Page 8: Guidelines for Canadian Drinking Water Quality: Guideline Technical Document – Tetrachloroethylene

7.0 Treatment technology and distribution system considerations

7.1 Municipal scale

Municipal drinking water treatment plants that rely on conventional treatment techniques (coagulation, sedimentation, filtration and chlorination) have generally been found to be ineffective (0-29%) in decreasing concentrations of VOCs in drinking water (Love et al., 1983; Robeck and Love, 1983). Some incidental removal of VOCs may occur as a result of volatilization in open basins (Health and Welfare Canada, 1993). U.S. EPA (1991b) reported granular activated carbon (GAC) adsorption and packed tower aeration (PTA) as the best available technologies (BATs) for the reduction of tetrachloroethylene concentrations in drinking water. These treatment technologies have proven effective for the reduction of an influent concentration of tetrachloroethylene in the range of 30-500 μg/L to an effluent concentration of below 1 μg/L (Love and Eilers, 1982; Chrobak et al., 1985; Reijnen et al., 1985; Hand et al., 1988; AWWA, 1991; Lykins and Clark, 1994; Dyksen et al., 1995).

Various advanced oxidation processes (AOPs), such as ozonation in combination with hydrogen peroxide, have also been found to be effective treatment methods for tetrachloroethylene (Aieta et al., 1988; Zeff, 1991; Dyksen et al., 1992; Topudurti, 1993; Hirvonen et al., 1996b; Karimi et al., 1997; Hirvonen et al., 1998).

The selection of an appropriate treatment process for a specific water supply will depend on many factors, including the characteristics of the raw water and the operational conditions of the specific treatment method.

7.1.1 Activated carbon adsorption

Activated carbon is used in water treatment processes either as GAC or as powdered activated carbon (PAC). The adsorption capacity of activated carbon to remove VOCs is affected by a variety of factors, such as competition from other contaminants, preloading with natural organic matter (NOM), temperature, the physicochemical properties of the VOCs and the carbon media (Speth, 1990; AWWA, 1991). The PAC application, most suitable for conventional treatment systems treating surface waters, may decrease occasional low concentrations of VOCs, including tetrachloroethylene, when it is applied at the treatment plant, allowing sufficient contact time and a proper mixing. In conventional treatment plants, the common points of PAC application are at the plant intake, during the rapid mix process and in the filter influent. Sufficient contact time, a function of the characteristics and the concentration of the contaminant to be adsorbed, is necessary (Najm et al., 1991). In general, PAC adsorption is found to be less efficient than GAC adsorption for VOC removal largely due to its use in coagulation/ sedimentation basins where the adsorption sites can be blocked due to floc formation; the fact that it will not have the necessary time to reach its maximum adsorption capacity; and the fact that the equilibrium liquid-phase concentration (concentration gradient driving force) decreases during the adsorption process. A full-scale study demonstrated that PAC application was capable of decreasing an influent tetrachloroethylene concentration of 0.7 μg/L to 0.3 μg/L, with an applied PAC dose of 7.1 mg/L (Singley et al., 1979).

Higher concentrations of VOCs are found in groundwater. For the continuous removal of VOCs from groundwater, GAC adsorption is a commonly used process (Snoeyink, 1990). The process uses a contactor packed with GAC. As the water passes through the GAC contactor, the contaminants diffuse to the adsorbent granules and accumulate on their surfaces (Crittenden et al., 2005). Packing the carbon in columns allows more complete contact between the water and the media, greater adsorption efficiency and greater process control than with PAC (Snoeyink, 1990).

The choice of GAC application for removing VOCs from drinking water supplies involves the following process design considerations: carbon usage rate, empty bed contact time (EBCT), pretreatment of the raw water, contactor configuration and method of GAC replacement or regeneration. During the operation time, and depending on a variety of factors discussed above, organic contaminants will "breakthrough" the carbon bed. Initial breakthrough is defined as the time when the contaminant concentration in the effluent exceeds the treatment objective. In systems with multiple beds, the individual beds can be operated beyond the time of initial breakthrough, provided the contaminant concentration in the blended effluents still meets the treatment objectives. Once the GAC is exhausted, it is removed from the contactor and replaced with fresh or regenerated GAC. For practical reasons, PAC is not recovered and reactivated; thus, its carbon use rate can be high when compared with that of GAC (Chowdhury et al., 2013). The regeneration and/or replacement of exhausted media are important economic considerations in achieving the contaminant treatment goal.

Common operating problems when using GAC adsorption contactors include biological growth and the increase in heterotrophic plate counts in the effluent, clogging and fouling of the carbon bed by chemical and bacterial precipitants (AWWA, 1991). Operating considerations include the needs to ensure a proper backwash, maintain the bed depth and bed density after backwashing and control the flow rate. To prevent clogging of the bed, pretreatment of the water prior to the GAC contactor is often required (Snoeyink, 1990; Speth, 1990; Crittenden et al., 2005).

An influent tetrachloroethylene concentration in the range of 30-194 µg/L could be decreased to below 1 µg/L in drinking water under reasonable operating conditions (Love and Eilers, 1982; Chrobak et al., 1985; Hand et al., 1988; U.S. EPA, 1990; AWWA, 1991; Dyksen et al., 1995). Operational data from a municipal-scale GAC treatment facility indicate that an influent concentration of tetrachloroethylene of 30 µg/L can be decreased to below 1 µg/L using a single contactor with a hydraulic loading rate of 2.56 gpm/ft2 (6.2 m/h) and an EBCT of 12.7 minutes, resulting in a carbon use rate of 0.0167 kg/m3 (kilograms of GAC per cubic metre of water treated; Hand et al., 1988). Other full-scale results demonstrated that two GAC contactors in parallel mode with a hydraulic loading rate of 7.1 gpm/ft2 (17.3 m/h) each and an EBCT of 10.5 minutes were capable of decreasing a tetrachloroethylene concentration of 194 µg/L to less than 1 µg/L with a carbon use rate of 102 lb GAC per million gallon of treated water (0.0122 kg /m3; Chrobak et al., 1985). A survey of treatment plants using GAC for the removal of tetrachloroethylene in the United States found that influent concentrations ranging from 5 to 150 µg/L could be decreased to below 1 µg/L under a variety of operating conditions (AWWA, 1991; Dyksen et al., 1995).

The GAC process is the most widely used for small water treatment systems due to its simplicity and ease of operation (Snoeyink, 1990).

7.1.2 Air stripping: packed tower aeration

Air stripping treatment technology is widely used to decrease the concentrations of VOCs in drinking water (Dyksen et al., 1984; Cummins and Westrick, 1990; U.S. EPA, 1991a; Dzombak et al., 1993; Dyksen, 2005; WHO, 2011). An air stripping process brings water and air into contact, allowing the transfer of volatile contaminants from the water to the air, as the driving force of the process is the contaminant concentration gradient between the two phases.

Although various air stripping equipment configurations exist, the U.S. EPA considers PTA as the best technology, achieving 99% removal of VOCs from drinking water (U.S. EPA, 1991b). PTA provides an optimum system for the removal of VOCs from water, as it allows for greater air-to-water ratios than other aeration systems. In the PTA column, the contaminated water flows downward by gravity over a bed of packed material, while the air is introduced into the tower below the packed bed and flows upward countercurrent to the water flow. As the PTA transfers VOCs from water to air, treatment of the stripping tower off-gas to decrease the contaminant concentrations prior to discharge into the atmosphere may be necessary (Crittenden et al., 1988; Adams and Clark, 1991).

Several design factors affect the stripping rate of VOCs: air-to-water ratio; available area of mass transfer, hydraulic loading rate, temperature of the water and air, and physical and chemical properties of the contaminant (AWWA, 1991; Crittenden et al., 2005; Dyksen, 2005). Diffused aeration, multistage bubble aerators, tray aeration and shallow tray aeration have been identified as alternative air stripping treatment technologies for the removal of tetrachloroethylene in drinking water for small systems (U.S. EPA, 1998).

A common operating problem is scaling and fouling of the column. The main causes of fouling are calcium carbonate and/or calcium sulphate scale, iron oxidation and microbial growth. Methods to prevent the fouling of the column include pH suppression of the influent, use of scale inhibitors or iron removal prior to the PTA application (U.S. EPA, 1984; Dyksen, 2005). Algal growth can also be a problem in locations where light could be introduced into the tower. Post treatment, such as the use of a corrosion inhibitor, may also be required to reduce corrosive properties of the water due to increased dissolved oxygen from the aeration process. Environmental conditions, such as water temperature, may impact the packed tower performance. Although temperatures below freezing can cause operational issues, contact between water and air in PTA will result in a change in the air temperature until it approaches the water temperature. The temperature influences both the Henry's Law constant and the rate of mass transfer coefficient of the contaminant. These parameters impact the size of the equipment and the removal efficiency of the VOCs (Crittenden et al., 2005).

Using PTA, an influent tetrachloroethylene concentration in the range of 60-500 µg/L could be decreased to below 1 µg/L in drinking water under reasonable operating conditions (Reijnen et al., 1985; Hand et al., 1986, 1988; AWWA, 1991; Dyksen et al., 1995). Full-scale PTA data indicates that an influent flow rate of 8.1 ML/day with a concentration of tetrachloroethylene of 60 µg/L can be decreased to 0.4 μg/L using an air-to-water ratio of 61, an air stripper length of 7.5 m, hydraulic loading rate of 15.5 kg/m2/s, and a packed column diameter of 2.4 m (Hand et al., 1988). Influent concentrations as high as 500 µg/L have been decreased to below 1 µg/L using an air-to-water ratio of 35, an air stripper length of 7.3 m, a hydraulic loading rate of 27.2 gpm/ft2 (66.3 m/h) and a packed column diameter of 2.3 m (Dyksen et al., 1995). Results of a survey conducted on full-scale drinking water treatment plants indicated that PTA is commonly capable of decreasing influent concentrations of tetrachloroethylene of up to 170 µg/L to levels below 1 µg/L in the finished water under a variety of operating conditions (Reijnen et al., 1985; AWWA, 1991).

Generally, diffused aeration achieves lower removal efficiencies and has higher power requirements than PTA systems (AWWA, 1991). Typical diffused aeration performance ranges from 73% to 95% removal of tetrachloroethylene (U.S. EPA, 1984, 1991a). A pilot-scale diffused aeration study indicated that influent concentrations of tetrachloroethylene can be decreased from 636 µg/L to less than 1 µg/L using an air-to-water ratio of 16 and a contact time of 10 minutes (Love and Eilers, 1982).

Cost evaluations of systems ranging from 1 to 100 ML/day were conducted by Adams and Clark (1991). These evaluations indicate that, in most cases, the use of PTA for the removal of tetrachloroethylene in drinking water is more cost effective than the use of GAC, even when vapour-phase GAC treatment of the stripping tower off-gas is required.

7.1.3 Combination of packed tower aeration and granular activated carbon

PTA technology combined with liquid-phase GAC adsorption has the potential to be effective for producing water with low effluent levels of VOCs. In a municipal-scale treatment plant combining these processes, air stripping was used for the bulk removal of VOC from water, and activated carbon adsorption was used in the second step to further decrease the residual VOC concentrations (McKinnon and Dyksen, 1984; Stenzel and Sen Gupta, 1985). In addition, the use of an air stripping process preceding liquid-phase GAC adsorption can extend the carbon bed life. A packed tower column combined with a GAC adsorber demonstrated that an influent concentration of tetrachloroethylene of 100 µg/L could be decreased to less than 1 μg/L using an air-to-water ratio of 90 and a packing height of 3.0 m (AWWA, 1991). No information was provided on the operational conditions of the GAC adsorber used in this study.

7.1.4 Ozonation

The reaction kinetics of ozonation are generally not considered to be favourable for the treatment of tetrachloroethylene in drinking water due to slow reaction rates and the need to achieve low effluent concentrations (Dyksen et al., 1992; Hirvonen et al., 1996b). However, pilot-scale studies have demonstrated that a 60-75% removal of tetrachloroethylene is achievable using ozone doses between 6 and 9 mg/L (Fronk, 1987; Zeff, 1991).

7.1.5 Advanced oxidation processes

AOP refers to the use of appropriate combinations of ultraviolet (UV) light, chemical oxidants and catalysts (ozone/hydrogen peroxide, ozone/UV, UV/hydrogen peroxide, UV/titanium dioxide, ozone/UV/titanium dioxide, ozone oxidation at elevated pH) to generate highly reactive radicals, such as hydroxyl radicals, which are strong oxidants and react rapidly and non-selectively with organic contaminants.

Physical and chemical properties of the water matrix have a major impact on AOPs. Water quality parameters that may impact the effectiveness of the AOPs are alkalinity of the water, pH, NOM, reduced metal ions (iron and manganese) and turbidity. The primary advantage of AOPs is their capability to completely convert the organic compounds into carbon dioxide and mineral acids (Crittenden et al., 2005), whereas the adsorption processes and the air stripping techniques transfer the contaminants from one phase to another phase and may require additional treatment (Glaze and Kang, 1988; Topudurti, 1993; Crittenden et al., 2005). Pilot-scale studies indicated that an in-line ozone/hydrogen peroxide process was capable of decreasing an influent tetrachloroethylene concentration of 18.6 µg/L to an effluent concentration of 1 µg/L using an applied ozone dose of 6.0 mg/L, a hydrogen peroxide to ozone ratio of 0.5 and a contact time of 3 minutes (Dyksen et al., 1992). Other full-scale data demonstrated that the use of an ozone/hydrogen peroxide process was effective in decreasing an influent tetrachloroethylene concentration of 10 µg/L to below 1 µg/L using an ozone dose of 4.7 mg/L and a hydrogen peroxide to ozone ratio of 0.57 (Karimi et al., 1997). Results from both studies suggest that the ozone dosage appears to have a greater impact on the removal of tetrachloroethylene than the contact time of the reaction. Similarly, full-scale data demonstrated that a combined UV/hydrogen peroxide/ozone oxidation process was capable of decreasing an influent tetrachloroethylene concentration of 7 µg/L to less than 1 µg/L (Zeff, 1991).

Full-scale AOPs using UV radiation and hydrogen peroxide oxidation treatment have been assessed for the removal of tetrachloroethylene. Medium-pressure UV lamps and hydrogen peroxide doses of 15-70 mg/L decreased an influent concentration of tetrachloroethylene in the range of 70-150 μg/L to below 1 μg/L in the treated water (Topudurti, 1993). Field-scale results, reported by Hirvonen et al. (1998), showed that low-pressure mercury lamps and hydrogen peroxide doses of 83-138 mg/L could decrease an influent concentration of tetrachloroethylene in the range of 76-139 μg/L to an effluent concentration of below 0.5 µg/L.

Specific operational issues should be considered when using each of the above-described AOP technologies. The common operating issues related to the use of UV radiation are UV lamp replacement, regular removal of the suspended particles that coat the quartz tubes housing the UV lamps and ensuring low levels of colour and turbidity of the water. The application of a UV/hydrogen peroxide process requires a high initial dosage of hydrogen peroxide in order to efficiently utilize the UV light to produce hydroxyl radicals, which results in a high effluent hydrogen peroxide concentration in the finished water. Other operating concerns when using ozone/hydrogen peroxide and UV/ozone include the stripping of the VOCs from the ozone contactor.

The formation of by-products from the oxidation and/or advanced oxidation of tetrachloroethylene or other inorganic or organic compounds in the source water should be considered when using these processes. AOPs produce reactive peroxy organic radicals, which undergo radical chain reactions and result in a variety of oxygenated by-products. The typical by-products produced by AOPs are aldehydes, ketones and carboxylic compounds. The presence of by-products may require additional treatment following AOPs and/or process optimization to minimize by-product formation.

The formation of low concentrations of trichloroacetic acid (TCA) (0.1 µg/L) and dichloroacetic acid (DCA) in the range of 1-6 µg/L from the UV/hydrogen peroxide oxidation of tetrachloroethylene has been observed in the treated water (Hirvonen et al., 1996a, 1998). However, the reported concentrations of these compounds were below the World Health Organization (WHO) drinking water guidelines of 200 µg/L for TCA and 50 µg/L for DCA (WHO, 2011). Health Canada's guideline for total haloacetic acids in drinking water is 80 µg/L.

Other studies found that with the exception of assimilable organic carbon (AOC) and bromate, no by-products were produced following ozone/hydrogen peroxide oxidation. The average concentration of an AOC of 53.5 µg/L in the groundwater was increased to about 239 µg/L in the treated water. An influent bromide concentration of 0.21 mg/L resulted in an effluent bromate concentration in the range of 0.029-0.11 mg/L (Karimi et al., 1997). The formation of bromate was found to be dependent on the applied ozone dosage.

7.1.6 Reverse osmosis

Reverse osmosis technology has shown some promise for its potential to remove tetrachloroethylene from drinking water (U.S. EPA, 1991a). Bench-scale investigations demonstrated that influent concentrations of tetrachloroethylene ranging from 6 to 153 µg/L were decreased up to 92% using thin film composite membranes (Lykins et al., 1988). The ability of reverse osmosis to remove synthetic organic chemicals has been found to be dependent on a variety of system components, including type of membrane, flux, recovery, synthetic organic chemical solubility, charge and molecular weight (Taylor et al., 2000).

7.1.7 Emerging and other treatment technologies

New drinking water treatment technologies for tetrachloroethylene are being developed but are still primarily in the experimental stage and/or have no published information on the effectiveness of full-scale application. Some of the emerging technologies are as follows:

  • Layered upflow carbon adsorption
    • Alternative GAC contactor configurations have shown some success for the removal of tetrachloroethylene in drinking water at decreased carbon usage rates. A pilot-scale layered upflow carbon absorber was capable of decreasing a tetrachloroethylene concentration of 13 µg/L to below 5 µg/L in the finished water, with an EBCT of 6 minutes and a carbon usage rate of 0.010 kg/m3 (Munz et al., 1990).
  • Membrane technologies
    • Membrane air stripping: Air stripping of VOCs with microporous polypropylene hollow fibre membranes has been introduced as an alternative method to PTA (Semmens et al., 1989; Castro and Zander, 1995). Pilot-scale studies demonstrated up to 80% decrease of tetrachloroethylene concentrations and greater mass transfer coefficients than with the use of traditional air stripping towers (Zander et al., 1989).
    • Nanofiltration: Laboratory-scale studies examining the effectiveness of various nanofiltration membrane characteristics demonstrated an average 93% removal of tetrachloroethylene with an influent concentration of 400 µg/L in synthetic water (Ducom and Cabassud, 1999).
    • Pervaporation: Pervaporation has been considered as an emerging polymeric membrane-based technology for the removal of organic contaminants from water. Pervaporation requires dense and selective membranes and the separation is based on the relative solubility and diffusivity of each component in the membrane material (Khayet and Matsuura, 2004). In pervaporation, one side of the membrane is in contact with the contaminated water, while the other side is exposed to a vacuum. The process includes sorption of the contaminant onto the membrane, permeation through the membrane and evaporation into the vapour phase. However, no information was found specifically related to the removal of tetrachloroethylene from water (Lipski and Cote, 1990; Uragami et al., 2001).
    • Vacuum membrane distillation: The vacuum membrane distillation process uses porous and hydrophobic membranes that act only as support for the vapour-liquid interface (Khayet and Matsuura, 2004). The liquid stream vaporizes on the feed side of the membrane; the vapour diffuses through the membrane pores and is condensed outside the membrane. Mass transfer through the membrane is improved by applying a vacuum on the permeate side (Couffin et al., 1998). Couffin et al. (1998) reported that the vacuum membrane distillation process appears to be a promising treatment technology for the removal of low concentrations of tetrachloroethylene from water.
  • Other AOPs
    • A pilot-scale photocatalytic oxidation system was capable of decreasing influent concentrations of tetrachloroethylene from 125 µg/L to 5 µg/L in the finished water. The oxidation system utilized UV radiation with a titanium dioxide semiconductor combined with the addition of 70 mg/L of hydrogen peroxide and 0.4 mg/L of ozone (Topudurti et al., 1998).
    • The Fenton oxidation reaction, based on strong oxidizing (OH·) radicals produced by the reaction of hydrogen peroxide with iron sulphate, is also an effective process for the degradation of tetrachloroethylene in water. Laboratory experiments demonstrated 95% degradation of tetrachloroethylene with a high initial concentration of 162 mg/L. The formation of by-products such as TCA should be considered when using AOPs such as Fenton oxidation (Yoshida et al., 2000).
    • Use of a high-energy electron beam (e-beam) has been shown to be an effective process for the destruction of tetrachloroethylene in aqueous solutions at a large scale. An energy dose in the range of 299-776 krad was required to achieve a 99% removal of tetrachloroethylene in an aqueous solution with a pH of 7 (initial concentration ranging from 0.1 to 4.5 mg/L) (Cooper et al., 1993).

7.1.8 Distribution system

A vinyl-toluene lining used in the 1960s and 1970s to minimize the corrosion of asbestos-cement pipes (primarily in New England) was a source of tetrachloroethylene in drinking water. However, use of this type of lining material was discontinued after it was discovered that tetrachloroethylene leached into the drinking water (Larson et al., 1983). Tetrachloroethylene was added to thin the resin used to line the inside of pipes prior to spraying. Leaching of tetrachloroethylene was observed when insufficient time was allowed for the resin to dry (cure) after its application. The highest tetrachloroethylene concentrations measured (up to 3500 µg/L) were most often found in dead-ends or in areas of low flow in the distribution system. Remedial actions taken to decrease tetrachloroethylene concentrations included the removal or replacement of lined pipes, the flushing of distribution systems and the installation of bleeders. Moreover, concentrations in water were found to decrease as the pipes aged (Larson et al., 1983).

Drinking water contamination incidents resulting from permeation of organic chemicals in soil through water piping and piping components have been reported in the literature (Glaza and Park, 1992; Bromhead, 1997; Goodfellow et al., 2002; Ong et al. 2008).

Plastic pipes and piping components are durable and have a good resistance to chemicals, such as chlorine, used in the water treatment. However, plastic pipes and piping components may come in contact with contaminated soils as a result of leaks from underground storage tanks, chemical spills, and improper disposal of used chemicals. These contaminants may pose serious threats to the longevity and structural integrity of plastic pipes and elastomeric gaskets and can affect the water quality in the distribution system.

A survey (2004) of water utilities on permeation incidents involving plastic pipes and elastomeric gasket provided information on utilities' experience. Of the 151 water utilities in the US and Canada that responded to the survey, 70% reported using plastic pipes. Polyvinyl chloride (PVC) and polyethylene (PE) pipes accounted for 18% and 0.18% of the mains, respectively. The survey also reported that ductile iron (DI) pipes accounted for 16% of the mains. Generally, permeation incidents involving potable water distribution pipes are rarely reported. The survey reported 6 water main permeation incidents; three of these incidents involved gasoline, one involved a chlorinated solvent and two were associated with unknown chemicals. The pipe materials involved in permeation incidents were PVC (4), asbestos cement (1) and cast iron (1). A case reported a cast iron pipe with lead joints exposed to chlorinated solvents. Concentrations of tetrachloroethylene were detected in the soil and groundwater (1,560 mg/kg and 0.44 mg/L, respectively) but not in the drinking water. Plastic service connections were used by 49% of the utilities, with PVC and PE accounting for 5% and 6%, respectively. In addition, the survey reported that of the 44 service connection incidents of permeation, two involved chlorinated solvents and DI pipes with styrene butadiene rubber gaskets. In an incident of permeation involving tetrachloroethylene and ductile iron pipes with styrene butadiene rubber gaskets, ranges of concentrations of tetrachloroethylene were 0.04-3.6 mg/kg in the contaminated soil, 0.9-15.4 mg/L in the groundwater and not detectable to 6.6 µg/L in the potable water (Ong et al., 2008).

Bromhead (1997) reported permeation incidents of tetrachloroethylene with polyethylene service pipes at contaminated locations in Rotterdam, Netherlands. Of the 143 water samples, 98% had tetrachloroethylene concentrations between <0.05 and 10 µg/L, and 2% had concentrations above 10 µg/L, with a maximum of 14.4 µg/L after 8 hours of stagnation.

7.2 Residential

Generally, it is not recommended that drinking water treatment devices be used to provide additional treatment to municipally treated water. 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 decreasing tetrachloroethylene concentrations in drinking water.

A number of residential treatment devices from various manufacturers are available that can remove tetrachloroethylene from drinking water to concentrations below 5 µg/L. Filtration systems may be installed at the faucet (point-of-use) or at the location where water enters the home (point-of-entry). From a health perspective point-of-entry systems are preferred for VOCs such as tetrachloroethylene because they provide treated water for bathing and laundry as well as for cooking and drinking. Certified point-of-use treatment devices are currently available for the removal of VOCs, including tetrachloroethylene. Where certified point-of-entry treatment devices are not available for purchase, systems can be designed and constructed from certified materials. Periodic testing by an accredited laboratory should be conducted on both the water entering the treatment device and the water it produces to verify that the treatment device is effective. Devices can lose their removal capacity through usage and time and need to be maintained and/or replaced. Consumers should verify the expected longevity of the components in their treatment device as per the manufacturer's recommendations.

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 (SCC, 2015):

  • CSA Group (www. csagroup.org);
  • NSF International (www.nsf.org);
  • Water Quality Association (www.wqa.org);
  • Underwriters Laboratories Inc. (www.ul.com);
  • Bureau de normalisation du Québec (www.bnq.qc.ca - available in French only); and
  • International Association of Plumbing and Mechanical Officials (www.iapmo.org).

An up-to-date list of accredited certification organizations can be obtained from the SCC (2015).

Treatment devices to remove tetrachloroethylene from untreated water (e.g., a private well) can be certified for either the removal of tetrachloroethylene alone or the removal of a variety of VOCs, including tetrachloroethylene.

Treatment devices that are certified to remove tetrachloroethylene or VOCs under NSF/ANSI Standard 53 are generally based on activated carbon adsorption technology. For a drinking water treatment device to be certified to NSF/ANSI Standard 53 (Drinking Water Treatment Units—Health Effects) for the reduction of tetrachloroethylene concentrations, the device must be capable of reducing an average influent concentration of 0.015 mg/L to a maximum finished effluent concentration of 0.005 mg/L or less (NSF/ANSI, 2011). Treatment devices certified for the reduction of organic chemicals included in NSF/ANSI Standard 53 by surrogate testing must be capable of reducing the concentration of tetrachloroethylene by greater than 99% from an influent (challenge) concentration of 0.081 mg/L to a maximum final (effluent) concentration of less than 0.001 mg/L (NSF/ANSI, 2011).

Reverse osmosis systems certified to NSF/ANSI Standard 58 (RO) may also be certified for the reduction of tetrachloroethylene concentrations to achieve a final concentration of less than 0.001 mg/L (NSF/ANSI, 2012). Reverse osmosis systems should only be installed at the point of use as the water they have treated may be corrosive to internal plumbing components.

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