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2.14 Dioxins and dioxin-like polychlorinated biphenyls (PCBs)
The term “dioxins” encompasses a group of 75 polychlorinated dibenzo-p-dioxin (PCDD) and 135 polychlorinated dibenzofuran (PCDF) congeners. Although dioxins are not produced by intention except for research and analytical purposes, these contaminants have a ubiquitous distribution due to their formation as unwanted and often unavoidable by-products in a number of anthropogenic activities. PCDDs and PCDFs are formed during incomplete combustion processes, industrial as well as natural. They occur also as contaminants during various industrial processes, eg, the chemical manufacture of some chlorinated compounds and chlorine bleaching of paper pulp.
The toxicity of individual dioxin congeners differs considerably. The congeners that are of toxicological importance are substituted in each of the 2-, 3-, 7- and 8-positions. Thus, from 210 theoretically possible congeners, only 17 are of toxicological concern. These compounds have a similar toxicological profile to that of the most toxic congener, 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD). Other compounds, such as selected polychlorinated biphenyls (PCBs) and polycyclic aromatic hydrocarbons, also exhibit “dioxin-like” toxicity. The current review considers the 17 dioxin congeners of toxicological concern and 12 dioxin-like PCBs; PAHs were considered earlier in this report.
Several comprehensive reviews of the toxicity of dioxins and dioxin-like PCBs have been undertaken (ATSDR, 1998; EC-SCF, 2000; Van Leeuwen and Younes, 2000; FSA, 2001; FAO/ WHO, 2002; US EPA, 2003). The discussion below summarises relevant data from these reviews. Particular attention is given to those studies that have been used in deriving reference health standards. Readers are referred to the original reviews for more details on adverse health effects.
2.14.1 Toxicological status
The most widely studied of all the dioxin-like compounds is 2,3,7,8-tetrachlorodibenzo-p-dioxin (2,3,7,8-TCDD, or TCDD). It has been shown to affect a wide range of organ systems in many animal species and can induce a wide range of adverse biological responses. The binding of TCDD to the so-called aryl hydrocarbon (Ah) receptor in cells appears to be the first step in a series of events that manifest themselves in biological responses, including changes at the biochemical, cellular and tissue levels. While there is only limited data on the toxic effects of other dioxins and dioxin-like PCBs, it may be inferred that biochemical, cellular and tissue-level effects that are elicited by exposure to TCDD are also induced by other chemicals that have a similar structure and that bind to the Ah receptor.
The acute toxicity of TCDD and related compounds can vary widely between and among species. For example in guinea-pigs, an LD50 of 0.6 µg/kg bw was recorded after oral administration, as compared with an LD50 of > 5000 µg/kg bw in Syrian hamsters. Explanations for this variation include differences in the Ah receptor, such as size, transformation and binding to the dioxin response element, pharmacokinetics (metabolic capacity, tissue distribution), and body fat content (FAO/WHO, 2002).
There is evidence of toxicity in humans as a result of acute high-dose and repeated or long-term exposure to dioxins. The most widely recognised and consistently observed effect following high-dose exposure to TCDD is chloracne. The condition can disappear after termination of exposure or can persist for many years. Other effects on the skin include hyperpigmentation and hirsutism, while other effects of PCDD/Fs and dioxin-like PCBs include elevated levels of liver enzymes and other disturbances in liver function, increased death rate from non-malignant liver disease, changes in thyroid function, impaired immunological function, effects on the cardiovascular system, influences on reproductive hormones and reproductive outcomes, and in some children, neuro-developmental delays (EC-SCF, 2000; van Leeuwen and Younes, 2000; FAO/WHO, 2002). Of the range of non-cancer health effects evaluated in exposed adult populations, some appear to be transient and were not observed when exposure ceased, whereas other effects persist for some years.
There is agreement amongst expert groups in recent years that the toxicological findings in laboratory animals of most relevance to humans are effects on the immune system, and on reproduction and development (FAO/WHO, 2002). At doses lower than those at which such effects have been observed (body burdens of 3–10 ng/kg bw), TCDD has been found to elicit biochemical or functional actions (eg, inducing liver enzymes), although these are classified as early expression of events that may or may not result in adverse effects (van Leeuwen and Younes, 2000).
Immunotoxic effects of TCDD have been observed in several species at multiple targets in the immune system. The main target of immunotoxicity of TCDD is the thymus, where cellular depletion is observed, with consequent reduced production of T lymphocytes and depression of cell-mediated immunity. Effects of PCDDs and PCDFs on the thyroid appear to be mediated through hormone metabolism, while dioxin-like PCBs may have a direct effect on the thyroid (FAO/WHO, 2002). Cell-mediated and humoral immune responses are also suppressed following TCDD, suggesting the thymus is not the only target within the immune system (EC-SCF, 2000). The severity of TCDD-induced immunotoxic effects varies among species and depends largely on the endpoint investigated (EC-SCF, 2000).
TCDD induces a distinct series of developmental effects, including foetal mortality, structural malformations and postnatal functional alterations, in a variety of species at doses below those associated with maternal toxicity. TCDD can induce significant embryo lethality (early or late resorptions, abortions, stillbirths), which is usually associated with indications of maternal toxicity. The timing of dosing and the age of the embryo or foetus have been shown to be major determinants of TCDD-induced prenatal mortality. Developmental effects that occur at doses not associated with maternal toxicity include induction of cleft palate and hydronephrosis, with the developing urogenital system of rodents, especially in males, being particularly sensitive to perturbation by TCDD and dioxin-like compounds. Effects include reductions in prostate growth and development (rats and mice), decreased testicular and epididymal sperm numbers (rats, mice, hamsters), and decreased numbers of ejaculated sperm (rats and hamsters) (FAO/WHO, 2002).
Endometrial effects in rhesus monkeys arising from exposure to dioxins had been reported (EC‑SCF, 2000 citing Rier et al, 1993), but follow-up studies, which included analysis of serum concentrations of dioxins and other chlorinated compounds, have given rise to uncertainties about the relationship between exposure to TCDD and endometriosis (EC-SCF, 2001).
Experimental studies demonstrate that TCDD is carcinogenic in all species and strains of laboratory animals tested, with cancers occurring at many sites. It has been characterised as a multi-site carcinogen. Several short-term assays for genotoxicity with TCDD covering various endpoints gave primarily negative results. Furthermore, TCDD did not bind covalently to mouse liver DNA. This data indicates that TCDD is not an initiator of carcinogenesis. Several studies have shown that TCDD is a potent tumour promoter. Several modes of action have been hypothesised, including increased expression of genes involved in cell growth and differentiation through binding of TCDD to the Ah receptor, induction of specific cytochrome P-450 (CYP1A1 and CYP1A2) resulting in oxidative stress, and inhibition of apoptosis (EC‑SCF, 2000).
In epidemiological studies, the strongest evidence for carcinogenicity of TCDD is associated with an increased risk for all cancers, rather than any specific site (IARC, 1997). Specific cancers associated with dioxin exposure are lung cancer, non-Hodgkin’s lymphoma and soft-tissue sarcoma (IARC, 1997). In a 2006 US National Academy of Sciences review of the health effects of Agent Orange, the committee found sufficient evidence of an association with herbicides (2,4-D, 2,4,5-T, picloram and cacodyllic acid) and/or TCDD for four cancers: soft tissue sarcoma, non-Hodgkin’s lymphoma, Hodgkin’s disease and chronic lymphocytic leukaemia; and limited evidence of an association with laryngeal cancer; cancer of the lung, bronchus, or trachea; prostatic cancer; and multiple myeloma (IOM, 2007). In contrast, other authors have attributed the occurrence of non-Hodgkin’s lymphoma, soft-tissue sarcoma and multiple myeloma to exposure to pentachlorophenol and not dioxin contamination (Cooper and Jones, 2008). Further, some authors argue that the evidence for human carcinogenicity is debatable and that TCDD will eventually be recognised as not carcinogenic to humans (Coles et al, 2003).
The biochemical and toxicological effects of PCDDs, PCDFs and coplanar PCBs are directly related to their concentrations in tissues, and not to the daily dose. The body burden, which is strongly correlated with the concentrations in tissue and serum, integrates the differences in half-lives between species. The half-life of TCDD varies considerably between species, with half-lives in mice, rats, and monkeys reported to be 12, 20, and 400 days, respectively; and a representative half-life in humans being 7.5–7.6 years, although a range of 3–16 years has been reported. Thus, rodents require appreciably higher daily doses (100–200-fold) to achieve a body burden at steady state that is equivalent to that recorded in humans exposed to background concentrations. Toxicokinetically, estimates of body burden are considered more appropriate measures of dose for interspecies comparisons than the daily dose.
Equivalence factors
Dioxins and dioxin-like compounds have a common mode of action, notably mediation of toxic effect through binding to the Ah receptor. This enables estimation of the cumulative risk of exposure to dioxins through expression of the toxicity of individual congeners relative to that of 2,3,7,8-TCDD, the most toxic congener. The relative toxicity is expressed as toxic equivalence factors (TEFs), estimated from the weaker toxicity of the respective congener in relation to the most toxic congener 2,3,7,8-TCDD, which is assigned the arbitrary TEF of 1. By multiplying the analytically determined amounts of each congener by the corresponding TEF and summing the contribution from each congener, the total toxic equivalent (TEQ) value of a sample can be obtained using the following equation:
TEQ = (PCDDi × TEFi) + (PCDFi × TEFi) + (PCBi × TEFi).
Several different TEF schemes have been proposed: the International TEFs (I-TEFs) (NATO/CCMS, 1988 cited in EC-SCF 2000), which provided TEFs for PCDDs and PCDFs, and Ahlborg et al (1994 cited in EC-SCF 2000) for dioxin-like PCBs; the 1998 WHO-TEFs, which were the consensus from an international meeting in 1997 for human, fish and wildlife risk assessment (van den Berg et al, 1998); and most recently evaluated, the 2005 WHO-TEFs, which considered additional data available since the previous evaluation (van den Berg et al, 2006). The primary difference in the 1998 and 2005 WHO evaluations was the use of half order-of-magnitude increments on a logarithmic scale to estimate TEFs. The use of different TEFs (Table 72) will give rise to different TEQ (toxic equivalent) values from the same analytical raw data. These differences have to be taken into account when results calculated with different TEF models are compared. In food and human samples, dioxin TEQ values based on WHO-TEFs (van den Berg et al, 1998) are approximately 10–20% higher than those obtained by using the I-TEFs (EC-SCF, 2000), and the change in dioxin TEQ values based on 2005 WHO-TEFs (WHO, 2005) being approximately 10–20% lower than those obtained by using the 1998 WHO-TEFs (van den Berg et al, 2006), ie, similar to the TEQs calculated using I‑TEFs. Based on dioxin contamination at 13 sawmill sites in New Zealand, TEQ values based on I-TEFs were on average 20% higher (range: 5–60% higher) than those obtained by using TEQ values based on 2005 WHO-TEFs, (pers. comm., A. Bingham, JCL Air and Environment).
Table 72: Comparison of TEFs for dioxins established at various times1
| Compound | Abbreviation |
I-TEF (1988/1994)2 |
WHO (1998) |
WHO (2005) |
|---|---|---|---|---|
Polychlorinated dibenzodioxins |
|
|
|
|
2,3,7,8-Tetrachlorodibenzodioxin |
TCDD |
1 |
1 |
1 |
1,2,3,7,8-Pentachlorodibenzodioxin |
1,2,3,7,8-PeCDD |
0.5 |
1 |
1 |
1,2,3,4,7,8-Hexachlorodibenzodioxin |
1,2,3,4,7,8-HxCDD |
0.1 |
0.1 |
0.1 |
1,2,3,6,7,8-Hexachlorodibenzodioxin |
1,2,3,6,7,8-HxCDD |
0.1 |
0.1 |
0.1 |
1,2,3,6,7,9-Hexachlorodibenzodioxin |
1,2,3,6,7,9-HxCDD |
0.1 |
0.1 |
0.1 |
1,2,3,4,6,7,8-Heptachlorodibenzodioxin |
1,2,3,4,6,7,8-HpCDD |
0.01 |
0.01 |
0.01 |
Octachlorodibenzodioxin |
OCDD |
0.001 |
0.0001 |
0.0003 |
Polychlorinated dibenzofurans |
|
|
|
|
2,3,7,8-Tetrachlorodibenzofuran |
2,3,7,8-TCDF |
0.1 |
0.1 |
0.1 |
1,2,3,7,8-Pentachlorodibenzofuran |
1,2,3,7,8-PeCDF |
0.05 |
0.05 |
0.03 |
2,3,4,7,8-Pentachlorodibenzofuran |
2,3,4,7,8-PeCDF |
0.5 |
0.5 |
0.3 |
1,2,3,4,7,8-Hexachlorodibenzofuran |
1,2,3,4,7,8-HxCDF |
0.1 |
0.1 |
0.1 |
1,2,3,6,7,8-Hexachlorodibenzofuran |
1,2,3,6,7,8-HxCDF |
0.1 |
0.1 |
0.1 |
1,2,3,7,8,9-Hexachlorodibenzofuran |
1,2,3,7,8,9-HxCDF |
0.1 |
0.1 |
0.1 |
2,3,4,6,7,8-Hexachlorodibenzofuran |
2,3,4,6,7,8-HxCDF |
0.1 |
0.1 |
0.1 |
1,2,3,4,6,7,8-Heptachlorodibenzofuran |
1,2,3,4,6,7,8-HpCDF |
0.01 |
0.01 |
0.01 |
1,2,3,4,7,8,9-Heptachlorodibenzofuran |
1,2,3,4,7,8,9-HpCDF |
0.01 |
0.01 |
0.01 |
Octochlorodibenzofuran |
OCDF |
0.0001 |
0.0001 |
0.0003 |
“Non-ortho” polychlorinated biphenyls |
|
|
|
|
3´,4,4´-Tetrachlorobiphenyl (PCB 77) |
3,3´,4,4´-TCB |
0.0005 |
0.0001 |
0.0001 |
3,4,4´,5,-Tetrachlorobiphenyl (PCB 81) |
3,4,4´,5-TCB |
– |
0.0001 |
0.0003 |
3,3´,4,4´,5-Pentachlorobiphenyl (PCB 126) |
3,3´,4,4´,5-PeCB |
0.1 |
0.1 |
0.1 |
3,3´,4,4´,5,5´-Hexachlorobiphenyl (PCB 169) |
3,3´,4,4´,5,5´-HxCB |
0.01 |
0.01 |
0.03 |
“Mono-ortho” polychlorinated biphenyls |
|
|
|
|
2,3,3´,4,4´-Pentachlorobiphenyl (PCB 105) |
2,3,3´,4,4´-PeCB |
0.0001 |
0.0001 |
0.0003 |
2,3,4,4´,5-Pentachlorobiphenyl (PCB 114) |
2,3,4,4´,5-PeCB |
0.0005 |
0.0005 |
0.0003 |
2,3´,4,4´,5-Pentachlorobiphenyl (PCB 118) |
2,3´,4,4´,5-PeCB |
0.0001 |
0.0001 |
0.0003 |
2,3´,4,4´,5’-Pentachlorobiphenyl (PCB 123) |
2,3´,4,4´,5´-PeCB |
0.0001 |
0.0001 |
0.0003 |
2,3,3´,4,4´,5-Hexachlorobiphenyl (PCB 156) |
2,3,3´,4,4´,5-HxCB |
0.0005 |
0.0005 |
0.0003 |
2,3,3´,4,4´,5´-Hexachlorobiphenyl (PCB 157) |
2,3,3´,4,4´,5´-HxCB |
0.0005 |
0.0005 |
0.0003 |
2,3´,4,4´,5,5´-Hexachlorobiphenyl (PCB 167) |
2,3´,4,4´,5,5´-HxCB |
0.00001 |
0.00001 |
0.0003 |
2,3,3´,4,4´,5,5´-Heptachlorobiphenyl (PCB 189) |
2,3,3´,4,4´,5,5´-HpCB |
0.00001 |
0.00001 |
0.0003 |
1 Bolding indicates which values have changed from the previous reassessment.
2 TEFs from NATO/CCMS (1988 cited in EC-SCF, 2000) for PCDDs, PCDFs, and Ahlborg et al (1994 cited in EC‑SCF, 2000) for PCBs.
2.14.2 New Zealand classification
ERMA NZ has not classified dioxins or furans, as these substances are not deliberately manufactured, and have no known technical use (other than in laboratory standards). They may be present as contaminants in other substances. When this is the case, the concentration of these contaminants and the contribution they make to the hazard of the substance would be taken into account in the approval of the main component in which they are a contaminant.
2.14.3 Reference health standards
Ingestion
Tolerable intakes of dioxins and dioxin-like PCBs have been extensively evaluated over the last 10 years (WHO, 1998; EC-SCF, 2000; 2001; FSA, 2001; FAO/WHO, 2002; US EPA, 2003). With the exception of the US EPA, all agencies have derived tolerable intakes (variably expressed as daily, weekly or monthly intakes) (Table 73) while the US EPA treated cancer as a non-threshold effect and has used benchmark dose modelling to determine cancer potency (US EPA, 2003) (Table 74).
The New Zealand Timber Treatment Guidelines (MfE and MoH, 1997) use a maximum allowable intake of 10 pg/kg bw/day (TEQ) based on the Pentachlorophenol Risk Assessment Pilot Study (NTG, 1992 cited in MfE and MoH 1997) for deriving interim soil guideline values for dioxins. This value was based on the conventional approach of applying uncertainty factors to a relevant NOAEL, which is the approach used by CCME (2000) in their latest evaluation of the toxicity of dioxins, to derive a TDI of 10 pg/kg bw based on reproductive effects in rats. Both of these sources use daily intake as the dose metric, as does the ATSDR (1998) in its derivation of a chronic-duration MRL of 1 pg/kg bw/day (TEQ), based on behavioural study in monkeys. It is notable that EC-SCF (2000) considered that the results of the study used by ATSDR were of doubtful significance for humans.
In contrast, most other agencies have adopted the approach that seems to arise out of the 1997 WHO consultation on dioxins, in which body burden was considered to be the most relevant dose metric (van Leeuwen and Younes, 2000). The rationale for changing to body burden was that, from a pharmaco-kinetic point of view, this was more appropriate for interspecies comparisons given the long half-lives of dioxins in humans and the difference in half-lives between humans and animals (EC-SCF, 2000; van Leeuwen and Younes, 2000). In the 1997 WHO evaluation, an additional uncertainty factor of 10 was applied to the body burdens of animals giving rise to the most sensitive adverse effects, to account for potential differences in susceptibility within the human population, the comparative susceptibility of humans and animals, and the variation in the half-lives of individual components of the dioxin-mix (van Leeuwen and Younes, 2000).
Furthermore, in the evaluation by WHO in 1997 (van Leeuwen and Younes, 2000) and subsequently by EC-SCF (2000), it was considered that there was no scientific basis for selecting any one particular study or effect; thus a range of tolerable intakes were derived. WHO established a TDI range of 1–4 pg TEQ/kg bw, while EC-SCF (2000) considered it more appropriate to express the tolerable intakes on a weekly basis, given the long half-lives in humans, and selected the lower end of the range and established a temporary tolerable weekly intake of 7 pg/kg bw.
In contrast to the 1997 WHO evaluation, EC-SCF (2000) used pharmaco-kinetic principles to convert animal body burdens into equivalent human daily intakes that on a chronic basis would lead to similar body burdens in humans, given by:
Body burden at steady state (ng/kg body weight) = f X intake (ng/kg bw/day) X half-life (days) / ln(2)
where f is the fraction of the dose absorbed, assumed to be 50% from food for humans, and the half-life of TCDD is 2750 (7.5 years: EC-SCF, 2000). This approach was also used by JECFA in their evaluation in 2001, although they used a half-life for TCDD of 2776 (7.6 years, FAO/WHO, 2002). Similarly, the US EPA used this approach in their draft reassessment although they used a half-life of 2593 (7.1 years, US EPA, 2003).
In 2000 data was published that allowed the calculation of the total amount of dioxin in the foetus associated with maternal exposure at steady state (Hurst et al, 2000b cited in EC-SCF, 2001). Based on this information, EC-SCF revised its earlier tolerable intake and established a tolerable weekly intake of 14 pg/kg bw, based on the midpoint of estimates derived from the lowest LOAEL and NOAEL for developmental effects in male rat offspring. JECFA also used this approach to derive a provisional tolerable monthly intake of 70 pg/kg bw, also based on the midpoint of (different) estimates derived from the same studies (FAO/WHO, 2002).
Based on the evaluations of the WHO’s European Centre for Environmental Health and International Programme on Chemical Safety, ECEH-IPCS (WHO, 1998), EC-SCF (2001), JECFA (FAO/WHO, 2002), and considering the IARC classification of dioxins as human carcinogens, the Australian NHMRC (2002) established a tolerable monthly intake of 70 pg/kg bw. DEFRA and EA (2002) similarly considered these evaluations in addition to an evaluation by the UK Committee on Toxicity of Chemicals in Food and the Environment (COT) (FSA, 2001), and recommended a tolerable daily intake of 2 pg/kg bw, based on the COT recommendations. In contrast to these agencies, the New Zealand Ministry of Health established an interim monthly maximum intake of 30 pg/kg, based on the lower end of the TDI recommended by WHO of 1–4 pg/kg bw (MoH, 2002). This value was recommended by the Organochlorines Technical Advisory Group, and was adopted by the Ministry as it also endorsed the precautionary approach recommended by Smith and Lopipero (2001) and further recognised the desirability of ongoing reduction in dioxin intake. It is unclear as to whether the EC-SCF (2001) and JECFA (FAO/WHO, 2002) evaluations were considered in this recommendation. This value was considered interim because further research on dioxins is being undertaken, and it was also noted that the margin between current exposures – even in New Zealand, which is low by international standards – and intakes that cause toxic effects in animals is undesirably small.
In contrast to all other agencies, the US EPA has consistently used cancer potency estimates as the primary basis for assessing the toxicity of dioxins and dioxin-like PCBs, with the difference in approach suggested to reflect differences in science policy (US EPA, 2003). The US EPA further considered that there was little point in setting an RfD because it would likely be below current background levels (US EPA, 2003). However, both the EPA’s Scientific Advisory Board and the National Academy of Sciences (NAS), in its review of EPA’s reassessment (NAS, 2006), recommended the derivation of an RfD. Furthermore, the NAS considered that the EPA reassessment needed substantial work, particularly the risk characterisation including the dose-response modelling for both cancer and non-cancer endpoints (NAS, 2006). Despite the NAS urging the EPA to finalise the reassessment quickly, to date it has not been finalised although a workshop was held in December 2008 to address the NAS concerns (US EPA, 2009).
Toxic equivalents
In all cases the tolerable intake or the potency estimates are considered to apply to the toxic equivalent concentration of dioxins and dioxin-like PCBs. In most cases, the WHO-TEFs (van den Berg et al, 1998) (see earlier) are generally the recommended TEFs, although this is most likely to be the case because the evaluations were undertaken before the WHO (2005) re‑evaluation.
Table 73: Summary of oral reference health standards for dioxins and furans (as TEQ) as a threshold contaminant, used by different international agencies
| Jurisdiction | Tolerable daily intake1 (pg/kg bw) |
Tolerable weekly intake1 (pg/kg bw) |
Tolerable monthly intake1 (pg/kg bw) |
Key study2 |
Critical effect2 |
Basis of value2 |
Reference |
|---|---|---|---|---|---|---|---|
New Zealand |
10 |
70 |
300 |
NTP (1991) |
Not stated |
NTP (1991) |
MfE and MoH (1997) |
New Zealand Ministry of Health |
1 |
7 |
30 |
Not stated |
Not stated |
Lower end of recommended TDI derived by FAO/WHO (1998) converted to a monthly intake |
MoH (2002) |
Joint FAO/WHO Expert Committee on Food Additives (JECFA) |
2.3 |
14 |
70 |
Faqi et al (1998); Ohsako et al (2001) |
Reproductive developmental effects in male offspring of treated females |
Midpoint of four estimates of provisional tolerable monthly intake (PTMI) determined by toxicokinetic modelling using a linear model and a power model to determine the maternal body burden after repeated dosing, which would result in the same body burden in foetuses as after administration of a bolus dose on day 15 of gestation at the LOAEL (Faqi et al, 1998) and NOAEL (Ohsako et al, 2001) reported in the studies, conversion to the equivalent human monthly intake, subtraction of the background body burden, and application of uncertainty factor of 3.2 to account for toxicokinetic differences between humans (both studies) and 3 for extrapolation from a marginal LOAEL to a NOAEL (Faqi et al, 1998 only) |
FAO/WHO (2002) |
European Food Safety Authority (EFSA) |
2 |
14 |
70 |
Faqi et al (1998) |
Reproductive developmental effects in male offspring of treated females |
LOAEL of 20 pg/kg bw (equivalent human dietary intake calculated using pharmaco-kinetics and the foetal body burden), and application of uncertainty factor of 3.2 to account for toxicokinetic differences between humans and 3 for extrapolation from a marginal LOAEL to a NOAEL |
EC-SCF (2001) |
Australia |
2.3 |
14 |
70 |
Evaluation of several previous evaluations |
Hormonal, reproductive and/or developmental effects |
WHO-ECEH/IPCS (1998) consultation, the EC-SCF (2001) opinion, and the JECFA evaluation (FAO/WHO, 2002) |
|
UK |
2 |
14 |
70 |
Evaluation of several previous evaluations |
Reproductive effects |
FSA (2001) |
DEFRA and EA (2002) |
The Netherlands – current |
10 |
70 |
300 |
Not stated |
Not stated |
Janssen et al (1995) |
Baars et al (2001) |
The Netherlands – proposed3 |
1 |
7 |
30 |
Not stated |
Not stated |
Lower end of recommended TDI derived by FAO/WHO (van den Berg et al, 1998) |
Baars et al (2001) |
Canada |
10 |
70 |
300 |
Murray et al (1979) |
Reproductive effects in rats |
NOAEL of 0.001 µg/kg bw/day and application of 100-fold uncertainty factor to account for inter- (10) and intra-species (10) variation |
CCME (2000) |
US ATSDR – chronic duration MRL – dioxins |
1 |
7 |
30 |
Schantz et al (1992) |
Altered behaviour in monkeys |
LOAEL of 1.2 x 10–4 µg/kg bw/day and application of uncertainty factors of 900 comprised of 3 for the use of a minimal LOAEL, 3 for inter-species variation and 10 for human variability |
ATSDR (1998) |
1 Bold values indicate the reference health standard adopted by the specific agency; the other values are shown for comparative purposes.
2 As reported in the reference cited in the reference column.
3 This value is yet to be officially adopted.
Table 74: Summary of oral reference health standards for dioxins and dioxin-like compounds as non-threshold contaminants, established by the US EPA
| Jurisdiction | Acceptable risk level1 |
Risk-specific dose (pg/kg bw/day) |
Cancer slope factor (per pg/kg bw/day) |
Key study2 |
Critical effects2 |
Basis of value2 |
Reference |
|---|---|---|---|---|---|---|---|
US EPA |
10–6 |
0.001 |
0.001 |
Becher et al (1998) |
All cancers |
Linear extrapolation of ED01 determined using benchmark dose modelling of data on occupationally exposed humans |
US EPA (2003) (draft) |
1 Where the acceptable risk level for a given jurisdiction is not 10–5, the risk-specific dose for a risk of 10–5 is shown in square brackets.
2 As reported in the reference cited in the reference column.
Inhalation
Inhalation will be a negligible route of exposure as dioxins and dioxin-like PCBs have limited volatility and the amount of dust considered to be inhaled typically represents a very small fraction of exposure (see section 1.1.3), so is not considered further.
Dermal absorption
The skin absorption factor is the only contaminant-specific parameter required for the dermal absorption pathway. Dermal absorption of individual dioxins and dioxin-like PCBs will be dependent on the physico-chemical property of the individual substance. Roy et al (2008) estimated that 1.9% of TCDD in a low-organic soil and 0.24% of TCDD in a high-organic soil would be absorbed by human skin. These estimates were based on in vivo (rat) and in vitro (rat and human) dermal absorption trials. Jackson et al (1993) examined in vivo dermal absorption of TCDD, OCDD and a number of furans in rats. These authors report dermal absorption of TCDD in a solvent carrier over 72 hours was 17%, and for OCDD was 4.3%, while absorption of PeDF and TCDF ranged from 25 to 45%. Assuming absorption is linear over the three days, this gives an average 24-hour absorption of 5.6% for TCDD, 1.6% for OCDD, and 8.3–15% for the furans. Given these estimates are for the compounds in a solvent carrier, they likely over-estimate absorption from soil. Further, rat skin is considered to be more permeable than human skin, and was 3–4 times more permeable when dermal absorption of TCDD was examined (Roy et al, 2008). Assuming this is applicable for all PCDDs and PCDFs, revised absorption values from the Jackson et al (1993) study are 1.9% for TCDD, 0.5% OCDD, and 2.8–5% for PeDF and TCDF.
Dermal uptake of PCBs is suggested to be greater than that of PCDDs and PCDFs, and is variable depending on the chlorine content and position (Garner et al, 2006). Roy et al (2009) estimated that 7.4% of PCB77 (3,3,4,4,-tetrachlorobiphenyl) in a low-organic soil would be absorbed by human skin and that 9.6% of PCB77 would be absorbed by rat skin in a high-organic soil – rat skin was estimated to be fourfold to ninefold more permeable than human skin. These estimates were based on in vivo (rat) and in vitro (rat and human) dermal absorption trials. Wester et al (1993) found that approximately 14% of two PCB mixtures (Aroclor 1242 and 1254) applied to soil (0.9% organic matter) was percutaneously absorbed by rhesus monkeys over a 24-hour period. However, they also indicate that a reduced amount (1.6% to 2.6%) was absorbed into human skin from soil over the same period, although minimal partitioning into human plasma occurred, ie, the PCBs remained bound to the skin. Mayes et al (2002) found that approximately 4% of Aroclor 1260 applied to soil (5–6% organic carbon) was percutaneously absorbed by rhesus monkeys over a 24-hour period.
For the PCDDs, it is recommended that an absorption factor of 0.02 is used – this is based on absorption of TCDD from soil as reported by Roy et al (2008), and provides a conservative estimate of the absorption of higher molecular weight PCDDs. Higher dermal absorption factors are recommended for the furans. In this case, 0.05 is recommended for TCDF based on the adjusted dermal absorption reported in Jackson et al (1993). This is also expected to be a conservative estimate as it is based on dermal absorption from a solvent carrier. For dioxin-like PCBs it is recommended that a dermal absorption factor of 0.07 is used – this is based on the absorption of PCB77 from soil as reported by Roy et al (2009), and provides a conservative estimate of the absorption of higher molecular weight PCBs. In comparison, US EPA (2004) recommends a dermal absorption factor of 0.03 for TCDD and other dioxins, and 0.14 for Aroclors 1242, 1254 and other PCBs.
Other routes of exposure – background exposure
The major route of exposure of humans to dioxins is estimated to be through the diet, primarily meat products (EC-SCF, 2000). There is limited data on the background exposure of New Zealanders to dioxins and dioxin-like PCBs, with Buckland et al(1998) providing the most extensive data. These authors determined the intake of dioxins from simulated diets for an adult male (25–44 years) consuming median energy (10.8 MJ per day) and adolescent males (15–18 years with a high energy intake (21.5 MJ/day) (Table 75). Intakes based on assuming substances not detected were present at half the limit of detection (½ LOD) – which is consistent with the Total Diet Surveys (Vannoort and Thomson, 2005) – are markedly higher than intake based on excluding substances not detected (excluding LOD). However, for consistency with other contaminants the ½ LOD results are used. It should also be noted that the simulated diets used in Buckland et al (1998) are slightly different to those used in the Total Diet Surveys (Vannoort and Thomson, 2005), which have been used to estimate dietary intakes of other contaminants considered in this review. The intakes of dioxins and dioxin-like PCBs were estimated using the I-TEFs, which appear to give dioxin TEQ values approximately 20% higher than those determined using the WHO (2005) TEFs (see earlier) although Smith and Lopipero (2001) recalculated the dietary data using the WHO (van den Berg et al, 1998) TEFs, and obtained a mean dietary intake for an adult male of 0.37 pg/kg bw/day (½ LOD), which is slightly higher than that calculated using the I-TEFs (Table 75). Ideally the intake values would be recalculated using WHO (2005) TEFs.
Smith and Lopipero (2001) also calculated an estimated average lifetime daily exposure (ALDE) of 1.4 pg/kg bw/day for New Zealanders aged 15 and over, based on serum fat concentrations and assuming a half-life of 7.5 years and that the fraction of the dose adsorbed is 90%. The ALDE estimates are higher than estimated dietary intakes as they includes intakes from dietary and non-dietary pathways, and contributions from historical and current exposures, while the dietary intake estimate represents exposure at a single point in time.
Inhalation of dioxins is expected to be a negligible route of exposure. Data from a report on organochlorine concentrations in air indicates that the mean concentrations of PCDD and PCDFs (expressed as I-TEQs) range from 28.1 to 83.9 fg I-TEQ/m3 in urban areas in New Zealand (Buckland et al, 1999). Assuming a 15-kg child inhales 6.8 m3/day and a 70-kg adult 13.3 m3/day (Proffitt and Cavanagh, 2008) of air at the maximum mean dioxin concentration, this gives rise to intakes of 0.038 and 0.015 pg/kg bw/day for a child and adult, respectively.
For the purposes of this work, a dietary intake based on an adult male is used as the best estimate of background exposure, as there are uncertainties in the calculation of ALDE (assumed half-life of TCDD, % absorption), dietary intake is estimated to be the most significant route of exposure, and finally this is most consistent with the approach adopted for other contaminants.
Table 75: Summary of dietary intakes of dioxins and dioxin-like PCBs for an adult male and adolescent male
|
Intake of dioxins (I-TEQ pg/kg bw/day) |
Intake of dioxins and dioxin-like PCBs (I‑TEQ pg/kg bw/day) |
||
|---|---|---|---|---|
Including ½ LOD |
Excluding LOD |
Including ½ LOD |
Excluding LOD |
|
Adult male (80 kg) |
0.18 |
0.047 |
0.33 |
0.15 |
Adolescent male (70 kg) |
0.44 |
0.14 |
0.76 |
0.34 |
2.14.4 Summary of effects
Notwithstanding a common mechanism of action as outlined earlier, it is noted that there are considerable species and strain differences in the acute toxicity of dioxins. Adverse effects reported in animals following exposure to dioxins include immunotoxicity, developmental and behavioural effects in offspring of treated rhesus monkeys, and developmental effects in rats. A summary of effects is provided in Table 76. Many of the original toxicity studies report only the doses used, while the WHO (FAO/WHO, 2002) and EC-SCF (2000; 2001) evaluations estimated the body burden associated with toxic effects reported in the original papers. Body burden is considered to be the most appropriate dose metric to assess the health effects of dioxins. Further, all studies were undertaken using TCDD, the most toxic congener, thus Table 76 provides a summary of the health effects associated with TCDD (unless otherwise stated).
Table 76: Summary of the health effects of TCDD
| Dose (ng/kg)1 | Type of poisoning |
Body burden2 (ng/kg bw) |
Effects |
|---|---|---|---|
10 ng/kg/day |
Chronic |
294 (EC-SCF, 2000) |
LOAEL for liver tumour formation in rats |
100 |
Single dose at gestation day 15 |
60 (EC-SCF, 2000) |
LOAEL for delayed hypersensitivity suppression in male offspring from exposed female rats |
0.15 ng/kg/day |
Chronic |
25–37 (EC-SCF, 2000) |
Subtle, non-persistent neurobehavioural effects in offspring of exposed female monkeys |
5 ng/kg/week |
Maintenance of 25 ng/kg bw |
25 (FAO/WHO, 2002) |
Lowest LOAEL developmental effects of reproductive system of male offspring (decreased sperm production) from exposed female rats |
12.5 |
Single dose at gestation day 15 |
13 (FAO/WHO, 2002) |
NOAEL developmental effects of reproductive system of male offspring (decreased anogenital distance) from exposed female rats |
1 Unless otherwise stated.
2 As shown in either FAO/WHO (2002), EC-SCF (2000), or EC-SCF (2001), calculated as the maternal body burden for developmental effects and at gestational day 15 (rats) or maternal body burden at delivery after 16.2 and 36.3 months of exposure (monkeys); or the body burden of the exposed animal (cancer studies).
2.14.5 Weight of evidence
- 2,3,7,8-TCDD is considered a known human carcinogen (Group 1) by the IARC (1997) and a human carcinogen (Class A) by the US EPA (2003), based on soft-tissue cancers in humans, although there is insufficient data to assess the carcinogenicity of all other congeners. Based on animal studies TCDD is considered a multi-site carcinogen and acts as a tumour promoter.
- TCDD is not genotoxic, and mechanistic data suggests a threshold interpretation of TCDD-induced carcinogenicity (EC-SCF, 2001; FAO/WHO, 2002).
- The toxicity of dioxins is mediated through binding to the Ah receptor (EC-SCF, 2000; FAO/WHO, 2002).
- Developmental effects on the reproductive system in male offspring of exposed pregnant females is considered to be the most sensitive toxicity endpoint and is also considered to be protective against carcinogenic effects of dioxins (EC-SCF, 2001; FSA, 2001; FAO/WHO, 2002).
- Body burden is considered to be the most suitable dose measure (EC-SCF, 2001; FAO/WHO, 2002; US EPA, 2003).
2.14.6 Recommendations for toxicological intake values
There is general agreement between the various expert committees that have reviewed dioxins that tolerable intakes are appropriate for use for dioxins and dioxin-like PCBs. The monthly intake value generally adopted is 70 pg/kg bw (also variously expressed as daily (2 pg/kg bw) or weekly (14 pg/kg bw) intakes). Given the long half-lives of dioxins, and thus the likely lack of effect of small excursions of a daily or even weekly intake, it is recommended that a monthly intake toxic-equivalent dose (TEQ) is used.
The Ministry of Health has confirmed it will retain its policy on a maximum monthly intake value of 30 pg/kg bw and therefore this value is recommended (Table 77). This value is based on the lower end of the range of tolerable intakes determined by WHO in 1998 (van den Berg et al, 2000), adopts a precautionary approach, and recognises the desirability of ongoing reduction of dioxin intake.
Table 77: Recommended toxicological criteria for dioxins
| Parameter | Value |
Basis |
Contaminant status |
Threshold |
See weight of evidence |
Oral intake dose (pg TEQ/kg bw/month) |
30 |
MoH (2002) |
Inhalation intake |
NA |
Low volatility of dioxins and dioxin-like PCBs suggests that inhalation exposure is negligible |
Skin absorption factor |
0.02 (PCDDs) |
Roy et al (2008) |
Background exposure (pg I-TEQ/kg bw/month) |
10.0 |
Daily dietary intake of an adult male determined in Buckland et al (1998) and extrapolated to a month. This value is considered applicable to children in the absence of any other data |
NA – not applicable.
Further it is recommended that WHO (2005) TEFs are used to calculate TEQs (Table 78), as these are based on the latest re-evaluation by WHO, and thus are likely to become the international standard.
Table 78: Recommended TEFs for dioxins and dioxin-like PCBs
| Compound | Abbreviation |
WHO (2005) |
|---|---|---|
Polychlorinated dibenzodioxins |
|
|
2,3,7,8-Tetrachlorodibenzodioxin |
TCDD |
1 |
1,2,3,7,8-Pentachlorodibenzodioxin |
1,2,3,7,8-PeCDD |
1 |
1,2,3,4,7,8-Hexachlorodibenzodioxin |
1,2,3,4,7,8-HxCDD |
0.1 |
1,2,3,6,7,8-Hexachlorodibenzodioxin |
1,2,3,6,7,8-HxCDD |
0.1 |
1,2,3,6,7,9-Hexachlorodibenzodioxin |
1,2,3,6,7,9-HxCDD |
0.1 |
1,2,3,4,6,7,8-Heptachlorodibenzodioxin |
1,2,3,4,6,7,8-HpCDD |
0.01 |
Octachlorodibenzodioxin |
OCDD |
0.0003 |
Polychlorinated dibenzofurans |
|
|
2,3,7,8-Tetrachlorodibenzofuran |
2,3,7,8-TCDF |
0.1 |
1,2,3,7,8-Pentachlorodibenzofuran |
1,2,3,7,8-PeCDF |
0.03 |
2,3,4,7,8-Pentachlorodibenzofuran |
2,3,4,7,8-PeCDF |
0.3 |
1,2,3,4,7,8-Hexachlorodibenzofuran |
1,2,3,4,7,8-HxCDF |
0.1 |
1,2,3,6,7,8-Hexachlorodibenzofuran |
1,2,3,6,7,8-HxCDF |
0.1 |
1,2,3,7,8,9-Hexachlorodibenzofuran |
1,2,3,7,8,9-HxCDF |
0.1 |
2,3,4,6,7,8-Hexachlorodibenzofuran |
2,3,4,6,7,8-HxCDF |
0.1 |
1,2,3,4,6,7,8-Heptachlorodibenzofuran |
1,2,3,4,6,7,8-HpCDF |
0.01 |
1,2,3,4,7,8,9-Heptachlorodibenzofuran |
1,2,3,4,7,8,9-HpCDF |
0.01 |
Octochlorodibenzofuran |
OCDF |
0.0003 |
“Non-ortho” Polychlorinated biphenyls |
|
|
3´,4,4´-Tetrachlorobiphenyl (PCB 77) |
3,3´,4,4´-TCB |
0.0001 |
3,4,4´,5,-Tetrachlorobiphenyl (PCB 81) |
3,4,4´,5-TCB |
0.0003 |
3,3´,4,4´,5-Pentachlorobiphenyl (PCB 126) |
3,3´,4,4´,5-PeCB |
0.1 |
3,3´,4,4´,5,5´-Hexachlorobiphenyl (PCB 169) |
3,3´,4,4´,5,5´-HxCB |
0.03 |
“Mono-ortho” polychlorinated biphenyls |
|
|
2,3,3´,4,4´-Pentachlorobiphenyl (PCB 105) |
2,3,3´,4,4´-PeCB |
0.0003 |
2,3,4,4´,5-Pentachlorobiphenyl (PCB 114) |
2,3,4,4´,5-PeCB |
0.0003 |
2,3´,4,4´,5-Pentachlorobiphenyl (PCB 118) |
2,3´,4,4´,5-PeCB |
0.0003 |
2,3´,4,4´,5’-Pentachlorobiphenyl (PCB 123) |
2,3´,4,4´,5´-PeCB |
0.0003 |
2,3,3´,4,4´,5-Hexachlorobiphenyl (PCB 156) |
2,3,3´,4,4´,5-HxCB |
0.0003 |
2,3,3´,4,4´,5´-Hexachlorobiphenyl (PCB 157) |
2,3,3´,4,4´,5´-HxCB |
0.0003 |
2,3´,4,4´,5,5´-Hexachlorobiphenyl (PCB 167) |
2,3´,4,4´,5,5´-HxCB |
0.0003 |
2,3,3´,4,4´,5,5´-Heptachlorobiphenyl (PCB 189) |
2,3,3´,4,4´,5,5´-HpCB |
0.0003 |
Inhalation exposure to dioxins and dioxin-like PCBs is likely to be negligible on contaminated sites due to their low volatility. Dermal absorption of these compounds is dependent on the physico-chemical properties of the individual congeners. It is recommended that a dermal factor of 0.02 is used as a conservative estimate of dermal absorption of PCDDs, based on dermal absorption of TCDD from soil (Roy et al, 2008). A higher absorption factor of 0.05 is recommended for PCDFs, based on adjustment of dermal absorption estimates from Jackson et al, 1993). For dioxin-like PCBs it is recommended that a dermal absorption factor of 0.07 is used as a is used as a conservative estimate of dermal absorption of PCBs, based on dermal absorption of a tetrachloropbiphenyl from soil Roy et al (2009).
Dietary intake is the primary source of background exposure to dioxins and dioxin-like PCBs and was estimated to be 0.33 pg I-TEQ/kg bw/day or 10.0 pg I-TEQ/kg bw/month for an adult, and is extended to children. Ideally, these intakes should be expressed on the basis of 2005 WHO-TEQs, although it is anticipated it will only make a marginal difference in the estimated intakes.
