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

Part II. Science and Technical Considerations - Continued

7.0 Treatment technology

7.1 Municipal scale

Generally, conventional water treatment processes (coagulation, flocculation and clarification) have only a small effect on reducing the levels of ammonia in drinking water. Some removal may occur if ammonia is sorbed to colloidal particles (Department of National Health and Welfare, 1993; Kurama et al., 2002).

Treatment technologies and strategies to remove ammonia in drinking water include biological treatment (controlled nitrification) and physicochemical processes such as breakpoint chlorination, ion exchange, membrane filtration and air stripping.

Free chlorine and chloramine are two secondary disinfectants used for distributed water. The type of disinfection method used by utilities may influence the treatment technology to remove ammonia from drinking water. Some utilities form chloramine as a strategy to remove naturally occurring ammonia in the raw water supply.

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, the source and the concentration of ammonia (including variation), the operational conditions of the specific treatment method and the utility's treatment goal.

7.1.1 Biological treatment (controlled nitrification)

Biological treatment processes are based on the ability of microorganisms (non-pathogenic bacteria) to catalyse the biochemical oxidation or reduction of drinking water contaminants and produce biologically stable water (Rittmann and Snoeyink, 1984).Biological treatment processes have been used in Europe for several years for the removal of ammonia from drinking water (Goodall, 1979; Rittmann and Snoeyink, 1984; Rogalla et al., 1990; Janda and Rudovský, 1994) and have more recently gained acceptance for use in North America (Andersson et al., 2001; Lytle et al., 2007; White et al., 2009; McGovern and Nagy, 2010).

Several authors have reported on full-scale biological treatment to oxidize ammonia in the source water, achieving an oxidation rate greater than 90% (Rittmann and Snoeyink, 1984; Rogalla et al., 1990; Janda and Rudovský, 1994; Andersson et al., 2001; Hossain et al., 2007; Lytle et al., 2007; White et al., 2009). The nitrification process is regarded as the pathway to oxidize ammonia in the biological treatment. As ammonia-oxidizing bacteria (AOB) and nitrite-oxidizing bacteria (NOB) (i.e., nitrifiers) are slow-growing organisms, biologically active filters require a period of colonization before efficient ammonia removal is reached. During this period, ammonia breakthrough and nitrite formation can have adverse impacts on water quality (Lytle et al., 2007; McGovern and Nagy, 2010). Based on pilot study results, Lytle et al. (2007) reported that a colonization to obtain complete nitrification can be achieved in new filters in less than 3 months. This was achieved by constantly running aerated raw water through the filters to promote bacterial regrowth. In order to have complete nitrification a stoichiometric oxygen (O2) demand of 4.33 mg O2/mg NH4+ -N is required. At ammonia concentrations exceeding this oxygen demand, the biological treatment process requires a constant oxygen feed (Lytle et al., 2007; White et al., 2009).

The process may increase the level of nitrate and may release bacteria into the finished water. The finished water typically requires polishing (e.g., granular activated carbon [GAC] filtration) and post-treatment, such as disinfection, to ensure that neither undesirable organisms nor growth products pass into the distribution system (Wilczak, 2006a).

Critical factors that ensure optimized performance for biological treatment include high dissolved oxygen concentrations, phosphorus, optimal temperature for the selected biomass, a large surface area for accumulating the slow-growing nitrifying biomass, appropriate hydraulic loading rates and maintenance of a long solids retention time (a biomass hold-up in the filter) (Rittmann and Snoeyink, 1984; Bablon et al., 1988; Janda and Rudovský, 1994; Kors et al., 1998; Andersson et al., 2001; Kihn et al., 2002; Hossain et al., 2007; Lytle et al., 2007).

There are different configurations for biological water treatment processes. Most of the systems operate in a fixed biofilm configuration, which includes a biogrowth support medium for the bacterial activity (Rittmann and Snoeyink, 1984; Rogalla et al., 1990; Muramoto et al., 1995; Kors et al., 1998; Andersson et al., 2001; Lytle et al., 2007). Other systems operate in a suspended growth mode, where bacteria are hydraulically maintained in suspension within a reactor such as a fluidized bed filter (Goodall, 1979; Gauntlett, 1981). Gauntlett (1981) reported that fluidized beds had a higher reaction rate per unit volume, shorter residence time, better bacterial control and an absence of blocking or channelling compared with the fixed bed configurations. A pilot-scale study using a fluidized bed achieved an ammonia reduction greater than 95 % of an influent concentration of 3 mg NH3-N/L (Gauntlett, 1981).

Lytle et al. (2007) reported achieving an ammonia removal of greater than 95% using biological treatment in a full-scale plant (average 0.6 million gallons per day [MGD] [2270 m3/day]). The plant was designed for iron removal, and the filters had been in operation since the 1980s. Three parallel gravity flow sand filters, each operated with a hydraulic loading rate of 2 gallons per minute [gpm] per square foot (4.9 m/h), were capable of reducing an influent ammonia concentration of 1.11 mg NH3-N/L in pre-aerated groundwater to below the detection limit of 0.1 mg NH3-N /L in the blended post-filtration water. Filtered water was chlorinated and had a free chlorine residual of 0.9 mg Cl2/L and a stable pH (Lytle et al., 2007). The authors reported a rise in the nitrate-nitrogen concentration (NO3-N) from below 0.04 mg/L to 1.11 mg/L in the filtered water. No nitrite was detected in the filtered water, confirming a complete oxidation of ammonia to nitrate through the filters.

Sand covered with manganese dioxide has been reported to be an effective support for the attachment of nitrifying bacteria. Pilot-scale and full-scale studies reported that sand filters coated with manganese oxides achieved an ammonia oxidation in the range of 95-98% (Janda and Rudovský, 1994; Stembal et al., 2005). Two water treatment plants, each using a single sand filter coated with manganese dioxide, demonstrated a reduction of influent ammonia concentrations of 3.82 and 1.76 mg/L in pre-aerated groundwater to 0.21 and 0.08 mg/L in finished water, respectively, using an air:water ratio of 50. Each filter operated with a hydraulic loading rate up to 5 m/h. The authors observed ammonia breakthrough and nitrite in the finished water (concentrations not specified) when the filters operated at hydraulic loading rates above 5 m/h. The authors suggested that the ammonia removal occurred by nitrification and by sorption on hydrated manganese dioxide (Janda and Rudovský, 1994). Another full-scale study using sand filters coated with manganese dioxide demonstrated that a two-step nitrification process, each step consisting of aeration/filtration, was capable of reducing an average influent ammonia concentration of 4.38 mg/L to 0.13 mg/L in the finished water (Janda and Rudovský, 1994). Muramoto et al. (1995) reported complete oxidation of an average influent ammonia concentration of 0.48 mg/L in a full-scale biological activated carbon filter with an empty bed contact time (EBCT) of 15 minutes.

Andersson et al. (2001) and Kihn et al. (2002) investigated the impact of temperature on controlled nitrification. The studies used open superstructure (i.e., chemical activated) and closed superstructure (i.e., physical activated) GAC filters. The filters had been in service since 1990 for open superstructure GAC and since 1984 for closed superstructure GAC. Each filter operated with hydraulic loading rates in the range of 3.9-5.0 m/h and EBCT between 20 and 30 minutes. Both filters were fed with pre-filtered and ozonated water with influent ammonia concentrations in the range of 0.02 to 0.12 mg NH4+-N/L. The study reported a 98% and a 90% ammonia removal for the open superstructure and for the closed superstructure GAC filters, respectively, at temperatures of 16ºC and higher. Both filters achieved up to 30% ammonia oxidation at temperatures below 4ºC (Andersson et al., 2001). This lowered oxidation rate is most likely due to the fact that low temperatures decrease the bacterial activity (Bablon et al., 1988; Groeneweg et al., 1994; Andersson et al., 2001; Kihn et al., 2002; Hossain et al., 2007).

A full-scale study compared a single-medium (sand) filter with a dual-media (sand and GAC) filter for the removal of an influent ammonia concentration below 0.2 mg NH4+/L (0.15 mg NH3-N/L) at low temperature. The dual-media filter showed no ammonia breakthrough at a temperature of 2ºC, whereas the single-medium filter allowed approximately 20% of the influent ammonia to pass through. However, the dual-media filter provided no advantages over the single-layer filter at temperatures greater than 7ºC (Bablon et al., 1988).

As nitrite is an intermediate compound in the oxidation of ammonia to nitrate in biological filters, utilities should ensure that their system is optimized such that the biological process is complete and nitrite is not present in the treated water.

7.1.2 Breakpoint chlorination

Breakpoint chlorination can eliminate ammonia from water through the formation of a free chlorine residual. Breakpoint chlorination is described as a process in which chlorine demand is satisfied, combined chlorine compounds are destroyed, ammonia is oxidized to form nitrogen gas and free chlorine residual is achieved when additional chlorine is added. The process requires frequent monitoring of ammonia concentrations and the various forms of chlorine (combined, total chlorine and free chlorine residual) to ensure that breakpoint chlorination is achieved at all times. It is necessary to generate a breakpoint curve for every plant and to monitor the fluctuation of ammonia to ensure that breakpoint chlorination is always achieved.

Utilities use breakpoint chlorination to remove excess ammonia in the source water and to control nitrification episodes in the distribution system. In distribution systems, breakpoint chlorination can be an effective method to control ammonia-oxidizing bacterial activity in the short term, but it may not prevent the establishment of nitrifying biofilm on return to chloramination (Kirmeyer et al., 1995; Odell et al., 1996; Zhang and DiGiano, 2002; Pintar and Slawson, 2003).

Breakpoint chlorination requires chlorine doses approximately 8-10 times higher (on a weight basis) than the ammonia concentration to achieve a free chlorine residual. The process is a series of reactions in which monochloramine is formed first. The reaction rate of monochloramine formation depends on pH, temperature and the chlorine-to-ammonia-nitrogen (Cl2:NH3-N) weight ratio, preferably in the range of 3:1 to 5:1. Once monochloramine is formed and Cl2:NH3-N is greater than 5:1, breakpoint chlorination proceeds through two main groups of reactions: 1) disproportionation (acid-catalysed reactions) of monochloramine to form dichloramine and 2) decomposition of dichloramine. Both groups of reactions require an excess of free chlorine (Kirmeyer et al., 2004). Dichloramine undergoes a series of decomposition and oxidation reactions to form nitrogen-containing products, including nitrogen, nitrate, nitrous oxide gas and nitric oxide (AWWA, 2006). Trichloramine, or nitrogen trichloride, is an intermediate during the complete decomposition of chloramines. Its formation depends on pH and the Cl2:NH3-N weight ratio and may appear after the breakpoint (Kirmeyer et al., 2004; Hill and Arweiler, 2006; Randtke, 2010). At Cl2:NH3-N of 7.6:1, the free ammonia is oxidized to nitrogen and chlorine is reduced to chloride. An increase of the Cl2:NH3-N weight ratio greater than 7.6:1, free chlorine is the predominant chlorine residual.

The reaction rate of breakpoint chlorination is determined by the formation and decay rates of dichloramine, reactions that are highly dependent on pH. Ideally, the reaction takes place at a pH in the range of 7.0-8.0 (Kirmeyer et al., 2004). The theoretical Cl2:NH3-N weight ratio for breakpoint chlorination is 7.6:1; the actual Cl2:NH3-N ratio varies from 8:1 to 10:1, depending on pH, temperature and the presence of reducing agents. The presence of iron, manganese, sulphide and organic chlorine demand compounds will compete with the free chlorine added, potentially limiting the chlorine available to react with ammonia (Kirmeyer et al., 2004; AWWA, 2006; Muylwyk, 2009). A contact time of 30 minutes or longer is necessary for the reaction to go to completion (Kirmeyer et al., 2004; Hill and Arweiler, 2006). The breakpoint ratio should be determined experimentally for each water supply (Hill and Arweiler, 2006).

Chlorine compounds certified to NSF International (NSF)/American National Standards Institute (ANSI) Standard 60 should respect the maximum use limit (MUL) stated in the standard. This ensures that any potential trace contaminants do not exceed their respective health-based limits even if used at the maximum dose stated for the additive. The responsible authority may choose to allow a utility to exceed the MUL in order to achieve breakpoint chlorination and disinfection goals. As exceeding the MUL could invalidate the certification, consultation with the body that has certified the chlorine compound is recommended. This will help ascertain what potential trace contaminants might be present and thus help determine what additional monitoring of hypochlorite-related contaminants might be triggered when the MUL is exceeded. As breakpoint chlorination requires relatively high concentrations of chlorine, this can cause other problems, such as the formation (or increased concentrations) of disinfection by-products in the presence of organic matter. However, efforts to limit the formation of disinfection by-products must not compromise the effectiveness of disinfection. An advanced treatment, such as GAC adsorption, may be considered following breakpoint chlorination to remove resulting taste and odour compounds as well as chlorination by-products (Janda and Rudovský, 1994; Wilczak, 2006a).

Breakpoint chlorination, relative to nitrification control in the distribution system, is not considered an effective long-term strategy. Utilities should consider more permanent control strategies, such as changes in operation or engineering improvement (Kirmeyer et al., 1995; Hill and Arweiler, 2006).

7.1.3 Ion exchange

Ion exchange is a physicochemical process that employs an exchange of ions (cations or anions) in the water to be treated with ions sorbed at the solid phase of the natural or synthetic resins. Cation exchange is capable of removing ammonia from drinking water.

Studies have investigated natural zeolites, such as clinoptilolite, bentonite, sepiolite and mordenite (Hodi et al., 1995; Demir et al., 2002; Park et al., 2002; Weatherley and Miladinovic, 2004; Wang et al., 2007), and synthetic resins (Lin and Wu, 1996; Abd El-Hady et al., 2001) for the removal of ammonium ions from water. Factors such as pH, pretreatment of the natural zeolites, media particle size, influent ammonium concentration and competing cations, such as calcium, magnesium and potassium, in the water affect the efficiency of ammonium removal. Ion exchange processes do not result in a constant percentage of removal of contaminants (e.g., ammonium ion) with time, because they will break through as the resin reaches its capacity. Once the resin's capacity is reached, contaminant concentrations will increase in the finished water, and the resin must be regenerated. Ion exchange technology may be inconvenient for a treatment plant with a capacity above 80 000 m3/day because of the large footprint required for the ion exchange columns (Kurama et al., 2002).

Clinoptilolite is the most abundant natural zeolite and has been shown to have a high selectivity for ammonium ion. Although it has been applied primarily in wastewater treatment, this technique has recently been studied for the reduction of ammonium concentrations in drinking water. Studies reported that the pretreatment of natural clinoptilolite increased both the ion exchange capacity of the clinoptilolite and the ammonium removal efficiency in aqueous solutions (Haralambous et al., 1992; Turan and Celik, 2003; Vassileva and Voikova, 2009; Siljeg et al., 2010).

Laboratory-scale and pilot-scale cation exchange experiments have been shown to reduce ammonia concentrations in drinking water. This technology seems to be effective when natural zeolites are used as the cation exchange material and the water has a low hardness (Haralambous et al., 1992; Weatherley and Miladinovic, 2004).

A pilot-scale study (Gaspard et al., 1983) evaluated the capability of clinoptilolite to remove ammonium ions in tap water. An average influent concentration of 2.25 mg NH4+/L (1.75 mg NH4+-N/L) was reduced to a predefined breakthrough level of 0.5 mg NH4+/L (0.39 mg NH4+-N/L), achieving an ion exchange capacity of 0.108 milliequivalents of ammonium ion per gram of clinoptilolite (1.47 mg NH4+-N/g) and 750 bed volumes (BV).

A laboratory column study using sodium clinoptilolite (Na+- clinoptilolite) achieved an exchange capacity of 0.47 mg NH4+/g clinoptilolite (0.37 mg NH4+-N/g) and 600 BV at pH of 8.26. An average influent concentration of 0.86 mg NH4+/L (0.67 mg NH4+-N/L) in groundwater was reduced to 0.15 mg NH4+/L (0.12 mg NH4+-N/L) (Hodi et al., 1995).

In another laboratory study, Weatherley and Miladinovic (2004) evaluated the performance of Na+-clinoptilolite and Na+-mordenite for ammonium removal from aqueous solution. The experiments were conducted with feed concentrations from 1.0 mg/L NH4+/L (0.78 mg NH4+-N/L) to 200.0 mg NH4+/L (155.6 mg NH4+-N/L) while maintaining the pH below 7.5. Equilibrium data demonstrated that Na+-clinoptilolite achieved a 98.8% reduction of an influent concentration of 10 mg NH4+/L (7.8 mg NH4+-N/L), in the absence of other ions in solution. However, in the presence of 40 mg/L each of calcium, magnesium and potassium, the resin achieved reduction of 93.7%, 94.7% and 95.9% of ammonia, respectively. Similarly, equilibrium data for Na+-mordenite showed that a reduction of 92.3% of an influent concentration of 10 mg NH4+/L (0.78 mg NH4+-N/L) was achieved in the absence of other ions in solution. However, in the presence of 40 mg/L each of calcium, magnesium and potassium, Na+-mordenite achieved 91.8%, 92.2% and 86.3% ammonia reductions, respectively. The presence of calcium, magnesium and potassium thus decreased the ammonium removal efficiency for both zeolites (Weatherley and Miladinovic, 2004).

Laboratory column tests (Turan and Celik, 2003) studied the impact of ammonia (form not specified) concentration on the ion exchange capacity of clinoptilolite and the effectiveness of clinoptilolite regeneration on column performance. The results showed that an increase in influent ammonia concentrations decreased the ammonia reductions. Initial concentrations of 10, 15 and 20 mg/L were reduced by 96%, 94% and 87%, respectively, after 12 hours of operation. The study reported that natural clinoptilolite achieved a 65.0% reduction of an initial ammonia concentration of 10 mg/L after 23 hours of operation, whereas twice-regenerated clinoptilolite achieved a 98.0% reduction under the same operating conditions.

Abd El-Hady et al. (2001) evaluated a synthetic strong acid cationic resin for removing ammonium ions in laboratory experiments. Three initial ammonium concentrations of 10 mg NH4+/L (7.8 mg NH4+-N/L ), 5 mg NH4+/L (3.9 mg NH4+-N/L) and 2 mg NH4+/L (1.6 mg NH4+-N/L) were reduced to below a predefined breakthrough concentration of 0.5 mg/L. Adsorption capacities of 0.156 mol/L (2.2 mg NH4+-N/ml resin), 0.085 mol/L (1.2 mg NH4+-N/ml resin) and 0.0317 mol/L (0.4 mg NH4+-N/ml resin) and BVs of 295, 340 and 380 were reported for the above three initial concentrations, respectively.

The major considerations when using ion exchange treatment include chromatographic peaking, disposal of the resin regenerant (Clifford, 1999) and the possible increased corrosivity of the treated water (Schock and Lytle, 2010). Regeneration results in a brine waste stream that contains high ammonium concentrations and must be disposed of appropriately, thus increasing the cost of this process. The exchange of ions can cause mineral imbalances that could increase the corrosive nature of the treated water (Schock and Lytle, 2010). In some cases, post-treatment corrosion control measures may need to be taken, to ensure that corrosion problems do not occur following treatment.

7.1.4 Membrane filtration

The available scientific information on the removal of ammonia from water supplies by membrane technologies is limited. These processes are based on forcing water across a membrane under pressure while the ionic species, such as ammonium, are retained in the waste stream. Reverse osmosis (RO) treatment systems typically require pre-filtration for particle removal and often include other pretreatment steps, such as the addition of anti-scaling agents, prechlorination/dechlorination and softening. Post-treatment steps typically include pH adjustment, corrosion inhibitor addition and disinfection (Cevaal et al., 1995).

RO and, to a lesser extent, nanofiltration (NF), can be effective technologies for reducing ammonia concentrations in drinking water (Koyuncu et al., 2001; Koyuncu, 2002; Kurama et al. 2002; Quail, 2008).

Koyuncu (2002) conducted a pilot-scale study to evaluate the effectiveness of nanofiltration and low-pressure reverse osmosis (LPRO) membranes for ammonia removal under different operating parameters. A spiral wound module was operated at feed influent ammonia concentrations in the range of 10-15 mg/L. The LPRO membrane showed a higher performance than the nanofiltration membrane under the tested conditions. The rejection of ammonia was increased with an increase in the pressure for both membranes. The LPRO membrane was capable of rejecting from 90% to 95% of ammonia concentrations using a feed pressure in the range of 3 to 6 bar (43.5-87.0 pounds per square inch [psi]) and temperature in the range of 15-25ºC. The nanofiltration membrane achieved up to 90% rejection at the same tested conditions. At temperatures above 25ºC, the LPRO membrane showed a decrease of the rejection rate, whereas the rejection rate of the nanofiltration membrane was slightly affected (Koyuncu, 2002). Both membranes had a negative charge at neutral and high pH and neutral or slightly positive charge at low pH. The study reported that a neutral pH was optimal for ammonia rejection by both membranes.

An earlier pilot-scale study by Koyuncu et al. (2001) evaluated the efficiency of brackish water reverse osmosis (BWRO) and saline water reverse osmosis (SWRO) membranes for the removal of ferroammonium sulphate (Fe.(NH4OH).(SO4)2.H2O) and ammonium hydroxide (NH4OH) in aqueous solution. Both membranes demonstrated a higher rejection rate for the ammonium complex than for ammonium hydroxide. The SWRO membrane was capable of rejecting 99% and 90% of the ammonium complex from influent concentrations of 8.5 mg/L and 100 mg/L, respectively. The BWRO membrane achieved a rejection rate of 96% and 83% of the ammonium complex from influent concentrations of 10 mg/L and 30 mg/L, respectively. The SWRO membrane was capable of rejecting 30-40% of ammonium hydroxide from influent concentrations of 2-90 mg/L. The BWRO membrane achieved rejection rates in the range of 10-20% of ammonium hydroxide from influent concentrations ranging from 1 to 186 mg/L. The study also reported rejection rates of 95% and 60% of ammonia in surface water by BWRO and SWRO membranes, respectively. The feed water had influent concentrations in the range of 3-4 mg/L and neutral pH. No fouling and no reduction in flux were observed during the 5-hour experimental runs (Koyuncu et al., 2001).

Laboratory testing of RO membranes found that two membranes (Desal-3LP and Desal-3b SE, Osmonics) were capable of effectively rejecting NH4+. An average rejection of 95% was achieved from 6.5 mg NH4+/L (5.05 mg NH4+-N/L) as a feed concentration. The study has found that tested nanofiltration membranes achieved up to 26% removal (Kurama et al. 2002).

Considerations when using RO treatment include disposal of the reject water and the possible increased corrosivity of the treated water (Schock and Lytle, 2011). RO rejects a significant portion of the influent water as contaminant-rich brine (Taylor and Wiesner, 1999). The concentrate discharge must be considered and disposed of appropriately. The removal of contaminants can cause mineral imbalances that could increase the corrosive nature of the treated water (Schock and Lytle, 2011). In some cases, post-treatment corrosion control measures may need to be taken.

7.1.5 Combination of reverse osmosis and biological treatment

Nagy and Granlund (2008) Quail (2008) and McGovern and Nagy (2010) presented a combined process of an RO (75% water treated) system and biological treatment (25% water treated) to remove inorganic contaminants found in groundwater simultaneously and to address copper corrosion control. The maximum design capacity of the water treatment plant was 6.5 MGD (24 605 m3/day). A spiral wound polyamide thin film composite RO membrane was capable of reducing of an ammonia concentration of 2.0 mg NH4+-N/L to 0.08 mg NH4+-N/L, achieving greater than 96% reduction at a system recovery of 82% and a feed pressure of 130 psi. Pretreatment included pH adjustment and addition of anti-scalant. While maintaining a dissolved oxygen concentration of 3 mg/L, the ammonia biological filters were seeded with backwash water from the existing wastewater plant. The nitrification process, established within 2 months, was capable of reducing an average influent ammonia concentration of 2.0 mg NH4+-N/L to an average effluent concentration of less than 0.16 mg NH4+-N/L, achieving reduction greater than 92% (McGovern and Nagy, 2010). The filters operated at a hydraulic loading rate of 4 gpm/ft2 (9.8 m/h) and had a media depth of 1.5 m. Blended water had an ammonia concentration of less than 0.16 mg NH4+-N/L.

7.1.6 Air stripping

Although air stripping is a common practice for removing ammonia from wastewater, its treatment efficiency for drinking water is expected to be marginal due to the low Henry's Law constant (0.0006 at 20ºC) in combination with relatively low concentrations of ammonia encountered in source water (Crittenden et al., 2005).

Separation of non-ionized ammonia from water can be achieved with air stripping in a packed tower by raising the pH of the water above 10 and increasing the temperature. As ammonia is soluble in water, a high air to water ratio is required; pH adjustment after the aeration is required for subsequent processes (U.S. EPA, 2000).

7.1.7 Emerging technologies

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

  • Trickling filters: A pilot-scale study evaluated trickling filters for simultaneous biological removal of ammonia, iron and manganese from potable water. Influent ammonia concentrations in the range of 0.5-3.0 mg/L were reduced up to 82% in the finished water under a variety of operating conditions (Tekerlekopoulou and Vayenas, 2007, 2008).
  • Electrochemical removal: A pilot-scale charge barrier capacitive deionization process is reported as effective in removing total dissolved solids, nitrate and ammonia from water. The process employs an adsorption of ions on the surface of two oppositely charged electrodes. The process achieved ammonia removal up to 88.1% at 1000 mg/L as feed concentration (Broseus et al., 2009).
  • Submerged membrane bioreactors:Although the use of membrane bioreactors has been applied primarily in wastewater treatment, this technique has recently been considered as a new technology in drinking water treatment. Laboratory studies examined the effectiveness of hollow fibre membrane modules directly immersed inside the activated sludge reactors for ammonia removal. Removal efficiencies in the range of 89-98% were achieved by the submerged membrane bioreactors through biological nitrification.The influent NH3-N concentrations have been reported in the range of 2.00-4.24 mg/L (Li and Chu, 2003; Tian et al., 2009).

7.1.8 Nitrification in the distribution system

One of the main concerns related to the presence of ammonia in drinking water is the potential for the formation of nitrite and nitrate, parameters with health risks and drinking water quality guidelines. Nitrite and nitrate are the products of nitrification, a two-step process that oxidizes ammonia either in natural water or in water that has been disinfected by chloramine. The occurrence of nitrification in chloraminated distribution systems has been well documented (Skadsen, 1993; Odell et al., 1996; Wilczak et al. 1996). According to Kirmeyer et al. (1995) and Wilczak et al. (1996), nitrification may occur at 63% of utilities that use chloramine as a secondary disinfectant. In a survey of 56 utilities using monochloramine, 48% of them reported that they had experienced nitrification (Kirmeyer et al., 2004).

Nitrification can occur irrespective of pipe material--plastic, polyvinyl chloride, asbestos-cement, ductile iron and cast iron. Certain pipe materials, such as unlined cast iron pipes or old mortar-lined iron pipes, may provide more favourable conditions for nitrification to occur (Cohen et al., 2001). Accumulated sediment and biofilm can protect the ammonia-oxidizing bacteria from chloramine residual. Higher concentrations of ammonia-oxidizing bacteria were detected in reservoir and pipe sediment materials than in pipe biofilm samples (Wolfe et al., 1990).

Nitrification in the distribution systems can have adverse impacts on water quality. These impacts include increased nitrite and nitrate levels, reduced chloramine residuals, increased bacterial regrowth (i.e., increased heterotrophic plate count [HPC], with a possible detection of Escherichia coli), as well as a reduction of pH and dissolved oxygen (Kirmeyer et al., 1995, 2004; Odell et al., 1996; Wilczak et al., 1996; Bremer et al., 2001; U.S. EPA, 2002; Lytle et al., 2007; Muylwyk, 2009; Zhang et al., 2009). Studies have also reported a link between corrosion problems and nitrification (Edwards and Dudi, 2004; Douglas et al., 2004; Zhang et al., 2008, 2010).

The potential increase of nitrite in the distribution system due to nitrification is significant, as it may exceed 1 mg/L NO2-N. However, when nitrite concentrations increase as a result of nitrification, the primary concern for utilities is that nitrite consumes chlorine and decomposes chloramines, which results in an increase in microbial counts, including an increase in the potential presence of coliform bacteria in the distribution system (Baribeau, 2006; Smith, 2006). Harrington et al. (2002) and the U.S. EPA (2002) noted that increases in nitrite up to 1 mg NO2-N/L due to nitrification could theoretically occur in any system in which the total ammonia concentration entering the distribution system is greater than 1 mg-N/L.

Factors contributing to nitrification in the distribution system include warm water temperatures, pH, a low Cl2:NH3-N ratio and the concurrent increase of free ammonia concentrations and chloramine residual. A number of distribution system parameters, such as detention time, reservoir design and operation, sediment and tuberculation in piping, biofilm and the absence of sunlight, can affect the nitrification (Skadsen, 1993; Kirmeyer et al., 1995, 2004; U.S. EPA, 1999; Lytle et al., 2007; Fleming et al., 2008; Baribeau, 2010).

The optimum temperature for nitrifiers to grow ranges between 20°C and 30°C (Baribeau, 2006); however, regrowth and nitrification can occur at temperatures as low as 5°C or even less in systems with long detention times (Pintar et al., 2000). Kors et al. (1998) discussed a case of nitrification under extreme cold-water conditions (below 4°C). The increase in temperature will increase the chloramine decomposition rate, which will promote nitrification, as more free ammonia will be released (Baribeau, 2006).

Although the optimum pH range for nitrifiers to grow is 7.5-8.0, nitrification can occur at pH 6.6-9.8 (Kirmeyer et al., 1995; Odell et al., 1996; Wilczak et al., 1996; Baribeau, 2006; Wilczak, 2006b). The pH may decrease during nitrification in low-alkalinity water. If the pH decreases below 8.0, chloramine decomposition may be accelerated. The pH data should be evaluated carefully, because pH may vary throughout the system depending on factors other than nitrification such as corrosion. Theoretical oxygen concentration (O2) required for biological oxidation of 1 g NH4+-N to NO2--N is 3.22 g O2, and 1.11 g O2 to oxidize 1 g NO2--N to NO3- -N. Thus the total theoretical O2 requirement is 4.33 g O2 to oxidize 1 g NH4+-N to NO3- -N (Baribeau, 2006).

The initial Cl2:NH3-N weight ratio used to form monochloramine (the preferred chloramine species) affects the level of the free ammonia available in the distribution system (Fleming et al., 2005, 2008). Free ammonia may enter the distribution system from the treatment plant due to the overdosing of ammonia or incomplete reaction with free chlorine. The measurement of free chlorine immediately upstream of the point of ammonia addition is critical to the proper dosing of ammonia at the treatment plant. Minimizing free ammonia entering the distribution system is extremely important (Cohen and Friedman, 2006; Wilczak, 2006b). A weight ratio of Cl2:NH3-N should generally be maintained between 4.5:1 and 5:1 in the plant effluent to enhance the formation of monochloramine and reduce the concentration of free ammonia entering into the distribution system (Harrington, 2003; Kirmeyer, 2004; Skadsen and Cohen, 2006). However, the water quality parameters and utility-specific chlorine demand must be considered when selecting the target ratio (Skadsen and Cohen, 2006). Kirmeyer et al. (2004) and Skadsen and Cohen (2006) suggested that minimizing free ammonia entering the distribution system to a concentration below 0.1 mg NH3-N/L and preferably below 0.05 mg NH3-N/L is an important optimization goal to reduce the potential for nitrification.

When free chlorine is the desired residual disinfectant in the distribution system, the removal of naturally occurring ammonia in the source water is beneficial to reduce chlorine demand and avoid chloramine formation. It is important to be aware that monochloramine may interfere with the N,N-diethyl-p-phenylendiamine (DPD) method used to monitor free chlorine and can create a false positive reading (Smith, 2006; Pon, 2008). For utilities practicing chloramination, it is important to take into consideration the ammonia concentration in the source water when establishing the ammonia dosage for chloramine formation (Skadsen and Cohen, 2006; Muylwyk, 2009; Shorney-Darby and Harms, 2010). Wolfe et al. (1990) reported that using Cl2:NH3-N ratio of 3:1 results in approximately 0.2 mg/L free ammonia when maintaining a total chlorine concentration of 1.5 mg/L in the distribution system. Bouwer and Crowe (1988) demonstrated that an ammonia-nitrogen concentration of 0.25 mg/L would promote the growth rate of nitrifying organisms in both the treatment plant and the distribution system. An optimization of Cl2:NH3-N ratio should ensure that Health Canada's guideline for chloramines is not exceeded (Health Canada, 1995).

Although chloramine is more stable than free chlorine, it decomposes and releases free ammonia. An understanding of chloramine chemistry is critical in order to maintain chloramine residual, prevent the release of free ammonia in the distribution system and prevent or control nitrification. The rate of chloramine residual loss in the distribution system is affected by reactions with natural organic matter (NOM) and inorganic constituents (chloramine demand) and a combination of hydrolysis and acid-catalysed disproportionation reactions (chloramine decay). Chloramine demand and decay in the distribution system release free ammonia, which, along with the ammonia entering the system, provides substrate for ammonia-oxidizing bacterial growth and promotes nitrification (Skadsen, 1993; Vikesland et al., 2001, 2006; Kirmeyer et al., 2004; Chowdhury et al., 2006; Wilczak, 2006b). Chlorine/chloramine demand should be satisfied as much as possible within the treatment plant, and chloramine decay should be minimized in the distribution system, as these reactions increase the free ammonia concentration in the distribution system and trigger nitrification (Baribeau, 2006; Wilczak, 2006b). It is important to note that even the stringent control of excess free ammonia and the maintenance of a proper Cl2:NH3-N ratio may not always be effective in preventing nitrification. This is due to the fact that chloramine in the distribution system will start to decay based on water quality conditions and water age, releasing free ammonia into the water (Cohen and Friedman, 2006).

The presence of bromide in chloraminated water complicates system chemistry by reacting with chlorine and chloramine species to form bromamines. The bromamines are capable of accelerating chloramine decay and may also be able to combine with organic contaminants to form halogenated organics, which remain poorly understood to date (Vikesland et al., 2001; Kirmeyer et al., 2004).

N-Nitrosodimethylamine (NDMA) is a nitrogen-containing disinfection by-product that may be formed during the treatment of drinking water, particularly during chloramination and, to a lesser extent, chlorination (Richardson, 2005; Charrois and Hrudey, 2007; Nawrocki and Andrzejewski, 2011). The key to controlling the formation of NDMA lies in limiting its precursors, including dichloramine. As such, optimization and control of free ammonia are important elements in preventing NDMA formation. Additional information on NDMA is available in the Guideline Technical Document on NDMA (Health Canada, 2011).

A research study (Kirmeyer et al., 1995) based on literature reports, case studies, an analytical survey and a phone survey of large chloraminated systems obtained conflicting results regarding the water quality and the treatment factors that affect nitrification episodes. In combination with the distribution system hydraulics, the importance of one factor over another factor causing nitrification was specific to each system. In general, free ammonia promotes nitrification in the distribution system and is available either through ammonia feed overdose or through release of free ammonia from chloramine demand and decay (Kirmeyer et al., 1995).

The treatment plant, the distribution systems and storage facilities all require monitoring for specific parameters. Parameters that can be monitored for potential causes of nitrification include chloramine residual, Cl2:NH3-N ratio, free ammonia concentration entering the distribution system, pH and temperature. Products of nitrification that can be monitored include nitrite/nitrate and HPC at the entry point of the distribution system and throughout the system (Odell et al., 1996; Wilczak et al., 1996).

The concentration of free ammonia entering the distribution system and at key locations in the system, such as storage facilities and areas with long water detention times (e.g., dead ends), in addition to parameters such as total chlorine residual and nitrite, is a very useful parameter to monitor for nitrification control. In particular, Smith (2006) suggested that a free ammonia concentration greater than 0.1 mg NH4+-N/L at storage facilities can be used as an indicator of nitrification requiring further investigation (i.e., alert level).

A site-specific evaluation is necessary to establish a nitrification monitoring program. The program should identify system-specific alert and action levels, which can be used to determine the appropriate level of nitrification response. The monitoring frequency of the parameters depends on the location and the purpose of the data. Distribution system nitrification parameters considered to be of higher priority are total chlorine residual, nitrite and nitrate. Changes in the trend of these nitrification parameters should trigger more frequent monitoring of other parameters, such as free ammonia.

There are several preventive and corrective measures that can be taken to address nitrification (AWWA, 2006). Preventive measures include:

  • Control of water quality parameters (pH, free ammonia entering the distribution system, organic matter) and operating parameters (Cl2:NH3-N weight ratio and chloramine residual):
    • Establishing the proper pH level is essential for maintaining chloramine residual in the distribution system and limiting nitrification (Wilczak, 2006b).
    • A minimization of free ammonia entering the distribution system to concentrations below 0.1 mg NH3-N /L and preferably below 0.05 mg NH3-N/L is an important optimization goal to reduce the potential for nitrification (Kirmeyer et al., 2004).
    • In general, chloramine residuals, greater than 2.0 mg/L (leaving the treatment plant) appear to be effective in preventing nitrification by limiting the growth of ammonia-oxidizing bacteria (Kirmeyer et al., 1995; Odell et al., 1996; U.S. EPA, 1999; Harrington et al., 2003). The chloramine residual concentration leaving the treatment plant will vary depending on the size of the distribution system and the water quality characteristics (U.S. EPA, 1999; Skadsen and Cohen, 2006). However, once nitrification is under way, the high chloramine residual (up to 8 mg/L) may not control nitrification (Skadsen, 1993). Increasing the chloramine concentration during a nitrification event may exacerbate the process, because it leads to an increase in the level of free ammonia as a result of chloramine decay (Woolschlager et al., 2001; Harrington et al., 2003; Hill and Arweiler, 2006).
  • Corrosion control programs: These may help minimize pipe biofilms and sediment, limit attachment of microorganisms, reduce the reaction between chloramine and corrosion products and thus reduce chloramine demand (Wilczak, 2006b).
  • Distribution system pipe flushing: Sediment flushing in the pipe network, reservoir turnover and cleaning will prevent or delay the onset of nitrification (Hill and Arweiler, 2006; Wilczak, 2006b). However, once nitrification occurs, flushing alone may be limited in effectiveness (Skadsen and Cohen, 2006).
  • Booster chlorination or chloramination stations: Attention is given to recombining the released (increased) free ammonia in the distribution system by booster chlorination to maintain the ratio near 5:1 throughout the system (Wilczak, 2006b). Free ammonia residual needs to be measured before chemical addition. If sufficient free ammonia is still present, only chlorine needs to be added.
  • Temporary/seasonal free chlorination (breakpoint chlorination): Periodic switching to free chlorine is a preventive and/or effective control method practised by water utilities. However, a temporary switch to free chlorination in the distribution system has been associated with numerous problems, including a temporary increase in HPC, coliform-positive samples (related to the sloughing of existing biofilm layers) (Odell et al., 1996), potential taste and odour problems, and potential disinfection by-product problems (Skadsen, 1993; Hill and Arweiler, 2006). Studies by Kirmeyer et al. (1995) and Odell et al. (1996) suggested that a return to chloramination following a free chlorination period led to subsequent nitrification within a short period.
  • Chlorite addition: It appears that chlorite is effective for nitrification prevention (McGuire et al., 1999; Baribeau, 2006; Wilczak, 2006b). The latest research demonstrates that chlorite addition is less effective in areas where nitrification has been substantially developed before the chlorite application. Chlorite application prior to nitrification development is a strategy for nitrification prevention for utilities with significant seasonal changes in their finished water temperature (McGuire et al., 2009; Zhu et al., 2010). However, chlorite addition is considered to be controversial, as chlorite is a regulated contaminant, and its presence can also lead to the formation of chlorate (Skadsen and Cohen, 2006). Utilities wishing to use chlorite addition as a control strategy should ensure that the Guidelines for Canadian Drinking Water Quality for chlorite and chlorate (Health Canada, 2008b) are not exceeded.

Corrective measures are similar to the preventive measures and include:

  • distribution system pipe flushing;
  • temporary/seasonal free chlorination (breakpoint chlorination);
  • reservoir cycling to limit water age. During severe nitrification episodes, reservoir cleaning, as well as drainage and disinfection, may be needed; and
  • chlorite addition.

The different measures used to control the nitrification episodes vary in their effectiveness and their ability to provide long-term improvements in nitrification problems. For these reasons, comprehensive strategies aimed at the prevention of nitrification episodes are recommended over strategies aimed at controlling nitrification as it occurs. Any strategy should also ensure that the relevant Guidelines for Canadian Drinking Water Quality (e.g., chloramines) are not exceeded. Detailed information on nitrification control and prevention measures is available in reports and reviews by Kirmeyer et al. (1995), AWWA (2006) and Zhang et al. (2009).

7.2 Residential scale

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 reducing ammonia concentrations in drinking water. Although no certified residential treatment devices are currently available for the reduction of ammonia levels in drinking water, treatment devices using reverse osmosis or ion exchange may be effective for the reduction of ammonia concentrations in drinking water.

Before a treatment device is installed, the water should be tested to determine general water chemistry and verify the presence and concentration of ammonia. 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 use 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 strongly recommends that consumers use devices that have been certified by an accredited certification body as meeting the appropriate NSF/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, 2011):

  • Canadian Standards Association International (www.csa-international.org);
  • NSF International (www.nsf.org);
  • Water Quality Association (www.wqa.org);
  • Underwriters Laboratories, Inc. (www.ul.com);
  • Quality Auditing Institute (www.qai.org); and
  • International Association of Plumbing & Mechanical Officials (www.iapmo.org).
  • An up-to-date list of accredited certification organizations can be obtained from the SCC (www.scc.ca).

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