Aluminum salts final content: chapter 3.1

3. Assessment of toxic under CEPA 1999

3.1 CEPA 1999 64(a) and 64(b): environment

3.1.1 Environmental risk characterization

The approach taken in the ecological component of this risk assessment was to review new information relevant to the three aluminum salts recommended for assessment by the Ministers' Expert Advisory Panel (that is, aluminum chloride, aluminum nitrate, aluminum sulphate), and to evaluate this information with reference to the original characterization of potential risk presented in Environment Canada and Health Canada (2000).

Environment Canada and Health Canada (2000) identified the pelagic, benthic and soil compartments as primary media of potential exposure for aluminum derived from the three salts subject to assessment, and conducted an analysis of potential risk for each compartment. This analysis is provided in the sections below, along with additional information collected subsequent to the publication of the 2000 assessment and deemed relevant to the evaluation of potential risk.

3.1.1.1 Aquatic organisms

3.1.1.1.1 Pelagic

Environmental exposure in water to aluminum from the three aluminum salts is expected to be greatest in areas near direct releases of process wastewater to the aquatic environment. Unfortunately, few measured data are available for receiving environments following direct releases from water treatment facilities or pulp and paper mills. In addition, measurements of total concentrations of a metal can rarely be correlated directly with their biological effects. Metal in particulate form is generally considered to be less available for uptake by organisms, and the formation of complexes with inorganic (for example, OH^-, SO₄^2-) or organic (for example, fulvic acid) ligands can reduce the available fraction of the dissolved form of a metal. Speciation modelling using the estimation models MINEQL+ and WHAM was conducted in order to estimate the level of dissolved inorganic monomeric aluminum present in rivers following release of wastewater from eight drinking water treatment plants (DWTPs) and two pulp and paper plants (Germain et al. 2000). The modelling provided results in the pH range of 6.56 to 8.38 and therefore the dissolved monomeric aluminate ion, Al(OH)₄^¯, would be the predominant aluminum species present (see Figure 2.1). As indicated in Section 2.4.1, dissolved inorganic monomeric aluminum is considered to have the highest bioavailability to aquatic species and to present the greatest risk of adverse effects to pelagic organisms. The level of dissolved inorganic monomeric form of aluminum was calculated, using aluminum levels estimated in effluents (Fortin and Campbell 1999) and assuming a 1:10 dilution. For the DWTPs considered, average concentrations of dissolved inorganic monomeric forms of aluminum (which are assumed to be the bioavailable forms) at saturation varied from 0.027 to 0.348 mg/L during backwash events, assuming that microcrystalline gibbsite is controlling the aluminum solubility. According to Hem and Robertson (1967), microcrystalline gibbsite controls aluminum solubility at pH values of less than 7, while the precipitate formed when the potential of hydrogen (pH) of water is in the 7.5-9.5 range has a solubility similar to that of boehmite. This precipitate will evolve to bayerite, a more stable and insoluble form of aluminum hydroxide, within a week. If it is assumed that boehmite is controlling the solubility, dissolved aluminum levels would be lower, ranging from 0.005 to 0.059 mg/L (Fortin and Campbell 1999). For the two pulp and paper mills considered, the dissolved aluminum values were among the lowest, whatever form is controlling the aluminum solubility.

The calculated dissolved aluminum concentration of 0.348 mg/L represents the saturation concentration, assuming that microcrystalline gibbsite controls solubility when aluminum salts are used to treat drinking water. This value was calculated for a location in the Canadian Prairies, where the pH of receiving waters (8.38) and solubility were the highest of all sites examined (Fortin and Campbell 1999). Backwash events can be considered to last for about 30 minutes and occur every 48 to 72 hours for each filter at a DWTP (Environment Canada and Health Canada 2000). If it is assumed that most DWTPs have about 20 filters (small DWTPs have fewer filters), it is estimated that concentrations in receiving waters near the point of discharge could be as high as 0.348 mg/L as much as 10% of the time. The rest of the time, aluminum concentrations would approach background values, which, for locations on the Prairies, are likely on average to be about 0.022 mg/L as monomeric inorganic aluminum (Environment Canada and Health Canada 2000). The temporally weighted concentration of dissolved monomeric aluminum at this location averaged over a period of several days would therefore be about 0.055 mg/L. This concentration was taken as a conservative (reasonable worst-case) Predicted Environmental Concentration (PEC) for waters close to discharge points.

Because aluminum releases reported by DWTPs occur in circumneutral to neutral waters, two critical toxicity values (CTVs) corresponding to the pH of waters where releases occur could be chosen. The work of Neville (1985) provides a no-observed-effect concentration (NOEC) of 0.075 mg/L as inorganic monomeric aluminum, based on the absence of deleterious effects on ventilation and respiratory activity of rainbow trout at pH 6.5. This CTV is considered valid for the pH range 6.5-8.0. A second CTV for alkaline conditions (pH > 8.0) is based on the work of Gundersen et al. (1994), who determined similar LC₅₀s (~ 0.6 mg dissolved Al/L) during several experiments in the pH range 8.0-8.6 and water hardness range 20 to 100 mg/L (as calcium carbonate). A NOEC for mortality of 0.06 mg dissolved Al/L can be derived for rainbow trout from data given for one of the 16-day exposures at 20 mg/L hardness and pH 8.0. The chemical concentrations in Gundersen et al. (1994) are expressed as "total" and "dissolved" aluminum; there was, unfortunately, no attempt to identify the forms of dissolved aluminum present. At the experimental pH, it is probable that a good proportion of the dissolved aluminum was the monomeric aluminate ion as the predominant species.Since the pH in waters for which the predicted environmental concentration (PEC) was estimated is 8.38, the corresponding CTV is 0.06 mg/L as dissolved inorganic monomeric aluminum.

It is possible that effects may be elicited at concentrations below that of the selected CTV of 0.06 mg/L. Wold et al. (2005) reported a 21-day lowest-observed-effect Concentration (LOEC) for reduced survival and reproduction in Daphnia pulex at a lowest test concentration of 0.05 mg/L. Testing was conducted at a pH of 7 ± 1, suggesting that the observed effects were due to the presence of aluminum hydroxide rather than the dissolved inorganic monomeric aluminum that is usually associated with toxicity. Recent studies (for example, Verbost et al. 1995; Kádár et al. 2002; Alexopoulos et al. 2003) provide evidence that the particulate and/or colloidal forms of aluminum, such as may be present under the transition conditions of mixing zones, are bioavailable and can exert adverse effects on organisms. Impaired oxygen consumption, gill damage, and reduced feeding behaviour have been reported in aquatic invertebrates and fish present in waters containing freshly neutralized aluminum (that is, aluminum in transition from ionic species to polymers or precipitating hydroxides), although it is not clear whether these effects result from physical damage to structures such as the gills, or from direct chemical toxicity. Therefore, while there may be circumstances or conditions under which particulate and colloidal forms of aluminum can exert adverse effects on aquatic organisms, these conditions are likely to be localized and/or transitory in nature, and the selected CTV of 0.06 mg/L, based on the inorganic monomeric form, is considered sufficiently representative of the overall potential for adverse impacts in aquatic species.

In determining predicted no-effect concentrations (PNECs) for aluminum, the nature of the biological response was considered, since some organisms respond to a narrow aluminum concentration range. This results in an abrupt "threshold" where an evident biological response occurs, with no observable effects at slightly lower concentrations (Hutchinson et al. 1987; Roy and Campbell 1995). Consequently, since the CTV chosen is a NOEC, the application factor used to derive a PNEC from the CTV was 1. Aluminum being a natural element, it is also useful to consider whether the PNEC is within the range of natural background concentrations. Although based on limited data, on an overall basis, the 90th-percentile value for dissolved aluminum at sampling stations located upstream of points of discharge of aluminum salts is 0.06 mg/L (Germain et al. 2000). It should be noted that only a portion of this dissolved aluminum is in inorganic monomeric forms (corresponding to the PNEC). Thus, the 90th-percentile value for inorganic monomeric aluminum in uncontaminated water is expected to be less than 0.06 mg/L.

The reasonable worst-case quotient for receiving water can therefore be calculated as follows:

Quotient:

= PEC ÷ PNEC

= 0.055 mg/L ÷ 0.06 mg/L

= 0.92

Since this conservative quotient is relatively close to 1, it is helpful to consider further the likelihood of biota being exposed to such concentrations in Canada.

It is likely that chemical equilibrium modelling overestimates inorganic forms of aluminum in solution, since it appears to overestimate dissolved aluminum. One reason for the overestimate is that a very large fraction of the aluminum released from DWTPs during backwash events is most probably in solid form, while calculations used to estimate the PEC assumed that all of the aluminum was in dissolved form (Germain et al. 2000). Although the modelling assumed that saturation was achieved instantly, this "solid" aluminum may take a relatively long time to dissolve such that aluminum levels in receiving waters do achieve saturation. In fact most of the aluminum solids released are expected to settle relatively quickly to bottom sediment. Dissolved concentrations may also be overestimated because of the assumption that the solubility of aluminum is controlled by microcrystalline gibbsite. Based on limited data on concentrations of dissolved aluminum at different treatment steps at one Canadian DWTP, solubility may be controlled by less soluble forms of aluminum hydroxide, such as boehmite (Fortin and Campbell 1999).

The possibility that modelled concentrations overestimate actual values is further supported by data for two sites on the North Saskatchewan River, where the dissolved inorganic aluminum concentrations predicted by modelling are 0.110 and 0.099 mg/L, while the measured concentrations at these sites are 0.005 and 0.010 mg/L (Roy 1999b).

Srinivasan et al. (1998) studied the speciation of aluminum at six different stages of water treatment at Calgary's DWTP. The total aluminum concentration ranged from 0.038 to 5.760 mg/L, and the dissolved inorganic aluminum concentration varied from 0.002 to 0.013 mg/L. George et al. (1991) measured < 0.06 mg monomeric Al/L in alum sludge from ten different DWTPs containing up to 2,900 mg total Al/L. These results show that the concentration of dissolved aluminum in process wastewaters is less than the PNEC.

Finally, while the potential for aluminum to influence the cycling and availability of phosphorus and other trace elements in aquatic systems is recognized (see Section 2.3.1; Environmental Fate), no empirical data were found to suggest the occurrence of this process in Canadian surface waters and, in particular, as a result of aluminum released from the three aluminum salts that are the subject of this assessment. For this reason, the potential for risk from this source will not be evaluated further here.

3.1.1.1.2 Benthic

Acute toxicity to benthic and pelagic organisms resulting from exposure to potentially high concentrations of aluminum in aluminum-based sludge is unlikely, because of the solubility constraints in receiving waters discussed above. Filtrates obtained from alum sludge were toxic to freshwater algae in waters with low pH (less than 6) or low hardness (less than 35 mg/L CaCO3/L); however, the available information indicates these conditions are not prevalent in Canadian waters that receive large inputs of aluminum from the three aluminum salts being assessed. AEC (1987) determined that aluminum was effectively bound to sludge within the pH range of 4.5 to 10.0, with less than 0.02% of the total aluminum released in waterwaters dissolved in the liquid phase associated with the sludge.

Hall and Hall (1989) reported delayed and reduced reproduction in Ceriodaphnia dubia following exposure to undiluted alum sludge effluent, suggesting that sublethal effects may be possible in the environment. However, effluent dilution occurs immediately upon release into a receiving water body. In addition, any observed ecosystem impacts would be difficult to link directly to the presence of aluminum given the potentially large number of contaminants that may also be present in the sludge.

There is evidence that aluminum sludge released from DWTPs can deposit and form a blanket over sediments in rivers with slow water velocity, and macroinvertebrate populations may be stressed due to a lack of oxygen and carbon sources on which to feed. For this reason, George et al. (1991) recommended that sludge be discharged during periods of fast water movement as this may be less detrimental to primary producers and benthic communities. AEC (1984) reported smothering effects related to settled sludge on sediments following disposal to rivers in Alberta may occur but concluded that while there is potential for adverse impacts resulting from the deposition of alum sludge in receiving waters, further research is needed. The study recommended alternative treatment and disposal methods for alum sludge be considered, including reduction in the quantities produced through substitution with alternative coagulants, routing of the sludge through sanitary sewer, lagooning, and landfilling or land application.

The City of Ottawa (2002) found depressed abundance of benthic organisms downstream from the Britannia DWTP up to 1,500 m from the discharge site compared to upstream sampling sites. Areas of sediment with an appearance similar to depositons at the outfall from the Britannia DWTP and with higher levels of aluminum were found 1,500 m downstream from the outfall while sampling sites closer to the discharge did not exhibit such strong similarities, and had lower concentrations of aluminum which approached aluminum concentrations found in the sediment 150 m upstream from the discharge.  This study thus showed that sludge sediment from the the Britannia DWTP can travel to distant locations from the point of discharge where deposition may occur due to site specific hydrological characteristics. In this study, it was also unclear whether the identified impacts were a result of the physical composition of the sediments (for example, grain size), on-going blanketing of the area, and/or toxicity of dissolved aluminum leaching out of sediment and into the water column.

In their environmental risk assessment guidance document for metals, ICMM (2007) indicate that trace metals discharged into aquatic ecosystems are most likely to be scavenged by particles and removed to sediments. Once associated with surface sediments, the metals are subjected to many types of transformation reactions, including formation of secondary minerals, and binding to various sediment fractions (for example, sulphides, organic carbon, iron hydroxides). For this reason, it may be difficult to establish clear relationships between measured concentrations of a metal in sediment and the potential for impacts to benthic organisms.

Overall, the greatest potential for risk to the benthic environment resulting from the release of aluminum-based effluents and sludges likely relates to the physical effects of blanketing and smothering of benthic communities in the vicinity of the outfall. While this impact does not constitute direct aluminum toxicity, the presence of aluminum coagulants and flocculants in water treatment processes results in the formation of substantial quantities of sludge, which may then be released into the environment. It is reasonable to expect that physical impairment of bethic populations would not be limited to alumimum coagulants sludge, but could also result from any other chemical coagulant used for the treatment of drinking water.  However although the potential for local impacts to benthic organisms exists, there are relatively few reports of such damage.

In recognition of the potential for adverse ecosystem effects, many Provinces have implemented strategies designed to reduce or eliminate the release of water treatment plant effluents and sludges to receiving water bodies (see Section 2.2.2). It is expected that addressing issues relating to overall effluent and sludge concerns, most notably the extremely high levels of total suspended solids (TSS) should also effectively deal with physical and chemical aspects of aluminum sludge toxicity in the aquatic receiving environment.

3.1.1.2 Terrestrial organisms

Terrestrial organisms are exposed to added aluminum when alum sludge from water treatment facilities, primarily MWWTPs, is applied to agricultural soils.

The lowest level of dissolved aluminum reported to adversely affect terrestrial organisms is 0.135 mg/L, which can reduce root and seedling growth in sensitive grain and forage crops. This concentration was therefore selected as the CTV, assuming that most of the dissolved aluminum was in inorganic monomeric forms. Considering that this CTV was derived from experiments using solution cultures, the effects data on which the CTV is based could overestimate the sensitivity of crops grown in soils in the field. Because of that, the fact that many species were affected at the same low level and the fact that aluminum is naturally present in soil, an application factor of 1 was applied to the CTV to derive the PNEC. The conservative PNEC for soil-dwelling organisms is therefore 0.135 mg dissolved monomeric Al/L.

No data were identified on concentrations of dissolved aluminum in soils that have received applications of alum sludge. However, as was noted in section 2.2.2.2, spreading on agricultural land is permitted in Canada only when the pH is greater than 6.0 or when liming and fertilization (if necessary) are done. Thus, the pH of receiving soils will likely be in the circumneutral range, where the solubility of aluminum is at a minimum. Based on results of equilibrium modelling, with the total dissolved aluminum concentrations being controlled by the precipitation of microcrystalline gibbsite, total dissolved aluminum concentrations would not exceed the PNEC unless soil pHs were less than about 5.1 (Bélanger et al. 1999). Because it is very unlikely that the pH of soils receiving alum sludge applications will be this low, it is very unlikely that the PNEC of 0.135 mg/L is exceeded in Canadian soils receiving such applications. In addition, while a shift in soil pH at the site of sludge application could mobilize the aluminum present in the sludge, the events causing such a shift (for example, storm events) and the resulting impacts are likely to be local and transitory in nature.

The expectation that the solubility and hence bioavailability of aluminum in sludges applied to agricultural soils will be extremely limited is supported by data on aluminum levels in plants growing on such soils. For example, aluminum in yellow mustard seed (Sinapsis alba) and Durum wheat seed (Triticum turgidum var. durum) collected from plants grown in soil amended with alum sludge from Regina's DWTP were found to be not statistically different from those of seeds collected in control plots (Bergman and Boots 1997).

Finally, although it has been noted that aluminum in the sludge can fix labile phosphorus by forming stable aluminum-phosphorus complexes and hence make it unavailable to plants, causing deficiencies (Jonasson 1996; Cox et al. 1997), this is unlikely to occur when soil receiving sludge is also fertilized as required in Canada.

3.1.2 Other lines of evidence relating to aluminum salts

Trends in production and use

An apparent increase in production and use of aluminum salts occurred over the period 1995 to 2000; however, from 2000 to 2006, user demand remained relatively constant and the total amount of aluminum contained in the salts (that is, aluminum chloride, aluminum nitrate, aluminum sulphate, polyaluminum chloride (PAC), polyaluminum silicate sulphate (PASS), aluminum chlorohydrate (ACH) and sodium aluminate), and therefore available for release to the Canadian environment, appeared stable at around 16,000 tonnes per year (Cheminfo Services Inc. 2008). Water treatment applications continued to be the primary consumer of sulphate and chloride salts in the years following publication of the original State of the Science report (Environment Canada and Health Canada 2000), with lesser quantities used in the pulp and paper sector.

Despite the proportionally higher demand for aluminum sulphate in comparison with the other aluminum salts (86% of the total demand in 2006), aluminum producers reported declining use of alum (and sodium aluminate) over the period 2000 to 2006, with increased use of other aluminum-based products, such as PAC, ACH and PASS, as well as non-aluminum products such as iron salts. PAC and iron chlorides were increasingly used as substitute coagulants/flocculants for alum in drinking water treatment, the former substance for its superior settling properties in colder water temperatures and the latter due to awareness of residual aluminum issues and superior performance in floc settling and dewatering of sludge (Cheminfo Services Inc. 1008). PAC is also particularly effective at water treatment facilities experiencing large fluctuations in water temperature, turbidity, pH and alkalinity. ACH, which is a highly concentrated and highly charged type of PAC, is sometimes used preferentially over alum because of its better buffering capacity, and PASS is very effective at removing phosphorus in cold waters with lower dosing rates and less sensitivity to variable conditions of alkalinity, pH, temperature and suspended solids (Cheminfo Services Inc. 2008). Physical process changes, such as conversion from acid to alkaline paper-making, have also contributed to reduced demand for alum.

Trends in sources and releases to the environment

No evidence of significant new sources of aluminum derived from the three salts that are the subject of this assessment has been identified.

Data provided in the study by Cheminfo Services Inc. (2008) indicated that while a slight decrease in Canadian consumption of aluminum salts occurred over the period from 2000 to 2006, the total amount of aluminum contained in these salts remained virtually unchanged, and this suggests that overall concentrations and total entry of aluminum into the environment have remained relatively constant.

Information collected since the publication of Environment Canada and Health Canada (2000) indicates that primary exposure routes for aluminum derived from the three salts have also remained unchanged. For drinking water treatment, releases are primarily to surface waters, with lesser proportions of aluminum released to sewer for subsequent wastewater treatment or present in sludge that is directed to landfill. While low levels of aluminum have been measured in final effluents leaving municipal wastewater treatment plants, the majority of the metal appears to remain within sludge which is then transferred to landfill or processed for landfarming. Releases related to industrial applications have decreased in recent years, largely due to lower aluminum use in the pulp and paper sector and therefore lower quantities entering receiving waters from industrial treatment plants and reduced quantities sent to landfill in paper products (Cheminfo Services Inc. 2008).

3.1.3 Sources of uncertainty

There are a number of uncertainties in this risk characterization. Regarding effects of aluminum on pelagic organisms, there are only a few acceptable studies conducted at circumneutral pH (6.5-8.0), conditions similar to those of aquatic environments receiving releases from DWTPs. There are also uncertainties associated with the decision to use an application factor of 1 to derive a PNEC for pelagic organisms, a choice that was made considering concentrations of aluminum in uncontaminated waters and the biological response of organisms to a narrow concentration range, resulting in an abrupt "threshold" where biological response occurs.

There are uncertainties associated with levels of aluminum released by DWTPs and with the levels and form of aluminum present in the aquatic environment. The use of the MINEQL+ and WHAM models provided aluminum results higher than those measured in the receiving environments when calculations were done assuming that aluminum solubility is controlled by microcrystalline gibbsite. When calculations were done with the boehmite form of aluminum hydroxide, levels were much lower than what was calculated with the microcrystalline gibbsite form (Fortin and Campbell 1999). Direct measurement and determination of aluminum speciation in final effluents from water treatment plants would confirm the estimated levels and forms provided by MINEQL+ and WHAM models.

Other uncertainties exist relating to the impact of aluminum sludge releases on benthic organisms. There are some indications that sludge releases, whatever the coagulant or flocculant used, may have a smothering effect on benthos. In recognition of the potential for adverse ecosystem effects, many Provinces have implemented strategies designed to reduce or eliminate the release of water treatment plant effluents and sludges to receiving water bodies (see Section 2.2.2). It is expected that addressing issues relating to overall effluent and sludge concerns, most notably the extremely high levels of total suspended solids (TSS) should also effectively deal with physical and chemical aspects of aluminum sludge toxicity in the aquatic receiving environment.

In relation to terrestrial organisms, there are uncertainties associated with the limited data available for effects on soil-dwelling organisms other than plants. The lack of information on aluminum levels in pore waters of soils receiving applications of alum sludge is not considered critical, since these levels are constrained by theoretical limits on solubility that are below the PNEC for sensitive vegetation.