Page 3: Guidelines for Canadian Drinking Water Quality: Guideline Technical Document – 2-Methyl-4-chlorophenoxyacetic Acid (MCPA)

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Part II. Science and Technical Considerations

4.0 Identity, use, sources and fate in the environment

MCPA is a phenoxyacetic acid herbicide, with various formulations: as the free acid, as a dimethylamine salt (MCPA-DMAS), as a sodium salt (MCPA sodium salt) and as an ester (MCPA 2-ethylhexyl ester, or MCPA-EHE) (U.S. EPA, 2004b,d). Although MCPA may be applied in various forms, a single common functional group (the parent acid) is the active portion of the herbicide formulation (U.S. EPA, 2004c,e; PMRA, 2006). Diethanolamine MCPA, another amine salt formulation, is not discussed in this document, as there is no information available on its toxicology and Health Canada's Pest Management Regulatory Agency (PMRA) is proposing that it be phased out (PMRA, 2006).

MCPA acts as a plant growth regulator and is used to control broadleaf weeds in post-emergence in agriculture and in urban, forestry and aquatic environments (IARC, 1983; Weed Science Society of America, 1989; HSDB, 2003). MCPA is absorbed through both leaves and roots and is translocated throughout the plant. By stimulating nucleic acid and protein synthesis, MCPA affects enzyme activities, respiration and cell division. As a result, treated plants exhibit malformed leaves, stems and roots (U.S. EPA, 1990).

Those physicochemical properties of MCPA acid and its related forms that are relevant to their behaviour in the environment are summarized in Table 1.

The physicochemical properties of these other forms of MCPA vary widely according to the formulation. In general, the salts are water soluble, whereas the esters are lipophilic and less water soluble (U.S. EPA, 2004d,e). Based on its Henry's law constant, MCPA is not expected to volatilize from water or moist surfaces. Its vapour pressure also indicates a low potential to volatilize, and its dissociation constant indicates that it will dissociate rapidly at environmental pH (PMRA, 2006). As available environmental fate data show that all forms of MCPA will revert to MCPA acid, the physicochemical characteristics of MCPA in the environment will be those associated with MCPA acid (U.S. EPA, 2004a).

It should be noted that MCPA is usually applied along with other phenoxy herbicides, such as 2,4-dichlorophenoxyacetic acid (2,4-D), 4-(2,4-dichlorophenoxy)butyric acid (2,4-DB), (R)(+)-2-(4-chloro-2-methylphenoxy)propanoic acid (also known as mecoprop-p or MCPP-p) and 4-(2-methyl-4-chlorophenoxy)butyric acid (MCPB) (U.S. EPA, 2004b).

MCPA is registered for use in Canada for agricultural sites, for fine turf (parks, playgrounds, golf courses, zoos, botanical gardens and athletic playing fields), for lawns (residences, public and commercial buildings) and sod (grown in sod farms harvested for transplanting), in forestry (spruce seedlings for reforestation) and at industrial sites (vegetation control) (PMRA, 2005a). In Canada, the major applications of MCPA are in agriculture, where it is used in the production of cereals (barley, oats, rye, wheat and flax) and in the production of canary seeds, legumes and grasses, asparagus, field corn and sweet corn. MCPA is also applied to stubble/summerfallow fields and pastures/rangeland (PMRA, 2005a). MCPA is used everywhere in Canada, particularly in the Prairies (PMRA, 2005a).

MCPA is among the top 10 pesticides sold in Canada. Table 2 presents the limited data available on the sales/use of MCPA from recent years. Brimble et al. (2005) reported that Saskatchewan is responsible for approximately 36% of all pesticide sales in Canada, making it the largest pesticide user in Canada. Although sales data for Saskatchewan are not currently available, a database to collect this information is under development. In Quebec, legislation bans the application of MCPA (and other specific pesticides) around all daycare facilities, schools, some public lands and private green spaces (Gouvernement du Québec, 2006).

4.1 Environmental fate

4.1.1 Soil

MCPA is not persistent in soil (U.S. EPA, 2004a), with a half-life varying between 15 and 50 days (Soderquist and Crosby, 1975; Sattar and Paasivirta, 1980). The rate of degradation depends upon several factors, such as soil type, soil pH, soil moisture, concentration of MCPA, climatic conditions and organic matter content (Sattar and Paasivirta, 1980). Degradation of MCPA occurred within 5-9 weeks in acidic soil compared with 1 week in neutral soil (pH 6.3 and above) (Sattar and Paasivirta, 1980).

Microbial degradation is the most important transformation process for MCPA in soil (Caux et al., 1995); the presence of both oxygen and proper moisture is important for its biodegradation (Sattar and Paasivirta, 1980). In the absence of oxygen, the biotransformation of MCPA in soil is negligible (PMRA, 2006). Photodecomposition and hydrolysis in soil are not important degradation processes for MCPA (U.S. EPA, 2004a,b; PMRA, 2006). When MCPA was applied to soil surfaces and irradiated with natural sunlight, its calculated half-life was 67 days, demonstrating that MCPA photodegraded very slowly (U.S. EPA, 2004a,b).

MCPA was shown to be mobile in soil in laboratory studies (U.S. EPA, 2004a). However, field studies indicate that MCPA does not leach appreciably below soil depths of 15 cm (PMRA, 2007). Its mobility seems to be related to the soil's organic matter content, increasing as the organic matter content decreases (WHO, 2003). Surface water may be contaminated via spray drift and runoff, whereas groundwater can be potentially contaminated because MCPA is mobile in soil and, as a result, may have a potential for leaching (U.S. EPA, 2004a). The mobility and leaching of the non-acid forms of MCPA (i.e., amine and sodium salts, esters) have not been determined.

Field studies with MCPA-EHE have shown that, under normal conditions, a large proportion converts to MCPA acid on the day of application, and the MCPA-EHE is nearly completely converted by day 3 (U.S. EPA, 2004a). However, under dry conditions, MCPA-EHE has been found to persist for days, with greater than 90% present after 48 hours (Smith and Hayden, 1980). In a non-sterile soil:calcium chloride system at pH 5.6 and pH 6.8, MCPA-EHE adsorbed to the soil particles but was available for degradation to MCPA, with a half-life of ≤ 12 hours (U.S. EPA, 2004a).

4.1.2 Water

Agricultural herbicides such as MCPA have been detected in lakes, rivers and reservoirs (dugouts), which may serve as sources for drinking water. Their migration from soil to water is the result of direct or indirect transport mechanisms, including non-target drift from aerial or ground boom spraying (vaporization in air), deposition in rain, erosion of soil particles by wind or water, surface runoff and leaching. In specific situations, MCPA can be found in water as a result of spills, deliberate dumping of tank residues or equipment washing operations (Caux et al., 1995; Murray et al., 2004). MCPA is not expected to volatilize from water based on its vapour pressure (8.18 × 10^-5 to 1.36 × 10^-4 Pa) and its Henry's law constant (7.46 × 10^-5 Pa·cm³/mol). By contrast, MCPA-EHE is expected to volatilize from water based on its low to intermediate vapour pressure (3.43 × 10^-4 to 1.63 × 10^-3 Pa) and its Henry's law constant (2.56 Pa·m³/mol) (PMRA, 2007).

In water, biological degradation (under aerobic conditions) is an important process affecting MCPA's environmental fate (Soderquist and Crosby, 1975; Sattar and Paasivirta, 1980; Smith and Hayden, 1981; PMRA, 2006). MCPA in rice paddy water under dark conditions is totally degraded by aquatic microorganisms in 13 days (Soderquist and Crosby, 1975). However, in anaerobic aquatic systems (sediment/water), the biodegradation of MCPA was negligible (PMRA, 2006). Hydrolysis and photodegradation are not important routes in the degradation of MCPA in water (PMRA, 2006). In an aqueous solution (pH 8.3), MCPA had a photolytic half-life of 20-24 days in sunlight. Photolysis of dilute aqueous MCPA solutions yielded 4-chloro-2-methylphenol as the major by-product; o-cresol and 4-chloro-2-formylphenol were also identified (Soderquist and Crosby, 1975). During periods of cold weather and low light, the degradation of MCPA via biological degradation or photolysis is rather limited (Byrtus et al., 2004). MCPA acid does not readily hydrolyse in sterile buffer solution at pH 5-9 (U.S. EPA, 2004a). MCPA has not been shown to bind significantly to sediments (Caux et al., 1995).

The derivatives of MCPA have been shown to dissociate in water to MCPA acid. In deionized water, MCPA-DMAS completely dissociated to MCPA acid and dimethylammonium ion within 1.5 min (U.S. EPA, 2004a). In sterile buffers, the hydrolysis of MCPA-EHE to MCPA acid was pH dependent (half-life < 117 hours at pH 9, but no hydrolysis at pH 5 and pH 7).

4.1.3 Atmosphere

Pesticides can enter the atmosphere via application drift, evaporation, sublimation or erosion of treated soil. Once in the air, pesticides can be transported and redistributed, degraded or returned to the earth's surface via wet deposition (dissolved in rain) and dry deposition (as gases or aerosols or sorbed to solid macro-particles) (Hill et al., 2002; Waite et al., 2005). However, little information on the fate of MCPA is available. Waite et al. (2005) demonstrated, by high-volume air sampling at various heights above ground level in the Canadian Prairies, that atmospheric concentrations of MCPA are strongly influenced by regional atmospheric transport and that its primary transport mechanism is via adsorption to solid particles in the atmosphere. The volatilization of MCPA in the acid or salt form is low (Weed Science Society of America, 1989; Caux et al., 1995). The half-life of MCPA due to photooxidation was estimated to be 2.2 days (Caux et al., 1995).

5.0 Exposure

Canadians can be exposed to MCPA through its presence in drinking water, air and food. In addition, certain segments of the population can be exposed through the use of specific consumer products or in occupational settings related to pesticide use and application. Although some exposure data are available, they are considered insufficient to justify modifying the default allocation factor for drinking water (i.e., the proportion of the total daily intake of MCPA allocated to drinking water) of 20%.

5.1 Air

During 1989 and 1990, MCPA was detected in only 4% of atmospheric samples at two dugout sites near Regina, Saskatchewan, where it was being applied locally, with a maximum concentration of 0.39 ng/m³ (Waite et al., 2004). It was detected in 24% of bulk (dry plus wet) atmospheric deposits at these same sites, with a mean deposition rate of 129 ng/m³ per day.

In a small southern Manitoba watershed, MCPA was detected at concentrations up to 13 ng/m³ in atmospheric samples between 1993 and 1996 (Rawn et al., 1999).

In a province-wide survey of herbicide concentrations in rainfall conducted in Alberta between April and September (1999-2000), MCPA was detected in less than 50% of samples taken (Hill et al., 2002). Among the major pesticides, MCPA was detected in the second highest total amount, with maximum concentrations ranging from 38 to 84 µg/m² (based on an amount basis). The total amount of MCPA deposited in rainfall varied over the season, with levels highest during June and July (spraying season), and varied with the region sampled: the highest levels were observed in central rural Alberta, the second highest in southern rural Alberta and the lowest in remote regions and city locations.

In Alberta in 2000, MCPA was detected in 7 of 68 samples (10%) at four sites across Alberta in ambient air samples (Kumar, 2001). Concentrations ranged from 0.03 to 0.46 ng/m³. MCPA was detected at relatively low concentrations and detection frequencies compared with other pesticides that had higher volatility.

5.2 Water

5.2.1 Drinking water

In Alberta, MCPA was seldom detected (11%) in a treated water survey program during the period 1995-2003 (Byrtus et al., 2004). MCPA was detected in 190 of a total of 1788 water samples collected for analysis throughout this period. MCPA was detected more at treatment facilities with surface water sources (13.7%) than at facilities with groundwater sources (2%); the maximum concentrations detected were 571 ng/L and 5 ng/L, respectively. The data showed no overall or seasonal trend over the 9 years in detection frequency or concentration, suggesting that MCPA could be detected in treated water supplies all year round.

MCPA was detected in 2 of 403 samples of municipal drinking water in surveys conducted in Manitoba from 1981 to 1999 (Manitoba Conservation, 2000). The maximum concentration found was 49 µg/L.

In various towns in Ontario, MCPA was detected in 4 of 29 raw water samples for 2004, at concentrations ranging from 27 to 49 ng/L, and in 2 of 132 treated water samples for the year 2004 and January 2005, at concentrations ranging from 27 to 66 ng/L (Ontario Ministry of the Environment, 2005). The detection limit was 20 ng/L.

In Nova Scotia, MCPA was not detected in raw or treated water from Dartmouth (detection limit 2 µg/L) or Stellarton (effective quantification limit 50 µg/L) during sampling done in June 2003 (Nova Scotia Department of Environment and Labour, 2005).

In Newfoundland and Labrador, MCPA was not detected in two wells sampled in 1982 (detection limit 2 µg/L) (Newfoundland and Labrador Department of Environment and Conservation, 2005).

MCPA was not detected in any of the 10 groundwater samples taken in Prince Edward Island in 1998 (Blundell and Harman, 2000).

5.2.2 Potential sources of drinking water

In British Columbia, MCPA was detected in 1 of 13 samples of runoff in the Lower Fraser Valley in 2003. The maximum concentration was 110 ng/L (Tuominen et al., 2005).

In Alberta, MCPA was detected in 1150 of 3062 samples (37.6%) collected from surface waters (rivers, streams, lakes, etc.) between 1995 and 2002 (Anderson, 2005). Concentrations ranged up to 8.5 µg/L, with the median concentration at 0.02 µg/L (detection limit 0.005 µg/L).

In Saskatchewan, runoff water from flood-irrigated fields enters a system of drainage ditches, which then drain into the South Saskatchewan River, potentially affecting downstream water quality. The concentrations of MCPA and other herbicides in the drainage water from two drainage ditches in an irrigation district that emptied into the South Saskatchewan River were sampled during the 1994-1996 growing seasons (Cessna et al., 2001). The maximum concentration of MCPA in the drainage water was 2.8 µg/L. Input of MCPA into the river was approximately 0.06% of the amounts applied; in all 3 years, MCPA was applied to flood-irrigated land irrigated by both drainage ditches. No significant increase in herbicide concentration was seen in the river as a result of herbicide input, because of the large flows seen in the river.

In a preliminary 2003 spring/summer pesticide surveillance study, MCPA was detected in all of 15 small reservoirs (dugouts) sampled in the three Prairie provinces (Murray et al., 2004). Mean MCPA concentrations in most of the dugouts ranged from 13 to 108 ng/L. The detection limit was 0.58 ng/L. Two reservoirs had higher levels, ranging from 200 to 320 ng/L, possibly a result of commercial formulations (which include MCPA) being applied in the watershed or catchment (Murray et al., 2004).

MCPA was detected in 10 of 49 samples of surface water taken from three of the Great Lakes--Lake Ontario (1998-1999), Lake Erie (1994, 1998, 2000) and Lake Huron (1999-2000)--with a maximum concentration of 5.4 ng/L, whereas it was not detected in nine samples from Lake Superior (1996-1997) (Struger et al., 2004). The detection limit was between 0.27 and 0.30 ng/L. Spatial and seasonal variations observed were related to MCPA's use patterns.

In sampling around the Great Lakes Areas of Concern in 2004, MCPA was detected in 127 out of 217 samples, with a maximum concentration of 350 ng/L (Struger et al., 2005). The detection limit was 5.8 ng/L. MCPA was detected in 120 out of 171 samples of surface water around the Great Lakes Areas of Concern and small streams in the Niagara and Burlington areas in 2003; the maximum concentration was 1.23 µg/L, with a mean of 0.0139 µg/L. MCPA was detected in 27 out of 59 samples of surface water in Great Lakes Areas of Concern and connecting channels in 2002, with a maximum concentration of 40 ng/L (mean 2.9 ng/L). In Lake Huron tributaries sampled in 2002, MCPA was detected in 11 out of 47 samples, with a maximum concentration of 80 ng/L (Struger et al., 2005). MCPA was not detected in all 133 samples taken in 1998-2000 in the area of the Don and Humber river watersheds (detection limit 100 ng/L) (Struger et al., 2002).

In Quebec, MCPA was seldom detected (11%) during the summer of 2003 in the tributaries of the St. Lawrence River; the maximum concentration seen was 120 ng/L (Murray et al., 2004). These tributaries drain from agricultural areas where MCPA is used in the cultivation of cereals, vegetables and grassland.

MCPA was detected in 37% of 481 samples of surface water taken from four rivers in the corn and soy growing areas of Quebec between 2002 and 2004; the maximum concentration found was 4.9 µg/L, and the detection limit was 0.01 µg/L (Giroux et al., 2006). This same percentage (37% of 1447 samples) was also seen in these same four rivers during previous sampling years (1993-2001); the maximum concentration found was slightly higher, 7.0 µg/L, and the detection limit varied from 0.02 to 0.05 µg/L (Giroux et al., 1997; Giroux, 1999, 2002). MCPA was also detected in 69 out of 122 samples from five other surface waters in agricultural areas in Quebec surveyed in 1996 and 1997, with a maximum concentration of 2.6 µg/L; the detection limit was 0.05 µg/L (Giroux, 1998).

In the Maritime provinces, MCPA was not detected in any of the 60 surface water samples taken in New Brunswick (2003 and 2004), Nova Scotia (2004) and Prince Edward Island (2003 and 2004); the detection limit was 1 µg/L (Murphy and Mutch, 2005). Nor was MCPA detected in the rivers of eight different municipalities with agricultural or urban activities in New Brunswick in 2004 (New Brunswick Department of Health, 2005).

5.3 Food

There are limited Canadian data available on the concentration of MCPA in food. Between 1980 and 1985, 354 composite vegetable samples representing nine vegetable commodities were collected from Ontario farm deliveries. MCPA was detected in one of six samples of asparagus, at a concentration of 0.06 mg/kg, below the maximum residue limit (MRL) of 0.1 mg/kg (Frank et al., 1987a). Monitoring studies in 1980-1982 in Ontario found MCPA levels of 0.01 mg/kg in peaches and 0.02 mg/kg in strawberries, both below the MRL (Frank et al., 1987b).

No current U.S. data were identified. Historically, during monitoring of pesticides in foods in the Total Diet Study program, in which ready-to-eat foods were sampled (June 1964-April 1969), MCPA was found very infrequently at low levels (maximum concentration 0.4 mg/kg) in products such as dairy products, grains, cereals and sugars (Bovey, 1980a). MCPA was not detected in any food composites collected from June 1969 to 1976 (IARC, 1983; HSDB, 2003).

MCPA-DMAS was not detected in grain 56 days after application to wheat treated in the tillering stage (Chow et al., 1971).

6.0 Analytical methods

The only U.S. Environmental Protection Agency (EPA)-approved method for measuring MCPA in drinking water is EPA Method 555 (Rev. 1.0). This method employs high-performance liquid chromatography (HPLC) with a photodiode array ultraviolet detector (PDA-UV). Samples are hydrolysed, acidified, filtered and extracted prior to separation and analysis using HPLC with a PDA-UV. The method detection limit is 0.8 µg/L (U.S. EPA, 1992).

The World Health Organization (WHO) recognizes liquid-liquid extraction with dichloromethane and derivatization with diazomethane using gas chromatography (GC) with either mass spectrometry (MS) or an electron capture detector. The method detection limit is approximately 0.1 µg/L (WHO, 2003). It is considered a very common method of analysis for MCPA (Hernandez et al., 1993; Tekel and Kavacicová, 1993).

The Centre d'expertise en analyse environnementale du Québec (CEAEQ, 1999) reported a method (MA.403-P.Chlp 2.0) for measuring MCPA in drinking water with a detection limit of 0.01 µg/L. The method employs solid-phase extraction (SPE) of an acidified sample on a C-18 column, followed by esterification and purification of the dried extract. The sample is analysed using a GC system equipped with a mass spectrometer (GC-MS).

7.0 Treatment technology

7.1 Municipal scale

Treatment technologies reported to be effective for the reduction of MCPA in drinking water include activated carbon adsorption and/or ozonation, membrane filtration, UV irradiation and advanced oxidation processes (AOPs). Full-scale and pilot-scale studies have demonstrated that effluent concentrations of MCPA well below the MAC of 0.1 mg/L (100 µg/L) are achievable using the methods described below. The selection of an appropriate treatment process for a specific water supply will depend on many factors, including the characteristics of the raw water supply and the operational conditions of the specific treatment method.

7.1.1 Conventional treatment

Limited scientific literature has been published on the effectiveness of conventional surface water treatment techniques for the reduction of MCPA in drinking water. Conventional treatment techniques are reported to be ineffective in reducing the concentration of a variety of classes of pesticides in drinking water, in particular hydrophilic or lipophobic compounds (Robeck et al., 1965; Miltner et al., 1989; Foster et al., 1991, 1992; Haiste-Gulde et al., 1993). The concentrations of organic compounds, such as pesticides, may be reduced through coagulation/flocculation if they are hydrophobic or have low molecular weight acidic functional groups (Randtke, 1988). The chemical properties of MCPA (moderately lipophilic; substituted acetic acid) may result in limited removal by conventional water treatment.

Operational data collected from a conventional water treatment process demonstrated a 50% reduction of MCPA to levels below 0.033 µg/L (Byrtus et al., 2004). Foster et al. (1992) observed some reduction of low levels of chlorophenoxy acid in a conventional surface water treatment plant. Similarly, Lambert and Graham (1995) reported a reduction of influent MCPA concentration (level not provided) to below 1 µg/L using slow sand filtration.

7.1.2 Adsorption

Activated carbon adsorption has been widely used to reduce the concentration of pesticides in drinking water and is recognized by WHO as a suitable treatment technology for the removal of MCPA to levels below 0.1 µg/L (Robeck et al., 1965; Miltner et al., 1989; Foster et al., 1992; Duguet et al., 1994; WHO, 2004). The capacity of activated carbon to remove pesticides by adsorption is affected by a variety of factors, such as concentration, pH, competition from other contaminants or natural organic matter, contact time and the physicochemical properties of the pesticide. The effectiveness of granular activated carbon (GAC) filtration is also a function of the empty bed contact time (EBCT), flow rate and filter run time. The effectiveness of powdered activated carbon is also a function of the carbon dosage.

Operational data from a municipal-scale treatment plant using conventional treatment with a GAC filter-adsorber demonstrated that this type of treatment system can reduce low influent MCPA levels of 0.47 µg/L to below 0.02 µg/L (Frick and Dalton, 2005). No information was provided on the operational conditions of the GAC adsorber used in this study. Pilot-scale study results indicated that post-filtration GAC adsorbers can reduce influent MCPA levels of 2 µg/L to below 0.1 µg/L. No breakthrough of MCPA was observed over a period greater than 400 days using an EBCT of 20 minutes (Schippers et al., 2004). Other pilot-scale studies also found that GAC filtration was capable of reducing influent MCPA concentrations of 1 µg/L to below 0.1 µg/L; however, the bed life was not determined, as MCPA was only periodically present in the water supply (Hart, 1989; Hart and Chambers, 1991).

7.1.3 Ozonation and ozonation/adsorption

WHO (2004) also recognizes ozonation as a suitable treatment technology for the reduction of MCPA to levels below 0.1 µg/L. Bonné et al. (2000) observed a 65% reduction of MCPA from influent levels of 20 µg/L to below 7 µg/L using ozonation alone and a reduction to effluent levels below 0.1 µg/L using combined ozonation and biological activated carbon filtration treatment. Additional pilot-scale studies using ozonation followed by GAC found 100% removal at ozone doses of 2-4 mg/L and contact times of less than 15 minutes (Foster et al., 1992). The potential presence of aliphatic acids, aldehydes, ketones and other ozonation by-products such as bromate and assimilable organic carbon (AOC) may require additional treatment following ozonation and/or process optimization to minimize by-product formation. Camel and Bermond (1998) reported that mineralization of pesticides using ozonation or other oxidation techniques is generally incomplete; as a result, water treatment process design using this method should include sand or GAC filtration following oxidation. It has also been suggested that the use of ozonation prior to GAC adsorption increases the biological activity of the filter bed and may extend the bed life for pesticide removal (Lambert and Graham, 1995).

7.1.4 Ultraviolet

Pilot-scale studies evaluating the effectiveness of UV treatment for the removal of pesticides from groundwater achieved a removal efficiency of 94% to effluent levels below 0.1 µg/L for MCPA with UV energy inputs of 300 Wh/m³. Removal efficiency was found to be dependent on the energy dose and was not affected by the initial pesticide concentration. In addition, it was reported that the formation of by-products from the UV irradiation of MCPA may include organics as well as nitrite (Bourgine et al., 1997). Bench-scale experiments examining the effectiveness of UV irradiation observed complete degradation of MCPA using a low-pressure lamp and a wavelength of 254 nm. Removals of 90% and 100% of the initial reaction concentrations (50 mg/L) of MCPA following 20 and 40 minutes of reaction time, respectively, were observed (Benitez et al., 2004).

7.1.5 Ultraviolet and hydrogen peroxide oxidation

Full-scale AOPs such as UV and hydrogen peroxide oxidation treatment have been assessed for the reduction of different classes of pesticides (Kruithof et al., 2002a,b). Although MCPA was not studied specifically, other chlorophenoxy herbicides were studied. Results indicated that a variety of UV doses (< 1200 mJ/cm²) and hydrogen peroxide doses (< 15 g/m³) resulted in an 80% or greater degradation of several chlorophenoxy herbicides (e.g., 2,4-D) to effluent levels below 0.1 µg/L. Formation of bromate and metabolites was reported to be insignificant; however, GAC filtration was required following UV/hydrogen peroxide oxidation to remove AOC and residual hydrogen peroxide. This study also noted that the UV doses required for oxidation of pesticides were much higher than those needed for disinfection (Kruithof et al., 2002a,b).

Laboratory-scale studies found that the degradation of MCPA is accelerated using UV/hydrogen peroxide when compared with UV treatment alone. Within 12 minutes, 90% degradation was observed; however, initial concentrations were relatively high (50 mg/L), and the results did not include information on the final effluent levels (Benitez et al., 2004).

7.1.6 Membrane filtration

Various membrane filtration techniques, including ultrafiltration, nanofiltration and reverse osmosis, have been shown to be effective methods for the reduction of MCPA in drinking water (Taylor et al., 1995, 2000; Hofman et al., 1997; Schippers et al., 2004). Pilot-scale studies of ultrafiltration membranes have reported average MCPA removals of ≥ 99% to effluent levels below 0.1 µg/L (Schippers et al., 2004). Similarly, pilot-scale studies of reverse osmosis membranes demonstrated average MCPA removals ranging from 97% to 99%, resulting in effluent concentrations below 0.2 µg/L. Results of this study also indicated that pesticide removal did not decrease with increasing membrane age (Bonné et al., 2000). Taylor et al. (2000) conducted reverse osmosis pilot-scale studies to evaluate various factors affecting pesticide removal. An average MCPA removal of 97.5% was observed under varying flux, recovery and source water quality conditions, to produce effluent concentrations below 0.3 µg/L. Pesticide rejection was found to be dependent on a variety of system components, including type of membrane system, flux, recovery, pesticide solubility, charge and molecular weight. Results from this study also indicated that representative influent water quality and bench- or pilot-scale membrane units should be used for the selection of membranes for full-scale application.

7.1.7 Emerging treatment technologies

Several drinking water treatment technologies for MCPA are being developed but are still primarily in the experimental stage or do not have published information on the effectiveness of pilot- or large-scale application. Some of the emerging technologies include the following:

• Other AOPs -- Experiments examining the degradation pathways of MCPA using ozone and UV indicate that the addition of UV results in a more rapid degradation of MCPA than with ozone alone (Benoit-Guyod et al., 1986). Other AOPs, such as ozone combined with hydrogen peroxide, have been effective for the removal of several classes of pesticides, including other chlorophenoxy herbicides; however, no information was found on the removal of MCPA specifically (Ijpelaar et al., 2002; Kruithof et al., 2002b; Benitez et al., 2004). Laboratory studies have shown that the Fenton process is also an effective AOP for the degradation of MCPA (Fdil et al., 2003).
• Electrochemical degradation -- A variety of electrochemical degradation techniques have been examined. Laboratory studies found electrochemical degradation by peroxi-coagulation and photoperoxi-coagulation was effective for the removal of MCPA (Boye et al., 2003). Electrochemical degradation by anodic oxidation and electro-Fenton using a boron-doped diamond electrode was also found to be effective for MCPA removal (Brillas et al., 2004).
• Alternative adsorbents -- Adsorption of MCPA from aqueous solutions has also been demonstrated using acid-activated spent bleaching earth, organo-clays and an adsorbent produced from elutrilithe (Akçay and Yurdakoç, 2003; Mahramanlioglu and Güçlü, 2003; Mahramanlioglu et al., 2003).

7.2 Residential scale

Municipal treatment of drinking water is designed to reduce contaminants to levels at or below their guideline value. As a result, the use of residential scale treatment devices on municipally treated water is generally not necessary but primarily based on individual choice. In cases where an individual household obtains its drinking water from a private well, a private residential drinking water treatment device may be an option for reducing MCPA concentrations in drinking water.

Health Canada does not recommend specific brands of drinking water treatment devices, but it strongly recommends that consumers use devices that have been certified by an accredited certification body as meeting the appropriate NSF International (NSF)/American National Standards Institute (ANSI) drinking water treatment unit 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 drinking water devices and materials as meeting NSF/ANSI standards:

• Canadian Standards Association International;
• NSF International;
• Water Quality Association;
• Underwriters Laboratories Inc.;
• Quality Auditing Institute;
• International Association of Plumbing & Mechanical Officials.

An up-to-date list of accredited certification organizations can be obtained from the Standards Council of Canada.

Although no certified treatment devices are currently available for the reduction of MCPA from untreated water (such as a private well), treatment devices using activated carbon filters or reverse osmosis may be effective. Treatment devices certified under NSF International/American National Standards Institute (NSF/ANSI) Standard 53 (Drinking Water Treatment Units - Health Effects) for the reduction of the concentration of volatile organic compounds or NSF/ANSI Standard 58 (Reverse Osmosis Drinking Water Treatment Systems) for the reduction of the concentration of organic compounds may be suitable alternatives for the reduction of MCPA. Before a treatment device is installed, the well water should be tested to determine general water chemistry and to select the type of device that is appropriate.

Periodic testing by an accredited laboratory should be conducted on both the water entering the treatment device and the finished water to verify that the treatment device is effective. Devices can lose removal capacity through usage and time and need to be maintained or replaced. Consumers should verify the expected longevity of the components in their treatment device as per the manufacturer's recommendations and service it when required.

8.0 Kinetics and metabolism

MCPA is rapidly absorbed following ingestion, is not extensively metabolized and is eliminated primarily in the urine, mostly as unchanged MCPA, in rats, dogs and humans. However, dogs have a longer plasma half-life and slower elimination than other species, including humans. MCPA-DMAS and MCPA-EHE were rapidly converted to MCPA in rat studies, and the metabolism and toxicokinetics were similar to those of MCPA.

8.1 Absorption

MCPA is readily absorbed from the gastrointestinal tract (via gavage and following direct gastric intubation) in rats, dogs and humans (Elo, 1976; Fjeldstad and Wannag, 1977; Kolmodin-Hedman et al., 1983a,b; Jahanshahi and Stow, 1995; Hardwick, 1999, 2000; Lappin et al., 2002).

Oral administration of single MCPA doses of 5 or 100 mg/kg bw resulted in peak plasma concentrations being attained within 2-4 hours of dosing in rats and within 4.5-7 hours in beagle dogs (Lappin et al., 2002). Similar peak plasma concentrations were attained (within 2-3 hours of dosing) in rats with a single dose of 5 mg/kg bw of either MCPA-DMAS or MCPA-EHE (van Ravenzwaay et al., 2004). Human volunteers given an MCPA dose of 0.015 mg/kg bw per day exhibited a peak plasma concentration after 1 hour (Kolmodin-Hedman et al., 1983a). Timchalk (2004), in a comparison review of these metabolism studies in rats, dogs and humans, showed that dogs had a longer plasma half-life than rats and humans. At a dosage of 5 mg/kg bw, the plasma half-life was 63 hours in dogs, which was considerably higher than that in rats (6 hours) or humans (11 hours), but humans received a lower dose of MCPA (0.015 mg/kg bw).

8.2 Distribution

Once absorbed, MCPA diffuses readily across biological membranes and is widely distributed to various tissues and organs via the circulatory system. Concentrations decline rapidly; in rats, MCPA is rapidly excreted, mainly unchanged, in the urine (Elo, 1976; Jahanshahi and Stow, 1995; van Ravenzwaay et al., 2004). No significant accumulation of MCPA is observed in the tissues.

In rats administered [¹⁴C]MCPA by direct gastric intubation at a single dose of 11.5 mg/kg bw, the peak concentration in tissues was reached between 2 and 8 hours after dosing, after which the concentrations declined rapidly (Elo, 1976). The tissues/organs with the highest concentrations were the blood, kidney, suprarenal gland, lung, heart, liver, thyroid gland and bone marrow. In contrast, the brain, adipose tissue, testis and muscle contained the lowest concentrations.

In rats administered [¹⁴C]MCPA at single oral dose of 5 or 100 mg/kg bw, the radioactivity in the tissues and carcasses accounted for ≤ 2.3% of the dose at sacrifice (Jahanshahi and Stow, 1995; van Ravenzwaay et al., 2004). In sacrificed rats receiving the low-dose regimens (sacrificed on day 4), MCPA was non-detectable in most tissues, except in fat, skin and kidneys. In sacrificed rats receiving the higher dose (sacrificed on day 7), the level of radioactivity was also highest in fat, skin and kidneys, but MCPA was detected at higher concentrations and in more organs than at the lower dose. In addition, females had higher levels of radioactivity than males. The authors observed that small amounts of residual activity found in kidneys were consistent with continued excretion of the compound (van Ravenzwaay et al., 2004).

8.3 Metabolism

The metabolism and toxicokinetics of MCPA-DMAS and MCPA-EHE were investigated in rats administered single doses of 5 mg/kg bw each and compared with MCPA (van Ravenzwaay et al., 2004). Toxicokinetics and metabolism were indistinguishable from those of MCPA. MCPA, MCPA-DMAS and MCPA-EHE were not extensively metabolized in rats after oral administration.

In the fatal intoxication of a 23-year-old male, the forensic autopsy found 4-chloro-2-methylphenol in the body fluids and organ tissues, suggesting a metabolite of MCPA (Takayasu et al., 2008). In a study where five healthy human volunteers were given 15 µg/kg body weight of MCPA, the further investigation of free or conjugated MCPA in one individual showed that conjugation varied between 56-73% within the 72 hour collection period (Kolmodin-Hedman et al., 1983b).

Lappin et al. (2002) reported that rats orally exposed to 5 and 10 mg/kg/day MCPA excreted the parent compound mainly in the urine (approximately 65% of the dose). The only significant metabolite in the rat urine was the hydroxymethylphenoxyacetic acid (HMCPA); a trace of the glycine conjugate of MCPA was also detected, however, in an amount too small to reliably quantify. The dog urine contained a smaller proportion of MCPA (2-29%) than that seen in the rat, however, the glycine conjugate of MCPA was detected at much higher levels (up to 37% of the dose, and the taurine conjugate, which was not detected in the rat urine, was also detected (up to 7% of the dose).

8.4 Excretion

Renal excretion is reported as the major route of MCPA elimination in rats and dogs dosed orally with MCPA (Elo, 1976; Jahanshahi and Stow, 1995; Lappin et al., 2002; van Ravenzwaay et al., 2004). Dogs and rats have different recovery patterns. Studies showed that renal clearance in dogs was slower and less extensive than in rats (Lappin et al., 2002) or humans (Fjeldstad and Wannag, 1977; Kolmodin-Hedman et al., 1983a,b).

In rats, 75-80% of the administered dose was excreted in the urine over 24 hours, irrespective of the dose. In dogs, after oral administration of MCPA in single doses of 5 or 100 mg/kg bw, elimination was not complete at the end of 120 hours (Lappin et al., 2002).

In rats, the parent compound (MCPA) was the major compound excreted in the urine, along with lower levels of an oxidation product (4-chloro-2-hydroxymethyl-phenoxyacetic acid [HMCPA]) and trace amounts of the glycine conjugate (Lappin et al., 2002). In dogs, the parent compound was also present in the urine, but at lower levels than in the rat. In addition, three metabolites were recovered in the urine: the glycine conjugate, at higher levels than in the rat, and, to a lesser extent, HMCPA and a taurine conjugate (Lappin et al., 2002).

Faecal elimination was a minor route for both species; however, dogs had a higher proportion of MCPA in the faeces compared with rats. Radioactivity was not detected in expired air of rats orally dosed with [¹⁴C]MCPA at 5 mg/kg bw (van Ravenzwaay et al., 2004).

In human volunteers given MCPA at 0.015 mg/kg bw, an average of 40% of the given dose was excreted in the urine during the first 24 hours (Kolmodin-Hedman et al., 1983b). In another study, volunteers given 5 mg of MCPA excreted 50% in the urine within 48 hours and 55% within 96 hours; by the fifth day, the concentration in the urine was below the detection limit (Fjeldstad and Wannag, 1977). No attempt to analyse metabolites was seen in either human study, nor were the doses radiolabelled using [¹⁴C]MCPA.

In a comparison review of metabolism studies in rats, dogs and humans, Timchalk (2004) showed that dogs had a longer plasma half-life and slower elimination than rats and humans, which resulted in a substantially higher body burden of MCPA, at comparable doses, than in other species. In previous studies with organic acids with similar pharmacokinetic properties, the dog had a more limited capacity than other species to excrete organic acids via the kidney. The authors suggested that the following two mechanisms may be responsible for this decrease in renal clearance: saturation of renal secretion and increased renal tubule reabsorption. These differences in the pharmacokinetics of MCPA and other related organic acids between dogs and other species suggest that the use of dog toxicity data for determining human health risk may not be appropriate.

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