Page 3: Guidelines for Canadian Drinking Water Quality: Guideline Technical Document – Haloacetic Acids
4.0 Identity, use and sources in the environment
There are nine common HAAs: MCA, DCA, TCA, MBA, DBA, bromochloroacetic acid, bromodichloroacetic acid, chlorodibromoacetic acid and tribromoacetic acid. This document focuses on the first five HAAs on this list, referred to as either HAA5 or total HAAs.
HAAs belong to the family of halogenated aliphatic carboxylic acids. Although these chemical analogues will in most cases be referred to as "acids" in this document, it should be understood that when present in drinking water at normal pHs, they will in fact be present as salts and strictly should be called acetates (EC, 2003; U.S. EPA, 2003b). The physical-chemical properties of the HAA5 compounds shown in Table 2 apply to the acids.
MCA is described as a colourless solution or white crystal with a vinegar-like odour ( Budavari et al., 1996; CHEMINFO, 2003a, 2003b). It is used mainly as a chemical intermediate in the production of cellulose ethers (mainly carboxymethylcellulose), thioglycolic acid and herbicides (Morris and Bost, 2002). It is also used in the manufacture of glycine, phenoxyacetic acid, sarcosine, amphoteric surfactants, synthetic caffeine, various indigo dyes, pharmaceuticals, preservatives (ethylenediaminetetraacetic acid) and bacteriostats (Lewis, 2001; Koenig et al., 2002; Morris and Bost, 2002).
DCA is a colourless to slightly yellow liquid with a pungent odour (IARC, 1995; Budavari et al., 1996). It is used as a topical astringent, fungicide and medicinal disinfectant, as a test reagent for analytical measurements, to treat lactic acidosis and in the synthesis of organic materials, including pharmaceuticals (Budavari et al., 1996; Koenig et al., 2002; Morris and Bost, 2002).
TCA is a colourless to white deliquescent crystal with a sharp, pungent odour (Ashford, 1994; Budavari et al., 1996). TCA is used as an intermediate in the synthesis of organic chemicals and as a laboratory reagent, herbicide, soil sterilizer and antiseptic (Budavari et al., 1996; Lewis, 2001; Verschueren, 2001; Meister, 2002). It has been used as an etching or pickling agent, a swelling agent and a solvent in plastics and in textile finishing (Koenig et al., 2002). Clinically, TCA has been used in 10-25% aqueous solutions in the treatment of recurrent corneal disease (Grant and Schuman, 1993), for the treatment of external cervical root resorption in dentistry (Heithersay and Wilson, 1988; Lewinstein and Rotstein, 1992) and to treat various skin afflictions (Koenig et al., 2002). TCA has been used as a facial chemical peel and for other therapeutic applications, such as a cauterizing agent, in wart removal and as an astringent (NTP, 2003a).
MBA is a colourless hygroscopic crystalline solid (Ashford, 1994). It has been used in organic synthesis, abscission of citrus fruit in harvesting (Lewis, 2001), commercial letterpress printing and production of plastics, as well as in medical and surgical hospitals (NIOSH, 1990).
HAAs are formed in drinking water when chlorine disinfectants used in water treatment react with organic matter (e.g., humic or fulvic acids) and inorganic matter (e.g., bromide ion) naturally present in the raw water (IPCS, 2000). HAAs are the second most frequently occurring DBPs, after THMs.
Various water treatment methods lead to the formation of chlorinated and brominated acetic acids, including chlorination, ozonation and chloramination. In the case of chlorination, hypochlorous acid (HOCl) and the hypochlorite ion (OCl-) are formed, which in turn react with a bromide ion, if present, oxidizing it to hypobromous acid (HOBr-) and hypobromite ion (OBr-), respectively. Hypochlorous acid and hypobromous acid then react with natural organic matter (NOM) to form different DBPs, including HAAs. The chlorinated HAAs generally dominate; however, in high-bromide waters, the brominated HAAs may be more prevalent (IPCS, 2000). In the case of ozonation, brominated acetic acids (MBA, DBA) can be formed when organic matter and bromide are present in the source waters (U.S. EPA, 2005a). Chloramination also results in HAA production if chloramine is produced by chlorination followed by ammonia addition (IPCS, 2000).
HAA formation can be appreciable when drinking water is chlorinated under conditions of slightly acid pH (IPCS, 2000). Whereas THM formation increases with increasing pH, HAA formation decreases, hydrolysis likely being a significant factor (Krasner et al., 1989; Pourmoghaddas and Stevens, 1995). Despite the fact that HAAs and THMs have different pH dependencies, their formation appears to correlate strongly when treatment conditions are relatively uniform and when the water has a low bromide concentration (Singer, 1993).
Longer contact times and higher water temperatures are contributing factors in HAA formation. At higher water temperatures, reactions are faster and chlorine demand is higher (Nikolaou et al., 1999). Increased concentrations of NOM with aromatic content (humic acids) in raw water favour formation of HAAs (Reckhow et al., 1990; Nikolaou et al., 1999). Increased concentrations of NOM in raw water also increase the chlorine demand and favour the formation of chlorinated DBPs. In the presence of bromide, the chlorination process may also favour the formation of brominated DBPs, depending on the physical and chemical properties of the water. High chlorine concentrations also favour the formation of higher concentrations of TCA compared with MCA and DCA. However, if bromide levels are high in source waters, the formation of brominated and chloro-brominated HAAs is more likely to occur (Nikolaou et al., 1999). Bromide levels in surface water and groundwater may fluctuate seasonally and may occur as a result of saltwater intrusion or pollution as well as from natural sources (Richardson et al., 1999; IPCS, 2000).
The HAA5 compounds may be released into the environment through various waste streams following their production and use. MCA and TCA can be formed as combustion by-products of organic compounds (waste incineration) in the presence of chlorine (Juuti and Hoekstra, 1998). Other potential sources of atmospheric TCA are the photooxidation of tetrachloroethylene (PCE), trichloroethylene (TCE) and 1,1,1-trichloroethane (Reimann et al., 1996; Sidebottom and Franklin, 1996; Juuti and Hoekstra, 1998; Bakeas et al., 2003), as well as biomass burning and natural formation in the marine boundary layer (Hoekstra, 2003). Some atmospheric MCA may also be formed from the hydrolysis of monochloroacetanilide herbicides (Reimann et al., 1996;) and directly or indirectly from car exhaust (Bakeas et al., 2003). DCA is believed to be a minor atmospheric degradation product of TCE (Peters, 2003).
HAAs are present in raw water, possibly the result of chlorinated municipal waste effluent, drinking water inputs, precipitation, the degradation of herbicides and industrial inputs involving reactions between chlorine and organic material. Scott et al. (2002) found that HAA levels in raw surface water corresponded with the level of industrial activity in the surrounding area. Concentrations of HAAs in the Detroit River were as follows: MCA, <0.005-0.59 µg/L; DCA, 0.48-1.2 µg/L; TCA, 0.1-2.2 µg/L; MBA, <0.005-0.04 µg/L; and DBA, <0.005-0.26 µg/L. Lake Malawi (Africa), with little industry, had no detectable levels of HAAs, whereas the Laurentian Great Lakes had total concentrations of approximately 0.5 µg/L, consisting of TCA, DCA and MCA; no significant levels of bromoacetic acids were detected at either of these locations (Scott et al. 2002).
4.1 Environmental fate
Volatilization from water surfaces is not expected based upon the low vapour pressures and high water solubilities of the HAA5 compounds. Low pKa values indicate that these compounds will exist almost entirely in the ionized form at pH values found in drinking water.
Microbial degradation of MCA in water is most likely the main aquatic degradation pathway. MCA was biodegraded in stream water with 73% conversion to carbon dioxide in 10 days at 29°C under laboratory conditions at the highest concentration used (Boethling and Alexander, 1979). In comparison, DCA was found to be more persistent in the aquatic environment. At a concentration of 10 mg/L, only 14% and 8% degradation were reported for river
water and seawater, respectively, after 3 days of incubation (Kondo et al., 1988). TCA is likely to be relatively persistent in water given its high solubility and low vapour pressure. The estimated half-lives of TCA in a model river and a model lake were 1632 and 12 000 days, respectively (HSDB, 2003). In a study to determine the stability of HAAs incubated in river water and seawater (20°C) for 30 and 9 days, respectively, TCA concentrations did not decrease significantly, whereas MCA, DCA, MBA and DBA almost completely disappeared (Hashimoto et al., 1998). The same authors indicated that approximately half the decomposition was due to microbial activity.
Available data suggest that drinking water may be a significant source of exposure to HAAs, but there are few data available to determine the exposure from other media, such as food and air. Since HAAs are neither volatile nor absorbed significantly through the skin, exposure via dermal and inhalation routes is considered negligible.
In general, levels of HAAs are highest in treated water from sources with high organic matter content, such as rivers and lakes, and low when the source water is groundwater. Within a single distribution system, however, HAA levels can vary greatly, depending on both water quality (e.g., HAA precursors, pH, temperature, ammonia and carbonate alkalinity) and treatment conditions (e.g., disinfectant dose, contact time, removal of NOM before the point of disinfectant application, prior addition of disinfectant) (Nikolaou et al., 1999; CDBP Task Group, 2000; IPCS, 2000).
Health Canada has conducted a series of studies to characterize the presence of CDBPs, including HAAs, in drinking water from treatment plants of various sizes using surface water and groundwater and using different disinfection processes. A 1995 survey investigated 53 sites, covering nine Canadian provinces, to determine the concentrations of HAA5 in drinking water for larger communities (10 000-100 000 people). Treatment plants in the study used one of three disinfection processes: chlorine-chlorine (n = 35), chlorine-chloramine (n = 10) and ozone- chloramine (n = 7). Samples were collected during winter and summer months for raw water, treatment plant water (after final disinfection) and treated water from the distribution system (5-10 km from the treatment plant) (Health Canada, 1995).
All HAAs were non-detectable (<0.01 µg/L) or at very low levels in raw water. DCA and TCA were the major HAAs present in treatment plants and distribution systems (winter and summer) for all treatment processes; concentrations ranged from 0.2 to 163.3 µg/L and from <0.1 to 473.1 µg/L, respectively. MCA, MBA and DBA were detected at concentrations ranging from 0.03 to 9.7 µg/L, from <0.01 to 9.2 µg/L and from <0.01 to 1.9 µg/L, respectively. Most treatment plants and distribution systems had DCA concentrations below 50 µg/L. Generally, mean DCA concentrations in treatment plants and distribution systems were higher for the chlorine-chlorine disinfection process, and mean summer concentrations in treatment plants and distribution systems were slightly higher than those in winter. Most treatment plants and distribution systems had TCA levels below 50 µg/L, although a few facilities using the chlorine-chlorine process had relatively high values (>100 µg/L). As with DCA, mean TCA concentrations were higher in summer than in winter for all processes. A comparison of TCA concentrations for the chlorine-chlorine process for both seasons indicated that there was a marked increase going from the treatment plant to the distribution system (Health Canada, 1995).
Table 3 shows the results of another Health Canada study (Aranda-Rodriguez et al., 2002; Health Canada, 2003), in which CDBPs (including HAA5) were surveyed in treated drinking water from small systems located in 27 communities (<10 000 people) within nine provinces. Sixteen of the 27 systems used chlorine only, while the balance used chlorine combined with flocculation and filtration processes. A majority of the locations (n = 23) used surface water, whereas two used groundwater and two used a combination of surface water and groundwater. Samples were collected in the warm water season (August to September 1999) and cold water season (January to March 2000) from five sites at each location: raw water, treatment plant (T) and within the distribution system at 0.1-6 km (D1, close to treatment facility), 0.75-16 km (D2, midpoint of system) and 1-23 km (D3, far system site).
No CDBPs were detected in raw water samples. In the treated water, THMs and HAAs accounted for 80% of the CDBPs. DCA and TCA were the most prevalent HAAs, and their concentrations for all locations ranged from <0.3 to 231 µg/L and from <0.1 to 257 µg/L, respectively. Concentrations of MCA, MBA and DBA ranged from <0.3 to 17.4 µg/L, from <0.4 to 18 µg/L and from <0.1 to 4.6 µg/L, respectively.
DCA and TCA concentrations in summer for small treatment plants and distribution systems significantly exceeded those in winter, whereas MCA concentrations in summer slightly exceeded those in winter. In summer, mean concentrations of MCA, DCA and TCA peaked in the treatment plant (T), at the D1 site and at the D2 site, respectively, indicating different formation-degradation patterns for these compounds in warm water conditions (Table 3). In winter, mean concentrations of MCA, DCA and TCA all peaked at the D2 site. Concentrations of MBA and DBA were relatively constant, regardless of the site or season.
Average DCA levels at D2 (midway in the distribution system) during summer (57.4 µg/L, Table 3) were higher than the average system concentrations of DCA in larger facilities (19.0 µg/L, chlorine-chlorine). A similar comparison for the cold water season revealed that average DCA concentrations were higher in the small systems (41.5 µg/L, Table 3) than in the larger systems (15.6 µg/L, chlorine-chlorine). Generally, a greater fraction of small systems had DCA values above 50 µg/L, and concentrations tended to increase in the distribution system after treatment (Table 3), whereas DCA concentrations appeared to level off to a greater extent in the distribution systems of larger facilities (chlorine-chlorine).
|HAA concentrations (µg/L)|
Average TCA levels were always higher in the distribution system than in the treatment plant, regardless of system size or season. A comparison of mean TCA values during summer in the distribution systems of small and large facilities indicated that smaller facilities (65.9 µg/L, D2, Table 3) had higher concentrations than large facilities (48.9 µg/L, chlorine-chlorine). In wintertime, there were higher concentrations in the distribution system of larger facilities (56.7 µg/L, chlorine-chlorine) than in the small facilities (42.6 µg/L, D2, Table 3).
The above Health Canada studies indicated that of the five HAAs, DCA and TCA were present in the highest concentrations. DCA and TCA levels ranged from <0.3 to 231 µg/L and from <0.1 to 473 µg/L, respectively, for both studies. Generally, concentrations of both compounds peaked in the distribution system (chlorine treatment) and decreased in the extremities of the system, were higher in summer than in winter and were higher in smaller facilities than in larger ones. Frequently, DCA peaked before TCA, indicating that the former may have a faster rate of formation and degradation. The remaining HAA5 compounds, MCA, MBA and DBA, were found at concentrations ranging from <0.01 to 18 µg/L. A comparison of HAA5 levels for the different disinfection processes indicated that levels were generally higher for plants using chlorine. Since HAA5 concentrations vary within and between distribution systems, depending on different factors, including water quality characteristics (e.g., HAA precursors, pH, season, temperature) and treatment conditions (e.g., disinfectant type, disinfectant dose, contact time), it is recommended that monitoring samples be taken at the water treatment plant and at points in the distribution system where historical data show the highest HAA concentrations.
The spatial variation in HAA5 concentrations in the distribution systems noted in these studies may be explained in part by differences between disinfectant residuals (chlorine versus chloramine) and the susceptibility of individual HAAs to microbial biodegradation. A U.S study (Williams et al., 1998) reported unexpectedly low HAA concentrations at the maximum residence time locations in distribution systems. Analysis of water quality parameters revealed that water at the maximum residence time locations had low levels of free chlorine and high heterotrophic plate counts. Others have previously identified specific bacteria and haloacid dehalogenase as being capable of degrading DCA (Uchiyama et al., 1992; Meusel and Rehm, 1993). Research on haloacid dehalogenase has shown it to have some degree of substrate selectivity, where MCA, DCA, MBA and DBA were degraded while TCA was not (Ploeg et al., 1991). Another factor that may affect the spatial variation of HAA5 in distribution systems is the pH. The rate of formation of TCA is significantly favoured by low pHs (<pH 7), whereas the rate of DCA formation is only slightly higher at low pHs (Miller and Uden, 1983).
Data from 193 communities in Newfoundland and Labrador for the period 1999-2003 indicated that DCA and TCA were the main HAAs present in treated distributed water. Samples had the following concentrations of HAA5: TCA, <1-600 µg/L (average 66.2 µg/L); DCA, <1-499 µg/L (average 50.2 µg/L); MCA, <1-15 µg/L (average 1.1 µg/L); DBA, <1-13 µg/L (average 0.4 µg/L); and MBA, <1-4 µg/L (average 0.1 µg/L). Total HAA5 concentrations for all communities ranged from <1 to 1114 µg/L and averaged approximately 111 µg/L (Newfoundland and Labrador Department of Environment, 2003).
Monitoring data (1999-2003) for 178 Ontario communities similarly indicated that DCA and TCA were the main HAAs present in treated distributed water. Concentration ranges for HAA5 compounds were as follows: TCA, <0.05-141 µg/L; DCA, 0.2-95.9 µg/L; MCA, 0.5-30.5 µg/L; MBA, 0.05-26.6 µg/L; and DBA, 0.05-17.0 µg/L. Total average HAA5 concentrations (based on individual averages for each compound) for all communities ranged from approximately 1.2 to 142.8 µg/L (Ontario Ministry of Environment and Energy, 2003).
Monitoring data from 37 communities in Manitoba for 2000 indicated that DCA, MCA and TCA were the main HAAs present in treated plant water. Samples (n = 47) had the following concentrations: DCA, <0.5-210 µg/L (average 63 µg/L); MCA, <1-51 µg/L (average 7.9 µg/L); TCA, <0.5-35 µg/L (average 6.7 µg/L); DBA, <0.5-5.4 µg/L (average 0.9 µg/L); and MBA, <0.5-3.1 µg/L (average 0.9 µg/L). Total HAA5 concentrations for all communities ranged from 2.5 to 268 µg/L and averaged approximately 80 µg/L (Manitoba Department of Conservation, 2004).
Non-ingestion exposure (dermal and inhalation) to HAAs via showering and bathing was found to be insignificant, because these compounds are neither volatile nor lipophilic (Xu et al., 2002; Xu and Weisel, 2003).
5.1.1 Analysis of HAA5 data
In order to obtain a better understanding of how average HAA5 data for surface water varied in magnitude and if there were any major differences according to community size, an analysis of provincial and territorial monitoring data from 1990 to 2004Footnote * was carried out. Average HAA5 values for each location were calculated based on data (n = 1-24) provided for seasonal quarters (January-March, April-June, July-September, October-December) for the period 1999-2004Footnote *. The data used were not necessarily quarterly averages because of data scarcity; some locations had data for only one season, and others had data for many seasons. The location of sampling between sites also varied.
126.96.36.199 Communities with >5000 persons
Health Canada received HAA data from 135 water treatment plants (distribution systems) serving communities of greater than 5000 persons, representing a total population of approximately 19.3 million. The majority of these systems were located in Ontario, Quebec, Nova Scotia and Newfoundland and Labrador, with a few plants located in Alberta, British Columbia, Manitoba, Saskatchewan and the Yukon. Of these, 88% had average HAA5 concentrations below 80 µg/L, while 12% exceeded this level (Table 4). On average, DCA accounted for 46% of the total concentration of HAA5. However, for a significant percentage of systems (26%), DCA accounted for 50-59% of HAA5.
|Province/ territory||No. of systems per province/ territory||No. of systems below 80 µg/ L HAA5||No. of systems above 80 µg/ L HAA5|
|Newfoundland and Labrador||10||5||5|
|Total no. of systems||135||119||16|
188.8.131.52 Communities with <5000 persons
Health Canada received HAA data from 312 systems serving communities of fewer than 5000 persons, representing a total population of approximately 333 300. The systems were located in Ontario, Quebec, Nova Scotia and Newfoundland and Labrador. Of these, 56% of systems had average HAA5 concentrations below 80 µg/L, whereas 44% of systems exceeded this level. On average, DCA accounted for 42% of the HAA5. However, for a significant percentage of water treatment plants (25%), DCA accounted for 50-59% of HAA5.
|Province/ territory||No. of systems per province||No. of systems below 80 µg/ L HAA5||No. of systems above 80 µg/ L HAA5|
|Newfoundland and Labrador||220||108||112|
|Total no. of systems||312||174||138|
Stack gases of municipal waste incinerators have been reported to contain 0.37-3.7 µg TCA/m3 and 3.2-7.8 µg MCA/m3 (Mowrer and Nordin, 1987).
Air samples taken in rural Scotland and the Netherlands contained DCA and TCA concentrations of #0.0007 µg/m3 (Heal et al., 2003; Peters, 2003), whereas atmospheric particulate measurements of MCA, DCA and TCA from Athens, Greece, ranged from 0.01 to 2.01 ng/m3, from 0.0006 to 0.46 ng/m3 and from 0.0009 to 0.125 ng/m3, respectively (Bakeas et al., 2003).
The detection of chlorinated acetic acids in rain water is an indication of their presence in the atmosphere. Reimann et al., (1996) reported the following levels in rainwater: MCA, 0.05- 9 µg/L; DCA, 0.05-4 µg/L; and TCA, 0.01-1 µg/L. Rainwater in Germany contained 1.35 µg DCA/L and 0.1-20 µg TCA/L (IARC, 1995). Sidebottom and Franklin (1996) reported that TCA concentrations in rainwater in remote areas (Antarctic, Arctic and sub-Arctic regions) generally ranged from 10 to 100 ng/L.
No information was available on exposure to MBA or DBA in air (U.S. EPA, 2003b). No Canadian data were available.
It is speculated that MCA, DCA and TCA may be found in meat and other food products
(U.S. EPA, 2003b). This would result from the use of chlorine in food production and processing, including the disinfection of chicken; processing of seafoods, poultry and red meats; sanitizing of equipment and containers; and oxidizing and bleaching in the flour industry (U.S. EPA, 1994).
MCA and TCA can be taken up from cooking water (Raymer et al., 2001). In addition, there is evidence that TCA may be taken up by plants via the roots or by leaves via uptake from the air (Schroll et al., 1994; Sutinen et al., 1995). TCA was found in food (vegetables and fruits) at concentrations of 0.01-0.19 mg/kg following irrigation (Demint et al., 1975). Reimann et al. (1996) examined the concentrations of MCA, DCA and TCA in a limited number of samples of several vegetables, fruits, grain and beer. MCA concentrations ranged from <0.7 to 5.3 µg/kg in vegetables, from 1.7 to 13.2 µg/kg in grains, from 2.3 to 11.8 µg/kg in flours/breads and from 0.2 to 2.6 µg/L in beer. DCA concentrations ranged from <0.9 to 3.5 µg/kg in vegetables, from <0.6 to 11.1 µg/kg in grains, from 0.8 to 19.8 µg/kg in flours/breads and from 1.5 to 15.2 µg/L in beer. TCA concentrations ranged from <0.2 to 5.9 µg/kg in vegetables and from <1.6 to 4.1 µg/kg in grains. TCA was below the detection limit of 1.5 µg/kg in breads and was not analysed in the flours or beer. None of these compounds was detected in fruits or tomatoes.
No information on concentrations of MBA or DBA in food was located (U.S. EPA, 2003b).
5.4 Contribution of drinking water to total exposure
The data for HAA5 in water suggest that drinking water has the potential to be a significant source of these compounds. While data for air and food indicate that these media are also potential sources of HAA5, the data are insufficient to quantify their relative contributions with reasonable certainty. On this basis, a default value of 20% can be used to describe the contribution of drinking water to total daily intake. The U.S. Environmental Protection Agency (U.S. EPA, 2003b) came to a similar conclusion with respect to lack of data for air and food and selected a default relative source contribution of 20% for MCA and TCA. No relative source contribution value was assigned specifically to DCA and DBA because it is classified as a carcinogen.
6.0 Analytical methods
HAAs are relatively non-volatile and hydrophilic organic compounds. These properties make common analytical methods (e.g., purge and trap, headspace and liquid-liquid extraction) less effective for HAA separation. To facilitate analysis using gas chromatography with electron capture detector (GC/ECD), these acids must be chemically converted into methyl esters (methylated HAAs).
Three U.S. EPA methods (EPA Method 552.1, EPA Method 552.2 and EPA Method 552.3) are approved for measuring HAA5 in drinking water (U.S. EPA, U.S. EPA, 1992, 1995, 2003d). Method detection limits (MDLs) for these methods vary depending on the analyte being measured, as outlined in Table 6. The U.S. EPA also recognizes American Public Health Association (APHA) Standard Method 6251B (APHA et al., 2005) as being equivalent to the U.S. EPA standard methods for measuring HAA5 in drinking water (U.S. EPA, 2003a). All methods include sample extraction, methylation and GC/ECD analysis (using a capillary GC/ECD) steps.
EPA Method 552.1 (U.S. EPA, 1992) employs solid-phase extraction via ion exchange resins. EPA Method 552.2 (U.S. EPA, 1995) and APHA Standard Method 6251B (APHA et al., 2005) both employ micro liquid-liquid extraction with methyl tert-butyl ether (MTBE) at acidic conditions. Sodium sulphate and sulphuric acid are added to samples to increase extraction efficiency. EPA Method 552.3 (U.S. EPA, 2003e) provides the option of performing an acidic extraction using either MTBE or tertiary amyl methyl ether (TAME) before adding acidic methanol to the extract, followed by heating.
The MDLs and practical quantitation limits (PQLs) for the various methods are summarized in Table 6.
|EPA Method 552.1||EPA Method 552.2||EPA Method 552.3||APHA Standard Method 6251B|
|Analyte||MDL (µg/L)Tableau 6 note de bas de page a||PQL (µg/L)||MDL (µg/L)Tableau 6 note de bas de page b||PQL (µg/L)||MDL (µg/L)Tableau 6 note de bas de page c||PQL (µg/L)||MDL (µg/L)Tableau 6 note de bas de page d||PQL (µg/L)|
7.0 Treatment technology
Although HAA formation in water is largely a function of the amount of organic compounds in water and their contact time with chlorine, it is important to recognize that the use of chlorination and other disinfection processes has virtually eliminated waterborne microbial diseases. In order to reduce HAA levels in the finished water, it is important to characterize the source water to ensure that the treatment process is optimized for precursor removal.
7.1 Municipal scale
There are three approaches to limiting the concentrations of HAAs in municipally treated drinking water:
- treatment of water to remove HAA precursors prior to disinfection;
- the use of alternative disinfectants and disinfection strategies; and
- treatment of water to remove HAAs after their formation.
The majority of changes occurring in the water industry now focus on strategies to remove DBP precursors prior to disinfection and the use of alternative disinfectants and alternative disinfection strategies.
7.1.1 Removal of precursors prior to municipal disinfection
The removal of organic precursors is the most effective way to reduce the concentrations of all DBPs, including HAAs, in finished water (U.S. EPA, 1999c; Reid Crowther & Partners Ltd., 2000). These precursors include synthetic organic compounds and NOM, which can react with disinfectants to form HAAs. Removing HAA precursors will also result in the formation of lower concentrations of HAAs (Reid Crowther & Partners Ltd., 2000). Conventional municipal-scale water treatment techniques (coagulation, sedimentation, dissolved air flotation, precipitative softening and filtration) can reduce the amount of HAA precursors, but are ineffective in removing HAAs once they are formed. Granular activated carbon (GAC), membranes and ozone-biofiltration systems can also remove organic matter from water. The U.S. EPA has identified precursor removal technologies such as GAC and membrane filtration, specifically nanofiltration, as Best Available Technologies (BAT) for controlling DBP formation (U.S. EPA, 2005b).
Potassium permanganate can be used to oxidize organic precursors at the head of the treatment plant, thus minimizing the formation of by-products at the disinfection stage (U.S. EPA, 1999a). The use of ozone for oxidation of precursors is currently being studied. Early work has shown that the effects of ozonation, prior to chlorination, depend on treatment design and raw water quality and thus are unpredictable. The key variables that seem to determine the effect of ozone are dose, pH, alkalinity and the nature of the organic material in the water. Ozone has been shown to be effective at reducing precursors at low pH. However, at pH levels above 7.5, ozone may actually increase the production of CDBP precursors (U.S. EPA, 1999a).
7.1.2 Alternative municipal disinfection strategies
The use of alternative disinfectants, such as chloramines (secondary disinfection only), ozone (primary disinfection only) and chlorine dioxide (primary and secondary disinfection), is increasing. However, each of these alternatives has also been shown to form its own set of DBPs. Combinations of disinfectants, when optimized, can help control HAA formation. Preozonation is feasible for water sources that have turbidity levels below 10 nephelometric turbidity units (NTU) and bromide concentrations below 0.01 mg/L, to minimize the formation of bromate (Reid Crowther & Partners Ltd., 2000). Ultraviolet (UV) disinfection is also being used as an alternative disinfectant. Since UV disinfection is dependent on light transmission to the microbes, water quality characteristics affecting UV transmittance must be considered in the design of the system. UV irradiation at typical doses and wavelengths does not affect HAA formation in subsequent chlorination or chloramination steps (Reid Crowther & Partners Ltd., 2000). Neither ozone nor UV disinfection leaves a residual disinfectant, and both must therefore be used in combination with a secondary disinfectant to maintain a residual in the distribution system.
It is recommended that any change made to the treatment process, particularly when changing the disinfectant, be accompanied by close monitoring of lead levels in the distributed water. A change of disinfectant has been found to affect the levels of lead at the tap; in Washington, DC, for example, a change from chlorine to chloramines resulted in significantly increased levels of lead in the distributed drinking water. When chlorine, a powerful oxidant, is used as the disinfectant, lead dioxide scales formed in distribution system pipes reach a dynamic equilibrium in the distribution system. In Washington, DC, switching from chlorine to chloramines decreased the oxidation-reduction potential of the distributed water and destabilized the lead dioxide scales, which resulted in increased lead leaching (Schock and Giani, 2004). Subsequent laboratory experiments by Edwards and Dudi (2004) and Lytle and Schock (2005) confirmed that lead dioxide deposits could be readily formed and subsequently destabilized in weeks to months under realistic conditions of distribution system pH, oxidation-reduction potential and alkalinity.
7.1.3 Removal of HAAs after formation
Although precursor removal is considered the most effective approach to reduce HAA concentrations, removal of HAAs is also possible. Adsorption onto activated carbon is widely used to remove organic compounds such as HAAs from drinking water. This method involves pumping water through a bed of activated carbon onto which HAA molecules become attached (adsorbed). If an activated carbon filter bed is deep enough to allow sufficient contact time, it can be effective in removing HAAs from drinking water. Biofiltration may be an effective process for removing biodegradable organic matter and biodegradable DBPs from water. GAC, anthracite, sand and garnet are common media that support colonization by bacteria and can be used as biological filters. Information on the use of biofiltration for HAA removal is limited, although work has shown that bacteria-colonized GAC (biologically active carbon) is an effective process for HAA removal (Xie, 2004).
7.2 Residential scale
Generally, it is not necessary to use drinking water treatment devices with municipally treated water. In cases where municipal treatment has produced low concentrations of HAAs in drinking water, some residential-scale point-of-entry or point-of-use treatment technologies such as activated carbon, reverse osmosis or distillation systems may remove the HAAs from the drinking water. At this time, however, none is certified specifically for HAA removal.
NSF International (NSF) has developed several standards for residential water treatment devices designed to reduce the concentrations of various types of contaminants in drinking water, but HAAs are not currently included in any NSF standard. Research is ongoing in the private and public sectors to test and adopt efficient methods for the reduction of HAAs in drinking water. Products that use adsorption or reverse osmosis technology can lose removal capacity through usage and time and need to be maintained and/or replaced. Consumers should verify the expected longevity of the adsorption media or membrane in their treatment device as per the manufacturer's recommendations and service it when required.
Health Canada does not recommend specific brands of drinking water treatment devices, but it strongly recommends that consumers look for a mark or label indicating that the device has been certified by an accredited certification body as meeting the appropriate NSF/American National Standards Institute (ANSI) standards. These standards have been designed to safeguard drinking water by helping to ensure the material safety and performance of products that come into contact with drinking water. Certification organizations provide assurance that a product conforms to applicable standards and must be accredited by the Standards Council of Canada (SCC). In Canada, the following organizations have been accredited by the SCC to certify treatment devices and materials as meeting NSF/ANSI standards:
- Canadian Standards Association International;
- NSF International;
- Water Quality Association;
- Underwriters Laboratories Inc.;
- Quality Auditing Institute; and
- International Association of Plumbing & Mechanical Officials.
An up-to-date list of accredited certification organizations can be obtained from the ;SCC.
8.0 Kinetics and metabolism
In the following studies, and in those in subsequent sections, HAAs were administered as the free acid, as the sodium salt or as a neutralized solution, depending on the study methodology. The sodium salt and neutralized solution of the HAA both result in the salt being formed and are designated as, for example,
"MCA (sodium salt)." The free acid is designated as, for example,
"MCA (acid)." The form of HAA used in each study is noted, because the form can influence the effects seen in the test systems. When it is described in the following studies that the HAA was neutralized, it signifies that sodium hydroxide was the base used to adjust the pH. In the event that another base was used, this is noted in the description of the study.
8.1.1 Monochloroacetic acid
Experiments with buffered solutions of MCA at pH 7 across human skin using in vitro diffusion chambers failed to show evidence of significant dermal absorption (Xu et al., 2002). The authors stated that ionization may be the most significant factor limiting the permeability of HAAs.
ECETOC (1999) reported that MCA given orally to rats or mice was rapidly and extensively absorbed.
8.1.2 Dichloroacetic acid
DCA is also readily absorbed into the bloodstream from the gastrointestinal tract following oral exposure in both rats and humans (Stacpoole et al., 1987, 1998a; James et al., 1998; Schultz et al., 1999). Dermal absorption in humans is minor both in vivo (Kim and Weisel, 1998) and in vitro (using diffusion chambers with a buffered solution of DCA) (Xu et al., 2002;). DCA exists primarily as an ionic species in drinking and swimming pool waters that are kept within a neutral pH range (Kim and Weisel, 1998), which limits its dermal absorption (Xu et al., 2002;).
8.1.3 Trichloroacetic acid
TCA is readily absorbed from the gastrointestinal tract following oral exposure in both rats and humans (Kim and Weisel, 1998; Schultz et al., 1999). The concentration of TCA in blood in rats following oral ingestion peaked at approximately 2 hours post-dosing (Schultz et al., 1999). No evidence of significant dermal absorption was seen with TCA in humans in vivo (Kim and Weisel, 1998) or using diffusion chambers in vitro (Xu et al., 2002;).
8.1.4 Monobromoacetic acid
No specific studies measuring the absorption of MBA following different routes of exposure were undertaken; however, acute oral studies have shown that MBA is absorbed and causes adverse effects (see Section 10.1).
8.1.5 Dibromoacetic acid
DBA is rapidly absorbed into the bloodstream from the gastrointestinal tract following oral exposure in rats; the blood concentration peaked at approximately 1 hour post-dosing (Schultz et al., 1999). Schultz et al. (1999) estimated the oral bioavailability of DBA (using only a single high dose) at only 30%. They postulated that it was due to first-pass metabolism. No other doses were used to confirm this value.
Other short-term studies (Linder et al., 1994a, b, 1995, 1997b; Parrish et al., 1996; Cummings and Hedge, 1998; Vetter et al., 1998; NTP, 1999b) report effects on the liver, kidney, spleen and male reproductive system, demonstrating that DBA is sufficiently absorbed to have caused adverse effects.
No evidence of significant dermal absorption was seen with DBA at pH 7 across human skin using in vitro diffusion chambers (Xu et al., 2002;). The authors stated that ionization may be the most significant factor limiting the permeability of HAAs, including DBA.
8.2.1 Monochloroacetic acid
Two different pathways have been proposed for the breakdown of MCA in biological systems (ECETOC 1999):
- formation of S-carboxymethyl glutathione and subsequently S-carboxymethyl cysteine, which is then metabolized to thiodiacetic acid (main route); and
- formation of glycolic acid following hydrolysis of the C-Cl bond; subsequent oxidation leads to the formation of oxalic acid and carbon dioxide. Other metabolic pathways suggested are via dehalogenation to form oxalate and glycine
and/or dehalogenation and reduction to thiodiacetic acid via glutathione conjugation (Bhat et al., 1990). MCA (acid) has also been reported to bind to lipids (Yllner, 1971; Bhat and Ansari, 1989; Kaphalia et al., 1992).
8.2.2 Dichloroacetic acid
DCA's principal metabolic pathway occurs via oxidative dechlorination to form glyoxylate (Keys et al., 2004). Glyoxylate can be further biotransformed to oxalate (by oxidation), to glycine (by transamination) and subsequently to glycine conjugates such as serine and/or 5,10 methylene tetrahydrofolate, or to glycolate (by reduction); all of these metabolites are excreted in variable quantities in the urine (Stacpoole, 1989; James et al., 1998; Stacpoole et al., 1998a; U.S. EPA, 2003c). Some DCA is also converted to carbon dioxide and eliminated via exhaled air (James et al., 1998). DCA can also be metabolized through reductive dechlorination to form MCA and subsequently thiodiacetate (James et al., 1998).
The enzyme that initially catalyses the glutathione-dependent oxygenation of DCA has been identified as glutathione-S-transferase-zeta (GST-zeta) and is found primarily in the cytosol (Tong et al., 1998a, b). Appreciable differences in the metabolism of DCA exist between species.
The half-lives of DCA in mice and rats following oral dosing were 1.5 hours and 0.9 hour, respectively (Larson and Bull, 1992). Repeat dosing with DCA has also shown an increased plasma elimination half-life in both rats and humans (Anderson et al., 1999).
Toxicokinetics studies indicate that DCA is able to inhibit its own metabolism (also known as suicide inhibition) by irreversibly inactivating the GST-zeta enzyme (U.S. EPA, 2003c; Keys et al., (2004). Prior treatment with DCA has been shown to inhibit the metabolic clearance of subsequent doses of DCA in rats (James et al., 1998), mice (Schultz et al., 2002) and humans (Curry et al., 1985; Stacpoole et al., 1998a).
Species- and age-related differences in GST-zeta activity were observed. The relative rate of DCA transformation was greater in mice and rat cytosol than in human hepatic cytosol (Tong et al., 1998a). Reduced liver metabolism was seen in young mice, accompanied by a decrease in immunoreactive GST-zeta, whereas the levels of that protein remained unchanged in aged mice (Schultz et al., 2002).
Pharmacokinetic models, created to help estimate concentrations of DCA in the liver, may be useful to refine the tissue dose-response for liver tumours. However, these models are limited, as they can provide estimates of liver concentrations only where the metabolism is not inhibited or is at its maximum inhibition. Partial inhibition is difficult to model, since concentrations may vary depending on GST-zeta activity (U.S. EPA, 2003c).
Carcinogenic and genotoxic effects have been associated with high doses of DCA, where its metabolism is inhibited (U.S. EPA, 2003c).
However, as reported in the U.S. EPA, (2003c) Toxicological Review of DCA, there are still many unanswered questions regarding DCA's metabolism, including whether there is more than one metabolic pathway, and its relevance to toxicity in laboratory animals and humans.
8.2.3 Trichloroacetic acid
A relatively small proportion of TCA is metabolized in the liver. The formation of carbon dioxide, glyoxylic acid, oxalic acid, glycolic acid and DCA was observed in rats and mice following oral administration of radiolabelled TCA (neutralized). It was suggested that TCA was metabolized by reductive dehalogenation to DCA (Larson and Bull, 1992). Further reductive dehalogenation of DCA to MCA and ultimately to thiodiglycolate has been proposed as a metabolic pathway (Bull, 2000). However, other investigators have suggested that metabolism to DCA may have been over-reported in earlier studies due to analytical methodologies that convert TCA to DCA due to the presence of a reagent (Ketcha et al., 1996; Lash et al., 2000).
8.2.4 Monobromoacetic acid
As part of a larger metabolism study (Jones and Wells, 1981), a group of three male Sprague-Dawley rats was orally administered MBA (sodium salt), equivalent to 50 mg MBA/kg bw, and the urine was collected for 24 hours. Unchanged MBA was excreted in the urine within 24 hours, along with N-acetyl-S-(carboxymethyl)cysteine. No other details were provided in the study.
8.2.5 Dibromoacetic acid
An in vitro metabolism study conducted by Tong et al. (1998a) demonstrated that GST-zeta enzyme catalysed the oxygenation of DBA to glyoxylic acid, a pathway shared by DCA. WHO (2004c) reported that glyoxylic acid can be metabolized to glycine, glycolate, carbon dioxide or oxalic acid, based on a study by Stacpoole et al. (1998b).
8.3.1 Monochloroacetic acid
After oral administration of a single toxic dose (225 mg/kg bw) of MCA (acid) to rats, levels initially remained below those seen following administration of a subchronic dose (10 mg/kg bw), because most of the toxic dose was retained in the stomach for up to 8 hours (a spasm of the pyloric sphincter prohibited further flux for several hours) (Saghir and Rozman, 2003).
Kaphalia et al. (1992) found that the liver, kidney, intestine and spleen were the organs containing the highest MCA levels when it was orally administered (as an acid) to rats as a single dose.
8.3.2 Dichloroacetic acid
DCA, when administered via gavage, is distributed initially to the liver and muscle and subsequently to other target organs (Evans, 1982; James et al. (1998). James et al. (1998) administered a single radiolabelled oral dose of DCA (sodium salt) to young adult rats, and the dose distributed mostly to the muscles (11.9%), liver (6.19%), gastrointestinal tract (3.74%), fat (3.87%) and kidney (0.53%). Other tissues, including plasma, spleen, heart, skin, bone, brain, lung and testes, accounted for 9.5% of the administered dose. DCA also exhibited low binding to plasma when given intravenously (Schultz et al., 1999). Schultz et al. (1999) noted that the lipophilicity of DCA was low when measured at a pH value close to that of blood (pH 7.4), which indicates that DCA would not tend to accumulate in fat.
8.3.3 Trichloroacetic acid
Following oral and intravenous administration in rats, TCA appears to bind significantly to plasma proteins and is also distributed to the liver (Templin et al., 1993; Schultz et al., 1999; Yu et al., 2000). Because of the significant binding to plasma, only the free TCA is available to tissues for uptake and elimination (Yu et al., 2000). Plasma protein binding has been found to vary across species and is highest in humans (Lumpkin et al., 2003).
8.3.4 Monobromoacetic acid
8.3.5 Dibromoacetic acid
DBA (acid) was detected in the plasma of male and female Sprague-Dawley rats following exposure via deionized drinking water in a range-finding reproductive/developmental study (Christian et al., 2001). DBA was not detected in the plasma of female B6C3F1 mice administered DBA in drinking water for 28 days (NTP, 1999b). This may be due to extensive metabolism and excretion and not to limited absorption (U.S. EPA, 2005a). Detectable and quantifiable levels of DBA were also found in the placenta, amniotic fluid and milk (Christian et al., 2001). According to Christian et al. (2001), no apparent accumulation of DBA was observed. DBA was also detected in testicular interstitial fluid when male Sprague-Dawley rats (number not given) were gavaged for 5 days with DBA (neutralized) at 250 mg/kg bw (Holmes et al., 2001). The concentration of DBA peaked at 30 minutes and exhibited a half-life of 1.5 hours.
The lipophilicity of DBA was low when measured at a pH value close to that of blood (pH 7.4), indicating that DBA would not tend to accumulate in fat. DBA also exhibited low binding to plasma when given intravenously (Schultz et al., 1999).
8.4.1 Monochloroacetic acid
Urination is reported as the major route of MCA elimination in rats when dosed orally or dermally with MCA (acid) (Saghir and Rozman, 2003). Approximately 90% of a single oral dose of MCA (acid) given to rats was excreted in the urine within 24 hours (Kaphalia et al., 1992).
8.4.2 Dichloroacetic acid
DCA is mostly eliminated either unchanged or by metabolic transformation primarily in expired air or in the urine. In the urine of rodents, the amount eliminated as unmetabolized DCA or as metabolites varies according to the dose. At low doses, DCA is almost completely eliminated in the urine as metabolites, while a higher percentage of unmetabolized DCA was seen using higher or repeat doses of DCA (Lukas et al., 1980; Lin et al., 1993; Gonzalez-Leon et al., 1997; Cornett et al., 1999), possibly due to the inhibition of its metabolism. In rodents and humans, variable levels of metabolites are found in the urine.
DCA is also eliminated from the lungs as carbon dioxide, but the levels may differ between species. Studies with rats and mice showed that carbon dioxide represented 24-30% and 2-45% of the total dose, respectively (Larson and Bull, 1992; Lin et al., 1993; Xu et al., 1995). Less than 2% of DCA was recovered in the faeces in animal studies (Larson and Bull, 1992; Lin et al., 1993).
DCA is a metabolite of TCE in humans and has been detected in the seminal fluid of some workers exposed to TCE (Forkert et al., 2003).
8.4.3 Trichloroacetic acid
In a limited metabolism study, three volunteers ingested TCA (sodium salt) at a dose of 3 mg/kg bw; the elimination half-life from the blood was 50.6 hours (Muller et al., 1974). In a longitudinal exposure pilot study, Bader et al. (2005) measured the elimination half-life of TCA in the urine of five volunteers. These volunteers were provided tap water (concentrations of TCA ranged from 50 to 180 µg/L) for the first 2 weeks and then TCA-free bottled water for the last 2 weeks. Individual TCA urinary elimination rates ranged from 2.1 to 6.3 days. TCA appears to persist for several days when a steady state is almost reached within the plasma, therefore reflecting average exposure over several days. The authors inferred that TCA in plasma may be a viable biomarker for drinking water exposure.
Since TCA is one of the major metabolites of TCE and PCE in humans (Monster et al., 1979; IARC, 1995; ACGIH, 2001; Forkert et al., 2003), several metabolism studies looked at the half-life of the parent compounds as well as their metabolites, such as TCA, and were included in this review.
After volunteers inhaled 50-100 ppm TCE (6 hours/day for 5-10 days), the elimination half-life of TCA from the blood ranged from 85 to 99 hours, a higher value than that obtained following ingestion of TCA (sodium salt) (Muller et al., 1974).
Allen and Fisher (1993) developed a physiologically based pharmacokinetic model for humans exposed to TCE with emphasis on the metabolite, TCA, and compared it with findings in mice and rats (Allen and Fisher 1993). The volume of TCA distribution in humans was found to be lower than in rats or mice. The model estimated that, in humans, 93% of total TCA eliminated was excreted unchanged in the urine, while the remainder may be metabolized or eliminated by other routes (Allen and Fisher 1993).
Another inhalation study looked at the comparative excretion rates of inhaled PCE and its metabolites in rats and humans (Volkel et al., 1998). The mean elimination half-life of TCA in the urine was 45.6 hours in humans, compared with 11.0 hours in rats, suggesting that elimination of TCA in rats is more rapid and may be due to differences in PCE metabolism (Volkel et al., 1998).
TCA, being one of the major metabolites of TCE and PCE in humans, has been used as a biomarker of occupational exposure to these chemicals (Monster et al., 1979; IARC, 1995; ACGIH, 2001; Forkert et al., 2003). TCA is also a metabolite of 1,1,1-trichloroethane, 1,1,2,2 tetrachloroethane and chloral hydrate (IARC, 1995).
8.4.4 Monobromoacetic acid
No studies on MBA excretion were identified.
8.4.5 Dibromoacetic acid
Only one study on DBA excretion was located. Schultz et al. (1999) exposed rats intravenously to a single dose of 109 mg DBA/kg bw, and the elimination half-life was calculated at 0.72 hour. The major route of elimination was believed to be via biotransformation. The urine and faeces were very minor contributors to overall blood clearance when DBA was given intravenously to rats, with the urine representing less than 1% of total clearance and the amount in the faeces being negligible (Schultz et al., 1999).
9.0 Health effects in humans
A limited population-based case-control study conducted in Nova Scotia and eastern Ontario did not find an association between HAA exposures and stillbirth risk when controlling for total THM exposures (King et al., 2005). The analysis included 112 stillbirth cases and 398 live birth controls that occurred between 1999 and 2001. No other specific information was available on the teratogenic, reproductive or embryotoxic effects of chlorinated HAAs in humans (CHEMINFO, 2003a, b, c, d, e). Epidemiological studies have been conducted with CDBPs to determine if exposure contributes to reproductive and developmental effects in humans (Bove et al., 1995; Mills et al., 1998). No other information was available to assess the effects of individual HAAs, nor have HAAs been satisfactorily dissociated from the many other by-products (hundreds, if not thousands) that are produced by chlorination. Therefore, it is difficult to infer causality between specific HAAs and adverse reproductive and developmental health outcomes in humans based on one limited study.
Epidemiological studies on the incidence of cancer have been conducted with CDBPs, but not specifically with chlorinated HAAs. Cohort and descriptive studies have also been conducted with TCE and PCE. Biological monitoring of exposure to both these compounds was done by measuring the urinary levels of DCA or TCA, which have been identified as metabolites of these two compounds (TCE and PCE) in humans (IARC, 1995). TCA has also been identified as a metabolite of other chlorine-containing ethanes and ethylenes (ACGIH, 2001).
Most chlorinated HAAs are found in drinking water (where the pH range of 6-9 is close to neutral) almost exclusively in the ionized form (anion) due to their very low pKa values (IPCS, 2000). HAAs are found in drinking water at very low concentrations, which means that the dilution of HAAs in drinking water and the buffer capacity of drinking water would mostly counter the irritative properties described below for concentrated acid solutions of HAAs.
9.1 Monochloroacetic acid
The probable lethal oral dose for MCA (acid) in humans is in the range of 50-500 mg/kg bw (Gosselin et al., 1984). In one fatal case, a 5-year-old girl ingested a teaspoon (5-6 mL) of a wart remover composed of 80% MCA (acid); she began vomiting, collapsed and died 8 hours later. The cause of her death was attributed to metabolic acidosis and cardiac arrhythmia. The autopsy revealed that the liver was damaged, and the stomach had signs of marked irritation (Feldhaus et al., 1993; Rogers, 1995).
The sodium salt of MCA is not corrosive to the skin and has limited skin absorption (ECETOC, 2001). MCA does not easily form a vapour (CHEMINFO, 2003 a, b). No cases of acute intoxication from MCA (sodium salt) have been located in the literature (ECETOC, 2001).
Three adult volunteers who drank 300 mL of a 0.05% aqueous solution of MCA (acid) daily for a period of 60 days (corresponding to 2 mg/kg bw per day) experienced no adverse health effects (Morrison and Leake, 1941; NTP, 1992). Dilute solutions of MCA (acid) (i.e., up to 1% aqueous) failed to produce irritation of human skin (Morrison and Leake, 1941).
9.2 Dichloroacetic acid
Patients diagnosed with genetic disorders, such as familial hypercholesterolaemia (a common disorder of lipid metabolism associated with a high risk of early mortality from coronary artery disease) or various mitochondrial disorders and, as a result, treated daily with DCA (sodium salt) for periods longer than 4 months were found to develop peripheral neuropathy (loss of reflexes and muscle weakness) and in one case hepatomegaly (enlarged liver) (Moore et al., 1979; Spruijt et al., 2001; Izumi et al., 2003).
Humans exposed in an occupational setting to concentrated DCA mist or vapour may develop pulmonary oedema several hours after exposure (CHEMINFO, 2003c).
9.3 Trichloroacetic acid
In an occupational setting, concentrated solutions and/or mists of TCA (acid) have caused skin effects (redness, swelling, pain and possibly burns), eye effects (severe irritation or corrosive injury) or damage to the gastrointestinal tract due to accidental ingestion (CHEMINFO, 2003d, e). According to CHEMINFO (2003d, e), the severity of the injury (dermal, oral, ocular) increases with the concentration of the acid and the duration of exposure.
9.4 Monobromoacetic acid
No studies reporting human health effects from exposure to MBA were reported.
9.5 Dibromoacetic acid
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