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Tomas Trnovec, Anton Kocan, and Lubica Palkovicova Slovak Medical University, Bratislava, Slovak Republic
- 1 Introduction
- 2 Relevant exposure metrics
- 3 Health effects of short- and long-term exposures
- 4 What health endpoints might be quantified
- 5 Is there evidence of a threshold or "safe" level
- 6 What sub-groups of the population are most susceptible or otherwise will need special consideration in quantification
- 7 Polychlorinated biphenyls as biomarkers
- 8 References
- 9 See also
PCBs cover a group of 209 different PCB congeners which can be divided into two groups according to their toxicological properties. One group, consisting of 12 congeners, shows toxicological properties similar to dioxins, is therefore termed "dioxin-like PCBs" (DL-PCBs). The other PCBs are referred to as "non dioxin-like PCBs" (NDL-PCBs). Both groups of PCBs, NDL-PCBs as well as DL-PCBs, are usually found in the various components of the environment, man included. There are no known natural sources of PCBs. PCBs are either oily liquids or solids that are colorless to light yellow. Some PCBs can exist as a vapor in air.
Many commercial PCB mixtures have been used as dielectric fluids, coolants and lubricants in transformers, capacitors, and other electrical and mechanical equipment. The manufacture of PCBs was stopped in the U.S. in 1977 and afterwards in many other countries, because of evidence they build up in the environment and can cause harmful health effects (Faroon OM et al., 2003; U.S. EPA/625/3-91/020. 1991).
In the past, PCBs were released to wastewater from its industrial uses. Today, PCBs are still detected in water due to the environmental recycling of the compound. Most of the PCBs in water are bound to the soil and sediments and may be released to the water slowly over a long period of time. Re-suspension of dried particles plays an important role in spreading of this pollutant. PCBs may enter the food chain through ingestion by aquatic organisms and fish.
Relevant exposure metrics
More than 90% of the NDL-PCB exposure in the general population is via food. Congener patterns in feed, particularly that of plant origin and in edible tissue may differ considerably. Due to the different sources of contamination, different origins of the feed and of food commodities, there is generally no correlation between the concentrations of NDL-PCBs and DL-PCB Toxic Equivalents (TEQ) or the total TEQ (polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs) and DL-PCBs) with the exception of samples where the circumstances of contamination are known. Average daily dietary intakes of total NDL-PCBs can be estimated to be in the range of 10-45 ng/kg body weight (b.w.) per day. Limited exposure data for young children, up to six years of age, indicates that the average intake (breastfeeding excluded) of total NDL-PCBs is about 27-50 ng/kg b.w. per day. However, where data on both adults and children within a specific population were available, in general children had exposure levels 2.5 fold higher than adults. In specific subpopulations with high dietary PCB exposure such as Baltic Sea fishermen the daily intake of the sum of the six NDL-PCBs from fish could be about 40 ng/kg b.w., corresponding to an intake of total NDL-PCBs of 80 ng/kg b.w. per day before taking into account the rest of the diet. Breastfed infants represent a group of high NDL-PCB intake which might be two orders of magnitude higher than adult exposure(Opinion of the scientific panel on contaminants, 2005).
The dietary exposure to PCDDs/PCDFs and DL-PCBs exceeds the Tolerable Weekly Intake (TWI) or the Tolerable Daily Intake (TDI) for a considerable part of the European population: the Scientific Committee on Food (SCF) of the EU adopted on 30 May 2001 an opinion on the Risk Assessment of dioxins and DL-PCBs in food. The Committee established a group TWI for dioxins and DL-PCBs of 14 pg WHO TEQ /kg b.w.. This TWI is in line with the provisional Tolerable Monthly Intake of 70 pg/kg body weight/month established by the Joint FAO/WHO Expert Committee on Food Additives (JECFA) at its fifty-seventh meeting (Rome, 5-14 June 2001) and concurs with the lower end of the range TDI of 1-4 pg WHO-TEQ/kg body weight, established by the World Health Organization (WHO) Consultation in 1998. Representative dietary intake data indicated that the average dietary intakes of PCDDs/PCDFs and DL-PCBs in the EU were in the range of 1.2-3 pg/kg b.w. and day which meant that a considerable part of the European population exceeded the TWI or TDI (COMMISSION OF THE EUROPEAN COMMUNITIES 2001)
Health effects of short- and long-term exposures
Several excellent review papers (Integrated Risk Information System; Opinion of the scientific panel on contaminants, 2005; Rice DC, 2005; Faroon OM et al., 2003; Damstra T et al., 2002; COMMISSION OF THE EUROPEAN COMMUNITIES 2001) have been published in which information on various health outcomes following exposure to PCBs in animals, either wild or laboratory, and humans can be found. Current epidemiological data indicates that adverse health effects might be associated with PCBs resulting from occupational exposure, accidental or environmental food contamination (Opinion of the scientific panel on contaminants, 2005). The relationship between exposure to PCBs and human health effects is reflective of the large variation in human exposure to the many different congeners and contaminants present in PCB formulations and to combustion by-products of PCBs (Faroon OM et al., 2003).
The following health outcomes following exposure of humans to PCBs have been outlined in the relevant reviews (Integrated Risk Information System; Opinion of the scientific panel on contaminants, 2005; Rice DC, 2005; Faroon OM et al., 2003; Damstra T et al., 2002): carcinogenicity (the human studies are being updated; currently available evidence is inadequate, but suggestive), genotoxicity, reproductive toxicity including impact on fertility, growth and development, immunological effects, neurological effects, irritation and sensitization, alteration of thyroid gland, cardiovascular system, and specific observations in infants and children (effects on perinatal growth and early postnatal development, nervous system development, thyroid gland, and immune functions). Recently, various other diseases were related to PCB exposure (e.g. diabetes (Vasiliu O et al., 2006) and arthritis (Lee DH et al., 2007).
PCBs belong among endocrine disruptors. Endocrine disruption is not considered a toxicological endpoint per se but a functional change that may lead to adverse effects. For the purposes of these scoping notes, endocrine disruptors are defined in a generic sense as follows: An endocrine disruptor is an exogenous substance or mixture that alters function(s) of the endocrine system and consequently causes adverse health effects in an intact organism, or its progeny, or (sub)populations (Damstra T et al., 2002).
What health endpoints might be quantified
PCB exposure-response functions for humans are rather largely unknown. This is why in risk assessment in connection with human PCB exposures the procedure relies mainly on data obtained with experimental animals. This requires introduction of various uncertainty factors when applying to human population.
Epidemiological studies are crucial to risk assessment, but offer only incomplete insight due to the limitations of observational studies, which depend, e.g., on study opportunities. Often the follow-up is too short, or the exposure estimates do not reflect the long induction time of the health outcomes under study. These problems tend to cause an underestimation of the true effects caused by PCBs. On the other hand, confounding from concomitant exposures to other toxic substances may be difficult to control for and may therefore cause an overestimation of the effects. Due to the high correlation among PCB congeners, such population studies are unlikely to reflect toxicity differences between congeners. Although the potential effects have been insufficiently studied so far, accumulating evidence suggests that clinically important deficits in neurobehavioral and immune functions may occur at widely prevalent exposure levels (Grandjean P, 2003 ). There is evidence of a real association between exposure to PCBs and human health effects and ways for its quantification will be commented further.
A number of types of cancer, as well as total cancer incidence, have been related to accidental and occupational exposure to PCDDs/PCDFs (mostly 2,3,7,8-TCDD). In addition, an increased prevalence of diabetes and increased mortality due to diabetes and cardiovascular diseases have been reported. In children exposed to PCDDs/PCDFs and/or PCBs in utero, effects on neurodevelopment, neurobehaviour and effects on thyroid hormone status have been observed at exposures at or near background levels. At higher exposures, due to accidental and occupational exposure, children exposed transplacentally to PCBs and PCDDs/PCDFs show skin defects (such as chloracne), tooth mineralization defects, developmental delays, behavior disorders, decrease in penile length at puberty, reduced height among girls at puberty and hearing loss. Although 2,3,7,8-TCDD is known as a human carcinogen, cancer is not considered to be the critical effect for the derivation of the Tolerable Intake. The critical effects are neurobehavioral changes, endometriosis and immunosuppression. PCBs are classified as probable human carcinogens and produce a wide spectrum of adverse effects in animals, including reproductive toxicity, immunotoxicity and carcinogenicity (COMMISSION OF THE EUROPEAN COMMUNITIES 2001).
The standard procedure for assessing noncancer risks associated with hazardous compounds has been to use a no observed adverse effect level (NOAEL) approach. Criticisms of the NOAEL approach has been widely spread during the last decade (Faustman EM 1996). The Environmental Protection Agency (EPA) Science Advisory Board has challenged the regulatory scientific community to develop improved methods for RfD (reference dose) calculation (USEPA, 1991; USEPA, 1995).
Development of benchmark dose methods is one approach that has been taken to address the challenge. The topic of benchmark dose analysis and its use has been treated by many authors, a sample of which is presented here (Benchmark Dose (BMD) Methodology; Benchmark Dose Technical Guidance Document, 2000; Bokkers BG, Slob W, 2005; Bokkers BGH, Slob W, 2007; Dakeishi M et al., 2006; Moerbeek M et al., 2004; Morales KH, Ryan LM, 2005; Sand S, 2005; Slob W, 2002; Slob W et al., 2005; van Wijngaarden E, et al., 2006). The BMD analysis has already been applied to environmental PCB exposurestwo times (Opinion of the scientific panel on contaminants, 2005).
Perspective of further application of the BMD approach is in the following: During the recent decade, in addition to the so far well documented highly PCB polluted sites (Anniston, Alabama, USA; Faroe Islands; Inuits residing in arctic regions of Quebec and Greenland), another region with significant PCB exposure was described in eastern Slovakia (Kocan et al., 2001; Kocan et al., 2004). Until now, papers were published and manuscripts were submitted or prepared on various health outcomes observed in this area: birth weight (Sonneborn D et al, 2007), thymus size (Park H.Y. et al., 2007 ), thyroid gland (Langer P et al.,1998; Langer P et al., 2004; Langer P et al., 2005; aLanger P et al., 2007), hearing impairment ((Trnovec T et al., submitted), neurobehavioral performance (Šovčíková E et al., 2003, Šovčíková E et al., 2004), dental enamel deficits (Jan J. et al., 2004; Jan J. et al., 2006), and glycaemia homeostasis (Radikova Z et al., 2004; bLanger P et al., 2007). In all the papers mentioned above, cited data on PCB serum concentrations is available. In no case surrogate exposure data were used. An attempt will be undertaken to analyze these data using the benchmark dose approach.
Is there evidence of a threshold or "safe" level
The applicability of the threshold approach to exposure to PCBs will be commented as follows:
The threshold of toxicological concern (TTC) (Kroes R et al., 2004) is a pragmatic risk assessment tool that is based on the principle of establishing a human exposure threshold value for all chemicals, below which there is a very low probability of an appreciable risk to human health. Polyhalogenated dibenzo-p-dioxins, dibenzofurans and biphenyls are excluded from consideration with respect to TTC, because the linearised low-dose method used for estimation of cancer risk is not appropriate.
A particular problem with dose-response relationships arises with chemicals that have endocrine disrupting properties. A major issue closely related to the threshold concept pertaining to endocrine disrupting chemicals (EDCs), to which PCBs belong (Damstra T et al., 2002; Pliskova M et al., 2005), and that must be resolved urgently, is whether or not they pose health hazards at low-dose levels.
What sub-groups of the population are most susceptible or otherwise will need special consideration in quantification
The toxic properties of PCDDs, PCDFs and PCBs seem to have been underestimated and new epidemiological, toxicological and mechanistic data have emerged in particular with respect to neurodevelopmental, reproductive and endocrine effects, which indicate that PCDDs, PCDFs and some PCBs have a broader impact on health than previously assumed, even in very low doses and in particular on the most vulnerable groups like breast-fed infants and developing foetus, which is directly exposed to the accumulated maternal body burdens throught transplacental transfer of these compounds (COMMISSION OF THE EUROPEAN COMMUNITIES 2001).
Polychlorinated biphenyls as biomarkers
- The text on this topic is taken from an equivalent page of the IEHIAS-project.
Sample collection and storage
PCBs are generally analyzed in blood (plasma or serum), adipose tissue, breast milk and cord blood
- Exposure generally occurs through food, mainly dairy products, meat and fish.
- In newborns, PCBs are transferred in utero and through breast feeding.
- Daily intake estimates of total PCBs range from 0,05-1 µg/kg bw.
- Excretion of PCB congeners is dependent on their rate of metabolism to more polar compounds.
- Half-lives vary between PCB congeners and range from a few days to 450 days.
- Breast milk, blood and other matrices should be stored using standard procedures.
- Samples are generally stored frozen at -20°C.
- PCBs are generally analysed using GC/ECD, high-resolution MS, or related techniques.
- Bioanalytical methods use receptor based assays.
- Levels of detection are around 11-19 ng/kg tissue.
The variation coefficients for interlaboratory PCB measurements range 7-14%
There are internationally accepted validated methods for PCB analysis in different matrices
- PCBs bioaccumulate in fatty tissues, so age is associated with higher body burden.
- As food is a major source of PCBs, dietary habits have an important influence on PCB exposure.
- There is also an 11% decline in total PCBs per 5-unit increase in BMI.
Concentrations reported in literature:
- Median concentrations of PCB-153 in human milk are in the range of 8-155 ng/g fat.
- Large differences among constituents, regions and times are observed
Dose-response relations are not easily made for PCBs because of the different toxicological properties of various congeners and isomers and the large variability in kinetics. PCBs are classified as carcinogens, particularly with regard to the liver. Also reproductive and developmental effects may be related to PCB exposure
Time trend, geographical variation, susceptibel groups:
No data available
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