Priority Substances List Assessment Report for Inorganic Chloramines

2.0 Summary of Information Critical to Assessment of "Toxic" under CEPA 1999

2.1 Identity, properties, uses, formation and analytical methods

2.1.1 Identity and uses

Inorganic chloramines consist of three chemicals that are formed when chlorine and ammonia are combined in water: monochloramine, dichloramine and trichloramine. Inorganic chloramines, free chlorine and organic chloramines are chemically related and are easily converted into each other; thus, they are not found in isolation. The chemistry of inorganic chloramines is complex. Site-specific conditions determine the dominant chlorine species that will be formed.

Synonyms for chloramine include chloroamine, combined available chlorine (CAC) and combined residual chlorine (CRC). CRC and CAC include both inorganic and organic chloramines (Kirk-Othmer, 1979).

Organic chloramines, or organochloramines, are a group of perhaps thousands of substances formed via the reaction of free chlorine and inorganic chloramines with various amino acids, peptides or proteins. Although a rigorous evaluation of organic chloramines was beyond the scope of this assessment, the current state of scientific knowledge on organic chloramines is presented in a separate supporting document (see El-Farra et al., 2000).

Total residual chlorine (TRC) includes both CRC and free chlorine. Free chlorine is also called free residual chlorine (FRC) or free available chlorine (FAC), and it includes hypochlorous acid and the hypochlorite ion. Table 1 lists the Chemical Abstracts Service (CAS) registry numbers, molecular formulas and molecular weights for inorganic chloramine and free chlorine compounds.

In the presence of bromide, such as in seawater and some groundwaters, reactive chlorine atoms can be completely or partially replaced by bromine atoms. The collection of reactive chlorine and bromine species is called chlorine-produced oxidants (CPO) or total residual oxidants (TRO).

Inorganic chloramines are formed in wastewaters and cooling waters as a result of a series of reactions that occur when free chlorine is added in the presence of sufficient amounts of aqueous ammonia. The proportion of each residual chlorine species depends on the amount of chlorine added, the conditions present in the water/influent as well as the chlorine contact time.

In potable water, inorganic chloramines (predominantly monochloramine) are intentionally produced, usually as a secondary disinfectant, for several reasons, including:

• persistence and the ability to maintain a chlorine residual or disinfection capability throughout the distribution system, thereby eliminating costs for water supply services associated with the need for chlorine booster stations;
• adequate disinfection capability for the protection of humans from pathogenic and other microorganisms;
• ability to penetrate and prevent biofilm accumulation in municipal distribution systems;
• lower tendency to form trihalomethanes (THMs) and other disinfection by-products in comparison with free chlorine; and
• minimization of objectionable tastes and odours (AWWARF, 1993).

Table 1 Chemical description of combined and free chlorine congeners

Item	Inorganic chloramine compounds			Free chlorine compounds
	Monochlor amine	Dichlor amine	Trichlor amine	Hypochlorous acid	Hypochlorite ion
CAS No.	10599-90-3	3400-09-0	10025-85-1	7790-92-3	-
Molecular formula	NH2Cl	NHCl2	NCl3	HOCl	OCl-
Molecular weight	51.48	85.92	120.37	52.46	51.45

Chlorination is also used to disinfect wastewater. Disinfection involves the oxidation of reactive organic material and the reduction or elimination of bacteria, viruses and protozoa by the chlorine residual. Consideration of disinfection efficiency, ease of application and cost has led to the use of chlorine as a primary disinfectant in food processing, seafood production and municipal wastewater treatment.

Chlorine and chloramine are used as biocides to reduce biofouling in water cooling towers and cooling systems of electrical generating stations, as well as at desalination, petrochemical, paint and metal fabricating facilities. Chlorination is also used as a treatment to remove slime and algae, bacteria and their extracellular excretions or to eliminate hydroids, barnacles, mussels, clams and oysters at water intakes for electrical generating facilities (Coulston et al., 1994).

2.1.2 Chemistry of chloramine formation

A variety of physical and chemical properties for chlorine species are presented in Table 2.

The formation of inorganic chloramines is adequately described using reactions involving aqueous chlorine and ammonia. In general, the reactions are governed by two main parameters: pH and the ratio of chlorine to ammonia. Time of reaction and temperature are also factors in these reactions. A thorough review of inorganic chloramine formation is presented in the supporting document prepared by El-Farra et al. (2000). Salient features of chloramine chemistry are presented below.

In aqueous solution, chlorine (Cl₂) is present as either hypochlorous acid (HOCl) or hypochlorite ion (OCl^-), although it is still usually referred to as chlorine. Similarly, ammonia (NH₃) may also be present as the ammonium ion (NH₄⁺) but is usually referred to as ammonia. The relative amounts of each of these species are determined by the pH of the water and the ionization constants for chlorine and ammonia (pK_a,HOCl = 7.54; pK_a,NH3 = 9.3; Snoeyink and Jenkins, 1980). The result of the reaction of these species is the formation of chlorinated species of ammonia:

monochloramine (NH₂Cl), dichloramine (NHCl₂) and trichloramine (NCl₃) (White, 1992).

Simplified reactions are often used to illustrate the complex effect of the chlorine-to-ammonia ratio on inorganic chloramine formation:

In looking at the reactions from left to right (the forward reactions), as the ratio of chlorine (hypochlorous acid) to ammonia increases, greater chlorine substitution is observed. These are equilibrium reactions, with the two arrows indicating that both forward a nd reverse reactions occur to an appreciable extent. Thus, there will always be at least small amounts of the materials shown on both sides of each equation present in solution. Also, the reverse reactions indicate that if chloramines are added to water, certain amounts of free chlorine and different chloramines or ammonia will be generated. These reverse reactions are called hydrolysis reactions because water is involved.

Table 2 Physical and chemical properties of combined and free chlorine congeners ¹

Parameter	NH2Cl	NHCl2	NCl3	HOCl	Cl2
Physical state at STP	Liquid	Liquid	Liquid	Liquid	Gas
Colour	Colourless	n/a	Bright yellow	Green-yellow	Yellow-green
Boiling point (°C)	n/a 2	n/a	70	n/a	-34.05
Melting point (°C)	-66	n/a	-40	n/a	-100.98
Water solubility	Soluble	Soluble	Limited to hydrophobic	7290 mg/L	Slightly soluble (1% at 9.6°C)
pK	14 ± 23	7 ± 3	n/a	7.40-7.54	n/a
Henry's law constant (Pa·m3/mol)	557 ± 61 at 20°C4	n/a	Very volatile 5	n/a	n/a
Other comments	May explode at room temperature; most stable in aqueous solution	Pungent odour	Powerful, irritating odour, highly explosive, extremely hydrophobic	Stable only in aqueous solution	Pungent odour

In reality, chloramine chemistry entails a complex series of reactions involving the species and pathways identified in Figure 1 and Figure 2. At hypochlorous acid-to-ammonia molar ratios of greater than approximately 1.5:1 to 2:1, oxidation reactions occur in addition to substitution reactions, with the net result being a decrease in the concentrations of chloramine species present. For the relatively neutral to slightly basic conditions encountered in most aquatic receiving environments, monochloramine and/or dichloramine are generally the chloramine species detected in greatest concentration. At pH values above 8 and hypochlorous acid-to-ammonia molar ratios of 1:1 or below, monochloramine is the only observed chloramine. Dichloramine and trichloramine are formed at higher molar ratios and at lower pH values. In slightly acidic water or when the hypochlorous acid-to-ammonia molar ratio is greater than 1:1, dichloramine may account for the largest fraction of the total chlorine concentration. However, trichloramine is the only chloramine observed below pH 3. Trichloramine proportions diminish up to pH 7.5 at hypochlorous acid-to-ammonia ratios greater than 2:1. Above pH 7.5, trichloramine is not detectable at any hypochlorous acid-to-ammonia ratio.

Figure 1: Summary of ammonia/chlorine reactions (adapted from Jafvert and Valentine, 1992)

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Certain wastewaters and cooling waters are "breakpoint-chlorinated" prior to discharge, which results in a discharge with low ammonia and residual chlorine concentrations. Under optimal conditions, the ammonia is completely oxidized to nitrogen gas (N2), and the chlorine is reduced to chloride ion (Cl-). The breakpoint phenomenon occurs quickly (within a few minutes) and to a significant extent in the pH range 6-9, the pH range for most natural waters (see Figure 3). It is generally described as a relationship between the ratio of chlorine to ammonia applied and the resulting TRC concentration (see Snoeyink and Jenkins, 1980; Montgomery, 1985; Metcalf and Eddy, Inc., 1991; Stumm and Morgan, 1996).

Figure 3 shows that increases in both the monochloramine concentration and combined chlorine residual occur for increases in the chlorine-to-ammonia molar ratio from 0:1 up to a maximum at approximately 1:1. It is generally thought that this is the region of the breakpoint curve that most often represents the mix of chlorine and ammonia in chlorinated receiving waters, making monochloramine the chloramine species expected to be in highest concentration.

The presence of certain organic amino compounds complicates the breakpoint process by providing an additional source of chlorine demand and by reacting with the added chlorine to form relatively stable organochloramines that cannot be completely oxidized. Organochloramine species form simultaneously with the inorganic chloramines. This results in a breakpoint that is not as sharp as that observed when reacting ammonia and free chlorine in isolation. Due to the higher chlorine demand, the breakpoint may be shifted to a chlorine-to-nitrogen molar ratio that is higher than 1.65:1 (see El-Farra et al., 2000).

2.1.3 Analytical methods

Concentrations of inorganic chloramines in aqueous solution may decrease very rapidly upon sampling. As a result, care must be taken to minimize chemical losses due to photolysis, volatilization and contamination. If storage is necessary, samples should be maintained at 4°C for no longer than 1 week (Environment Canada, 1999a). APHA et al. (1995) recommend that analysis for residual chlorine compounds be conducted immediately after sampling.

Figure 2: Principal chemical pathways for the reaction, degradation and fate of free chlorine in aquatic environments (adapted from Jolley and Carpenter, 1983)

Several analytical methods are used to determine the available chlorine in solution (Jolley and Carpenter, 1983). For CRC, the commonly used methods include N,N-diethyl-p-phenylenediamine (DPD) colorimetry, DPD titration and amperometric titration (Fava and Seegert, 1983; APHA et al., 1995; Harp, 1995).

The DPD colorimetric and the DPD titration methods are based on the same chemistry, and, for most samples, there are no clear advantages to the use of either method. The DPD colorimetric method is, however, faster and easier to operate. On the other hand, it is prone to interferences from sample colour and turbidity, while the DPD titration method does not suffer from these interferences.

The DPD methods can be used to estimate the separate monochloramine, dichloramine and combined fractions for natural and treated waters. Both methods are based on the total oxidizing capacity of the sample being analyzed; thus, both are readily subject to interferences from other oxidizing agents, such as chlorine dioxide, hydrogen peroxide, bromine and ozone. They cannot be used with marine water and some groundwaters where significant quantities of bromine will be present (Harp, 1995). Under ideal conditions, the DPD colorimetric method has a minimum detection limit of 0.01 mg/L as Cl₂, while the DPD titration method has a minimum detection limit of 0.018 mg/L as Cl₂(APHA et al., 1995).

Amperometric titration requires a higher degree of skill and care than both the DPD methods. It can be used to determine TRC and can differentiate between FRC and CRC; however, several factors may affect the determination of chlorine species. Metallic ions of silver, copper and iron have been reported as interferences or may diminish the electrode response. Oily or foamy surfactants may adhere to the electrodes, reducing their sensitivity. The violent stirring of some titrators may lower chlorine concentrations by volatilization. Also, oxidized forms of manganese may produce a falsely higher concentration of dichloramine (Harp, 1995). Interference can also occur in highly coloured waters. If present, various organic chloramines can be measured by amperometric titration as free chlorine, monochloramine or dichloramine, depending on the activity of chlorine in the organic compound. As with the DPD methods, amperometric titration cannot distinguish inorganic chloramines from organic chloramines (APHA et al., 1995). The minimum detection limit of amperometric titration is approximately 0.01 mg/L as Cl₂under ideal conditions.

Figure 3: Graphic representation of the ideal breakpoint phenomenon for the chlorination of ammonia at pH 6 and 8 (El-Farra et al., 2000)

High-pressure liquid chromatography (HPLC) with post-column electrochemical detection can be used to quantify inorganic chloramines in potable water, surface (fresh) water, groundwater, and industrial or municipal wastewater. Unlike the previously discussed methods, the HPLC method is not subject to interference from organic chloramines; however, its use requires a skilled analyst, and the instrumentation is substantially more expensive than that used in the traditional methods. HPLC provides an inherently mild condition (neutral pH) and employs direct analysis without sample alteration for detecting inorganic chloramines. This method can be used to calculate total organic chloramines in conjunction with the DPD ferrous ammonium sulfate (FAS) titration method (i.e., analysis of TRC and FRC by DPD FAS titration method, analysis of inorganic chloramines by HPLC: total organic chloramines = TRC - FRC -monochloramine - dichloramine). The HPLC method is currently unable to differentiate combined chlorine species from combined bromine species in marine and estuarine waters. This HPLC method has a detection limit of 0.01 mg/L as Cl₂(Environment Canada, 1999a).

2.2 Entry characterization

Chloramines are released to the Canadian environment mainly by municipal and industrial sources in potable water, effluents and cooling water. To quantify releases of chloramine into the Canadian environment, surveys were sent to municipalities and industries across Canada (Environment Canada, 1997b,c,d). Surveys were sent to all Canadian municipalities with populations exceeding 5000 (based on the 1994 Census) to determine chloramine production and release to the Canadian environment from chlorinating wastewater treatment facilities (which do not dechlorinate before discharge), as well as from drinking water treatment facilities and distribution systems (Environment Canada, 1997b,c). As per its request, Quebec was excluded from the municipal survey. The Municipal Water Use Database (MUD) (Environment Canada, 1996) was consulted to determine which municipalities in Quebec used chloramines as a secondary disinfectant for potable water and to obtain additional data with which to estimate loading from this province.

Most respondents reported concentrations as TRC, since the analytical methods used were not capable of differentiating between inorganic and organic chloramines. For chloraminated drinking water, the TRC concentration will be almost completely due to monochloramine. Effluent TRC will be predominantly composed of inorganic and organic chloramines in proportions dependent on site-specific conditions.

A survey was administered to industries across Canada that may be discharging chlorinated effluents or cooling waters under Section 16 of the Canadian Environmental Protection Act (CEPA). Industries were required to respond if their facility produced or released a trigger quantity of 1000 kg of chloramines per year (Environment Canada, 1997c).

2.2.1 Municipal potable water

The inventory of chloramines produced and released to the environment from potable water included chloramines intentionally formed for disinfection purposes. Potable water treated with free chlorine may also contain inorganic and organic chloramine species, depending on the concentrations of ammonia and organic nitrogen present at the time of chlorination. This latter source has not been considered in this source inventory. However, it is recognized that the assessment of fate and effects as described in this Assessment Report would also relate to releases of inorganic chloramines resulting from potable water treated with free chlorine.

In 1996, 64 Canadian municipalities with populations exceeding 5000 used potable water that was intentionally treated with chloramine. There were 29 facilities in Canada treating a total of approximately 1 220 000 000 m³ of drinking water per day, which serviced approximately 6.9 million Canadians in 1996. In 1995, there were 28 facilities that treated an approximate total of 1 200 000 000 m³/day. The survey revealed that some of Canada's most populated regions produce and/or use chloramine-treated drinking water. These included the Greater Toronto Area, Edmonton, the Regional Municipalities of Ottawa-Carleton, Hamilton-Wentworth and Peel, and the Capital Regional District of Victoria.

The largest producer of chloramines in potable water during 1996 was Ontario (67.7%), followed by Alberta (23.1%), Saskatchewan (4.9%) and British Columbia (2.9%). Newfoundland and Quebec produced substantially less (approximately 1.5%). Nationally, drinking water treatment facilities achieve chloramine concentrations of between 0.01 and 4.80 mg/L at the source and throughout the distribution system. The average chloramine concentration in drinking water (at the source and throughout the distribution system) was approximately 1.0 mg/L in 1995 and 1996. The average minimum concentrations were 0.61 mg/L and 0.60 mg/L in 1995 and 1996, respectively, whereas the average maximum concentrations were 1.66 mg/L and 1.46 mg/L in the same years. The national average, average minimum and average maximum chloramine concentrations in potable water refer to the flow-weighted arithmetic mean of all average, minimum and maximum chloramine or TRC concentrations reported from all surveys (Environment Canada, 1997b).

Of the 213 respondents to the drinking water survey, approximately 24% were able to provide some data regarding environmental releases. Accidental drinking water releases are typically episodic and unpredictable with respect t o their time, duration and location and o ccur fro m main breaks, leaks and overflows from treatment facility reservoirs. Most distribution main leaks and breaks have discharges of less than 0.01 m³/s and durations of less than 8 hours. However, releases of up to approximately 1.0 m³/s and lasting several days or weeks have occurred. An estimated 9% of the total volume of water treated was released accidentally to the environment in 1996.

Outdoor uses (e.g., for lawn/garden watering, driveway washing and car washing) accounted for an estimated 7.5% and 7% of the total volume of chloramine-treated drinking water in 1995 and 1996, respectively. Lesser releases of chloraminated drinking water also occur from fire fighting, street cleaning and main flushing.

Of the 15 facilities that were contacted for data characterizing chloramine-containing waste streams from municipal drinking water treatment plants, 10 facilities dechlorinated their waste streams, diverted them to sanitary sewers or did not discharge any waste streams at all. The remainder discharged directly to a freshwater environment. Based on the available data, WTP backwash wastes accounted for an estimated 22 100 000 m³ and a total TRC loading of approximately 8230 kg in 1997. Unlike typical effluents, these discharges are intermittent, lasting approximately 15-30 minutes per discharge and occurring up to approximately 6 times per day. Reported CRC concentrations in waste ranged from approximately 0.07 to 2.00 mg/L (approximate flow-weighted mean = 0.370 mg/L).

2.2.2 Municipal wastewater

The Environment Canada (1997c) municipal wastewater survey had an overall 49% response rate and a 61% response rate of municipalities that chlorinated their wastewater. To fill the remaining data gaps, various other sources were consulted to determine the loading of chlorinated wastewater to the Canadian environment (e.g., Environment Canada, 1996; OMEE, 1997; Alberta Environmental Protection, 1998). It was determined that 173 municipal wastewater treatment plants (WWTPs) disinfected their wastewater with chlorine and did not dechlorinate prior to discharging to aquatic systems in 1996. Dechloronation involves the removal of residual chlorine usually by physical or chemical processes. Many WWTPs dechlorinate effluents at all times when they chlorinated for disinfection purposes (e.g., all WWTPs in the Greater Vancouver Regional District).

TRC concentrations ranged between the detection limit (usually 0.01 mg/L) and 4.00 mg/L for 1995 and 1996. The national average TRC concentrations in chlorinated municipal effluent were 0.72 mg/L and 0.70 mg/L for 1995 and 1996, respectively. The national average maximum TRC concentrations were 1.45 mg/L and 1.36 mg/L for 1995 and 1996, respectively. Total discharges of chlorinated municipal wastewater effluent were approximately 1 770 000 000 m³ and 1 830 000 000 m³ for 1995 and 1996, respectively. The total average loading of TRC to surface water from municipal sewage treatment effluent was approximately 1 300 000 kg in both 1995 and 1996.

Based on the average TRC loading reported by municipalities in 1996, Ontario produced most of Canada's chlorinated municipal sewage treatment plant effluent (89.9%), followed by Saskatchewan (4.1%), Alberta (2.0%) and Nova Scotia (1.1%). The remaining provinces each produced less than 1% of the national chlorinated wastewater discharged. Production of chlorinated sewage was proportionally similar in 1995. The Yukon and Northwest Territories did not discharge any chlorinated wastewater effluent in 1996.

In 1996, approximately 98% of TRC loading was to a freshwater environment (23% to a river, 48% to a lake, 27% to an unspecified freshwater type) and 2% to a marine/estuarine environment. All marine discharges of chlorinated municipal wastewater occurred in Atlantic Canada.

2.2.3 Industrial effluents and cooling waters

In total, 54 facilities responded to the industrial survey (Environment Canada, 1997d).

2.2.3.1 Cooling waters

According to the industrial survey responses, 21 facilities in Canada used chlorine to treat cooling water and did not dechlorinate prior to discharge in 1995 and 1996. Many of these facilities were petroleum refineries (6 facilities in total), metal fabricators (4), chemical manufacturers (4) and electrical generating stations (2). In addition to the survey, all major electrical utilities, with the exception of Hydro-Québec, were contacted for detailed information regarding chlorination of cooling waters. It was found that EPCOR and Kirkland Lake Power Corporation each operated a generating station that produced or released chlorine or chloramine over the trigger quantity of 1000 kg per year in 1995 and/or 1996. Total discharge of chlorinated cooling water in Canada for 1996 was approximately 132 000 000 m³, and the total TRC loading was approximately 86 000 kg. The national average TRC concentration in cooling water was 0.77 mg/L for 1996 (range = 0.46-1.48 mg/L). In 1996, the largest proportion of discharged chlorinated cooling water was found in Alberta (TRC loading was approximately 33 400 kg), Quebec (26 400 kg) and Ontario (25 000 kg).

2.2.3.2 Zebra mussel control

In Ontario and Quebec, chlorine is used to inhibit the fouling of intake and outfall pipes by zebra mussels. Chlorination to control zebra mussel populations is required only between June and October, and then only when the plants are in operation (Environment Canada, 1993).

Ontario Hydro's Nuclear Division used chlorine to control zebra mussel fouling at their nuclear generation stations on the Great Lakes (Bruce, Darlington and Pickering). One facility in Quebec chlorinated discharge for zebra mussel control in 1995 and 1996. The average combined flow for all stations was approximately 6 350 000 000 m³/d. The average TRC concentration for 1996 was approximately 0.01 mg/L. The total TRC loading from all facilities was approximately 142 000 kg in 1996.

2.2.3.3 Industrial wastewater

The chemical composition of industrial wastewater varies widely depending on the nature of the industry. In contrast to the relatively consistent characteristics of domestic sewage, industrial wastewater often has quite different characteristics, even for similar industries. Industrial wastewater can include employees' sanitary wastes, process wastes from manufacturing, wash waters, and water from heating or cooling operations (Henry and Heinke, 1996). Of the industrial surveys received, 18 facilities were found to discharge a total of 22 800 000 m³ of chlorinated wastewater in 1996.

The total TRC loading from industrial wastewater in 1996 was approximately 4900 kg. The average TRC concentration for these facilities is 2.1 mg/L, with concentrations ranging between undetectable and 3.6 mg/L for 1995 and 1996.

Five facilities in Ontario discharge chlorinated wastewater on a continuous basis. Detailed information regarding these sites was not available. The province of Newfoundland has one industry that discharges chlorinated wastewater at an approximate rate of 655 000 m³/d. This facility's wastewater has an average TRC concentration of 1.0 mg/L, with concentrations ranging between 0.8 and 1.2 mg/L.

2.2.4 Summary

The source inventory found that most of the chloramine/TRC loading from all known sources occurred in Ontario (approximately 80% of total), followed by Quebec (8%) and Alberta (6%) (Table 3). As shown in Figure 4, approximately 99% of all chloramine and TRC discharges are to fresh water, and only approximately 1% are destined for a marine environment. Discharges to land amounted to 0.01% of all emissions. This release was too small to be shown in Figure 4. Since there are no comprehensive data available for the destination of potable water flows from distribution systems, these data have not been included in Figure 4.

2.3 Exposure characterization

2.3.1 Environmental fate

All releases of chloramine compounds into the Canadian environment reported by the municipal and industrial surveys (Environment Canada, 1997b,c,d) were in aqueous solution. Hence, chloramine fate is governed largely by water-phase processes. However, other phases, such as air and soils, are also involved.

2.3.1.1 Air

Studies describing the fate of chloramines in ambient air do not exist. In the air phase, it would be expected that chloramines would dissipate due to advection and dilution and would be subject to reaction, although no informat ion has been located characterizing reacti ons fo r chlo ramines in a gaseous state. Various studies indicate that chloramines are thermodynamically unstable and susceptible to photolysis (Gilbert et al., 1987; Gilbert and Smith, 1991; Lorberau, 1993). Monochloramine and dichloramine are very water soluble and are thus susceptible to removal from the atmosphere by rain. Gas-phase trichloramine is explosive in nature, particularly in the presence of monochloramine and dichloramine or when in vacuo. This has had an inhibiting effect on relevant scientific research (Gilbert et al., 1987).

2.3.1.2 Water

Chloramine and FRC species are easily transformed to one another in water, and various CRC and FRC species are usually present simultaneously. If FRC is released to a fresh surface water, inorganic or organic chloramines may be formed immediately, and the dominant inorganic chloramine species will be based on relevant site-specific conditions, particularly the pH and molar ratio of hypochlorous acid to ammonia. Conditions prevalent in natural fresh surface waters are conducive to the formation and presence of monochloramine and dichloramine (see Section 2.1.1). Monochloramine is, however, the principal inorganic chloramine in fresh waters due to its rapid formation and relative stability in comparison with other CRC and FRC species (Johnson, 1978; Margerum et al., 1978).

Trichloramine is rarely found in the environment, since its formation is dependent on uncommon natural conditions (i.e., pH <4.4; hypochlorous acid-to-ammonia ratio >7.6:1). Once formed, trichloramine is extremely volatile (Table 2) and will move quickly to the air phase.

The addition of FRC or CRC to water containing bromine will lead to the formation of bromamines. This is particularly an issue in seawater; however, some groundwater and fresh surface water also have sufficient amounts of bromine to produce bromamines (see El-Farra et al., 2000). Inorganic chloramines are thus viewed as being in dynamic equilibrium with several forms of residual oxidants. Since studies describing the fate of inorganic chloramines are few, and because most researchers describe the fate of TRC with little or no speciation, the analysis of chloramine fate has involved an examination of the behaviour of TRC to infer the behaviour of chloramines.

Table 3 Estimated chloramine and TRC loading to the Canadian environment from all sources in 1996

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Figure 4 Chloramine and TRC loading from chlorinated cooling water and municipal and industrial effluent (Environment Canada, 1997a,b,c)

Inorganic chloramine fate is governed largely by water-phase processes, including dilution, mixing, advection, chemical demand (reactions with organic and inorganic compounds), benthic demand, photodegradation, volatilization, sediment adsorption and reaction, and sediment-associated transport, deposition, burial and resuspension.

Mixing and dispersion of discharges containing chloramines depend on water body morphometry, as well as the magnitude and direction of water flows and currents. If the dilution of the effluent is small and/or if the current velocity is fast, complete mixing may not occur for several kilometres downstream from the source (e.g., Milne, 1991). The effluent may then be contained in a long narrow plume. For instance, discharges from the E.L. Smith and Rossdale WTPs to the North Saskatchewan River in Edmonton, Alberta, did not completely mix with the river, and significant transverse concentration gradients existed in a plume for considerable distances downstream (Milne, 1991). TRC concentrations measured in the Rossdale WTP waste discharge ranged from 2.05 to 2.16 mg/L; at 1402 m from the source, TRC concentrations ranged from <0.01 (detection limit) to 0.065 mg/L (Milne, 1991). A discussion of mixing, advection and dispersion in rivers, lakes, and estuarine and marine environments has been provided with reference to dispersion modelling in a supporting document entitled "Tier 2 Exposure Assessment of Effluents for Inorganic Chloramines" (McCullum et al., 2000).

Decay rate constants (k) of inorganic chloramines are highly variable, varying by 4 orders of magnitude depending on the type of water used (e.g., fresh or salt water, pH, surface or deionized water, etc.), chlorine/chloramine dose, study design (e.g., in situ versus laboratory) and experimental conditions. Generally, studies conducted under controlled laboratory conditions produce decay rates that are at least 1 order of magnitude lower than those produced in in situ studies (Milne, 1991). This is likely due to the use of controls that limit many environmental decay processes, such as volatilization, photodegradation and benthic demand.

Water column chloramine demand is produced by chemical reactions with inorganic (e.g., I^-, S^2-, Fe²⁺, Mn²⁺, HSO₃^- and NO₂^- ions) and organic (e.g., alkyl sulfides, amines and some nitrogen heterocyclic aromatics) substances, as well as adsorption to solids and colloidal matter (Morris and Isaac, 1980; Christman et al., 1983; Scully and White, 1992). Reactions in aquatic environments are affected by temperature, pH and turbulence (Heinemann et al., 1983; Abdel-Gawad and Bewtra, 1988; Milne, 1991).

Under controlled conditions (no sunlight or volatilization, 15°C) using deionized water mixed with various fresh surface waters (Norrish Creek, British Columbia; North Saskatchewan River, Alberta; and Grand River, Ontario), Environment Canada (1998a) determined decay rate constants ranging from approximately 0.017 to 0.413 per day (half-life of 1.67-40.8 days) for monochloramine. Using deionized water and seawater mixtures from Burrard Inlet, British Columbia, Environment Canada (1998b) determined monochloramine decay rate constants of 0.74-1.01 per day (half-life of 0.68-0.94 days). Abdel-Gawad and Bewtra (1988) determined a TRC decay rate constant of 0.48 per day (half-life of 1.44 days) using municipal wastewater-St. Claire River water mixtures at 20°C. Heinemann et al. (1983) determined chemical demand rate constants of 1.73-23.76 per day (half-life of 0.03-0.40 days) at 20°C also using effluent-surface water mixtures.

Lee et al. (1982) and Heinemann et al. (1983) indicated that TRC is very volatile and accounted for 20-80% of the lost chlorine from various Colorado rivers. There are no volatilization rate constants derived specifically for chloramines. TRC may be influenced by photodegradation; however, when all environmental decay processes are combined, its effect may be nullified (Milne, 1991).

The overall or gross decay rate incorporates different environmental factors, which, when taken together, may represent natural environmental conditions. A focussed review of the literature revealed that the first and third quartiles of reported overall decay rate constants for CRC, TRC and TRO were approximately 0.70 and 20.0 per day (half-life of 0.03-1.0 days), respectively (see Pasternak, 2000). Reckhow et al. (1990) derived higher rates of in situ monochloramine decay (i.e., 144 per day; half-life of 0.005 days). Wisz et al. (1978) undertook an evaluation of CRC decay from chlorinated municipal wastewater discharged to Aurora Creek (now called Tannery Creek), Ontario, and reported that monochloramine decay rate constants ranged from 4.97 per day in the winter to 19.54 per day in the summer (half-life of 0.04-0.14 days).

2.3.1.3 Sediment

Inorganic chloramine loss from the water column may occur via adsorption and reaction with suspended solids and bottom sediments (Milne, 1991). Environment Canada (1998b) found that sediments at a concentration of 5000 mg dry weight/L from the Grand River, North Saskatchewan River and Downes Creek produced monochloramine decay rate constants of 0.50, 0.28 and 14.83 per day (half-lives of 1.4, 2.5 and 0.05 days), respectively. The highest decay rates were associated with sediments that had higher organic nitrogen and carbon and that were suspected to contain biologically active materials (i.e., stream scum).

Stream beds may be c overed w ith acti ve biological materials in the form of slimes, sludges and algae, particularly at wastewater outfalls. This biological material has a capacity for uptake of residual chlorine (Krenkel and Novotny, 1980). The rate of pollutant uptake in this layer (or the rate produced by benthic demand) will be influenced by the type of biological material, temperature, flow and sediment characteristics and depth (Milne, 1991). Due to its dependence on site-specific conditions, it is very difficult to make generalizations regarding chloramine loss rate due to benthic demand, except that it may be extremely rapid.

One study provided a preliminary estimate of benthic demand on TRC. Milne (1991) undertook in situ benthic demand tests in the North Saskatchewan River, just upstream of the E.L. Smith Water Treatment Plant, Edmonton, Alberta, in September and October of 1990 and September of 1991. He observed a TRC loss rate constant (geometric mean) of 448-591 per day (half-life of 0.001-0.002 days). Benthic demand was in fact higher than the overall measured TRC decay rate constants for the North Saskatchewan River (i.e., 20.0 and 28.0 per day, or half-life of 0.03 and 0.04 days). It is unknown whether prior exposure to residual chlorine would affect benthic uptake of chloramine or whether adsorption or reaction was occurring, although the latter seems apparent given the organic nature of the benthic material.

2.3.1.4 Soils

There are no studies evaluating the fate of inorganic chloramines on/in soils. Based on related information on fate associated with sediments and surface waters, inorganic chloramines would experience chemical reaction with particulates, volatilization and photolysis at the soil surface and chemical reaction and adsorption within the soil matrix. Inorganic chloramine may oxidize surface layer soil organic matter (Bodek et al., 1988), particularly materials composed of organic nitrogen compounds, such as alkyl sulfides, amines and some nitrogen heterocyclic aromatics (e.g., see Christman et al., 1983; Scully and White, 1992). Zellmer et al. (1987) speculated that the components of a fine-silty clay loam soil had completely bound or inactivated sodium hypochlorite in the C-8 mixture (15% perchloroethylene, 8% calcium hypochlorite, 1% emulsifier and 76% water). They simulated spill conditions and found no residual chlorine concentrations at any depth of a 68-cm soil core after 144 hours using the starch-iodine method of analysis.

2.3.1.5 Biota

Accumulation of inorganic chloramine in biota is not likely, since inorganic chloramines are known to be transient and highly reactive with organic substances.

2.3.2 Environmental concentrations

There are no data regarding inorganic chloramine concentrations in ambient air, groundwater, sediments, soils or biota. TRC, TRO, monochloramine and dichloramine have been measured in effluents containing municipal wastewater and drinking water, in cooling water, and in potable water, as well as in surface water near WWTP outfalls (see Pasternak and Powell, 2000). The following summarizes TRC and CRC concentrations in municipal wastewater and drinking water and in surface waters near WWTP outfalls.

2.3.2.1 Effluent releases

Environment Canada (1997e) measured TRC, FRC, monochloramine, dichloramine and total organic chloramine concentrations at three sewage treatment plants in British Columbia in 1997. These three plants perform dechlorination prior to d ischarging treated wastewater. During this study, wastewater samples collected after chlorination and prior to dechlorination were analyzed for CRC species. Table 4 indicates that TRC was composed predominantly of inorganic and/or organic chloramines. This is substantiated by other studies presented in this section.

Based on a screening of approximately 50 riverine sites across Canada receiving chlorinated municipal WWTP effluent using average 1996 effluent data and mean annual river flow rates, it was found that effluents were diluted by a median value of approximately 940 times. During low flow, the median dilution was approximately 310 (range 2-50 000) (Pasternak and Powell, 2000).

As a result, chloramine concentrations in effluents may be quickly diluted to non-detectable levels (approximately <0.01 mg/L) if rapid mixing occurs in a sufficient volume of surface water. However, at low dilutions and under conditions of slow chloramine decay (e.g., water with low organic material, low temperature and low total suspended solids), elevated chloramine concentrations may persist for great distances downstream from their source.

During the summer of 1998, Environment Canada contracted sampling studies at three riverine sites with low dilution (Sheep River, Don River and Lynne River) and at one lake site (Lake Ontario) with high loading and high dilution. In the Sheep River, at the Okotoks WWTP, Okotoks, Alberta, the CRC concentration at the sewage treatment plant's outfall was measured at 0.330 mg/L. At 5 m from the outfall, the CRC concentration ranged between <0.005 mg/L (the detection limit) and 1.350 mg/L. At 150 m downstream, CRC concentrations ranged from below the detection limit to 0.150 mg/L (Golder Associates Ltd., 1998).

At the Ashbridges Bay WWTP, the CRC concentration in the effluent was 1.250 mg/L, which dropped to 0.090-0.180 mg/L at 100 m from the outfall in Ashbridges Bay, Lake Ontario. The CRC concentration 1300 m away from the outfall ranged from 0.020 to 0.030 mg/L. At the North Toronto WWTP, the CRC concentration in the effluent was measured at 1.040 mg/L. The CRC concentration in the Don River 30 m downstream from the outfall was found to be <0.005-0.053 mg/L. At 200 m downstream, the CRC concentration was measured at <0.005-0.013 mg/L. The Simcoe WWTP effluent's CRC concentration ranged from 0.170 to 0.343 mg/L. In the Lynn River 3 m downstream from the plant's outfall, the CRC concentration ranged from <0.005 to 0.260 mg/L. At approximately 500 m downstream from the outfall, the CRC concentration in the Lynn River was <0.005 mg/L. Detectable chloramine concentrations were then measured at approximately 300 m downstream from the outfall. A subsequent, confirmatory sample at 600 m downstream, however, contained CRC concentrations of 0.020-0.023 mg/L. Sampling at the 300-m site disturbed the bottom sediments and may have caused the resuspension of chloramines into the water column. This resuspended chloramine may have been measured at the 600-m site. These observations suggest that chloramines may be stored in the sediments of a river and could be released if the sediments are disturbed (Gartner Lee Ltd., 1998).

Table 4 Chloramine speciation at three municipal WWTPs in British Columbia ¹

Item	Concentration (mg/L)
	Ladysmith Sewage Treatment Plant	Joint Abbotsford-Matsqui Environmental System	Kamloops Sewage Treatment Plant
	Primary treatment	Secondary treatment	Tertiary treatment (P removal)
Free residual chlorine	<0.02	<0.02	<0.02
Monochloramine 2	0.407	0.026	0.414
Dichloramine 2	0.02	<0.02	0.374
Total organochloramine 3	0.783	0.058	0.017
Total res idual chlo rine	1.21	0.084	0.805

Wisz et al. (1978) studied the decay of residual chlorine in the receiving waters at four Ontario municipal WWTPs (at Aurora, Bolton, Brantford and Alliston). The downstream persistence of residual chlorine was noted to depend largely on dilution. Relatively fast degradation of residual chlorine occurred in receiving waters where the dilution ratio was greater than 20:1. During the summer study period, under low-flow stream conditions, the dilution ratios at the selected facilities were as follows: Brantford 43:1, Alliston 23:1, Aurora 1:1.3 and Bolton 24.6:1. Two receiving streams showing rapid degradation were the Grand River (Brantford WWTP) and the Boyne River (Alliston WWTP). TRC concentrations at the two outfalls of the Brantford WWTP were 0.880-2.288 mg/L and 0.920-2.272 mg/L, respectively. At 107 m downstream, the TRC concentration in the Grand River was already approaching the detection limit of 0.002 mg/L. At 942 m, there were no measurable concentrations of TRC. At the Alliston WWTP outfall, the TRC concentrations ranged from 0.768 to 1.408 mg/L. Downstream at 61 m, the TRC concentration had declined to <0.002-0.045 mg/L; at approximately 1.5 km, the TRC concentration was below the detection limit.

Wisz et al. (1978) showed that chloramine decay is substantially slower in the winter than during the summer. In the summer of 1976, the monochloramine and dichloramine concentrations at the Aurora WWTP outfall ranged from 1.120 to 1.440 mg/L and from below the detection limit to 0.144 mg/L, respectively. At 91 m downstream in Aurora Creek, monochloramine concentrations ranged from 0.364 to 0.632 mg/L. At 2900 m downstream, monochloramine concentrations fell below detection limits (<0.002 mg/L). Dichloramine was measured at approximately 2900 m downstream from the outfall (0.013-0.018 mg/L). In the winter of 1977, monochloramine and dichloramine concentrations measured at the Aurora Creek outfall were 1.200 and 0.320 mg/L, respectively. At 4800 m downstream from the outfall, the concentrations of monochloramine and dichloramine were 0.216 and 0.080 mg/L, respectively.

2.3.2.2 Drinking water releases

Sampling studies were conducted by Environment Canada and the Fraser Valley Regional District (FVRD) in Mission and Abbotsford, British Columbia, to determine whether chloramines from outdoor residential water use and from industrial washdown activities could be measured in local waterways (Vanden Berg and Wade, 1997; FVRD, 1998, 1999; Pasternak et al., 1998, 1999). Measurable concentrations of CRC and FRC were not found at outlets of storm drains and in local streams using HPLC, DPD FAS and amperometric titration methods. These studies did not substantiate the findings of Triton Environmental Consultants Ltd. (1995), who found measurable TR C and FRC concentrations at several of the same streams using the Hach Kit DPD colorimetric method. The reports, with the exception of Triton Environmental Consultants Ltd. (1995), concluded that low-discharge uses of chloramine-treated water (e.g., lawn and garden watering, car washing, driveway washing and equipment washing) do not result in measurable levels of chloramines in surface waters of the FVRD.

Two major salmonid and invertebrate kill events resulting from the release of chloramine-treated drinking water via main breaks to Fergus Creek (Surrey, British Columbia) indicate that levels of inorganic chloramines that are acutely lethal to salmonids and invertebrates have occurred in surface waters (see Section 2.4.1.3). TRC concentrations in the chloraminated drinking water from the area of each main break were 2.53 and 2.75 mg/L (Nikl and Nikl, 1992). These events resulted in Fisheries Act convictions.

2.4 Effects characterization

2.4.1 Ecotoxicology

A thorough review of the international literature regarding the environmental toxicology of inorganic chloramines has been completed in which relevant toxicological information was summarized, important data gaps were identified and numerical data quality rankings were assigned to relevant articles (see Farrell and Wan, 2000: Appendix B).

2.4.1.1 Mode of action

Monochloramine affects protein-associated processes in bacteria, and the mode of action appears to be a "multiple-hit mechanism" involving the inactivation of several sites before cell death or irreversible injury occurs (Jacangelo et al., 1991). Amino acids, especially those with sulfur groups and tryptophan, were found to be more reactive than the nucleic acids in disinfection studies (Ingols et al., 1953; Boyce, 1963). Monochloramine inhibited bacterial growth as well as DNA, RNA and protein synthesis at levels of 0.515 mg/L. At a monochloramine concentration of 5.15 mg/L, enzymes that contained sulfhydryl groups were significantly inhibited (Kohl et al., 1980). Monochloramine inhibition was found to be irreversible with the addition of sulfhydryl-containing chemicals after the addition of monochloramine. It was speculated that monochloramine probably acts on other sites in addition to the sulfhydryl groups (Boyce, 1963; Kohl et al., 1980).

Studies on the mode of action of monochloramine in viruses are limited. Two mechanisms have been shown. One involves the interaction of monochloramine with the RNA in a bacteriophage, and the other involves the interaction of a mixture of inorganic chloramines with the protein coat of the virus. The mode of action in viruses may vary with virus type and chloramine concentration (Fujioka et al., 1980; Olivieri et al., 1980).

Chloramines appear to cross the fish gill epithelium quite readily and do not cause significant cellular damage in comparison with free chlorine. In fish, inorganic chloramines affect transport of oxygen in blood by reacting with the hemoglobin of the red blood cells to form methemoglobin, inhibiting the cell's ability to bind oxygen (Buckley, 1976). The percentage of methemoglobin within the hemoglobin increases significantly with exposure to chloramines, and the fish begins to exhibit signs of hemolytic anemia. Severe hemorrhaging occurs throughout the body and from the fins. In addition, the body of the fish becomes covered with a mucous coating, and the fish shows increased "coughing" and erratic swimming (Grothe and Eaton, 1975; Buckley, 1977; Travis and Heath, 1981).

Inorganic chloramines have also been found to reduce filtration and reproduction in rotifers, lobsters and fish, but the underlying mechanisms for the responses are not clear (Capuzzo et al., 1976, 1977; Capuzzo, 1977, 1979a).

Certain fish are able to avoid chloramines at levels ranging from 0.0057 to 0.44 mg/L (Fava and Tsai, 1978; Cherry et al., 1979; Hidaka and Tatsukawa, 1985). However, some fish species have demonstrated a reduced avoidance response at a preferred temperature. Warmer temperatures may counteract the avoidance response and result in fish attraction (Hall et al., 1982, 1983). The excess ammonia that is used to create inorganic chloramine solutions may also attract some fish species, thereby mitigating the avoidance response to inorganic chloramine (Cherry et al., 1982). Finally, some fish are capable of detecting lower inorganic chloramine levels if allowed longer exposure periods, during which time they acquire avoidance response skills (Fava and Tsai, 1978).

2.4.1.2 Terrestrial organisms

Studies of inorganic chloramine toxicity (acute or chronic) to plants are limited. A study on horse bean (Vicia faba) seeds showed that a 60-minute exposure to monochloramine at a concentration of 5.15 mg/L resulted in 24% abnormal anaphases. It appears that monochloramine can induce chromosome damage at concentrations that do not cause visible damage to the plant (Sekerka, 1981).

There are very few inorganic chloramine toxicity data (acute or chronic) available for terrestrial animals, with the exception of rats (Farrell and Wan, 2000: Appendix A). Only one study was found examining the effects of inorganic chloramines on amphibians. This study, which utilized eggs from the urodele amphibian, Pleurodeles waltl, found that monochloramine caused mutations at a concentration of 0.15 mg/L but not at 0.05 or 0.01 mg/L (Fernandez et al., 1993).

2.4.1.3 Aquatic organisms

The acute toxicity of inorganic chloramines to aquatic organisms is species-specific and is a function of life stage, chemical species, exposure duration, pH and temperature. Variability in test conditions (e.g., differences in pH, temperature, exposure duration and inorganic species composition) and data quality makes comparisons between historical inorganic chloramine toxicity values quite difficult. Particular concern exists over the variable and imprecise ways in which inorganic chloramine concentrations were reported. Also, synergism between free chlorine toxicity and inorganic chloramine toxicity may occur (Farrell and Wan, 2000: Appendix A).

To summarize the available data, determine sensitive aquatic species and identify data gaps, a meta-analysis approach (Mattice and Zittel, 1976; Mattice, 1977) was used in which all aquatic toxicity data were graphically represented by plotting LC₅₀ values as a function of exposure time according to biological and environmental categories (see Farrell and Wan, 2000: Appendix C). These graphs were intended to establish a lower-boundary concentration line above which lay all acute toxicity data, including those for sensitive species. However, this objective was not realized, because no single species had a sufficiently comprehensive data set to allow a lower boundary line for inorganic chloramine concentration to be set with confidence.

Consequently, supplementary acute toxicity testing, supported by the best available analytical chemistry, was performed with representative freshwater fish (juvenile chinook salmon, Oncorhynchus tshawytscha) and invertebrates (Ceriodaphnia dubia and Daphnia magna) and marine invertebrates (Amphiporeia virginiana and Eohaustorius washingtonianus). Time-to-lethality (e.g., LT₁₀₀, LT₅₀, LT₂₀, LT₀) reference lines were determined for these species.

2.4.1.3.1 Freshwater and marine algae and plankton

Inorganic chloramines are used for the control of freshwater and marine fouling, bacterial growth and planktonic growth. According to Farrell and Wan (2000: Appendix A), the lowest reported observed effect concentration produced by residual chlorine on a species of algae is 0.01 mg/L. This corresponded to a 15-minute EC₅₀ (carbon uptake) for the unicellular alga Pyramimonas virginica (Bender et al., 1977). Maruyama et al. (1988) found 10-day EC₅₀s (growth) of 0.014 and 0.02 mg/L for the multicellular red alga Porphyra yezoensis.

2.4.1.3.2 Freshwater invertebrates

The LC₅₀ values for invertebrates ranged from 0.011 mg/L for the freshwater water flea, D. magna (24-hour LC₅₀), to 0.96 mg/L for the freshwater crayfish, Oronectes nais (96-hour C₅₀) (Ludwig, 1979; Kaniewska-Prus, 1982). The reported LC₅₀ values for D. magna varied considerably (e.g., 24-hour LC₅₀ values ranged from 0.011 to 0.110 mg/L). Therefore, supplementary acute toxicity tests were performed with D. magna in conjunction with the best available analytical chemistry to ascertain the reliability of literature values; the results of these tests are summarized i n Farrell and Wan (2000). The estimated 24-hour and 48-hour LC₅₀ values for D. magna were 0.019 mg/L and 0.017 mg/L for inorganic chloramines, respectively, in 20°C water at pH 8. These LC₅₀ values were comparable to the lowest of the existing LC₅₀ values for D. magna published in the literature.

However, the rotifer, Keratella cochlearis (24-hour LC₅₀ 0.0135 mg/L), and the Australian water flea, C. dubia (24-hour LC₅₀ 0.012 mg/L), were more sensitive than D. magna to continuous chloramine exposures (Beeton et al., 1976; Taylor, 1993). In contrast, the Asiatic clam, Corbicula fluminea, was very resistant, with an LC₅₀value greater than 2 mg/L (Belanger et al., 1991).

Using data from the open literature, too few data points existed for a single sensitive species to permit a lower-boundary concentration line for continuous chloramine toxicity to freshwater invertebrates to be defined with confidence. Therefore, comprehensive time-to-50%-lethality (LT₅₀) tests with C. dubia, the second most sensitive freshwater invertebrate, were conducted. These supplementary tests were supported by the best available analytical chemistry; although static exposures were used, water replacement every hour limited chloramine degradation over time in the test chambers (Farrell and Wan, 2000: Appendix E). For exposures up to 3200 minutes for 26 monochloramine concentrations, the LC₅₀ for third-generation neonate (12-24 hours old) C. dubia was predicted by a simple exponential equation:

LC50(mg/L) = 61.6t -1.08

(1)

where t = exposure time in minutes (R² = 0.95).

For times to lethality for 20% of C. dubia, the following exponential equation was derived by Farrell and Wan (2000: Appendix C):

LC20 (mg/L) = 53.9t -1.10

(2)

where t = exposure time in minutes (R² = 0.92).

The LT₅₀and the LT₂₀curves for C. dubia are depicted in Figures 5 and 6.

2.4.1.3.3 Marine invertebrates

The available acute toxicity data for marine/ estuarine invertebrates are highly variable.

Figure 5 LT₅₀ curve fitted to data for the effects of inorganic chloramines on Ceriodaphnia dubia (Farrell and Wan, 2000: Appendix C)

Figure 6 LT₂₀ curve fitted to data for the effects of inorganic chloramines on Ceriodaphnia dubia (Farrell and Wan, 2000: Appendix C)

Nevertheless, this grouping contained data for species that seemed to be extremely sensitive to inorganic chloramines (i.e., 48-hour LC₅₀ of <0.01 mg/L for juveniles and larvae of an oyster, Crassostrea virginica, and 48-hour LC₅₀ of 0.001 mg/L for larvae of a clam, Mercenaria mercenaria) (Bender et al., 1977; Capuzzo, 1979b). The acute toxicity of CPO to two marine invertebrates, A. virginiana and E. washingtonianus, was studied (Farrell and Wan, 2000: Appendix E) using the best available resources (test species and analytical chemistry).

The estimated 48-hour LC₅₀ values for A. virginiana and E. washingtonianus were 0.567 mg/L and 0.626 mg/L, respectively, while the 168-hour LC₅₀ values were 0.043 mg/L and 0.134 mg/L, respectively, in 10°C and 15°C seawater at pH 7.5.

2.4.1.3.4 Freshwater fishes

In terms of acute toxicity at comparable exposure durations, the most sensitive freshwater invertebrate species was almost 10 times more sensitive to inorganic chloramines than the most sensitive fish species. Ninety-six-hour LC₅₀ values for fish ranged from 0.07 mg/L for coho salmon (Oncorhynchus kisutch) to 1.72 mg/L for carp (Cyprinus carpio) (Buckley, 1976; Heath, 1977).

Fish species have shown an inverse relationship between temperature and monochloramine resistance (Roseboom and Rishey, 1977; Seegert et al., 1979; Elmore et al., 1980). The temperature at which this relationship starts to take effect is very dependent on species and the temperature range within which the species functions most effectively. Some organisms may show eurythermal adaptation, which is the ability to shift lethal limits, reproduction and metabolic activities to allow tolerance of a wi de range of thermal stresses. Several instan ces of greater chloramine to lerance by cold-water fish species, such as brook trout (Salvelinus fontinalis), in comparison with warm-water fish species, such as channel catfish (Ictalurus punctatus), have been observed (Heath, 1977). This is in direct conflict with the general supposition that cold-water salmonids are the most sensitive to chlorine compounds (Farrell and Wan, 2000: Appendix A).

The evaluation of the effects of pH, by regression analyses of chlorine bioassay data, on CRC toxicity in fresh water (pH 5-8) indicates that changes in pH modify the extent of effluent toxicity only slightly (Farrell and Wan, 2000: Appendix A).

Dichloramine solutions appear to be more toxic to fish than monochloramine solutions, although a few exceptions have been found. Trichloramine species are very rarely found in the environment (Thomas et al., 1980; Brooks and Bartos, 1984).

The effects of inorganic chloramines on various life stages (alevin, fry, juvenile) of brook trout and coho salmon have been examined. The alevin life stage was more tolerant of inorganic chloramines than the fry, with larger, older alevins being less tolerant than the newly hatched alevins. The greatest sensitivity to inorganic chloramines was observed just after the fry stage. This may be due in part to the physiological and behavioural stresses related to the change in feeding patterns (i.e., from using a yolk sac to foraging for food) that occurs at that point in the fish life cycle.

Post-exposure mortality was limited to toxicity tests of less than 200 minutes in duration. Intermittent chloramine exposure was either less toxic than or as toxic as continuous exposure.

The lowest No-Observed-Adverse-Effect Concentration (NOAEC) for inorganic chloramines for fish was 0.0165 mg/L for the fathead minnow (Pimephales promelas). The threshold for growth reduction in coho salmon was 0.011-0.023 mg inorganic chloramines/L. By comparison, avoidance behaviour data existed for 12 species. Thresholds for adult coho salmon and rainbow trout (Oncorhynchus mykiss) appeared to be at 0.090-0.110 mg/L for a 10-minute exposure, but alewife (Alosa pseudoharengus) detected chloramines at 0.002 mg/L in a 45-minute exposure.

To define a lower-boundary concentration line for continuous chloramine toxicity, LT₅₀ tests with juvenile chinook salmon were conducted (Farrell and Wan, 2000). There were insufficient data from the literature to allow a lower-boundary concentration line for continuous chloramine toxicity to be defined with confidence without these further tests. For exposures up to 10 days, the LC₅₀ was predicted by a simple exponential equation:

LC50(mg/L) = 7.24t -0.452

(3)

where t = exposure time in minutes (R² = 0.94).

Figure 7 LT₂₀ and LT₅₀ curves fitted to data for the effects of inorganic chloramines on juvenile chinook salmon (Oncorhynchus tshawytscha) (Farrell and Wan, 2000: Appendix B)

The projected incipient lethal concentration for 50% mortality of juvenile chinook salmon was 0.09 mg/L for exposure durations of no longer than 10 days. Chloramine exposures of <0.67 mg/L for up to 3 hours produced no post-exposure mortality.

For chinook salmon exposed up to 10 days, the LC₂₀ was fitted to the observed toxicity data by the exponential equation:

LC20 (mg/L) = 6.97t -0.488

(4)

where t = exposure time in minutes (R ² = 0.92).

The LT₅₀and LT₂₀curves for chinook salmon are presented in Figure 7. When equations 1 and 2 for C. dubia are compared with equations 3 and 4 for chinook salmon, it is clear that juvenile chinook salmon are less sensitive to inorganic chloramines than C. dubia.

2.4.1.3.5 Marine/estuarine fishes

Toxicity data regarding marine and estuarine fishes are few. The silverside (Menidia menidia) was the most sensitive marine/estuarine species (96-hour LC₅₀ 0.040 mg/L) (Bender et al., 1977).

The effects of inorganic chloramines (measured as total chlorine) on three species of juvenile marine fish - winter flounder (Pseudopleuronectes americanus), scup (Stenotomus versicolor) and killifish (Fundulus heteroclitus) - were investigated by Capuzzo et al. (1977) at 25°C and 30°C. Inorganic chloramines were found to be the most toxic to killifish (100% mortality at 1.20 mg residual chloramine/L), followed by winter flounder (100% mortality at 2.55 mg/L) and scup (100% mortality at 3.10 mg/L). Stress was observed at 0.65 mg residual chloramine/L for killifish, 1.50 mg/L for winter flounder and 2.20 mg/L for scup. A synergistic effect between inorganic chloramine toxicity and temperature was observed, as increasing the temperature to 30°C caused 100% mortality to be observed at lower inorganic chloramine concentrations.

2.4.2 Abiotic atmospheric effects

Inorganic chloramines are formed in and released to aquatic environments. Although they are known to volatilize, there are no literature-reported chloramine concentrations in ambient air. Inorganic chloramines absorb radiation in the 200-300 nm wavelength region of the spectrum, and they are susceptible to photolysis in water (Hand and Margerum, 1983; Lin et al., 1983; Reckhow et al., 1990). There are no data pertaining to inorganic chloramine fate in the atmosphere, although there are reports that they are very unstable and not persistent in the atmosphere (Kirk-Othmer, 1979; Lide, 1998). Monochloramine and dichloramine are very water soluble and hence susceptible to removal from the atmosphere by scavenging rain and subsequent deposition to soil and water. The combined effect of atmospheric reactivity and rain scavenging would inhibit the involvement of monochloramine and dichloramine in stratospheric ozone layer depletion. Also, the available information indicates that chloramines would make a negligible contribution to tropospheric ozone formation.