1.1. Identification of the agent

1.1.1. Nomenclature

Polychlorinated biphenyls (PCBs) are a class of aromatic chemical compounds in which some or all hydrogen atoms attached to the biphenyl ring are substituted by chlorine atoms (m + n = 1–10) (Fig. 1.1). Synonyms for PCBs include chlorinated biphenyls, chlorinated diphenyls, chlorobiphenyls, or polychlorobiphenyls.

Fig. 1.1

Chemical structure of PCBs and the IUPAC numbering system

The general chemical formula is C₁₂H_(10-m-n)Cl_(m+n), where (m + n) is the number of chlorine atoms on the two rings. Depending on the position and number of the chlorine atoms, there are theoretically 209 individual PCB compounds (congeners). The carbon positions are numbered 1 to 6 on one ring, and 1′ to 6′ on the other. While positions 2,2′,6, and 6′ are called “ortho,” positions 3,3′,5 and 5′ are named “meta” and positions 4 and 4′ are called “para.”

Two different but correlated nomenclature systems are currently used. According to the International Union of Pure and Applied Chemistry (IUPAC) and in particular rule A-52.3 related to hydrocarbon systems, an unprimed number is considered lower (higher priority) than the same number when primed. Assemblies of unprimed and primed numbers are arranged in ascending numerical order. For a given PCB congener, the name lists the numbers sequentially [e.g. the PCB congener with chlorines on carbons 2,4,5, and 3′,4′ is identified as 2,3′,4,4′,5 (and not 2′,3,4,4′,5′)]. A deviation in that system lists the unprimed and primed chlorinated ring positions separately, sometimes eliminating the prime symbols and the commas for clarity and ease of typing (e.g. 245–3′4′5′ or 245–345).

In an additional strategy proposed by Ballschmiter & Zell (1980), a number (called “BZ number”) is attributed to each individual congener. This number correlates the structural arrangement of the PCB congener and ascending order of number of chlorine substitutions within each sequential homologue (Ballschmiter & Zell, 1980). This results in the congeners being numbered from PCB-1 to PCB-209. This shorthand nomenclature has become quite popular and is convenient for many uses, although it is important to note that it obscures the chemical identity of the congener and does not strictly follow the IUPAC rules.

Slight changes in the original BZ congener-numbering system were later recommended to correct some errors (Schulte & Malisch, 1983; Ballschmiter et al., 1992), and this resulted in the renumbering of BZ numbers 199–201. Guitart et al. (1993) used a computer program to systematically renumber the PCBs according to the strict IUPAC rules. As a result, they recommended that the congeners previously numbered 107, 108, 109, 199, 200, and 201 be renumbered 109, 107, 108, 200, 201, and 199, respectively (reviewed in Mills et al., 2007). The nomenclature for PCB congeners based on this report is shown in Table 1.1 and will be preferred in this Monograph. However, in the scientific literature, the revised numbering of congeners 107–109 has not been adopted systematically; the numbering system commonly used has been that proposed by Ballschmiter et al. (1992) where only the original BZ numbers 199–201 are changed.

Table 1.1

Correspondance between BZ number^a and position of chlorine atoms on each phenyl ring of the PCBs^b.

PCBs can be categorized by degree of chlorination (number of chlorine atoms) in 10 homologue groups (Table 1.2) from monochlorobiphenyls to decachlorobiphenyls. More than 60% of the PCBs are tetra- to hexachlorophenyls.

Table 1.2

Physical and chemical properties of PCBs according to homologue group.

In the biphenyl molecule, the two aromatic rings can rotate about the connecting single 1,1′-bond (Fig. 1.1). As with all molecules, there is a low-energy preferred conformation. With PCBs, this conformation is dependent on the degree of chlorine substitution, since chlorine is larger than hydrogen and creates more steric hindrance to the rotation (Erickson, 2001). The two extreme theoretical configurations are “planar” or “coplanar,” in which the two benzene rings are in the same plane, and “non-planar” in which the benzene rings are at a 90° angle to each other (Faroon et al., 2000). The probability of attaining a planar configuration is essentially determined by the number of substitutions in the ortho positions (2,2′,6,6′): the benzene rings of non-ortho substituted PCBs as well as mono-ortho substituted PCBs can assume a planar configuration and are referred to as “planar” or “coplanar” congeners (Erickson, 1997). The replacement of hydrogen atoms in the ortho positions with larger chlorine atoms forces the aromatic rings to rotate out of the planar configuration (Fig. 1.2); such structures are referred to as “non-planar” or “non-coplanar” congeners. [The Working Group does not recommend the use of this terminology, which is not technically appropriate since these PCBs do not easily assume a planar conformation.]

Fig. 1.2

Tridimensional chemical structures of selected PCBs

The relationship between PCB congener number and the Chemical Abstracts Service (CAS) registry number is given in Table 1.3. The congener numbering presented in this table follows that in Table 1.1, with the revised numbering of congeners 107–109. The congener lipophilicity is given in the same table, and was expressed against capacity to partition in octanol and water (K_ow) (see Section 1.1.2). Congeners can also be characterized by descriptors (CP0, CP1, 4Cl, PP, 2M) that give rapid access to geometry and substituent positions. The first descriptor, CP0, characterizes 20 congeners that are referred to as non-ortho congeners, consisting of those with chlorine substitution at none of the ortho positions on the biphenyl backbone. The second descriptor, CP1, comprises 48 congeners that are referred to as mono-ortho congeners and include those with chlorine substitution at only one of the ortho positions; CP0 and CP1 congeners can adopt a planar configuration. The 4Cl descriptor designates 169 congeners that have a total of four or more chlorine substituents, regardless of position. There are 54 PP congeners that have both para positions chlorinated. The 2M group contains 140 congeners that have two or more of the meta positions chlorinated. A total of 11 congeners have no descriptor.

Table 1.3

Relationship between BZ number, CAS number, IUPAC name,^a congener descriptor, and log K_ow for individual PCBs.

The twelve congeners that display all descriptors are referred to as “dioxin-like” (Table 1.4). These twelve PCBs, namely PCB-77, PCB-81, PCB-105, PCB-114, PCB-118, PCB-123, PCB-126, PCB-156, PCB-157, PCB-167, PCB-169, and PCB-189, have been assigned toxicity equivalency factors (TEFs, assigned by WHO in 1998 and revised in 2005) (Van den Berg et al., 2006). [The Working Group stressed that the activities of these PCB congeners are not solely dioxin-like.]

Table 1.4

The 12 dioxin-like PCBs, with corresponding CAS number, IUPAC name, and individual WHO₁₉₉₈-TEF and WHO₂₀₀₅-TEF values.

Depending on the context of the study or investigation, specific congeners may be monitored. For instance, the Stockholm Convention on Persistent Organic Pollutants (POPS) recommends measurement of six indicator PCBs (PCB-28, PCB-52, PCB-101, PCB-138, PCB-153, and PCB-180) to characterize contamination by PCBs. These congeners were chosen because they are found at higher concentrations in the environment, in food, or in human fluids/tissues. Depending on country and context, different lists of varying numbers of congeners may be used, e.g. 36 congeners for the Centers for Disease Control and Prevention, USA, or only PCB-138, PCB-153, and PCB-180 most frequently in epidemiological studies with human blood (see Section 2).

Of the 209 PCB congeners, 78 display axial chirality. Only 19 of these congeners, those with three or more chlorine atoms in the ortho position, exist as two mirror-image atropisomers, i.e. two chiral atropisomers (Lehmler & Robertson, 2001; Kania-Korwel & Lehmler, 2013). The IUPAC nomenclature and BZ number for the 19 atropisomeric PCBs are listed in Table 1.5. They are stereoisomers resulting from hindered rotation around single bonds where the steric-strain barrier to rotation is high enough to allow for the isolation of the enantiomers (Haglund & Wiberg, 1996; Harju & Haglund, 1999). Both atropisomers have the same chemical and physical behaviour, except for optical rotation (Lehmler et al., 2010). They are stable at 25 °C, but at elevated temperatures it is necessary to separate the enantiomers via high-resolution chiral gas chromatography (GC) (Schurig & Reich, 1998; Harju & Haglund, 1999).

Table 1.5

PCB congeners that exist as chiral atropisomers.

1.1.2. Chemical and physical properties of PCBs

Pure single PCB congeners are mostly colourless or slightly yellowish, often odourless, crystalline compounds. Commercial products, however, are viscous liquid mixtures of these compounds, with viscosity increasing with degree of chlorination, and colour ranging from light yellow to a dark colour. For example, Aroclor 1242 is a “mobile liquid” and Aroclor 1260 is a “sticky resin” (Erickson, 2001). These products do not crystallize at low temperatures, but turn into solid resins. An important property of PCBs is their general inertness; they resist acids, alkalis and oxidants and are fire-resistant because of their high flash-points (IPCS, 2003). However, under certain conditions, they may be destroyed by chemical, thermal and biochemical processes. PCBs show excellent dielectric (insulating) properties. This has made them useful in a wide variety of applications, including as dielectric fluids in transformers and capacitors, heat-transfer fluids, and lubricants.

The physical properties of PCBs are important in understanding their analytical, physiological, and environmental properties. However, the interactions of the various physical properties can be extremely complex (Erickson, 2001). Chemical and physical properties such as solubility, vapour pressure, and Henry’s law constant have been reported for individual congeners (Shiu & Mackay, 1986; Murphy et al., 1987; Sabljić & Güsten, 1989; Dunnivant et al., 1992; Falconer & Bidleman, 1994). Data for homologue groups and for a selection of PCBs are presented in Table 1.2, Table 1.3, and Table 1.6. Melting points range from 25 °C (PCB-2, PCB-7 and PCB-9) to 306 °C (PCB-209). Boiling points increase from low (monochlorobiphenyl, 285 °C) to highly (decachlorobiphenyl, 456 °C) chlorinated congeners (Hutzinger et al., 1974; Shiu & Mackay, 1986).

Table 1.6

Physical and chemical data for a selection of PCB congeners.

The solubility of PCBs in water is extremely low, ranging from an average of 0.0012 to 4830 μg/L for the chlorobiphenyl congeners that occur commonly. The high solubility of the ortho-chlorinated congeners (4.8 mg/L for PCB-1) may be due to hydrogen bonding associated with the more polar character of these molecules. Solubility decreases rapidly in ortho-vacant congeners, especially as the para positions are filled, which may result in greater and more uniform perimeter electronegativity and interference with hydrogen bonding. PCBs are freely soluble in non-polar organic solvents, oils and biological lipids, and the shift from water to lipid solubility is linked to the degree of chlorination (Hutzinger et al., 1974; Shiu & Mackay, 1986; ATSDR, 2000; IPCS, 2003).

The octanol/water partition coefficient (K_ow) is defined as the ratio of a chemical’s concentration in the octanol phase to its concentration in the aqueous phase of a two-phase octanol/water system; values of K_ow are thus unitless (Table 1.3 and Table 1.6). The reported log K_ow values have been reviewed by Shiu & Mackay (1986). Fig. 1.3 shows the remarkable correlation between log K_ow (lipophilicity) and number of chlorine atoms (BZ numbers); log K_ow values ranged from 4.5 to 8.3. This partitioning plays a key role in environmental fate and transport. PCBs tend to favour the non-polar phase and will partition away from water to most solids, the organic portion being the preferred site (Erickson, 2001).

Fig. 1.3

Octanol/water partition coefficient (K_ow) of PCB congeners according to the degree of chlorination (BZ number)

PCBs are characterized by Henry’s law constants [a measure of the equilibrium distribution coefficient between air and water] that tend to decrease with a higher degree of chlorination. Less chlorinated PCB congeners have a considerably higher vapour pressure (1–2 Pa at 25 °C for monochlorobiphenyls) than the more highly chlorinated congeners (1.4 × 10⁻⁶ Pa for decachlorobiphenyl) (Shiu & Mackay, 1986). Therefore, the composition in air is dominated by the less chlorinated congeners and atropoisomers.

At high temperatures, PCBs are combustible, and the products of combustion include polychlorinated dibenzofurans (PCDFs) and hydrogen chloride, and polychlorinated dibenzodioxins (PCDDs) (IPCS, 1993; ATSDR, 2000).

Photochemical degradation may be one route for the breakdown of PCBs in the environment: photochemical experiments conducted under simulated natural conditions on several pure chlorobiphenyls and on commercial PCB products have indicated several degradative reactions, such as dechlorination, polymerization and solvolysis.

1.1.3. Trade names and composition of commercial products

PCBs have never been used as single compounds, but rather as complex mixtures. The commercial products were manufactured to yield a certain degree of chlorination to fulfil technical requirements, generally between 21% and 68% chlorine.

Trade names for commercial products are given in Table 1.7. The most well known are Aroclor, Clophen, Phenochlor, Kanechlor, Pyralene, Fenclor, and Delor. The Aroclors, which were manufactured in the USA, are identified by a four-digit numbering code in which the first two digits indicate the type of mixture and the last two digits indicate the approximate chlorine content by percentage weight. Thus Aroclor 1242 is a chlorinated biphenyl mixture with an average chlorine content of 42%. The exception to this code is Aroclor 1016, which has an average chlorine content of 41% (Hutzinger et al., 1974). Similarly, the Kanechlors are identified by a three-digit value indicating the average chlorine content (300 for 30%). Other products of similar chlorination content have been produced by different companies in Europe, Japan, and China.

Table 1.7

Trade names for commercial PCB products^{a, b}.

Table 1.8 indicates equivalencies between main commercial formulations of PCBs. [The Working Group noted that these should be considered as approximate.] Since different production yield slight differences in the congener mixture, mixtures with comparable chlorine content but from different manufacturers (e.g. Aroclor 1260 and Clophen A60) show varying compositions, although with strong similarities (Johnson et al., 2000).

Table 1.8

Comparison of commercial PCB products based on percentage chlorination.

The homologue composition of the commercial PCB products varies greatly according to chlorination degree achieved (Table 1.9). For example, Aroclor 1242 is a mixture of mono- to heptachlorobiphenyls, while Aroclor 1260 contains penta- to octachlorinated homologues. The concentrations of single congeners within each homologue group also differ between different products and batches (Fig. 1.4). About 130 of the 209 congeners have been identified in commercial formulations at concentrations above 0.05%. Generally, commercial PCB products consist of about 100–140 PCB congeners, with mono- and non-ortho substituted PCBs as minor or trace constituents (Frame et al., 1996a, b; Johnson et al., 2000).

Table 1.9

Homologue composition and physical properties of selected commercial PCB products.

Fig. 1.4

Congener-specific composition of Aroclor formulations

An archetypal distribution of PCB congeners was detected in Aroclor 1254, lot 124–191 (corresponding to the historical G4 production process), while lot 6024 showed a profile characteristic of the A4 production process used between 1974 and 1976 (Kodavanti et al., 2001). Indeed, Aroclor 1254 was produced by two different chlorination procedures (two-step versus single-step chlorination) (Frame et al., 1996a, b). The differences in composition of the two lots are given in Table 1.10. Although Aroclor 1254 A4 probably represented less than 1% of the total production of Aroclor 1254, this PCB product was extensively used by standard suppliers and thus by researchers (Frame, 1999).

Table 1.10

Chemical profile and impurities (polychlorodibenzodioxins, polychlorodibenzofurans and polychlorinated naphthalenes) in lots 124–191 and 6024 of Aroclor 1254.

Chiral PCB congeners are important constituents of both technical and environmental mixtures of PCBs. For example, the total concentration of chiral PCB congeners in the commercial mixtures Aroclor 1242 and Aroclor 1260 is 6% and 30% by weight, respectively (Kania-Korwel et al., 2007). Chiral enantiomers may have different biological and toxicological properties (Püttmann et al., 1989; Rodman et al., 1991). There is evidence that PCB atropisomers differ in their biological activities (Kania-Korwel et al., 2006, 2008). They have been found in non-racemic proportions in many species (Lehmler et al., 2010; Wong & Warner, 2009). While physical and chemical processes in the environment generally affect the two enantiomers of a known compound at the same rate, biological processes may result in the enrichment of one of the enantiomers, because of enantio-selective interactions with biological macromolecules (Buser & Mueller, 1993).

1.1.4. Contaminants and impurities of commercial products

Commercial PCB products have been reported to be contaminated with other chlorinated aromatic compounds, such as polychlorinated naphthalenes and PCDFs (IARC, 1978). Vos & Koeman (1970) were able to identify tetrachlorodibenzofurans, pentachlorodibenzofurans, and chlorinated naphthalenes in samples of Phenoclor DP-6 and Clophen A60. Bowes et al. (1975) examined samples of Aroclor 1248, 1254 and 1260 produced in 1969, samples of Aroclor 1254 from 1970 and Aroclor 1016 from 1972, and samples of Aroclor 1260, Phenoclor DP-6 and Clophen A60. They found PCDFs in all Aroclor preparations except Aroclor 1016, and in Clophen A60 and Phenoclor DP-6 (Table 1.11). The levels of PCDFs were in the low microgram per gram range (Erickson, 2001), but additional PCDFs may be formed from PCBs on heating. Impurities such as 2,3,7,8-tetrachlorodibenzofuran and 2,3,4,7,8-pentachlorodibenzofuran have been reported in different amounts under various manufacturing conditions in Aroclor 1248, Aroclor 1254, Clophen A-60, Phenoclor DP-6, and Kanechlor 400 (de Voogt & Brinkman, 1989). Rappe & Gara (1977) confirmed by capillary gas chromatography–mass spectrometry (GC–MS) that 2,3,7,8-tetrachlorodibenzofuran was one of the main PCDFs in “Yusho oil,” as reported by Nagayama et al. (1976).

Table 1.11

Concentrations of chlorodibenzofurans in Aroclor, Clophen, and Phenoclor.

The proportion of impurities may vary between batches. For example, Aroclor 1254 with lot numbers 6024 and 124–191, which were produced by the same company by two different production processes, showed a 3.4-fold difference in the total concentration of PCDFs (Table 1.10).

It is important to note that PCDDs are not found in commercial PCB preparations (Erickson, 2001).

Overall, differences in composition as well as the presence of toxicologically relevant impurities may have had a significant impact on the results of toxicological studies with commercial PCB products and mixtures (EFSA, 2005). Consistent interpretation of the results of such studies, especially differentiation of the effects caused by respective PCBs, may only be achieved if the congener composition of these mixtures is known. The determination of the content in specific congeners was not feasible in most cases due to the lower sensitivity of analytical techniques available in the past.

1.2. Analysis

1.2.3. Analysis of environmental samples

Selected methods for analysis of PCBs in environmental matrices are presented in Table 1.12.

Table 1.12

Selected methods of analysis of PCBs in environmental matrices.

Supplementary material on analysis of PCBs in soil and sediment is available online at: http://monographs.iarc.fr/ENG/Monographs/vol107/suppl_S1.pdf.

(a) Air and dust

Both active and passive sampling are used for PCB analysis in air. Passive sampling has been applied to the analysis of outdoor air using semi-permeable membrane devices (Ockenden et al., 2001) and polyurethane foam (Mari et al., 2008). Vegetation is used as a natural passive sampler, for example tree bark integrating atmospheric PCB concentrations over the life time of the tree (Hermanson and Hites, 1990) or pine needles reflecting up to several years of PCB exposure (Kylin et al., 1994).

Polyurethane foam has also been used for indoor air collection (Hazrati & Harrad, 2006), but active sampling is often the preferred method (EPA, 1999; Kohler et al., 2005). To account for concentration differences and the limited air volume in an indoor setting, outdoor air is usually sampled by high-volume sampling, while low-volume sampling is used for indoor air.

Once retained on a solid matrix (filter, sorbent), PCBs are solvent-extracted using the same techniques as commonly applied for soil, sediment or biota. Before extraction, recovery/internal standards are added, e.g. PCB congeners that are not present in the environment, or ¹³C-labelled PCB congeners. Extraction is often performed by Soxhlet (EPA, 1999; Menichini et al., 2007). Ultrasonic extraction and pressurized liquid extraction (PLE) have also been described (Aydin et al., 2007; Mari et al., 2008). Barro et al. (2005) applied headspace-solid phase micro extraction, which does not involve solvents.

Whether or not clean-up steps are required depends on potential interferences from the matrix (e.g. particles) and co-extracted compounds as well as on expected concentrations. Adsorption chromatography can be applied, for example using alumina (Zhang et al., 2011a) or silica. As all PCB congeners are acid stable, acid treatment is possible.

Dust for PCB analysis has been collected in several ways, for example from the residents’ vacuum cleaner bags (Franzblau et al., 2009; Knobeloch et al., 2012), by vacuuming (Wilson et al., 2001; Harrad et al., 2009) and from air conditioning units (Tan et al., 2007). Dust samples originating from vacuum bags might be sieved, but cut-off sizes differ, e.g. 150 µm (Wilson et al., 2001) and 1 mm (Knobeloch et al., 2012) have been described.

Extraction techniques are basically the same as described for sorbents, including Soxhlet extraction (Dirtu & Covaci, 2010), PLE (Harrad et al., 2009) and ultrasonic extraction (Wilson et al., 2001). Before extraction, internal/recovery standards should be added, as described for air samples. Due to interferences from the matrix and co-extraction of other compounds, clean-up of dust samples will be required. PLE can be combined with simultaneous clean-up by adding adsorption materials to the cells; however, additional clean-up steps may be necessary (Harrad et al., 2009). Various sorbents have been used for clean-up of dust samples, including Florisil (Wilson et al., 2001; Harrad et al., 2009), silica gel (Dirtu & Covaci, 2010), and combinations of both (Knobeloch et al., 2012). As described for air samples, acid treatment has also been applied (Harrad et al., 2009).

(b) Water

The PCB content in a water sample is strongly influenced by the amount of suspended particulate matter (SPM) that adsorbs PCBs. Depending on the objectives of the analysis, different approaches can be chosen, resulting in different fractions to be analysed:

• Unfiltered water includes dissolved components and those bound to colloids and SPM.
• Filtered water gives PCB concentration on SPM (residue on the filter) and dissolved or bound to colloids (filtrate).
• Passive sampling targets the dissolved fraction.

Passive sampling devices integrate PCB concentrations over time, which reduces temporal variability. Common formats for water sampling include semipermeable membrane devices, low density polyethylene, and silicone rubber (Lohmann et al., 2012). Passive sampling techniques have been applied for analysis of PCBs in river water (Grabic et al., 2010) and seawater (Granmo et al., 2000; Fernandez et al., 2012).

In water bodies with a low SPM content, e.g. seawater, PCB concentrations will likely be low, and large amounts of water will have to be sampled and processed, while avoiding contamination. Guidelines for seawater sampling and the subsequent analysis of organic contaminants have been established by OSPAR (OSPAR, 2013). Studies have shown that the critical part of such analysis occurs outside the laboratory, i.e. during sampling, transport, and storage (Wolska et al., 2005). As described for air and dust, recovery/internal standards should be added before extraction.

PCBs from water samples are typically extracted by either liquid–liquid extraction (LLE), i.e. the direct extraction of PCBs with a non-polar solvent (Hope et al., 1997), or solid-phase extraction (SPE), where PCBs are retained on a solid phase and subsequently eluted with a non-polar solvent (Russo et al., 1999). The United States Environmental Protection Agency (EPA) method 1668B for determination of PCBs in several matrices describes SPE, continuous LLE, and separatory funnel extraction as suitable extraction methods for aqueous samples (EPA, 2008a).

The amount of SPM in the sample is a critical factor, as LLE might be insufficient and SPE cartridges might become blocked by samples with a high SPM content (Erger et al., 2012). Alternatively, SPM might be removed by filtration and analysed separately, for example by Soxhlet or ultrasonic extraction. This could be the method of choice for water samples with a high SPM content, for example wastewater samples or landfall leachate (Zorita & Mathiasson, 2005).

To what extent purification is necessary depends on the nature of the sample, its SPM content, PCB concentration and that of interfering compounds. Although sampling only freely dissolved PCBs, some passive sampling approaches add a clean-up step after extraction, for example using acid silica or aluminium (Grabic et al., 2010). Surface-water samples, however, have often been analysed without clean-up (Hope et al., 1997; Erger et al., 2012), while other studies have included adsorption chromatographic steps (Khim et al., 2001). Gel permeation chromatography (GPC) may be used for water extracts that contain organic compounds of high relative molecular mass (EPA, 2008a).

PCB exposure from snow can be considered insignificant, with the exception of polar regions where snow may be a source of drinking-water. Analytical methods are similar to those for water (Carrera et al., 1998).

1.2.4. Analysis of biological samples

Several matrices have been analysed to determine internal exposure to PCBs, or body burden, including adipose tissue, meconium, placenta, blood, umbilical cord blood, human milk, and hair (Table 1.13).

Table 1.13

Selected methods for analysis of PCBs in biological matrices.

(a) Tissues (adipose tissue and placenta)

The analytical methods applied to the analysis of PCBs in tissues such as adipose and placenta are similar to those used for environmental samples. The characteristically high lipid content of adipose and other tissues, however, requires rigorous lipid removal before instrumental analysis.

Different ways of sample pretreatment have been applied after or as part of the homogenization procedure, for example sample drying with Na₂SO₄ or hydromatrix (Covaci et al., 2002a; Saito et al., 2004), melting of fat (De Saeger et al., 2005), mixing with base (Kim & Fisher, 2008) and addition of ethanol for protein precipitation (Whitcomb et al., 2005).

Extraction is generally carried out with a non-polar solvent such as toluene or hexane, in some cases in a mixture with acetone, dichloromethane or propanol (Guvenius et al., 2002; Saito et al., 2004; Fernandez et al., 2008). The extraction could often proceed by shaking or rotating, for example in an Ultra Turrex or Vortex (Guvenius et al., 2002). Other extraction techniques are the same as those applied in environmental analyses, including ultrasonic extraction (Suzuki et al., 2005), Soxhlet (Fernandez et al., 2008), PLE (Saito et al., 2004), and MAE (Li et al., 2006). Supercritical fluid extraction with carbon dioxide (sometimes modified with dichloromethane) has also been applied (Stellman et al., 1998). For the extraction of placenta, Gómara et al. (2012) additionally described the preparation of a suspension that was liquid–liquid extracted. As for other matrices, recovery/internal standards are generally added before extraction.

A common method for lipid removal is treatment of the sample with acid, usually sulfuric acid (Whitcomb et al., 2005). GPC is another suitable method (Ma et al., 2012), but may not achieve complete removal of lipids. The use of partially deactivated neutral aluminium for lipid removal has also been described (Stellman et al., 1998).

For further clean-up of the extracts, the same techniques are applied as in the environmental analyses, either individually or in combinations. These include silica gel (Suzuki et al., 2005; Fernandez et al., 2008), alumina (Covaci et al., 2002a), Florisil (Whitcomb et al., 2005) and GPC (Saito et al., 2004). Impregnating the silica gel with acid is a common way of combining adsorption chromatography with lipid removal (Covaci et al., 2002a; Fernandez et al., 2008).

Some studies have analysed PCB metabolites in adipose tissue and placenta, e.g. hydroxylated PCBs and methylsulfonyl-PCBs. These methods usually included a fractionation by adsorption chromatography and elution with different solvents (Guvenius et al., 2002; Saito et al., 2004). In the method by Gómara et al. (2012), hydroxylated PCBs were separated from the parent compounds during liquid-liquid extraction (LLE). After derivatization, the fraction containing hydroxylated PCBs was cleaned up in the same way as described for the parent compounds.

(b) Blood (including umbilical cord blood)

Numerous studies have analysed PCBs in blood, mostly in serum, but also in plasma (Schettgen et al., 2011). The analytical methods used generally do not differ for serum and plasma. Given the low lipid content of blood, PCB concentrations are generally low and the sample amount available for analysis may be a challenge. Most studies work with volumes of 0.5–2 mL. Methods have recently been developed to extract PCBs from only 50 µL of plasma and from dried blood spots (Lu et al., 2012).

Umbilical cord blood has often been analysed in combination with maternal blood, using the same methods. Given the lower lipid content and usually lower PCB concentrations in cord blood, adjustments of the sample intake might be useful; however, sample availability is usually the limiting factor.

Apart from the addition of internal standards, the first step in PCB analysis of serum, plasma or cord blood is generally the denaturation of protein, e.g. by addition of formic acid (Kang et al., 2008), methanol (Korrick et al., 2000), or acetonitrile (Agudo et al., 2009). Different extraction techniques have been described, among which the simple mixing of the sample with solvent (Apostoli et al., 2005; Schettgen et al., 2011). LLE has also been used (Kawashiro et al., 2008; Bachelet et al., 2011) as well as SPE on C₁₈ or hydrophilic–lipophilic balanced reversed phase sorbent (Covaci & Schepens, 2001; Lee et al., 2011). Guvenius et al. (2003) used Lipidex 5000, a lipophilic gel, for extraction of PCBs from cord blood.

Since they are present at low concentration, lipids are not always removed from the extract (Lu et al., 2012). Lipids can be removed by direct addition of acid to the extracts (Atuma & Aune, 1999) or by clean-up methods on acidified silica (Covaci & Schepens, 2001). Further clean-up sorbents include Florisil (Whitcomb et al., 2005), alumina (Stellman et al., 1998), neutral silica gel (Atuma & Aune, 1999), or combinations of these (Guvenius et al., 2003; Apostoli et al., 2005).

To account for the low concentrations of PCBs in blood, extracts are often reduced to very small volumes, e.g. 50 µL (Covaci & Schepens, 2001). This is achieved by addition of non-volatile keepers (Covaci & Schepens, 2001), or by evaporation to dryness (Apostoli et al., 2005) and reconstitution in the desired solvent or a solution of syringe standards in this solvent. Evaporation to dryness carries the risk of loss of volatile PCB congeners.

Hydroxylated PCB metabolites have been analysed in blood and umbilical cord blood. Guvenius et al. (2003) used the same extraction method as for parent PCBs, but obtained hydroxylated PCBs in an isolated fraction, based on different elution solvents. Park et al. (2009) treated the sample with hydrochloric acid and 2-propanol, and extracted hydroxylated PCBs by LLE. Hydroxylated PCBs require derivatization to non-polar molecules before separation by GC (Sandau et al., 2000).

(c) Urine

A few studies have assessed PCB metabolites (hydroxylated PCBs) in urine samples. Hydroxylated PCBs are more polar than their parent compounds and act as weak acids, which has to be taken into account in extraction, clean-up, and separation by GC.

Hong et al. (2005a, b) presented two methods for the extraction of hydroxylated PCBs from urine. The first method combined SPE testing of four different phases, with five derivatization methods. Best recoveries and GC separations were found for hydroxylated PCBs extracted on a C₂ phase and derivatized with iodopropane under basic conditions (Hong et al., 2005a). The second method used headspace solid-phase microextraction and on-fibre derivatization, achieved by placing the needle in the headspace of a solution of bis(trimethylsilyl)trifluoroacetamide (BSTFA) (Hong et al., 2005b). The derivatized hydroxylated PCBs were transferred to the GC injector by thermic desorption. Several fibre materials were tested, of which polydimethylsiloxane-divinylbenzene (PDMS-DVB) gave the highest signal in the analysis.

(d) Human milk

Breast milk is the most extensively analysed matrix for the estimation of PCB body burden in humans. The first studies date back to the 1970s (Musial et al., 1974), and programmes for the biomonitoring of human milk have been established in several countries or regions (Wilhelm et al., 2007; Krauthacker et al., 2009; Cerná et al., 2012). Analytical methods are very diverse: the milk samples may be treated as liquid, or the lipid phase may be isolated, or the sample may be freeze-dried and treated as solid.

When the whole milk sample is treated as a liquid, the first steps usually include protein denaturation and dispersion of fat globules by addition of sodium oxalate, or acetic acid and methanol, sometimes in combination with ultrasound treatment (Dmitrovic & Chan, 2002; Fürst, 2006). Before or after this step, internal standards are usually added, and the sample is extracted by LLE (Chovancová et al., 2011), or SPE (Covaci et al., 2001; Dmitrovic & Chan, 2002). Hexane is a commonly used solvent, although a large variety of solvent combinations and solvent sequences have been described in the literature.

In some studies, the lipid phase of the milk sample is separated or extracted and a defined amount of fat used for further analysis (Fürst, 2006; Pérez et al., 2012).

In the third approach, milk samples are freeze-dried and a defined amount is extracted with techniques commonly applied to solid samples, e.g. Soxhlet extraction (Duarte-Davidson et al., 1991) and PLE (She et al., 2007). Matrix solid-phase dispersion has also been described (Gómara et al., 2011). However, freeze-drying always runs the risk of loss of volatile PCBs and cross-contamination.

As described for other human matrices, lipids in the extract are removed before instrumental analysis. Furthermore, the extracts usually contain co-extracted compounds that are likely to interfere with PCBs in the instrumental analysis. The clean-up techniques therefore generally include lipid destruction by acid treatment, either directly in the extract (Duarte-Davidson et al., 1991), or by acidified silica gel (Covaci et al., 2001). Alternatively, GPC has been used, but usually in combination with acid treatment (She et al., 2007). Further clean-up techniques include adsorption chromatography on neutral or basic silica (She et al., 2007), alumina (Chovancová et al., 2011), and Florisil (Pérez et al., 2012), also in combinations (Fürst, 2006). Ingelido et al. (2007) described clean-up by supercritical fluid extraction.

(e) Human hair

With a lipid content of about 2% (Altshul et al., 2004), hair accumulates lipophilic compounds such as PCBs and has the advantage of being sampled non-invasively. However, to what extent hair PCB content reflects internal exposure to PCBs is difficult to determine, even if the hair is washed before analysis to avoid co-extraction of dust particles. Comparisons of serum and hair samples showed weak correlations for most PCB congeners and considerably higher PCB concentrations in hair, also on a lipid-normalized basis (Altshul et al., 2004). Effects of hair colour (natural or dyed) cannot be ruled out (Covaci et al., 2002b).

Sample amounts of less than 1 g are sufficient for detection of PCBs. The hair samples are washed, and cut or pulverized, and then spiked with internal or recovery standards and incubated with hydrochloric acid (Covaci et al., 2002b). Extraction techniques applied in hair analyses include LLE (Covaci et al., 2002b), Soxhlet (Zhang et al., 2007), and ultrasonic extraction (Barbounis et al., 2012). The same methods for lipid removal and extraction clean-up as for other biological matrices have been used, e.g. adsorption chromatography on acidified silica gel, alumina (Covaci et al., 2002b), and Florisil (Zhang et al., 2007). A comparison between three laboratories analysing the same hair sample but using different internal standards, extraction techniques and analytical instruments (GC-ECD, GC-LRMS and GC-HRMS) showed good agreement, with a relative standard deviation of 15% (Gill et al., 2004).

1.3. Production and uses

1.3.2. Production volumes and trends

Although the commercial production of PCBs began in the 1920s, it was not until after 1945 that production reached substantial volumes. Production peaked in the 1960s and 1970s, and had ceased in most countries by the end of the 1970s or early 1980s.

Estimates of the total cumulative worldwide production of PCBs indicate that 1 to 1.5 million tonnes (or more) of commercial PCB products were manufactured. Production volumes from former Czechoslovakia, France, Germany, Italy, Japan, China, Poland, the Russian Federation and the former Soviet Union, Spain, the United Kingdom, and the USA, as reported by Tatsukawa (1976), de Voogt & Brinkman (1989), Jiang et al. (1997), AMAP (2000), Holoubek et al. (2001a), and Sułkowski et al. (2003), add up to around 1 325 000 tonnes for 1930–1993 (Table 1.14).

Table 1.14

Volume and duration of PCB production in countries with known production (by production volume).

In the USA, annual production peaked in 1970 with a total volume of 39 000 tonnes. From 1957 to 1971, 12 different types of Aroclor with chlorine contents ranging from 21% to 68% were produced in the USA by Monsanto Chemicals Co. (see Section 1.1). In addition, Geneva Industries produced a smaller amount of PCBs from 1972 to 1974 (EPA, 2008b).

In China, the production of PCBs began in 1965 and was gradually stopped between 1974 and the 1980s. According to preliminary investigation and analysis, 7000–10 000 tonnes of PCBs were produced in China from 1965 to 1974, with 9000 tonnes as PCB₃ [similar to Aroclor 1242] and 1000 tonnes as PCB₅ [similar to Aroclor 1254] (Xing et al., 2005; NIP China, 2007).

Information from the Democratic People's Republic of Korea (NIP Korea DPR, 2008) indicated that production of PCBs has been ongoing at two sites since the 1960s. The initial production capacity for PCBs was 1200 tonnes per year, with a tendency to increase until the 1980s; however, capacity has decreased since the early 1990s, and the average annual production volume in 2001–2006 was 411.6 tonnes. The total amount produced up to 2006 could be estimated at around 30 000 tonnes. According to this report, the Democratic People's Republic of Korea planned to reconsider its production of PCBs in 2012.

The commercial products were marketed under more than one hundred different trade names, depending on place of manufacture, production process, and chlorine content. Aroclors comprised at least 10 different commercial PCB products, under the names Aroclor 1016, 1221, 1232, 1242, 1248, 1254, 1260, 1262, 1268, and 1270. It should be noted that Aroclor 5460 was not a PCB product, but consisted of polychlorinated terphenyls. Other commercial PCB products include Clophen (four products), Delor (three products), Fenclor (five products), Kanechlor (five products), Phenochlor (four products), Pyralene (three products), Sovol, and Therminol (see Section 1.1.3).

1.3.3. Uses

Due to the physical and chemical properties of PCBs, such as non-flammability, chemical stability, high boiling point, and high dielectric constants, PCBs were widely used in several industrial and commercial open and closed applications (Table 1.15). PCBs have also been used in corresponding military applications, but detailed information on military use is typically very scarce.

Table 1.15

Industrial uses of PCBs.

As a result of the production process, PCBs were never used as individual congeners, but as technical products composed of multiple congeners. The commercial PCB products were generally used as such, but mixtures with other compounds were also produced to obtain specific properties. For example, the PCB product Sovol may have been mixed with α-nitronaphtalene to increase volatility, and sold as Nitrosovol (UNEP, 1988). Similarly, Galbestos was a mixture of PCBs and asbestos used on galvanized steel and galvanized corrugated sliding panels in various industrial and military applications.

(a) Closed applications

The predominant applications for PCBs were in dielectric fluids in capacitors and transformers. These applications are considered to be closed applications, since PCBs are not expected to leak out of the system. However, transformers had occasionally to be topped up with PCBs so that these systems were not completely closed.

While applications in hydraulic and heat transfer, and cooling systems are also usually considered to be closed applications, there have been reports of accidental leaks from such systems, and thus these applications are often referred to as “normally closed.”

During the 1960s, dielectric fluid in capacitors and transformers represented 50–60% of the sales of PCBs in the USA (IARC, 1978). In 1972, Monsanto restricted its future sale of PCBs to capacitor and transformer applications (Erickson, 2001); after this date, these applications represented some 99% of the total use of PCBs in the USA (Durfee et al., 1976). In China, PCB₃ [similar to Aroclor 1242] was used primarily in power capacitors applied in electricity production, distribution and transmission, while PCB₅ [similar to Aroclor 1254] was used mainly as a paint additive (see Table 1.8).

As production and use of PCBs became banned, outdated PCB-containing equipment (equipment filled with PCBs as dielectric fluid) was generally removed from use and stored for disposal (Xing et al. 2005). In this equipment, about 6000 tonnes of PCBs came from capacitors (NIP China, 2007). According to Durfee et al. (1976), these applications represented some. It should also be noted that

(b) Open applications

PCBs were also used in several open applications as a major constituent of permanent elastic sealants and as flame-retardant coatings (Heinzow et al., 2007).

The use as plasticizer in sealants (caulking material) and flooring material was common in many countries, representing up to 15–20% of the total use of PCBs in Sweden (Jansson et al. 1997). The sealants were mainly used in outdoor applications, but indoor use was not uncommon. Use in flooring material was limited to indoor use.

Sealants that were mixed with PCBs were mainly of the polysulfide type. The mixing was often performed on site. Information on concentrations to be used were not available to the Working Group; however, from a technical point of view, PCB concentrations were likely to be above 5%. Sealants analysed some 40 years after application often contained concentrations of PCBs of 5–15%, with concentrations of up to 35% being reported. The concentration may vary not only between sites, but also within a building. These variations may be the result of use of sealants with different PCB content, or of secondary processes, such as migration out of the matrix. There are reports indicating that inner parts of sealants could contain higher concentrations than the superficial parts (Johansson et al., 2003).

In addition to the use as sealants and flame-retardant coatings, PCBs have also been used in other open applications, such as in inks, adhesives, microencapsulation of dyes for carbonless duplicating paper, conveyor belts, rubber products, paints, pesticide fillers, plasticizers, polyolefin catalyst carriers, immersion oil for microscopes, cutting and lubricating oils, surface coatings, wire insulators, and metal coatings (ATSDR, 2000; Erickson, 2001; Erickson & Kaley, 2011). Also, use in small ballasts for fluorescent lights could be regarded as an open application, especially after long-lasting usage.

(c) Disposal of equipment containing PCBs

Improper handling of electronic waste (e-waste) has been identified as a source of environmental contamination with PCBs, especially for old equipment (Leung et al., 2006). Dismantling of ships has also been identified as a potentially important source of occupational exposure to and environmental contamination with PCBs (Basel Convention, 2003).

With the complete ban on the use of PCBs, stockpiles awaiting elimination have successively appeared in many countries.

In 2000, 23 companies worldwide had facilities for the disposal of equipment containing PCBs, of which 11 were in Europe. The use of solvent for decontamination represents the most common procedure of disposal, followed by destruction by incineration, dechlorination with sodium, retrofilling and vitrification. The most common technology used for destruction of PCBs is by incineration, with an efficiency of between 99% and 99.99999% (IOMC, 1998). For exemple, France has an installed capacity for incineration of PCB residues amounting to around 20 000 tonnes per year (INERIS, 2013).

1.4. Environmental occurrence and exposure

PCBs are found worldwide at measurable levels in all environmental media (soils and sediments, water, air), in wildlife, and also probably in the body of every human. Human exposure to PCBs occurs mostly via ingestion of contaminated food (see Section 1.4.7), but also via inhalation and dermal absorption.

Soils are natural sinks for persistent and lipophilic compounds such as PCBs; PCBs are absorbed by the organic carbon of the soil, and once absorbed they are relatively persistent (Buckley-Golder, 1999) (see Section 1.4.5). PCBs enter the soil via different pathways: industrial releases from manufacture, use and disposal, accidental releases, atmospheric deposition, application of sewage sludge, and erosion and leachate from nearby contaminated areas. PCBs in organic liquids may be dissolved by soils and then migrate with the solvent.

The congener patterns of PCBs in soils and sediments change over time as a result of the activity of aerobic bacteria (that degrade less chlorinated congeners) and anaerobic bacteria (that can cause partial dechlorination of more highly chlorinated congeners) (Hardell et al., 2010). The patterns found in environmental biota are often referred to as “weathered,” since they result from alterations in the composition of a mixture (e.g. resulting from bio accumulative and metabolic processes in higher biological organisms and through bacterial action, exposure to ultraviolet radiation, etc.). “Weathering” processes result in PCB patterns with either a higher chlorinated fraction or congeners with higher bioaccumulative properties compared with the commercial products. “Weathering” must be considered when assessing PCB-associated risks based on studies with experimental animals exposed to commercial PCB products.

Water is a major pathway for migration of PCBs, both in solution and particulate-bound, although PCBs are lipophilic and generally not very soluble in water (see Section 1.4.6). Less chlorinated PCB congeners have greater solubility than more highly chlorinated congeners.

Air is another major pathway for PCB migration (see Sections 1.4.3 and 1.4.4). PCBs are semivolatile compounds and, as with water solubility, less chlorinated congeners are more volatile than more highly chlorinated ones (Totten et al., 2006). There is extensive evidence that PCBs in aquatic systems exchange with PCBs in air (Bamford et al., 2002). Air transport of PCBs can occur in either the vapour phase or particulate-bound, thus contributing to global pollution and PCB contamination of remote regions of the earth. PCBs in air come from several direct or indirect sources, including industrial facilities, military sites, contaminated bodies of water, landfills and hazardous waste sites, electric arc furnaces, incineration and other forms of combustion, sewage sludge applied to agricultural lands, and construction materials, including in paints (Hu & Hornbuckle, 2010), caulking, light ballasts, floor sealants, and adhesives and plasticizers in older buildings (Wallace et al., 1996).

PCBs from soil, sediment, air and water enter the food-chain by uptake and bioaccumulation in plants and animal fats. There is significant biological magnification of PCB concentration as PCBs move up the food-chain. PCB concentrations vary depending on the degree of bioaccumulation, and are usually highest in carnivorous fish coming from contaminated waters. PCBs are found in the fat of all meat animals, in all dairy products containing fat, and in eggs (ATSDR, 2000; IOM, 2003), sometimes at high concentrations due to local contamination of grasses, and feeding practices in some countries (see Section 1.4.2). Also, it is not uncommon to feed domestic animals with fish meal or oil, or waste animal fats, which results in recycling of PCBs (IOM, 2003). For example, farmed salmon fed with concentrated fish meal or fish oil containing significant amounts of PCBs showed elevated concentrations of PCBs (Hites et al., 2004). PCBs found in food are typically of higher chlorination, since they are less volatile and more biologically persistent in plants and animals than the lower congeners.

Another important route of exposure to PCBs is inhalation; however, it is difficult to determine the relative contribution of inhalation compared with ingestion. Harrad et al. (2006) have suggested that inhalation may account for 4–63% (median, 15%) of overall exposure in humans. PCBs may be attached to indoor dust, which can be either ingested or inhaled. Individuals who spend significant periods of time in the presence of either outdoor or indoor vapour-phase PCBs will have continuous exposure that is not reflected in measurements of “total” PCBs, because the less chlorinated congeners are more rapidly metabolized and excreted by the human body (Fig. 1.5; Johansson et al., 2003). Concentrations of different PCB congeners were measured in blood from individuals living in houses where PCB-containing sealant was used. Concentrations of most congeners were only slightly elevated (1.2 to 3.2 times), but the two congeners with a low level of chlorination (PCB-28 and PCB-66) were detected at much higher concentrations (30 and 9 times, respectively) in contaminated flats than in control flats.

Fig. 1.5

Blood PCB concentrations in individuals living in PCB-contaminated flats relative to individuals living in control flats

Dermal absorption of PCBs may occur primarily in the occupational setting, but also through contact with contaminated sediments or other applications to the skin (Wester et al., 1987, 1993). Less chlorinated congeners are more rapidly absorbed through the skin than more highly chlorinated congeners (Garner & Matthews, 1998).

Congener patterns in the general human population are always different from any pattern found in commercial PCB products (Patterson et al., 2009). The factors that may explain this are:

• The general public is exposed to multiple sources of PCBs, only rarely to a single commercial product.
• There may be more than one route of exposure for almost all matrices/animals/humans.
• Dechlorination occurs to varying degrees in sediments, soils, water and air. Commercial PCB products will volatize to some degree, and in doing so, will lose less chlorinated congeners.
• PCBs ingested by fish and animals will be metabolized (to less chlorinated and hydroxylated congeners) to different degrees. Thus most food stuffs will demonstrate a shift in the congener profile compared to the commercial product.
• When inhalation is the major route of exposure, there is selective exposure to the more volatile, less chlorinated and less persistent congeners.
• Genetic differences among individuals may confer differences in metabolic activity and selective metabolism of different congeners.

1.4.3. Outdoor air

PCBs in outdoor air may be a significant source of exposure. Concentrations of PCBs in air depend on a variety of factors, including temperature and proximity to local sources. Temperature is particularly important in controlling the cycle of volatilization and precipitation. Proximity to local sources, such as industrial facilities, landfills, or contaminated bodies of water, results in elevated air concentrations of both vapour phase and particulate-bound PCBs that dissipate with distance at different rates, resulting in both local and distant contamination. Combustion and other high-temperature processes generate PCBs, in particular during combustion of highly chlorinated compounds; however, this route of unintentional formation is considered to contribute little to total airborne PCBs. Migration to the outdoor environment has also been shown to occur as a result of erosion of exposed sealants.

(a) North America

PCB concentrations in outdoor air vary greatly between urban and rural sites in North America, and may be very high near industrial facilities and other contaminated sites (Table 1.16). These differences reflect primarily the impact of local sources and dilution in air, but also the deposition of PCBs at lower temperatures.

Table 1.16

PCB concentrations in outdoor air in North America.

The major sources in Chicago are from landfills, sewage sludge drying beds, and transformer storage yards (Hsu et al., 2003). Shen et al. (2006) found large relative differences in air PCB concentrations between urban, rural and remote sites, with the highest concentrations in Toronto, Canada, and the Eastern third of the USA [absolute concentrations could not be quantified] using results from passive air samplers in 31 stations in Canada and the USA.

(b) Europe

In Europe, the reported PCB concentrations in outdoor air range from ~10 up to ~1000 pg/m³ in western European countries and from ~50 up to ~9000 pg/m³ in eastern European contries.

Measurement in the Baltic region showed PCB concentrations in southern Norway to be rather high and similar to those in urban areas (Backe et al., 2000; Agrell et al., 2001). Results from the Czech national monitoring system and European Monitoring and Evaluation Programme (EMEP) background monitoring stations also showed relatively high PCB concentrations in this country (EC, 2004). Typical values for background sites usually range up to ~100 pg/m³ and up to several 100s pg/m³ for contaminated areas (Kocan, 2000, 2001).

PCB concentrations in outdoor air may also be measured in precipitation as total deposition rates (ng/m² per day). In southern Sweden (Backe et al., 2002), PCB concentrations ranged from 1.18 to 81.4 ng/L, with no seasonal trends. In Paris, France, average PCB concentrations (sum of seven congeners) in rain during 1986–2001 remained approximately constant at about 40 ng/L (Chevreuil et al., 2001).

Declining concentrations of PCBs have been observed since the early 1960s and 1970s, decreasing by 67% in France (EC, 2004) and by 78% in the United Kingdom (CITEPA, 2013) over 20 years. The difference observed between the steady concentrations in rain and the decrease in general atmospheric emissions may be partly explained by water solubility limits and differences between point sources and global emissions.

Air concentrations of the seven indicator PCBs 28, 52, 101, 118, 153, 138 and 180 were measured at four locations in the Czech Republic, Finland, Sweden, and the Netherlands, from 1996 to 2001. Measured values did not vary noticeably during this period at any location (Fig. 1.6). This suggests that a steady-state has been reached between degradation and environmental cycling, with an ongoing low-level input from existing equipment and contaminated material (Holoubek et al., 2003).

Fig. 1.6

Annual average atmospheric concentrations of seven indicator PCBs (PCB₇) from four European Monitoring and Evaluation Programme stations in Europe, 1996–2001

(c) Asia

Limited information on the concentrations of PCBs in air and dust has been reported in Asian countries (Table 1.17). One of the most extensive studies reported results for outdoor air samples from 55 sites in Japan, 20 in China, 30 in the Republic of Korea, and 1 in Taiwan, China. The range of concentrations was 100–1000 pg/m³.

Table 1.17

PCB concentrations in outdoor air and dust in Asian countries.

(d) South and Central America

Shen et al. (2006) found large relative differences in air PCB concentrations between urban, rural and remote sites using passive air samplers in 4 stations in Mexico, Belize, and Costa Rica. One site in Mexico had higher concentrations than sites in Central America and in Canada.

Li et al. (2010) reviewed information from various research groups around the world and reported the average concentration of the sum of PCBs in air to be 66 pg/m³ (range, 9–670 pg/m³) for South America, and 59 pg/m³ (range, 17–150 pg/m³) for Central America.

(e) Africa

Only recently have data from passive air samplers deployed on the African continent become available. PCB concentrations have been reported as very high in Senegal (500 pg/m³) (Klánová et al., 2009), Côte d'Ivoire, and the Gambia (up to 300 pg/m³) (Gioia et al., 2011). Concentrations in some areas in South Africa, Kenya, Egypt, the Democratic Republic of the Congo, Ghana, Mali, and the Sudan were also high, and comparable to those in urban areas in more developed countries. These levels could not be explained by biomass burning or primary emissions, and were probably due to e-waste dumps. Lower concentrations have been measured in the Congo, Ethiopia, Mauritius, Nigeria, the Togolese Republic, Tunisia, and Zambia.

(f) Vegetation used for monitoring studies

Plant foliage is a reliable proxy for monitoring levels of vapour-phase compounds in outdoor air since it bioaccumulates organic pollutants. Several researchers have used vegetation, grass, conifer needles, mosses, pollen, and leafy vegetable species (cabbage and lettuce) as biomonitors to evaluate patterns of PCB contamination (Larsen et al., 1985; Reischl et al., 1989; Kylin, 1994; Simonich & Hites, 1995). This method has been employed in high-mountain ecosystems (Daly & Wania, 2005), and in several countries, including the Czech Republic (Holoubek et al., 1994), Poland (Migaszewski, 1999), western Finland (Sinkkonen et al., 1995), Germany (Reischl et al., 1987), Italy (Gaggi et al., 1985), and France (Granier & Chevreuil, 1992).

1.4.4. Indoor air

PCBs have been shown to migrate into surrounding materials, such as concrete or wood, and to indoor air. The major sources are PCB-containing caulk, paint (where PCB-11 is the main marker), floor sealants, and ballasts in lighting devices. Outgassing from contaminated dust may also contribute. Joint sealants are increasingly recognized as important diffuse sources of indoor air contamination by PCBs.

(a) North America

PCBs have been measured in indoor air in several studies (Vorhees et al., 1997; Vorhees et al., 1999; Herrick et al., 2004; Colt et al., 2005; Franzblau et al., 2009; Harrad et al., 2009). In the USA it was reported that indoor air concentrations of PCBs were 5–300 times greater than those in outdoor air (Wallace et al., 1996), and that concentrations were higher in older buildings. The concentrations of PCBs in indoor air in North America are summarized in Table 1.18.

Table 1.18

PCB concentrations in indoor air in North America.

(b) Europe

The highest indoor concentrations (up to 7500 ng/m³) have been reported in buildings constructed between 1960 and 1975 from prefabricated concrete elements sealed with elastic materials containing PCBs (Balfanz et al., 1993). Joint sealants containing PCB were discovered in various public buildings in Europe (Kohler et al., 2005; Wilkins et al., 2002). Estimated indoor PCB concentrations in contaminated sections were the lowest in microenvironments such as cars (8.92 ng/m³), and were inversely related to the degree of chlorination of the PCB mixtures used (Hammar 1992; Harrad et al., 2006; Kuusisto et al., 2006, 2007; Frederiksen et al., 2012). The concentrations of PCBs in indoor air in Europe are summarized in Table 1.19.

Table 1.19

PCB concentrations in indoor air in Europe.

(c) Asia

Indoor floor dust samples (n = 43) collected from rural homes and mosques in Gujarat, Pakistan, showed median total PCB concentrations of 0.67 ng/g (range, 0.3–6.1 ng/g) (Ali et al., 2012). The PCB profile was dominated by PCB-153 (> 60% of the sum of PCBs), with concentrations between < 0.2 and 2.4 ng/g. These PCB concentrations were 10 times lower than those reported in house dust in Singapore (Tan et al., 2007).

1.4.6. Water

Inputs of PCBs to the hydrological cycle are principally via discharges of sewage and industrial effluents, urban run-off, leachates from solid waste landfill sites, atmospheric deposition and, of increasing concern, via agricultural run-off (Scrimshaw et al., 1996).

Water can contain PCBs either in solution or bound to particulates. While PCBs are not very water-soluble, water can be a significant source of exposure to less chlorinated congeners that have a greater solubility than more highly chlorinated congeners. PCB concentrations in sea and freshwater are summarized in Table 1.20.

Table 1.20

PCB concentrations in various types of water around the world.

(a) North America

(i) Drinking-water

In the USA, the EPA has set a goal for PCBs in drinking-water of zero, and a maximum contaminant concentration of 500 ng/L (500 ppt), with sources being primarily landfills, and discharge of waste chemicals (EPA, 2014). While conventional treatment of drinking-water will remove particulate-bound PCBs, those that are soluble are often not completely removed. Solubilities of individual PCB congeners vary from about 4 ppm for monochlorobiphenyl to as low as 0.0007 ppm for the decachlorobiphenyl (Erickson, 1997). Thus under certain circumstances, drinking-water can still be a source of exposure to less chlorinated congeners.

(ii) Sea and freshwater

The USA–Canadian Great Lakes are contaminated by multiple sources of PCBs (Bhavsar et al., 2007; Turyk et al., 2012). It has been shown that industrial sites on rivers feeding Lake Erie received the largest quantities of PCBs, with 26% derived from atmospheric deposition (Kelly et al., 1991). The Hudson River in New York is highly contaminated with PCBs because of releases from two large capacitor plants (Carpenter & Welfinger-Smith, 2011), the Fox River in Wisconsin is highly contaminated because of releases from a manufacturer of carbonless copy paper, and a paper mill (Imamoglu et al., 2004), and the St Lawrence River and several of its tributaries have been contaminated by releases from aluminium foundries operated by companies that discarded hydraulic fluids containing PCBs in drains (Fitzgerald et al., 1996). The Hudson and Fox Rivers are being dredged to remove these contaminants.

(b) Europe

(i) Sea

An extensive review of data obtained during the 1980s has been published (Tolosa et al., 1995). In general, the concentrations of PCBs for all the investigated areas in the Mediterranean Sea were similar except in the Ligurian Sea where concentrations were higher. Predictably, the highest concentrations were reported in urban and industrial wastewaters (e.g. from Marseille and Barcelona) as well as in river discharges (e.g. from the Rhone), and decreasing concentration gradients have been found in transects offshore from these sources. PCB concentrations in the suspended particulate matter from coastal and open Western Mediterranean waters were of 5–35 pg/L in 1990, of the same order of magnitude as those reported in other regions, e.g. North Sea and North Atlantic. A more recent study covering the whole Western basin also shows a spatial gradient from the continental shelf (3.5–26.6 pg/L) towards the open sea (1.7–6.6 pg/L); a relatively important enrichment (8.4 pg/L) in open sea stations located in higher productivity frontal zones was observed (Dachs et al., 1997). The dissolved PCBs (Σ12 congeners) amounted to 28–63 pg/L. Total concentrations of PCBs in estuarine and coastal sediment samples of the Mediterranean Sea ranged from 0.04 to 1684 ng/g dw (Koci, 1998; Vale et al., 2002; Vojinovic-Miloradov et al., 2002; Cardellicchio et al., 2007).

During 1974–82, PCB concentrations decreased by a factor of 3 in offshore Monaco (Burns & Villeneuve, 1987), while the surface sediments of the Adriatic coast did not show a temporal trend (Picer & Picer, 1991).

Concentrations of PCBs in ocean water are usually in the low picogram per litre range. The general trend for concentrations in the Baltic Proper suggests an increase in PCB concentrations from the early 1970s onwards (ICES, 2000). This is an opposing trend to the decreasing concentration trends for PCBs in biota from the Baltic Proper (HELCOM, 1996; Roots, 1996).

The monitoring of PCBs in coastal areas may be based on measurements in mussels. Trends in PCB concentrations in the Seine estuary in France are reported in Fig. 1.7 (RNO, 2012). The rate of decrease was 3.5% per year. As reported by the Arctic Monitoring and Assessment Programme (AMAP), several time-series of PCB-153 concentrations in blue mussels from around Iceland showed significant decreasing trends; however, one time-series from a fjord system showed a significant increase (Rigét et al., 2010). Active mussel watching (mussel transplantation) has also been applied in monitoring programmes in Africa (see Table 1.20).

Fig. 1.7

PCB contamination along the coast of France

(ii) Freshwater

The major source of freshwater contamination in Europe comes from diffuse leaching of products from users, households, and industries into wastewater streams (UNEP, 2002). The areas most polluted by flood disasters are in Poland (the River Odra) (Wolska et al., 1999; Protasowicki et al., 1999) and in the Czech Republic. The River Danube is a major source of contamination to the Black Sea; however, many chlorinated hydrocarbons have been banned by several European and other countries in the past 10 years (Winkels et al., 1998; Covaci et al., 2002c; Fillmann et al., 2002; see Table 1.20).

Industrial contamination is known to have occurred in Germany (the Rivers Elbe and Rhine and their tributaries) (Brauch, 1993), in former Czechoslovakia (the Sumava Lakes) (Nondek & Frolikova, 1991), in England and Ireland (where however approximately a 50% decline in concentrations between 1970 and 1990, was recorded) (Sanders et al., 1992; Harrad et al., 1994) and in Slovenia through the dumping of industrial waste in the Krupa river during the manufacture of transformers. PCB contamination also occurred in the Balkan area, in the cities of Pancevo, Novi Sad, Belgrade, Kragujevac, in Serbia, after military intervention by NATO in spring 1999.

1.4.7. Food products

Since the early 1990s, food has been identified as the major route of human exposure to lipophilic and persistent organochlorines such as PCBs, PCDDs, and PCDFs. In populations that are not exposed to other known sources, dietary intake contributes to about 90% of the total daily intake of dioxin-like compounds including dioxin-like PCBs, and of this, food of animal origin contributes about 90% in various regions of the world (Schecter et al., 1997; Büchert et al., 2001; Llobet et al., 2003a, b; Päpke & Fürst, 2003; Schecter et al., 2003a, b; Charnley & Doull, 2005; Huwe & Larsen, 2005).

Similarly, it is generally accepted that the major route of exposure to non-dioxin-like PCBs, namely to PCB₆, is dietary intake, by consumption of fatty foodstuffs (IARC, 1978; IPCS, 1993; EFSA, 2005; Lindell, 2012). However, inhalation can also be a significant source of exposure (see Section 1.4.4).

Human food can become contaminated by PCBs via three main routes:

• uptake from the environment, by fish, birds, livestock (via food-chains), and crops;
• contamination of animal feed, by regular practices or accidentally;
• direct contamination of food, accidentally.

Data on PCB concentrations in food are reported in many different ways, making comparisons difficult. The number of congeners analysed differs between studies and often congeners are summed according to groups, such as indicator PCBs, DL-PCBs, or some other number of congeners. When using TEQs, the scheme used should be noted; also some studies report TEQ on the basis of bioassays such as the CALUX system as biological equivalents (BEQ). Results have been reported with different reference units (wet weight, dry weight, or lipid weight). Further difficulties in interpretation arise since different parts of fish or seafood are analysed (muscle, liver, skin, etc.) and PCB concentrations are also sometimes reported on the basis of prepared food (to account for changes by cooking or frying). Finally, the objectives of a study may bias the sampling strategy, often resulting in reporting of higher concentrations.

(a) PCB concentrations in food

Concentrations of DL-PCBs in various meats and dairy products from selected countries and regions are presented in Table 1.21.

Table 1.21

Concentrations of dioxin-like PCBs in selected foods from various countries and regions.

(i) Polar regions and North America

PCB concentrations in food for polar regions and North America are summarized in Table 1.22. Domingo & Bocio (2007) reviewed the concentrations of PCB and PCDD/PCDF in marine species and human intake through fish and seafood consumption by different region-specific sections.

Table 1.22

PCB concentrations in marine foods and estimated dietary intake in polar regions and North America.

The traditional food items for indigenous peoples in the Arctic include lipid-rich tissue of high trophic-level animals. After long-range transport and biomagnification of PCBs in the Arctic marine food-chain, PCBs accumulate in edible animals like fish, seals and whales (AMAP, 2004). This dietary exposure led to PCB concentrations in Arctic inhabitants that exceeded those of individuals living at temperate latitudes (Dewailly et al., 1993), but levels have been shown to decrease (AMAP, 2009). Likewise, PCBs in traditional food items have generally decreased (Rigét et al., 2010).

(ii) Africa

Loutfy et al. (2006) investigated levels of WHO-TEQs from diet in Egypt, and determined a range of 6.59–9.98 pg TEQ/kg per day, with about 40% of this value due to DL-PCBs. This value exceeds the maximum WHO tolerable daily intake (TDI) of 4 pg TEQ/kg per day. The primary source was found to be dairy products, in which PCB concentrations were several times higher than in such products in more developed countries. Loutfy et al. (2007) determined the concentrations of PCDD/PCDF and dioxin-like PCBs in samples of fish and seafood (mullet fish, bolti fish, bivalves and crab) randomly acquired in local markets in Egypt. The upper-bound concentrations of dioxin-like PCBs ranged from 0.14 (bivalves) to 0.76 (mullet) pg WHO-TEQ/g wet weight, respectively.

Adu-Kumi et al. (2010) reported an average TEQ for dioxin-like PCBs in fish from two lakes in Ghana to be 0.7 pg WHO-TEQ/g.

(iii) Australia and New Zealand

In 2000–2001, 168 samples of 22 foods collected for the Australian Total Diet Survey were analysed for DL-PCBs and compared with those from other areas of the world (Table 1.21; Food Standards Australia New Zealand, 2004).

A more recent study reported PCB concentrations from composite samples of Australian farmed yellowtail kingfish (mean, 21 μg/kg; range, 8.6–29 μg/kg), mulloway (mean, 5.4 μg/kg; range, 4.7–6 μg/kg) and manufactured feed (Padula et al., 2012). The mean concentration of DL-PCBs was 2.1 pg TEQ/g (range, 1.2–2.8 pg TEQ/g) in kingfish, and 0.51 pg TEQ/g (range, 0.41–0.61 pg TEQ/g) in mulloway.

(iv) Asia

Concentrations of specific PCB congeners in samples of food from Asia are summarized in Table 1.23. In Japan, a study sponsored by the Ministry of Health and Welfare showed a more than 50% decrease in concentrations of three non-ortho substituted PCBs in human milk samples between 1973 and 1996 (Environment Agency of Japan, 1999). A report from the Republic of Korea demonstrated regular dietary exposure (Son et al., 2012; Table 1.23). In China, Liu et al. (2011) determined concentrations of seven indicator PCBs in marine fish. The sum of PCB₇ ranged from 0.3 to 3.1 μg/g wet weight, with median and mean values of 6.4 ng/g wet weight and 398 ng/g wet weight, respectively (Table 1.23). The average concentrations and contributions of the seven specific congeners at four different sites are presented in Table 1.24. [It was noted that the concentrations found in this study were higher than in other parts of the world.]

Table 1.23

PCB concentrations in food in Asia.

Table 1.24

PCB concentrations in marine fish from China.

(v) Europe

The major contributors to total exposure in Europe appeared to be milk and dairy products for almost all groups of infants and toddlers (Barr et al., 2006; Becker et al., 2009), and fish and seafood products for most of the adolescents, adults, elderly and very elderly groups (Langer et al., 2007; Fréry et al., 2009; ANSES, 2011).

The most comprehensive assessment of PCB concentrations in food was undertaken by the European Food Safety Agency (EFSA) (EFSA, 2005, 2010, 2012). For the 27 European Union Member States, and Switzerland and Norway, in a report that took all food groups together, the upper bound (lower bound) for the 50th, 90th and 95th percentiles were < 0.005 (< 0.005), 0.02 (0.01) and 0.03 (0.01) pg WHO₂₀₀₅-TEQ/g wet weight for PCDD/PCDF, respectively. For the total TEQ, the upper bound (lower bound) concentrations were 0.01 (< 0.005), 0.04 (0.02) and 0.07 (0.04) pg WHO₂₀₀₅-TEQ/g wet weight, respectively (EFSA CONTAM, 2012). Infant formulae showed upper bound concentrations below the current maximum levels (0.2 pg WHO₂₀₀₅-TEQ/g wet weight), with highest concentrations found in ready-to-eat meals containing fish or meat. Overall, a decrease in concentrations of DL-PCBs was observed for the three food groups available: “raw milk and dairy products,” “hen eggs and egg products” and “muscle meat from fishes other than eels.” Feed and food of animal origin contained higher concentrations of PCDD/PCDF and DL-PCBs combined (the non-ortho PCBs were the main contributors to the total TEQs) than foods from plant origin. PCB-153, PCB-138, and PCB-180 represented altogether 36.9–97.8% of the sum of PCB₆. The maximum levels were exceeded in 9.7% of the food samples and 2.3% of the feed samples for PCDD/PCDF and DL-PCBs combined, and in 3.0% of the food samples and 2.4% of the feed samples for the PCB₆. With respect to food categories, lower PCB concentrations were found in meat from sheep, eggs from battery rearing, farmed salmon and trout, and farm milk (which however showed higher concentrations of PCDD/PCDF and DL-PCBs combined than milk from bulk) (EFSA, 2012).

The Baltic Sea area is heavily contaminated with persistent organochlorine compounds, including PCBs (Kiviranta et al., 2003), as is clearly attested by samples of fatty fish from the eastern coast in Sweden (Svensson et al., 1995). In the most contaminated feed group, the highest relative contribution to the WHO₂₀₀₅-TEQ_total came from non-ortho PCBs, up to twice the average contribution (EFSA, 2012).

(b) Estimated daily dietary intake

In Europe, more than 90% of PCB exposure in the general population is via food consumption (EFSA, 2005; Table 1.25). Average daily dietary intakes of the sum of PCB₆ are in the range of 10–45 ng/kg bw for adults, and two and a half times higher in children. Limited exposure data for young children indicate that the average daily intake (breastfeeding excluded) of the sum of PCB₆ is about 27–50 ng/kg bw. Overall, the non-ortho PCBs represented 21.0–74.9% of the WHO₂₀₀₅-TEQ_total of PCDD/PCDF and DL-PCBs combined in food (EFSA, 2012), and the mono-ortho PCBs represented no more than 12% of the WHO₂₀₀₅-TEQ_total. In the most contaminated samples, such as products from aquatic animals and from ruminants, the relative contribution of the non-ortho PCBs ranged from 34.2% to 86.1%. Most likely due to an effect of the European risk management measures, a decrease in exposure to the sum of PCB₆ was observed between 2002–2004 and 2008–2010 in most but not all population groups, and it was estimated between 2.0% and 75.6%.

Table 1.25

Dietary exposure to PCBs for an average consumer on the European market.

In the USA, the daily dietary intake of PCBs for adults decreased from 1978 (0.027 µg/kg bw) until 1986–1991 (< 1 ng/kg bw) (IPCS, 2003). Mean daily intakes for infants during the same period decreased from 11 to < 1 ng/kg bw. However, trends during 1991–1997 did not appear to decrease, and ranges of daily dietary intake were 3–5 ng/kg bw for adults, and 2–12 ng/kg bw for children of different ages (IPCS, 2003).

Daily dietary intake of PCBs from countries in Asia are presented in Table 1.26. In China, the estimated daily intake from four food groups of animal origin ranged from 0.09 to 0.59 pg TEQ/kg bw for DL-PCBs, which is lower than the daily intake in some developed countries (Liu et al., 2013). A survey of food items on the market and typical consumption patterns in Japan reported a daily intake for the general population of 2.60 pg TEQ/kg bw per day (Koizumi et al., 2005). Of these, 2.41 pg TEQ/kg bw per day was from ingestion of food, while inhalation and soil ingestion contributed only to 0.19 pg TEQ/kg bw per day. A “typical” Japanese person receives 120.7 pg TEQ per day through food consumption (mainly fish/shellfish, followed by meat/eggs).

Table 1.26

Estimated daily dietary intake of PCBs in Asia.

In specific subpopulations with high dietary PCB exposure, such as Baltic Sea fishermen, the daily intake from fish of the sum of PCB₆ was estimated at 40 ng/kg bw, corresponding to a total daily intake of the sum of non-dioxin-like PCBs of 80 ng/kg bw, before taking into account the rest of the diet (Lindell, 2012).

In breastfed infants, the most recent WHO study of PCB exposure reported a mean daily intake of about 1600 ng/kg bw (range, 230–7300 ng/kg bw per day) for total PCB₆. Thus, exposure of infants to PCB₆ (and DL-PCBs) through human milk is about two orders of magnitude higher than the average daily intake by adults.

1.4.9. Population biomonitoring

(a) Blood

The presence of PCBs in serum or blood may reflect exposure from any source (Dewailly et al., 1988). Results from different studies in humans have indicated that measurements of PCBs in serum generally reflect cumulative past exposure. Many PCB congeners can remain in the body for years after exposure, although some of the less chlorinated congeners are more volatile and consequently show shorter residence times.

(i) North America

Hopf et al. (2009a) provided an extensive review of reports on background levels of PCBs in the USA population. They concluded that serum concentrations increased up to 1979 and decreased after that, but that the background levels are still of concern. The NHANES survey over the period 2002–2004 reported increasing concentrations of PCBs with age, and concentrations were higher in men than in women, and higher in African-Americans and Caucasians than in Mexican-Americans (Patterson et al. 2009). Sjödin et al. (2004) showed a decline in concentrations of PCB-153 between 1985 and 2002 in pooled samples from the NHANES study.

Several studies have looked at specific populations living near specific contaminated sites or eating contaminated fish (Table 1.27).

Table 1.27

Serum concentrations of PCBs after consumption of PCB-contaminated fish, North America.

Serum concentrations for the sum of 17 congeners in Viet Nam veterans were 167.5 ng/L lipid adjusted, of which the major portion (116.6 ng/L) were di-ortho congeners (Schecter et al., 1996).

Jarrell et al. (2005) determined the sum of 24 congeners in pregnant women in Canada, and reported a mean value of 0.78 ng/L wet weight.

Because the less chlorinated PCBs are more volatile, teachers working in a school where caulk containing PCBs was used showed serum congener profiles that were enriched in less chlorinated congeners (Herrick et al., 2011).

DeCaprio et al. (2005) reported finding a pattern of PCB congeners in serum specific of young native Americans living near a PCB-contaminated waste site. This pattern was not clearly observed in older individuals because it was obscured by the greater concentrations of more persistent congeners, coming primarily from dietary exposure.

(ii) Europe

Several European studies on human biomonitoring have reported blood PCB concentrations in adults or children (summarized in Table 1.28). Past environmental contamination in industrial areas has polluted surrounding soils and forage, leading in turn to high blood PCB concentrations in the adult population. Age-related accumulation of PCBs has been observed in many studies (Patterson et al., 1994; Apostoli et al., 2005; Park et al., 2007), and may be partially explained by historical high levels of exposure in the 1970s.

Table 1.28

Blood concentrations of PCBs in various European countries.

In Germany, Environmental Surveys (GerES) were carried out in 1998 (Becker et al., 2002) and during 2003–2006 (Becker et al., 2009). GerES data show mean blood concentrations for the sum of PCBs of 1.3–1.7 µg/L in 1998 and of 286 ng/L in the more recent survey, with strong difference (factor of 5.6) between age groups 18–25 and 66–69 years. In Belgium in 2007–2011 (Schoeters et al., 2011), the Flemish Human Environmental Survey reported average blood PCB concentrations of 333 ng/g lipid. Average concentrations in the United Kingdom in 2003 were 170 ng/g lipid (Thomas et al., 2006). In Spain in 2004–2008, concentrations of the most common PCBs were in the range of 21.8 to 38.9 ng/g lipid (Ibarluzea et al., 2011). In France, blood analysis in the general adult population was first carried out in 1986 (Dewailly et al., 1988) and then in 2006–2007 (French Nutrition and Health survey; Fréry et al., 2013). The reported blood PCB concentrations in populations in industrial polluted areas such as Italy (Turci et al., 2004; Apostoli et al., 2005; Turrio-Baldassarri et al., 2008) and Slovakia (Jursa et al., 2006) were high compared with those in non-occupationally exposed populations such as in Sweden (Salihovic et al., 2012). In the Faroe Islands (Denmark), high concentrations of PCBs and hydroxylated PCBs in serum samples from pregnant women were attributed to the traditional diet, made of pilot whale meat, blubber and other marine food (Fängström et al., 2002).

The most frequently detected di-ortho-chlorine-substituted PCBs in population studies are PCB-138, PCB-153, and PCB-180 (Glynn et al., 2000), accounting for 65–78% of the measured sum of total PCBs (Needham et al., 2005). The seven PCB indicator congeners (118, 138, 153, 156, 170, 180, and 194) contributed to 99% of the total PCB levels, with a modest contribution from dioxin-like congeners (Apostoli et al., 2005).

In several countries in the European Union, a clear decrease in blood concentrations of PCBs has been observed in the last two decades. Overall, mean whole blood concentrations of PCB-138, PCB-153, and PCB-180 appear to have decreased by approximately 80% in 20 years (Link et al., 2005; Hagmar et al., 2006; Agudo et al., 2009; AMAP, 2009). Nevertheless, compared with North America (CDC, 2005), serum concentrations of PCB-138, PCB-153, and PCB-180 were higher by two- to fivefold in Germany in 1998 (Heudorf et al., 2002), or Italy in 2001–2003 (Turci et al., 2004; Apostoli et al., 2005; Needham et al., 2005). Similarly, serum concentrations of hydroxylated PCBs and methylsulfonyl-substituted metabolites of PCBs were higher by two to threefold in a contaminated area in a study in Slovakia (Hovander et al., 2006).

(iii) Asia

In Asia, PCB concentrations in several biological samples (including serum or whole blood, umbilical cord blood, hair, breast milk, adipose tissue, liver, kidney, and lung tissues) showed a wide range (Table 1.29; Schecter et al., 2003a). Data specific to the Yusho and Yucheng patients are presented in Table 1.30 and Table 1.31, respectively.

Table 1.29

PCB concentrations in biological samples from populations in Asia.

Table 1.30

PCB, PCDF, and PCDD concentrations in biological samples from the Yusho population, Japan.

Table 1.31

PCB concentrations in biological samples from Yucheng patients, Taichung County, Taiwan, China.

(iv) South and Central America

Rodríguez-Dozal et al. (2012) analysed serum samples from pregnant women in Mexico for 19 congeners and Aroclor 1260. For Aroclor 1260 [calculated as the sum of PCB-138 and PCB-153 multiplied by 5.2], they reported regional differences (mean concentration, 31.1 ng/g lipid) and elevated concentrations from residents of Merida (maximum, 546.2 ng/g lipid). Trejo-Acevedo et al. (2012) measured serum PCB concentrations (sum of 14 congeners) from children living in a malaria-endemic area of Mexico, and reported a mean serum PCB concentration of 5892 ± 3895.7 ng/g lipid. In an analysis of PCB congeners in maternal blood of women in Sao Paulo State, Brazil, PCB-118, PCB-138, and PCB-153 were detectable in more than 70% of samples, and their concentrations were almost double in women from industrial areas compared with women from rural areas (Rudge et al., 2012).

(v) Africa

Röllin et al. (2009) reported overall low blood concentrations of PCBs (99, 118, 138, 153, 170, 180 and 187) in delivering mothers from seven geographical regions in South Africa. Large regional differences were observed, with women from rural areas having the lowest levels of PCBs. PCB-138 and PCB-153 were found in the blood of mothers from all of the 61 sites studied at geometric mean concentrations of 3.56 and 3.2 ng/g lipid, respectively. Ahmed et al. (2002) reported the sum concentration of 29 congeners in blood from Egyptian women to be 61.9 ng/g. Weiss et al. (2006) reported concentrations of PCB-153 in infertile women in the United Republic of Tanzania to be 0.17 µg/kg. Sum PCB concentrations in serum samples from Bizerte, Tunisia, ranged from 37.5 to 284.6 ng/g lipid, with mean and median value of 136.1 ng/g lipid and 123.2 ng/g lipid, respectively. The PCB profile consisted mainly of persistent congeners such as PCB-138, PCB-153, and PCB-180 (82.7% of the sum of PCBs). PCB concentrations were significantly higher in men (P < 0.05) than in women (Ben Hassine et al., 2014).

(b) Human milk

Due to its high fat content, human milk can accumulate large amounts of PCBs, thus making it an ideal matrix for the determination of concentrations of PCBs and other lipophilic compounds, and can be sampled using non-invasive techniques. In addition, human milk represents a good indicator of the body burden of lipophilic non-metabolized PCBs, since fat is mobilized for the production of milk during lactation. Animal studies and mass balance studies for humans have revealed that large amounts of PCBs can be eliminated through lactation (Lindell, 2012). Data are summarized in Table 1.32.

Table 1.32

PCB concentrations in human milk, by country.

(i) Global assessment

The transfer of PCBs from mother to infants via breast milk is an important source of exposure, and several factors (including maternal residence, age, and parity) can potentially affect levels of contaminants in breast milk. Because of the importance of breastfeeding for infants, contamination of human milk is of specific public concern.

The Stockholm Convention on Persistent Organic Pollutants (POPs) is a guidance document, the objective of which is to document the effectiveness of the implementation of the obligations under the Convention. The World Health Organization (WHO) introduced worldwide measurement campaigns to determine the exposure of infants to dioxin-like PCBs (UNEP, 2012). The evaluation of the Stockholm Convention has been applied (with slight changes) for five rounds of the UNEP/WHO survey. Often, human milk from primiparae mothers (for detail, see UNEP, 2012) is preferred to human blood, since sampling is non-invasive and PCBs are easier to detect (due to the higher lipid content of milk). It should be noted that for global assessment, the concentrations of dioxin-like PCBs (DL-PCBs) on a TEQ basis for the last three rounds of the UNEP/WHO survey on mothers’ milk may be lower by 30% (range, 2–60%) if WHO toxic equivalency factors (TEFs) for 2005 (WHO₂₀₀₅-TEF) are applied, rather than those for 1998 (WHO₁₉₉₈-TEF). International chemical assessments report that the average concentration of PCBs in human milk fat ranges from 0.5 to 4 µg/g (IPCS, 2003). For the sum of PCB₆, the median is between 10.8–30.7 ng/g lipid, and maxima are between 37.1–65.8 ng/g lipid. Overall, the UNEP/WHO survey showed a correlation between maternal age and concentrations of DL-PCBs in breast milk, and lower concentrations of PCBs in breast milk of multiparous women when compared with primiparous women.

(ii) Americas

The mean concentration of PCBs in whole breast milk in Canadian women steadily increased from 6 μg/kg in 1970 to 12 μg/kg in 1975, and to 26 μg/kg in 1982, before declining to 6 μg/kg in 1986 (IPCS, 2003).

Recent data on concentrations of PCDD/PCDFs and DL-PCBs in human milk from South America were reported only for Brazil (Paumgartten et al., 2000).

(iii) Europe

In Europe, concentrations of DL-PCBs (on a TEQ basis) and PCB indicators in human milk are considerably higher than in other regions of the world, a legacy from past exposures. For the sum of PCB₆, the median of 115.3 ng/g lipid is between 3.8 times and 10.7 times higher than in other regions of the world, and maximum concentrations are up to 14.9 times higher (IPCS, 2003). However, a WHO survey identified a decrease in WHO-TEQ PCDD/PCDF and PCB concentrations in human milk over the last decade (Van Leeuwen et al., 2002). It was assumed that this decrease was the result of the ban of PCB use in open systems, and the strict regulations on the use of PCBs and on their disposal in closed systems.

Norén & Meironyté (referenced in IPCS, 2003) reported a steady decrease (from 910 to 324 ng/g lipid) in total PCB concentrations in the breast milk of Swedish women between 1967 and 1997. A declining trend could be observed for the sum of the PCB₆ in Germany, and mean values for the congeners PCB-138, PCB-153 and PCB-180 were approximately 60–70% lower in 2000 than in 1984 (Fürst, 2001; Fig. 1.8). An approximately 74% decrease in DL-PCB concentrations during the last decade was reported in Italy (Di Domenico & Turrio Baldassarri, 1990; Weiss et al., 2003; Abballe et al., 2008). Analyses of milk samples from the Czech Republic also revealed a decline in median concentrations between 1994 and 2000, the strongest decrease being observed between 1994 and 1997 (Cerná et al., 2008). Nevertheless, it should be noted that concentrations in areas with heavy contamination did not show a significant decline in exposure over the past 10 years.

Fig. 1.8

PCB concentrations in human milk in Germany, 1984–2000

(iv) Asia

In Japan, a time-trend study showed that average PCB concentrations in human milk increased from 1.3 ng/g in 1972 to a peak of 1.5 ng/g in 1974, and then decreased by about 13% in 1998 (Environment Agency of Japan, 1999). In contrast, daily intake of PCBs from breast milk was estimated to decrease from 22.3 µg/g to 0.31 µg/g during this same period. [This trend reflects a change in PCB concentrations in food, due to both a decrease in contamination and more dependence on imported foods, which were less contaminated than domestic foods (IPCS, 2003), and is consistent with the observed decline in PCB concentrations in the environment and in human tissues.]

In China, a national investigation of individuals in 12 provinces representing approximately 50% of the total Chinese population reported PCDD/PCDF-TEQ and total-TEQ in human milk from rural areas to be lower than those from urban areas (Li et al., 2009). Positive correlations were found between total-TEQ in human milk and the consumption of aquatic food and meat.

PCB levels in breast milk samples from women in Asia are summarized in Table 1.29 and Table 1.30.

(c) Adipose tissue

(i) North America

Lordo et al. (1996) reported PCB concentrations (sum of tetra- to octochlorobiphenyls) in pooled adipose tissue to be 672 ng/g in 1986, compared with 407 ng/g in 1982, and 508 ng/g in 1984. Stellman et al. (1998) reported a total PCB concentration of 267 ng/g in breast adipose tissue of healthy women from Long Island, New York. An approximation of Aroclor 1260 [summed concentrations of PCB-138 and PCB-153 multiplied by 5.2] measured in breast adipose tissue, was reported to be 870 ng/g (Aronson et al., 2000). Muscat et al. (2003) measured PCB concentrations (sum of 14 congeners) in breast adipose tissue in women without metastatic breast cancer to be 361 ± 235.9 ng/g and 395.4 ± 279.3 ng/g, in women who did not have recurrence and women who did have recurrence, respectively.

(ii) Europe

The results of a study conducted in 1993–94 suggested that concentrations of PCBs in adipose tissue are the best indicator of long-term exposure or of total body burden of PCBs, compared with human milk or blood (Kocan et al., 1994).

PCB concentrations in adipose tissue of the general population in industrialized countries vary very widely, ranging from < 1000 to 5000 ng/g fat (Falandysz et al., 1994; Holoubek et al., 1995, 2001b). In a comparative study in Europe (Van Bavel et al., 2003), PCB concentrations in the population in Sweden were one third (mean, 661.9 ng/g fat; range, 247.2–1651.2 ng/g fat; Σ37 PCBs) of those in the Hungarian samples.

(iii) South and Central America

Breast adipose tissue in 76 women from an agricultural region of north-eastern Argentina contained eight PCBs at very low levels (only 1.3% above detection limits), but high levels of p,p-dichlorodiphenyldichloroethane (DDE) and other pesticides (Muñoz-de-Toro et al., 2006).

The sum of four PCB congeners in children from Nicaragua was 530 ng/g lipid weight (2.0 ng/g wet weight) in those living and working near a waste-disposal site and eating fish from contaminated Lake Managua, 230 ng/g lipid weight (0.9 ng/g wet weight) in those living nearby but not working at the waste site and not eating fish, and 160 ng/g lipid weight (0.6 ng/g wet weight) in those living at a distance from the waste site and not eating fish (Cuadra et al., 2006).

(iv) Asia

PCB concentrations in adipose tissue were reported from Yusho and Yucheng patients (see Table 1.30 and Table 1.31).

(v) Adipose versus serum measurements

Arrebola et al. (2012a, b) measured concentrations of three PCB congeners in serum and adipose tissue in adults from Bolivia. PCB-138 had median concentrations of 0.2 ng/mL in serum [33.7 ng/g lipid] and 84 ng/g in adipose tissue [105 ng/g lipid]. The median values for PCB-153 was 0.3 ng/mL in serum [59.0 ng/g lipid], and 52.7 ng/g in adipose tissue [65.8 ng/g lipid]. PCB-180 had median values of 0.1 ng/mL in serum [26.7 ng/g lipid] and 32.8 ng/g in adipose tissue [41.0 ng/g lipid].

(d) Umbilical cord blood, placenta, and fetal tissue

(i) North America

Stewart et al. (2000) reported cord blood PCB concentrations from women living along Lake Ontario and eating contaminated fish. The average cord blood PCB concentration was 0.525 ng/g wet weight [25th percentile, 0.174 ng/g wet weight; 75th percentile, 1.11 ng/g wet weight]. In plasma from umbilical cord in Inuit women from northern Canada, the geometric mean for the sum of 14 PCB congeners was 279.9 ng/g lipid (range, 70.8–1420.1 ng/g lipid) (Muckle et al. 2001). Dallaire et al. (2003) reported changes in concentrations in umbilical cord blood in this population over time, and found a 7.9% annual decrease between 1994 and 2000. Choi et al. (2006) measured 51 congeners in cord blood from women living near a PCB-contaminated site in Massachusetts, and reported a geometric mean of 0.40 ng/g (range, 0.068–18.14), with no consistent relationship with residential distance from the waste site. Consumption of meat and local dairy products (but not fish) were associated with higher cord blood PCB concentrations.

In women from New York state, Schecter et al. (1998) reported the concentration of three dioxin-like PCBs to be 18.2 pg/g lipid in placenta, giving a TEQ of 1.05. The concentrations of 14 single PCB congeners in plasma from Inuit women from Nunavik and southern Quebec were highly correlated with those in placenta (Pearson’s r = 0.77–0.97; P < 0.001), and concentrations in Inuit women were on average four times higher than in women from southern Quebec (Pereg et al., 2002). Doucet et al. (2009) analysed placenta from Canadian women having elective abortions in 1998–2006 and reported annual average total PCB concentrations ranging from 7 to 70 ng/g lipid, with no clear time trend.

(ii) Europe

Koopman-Esseboom et al. (1994) used the concentrations of four congeners (PCB-118, PCB-138, PCB-153, and PCB-180), as measured in umbilical cord blood and in breast milk, as indicators of exposure of the developing fetus and breastfed infant. For these congeners, the correlation coefficients between maternal plasma, cord plasma and human milk were highly significant.

Soechitram et al. (2004) analysed PCBs (PCB-118, PCB-138, PCB-146, PCB-153, PCB-156, PCB-180) and hydroxylated metabolites of PCBs (PCB-107, PCB-136, PCB-146, PCB-153, PCB-172, PCB-187) in samples of maternal plasma and corresponding cord blood in the Netherlands. The calculated ratio for cord versus maternal blood was 1.28 ± 0.56 for PCBs and 2.11 ± 1.33 for hydroxylated PCBs, expressed per gram of lipid. A significant correlation between the respective maternal and cord concentrations for both PCBs and hydroxylated PCBs was found. The results indicated that approximately 50% and 30% of hydroxylated PCBs and PCBs, respectively, was transferred across the placenta to the fetus.

(e) Hair

(i) North America

Altshul et al. (2004) reported median PCB concentrations (sum of 57 congeners) in hair of 2640 ng/g fat (range, 1180–3620 ng/g fat) in a population of students in Boston, USA. Washing hair with shampoo decreased concentrations of PCBs by 25–33% on average, and up to 62% for less chlorinated congeners. [The Working Group considered that the analytical method was reliable and reproducible.] The concentrations of PCBs in hair were higher than in serum. Correlation between concentrations in hair and blood was moderate for the more persistent PCB congeners, with no or little correlation for the other congeners.

(ii) Europe

Covaci et al. (2002b) assessed PCB exposure in hair samples from Greece, Romania, and Belgium. Mean PCB concentrations in samples from Belgium were up to 14 ng/g hair, while concentrations in samples from Greece were about three times lower. Similar ratios of PCB-153 over total PCBs were found for all three countries.

(iii) Asia

One study measured PCB concentrations in the hair of pregnant women in various cities in the Islamic Republic of Iran (see Table 1.29).

1.5. Occupational exposure to PCBs

In 1978, an estimated 12 000 persons in the USA were exposed occupationally to PCBs (Lloyd et al., 1976; NIOSH, 1977). Since the previous IARC evaluations of PCBs (IARC, 1978, 1987), occupational exposures to PCBs have changed, since most industrial countries have banned or partially banned their use. Nevertheless, the earlier references cited previously have been incorporated in the present monograph.

Earlier occupational exposures to PCBs occurred during PCB manufacture, capacitor and transformer manufacture and repair, production of carbonless copy paper, and accidental releases from these processes. More recent occupational exposures to PCB usually occur through PCB emissions via waste incineration, fires, and waste recycling.

1.5.2. Capacitor manufacture

Before PCBs were banned, capacitor manufacturers filled (impregnated) casings with wound paper and foil/plastic with the PCB-containing oil before the top was fastened (crimped, sealed, soldered). PCB exposure (probably via the dermal route) occurred during filling: the capacitors were either flood-filled or manually filled, resulting in spills and worker exposures. The brand of PCB oil used differed geographically (Aroclors were used in the USA, Pyralene/Phenochlor in Sweden and Italy) and temporally (the percentage chlorination was reduced, e.g there was a switch from Aroclor 1254, with 54% chlorination, to Aroclor 1242, with 42% chlorination.

Other chemical exposures in capacitor manufacturing were possible, such as from other impregnation oils (e.g. mineral oils), degreasing agents such as trichloroethylene (Brown & Jones, 1981; Bertazzi et al., 1987), dibenzofurans (Gustavsson et al., 1986), chlorinated naphthalenes, lead solder, epoxies, and methyl ethyl ketone (MEK) (Mallin et al., 2004; Persky et al., 2012). Ageing capacitors can from time to time explode [due to the physical stress of metal ageing], thus further exposing workers.

Fig. 1.9 shows ranges of air PCB concentrations in capacitor-manufacturing sites (Ouw et al., 1976; Brown & Jones, 1981; Bertazzi et al., 1982; Fischbein et al., 1982; Lawton et al., 1985; Gustavsson et al., 1986). The earlier concentrations (measured in the 1950s–1960s) were highest (Bertazzi et al., 1982), and decreased in later years.

Fig. 1.9

Air PCB concentrations in capacitor manufacturing plants (μg/m³) by year

In the 1950s, maximum PCB concentrations in workroom air in several plants in the USA (Massachusetts) were reported to be 200–10 500 μg/m³ (Elkins, 1959). No details on the number of plants surveyed, nor the number of samples collected, or the work performed in the plants were given, but four different jobs were surveyed: for impregnating with PCBs, average air concentrations ranged from 200 to 5800 μg/m³; for soldering, 800 μg/m³; mixing oil, 600 μg/m³, and regulator filling, 100 μg/m³. No toxic effects were noted at these concentrations; however, it was noted that air PCB concentrations of > 10 000 μg/m³ were “unbearably irritating.” This is contrary to a report in a Japanese capacitor-manufacturing plant where a dermatitis outbreak occurred when air PCB concentrations reached 100 μg/m³ (Meigs et al., 1954). Air PCB concentrations between 1953 and 1957 in a Japanese capacitor-manufacturing factory ranged from 400 to 6700 μg/m³ (NIEHS, 1976).

Ouw et al. (1976) reported that workers in the electrical industry in Australia were exposed to Aroclor 1242 at air concentrations of 320–2220 μg/m³, with a mean of 1270 μg/m³; and were found to have PCB blood concentrations of approximately 0.4 g/kg bw. Contact with PCBs was primarily via the skin.

Brown & Jones (1981) measured air concentrations of Aroclor 1016 in two plants in the USA (New York and Massachusetts) plants in 1977. The time-weighted averages (TWA) were different for the two plants, with air concentrations at the New York plant being lower than at the Massachusetts plant. For the New York plant, PCB concentrations in personal air samples ranged from 24 to 393 μg/m³ (n = 28), and in area air samples from 3 to 476 μg/m³ (n = 19). For the Massachusetts plant, PCB concentrations in personal air samples ranged from 170 to 1260 μg/m³ (n = 29), and in area air samples from 50 to 810 μg/m³ (n = 25). Air PCB concentrations (TWA) were extremely high during capacitor impregnation (New York: 160 μg/m³, Massachusetts: 850 μg/m³), degreasing (Massachusetts; 1260 μg/m³), and sealing/soldering (New York: 393 μg/m³, Massachusetts: 720 and 1060 μg/m³). Capacitors that failed were sent for repair where they were re-opened and manually drained. Repair workers’ PCB exposures were measured as 298 μg/m³ (recovery), and 50 μg/m³ (repair) in the New York plant. [These workers would also have had extensive dermal exposures, which were not assessed.]

Eight studies have reported PCB concentrations in workers’ blood in Australia, Finland, Italy, Germany, and the USA (Karppanen & Kolho, 1972; Ouw et al., 1976; Maroni et al., 1981a; Bertazzi et al., 1982; Acquavella et al., 1986; Wolff et al., 1992; Kannan et al., 1994; Seegal et al., 2011; Persky et al., 2012). The reporting of PCB blood concentrations was not uniform, which hindered comparison across studies.

Karppanen & Kolho (1972) compared blood PCB concentrations in workers in a capacitor factory in Finland where Aroclor 1242 had been used as the impregnating fluid: the groups comprised laboratory workers handling PCBs (n = 6), impregnation workers (n = 11) employed for 4 years, and a control group (n = 9) that had never been professionally exposed to PCBs. Blood PCB concentrations were approximately 50 times greater in impregnation workers (0.07–1.9 μg/g) than in the control group (0.003–0.012 μg/g), and were also higher in laboratory workers (0.036–0.062 μg/g) than in the controls. The pattern of PCB congeners in the exposed workers differed markedly from that of the PCBs actually used. More highly chlorinated PCBs persisted in the blood, while the less chlorinated PCBs contained in Aroclor 1242 had been eliminated from the body. [Consequently, the total PCB intake must have been higher than that reflected by the levels detected in blood.]

Total serum PCB concentrations have been reported to be 18.2 ppb (Acquavella et al., 1986) in a clinical survey among 205 workers at a capacitor-manufacturing plant in the USA. This mean value represented workers (n = 205) with (39%) and without (61%) potential for PCB exposure in their jobs. [The Working Group noted that PCB concentrations were not reported separately for workers with and without occupational PCB exposure.] Log serum PCB concentrations were found to be significantly correlated with duration of employment, age, cumulative occupational exposure, and fish and wine consumption, as confirmed by multiple linear regression.

In a study of mortality in Italy (Bertazzi et al., 1982), workers in the autoclave room were exposed to air PCB concentrations of 5200–6800 μg/m³ in 1954 (n = 3), and 48–275 μg/m³ in 1977 (n = 9). Eighteen workplace surface-wipe samples showed extensive PCB contamination (0.2–159 μg/cm²), as did nine hand-wipe samples (0.3–9.2 μg/cm²). Workers’ serum PCB concentrations were reported by type of PCBs: for highly chlorinated PCBs (54% chlorination) (n = 67), the mean was 230.5 ppb (SD, 174.5), while for less chlorinated PCBs (42% chlorination) (n = 67) the mean was 114.1 ppb (SD, 79.6). In a later study (Bertazzi et al., 1987), the corresponding values were 202.8 ppb (SD, 111.7; n = 37) and 42.9 ppb (SD, 34.7; n = 37), respectively.

Wolff et al. (1992) studied PCB blood concentrations in capacitor workers in 1976 and 1979 in the USA. For the first sampling year, the mean concentration of the less chlorinated PCBs (di-, tri-, and tetrachlorobiphenyls) was 55 ng/mL (range, 6–2257 ng/mL), and for the latter year was 41 ng/mL (range, 6–350 ng/mL). Mean concentrations of highly chlorinated PCBs were 10 ng/mL (range, 1–308 ng/mL) in 1976, and 13 ng/mL (range, 2–350 ng/mL) in 1979. These capacitor workers (n = 60) were also surveyed by the National Institute for Occupational Health and Safety in 1977 (NIOSH, 1977), when the following blood PCB concentrations were reported as follows: less chlorinated PCBs (quantified as Aroclor 1242), 2–3300 ppb (ng/mL), and more highly chlorinated PCBs (quantified as Aroclor 1254), 5–250 ng/mL.

About 30 years later (in 2003–2006), Seegal et al. (2011) measured blood concentrations of individual PCB congeners in some of these former capacitor workers, and found that concentrations had dropped statistically significantly: mean concentration of less chlorinated PCBs (PCBs 28, 56, 66, 74, 99, 101) was 2.84 ng/g or 0.45 μg/g lipid in men, and 2.29 ng/g or 0.34 μg/g in women; mean concentration of highly chlorinated PCBs (PCBs 105, 118, 138, 146, 153, 156, 167, 170, 172, 174, 177, 178, 180, 183, 187, 199, 203) was 4.09 ng/g or 0.65 μg/g lipid in men, and 3.21 ng/g or 0.47 μg/g lipid in women; and total PCB concentration was 7.47 ng/g or 1.19 μg/g lipid in men, and 5.81 ng/g or 0.86 μg/g lipid in women.

Maroni et al. (1981a) carried out a study in two Italian electrical-capacitor manufacturing plants using PCBs as a dielectric fluid. Plant A produced electric capacitors filled with a mixture of mineral oils and PCBs. PCBs with 54% chlorination were used from 1949 to 1965, and subsequently replaced with Pyralene 3010 with 42% chlorination. The power-capacitor casings were filled with PCBs in autoclaves, and were manually removed when cooled from 70 °C to 40 °C before they were welded, tested, and finished externally. Electric “filters” (small capacitor systems used in electrical household appliances) were impregnated with PCBs. Plant B performed short-circuit testing of high-power capacitors filled with Apirolio, a PCB mixture with 42% chlorination. Stress-testing the capacitors often included explosions. Airborne PCBs were mainly trichlorobiphenyls and concentrations ranged from 48 μg/m³ (filter operations) to 275 μg/m³ (power-capacitor manufacturing). Surface-wipe samples showed both tri- and pentachlorobiphenyl mixtures, with the highest amounts being found on the capacitor basket rolling carrier: trichlorobiphenyls, 127 mg; and pentachlorobiphenyls, 15 mg. Plant A employed 67 workers (40 women and 27 men): 48 were currently employed in the capacitor-manufacturing departments, 16 had been employed there for at least 6 months before the beginning of the study, and 3 had always been employed in other non-manufacturing departments without direct exposure to PCBs. PCB recovery from the palms of the hands of power-capacitor workers (plant A) showed total PCB (tri- and pentachlorinated biphenyls) skin-surface concentrations to be 4–28 μg/cm². Mean (± SD) blood PCB concentrations differed between current (377 ± 190 μg/kg) and past exposed workers (292 ± 161 μg/kg); workers with occasional exposure had the lowest mean total PCB exposures (110 ± 31 μg/kg). Blood PCB concentrations by job performed were highest for welders (1259 μg/kg), followed by impregnation workers (556 ± 337 μg/kg), assembly of capacitors (406 ± 173 μg/kg), and finally assembly of filters (246 ± 130 μg/kg). The blood PCB concentrations were not correlated with duration of exposure, but with the percentage ratio of hours per year spent with direct exposure to PCBs. Plant B included 13 workers (all men) exposed to PCBs during handling of the capacitors contaminated with Apirolio, dispersed from explosions sometimes caused by stress-testing. Blood PCB concentrations in currently exposed workers in plant B (200 ± 146 μg/kg) were between occasionally exposed (110 ± 31 μg/kg) and past exposed workers (292 ± 161 μg/kg) in plant A. Although the PCB mixture used in both plants had a chlorine content of 42%, the workers differed in their ratio of penta- to trichlorobiphenyls; plant A workers had higher concentrations of pentachlorobiphenyls than of trichlorobiphenyls, while the reverse was true in plant B workers. This difference was attributed to the heavy past exposure to highly chlorinated PCBs used until 1965 in plant A. Workers with abnormal liver findings (n = 16) had twice the concentrations of tri- (215 ± 95 μg/kg) and pentachlorobiphenyls (308 ± 306 μg/kg) compared with workers (n = 64) without abnormal liver findings (tri- and pentachlorobiphenyl concentrations were 92 ± 64 and 176 ± 108 μg/kg, respectively) (Maroni et al., 1981b). Duration of exposure did not explain this observed difference.

One German capacitor-manufacturing worker was reported to have a blood PCB-169 concentration of 11 ng/g (Kannan et al., 1994).

After a capacitor explosion at a Finnish paper mill, workers’ (n = 15) blood PCB concentrations were 3.5–48.3 μg/L (Luotamo et al., 1984). [These levels were much lower than during capacitor manufacturing itself.]

In a recent cross-sectional study, Persky et al. (2012) reported blood PCB concentrations separately for diseased (having diabetes) and non-diseased (without diabetes) workers. In diseased workers, the concentrations were: DL-PCBs, 2.5 ng/g; NDL-PCBs, 17.0 ng/g; estrogenic PCBs [PCB-52, 99, 101, 110, 153], 3.6 ng/g; anti-estrogenic PCBs [PCB-105, PCB-156], 3.6 ng/g; and PCB-74, 4.9; PCB-99, 1.0 ng/g; PCB-118, 1.4 ng/g; PCB-138, 2.5 ng/g; PCB-146, 0.4 ng/g; PCB-153, 2.8 ng/g; PCB-156, 0.6 ng/g; PCB-170, 0.7 ng/g; PCB-180, 1.1 ng/g; PCB-187, 0.3 ng/g; PCB-194, 0.2 ng/g; PCB-201, 0.2 ng/g; PCB-203, 0.2 ng/g; and PCB-206, 0.1 ng/g. In non-diseased workers, the concentrations were: DL-PCBs, 0.4 ng/g; NDL-PCBs, 4.3 ng/g; estrogenic PCBs, 1.0 ng/g; anti-estrogenic PCBs, 0.1 ng/g; PCB-74, ng/g; 0.8, PCB-99; 0.3 ng/g; PCB-118, 0.2 ng/g; PCB-138, 0.6 ng/g; PCB-146, 0.1 ng/g; PCB-153, 0.8 ng/g; PCB-156, 0.1 ng/g; PCB-170, 0.2 ng/g; PCB-180, 0.4 ng/g; PCB-187, 0.1 ng/g; PCB-194, 0.1 ng/g; PCB-201, 0.1 ng/g; PCB-203, 0.1 ng/g; and PCB-206, 0.04 ng/g.

1.5.12. Firefighters and rescue workers

Firefighters and rescue workers have also been surveyed for PCB exposure in several recent studies, demonstrating a wide variability in serum PCB concentrations (Table 1.33; Kelly et al., 2002; Schecter et al., 2002; Dahlgren et al., 2007; Chernyak et al., 2009, 2012).

Table 1.33

Serum PCB concentrations in firefighters.

1.6. Exposure assessment of epidemiological studies

1.6.2. Studies of environmental exposure

Cohort studies of environmental exposure have used many approaches to assess exposure to PCBs. Exposure approaches include interview, questionnaires, cumulative PCB exposure, dietary intake of fatty fish, PCB concentrations in biological media such as blood, adipose tissue, and breast milk, and in the environment such as carpet dust, or any combinations of these. Biological measures of body burden have been used extensively (see Table 1.34).

Table 1.34

Common measures of exposure to PCBs and design of the exposure assessment in epidemiological studies in non-occupational settings.

1.7. Regulations and guidelines

1.7.2. Environmental regulations

(a) European Union and Member States

The Member States of the European Union have taken actions to eliminate the production, use, and release of PCBs since 1985. In 2004, to implement the Stockholm Convention on POPs, by regulation EC/850/2004 (EU 850/2004), the production, placing on the market, and use of PCBs were prohibited. Low POPs concentration limits were adopted through Council Regulation (EC) No. 1195/2006 (EU 1195/2006) amending Annex IV to Regulation (EC) No 850/2004. Within the European Union of 26 Member States, several measures have been adopted to reduce the presence of PCDDs, PCDDs, PCDFs, and PCBs in the environment, in food and in feed. These include:

• Commission Regulation (EC) No. 1883/2006 of 19 December 2006 laid down methods of sampling and analysis for the official control of levels of dioxins and DL-PCBs in certain foodstuffs;
• Commission Recommendation 2006/88/EC of 6 February 2006 concerning the reduction of the presence of dioxins, furans and PCBs in feedingstuffs and foodstuffs;
• Commission Recommendation 2006/794/EC of 16 November 2006 on the monitoring of background levels of dioxins, DL-PCBs and NDL-PCBs in foodstuffs.
• The most recent Commission Regulation (EU) No. 1259/2011 amended Regulation EU 1881/2006 as regards maximum levels for DL-PCBs and NDL-PCBs (EC, 2011a); it also changed the formerly used 1998 WHO TEFs to the scheme adopted in 2005 (referred to as WHO₂₀₀₅-TEFs) (Van den Berg et al., 2006) and includes maximum levels for NDL-PCBs in food.

See Table 1.35

Table 1.35

Maximum permitted levels for dioxin-like compounds and indicator PCBs in the European Food and Feed regulation.

(b) North America

(i) USA

The United States Food and Drug Administration has established tolerance levels in various foods in an attempt to reduce human exposure to PCBs (FDA, 2013). [These limit values were set in 1971 and 1977, before any epidemiological and most experimental studies were conducted, and have not been revised since.] The temporary tolerance levels for PCB residues are as follows:

• 1.5 ppm in milk (fat basis);
• 1.5 ppm in manufactured dairy products (fat basis);
• 3 ppm in poultry (fat basis);
• 0.3 ppm in eggs;
• 0.2 ppm in finished animal feed for food-producing animals (except the following finished animal feeds: feed concentrates, feed supplements, and feed premixes);
• 2 ppm in animal feed components of animal origin, including fishmeal and other by-products of marine origin and in finished animal feed concentrates, supplements, and premixes intended for food-producing animals.
• 2 ppm in fish and shellfish (edible portion). The edible portion of fish excludes head, scales, viscera, and inedible bones;
• 0.2 ppm in infant and junior foods;
• 10 ppm in paper food-packaging material intended for or used with human food, finished animal feed and any components intended for animal feeds. The tolerance does not apply to paper food-packaging material separated from the food therein by a functional barrier that is impermeable to migration of PCB.

The United States Environmental Protection Agency (EPA) has set a maximum contaminant level for PCBs of 0.0005 mg/L (500 ppt) in drinking-water. The EPA requires that spills or accidental releases into the environment of 1 pound (0.45 kg) or more of PCBs be reported to the EPA (ATSDR, 1996).

(ii) Canada

The import, manufacture, and sale (for re-use) of PCBs were made illegal in Canada in 1977. Release of PCBs to the environment was made illegal in 1985. However, use of PCB-containing equipment is allowed until the end of its service life. The storage of PCBs has been regulated since 1988. Export has been regulated since 1997. These provisions are maintained in the Chlorobiphenyls Regulations, under the Canadian Environmental Protection Act, 1999 (CEPA, 2011).

The regulation of waste is consistent with the Basel Convention's “Technical guidelines for the environmentally sound management of wastes consisting of, containing, or contaminated with persistent organic pollutants” (Basel Convention, 2007, 2015).

(c) Australia and New Zealand

(i) Australia

The Industrial Chemicals (Notification and Assessment) Act 1989 was amended to give effect to the Stockholm Convention (NICNAS, 1989).

The National Strategy for The Management of Scheduled Waste was endorsed by the Australian and New Zealand Environment and Conservation Council in 2003 (ANZECC, 2003) and provides for the safe management and disposal of organochlorine pesticides, PCBs and hexachlorobenzene. The PCB Management Plan provides treatment provisions for different types of PCB waste including liquid residues and discharges, gaseous emissions, solid residues and disposal (Australian Government, 2006, 2007).

(ii) New Zealand

The Hazardous Substances and New Organisms (HSNO) Act 1996 (as amended by the HSNO [Stockholm Convention] Act Amendment 2003), prohibits the production, use and import of the chemicals listed in Annex A of the Convention, including PCBs. Exempted use of PCBs as per the Toxic Substances Regulations 1983 is permitted, but subject to phase-out no later than December 2016. The HSNO Act 1996 is administered by The New Zealand Environment Risk Management Authority (ERMA) by: assessing new chemicals, pesticides or industrial chemicals currently in use that exhibit POP characteristics (Articles 3.3 and 3.4); permitting the appropriate use of POPs for laboratory-scale research or as a reference standard (Article 3.5); managing the existing exempted use and storage of PCBs (Article 3.6); prohibiting import, manufacture, or use of POPs (Article 3.1 and 3.2). The Imports and Exports (Restrictions) Act 1988, via the Imports and Exports (Restrictions) Prohibition Order (No. 2) 2004, prohibits export of POPs (except as conditionally provided under Article 3.2). Import and export are regulated under The Imports and Exports (Restrictions) Act 1988.

(d) Asia

(i) China

China implements an import and export registration system, included under its Regulations on Environmental Management of Chemicals and the Import and Export of Toxic Chemicals of 1994. In 2005, PCBs were included in the List of Toxic Chemicals Strictly Prohibited from Import and Export, by No. 116 Notice on the List of Goods Prohibited from Import (the Sixth Group). The National Implementation Plan under the Stockholm Convention entered into force for China in 2004, and also applied to the Special Administrative Regions of Hong Kong and Macao (NIP China, 2007). This plan aims to prohibit and prevent the production and import of PCBs, and to achieve the environmentally sound management of currently used equipment containing PCBs. China used to produce PCBs, but production was stopped in the 1970s. The plan called for establishing a system for the declaration, registration, and environmentally sound management of equipment in use containing PCBs by 2010. Identification of high-risk equipment currently in use across the country is to be achieved by 2015, with uses of PCBs eliminated by 2025.

Furthermore, China has also stipulated special administrative regulations and standards with regard to PCBs. The Notice on the Issues Concerning Prevention of Pollution Caused by Hazardous Polychlorinated Biphenyls was promulgated in 1979 to ban future imports of power equipment containing PCBs. The Notice on Enhancement of the Management over Waste Polychlorinated Biphenyl Power Capacitors was issued in 1990 to forbid trading and dismantling downstream capacitors containing PCBs. The Provisions on the Pollution Caused by Power Installations Containing Polychlorinated Biphenyls and Related Wastes of 1991 addresses the declaration, transfer, transport, import, treatment, disposal, sealing-up and storage of PCB wastes and other sources. The Control Standard on Polychlorinated Biphenyls for Wastes (GB13015–91) was implemented in 1991, in which the value of the control standard on PCBs wastes and the treatment methods for wastes containing PCBs are stipulated (NIP China, 2007).

(ii) Taiwan, China

Importation of PCBs was prohibited in 1980. The Environmental Protection Administration of Taiwan, China, banned the manufacture, sale, and use of PCBs in 1988. An extensive investigation of electrical devices in 1990–1991 indicated that more than 80 000 PCB-containing electrical devices were still in use, mainly capacitors and transformers. A full-scale ban on the use of PCBs, with the exception of experimental, research, and educational purposes, took effect in January 2001. This prohibited use of any electrical devices containing PCBs by the end of 2000, mandating immediate disposal at end of use of capacitors and transformers containing PCBs (Environmental Protection Administration, 1988). Furthermore, PCBs may not be detectable in effluents from business, sewage systems and building sewage-treatment facilities.

(iii) India

According to Schedule VI of the Hazardous Waste (Management, Handling and Transboundary Movement) Rules 2008, the import and export of hazardous wastes, substances and articles containing or consisting of or contaminated with PCBs are prohibited (Ministry of Environment and Forests, 2008).

(e) Africa

United Republic of Tanzania

The Industrial and Consumer Chemicals (Management and Control) Act of 2003 provides for the management and control of PCBs under the list of severely restricted/banned/eliminated chemicals in Schedule 8. The government of the United Republic of Tanzania issued an Environmental Management Act (Government of the United Republic of Tanzania, 2004) that specifically provides for the control and management of current and future POPs, requiring submission of an annual report on implementation.

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