHs (measured as EC20s)and their logKOW. Filled-in symbols represent experiments carried out in the darkand the solid line corresponds to regression line obtained using those data. Opensymbols represent experiments carried out under fluorescent light and the dottedline corresponds to regression line obtained using those data. Circles: mussel data;triangles: sea-urchin data; diamonds: ascidian data; hexagon: data from our pre-vious results with the copepod Acartia tonsa (Bellas and Thor, 2007).
Toxic Units0.1 1 10
Larv
al g
row
th (
%)
0
20
40
60
80
100
Fig. 5. Larval growth after exposure of P. lividus fertilized eggs to an equitoxicmixture of five PAHs (naphthalene, phenanthrene, pyrene, fluoranthene and fluo-rene) (n = 4). Symbols correspond to experimental data. Error bars represent stan-dard deviations. Solid line indicates the concentration–response relationship for theexperimental data. Dashed line represents the predicted mixture toxicity accordingto CA.
J. Bellas et al. / Marine Pollution Bulletin 57 (2008) 493–502 499
embryos, respectively. The effects of the studied PAHs varieddepending on the tested species: all PAHs were toxic at the exper-imental concentrations to sea-urchin embryos and all but fluoran-thene were toxic to mussel embryos; however, only naphthaleneexerted significant toxicity to ascidian embryos when exposureswere conducted in the dark. Comparison of our results with previ-ous research is not straightforward since studies on the effects ofPAHs to aquatic organisms carried out in dark conditions, as thosein the natural habitats of early embryonic stages of marine inver-tebrates, are scarce. A comprehensive literature search has re-turned only a few studies where PAHs exposure was conductedin the dark. For instance, S�thre et al. (1984) investigated the ef-fects of naphthalene to sea-urchin (Strongylocentrotus droebachien-sis) and fish (Gadus morhua) embryos obtaining EC50 valuesbetween 7 and 27 lM. Those values are in agreement with the re-sults obtained here with sea-urchin and ascidian embryos (EC50sbetween 15 and 37 nM), but mussels yielded higher values(52 nM). Recently, Bellas and Thor (2007) evaluated the toxicityof fluoranthene, phenanthrene and pyrene on the survival, egg pro-
duction and recruitment of the copepod Acartia tonsa, and EC50 val-ues ranged between 385 and 824 nM for fluoranthene, between1012 and 2436 nM for phenanthrene, and between 295 and306 nM for pyrene. In general, copepods seem to be more sensitiveto PAHs than the species tested here. In contrast, Kagan et al.(1985) did not find toxicity of fluoranthene and pyrene to embryosand larvae of freshwater organisms (brine shrimp Artemia salina,water flea Daphnia magna, mosquito Aedes aegypti, leopard frogRana pipiens, and fish Pimephales promelas) when incubations wereconducted in the dark, which is in agreement with previous studiessuggesting higher sensitivities of marine than freshwater inverte-brates to pesticides (Hutchinson et al., 1998; Leung et al., 2001;Robinson, 1999). In the present study, EC50 calculations for exper-iments conducted in the dark were only possible for naphthaleneand phenanthrene (with the exception of ascidian experiments),since pyrene, fluoranthene, fluorene and hydroxypyrene did notcause a 50% decrease on the biological responses below the maxi-mum aqueous solubility. The low aqueous solubility of most PAHsis an important factor to take into account in order to assess therisk of those compounds to the marine environment, since theirbioavailability and therefore the maximum lipid concentration at-tained in the organisms is constrained by their aqueous solubility(Di Toro et al., 2000). A solubility cut-off has been suggested so thataqueous solubility of compounds with logKOW P 5.3 is too low toresult in the lipid concentration necessary to cause deleterious ef-fects even in the most sensitive organisms (Veith et al., 1983; DiToro et al., 2000). The PAHs employed in our experiments all havelogKOW < 5.
However, a growing body of evidence suggests that toxicity ofsome intermediate molecular weight PAHs may be enhanced inthe presence of ultraviolet (UV) light. Although photoinduced tox-icity of PAHs has been demonstrated in many studies with fresh-water organisms since the early 1980s (reviewed by Arfstenet al., 1996), the first report of phototoxicity to marine species isrelatively recent (Pelletier et al., 1997). In that study, toxicity of flu-oranthene and pyrene to larvae and juveniles of the bivalve Mulinialateralis and the mysid Mysidopsis bahia was 12 to >50,000 timeshigher under UV light than under fluorescent light. Likewise, Spe-har et al. (1999) attained a 21–1880 times increase in fluoranthenetoxicity under UV light to several freshwater and marine speciescompared to incubations under fluorescent light. Lyons et al.(2002) also found that pyrene provoked ca. 50% impairment onthe embryonic development of the pacific oyster Crassostrea gigasat approximately 500 nM under fluorescent light, whilst underUV light 25 nM caused a 100% inhibition of the embryonic develop-ment. More recently, Peachey (2005) reported significantly highertoxicity of fluoranthene and pyrene to larvae of three crustaceans(Libnia dubia, Menippe adina and Panopeus herbstii) under UV light;however, this study used very high UV light regimes (UVA: 1455–4914 lW cm�2, UVB: 196-581 lW cm�2) in comparison with real-istic UV light intensities measured in aquatic habitats (Barronet al., 2000), and therefore its ecological relevance is questionable.On the other hand, Steevens et al. (1999) did not find an increase inphenanthrene toxicity to sea-urchin (Lytechinus variegatus) em-bryos under UV-B light exposure compared to ambient or artificiallight. Furthermore, the aforementioned studies and previous eval-uations of PAHs phototoxicity with aquatic organisms have onlycompared toxicity under fluorescent or ambient light and underUV light exposure, but those studies did not consider the PAHs tox-icity in the dark. Indeed, care must be taken regarding the interpre-tation of photoenhanced toxicity of PAHs under UV light inlaboratory toxicity tests, since UV light itself may cause damageto marine organisms, and particularly to the sensitive early lifestages (Browman et al., 2000; Kuhn et al., 2000) so that the effectsof PAHs to organisms under UV light exposure might be seen as theinteraction between toxicants. Also, UV light is rapidly dispersed
500 J. Bellas et al. / Marine Pollution Bulletin 57 (2008) 493–502
and attenuated in the first decimetres of the water column (Barronet al., 2000) and therefore, the ecological relevance of UV lightphotoactivation of PAHs may be low. In the present work we havedemonstrated that some PAHs were more toxic to marine pelagicorganisms in the presence of fluorescent light (70 lE m�2 s�1) thanin dark conditions. Fluoranthene and pyrene were significantlymore toxic to mussel and sea-urchin embryos and hydroxypyrenewas more toxic to ascidian embryos under fluorescent light. Phen-anthrene did not show enhanced toxicity under fluorescent lightfor mussel; however, a significant increase in toxicity was detectedfor ascidian and sea-urchin embryos. On the other hand, naphtha-lene and fluorene did not show significant phototoxicity. The irra-diance value used in our experiments may be easily reached oreven exceeded in the water column (Barron et al., 2000), indicatingthat photoenhanced toxicity of certain PAHs may occur in the mar-ine environment and therefore this factor should be taken into ac-count in current risk evaluation procedures for PAHs.
Previous studies have modeled the photoinduced toxicity ofPAHs attending to their molecular electronic structure, specificallythe HOMO–LUMO gap, i.e. ‘‘the necessary energy to elevate anelectron from the highest occupied molecular orbital (HOMO) tothe lowest unoccupied molecular orbital (LUMO)” (Mekenyanet al., 1994; Veith et al., 1995; Ribeiro and Ferreira, 2005). Mole-cules with HOMO–LUMO gaps below 8.1 eV are considered to bepotentially phototoxic and maximum phototoxicity occurs in com-pounds with gap energies between 6.7 and 7.5 eV. Results of thepresent study are in agreement with this model since fluoranthene,pyrene and hydroxypyrene which present HOMO–LUMO gaps of7.7, 7.2 and 7.1 eV exhibited photoinduced toxicity, whilst naph-thalene and fluorene (10.1 and 8.5 eV) did not show significantphototoxicity. However, phenanthrene (8.2 eV) photoenhancedtoxicity was observed for ascidian and sea-urchin embryos butnot for mussel embryos. A possible explanation, as suggested byBoese et al. (1998), may be that compounds such as phenanthrene,with HOMO–LUMO gaps near the upper limit for phototoxic mol-ecules, are accumulated in the tissues of ascidian and sea-urchinembryos in a sufficient large amount to cause a phototoxicresponse.
The toxicity of single PAHs tested here was inversely related totheir KOW values following the logKOW-dependent QSAR modelsderived for baseline toxicity of general narcotics. According to DiToro et al. (2000), the slope of the equation is related to the chem-ical properties of the target lipid in the organism and thereforeshould be the same regardless of the species tested. Despite usingdifferent species and endpoints we obtained a regression slope of1.09, which is close to unity, a value commonly found in the liter-ature (e.g. Lee et al., 2001; Barata et al., 2005), and very similar tothe universal narcosis slope found by Di Toro et al. (2000)(m = �0.97 or �0.95 after chemical class corrections), and to theregression slope found by Neff et al. (2005) (m = �1.16) for aro-matic hydrocarbons. The regression slope obtained here using tox-icity data from photoactivated PAHs was 1.47, which significantlydeparts from the general narcosis model. We have also calculatedthe excess toxicity (Te) as the quotient of the PAH toxicity pre-dicted by the log KOW-dependent QSAR model (baseline toxicity)to the measured toxicity of the PAHs in our experiments (Groteet al., 2005). The Te is therefore a measure of how much toxic thecompound actually is compared to the assumption of general nar-cosis. Compounds with Te values > 27 are considered reactivechemicals or specifically acting chemicals rather than general orpolar narcotics, whereas Te < 10 indicate compounds with a nar-cotic mode of action. Calculated average Te values for naphthalene,fluorene, phenanthrene, fluoranthene and pyrene were 11, 8, 10, 11and 10, respectively, indicating that, in the absence of photoactiva-tion, the toxicity of PAHs mainly results from general narcosismechanisms (Di Toro et al., 2000; Grote et al., 2005). The results
of the fluorescent light toxicity data yielded different results.Naphthalene and fluorene showed Te values of 5 and 11, respec-tively, indicating no influence of the fluorescent light on the toxic-ity. In contrast, Te values for phenanthrene, fluoranthene andpyrene were 36, 310 and 412, indicating a deviation from generalnarcosis due to photoinduced toxicity mechanisms under fluores-cent light exposure.
PAHs frequently occur in complex mixtures rather than singlyin marine ecosystems (Kennish, 1992), thus, mixture toxicity ofPAHs should be taken into account in order to assess the impactof PAHs to the marine environment. Since PAHs are consideredas general narcotics and therefore similarly acting substances,the mixture toxicity of PAHs is expected to be additive as has beenshown in several studies (Swartz et al., 1995; Di Toro et al., 2000;Landrum et al., 2003; Barata et al., 2005). Nevertheless, deviationfrom additive toxicity has been also reported. For instance, syner-gistic effects have been found in marine and freshwater inverte-brates (Boese et al., 1999; Verrhiest et al., 2001) and, inagreement with our results, less than additive effects were re-ported for the marine amphipod Rhepoxynius abronius, the fresh-water amphipod Hyalella azteca and D. magna (Swartz et al.,1997; Lee et al., 2001; Olmstead and LeBlanc, 2005), indicating thatconcentration addition slightly overestimate the toxicity of PAHsmixtures. Olmstead and LeBlanc (2005) suggest that the deviationfrom additivity may be explained on the basis of different toxicitymechanisms of PAHs, questioning the assumption of a commonmode of action. Again, we have to bear in mind that our studytested the PAH mixture in the dark, whereas most studies are con-ducted under fluorescent light, which may increase the toxicity ofPAHs and alter their mechanism of toxicity.
In order to estimate the risk associated with the occurrence ofPAHs in the marine environment we have compared the reportedmaximum concentrations in polluted estuaries (Cmax) and the pre-dicted no effect concentrations (PNEC) of PAHs for the speciestested here, which were derived from the toxicity thresholds (esti-mated as the EC10) for each compound, applying an assessmentfactor of 10 (OECD 1992). Risk quotients (RQ) greater than 1, calcu-lated as Cmax/PNEC, indicate that adverse effects for these speciesare likely to occur. Although PAH water concentrations in offshoresites are usually below values reported in literature to be acutelytoxic to aquatic organisms (
PPAHs < 0.1 lg l�1, Kennish, 1992),
higher concentrations have been found in polluted coastal andestuarine areas. For instance, PAHs levels in the Galician coast in-creased up to 30–40 times the background levels after the Prestigeoil-spill (González et al., 2006; Franco et al., 2006), and PAH con-centrations as high as 30 lg/L have been measured in pollutedestuaries (e.g. Law et al., 1997; Maskaoui et al., 2002; Zhou andMaskaoui, 2003). Moreover, PAHs present a great affinity for sedi-ment particles and accumulate in the bottom of estuaries andcoastal zones close to urban and industrial areas, reaching valuesof 0.3–528 mg/kg in highly polluted sites (Simpson et al., 1996).Those pollutants may be available to benthic organisms by inges-tion of sediment particles or to pelagic organisms when they arereleased to the water column by biological, chemical and/or phys-ical processes (Fichet et al., 1998; Long et al., 1996; Bellas et al.,2007). On the basis of the present data, and ignoring the photoac-tivation of PAHs, worst-case environmental concentrations ofnaphthalene, fluorene and pyrene are not considered to pose a riskto sea-urchin, mussel or ascidian larvae (RQ = 0.04–0.39), whilstphenanthrene is approaching 1 for mussel larvae (0.73) and RQ va-lue of fluoranthene for sea-urchin is 1.1, indicating a risk for thesespecies. The RQ values increase when we consider the fluorescentlight exposure, except for naphthalene and fluorene. Thus, RQ offluoranthene and pyrene for sea-urchin and mussel larvae are0.85 and 1.4, and 3.0 and 1.1, respectively, indicating a higher riskfor those species.
J. Bellas et al. / Marine Pollution Bulletin 57 (2008) 493–502 501
The present work reports acute effects of PAHs on the early lifestages of marine invertebrates of ecological and commercial rele-vance. Although these effects were detected at levels above the re-ported typical concentrations in coastal areas and therefore littlerisk is predicted to occur, a higher level of risk was found forPAH polluted estuaries. Moreover, photoactivation of PAHs at nat-urally relevant levels may increase the risk of those compounds inthe marine environment.
Acknowledgements
We wish to thank Rocío Rendo and María López for technicalassistance. This study was supported by a Juan de la Cierva Con-tract from the Spanish Government (Ministerio de Educación yCiencia) to Juan Bellas and a predoctoral fellowship (Xunta de Gali-cia) to Liliana Saco-Álvarez. Research was funded by the projectPGIDIT05RMA31201PR (Xunta de Galicia) and by the projectVEM2003-20068-C05-02 from the Spanish Government (Ministe-rio de Ciencia y Tecnología).
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