Archived publication
This publication is no longer current or has been superseded.
2.1 Arsenic (As)
Observations of the toxicological effects of arsenic primarily come from epidemiological investigations of human populations exposed to arsenic in drinking water. Several comprehensive reviews of the toxicity of arsenic have been undertaken (NRC, 1999; 2001; EA, 2009; FPTCDW, 2006; ATSDR, 2007; Fowler et al, 2007). 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.
2.1.1 Toxicological status
Arsenic exists in organic and inorganic forms. Inorganic forms, particularly trivalent forms (arsenites), are considered to be the most toxic. However, it has recently been suggested that methylated metabolites may be more toxic than previously thought (NRC, 1999; 2001; Rossman, 2003) with dimethylarsenous acid (DMAIII) demonstrating greater toxicity than arsenite in some bioassays.
Arsenic can cause cancerous and non-cancerous effects. A large number of the latter are associated with exposure to arsenic and include dermal lesions, pigmentation, keratoses, peripheral vascular disease (eg, blackfoot disease), and cardiovascular effects (US EPA, 2001; WHO, 2001; ATSDR, 2007). Skin lesions are considered to be a sensitive indicator of chronic arsenic exposure.
Arsenic is classified as a known human carcinogen by the International Agency for Research (IARC, 1987; 2004 – Group 1) and by the US EPA (1993 – Group A). Skin cancer is a well-documented feature in human populations exposed to arsenic via drinking water with naturally high concentrations of arsenic (WHO, 2001). More recently the US National Research Council (NRC) concluded that there was sufficient evidence from epidemiological studies that chronic ingestion of inorganic arsenic causes bladder and lung cancer, in addition to skin cancer (NRC, 1999; 2001). The increase in cancer risk observed in epidemiological studies is primarily attributed to the presence of arsenite. Any arsenate that is present is rapidly reduced to arsenite once it enters the cells (Rossman, 2003). In humans, arsenic compounds are metabolised by methylation, primarily in the liver, and it has recently been suggested that these metabolites may play a role in bladder and perhaps some other cancers (Rossman, 2003). Inhalation of arsenic primarily results in tumours in the lung (WHO, 2001).
While carcinogenicity is considered to be the primary toxicological effect of concern in humans, in animal studies carcinogenicity of arsenic is often not found. There is considerable debate over the mechanism of carcinogenicity of arsenic (eg, Basu et al, 2001; Hughes, 2002; Rossman, 2003) and, therefore, whether it should be treated as a threshold or non-threshold contaminant in the context of deriving soil guideline values. NRC (1999) concluded that the most plausible explanation for the mode of action of arsenic carcinogenesis is that it induces chromosomal abnormalities without interacting directly with the DNA. Such indirect effects are typically considered to give rise to sublinear dose responses, ie, act as a threshold contaminant. Baars et al (2001) also considered the general consensus to be that the carcinogenic action of arsenic is based on non-genotoxic mechanisms, and arsenic should, therefore, be considered a threshold contaminant. This conclusion is supported by other studies on epidemiological data relating arsenic ingestion to skin and internal cancers (eg, Rudel et al, 1996) and studies that indicate arsenic is active late in the carcinogenic process, ie, acts as a cancer promoter (Basu et al, 2001). Similarly, Mead (2005) states: “there is general agreement that arsenic does not interact directly with DNA, and that its toxic effects occur through indirect alteration of gene expression”. An emerging consensus is also that arsenic is not an initiator of cancer, but rather works with other factors (eg, smoking, UV-radiation) to promote cancer (NRC, 1999; Mead, 2005).
2.1.2 New Zealand classification
The Hazardous Substances and New Organisms Act 1996 (HSNO) classification of arsenic set by the Environmental Risk Management Authority New Zealand (ERMA NZ) is shown in Table 1, but only has meaning when expressed in respect to a particular form of the element, ie, arsenic metal, whereas the form present in the environment as a soil contaminant is most likely to be arsenic oxide or a salt such as sodium arsenate/arsenite. Overall, arsenic is of relatively high toxicity (6.1B oral classification) with the following long-term endpoints: it is mutagenic (6.6B), a proven human carcinogen (6.7A), and is highly toxic from chronic exposures (6.9A). These endpoints are believed to be the most relevant findings concerning most chemical forms of arsenic likely to be encountered.
Table 1: HSNO classification of arsenic metal
| Hazardous property | HSNO classification |
|---|---|
Acute toxicity |
6.1B |
Skin irritation |
ND |
Eye irritation |
ND |
Sensitiser |
– |
Mutagenicity |
6.6B |
Carcinogenicity |
6.7A |
Reproductive/developmental toxicity |
– |
Target organ systemic toxicity |
6.9A |
ND – no classification due to no data/insufficient data/inconclusive data; – not assigned.
2.1.3 Reference health standards
Numerous studies have reported the effects of chronic exposure to arsenic in populations in regions with elevated concentrations of arsenic in drinking water. WHO (2001) cites the following places and references: Cordoba, Argentina (Arguello et al, 1938; Bergoglio, 1964), Antofagasta, Chile (Borgono et al, 1977; Zaldivar and Guiller 1977; Zaldivar et al, 1981), Mexico (Cebrian et al, 1983), and south-western Taiwan (Tseng et al, 1968; Tseng, 1977). These epidemiological studies form the basis for the development of reference health standards by different agencies.
Normally, most arsenic in drinking water is present as the two inorganic species arsenate and arsenite, with only low proportions of various organoarsenic compounds. Of the two inorganic forms, arsenate usually dominates in oxidised (including chlorinated) waters, but the more toxic arsenite form can become dominant in some circumstances as a result of chemical reduction reactions that occur in the environment. For exposure studies of arsenic in drinking water, the total arsenic dose therefore represents the mix of arsenic species that a given population was exposed to in drinking water.
The most comprehensive dataset is that from south-western Taiwan, initially reported by Tseng et al (1968; 1977 cited in WHO, 2001) who focused on skin cancer and skin disease. As cited in NRC (1999), Chen et al (1985; 1986; 1992) and Wu et al (1989) subsequently reworked these data to examine the risk of internal cancers. This study has been recommended for risk quantification for several reasons, including the large stable population (>40,000 people) that had lifetime exposures to arsenic, pathology data collection that was “unusually thorough”, and populations that were reasonably homogenous with respect to lifestyle (NRC, 2001; US EPA, 2001). Nonetheless, there are recognised weaknesses, including the use of median exposure data at the village level, the low income and relatively poor diet of the study population, and high exposures to arsenic via food. Further, NRC (2001) cites two recent studies (Ferreccio et al, 2000; Chiou et al, 2001) that are also of sufficient quality to warrant consideration in quantitative risk assessment for arsenic in drinking water.
Ingestion
Numerous agencies have derived reference health standards for the ingestion of arsenic, but have variously considered arsenic to be either a threshold or non-threshold contaminant. A summary of these values, and the bases for their derivation, are shown in Tables 2 and 3. To make it easier to compare the non-threshold values, the risk-specific dose at the acceptable excess risk level adopted in the given jurisdiction is given first, followed by that at an acceptable excess risk level of 10–5 (shown in brackets).
The majority of agencies that derived threshold values for arsenic based their reference health standards on the JECFA (FAO/WHO, 1988) value (Table 2). It should be noted that no quantitative assessment was undertaken, and that the JECFA TDI is based on a lowest observable adverse effects level (LOAEL): 25 out of 86 people had symptoms possibly associated with arsenic poisoning. In contrast, the US EPA (1993) quantitatively evaluated five epidemiological studies, including some of those evaluated by JECFA, to derive their reference dose. The US EPA (1993) ultimately considered that studies on the south-western Taiwanese population (WHO, 2001 citing Tseng et al, 1968; Tseng, 1977) were superior due to the provision of appropriate data, eg, exposure times, arsenic concentrations, and the large number of people studied. The ATSDR (2007) used an identical approach to that of the US EPA to derive its chronic oral risk level.
The majority of agencies have considered arsenic as a non-threshold contaminant. As (Table 3) shows, US EPA (1988) is the basis for the subsequent derivation of many drinking water and soil guideline values. However, despite a common source of data being used, the toxicological intake values used by the different jurisdictions are different. Some of this discrepancy seems to arise from the original (US EPA, 1988) study itself. Specifically, this study determined that maximum likelihood estimates (MLE) of cancer risk, for a 70-kg person who consumes 2 L of water per day contaminated with 1 μg/L of arsenic, range from 3 × 10–5 (on the basis of Taiwanese females) to 7 × 10–5 (based on Taiwanese males). The midpoint of this range gives rise to the US EPA drinking water unit risk of 5 × 10–5. The study also states the MLE of 1 μg/kg bw/day of arsenic from water ranges from 1 × 10–3 to 2 × 10–3. The midpoint of this range gives rise to the US EPA oral slope factor of 1.5 per mg/kg bw per day. However, US EPA risk assessment guidance documents (US EPA, 1989) indicate that a drinking water unit risk can be converted to a slope factor by multiplying by bodyweight (70 kg), and dividing by the volume of water drunk (2 L) and a conversion factor (1000) to convert micrograms into milligrams. Using this latter approach gives rise to a slope factor of 1.75, which is the slope factor used by Environment Canada (1999).
WHO drinking water guidelines (pre-2003) have also based risk estimates on US EPA (1988). Specifically WHO (1996) indicates that the drinking water guideline of 10 μg/L nominally gives rise to a skin cancer risk of 6 × 10–4 based on data for males provided in US EPA (1988). This is stated to give rise to estimated lifetime skin cancer risks of 10–5 and 10–6 for arsenic concentrations of 0.17 and 0.017 μg/L, respectively. It is unclear how these are derived, as US EPA (1988) indicates that, based on data for Taiwanese males, estimated skin cancer risks for a 70-kg adult consuming 2 L per day of water containing 1 μg/L of arsenic is 7 × 10–5, giving rise to a risk of 10–5 at 0.14 μg/L; arguably this difference is attributable to rounding errors. However, it should also be noted that WHO (2003) cites risk estimates from more recent studies (NRC, 2001); specifically, that the maximum likelihood estimates for the incidence of bladder and lung cancer in US populations exposed to 10 μg/L of arsenic range from 12 to 23 per 10,000, or a risk of 12 × 10–4 to 23 × 10–4. The New Zealand Drinking Water Guidelines datasheet for arsenic states the old risk estimates (MoH, 2005). The WHO drinking water guideline is used in MfE and MoH (1997) to derive a slope factor for arsenic of 0.15 per mg/kg bw/day, which also takes into account a mortality rate from skin cancer of 7%.
More recent evaluations of arsenic in drinking water have focused on the risk of internal cancers, specifically bladder and lung cancers (NRC, 1999; 2001; US EPA, 2001; FPTCDW, 2006). The south-western Taiwanese population still forms the basis for new risk estimates, although data from Chen et al (1988; 1992) and Wu et al (1989) (all cited in NRC, 1999) are used instead of Tseng et al (1968) and Tseng (1977) (both cited in WHO, 2001) as the former are considered to provide better estimates of arsenic exposure and are focused on internal cancers. The risk models developed by Morales et al (2000) form the basis for dose-response models used by US EPA (2001) and Health Canada (2003, 2005) to develop risk estimates for arsenic in drinking water. The risk models used can have a marked influence on the derived risk estimates, in addition to the extrapolation of exposure data for a Taiwanese population to other populations. Both agencies used a Poisson model, a Taiwanese to Canadian/US conversion factor, and the same risk metrics to develop their risk estimates for drinking water. However, Health Canada (2005) included an external unexposed comparison population, while US EPA (2001) didn’t. Their justification was that models with no comparison population were more reliable, as models with comparison populations resulted in supralinear (higher than a linear) dose-response, and there is no biological data to support a supralinear curve as being biologically plausible (US EPA, 2001). In contrast, Health Canada (2005) included a comparison population based on the recommendations of NRC (2001) and the fact that “an external comparison population is classically used in the analysis of cohort data (Breslow and Day 1987), since it provides a more accurate estimate of the baseline cancer rates and minimises the impact of exposure misclassification in the low dose range within the study population” (FPTCDW, 2006).
The USEPA circulated a memorandum in 2008 containing a human health risk assessment relating to a decision on the re-registration eligibility of inorganic arsenicals as wood preservatives (US EPA, 2008). The human cancer risk assessment employed an oral cancer slope figure of 3.67 per mg/kg bw/day, based on earlier risk modelling by the agency in developing new rules for arsenic in drinking water. However, it is unclear as to exactly where this information was presented. From this slope factor, an extra cancer risk of 1 in 100,000 would be associated with an oral arsenic dose level of 0.003 μg/kg bw/day.
Based on recent evaluations of the carcinogenicity of arsenic, the EA (2009) considered that an oral index dose based on excess lifetime cancer risk would lie in the range of 0.0006 to 0.003 μg/kg bw/day. However, their final recommended oral index dose for deriving soil guideline values was 0.3 μg/kg bw/day, based on equivalence with the UK drinking water standard of 10 mg/L assuming 2 L/day is consumed by a 70-kg adult, to avoid disproportionately targeting soil exposures.
It is interesting to note that the risk-specific doses determined from internal cancers are higher than determined for skin cancer from US EPA (1988) (Table 3), ie, arsenic nominally has the potential to cause skin cancer at doses lower than required for internal cancers, yet internal cancers are the focus of recent studies. The rationale for this is not explicitly stated, although it may be attributable to a higher mortality rate from internal cancers. Further, NRC (1999) questioned the validity of the US EPA (1988) results in light of new information that adds more uncertainty to the data used in this report. Specifically, it has been recognised that arsenic exposure among persons and villages grouped together in the data reported in the Tseng studies is more variable than previously realised. New risk estimates are based on studies by Chen et al (1988; 1992) and Wu et al (1989) (all cited in NRC, 1999), which provide more detailed estimates for exposure; thus it is anticipated that these estimates are more robust.
Table 2: Summary of oral reference health standards for arsenic as a threshold contaminant, used by different international agencies
| Jurisdiction | Tolerable daily intake |
Key study1 |
Critical effect1 |
Basis of value1 |
Reference |
|---|---|---|---|---|---|
Joint FAO/ WHO Expert Committee on Food Additives (JECFA) |
2.1 |
Grantham and Jones (1977) |
Not stated – basis was that of 33 people in Nova Scotia using water with arsenic concentrations >0.1 mg/L, 23 (70%) had mild symptoms and signs possibly attributable to arsenic poisoning, whereas only 25 out of 86 people (29%) consuming water with arsenic at 0.05–0.1 mg/L were similarly affected Additional epidemiological studies were used to support evidence for arsenic toxicity |
JECFA (FAO/WHO, 1983) concluded that, based on the available epidemiological evidence, water supplies containing concentrations of 0.1 mg/L may give rise to presumptive signs of toxicity. An assumed daily water consumption of 1.5 L was used to convert this value to a daily intake of 0.15 mg, and a bodyweight of 75 kg converts this to a provisional daily intake of 2 μg/kg bw This value was “confirmed” by JECFA (FAO/WHO, 1988) by assigning a PTWI of 0.015 mg/kg bw, with the “clear understanding that the margin between the PTWI and intakes reported to have toxic effects in epidemiological studies was narrow” |
FAO/WHO (1983; 1988) |
Australia |
2.1 |
Grantham and Jones (1977) |
Not stated |
FAO/WHO (1988) |
NEPC (1999) |
The Netherlands – current |
2.1 |
Grantham and Jones (1977) |
Not stated |
FAO/WHO (1988) |
Baars et al (2001) |
The Netherlands – proposed2 |
1 |
Grantham and Jones (1977) |
Not stated |
FAO/WHO (1988) with an additional safety factor of 2 to account for observation errors inherent in epidemiological studies |
Baars et al (2001) |
US ATSDR – chronic duration MRL and US EPA |
0.3 |
Tseng et al (1968), Tseng (1977), both cited in WHO (2001) |
Hyperpigmentation, keratosis and possible vascular complications Human chronic oral exposure in drinking water |
No observable adverse effect level (NOAEL) 0.009 mg/L converted3 to 0.0008 mg/kg bw/day Uncertainty factor (UF) of 3 applied to account for both the lack of data to preclude reproductive toxicity as a critical effect and to account for some uncertainty in whether the NOAEL of the critical study accounts for all sensitive individuals |
ATSDR (2007) US EPA (2005) |
1 As reported in the reference cited in the reference column.
2 This value is yet to be officially adopted.
3 Conversion assumed a water intake of 4.5 L/day, a bodyweight of 55 kg, and a daily intake of 0.002 mg As/kg from food: NOAEL – [9 μg/L x 4.5 L/day + 2 μg/day (contribution of food)] x (1/55 kg) = 0.8 μg/kg bw/day.
Table 3: Summary of oral reference health standards for arsenic as a non-threshold contaminant, used by different international agencies
Inhalation
Inhalation will be a negligible route of exposure as arsenic is not volatile and the amount of dust considered to be inhaled typically represents a very small fraction of exposure (see section 1.1.4), so is not considered further.
Dermal absorption
The skin absorption factor is the only contaminant-specific parameter required for the dermal absorption pathway. Despite the fact that skin cancer is a primary toxicological effect of concern as a result of exposure to arsenic, dermal absorption of arsenic is generally considered to be negligible. US EPA (2004) guidance uses a dermal absorption factor of 3% based on Wester et al (1993), who examined the dermal uptake of arsenic in solution. However, recent studies on the dermal absorption of soil-absorbed arsenic in rhesus monkeys indicate that the mean dermal absorption is 0.5%, ie, negligible (Lowney et al, 2007).
Other routes of exposure – background exposure
For threshold contaminants it is important to account for background exposure. As arsenic has been considered a threshold contaminant by some agencies, such exposure is considered here.
Dietary intake of arsenic is considered to be the primary source of exposure. Vannoort and Thomson (2005) estimated that dietary intake of arsenic comprised 15% of the provisional tolerable weekly intake (PTWI: 15 μg/kg bw/week) for a young child (1–3 years, 13 kg) and around 10% for young males (25+ years, 82 kg) and females. This suggests a daily intake of inorganic arsenic of approximately 0.32 μg/kg bw for a young child and 0.21 μg/kg bw for young males and females. Davies et al (2001) found arsenic in 152 drinking water zones (18% of those assessed), although they noted that most results were comfortably below 50% of the maximum acceptable value (MAV) for drinking water (10 μg/L). Data on the detection of arsenic up to 50% of the MAV indicates that the majority of detections are 0–10% of the MAV.
Assuming a daily consumption of 1 L per day at arsenic concentrations of 1 μg/L by a young child gives rise to a daily arsenic intake from drinking water of 0.067 μg/kg bw, and a total daily arsenic intake of 0.38 μg/kg bw. Consumption of 2 L per day for a young male gives rise to a daily arsenic intake from drinking water of 0.024 μg/kg bw and a total daily arsenic intake of 0.23 μg/kg bw.
It is notable that these intakes are higher than all risk-specific doses shown in Table 3, suggesting that the acceptable increased risk level of 1 in 100,000 is already exceeded by intake from food. Similarly, estimated arsenic intake for a child exceeds the US EPA reference dose, while arsenic intake for an adult male comprises ~75% of the reference dose.
However, it should be noted that dietary intake in the New Zealand Total Diet Survey (Vannoort and Thomson, 2005) is likely to be overestimated for two main reasons: assumptions regarding the proportion of inorganic arsenic, and detection limits. Only total arsenic is determined in the NZTDS, and assumptions regarding the proportion of inorganic arsenic in different foods are made to estimate inorganic arsenic intake. Specifically, the NZTDS uses US Food and Drug Administration (US FDA) assumptions, which are acknowledged to be conservative, to determine the proportion of inorganic arsenic in the diet. The US FDA assumptions are that 10% of total arsenic in fish/seafood is inorganic, and that the arsenic in all other foods is 100% inorganic. A survey of total and inorganic arsenic in 40 different foods (Schoof et al, 1999 cited in Vannoort and Thomson, 2005) reported that of the total arsenic, inorganic arsenic ranged from <1% to 100% as follows: marine fish (<1%), beef, chicken (1%), tomatoes (10%), rice (24%), potatoes (28%), apples and grapes (38%), carrots (53%), spinach and peas (100%). These assumptions also alter the importance of different food types; for example, for a young (19–24 years) male, fish forms approximately 83% of total arsenic intake and approximately 50% of the inorganic arsenic intake, while rice forms 2% of total arsenic intake and 13% of inorganic arsenic intake. Detection limits influence estimated dietary intake of an element, as, where a contaminant is determined to be below the detection limit for a particular food, half the detection limit is assigned to that food to estimate the total intake. This approach is commonly used and has been used in previous New Zealand dietary surveys. If the detection limit is high for a particular food and arsenic is not detected in this food, then it is likely the contribution of this food to total arsenic intake is overemphasised. Irrespective of the detection limit, inclusion of these “non-detects” is likely to overestimate the total arsenic intake. Nonetheless, this approach is used in the NZTDS for consistency with previous and international studies.
Summary of effects
Arsenic exposure can have numerous cancerous and non-cancerous effects including dermal lesions, pigmentation, keratoses, peripheral vascular disease (eg, blackfoot disease), and cardiovascular effects, skin cancer and internal cancers (bladder, lung, liver) (US EPA, 2001; WHO, 2001; ATSDR, 2007). Table 4 provides a summary of the effects at different levels of arsenic exposure and has primarily been sourced from ATSDR (2007). Cancer is the endpoint most consistently seen as a consequence of long-term chronic exposure, and is also the endpoint with the most extensive quantitative information available on dose-response.
Table 4: Summary of the health effects of arsenic
| Dose (mg /kg/day | Type of poisoning |
Effects |
|---|---|---|
>2 |
Acute |
Vomiting, diarrhoea, and abdominal pain, headache, lethargy, mental confusion, hallucination, seizures, and coma |
>0.065–0.14 |
Chronic |
Cardiovascular diseases such as “blackfoot disease”, which is endemic in an area of Taiwan where average arsenic concentrations in drinking water range from 0.17 to 80 μg/L |
0.03–0.1 |
Chronic |
Peripheral neuropathy, characterised initially by numbness of the hands and feet and a “pins and needles” sensation and progressing to muscle weakness, wrist-drop and/or ankle-drop, diminished sensitivity, and altered reflex action. Reports of neurological effects at lower arsenic levels (0.004–0.006 mg/kg per day) have been inconsistent |
0.01 |
Chronic |
Vomiting, diarrhoea, and abdominal pain – symptoms diminish after cessation of exposure |
>0.002 |
Chronic |
Skin lesions (hyperkeratinisation and hyperpigmentation) |
0.0012 |
Chronic |
Lowest reported dosage associated with increased incidence of skin lesions |
0.0000086 |
Chronic |
Bladder and lung cancers – negligible risk based on consumption of 0.3 μg/L in drinking water (FPTCDW, 2006) |
0.000001–0.000002 |
Chronic |
Bladder and lung cancers (NRC, 2001) |
Weight of evidence
- Arsenic is a known human carcinogen, based on human epidemiological studies that show increases in skin cancers and internal cancers (particularly bladder, liver and lung) as a result of chronic exposure to inorganic arsenic in drinking water (IARC, 1987; 2004; NRC, 1999; 2001; WHO, 2001; Mead, 2005; FPTCDW, 2006).
- Inorganic arsenic does not induce point mutations in bacterial and mammalian assays, although it does induce clastogenic effects (chromosome breakage, chromosomal aberrations and sister-chromatid exchange) in in vitro studies (ie, is genotoxic) (Rudel et al, 1996; Basu et al, 2001; Hughes, 2002; Rossman, 2003).
- Arsenic enhances the clastogenicity and mutagenicity of other DNA-damaging agents, ie, acts as a co-carcinogen (Rudel et al, 1996; NRC, 1999; Hughes, 2002; Rossman, 2003).
- The mechanisms of genotoxicity and carcinogenicity of arsenic are unknown, although several modes of action have been suggested including inhibition of DNA repair, altered DNA methylation patterns, and oxidative stress (Basu et al, 2001; Hughes, 2002; Rossman, 2003).
- While there is a general consensus that arsenic is likely to act indirectly on DNA in a sub-linear or threshold manner (eg, Rudel et al, 1996; Mead, 2005), it is also generally considered that there is insufficient data available to determine a “well-defined non-linear dose-response” (US EPA, 2001). Hence, a linear dose-response is assumed, ie, arsenic is treated as a non-threshold contaminant (eg, NRC 1999; 2001; US EPA, 2001; EA, 2009; FPTCDW, 2006).
- Given the consensus of international regulatory agencies, it is recommended that until further data is available on its mode of action, arsenic should be treated as a non-threshold contaminant in New Zealand.
Recommendations for toxicological intake values
Classification of arsenic as a non-threshold contaminant is consistent with its classification by ERMA NZ, since a mutagenicity (6.6B) classification has been applied indicating arsenic is a genotoxic carcinogen. As such, a risk-specific dose is proposed.
The risk-specific dose of 0.0086 μg/kg bw/day, derived from the arsenic concentration in drinking water determined to represent “negligible risk” (0.3 μg/L) by Canadian agencies (FPTCDW, 2006), is recommended (Table 5). This value is based on the most current risk-modelling data, and includes an external comparison population. It should be noted that as the risk estimates used in FPTCDW (2006) were based on exposure to arsenic via drinking water, the recommended risk-specific dose is likely a conservative estimate for intake via contaminated soil, as arsenic in contaminated soil will be less bioavailable than arsenic in drinking water. However, there is insufficient data to be able to quantitatively take account of this. Dermal absorption is considered to be negligible, although the skin absorption factor of 0.5% (Lowney et al, 2007) could be used as a refinement in the development of soil guideline values. An inhalation dose is not considered relevant for soil contamination by non-volatile substances (section 1.13). Because it is recommended that arsenic should be considered as a non-threshold contaminant, background exposure is irrelevant: exposure from all sources should be as low as reasonably practicable.
Table 5: Recommended toxicological criteria for arsenic
| Parameter | Value |
Basis |
|---|---|---|
Contaminant status |
Non-threshold |
See weight of evidence |
Oral risk-specific dose (μg/kg bw/day) |
0.0086 |
From Health Canada (2005) risk-modelling for internal cancers, which includes an unexposed comparison population |
Inhalation intake |
NA |
Lack of volatility of arsenic indicates inhalation exposures are minimal |
Skin absorption factor |
0.005 |
Lowney et al (2007) |
Background exposure (μg/kg bw/day) |
NA |
Exposure to non-threshold contaminants from all sources should be as low as reasonably practicable |
NA – not applicable.
