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Levels and toxicity of polycyclic aromatic hydrocarbons in marine sediments
Trends in Analytical Chemistry, Vol. 28, No. 6, 2009
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Levels and toxicity of polycyclic aromatic hydrocarbons in marine sediments Anastasia Nikolaou, Maria Kostopoulou, Giusy Lofrano, Sureyya Meric, Andreas Petsas, Maria Vagi The occurrence of polycyclic aromatic hydrocarbons (PAHs) in the marine environment has attracted the attention of the scientific community, as these compounds are frequently detected in seawater and sediments at increasing levels and can have adverse health effects on marine organisms and humans. Several PAHs are potential human carcinogens and are included in the priority list of the European UnionÕs Water Framework Directive (2000/60/EC). Research regarding their environmental levels requires their determination by gas-chromatography and liquid-chromatography techniques, which have been developed and optimized, especially for marine-sediment samples. Results of sample analyses reveal the increasing occurrence of many species of PAHs worldwide, especially in marine sediments, where they finally accumulate, mostly in areas near intense industrial activities. In parallel, research on the toxicity of PAHs and their mixtures is continuing and is aiming to provide more insight into the health risks associated with the levels of PAHs in the environment. ª 2009 Elsevier Ltd. All rights reserved. Keywords: Health risk; Human carcinogen; Marine organism; Marine sediment; PAH; Polycyclic aromatic hydrocarbon; Priority list; Sample analysis; Seawater; Toxicity
1. Introduction Anastasia Nikolaou*, Maria Kostopoulou, Andreas Petsas, Maria Vagi University of the Aegean, Faculty of Environment, Department of Marine Sciences, University Hill, 81100 Mytilene, Greece Giusy Lofrano University of Salerno, Department of Civil Engineering, 84084 Fisciano (SA), Italy Sureyya Meric Present address: University of Naples, Federico II, Department of Biological Sciences, Section of Physiology and Hygiene, Ecotoxicology Research Laboratory (ERL-UNINA), I-80134 Naples, Italy
*
Corresponding author. Tel.: +30 2251036848; Fax: +30 2251036809; E-mail: [email protected]
0165-9936/$ - see front matter ª 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.trac.2009.04.004
The occurrence of polycyclic aromatic hydrocarbons (PAHs) in the environment has been documented for several decades. During the 1950s, they were reported to be present in food and cigarette smoke, and afterwards they were also detected in air samples, due to traffic-exhaust gases. In the marine environment, PAHs have been detected after oil spills since 1967. Monitoring programs have been introduced in order to evaluate the background levels of PAHs, as they can also derive from natural sources, and to estimate levels of environmental pollution from PAHs [1]. Many laboratories in the world are analyzing samples and revealing the increasing presence of PAHs in many environmental matrices, especially marine sediments [2–27] (Table 1). Several PAHs have been identified as chemical carcinogens. In 1775, an association between the incidence of scrotal cancer in chimney sweeps and their exposure to soot was noted by the British
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Table 1. Recent studies of determination of PAHs in marine sediments
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surgeon Sir Percival Pott [28]. In 1915, Japanese workers induced skin tumors in rabbits by applying coal tar [29]. The principal carcinogenic component of coaltar pitch was identified as benzo[a]pyrene in 1933 [1]. In the decades that followed, the evolution of chromatographic techniques provided the opportunity for the determination of many other PAH species in aqueous samples and more complex environmental matrices (e.g., marine sediments [2–27]). Procedures for the analysis of PAHs from sediments include Soxhlet extraction [30,31] ultrasonic extraction, microwave dissolution, pressurized liquid extraction (PLE) and supercritical fluid extraction (SFE) [32–38]. Solid-phase microextraction (SPME) techniques have also started to attract interest for sediment samples [19]. Clean-up procedures [e.g., column chromatography using a variety of adsorbents (e.g., acid-modified and base-modified silica gel, alumina and Florisil)] and addition of chemical modifiers (e.g., Na4EDTA) have also been developed to increase the extraction efficiency for PAHs from sediments [14]. The final extract is typically analyzed by gas-chromatography (GC) or liquid-chromatography (LC) techniques [1–27,39–41]. Scientific interest in the quality of marine sediments is quite recent and has especially increased in the past 10 years in relation to application of the European Union (EU)Õs Water Framework Directive (WFD) (2000/60/EC) [42]. One of the main objectives of the WFD is achievement and preservation of ‘‘Good Chemical Status’’ of
surface waters of the EU member states by 2015, while, in parallel, the monitoring of their quality is required. In coastal areas where various human activities take place, the effects have already become obvious, having social, economic and environmental impacts [5,12]. However, the quality of marine waters is directly related to the quality of sediments, which are the final compartment of storage of a large number of xenobiotics, including many PAH species [2–27]. According to the WFD, PAHs are considered priority substances due to their environmental behavior and their toxic effects [41]. Some PAHs are considered potentially carcinogenic for humans, particularly benzo[a]anthracene, chrysene, benzo[b]fluoranthene, benzo[a]pyrene and benzo[ghi]perylene [43] (Table 2). EU Directive 98/83/EC, which is relevant to water intended for human consumption, has set a limit of 0.10 lg/L for the total concentration of benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[ghi]perylene and indeno[1,2,3-cd]pyrene, while EU Decision 2455/ 2001/C included PAHs in a WFD list of priority substances and EU Directive 2008/105/EC set regulatory limits for PAHs in inland, transitional and coastal waters [44]. The Convention for the Protection of the Marine Environment of the North-East Atlantic (OSPAR Convention) has been in force since 1998 and has been ratified by Belgium, Denmark, Finland, France, Germany, Iceland, Ireland, Luxembourg, Netherlands, Norway, Portugal, Sweden, Switzerland and United Kingdom, and approved by the European Community
Table 2. Carcinogenic action of PAHs [46,47] PAHs
Indicator of carcinogenesis1
Total estimation2
Acenaphthylene Anthracene Benzo[a]anthracene Benzo[b]fluoranthene Benzo[k]fluoranthene Benzo[b]fluorene Benzo[g,h,i]perylene Benzo[a]pyrene Benzo[e]pyrene Chrysene Fluoranthene Fluorene Indeno[1,2,3-c,d]pyrene Perylene Phenanthrene Pyrene Dibenzo[a,h]anthracene Dibenzo[a,h]pyrene Benzo[c]fluorene Benzo[j]fluoranthene Dibenzo[a,e]pyrene
I S S S I I S I L I I S I I I S S I S S
3 2A 2B 2B 3 3 2A 3 3 3 3 2B 3 3 3 2A 2B 3 2B 2B
EPA Classification (EPA, 1994)3
Toxicity Equivalency Factor (TEF) (EPA; 1993)
D D B2 B2 B2
0.001 0.01 0.1 0.1 0.1
D B2
0.01 1
B2 D D B2
0.01 0.001 0.001 0.1
D D B2
0.001 0.001 5
1
No adequate data for humans.For animals: I, Insufficient data; L, Limited data; S, Sufficient data. 1, Carcinogen for humans; 2A, Probable carcinogen for humans; 2B, Possible carcinogen for humans; 3, Not classified regarding carcinogenicity for humans. 3 D, not classifiable as to human carcinogenicity; B2, probable human carcinogen. 2
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and Spain [45]. However, no limits have yet been set for marine sediments, although they are the final receivers of PAHs in the marine environment. The US Environmental Protection Agency (EPA) has determined that contaminated sediments pose both ecological and human health risks throughout the USA. It is estimated that roughly 10% of the sediments from US lakes, rivers and bays are contaminated with toxic chemicals that can adversely affect aquatic organisms or impair the health of wildlife or humans, who consume contaminated fish or shellfish [46,47]. Contaminated sediments are important sources of pollution and may result in ecotoxicological effects, which can occur at all levels of biological organization, from molecular to ecosystem [48,49]. Moreover, remobilization of toxic pollutants, which increases their bioavailability, can occur when contaminated sediments are disturbed and dredged [50]. US EPA has listed as priority pollutants 16 PAHs in wastewaters and 24 PAHs in soils, sediments, hazardous solid wastes, and groundwater [51]. Long et al. have proposed sediment-quality guidelines for PAHs in the USA [52]. 2. Sources and fate of PAHs in the marine environment The introduction of PAHs into the marine environment is performed via different processes [e.g., combustion of organic matter (pyrolytic origin), slow transformation of
organic matter on the geothermal scale (petroleum hydrocarbons), and degradation of biogenic material (diagenesis)]. Naturally-formed PAHs are biosynthesis products or come from oil welling up, and they usually occur in marine sediments at very low levels in the range 0.01–1 ng/g dry weight (d.w.) (background concentrations), but, in some cases, their background levels in sediments can be much higher (e.g., in anoxic sediments, perylene can occur at levels >400 ng/g [53] and background concentrations of PAHs >1300 ng/g have been measured in Alaskan sediments [54]). PAHs also originate from anthropogenic sources (e.g., industrial production, transportation and waste incineration). Human activities are important sources of a number of PAHs in the aqueous environment with the highest values being recorded in estuaries and coastal areas, and in areas with intense vessel traffic and oil treatment. Based upon diagnostic ratios and/or predominance of different PAH congeners, Zaghden et al. [55] classified three sources of PAHs: petrogenic; pyrolytic; and, natural oil seeps of diagenetic origin [56]. Petrogenic PAHs are related to petroleum, including crude oil and its refined products. Biogenic PAHs are generated by biotic processes or by early stages of diagenesis in marine sediments (e.g., perylene). Pyrogenic PAHs are generated by combustion of fossil fuels (coal and oil) and recent organic material (e.g., wood) [20]. The use of ratios of PAH components of the same molecular mass has been
Sediment sample
Extraction Soxhlet Ultrasonic SPE, SFE, microwave
Clean up Solvent exchange SPE cartridges
GC-MS
HPLC-DAD
Figure 1. The typical procedure followed for the analysis of PAHs in marine sediments.
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established as a reliable method of inferring PAH sources and diagenesis of PAHs according to: (i) the different thermodynamic stability of congeners; (ii) the characteristic composition of different PAH sources; and, (iii) changes in the relative abundances of PAHs between source and sediment [56]. Fluoranthene and pyrene (mass 202) are considered good indicators of petroleum combustion while phenantrene and anthracene (mass 178) are commonly used to distinguish between combustion and petroleum sources. PAHs of molecular masses 228 (benzo[a]anthracene and chrysene) and 276 (indeno[1,2,3-cd]pyrene and benzo[ghi]perylene) are used less frequently as PAH-source indicators and few guidelines have been established for their interpretation [57,58]. PAHs take part in various physical, chemical and biological processes in the marine environment, the most common of which are: adsorption and desorption; sedimentary deposition; atmospheric release; biodegradation; and, abiotic degradation Higher molecular-weight compounds are associated primarily with particles and are likely to be removed by dry deposition, while lower molecular-weight compounds are found primarily in the gas phase and are subject to transformation or removal by photochemical degradation [11]. The adsorption of PAHs occurs on the surface of organic and inorganic particles. The rate of adsorption is positively affected by the content of organic matter in the particles. The final fate of PAHs is generally sedimentary deposition, after transport in the water column, as reported for material collected in sediment traps [55]. However, in situ factors (e.g., portioning of PAHs between sorbed and aqueous phases, bioturbation and selective microbial degradation) may affect the observed composition of PAHs that results [59]. PAH molecules adsorbed on the sediments can be subject to slow-rate biodegradation and transformation to other forms via the actions of benthic organisms. The half-life of fluoranthene in surface sediments ranges from a few days to some years, depending on the environmental conditions. Degradation of compounds containing more than six rings has not been documented, while there is no evidence of degradation of PAHs in deep sediments [60,61]. The dissolution of PAHs in water is low, especially higher molecular-weight compounds. The dissolved molecules can be subject to photolysis and to chemical and biological oxidation, with both processes being inversely related to the content of dissolved organic matter [59]. Abiotic removal due to photochemical reactions of PAHs is important. Photo-oxidation is faster, with a half-
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life of a few hours in the presence of intense sunlight. PAHs containing three or more rings in their molecules show intense ultraviolet (UV) absorbance at wavelengths >300 mm, which exist in sunlight, resulting in oxidation of the molecules. The photo-oxidation reactions of dissolved PAH molecules include transfer of energy from the triplet (excited) state of the aromatic system, resulting in production of atomic oxygen, which reacts with the compound producing peroxide. Photolysis or pyrolysis of endoperoxides produces various reaction products via dealkylation and breaking the ring [61–63]. By contrast, photo-oxidation of PAHs in the adsorbed phase is not performed via peroxides. The compounds are photo-oxidized at a higher rate in adsorbed phase than in solution. Since PAHs in the environment occur combined with particles, this kind of photo-oxidation has greater environmental importance. The photo-oxidation half-life for benzo[a]pyrene under different conditions has been reported to be <1 day. Non-substituted anthracenes easily form photodimers by reaction of a molecule in an excited state with another molecule in the basic state. Atmospheric deposition, river run off, domestic and industrial outfalls, and direct spillage of petroleum or petroleum products are the main sources of anthropogenic PAHs in the marine environment. Knowledge of sources and possible transport pathways in aquatic sediments is the first step to effective pollution control. Vessels make a significant contribution by transporting these compounds in the marine environment. Especially in harbors, where the renewal of waters through contact with the open sea is limited, the accumulation of PAHs can be significant [2,64,65]. The processes of industrialization and urbanization that are rapidly gathering pace in some countries (e.g., India and China) increase the potential for associated increases in anthropogenic PAHs [63]. Studies focused on heavily contaminated, developing coastal regions show that contamination levels in those regions are growing to those of polluted industrialized zones in developed countries [59]. Once PAHs are introduced into the marine environment, physical transport and mechanical factors are mostly responsible for their observed distribution in sediments [59]. They are present in both dissolved and particulate phases. Due to their low solubility and their hydrophobic nature, they readily associate with inorganic and organic suspended particles and may accumulate in sediments at high concentrations [9].
3. Analytical methods Fig. 1 shows a generalized typical procedure for the determination of PAHs in marine sediment. Table 3 summarizes information on the analytical methods applied by different researchers and the ana-
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Sample preparation
Column and temperature program used
Range of total concentration of PAHs
Ref.
Naphthalene; Acenaphthylene; Acenaphthene; Fluorene; Phenanthrene; Anthracene; Fluoranthene; Pyrene; Benz[a]anthracene; Chrysene; Benzo[b]fluoranthene; Benzo[k]fluoranthene; Benzo[a]pyrene; Benzo[ghi]perylene; Dibenz[a,h]anthracene; Indeno[1,2,3-cd]pyrene Anthracene; Phenanthrene; Fluoranthene; Pyrene; Benzo[a]anthracene; Chrysene; Benzo[b]fluoranthene; Benzofluoranthene; Benzo[e]pyrene; Benzo[a]pyrene; Perylene; Indeno[1,2,3-cd]pyrene; Benzo[ghi]perylene; Coronene
Soxhlet extraction with dichloromethane for 16 h
DB-5MS (70C for 4 min, 10C/min to 300C, held for 10 min)
3.15–144.89 lg/kg d.w.
[2]
Pressurized solvent extraction with a mixture of dichloromethane and acetone (3:1, v/v) at 175oC and 10.5 MPa for 5 min. Rotary preconcentration, clean up with 5% H2O deactivated silica-gel column, elution with eluted with 20 mL hexane/dichloromethane (3:1, v/v). Activated copper treatment, pass through fully activated silica-gel column, elution with hexane/ dichloromethane (3:1, v/v). Ultrasonic extraction with dichloromethane for 30 min, then shaken overnight, and again ultrasonication for 30 min. Centrifugation for 10 min at 3000 rpm, rotary evaporation to 3–6 ml, nitrogen evaporation to 0.5 ml. Clean up with Enviropack 18 column (3 ml/ 500 mg), elution with 10 ml of 75:25 (v/v) pentane: dichloromethane mixture. Concentration with nitrogen to 0.3 ml, addition of 2 ml hexane and concentration again to 0.5 ml. Centrifugation for 10 min, addition of Na2SO4 and activated copper. A. Soxhlet extraction for 24 h, using dichloromethane-pentane 1:1 solvent mixture. Filtration of the extracts through a pre-cleaned Pasteur pipette filled with solvent-rinsed glass wool and pre-cleaned anhydrous Na2SO4, (previously rinsed with dichloromethane) and concentration in a rotary evaporator to final volume around 2 ml. Evaporation with nitrogen to dryness and dissolution in 1 ml solution containing the perdeuterated internal standards in cyclohexane (0.2 mg/L each): acenaphthene d10; phenanthrene d10, chrysene d12 and perylene d12. B. Ultrasonication with pentane-dichloromethane 1:1 v/v C. Ultrasonication with dichloromethane
HP5-MS (70C for 2 min, 30C/min to 150C, increased to 310C at 4C/ min and held for 10 min)
6–8399 ng/g d.w.
[5]
RTX5-MS (110C for 2 min, 25C/min to 310C, 5 min, held for 15 min)
0.383 lg/g d.w.
[4]
Equity-5 (40C for 2 min, 40C/min to 100C, 10C/ min to 200C, 30C/min to 325C, held for 8 min)
72–18381 lg/kg d.w.
[9]
Naphthalene; Acenaphthylene; Acenaphthene; Fluorene; Phenanthrene; Anthracene; Fluoranthene; Pyrene; Benzo(a)anthracene; Chrysene; Benzo(b,k)fluoranthene
Benzo(a)pyrene; Benzo(ghi)perylene; Dibenzo(ah)anthracene; Indeno(1,2,3-cd)pyrene
Naphthalene; 2 methyl naphthalene; 1 methyl naphthalene; Acenaphthylene; Acenaphthene; Fluorene; Phenanthrene; Anthracene; 2 methyl anthracene; 9 methyl anthracene; Fluoranthene; Pyrene; 1 methyl pyrene; Benzo(a)anthracene; Chrysene; Benzo(b)fluoranthene; Benzo(k)fluoranthene; Benzo(a)pyrene; Perylene; Indeno(1,2,3-cd)pyrene; Dibenzo(a,h)anthracene; Benzo(ghi)perylene
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PAHs studied
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Table 3. Analytical methods and concentration range reported by different researchers for PAHs in marine-sediment samples
Soxhlet extraction with dichloromethane/ methanol (v/v: 2:1) for 48 h. Clean up with silica/ alumina column chromatography
Naphthalene; Acenaphthylene; Acenaphthene; Fluorene; Phenanthrene; Anthracene; Fluoranthene; Pyrene; Benz[a]anthracene; Chrysene; Benzo[b+k]fluoranthenes; Benzo[e]pyrene; Benzo[a]pyrene; Indeno[1,2,3cd]pyrene; Dibenz[a,h]anthracene; Benzo[g,h,i]perylene
Soxhlet extraction with acetone–hexane (1:1), desulfurization with activated copper, evaporation with nitrogen, purification by liquid chromatography on Florisil cartridge, further clean up on a silica-gel column
Naphthalene; Acenaphthylene; Acenaphthene; Fluorene; Phenanthrene; Anthracene; Fluoranthene; Pyrene; Benz[a]anthracene; Chrysene; Benzo[b]fluoranthene; Benzo[k]fluoranthene; Benzo[a]pyrene; Indeno(1,2,3cd)pyrene; Dibenzo(a,h)anthracene; Benzo[ghi]perylene; 2-methylnaphthalene; 1-methylnaphthalene; 2,6dimethylnaphthalene; 1,3,5-trimethylnaphthalene; 1methylnaphthalene; Biphenyl; Benzo[e]pyrene; Perylene Naphthalene; 2-methylnaphthalene; Acenaphthylene; Acenaphthene; 2,3,5-trimethylnaphthalene; Fluorene; Phenanthrene; Anthracene; 2-methylanthracene; Fluoranthene; Pyrene; 9,10-dimethylanthracene; Benzo(c)phenanthrene; Benzo(a)anthracene; Chrysene; Benzo(b)fluoranthene; Benzo(k)fluoranthene; 7,12dimethylbenzo(a)anthracene; Benzo(a)pyrene; 3methylchloranthrene; Indeno(1,2,3-cd)pyrene; Dibenzo(a,h)anthracene; Benzo(g,h,i)perylene; Dibenzo(a,l)pyrene; Dibenzo(a,h)pyrene
Soxhlet extraction with methylene chloride. Extensive clean up by Si/Al column chromatography and HPLC with size-exclusion column
Ultrasonication with dichloromethane, shaken overnight and ultrasonicated again, centrifugation, evaporation under a nitrogen stream, clean up with solid-phase extraction column
HP-5 (50oC for 1 min, 10oC/ min to 180oC, for 7 min, 10oC/min to 230oC, for 25 min, 20oC/min to 280oC, for 5 min) PTE-5 GC-MS, EI mode (70 eV electron energy), with ion source, quadrupole and transfer line temperatures of 200C, 100C and 290C, respectively DB-5MS 1.5 min at 60C, first rate 4C/min to 300C, isothermal pause 10 min at 300C
DB5MS 60C to 310C at a rate of 15C/min, final holding time 13 min.
68–1500 ng/g d.w.
[27]
380–12,750 lg/kg d.w.
[68]
8.80–18 500 ng/d.w.
[59]
2005: 36 ± 3– 1908 ± 114 ng/g d.w. 2004: 28 ± 3– 312 ± 24 ng/g d.w.
[67]
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Fluoranthene; Pyrene; Benzo(a)anthracene; Chrysene; Benzofluoranthene; Benzo(e)pyrene; Indeono[1,2,3cd]pyrene; Benzo[ghi]perylene
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lytical conditions used in each case. As can be seen from Table 3, the HP-5 chromatographic column or equivalent is used most (in GC-MS analysis). Clean-up procedures are applied by many researchers in order to increase the sensitivity of the methods, usually by silica/alumina columns, after desulfurization by activated copper [66,67,59,68]. Some studies still utilize the traditional time-consuming, but effective, Soxhlet extraction, while HPLC instrumentation is also used in some cases with relatively good results.
4. Levels of PAHs in marine sediments PAHs are widespread contaminants in oceans. However, their concentrations in sediments of coastal bays, estuaries and continental shelves are often much higher due to greater pressures from specific anthropogenic inputs, suggesting direct influence of these sources on pollutantdistribution patterns [56]. Due to the chemical composition of seawater, the occurrence of PAHs, which are very hydrophobic compounds, will be at very low levels of concentration (<1 ng/L), in contrast to concentration in other aqueous matrices (e.g., wastewater, river water, rainwater and sediments), where, depending on the study area, the corresponding values are in the range 1 ng/g d.w. to >100 ng/g d.w. [12]. Due to their hydrophobic character, PAHs tend to adsorb onto the particulate matter so that they are transported to and accumulate in sediments. Sprovieri et al. [56] found that the priority PAHs in samples of surface sediment collected from Naples harbor were in the range 9–31,774 ng/g. Three-ring and fourring PAHs appeared dominant in the sediments studied, with median values of concentration generally greater than 60–70% of the total concentration of the PAHs. There was no correlation between grain size or total organic carbon and the distribution of PAHs or single congeners. Te Naples harbor showed a median and a range of variability greater than most other Mediterranean and European ports (e.g., Sardinia, Corsica, Spain, France and England) but similar to those measured in commercial harbors in the USA, the Middle East and Australia. Total PAH concentrations measured by Vane at al. [74] in sediments from the Mersey Estuary (UK) were in the range 626–3766 mg/kg, which were intermediate compared to other UK estuaries with similar histories of industrialization and urbanization. The molecular indices in this case suggested mainly pyrolitic inputs, augmented by a variety of industrial or petrogenic sources. A major part of recent studies on hydrocarbon pollution has focused on the northwestern part of the Mediterranean Sea [69,70], but there is a tremendous lack of information regarding the eastern Mediterranean Sea 660
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[71] and the southern Mediterranean Sea, and data are scarce for the coasts of Algeria and Egypt [72,73]. There is therefore an urgent need for better assessment of hydrocarbon contamination for the whole Mediterranean Sea.
5. Toxicity of PAHs in marine sediments The ecotoxicity of individual PAHs or mixtures of some of PAHs has been investigated using different bioassays. The most indicative toxicity end-points have been changed in line with the species tested. In one toxicity study of individual PAHs, the effect of benzo[a]pyrene (BaP) on Mytilus edulis fed daily with Isochrysis galbana algae that had previously been kept in the presence of BaP for 24 h was investigated in terms of blood-cell lysosomal membrane damage (neutral red-dye retention assay) and induction of digestive gland microsomal mixed-function oxygenase (MFO) parameters [BaP hydroxylase (BPH) and NADPH-cytochrome c (P450) reductase activities]. It was reported that the bioaccumulation of BaP in tissues (and to a lesser extent in digestive gland microsomes) of M. edulis increased with both increasing BaP and algal exposure concentrations, and over time. This pointed to impact on M. edulis in line with BaP exposure possibly due to a direct effect of BaP on the integrity of the blood-cell lysosomal membrane. An increase in NADPH-cytochrome c reductase activity was explained due to a transient response of the digestive gland microsomal MFO system to BaP exposure [74]. BaP was also reported to induce infertility in the male reproductive system. Microsomes isolated from liver and testes of rat, mouse, hamster, ram, boar, bull and monkey were incubated with BaP. Post incubation, the ethyl-acetate-extracted samples were analyzed for BaP and metabolites by reversed-phase HPLC with fluorescence detection. Given the ability of BaP 7-8-diol 9, 10epoxide, 3-, and 9-hydroxy BaP to bind with DNA and form adducts, Smith et al. concluded that a risk was likely to arise from accumulation of BaP metabolites in testicular tissues [75]. Ethoxy resorufin dealkylase (EROD) inducing potency of 10 PAHs, generally used in risk assessment studies, was measured in H4IIE rat hepatoma cell line in vitro bioassays. The responses of the PAHs investigated to EROD activity varied: anthracene (Ant) and phenanthrene (Phe) exhibited no response; naphthalene (Nap) gave no or a very weak response; fluoranthene (Fla) and benzo[ghi]perylene (BghiP) showed weak responses at the highest doses; indeno [1,2,3-cd]pyrene (IP), benz[a]anthracene (BaA), benzo[a]pyrene (BaP), chrysene (Chr) and benzo[k]fluoranthene (BkF) displayed full bell-shaped dose-response curves.
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Table 4. Toxicity reference values for Mytilus larval development [77] Compound Anthracene* Fluorene*
Fluoranthene* Phenanthrene* Pyrene* Quinoline
Maximum observed porewater concentration (lg/L)
48-h EC50 (lg/L)
Ref.
0.3 <0.1 1.1 <0.1 1.0 <0.1
4260 1260 58.8 410 >11900 4300
[83] [84] [83] [84] [83] [84]
Note: Asterisks indicate a compound identified as a COPC based on comparison of bulk sediment concentrations with numerical sediment quality criteria (BCWLAP, 2003). PAH TRVs include a 10-fold uncertainty factor to address potential interspecies differences.
BaP EROD induction equivalency factors (BaP-IEFs) were calculated and increased in the order Ant < Phe < Fla < Nap < BghiP < IP < BaA < BaP < Chr < BkF [76]. Although there has been a severe toxic outcome from individual PAHs, as reported in Table 4 for the toxicity of some individual PAHs to Mytilus galloprovincialis [77], there was found almost no dose-response relationship for PAHs in the presence of complex mixtures (e.g., sediments). A positive relationship was found in a study performed on the sediments collected in the Izmit Bay, Turkey, where bulk and elutriate samples were toxic to Phaeodactylum tricornutum tested in batch bioassays. The sediments collected from the inner sites of the bay were reported to be contaminated with Cd, Hg, As and PAHs, which were consistent with the high organic carbon contents. Bulk sediments displayed toxicity to the microalgae throughout the bay [78]. However, in a survey conducted on marine sediments in Santa Monica bay, spatial and temporal patterns indicated that toxicity was most strongly associated with the historical disposal of municipal wastewater sludge. The statistical significance between response by bioassays and chemical analyses was negative (e.g., seaurchin fertilization correlated negatively with concentration of zinc, silver, copper, cadmium, total PCBs and total PAHs, while amphipod survival had a significant negative correlation only for silver [79]). In the extracts of freeze-dried surface sediment samples from the southern North Sea (fraction <63 lm), the concentrations of the PAHs measured were in the range 2.6–200 lg/kg d.w., with the highest concentrations at near-shore locations and river mouths. Responses in the multi-bioassays toxicity tests were found to vary. The responses in Microtox and Mutatox genotoxicity assays were low with higher responses observed at near-shore locations. Relevant responses were obtained in umu-C genotoxicity and ER-CALUX assays for estrogenicity. At the oyster grounds, DR-CALUX responses appeared to be linked to the occurrence of larger PAHs (4–6 rings). By applying a non-destructive clean-up procedure, DR-CALUX responses were significantly higher than those obtained in the current protocol [80].
The 16 US EPA priority pollutant PAHs were investigated in sediments of the Niger Delta comprising 2–6ring congeners with molecular mass of 128–278. Concentrations of total PAHs in the range 20.7– 72.1 ng/g d.w. showed no correlation with the toxicity bioassays of Vibrio fischeri and Lemna minor [81]. Similar results were reported by Bihari et al. [82], who found that the organic extracts of the superficial sediments collected from the coastal area of Rovinj (Northern Adriatic, Croatia) showed toxic potential that was consistent with the sediment type, but no correlation was observed between toxicity measured by Microtox assay and concentrations of total PAHs and PAH congeners. The genotoxicity assessment of the organic extracts from the sediment showed no significant genotoxic potential in the bacterial umu-test. By using hemolymph of M. galloprovincialis, the DNA damage was positively related to total PAHs at four sampling sites, but the highest DNA damage was not observed at the site with the highest total PAH content in the sediment [82].
6. Conclusions and outlook Due to continuously increasing anthropogenic activities, management of the pollution from sediments in coastal areas is attracting more attention. PAHs are characterized by high toxicity and have been included in the priority substances of the WFD. The distributions of PAHs in the environment and potential human health risks have become the focus of much attention. Higher molecular-weight compounds are associated primarily with particles and are likely to be removed by dry deposition, while lower molecular-weight compounds are found primarily in the gas phase and are subject to transformation or removal by photochemical degradation. Their presence in marine sediments combined with other potentially toxic compounds can result in negative effects, which have yet to be investigated to any great extent, mainly due to lack of appropriate methodology and the complexity of the subject matter. At present, significant research is being devoted to optimizing ana-
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lytical methodologies for the determination of trace concentrations of PAHs in complex matrices (e.g., sediments). For the southern Mediterranean Sea, in particular, due to lack of data, there is an urgent need to assess hydrocarbon contamination.
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Trends in Analytical Chemistry, Vol. 28, No. 6, 2009 [60] J.M. Neff, Polycyclic Aromatic Hydrocarbons in the Aquatic Environment: Sources, Fates and Biological Effects, Applied Science Publishers Ltd., Essex, England, 1979, 262 pp. [61] J.W. Moore, S. Ramamoorthy, Aromatic hydrocarbons-polycyclics, in: J.W. Moore, S. Ramamoorthy, Organic Chemicals in Natural Waters: Applied Monitoring and Impact Assessment. Springer-Verlag, New York, NY, USA, 1984, pp. 67–87. [62] N.K. Nagpal, Ambient Water Quality Criteria For Polycyclic Aromatic Hydrocarbons (PAHs), Ministry of Environment, Lands and Parks, Province of British Columbia, February 1993 (http:// www.env.gov.bc.ca/wat/wq/BCguidelines/pahs/pahs-02.htm, Last accessed 7 August 2008). [63] J. Xu, Y. Yu, P. Wang, W. Guo, S. Dai, H. Sun, Chemosphere 67 (2007) 1408. [64] M.I. Bakker, B. Casado, J.W. Koerselman, J. Tolls, C. Kolloffel, Sci. Total Environ. 263 (2000) 91. [65] A.L. NÕGuessan, J.S. Levitt, M.C. Nyman, Chemosphere 55 (2004) 1413. [66] B. Men, M. He, L. Tan, C. Lin, X. Quan, Mar. Pollut. Bull. (2009) (in press). [67] A. Curtosi, E. Pelletier, C.L. Vodopivez, W.P. Mac Cormack, Sci. Total Environ. 383 (2007) 193. [68] N. Cardellicchio, A. Buccolieri, S. Giandomenico, L. Lopez, F. Pizzulli, L. Spada, Mar. Pollut. Bull. 55 (2007) 451. [69] E. Lipiatou, I. Tolosa, R. Simo´, I. Bouloubassi, J. Dachs, S. Marti, M.-A. Sicre, J.M. Bayona, J.O. Grimalt, A. Saliott, J. Albaiges, Deep Sea Res., Part II 44 (1997) 881. [70] J. Dachs, J.M. Bayona, J. Fillaux, A. Saliot, J. Albaige´s, Mar. Chem. 65 (1999) 195. [71] A. Gogou, I. Bouloubassi, E.G. Stephanou, Mar. Chem. 68 (2000) 265. [72] T.A.T. Aboul-Kassim, B.R.T. Simoneit, Mar. Chem. 54 (1996) 135. [73] T.A.T. Aboul-Kassim, B.R.T. Simoneit, Mar. Pollut. Bull. 30 (1995) 63. [74] C.H. Vane, I. Harrison, A.W. Kim, Sci. Total Environ. 374 (2007) 112. [75] T.L. Smith, S.T. Merry, D.L. Harris, J.J. Ford, J. Ike, A.E. Archibong, A. Ramesh, Toxicol. in Vitro 21 (2007) 753. [76] A.T.C. Bosveld, P.A.F. de Bie, N.W. van den Brink, H. Jongepier, A.V. Klomp, Chemosphere 49 (2007) 75. [77] B.G. McDonald, Ecotoxicol. Environ. Safe. 62 (2005) 383. [78] L.G. Tolun, O.S. Okay, A.F. Gaines, M. Tolay, H. Tufekci, N. Kirati, Environ. Int. 26 (2001) 63. [79] D. Greenstein, S. Bay, A. Jirik, J. Brown, C. Alexander, Mar. Environ. Res. 56 (2002) 277. [80] H.J.C. Klamer, P.E.G. Leonards, M.H. Lamoree, L.A. Villerius, J.E.A. Kerman, J.F. Bakker, Chemosphere 58 (2005) 1579. [81] A. Olajirea, R. Altenburger, E. Kqster, W. Brack, Sci. Total Environ. 340 (2005) 123. [82] N. Bihari, M. Fafandel, B. Hamer, B. Kralj-Bilen, Sci. Total Environ. 366 (2006) 602. [83] M.C. Pelletier, R.M. Burgess, K. Ho, A. Kuhn, R.A. McKinney, S.A. Ryba, Environ. Toxicol. Chem. 16 (1997) 2190. [84] M.C. Pillai, C.A. Vines, A.H. Wikramanayak, G.N. Cherr, Toxicology 186 (2003) 93. Anastasia Nikolaou obtained her BSc (1997) and PhD degree (2002) in the Department of Environmental Studies, University of the Aegean, Mytilene, Greece. She has been a lecturer in the Department of Marine Sciences of the University of the Aegean (Mytilene, Greece) since 2005. Her research interests include analytical methods for the determination of toxic priority substances in water and sediment, water quality and treatment/disinfection by-products, investigation of fate and toxicity of organic compounds in the environment. She has participated in many related national and international projects, authored or co-authored a number of refereed papers and technical reports, and has been editor of two books and guest editor for Desalination Journal.
Trends Maria Kostopoulou obtained her BSc (1978) in Biochemistry in the University of Paris VII (France), and her BSc (1981) in Chemistry and PhD degree (1987) in the Department of Chemistry, University of Athens, Greece. She has been an external Scientific Researcher in the Department of the Environment of the University of the Aegean since 1992 and she has been working as an Assistant Professor in the Department of Marine Sciences of the University of the Aegean (Mytilene, Greece) since 2001. Her research interests include analytical methods for the determination of priority substances in environmental matrices (wastewater, water and sediments), as well as toxicity evaluation of individual compounds and mixtures by the use of marine microalgae. She has participated in many related national and international projects. Giusy Lofrano holds BS and MSc Degrees (2003) in Civil Engineering and PhD degree (2007) in Environmental Engineering from University of Salerno, Italy. She is currently working on the application of Fenton oxidation, photo-Fenton oxidation and ozonation to high-strength wastewaters, particularly in the leather-tanning industry, to achieve an economically feasible and environmentally friendly integrated wastewater treatment. Her research interests are in the fields of biological treatment of wastewater, advanced oxidation processes, and industrial pollution control. She has been co-author of articles on constructed wetlands and advanced oxidation processes, she has participated in related national and international projects and she is currently assisting the didactic activity of Wastewater Treatment Plant-I and II courses at the Engineering Faculty of University of Salerno. Sureyya Meric¸ holds BS, MSc and PhD degrees in Environmental Engineering from Istanbul Technical University (Istanbul, Turkey). She has been working in this field for 20 years. Her post-doc studies were focused on ecotoxicology of many chemicals and complex mixtures. She has developed expertise in chemical and biological treatment of wastewater, wastewater reuse, water treatment and disinfection, toxicity of disinfection by-products, activated sludge modeling, inhibition, industrial pollution control, water quality and management, aquatic and sediment toxicity monitoring. Impact assessment of priorityemerging pollutants, xenobiotics and their removal by advanced oxidation processes, and groundwater remediation. She has been investigator or co-ordinator of many national and international projects, and a member of environmental organizations, committees and the editorial boards of international journals. Andreas Petsas holds BS (1998) and PhD (2006) degrees in Environmental Studies from the University of the Aegean, Faculty of the Environment. He has worked on kinetics of removal (hydrolysis, photodegradation) of organophosphorous insecticides from various environmental matrices, as well as on the toxicity of these compounds to marine algae. He has participated in several national and international projects on these subjects. Since 1999, he has been a scientific collaborator in the Water and Air Quality Laboratory of the Department of Environmental studies of the University of the Aegean, and, since 2006, he is an adjunct lecturer at the Department of Marine Sciences of the University of the Aegean. His research interests include development and optimization of gas-chromatography analytical methods for determination of organic pollutants in environmental samples and he has co-authored several papers in international scientific journals. Maria Vagi holds a BS degree in Chemistry (University of Thessalonica, 1997), and a PhD degree (2007) in Environmental Studies, University of the Aegean, Faculty of the Environment. She has worked on kinetics of removal (hydrolysis and photodegradation) of several categories of insecticides and organic pollutants from various environmental matrices, as well as on the toxicity of these compounds to marine algae. She has participated in several national and international projects on these subjects. She has been a researcher in the Water and Air Quality Laboratory from 1997 until 2007, and, since
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2000, she has been working as a laboratory-technical staff member at the Department of Marine Sciences. Her research interests include development and optimization of gas-chromatography analytical
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methods for determination of organic pollutants in environmental samples and she has co-authored several papers in international scientific journals.
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Levels and toxicity of polycyclic aromatic hydrocarbons in marine sediments Anastasia Nikolaou, Maria Kostopoulou, Giusy Lofrano, Sureyya Meric, Andreas Petsas, Maria Vagi The occurrence of polycyclic aromatic hydrocarbons (PAHs) in the marine environment has attracted the attention of the scientific community, as these compounds are frequently detected in seawater and sediments at increasing levels and can have adverse health effects on marine organisms and humans. Several PAHs are potential human carcinogens and are included in the priority list of the European UnionÕs Water Framework Directive (2000/60/EC). Research regarding their environmental levels requires their determination by gas-chromatography and liquid-chromatography techniques, which have been developed and optimized, especially for marine-sediment samples. Results of sample analyses reveal the increasing occurrence of many species of PAHs worldwide, especially in marine sediments, where they finally accumulate, mostly in areas near intense industrial activities. In parallel, research on the toxicity of PAHs and their mixtures is continuing and is aiming to provide more insight into the health risks associated with the levels of PAHs in the environment. ª 2009 Elsevier Ltd. All rights reserved. Keywords: Health risk; Human carcinogen; Marine organism; Marine sediment; PAH; Polycyclic aromatic hydrocarbon; Priority list; Sample analysis; Seawater; Toxicity
1. Introduction Anastasia Nikolaou*, Maria Kostopoulou, Andreas Petsas, Maria Vagi University of the Aegean, Faculty of Environment, Department of Marine Sciences, University Hill, 81100 Mytilene, Greece Giusy Lofrano University of Salerno, Department of Civil Engineering, 84084 Fisciano (SA), Italy Sureyya Meric Present address: University of Naples, Federico II, Department of Biological Sciences, Section of Physiology and Hygiene, Ecotoxicology Research Laboratory (ERL-UNINA), I-80134 Naples, Italy
*
Corresponding author. Tel.: +30 2251036848; Fax: +30 2251036809; E-mail: [email protected]
0165-9936/$ - see front matter ª 2009 Elsevier Ltd. All rights reserved. doi:10.1016/j.trac.2009.04.004
The occurrence of polycyclic aromatic hydrocarbons (PAHs) in the environment has been documented for several decades. During the 1950s, they were reported to be present in food and cigarette smoke, and afterwards they were also detected in air samples, due to traffic-exhaust gases. In the marine environment, PAHs have been detected after oil spills since 1967. Monitoring programs have been introduced in order to evaluate the background levels of PAHs, as they can also derive from natural sources, and to estimate levels of environmental pollution from PAHs [1]. Many laboratories in the world are analyzing samples and revealing the increasing presence of PAHs in many environmental matrices, especially marine sediments [2–27] (Table 1). Several PAHs have been identified as chemical carcinogens. In 1775, an association between the incidence of scrotal cancer in chimney sweeps and their exposure to soot was noted by the British
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Table 1. Recent studies of determination of PAHs in marine sediments
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surgeon Sir Percival Pott [28]. In 1915, Japanese workers induced skin tumors in rabbits by applying coal tar [29]. The principal carcinogenic component of coaltar pitch was identified as benzo[a]pyrene in 1933 [1]. In the decades that followed, the evolution of chromatographic techniques provided the opportunity for the determination of many other PAH species in aqueous samples and more complex environmental matrices (e.g., marine sediments [2–27]). Procedures for the analysis of PAHs from sediments include Soxhlet extraction [30,31] ultrasonic extraction, microwave dissolution, pressurized liquid extraction (PLE) and supercritical fluid extraction (SFE) [32–38]. Solid-phase microextraction (SPME) techniques have also started to attract interest for sediment samples [19]. Clean-up procedures [e.g., column chromatography using a variety of adsorbents (e.g., acid-modified and base-modified silica gel, alumina and Florisil)] and addition of chemical modifiers (e.g., Na4EDTA) have also been developed to increase the extraction efficiency for PAHs from sediments [14]. The final extract is typically analyzed by gas-chromatography (GC) or liquid-chromatography (LC) techniques [1–27,39–41]. Scientific interest in the quality of marine sediments is quite recent and has especially increased in the past 10 years in relation to application of the European Union (EU)Õs Water Framework Directive (WFD) (2000/60/EC) [42]. One of the main objectives of the WFD is achievement and preservation of ‘‘Good Chemical Status’’ of
surface waters of the EU member states by 2015, while, in parallel, the monitoring of their quality is required. In coastal areas where various human activities take place, the effects have already become obvious, having social, economic and environmental impacts [5,12]. However, the quality of marine waters is directly related to the quality of sediments, which are the final compartment of storage of a large number of xenobiotics, including many PAH species [2–27]. According to the WFD, PAHs are considered priority substances due to their environmental behavior and their toxic effects [41]. Some PAHs are considered potentially carcinogenic for humans, particularly benzo[a]anthracene, chrysene, benzo[b]fluoranthene, benzo[a]pyrene and benzo[ghi]perylene [43] (Table 2). EU Directive 98/83/EC, which is relevant to water intended for human consumption, has set a limit of 0.10 lg/L for the total concentration of benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[ghi]perylene and indeno[1,2,3-cd]pyrene, while EU Decision 2455/ 2001/C included PAHs in a WFD list of priority substances and EU Directive 2008/105/EC set regulatory limits for PAHs in inland, transitional and coastal waters [44]. The Convention for the Protection of the Marine Environment of the North-East Atlantic (OSPAR Convention) has been in force since 1998 and has been ratified by Belgium, Denmark, Finland, France, Germany, Iceland, Ireland, Luxembourg, Netherlands, Norway, Portugal, Sweden, Switzerland and United Kingdom, and approved by the European Community
Table 2. Carcinogenic action of PAHs [46,47] PAHs
Indicator of carcinogenesis1
Total estimation2
Acenaphthylene Anthracene Benzo[a]anthracene Benzo[b]fluoranthene Benzo[k]fluoranthene Benzo[b]fluorene Benzo[g,h,i]perylene Benzo[a]pyrene Benzo[e]pyrene Chrysene Fluoranthene Fluorene Indeno[1,2,3-c,d]pyrene Perylene Phenanthrene Pyrene Dibenzo[a,h]anthracene Dibenzo[a,h]pyrene Benzo[c]fluorene Benzo[j]fluoranthene Dibenzo[a,e]pyrene
I S S S I I S I L I I S I I I S S I S S
3 2A 2B 2B 3 3 2A 3 3 3 3 2B 3 3 3 2A 2B 3 2B 2B
EPA Classification (EPA, 1994)3
Toxicity Equivalency Factor (TEF) (EPA; 1993)
D D B2 B2 B2
0.001 0.01 0.1 0.1 0.1
D B2
0.01 1
B2 D D B2
0.01 0.001 0.001 0.1
D D B2
0.001 0.001 5
1
No adequate data for humans.For animals: I, Insufficient data; L, Limited data; S, Sufficient data. 1, Carcinogen for humans; 2A, Probable carcinogen for humans; 2B, Possible carcinogen for humans; 3, Not classified regarding carcinogenicity for humans. 3 D, not classifiable as to human carcinogenicity; B2, probable human carcinogen. 2
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and Spain [45]. However, no limits have yet been set for marine sediments, although they are the final receivers of PAHs in the marine environment. The US Environmental Protection Agency (EPA) has determined that contaminated sediments pose both ecological and human health risks throughout the USA. It is estimated that roughly 10% of the sediments from US lakes, rivers and bays are contaminated with toxic chemicals that can adversely affect aquatic organisms or impair the health of wildlife or humans, who consume contaminated fish or shellfish [46,47]. Contaminated sediments are important sources of pollution and may result in ecotoxicological effects, which can occur at all levels of biological organization, from molecular to ecosystem [48,49]. Moreover, remobilization of toxic pollutants, which increases their bioavailability, can occur when contaminated sediments are disturbed and dredged [50]. US EPA has listed as priority pollutants 16 PAHs in wastewaters and 24 PAHs in soils, sediments, hazardous solid wastes, and groundwater [51]. Long et al. have proposed sediment-quality guidelines for PAHs in the USA [52]. 2. Sources and fate of PAHs in the marine environment The introduction of PAHs into the marine environment is performed via different processes [e.g., combustion of organic matter (pyrolytic origin), slow transformation of
organic matter on the geothermal scale (petroleum hydrocarbons), and degradation of biogenic material (diagenesis)]. Naturally-formed PAHs are biosynthesis products or come from oil welling up, and they usually occur in marine sediments at very low levels in the range 0.01–1 ng/g dry weight (d.w.) (background concentrations), but, in some cases, their background levels in sediments can be much higher (e.g., in anoxic sediments, perylene can occur at levels >400 ng/g [53] and background concentrations of PAHs >1300 ng/g have been measured in Alaskan sediments [54]). PAHs also originate from anthropogenic sources (e.g., industrial production, transportation and waste incineration). Human activities are important sources of a number of PAHs in the aqueous environment with the highest values being recorded in estuaries and coastal areas, and in areas with intense vessel traffic and oil treatment. Based upon diagnostic ratios and/or predominance of different PAH congeners, Zaghden et al. [55] classified three sources of PAHs: petrogenic; pyrolytic; and, natural oil seeps of diagenetic origin [56]. Petrogenic PAHs are related to petroleum, including crude oil and its refined products. Biogenic PAHs are generated by biotic processes or by early stages of diagenesis in marine sediments (e.g., perylene). Pyrogenic PAHs are generated by combustion of fossil fuels (coal and oil) and recent organic material (e.g., wood) [20]. The use of ratios of PAH components of the same molecular mass has been
Sediment sample
Extraction Soxhlet Ultrasonic SPE, SFE, microwave
Clean up Solvent exchange SPE cartridges
GC-MS
HPLC-DAD
Figure 1. The typical procedure followed for the analysis of PAHs in marine sediments.
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established as a reliable method of inferring PAH sources and diagenesis of PAHs according to: (i) the different thermodynamic stability of congeners; (ii) the characteristic composition of different PAH sources; and, (iii) changes in the relative abundances of PAHs between source and sediment [56]. Fluoranthene and pyrene (mass 202) are considered good indicators of petroleum combustion while phenantrene and anthracene (mass 178) are commonly used to distinguish between combustion and petroleum sources. PAHs of molecular masses 228 (benzo[a]anthracene and chrysene) and 276 (indeno[1,2,3-cd]pyrene and benzo[ghi]perylene) are used less frequently as PAH-source indicators and few guidelines have been established for their interpretation [57,58]. PAHs take part in various physical, chemical and biological processes in the marine environment, the most common of which are: adsorption and desorption; sedimentary deposition; atmospheric release; biodegradation; and, abiotic degradation Higher molecular-weight compounds are associated primarily with particles and are likely to be removed by dry deposition, while lower molecular-weight compounds are found primarily in the gas phase and are subject to transformation or removal by photochemical degradation [11]. The adsorption of PAHs occurs on the surface of organic and inorganic particles. The rate of adsorption is positively affected by the content of organic matter in the particles. The final fate of PAHs is generally sedimentary deposition, after transport in the water column, as reported for material collected in sediment traps [55]. However, in situ factors (e.g., portioning of PAHs between sorbed and aqueous phases, bioturbation and selective microbial degradation) may affect the observed composition of PAHs that results [59]. PAH molecules adsorbed on the sediments can be subject to slow-rate biodegradation and transformation to other forms via the actions of benthic organisms. The half-life of fluoranthene in surface sediments ranges from a few days to some years, depending on the environmental conditions. Degradation of compounds containing more than six rings has not been documented, while there is no evidence of degradation of PAHs in deep sediments [60,61]. The dissolution of PAHs in water is low, especially higher molecular-weight compounds. The dissolved molecules can be subject to photolysis and to chemical and biological oxidation, with both processes being inversely related to the content of dissolved organic matter [59]. Abiotic removal due to photochemical reactions of PAHs is important. Photo-oxidation is faster, with a half-
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life of a few hours in the presence of intense sunlight. PAHs containing three or more rings in their molecules show intense ultraviolet (UV) absorbance at wavelengths >300 mm, which exist in sunlight, resulting in oxidation of the molecules. The photo-oxidation reactions of dissolved PAH molecules include transfer of energy from the triplet (excited) state of the aromatic system, resulting in production of atomic oxygen, which reacts with the compound producing peroxide. Photolysis or pyrolysis of endoperoxides produces various reaction products via dealkylation and breaking the ring [61–63]. By contrast, photo-oxidation of PAHs in the adsorbed phase is not performed via peroxides. The compounds are photo-oxidized at a higher rate in adsorbed phase than in solution. Since PAHs in the environment occur combined with particles, this kind of photo-oxidation has greater environmental importance. The photo-oxidation half-life for benzo[a]pyrene under different conditions has been reported to be <1 day. Non-substituted anthracenes easily form photodimers by reaction of a molecule in an excited state with another molecule in the basic state. Atmospheric deposition, river run off, domestic and industrial outfalls, and direct spillage of petroleum or petroleum products are the main sources of anthropogenic PAHs in the marine environment. Knowledge of sources and possible transport pathways in aquatic sediments is the first step to effective pollution control. Vessels make a significant contribution by transporting these compounds in the marine environment. Especially in harbors, where the renewal of waters through contact with the open sea is limited, the accumulation of PAHs can be significant [2,64,65]. The processes of industrialization and urbanization that are rapidly gathering pace in some countries (e.g., India and China) increase the potential for associated increases in anthropogenic PAHs [63]. Studies focused on heavily contaminated, developing coastal regions show that contamination levels in those regions are growing to those of polluted industrialized zones in developed countries [59]. Once PAHs are introduced into the marine environment, physical transport and mechanical factors are mostly responsible for their observed distribution in sediments [59]. They are present in both dissolved and particulate phases. Due to their low solubility and their hydrophobic nature, they readily associate with inorganic and organic suspended particles and may accumulate in sediments at high concentrations [9].
3. Analytical methods Fig. 1 shows a generalized typical procedure for the determination of PAHs in marine sediment. Table 3 summarizes information on the analytical methods applied by different researchers and the ana-
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Sample preparation
Column and temperature program used
Range of total concentration of PAHs
Ref.
Naphthalene; Acenaphthylene; Acenaphthene; Fluorene; Phenanthrene; Anthracene; Fluoranthene; Pyrene; Benz[a]anthracene; Chrysene; Benzo[b]fluoranthene; Benzo[k]fluoranthene; Benzo[a]pyrene; Benzo[ghi]perylene; Dibenz[a,h]anthracene; Indeno[1,2,3-cd]pyrene Anthracene; Phenanthrene; Fluoranthene; Pyrene; Benzo[a]anthracene; Chrysene; Benzo[b]fluoranthene; Benzofluoranthene; Benzo[e]pyrene; Benzo[a]pyrene; Perylene; Indeno[1,2,3-cd]pyrene; Benzo[ghi]perylene; Coronene
Soxhlet extraction with dichloromethane for 16 h
DB-5MS (70C for 4 min, 10C/min to 300C, held for 10 min)
3.15–144.89 lg/kg d.w.
[2]
Pressurized solvent extraction with a mixture of dichloromethane and acetone (3:1, v/v) at 175oC and 10.5 MPa for 5 min. Rotary preconcentration, clean up with 5% H2O deactivated silica-gel column, elution with eluted with 20 mL hexane/dichloromethane (3:1, v/v). Activated copper treatment, pass through fully activated silica-gel column, elution with hexane/ dichloromethane (3:1, v/v). Ultrasonic extraction with dichloromethane for 30 min, then shaken overnight, and again ultrasonication for 30 min. Centrifugation for 10 min at 3000 rpm, rotary evaporation to 3–6 ml, nitrogen evaporation to 0.5 ml. Clean up with Enviropack 18 column (3 ml/ 500 mg), elution with 10 ml of 75:25 (v/v) pentane: dichloromethane mixture. Concentration with nitrogen to 0.3 ml, addition of 2 ml hexane and concentration again to 0.5 ml. Centrifugation for 10 min, addition of Na2SO4 and activated copper. A. Soxhlet extraction for 24 h, using dichloromethane-pentane 1:1 solvent mixture. Filtration of the extracts through a pre-cleaned Pasteur pipette filled with solvent-rinsed glass wool and pre-cleaned anhydrous Na2SO4, (previously rinsed with dichloromethane) and concentration in a rotary evaporator to final volume around 2 ml. Evaporation with nitrogen to dryness and dissolution in 1 ml solution containing the perdeuterated internal standards in cyclohexane (0.2 mg/L each): acenaphthene d10; phenanthrene d10, chrysene d12 and perylene d12. B. Ultrasonication with pentane-dichloromethane 1:1 v/v C. Ultrasonication with dichloromethane
HP5-MS (70C for 2 min, 30C/min to 150C, increased to 310C at 4C/ min and held for 10 min)
6–8399 ng/g d.w.
[5]
RTX5-MS (110C for 2 min, 25C/min to 310C, 5 min, held for 15 min)
0.383 lg/g d.w.
[4]
Equity-5 (40C for 2 min, 40C/min to 100C, 10C/ min to 200C, 30C/min to 325C, held for 8 min)
72–18381 lg/kg d.w.
[9]
Naphthalene; Acenaphthylene; Acenaphthene; Fluorene; Phenanthrene; Anthracene; Fluoranthene; Pyrene; Benzo(a)anthracene; Chrysene; Benzo(b,k)fluoranthene
Benzo(a)pyrene; Benzo(ghi)perylene; Dibenzo(ah)anthracene; Indeno(1,2,3-cd)pyrene
Naphthalene; 2 methyl naphthalene; 1 methyl naphthalene; Acenaphthylene; Acenaphthene; Fluorene; Phenanthrene; Anthracene; 2 methyl anthracene; 9 methyl anthracene; Fluoranthene; Pyrene; 1 methyl pyrene; Benzo(a)anthracene; Chrysene; Benzo(b)fluoranthene; Benzo(k)fluoranthene; Benzo(a)pyrene; Perylene; Indeno(1,2,3-cd)pyrene; Dibenzo(a,h)anthracene; Benzo(ghi)perylene
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PAHs studied
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Table 3. Analytical methods and concentration range reported by different researchers for PAHs in marine-sediment samples
Soxhlet extraction with dichloromethane/ methanol (v/v: 2:1) for 48 h. Clean up with silica/ alumina column chromatography
Naphthalene; Acenaphthylene; Acenaphthene; Fluorene; Phenanthrene; Anthracene; Fluoranthene; Pyrene; Benz[a]anthracene; Chrysene; Benzo[b+k]fluoranthenes; Benzo[e]pyrene; Benzo[a]pyrene; Indeno[1,2,3cd]pyrene; Dibenz[a,h]anthracene; Benzo[g,h,i]perylene
Soxhlet extraction with acetone–hexane (1:1), desulfurization with activated copper, evaporation with nitrogen, purification by liquid chromatography on Florisil cartridge, further clean up on a silica-gel column
Naphthalene; Acenaphthylene; Acenaphthene; Fluorene; Phenanthrene; Anthracene; Fluoranthene; Pyrene; Benz[a]anthracene; Chrysene; Benzo[b]fluoranthene; Benzo[k]fluoranthene; Benzo[a]pyrene; Indeno(1,2,3cd)pyrene; Dibenzo(a,h)anthracene; Benzo[ghi]perylene; 2-methylnaphthalene; 1-methylnaphthalene; 2,6dimethylnaphthalene; 1,3,5-trimethylnaphthalene; 1methylnaphthalene; Biphenyl; Benzo[e]pyrene; Perylene Naphthalene; 2-methylnaphthalene; Acenaphthylene; Acenaphthene; 2,3,5-trimethylnaphthalene; Fluorene; Phenanthrene; Anthracene; 2-methylanthracene; Fluoranthene; Pyrene; 9,10-dimethylanthracene; Benzo(c)phenanthrene; Benzo(a)anthracene; Chrysene; Benzo(b)fluoranthene; Benzo(k)fluoranthene; 7,12dimethylbenzo(a)anthracene; Benzo(a)pyrene; 3methylchloranthrene; Indeno(1,2,3-cd)pyrene; Dibenzo(a,h)anthracene; Benzo(g,h,i)perylene; Dibenzo(a,l)pyrene; Dibenzo(a,h)pyrene
Soxhlet extraction with methylene chloride. Extensive clean up by Si/Al column chromatography and HPLC with size-exclusion column
Ultrasonication with dichloromethane, shaken overnight and ultrasonicated again, centrifugation, evaporation under a nitrogen stream, clean up with solid-phase extraction column
HP-5 (50oC for 1 min, 10oC/ min to 180oC, for 7 min, 10oC/min to 230oC, for 25 min, 20oC/min to 280oC, for 5 min) PTE-5 GC-MS, EI mode (70 eV electron energy), with ion source, quadrupole and transfer line temperatures of 200C, 100C and 290C, respectively DB-5MS 1.5 min at 60C, first rate 4C/min to 300C, isothermal pause 10 min at 300C
DB5MS 60C to 310C at a rate of 15C/min, final holding time 13 min.
68–1500 ng/g d.w.
[27]
380–12,750 lg/kg d.w.
[68]
8.80–18 500 ng/d.w.
[59]
2005: 36 ± 3– 1908 ± 114 ng/g d.w. 2004: 28 ± 3– 312 ± 24 ng/g d.w.
[67]
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Fluoranthene; Pyrene; Benzo(a)anthracene; Chrysene; Benzofluoranthene; Benzo(e)pyrene; Indeono[1,2,3cd]pyrene; Benzo[ghi]perylene
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lytical conditions used in each case. As can be seen from Table 3, the HP-5 chromatographic column or equivalent is used most (in GC-MS analysis). Clean-up procedures are applied by many researchers in order to increase the sensitivity of the methods, usually by silica/alumina columns, after desulfurization by activated copper [66,67,59,68]. Some studies still utilize the traditional time-consuming, but effective, Soxhlet extraction, while HPLC instrumentation is also used in some cases with relatively good results.
4. Levels of PAHs in marine sediments PAHs are widespread contaminants in oceans. However, their concentrations in sediments of coastal bays, estuaries and continental shelves are often much higher due to greater pressures from specific anthropogenic inputs, suggesting direct influence of these sources on pollutantdistribution patterns [56]. Due to the chemical composition of seawater, the occurrence of PAHs, which are very hydrophobic compounds, will be at very low levels of concentration (<1 ng/L), in contrast to concentration in other aqueous matrices (e.g., wastewater, river water, rainwater and sediments), where, depending on the study area, the corresponding values are in the range 1 ng/g d.w. to >100 ng/g d.w. [12]. Due to their hydrophobic character, PAHs tend to adsorb onto the particulate matter so that they are transported to and accumulate in sediments. Sprovieri et al. [56] found that the priority PAHs in samples of surface sediment collected from Naples harbor were in the range 9–31,774 ng/g. Three-ring and fourring PAHs appeared dominant in the sediments studied, with median values of concentration generally greater than 60–70% of the total concentration of the PAHs. There was no correlation between grain size or total organic carbon and the distribution of PAHs or single congeners. Te Naples harbor showed a median and a range of variability greater than most other Mediterranean and European ports (e.g., Sardinia, Corsica, Spain, France and England) but similar to those measured in commercial harbors in the USA, the Middle East and Australia. Total PAH concentrations measured by Vane at al. [74] in sediments from the Mersey Estuary (UK) were in the range 626–3766 mg/kg, which were intermediate compared to other UK estuaries with similar histories of industrialization and urbanization. The molecular indices in this case suggested mainly pyrolitic inputs, augmented by a variety of industrial or petrogenic sources. A major part of recent studies on hydrocarbon pollution has focused on the northwestern part of the Mediterranean Sea [69,70], but there is a tremendous lack of information regarding the eastern Mediterranean Sea 660
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[71] and the southern Mediterranean Sea, and data are scarce for the coasts of Algeria and Egypt [72,73]. There is therefore an urgent need for better assessment of hydrocarbon contamination for the whole Mediterranean Sea.
5. Toxicity of PAHs in marine sediments The ecotoxicity of individual PAHs or mixtures of some of PAHs has been investigated using different bioassays. The most indicative toxicity end-points have been changed in line with the species tested. In one toxicity study of individual PAHs, the effect of benzo[a]pyrene (BaP) on Mytilus edulis fed daily with Isochrysis galbana algae that had previously been kept in the presence of BaP for 24 h was investigated in terms of blood-cell lysosomal membrane damage (neutral red-dye retention assay) and induction of digestive gland microsomal mixed-function oxygenase (MFO) parameters [BaP hydroxylase (BPH) and NADPH-cytochrome c (P450) reductase activities]. It was reported that the bioaccumulation of BaP in tissues (and to a lesser extent in digestive gland microsomes) of M. edulis increased with both increasing BaP and algal exposure concentrations, and over time. This pointed to impact on M. edulis in line with BaP exposure possibly due to a direct effect of BaP on the integrity of the blood-cell lysosomal membrane. An increase in NADPH-cytochrome c reductase activity was explained due to a transient response of the digestive gland microsomal MFO system to BaP exposure [74]. BaP was also reported to induce infertility in the male reproductive system. Microsomes isolated from liver and testes of rat, mouse, hamster, ram, boar, bull and monkey were incubated with BaP. Post incubation, the ethyl-acetate-extracted samples were analyzed for BaP and metabolites by reversed-phase HPLC with fluorescence detection. Given the ability of BaP 7-8-diol 9, 10epoxide, 3-, and 9-hydroxy BaP to bind with DNA and form adducts, Smith et al. concluded that a risk was likely to arise from accumulation of BaP metabolites in testicular tissues [75]. Ethoxy resorufin dealkylase (EROD) inducing potency of 10 PAHs, generally used in risk assessment studies, was measured in H4IIE rat hepatoma cell line in vitro bioassays. The responses of the PAHs investigated to EROD activity varied: anthracene (Ant) and phenanthrene (Phe) exhibited no response; naphthalene (Nap) gave no or a very weak response; fluoranthene (Fla) and benzo[ghi]perylene (BghiP) showed weak responses at the highest doses; indeno [1,2,3-cd]pyrene (IP), benz[a]anthracene (BaA), benzo[a]pyrene (BaP), chrysene (Chr) and benzo[k]fluoranthene (BkF) displayed full bell-shaped dose-response curves.
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Table 4. Toxicity reference values for Mytilus larval development [77] Compound Anthracene* Fluorene*
Fluoranthene* Phenanthrene* Pyrene* Quinoline
Maximum observed porewater concentration (lg/L)
48-h EC50 (lg/L)
Ref.
0.3 <0.1 1.1 <0.1 1.0 <0.1
4260 1260 58.8 410 >11900 4300
[83] [84] [83] [84] [83] [84]
Note: Asterisks indicate a compound identified as a COPC based on comparison of bulk sediment concentrations with numerical sediment quality criteria (BCWLAP, 2003). PAH TRVs include a 10-fold uncertainty factor to address potential interspecies differences.
BaP EROD induction equivalency factors (BaP-IEFs) were calculated and increased in the order Ant < Phe < Fla < Nap < BghiP < IP < BaA < BaP < Chr < BkF [76]. Although there has been a severe toxic outcome from individual PAHs, as reported in Table 4 for the toxicity of some individual PAHs to Mytilus galloprovincialis [77], there was found almost no dose-response relationship for PAHs in the presence of complex mixtures (e.g., sediments). A positive relationship was found in a study performed on the sediments collected in the Izmit Bay, Turkey, where bulk and elutriate samples were toxic to Phaeodactylum tricornutum tested in batch bioassays. The sediments collected from the inner sites of the bay were reported to be contaminated with Cd, Hg, As and PAHs, which were consistent with the high organic carbon contents. Bulk sediments displayed toxicity to the microalgae throughout the bay [78]. However, in a survey conducted on marine sediments in Santa Monica bay, spatial and temporal patterns indicated that toxicity was most strongly associated with the historical disposal of municipal wastewater sludge. The statistical significance between response by bioassays and chemical analyses was negative (e.g., seaurchin fertilization correlated negatively with concentration of zinc, silver, copper, cadmium, total PCBs and total PAHs, while amphipod survival had a significant negative correlation only for silver [79]). In the extracts of freeze-dried surface sediment samples from the southern North Sea (fraction <63 lm), the concentrations of the PAHs measured were in the range 2.6–200 lg/kg d.w., with the highest concentrations at near-shore locations and river mouths. Responses in the multi-bioassays toxicity tests were found to vary. The responses in Microtox and Mutatox genotoxicity assays were low with higher responses observed at near-shore locations. Relevant responses were obtained in umu-C genotoxicity and ER-CALUX assays for estrogenicity. At the oyster grounds, DR-CALUX responses appeared to be linked to the occurrence of larger PAHs (4–6 rings). By applying a non-destructive clean-up procedure, DR-CALUX responses were significantly higher than those obtained in the current protocol [80].
The 16 US EPA priority pollutant PAHs were investigated in sediments of the Niger Delta comprising 2–6ring congeners with molecular mass of 128–278. Concentrations of total PAHs in the range 20.7– 72.1 ng/g d.w. showed no correlation with the toxicity bioassays of Vibrio fischeri and Lemna minor [81]. Similar results were reported by Bihari et al. [82], who found that the organic extracts of the superficial sediments collected from the coastal area of Rovinj (Northern Adriatic, Croatia) showed toxic potential that was consistent with the sediment type, but no correlation was observed between toxicity measured by Microtox assay and concentrations of total PAHs and PAH congeners. The genotoxicity assessment of the organic extracts from the sediment showed no significant genotoxic potential in the bacterial umu-test. By using hemolymph of M. galloprovincialis, the DNA damage was positively related to total PAHs at four sampling sites, but the highest DNA damage was not observed at the site with the highest total PAH content in the sediment [82].
6. Conclusions and outlook Due to continuously increasing anthropogenic activities, management of the pollution from sediments in coastal areas is attracting more attention. PAHs are characterized by high toxicity and have been included in the priority substances of the WFD. The distributions of PAHs in the environment and potential human health risks have become the focus of much attention. Higher molecular-weight compounds are associated primarily with particles and are likely to be removed by dry deposition, while lower molecular-weight compounds are found primarily in the gas phase and are subject to transformation or removal by photochemical degradation. Their presence in marine sediments combined with other potentially toxic compounds can result in negative effects, which have yet to be investigated to any great extent, mainly due to lack of appropriate methodology and the complexity of the subject matter. At present, significant research is being devoted to optimizing ana-
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lytical methodologies for the determination of trace concentrations of PAHs in complex matrices (e.g., sediments). For the southern Mediterranean Sea, in particular, due to lack of data, there is an urgent need to assess hydrocarbon contamination.
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Trends in Analytical Chemistry, Vol. 28, No. 6, 2009 [60] J.M. Neff, Polycyclic Aromatic Hydrocarbons in the Aquatic Environment: Sources, Fates and Biological Effects, Applied Science Publishers Ltd., Essex, England, 1979, 262 pp. [61] J.W. Moore, S. Ramamoorthy, Aromatic hydrocarbons-polycyclics, in: J.W. Moore, S. Ramamoorthy, Organic Chemicals in Natural Waters: Applied Monitoring and Impact Assessment. Springer-Verlag, New York, NY, USA, 1984, pp. 67–87. [62] N.K. Nagpal, Ambient Water Quality Criteria For Polycyclic Aromatic Hydrocarbons (PAHs), Ministry of Environment, Lands and Parks, Province of British Columbia, February 1993 (http:// www.env.gov.bc.ca/wat/wq/BCguidelines/pahs/pahs-02.htm, Last accessed 7 August 2008). [63] J. Xu, Y. Yu, P. Wang, W. Guo, S. Dai, H. Sun, Chemosphere 67 (2007) 1408. [64] M.I. Bakker, B. Casado, J.W. Koerselman, J. Tolls, C. Kolloffel, Sci. Total Environ. 263 (2000) 91. [65] A.L. NÕGuessan, J.S. Levitt, M.C. Nyman, Chemosphere 55 (2004) 1413. [66] B. Men, M. He, L. Tan, C. Lin, X. Quan, Mar. Pollut. Bull. (2009) (in press). [67] A. Curtosi, E. Pelletier, C.L. Vodopivez, W.P. Mac Cormack, Sci. Total Environ. 383 (2007) 193. [68] N. Cardellicchio, A. Buccolieri, S. Giandomenico, L. Lopez, F. Pizzulli, L. Spada, Mar. Pollut. Bull. 55 (2007) 451. [69] E. Lipiatou, I. Tolosa, R. Simo´, I. Bouloubassi, J. Dachs, S. Marti, M.-A. Sicre, J.M. Bayona, J.O. Grimalt, A. Saliott, J. Albaiges, Deep Sea Res., Part II 44 (1997) 881. [70] J. Dachs, J.M. Bayona, J. Fillaux, A. Saliot, J. Albaige´s, Mar. Chem. 65 (1999) 195. [71] A. Gogou, I. Bouloubassi, E.G. Stephanou, Mar. Chem. 68 (2000) 265. [72] T.A.T. Aboul-Kassim, B.R.T. Simoneit, Mar. Chem. 54 (1996) 135. [73] T.A.T. Aboul-Kassim, B.R.T. Simoneit, Mar. Pollut. Bull. 30 (1995) 63. [74] C.H. Vane, I. Harrison, A.W. Kim, Sci. Total Environ. 374 (2007) 112. [75] T.L. Smith, S.T. Merry, D.L. Harris, J.J. Ford, J. Ike, A.E. Archibong, A. Ramesh, Toxicol. in Vitro 21 (2007) 753. [76] A.T.C. Bosveld, P.A.F. de Bie, N.W. van den Brink, H. Jongepier, A.V. Klomp, Chemosphere 49 (2007) 75. [77] B.G. McDonald, Ecotoxicol. Environ. Safe. 62 (2005) 383. [78] L.G. Tolun, O.S. Okay, A.F. Gaines, M. Tolay, H. Tufekci, N. Kirati, Environ. Int. 26 (2001) 63. [79] D. Greenstein, S. Bay, A. Jirik, J. Brown, C. Alexander, Mar. Environ. Res. 56 (2002) 277. [80] H.J.C. Klamer, P.E.G. Leonards, M.H. Lamoree, L.A. Villerius, J.E.A. Kerman, J.F. Bakker, Chemosphere 58 (2005) 1579. [81] A. Olajirea, R. Altenburger, E. Kqster, W. Brack, Sci. Total Environ. 340 (2005) 123. [82] N. Bihari, M. Fafandel, B. Hamer, B. Kralj-Bilen, Sci. Total Environ. 366 (2006) 602. [83] M.C. Pelletier, R.M. Burgess, K. Ho, A. Kuhn, R.A. McKinney, S.A. Ryba, Environ. Toxicol. Chem. 16 (1997) 2190. [84] M.C. Pillai, C.A. Vines, A.H. Wikramanayak, G.N. Cherr, Toxicology 186 (2003) 93. Anastasia Nikolaou obtained her BSc (1997) and PhD degree (2002) in the Department of Environmental Studies, University of the Aegean, Mytilene, Greece. She has been a lecturer in the Department of Marine Sciences of the University of the Aegean (Mytilene, Greece) since 2005. Her research interests include analytical methods for the determination of toxic priority substances in water and sediment, water quality and treatment/disinfection by-products, investigation of fate and toxicity of organic compounds in the environment. She has participated in many related national and international projects, authored or co-authored a number of refereed papers and technical reports, and has been editor of two books and guest editor for Desalination Journal.
Trends Maria Kostopoulou obtained her BSc (1978) in Biochemistry in the University of Paris VII (France), and her BSc (1981) in Chemistry and PhD degree (1987) in the Department of Chemistry, University of Athens, Greece. She has been an external Scientific Researcher in the Department of the Environment of the University of the Aegean since 1992 and she has been working as an Assistant Professor in the Department of Marine Sciences of the University of the Aegean (Mytilene, Greece) since 2001. Her research interests include analytical methods for the determination of priority substances in environmental matrices (wastewater, water and sediments), as well as toxicity evaluation of individual compounds and mixtures by the use of marine microalgae. She has participated in many related national and international projects. Giusy Lofrano holds BS and MSc Degrees (2003) in Civil Engineering and PhD degree (2007) in Environmental Engineering from University of Salerno, Italy. She is currently working on the application of Fenton oxidation, photo-Fenton oxidation and ozonation to high-strength wastewaters, particularly in the leather-tanning industry, to achieve an economically feasible and environmentally friendly integrated wastewater treatment. Her research interests are in the fields of biological treatment of wastewater, advanced oxidation processes, and industrial pollution control. She has been co-author of articles on constructed wetlands and advanced oxidation processes, she has participated in related national and international projects and she is currently assisting the didactic activity of Wastewater Treatment Plant-I and II courses at the Engineering Faculty of University of Salerno. Sureyya Meric¸ holds BS, MSc and PhD degrees in Environmental Engineering from Istanbul Technical University (Istanbul, Turkey). She has been working in this field for 20 years. Her post-doc studies were focused on ecotoxicology of many chemicals and complex mixtures. She has developed expertise in chemical and biological treatment of wastewater, wastewater reuse, water treatment and disinfection, toxicity of disinfection by-products, activated sludge modeling, inhibition, industrial pollution control, water quality and management, aquatic and sediment toxicity monitoring. Impact assessment of priorityemerging pollutants, xenobiotics and their removal by advanced oxidation processes, and groundwater remediation. She has been investigator or co-ordinator of many national and international projects, and a member of environmental organizations, committees and the editorial boards of international journals. Andreas Petsas holds BS (1998) and PhD (2006) degrees in Environmental Studies from the University of the Aegean, Faculty of the Environment. He has worked on kinetics of removal (hydrolysis, photodegradation) of organophosphorous insecticides from various environmental matrices, as well as on the toxicity of these compounds to marine algae. He has participated in several national and international projects on these subjects. Since 1999, he has been a scientific collaborator in the Water and Air Quality Laboratory of the Department of Environmental studies of the University of the Aegean, and, since 2006, he is an adjunct lecturer at the Department of Marine Sciences of the University of the Aegean. His research interests include development and optimization of gas-chromatography analytical methods for determination of organic pollutants in environmental samples and he has co-authored several papers in international scientific journals. Maria Vagi holds a BS degree in Chemistry (University of Thessalonica, 1997), and a PhD degree (2007) in Environmental Studies, University of the Aegean, Faculty of the Environment. She has worked on kinetics of removal (hydrolysis and photodegradation) of several categories of insecticides and organic pollutants from various environmental matrices, as well as on the toxicity of these compounds to marine algae. She has participated in several national and international projects on these subjects. She has been a researcher in the Water and Air Quality Laboratory from 1997 until 2007, and, since
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2000, she has been working as a laboratory-technical staff member at the Department of Marine Sciences. Her research interests include development and optimization of gas-chromatography analytical
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methods for determination of organic pollutants in environmental samples and she has co-authored several papers in international scientific journals.