Priority Substances List Assessment Report for Ammonia in the Aquatic Environment

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

2.1 Identity and physical/chemical properties

Ammonia terminology has been a source of confusion within the technical literature for years. Terms such as free ammonia, total ammonia, non-dissociated ammonia and un-ionized ammonia nitrogen are commonly encountered and tend not only to confuse the reader, but also to make comparison of the data difficult. To solve this problem, the definitions given below will be adhered to throughout this document.

In aqueous solutions, a chemical equilibrium is established between un-ionized ammonia (NH₃), ionized ammonia (NH₄⁺) and hydroxide ions (OH^-). The equilibrium for these chemical species can be expressed in simplified form by the equation:

NH₃ + H₂O ⇔ NH₃·H₂O ⇔ NH₄⁺ + OH^- or NH₄ OH

The term "un-ionized ammonia" refers to all forms of ammonia in water other than the ammonium ion (NH₄⁺). In the literature, this has been designated as NH₃ , NH₄ OH and NH₃·H₂O. The expression NH₃ is used throughout this report to represent un-ionized ammonia, except in instances where one or the other expression is more appropriate to the context (e.g., see the chemical equilibrium discussion later in this section).

The term "ionized ammonia" refers to the ammonium ion, NH₄⁺.

The terms "total ammonia" and "ammonia" refer to the sum of un-ionized ammonia and ionized ammonia (NH₃ + NH₄⁺).

Ammonia is a colourless alkaline gas, lighter than air and possessing a unique penetrating odour. The name "ammonia" is a general term and refers to anhydrous ammonia, ammonia gas and ammoniac anhydre. Ammonia has the Chemical Abstracts Service (CAS) registry number 7664-41-7. The substance has the molecular formula NH₃ and a molecular weight of 17.03 (Grayson and Eckroth, 1978).

Considerable information about the properties of ammonia is given in WHO (1986) and Environment Canada (1984). The physical and chemical properties relevant to the environmental fate of ammonia are summarized in Table 1.

Using their pK_a calculations, Emerson et al. (1975) developed a table (Table 2) describing the percentage of NH₃ in fresh or soft water for temperatures between 0 and 30°C and for pHs in the range 6.0-10.0.

The relationship illustrated in Table 2 holds in most fresh waters. However, the concentration of un-ionized ammonia will be lower at the higher ionic strengths (total dissolved solids) of very hard fresh waters or saline waters. For a given total ammonia concentration, the concentration of un-ionized ammonia decreases slightly with increasing salt content, and this effect can be significant in estuarine and marine waters. Using the appropriate activity coefficients, this relationship can be restated for seawater with an ionic strength of 0.7 as follows (API, 1981):

f = 1/[10 ^{(pKa - pH) + 0.221)}+ 1]

At 25°C, the pKa is 9.24 (Emerson et al., 1975), so at pH 8, the above equation shows that 3.3% of the total ammonia in seawater would exist in the un-ionized form. The corresponding value in fresh water is 5.4%. At this pH and temperature, seawater with an ionic strength of 0.7 contains 38% less un-ionized ammonia than does fresh water.

Table 1 Physical and chemical properties of ammonia

Property	Value
Boiling point at 100 kPa	−33.42°C
Melting point at 100 kPa	−77.74°C
Density (liquid) at −33.7°C and 100 kPa	682.8 kg/m3
Density (gas) at 25°C	0.7067 kg/m3
Vapour pressure
at 15.5°C	640 kPa1
at 21°C	880 kPa1
at 25°C	1000 kPa
pKa (25°C, pH 8)	9.24
Solubility in water, 101 kPa
at 0°C	895 g/L
at 20°C	529 g/L
at 40°C	316 g/L
at 60°C	168 g/L

¹Figure from literature review by Environment Canada (1984).

2.2 Entry characterization

2.2.1 Production and use

According to a marketing research report by Lauriente in 1995, Canada produced 3.0 million tonnes of ammonia in 1990 and 3.4 million tonnes in 1993. Exports of ammonia from Canada were significant compared with imports. In 1989 and 1993, respectively, 1.1 million tonnes and 0.84 million tonnes of ammonia were exported, while 2000 tonnes and <500 tonnes were imported (Lauriente, 1995).

The primary industrial use of ammonia is as the nitrogen source in fertilizers, with direct application of anhydrous ammonia being the largest single method of consumption. The Prairie provinces are the largest users of these products, consuming 81% of the nitrogen content sold (Korol and Rattray, 1998). Ammonium sulphate, ammonium nitrate, urea and ammonium phosphate are fertilizers produced from ammonia. To a lesser extent, ammonia is used in many industrial applications. Ammonia can be regarded as reduced nitric acid and is used in the production of many other substances.

At petroleum refineries, ammonia is formed from catalyst regenerators in the fluid catalytic cracking process. Other uses of ammonia include the following:

• Manufacture of synthetic fibres (caprolactam for nylon), plastics and glues; pharmaceuticals, vitamins, amino acids, dentifrices, lotions and cosmetics; household ammonia, detergents and cleansers; numerous organic and inorganic chemicals, such as nitric acid, cyanides, amides, amines, nitrates, nitriles, hexamethylene diamine, ethanolamines, ammonium thiosulphate and dye intermediates.
• Production of explosives, rocket fuel, beer and nitrogen oxides required for manufacturing sulphuric acid, sugar purification and treatment and refinement of metals.
• Use as a refrigerant in both compression and absorption systems; a neutralizing agent for acids in oil protecting refinery equipment from corrosion; a flux for soldering; a treatment for wheat and barley straw as a supplement to sheep and cattle feed; a food additive; for growth and control of pH in yeast cultures; a latex preservative; a flame-proofing agent; a curing agent in leather making; a mothproofing agent; and as a reducing agent for nitrogen oxides in flue gas during steel production.
• Use as a dyeing agent, and preventing afterglow in matches.

Table 2 Percentage of total ammonia present as NH₃ in aqueous ammonia solutions

Temp. (°C)	pH
	6.0	6.5	7.0	7.5	8.0	8.5	9.0	9.5	10.0
0	0.008	0.026	0.082	0.261	0.820	8.55	7.64	20.7	45.3
5	0.012	0.039	0.125	0.394	1.23	3.80	11.1	28.3	55.6
10	0.018	0.058	0.186	0.586	1.83	5.56	15.7	37.1	65.1
15	0.027	0.086	0.273	0.859	2.67	7.97	21.5	46.4	73.3
20	0.039	0.125	0.396	1.24	3.82	11.2	28.4	55.7	79.9
25	0.056	0.180	0.566	1.77	5.38	15.3	36.3	64.3	85.1
30	0.080	0.254	0.799	2.48	7.46	20.3	44.6	71.8	89.0

2.2.2 Sources and releases

2.2.2.1 Natural sources

Much of the ammonia in the atmospheric environment is from natural sources. Since ammonia is continually released throughout the biosphere by the breakdown or decomposition of organic waste matter, any natural or industrial process that concentrates and makes nitrogen-containing organic matter available for decomposition represents a potential source of high local concentrations of ammonia.

Releases from natural processes that can be accounted for are approximately double the releases from the animal husbandry industry. Natural releases are estimated at over 500 000 tonnes while animal husbandry industry accounts for 294 000 tonnes to the atmosphere (Appendix A). The estimates of natural production and release to air are very approximate (Geadah, 1980; Environment Canada, 2000b).

2.2.2.2 Anthropogenic sources

2.2.2.2.1 Industrial

Environment Canada conducts an annual survey of Canadian industries (National Pollutant Release Inventory [NPRI]) likely to be using or releasing pollutants, including ammonia (NPRI, 1996). Total reported industrial releases of ammonia in 1996 were 32 037 tonnes. This makes ammonia the top-ranked NPRI substance in terms of amounts released in Canada. The NPRI has strict reporting criteria, such that most municipal sewage treatment plants, very few animal husbandry systems and no transportation systems had to report. These are known to be some of the major anthropogenic sources of ammonia released to the Canadian environment.

Industries in Alberta released more ammonia than industries in any other provinces in 1996, accounting for a third of the releases (9891 tonnes not including deep well disposal). This is due to the large number of fertilizer manufacturing facilities, pulp and paper mills and petroleum refineries in the province. Ontario released 7552 tonnes, and Quebec released 1914 tonnes (see Appendix B for details).

Industrial releases directly to watercourses totalled 5972 tonnes in 1996 and are typically from companies that are resource-based, such as pulp and paper, mining and coal-fired power generation, although a few significant releases come from heavy industries located in cities and from food processing. Fourteen companies were involved in pulp and paper manufacture in 1996, and these released a total of 1371 tonnes of ammonia that year. Three steel mills released 775 tonnes of ammonia, a single food processor released 504 tonnes, two fertilizer manufacturers released 180 tonnes, five mines released 537 tonnes and a coal-fired power generation plant released 62 tonnes. The rest (2543 tonnes) was released by many other industries.

Large releases in a city are usually due to one or two facilities. Hamilton, Ontario, is a major release site due to three steel-producing facilities. Other cities with significant industrial releases of ammonia are:

• Maitland, Ontario (fertilizer and chemical manufacture) -- air and water releases,
• Toronto, Ontario (chemical and paper manufacture) -- air and waste treatment releases,
• Medicine Hat, Alberta (fertilizer and chemical manufacture) -- air and waste treatment releases,
• Brandon, Manitoba (fertilizer and chemical/pharmaceutical manufacture) -- air and waste treatment releases, and
• Fort Saskatchewan, Alberta (fertilizer and chemical manufacture) -- air and water releases.

The fertilizer industry is the largest industrial releaser of ammonia in Canada. Of the 10 largest industrial sources of ammonia listed in the 1996 NPRI report, six are fertilizer manufacturers. Three are located in Alberta, at Redwater, Medicine Hat and Calgary, one in Manitoba, at Brandon, and two in Ontario, at Courtright and Maitland. Together they released 12 302 tonnes out of 32 037 tonnes reported, which is 38% of the total released. Most of these releases are to air.

The metal foundry industry is located primarily in Ontario. Of the largest ammonia sources in this sector, the top releasers are Algoma Steel in Sault Ste. Marie, Ontario (676 tonnes); the Cobalt Refinery Company, Fort Saskatchewan, Alberta (528 tonnes); the Inco Nickel Refinery, Copper Cliff, Ontario (297 tonnes); the Stelco refinery, Hamilton, Ontario (182 tonnes); and the Dofasco refinery, Hamilton, Ontario (180 tonnes). There are significant differences in the releases, however. The Inco and Cobalt facilities release nearly all of their ammonia to the air, Algoma Steel releases most of its ammonia to water, and the Stelco and Dofasco facilities split their releases between air and water. Other metal-working facilities, such as Stelwire of Hamilton, Ontario, produce large quantities of ammonia (245.5 tonnes), but they send all of it to the local municipal wastewater treatment plant (WWTP), so their releases will show under the municipal wastes for Hamilton.

The petroleum extraction and refining industry is a relatively large source of ammonia. The largest releasers within this sector in 1996, all in Alberta, were the Shell Scotford refinery, Fort Saskatchewan (2488 tonnes), Petro Canada refinery, Edmonton (1718 tonnes), and Imperial Oil Limited Strathcona refinery, Edmonton (1130 tonnes). However, their releases of ammonia are primarily to deep-well disposal, very little being released to surface waters. Following these are the Syncrude Canada Mildred Lake Site near Fort McMurray, Alberta (454 tonnes), and the Ultramar Ltée Raffinerie de St-Romuald, in St-Romuald, Quebec (229 tonnes). Syncrude does not have a reported release, as it uses massive retention ponds.

2.2.2.2.2 Municipal

Four sources of information were used to determine releases of ammonia from municipal sewage treatment plants. Environment Canada issued a voluntary survey to municipalities in all provinces (Environment Canada, 1997b) except Quebec (at their request) to collect information on effluent flow rates and releases of ammonia to local watercourses. Also requested was information on the quantities and ammonia content of sewage sludge and sludge disposal methods. The Ontario Ministry of Environment and Energy (OMEE, 1997) provided a copy of its municipal discharge database, which includes data on ammonia concentrations from all municipalities in Ontario. This was combined with Environment Canada's Municipal Water Use Database (Environment Canada, 1997c), which contains information on flow rates with which to calculate loading rates of ammonia from Ontario municipalities. A survey of 15 communities in Quebec provided ammonia release data and flow rates for a 3-day period in 1996 and 1997 (MEFQ, 1998). The average ammonia concentration in sewage effluents reported to Environment Canada (1997b) was 13.89 mg/L.

A survey of water usage by Canadian municipalities indicates that sewage treatment plants are a large source of ammonia to the aquatic environment. Average daily flow rates from all municipalities in Canada for 1994 were 12.3 × 10⁶ m³/day (Environment Canada, 1997c). This equates to 4.49 × 10¹² L/year. With an estimated average total ammonia concentration of 13.89 mg/L in domestic sewage (Environment Canada, 1997b), the estimated load to aquatic systems is 62 000 tonnes/year.

As the Environment Canada survey of municipalities was not exhaustive, being voluntary, the tonnage of ammonia released is a conservative estimate. Also, many treatment facility operators did not know the concentration of ammonia in their effluent and so did not provide those data. A figure of 13.89 mg/L (the national average) was used to estimate releases from these facilities.

The current knowledge of the quantities of ammonia disposed of in sewage sludge is not very good. The municipal survey did include questions on quantities of sludge produced and ultimate disposal methods; however, the answers received were of low quality. Many facilities record quantities of sludge in volumes, while others record weights. Many do not know the concentrations of ammonia in the sludge. An estimate of 5722 tonnes of ammonia disposed of in sludge was generated, based on 1222 tonnes of ammonia reported and 4500 tonnes extrapolated from the reports. The average ammonia concentration reported in sludge was 2200 mg/kg, with a range of 0.29-38 600 mg/kg.

Appendix C summarizes the ammonia loading rates in municipal wastewater for the top 12 urban centres in 1995-1996. Montréal has the highest ammonia loading rate of any city in Canada. This is due to an effluent flow rate twice that of Toronto, although its effluent is surprisingly dilute for a primary treatment system. Montréal's loading rate of 6128 tonnes/year to the St. Lawrence River is nearly the same as Toronto's, at 5938 tonnes/year released to Lake Ontario. The Greater Vancouver Sewerage and Drainage District (GVS&DD) is the next largest releaser of ammonia (5741 tonnes/year to the Fraser River and Strait of Georgia), as its effluents are moderately high in ammonia and have a high flow rate. The four Vancouver facilities have either primary or secondary treatment. Winnipeg is fourth because of its very high ammonia concentration, averaging 26 mg/L in 1995, giving it a loading rate of 2152 tonnes/year to the Red River. Edmonton is fifth, but by 2005 Edmonton's loading rate should be at least half of what it is currently, as the city is installing a nitrification/ denitrification process. Many urban centres have several sewage treatment facilities; for these calculations, their effluent flows and ammonia concentrations have been flow weighted and aggregated.

A comparison was made of industrial releases directly to water and municipal releases to water. The ammonia from municipal effluents in any province far outweighs the industrial discharges of ammonia to water.

2.2.2.2.3 Combined sewer overflows

Combined sewer overflows occur when stormwater drains are routed into the sewage system, so that the sewage treatment system is overloaded during a large rain event. When overloading occurs, raw sewage is diverted directly into the receiving water along with the stormwater. Mean concentrations of ammonia in stormwater estimated for three Ontario cities -- Sarnia, Sault Ste. Marie and Windsor -- were 0.5, 0.7 and 0.3 mg NH₃/L, respectively (Marsalek and Ng, 1989). Annual average ammonia concentrations in municipal wastewater effluents were 75.4, 181.5 and 27.7 mg/L, respectively. Concentrations were calculated from point source loadings that were divided by the annual volume of stormwater runoff (UGLCC, 1988a,b,c). When comparing loadings in the stormwater with those in combined sewer overflows, overflows exceeded stormwater in loadings of ammonia in both Sarnia and Windsor, despite the fact that stormwater discharges in Sarnia and Windsor were 6.7 × 10⁶ and 22.3 × 10⁶ m³/year, respectively, and the volumes of combined sewer overflow in Sarnia and Windsor were 1.0 × 10⁶ and 5.2 × 10⁶ m³/year, respectively (Marsalek and Ng, 1989).

2.2.2.2.4 Agricultural

Manure application

Few, if any, agricultural operations track ammonia emissions. Livestock manure is considered to be the major source of NH₃ emission to the atmosphere (Ryden et al., 1987); however, quantification of this source is difficult and comes down to estimates of loss rates, ammonia concentrations in manure and numbers of animals (refer to Appendix A). Canada has a large population of farm animals (about 114 million, mostly cattle, swine and poultry). Emission factors range from 0.32 kg NH₃ per animal per year for poultry to 40 kg NH₃ per animal per year for beef cattle (U.S. EPA, 1994). This source generated an estimated 294 000 tonnes of ammonia in 1995 (Environment Canada, 2000b).

Manure is spread on land mainly as a way of disposing of farm animal waste. Land application of manure is higher in regions where farm animal production is high and the manure can be collected and distributed easily. In the Prairies, cattle are concentrated only in certain areas, and fields are large and require large amounts of manure for complete coverage. Although Alberta's large cattle population produces over 25% of all the animal manure in Canada (Patni, 1991), only a small proportion is confined at any time, when manure collection and land application are practical. Mixed farms and ranches are not well suited for collection and application of manure, except from cow-calf operations and feedlots. Manure in British Columbia and central and eastern Canada comes mostly from dairy and poultry farms, which are numerous and scattered throughout crop production regions. This makes it easier to get the manure to the fields where it is needed. Many dairy farms grow silage corn as cattle feed, and the manure from the cattle is applied to the cornfields, creating an on-farm nutrient cycling system.

Agricultural research shows that 10-75% of the ammonia in cattle manure can volatilize if the manure is not incorporated into the soil within a week. During hog production, 40-95% of the excreted nitrogen may be lost before the manure reaches the field. This nitrogen is lost primarily as ammonia volatilized from barns, from manure storage facilities and following field application. Many of the new hog production facilities in Canada, which include lagoon storage and slurry irrigation, will probably result in NH₃ emission losses up to 75% of the excreted nitrogen (Paul, 1997). Lockyer and Pain (1989) showed losses of up to 83% from poultry slurry, 21% from air-dried poultry manure, 36-75% from pig manure and 41% from cattle manure when applied to turf and not incorporated. In most cases, 80% or more of the ammonia loss occurred within 48 hours of application. Air-drying of poultry manure reduced ammonia losses to 12% from the poultry house and from application. Lockyer and Whitehead (1990) conducted ammonia loss experiments with cattle urine applied to soil. They found that 3.7-26.9% of the ammonia in the urine was lost within 15 days of application. Most of the loss occurred within the first 4 days. The temperature of the soil was the most important factor in determining the amount of ammonia lost to the atmosphere. They estimated average ammonia losses to air for grazing systems to be 37 kg N/ha or 12% of the ingested nitrogen content in the forage.

The quantities of ammonia lost from the soil decline considerably if the manure is liquefied and injected under the surface. Hoff et al. (1981) showed that liquid swine manure lost 11-14% of the ammonia when applied to the surface and only 2.5% when injected. Ryden et al. (1987) showed that cattle manure lost 16-32% of the ammonia when applied to the surface and only 0.9% of the ammonia when injected under the surface.

Storage systems for manure are a source of ammonia loss as well. Under acid conditions, NH₄⁺ , which is relatively non-volatile, predominates, while under basic conditions, NH₃ , which will evaporate readily, predominates. In anaerobic decomposition of poultry manure, low-pH conditions (pH 5-6) led to a low (1%) loss of ammonia. Under aerobic decomposition, the basic pH that developed (pH 8.4-8.9) promoted loss of ammonia (9--44%). Losses of ammonia from storage tanks were reduced by up to 85% simply by covering the tanks; even a tarp sufficed (de Bode, 1990). There are many ways of handling, storing and applying manure from various animals that will reduce the loss of ammonia. The most obvious are quick storage of manure, covering manure pits to prevent volatilization, promoting anaerobic storage conditions, acidifying manure to prevent formation of NH₃ from NH₄⁺ , tillage of soil prior to surface application, injection of manure slurries into soil rather than surface application, and application during wet or cool weather (McGinn and Pradhan, 1997).

Mineral fertilizer application

Ammonia loss by volatilization from mineral fertilizers depends in large part on the soil pH, due to the overriding dependence of ammonia ionization on pH. NH₃ will evaporate readily, whereas NH₄⁺ will not.

Some inorganic nitrogen fertilizers are acidic, so that NH₃ loss from these materials depends on soil chemical reactions or on the inherent alkalinity of the soil. Representative inorganic nitrogen fertilizers in this category are ammonium nitrate, diammonium sulphate and ammonium chloride. If the soil is sufficiently alkaline (with calcium carbonate usually), the reactions will form diammonium carbonate, which is unstable and decomposes, producing NH₃ and carbon dioxide gases (Fenn and Hossner, 1985).

The concern over nitrogen loss from industrial fertilizers has resulted in massive amounts of scientific literature on many relevant aspects. Top-applied ammonium sulphate has been measured, in both field and laboratory, to lose up to 55% of the NH₃ -N. Losses from unincorporated urea can be very high, up to 60% over a period of 4 days in surface-applied pastures (Fenn and Hossner, 1985). One study by Touchton and Hargrove (1982) found that a 270 kg N/ha application of urea-ammonium nitrate to the surface resulted in less nitrogen uptake by corn than from an application of 90 kg N/ha with incorporated urea-ammonium nitrate. This equated to a loss of 67% of the NH₃ -N when urea-ammonium nitrate was not incorporated.

The Environment Canada Ammonia Air Emissions Inventory for 1995 (Appendix A) estimates ammonia lost from the application of individual types of fertilizer, using appropriate emission factors and sales data. The amounts lost to the atmosphere are greatest from urea, accounting for an estimated 72% of the losses (130 217 tonnes) from applied fertilizers (Environment Canada, 2000b). The Canadian Fertilizer Institute reported 3 million tonnes of urea produced that year (CFI, 1997), so a loss of 130 217 tonnes would be 4.3% of the total applied. This would be consistent with reported losses for incorporated urea, which is the recommended method of application.

Plant tissues

There is some evidence that plants play an important role in the concentrations of ammonia in the atmosphere. The maintenance of low ambient concentrations depends in part on the existence of an NH₃ compensation point, i.e., an atmospheric NH₃ concentration above which plants will absorb NH₃ from the air and below which they will release it. Denmead et al. (1977) found that, even though there was considerable release of NH₃ from the ground in a grass-clover pasture, almost none of it escaped to the atmosphere above the canopy. The effect of plant absorption was to reduce the NH₃ concentration in the air from >16 µg/m³ near the soil surface to 1 µg/m³ at the top of the canopy. Other researchers report a similar phenomenon in a field of quack grass (Agropyron repens), reducing the NH₃ concentration from 40 µg/m³ above the grass canopy to 3 µg/m³ within it (Lemon and van Houtte, 1980). Accurately estimating losses from this source would be extremely difficult.

Runoff of ammonia from soil

Runoff of nutrients, including ammonia, from various land use types, including intensive livestock operations and crops, has been studied to the extent that we know the vast majority of nutrients in runoff are associated with either soluble phosphorus or nitrate. Both of these are water-soluble and easily transported in solution. Ammonia in ionized form, on the other hand, is typically tightly bound to soil colloids and is not easily transported in solution once it contacts soil. Once ammonia binds with soil, it will travel with soil particles during erosive events. Minor amounts of ammonia will travel in solution if there is freely available ammonia on the soil surface. In spring, however, considerable quantities of ammonia can be liberated as runoff from melting snow. This is due to the accumulation of ammonia trapped in snow or from deposited manure. The soil is still frozen, so that little ammonia will be absorbed and bound; what does not evaporate over winter travels with runoff and enters waterways in the freshet. Data on levels of ammonia in runoff are reported in Section 2.3.2.4.

2.3 Exposure characterization

2.3.1 Environmental fate

The nitrogen cycle is an attempt to describe the natural cycling of nitrogen from the atmosphere through incorporation into living organisms and from them back into the abiotic environment through degradative processes. Figure 1 illustrates the nitrogen cycle (after Manahan, 1994).

Several processes can create nitrogenous compounds usable by organisms from nitrogen gas (N₂). Lightning and cosmic radiation combine atmospheric nitrogen and oxygen into nitrates, which are carried to the earth's surface in precipitation. A few nitrogen-fixing bacteria, symbiotic mycorrhizal fungi living on the roots of plants, cyanobacteria, and certain lichens and epiphytes in tropical forests can split N₂ and make the nitrogen molecule available for amino acid synthesis. Ammonia is formed either as a waste product or when plants and animals die. Another set of microorganisms is capable of using NH₃ and eventually forming nitrate (NO₃^-) and nitrous oxide (N₂O).

2.3.1.1 Air

Ammonia is released into the atmosphere by agricultural, waste disposal and industrial activities. There is no known photochemical reaction by which ammonia could be produced in the atmosphere (WHO, 1986). Atmospheric ammonia undergoes four primary types of reactions: gas-phase, liquid-phase, thermal and photochemical. The first two are the most important types of reactions. From various studies consulted, the main reactions of interest appear to be those associated with the following combinations of reactants, since there is a high availability of nitric acid (HNO₃), hydrochloric acid (HCl), sulphur dioxide (SO₂) and sulphuric acid (H₂SO₄) in the atmosphere as a result of industrial and urban emissions:

• ammonia/nitric acid/ammonium nitrate
  (NH₃/HNO₃/NH₄NO₃),
• ammonia/hydrochloric acid/ammonium chloride
  (NH₃/HCl/NH₄Cl),
• ammonia/nitric acid/sulphuric acid
  (NH₃/HNO₃/H₂SO₄), and
• ammonia/sulphur dioxide (NH₃/SO₂).

Figure 1 The nitrogen cycle

In a polluted atmosphere, ammonia reacts with nitric acid and/or hydrochloric acid, which results in the formation of ammonium nitrate and/or ammonium chloride. These ammonium salts account for 10-30% of the fine aerosol (solid or liquid particles suspended in a gas with a particle diameter <0.5 µm) in a polluted atmosphere. These aerosols are very sensitive to temperature and relative humidity. In a polluted atmosphere, ammonia can also react with more than just one pollutant (Bassett and Seinfeld, 1983):

(1)

3NH₃ (g) + 2HNO₃ (g) + H₂SO₄ (g) ⇔ (NH₄)₂SO₄·2NH₄ NO₃ (s)

and

(2)

5NH₃ (g) + 3HNO₃ (g) + H₂SO₄ (g) ⇔ (NH₄)₂SO₄·3NH₄ NO₃ (s)

An intensive study (Esmen and Fergus, 1977) of the NH₃/SO₂ reaction system in a dry atmosphere led to the confirmation of at least eight reactions, the most important of which are:

(3)

NH₃ + SO₂ + H₂O ⇔ NH₄HSO₃

and

(4)

NH₂ HSO₃ + NH₃  ⇔ NH₂·SO₃ NH₄

The above reactions were found to be very rapid, on the order of milliseconds. In the atmosphere, gas-phase reactions are expected to occur in the presence of water droplets. Aerosol formed through gas-phase reactions is therefore expected to act as condensation nuclei or to be captured by available drops to contribute high local concentrations of bisulphite ion (HSO₃^-). This is significant in the eventual formation of the ammonium sulphate [(NH₄)₂SO₄] aerosol in liquid-phase reactions. For this reason as well, reaction (3) is very important in the presence of water.

With respect to the liquid-phase reactions of the NH₃/SO₂ system, various studies (Moller and Schieferdecker, 1985; Behra et al., 1989; Plass et al., 1993) concluded that the function of NH₃ in this system is to neutralize the hydrogen ions formed in the absorption of sulphur dioxide and its subsequent oxidation to sulphate. Thus, NH₃ maintains sulphur dioxide solubility and the rate of sulphate production by buffering the pH to between 4 and 5.

Of all known atmospheric ammonia reactions, one of the most important seems to be that involving conversion of ammonia to ammonium (NH₄⁺) particulate (see reaction 4 above). This conversion occurs in the lowest 100 m of the atmosphere at rates in the range 1 × 10^-3/s to 5 × 10^-5/s (/s is indicative of a first-orderreaction, which means the reaction rate is dependent on the concentration of one reactant, namely NH₃), and daytime conversion is much faster than that at night. The reaction is dependent on temperature, relative humidity and pH (Fangmeier et al., 1994).

All studies consulted conveyed the opinion that the main factors that hindered long-range transport of ammonia in the atmosphere, both vertically and horizontally, were rapid conversion to ammonium aerosol and the relatively high dry deposition velocity of ammonia.

Because of the rapid reaction rates of ammonia in air, anywhere from 56% (ECETOC, 1994) to 94% (Moller and Schieferdecker, 1985; Quinn et al., 1988; ECETOC, 1994) of atmospheric ammonia is converted to ammonium particulate/ammonium aerosol. Over oceans, ammonium particulate/ammonium aerosol has an estimated atmospheric residence time of 22 hours (Quinn et al., 1988); over land, the estimated residence time is in the range 7-19 days (Moller and Schieferdecker, 1985; Fangmeier et al., 1994). In comparison, the estimated residence time of atmospheric NH₃ is 3.6 hours over oceans (Quinn et al., 1988) and in the range of 2.8 hours to 4 days over land (Fangmeier et al., 1994). These short residence times are primarily due to the rapid conversion to ammonium particulate/ammonium aerosol and the high dry deposition velocities of ammonia (Asman and Janssen, 1987; Asman et al., 1989).

Figure 2 summarizes the chemistry, distribution, transport and deposition of atmospheric ammonia. Depending on the atmospheric conditions, anywhere from 56% to 94% of atmospheric ammonia is converted to ammonium particulate/ammonium aerosol, and less than 1% is converted to nitric oxide (NO). The balance remaining, 6-44%, is gaseous ammonia.

Figure 2 Fate analysis of ammonia in the atmosphere

V_d = net dry deposition velocity

[Emission: 100%] [< 5 km: Dry deposition] [10's - 1000s km: Wet deposition]

It is known from measurements that gaseous ammonia concentration rapidly decreases with height and distance from ground-level emission sources. Results of four studies on the decrease in ammonia concentration with an increase in distance from a ground-level emission source showed that 50-75% of the gaseous ammonia detected was deposited between 500 and 4000 m from the source (Denmead et al., 1982; Asman et al., 1989; Fangmeier et al., 1994; Janzen et al., 1997).

Gaseous ammonia is removed from the atmosphere via dry deposition, whereas ammonium aerosol is removed via both dry and wet deposition. Dry deposition is most significant in regions with high ammonia emissions and is indicative of short-range transport of less than about 5 km. In contrast, wet deposition is most significant in regions with low ammonia emissions and is indicative of long-range transport, ranging from tens to thousands of kilometres distant (ECETOC, 1994; Fangmeier et al., 1994).

Ammonia may be a significant local pollutant and, as a precursor of nitric oxide and ammonium aerosols, can have long-range impacts.

2.3.1.2 Surface water

A fate analysis for ammonia in the aquatic environment is displayed schematically in Figure 3 and is a composite from several reviews (NRC, 1979; API, 1981; WHO, 1986). Ammonia has a critical role in the nitrogen cycle, so that when it is introduced into aquatic systems, it is usually rapidly transformed into other nitrogenous forms (e.g., nitrates and organically bound nitrogen). The major processes include fixation, assimilation, ammonification, nitrification and denitrification. In the sections below, these major processes are discussed along with reference to decreases in ammonia concentration due to dilution.

Figure 3 Fate analysis of ammonia in aquatic environments

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Figure 3 also presents the relative importance of input sources and depletion routes for ammonia in the aquatic environment. Nitrification and volatilization are the important and competitive fate processes in surface waters that are not ice covered. Under ice cover, both processes are greatly reduced. Volatilization is the predominant removal process for industrial effluents until the effluents are diluted to a concentration that is not harmful to nitrifying bacteria. These bacteria require a substrate on which to grow, typically suspended solids in the water. Nitrification processes are more likely to be significant in lakes, slow-moving rivers, estuaries and sewage effluents. Nitrification is important in preventing the persistence or accumulation of high ammonia levels in water receiving sewage effluent or runoff. However, winter conditions will inhibit bacterial growth so that under restricted water flow conditions, ammonia can build up in receiving waters. Conditions of high nitrification may contribute to low levels of dissolved oxygen, as nitrification is an oxygen-consuming process (WHO, 1986).

Nitrification of ammonia can also have significant impacts on water systems by promoting acidification. In a greenhouse, seven identical mini-ecosystems, simulating soft-water ponds, were exposed to different types of artificial rainwater. Although ammonium sulphate deposition was only slightly acidic, due to nitrification it acted as an important acid source, causing acidification to pH 3.8. Under acidified conditions, ammonium sulphate deposition led to a luxuriant growth of Juncus bulbosus and Agrostis canina. In the mini-ecosystems, sulphuric acid deposition with a pH of 3.5 decreased the pH of the water to only 5.1 within 1 year (Schuurkes et al., 1986).

Loss of ammonia to the atmosphere at elevated pH is another mechanism for ammonia removal. It has been estimated (API, 1981) that volatilization could account for 67.5% of the observed loss of ammonia below an industrial discharge to the Wabash River in the United States. It was also estimated that 20% of the ammonia discharged by a fertilizer plant was lost to the atmosphere, and a 55% loss for 10-year, 7-day low-flow conditions was predicted.

2.3.1.3 Soil and groundwater

A schematic of the terrestrial nitrogen cycle is represented in Figure 1. Ammonium is an important intermediate in the assimilation of nitrogen from the soil by plants. Nitrogen is present in the soil largely in the organic form and is unavailable to plants. Microbial processes must mineralize it. As nitrification is an energy-yielding process, the rates of conversion are rapid, so that ammonium rarely accumulates in soil while bacteria are active. Organic nitrogen compounds are reduced to ammonium, which is converted to nitrite (NO₂^-) by Nitrosomonas and then to nitrate by Nitrobacter (API, 1981; WHO, 1986). Most plants can assimilate the ammonium ion, but it is usually oxidized to the nitrate ion, the most common form of mineralized nitrogen in soil, which may be assimilated by plants as well (NRC, 1979; WHO, 1986).

Another source of mineralized nitrogen is nitrogen fixation, where gaseous nitrogen is transformed to ammonium ion, usually by metabolic processes. Nitrogen fixation occurs in blue-green algae and a few genera of microorganisms, which include aerobic bacteria, such as Azotobacter species, anaerobic bacteria, such as Clostridium species, and organisms in symbiotic association with higher plants, such as Rhizobium species found in legumes. Volatilization, adsorption and chemical transformation will also affect the fate of ammonia in soil (NRC, 1979; WHO, 1986).

Ammonia is bound in soil by the attraction of the positive charge on the ammonium ion to the negatively charged soil micelles. In soil, ammonium is adsorbed primarily by four mechanisms: chemical (exchangeable), fixation (non-exchangeable), reaction with organic matter and physical attractive forces.

Since ammonia is so poorly mobile in soil, it is unlikely to leach to groundwater except under unusual circumstances, such as when the cation exchange capacity of the soil is exceeded. The worst situation for ammonium leaching would probably occur when the soil is at field capacity with respect to water. In this case, ammonium ions can penetrate the soil and continue downward, with only small amounts remaining as part of the interstitial fluid. Moisture that is present in the soil or added as precipitation will dilute ammonia on the surface and reduce its rate of evaporation.

If ammonium ions reach the groundwater table, they will continue to move in the direction of groundwater flow and will be diluted slowly through diffusion or will be adsorbed by soil and mineral particles. It is possible that deep-soil bacteria utilize ammonia for amino acid synthesis in the presence of oxygen.