Chapter 5 

Polycyclic Aromatic Hydrocarbons 

a Constituent of Petroleum: 

Presence and Influence in 

the Aquatic Environment 

Daniela M. Pampanin and Magne O. Sydnes 

Additional information is available at the end of the chapter 

http://dx.doi.org/10.5772/48176 

1. Introduction 

Crude oil is a complex mixture of hydrocarbons containing more than 17000 compounds [1]. 

Among the constituents of crude oil there is a group of substances called polycyclic aromatic 

hydrocarbons (PAHs). PAHs are aromatic compounds containing from two to eight 

conjugated ring systems. They can have a range of substituents such as alkyl, nitro, and 

amino groups in their structure [2]. Nitrogen, sulfur, and oxygen atoms can also be 

incorporated into their ring system [2,3]. The precursors for PAHs found in crude oil are 

natural products, such as steroids, that have been chemically converted to aromatic 

hydrocarbons over time [4]. 

The PAHs that are present in the marine environment in relevant concentrations are 

divided into two groups depending on their origin, namely pyrogenic and petrogenic [5]. 

Pyrogenic PAHs are formed by incomplete combustion of organic material while the 

petrogenic PAHs are present in oil and some oil products [4,6,7]. In general the pyrogenic 

PAHs are composed of larger ring systems then the petrogenic PAHs. Sources for 

pyrogenic PAHs are forest fires [6,7,8], incomplete combustion of fossil fuels [6,7,8], and 

tobacco smoke [6,7]. A range of PAHs are naturally present in crude oil [4,9,10] and coal 

[10,11] and these compounds are referred to as petrogenic PAHs. In the costal zones PAHs 

enters the water primarily from sewage, runoff from roads [12], the smelter industry 

[13,14,15] and oil spills [16,17], while offshore PAHs chiefly enter the water through 

oil seeps [18], oil spills [16], and produced water discharge from offshore oil installations 

[19]. 
84 Hydrocarbon 

2. Oil as a source of polycyclic aromatic hydrocarbons to the aquatic 

environment 

Hydrocarbons in the form of crude oil, and therefore also PAHs, have and are entering the 

environment naturally through oil seeps, which is oil leaking naturally from oil reservoirs. 

Oil seeps are found scattered all over the globe with a higher concentration in certain 

regions of the world [18]. Numerous times the presence of a natural oil seep has resulted in 

the discovery of oil reservoirs that are large enough for commercial oil production [20], 

however, the presence of an oil seep does not guarantee the discovery of a production 

worthy oil reservoir [21]. Oil seeps vary in size with macroseeps resulting in visible oil slicks 

on the water surface, when the oil seep is situated on the seafloor, and microseeps that are 

invisible at the surface [22]. Estimates for the world-wide seepage rate vary between 0.02-2.0 

x 106 tons per year with the most realistic estimate being 0.2 x 106 tons per year [23]. The 

presence of natural oil seeps also results in local presence of hydrocarbon-eating 

microorganisms [24,25], a fact that gives these regions an advantage in the case of an 

accidental oil spill [24,26]. For example resent research showed that the presence of oil 

eating microorganisms due to natural oil seeps in the Gulf of Mexico in addition to 

favorable water currents resulted in a quicker degradation of oil in the region after the 

Deepwater Horizon accident than otherwise would be expected [26,27]. As expected, it is the 

lighter fractions, viz. short chain alkanes, of the oil that are first degraded by 

microorganisms [28]. The easily accessible energy source, short chain alkanes, results in an 

explosion like increase in the number of the oil degrading microbes. After some time these 

microorganisms also start degrading the more complex molecules such as long chain 

alkanes and aromatic compounds like PAHs [29]. A range of PAH degrading bugs has been 

found naturally in the environment [30-34]. These microorganisms have been isolated, 

sequenced and studied extensively in the laboratory, and their mechanism of degrading 

PAHs is fairly well understood [29]. 

For monitoring purposes The US Environmental Protection Agency has made a list of 16 

unsubstituted PAHs that are on a priority pollutant list [35]. These PAHs are usually 

referred to as the EPA 16 PAHs (Figure 1) and are the PAHs most commonly analyzed for. 

The concentration of these 16 PAHs and other PAHs, and the ratios between the various 

compounds differ from oil to oil [36,37]. This fact is quite clearly presented in the work of 

Kerr et al. where they have analyzed the concentration of the prioritized 16 PAHs in 48 

crude oils from around the globe (North America, South America, Africa, and Asia) [36]. 

Their results are summarized in Table 1 and show an enormous variation of the content of 

the various PAHs in crude oil from different sites. Crude oil from the North Sea is 

reported to have a PAH concentration of 0.83% [38] while for example crude oil that 

leaked out of Exxon Valdez (referred to as Exxon Valdez crude oil hereafter) had a PAH 

content of 1.47% [39]. 

Naphthalene, one of the 16 EPA PAHs, is present in the highest concentration in crude oil, 

nevertheless this quantity does not give the full details of the naphthalene content of the oil. 

In fact, if the content of methylated naphthalene, a normal constituents of crude oil, is also 
Polycyclic Aromatic Hydrocarbons a Constituent of Petroleum: 

Presence and Influence in the Aquatic Environment 85 

included in the analysis the total naphthalene concentration is much higher as illustrated in 

Table 2. A similar situation is also found for phenanthrene [38,39] and chrysene [38], while 

for the less abundant PAHs in crude oil excluding the alkylated derivatives of the parent 

compounds does not alter the total quantity significantly. In fact the alkylated PAHs found 

in crude oil have been reported to be more toxic then their unsubstituted congeners [41-43]. 

Crude oil 48 different crude oils [36] North Sea [38] Goliat [40]b 

PAH Minimum

mg/kg oil

Maximum

mg/kg oil

Mean 

mg/kg oil

mg/kg oil mg/kg oil 

Naphthalene 1.2 3700 427 1169 1030 

Acenaphthene 0 58 11.1 18 12 

Acenaphthylene 0 0 0 11 * 

Fluorene 1.4 380 70.34 265 75 

Anthracene 0 17 4.3 1.5 * 

Phenanthrene 0 400 146 238 175 

Fluoranthene 0 15 1.98 10 6 

Pyrene 0 9.2 - 20 * 

Benzo[a]anthracene 0 16 2.88 11 Nac

Chrysene 4 120 30.36 26 Nac

Benzo[b]fluoranthene 0 14 4.08 4.2 Nac

Benzo[k]fluoranthene 0 1.3 0.07 Nda Nac

Benzo[a]pyrene 0 7.7 1.5 1.3 Nac

Dibenz[a,h]anthracene 0 7.7 1.25 Nda Nac

Benzo[g,h,j]perylene 0 1.7 0.08 1 Nac

Indeno[1,2,3-cd]pyrenec 0 1.7 0.08 Nda Nac

Table 1. Minimum, maximum, and mean PAH content in 48 different crude oils [36], and PAH content 

in two North European crude oils, North Sea crude oil [38], Goliat crude oil [40] (aNd = not detected; 

bGoliat is situated in the Barents Sea; cNa = not analyzed for). 

Crude oil North Sea [38] Goliat [41] Exxon Valdez crude oil [39] 

PAH mg/kg oil mg/kg oil mg/kg oil 

Naphthalene 1169 1030 720 

1-Methylnaphthalene 2108 2700 1330 

2-Methylnaphthalene 2204 4200 1020 

3-Methylnaphthalene 1172 2800 Naa 

Table 2. Concentration of naphthalene and the three regioisomers of methylnaphthalene in crude oil 

from three different locations (aNa = not analyzed). 

PAHs represent the group of compounds in oil that has received the greatest attention due 

to their carcinogenic and mutagenic properties [38,44]. More precisely, intermediates are 

formed that are far more toxic than the mother compounds during cellular detoxification of 

the PAHs in vivo [45]. Due to the toxicity of the PAH metabolites the oil industry in some 
86 Hydrocarbon 

areas of the world (e.g. North Sea, Mediterranean Sea, Australian Northwest Shelf, Gulf of 

Mexico) are required to monitor their discharges to the aquatic environment [46-49].In the 

North Sea this is taken care of through the Water Column Monitorin programs [48]. For 

monitoring purposes the 16 EPA compounds has been chosen as the most important PAHs 

to analyze for. 

Figure 1. Chemical structure of the EPA selected 16 PAHs. 

Unfortunately, accidental oil spills take place from time to time, most recently the Deapwater 

Horizon accident in the Mexican Gulf in 2010 (April 20th). This incident resulted in the release of 

779 million liters of crude oil to the Gulf [50]. The Exxon Valdez oil spill in Prince William 

Sound, Alaska, USA in 1989 (March 24th), which took place after the oil tanker ran ashore, 

resulted in the release of 42 million liters of crude oil [16]. In European waters the Erika

accident in 1999 (December 12th) released 18000 tons of crude oil into French coastal waters 

and in Spain Prestige spilt 60000 tons of heavy fuel oil into the waters outside of Galicia 

in 2002 (November 13th) [17]. These are unfortunately only a few examples of accidental release 

of large quantities of oil to the aquatic environment. The environmental consequences of the 

Exxon Valdez spill is the most studied oil spill ever [51,52], however, the influence of the 

Deapwater Horizon oil release will probably be just as well studied, or even more studied [53]. 

3. Produced water as a source of polycyclic aromatic hydrocarbons to the 

aquatic environment 

Oil production offshore (Figure 2) (and onshore) also results in the production of large 

volumes of water, so called produced water (PW), in addition to crude oil. PAHs contained in 
Polycyclic Aromatic Hydrocarbons a Constituent of Petroleum: 

Presence and Influence in the Aquatic Environment 87 

PW are receiving much attention due to their potential for causing adverse effects in the 

marine environment [54,55]. PW generally consists of a mixture of: 1) formation water 

contained naturally in the reservoir; 2) injected water used for the recovery of oil; and 3) 

treatment chemicals added during production [19]. Data from offshore oil production 

platforms in the North Sea have showed that the major aromatic compounds in PW are BTEX 

(benzene, toluene, ethylbenzene and xylene) (97%), 2- and 3-ring PAHs (3%) named NPD 

(naphthalenes, phenanthrenes and dibenzothiophenes) and larger PAHs (<0.2%) [19,56,57]. 

The average concentrations of NDPs and the remaining 14 PAHs from the EPA 16 list are 

shown in Table 3. With regard to the NPD group, the lowest molecular weight substance, 

naphthalene represents the largest fraction (92%). Even PAHs with more than 3-rings 

constitute less than 0.2% of the aromatic content of PW, with the concentrations of the 

individual compounds decreasing with the number of rings in their structure, by about one 

order of magnitude for each additional ring. 

16 PAHs from 

the EPA list 

Oil installation Gas installation Unspecified 

installation 

Naphthalene 145 115 108 

Phenanthrene 13.6 20.9 10.7 

Fluorene 8.3 13.1 6.7 

Acenaphthene 2 50.1 1.78 

Acenaphthylene 0.86 12.6 2.35 

Fluoranthene 0.26 35.4 0.29 

Anthracene 3.74 110 1.17 

Pyrene 0.63 8 0.47 

Benzo[a]pyrene 0.52 - 0.022 

Chrysene 0.84 1 0.52 

Benzo[a]anthracene 0.23 1 0.25 

Benzo[b]fluoranthene 0.028 - 0.031 

Benzo[k]fluoranthene 0.007 - 0.007 

Dibenz[a,h]anthracene 0.005 - 0.005 

Benzo[g,h,j]perylene 0.029 - 0.019 

Indeno[1,2,3-cd]pyrenec 0.005 - 0.006 

Table 3. Mean concentrations of aromatic compounds in produced water (expressed as μg/L) (data 

from OLF, UKOOA and OGP companies, [58]). 

PAHs in PW are present in both dissolved and dispersed (oil droplets) forms. Existing 

oil/water separators, such as hydrocyclones, are efficient in removing droplets, but not 

dissolved hydrocarbons from PW. Therefore, much of the PAHs discharged in to the marine 

environment are dissolved low molecular weight compounds. Nevertheless, since the 

treatment procedures are not 100% effective, discharged PW still contains some dispersed 

oil (droplet size from 1 to 10 μm) [59]. Concentration of total PAHs in PW typically range 

from 0.040 to 3 mg/L (Table 4) and consist primarily of the most water-soluble congeners 
88 Hydrocarbon 

such as naphthalene, phenanthrene and their alkylated homologues (2- and 3-ring PAHs). 

The abundance of these alkyl substituted PAHs is also higher than for the parent 

compounds (the non-alkylated homologues). Higher molecular weight PAHs (up to 6-ring) 

are in fact rarely detected in properly treated PW [60,61]. The ratio between alkylated-PAHs 

and the corresponding parent compounds is therefore used to confirm the nature of the 

pollution source in field studies [54,62]. 

Due to their lipophilic properties, PAHs are mainly associated with dispersed oil droplets 

[59,63]. In fact, it has been documented that up to 10% of the total PAHs in PW from a 

platform on the Northwest Shelf of Australia were in the dissolved fraction being formed 

mainly by alkylnaphthalenes and alkylphenanthranes [64]. Moreover, these droplets also 

contained almost all the dibenzothiophenes, fluoranthenes/pyrenes and chrysenes present in 

the PW. 

Figure 2. The sources of PAHs entering the marine environment offshore are predominantly natural oil 

seeps, oil spills from boats or platforms, and produced water discharge from oil and gas producing 

installations such as the one shown. PAHs in produced water and oil seeps represent a chronic release 

to the marine environment.

As a chronic source of PAH contamination in the marine environment, PW is also a source 

of concern with respect to possible long term impact on the environment [65]. Estimates of 

the PW discharge volumes on the Norwegian shelf predict an increase until 2010–2014, 

reaching a maximum of about 200 million L/year [66]. Therefore, this offshore discharge is 

currently under periodical monitoring [48,67]. 
Polycyclic Aromatic Hydrocarbons a Constituent of Petroleum: 

Presence and Influence in the Aquatic Environment 89 

Compound Gulf of 

Mexico 

North Sea Scotia 

Shelf 

Grands Bank 

(Canada) 

Naphthalene 5.3-90.2 237-394 1512 131 

C1-Naphtalenes 4.2-73.2 123-354 499 186 

C2-Naphtalenes 4.4-88.2 26.1-260 92 163 

C3-Naphtalenes 2.8-82.6 19.3-81.3 17 97.2 

C4-Naphtalenes 1.0-52.4 1.1-75.7 3.0 54.1 

Acenaphthylene ND-1.1 ND 1.3 2.3 

Acenaphthene ND-0.1 0.37-4.1 ND ND 

Biphenyl 0.36-10.6 12.1-51.7 ND ND 

Fluorene 0.06-2.8 2.6-21.7 13 16.5 

C1-Fluorene 0.09-8.7 1.1-27.3 3 23.7 

C2-Fluorene 0.20-15.5 0.54-33.2 0.35 4.8 

C3-Fluorene 0.27-17.6 0.30-25.5 ND ND 

Anthracene ND-0.45 ND 0.26 ND 

Phenanthrene 0.11-8.8 1.3-32.0 4 29.3 

C1-Phenanthrene 0.24-25.1 0.86-51.9 1.30 45.0 

C2-Phenanthrene 0.25-31.2 0.41-51.8 0.55 37.1 

C3-Phenanthrene ND-22.5 0.20-34.3 0.37 24.4 

C4-Phenanthrene ND-11.3 0.50-27.2 ND 13.2 

Fluoranthene ND-0.12 0.01-1.1 0.39 0.51 

Pyrene 0.01-0.29 0.03-1.9 0.36 0.94 

C1-Fluoranthene/Pyrenes ND-2.4 0.07-10.3 0.43 5.8 

C2-Fluoranthene/Pyrenes ND-4.4 0.21-11.6 ND 9.1 

Benzo[a]anthracene ND-0.20 0.01-0.74 0.32 0.60 

Chrysene ND-0.85 0.02-2.4 ND 3.6 

C1-Chrysenes ND-2.4 0.06-4.4 ND 6.3 

C2-Chrysenes ND-3.5 1.3-5.9 ND 18.8 

C3-Chrysenes ND-3.3 0.68-3.5 ND 6.7 

C4-Chrysenes ND-2.6 ND ND 4.2 

Benzo[b]fluoranthene ND-0.03 0.01-0.54 ND 0.61 

Benzo[k]fluoranthene ND-0.07 0.006-0.15 ND ND 

Benzo[e]pyrene ND-0.10 0.01-0.82 ND 0.83 

Benzo[a]pyrene ND-0.09 0.01-0.41 ND 0.38 

Pyrelene 0.04-2.0 0.005-0.11 ND ND 

Indeno[1,2,3-cd]pyrenec ND-0.01 0.022-0.23 ND ND 

Dibenz[a,h]anthracene ND-0.02 0.012-0.10 ND 0.21 

Benzo[g,h,j]perylene ND-0.03 0.01-0.28 ND 0.17 

Total PAHs 40-600 419-1559 2148 845 

Table 4. Concentrations of individual PAHs or alky congener groups in produced water from various 

areas (expressed as μg/L) (ND = not detected) (from [61]). 
90 Hydrocarbon 

4. Environmental monitoring of polycyclic aromatic hydrocarbons 

There are various environmental monitoring methods which may be performed in order to 

assess risks of PAH contamination for organisms and to classify the environmental quality 

of ecosystems. Five approaches are hereby reported: 1) chemical monitoring: exposure 

assessment by measuring levels of a selected set of compounds in abiotic environmental 

compartments; 2) bioaccumulation monitoring: exposure assessment by measuring PAH 

levels in biota or determining the critical dose at a critical site (bioaccumulation); 3) 

biological effect monitoring: exposure and effect assessment by determining the early 

adverse alterations that are partly or fully reversible (biomarkers); 4) health monitoring: 

effect assessment by examining the occurrence of irreversible diseases or tissue damage in 

organisms; 5) ecosystem monitoring: assessment of the integrity of an ecosystem by making 

an inventory of, for instance, species composition, density and diversity [68-70]. All these 

methods are currently in use to monitor the aquatic environment contamination from PAH 

compounds. 

Since the occurrence and abundance of PAHs in aquatic environments represent a risk to 

aquatic organisms and ultimately to humans (through fish and shellfish consumption), there 

is a constant need for their determination and quantification around the world [71]. The 

monitoring of PAH presence in the aquatic environment is therefore a world-wide activity. 

Since some of these compounds are well known carcinogens and mutagens [44,72], this 

contaminant class has been generally regarded as high priority for environmental pollution 

monitoring. In fact, the European Union included these pollutants in the list of priority 

hazardous substances for surface waters in the Water Framework Directive 2000/60/EC [73]. 

Moreover, for PAH content in biota several guidelines exist: the commission regulation (EC 

2005) which stipulates a maximum concentration of 10 μg/kg (w/w) of benzo[a]pyrene in 

edible molluscs (this regulation also limits the benzo[a]pyrene concentration in other 

alimentary products), the OSPAR Commission which has developed eco-toxicological 

assessment criteria (i.e., concentrations levels above which concern is indicated) for different 

PAHs in fish and mussels [74], and the Oregon Health Division which has derived riskbased criteria for PAHs [75] assessing the risk in terms of benzo[a]pyrene equivalents. 

Many studies have been carried out to determine the PAH distribution in marine organisms 

in different geographical areas using different approaches. Different authors have reported 

the presence of PAHs in waters, marine organisms, and sediments using chemical and 

biological markers [76-78]. Current monitoring techniques employed to determine 

environmental quality include the chemical analyses of sediment and water samples for 

determining the concentration of PAH parent compounds. PAHs are sparingly soluble in 

water and are difficult to detect although they show a much greater association with 

sediments. In fact, due to their hydrophobic character, these compounds rapidly tend to 

become associated with particles and end up in the sediments which act as a sink for them 

[71,79,80]. High concentrations of pyrogenic PAH mixtures in sediment samples have been 

found in several freshwater, estuarine and marine regions with heavy vessel traffic or at 

locations with PAH-containing effluents from industrial areas [81]. Finally, chemical 
Polycyclic Aromatic Hydrocarbons a Constituent of Petroleum: 

Presence and Influence in the Aquatic Environment 91 

analyses of these materials are laborious and costly due to the lengthy extraction 

methodology needed. Since monitoring of a large number of PAHs is expensive, time 

consuming and analytically demanding, a selection of compounds could be monitored to 

give an indication of the overall contamination. 

Standard procedure for determining PAHs in the aquatic environment have a long history 

that covers at least three decades. A few examples are reported here (see also Table 5). One 

of the first official procedures was published in 1985 by HMSO (Her Majesty’s Stationery 

Office) and provided the analysis of six PAHs in water samples [82]. In 1986, the EPA 

published EPA procedure 8310 for determining PAHs by HPLC [83]. In 2007, the EPA 

published the EPA procedure 8270D for determination of 16 PAHs applying gas 

chromatography coupled with mass spectrometry (GC-MS) [84]. To close, the standard 

method for PAH determination in water recommended by the International Organization 

for Standardization (ISO), ISO 17993:2002, is based on liquid-liquid extraction and final 

determination by HPLC [85]. 

Method 

name Application Sample preparation Determination technique 

EPA-610 

PAHs in municipal 

and industrial 

wastewaters 

About 1L of water samples is 

extracted with dichloromethane, 

the extract is then dried and 

concentrated to a final volume of 

less than 10 mL. 

HPLC or GC (with a 

packed column) 

EPA-8310 Groundwater and 

wastewater 

EPA 3500 series methods (version 

3, 2000) HPLC 

EPA-8270D Aqueous samples EPA 3500 series methods (version 

3, 2000) 

GC-MS with use of 

high resolution 

capillary columns and 

deuterated internal 

standard 

ISO 

17993:2002 

Water quality 

(determination of 

15 PAHs) 

Water samples collected in brown 

glass bottles are stabilized by 

adding sodium thiosulfate. 1 L of 

sample is extracted using hexane. 

The extract is dried with sodium 

sulfate and then enriched by 

removal of hexane by rotary 

evaporation 

HPLC and 

fluorescence detector 

Table 5. Standard procedures for determining of PAHs in aqueous samples (modified from [86]). 

Modern analyses have at their disposal a wide spectrum of tools, specialized analytical 

equipment and standard procedures that are capable of providing high quality results even 

if intercalibration studies and commonly accepted SOP are needed [86]. 
92 Hydrocarbon 

Chemical fingerprinting of a PAH mixture can lead to the identification of the 

contamination source, for example an oil spill accident [87]. PAH compounds in a 

petrogenic exposure often contain one or more methyl- ethyl- or butyl-(and sometimes 

higher alkyl-) groups on one or more of the aromatic carbons [88]. For example, the ratio 

between alkyl-PAH compounds and the corresponding parent compounds is commonly 

used to confirm the nature of the pollution source in field studies from PW discharges (i.e. 

the abundance of alkyl substituted PAHs is higher than for the parent compounds (the nonalkylated homologues)). It has been shown that mussels caged down-stream of PW 

discharges from oil platforms accumulate higher concentrations of alkylnaphthalenes, 

alkylphenanthrenes and alkyldibenzothiophenes, than their respective parent compounds 

[62]. In a recent study, the ratio of alkylated over non-alkylated PAHs indicated a diffuse 

petrogenic contamination in the Ekofisk area (up to 2000 m from PW fallout) [55]. 

PAHs co-occur in various amounts and the composition of the mixture differs depending 

upon the source from which they are derived and their subsequent degradation. For 

example a recent study clearly demonstrated the correlation between PAH contamination in 

river water and the nearby activity of textile factories [89]. A widespread PAH 

contamination in the San Francisco Bay has been reported after a 10 year monitoring of 

water, sediment and biota samples [90]. Since the process of industrialization and 

urbanization is growing rapidly in South America, the potential increase in PAH 

contamination is under evaluation through a number of surveys in the coastal areas, 

collecting information about the PAH concentration in water, sediment and biota [91-93]. Of 

course, oil spills are well-known examples of PAH contamination in the aquatic 

environment. Unfortunately, many cases of studies are reported in the literature from 

different parts of the world, such as the Gulf of Mexico in [94], the Mediterranean Sea [95], 

the Spanish coasts [96], and the Philippines islands [97]. 

PAH contamination is also constantly under the attention of different countries in relation to 

oil and gas explorations [98]. For example the Water Column Monitoring program financed 

by Oljeindustriens Landsforbund (OLF), has provided information about PAH 

contamination from platform discharges in the North Sea since 2001, almost yearly based 

[48]. This monitoring program is a good example of integration of chemical monitoring 

methods with biological effect monitoring approaches in PAH monitoring. 

Passive sampling devices, such as semipermeable membrane devices (SPMDs) and polar 

organic chemical integrative samplers (POCIS) also provide a useful contribution to the 

monitoring of PAH contaminants in the aquatic environment [99-101]. The principle of the 

passive sampling technique is the placement of a device in the environment for a fixed 

period of time, where it is left unattended to accumulate contaminants by diffusive and/or 

sorptive processes. The main advantages of using passive sampling devices over traditional 

discreet spot water samples are: 1) concentrations are time-integrative during exposure, 

compensating for fluctuations in discharges; 2) lower detection limits are normally 

achievable as a larger sample has been taken and; 3) only the freely dissolved and thus more 

readily bioavailable fraction is measured. Furthermore, there is a reduction in the need for 

the use of animals in scientific experiments. However, relating passive sampling device 
Polycyclic Aromatic Hydrocarbons a Constituent of Petroleum: 

Presence and Influence in the Aquatic Environment 93 

accumulations to the overall ecological relevance of contaminants is complicated and effects 

can only be inferred. 

Nevertheless, the chemical approach does not provide information about PAH 

bioavailability, and the toxic potential and the risk posed by the potentially much more toxic 

daughter compounds produced by many organisms as a result of metabolism and 

biotransformation of the parent chemicals is not taken into account [102]. Therefore, a 

biological effect monitoring approach using living organisms for PAH environmental 

monitoring has been developed during the last couple of decades. Metabolites arising from 

biotransformation processes maybe concentrated in body fluids, tissues or excreta and the 

analysis of such biological compartments provides an opportunity to detect and measure 

exposure of organisms to bioavailable contaminants [103]. On the other hand, PAHs are 

known to induce toxic effects at the individual level [104,105], and integration of chemical 

analyses with biomarker responses in organisms has been recommended for monitoring of 

PAH contamination, in particular in oil related activities (e.g. offshore exploitation activities) 

[48]. The feasibility of using tissue concentrations of PAH compounds in marine species as a 

marker of environmental contamination depends on the relative rates of uptake, 

biotransformation and excretion of the organism. Invertebrate filter feeders, such as Mytilus

spp., are highly efficient accumulators and bioconcentrators of PAHs and therefore 

commonly used [54,106-108]. Total amounts of the EPA 16 PAHs between 10 μg/kg and 20 

μg/kg, have been found in mussels caged in the vicinity of Norwegian platforms [48,54,109]. 

Fish and other invertebrates, rapidly biotransform PAHs and their presence in tissues is low 

and no representative of the overall contamination. PAHs related to offshore operational 

discharges in fact are generally not found in muscle of wild specimens of fish collected in 

regions with oil and gas activity [110]. 

Some biomarkers are widely used as sensitive and early warning signals of exposure to 

PAHs. For example, many studies indicated that PAH compounds were detectable several 

kilometers away from North Sea oil production platforms using in vitro bioassays and 

biomarkers [111]. Currently the induction of ethoxyresorufin-O-deethylase (EROD) activity, 

the production of bile metabolites and the formation of DNA adducts have shown the 

greatest potential for identifying level of exposure to PAHs following contamination of the 

aquatic environment [87,112-114]. The historically commonly used marker is the induction 

of Cytochrome P450 1A (CYP1A) in fish measured by the catalyzed O-deethyulation of 

ethoxyresorufin in hepatic microsomes [104,115-118]. The induction of CYP1A in fish 

following exposure to certain classes of organic contaminants has been the basis of the use of 

the cytochrome P450 system as a biomarker in pollution monitoring since the 70’s [104]. 

Many studies reported the used of this biomarker in various monitoring surveys since the 

‘80s [104,119,120]. In many cases, EROD activities or CYP1A protein levels were correlated 

with environmental levels of CYP1A-inducing chemicals such as PAHs. Nevertheless, a 

linear dose–response relationship cannot always be found between the PAH concentration 

and the CYP1A content and/or activity in the natural environment, where a mixture of both 

inducers and inhibitors of CYP1A may act simultaneously [121]. Moreover, other factors 

(e.g. temperature, season or sexual hormones) can also modulate the responsiveness of the 
94 Hydrocarbon 

CYP1A system in fish [122]. Since PAH compounds are absorbed via gills and may be 

metabolized before reaching the liver, hepatic EROD activity may not be the only and most 

sensitive organ to reflect the presence of CYP1A inducing agents (such as PAHs) in water. 

Therefore, a sensitive method to determine EROD activity in gill filaments has also been 

developed [123]. 

A very efficient tool to assess PAH exposure in fish is the determination of PAH metabolites 

in fish bile. They can be measured using several analytical methods from the simple and fast 

fluorescence assay (fixed fluorescence detection or synchronous fluorescence spectrometry) 

to the HPLC with fluorescence detection (HPLC-F) after deconjugation, extraction and 

derivatization of the bile samples, to the extremely sensitive and advanced LC-MS/MS and 

GC-MS/MS methods. These methods are very different both in regard to their analytical 

performances towards different PAH metabolites as well as in technical demands and 

monitoring strategies. A recent review reported the state of the art for the different methods 

for determining metabolites of PAH pollutants in fish bile [87]. This approach has also been 

developed for crustaceans. Metabolites in crabs urine have been analyzed to monitor 

environmental contamination from PAH with success [103,124]. Regarding DNA adduct 

analysis as a biomarker of exposure to PAHs, a description of methods is reported in the 

section DNA adducts from polycyclic aromatic hydrocarbons. 

5. Polycyclic aromatic hydrocarbon metabolites 

The strongly hydrophobic PAHs accumulate in fatty tissue such as liver, where they 

penetrate the cells by means of passive diffusion. Inside the hepatocytes, PAHs are oxidized 

and hence made more water-soluble and more reactive by enzymes with aryl hydrocarbon 

hydroxylase activity to form epoxides and diols according to the general route outlined in 

Figure 3 [125]. CYP1A is the most important and best described aryl hydrocarbon 

hydroxylase enzyme participating in the first step (Phase I) of xenobiotic detoxification 

[126,127]. CYP1A-derived metabolites generally have a high affinity to nucleic acids and 

proteins, which may result in adduct formation and possible impaired function of these 

biomolecules. 

Figure 3. General outline of the metabolic degradation of PAHs. 

PAH PAH

[O]-enzyme

[O]-enzyme

H2O-enzyme

H+ H+

O

PAH

OH

OH

PAH

OH

OH

O

PAH

PAH

OH OH
Polycyclic Aromatic Hydrocarbons a Constituent of Petroleum: 

Presence and Influence in the Aquatic Environment 95 

The vast majority of Phase I-derived PAH metabolites are passed on to Phase II (most 

notably glutathione transferase) and Phase III (most notably the transmembrane ATPbinding cassette exporter proteins) to eventually become excreted with the bile fluid [126]. 

In this context, the presence of PAH metabolites in the bile of fish is a highly regarded 

biomarker of recent PAH exposure (exposure that has taken place within a few days prior to 

sampling and analysis) (vide supra) [37,87].

The various PAHs form a range of metabolites in vivo. Most of the studies mapping these 

metabolites have been performed on humans, rats, mice, and hamsters [128]. However, 

some studies have been conducted on fish and/or fish cells. The results from these studies 

show that the point of oxidation varies from species to species for the same PAH due to the 

presence of various cytochrome P450 (CYP) isoforms [128]. For fish it is predominantly 

CYP1A’s role in PAH metabolism that has been studied, however, for humans, in particular, 

other isoforms of CYP have also been well investigated. For a general discussion regarding 

the various CYP families found in fish, see Uno et al. [129]. Figure 4 summarizes the data 

reported for oxidation site for a range of PAHs, phenanthrene [130,131], chrysene [131,132], 

pyrene [130,133], benz[a]pyrene [131,134], benzo[c]phenanthrene [135,136], and 

dibenzo[a,l]pyrene [135-137], based on in vitro tests with fish CYP1A. Most of the PAHs 

shown in Figure 4 have predominantly one major site where oxidation takes place, thus, 

indicating a high regioselectivity in the enzymatic oxidation by CYP. 

Figure 4. Point of oxidation with distribution intervals. 

The absolute stereochemistry of PAH metabolites, derived from metabolism of PAHs in the 

liver, influences the toxicity of the metabolite [138,139]. In particular the diols with R,Rconfiguration and the R,S-diol-S,R-epoxides show high carcinogenic activity [139]. As 

shown in Figure 4 for chrysene the predominant diol formed is the 1,2-diol (formed in 58%), 

with 3,4- and 5,6-diol being formed in 24% and >1%, respectively (structures for the diols are 

shown in Figure 5) [132]. The data for chrysene depicted in Figure 4 were based on in vitro

tests utilizing liver microsomes from brown bullhead, however, these findings were later 

confirmed by Jonsson et al. in in vivo tests with Atlantic cod [140]. Close to 90% of the 

chrysene 1,2-diol is formed with the R,R-configuration (structure shown in Figure 5) and 
96 Hydrocarbon 

slightly more than 10% is formed with the S,S-configuration [132]. For the 3,4-diol 97% is 

formed with the R,R-configuration (structure shown in Figure 5) [132], thus indicating that 

in the fish brown bullhead predominantly the most toxic metabolites are formed. The diols 

derived from chrysene are only considered to have weak carcinogenic activity [141-143], 

however, 1,2-dihydroxy-1,2-dihydrochrysene is the starting point for the biosynthesis of the 

most carcinogenic chrysene metabolite 1,2-dihydroxy-3,4-epoxy-1,2,3,4-tetrahydrochrysene 

[144]. 

Figure 5. Chemical structure of chrysene metabolites. 

Figure 6 and 7 summarizes the structure of the metabolites formed by in vivo oxidation of 

benzo[c]phenanthrene [145], and benzo[α]pyrene, respectively [146,147]. The two major 

metabolites formed from benzo[c]phenanthrene are compounds 1 and 2. Biological testing 

revealed that metabolite 3 from benzo[a]pyrene was not carcinogenic while compound 4

was carcinogenic (Figure 7) [146,147]. The carcinogenic metabolite was also the major 

compounds formed biosynthetically. 

Figure 6. Chemical structure of benzo[c]phenanthrene metabolites. 

Figure 7. Chemical structure of 7,8-diol-9,10-epoxide-benzo[a]pyrene. 

In vivo oxidation in rodents of dibenzo[a,l]pyrene, which is considered as the most potent 

carcinogenic PAHs, results in a range of metabolites as outlined in Figure 8 [148]. The study 

showed that predominantly the (-)-11R,12R-enantiomer was formed and that the genotoxic 

events mainly took place by stereoselective activation of that enantiomer. 

The PAHs bay region diol epoxide has been singled out as the cause for this group of 

compounds carcinogenic activity [149-151]. The diol epoxides have been found to react with 

cellular macromolecules of paramount importance, namely DNA and proteins [152-154]. 
Polycyclic Aromatic Hydrocarbons a Constituent of Petroleum: 

Presence and Influence in the Aquatic Environment 97 

Miller proposed in 1970 that PAH metabolites are electrophiles that react with nucleophiles 

in vivo, thus delivering their biological effect [155]. Later it has been proposed that 

mechanistically the diol epoxides are electrophiles that alkylate the purine bases in DNA via 

an SN1-like epoxide ring opening process [156,157]. 

Figure 8. Chemical structure of bibenzo[a,l]pyrene metabolites formed in vivo [148]. 

In the worst scenario, however, adducts are formed between a PAH metabolite and an 

oncogene in the genomic DNA, hence switching off the cell’s ability to enter into 

programmed cell death (apoptosis) and provoking the onset of cancerogenesis. The best 

described example is DNA adduct formation between the pro-apoptotic p53 gene and 

benzo[a]pyrene-diol-epoxide after metabolic transformation of benzo[a]pyrene by the 

combined actions of CYP1A and epoxide hydrolase, another Phase I enzyme (this is also 

the mechanism by which cigarette smokers develop lung cancer). The presence of DNA 

adducts in liver tissue is, unlike CYP1A-induction or accumulation of PAH metabolites in 

bile, the result of cumulative exposure over weeks or even months [158,159]. Hence, 

hepatocytic CYP1A induction, accumulation of PAH metabolites in bile and elevated liver 

DNA adducts represent a chain of events that, although partly separated in time, are 

tightly interrelated from a mechanistic point of view. These three biomarkers have, 

therefore, received much attention and represent valuable biomarkers of PAH-exposure 

and effect in fish [159-162]. Analogous with several studies on mammals, responses of 

these core biomarkers of oil exposure have been associated with genotoxic effects such as 

liver neoplasia [163]. 
98 Hydrocarbon 

6. DNA adducts from polycyclic aromatic hydrocarbons 

The toxicity of PAHs is a continuous subject of intense investigation. The carcinogenic 

potential of PAHs was recognized as early as 1933 by Cook et al. [164], who isolated a 

cancer-producing hydrocarbon from coal tar. Many PAHs act as potent carcinogens and/or 

mutagens via DNA adduct formation. Aquatic vertebrates such as fish are capable of 

metabolizing PAHs (vide supra), producing reactive intermediates, occasionally with the 

formation of hydrophobic DNA adducts as an end result. 

Metabolic activation of PAHs to reactive intermediates is a prominent mode of their toxic 

action. Xue and Warshawsky described the principal metabolic pathways that yield reactive 

PAH intermediates [165]. Two pathways in particular produce electrophiles that may 

covalently bind to DNA (forming a DNA adduct): 1) electrophilic diol-epoxides from 

sequential PAH oxidation by cytochrome P450 (CYP) enzymes, hydrolysis of the resulting 

arene oxides by microsomal epoxide hydrolase, and a second CYP-catalyzed oxidation; 2) 

one electron oxidation of PAHs by CYP peroxidase yields the radical anion. 

There are 18 potential sites for adduct formation in DNA. The specificity of reactions at 

different sites depends on the reactive species, nucleophilicity of the DNA site and steric 

factors. The spectra of DNA adducts resulting from PAHs are considerably different from 

the ones formed by small alkylating agents. For example the dihydrodiol epoxide 

metabolites of PAHs react predominantly at the exocyclic amino groups of guanine and 

adenine. The major DNA adduct of the carcinogen, benzo[a]pyrene-7,8-dihydrodiol-9,10-

oxide, occurs at N2 of guanine [166]. DNA binding basically depends on its molecular 

structure and functional state (accessibility of nucleophilic target sites) while physiological 

and biochemical features determine differences in adduct formation among tissues and 

across species. 

DNA adducts caused by PAH metabolites are known to be crucial factors in the aetiology of 

cancer development. For this reason their presence and formation has been profusely 

studied. Adducts with benzo[a]pyrene (recognized as a model compound for the PAH 

group) are the most frequently reported [167,168], but adducts can also be formed with low 

molecular weight PAHs, like chrysene, which is a constituent in most mineral oils [169]. A 

schematic outline of the adduct formation process is outlined in Figure 9. 

DNA adducts have been used as a biomarker of exposure to PAHs since the ‘90s. They 

represent a very important endpoint, being a marker of genotoxicity. Numerous monitoring 

studies have reported the formation of DNA adduct formation in aquatic organisms (e.g. 

fish and bivalves) due to exposure to PAHs. For example, Lyons et al [171] and Harvey et al. 

[172] reported the genotoxic impact of the Sea Empress oil spill accident on different fish 

species (Lipophorys pholis and Limanda limanda) as well as on invertebrates (Halichondira 

panicea and Mytilus edulis). The DNA adduct patterns of fish liver exhibited the typical 

diagonal radioactive zone (DRZ) even 17 months after the spill took place. Detection of 

DNA adducts has been used to assess the impact of the Erika oil spill along the coasts of 

French Brittany. To confirm that the DNA adducts were really related to the Erika
Polycyclic Aromatic Hydrocarbons a Constituent of Petroleum: 

Presence and Influence in the Aquatic Environment 99 

petroleum, human hepatocyte (HePG2 cells) were exposed to an Erika fuel extract. 

Incubation of HePG2 cells lead to the formation of DNA adducts with similar patterns to the 

ones observed in the monitoring study using fish (Solea solea). These data indicates that 

human hepatocytes biotransform Erika fuel into genotoxic metabolites similarly to hepatic 

cells of fish and confirmed that the adducts observed in the monitoring study were related 

to the contamination of the sediment by the oil spill [173]. 

Figure 9. Metabolism of PAH leading to protein and DNA adducts. Figure adapted from reference 

[170]. 

It has been shown that DNA adducts persisted in vertebrates species due to the low 

efficiency of repair systems, representing a parameter for long term exposure [174]. These 

adducts are very persistent in fish liver [175-178]. French et al. [174] observed a steady 

increase in DNA adduct levels during a chronic exposure of sole (Pleuronectes vetulus) to 

PAH contaminated sediment for 5 weeks, which were persistent even after a depuration 

period. The persistency of DNA adducts has been demonstrated also in Atlantic cod (Gadus 

morhua) [38]. In this study, hepatic DNA adducts appeared after 3 days of exposure to low 

concentration of crude oil (0.06 ppm) and increased steadily during the entire exposure 

period of 30 days. 

Several techniques (e.g. immunoassay, fluorescence assay, gas chromatography-mass 

spectroscopy (GC-MS), 32P-postlabelling and mass spectrometry (MS) analysis) have been 

developed for the analysis of PAH derived DNA adducts. At present, the most sensitive and 

frequently applied technique in aquatic organisms is the 32P-postlabelling assay [179]. Its 

high sensitivity is unique and achueves the determination of 1-100 adducts in 109

nucleotides [180] 
100 Hydrocarbon 

The 32P-postlabelling assay appeared in the early ‘80s and has been applied with a range of 

protocols in order to detect DNA adducts produced by known carcinogens and complex 

mixtures [181]. Briefly, the assay involves DNA purification, digestion to normal and 

adduct-modified 3’-mononucleotides, removal of normal nucleotides (via enzymatic 

digestion, solvent extraction or chromatographic methods), 32P-postlabelling at the 5’ 

position of adducted nucleotides followed by chromatographic separation, detection and 

quantification (via autoradiography, scintillation counting or phosphor screen imaging 

analysis) [182]. Following this assay, PAHs cause the appearance of the diagonal radioactive 

zone (DRZ) (Figure 10) [173,178,183,184]. 

Figure 10. Example of the bulky DNA adduct zone (DRZ) typical of a contamination by PAHs detected 

by the 32P-postlabelling method (liver sample of fish collected in a PAH contaminated coastal area) 

(Pampanin unpublished data). 

MS/MS analysis has recently emerged as a powerful tool in the detection and structure 

elucidation of DNA adducts as well as for their quantification at very low concentrations (as 

often present in biological samples). In fact, electrospray ionization tandem mass 

spectrometric (ESI-MS/MS) analysis was capable of revealing DNA adducts in different 

aquatic organisms (e.g. fish (Oreochromis mossambicus) and mussel (Perna perna) soft tissue) 

[185,186]. This MS/MS approach provided a rapid determination and discrimination of 

structurally different phenanthrene derived DNA adducts in fish bile samples [186]. This 

technique is able to detect one modified base in 106-1012 unmodified bases [187]. 

Development of methodologies to detect DNA damages induced by PAHs is of constant 

concern in aquatic ecotoxicology. Direct chemical methods, such as high performance liquid 

chromatography with electro-chemical detection (HPLC-EDC), GC-MS and the 32Ppostlabelling, are highly sensitive and specific [180,188], however, they are very time and 

money consuming. A number of antibodies have been generated against carcinogenic 

products of DNA modifications, including those generated by PAHs. Immunoassays 

(immunohistochemistry or ELISA) are routinely employed to detect DNA adducts in 

humans, while the use of such approaches is more limited in aquatic species [189]. An 

immunoperoxidase method for revealing 7,8-dihydro-8-oxodeoxyguanosine (8-oxo-dG) in 

marine organisms has been described [189]. This work was also followed by the use of 

immunofluorescence and antibodies for DNA adduct detection in both vertebrates (fish, 
Polycyclic Aromatic Hydrocarbons a Constituent of Petroleum: 

Presence and Influence in the Aquatic Environment 101 

Anguilla anguilla) and invertebrated (mussel, Mytilus galloprovincialis). The 

immunohistochemical approach demonstrated a good selectivity, low cost, is easy to 

perform and readily allowed the analysis of a large number of samples. Nevertheless, it 

does not reach the high sensitivity of other methods [190]. 

7. Protein adducts from polycyclic aromatic hydrocarbons 

As discussed earlier (vide infra) CYP1A derived PAH metabolites have a high affinity to 

nucleic acids and proteins, which may result in adduct formation (Figure 11). It is likely that 

PAH protein adducts are formed after an initial docking, viz. protein-ligand interaction, of 

the PAH metabolite to the protein followed by reaction with nucleophilic sites in the vicinity 

of the docking site [191,192]. This has been shown to be the case for human serum albumin, 

which is predominantly alkylated at histidine146 by diol epoxides of fluoranthene and 

benzo[a]pyrene [193]. It is highly likely that the same type of mechanism is operating in 

animals and fish. 

Figure 11. General mechanism for PAH protein adduct formation. 

Naturally, adduct formation between PAH metabolites and human proteins and in 

particular human hemoglobin and serum albumin have been very well documented 

[170,192-199]. However, there are a few studies concerning protein adducts in fish. Plasma 

albumin adducts have been isolated from two species of fish (Brook trout (Salvelinus 

fantinalis) and Arctic charr (Salvelinus alpinus)) [200]. In the studies conducted by Padrόs and 

Pelletier, which were conducted by injection of benzo[a]pyrene, it was found that only the 

(+)-anti-diol-epoxide of benzo[a]pyrene generated adducts with serum albumin. In that 

study it was also found that there was no accumulation of the adduct upon repeated 

injections, thus indicating a relatively short half-life of less than two days for serum albumin 
102 Hydrocarbon 

in fish. In humans the half-life of serum albumin has been reported to 20 days [201]. In fish 

the presence of this adduct would be an indication of a very recent exposure to 

benzo[a]pyrene, while in humans this would also function as a marker of longer term 

exposure to the PAH. It has been found that the point of adduct formation between human 

serum albumin and benzo[a]pyrene anti-diol epoxide is dependent on the stereochemistry of 

the PAH metabolite [194]. The (+)-enantiomer generates a carboxylic ester adduct with 

Asp187 or Glu188 and that the (-)-enantiomer forms an adduct with His146. 

The different isoforms of hemoglobin present in different species results in the formation of 

different adducts. For example, rat hemoglobin possesses a reactive β-cysteine in position 

125 not present in human hemoglobin [202]. syn And anti fluoranthene diol-epoxides form 

adducts with this particular cysteine in rats. The presence of different isoforms of the same 

protein in different species results in the possibility of generating different adducts for the 

same PAH in the different species. The point of adductation most likely reflects on the 

proteins ability to function. Thus, the adduct formation of a specific protein might affect one 

species more severely than another. 

8. Future perspective 

Environmental research related to PAHs has to date, with a few exceptions, predominantly 

been concerned with finding metabolites of the compounds and detecting the presence of 

PAH DNA adducts. However, based on the discussion herein it is clear that the next step 

has to be towards analysis that can provide clear answers regarding the stereochemical 

outcome of the oxidation processes taking place in vivo. By such means it is easier to 

evaluate the toxicity of the various PAHs. This is a rather large task since different species 

metabolize PAHs differently resulting in dissimilar distributions between the stereoisomers. 

PAH protein adducts have been studied extensively for humans (vide supra) and rodents, 

however, for fish and other aquatic animals this topic is barely touched upon. Studies of 

adduct formation in fish will further aid the evaluation of the toxicity of the different PAHs. 

In the PAH protein adduct studies that have been conducted on other species we have seen 

that the point where the adduct is formed in a specific protein varies from species to species. 

Generating new knowledge as to where adducts are formed with the same PAH in the same 

protein in other species might in addition to providing increased knowledge as to the 

impact of adduct formation also possibly generate new interesting research questions. 

Hemoglobin in fish has a very short half-life so adduct formation on hemoglobin might not 

have such a great health impact on the fish. However, other less abundant proteins in the 

blood are also most likely susceptible to adduct formation with PAH metabolites. Detecting 

the proteins affected and determining the site of adduct formation will aid in the overall 

judgment of the toxicity of the PAH responsible for the adduct formed. Adducts with 

proteins present in the bile may also be of value in assessing the toxicity of PAHs. 

In human health care proteomics has for some time been extensively used for diagnostics 

[203,204] and these techniques are also slowly making their way into ecotoxicology 

[205,206]. It has been found in human amniotic epithelial cells exposed to anti-7,8-
Polycyclic Aromatic Hydrocarbons a Constituent of Petroleum: 

Presence and Influence in the Aquatic Environment 103 

dihydroxy-9,10-epoxy-7,8,9,10-tetrahydrobenzo[a]pyrene, a compound that causes adduct 

formation on DNA and oxidative damage on DNA, resulting in alternations of the 

expression of three proteins [207]. This result highlights that proteomics and the study of 

expression rates of particular proteins can be a powerful method in the future in order to 

determine if marine animals have been exposed to PAHs present in oil. 

Author details 

Daniela M. Pampanin 

Biomiljø, International Research Institute of Stavanger, Mekjarvik, Randaberg, Norway 

Magne O. Sydnes 

Faculty of Science and Technology, University of Stavanger, Stavanger, Norway 

Acknowledgement 

Funding from the University of Stavanger is gratefully acknowledged. Andrea Bagi, 

University of Stavanger, is thanked for valuable literature input for the introduction part of 

the chapter. Emily Lyng, International Research Institute of Stavanger, is acknowledged for 

her careful proof reading of the manuscript. 

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