ARCHIVED - PSL Assessment Report for Aluminum Salts

Priority Substances List
Assessment Report
Follow-Up to the State of Science Report, 2000

Aluminum Chloride
Aluminum Nitrate
Aluminum Sulphate

Chemical Abstracts Service Registry Numbers
7446-70-0
13473-90-0
10043-01-3

Environnement Canada
Health Canada

November 2008

View Final decision after screening assessment (January 2010)

Figure 2.1 Solubility of aluminum species (and total aluminum, Alt) in relation to pH in a system in equilibrium with microcrystalline gibbsite
Figure 2.2 Mean aluminum concentrations in PM10 in outdoor air from provinces and territories across Canada (µg/m³) (1996-2006)
Figure 2.3 Mean total aluminum concentrations in aluminum-treated drinking water from provinces and territories across Canada (µg/L) (1990-2007)
Figure 2.4 Comparison of mean total aluminum concentrations in soils from provinces across Canada (mg/kg) (1987-2007)
Figure 3.1 Compilation of the LOEL values from the two major subsets of studies (Adult exposure > 90 days and Reproductive/developmental) considered in the exposure-response analysis.

Table 2.1 Physicochemical properties of aluminum chloride, aluminum nitrate and aluminum sulphate ¹
Table 2.2 Estimated production, import, export and consumption of aluminum in the form of aluminum salts in Canada for 2006
Table 2.3 Estimated total releases in Canada of aluminum from aluminum salts1 for 2006, by application
Table 2.4 Estimated total releases of aluminum, by salt, for 2006
Table 2.5 Mean total aluminum concentrations in various food groups based on the fifth Canadian Total Diet Study (2000-2002)
Table 2.6 Range of total aluminum concentrations in various cosmetic products sold in Canada
Table 2.7 Ranges of estimated aluminum bioavailability for various routes of exposure in humans and/or animals
Table 3.1 Estimated mean daily intake of total aluminum based on Canadian data
Table 3.2 Contribution (%) of each source of exposure based on Canadian mean daily intake of total aluminum
Table 3.3 Average estimated daily intake (EDI) and margin of exposure (MOE) for different age classes in the Canadian population

Aβ

amyloid beta

ACH

aluminum chlorohydrate

AD

Alzheimer's disease

AD_AF

chemical-specific animal to human toxicodynamic adjustment factor

ADP

adenosine diphosphate

ADRDA

Alzheimer's Disease and Related Disorders Association

AK_AF

chemical-specific animal to human toxicokinetic adjustment factor

AMS

accelerator mass spectrometry

ApoE

apolipoprotein E

ASFA

Aquatic Sciences and Fisheries Abstracts (World Health Organization Food and Agricultural Organization)

ATP

adenosine triphosphate

ATPase

class of enzymes that catalyze the decomposition of ATP into ADP

ATSDR

Agency for Toxic Substances and Disease Registry (U.S. Department of Health and Human Services)

BAF

bioaccumulation factor

BCF

bioconcentration factor

BIOSIS

Biosciences Information Services

CAplus

Chemical Abstracts Plus

CAB

Commonwealth Agricultural Bureaux

cAMP

cyclic adenosine monophosphate

CAS

Chemical Abstracts Service

CASReact

CAS Reaction

CBTB

Canadian Brain Tissue Bank

CEPA

Canadian Environmental Protection Act

CEPA 1999

Canadian Environmental Protection Act, 1999

CESARS

Chemical Evaluation Search and Retrieval System (Ontario Ministry of the Environment and Michigan Department of Natural Resources)

cGMP

cyclic guanosine monophosphate

ChemCats

Chemical Catalogs online

Chemlist

regulated Chemicals Listing

CHRIS

Chemical Hazard Release Information System

CI

confidence interval

CSHA

Canadian Study of Health and Aging

CT

computer tomography

CTV

Critical Toxicity Value

D_a

administered dose

D_b

base diet dose

D_c

cumulative dose

DIN

Drug Identification Number

DNA

deoxyribonucleic acid

DOC

dissolved organic carbon

DOM

dissolved organic material

DSM

Diagnostic and Statistical Manual of Mental Disorders (American Psychiatric Association)

dw

dry weight

DWTP

drinking water treatment plant

EC₅₀

median effective concentration

ECETOC

European Centre for Ecotoxicology and Toxicology of Chemicals

EDI

estimated daily intake

EEM

Environmental Effects Monitoring

ELIAS

Environmental Library Integrated Automated System (Environment Canada library)

EPA

Environmental Protection Agency (U.S.)

ETAAS

electrothermal atomic absorption spectroscopy

FAO

Food and Agricultural Organization (United Nations)

GD

gestational day

GEOREF

Geo Reference Information System (American Geological Institute)

GLP

good laboratory practice

GVRD

Greater Vancouver Regional District

GVS&DD

Greater Vancouver Sewerage & Drainage District

HDAF

chemical-specific human variability toxicodynamic adjustment factor

HKAF

chemical-specific human variability toxicokinetic adjustment factor

HSDB

Hazardous Substances Data Bank (U.S. National Library of Medicine)

ICD

International Classification of Diseases (World Health Organization)

IM

intramuscular

IPCS

International Programme on Chemical Safety (World Health Organization)

IV

intravenous

JECFA

Joint FAO/WHO Expert Committee on Food Additives

KASAL

alkaline aluminum phosphate and dibasic sodium phosphate

LC₅₀

median lethal concentration

LD₅₀

median lethal dose

LOAEC

lowest observed adverse effect concentration

LOEC

lowest observed effect concentration

LOEL

lowest observed effect level

LSA

Ontario Longitudinal study of Aging

MEDLINE

Medical Literature Analysis and Retrieval System Online (U.S. National Library of Medicine)

MINEQL+

chemical equilibrium modeling software

MMAD

mass median aerodynamic diameter

MMSE

Mini-Mental State Examination

MOE

margin of exposure

mRNA

messenger ribonucleic acid

MS

multiple sclerosis

MWWTP

municipal wastewater treatment plant

NADPH

nicotinamide adenine dinucleotide phosphate

NATES

National Analysis of Trends in Emergencies System

NEMISIS

National Enforcement Management Information System and Intelligence System

NFT

neurofibrillary tangles

NICNAS

National Industrial Chemicals Notification and Assessment Scheme (Australian Government Department of Health and Aging)

NIH

National Institutes of Health (U.S. Department of Health and Human Services)

NINCDS

National Institute of Neurological and Communicative Disorders and Stroke

NOEC

no observed effect concentration

NOEL

no observed effect level

NTIS

National Technical Information Service (U.S. Department of Commerce)

OECD

Organisation for Economic Co-Operation and Development

OR

odds ratio

p value

the probability of obtaining a value of the test statistic at least as extreme as the one that was actually observed, given that the null hypothesis is true

PAC

polyaluminum chloride

PAQUID

Principle lifetime occupation and cognitive impairment in a French elderly

PAS

polyaluminum sulphate

PASS

polyaluminum silicate sulphate

PEC

Predicted Environmental Concentration

PM

particulate matter

PM_2.5

particulate matter less than 2.5 micrometers in aerodynamic diameter

PM₁₀

particulate matter less than 10 micrometers in aerodynamic diameter

PND

postnatal day

PNEC

Predicted No-Effect Concentration

POLTOX

Cambridge Scientific Abstracts (U.S. National Library of Medicine)

PPP2

Protein Phosphatase 2

PSL

Priority Substances List

PSL2

Second Priority Substances List

PTEAM

Particle Total Exposure Assessment Methodology

PubMed

free Internet access to MEDLINE

RMOC

Regional Municipality of Ottawa-Carleton

RR

relative risk

RTECS

Registry of Toxic Effects of Chemical Substances (U.S. National Institute for Occupational Safety and Health)

SALP

sodium aluminum phosphate

SOS

State of the Science

t_½

half-life

TDI

total dietary intake

Tf

transferrin

TNF_-α

alpha tumour necrosis factor

TOXLINE

toxicology database (U.S. National Library of Medicine)

TRI93

Toxic Chemical Release Inventory (U.S. Environmental Protection Agency, Office of Toxic Substances)

TSS

total suspended solids

USEPA-ASTER

Assessment Tools for the Evaluation of Risk (U.S. Environmental Protection Agency)

Vd

volumes of distribution

WASTEINFO

Waste Management Information (Bureau of the American Energy Agency)

WHAM

Windermere Humic-Aqueous Model software designed to calculate equilibrium chemical speciation in surface and ground waters, sediments and soils

WHO

World Health Organization

The three aluminum salts, aluminum chloride, aluminum nitrate and aluminum sulphate, were included on the Second Priority Substances List (PSL2) under the Canadian Environmental Protection Act, 1999 (CEPA 1999) in order to assess the potential environmental and human health risks posed by exposure to aluminum derived from these three salts in Canada.

In December 2000, the PSL2 assessment of the three aluminum salts was formally suspended due to limitations in the available data for assessing health effects. At the same time, a State of the Science report (Environment Canada and Health Canada 2000) on the three aluminum salts was released, providing an in-depth review of toxicity and exposure information relating to human health and the environment. During the suspension period, additional health effects information was published in the scientific literature.  As well, a long-term study on the neurodevelopmental effects of aluminum, commissioned by an industry consortium, was initiated. The results of this study, however, will not be available until after the mandated publication deadline for this PSL assessment.

In Canada, municipal water treatment facilities are the major users of aluminum chloride and aluminum sulphate, accounting for 78% of the estimated 16.1 kilotonnes of the 2006 domestic consumption. Industrial water and wastewater treatment, and use in the pulp and paper industry, account for an additional 20 %. Aluminum sulphate and aluminum chloride are also used as ingredients in drugs and cosmetics, such as antiperspirants and topical creams.  Aluminum sulphate is permitted as a food additive in a limited number of products. Aluminum nitrate, used in far less quantities than the sulphate and chloride salts, may be used in fertilizers, and as a chemical reagent in various industries.  

Aluminum salts occur naturally in small quantities in restricted geological environments and aluminum can be released into the Canadian environment from these natural sources. However, since aluminum is present in relatively large amounts in most rocks, dominantly in aluminosilicate minerals, which weather and slowly release aluminum to the surface environment, the small amounts of aluminum in surface waters resulting from weathering of aluminum salts such as aluminum sulphate cannot be distinguished from other natural aluminum releases.

During their use in water treatment, aluminum salts react rapidly, producing dissolved and solid forms of aluminum with some release of these to Canadian surface waters. The amount of anthropogenic aluminum released nationally in Canada is small compared with estimated natural aluminum releases; however anthropogenic releases can dominate locally near strong point sources. Most direct release into surface waters of aluminum derived from the use of aluminum salts in water treatment processes originates from drinking water treatment plants (DWTPs). However, direct releases of process waters from DWTPs are regulated by many provincial and territorial authorities, and these releases typically occur in circumneutral water, where the solubility of aluminum is minimal. Disposal of sludge produced by municipal and industrial water treatment facilities on land through landfarming practices is a source of aluminum to the terrestrial environment. However, the presence of dissolved organic matter and inorganic chelating agents will lower the amount of bioavailable aluminum in both the terrestrial and aquatic environments.

While extensive recent data on total aluminum concentrations in Canadian surface waters are available, few data exist on levels in areas close to sites where releases occur. The situation for sediment and soil is similar, in that data exist for the Canadian environment in general, but not for areas where releases occur. A large number of environmental toxicity data are available for acidified environments, but relatively few exist for circumneutral environments similar to those where most releases occur.

Based on a comparison of highest measured and estimated aluminum levels present in both aquatic and terrestrial environments in Canada that receive direct inputs of aluminum from the use of the three aluminum salts, and Predicted No-Effect Concentrations (PNECs) derived from experimental data for aquatic and terrestrial biota, it is considered that, in general, it is unlikely that organisms are exposed to harmful levels of aluminum resulting from the use of aluminum salts in Canada. However, it is acknowledged that under some release conditions there is potential for local impacts to benthic organisms related to the settling of aluminum sludge from DWTPs onto the sediment surface. As such, it is proposed that the three aluminum salts (i.e., aluminum chloride, aluminum nitrate, aluminum sulphate) are not entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity.

With respect to the human health risk assessment, both epidemiological and experimental animal data were reviewed.  While positive significant associations between aluminum in drinking water and the development of Alzheimer's disease (AD) or other forms of dementia have been observed in some studies, a weight-of evidence analysis of all available data does not provide sufficient scientific support at this time for a causal relationship between aluminum in drinking water and AD.  Considering experimental animal studies, the dose at which neurotoxic, reproductive, and developmental effects have been repeatedly observed was used to establish an exposure level of concern.

The human health risk assessment, carried out on the basis of total aluminum exposure, suggests thata proportion of individuals, particularly in the younger age groups, may be exposed to levels of total aluminum that, when compared to the exposure level of concern (derived from studies in experimental animals measuring neurotoxic effects), results in margins that may not be adequate.  For most age groups, food is estimated to be the primary source of exposure to aluminum.

Aluminum chloride, aluminum nitrate, and aluminum sulphate are not significant contributors to aluminum in the diet. The use of aluminum sulphate and aluminum chloride for the treatment of drinking water contributes less than 1 % of the total aluminum intake in children, the most sensitive age group.  Thus, the three salts are not considered to be significant contributors to the overall health risk.

Food additives in Canada are regulated under the Food and Drugs Act. The Food Directorate of Health Canada is currently re-evaluating dietary exposure to aluminum and reviewing current uses of aluminum-containing food additives.

Based on the information available for human health and the environment, it is proposed that the three aluminum salts, aluminum chloride, aluminum nitrate, aluminum sulphate, are not entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity or that constitute or may constitute a danger to the environment on which life depends. It is also proposed that aluminum chloride, aluminum nitrate and aluminum sulphate, are not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health. It is therefore proposed that aluminum chloride, aluminum nitrate and aluminum sulphate do not meet the definition of "toxic" under section 64 of the Canadian Environmental Protection Act, 1999.

The Canadian Environmental Protection Act, 1999 (CEPA 1999) requires the Ministers of the Environment and of Health to prepare and publish a Priority Substances List (PSL) that identifies substances (including chemicals, groups of chemicals, effluents and wastes) that may be harmful to the environment or constitute a danger to human health. The Act also requires both Ministers to assess these substances to determine whether they meet or are capable of meeting the criteria as defined in section 64 of the Act. A substance meets the criteria under CEPA 1999 if it is entering or may enter the environment in a quantity or concentration or under conditions that:

1. have or may have an immediate or long term harmful effect on the environment or its biological diversity;
2. constitute or may constitute a danger to the environment on which life depends; or
3. constitute or may constitute a danger in Canada to human life or health.

For substances deemed to meet the criteria defined in section 64, risk management measures  are identified and implemented in consultation with stakeholders, in order to reduce or eliminate the risks posed to human health or the environment. These measures may include regulations, guidelines, pollution prevention plans or codes of practice to control any aspect of the life cycle of the substance, from the research and development stage through to manufacture, use, storage, transport and ultimate disposal.

Based on initial screening of readily accessible information, the rationale provided by the Ministers' Expert Advisory Panel in 1995 for including aluminum chloride, aluminum nitrate and aluminum sulphate on the Second Priority Substances List was as follows (Environment Canada and Health Canada 2000):

"Aluminum, from both natural and man-made sources, is widespread in the Canadian environment. Intakes of aluminum among the human population and ambient airborne concentrations in some parts of the country are close to those that have induced developmental and pulmonary effects in animal studies. Epidemiological studies have indicated that there may be a link between exposure to aluminum in the environment and effects in humans. Aluminum compounds are bioaccumulative, and can cause adverse ecological effects, especially in acidic environments. The Panel identifies three aluminum compounds as being of particular concern. An assessment is needed to establish the weight of evidence for the various effects, the extent of exposure and the aluminum compounds involved. If necessary, the assessment could be expanded to include other aluminum compounds."

A preliminary report was completed for the three aluminum salts and released as a State of the Science (SOS) report in December 2000. With respect to immediate or long term harmful effects of the three aluminum salts on the environment or its biological diversity, the report proposed that, based on measured and estimated aluminum levels in Canadian aquatic and terrestrial environments receiving direct inputs of aluminum from the use of aluminum salts and on the Predicted No-Effect Concentrations (PNECs)derived from experimental data for aquatic and terrestrial biota, it is in general unlikely that organisms are exposed to harmful levels of aluminum resulting from the use of aluminum salts in Canada.

With respect to human health, a conclusion regarding section 64(c) could not be reached in 2000, owing to the limitations in the available data for assessing health effects. Therefore, the assessment of aluminum salts was suspended in December 2000 for a period of six years to allow for the development of additional human health effects data in order that Health Canada could reach a conclusion on whether aluminum salts (chloride, nitrate and sulphate) and possibly other aluminum compounds should be considered as "toxic" under CEPA 1999.

In terms of this draft PSL2 assessment, the conclusions made under section 64 of CEPA 1999 relate directly to the three aluminum salts nominated by the Ministers' Expert Advisory Panel (chloride, nitrate, and sulphate). However, different approaches are taken by Environment Canada and Health Canada in evaluating the potential for risk.

In characterizing the potential for risk to the environment, data relevant to the entry of the three listed salts into the Canadian environment from local point sources (e.g., drinking water treatment plants) were examined in conjunction with data on environmental fate and exposure. The focus was on assessing potential for effects on the environment near point sources. This evaluation formed the basis for determining whether the three aluminum salts identified by the Ministers' Expert Advisory Panel (chloride, nitrate and sulphate) are "toxic" under section 64 of CEPA 1999. 

The human health risk characterization consists of a two-stage evaluation. In the first stage, exposure of the general Canadian population to total aluminum in air, drinking water, diet, and soil is evaluated. Although the relative bioavailabiilties of aluminum in these different media are also assessed, no distinction is made in the exposure assessment at this first stage with regard to the specific salts contributing to total aluminum exposure. This stage allows for a comparison of the total aluminum intake with the estimated exposure level of concern, and therefore provides a characterization of human health risk in relation to total aluminum exposure from the four media. In the second stage, the relative contribution of each of the three listed aluminum salts (chloride, nitrate, and sulphate) to this total aluminum exposure is evaluated, and a decision with respect to section 64(c) of CEPA is made for the three salts.

Health Canada chose this two-stage approach on the basis of both scientific and practical considerations. First, overall exposure to the aluminum moiety (Al³⁺), and not exposure to a particular aluminum compound, is the critical parameter for evaluating potential toxicological risk¹. Second, concentrations of aluminum in foods, soil, drinking water, and air are generally reported as total aluminum, and not in terms of specific salts.

The search strategies employed in the identification of relevant data are presented in Appendix A. All original studies that form the basis for determining whether aluminum salts are "toxic" under CEPA 1999 have been critically evaluated and are described in the assessment. For issues relevant to the environmental and human health effects of aluminum, but outside the scope of the present assessment, the information is summarized briefly and the reader is referred to recent critical reviews published in the scientific literature for a more detailed discussion.

The human health components of the present document were prepared by the Safe Environments Programme- Quebec Region, in collaboration with the Existing Substances Division of the Safe Environments Programme (National Capital Region).  The environmental components were prepared by the Existing Substances Division of the Science and Technology Branch.

The human health components of this assessment have been peer reviewed by the following external experts:

• Dr. Diane Benford, Food Standards Agency, United Kingdom
• Dr. Nicola Cherry, University of Alberta, Edmonton, Alberta
• Dr. Rajendra Chhabra, National Institute of Environmental Health Sciences, Research Triangle Park, North Carolina
• Dr. Herman Gibb, Sciences International, Arlington, Virginia
• Dr. Lesbia Smith, Environmental and Occupational Health Plus, Toronto, Ontario
• Dr. Robert Yokel, University of Kentucky, Lexington, Kentucky

Information relevant to environmental components of this assessment has been reviewed by the following external experts:

• Dr. Pierre-André Côté, Canadian Water and Wastewater Association, Quebec City, Quebec
• Mr. André Germain, Environment Canda, Monteal, Quebec.
• Mr. Robert Garrett, Geological Survey of Canada, Ottawa, Ontario
• Dr. William Hendershot, McGill University, Montreal, Quebec
• Mr. Christopher Lind, General Chemical Corporation, Newark, New Jersey
• Mr. Robert Roy, Fisheries and Oceans Canada, Mont-Joli, Quebec
• Mr. James Brown, Reynolds Metals Company, Richmond, Virginia
• Mr. Scott Brown, National Water Research Institute, Burlington, Ontario
• Mr. Christopher Cronan, University of Maine, Orono, Maine
• Dr. Lawrence Curtis, Oregon State University, Corvallis, Oregon
• Mr. Richard Lapointe, Société d'électrolyse et de chimie Alcan Ltée, Montreal, Quebec
• Dr. Stéphanie McFadyen, Canadian Water and Wastewater Association, Ottawa, Ontario
• Dr. Wayne Wagner, Natural Resources Canada, Ottawa, Ontario

________________________________________

¹ Note, however, that different aluminum salts are absorbed into the bloodstream to different degrees (Yokel et al. 2006) and this aspect is considered in this assessment within section 2.3.3.1.

• 2.1 Identity and physical/chemical properties
• 2.2 Entry characterization
• 2.2.1 Production, import, export and use
• 2.2.1.1 Aluminum chloride
• 2.2.1.2 Aluminum nitrate
• 2.2.1.3 Aluminum sulphate
• 2.2.2 Sources and releases
• 2.2.2.1 Natural Sources
• 2.2.2.2 Anthropogenic sources
• 2.3 Exposure Characterization
• 2.3.1 Environmental Fate
• 2.3.1.1 Air
• 2.3.1.2 Water
• 2.3.1.3 Sediment
• 2.3.1.4 Soil
• 2.3.1.5 Biota
• 2.3.2 Environmental concentrations
• 2.3.2.1 Air
• 2.3.2.1.1 Ambient air
• 2.3.2.1.2 Indoor air
• 2.3.2.2 Water
• 2.3.2.2.1 Surface water
• 2.3.2.2.2 Drinking water
• 2.3.2.3 Sediment
• 2.3.2.4 Soil
• 2.3.2.5 Biota
• 2.3.2.6 Food
• 2.3.2.7 Consumer products
• 2.3.2.7.1 Non-prescription drugs
• 2.3.2.7.2 Cosmetics
• 2.3.2.8 Vaccines
• 2.3.3 Toxicokinetics: human and experimental animals
• 2.3.3.1 Absorption
• 2.3.3.1.1 Oral absorption
• 2.3.3.1.2 Dermal absorption
• 2.3.3.1.3 Inhalation absorption
• 2.3.3.1.4 Parenteral administration
• 2.3.3.1.5 Summary of estimates of aluminum bioavailability
• 2.3.3.1.6 Integrating bioavailability in human health risk assessment
• 2.3.3.2 Distribution
• 2.3.3.3 Elimination
• 2.4 Effects characterization
• 2.4.1 Ecotoxicology
• 2.4.1.1 Aquatic organisms
• 2.4.1.1.1 Pelagic
• 2.4.1.1.2 Benthic
• 2.4.1.2 Terrestrial organisms
• 2.4.2 Experimental mammal studies
• 2.4.2.1 Acute toxicity
• 2.4.2.2 Short-term toxicity (duration of exposure less than 90 days)
• 2.4.2.3 Subchronic and chronic toxicity (exposure duration greater than 90 days, non-cancer endpoints)
• 2.4.2.4 Reproductive and developmental toxicity
• 2.4.2.5 Carcinogenicity
• 2.4.2.6 Genotoxicity
• 2.4.3 Human studies
• 2.4.3.1 Human case studies of exposure to aluminum
• 2.4.3.2 Epidemiological studies of aluminum exposure via drinking water
• 2.4.3.3 Epidemiological investigations of exposure to aluminum in antacids, antiperspirants or food
• 2.4.3.4 Epidemiological investigations of exposure to aluminum in vaccines
• 2.4.3.5 Epidemiological investigations of occupational exposure to aluminum
• 2.4.4 Mode of action of toxic effects of aluminum

Aluminum chloride is also known as aluminum trichloride, aluminum chloride (1:3) and trichloroaluminum (ATSDR 2006). It has the Chemical Abstracts Service (CAS) registry number 7446-70-0 and a chemical formula of AlCl₃. In its hydrated form, AlCl₃-6H₂O, it is called hexahydrated aluminum chloride (CAS No. 7784-13-6). Trade names include Aluwets, Anhydrol and Drichlor.

Synonyms for aluminum nitrate include aluminum trinitrate and aluminum (III) nitrate (1:3). The CAS registry number is 13473-90-0 and the chemical formula is Al(NO₃)₃. The nonahydrate aluminum nitrate, Al(NO₃)3-9H₂O (CAS No. 7784-27-2), is the stable form of this compound.

Aluminum sulphate can also be identified as alum, alumsulphate (2:3), aluminum trisulphate, dialuminum sulphate and dialuminum trisulphate. The CAS registry number for aluminum sulphate is 10043-01-3 and the chemical formula is Al₂(SO₄)₃. Alum is often represented as Al₂(SO₄)₃-14H₂O. It may be found in different hydrated forms. The commercial product, called cake alum or patent alum, is an octadecahydrate aluminum sulphate, Al₂(SO₄)3-18H₂O.

In addition to these three compounds, aluminum polymers such as polyaluminum sulphate (PAS) and polyaluminum chloride (PAC) are used in water treatment. The general formula for PAS is Al_a(OH)_b(SO₄)_c, where b + 2c = 3a; for PAC, the general formula is Al_a(OH)_bCl_c, where b/a is usually about 2.5 (e.g., Al₂(OH)₅Cl). Mixed aluminum polymers may also be used; their general formula is Al_a(OH)_bCl_c(SO₄)_d, and b/a varies between 0.4 and 0.6.

Physicochemical properties of the three aluminum salts are presented in Table 2.1..

2.2.1 Production, import, export and use

Aluminum sulphate and aluminum chloride are produced in Canada, while aluminum nitrate is imported. Information on sources and emissions of aluminum salts or aluminum resulting from the use of aluminum salts was initially obtained through an industry survey carried out under the authority of section 16 of CEPA (CEPA 1988; Environment Canada 1997). Information regarding the use of aluminum chloride and aluminum sulphate in water treatment plants was obtained on a voluntary basis from Canadian municipalities with the help of provincial and territorial authorities. In 2007, additional research was conducted in order to review use patterns and quantities of aluminum derived from sources identified in the original assessment, as well as to identify and quantify potential new sources of aluminum to the environment resulting from the application of aluminum salts in Canada (Cheminfo Services Inc. 2008).

Table 2.2 provides estimated production, import, export and consumption values for the year 2006, based largely on input from Canadian aluminum salt producers. Unless otherwise stated, quantities reported in Table 2.2 and the accompanying text represent the amount of elemental aluminum present in the respective salts rather than the total amount of the salt. Polymeric forms of the chloride and sulphate are detailed separately, as these salts were found to be commonly used individually or in combination with other salts in water treatment processes. No producers or users of aluminum nitrate were identified for 2006 and, therefore, while it is likely that very small quantities were being imported into Canada in that year for a variety of low volume applications, no numerical data were available. Total Canadian consumption of aluminum as aluminum salts in 2006 was estimated at 16.1 kilotonnes, with aluminum sulphate accounting for approximately 80% of this demand, and PAC for the majority of the remainder (Cheminfo Services Inc. 2008). Approximately 80% of the total aluminum demand was for the treatment of drinking water and wastewater at municipalities. Industrial fresh water and wastewater treatment facilities accounted for the majority of the remaining demand in Canada.

Five companies produced most of the aluminum salts used in Canada in 2006 (Cheminfo Services Inc. 2008). Imports and exports were roughly in balance, with imports representing approximately 10% of 2006 domestic consumption and exports representing approximately 14% of 2006 production. Alum, PAC and aluminum chlorohydrate (ACH) were the major imported aluminum salts, while PAC and alum were exported.

Total Canadian demand for aluminum salts remained relatively constant between 2000 and 2006 (Cheminfo Services Inc. 2008). Canada's salt producers indicate that the demand for alum and sodium aluminate declined during this period, while PAC, ACH and polyaluminum silicate sulphate (PASS) increased in use. While overall aluminum salts demand for municipal water treatment has increased slightly, use in the pulp and paper industry has dropped. The overall total amount of aluminum contained in the salts used in Canada has remained constant at close to 16 kilotonnes per year (Cheminfo Services Inc. 2008).

2.2.1.1 Aluminum chloride

Aluminum chloride is used in either anhydrous or hydrated form. In the anhydrous form, it is used as a catalyst, in Friedel-Crafts reactions, in the manufacture of rubber, the cracking of petroleum, and the manufacture of lubricants. In its hydrated form, it is used by the pharmaceutical industry as an active ingredient in deodorants and antiperspirants, as well as in wood preservation, and in the manufacture of adhesives, paint pigments, resins, fertilizers and astringents (Germain et al. 2000; Pichard 2005; Merck 2006). Polymeric forms, primarily polyaluminum chloride (PAC) and the more concentrated and highly charged aluminum chlorohydrate (ACH), are used as coagulants and flocculants in water treatment.

PAC has the highest Canadian production and use volumes of the three aluminum chloride salts. PAC demand increased over the period 2000 to 2006, with greatest quantities being used in the treatment of drinking water (Cheminfo Services Inc. 2008). Similar increased demand was evident in other applications, including industrial freshwater treatment, municipal and industrial wastewater treatment, and as a pulp and paper additive (Cheminfo Services Inc. 2008). Production and demand were substantially lower for both aluminum chloride and ACH. Canadian consumption of aluminum chloride remained stable from 2000 to 2006, while ACH demand increased substantially (Cheminfo Services Inc. 2008). Most of the increased demand was associated with increased applications in industrial wastewater treatment, with slower rates of growth in other applications.

2.2.1.2 Aluminum nitrate

Aluminum nitrate is used as a chemical reagent (catalyst), in the leather tanning industry, as an antiperspirant, as a corrosion inhibitor, and in the manufacture of abrasives, refractories, ceramics, electric insulation, catalysts, paper, candles, pots, artificial precious stones and heat-resistant fibres (Budaveri et al. 1989; Pichard 2005). It is also used as an adsorbent in chromatography for the production of filter membranes, in radiation protection dosimetry in the uranium extraction sector, and as a nitrating agent in the food industry (Merck 2006).

There are no known producers of aluminum nitrate in Canada, and only one user was identified in a survey done in 1997 by Environment Canada (1997). This user reported that less than 400 kg of aluminum nitrate was included in fertilizers for export to the United States. It is likely that very small quantities of aluminum nitrate are being imported into Canada for a variety of low volume applications, including laboratory uses, leather manufacturing, manufacturing of fire works, and other minor applications (Cheminfo Services Inc. 2008).

2.2.1.3 Aluminum sulphate

In Canada, aluminum sulphate is used primarily as a coagulant and flocculant in water and wastewater treatment. There are other applications, however, in the leather industry, the paper industry, as a mordant in dyeing, in the fireproofing and waterproofing of textiles, in resin manufacture, and in the preparation of fertilizers and paint pigments (Germain et al. 2000; Pichard 2005; Merck 2006). The Canadian Fertilizers Product Forum advises that aluminum sulphate (alum) is used as a soil pH adjuster in the Lawn and Garden industry (2008 email from The Canadian Fertilizers Product Forum to J. Pasternak, Environment Canada; unreferenced). Aluminum sulphate can also be used to waterproof concrete, decolorize petroleum products, and as a formulant in antiperspirants and pesticides (Budaveri et al. 1989). Aluminum sulphate or alum is used in the treatment of eutrophic or mesotrophic lakes, to reduce the amount of nutrients present in the water. Both alum (Al₂(SO₄)₃) and sodium aluminate (Na₂Al₂O₄) are highly effective coagulants and flocculants that adsorb and precipitate soluble phosphorus and other compounds such as organic matter, forming clumps that settle to the bottom of the lake. In saturated solutions, aluminum sulphate is considered a mild corrosive and can be applied to ulcers in concentrations of 5% to 10% to prevent mucous secretion (Pichard 2005). The substance is also used as a food additive and some foods, such as baking powder.

It is estimated that approximately 276 kilotonnes of aluminum sulphate (11.9 kilotonnes on an aluminum basis) were produced in Canada in 2006, 15 kilotonnes (0.6 kilotonnes of aluminum) were imported and 12 kilotonnes (0.5 kilotonnes of aluminum) exported (Table 2.2). Municipal drinking water and wastewater treatment plants were the main users, comprising almost 84% of the total demand for that year. Industrial water treatment facilities and the pulp and paper sector accounted for most of the remaining consumption (15.8%).

2.2.2 Sources and releases

Aluminum sulphate minerals such as aluminite and alunite occur naturally in Canada in certain restricted geological environments. Aluminum chloride and aluminum nitrate do not occur naturally in the environment. Aluminum can be released from natural aluminum sulphate minerals; however, since aluminum is a common constituent of rocks, where it occurs dominantly in aluminosilicate minerals (e.g.,kaolinite, boehmite, clay, gibbsite, feldspar, etc.), which weather and slowly release aluminum to the surface environment. Aluminum present in surface waters due to man-made applications cannot be distinguished from natural aluminum released during weathering of aluminum-bearing minerals.

While aluminum chloride, aluminum nitrate and aluminum sulphate have many commercial applications in Canada, releases of aluminum to the environment from most commercial applications are expected to be small. However there is potential for release of relatively large amounts of aluminum resulting from the use of aluminum chloride and aluminum sulphate in water treatment plants (industrial water, drinking water or wastewater). In this application, aluminum will react rapidly, producing sludge, usually in the form of aluminum hydroxide (Al(OH)₃). Most sludge produced by municipal wastewater treatment plants (MWWTPs) or industries is sent to landfills or spread on land, with the remainder being composted, held in permanent lagoons, or incinerated prior to landfilling (Germain et al. 2000). Most provinces control DWTP waste flows through their respective systems of permits and/or approvals.  Sludge purged from clarifiers or accumulated in sedimentation basins of drinking water treatment plants (DWTPs) cannot be released directly to the aquatic environment in many provinces. It may be sent to sewers, incinerated with wastewater sludge and landfilled, held in permanent lagoons, spread on land or landfilled. Likewise, backwash waters (used to clean filters) cannot be discharged directly into open water bodies in many provinces where these discharges are often subjected to requirements for pretreatment (e.g., diversion to sedimentation ponds) or diversion to MWWTPs.  While many provinces do not generally allow direct discharge to surface water of any DWTP effluents containing sludges or backwash waters (e.g., Alberta, Manitoba, Ontario and New Brunswick), some of their existing plants may continue to discharge effluents directly to surface waters. Communication with provincial agencies indicates that these provinces are generally requiring some type of environmental impact assessments of the subject discharges with consideration of alternatives to direct discharge.  Some existing large plants in these provinces have recently removed their DWTP direct discharges from surface water (e.g., Britannia DWTP and Lemieux Island DWTP in Ottawa, ON), or are developing plans for alternatives to direct discharge to surface waters (e.g., certain plants in Alberta).  In other provinces, direct discharge may be allowed through provincial approvals systems if it is shown that the discharge results in no adverse effects (defined based on varying criteria) on the receiving body of water (e.g., Saskatchewan, Nova Scotia and Newfoundland).  It should be noted that some provinces and territories either do not have any coagulant usage for drinking water treatment, or they only use very small amounts and have requirements for DWTP effluent treatment destined for surface water (e.g., Prince Edward Island, Yukon Territory, Northwest Territories and Nunavut Territory) (Environment Canada unpublished 2008a)

While most aluminum is released in particulate form, a certain proportion occurs as the dissolved metal and it is this form that is considered easily absorbed and therefore bioavailable to aquatic organisms. The following section therefore discusses aluminum releases in general, with additional emphasis given to dissolved forms. This approach was necessary because very few studies examine monomeric aluminum levels in the environment or in anthropogenic releases.

2.2.2.1 Natural Sources

Atmospheric deposition of aluminum on land or water is small compared with internal releases by weathering and erosion of rock, soil and sediment (Driscoll et al. 1994). Weathering and erosion of "alum"-containing rocks will release aluminum into soils and streams, in part as Al³⁺ and other dissolved cationic and anionic species, depending on pH and the availability of complexing ions (Garrett 1998). These releases will be small, however, in relation to releases from weathering and erosion of aluminosilicate minerals.

There are no reliable estimates of the quantities of aluminum released to the environment by natural processes on a global scale, most of which comes from natural aluminosilicate minerals. Quantification of total or dissolved aluminum releases in Canada and elsewhere is very difficult and can provide only a rough estimate. Using Garrels et al.'s (1975) proposed global stream flux of 2.05 g/m² per year, total aluminum releases (including particulate material) were estimated to be approximately 20.45 million tonnes per year for Canada. Studies of weathering flux in selected Canadian and U.S. catchments (e.g., Likens etal. 1977; Kirkwood and Nesbitt 1991) yield similar or somewhat lower estimates (2 to 20 million tonnes per year) when extrapolated to the whole of Canada.

2.2.2.2 Anthropogenic sources

Very limited information is available on historical releases of the three aluminum salts. Accidental releases are reported to Environment Canada's National Analysis of Trends in Emergencies System (NATES) database and, more recently, the National Enforcement Management Information System and Intelligence System (NEMISIS). Between 1974 and 1991, 24 events released 316.2 tonnes of aluminum sulphate, mainly to land, and approximately 80% of the spilled material was recovered. Four accidental releases of aluminum chloride occurred in 1986 and 1987, and the product was not recovered on two occasions, resulting in a total release of 18.18 tonnes (Environment Canada 1995). Six spills involving the three aluminum salts subject to this assessment were reported from 1992 to 2008, all for aluminum sulphate. Approximately 40,000 liters of aluminum sulphate were released during these events, to both land and surface water, with no identified recovery of the spilled material.  None of the reported incidents related to municipal or industrial effluent discharges (Environment Canada 2008b).

Municipal drinking water and wastewater treatment plants are the main users of aluminum sulphate, aluminum chloride and other aluminum-based polymeric products. Aluminum salts are used as coagulants and flocculants to cause fine materials that are suspended, soluble or both to agglomerate, for subsequent removal via sedimentation and filtration. As part of this agglomeration or coagulation process, most of the aluminum associated with the added aluminum salt hydrolyses to aluminum hydroxide, which precipitates and becomes part of the floc structure. As such, it makes up a part of the sludge generated by the treatment process. A small amount of the aluminum added may stay with the finished water in either colloidal particulate (Al(OH)₃) or soluble form (e.g., AlOH²⁺, Al(OH)₂⁺,  Al(OH)₃, Al(OH)₄^-), dictated by the conditions of the treatment process and in particular, the pH (see Figure 2.1  below and from Stumm and Morgan 1981) .

While no comprehensive inventory of releases of aluminum associated with commercial use of aluminum salts exists, order-of-magnitude estimates derived from information provided by Canadian producers and users confirm that most releases are associated with wastewater treatment processes (approximately 43% in 2006), with drinking water treatment plants accounting for the majority of the remainder (about 36%; Table 2.3; Cheminfo Services Inc. 2008). All other sources are relatively minor. Again, most quantities are reported in terms of the elemental aluminum present in the respective salts. Approximately three quarters of the releases are to land, including: landfill, application on farms, and permanent lagoons. It is estimated that 5% of the aluminum used at pulp and paper mills for paper sizing is released to water courses (rivers or lakes), while 95% is contained on the paper, which is assumed to receive eventual disposal to landfills and composting in a minor, but growing proportion (2008 email from Canadian Wastewater Association to J. Pasternak, Environment Canada; unreferenced).

Most of the aluminum releases are from the use of aluminum sulphate, which is the aluminum salt having the highest quantity of consumption in Canada (Table 2.4; Cheminfo Services Inc. 2008).

Approximately 2% of the total aluminum used by municipalities for drinking water treatment (6.8 kilotonnes; see Table 2.2) ends up in drinking water (Table 2.3; Cheminfo Services Inc. 2008). A survey of 102 Canadian water treatment facilities conducted in 2006 found that over 80% of drinking water treatment plants (DWTPs) that use aluminum salts as coagulants and flocculants measure the concentration of aluminum in the treated water. The survey considered data from municipal drinking water and wastewater treatment facilities across Canada, primarily from larger municipalities (population > 100,000), although a small sample of small-to-medium sized municipalities was included (population range 20,000-100,000; Cheminfo Services Inc. 2008). Outlet concentrations in drinking water at the surveyed DWTPs which used aluminum ranged from 0.005 to 0.2 mg/L, with an average value of 0.067 mg/L. For comparison, Health Canada's Guidelines for Canadian Drinking Water Quality are 0.1 mg/L for conventional treatment plants using aluminum-based coagulants and 0.2 mg/L for other treatment systems using aluminum-based coagulants (Health Canada 2007a).

Less than half of the aluminum used at drinking water plants is released to receiving waters - mostly as solid aluminum hydroxide sludge (Cheminfo Services Inc. 2008). Notable examples of this practice occur in water treatment plants in Toronto. Most of the remaining aluminum is contained in sludge that is sent to landfill. Some of the sludge from drinking water facilities (commonly called "filter backwash solids"), in dilute form, may also be sent to wastewater treatment facilities in the municipality. Results from the 2006 survey suggest that approximately 16% of the aluminum used at drinking water treatment facilities is contained in sludge sent to nearby wastewater treatment facilities. A very small portion (~2%) remains permanently stored in lagoons, which for assessment purposes has been assumed to be a land destination. The 2006 survey did not identify any sludge from drinking water treatment plants going to farms; however, it is possible that some disposal by this method may be occurring in Canada as a small proportion of DWTP sludge was identified for landfarming in the earlier survey conducted for 1995 and 1996 (Germain et al. 2000).

In a study done with sludge from Calgary and Edmonton, AEC (1987) found that less than 0.02% of aluminum bound with sludge (containing 78,187 mg Al/kg dw) was released in water (i.e., 0.20 to 0.32 mg/L). Srinivasan et al. (1998) studied the speciation of aluminum at six different stages of water treatment at Calgary's DWTP. Total aluminum concentrations ranged from 0.038 to 5.760 mg/L, and dissolved inorganic aluminum concentrations varied from 0.002 to 0.013 mg/L. George et al. (1991) measured monomeric aluminum concentrations of less than 0.06 mg/L in alum sludge from ten different DWTPs containing up to a total of 2,900 mg Al/L; Calgary's DWTP was one of the plants studied.

Calgary's DWTP reported the aluminum content in backwash water following the cleaning of its filters. Dissolved aluminum levels ranged from 0.07 to 0.44 mg/L, and total aluminum concentrations varied from 0.76 to 3.3 mg/L. The backwash waters from this DWTP were not released to the river but were treated and sold as fertilizer (Do 1999).

Most of the aluminum discharged from municipal wastewater treatment plants (MWWTPs) surveyed in the 2006 study is associated with sludge. Approximately two thirds of the aluminum in MWWTP sludge is applied to farmland, with most of the balance (around 30%) being sent to landfill. About 5% of total aluminum releases are to surface waters and a very small proportion (less than 1%) is stored permanently in lagoons (Table 2.3). In Quebec City, the sludge from the drinking water treatment plant is directed to MWWTP where the resulting sludge is dried and incinerated with residential waste (co-incineration). The mineral and non-combustible component of the sludge is then landfilled (2008 email from Canadian Wastewater Association to J. Pasternak, Environment Canada; unreferenced). In most cases, the sludge sent to landfills was first sent for anaerobic digestion (where methane gas is generated from the organic content and used for plant energy) and the remaining solids concentrated to remove excess water. Some provinces (e.g., Alberta, Ontario and Quebec) have guidelines for the disposal of sewage sludge on agricultural land; spreading on agricultural land is permitted only when the pH is greater than 6.0 or when liming and fertilization (if necessary) are done. Although not a common practice, a few of the municipalities participating in the 2006 survey provided measured concentrations for aluminum present in sludge solids from their plants. In general, these values were in the range of 10 to 60 mg per gram of solids (dry basis) (Cheminfo Services Inc. 2008).

Final effluent concentrations of aluminum were not always available for MWWTPs participating in the 2006 survey (Cheminfo Services Inc. 2008). Where data were available, reported concentrations ranged from 0.013 to 1.200 mg/L, with an average value (weighted by water volume treated) of 0.816 mg/L. The form of the aluminum measured was not specified. Many of the MWWTPs surveyed relied on substances other than aluminum to treat wastewater, such as iron salts (ferrous and ferric chloride) and/or polyacrylamides, while others did not use any chemicals in their water treatment process.

Only two respondents to the 2006 survey provided information on aluminum concentrations in receiving waters in the vicinity of their effluent outfalls. The typical background level of dissolved aluminum in Lake Ontario in the vicinity of Toronto was reported to be approximately 0.010 mg/L, while typical concentrations in the North Saskatchewan River near Edmonton were 0.020 to 0.040 mg/L (Cheminfo Services Inc. 2008). These data are insufficient to determine in a useful way the contribution of aluminum from aluminum salt consumption in receiving waters. In the original State of the Science (SOS) report (Environment Canada and Health Canada 2000), it was determined that while extensive data on total aluminum concentrations in Canadian surface water are available, few data exist in areas close to sites where releases occur. The situation for sediment and soil is similar, in that data exist for the Canadian environment in general, but not for areas where releases occur. The state of available relevant concentration data has not changed since 2000.

In addition, changes in policies and procedures relating to the direct release of treatment plant effluents into surface waters have occurred since the publication of the original SOS report. In 1993, a total aluminum concentration of 36 mg/L was measured just downstream of the Regional Municipality of Ottawa-Carleton's (RMOC) DWTP discharge pipe, while the concentration 200 m downstream of the plant was 0.5 mg/L (Germain et al. 2000). Similarly, in 1998, sediment concentrations in the Ottawa River were 125,160, 51,428 and 41,331 mg/kg dw at points closest to, 300 m, and 500 m downstream of the DWTP, respectively, and were significantly elevated compared with control and upstream values of  17,543 and 20,603 mg/kg dw, respectively. In 2008, all wastes from the plant were diverted to a nearby MWWTP, effectively eliminating the direct discharge of aluminum-bearing sludge into the river (Environment Canada 2008c). However, it will likely take some time before conditions in bottom sediment in the vicinity of the DWTP outfall return to those in line with non-impacted areas.

Germain et al. (2000) reported mean total aluminum levels in the effluent of some MWWTPs using aluminum salts. Concentrations varied from 0.03 to 0.84 mg/L, and the maximum value reported by one plant reached 1.8 mg/L. These figures are in the same order of magnitude as those reported by Orr et al. (1992) for 10 Ontario MWWTPs and by MEF and Environnement Canada (1998) for 15 Quebec MWWTPs, and agree well with those of Cheminfo Services Inc. (2008) reported above. Some plants do not use aluminum-based coagulants and flocculants but still reported aluminum levels in their effluents; their mean total aluminum levels ranged from 0.003 to 0.90 mg/L (Germain et al. 2000). Many wastewater treatment plants, such as those in Quebec, receive influents from combined sewers which collect both wastewater and stormwater. In these cases, part of the solids content of the influent will come from urban drainage that could contain aluminum-bearing solids from erosion processes and other sources. The content of wastewater treatment plant influents is determined by the nature and proportions of their primary inputs (i.e., residential, commercial, institutional, industrial) and contaminants present in these waters may also appear in the effluent, depending on the treatment process (2008 email from Canadian Wastewater Association to J. Pasternak, Environment Canada; unreferenced).

Federal, provincial/territorial and municipal governments all play a role in managing treated drinking water quality in Canada (Cheminfo Services Inc. 2008). Voluntary guidelines have been established for aluminum concentrations in drinking water, and while provincial/territorial and municipal government authorities recognize these guidelines, they have not been adopted as mandatory standards. For example, in British Columbia, Alberta, Newfoundland and Manitoba, the Guidelines for Canadian Drinking Water Quality - Technical Documents: Aluminum as specified by Health Canada (i.e., 0.1 mg/L for conventional treatment plants using aluminum-based coagulants and 0.2 mg/L for other treatment systems using aluminum-based coagulants) are recognized, but specific standards have not yet been fully incorporated into operating permits for treatment facilities. In Ontario, Certificates of Approval with a limit of 0.1 mg/L are issued to drinking water treatment plants; however, this limit is included as a guideline rather than a standard. In Quebec, no limits on aluminum content in drinking water are found in the provincial regulations (including the Regulation Respecting the Quality of Drinking Water), and operating approvals are not required by wastewater treatment facilities (Cheminfo Services Inc. 2008).

Similarly, no federal legislation specific to municipal wastewater effluent discharges is in place (Cheminfo Services Inc. 2008). The federal government enforces CEPA (1999) that governs the releases of toxic substances to the environment, and the Fisheries Act that protects Canadian waters against the deposit of deleterious substances into fish habitat. In recent years, federal, provincial, and territorial governments have been working to develop a Canada-wide Strategy for the Management of Municipal Wastewater Effluent through the Canadian Council of Ministers of the Environment (CCME 2008); however, release standards for aluminum are not currently proposed or under development under the Strategy.

Less information is available on industrial releases of aluminum salts. The pulp and paper sector is the primary industrial user of aluminum salts, with applications in water treatment and as a paper additive. Alum is more commonly used for water treatment at mills in the warmer months of the year, while polyaluminum chloride (PAC) and polyaluminum silicate sulphate (PASS) have been found to be more effective winter coagulants. Recent quantitative release data for industrial uses are not available, although average concentrations of residual aluminum in treated water are estimated to be in the range of 0.02 mg/L (Cheminfo Services Inc. 2008). A 35% to 40% decrease in use of aluminum salts as a pulp and paper additive has been reported for the period 2000 to 2006, indicating a significant reduction in demand for this application (Cheminfo Services Inc. 2008).

Germain et al. (2000) reported mean total aluminum levels ranging from 0.46 to 4.8 mg/L in wastewaters released into rivers by the pulp and paper industry over the period 1990 to 1997. Mean total aluminum levels measured for other types of industries ranged from 0.01 to 2.3 mg/L. Since 1995, pulp and paper mills have been subject to the Pulp and Paper Effluent Regulations passed in 1992 under the Fisheries Act. In Quebec, for example, implementation of these regulations has led to a mean reduction of approximately 60% in total aluminum concentrations present in effluents (Germain et al. 2000). Environmental Effects Monitoring (EEM) reports published by the pulp and paper industry provide information on the distance from point of discharge that is required to dilute an effluent to less than 1% in the receiving water body. In some cases, only a few metres were needed, while in others, up to 300 km was required. In these cases, water input from other watercourses was needed to achieve dilution to 1%.

Sludge containing aluminum from the salts used in industrial water treatment can be sent to landfill or to steam boilers and co-generation units that handle bark, sludge, or other fuels (Cheminfo Services Inc. 2008). Aluminum may be present in the fly ash after burning of the sludge, although a small portion may also be emitted to air along with particulate matter (PM) emissions. No data are available on aluminum concentrations in fly ash; however, potential PM emissions are usually controlled with baghouses, electrostatic precipitators or other PM control systems.

The use of sludge derived from aluminum-based water treatment facilities as a soil amendment is the primary pathway by which aluminum salts enter the terrestrial environment. It is likely that the amount of aluminum added to soil through this practice is small in comparison with aluminum naturally present in soil. Sludge disposal guidelines specifying maximum application rates and soil pH requirements exist for a number of provinces. In Ontario, sludge application rates cannot exceed 8 tonnes solids/ha/5 years and the pH of the receiving soil must be greater than 6.0 or liming is required (ME and MAFRA 1996). Still, potential exists for the release of aluminum into soil due to high amounts of the metal present in sludge residuals (Mortula et al. 2007). In addition, a shift in soil pH at the site of sludge application could mobilize aluminum in the sludge by shifting the chemical equilibrium towards more soluble forms of the metal. Soil acidification may occur during high water discharge events (e.g., storm events), when water entering the sludge deposition area has interacted with organic matter or travelled through more acidic upper mineral soils (Pellerin et al. 2002). Aluminum solubilized in this process is then available to be transported to adjacent soils or water bodies along shallow flow paths in the soil.

2.3.1 Environmental Fate

The sections below summarize the information available on the distribution and fate of aluminum and the three aluminum salts, aluminum chloride, aluminum nitrate and aluminum sulphate, in the environment. A more detailed discussion on environmental fate can be found in Bélanger et al. (1999), Germain et al.(2000) and Roy (1999a).

2.3.1.1 Air

In air, hydrated aluminum chloride will react with moisture to produce hydrochloric acid and aluminum oxide (Vasiloff 1991). Aluminum nitrate and aluminum sulphate are likely to react in the same way, forming nitric and sulfuric acids, respectively. As the three aluminum salts that are the subject of this assessment are not usually emitted to air, the amount of aluminum present in air due to these salts is expected to be negligible compared with amounts coming from the natural erosion of soil (Environment Canada and Health Canada 2000).

2.3.1.2 Water

Natural sources of aluminum release to aquatic systems include weathering of rocks, glacial deposits and soils and their derivative minerals, and atmospheric deposition of dust particles. The most obvious increases in aluminum concentrations have consistently been associated with environmental acidification (Driscoll and Schecher 1988; Nelson and Campbell 1991). For this reason, recently observed changes in global climate and alterations in the acidity of atmospheric and oceanic systems, both resulting at least in part from human activities, have the potential to influence the presence and mobility of aluminum in the environment (Pidwirny and Gow 2002; Crane et al. 2005). The relationship is complex, however, and more research is needed in order to elicit the nature of potential impacts and their consequences for biota. Crane et al.(2005) postulated that increasingly severe weather patterns occurring as a consequence of global climate change, such as an increased incidence of prolonged heavy rainfall in some areas, may intensify physical and chemical weathering processes. When combined with the effects of acidification of waters, this could lead to significant changes in the speciation and mobility of aluminum and other metals.

Soil minerals such as gibbsite (Al(OH)₃) and jurbanite (AlSO₄(OH)-5H₂O) are considered the primary sources of aluminum release to the aqueous environment, especially in poorly buffered watersheds (Driscoll and Schecher 1990; Campbell et al. 1992; Kram et al. 1995). In more buffered watersheds, a solid-phase humic sorbent in soil is involved in the release of aluminum (Cronan et al. 1986; Bertsch 1990; Cronan and Schofield 1990; Cronan etal. 1990; Seip et al. 1990; Taugbol and Seip 1994; Lee et al. 1995; Rustad and Cronan 1995).

The three aluminum salts -- chloride, nitrate and sulphate -- are highly soluble and will form various dissolved species on contact with water. The fate and behaviour of aluminum in the aquatic environment are very complex. Aluminum speciation, which refers to the partitioning of aluminum among different physical and chemical forms, and aluminum solubility are affected by a wide variety of environmental parameters, including pH, solution temperature, dissolved organic carbon (DOC) content, and the presence and concentrations of numerous ligands. Metals in solution may be present as dissolved complexes, as "free" or aquo ions, in association with particles, as colloids or as solids in the process of precipitating. Colloidal particles (i.e., those in the range of 0.001 to 1 µm) are important in the transport of metals in stream ecosystems (Kimball et al. 1995; Schemel et al. 2000), as well as the accumulation of metals in sediment (Church et al. 1997) and biofilm (Besser et al. 2001), and the transfer to biota. Farag et al. (2007) proposed that colloids and biofilm may play critical roles in the pathway of metals to the food chain. The reactivity of aluminum, as well as geochemical behaviour, bioavailability and toxicity, are dependent upon its speciation (Neville et al. 1988; Gagnon and Turcotte 2007).

There are two general types of ligands that can form strong complexes with aluminum in solution. Inorganic ligands include anions such as sulphate (SO₄^2-), fluoride (F^-), phosphate (PO₄^3-), bicarbonate (HCO₃^-) and hydroxide (OH^-), among others. Organic ligands include oxalic, humic and fulvic acids (Driscoll et al. 1980; Sparling and Lowe 1996). The relative concentrations of the inorganic and organic ligands generally determine the proportions and type of complexes that are formed in solution.

Interactions with pH (Campbell and Stokes 1985; Hutchinson and Sprague 1987; Schindler 1988; Driscoll and Postek 1996) and DOC (Hutchinson and Sprague 1987; Kullberg et al. 1993) are of primary importance to the fate and behaviour of aluminum. DOC will complex with aluminum in water, forming aluminum-organic complexes and reducing concentrations of monomeric forms of aluminum (Farag et al. 1993; Parent et al. 1996). At a pH of 4.5, a concentration of 1 mg DOC/L can complex approximately 0.025 mg Al/L, with this complexing capacity increasing as pH increases (Neville et al. 1988). Fractions of dissolved organic aluminum were estimated for various rivers in Canada using the MINEQL+ (Schecher and McAvoy 1994) and WHAM (Tipping 1994) models; the results suggested that the importance of complexation with dissolved organic material (DOM) decreased over the pH range 7.0 to 8.5, likely due to reduced concentrations of the Al³⁺ and AlOH²⁺ species which can associate with DOM (Fortin and Campbell 1999).

Aluminum is a strongly hydrolysing metal and is relatively insoluble in the neutral pH range (6.0-8.0) (Figure 2.1). In the presence of complexing ligands and under acidic (pH < 6) and alkaline (pH > 8) conditions, aluminum solubility is enhanced. At low pH values, dissolved aluminum is present mainly in the aquo form (Al³⁺). Hydrolysis occurs as pH rises, resulting in a series of less soluble hydroxide complexes (e.g., Al(OH)²⁺, Al(OH)²⁺). Aluminum solubility is at a minimum near pH 6.5 at 20°C and then increases as the anion, Al(OH)₄-, begins to form at higher pH (Driscoll and Schecher 1990; Witters et al. 1996). Thus, at 20°C and pH < 5.7, aluminum is present primarily in the forms Al³⁺ and Al(OH)²⁺. In the pH range 5.7 to 6.7, aluminum hydroxide species dominate, including Al(OH)²⁺ and Al(OH)²⁺, and then Al(OH)₃. Typically, at a pH of approximately 6.5, Al(OH)₃ predominates over all the other species. In this range, aluminum solubility is low, and availability to aquatic biota should also be low. At pH > 6.7, Al(OH)₄- becomes the dominant species. Aluminum-hydroxide complexes predominate over aluminum-fluoride complexes under alkaline conditions. However, the aluminum speciation determined for some rivers in Canada indicated that only one river, of pH less than 7, had a significant concentration (> 1%) of aluminum-fluoride complexes (Fortin and Campbell 1999). It is important to note that the various aluminum species described above are always present simultaneously at any pH value. The influence of pH in aquatic systems is mainly to change the proportion of all the species as the pH changes (2008 email from Canadian Wastewater Association to J. Pasternak, Environment Canada; unreferenced).

Mononuclear aluminum hydrolytic products combine to form polynuclear species in solution (Bertsch and Parker 1996). Aluminum begins to polymerize when the pH of an acidic solution increases to over 4.5:

2Al(OH)(H₂O)₅²⁺ Al₂(OH)₂(H₂O)₈ ⁴⁺ + 2H₂O

Polymerization gradually proceeds to larger structures, eventually leading to the formation of the Al13 polycation (Parker and Bertsch 1992a, 1992b). In nature, conditions that favour the formation of polynuclear forms of aluminum can occur during the liming of acidic aluminum-rich watersheds (Weatherley et al. 1991; Lacroix 1992; Rosseland et al. 1992) and possibly during the addition of alum to circumneutral waters (Neville et al. 1988; LaZerte et al. 1997).

Figure 2.1: Solubility of aluminum species (and total aluminum, Alt) in relation to pH in a system in equilibrium with microcrystalline gibbsite
(0.001 mM = 0.027 mg/L; Driscoll and Schecher 1990)

Click on image to enlarge.

Temperature has been shown to influence the solubility, hydrolysis and molecular weight distribution of aqueous aluminum species as well as the pH of solutions. Lydersen et al. (1990b) reported a higher degree of aluminum hydrolysis and greater polymerization to high molecular weight species in inorganic aluminum solutions stored for one month at 25°C compared with those stored for an equivalent period at 2°C. The researchers hypothesized that more advanced polymerization evident at the higher temperature resulted in more deprotonation and condensation reactions, possibly accounting for the observed lower pH of the 25°C test solutions (range 4.83 to 5.07 versus 5.64 to 5.78 in the solutions at 2°C). Solubility and sedimentation were significantly higher at 25°C, with dissolution controlled by microcrystalline gibbsite. While substantial amounts of high molecular weight aluminum species were present in the solution at 2°C, little sedimentation was observed. Dissolution at the lower temperature appeared controlled by an amorphous Al(OH)₃(s) with much higher solubility and, therefore, a high proportion of the high molecular weight inorganic aluminum species remained as colloids in the solution. The effects of low temperature on the coagulation efficiency of aluminum sulphate have been studied in relation to water treatment processes (Braul et al. 2001; Wobma et al. 2001; Kundert et al. 2004). The results provide further evidence that temperature-dependent fluctuations in the predominant aluminum species present in an aquatic system may occur in regions of Canada that experience marked seasonal fluctuations in temperature.

When released into water, for example within a drinking water treatment plant (DWTP), most of the aluminum associated with the aluminum salts considered in this report hydrolyses to form aluminum hydroxides (Hossain and Bache 1991). Reactions between aluminum salts, water and associated "impurities" result in the formation of a floc, which separates from the water phase to form alum sludge. A small fraction of the aluminum can stay in the water in either colloidal or dissolved form. Barnes (1985) describes the different reactions involved in the formation of aluminum hydroxide in aqueous solution; the overall reaction can be represented by the following equation:

Al₂(SO₄)₃ + 6H₂O 2Al(OH)₃0 + 3H₂SO₄

The aluminum hydroxide present in sludge is expected to remain mostly solid following release into surface water. Ramamoorthy (1988) showed that less than 0.2% of the aluminum hydroxide present in sludge was released in supernatant water at a pH of 6 and less than 0.0013% was released at pH 7.65. In both cases, aluminum hydroxide was present mostly in particulate form. At these pH values, aluminum solubility is low and kinetics favour the formation of solid aluminum hydroxide.

When used to treat sewage water, alum will also react with phosphate, as shown in the following reaction (Romano 1971; Barnes 1985):

Al₂(SO₄)₃ + 2PO43- AlPO₄(s)+3SO₄^2-

This process has been used for many years to treat phosphorus in wastewaters, as well as to reduce phosphorus levels in runoff from land fertilized with poultry litter and restore phosphorus-enriched eutrophic lakes (Lewandowski et al. 2003).

Kopácek et al. (2001) examined the possible role of aluminum in influencing the natural cycling of phosphorus, which is often a limiting nutrient in aquatic systems. The researchers postulated that aluminum from nearby lower pH soils may enter circumneutral water bodies during episodic acidification events, such as spring melt, leading to the formation of colloidal aluminum oxyhydroxide flocs which will strongly adsorb orthophosphate in the water column. The phosphate-bound particulate aluminum settles onto the lake bottom, removing the bioavailability of this phosphorus to organisms in the water column. The increasing sediment concentrations of aluminum-phosphorus floc disrupt the redox-dependent cycling of phosphorus in the lake, indicating that while aluminum does not enter directly into biotic cycles, it is capable of influencing the biogeochemical cycles of substances that are integral to living systems. Based on the solubility characteristics of aluminum (see Figure 2.1), this process may also occur when acidic waters, which generally contain the most aluminum (Gensemer and Playle 1999), enter downstream waters of higher pH.

The cycling and availability of other trace elements (e.g., nitrogen) and of organic carbon may also be influenced by the adsorption and coagulation properties of aluminum (Driscoll and Schecher 1990; Lee and Westerhoff 2006). Dissolved organic carbon (DOC) has been shown to provide an important weak acid/base buffering system that aids in the regulation of pH in dilute acidic waters and removal of DOC by adsorption to aluminum could adversely affect pH conditions in a water body (Johannessen 1980; Driscoll and Bisogni 1984). As well, coagulation and removal of DOC and other light attenuating materials may alter patterns of water column heating, resulting in decreased thermal stability in a water body (Almer et al. 1974; Malley et al. 1982). Changes to the heating pattern and thermal stratification of a lake can profoundly impact ecosystems by altering the vertical transport of solutes and restricting coldwater fisheries (Driscoll and Schecher 1990).

Aluminum is highly reactive in seawater and will be rapidly scavenged by particulate matter when released into this medium (Nozaki 1997). The mean oceanic residence time for aluminum is predicted to be short compared to some other elements, in the range of 100 to 200 years, with vertical distribution dictated by terrestrial and atmospheric inputs at the surface, intense particle scavenging throughout the water column, and some regeneration in bottom waters (Orians and Bruland 1985). The higher ionic strength and relative magnitude of individual ion concentrations in saline waters compared with freshwaters lead to differences in coagulation reactions with aluminum salts. Duan et al. (2002) identified distinctly different characteristics between the two water types with respect to colloid destabilization, coagulation mechanisms, and colloidal removal. These differences can become important when water treatment processes include release of effluent or backwash materials into marine or brackish waters.

2.3.1.3 Sediment

Sediment, where metals are generally considered less biologically available, is nonetheless an important medium for aluminum (Stumm and Morgan 1981; Campbell et al. 1988; Tessier and Campbell 1990). Aluminum occurs naturally in aluminosilicates, mainly as silt and clay particles, and can be bound to organic matter (fulvic and humic acids) in sediments (Stumm and Morgan 1981). At pH > 5.0, dissolved organic matter (DOM) can co-precipitate with aluminum, thereby controlling its concentrations in lakes with elevated concentrations of DOM (Urban et al. 1990). DOM plays a similar role in peatlands (Bendell-Young and Pick 1995). At pH < 5.0, the cycling of aluminum in lakes is controlled by the solubility of mineral phases such as microcrystalline gibbsite (Urban et al. 1990). Lakes receiving drainage from acidified watersheds can act as a sink for aluminum (Troutman and Peters 1982; Dillon et al. 1988; Dave 1992).

Experimental acidification of lakes and limnocorrals has shown that aqueous aluminum concentrations rapidly increase in response to acidification (Schindler et al. 1980; Santschi etal. 1986; Brezonick et al. 1990). Mass-balance studies have demonstrated that retention of aluminum by sediments decreases as pH decreases (Dillon et al. 1988; Nilsson 1988). Under such conditions, sediments in acidified watersheds can provide a source of aluminum to the water column (Nriagu and Wong 1986). Based on calculation of fluxes in acidic lakes, Wong et al. (1989) suggested that sediment is a source of aluminum to the overlying water column.

The release of aluminum hydroxide sludge from drinking water treatment plants (DWTPs) directly to surface waters is the primary pathway by which aluminum from aluminum salts enters sediment. If water velocity is low at the point of discharge, much of the released sludge will settle onto the surface of local sediment. Since, in Canada, the waters receiving such discharges are typically circumneutral, the solubility of aluminum in the sludge will generally be minimal (Environment Canada and Health Canada 2000).

2.3.1.4 Soil

Atmospheric deposition of aluminum to soil is attributed mostly to the deposition of dust particles and is generally low (Driscoll et al. 1994). Volcanic activity can also act as a major natural source of aluminum to soil (Pichard 2005). Aluminum is the third most abundant element in the earth's crust, making up approximately 8% of rocks and minerals and accounting for about 1% of the total mass of the Earth (Landry and Mercier 1992; Skinner and Porter 1989). Approximately 75% of Canada is covered by glacial till (Landry and Mercier 1992); examples of aluminum-bearing minerals inherited from glacial till (i.e., primary minerals) are feldspars, micas, amphiboles and pyroxenes. Transformation of primary minerals by chemical weathering reactions results in new solid phases (i.e., secondary minerals). Aluminum-bearing secondary minerals such as smectite, vermiculite and chlorite are often found in Canadian soils developed on glacial till.

Inputs of aluminum into soil solutions usually occur by mobilization of aluminum derived from the chemical weathering of soil minerals. The most important reaction in the chemical weathering of the common silicate minerals is hydrolysis. However, aluminum is not very soluble over the normal soil pH range; thus, it generally remains near its site of release to form clay minerals or precipitate as amorphous or crystalline oxides, hydroxides or hydrous oxides. Silica is much more soluble than aluminum at normal soil pH and is always in excess of the amount used to form most clay minerals, so that some is removed from the soil system in leachates (Birkeland 1984). In some parts of the world, the extent of chemical transformation by chelation is believed to exceed that by hydrolysis alone. In forest soils of cold and humid regions, such as those of eastern Canada, aluminum is believed to be transported from upper to lower mineral soil horizons by organic acids leached from foliage and the slow decomposition of organic matter in the forest floor (Courchesne and Hendershot 1997). The movement of aluminum-organic complexes stops when the soil solution becomes saturated (or when the aluminum-to-organic-carbon ratio reaches a critical value), thereby reducing their solubility. In pristine conditions, aluminum is normally retained within the B horizon of the soil. A third important reaction involving aluminum is the transformation of one mineral into another through the exchange of interlayer cations (Sposito 1996).

Although the dissolution and precipitation reactions of aluminum-bearing minerals are often good indicators of the solubility of aluminum in soils, they are by no means the only pedogenic processes controlling the concentrations of aluminum in soil solutions. Many other processes may partly control the uptake of aluminum by plants and soil organisms. Aluminum may be 1) adsorbed on cation exchange sites, 2) incorporated into soil organic matter, 3) absorbed by vegetation or 4) leached out of the soil system (Ritchie 1995). Aluminum can form stable complexes with various types of soluble and insoluble organic matter, from simple low-molecular weight organic acids to humic and fulvic acids (Vance et al. 1996; Ritchie 1995). Organic ligands play an important role in the speciation of aluminum in soil solutions (David and Driscoll 1984; Driscoll et al. 1985; Ares 1986).

In eastern Canada, the atmospheric deposition of strong acids, such as nitric acid and sulfuric acid, has accelerated the natural acidification of soil. The increased H+ activity (lower pH) in the soil solution creates a new equilibrium where more Al³⁺ is dissolved in the soil solution, cation nutrients (Ca²⁺, Mg²⁺ and K⁺) are replaced on the soil exchange complex by Al³⁺ and the base cations are eventually leached out of the soil.

There may be significant variation in Al³⁺ solubility with depth in a soil profile (Hendershot et al. 1995). In the surface horizons, the soil solutions tend to be undersaturated with respect to aluminum-bearing minerals; in the lower B and C horizons, aluminum in soil solutions can be expected to be near equilibrium with some aluminum solids. Although the equilibrium concentration is close to that which would be expected if gibbsite were controlling equilibrium, gibbsite has generally not been identified in Canadian soils. Other forms of aluminum, for example, hydroxy interlayered vermiculite, may control aluminum solubility at values close to those of gibbsite. Amorphous aluminum complexed with organic matter may also have a similar pH solubility curve that is a function of the pH-dependent variation in the number of binding sites.

Fluoride and hydroxide complexes are the two strongest groups of inorganic ion associations with aluminum in soil solutions (Nordstrom and May 1995). In very acidic soils, aluminum in the soil solution is present mainly as free Al³⁺; as pH increases, free Al³⁺ hydrolyses to form complexes with OH^- ions (e.g., AlOH²⁺, Al(OH)²⁺, Al(OH)₃0). Near pH 6.5, aluminum solubility is at a minimum, but it increases at neutral to alkaline conditions because of the formation of Al(OH)₄^- (Driscoll and Postek 1996). According to Lindsay et al. (1989), fluorine, the most electronegative and one of the most reactive elements, is released as fluoride ion through the dissolution of fluoride-bearing minerals. In acidic soils (pH < 5.5), low-ligand-number complexes such as AlF²⁺ are normally formed. In neutral to alkaline conditions, it is more difficult for F- to compete with OH^- for aluminum in the soil solution because of the increased level of OH^- and probably the presence of calcium that tends to link with fluoride (CaF₂). Consequently, aluminum-hydroxide complexes predominate over aluminum-fluoride complexes in alkaline conditions.

The complexation of aluminum with sulphate is weaker than that with fluoride. However, in acidic soils where the sulphate concentration is high, aluminum may also form aluminum-sulphate complexes (Driscoll and Postek 1996). At low sulphate concentrations, AlSO₄⁺ is the dominant aqueous form, whereas Al(SO₄)₂^- is predominant in soil solutions with higher sulphate concentrations. Brown and Driscoll (1992) showed that several aluminosilicate complexes, including AlSiO(OH)₃²⁺, are present in various regions of the eastern U.S. and Canada.

It has been shown that most dissolved aluminum in soil solution of the forest floor is organically bound and that these aluminum-organic complexes become less abundant with increasing soil depth (Nilsson and Bergkvist 1983; David and Driscoll 1984; Driscoll et al. 1985). In the Adirondacks of New York, David and Driscoll (1984) found that 82% and 93% of the total dissolved aluminum in the organic horizons of conifer and hardwood stands, respectively, were organically complexed. The proportion of organic to inorganic aluminum decreased at both sites from the organic to the upper mineral horizons and from the upper to the lower mineral horizons. In the soil solutions of the mineral horizons, aluminum-organic complexes accounted for 67% and 58% of the total aluminum in the conifer and hardwood sites, respectively, which indicates the importance of aluminum-organic complexes in humus-rich forest soils of eastern North America.

2.3.1.5 Biota

In general terms, a substance is considered to be bioavailable if, under the conditions of exposure, it can be taken up by organisms (Environment Canada 1996). The bioavailability of a substance is determined by its chemical form, the physical and chemical characteristics of the media (e.g., water, soil, food) in which it occurs, the receptor species, and the route of the exposure (e.g., dermal contact, ingestion, inhalation). For metals such as aluminum, the "free" or hydrated dissolved ions (i.e., Al³⁺, Al(OH)²⁺ and Al(OH)₂⁺) are normally considered to be the principal bioavailable forms (Newman and Jagoe 1994). However, there is evidence that some other forms of a metal, such as organometallic compounds (e.g., of mercury and tin), oxyanions of the metal (e.g., CrO₄^2-, AsO₄^3-), and dissolved organic and inorganic metal complexes (e.g., colloidal and polynuclear aluminum complexes) can also be taken up by organisms (Parker and Bertsch 1992b; Benson et al. 1994; Campbell 1995).

Bioavailability directly influences the potential for bioconcentration, bioaccumulation and biomagnification of a substance in organisms. ICMM (2007) defines bioconcentration as the increase in concentration of a substance in an organism (or specified tissues thereof) relative to the concentration of the substance in the environmental medium (generally water) to which it is exposed, bioaccumulation as the amount of a substance taken up by an organism from water (bioconcentration) as well as through ingestion via the diet and inhalation, and biomagnification as the process by which the tissue concentration of a bioaccumulated substance increases as it passes up the food chain through at least two levels (Parametrix 1995). The three processes are significant indicators of the propensity of a substance to impart toxicity to individual organisms and at higher trophic levels in the food chain. However, bioaccumulation of essential elements (such as some metals) in organisms is typically subject to metabolic regulation (ICMM 2007).

Bioconcentration factors (BCFs) and bioaccumulation factors (BAFs) are unitless values derived by dividing steady state tissue concentrations of a substance by the steady state environmental concentration (ICMM 2007). For synthetic organic compounds, the use of a BCF and BAF threshold value (such as that of 5000 specified in the CEPA 1999 Persistence and Bioaccumulation Regulations; Canada 2000) provides valuable information for the evaluation of hazard and risk. Bioaccumulation is more complex for naturally occurring inorganic substances such as metals, however, as processes such as adaptation and acclimation can modulate both accumulation and potential toxic impact (ICMM 2007). All biota will naturally accumulate metals to some degree without deleterious effect and as some metals are essential elements, bioaccumulation does not necessarily indicate the potential for adverse effects (McGreer et al. 2003). While metal bioaccumulation is homeostatically regulated for metals essential to biological function (Adams et al. 2000), non-essential metals may also be regulated to some degree as these homeostatic mechanisms are not metal-specific (ICMM 2007).

Thus, interpretation of the toxicological significance of bioaccumulation data for metals such as aluminum is complex. A more complete discussion of aluminum bioavailability and the implications for bioaccumulation and toxicity can be found in Roy (1999a) and Bélanger et al. (1999).

Few studies have examined the uptake and accumulation of aluminum by algae. While the algal bioassays conducted by Parent and Campbell (1994) were not specifically designed to determine the effect of pH on aluminum bioaccumulation, their data indicated that the accumulation of aluminum by Chlorella pyrenoidosa increased with the concentration of inorganic monomeric aluminum. In addition, the comparison of assays performed at the same concentration of aluminum but at different pH values showed that aluminum accumulation was suppressed at low pH (Parent and Campbell 1994). Aquatic invertebrates can also accumulate substantial quantities of aluminum, yet there is evidence that most of the metal is adsorbed to external surfaces and is not internalized (Havas 1985; Frick and Hermann 1990). Using the results of Havas (1985), the bioconcentration factor (BCF) for Daphnia magna varied from 10,000 at pH 6.5 down to 0 at pH 4.5. Similar results, i.e., decreasing accumulation of aluminum with decreasing pH, were reported for crayfish (Malley et al. 1988), caddisfly (Otto and Svensson 1983), unionoid clams (Servos et al. 1985) and a chironomid (Young and Harvey 1991). Other studies with clams and benthic insects showed no relationship between water pH and tissue accumulation (Sadler and Lynam 1985; Servos etal. 1985). Frick and Herrmann (1990) found that the largest portion (70%) of the aluminum was present in the exuvia of the mayfly, Heptagenia sulphurea, indicating that the metal was largely adsorbed and was not incorporated into the organism.

BCFs for fish were calculated to range from 400 to 1,365 based on results presented in Roy (1999a). Numerous field and laboratory studies have demonstrated that fish accumulate aluminum in and on the gill. It has been suggested that the rate of transfer of aluminum into the body of fish is either slow or negligible under natural environmental conditions (Spry and Wiener 1991). The initial uptake of aluminum by fish essentially takes place not on the gill surface but mainly on the gill mucous layer (Wilkinson and Campbell 1993). Fish may rapidly eliminate mucus and the bound aluminum following the exposure episode. For example, Wilkinson and Campbell (1993) and Lacroix et al. (1993) found that depuration of aluminum from the gills of Atlantic salmon (Salmo salar) was extremely rapid once fish were transferred into clean water. The authors suggested that the rapid loss is due to expulsion of aluminum bound to mucus.

Far fewer studies have examined aluminum accumulation in benthic organisms. However, chironomids do not appear to accumulate aluminum to the same degree as other aquatic invertebrates. Krantzberg (1989) reported that the concentration of aluminum in chironomids was < 0.3 nmol/g dw for the entire body and < 0.1 nmol/g dw for the internal structures. Most aluminum is either adsorbed externally or is associated with the gut contents of chironomids (Krantzberg and Stokes 1988; Bendell-Young et al. 1994).

BCFs for terrestrial plants were calculated based on data cited in the review by Bélanger et al. (1999). For both hardwood and coniferous species, the calculated BCF ranged from 5 to 1,300 for foliage and from 20 to 79,600 for roots in studies done with aluminum solutions. For those conducted with soil, BCFs were lower for both foliage (0.03-1.3) and roots (325-3,526). BCFs calculated for grain and forage crops ranged from 4 to 1,260 in foliage and from 200 to 6,000 in roots for experiments done with solutions. For soil experiments, the foliar BCF varied from 0.07 to 0.7.

2.3.2 Environmental concentrations

To determine aluminum concentrations in various environmental media in Canada, the most recent available data in Canada were used where possible, although data from other countries were examined as well. Concentrations in environmental media to be used as input into the human health risk assessment (i.e., air, drinking water, soil, and food) are estimated based on total aluminum. Bioavailability of aluminum in different media in relation to absorption in humans is considered separately in section 2.3.3. Data presented below are also relevant to the assessment of ecotoxicological effects.

2.3.2.1 Air

2.3.2.1.1 Ambient air

Ambient air at more than 40 Canadian sites, primarily in urban areas, was sampled over a period of ten years (1996-2006). More than 10,000 samples were measured at different sites throughout Canada, although the number varied from year to year. In 2006, only 25 sites were measured, resulting in 1,400 samples, 96% of which had levels greater than the detection limit (approximately 0.001 µg/m³).

Total aluminum concentrations measured in individual samples of PM10 (i.e., particulate matter smaller than 10 µm in diameter) ranged from the detection limit to 24.94 µg/m³, with the lowest concentration being measured in Saint John, New Brunswick and the highest in Vancouver, British Columbia (Dann 2007). Figure 2.2 shows estimated mean aluminum concentrations measured in ambient air for all sampling sites by province for the ten-year period. On the basis of these measurements from across Canada, the estimated provincial/territorial mean aluminum concentration in PM10 is 0.17 µg/m³. This value was used for the purpose of assessing exposure of the Canadian population to aluminum in ambient air.

Figure 2.2: Mean aluminum concentrations in PM10 in outdoor air from provinces and territories across Canada (µg/m³) (1996-2006)

Click on image to enlarge.

For most of the Canadian sites where PM10 measurements were carried out, data were also available for PM2.5 particles (i.e., smaller than 2.5 µm in diameter). Close to 20,000 measurements were available from 1998 to 2006, 77% of which had levels greater than the detection limit. Using all available data, the mean aluminum concentration in PM2.5 in Canada is approximately 0.069 µg/m³, with a maximum aluminum concentration of 9.24 µg/m³ measured in Vancouver, British Columbia (Dann 2007).

No published data were available on aluminum levels in ambient air in the vicinity of aluminum smelters or other industries in Canada, and limited data from other countries were identified. In an industrial area of the province of Turin in Italy, levels of 1.12 and 0.4 µg/m³ of aluminum were measured during industrial activity and during holidays, respectively, (Polizzi et al. 2007). According to JECFA (2007), the concentration of aluminum in ambient air of industrial areas may range from 25 to 2,500 µg/m³. It should be noted that the three aluminum salts -- chloride, nitrate and sulphate -- are unlikely to have contributed significantly to total concentrations measured in ambient air, as their use does not generally result in air emissions of aluminum.

2.3.2.1.2 Indoor air

Few data on aluminum concentrations in indoor air in residential dwellings were identified for Canada. Studies in the U.S. did provide data on aluminum in indoor air. These findings are summarized below.

In 1990, a Particle Total Exposure Assessment Methodology (PTEAM) study was conducted in Riverside, California, in which samples were collected from 178 non-smokers over ten years of age. In addition to the personal sampling (portable sampler), stationary samplers were set up inside the residential dwellings and outside near the entrance door. Airborne particle (PM10 and PM2.5) samples were collected for two 12-hour periods (nighttime and daytime), and more than 2,900 samples were analyzed (Clayton et al. 1993; Thomas et al. 1993). In this study, the aluminum concentrations exceeded the reporting limit of 0.5 µg/m³ in more than half of the personal PM10 samples taken during the two periods. In the case of PM2.5, only 20% of the measurements exceeded the reporting limit. Estimated daytime median concentrations of aluminum for the PM10 indoor, outdoor and personal exposure monitors were 1.9, 2.5 and 3.4 µg/m³, respectively; the corresponding nighttime median concentrations were 0.99, 1.7 and 1.0 µg/m³. Based on the average daytime and nighttime concentrations of aluminum in PM10 particles, the estimated mean concentration of aluminum in indoor air was about 1.49 µg/m³.

For the purpose of assessing exposure for the general Canadian population, this estimated mean concentration of aluminum in PM10 particles of 1.49 µg/m³ was considered to represent the typical indoor air concentration of aluminum in Canada. As in the case of ambient air, the three aluminum salts -- chloride, nitrate and sulphate -- are unlikely to have contributed significantly to total aluminum concentrations measured in indoor air.

2.3.2.2 Water

2.3.2.2.1 Surface water

Aluminum is a naturally occurring element and is present in all water bodies in Canada and elsewhere. Aluminum can be analysed under different forms, but historically results were reported mostly as total aluminum because of the low cost and ease of analysis. In many cases, results are also available for extractable or dissolved aluminum. Total aluminum represents all the aluminum present in a water sample, including the particulate fraction. Extractable aluminum includes both the "dissolved" fraction and weakly bound or sorbed aluminum on particles, and "dissolved" aluminum represents the fraction present in a sample filtered through a 0.45 µm membrane. All the bioavailable aluminum is considered to be present in this fraction, but not all the dissolved aluminum is bioavailable. Colloidal aluminum (0.01 to 0.1 µm) and organic aluminum (aluminum bound with soluble organic ligands) that are included in this fraction are generally thought to be less bioavailable than truly dissolved forms of the metal (Roy 1999a).

At reference lake and river sites across Canada that have not been influenced by effluents from facilities using aluminum salts, mean total aluminum concentrations ranged from 0.05 to 0.47 mg/L, with a maximum value of 10.4 mg/L, measured in British Columbia. Mean extractable aluminum concentrations ranged from 0.004 to 0.18 mg/L, with a maximum value of 0.52 mg/L found in a lake in the Abitibi region of Quebec. Mean dissolved aluminum concentrations varied from 0.01 to 0.08 mg/L and the highest dissolved aluminum value reported was 0.9 mg/L in British Columbia (Germain et al. 2000).

Aluminum was measured in water taken both upstream and downstream of facilities using aluminum salts and releasing aluminum or aluminum salts, but sampling stations were typically not located close enough to sources to allow the local impact of the effluents to be assessed. Mean total aluminum levels generally varied from 0.002 to 2.15 mg/L, with a maximum value of 28.7 mg/L, measured in the Oldman River, 40 km downstream of Lethbridge, Alberta. Total aluminum levels are usually higher in the Prairies, in rivers with high total particulate matter content. Mean extractable aluminum concentrations ranged from 0.03 to 0.62 mg/L, and the maximum value of 7.23 mg/L was reached in the Red Deer River, at Drumheller, Alberta. Mean dissolved aluminum concentrations were much lower, ranging from 0.01 to 0.06 mg/L. In surface water, the maximum dissolved aluminum concentration (0.24 mg/L) was measured in the Peace River, Alberta (Germain et al. 2000). Concentrations in downstream locations were not consistently elevated in relation to concentrations in upstream locations, suggesting that the impacts of releases of aluminum salts are mostly local.

Although information on the forms of dissolved aluminum present at these monitoring locations was not identified, results of equilibrium modelling suggest that most dissolved aluminum in waters with pH values of 8.0 and higher is in inorganic monomeric forms (Fortin and Campbell 1999). For the 12 Prairie locations where dissolved and total aluminum levels were reported, pH levels were 8.0 or higher, and dissolved aluminum represented less than 3% of total aluminum (Roy 1999b). The overall average concentration of dissolved aluminum at these sites was 0.022 mg/L, similar to levels of inorganic monomeric aluminum reported in comparatively pristine Adirondack surface waters (pH from ~5.8 to ~7.2), where most values were around 0.027 mg/L (Driscoll and Schecher 1990).

Empirical data indicating an increase in aluminum levels in ambient water receiving inputs of aluminum salts were available for only a few locations. A total aluminum concentration of 36 mg/L was attained just downstream of the discharge pipe of a Regional Municipality of Ottawa-Carleton's (RMOC) DWTP in water samples taken following a rountine release of backwash in 1993; samples taken 200 m downstream of the discharge pipe showed a total aluminum level of 0.5 mg/L In 1994, the total aluminum level reached 11.3 mg/L just downstream of the discharge. In 2008, all wastes previously destined for the Ottawa River from RMOC DWTPs were diverted completely to the local sewage treatment plant for treatment prior to discharge (Wier, pers. comm. 2008). In the Kaministiquia River, the increase in mean total aluminum noted from upstream to downstream stations corresponds approximately to the inputs from the pulp and paper mill located in Thunder Bay, Ontario. The mean difference of 0.071 mg/L observed in total aluminum concentrations for samples taken on the same day at both stations for the period 1990-1996 is equivalent to the predicted aluminum increase of 0.069 mg/L calculated with the aluminum releases reported by the mill (Germain et al. 2000). For the Ottawa and Kaministiquia rivers, estimated dissolved monomeric aluminum levels were 0.027 mg/L and 0.040 mg/L, respectively. These values were obtained using the MINEQL+ model and estimated concentrations in effluents, assuming solubility controlled by microcrystalline gibbsite (Fortin and Campbell 1999). Using boehmite as the controlling phase provides lower dissolved inorganic aluminum levels (0.005 mg/L and 0.007 mg/L, respectively).

The Quebec Environment Ministry, now Ministère du Développement durable, de l'Environnement et des Parcs, and Environment Canada examined the toxic potential of effluents generated by 15 municipal wastewater treatment plants in Quebec (Ministère de l'Environnement du Québec and Environment Canada 2001). The plants were considered to represent treatment methods used most commonly in Quebec and serviced over 50% of the province's population. Whole effluent sampling was conducted twice a year, during summer and winter operating conditions, over the period 1996 to 1999. Total aluminum concentrations in the effluents ranged from below the detection limit (0.002 to 0.1 mg/L) to 3.57 mg/L in summer and up to 4.25 mg/L under winter operating conditions. Concentrations remained at or below 1 mg/L year-round in all but two of the plants; however, 20 out of 45 summer readings and 25 out of 39 winter readings exceeded the maximum interim water quality guideline of 0.156 mg/L for the protection of freshwater life (water pH equal to or greater than 6.4) as recommended by CCME (2003). The study concluded that ammonia nitrogen and surfactants were mainly responsible for the observed effluent toxicity, with pesticides possibly a factor during summer months; however, the presence of aluminum in the effluents at levels above background may also have contributed to some extent. The results suggest that periodic episodes of aluminum toxicity are possible in some receiving waters; however, the nature of the collected data makes concluding on potential risk to the environment difficult. The study was designed to evaluate the toxic potential of whole effluents and did not include consideration of factors such as dilution effects, interactions between constituents in the effluents, and natural background levels of aluminum in the receiving environments. Therefore, while effluent concentrations may have exceeded the recommended water quality guideline, it is uncertain whether these guidelines were also exceeded in the surface waters receiving these effluents. In addition, it is likely that a large fraction of the total aluminum present in the effluents was associated with particulates that would settle out of the water column upon release into surface waters (Germain et al. 2000). This would substantially reduce the potential for adverse impacts to pelagic organisms, although negative impacts to benthic organisms could still occur. These impacts could relate directly to aluminum toxicity or be associated with physical aspects such as blanketing effects and/or the presence of other toxic contaminants.

Agencies such as the Greater Vancouver Regional District (GVRD; now Metro Vancouver) routinely monitor wastewater products generated at municipal treatment plants, in order to evaluate effluent quality and ensure compliance with provincial regulations such as the Environmental Management Act. Wastewater monitoring in the GVRD is conducted by the Greater Vancouver Sewerage & Drainage District (GVS&DD) and includes determination of total and dissolved aluminum concentrations in wastewater treatment plant influents and effluents, as well as estimates for influent and effluent loading of aluminum. Monthly data summaries are provided on the GVRD website and these are compiled annually into a Quality Control Report. For 2006, the latest report available on the website, influent concentrations measured at the five wastewater treatment plants operating in the GVRD ranged from 0.47 to 2.74 mg/L and 0.04 to 0.25 mg/L for total and dissolved aluminum, respectively (GVRD 2006), while effluent values were 0.05 to 0.97 mg/L and 0.02 to 0.16 mg/L. While influent concentrations of total aluminum were generally comparable between primary and secondary wastewater treatment plants, mean total aluminum concentrations were higher in primary treatment effluents as compared with those from plants using secondary treatment, likely reflecting greater removal of particulate aluminum from the water phase during the coagulation and flocculation process of secondary treatment. In general, influent concentrations of both total and dissolved aluminum were comparable between the two types of wastewater treatment. However, estimated loading rates varied widely between the plants and annually within each plant, with influents ranging from 7.8 to 1,380 kg/d total and 1.0 to 98 kg/d dissolved aluminum, and effluent rates 0.9 to 943 kg/d and 0.2 to 59 kg/d for total and dissolved aluminum, respectively. An analysis of total aluminum concentrations in treatment plant effluents from 1997 to 2006 indicated that levels had remained generally stable around 0.1 to 1.0 mg/L or decreased steadily during this period. A marked reduction in total aluminum was observed at two plants following the implementation of secondary treatment in 1998 and 1999, confirming the efficacy of this process in removing particulate aluminum from water.

2.3.2.2.2 Drinking water

Many drinking water treatment plants in Canada using surface water supplies add aluminum salts (aluminum sulphate, aluminum chloride or polymer forms) as a coagulant/flocculent to eliminate organic compounds, micro-organisms and suspended particulate matter. Treatment with aluminum salts may not necessarily increase the total aluminum concentration in finished drinking water, as the aluminum associated with suspended solids is removed. However, aluminum salt addition does appear to increase the concentration of low-molecular-weight, dissolved aluminum species, which may potentially present a higher bioavailability (Health Canada 1998b). More information on the bioavailability of aluminum from drinking water can be found in section 2.3.3.1.1.

For most provinces and territories, data on concentrations of aluminum in drinking water were obtained directly from municipalities that use aluminum salts in drinking water treatment (Health Canada 2007b). Data were also obtained from monitoring programs carried out in five provinces and territories from 1990 to 1998 (Environment Canada and Health Canada 2000). Over 10,000 drinking water samples from approximately 1,200 sites across Canada were analyzed over the past 20 years. The majority of the data analyzed was collected over ten years, in some cases up to 2007 (Health Canada 2007c).

In drinking water treatment systems in Canada that have surface water sources and use aluminum salts, the mean total aluminum concentration was estimated at 101 µg/L². Mean concentrations for the different provinces (see Figure 2.3) varied from 20.0 µg/L in New Brunswick (between 1995 and 2007) to 174 µg/L in Alberta (between 1990 and 2002).

In addition to the analysis of alum-treated drinking water, more than 2,800 samples of drinking water derived from groundwater sources from various Canadian municipalities were analyzed. Aluminum salts are not used in treatment of groundwater, except in the case of certain sites in the Northwest Territories. New Brunswick private wells had the highest mean total aluminum concentrations at approximately 40.0µg/L, whereas Ontario had the lowest concentrations of about 10.0µmg/L. On the basis of all the data from about 30 drinking water treatment systems in Canada, the mean aluminum concentration is estimated to be 25.2µg/L in groundwater sources, which is four times lower than that estimated for surface water treated with alum.

The average value of 101 µg/L, associated with the various aluminum salt-treated water supplies was used for the purpose of assessing exposure of the Canadian population to aluminum in drinking water.

Figure 2.3: Mean total aluminum concentrations in aluminum-treated drinking water from provinces and territories across Canada (µg/L) (1990-2007)

Click on image to enlarge.

2.3.2.3 Sediment

Based on limited data, total aluminum levels in Canadian sediments are of the same order of magnitude as those measured in soils (see Section 2.3.2.4), with levels varying between 0.9% and 12.8%. The highest levels were found in Lake St. Louis, Quebec. Of particular interest are aluminum levels measured in sediment of the Ottawa River less than 300 m downstream of a location where backwash water (from the Britannia DWTP)_ had been discharged for approximately 27 years (Environment Canada 2008c). In 1989, the mean total aluminum content of sediment collected from a control site situated 100 m off the treatment plant effluent plume was 17,543 mg/kg dw, while the value closest to the outfall was 125,160 mg/kg dw (Germain et al. 2000). Mean concentrations measured 300 m and 500 m downstream of the plant discharge point were 51,428 and 41,331 mg/kg dw, respectively, still elevated compared with the control site and that of an upstream location (mean concentration 20,603 mg/kg dw). In a follow-up study conducted in 2000 (City of Ottawa 2002), sampling confirmed that concentrations of aluminum were highest in riverbed sediment located at the discharge outlet of the Britannia DWTP (approximate mean of 150,000 mg/kg dw), then declined over 500 m to approximately 12,000 mg/kg dw. This concentration was not appreciably higher than the sampling location 150 m upstream from the discharge outlet (10,000 mg/kg). The aluminum concentration then increased to approximately 61,000 mg/kg at the 1,500 m sampling site indicating that this was likely a far-field zone of deposition. Waste discharges of aluminum-bearing sludge from Ottawa DWTPs previously destined for the Ottawa River have been diverted in 2008 to the local WWTP for treatment (Environment Canada 2008c).

2.3.2.4 Soil

Aluminum is the third most abundant element in the earth's crust, after oxygen and silicon, occurring as aluminosilicates and other minerals. The data on soil aluminum concentrations presented below come from soil surveys covering various geographic areas, and generally represent naturally-occurring aluminum concentrations.

In Canada, soil sampling has been carried out since the 1930s, but analysis for aluminum has only occurred in the past 20 years. Data for more than 40 studies based on over 40,000 soil samples across Canada from the past 20 years are thus available and were used to estimate the average total aluminum concentration in soil. Two studies cover all of Canada, while others focus on specific regions such as the Prairies, a province, or a municipality, in connection with local industries, types of soil, soil horizons, soil groups, or land use. In addition, some Canadian data on aluminum in dust from inside residential dwellings were available for consideration. More detailed information describing the available soil concentration data may be found in the supporting documentation for this assessment (Health Canada 2008a).

The estimated exposure to the Canadian population is based on data representing surface soil horizons, or in the first few decimetres, and not on data measured in the C horizon (primary environment; Reimann and Garrett 2005). The surficial concentrations of natural elements are, nonetheless, directly related to their concentration in the primary environment.

Some researchers have maintained that background concentrations³ should not be expressed as an absolute value but rather a range of values varying by sampling location and scale (Choinière and Beaumier 1997; Reimann and Garrett 2005). For the purposes of the present assessment, however, the concentration of aluminum in surface soil has been based on the arithmetic mean of all available data, and not based on a concentration range.

The mean total aluminum concentration in Canada is estimated to be 41,475 mg/kg⁴ . Figure 2.4 summarizes the mean total aluminum concentrations in soils by province and for Canada as a whole. The mean concentrations of total aluminum ranged from 12,000 mg/kg in Nova Scotia to 87,633 mg/kg in British Columbia. While a single estimate of aluminum concentration in soil has been calculated for the purpose of the present assessment, it is important to recognize that aluminum concentrations in soil vary extensively from one region to another.

In recent years, Health Canada initiated research in the Ottawa region comparing mean aluminum concentrations in residential gardens with concentrations in dust from inside residential dwellings. The results showed that mean aluminum concentrations were about 26,000 mg/kg inside residential dwellings, but more than double that (55,841 mg/kg) in gardens (Rasmussen et al. 2001).

Figure 2.4: Comparison of mean total aluminum concentrations in soils from provinces across Canada (mg/kg) (1987-2007)

Click on image to enlarge.

Measures of extractable and dissolved aluminum in soil

In general, unless the soil pH falls below 4, levels of the more soluble Al³⁺ form (i.e., the form considered to be more readily taken up by organisms) in the soil pore fluids are likely to be low. Hendershot and Courchesne (1991) measured aluminum in soil solution at St. Hippolyte, Quebec. The median total dissolved aluminum level was 0.570 mg/L, the median inorganic aluminum level 0.190 mg/L and the median Al³⁺ level 0.0003 mg/L in samples collected at a depth of 25 cm (pH = 5.5). Total dissolved aluminum was also measured in soil solution in the Niagara, Ontario, region; its level reached 1.214 mg/L (pH 4.2) in untreated soil. Following treatment with lime, aluminum was not detected in soil pore waters, and the pH increased to 4.8-5.5 prior to planting alfalfa (Medicago sativa L.). After three cuts of alfalfa, the pH was elevated to 6.0 in control plots and to 7.5-8.0 in limed plots; the mean total dissolved aluminum level was 0.335 mg/L in pore waters in the control plots and 0.016 to 0.397 mg/L in limed plots (Su and Evans 1996).

Turmel and Courchesne (2007) reported concentrations of 16.5 to 18.5 mg/kg dw total recoverable aluminum (from nitric acid digestion) in surface soil samples (pH 5.2) collected in 2005 from an abandoned agricultural field near a zinc plant in the Valleyfield area of Quebec. Soil collected under a nearby forest stand (pH 6.0) contained from 8.8 to 11.7 mg/kg dw total recoverable aluminum. The water soluble fraction of aluminum for the soils was 0.477 to 0.507 mg/L and 0.403 to 0.424 mg/L for the agricultural and forest soil samples, respectively.

Data relating to aluminum levels in soils treated with aluminum hydroxide sludges are limited. Near Regina, Saskatchewan, 1100 tonnes of alum sludge from a DWTP were spread on 16 ha of soil at a rate of 75 tonnes per hectare. There was no statistical difference in the mean acid-extractable aluminum level in both control (4.0%) and treated (4.1%) soil (Bergman and Boots 1997). In a study done for the American Water Works Association, Novak et al. (1995) measured the aluminum content of soil before (pH 4.7 and 5.5 at two sites) and after application of water treatment residuals. The PAC residual contained 2,330 mg Al/kg dw, and the alum residual, 6,350 mg/kg dw. In cropland soil treated according to the Mehlich III extraction procedure, which estimates the amount of aluminum available for uptake by organisms, concentrations of this available aluminum varied between 405 and 543 mg/kg dw (or 0.04% and 0.05%) before the application of the water treatment residuals. Addition of PAC and alum residuals resulted in an increase of available aluminum to 770 mg/kg dw and 1115 mg/kg dw, respectively. In another experiment, alum residual containing 150,000 mg Al/kg dw was applied to forest soil (pH 4.7). Soil analyses done 30 months later showed no differences between the control and the treatment plots for bioavailable and total aluminum.

2.3.2.5 Biota

Aluminum concentrations in vegetation related to the production or use of the aluminum salts considered in this report were available for only a few locations in Canada. Vasiloff (1991, 1992) reported aluminum levels in bur oak (Quercus macrocarpa) foliage collected from trees near an aluminum chloride producer in Sarnia, Ontario. Total aluminum levels ranged from 25 to 170 mg/kg dw in 1989 and from 57 to 395 mg/kg dw in 1991. Levels were higher in the foliage of trees closer to the aluminum chloride plant. These levels were below the Ontario Rural Upper Limit of Normal for aluminum in tree foliage (Vasiloff 1992). Fugitive emissions of aluminum chloride and subsequent hydrolysis, resulting in the formation of hydrochloric acid, were responsible for the damage to trees, including death that was observed at one location. The company ceased its operations in the mid-1990s. No such damage was reported near aluminum sulphate plants.

Novak et al. (1995) measured aluminum levels in soils before (pH 4.7 and 5.5 at two sites) and after the application of water treatment residuals (PAC and alum sludge), as well as aluminum contents in tissues of corn (Zea mays), wheat (Triticum aestivum) and loblolly pine (Pinus taeda) in control and treated soils. Statistical differences in aluminum contents were noted only in corn tissues. Aluminum levels were lower (15.1 mg/kg dw versus 18.6 to 19.6 mg/kg dw) in plants grown in soil treated with 2.5% of PAC water residual than in plants grown in soil treated with 1.34% alum or in controls; however, crop yields (kg/ha) were not lower. Aluminum levels in loblolly pine tissues were not statistically different in trees grown in control (270 mg/kg dw) and treated (152 to 170 mg/kg dw) soil.

No information was found relating concentrations in animals with aluminum entering the environment from direct production or use of the three salts subject to this assessment.

Morrissey et al. (2005) reported mean levels of 55 mg/kg dw in feathers and 2780 mg/kg dw in feces of American dippers (Cinclus mexicanus) residing in the Chilliwack watershed of British Columbia. The samples were collected over the period 1999 to 2001, and were considered to represent overall exposure to both natural and anthropogenic sources in the region. Benthic invertebrates (primarily insect larvae) and salmon fry, both key dietary items for the birds, contained mean concentrations of around 1,500 mg/kg dw and 165 mg/kg dw, respectively. Aluminum was present in all invertebrate (n = 30), fish (n = 9) and bird fecal samples (n = 14), but only 16% of the feather samples (n = 82). Based on a calculated total dietary intake (TDI) value of 26 mg/kg bw/d, derived using procedures described in CCME (1998), the researchers hypothesized that dipper populations in the region may be subject to chronic exposure effects of aluminum.

2.3.2.6 Food

Most foods, whether of plant or animal origin, contain a certain amount of aluminum originating from: (a) naturally-occurring aluminum in the soil, (b) the addition of aluminum salt-based food additives, and (c) the migration from aluminum-containing materials in contact with food (InVS-Afssa-Afssaps 2003). More than 80% of total aluminum concentrations found in foods and beverages range from 0.1 to 10 mg/kg wet weight. Some foods containing additives can exceed aluminum concentrations of 100 mg/kg⁵.

Selection of data for foods in Canada

Data on the concentrations of aluminum in Canadian foodstuffs are collected through Canadian Total Diet Studies, carried out by the Health Products and Food Branch of Health Canada, with the fifth Total Diet Study being the most recent. The Total Diet Study estimates the concentrations of more than 15 trace metals (both essential and non-essential) in foods commonly consumed by Canadians.

Estimating quantities of aluminum ingested by an individual is complicated by the fact that foods are composite materials, and the components have very different aluminum concentrations. In the Total Diet studies, foods bought in grocery stores are prepared to reflect the Canadian diet; hence raw meat is cooked, and vegetables are peeled, trimmed or otherwise cleaned for serving, if not cooked. Processed foods or mixes are prepared as directed.

While the Total Diet Study provides data on total aluminum concentrations in foods, it does not allow estimation of the proportion of naturally-occurring aluminum versus the proportion of added aluminum salts. Some qualitative information in this regard is, however, included below.

With respect to aluminum originating from the contact of food with packaging material, this source would be included in the total aluminum concentration measured in the food item in the Total Diet Study. Aluminum utensils, pots and pans are not used to prepare the food, and so this potential source is not reflected in the measured concentrations. Some information on this aspect from other studies is, however, included below.

Estimated exposure in this assessment was based on preliminary data from the first three years of the fifth Total Diet Study (2000-2002) conducted in Ottawa (2000), Saint John (2001) and Vancouver (2002) (Dabeka 2007).

Mean aluminum concentrations in Canadian foods

In Canada, some foods have naturally high total aluminum concentrations, including yeast, raisins, mollusks and shellfish as well as some spices and herbs, where concentrations greater than 400 mg/kg were found (e.g., black pepper and oregano) (Dabeka 1007). Although concentrations in some aromatic herbs and spices may be high, their overall contribution to the daily diet is very low as only small quantities are normally ingested.

Tea is frequently studied by researchers as the plant generally assimilates high concentrations of aluminum (Wu et al. 1997). The fifth Total Diet Study in Canada showed aluminum concentrations of about 4.3 mg/kg in infused tea. This can be compared to the concentrations in other beverages of 0.67 mg/kg in red wine, 0.51 mg/kg in beer and a much lower average concentration of 0.08 mg/kg in coffee (Dabeka 2007). For the Canadian data, all samples were analyzed as prepared for consumption (i.e., brewed tea and coffee).

In addition to natural aluminum in foods, aluminum-containing food additives are permitted for use as a colouring agent, firming agent, stabilizing agent, pH adjusting agent, anti-caking agent, dusting agent, emulsifier, and carrier. Specific maximum levels of use prescribed in the Canadian Food and Drugs Regulations range from 0.036% (or 360 mg/kg) for aluminum sulphate in some egg products to 3.5% (or 35,000 mg/kg) for sodium aluminum phosphate in creamed and processed cheese products (Health Canada 2004).

Table 2.5 summarizes mean total aluminum concentrations found in various food groups in Canada based on the fifth Total Diet Study performed between 2000 and 2002. Certain food groups include diverse items, such that aluminum concentrations may vary considerably within a food group. More detailed information on the concentrations in specific items is presented below and included in the supporting documentation for this assessment (Health Canada 2008a).

Cereal products are generally the primary source of dietary exposure to aluminum, followed by sugar-containing foods and dairy products. Other food categories account for less than 10% of the total aluminum dietary exposure. The mean total aluminum concentration in cereal products is a result of higher levels found in retail (ready-to-eat or mix) cakes, pancakes, muffins, Danish pastries, donuts, and cookies (concentrations ranging between 11 and 250 mg/kg). Such levels likely result from the direct addition of aluminum-based food additives, or from the use of baking powder in which aluminum-base food additves are also permitted (baking powder that is purchased in stores and used in home-cooking does not generally contain added aluminum salts.). Lower levels of aluminum are found in pasta, rice, bread, and cooked wheat, oatmeal and corn-based cereals, which are also included in the cereal products category.

Similarly, the mean aluminum concentration in the "Foods, primarily sugar" category is attributed to the level of aluminum found in chewing gum. Most food items included in that particular category such as candy, gelatine desserts, honey, jams, pudding, and syrup, contain very low levels of aluminum.

^* see text for details on specific food items in this category

Total Diet Studies in Canada have also examined various fast food products, where mean aluminum concentrations exceeding 1 mg/kg were found in french fries and pizza, and up to approximately 50 mg/kg in chicken burger (Dabeka 2007).

Two of the three salts specifically named on the PSL2 (chloride and nitrate) are not used as food additives. Aluminum sulphate (including its potassium and sodium salts) may be used as a food additive, but other aluminum-containing additives (basic and acidic sodium aluminum phosphate, sodium aluminosilicate) are much more widely used⁶ . This was confirmed through recent information gathered by Health Canada's Food Directorate from those members of the food industry who manufacture products in which aluminum-based food additives are permitted. This information indicates that aluminum sulphate (and its salts) are used as food additives in a limited number of food items, such as muffins, pizza, tortilla, burritos, egg products and some dry bakery mixes, and in quantities less than 0.5% of the final product weight.

Mean aluminum concentrations in Canadian infant formulas and in breast milk

Health Canada regularly tests infant formulas for metal concentrations as well as the water added to certain formulas as a point of comparison. Available data from the most recent Canada Total Diet Study as well as information from studies conducted by the Health Products and Food Branch are evaluated to estimate aluminum levels in bovine protein and soy-based infant formulas.

According to the fifth Canadian Total Diet Study conducted in 2000-2002, aluminum concentrations of 0.20 and 0.79 mg/kg were measured in the bovine protein and soy-based infant formulas, respectively. These concentrations were measured in the reconstituted infant formulas prepared for consumption.

Aluminum concentrations in several types of bovine protein and soy-based infant formulas were also measured in another Canadian study undertaken between 1999 and 2001 (Health Canada 2003). The mean concentrations in bovine protein formulas were about 0.13 mg/kg in liquid concentrates, 0.18 mg/kg in powdered formula to which a specified quantity of water was added and approximately 0.40 mg/kg in ready-to-use concentrates with iron added. Soy-based infant formulas had mean aluminum concentrations of approximately 0.73 mg/kg in the case of both ready-to-use concentrates and powdered formulas. Again, these concentrations were all measured in the reconstituted infant formulas prepared for consumption.

Two studies were undertaken in Canada to measure levels of aluminum in breast milk. They indicated that mean concentrations of aluminum in breast milk were of the same order of magnitude as elsewhere in the world. In one study in Quebec, which involved only five women, a mean concentration of aluminum in breast milk of 0.34 mg/kg was measured (Bergerioux and Boisvert 1979). In a second study, a median aluminum concentration of 0.014 mg/kg in 12 Albertan women was measured (Koo et al. 1988). Thus, the average concentration of aluminum in breast milk is considered to be approximately 0.11 mg/kg⁷.

Migration of aluminum from materials in contact with food

Aluminum concentrations in food generally increase when there is direct contact with aluminum packaging material or aluminum utensils, pots and pans, especially when food is cooked. Researchers have demonstrated that the migration of aluminum to food could depend on pH, container type, cooking time, purity of the aluminum used in the coating of utensils or aluminum pots, or salt addition to boiling water (Muller et al. 1993; Abercrombie and Fowler 1997; Gramiccioni et al. 1996; Gourrier-Fréry and Fréry 2004; Pennington 1988; InVS-Afssa-Afssaps 2003). For example, aluminum concentrations in coffee, soft drinks and beer increased from 0.02 mg/L to more than 0.25 mg/L when an aluminum percolator was used to brew coffee, or when soft drinks and beer were kept in aluminum cans for more than six months. A level up to 0.87 mg/L in drinks was also observed after 12 months of storage in cans (Muller et al. 1993; Abercrombie and Fowler 1997). Concentrations of up to 35 mg/L were found in acidified fruit juices after boiling in an aluminum pot (Liukkonen-Lilja and Piepponen 1992).

With respect to the uses of the three salts -- aluminum chloride, aluminum nitrate and aluminum sulphate -- in food packaging, aluminum sulphate is used as a component in metalized films and aluminum chloride is used as a component in a wax product that is applied as a coating on plastic films. While both of these products may be used in food packaging, the estimated amount of aluminum migrating from these films into the food would be negligible (Health Canada 2008b).

2.3.2.7 Consumer products

2.3.2.7.1 Non-prescription drugs

The major pharmaceutical uses of aluminum are: as an antacid and as phosphate binder for patients with chronic kidney failure (aluminum hydroxide); as a component of the prescription antiulcer medication, sucralfate (sucrose sulfate-aluminum complex), as a component in some vaccines and injections (e.g., alum precipitated allergen extracts, MMR vaccine) (see section 2.3.2.8), as a hemostatic agent to control bleeding from minor cuts (aluminum potassium sulfate (alum), aluminum chloride or aluminum sulfate), as a component in hydrated magnesium aluminum silicate in the antidiarrheal, attapulgite, and as astringents (there are numerous aluminum derivatives in antiperspirants and in some deodorants). Aluminum containing antacids, represent, by far, the largest potential exposure to aluminum in individuals consuming these drug products on a regular, prolonged, basis.

Concentrations of aluminum compounds in over-the-counter products sold in Canada were obtained from the Health Canada Drug Product Database⁸ . The Drug Product Database contains brand name, Drug Identification Number (DIN), ingredient and other information for approximately 23,000 drugs approved for use in Canada. Based on the concentrations of specific aluminum compounds, the elemental aluminum contents of orally administered over-the-counter products marketed in Canada are estimated to be 8,700 to 60,000 mg/kg for antacids (heartburn medication), 30,000 to 50,000 mg/kg for dental agents, and 3,500 mg/kg for attapulgite⁹.

2.3.2.7.2 Cosmetics

Compounds such as aluminum chlorohydrate, ammonium aluminum sulphate, aluminum hydroxide, aluminum starch octenylsuccinate, aluminum-based dyes and aluminum silicate are used in deodorants, antiwrinkle preparations, toothpastes, eye and face makeup, shampoo, lipstick, moisturizers and other cosmetic products sold in Canada. Data on concentrations of aluminum compounds in these products are available through Health Canada's Cosmetic Notification System, a mandatory system under which manufacturers must submit information including composition data on cosmetics prior to first sale in Canada.

Table 2.6 presents reported ranges of aluminum concentrations that may be contained in a wide variety of cosmetic products sold in Canada. However, it should be noted that the data on concentrations are available with respect to reporting categories (< 0.1%, 0.1% to 0.3%, 0.3% to 1.0%, 1% to 3%, 3% to 10%, 10% to 30% and 30% to 100%). Thus the maximum concentration represents an upper limit of a reporting category, and is therefore possibly an overestimate, by a factor of up to 3.3, of the actual maximum concentration in the product category.

^* Note that the maximum concentration corresponds to an upper limit for a reporting category (see text) and may thereby overestimate the maximum concentration by up to a factor of 3.3)
^** Maximum upper bound not available, as the upper limit of reporting category is 100%

2.3.2.8 Vaccines

Most of the vaccines authorized in Canada contain an aluminum salt adjuvant, according to the systematic vaccination schedule used for infants, young children, adolescents and adults (Canada Public Health Agency 2006). Various types of vaccine adjuvants are used by pharmaceutical companies, such as aluminum hydroxide, aluminum phosphate, aluminum sulphate and aluminum potassium sulphate. The quantity of aluminum ranges between 125 µg and 1,000 µg (aluminum hydroxide) per dose, depending on the vaccine. There is no standard or recommendation available in Canada with respect to the maximum quantity of aluminum or aluminum compound that may be used as an adjuvant in vaccines.

2.3.3 Toxicokinetics: human and experimental animals

An overview of the toxicokinetic processes of aluminum was carried out with the goal of highlighting the various factors influencing its pathway from the environment to target organs. Each toxicokinetic process is described below (absorption, distribution and elimination). Aluminum does not undergo phase I and II biotransformation reactions, occurring only in the +3 oxidation state. The metabolism of aluminum is therefore described in relation to its speciation, in the context of the distribution and elimination processes.

2.3.3.1 Absorption

Even at moderately elevated levels in the environment, exposure of aluminum leads to only small increases of aluminum in human tissues due to its low bioavailability through all routes of exposure.Bioavailability refers to the fraction of the total amount of the substance ingested, inhaled or in contact with the skin that reaches the systemic circulation. In this assessment, emphasis is placed on oral bioavailability, as the estimated daily intake (EDI) of the Canadian population shows that ingestion is the major route of exposure (see section 3.2.1); the bioavailability of aluminum with respect to other exposure routes (inhalation and dermal) is also reviewed. Bioavailability estimates for all exposure routes have been summarized in Table 2.7.

2.3.3.1.1 Oral absorption

The interpretation of aluminum oral bioavailability estimates requires the understanding of: (a) the methods used to calculate oral bioavailability, and (b) the physiological and biochemical factors that influence oral absorption. Since the ingested matrix to which aluminum is bound likely influences its potential absorption, the oral bioavailabilities of aluminum measured in drinking water, food and soil are distinguished.

Methods to calculate oral bioavailability

The methods to calculate the oral bioavailability in experimental studies are: (a) mass balance based on intake, and fecal and urinary excretion; (b) comparison of intake with urinary excretion; (c) concentration in a single blood sample and a calculated volume of distribution; (d) aluminum concentration in tissue; and (e) comparison of areas under the plasma concentration-time curve after oral and intravenous administration (Yokel and McNamara 2000). The most common method is comparison of intake with urinary excretion. This method is the simplest and least invasive, and is relatively reliable provided that the collection period is long enough to measure nearly all the aluminum excreted in the urine.

Prior to 1990, aluminum analyses were based on the quantification of the common isotope ²⁷Al (≈100% of the natural isotopes). As ²⁷Al in the environment is ubiquitous, contamination during sampling and analysis may easily occur, leading to overestimation of the tissue concentrations, particularly when the administered amounts of aluminum are near the baseline exposure. The relative contribution from endogenous ²⁷Al is minimized by administering doses that are much higher than the levels encountered in the environment. However, oral absorption may depend on dose. Thus, this approach increases the uncertainty in the estimation of bioavailability of environmental concentrations of aluminum. On this point, the observed relationship between dose and bioavailability is inconsistent: increased dose of aluminum decreased its bioavailability in the experimental studies of Greger and Baier (1983), Weberg and Berstad (1986), and Cunat et al. (2000) while opposite results were observed in other animal studies (Yokel and McNamara 1985; Ittel et al. 1993).

In recent years accelerator mass spectrometry (AMS) has been used to quantify the isotope ²⁶Al, administered as a tracer (Priest 2004). This analytical technique has allowed researchers to more accurately measure bioavailability of aluminum at levels comparable to the levels to which the general population is actually exposed, since it is possible to distinguish the aluminum in the administered dose (²⁶Al) from the aluminum already in the body (²⁷Al). However the cost and small number of facilities limit the sample analyses, which can result in the diminishing of the precision of the estimation and the information concerning the intra-individual variability (Yokel and McNamara 2000).

Factors influencing oral absorption

The principal mechanism of absorption of ingested aluminum seems to be a passive diffusion through the paracellular pathway (Zhou and Yokel 2005). This diffusion occurs predominantly in the small intestine (duodenum and jejunum) and, to a lesser extent, through the gastric mucosa in stomach (Powell and Thompson 1993; Walton et al. 1994). In addition to passive diffusion, Cunat et al. (2000) suggested that absorption of aluminum may occur by a transcellular and saturable route, which may explain the possible dependency of absorption on the dose level.

The rate of uptake, and consequently the cumulative absorption of aluminum, has been shown to vary depending on physiological and chemical factors. Krewski et al. (2007) summarized factors based on findings in both human and animal studies, including:

• Solubility: absorption is greater with more soluble aluminum compounds;
• Gastric pH: absorption is greater at pH 4 compared to pH 7, probably due to the generation of more soluble aluminum compounds;
• Carboxylic acids: increased absorption in the presence of carboxylic acids, particularly citrate that is naturally present in many foods and fruit juices;
• Silicon compounds: decreased absorption in the presence of silicon-containing compounds in the dietary intake, due to a possible formation of hydroxyaluminosilicate.

Among the factors cited above, particular attention has been given to the significant impact of citrate during the ingestion of aluminum. Oral bioavailability has been found to increase by a factor of 5 to 150 when aluminum is ingested with citrate solution, as verified with studies employing the same aluminum complex and under the same experimental conditions (Weberg and Berstad 1986; Yokel and McNamara 1988; Froment et al. 1989; Priest et al. 1996; Drueke et al. 1997; Schönholzer et al. 1997). Citrate probably facilitates the absorption by opening the tight junction between intestinal cells (Froment et al. 1989; Zhou and Yokel 2005). Zhou et al. (2008) recently explored the influence of citrate in drinking water at a similar molar concentration to aluminum. The researchers did not observe a significant enhancement of aluminum absorption for an Al:citrate molar ratio of 1:1, and suggested that aluminum absorption may depend on citrate dose.

The principal biochemical explanation for how the factors listed above influence absorption is the nature of the ligand to which the ion Al³⁺ is associated in the gastrointestinal fluid. In vitro studies using Caco-2 cells derived from the human lower intestine show differences between ligands in the uptake rate of aluminum; aluminum citrate and aluminum nitrilotriacetate were absorbed more rapidly than aluminum lactate (Alvarez-Hernandez et al. 1994) and the uptake rate of aluminum fluoride was higher than that of, in decreasing order, Al³⁺, aluminum maltolate, aluminum citrate and aluminum hydroxide (Zhou and Yokel 2005). Results from in vivostudies provided evidence for significant differences in the oral bioavailability calculated for different ingested aluminum complexes (Yokel and McNamara 1988; Froment et al. 1989). Cunat et al. (2000) concluded that the organic ligands enhance aluminum absorption, in comparison to the inorganic ligands (citrate > tartrate, gluconate, lactate > glutamate, chloride, sulphate, nitrate), based on the results of a study in which rat intestines were locally perfused with aluminum.

The pH of the exposure media may play an important role in the absorption of aluminum, as it affects aluminum speciation. In aluminum sulphate-treated water with low pH, the aluminum sulphate and Al³⁺ (very soluble) are the predominant complexes while, when increasing the pH from 6.3 to 7.8, the predominant complex is aluminum hydroxide (likely insoluble). At pH above 7.8, the solubility in water increased due to the presence of the negative ions of aluminum hydroxyl (Walton et al. 1994). As mentioned in section 2.3.2.2.2, while treatment with aluminum sulphate may reduce the total aluminum concentration in finished water as compared to the untreated water source, through the removal of suspended solids containing aluminum, there is evidence that treatment with aluminum salts also increases the concentration of low-molecular-weight, dissolved aluminum species (Health Canada 1998b).

The low pH of the gastric fluid creates a high potential for transformation of the ingested aluminum complex. This led Reiber et al. (1995) to argue that the aluminum in drinking water would not be more readily assimilated than other forms of aluminum, and that regardless of the form in which the aluminum is consumed, a substantial portion of it will likely be solubilized to monomolecular aluminum in the stomach. Other researchers, however, consider this to be an oversimplication, in light of the observed differences in the oral absorption of different aluminum compounds (Krewski et al. 2007).

Concurrent absorption of aluminum with other dietary nutrients has been shown to influence the intestinal absorption of this metal. For example, the presence of vitamin D likely favours the absorption of aluminum (Adler and Berlyne 1985; Ittel et al. 1988; Long et al. 1991; Long et al. 1994) and the consumption of folic acid supplementation is expected to diminish aluminum absorption and/or its accumulation in various organs (bone, kidney and brain) by a possible formation of folate-Al complex (Baydar et al. 2005). Domingo et al. (1993) investigated the effects of various dietary constituents, such as lactic, malic and succinic acids, on the levels of absorption and distribution of aluminum in drinking water and in the diet of mice, where they observed an enhanced absorption with these concurrent ingestions.

A few studies have been conducted to examine whether food composition or the presence of food in the stomach affect oral aluminum bioavailability, and the results have been mixed. The nature of the contents in the stomach influenced the absorption of aluminum in the study of Walton et al. (1994) in which adult Wistar rats were exposed to water treated with aluminum sulphate along with various beverages and foods. The aluminum concentrations in serum increased when the aluminum sulphate treated drinking water was taken with orange juice; the same phenomenon was observed, but to a lesser extent, with coffee. The authors note that the low levels of aluminum in these two beverages would not have contributed to this increase in aluminum levels. In comparison, when aluminum sulphate treated water was given with beer, tea or cola (beverages that may contain appreciable levels of aluminum) the serum concentration did not markedly rise. Meat and carbohydrate/cereal products decreased aluminum absorption. Drüeke et al. (1997) performed a study in rats using ²⁶Al to examine the effect of silicon contained in drinking water as well as solid food, on the absorption of aluminum. In their study, high Si concentrations in the drinking water failed to depress the ²⁶Al fraction absorbed, as estimated on the basis of skeletal accumulation and urinary excretion. In addition, absorption of ²⁶Al was approximately 15 times higher in the fasted state than in the non-fasted state. As part of a study conducted in rats with ²⁶Al, Yokel et al. (2001a) tested the hypothesis that the stomach contents affect aluminum absorption. According to the authors, although stomach contents delayed aluminum absorption, it did not significantly alter the extent of ²⁶Al absorption.

Estimation of the oral bioavailability of aluminum in drinking water

Experimental data for oral bioavailability of aluminum from drinking water, obtained in studies conducted in humans and animals, and based on varying calculation and quantification methods, were evaluated.

The compilation of central values (mean or median) of the results of different studies in humans results in a range of 0.010% to 0.52% for oral bioavailability of aluminum in drinking water, based on experiments involving more than one volunteer. The lower value is the mean value obtained from the data of two volunteers exposed to ²⁶Al-hydroxide in Priest et al. (1998). This experimental study observed the higher value of 0.52% as well when these two volunteers were exposed to ²⁶Al-citrate. In a much larger study with 29 subjects consuming an aluminum-controlled diet, the oral bioavailability from aluminum sulphate-treated municipal drinking water was estimated at 0.36% to 0.39% (Stauber et al. 1999).

As for the central values for the oral bioavailability for experimental animals, a range of 0.04% to 5.1% is reported in experimental studies with the isotope ²⁶Al, whereas the range based on ²⁷Al is 0.01% to 4.56%. The maximum central value of 5.1% for the animal experiments using ²⁶Al was obtained following ingestion of a concentrated solution of citrate (Schönholzer et al. 1997). The second highest value is 0.97%, based on the exposure to aluminum chloride (Zafar et al. 1997). The maximum central value of 4.56% for ²⁷Al was obtained for aluminum citrate ingested by rats with renal failure (Yokel and McNamara 1988). If only healthy animals had been considered, the maximum value would have been 2.18% for ²⁷Al-citrate.

Krewski et al. (2007) proposed a range for the oral bioavailability of aluminum in drinking water of 0.05% to 0.4% for rats and rabbits, and 0.1% to 0.5% for humans, with a most likely value of 0.3%. The approximate correspondence between the ranges and the most likely estimates in humans and animals for bioavailability from drinking water suggests that there is little interspecies difference in this respect.

Estimation of the oral bioavailability of aluminum in food

In spite of the important contribution of food in the total exposure to aluminum, the database for oral bioavailability of aluminum in food is limited. In an early investigation into the potential for the absorption of aluminum accumulated in food, Jones (1938) demonstrated that a large percentage of aluminum in bread made with aluminum-based baking powder was soluble in the gastric juice of dogs. Several decades later, Yokel and Florence (2006) confirmed that some aluminum from biscuits made with baking powder containing acidic ²⁶Al-sodium aluminum phosphate (SALP) reaches the systemic circulation. In this study, about 0.12% of the ingested aluminum crossed the gastrointestinal tract of exposed rats. Using the same experimental method¹⁰, Yokel et al. (2008) estimated oral bioavailabilities of ~0.1% and ~0.3% for basic ²⁶Al-SALP incorporated into cheese at concentrations of 1.5% and 3%, respectively.

The oral bioaccessibility¹¹ of aluminum encountered in different foods was measured by Lopez et al. (2002) and Owen et al. (1994). It is not possible, however, to directly compare their results, since their methodologies differed. Moreover, the bioaccessibility estimates, ranging from 0.3% to 0.9% by Owen et al. (1994) and 0.85 to 2.15% by Lopez et al (2002), cannot be directly used to estimate the oral bioavailability of aluminum, as the in vitro-in vivo relationship has not been established (Ruby et al. 1999). Nonetheless these bioaccessibility studies do provide evidence that oral bioavailability may change according to the nature of consumed foods. For example, the aluminum in bread, jam and tea appeared to be about 2.7 times more soluble than the aluminum in sponge cake (Owen et al. 1994). It is expected that the actual oral bioavailability of aluminum in food is lower than these bioaccessibility values, as solubility in the intestinal tract would not be the only factor limiting absorption.

The oral bioavailability of aluminum in food has also been estimated based on the comparison of aluminum intake in the general population with the urinary excretion and/or the body burden of aluminum (Ganrot 1986; Priest 1993, 2004; Powell and Thompson 1993; Nieboer et al. 1995). These estimates range from 0.1% to 0.8%. Note that the oral bioavailability estimate of 0.12% of Yokel (2006) for rats fed aluminum-containing biscuits falls in this range, as does the estimate of 0.53% by Stauber et al. (1999), based on a controlled diet in humans.

Bioavailability of aluminum in antacids (aluminum hydroxide) has been estimated in three studies in humans, measured alone or in combination with citrate, orange juice, bicarbonate, or calcium acetate (Mauro et al. 2001; Haram et al. 1987; Weberg and Berstad 1986). These measured bioavailabilities, ranging from 0.001% to 0.2% were generally comparable to the bioavailabilities measured in food.

The limited data concerning the oral bioavailability of aluminum in food do not allow for the determination, with good predictive value, of the potential absorption of aluminum in food. For the purpose of comparison with other media (Table 2.7), the interval of 0.1% to 0.8% is retained, with a most likely range of 0.1% to 0.3%, based on the recent work of Yokel and Florence (2006) and Yokel et al. (2008).

Estimation of the oral bioavailability of aluminum in soils

Another factor of importance in the human health risk assessment for aluminum is the oral bioavailability of aluminum in ingested soil, as soil ingestion is a significant exposure pathway for the toddler group (see section 3.2.1). No bioavailability data on soil were identified. Limited data, however, were found for the bioaccessibility of aluminum in soil, which, as noted above, is an in vitro measure of the soluble fraction of the substance available for absorption.

Shock et al. (2007) estimated the bioaccessibility of aluminum in different tundra soil samples contaminated by mining waste dust, by simulating gastric fluid in an in vitro experiment. The estimated values varied from 0.31% to 4.0%, according to the grain size and to the solid:fluid ratios used in the experiment. As expected, aluminum in the soil with lower grain size was most available for absorption.

As is the case for the bioaccessibility data of aluminum in food, these bioaccessibility estimates for aluminum in soil need to be tied to the in vivo bioavailability estimates from appropriatein vivo models (Ruby et al. 1999). Even if the experimental protocols used to measure food and soil aluminum bioaccessibility differed slightly, the data of Shock et al. (2007) suggest that the bioaccessibility of aluminum in soil is similar to that in food. In the absence of more relevant data, the range for the oral bioavailability of aluminum in soil is therefore assumed to be similar or less than that of food. The oral bioavailability of aluminum in soil is considered to be a major source of uncertainty for this exposure pathway.

2.3.3.1.2 Dermal absorption

Utilization of antiperspirant with aluminum would contribute to the body burden if aluminum passes through the skin barrier. There is some evidence from case studies, described below, that small amounts of aluminum do reach the systemic circulation. However, to date, no data for dermal bioavailability are available from controlled studies of more than one or two individuals.

In the study of Flarend et al. (2001), ²⁶Al-chlorohydrate (aluminum complex in antiperspirant) was applied to a single underarm of one man and one woman. The cumulative urinary excretion after 43 days following the application accounted for 0.0082% (male) and for 0.016% (female) of the applied dose. After correcting this fraction for the aluminum not excreted in urine (15% of the absorbed dose), this application was estimated to result in a dermal bioavailability of about 0.012%. On the basis of these data, the authors estimated that the amount of aluminum absorbed from regular use would be 0.25µg/d.

Guillard et al. (2004) reported on one clinical case in which a woman who used an antiperspirant cream with aluminum chlorohydrate over four years showed elevated levels of aluminum in plasma and urine (10.47µg/dL in plasma)¹². When the woman discontinued use, concentrations in her urine and plasma dropped to reported normal values after the third and eighth months, respectively.

2.3.3.1.3 Inhalation absorption

The ambient air of multiple occupational environments, such as the aluminum production industry and welders' factory (Priest 2004), may have high levels of aluminum. The higher urinary excretion of aluminum in exposed workers, compared to the general population, demonstrates that some inhaled aluminum can reach the systemic circulation (Sjogren et al. 1985; Sjogren et al. 1988; Pierre et al. 1995). This absorption depends on the form of aluminum in the ambient air (adsorbed to PM, vapour condensation fumes and flakes) and, in the case of particulate matter, also depends on the distribution of the sizes of the aerodynamic diameter of PM (PM_2.5 versus PM₁₀).

Priest (2004) estimated a deposited pulmonary fraction of 1.9% in a study of two volunteers who inhaled ²⁶Al-oxide adsorbed to particles with a mass median aerodynamic diameter (MMAD) of 1.2 µm. The last value is supported by animal studies showing a deposition of fly ash of aluminum into the lungs from 2% to 12% (Krewski et al. 2007). As well, Yokel and McNamara (2001) have proposed an absorption fraction of about 1.5% to 2%, on the basis of the relationship between the urinary excretion of aluminum-exposed workers and the concentrations of airborne soluble aluminum measured in their environment.

An investigation in New Zealand rabbits exposed via the nasal-olfactory pathway (sponge soaked in aluminum solutions inserted into nasal recess for four weeks) provided evidence that inhaled aluminum in the olfactory tract can cross the nasal epithelium to reach the brain directly through axonal transport (Perl and Good 1987). While an analytical protocol for quantifying the amount of aluminum transported along this pathway under environmental exposure conditions has been described (Divine et al. 1999), further experimental work is required to document transport of aluminum via this pathway to the olfactory bulb, and subsequently to other regions of the brain.

2.3.3.1.4 Parenteral administration

Intravenous injection of aluminum-containing products (e.g., intravenous feeding solutions) results in complete availability of the aluminum to the systemic circulation (Yokel and McNamara 2001; Priest 2004). In the case of intramuscular injection of aluminum species (e.g., via vaccination), potentially all of the aluminum injected may be absorbed into the bloodstream. However, the uptake rate from the muscle to blood circulation differs according to the aluminum complex. Evidence of this was provided in an experimental study, in which rabbits were injected with ²⁶Al-hydroxide and ²⁶Al-phosphate, two common vaccine adjuvants, at standard dose levels. After 28 days, 17 % of the aluminum hydroxide and 51% of the aluminum phosphate were absorbed (Flarend et al. 1997). The authors estimate that this dose, when administered in humans, would represent an increase of 0.4 µg/dL in plasma (see section 2.3.3.2 on distribution, for estimates of normal plasma concentrations).

2.3.3.1.5 Summary of estimates of aluminum bioavailability

The estimates of aluminum bioavailability presented for the different exposure routes in sections 2.3.3.1.1 to 2.3.3.1.4 are summarized in Table 2.7. The information available to generate these estimates varies considerably depending on the exposure route, and should be considered in any application of these estimates in risk assessment.

(a) Ranges based on a compilation of the central values of estimates of the oral bioavailability of aluminum from drinking water, obtained in numerous experimental studies conducted in humans and animals. Proposed likely estimate based on experimental work of Stauber et al. (1999) in humans and the critical review of experimental animal data in Krewski et al. (2007).
(b) Based on comparisons of estimates of aluminum intake and urinary excretion in humans and experimental animal data. The estimate of bioavailablility of aluminum in food is associated with greater uncertainty than that of drinking water, because of the limitations of the database. Proposed likely range based on Yokel and Florence (2006) and Yokel et al. (2008).
(c) Based on human data reported in three studies for the bioavailability of aluminum in antacids alone or in combination with citrate, orange juice, bicarbonate, or calcium acetate.
(d) Assumed to be similar to that in food as a default value in the absence of bioavailability data from soil ingestion; considered to be of low predictive value.
(e) Based on experimental results reported in one study following a dermal exposure in two individuals.
(f) Proposed absorption fraction by Yokel and McNamara (2001) on the basis of the results from two studies in aluminum-exposed workers.
(g) Includes both intravenous (IV) and intramuscular (IM) injection.

2.3.3.1.6 Integrating bioavailability in human health risk assessment

As discussed in previous sections, the generally low oral absorption of aluminum (< 1%) is well recognized. Nonetheless, there is considerable uncertainty associated with differences in oral bioavailability, in relation to:

• the bioavailability of aluminum in different environmental media (soil, different types of food, drinking water, air, dermal application);
• the bioavailability of aluminum in humans versus experimental animal species;
• the influence of dose and dosing regime (bolus dose versus repeated exposure via drinking water or food).

The characterization of human health risks consists of a comparison of exposure in humans via different media to a critical exposure level, most often documented in experimental animal studies. In this characterization, relative bioavailability rather than absolute bioavailability is the parameter of greatest interest. Relative bioavailability for a substance may, for example, refer to the ratio of absorbed fractions via two different exposure pathways, or it may refer to the ratio of total absorption by humans (all pathways considered) as compared to the total absorption in experimental animals in the critical study or studies.

Relative bioavailability can be established by directly measuring two absorption fractions and taking the ratio of the two, or potentially indirectly through the measurement of in vitro bioaccessibility and then by comparing in vitro bioaccessibilities (e.g., the fraction of a substance that is extracted through a weak acid solution simulating gastric fluid). In the case of aluminum, bioaccessibility would considerably overestimate bioavailability, as the available evidence indicates that only a fraction of the species dissolved in the stomach is eventually absorbed. However, to the extent that bioaccessibility is proportional to bioavailability, relative bioaccessibility will be approximately equivalent to relative bioavailability.

In the previous sections, experimental data were reviewed with respect to both bioavailability and bioaccessibility of aluminum salts in various media, in humans, and experimental animals. The discussion that follows reconsiders these data from the perspective of relative bioavailability.

The most comprehensive data concerns the bioavailability of aluminum dissolved in drinking water, as measured in both human and animal studies. In humans, measurements of oral absorption of aluminum (citrate, chloride, hydroxide or lactate complexes) generally varies between 0.01% and 0.65%, while in experimental animals the range of reported values is 0.01% to 5.1%. The ranges largely overlap and do not provide evidence for differences between humans and animals in the bioavailability of aluminum in drinking water. The proposed likely estimate for aluminum bioavailability in both humans and animals is 0.3%.

The data on bioavailability of aluminum in food are much more limited, both for humans and animals. Section 2.3.3.1.1 proposes a range of 0.1% to 0.8% for the bioavailability of aluminum salts in food (humans) and 0.02 to 0.3 in animals. These ranges have a high level of uncertainty because of the limited database, but do not provide evidence for differences between humans and animals in the bioavailability of aluminum in food.

The bioaccessibilities of aluminum in soil and food were also compared in section 2.3.3.1.1. These very limited data do not provide evidence for a difference in the amount of aluminum available for absorption of aluminum from these two media, and hence do provide a basis for concluding that there are differences in bioavailability between soil and food.

In comparing the bioavailability of aluminum in drinking water and food, in both animals and humans, the ranges of experimental values largely overlap, and the proposed likely value for drinking water is at the upper end of the proposed likely range for food. Thus the available data are insufficient for identifying a difference in bioavailability of aluminum in drinking water and food.

With regard to inhalation absorption of aluminum, there is again significant variability in the available data. These data do indicate that the bioavailability of aluminum from inhalation may be higher than from the oral route; however, since the concentrations of aluminum in ambient and indoor air are low, the absorption factor for the inhalation route would not significantly influence the evaluation of cumulative exposure from soil, air, drinking water, and food.

Although dermal absorption of aluminum salts is thought to be very low, the data is extremely limited (confined to two studies), each involving one or two individuals (see section 2.3.3.1.2). Therefore, no definitive conclusions can be drawn with respect to its relative bioavailability, although the information available suggests that it is lower than for other routes of exposure.

Consideration of bioavailability may considerably influence the conclusions of human health risk characterization if relative bioavailabilities for different salts, different exposure media and different species are greater than or less than one. In this assessment, however, the limited available data did not provide evidence for relative oral bioavailabilities significantly different from one, either with respect to comparisons of humans and experimental animals, or with respect to comparisons of water, food and soil. The bioavailability via inhalation, which is higher than oral bioavailability, would not significantly influence the estimated absorbed dose, because of the low estimated concentrations of aluminum in ambient and indoor air. Dermal exposure, which appears to be associated with a very low absorption, was considered only qualitatively in this assessment. For these reasons, the estimated values of bioavailability for different media were not explicitly integrated into the human health risk characterization present in this Draft Assessment (see section 3.2.4).

2.3.3.2 Distribution

Once absorbed into the systemic circulation, much of Al³⁺ is readily associated at the binding sites of transferrin (Tf), the plasma protein for iron transport. Since, under normal conditions, Tf in blood is only one-third saturated with iron, binding sites for the absorbed aluminum are available (Harris et al. 1996). Consequently, the Al-Tf complex is the predominant aluminum species in plasma, accounting for approximately 91% of the total aluminum in plasma (7% to 8% of aluminum is associated with citrate and less than 1% with phosphate and hydroxide) (Martin 1996). As well, Day et al. (1994) reported that, one hour after the ingestion of ²⁶Al-citrate, 99% of the ²⁶Al in blood was measured in plasma of which 80% was bounded to Tf, 10% to albumin and 5% to proteins having low molecular weight; after 880 days, 86% of aluminum in blood was bounded to plasma proteins (mostly to Tf) and the rest was associated with erythrocytes.

The major physiological compartment of aluminum is the skeleton. Krewski et al. (2007) suggest that approximately 58%, 26%, 11%, 3%, 0.95%, 0.3%, 0.25% and 0.2% of the aluminum body burden would be in the bone, lung, muscle, liver, brain, heart, kidney and spleen, respectively. Aluminum measured in the lungs may reflect deposition of airborne particles. In addition, a significant amount of aluminum analyzed in skin may result from unabsorbed aluminum deposited on skin surface (Priest 2004).

The transport of aluminum into the body and its deposition into the tissues and organs have been shown to vary widely (Priest 2004). This variability, yielding different aluminum concentrations in tissues and organs, can be explained by some of the same factors influencing aluminum absorption. For example, the presence of citrate seems to enhance the distribution of aluminum into the tissue before being associated with Tf (Quartley et al. 1993; Maitani et al. 1994). According to Jouhanneau et al. (1997), the concomitant ingestion of citrate increases aluminum absorption, but does not appear to modify the relative distribution of ²⁶Al in bone, brain and liver in comparison with ingestion without citrate.

Experimental studies have reported volumes of distribution (Vd) for aluminum, describing its potential to be distributed in tissues and organs. Most of these studies suggested that the initial Vd is approximately the blood volume (Krewski et al. 2007). However, longer collection periods lead to higher Vd, indicating a possible dependency between elimination rate and blood concentrations of aluminum (Krewski et al. 2007) (see section 2.3.3.3). Calculating the oral bioavailability of aluminum using blood volume, instead of Vd, may consequently lead to an underestimation (see section 2.3.3.1).

As neurological and reproductive/developmental endpoints are of greatest concern with respect to the environmental exposures evaluated in this assessment (see section 3.2.3.2), particular attention is paid to the distribution processes leading to accumulation in the brain and in the foetus. As well, aluminum retention in bone was investigated, as it plays an important role in the kinetics of aluminum. The principal observations with regard to retention in these tissues as well as measurements of plasma aluminum levels are briefly described below.

Plasma

In a review of blood aluminum concentrations for healthy individuals, plasma or serum measurements varying between 0.19 and 1.02µg/dL in 11 studies were reported (Nieboer et al. 1995). However, according to the authors, potential problems of controlling contamination and analytical sensitivity influenced the estimates of earlier reports such that the true value more likely lies in the range of 0.11 to 0.32 µg/dL (0.04 to 0.12 µmol/L).Valkonen and Aitio (1997) reported a mean aluminum concentration of 0.16 µg/dL (0.06 µmol/L) in the serum of a healthy, non-exposed population (n = 44) who did not use antacid drugs. In another study, the mean level of aluminum in serum in 18 healthy subjects not using aluminum-containing medicines was 0.099 µg/dL (Razniewska and Trzcinka-Ochocka 2003). Liao et al. (2004) reported blood aluminum levels in workers from three optoelectronic companies in Taiwan, China. The median aluminum concentration measured was 0.36 µg/dL in the exposed workers (n = 103) and 0.32 µg/dL in the non-exposed office workers (n = 67). Higher levels of aluminum were found in aluminum welders, with mean plasma aluminum levels of 1.25 to 1.39 µg/dL (pre-shift) and 1.48 to 1.86 µg/dL plasma (post-shift) (Kiesswetter et al. 2007).

Some data on measured serum aluminum levels in animals exposed only through the normal laboratory diet were identified. Kohila et al. (2004), Johnson et al. (1992), Gonzalez-Munoz et al. (2008) and Kaneko et al. (2004) reported values ranging from approximately 0.15 to 0.66µg/dL in different strains of rats and mice. Note that some variation in serum levels would be due to the high variability in aluminum concentration in different brands and lots of laboratory chow.

No studies were identified in which both animal and human serum levels were compared within a single study, using the same analytical methodology. The aluminum content of the standard laboratory animal diet is significantly higher than that of the typical human diet, however, so it would not be unexpected that serum aluminum concentrations observed in laboratory animals would be generally higher than reported levels in humans.

Bone

Bone exhibits more affinity to aluminum than does the brain; for example the aluminum concentrations in bone are about five-fold greater than those in the brain after repeated exposure in rats and rabbits (DuVal et al. 1986; Fiejka et al. 1996; Garbossa et al. 1998). However, the slower elimination of aluminum from the brain, as compared to bone, may be attributed in part to the bone-cell turnover and the lack of neuron turnover (Krewski et al. 2007).

In general, aluminum in bone is principally captured in the mineralization front and in the osteoid (Boyce et al. 1981; Cournot-Witmer et al. 1981; Ott et al. 1982; Schmidt et al. 1984). There are three probable mechanisms of aluminum deposition in bone that govern the elimination rate of aluminum in this matrix (Priest 2004). First, aluminum can be attached to the bone surface by heterionic exchange with calcium; this aluminum can be easily released to the fluids close to the bone surface, and then bound to Tf. Second, aluminum can be incorporated into the structure of the developing hydroxyapatite crystal during the formation of the mineral lattice; this strongly binds the molecule to bone cells and there is little subsequent release of aluminum from the bone matrix. Third, aluminum can be complexed to organic components at the surface of bone; in this case, the migration of aluminum through its deposition at the mineralization front can occur, leading to a slow turnover.

Brain

The concentrations measured in the brains of exposed rats ranged from 0.0006% to 0.009% of aluminum administered dose per gram of brain, after intravenous or intraperitoneal injection (Krewski et al. 2007). It was suggested that 90% of the aluminum in brain is associated with citrate, 5% with hydroxide, 4% with Tf and 1% with phosphate (Yokel 2001). In humans, the aluminum accumulation is higher in the cerebral cortex and hippocampus than in other brain structures (Gupta et al. 2005).

There are two ways by which aluminum can reach the central nervous system, either through the blood-brain barrier or through the choroid plexus in the cerebrospinal fluid of the cerebral ventricles. Although there is some evidence that aluminum crosses the blood-brain barrier by Tf-receptor mediated endocytosis of the Al-Tf complexes (Roskams and Connor 1990), other mechanisms of uptake, independent of Tf, may be involved as well (Yokel and McNamara 1988; Allen et al. 1995; Radunovic et al. 1997), such as diffusion of the low molecular weight aluminum species or other carrier-mediated processes. In addition, aluminum may reach the brain through the nasal epithelium by axonal transport (Perl and Good 1987; Zatta et al. 1993), although the potential magnitude of this pathway has not been quantified. Axonal transport, however, would not be expected to contribute significantly to exposure in the general population due to the low concentration of aluminum in ambient air, outside of particular occupational settings (see section 2.3.2.1).

The transport of aluminum out of the brain seems to occur by its association with citrate (Yokel 2000). The ability to remove aluminum from the brain is low (Krewski et al. 2007). For instance, in a study in which ²⁶Al-Tf was administered intravenously in rats, Yokel et al. (2001b) reported that brain concentrations of aluminum did not significantly decrease 128 days after administration.

Placenta and foetus

Aluminum distributes to the placenta and foetus, as has been demonstrated by experimental studies in which aluminum was administered by different routes to rabbits, mice and guinea pigs during gestation (Yokel 1985; Cranmer et al. 1986; Golub et al. 1996b; Yumoto et al. 2000). Yumoto et al. (2000) estimated that approximately 0.2% of the subcutaneous injected dose of ²⁶Al-chloride was transferred to the foetus as well as to the placenta. In the study of Cranmer et al. (1986), fetal aluminum content was significantly increased following both intraperitoneal and oral administration, although the increase was greater with intraperitoneal dosing. No study investigating the level of aluminum in the human placenta was identified.

Milk

Aluminum is efficiently transferred from blood to milk in exposed lactating animals (Yokel and McNamara 1985; Muller et al. 1992; Yumoto et al. 2000) as well as in human lactating mothers (see section 2.3.2.5). According to the calculations of Findlow et al. (1990), almost all the aluminum in milk (human and bovine) should be associated with citrate, with approximately 88% as Al(citrate)(OH)₂^-2 and approximately 11% as Al(citrate)(OH)^-1.

2.3.3.3 Elimination

The principal organ of aluminum excretion is the kidney, accounting for more than 95% of the total excretion (Exley et al. 1996; Krewski et al. 2007). The urinary excretion is believed to occur by passive filtration through the glomerulus, instead of active secretion by the proximal tubules. This hypothesis is based on the results of animal studies demonstrating that when only the free fraction of aluminum was assumed to be removed from blood, the elimination rate of aluminum is approximately the same as the glomerular filtration rate (Henry et al. 1984; Yokel and McNamara 1985, 1988). If this hypothesis is true, then the factors influencing glomerular filtration rates (such as kidney disease, pregnancy and age) should also influence the rate of elimination of aluminum (Guyton 1991). Indeed, it has been observed that individuals with renal failure have lower capacity of elimination (Nieboer et al. 1995; Krewski et al. 2007).

A small portion of the absorbed aluminum appears to be eliminated through other excretion routes. The second most important route would likely be biliary excretion. Most of the experimental studies with animals have demonstrated that less than 1.5% of the total eliminated aluminum occurred by biliary excretion (Krewski et al. 2007). As well, sweat, saliva and seminal fluid can contribute, to a much lesser extent, to the elimination of aluminum from the body (Krewski et al. 2007).

The elimination rate of aluminum appears to be regulated by the presence of various aluminum complexes in the body's systemic circulation. aluminum citrate complexes are eliminated more easily than Al-Tf (Maitani et al. 1994), most likely because the lower molecular weight of the aluminum citrate complex would facilitate glomerular filtration. This may explain why the presence of citrate can enhance renal elimination (Van Ginkel et al. 1993; Cochran et al. 1994). Also, the concomitant presence of aluminum and silicon yields a filterable complex (probably the same observed in the gastrointestinal tract); this complex seems to favour renal excretion by limiting the renal reabsorption of aluminum (Bellia et al. 1996; Birchall et al. 1996). As well, fluoride is a natural element which contributes to the rapid elimination of aluminum (Chiba et al. 2002).

Some animal studies have shown lower clearances of aluminum from the body, and consequently higher elimination half-lives (t½), after increasing the aluminum dosages (Höhr et al. 1989; Pai and Melethil 1989; Xu et al. 1991). This observation is probably explained by the fact that the fraction of ultrafilterable aluminum complexes decreased when the aluminum concentrations in blood increased (Xu et al. 1991; Yokel and McNamara 1988). Also, Greger and Radzanowski (1995) obtained a positive correlation between the t½ of aluminum in tibia and kidneys and the age of exposed rats, indicating that the ability to remove aluminum may diminish with time.

Priest et al. (1995) and Talbot et al. (1995) investigated the elimination rates of aluminum in humans, on the basis of the time-profiles of aluminum in blood and urine of seven volunteers who had received intravenous injection of ²⁶Al-citrate. Blood, urine and feces were collected during the five days following injection, except for the volunteer in Priest et al. (1995) for which the follow-up was at 13 days. Around 59.1% (46.4% to 74.42% range) of the uptake was excreted in the cumulative urine collected during 24 hours following injection whereas after five days, around 71.8% of the dose was recovered in urine (62.3% to 82.9% range). These results are considerably different than those reported in a study by Steinhausen et al. (2004), in which two volunteers received an IV injection of ²⁶Al-chloride, where the five-day urinary excretion accounted only for 25% of the dose.

Priest et al. (1995) and Talbot et al. (1995) described the whole-body retention of aluminum, blood concentration and urinary excretion, after the first day of injection, by a power function (e.g., C_b(t) = 0.37t^-0.9, expressed as a percent of injection/L). However, in a study with a follow-up period of 11 years, Priest (2004) demonstrated that the pattern of the whole-body retention of aluminum must be represented by a multiple-exponential equation.¹³ Numerous studies have actually shown that the rate of aluminum clearance in blood diminishes with time following aluminum administration, and thus a single elimination half-life (t½) cannot describe the whole-body elimination of aluminum (Priest 2004). Some authors have attempted to calculate specific t½ of aluminum for the tissues and organs of rats (Greger et al. 1994; Greger and Radzanowski 1995; Rahnema and Jennings 1999). In general, it was shown that aluminum deposited in well-perfused tissues/organs (e.g., kidneys and lungs) is released more rapidly than aluminum in slowly-perfused tissues (e.g., bone and spleen). These t½ values varied from 2.3 to 113 days. However, even if the brain is well-perfused, the retention of aluminum appears to be strong (see section 2.4.2.2). According to the experimental data in animals, Krewski et al. (2007) estimated that the t½ of aluminum deposited in brain is from 13 to 1,635 days.

A multicompartmental model was developed to describe the kinetics of aluminum in humans, based on the retention of ²⁶Al in the volunteer of the Priest et al. (1995) study, who was followed over more than ten years (Priest 2004). Five compartments are used to describe aluminum accumulation in the different organs and tissues; for each compartment, specific tissues or organs are indicated with a specific elimination half-life. These compartments are fed by the compartment of blood and extracellular fluids. As well, Nolte et al. (2001) proposed an open compartmental model to describe the kinetics of aluminum in humans based on the binding of aluminum with transferrin and citrate; this model was used by Steinhausen et al. (2004).

² An arithmetic mean was made with all available data per province or territory which provided total aluminum concentrations from water treatment systems that have surface water sources and use aluminum salts. Then an average of these values from the nine provinces/territories was calculated to represent the Canadian average (101 µg/L).
³ Background concentration is a term used in geochemical exploration that refers to the natural abundance of a sterile element from the Earth's crust (Hawkes and Webb 1962).
⁴ Average of the results obtained from over 40 studies covering ten provinces.
⁵ Estimate based on data pooled from the fourth and fifth Total Diet Studies.
⁶ Refer to http://www.hc-sc.gc.ca/fn-an/alt_formats/hpfb-dgpsa/pdf/legislation/e_c-tables.pdf for food-additive uses.
⁷ Weighted mean from the two Canadian studies (Bergerioux and Boisvert 1979; Koo et al. 1988). Human milk density = 1,030 g/L (Health Canada 1998a).
⁸ Note that that many aluminum containing products (e.g. antiacids, antiperspirants) are now considered Natural Health Products in Canada
⁹ www.hc-sc.gc.ca/dhp-mps/prodpharma/databasdon/index-eng.php
¹⁰ Bioavailability is determined by comparing the areas under the serum concentration x time curve (AUC) for the 26Al given orally and the 27Al administered intravenously (Yokel et al. 2008).
¹¹ The oral bioaccessibility is the soluble fraction of the substance in the gastrointestinal system that is available for absorption (Ruby et al. 1999).
¹² Guillard et al. (2004) indicated that the normal range of aluminum in blood plasma would be < 1.0 µg/dL.
¹³ The equation of the retention is
R(t) = 29.3e^-0.595∙t + 11.4e^-0.172∙t + 6.5e^{-0.000401∙t};
the corresponding elimination half-lives are 1.4, 40 and 1,727 days.

2.4.1 Ecotoxicology

Below, a brief summary of effects data for the most sensitive aquatic and terrestrial organisms is presented. More extensive descriptions of environmental effects are provided in several reviews (e.g., ATSDR 2006; Bélanger et al. 1999; Roy 1999a).

When aluminum salts are added to water, they hydrolyse, and monomeric aluminum can be formed in the dissolved fraction. It is the monomeric aluminum, and not the salts, that can adversely affect organisms (Driscoll et al. 1980; Parker et al. 1989; Baker et al. 1990). The following summary focuses, therefore, on the effects of the dissolved (particularly monomeric) forms of aluminum that are produced when aluminum salts dissociate.

2.4.1.1 Aquatic organisms

Most of the research on the impact of aluminum on aquatic life has been related to the impacts of acid rain. In this report, emphasis was placed on the potential toxic impacts of aluminum in waters of neutral or near-neutral pH as the available information suggests that releases associated with the three aluminum salts being assessed occur primarily into waters of circumneutral pH (Roy 1999b; Germain et al., 2000). As described below, because of this consideration, the most relevant effects data identified were for fish. This assessment report does not provide a detailed examination of potential effects from exposure to polymeric aluminum, as polymeric aluminum is most likely to form, and to cause toxicity, during the neutralization of acidic aluminum-rich waters and this is unlikely to occur in the release scenarios considered in this assessment (Roy 1999b).

The gills are the primary target organ for aluminum in fish (Dussault et al. 2001). Aluminum binds to the gill surface, causing swelling and fusion of the lamellae and increased diffusion distance for gas exchange (Karlsson-Norrgren et al. 1986; Tietge et al. 1988). The resulting damage leads to loss of membrane permeability, reduced ion uptake, loss of plasma ions, and changes in blood parameters relating to respiration. Fish death may result from ionoregulatory or respiratory failure, or a combination of both, depending upon the pH of the water and concentration of waterborne aluminum (Neville 1985; Booth et al. 1988; Gensemer and Playle 1999). Ionoregulatory disturbances prevail at lower pH (e.g., below 4.5) and relate to decreased levels of plasma Na+ and Cl¯ ions (Neville 1985; Gensemer and Playle 1999). At pH levels above 5.5, binding of the positively charged aluminum species to negatively charged sites on the gill surface, with subsequent aluminum polymerization, leads to mucous secretion, clogging of the interlamellar spaces and hypoxia (Neville 1985; Poléo 1995; Poléo et al. 1995; Gensemer and Playle 1999).

Aluminum exposure may also disrupt ionic balance and osmoregulation in aquatic invertebrates (Otto and Svensson 1983). Reduced Na+ and/or Ca²⁺ uptake in response to aluminum exposure have been documented in crayfish (Appleberg 1985; Malley and Chang 1985), mayfly nymphs (Herrmann 1987) and the water boatman, Corixa sp. (Witters et al. 1984). Aluminum reduced Na+ influx and, to a lesser extent, increased outflux, in Daphnia magna, thereby impairing osmoregulation (Havas and Likens 1985). Aluminum may disrupt the respiratory organs of some invertebrates, such as the anal papillae of the phantom midge, Chaoborus sp. (Havas 1986). Respiratory effects can occur when acidic waters are rapidly neutralized, such as when an acidic tributary enters a larger, neutral receiving stream, leading to the formation of mononuclear and polynuclear aluminum species from the dissolved ion (Gensemer and Playle 1999). These species may bind to or precipitate onto the bodies of invertebrates, creating a physical barrier to respiration. Aluminum has been reported to impair reproduction in Daphnia magna (Beisinger and Christensen 1972), although recent work with Daphnia pulex suggests that adaptive strategies which heighten survivorship and fecundity may occur following long-term exposure to sublethal levels (Wold et al. 2005). Hall et al. (1985) reported that aluminum may reduce the surface tension of water, affecting egg deposition, emergence, feeding and mating behaviour of some stream invertebrates.

2.4.1.1.1 Pelagic

Water pH is known to have a significant effect on the toxicity of dissolved aluminum. Under acidic conditions, aluminum is most toxic in the pH range 5.0-5.5. At more acidic pH, its toxicity decreases, while at still lower pH, aluminum can offer transitory protection against the toxicity of H⁺ (Muniz and Leivestad 1980; Baker 1982; van Coillie et al. 1983; Roy and Campbell 1995). Elevated concentrations of the cations Ca²⁺ and Mg²⁺ reduce the toxicity of metals (Pagenkopf 1983; Campbell 1995), yet there are relatively few results examining the effects of elevated calcium on aluminum toxicity. In fish exposed to aluminum at low pH, elevated calcium has been shown to improve survival (Booth et al. 1988; Mount et al. 1988; Sadler and Lynam 1988), reduce losses of plasma ions (Brown 1981; Sadler and Lynam 1988; McDonald et al. 1989) and reduce accumulation of aluminum on gills (Wood et al. 1988a,b). However, Duis and Oberemm (2001) reported low hatching success and high embryo mortality in vendace, Coregonus albula, exposed to high aluminum concentrations of 2.1 and 2.4 mg/L at low pH (4.75, 5.00) and in the presence of 111 to 117 mg/L calcium. Increasing calcium concentrations to 233 to 256 mg/L had no influence on hatching and survival percentages, suggesting that the toxic effect of high aluminum levels can exceed the protective effect of high calcium.

The toxicity of dissolved aluminum is reduced in the presence of inorganic ligands, such as fluorides, sulphates and silicates, as well as organic ligands, such as fulvic and humic acids (Roy 1999a). It is well established that DOM in particular influences the speciation and absorption of aluminum. In laboratory studies with fish, the toxicity of aluminum was reduced in the presence of organic acids, such as citric acid (Driscoll et al. 1980; Baker 1982), salicylic or oxalic acid (Peterson et al. 1989), humic acid (van Coillie et al. 1983; Parkhurstet al. 1990; Peuranen et al. 2002) and fulvic acid (Neville 1985; Lydersen et al. 1990a; Witters et al. 1990; Roy and Campbell 1997). In laboratory studies with amphibians (frog eggs and tadpoles), LC₅₀s for aluminum increased (i.e., toxicity was reduced) in the presence of DOM. However, in the field, the effects of DOM in attenuating aluminum toxicity are difficult to separate from the influences of pH and aluminum concentration (Clark and Hall 1985; Freda 1991).

Most aquatic toxicity studies involving aluminum have been conducted under conditions of low pH, and a number of these accounted for the solubility of the metal in the experimental design. The general conclusion of these studies is that aluminum toxicity is related to the concentration of dissolved inorganic monomeric aluminum (Roy 1999a).

At pH < 6.0, fish, the salmonids in particular, are among the most sensitive organisms to dissolved aluminum. In soft acidic waters, the LC₅₀ can be as low as 54 µg/L (for Atlantic salmon at pH 5.2), while in chronic studies, a Lowest-Observed-Effect Concentration (LOEC) of 27 µg/L was determined for growth (for brown trout [Salmo trutta] at pH 5.0). Some species of algae show a comparable sensitivity. Parent and Campbell (1994) determined a LOEC of 150 µg/L (as inorganic monomeric aluminum) at pH 5.0 with the alga Chlorella pyrenoidosa. While many invertebrates tolerate elevated levels of aluminum, Havens (1990) found that exposures to 200 µg Al/L at pH 5.0 were extremely toxic to Daphnia galeata mendotae and Daphnia retrocurva. France and Stokes (1987) concluded that stress from aluminum exposure was secondary to the stress of low-pH exposure for survival of Hyalella azteca. Results of other studies also suggest that invertebrates are more sensitive to low pH than to aluminum. Amphibians show a similar sensitivity. Freda (1991) summarized her work by concluding that aluminum can be lethal to amphibians that inhabit soft acidic (pH 4 to 5) waters if concentrations exceed 200 µg inorganic Al/L.

At pH 6.0 to 6.5, there are few studies that provide effects estimates in terms of inorganic monomeric aluminum. At pH 6.0, a LOEC of 8 µg/L (inorganic monomeric aluminum) for growth of the alga C. pyrenoidosa can be estimated from the data of Parent and Campbell (1994). Growth of the alga was reduced at this single exposure concentration in media without phosphate. This LOEC is, however, well within the likely range of natural concentrations of inorganic monomeric aluminum in surface water. In comparison, Neville (1985) observed that 75 µg Al/L (as inorganic monomeric aluminum) caused physiological distress to rainbow trout (Oncorhynchus mykiss) at pH 6.1 but not at pH 6.5.

At pH 6.5 to 8.0, there are few effects data available. At neutral or near-neutral pH, aluminum has a tendency to precipitate, and the chemistry of these solutions is difficult to control. While the toxicity of alum in neutral-pH waters has been the subject of many studies, the results are unreliable, due to extreme variation between replicates of the same exposure concentration and between duplicate experiments (Lamb and Bailey 1981; Dave 1985; George et al. 1995; Mackie and Kilgour 1995). However, a No-Observed-Effect Concentration (NOEC) for respiratory activity at pH 6.5 is provided by the results of the study by Neville (1985), who found that rainbow trout tolerated 75 µg Al/L (as inorganic monomeric aluminum) during exposures at this pH. Wold et al. (2005) reported a LOEC of 0.05 mg/L Al for reduced survival and reproduction in Daphnia pulex exposed for 21 days to concentrations ranging from 0.05 to 0.50 mg Al/L (nominal) as aluminum sulphate. The test water was maintained 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. In addition, the study reported that clonal populations of D. pulex derived from a lake with ongoing alum treatment showed higher age-specific survivorship, higher fecundity and faster growth rates than those collected from waters having less recent or no prior alum exposure. The researchers hypothesized that Daphnia may be capable of exhibiting adaptive strategies that heighten survivorship and fecundity when exposed to sublethal chemical stresses.

Gopalakrishnan et al. (2007) reported a lowest 24-hour EC₅₀ value of 0.210 mg/L for development of the trochophore larva in the marine polychaete, Hydroides elegans. The study was conducted at a pH of 8.1 and aluminum concentrations (measured using atomic absorption spectrophotometry) were well maintained within 2% to 15% of nominal values. Differential sensitivities were observed during embryogenesis and larval development, with lowest toxicity evident at the stage of the fertilization membrane and successively higher toxicity at the blastula and trochophore stages, respectively.

At pH > 8.0, LOECs for survival of rainbow trout are ≥ 1.5 mg/L as total aluminum (Freeman and Everhart 1971). In a more recent study, Gundersen et al. (1994) reported LC₅₀s for exposures of rainbow trout in the pH range 8.0-8.6. The LC₅₀s at all pHs were approximately the same value, ~0.6 mg/L (range: 0.36-0.79 mg/L) as dissolved aluminum (i.e., filterable through a 0.4-µm filter), and were similar in both acute (96-hour) and longer-term (16-day) exposures at hardness levels ranging from 20 to 100 mg/L (as calcium carbonate). A NOEC for mortality of 0.06 mg dissolved Al/L can be derived from data given for one of the 16-day exposures conducted at 20 mg/L hardness and pH 8.0. Although these concentrations were measured as dissolved aluminum, it is probable that the monomeric aluminate ion, AlOH4- , predominated at this pH.

In contrast, Poléo and Hytterød (2003) reported that juvenile Atlantic salmon, Salmo salar,exposed under alkaline (pH 9.5) conditions to concentrations of around 0.35 mg/L (predominantly aluminate ion) showed no acute toxicity effects. The researchers noted that the aluminum concentrations used in their study were lower than those of Freeman and Everhart (1971) and Gundersen et al. (1994), and hypothesized that more environmentally relevant concentrations of aluminum do not have any acute effect on salmonids under alkaline conditions, while very high concentrations of aluminum might have. While no acute effects were observed, physiological responses in the form of elevated blood glucose and hematocrit levels and a decrease in plasma Cl^-, were evident after a three-week exposure period and were considered indicative of a stress response in the fish. The authors concluded that the combination of high pH and aluminum may impose some stress but this is unlikely to represent a serious problem unless the exposure continues for a long period of time. High alkalinity conditions such as those used in the study can occur in water bodies during periods of intense photosynthetic activity in the summer months. At these times, concentrations of aluminum present in the water would also be expected to rise as the solubility of the substance increases over that at lower pH.

While toxicity is most commonly associated with inorganic monomeric aluminum species, there is evidence that aluminum undergoing transition from one species to another is also bioavailable and can exert adverse effects on organisms. Such transition conditions can occur in mixing zones, for example, when acidic waters enter a larger, more neutral receiving system or during the liming of acidic waters. Berkowitz et al. (2005) found that the addition of alum to lake water samples (pH 8.22 to 9.08) resulted in a rapid initial decrease in pH and alkalinity followed by a gradual recovery in pH over several weeks. Dissolved Al concentrations increased following treatment, and then decreased after 150 days. Soucek (2006) determined that freshly neutralized aluminum (i.e., aluminum in transition from ionic species in acidic waters to polymers or precipitating hydroxides after a rapid pH increase) impaired oxygen consumption in Daphnia magna and the perlid stoneflies, Perlesta lagoi and Acroneuria abnormis (lowest LOEC for the study 0.5 mg/L, which was also the lowest concentration tested). Alexopoulos et al. (2003) reported that freshly neutralized aluminum at a concentration of 0.5 mg/L associated specifically with the gills of the freshwater crayfish, Pacifastacus leniusculus, creating a physical barrier during precipitation that resulted in impaired respiration and asphyxiation. Particulate aluminum has been shown to decrease filter feeding in the freshwater bivalve, Anodonta cygnea, presumably as an avoidance response to the toxicant (Kádár et al. 2002). Poléo and Hytterød (2003) examined toxicity under steady-state (pH retained at 9.5) and non-steady state (pH lowered from 9.5 to 7.5) conditions in order to evaluate the possible impact of transient aluminum chemistry on Atlantic salmon, Salmo salar. No increase in toxicity occurred under the non-steady state conditions (i.e., where aluminum solubility was lowered as the pH decreased) and the physiological disturbances observed at high pH were mitigated. The results contrasted with those obtained in studies where aluminum solubility was lowered by raising the pH of aluminum-rich water. In these cases, toxicity to fish increased as the solubility of aluminum was decreased and aluminum precipitated onto the gills (e.g., Poléo et al. 1994; Poléo and Bjerkely 2000).

Verbost et al. (1995) reported enhanced toxicity in a mixing zone of acid river water containing aluminum (pH 5.1, aluminum 345 µg/L) with neutral lake water (pH 7.0, aluminum 73 µg/L). The resulting water (pH of 6.4, aluminum 235 µg/L) was expected to have low toxicity; however, the freshly mixed water was highly toxic to brown trout, Salmo trutta, with necrosis and apoptosis of the gills evident in exposed fish. A clear gradient in the deleterious effects occurred with increasing distance from the mixing area, with fish furthest from the mixing zone exhibiting only mild effects. The researchers concluded that freshly mixed acid and neutral water contains toxic components during the first seconds to minutes after mixing, and that even short exposure to this toxic mixing zone is detrimental to migrating trout. Farag et al. (2007) hypothesized that colloids formed in mixing zones may contribute to aluminum toxicity in fish by providing a direct route of the metal to the gills.

Finally, in a study done with DWTP sludge from Calgary and Edmonton, Alberta, AEC (1987) concluded that all sludges tested were non-toxic using a microbial test and acutely and subacutely non-toxic to rainbow trout. However, delayed release of first broods and significantly reduced reproduction were reported in the freshwater cladoceran, Ceriodaphnia dubia, exposed for 7 days to 100% aluminum sludge effluent collected from a DWTP in the U.S. (Hall and Hall 1989). The researchers considered that the effects were likely due to the combined effects of reductions in pH and dissolved oxygen concentrations, physical stress due to high levels of suspended solids, and possibly the presence of aqueous aluminum. Aqueous aluminum alone was probably not the factor exerting sub-lethal toxicity in 100% effluent since similar aqueous aluminum concentrations were observed in the 50% effluent where delays and significant reductions in reproduction were not observed. The same study observed significant mortality in fathead minnow, Pimephales promelas, exposed to the 100% effluent, as well as at the lowest test concentration of 6.3%. Mortality in the intervening concentrations was not statistically different from that in the controls. Mortality at 100% effluent was attributed to physical stress resulting from high levels of suspended solids. While a causative agent for the observed mortality at 6.3% could not be identified, the researchers noted that this test concentration had the highest concentration of aqueous aluminum, with measured levels up to 0.43 mg/L as compared with 0.05 to 0.31 mg/L at the other test concentrations. No sublethal impacts were evident in the fish testing.

2.4.1.1.2 Benthic

Alum can be used to treat eutrophic lakes to reduce the amount of phosphorus present in water or prevent its release from sediment. Lamb and Bailey (1981) concluded that a well-planned and controlled alum treatment would not result in significant mortality in benthic insect populations. Connor and Martin (1989) measured no detrimental effects on midge or alderly larvae following treatment of Kezar Lake, New Hampshire, sediment, and long-term effects on benthic invertebrates were minimal. Narf (1990) reported that benthic population diversities and numbers increased or remained the same following lake treatment with alum. Smeltzer (1990) observed a temporary impact on benthos after treatment of Lake Morey, Vermont, with an alum/sodium aluminate mixture. Benthos density, already low in the year prior to treatment, and richness were lower following treatment. However, changes were not significant, the benthic community recovered, and two new chironomids appeared the following year.

The Sludge Disposal Committee examined the impact of alum sludge discharge in aquatic environments and concluded that residue will tend to deposit near the point of discharge if the water velocity is low (Cornwell et al. 1987) and that it could have adverse effects, including development of anaerobic conditions. Roberts and Diaz (1985) related the reduction in phytoplanktonic productivity observed during alum discharge in a tidal stream in Newport News, Virginia, to the reduction in light intensity. Lin et al. (1984) and Lin (1989) found no buildup of sludge in pooled waters in the Vermillion and Mississippi rivers following sedimentation basin cleaning of DWTPs in St. Louis, Missouri. There were no significant differences in types and densities of macroinvertebrates in bottom sediments, and even higher density and diversity were found in some sites.

George et al. (1991; 1995) reported that macroinvertebrates located downstream of four DWTPs appeared to be stressed by alum discharges. In the Ohio River, effects seemed temporary and were limited in space. In addition, organisms collected from upstream locations indicated that environmental factors other than the aluminum sludge discharge may also have been affecting the system. A water-sediment microcosm study done with bottom sediment from the receiving rivers over a 72-day period showed significantly lower oligochaete content in bottom sediment treated with alum sludge. Testing with bentonite gave the same results, and the authors concluded that aluminum sludge deposits on sediment may have the potential to detrimentally affect benthic macroinvertebrate populations by limiting their access to oxygen or food and, therefore, the smothering effect from sludge may prove to be more important to aquatic organisms than aluminum content. However, in laboratory testing, filtrates obtained from aluminum sludge were toxic to the freshwater alga, Selenastrum capricornutum, in waters with low pH or a hardness of less than 35 mg/L CaCO₃, suggesting that water-soluble constituents from the aluminum sludge may be capable of affecting algal growth. The study recommended that further toxicity testing be conducted to more fully ascertain potential toxic effects, and that aluminum sludge not be discharged into soft surface waters (i.e., hardness < 50 mg CaCO₃/L) or those with a pH of less than 6.

A study has been undertaken to examine the environmental impact of filter backwash and basin cleaning effluents to the Ottawa River from the Britannia and Lemieux Island DWTPs in Ottawa (RMOC 2000; City of Ottawa 2002). In this study, riverine characteristics downstream of the Britannia site were reported to be beneficial for the sampling of benthic invertebrates due to the slow water velocities of a bay environment. Unlike the Britannia site, the Ottawa River in the vicinity of the Lemieux Island DWTP was characterized by strong currents and an absence of natual benthic habitat. To examine the impact of effluents from the Lemieux Island facility, artificial habitat was installed for benthic organisms at both upstream and downstream locations from the discharge site. The results of the sampling showed that species abundance and diversity was depressed at both sites downstream from the effluent discharges in comparison to sampling sites located upstream.  At sites located 150 and 6,000 m upstream from the Britannia DWTP outfall, approximately 160 and 250 organisms were counted, whereas downstream sites located at 0, 300, 500 and 1,500 m (furthest sampling location) had between 3 (at 0 m) and approximately 100 organisms (at 1,500 m) (diversity of organisms not provided for Britannia site). At the artificial sampling sites 30 and 110 m downsteam of the Lemieux Island DWTP, approximately 250 and 1,000 organisms were counted representing 17 and 21 taxa, respectively.  The site located 90 m upstream from the Lemiux discharge had approximately 1,800 organisms representing 24 taxa.

Toxicity of basin sediment from each of the Britannia and Lemieux Island DWTPs was also examined. The studies showed complete mortality of midge larvae (Chironomus riparius) within the 10 day test exposure, while survival of Hyalella azteca (14 day exposure) was not significantly different from that of the control animals. The study could not determine whether the mortality was attributable to the physical characteristics of the sludge (e.g., particle size) or the presence of chemical contaninants.  The sludge from the Lemieux Island DWTP was shown to inhibit growth of Hyalella azteca over the 14 day exposure period, but the Britannia DWTP sludge resulted in no observed effect. The study did not suggest why one sludge demonstrated growth effects, but not the other (methodology and experimental conditions were not provided).

Ultimately, the cause of the the depressed levels of organisms downstream in the Ottawa River from Britannia and Lemieux DWTPs was not due to one causal factor, rather may have resulted from a number of attributes including: physical composition of the sediment and its ability to support life; ongoing blanketing of the area due to new discharges; and toxicity of dissolved aluminum leaching out of the sediment into the water column (City of Ottawa 2002).

In studies related to wastewater releases by DWTPs, AEC (1984) reported there is potential for smothering effects on benthic organisms related to settled sludge on sediments following their release to rivers in Alberta. A number of other possible adverse impacts resulting from the discharge of aluminum sludge to receiving waters were identified, including: formation of sludge deposits in quiescent areas of streams; toxic effects on aquatic organisms from other contaminants present in the sludge; periodic high oxygen demand if water treatment plant sludge is discharged in large slugs or if previously deposited sludge is periodically re-suspended due to increased stream velocity; increased aluminum concentrations of downstream water supplies; and aesthetic problems where stream flow, stream turbidity, and/or sludge dilution are low. The researchers concluded that aluminum sludge exhibits a wide range of characteristics which depend on the raw water characteristics (turbidity, etc.) and other factors and, therefore, while numerous suspicions have been expressed regarding the potential for adverse effects resulting from the discharge of alum sludges to receiving waters, there appeared to be a lack of good scientific evidence to substantiate these concerns. Recommendations of the report included the acquisition of baseline data through bioassay testing and other studies, as well as consideration of alternatives to direct stream disposal practices such as reduction of the quantities of alum sludge produced through substitution with other coagulants, discharge at controlled rates to a sanitary sewer, lagooning with natural freeze-thaw dewatering, thickening and dewatering followed by landfilling, and land application.

A subsequent study examining the binding, uptake and toxicity of aluminum sludges from three water treatment systems in Edmonton and Calgary determined that aluminum was effectively bound to sludges within the pH range 4.5 to 10.0, with more than 99.98% of the total aluminum being in the form of sludge (AEC 1987). Sludge collected from the three plants was found to be non-toxic to rainbow trout, Long Evans rats, and the microbial toxicity test system, Microtox.

2.4.1.2 Terrestrial organisms

Research on the effects of aluminum to soil organisms has concentrated largely on screening for aluminum-tolerant strains of root nodulating bacteria and mycorrhizal fungi, due to the importance of these species in improving crop production (Bélanger et al. 1999). In general, toxicity threshold values for bacterial species fall in the range of 0.01 to 0.05 mM (pH 4.5 to 5.5), while those of mycorrhizal fungi range from 0.1 to 20 mM (pH 3.4 to 4.5) when based on hyphal growth inhibition and 30 to 157 mg/kg soil (pH 4.5 to 5.0) when based on reduced spore germination. For soil macroinvertebrates, growth of newly hatched earthworm, Dendrodrilus rubidus, was significantly reduced at 10 mg Al/kg soil (soil pH 4.2 to 4.9; Rundgren and Nilsson 1997), while significantly inhibited growth and cocoon production were reported for the earthworm, Eisenia andrei, at concentrations ranging from 320 to 1000 mg/kg dry soil, with toxicity decreasing as soil pH increased from 3.4 to 7.3 (van Gestel and Hoogerwerf 2001). A more complete examination of potential impacts to soil-dwelling microorganisms, fungi and invertebrates can be found in Bélanger et al. (1999).

The remainder of this section focuses on the effects of aluminum on sensitive plant species. It should be noted, however, that the problem with alum sludge may be associated not only with the direct toxic effects of aluminum on plants, but also with indirect effects related to phosphorus deficiencies (Jonasson 1996; Cox et al. 1997; Quartin et al. 2001). Aluminum's capacity to fix labile phosphorus by forming stable aluminum-phosphorus complexes and hence make it unavailable to plants can be responsible for the observed effects. In addition, toxic substances captured by the floc during water treatment may be available for uptake by soil species and exert adverse effects.

The presence of aluminum in solution, soil solution or soil resulted in a decrease in seedling growth, elongation or branching of roots of hardwood and coniferous species at varying levels (Horst et al. 1990; Bertrand et al. 1995; McCanny et al. 1995; Schier 1996). The most sensitive species was honeylocust (Gleditsia triacanthos) (Thornton et al. 1986a, 1986b). All measures of growth, except root elongation, consistently declined as solution aluminum increased, 0.05 mM or 1.35 mg/L being the critical value for a 50% general decrease (pH = 4.0). Since honeylocust is not an important species in Canadian forests and since the results obtained by Thornton et al. (1986b) contradict the results obtained for this species by other researchers, it was decided that the two next Lowest-Observed-Adverse-Effect Concentrations (LOAECs) are more relevant. Hybrid poplar (Populus hybrid) (Steiner et al. 1984) and red oak (Quercus rubra) (DeWald et al. 1990) showed a 50% decline in root elongation at an aluminum solution level of 0.11 mM (2.97 mg/L). The most sensitive coniferous species is pitch pine (Pinus rigida) (Cumming and Weinstein 1990). Seedlings inoculated with mycorrhizal fungus, Pisolithus tinctorius, showed increased tolerance to aluminum, whereas non-mycorrhizal seedlings exposed to 0.1 mM (2.7 mg/L) (pH 4.0) aluminum exhibited decreased root and shoot growth.

In an experiment done with scots pine (Pinus sylvestris), Ilvesniemi (1992) found that when nutrition was optimal, pines tolerated high levels of aluminum, but in nutrient-poor solution, their tolerance to aluminum was reduced tenfold. Hutchinson et al. (1986) and McCormick and Steiner (1978) also observed that pines were tolerant of high levels of aluminum in optimal nutrient solution.

Grain crop and forage crop species were also affected by different levels of aluminum (Bélanger et al. 1999). Wheeler et al. (1992) found that two barley (Hordeum vulgare) cultivars and eight common wheat (Triticum aestivum) cultivars were particularly sensitive, growth being decreased by more than 50% at aluminum levels as low as 0.005 mM (0.135 mg/L) (pH 4.5). Wheeler and Dodd (1995) also showed a 50% decline in growth of clover species, Trifolium repens, Trifolium subterraneum and Trifolium pratense, at 0.005 mM (0.135 mg/L) aluminum (pH 4.7). In a solution culture study, Pintro et al. (1996) found that the root elongation rate of maize (Zea maize HS777 genotype) was also negatively affected at an aluminum level of 0.005 mM (0.135 mg/L) (pH 4.4). In a study done on barley, Hammond etal. (1995) found significant amelioration of the toxic effects of aluminum on root and shoot growth when silicon was added to the solution medium. Silicon amelioration of aluminum toxicity in maize has also been reported (Barcelo et al. 1993; Corrales et al. 1997). In the presence of silicon, aluminum uptake seems to be decreased because of the formation of aluminum-silicon complexes, thus leading to a decrease in absorption of aluminum. In addition, complexes formed with organic anions, sulphate and phosphate appear to be non-toxic to plants (Kinraide 1997; Takita et al. 1999; Matsumoto 2000), while the aluminum-hydroxy species was reported to be phytotoxic in early studies (Alva et al. 1986; Wright et al. 1987; Noble et al. 1988a) but not in more recent ones (Kinraide 1997). Complexation with fluoride has been shown to ameliorate the phytoxic effects of aluminum in nutrient solutions (Cameron et al. 1986; Tanaka et al. 1987; MacLean et al. 1992); however, the aluminum-fluoride complex may also become toxic at high concentrations, with toxicity linked to the proportion and concentration of the different types of aluminum-fluoride species present in solution (Kinraide 1997; Stevens et al. 1997). Manoharan et al. (2007) reported severely restricted root growth in barley exposed to fluoride and aluminum in acidic soils (pH 4.25 to 5.48). Toxicity was attributed the activities of AlF₂⁺ and AlF²⁺ complexes formed in the soil. Fluoride may enter soil through the application of phosphate fertilizers, which usually contain 1% to 4% fluoride as an impurity (Loganathan et al. 2003). Calcium supplementation has also been reported to alleviate aluminum toxicity in barley, possibly by reducing cellular absorption of the metal and enhancing protection through increased activity of antioxidant enzymes (Guo et al. 2006).

Wheeler and Dodd (1995) investigated the effect of aluminum on yield and nutrient uptake of some temperate legumes and forage crops using a low ionic strength solution. The solution aluminum levels at which top yield and root yield of 58 white clover cultivars were reduced by 50% ranged from approximately 0.005 to 0.02 mM (0.135 to 0.540 mg/L) (pH 4.5 to 4.7).

Although inorganic monomeric forms of dissolved aluminum (Al³⁺, Al(OH)²⁺ and Al(OH)₂⁺) are believed to be the most bioavailable and responsible for most toxic effects (Alva et al. 1986; Noble et al. 1988b), information on the concentrations of different dissolved aluminum complexes was not reported in many of the effects studies reviewed. For studies indicating particular sensitivity that were carried out in the laboratory in artificial solutions, it is likely that the majority of the aluminum present in these key studies was in inorganic monomeric forms. Considering that solution culture experiments gave lower LOEC values than did sand culture experiments in forest species studies, the effects data reviewed are considered to be conservative estimates of the effects levels for vegetation grown in natural soils.

2.4.2 Experimental mammal studies

The scientific literature concerning the effects of aluminum exposure in experimental mammals is large, including studies with a variety of administration routes (ingestion, inhalation, dermal, intraperitoneal, intravenous, intracisternal). The characterization of effects presented below includes studies of oral, inhalation and dermal administration, with emphasis on the oral exposure studies. This reflects the importance of the oral route in environmental exposures within the general Canadian population, as compared to dermal and inhalation as well as the research emphasis on oral studies within the scientific community. For more detailed discussion of other routes of exposure, the reader may consult the comprehensive reviews cited, in particular Krewski et al. (2007).

Health Canada considers neurotoxicity and reproductive/developmental toxicity as the categories of effects of greatest potential concern for the general population, in light of the evidence from case studies and epidemiological investigations, discussed in section 2.4.3. Recent comprehensive reviews also collectively support this conclusion (InVS-Afssa-Afssaps 2003; ATSDR 2006; JECFA 2006; Krewski et al. 2007; EFSA 2008). Thus, most of the studies presented in this section focus on neurotoxicity or reproductive/developmental toxicity in which aluminum is administered to the experimental animals through diet, drinking water or gavage.

Various aluminum salts, including chloride, nitrate, sulphate, lactate, citrate, maltolate, fluoride and hydroxide have been used in experimental animal studies to investigate the effects of Al³⁺ absorbed in the bloodstream and distributed to target organs. Aluminum speciation (i.e., the ligands associated with aluminum) and the overall composition of the diet may influence toxicokinetics and consequently the subsequent toxicity of Al³⁺ (see section 2.3.3.1.1). With respect to absorption, however, no one aluminum salt is representative of the mix of aluminum compounds in the human diet that contribute to the Al³⁺ reaching the bloodstream. Therefore, for the purpose of characterizing effects of total aluminum, all oral studies were examined, regardless of the aluminum salt administered. Relative bioavailability of particular salts is then considered in the exposure-response analysis of section 3.2.3.

A number of the experimental animal studies are designed to explore the influence of factors that may potentially exacerbate the toxic effects of aluminum (e.g., restraint) or provide protection (e.g., therapeutic substances such as Gingko). The results reported in this section, however, focus on the differences between aluminum-treated animals and controls, rather than the influence of these other factors.

In most of the studies consulted, there is a lack of data on the aluminum concentration in the base diet. Studies on different brands of commercial laboratory animal chow show that aluminum levels in the chow can be significant relative to the administered doses, and also highly variable between brands and even between different lots of the same brand (ATSDR 2006). Typical levels of 250 to 350 ppm of aluminum in rodent chow (ATSDR 2006) would contribute approximately 13 to 18 mg Al/kg/d in rats and 33 to 46 mg Al/kg/d in mice, on the basis of default reference values for animal intake and body weight proposed in Health Canada (1994). While it may be hypothesized that the absorption of the base diet aluminum may differ from (and be significantly less) than the absorption of the administered aluminum, there are little relevant experimental data on this question (see section 2.3.3). Therefore the lack of data on base diet aluminum in many of the toxicity studies must be considered as a major uncertainty in the overall database, when considering these studies in the exposure-response analysis and risk characterization.

Notwithstanding the importance of quantifying total aluminum exposure in animal studies, in order to provide a qualitative summary of the literature for the purpose of hazard identification, all studies have been evaluated, regardless of whether the base diet aluminum concentration is reported. In the exposure-response analysis (section 3.2.3), however, administered and combined doses are distinguished and the influence of this factor is considered.

The description of the studies in this section is focused on the nature of the effects investigated and observed, rather than the exposure-response relationship. The database is large (138 studies) and the experimental conditions (e.g., administered salts and dosing regimen) vary, and in the majority of the studies only one dose was tested. Thus direct comparisons of the dose-effect data may be misleading. While some information on the lowest observed dose at which effects occurred is provided¹⁴ as well as the highest dose at which no effects were observed, a more detailed discussion of the exposure-response analysis is presented in section 3.2.3. The details of the studies considered in that analysis are summarized in Tables C1 and C2 (Appendix C). Tables summarizing the full dataset are available in the Health Canada Supporting Document, prepared for this draft assessment (Health Canada 2008a).

2.4.2.1 Acute toxicity

Oral exposure

The oral LD₅₀ (lethal dose, 50% kill, single administration) for different aluminum salts, as measured in different strains of mice, rats, guinea pigs and rabbits, varies according to the aluminum salt administered as well as according to the experimental animal species. In an early review an LD50 of apparently 6,200 mg Al/kg bw was reported for Al₂(SO₄)₃ and of 3,850 mg Al/kg bw for Al(Cl)₃ administered to mice (Sorenson et al. 1974), although it is unclear from the review article if these values refer to the dose in terms of aluminum or the dose in terms of the salts. Sorenson et al. (1974) also reported LD₅₀ values from 260 to 4,280 mg/kg bw for Al(NO₃)₃-9H₂O in two separate studies on rats. The lower value of 260 mg/kg Al(NO₃)₃-9H₂O clearly underestimates the LD₅₀ (i.e., overestimates the toxicity), as Colomina et al. (2002), Colomina et al. (2005) and Domingo et al. (1996) have shown. These research groups tested administered doses of 50 to 100 mg Al/kg bw/d, equivalent to approximately 700 to 1,400 mg Al(NO₃)₃-9H₂O/kg bw/d, and the effects were limited to alterations in weight gain and subtle neurological effects (see sections 2.4.2.2 to 2.4.2.4 and section 3.3 for more detailed discussion of these studies).

In a study of oral and intraperitoneal administration during 14 days, Llobet et al. (1987) estimated the acute oral toxicity of aluminum chloride, nitrate and sulphate in Sprague-Dawley rats and Swiss mice. Aluminum chloride and nitrate produced acute toxicities of similar magnitude (LD50 of 222 to 370 mg Al/kg) in the mice and rats, whereas the toxicity of aluminum sulphate was considerably lower (LD₅₀ > 730 mg Al/kg in both species).

Inhalation exposure

In Golden Syrian hamsters and New Zealand rabbits exposed over a short duration (four to six hours per day for three to five days at levels of 7 to 200 mg/m³) to aluminum chlorohydrate through inhalation, the effects observed are those typically associated with inhalation of particulate matter, including alveolar wall thickening, increased number of macrophages and increased lung weight (ATSDR 2006). A more detailed discussion of the pulmonary effects in experimental animals of inhalation exposure to aluminum oxide dust and refractory alumina fibres, and aluminum hydroxide is provided by Krewski et al. (2007). The observed responses to various species of aluminum are described as "typical of foreign body reaction", including alveolar proteinosis and wall thickening, and some nodule formation.

Dermal exposure

Dermal effects of aluminum compounds (10% w/v chloride, nitrate, chlorohydrate, sulphate, hydroxide) applied to skin of mice, rabbits and pigs over five-day periods (once per day) include epidermal damage, hyperkeratosis, acanthosis and microabscesses (ATSDR 2006; Krewski et al. 2007).

2.4.2.2 Short-term toxicity (duration of exposure less than 90 days)

Oral exposure

The results of 40 short-term studies in adult mice, rats and rabbits (exposure duration between 3 and 13 weeks) are summarized below. In all the studies considered, aluminum was administered orally in drinking water, in the diet or by gavage. The aluminum salts include lactate, chloride, sulphate, nitrate and hydroxide. In some studies citrate was administered with the aluminum salt in order to enhance absorption.

As discussed in section 2.4.2, many of the short-term studies did not quantify the concentration of aluminum in the base diet. In these cases the value of the actual combined dose is highly uncertain, particularly in the studies where the administered dose was significantly less than the possible baseline dose in the diet (e.g., Basu et al. 2000; El-Demerdash 2004; Kaizer et al. 2005; Kaur and Gill 2005, 2006; Jyoti and Sharma 2006; Sparks et al. 2006; Kaur et al. 2006). In three studies (Thorne et al. 1986; Shakoor et al. 2003; Campbell et al. 2004), ambiguities in the reporting of the doses precluded consideration of the dose-response relationship; however the qualitative observations from these studies are included in the following summary of effects.

Neurobehavioural effects in adult rats and mice following oral administration from 21 to 90 days included decreased performance in the rotarod test (Bowdler et al. 1979; Shakoor et al. 2003; Kaur et al. 2006), decreased performance in passive and active avoidance tests (Commissaris et al. 1982; Connor et al. 1988; Connor et al. 1989; Kaur et al. 2006), reduced motor activity (Commissaris et al. 1982; Golub et al. 1989; Shakoor et al. 2003), decreased forelimb and hindlimb grip strength (Oteiza et al. 1993), increased sensitivity to flicker (Bowdler et al. 1979) and air puff startle response (Oteiza et al. 1993), and reduced recovery in neurological function following spinal cord injury (Al Moutaery et al. 2000).

Of the above studies, the lowest administered dose at which effects occurred was observed by Kaur et al. (2006), in which male Wistar rats were administered 10 mg Al/kg bw/d as aluminum lactate for up to 12 weeks, with testing at 0, 4, 8 and 12 weeks. A significant decrease in performance between exposed and control groups was observed at four weeks and became more pronounced following eight weeks of exposure. Decreased performance in memory function tests (passive and active avoidance responses) was also observed in the exposed animals tested at 12 weeks.

In contrast, no alterations in passive or active avoidance test results were reported in aluminum-exposed animals, at doses of 67 mg Al/kg bw/d of aluminum chloride administered by gavage to male Sprague-Dawley rats for 28 days (Bowdler et al. 1979) and 600 mg Al/kg bw/d of aluminum nitrate administered in drinking water for 14 days to male CD mice (Colomina 1999).

Reduced body weight among aluminum-exposed animals was observed by Bataineh et al. (1998), at a dose of 15 mg Al/kg bw/d of aluminum chloride administered to male Sprague-Dawley rats in drinking water for 12 weeks. On the other hand, Colomina et al. (1999) observed a reduction in body weight only at 600 mg Al/kg bw/d of aluminum nitrate, and no effect at 300 mg Al/kg bw/d, in mice administered aluminum via drinking water for 14 days. In other short-term studies, the authors either did not observe this effect, at a dose of100 mg Al/kg bw/d administered in the diet of Swiss Webster mice (Donald et al. 1989; Golub and Germann 1998), or did not report differences in body weight between exposed and control groups.

The most extensive histopathological changes in the short-term studies were reported by Roy et al. (1991a) in which male rats were given doses of 17 to 172 mg Al/kg bw/d as aluminum sulphate via gavage. The concentration of aluminum in the base diet was not quantified. Multifocal neuronal degeneration, abnormal and damaged neurons, and reduced neuronal density were identified in specific brain regions (e.g., cerebral cortex, subcortical region and base of brain) at 29 mg Al/kg bw/d. In the liver, Roy et al. (1991a) observed cytoplasmic degeneration in the periphery of the hepatic lobule at all doses. With increasing doses, multifocal degeneration of the entire liver tissue was observed, followed by fibrous tissue proliferation. Kidney effects observed in this study at 22 mg Al/kg bw/d included increased swelling and degeneration of the cortical tubules.

Other histopathological effects reported in different strains of rats include necrosis-like changes in hippocampal CA1 cells and accumulation of synaptic vesicles in presynaptic terminals (Jyoti and Sharma 2006), congestion of cerebral and meningeal blood vessels, multifocal neuronal degeneration, neurofibrillary degeneration and foci of demyelination (El-Rahman 2003), increased membrane fluidity and decreased cholesterol/phospholipid ratio in synaptosomes (Silva et al. 2002), increased number of vacuolated spaces in the matrix of the cerebral cortex (Basu et al. 2000), decreased NADPH-diaphorase positive neurons in the cerebral cortex (Rodella et al. 2001) and increased hippocampal muscarinic receptors (Connor et al. 1988). The lowest administered doses at which such changes occurred were in the studies of Jyoti and Sharma (2006) in which exposed male Wistar rats received a dose of 10 mg Al/kg bw/d of aluminum chloride in drinking water for five weeks, and of Basu et al. (2000), in which male Sprague-Dawley rats received 10 mg Al/kg bw/d of aluminum chloride via gavage for 40 days.

The biochemical changes to the brains of adult rodents resulting from oral administration of aluminum salts for periods of less than 90 days included effects on cholinergic neurotransmission (Kumar 1998; Shakoor et al. 2003; El-Demerdash 2004; Kaizer et al. 2005; Kaur and Gill 2006) as well as changes in the levels of other neurotransmitters and signalling proteins (Flora et al. 1991; Tsunoda and Sharma 1999b; Kumar 2002; El-Rahman 2003; Becaria et al. 2006), alterations in calcium transfer, binding and signalling in the brain (Kaur et al. 2006; Kaur and Gill 2005), evidence of oxidative stress in different regions of the brain (Fraga et al. 1990; Katyal et al. 1997; Abd el-Fattah et al. 1998; El-Demerdash 2004; Nehru and Anand 2005; Becaria et al. 2006; Jyoti and Sharma 2006), changes in ATPase activity (Katyal et al. 1997), alterations to cyclic AMP second messenger systems (Johnson and Jope 1987), increased levels of amyloid precursor protein (Becaria et al. 2006) and increased TNF-µ (alpha tumour necrosis factor) mRNA expression in the brain (Tsunoda and Sharma 1999a; Campbell et al. 2004). The lowest administered dose at which such effects were observed was 10 mg Al/kg bw/d administered to rats as aluminum lactate via gavage or as aluminum chloride via drinking water in Kaur and Gill (2006) and Basu et al. (2000).

Inhalation exposure

The toxicological literature for short-term inhalation exposure studies is limited compared to that for oral exposure. The most recent comprehensive reviews of this literature can be found in ATSDR (2006) and Krewski et al. (2007). The most sensitive and best documented endpoints concern the respiratory system. The observed effects were those commonly associated with particle inhalation exposure (> 7 mg/m³), including a thickening of the alveolar walls, an increase in alveolar macrophages and heterophils, granulomatous nodules and lesions, and increased lung weight (ATSDR 2006).

2.4.2.3 Subchronic and chronic toxicity (exposure duration greater than 90 days, non-cancer endpoints)

Oral exposure

The results of 49 subchronic and chronic toxicity studies (exposure greater than 90 days) in adult mice, rats, rabbits, monkeys and dogs are summarized below. In all the studies considered, aluminum was administered orally in drinking water, in the diet or by gavage. The aluminum salts include lactate, chloride, sulphate, nitrate, hydroxide, citrate, maltolate, fluoride and KASAL (basic sodium aluminum phosphate).

As in the case of the short-term studies, many of the subchronic and chronic toxicity studies did not quantify the concentration of aluminum in the base diet. In those studies where the administered dose was substantially less than the possible baseline dose in the diet, the uncertainty associated with the actual combined dose was increased (see, for example, Krasovskii et al. (1979); Fleming and Joshi (1987); Bilkei-Gorzo (1993); Varner et al. (1993); Varner et al. (1994); Varner et al. (1998); Sahin et al. (1995); Somova et al. (1997); Jia et al. (2001a); Pratico et al. (2002); Abd-Elghaffar et al. (2005); Hu et al. (2005); Becaria et al. (2006); and Li et al. (2006)).

Neurobehavioural effects in adult mice and rats, following oral exposure for 90 days or more, included decreased spontaneous motor activity (Commissaris et al. 1982; Lal et al. 1993; Jia et al. 2001a; Jia et al. 2001b; Hu et al. 2005). The lowest administered dose associated with this effect was 1 mg Al/kg bw/d as observed by Huh et al. (2005) in male Sprague-Dawley rats who received aluminum maltolate at this dose in drinking water over a period of one year¹⁵ . In contrast Domingo et al. (1996) and Colomina et al. (2002) found no differences in field activity of Sprague-Dawley rats, where animals received an administered dose of 100 mg Al/kg bw/d of aluminum nitrate (with citrate) in drinking water for periods of four to six months. Decreased motor coordination as measured by performance in the rotarod test (Sahin et al. 1995), decreased grip strength, and effects on temperature sensitivity and negative geotaxis (Golub et al. 1992a) were also observed.

Other observed neurobehavioural effects included learning and memory deficits (maze performance, passive avoidance tests) reported by Bilkei-Gorzo (1993), Lal et al. (1993), Gong et al. (2005), Gong et al. (2006) and Li et al.(2006). The lowest administered dose associated with such effects was 6 mg Al/kg bw/d, observed by Bilkei-Gorzo (1993) in Long Evans rats exposed for 90 days to aluminum chloride (plus citrate) via gavage, although there was some ambiguity in the reporting of doses in this study. In contrast, no effects on similar learning or memory tests were observed by Varner et al. (1994), Domingo et al. (1996), Colomina et al. (2002) and von Linstow Roloff et al. (2002). In the study of von Linstow Rolloff et al. (2002) an administered dose of 140 mg Al/kg bw/d was administered to male Lister hooded rats as aluminum sulphate in drinking water.

With respect to body weight, Pettersen et al. (1990), Gupta and Shukla (1995), Colomina et al. (2002) and Kaneko et al. (2004) observed reductions in body weight in aluminum-exposed animals (rodents and dogs) at doses ranging from 25 mg Al/kg bw/d of aluminum maltolate administered in drinking water to mice for up to 120 days (Kaneko et al. 2004) to 94 mg Al/kg bw/d of aluminum nitrate administered in drinking water to rats for 114 days (Colomina et al. 2002). In the Kaneko et al (2004) study, aluminum chloride was administered to another exposure group at the same dose as aluminum maltolate, and no difference in body weight between aluminum-exposed animals and controls was observed. The authors attributed the contrasting observations to the greater bioavailability of aluminum maltolate as compared to chloride, documented as well by the greater accumulation of aluminum in the brain, liver, kidney and spleen in mice exposed to aluminum maltolate.

Histopathological effects reported in rats and mice included increased damaged or abnormal neurons in specific brain regions (e.g., cerebral cortex and hippocampus) (Varner et al. 1993; Varner et al. 1998; Abd-Elghaffar et al. 2005), neurofibrillary degeneration and vacuolization of nuclei (Somova et al. 1997), and vacuolated astrocytes and vacuolization of neuronal cytoplasm (Florence et al. 1994). The lowest administered dose in which these effects were observed was less than 1 mg Al/kg bw/d in the Varner et al. (1998) and Varner et al. (1993) studies in which aluminum nitrate and sodium fluoride (to form aluminum fluoride) was administered in drinking water to male Long Evans rats for periods of 45 to 52 weeks¹⁶.

Petterson et al. (1990) observed mild to moderate histopathological effects in testes, liver and kidney, including hepatocyte vacuolization, seminiferous tubule germinal epithelial cell degeneration and tubular-glomerularnephritis in beagle dogs receiving a dose of 75 mg Al/kg bw/d of sodium aluminum phosphate. In this same study, no significant differences between exposure groups and controls were observed at the lower doses of 4 to 27 mg Al/kg bw/d.

The biochemical endpoints examined in subchronic and chronic experimental studies are considerably varied, as are the methodologies used to investigate these endpoints. The observed effects included a decrease in nitrergic neurons in the somatosensory cortex (Rodella et al. 2006), perturbations in ATPase activity in the brain (Lal et al. 1993; Sarin et al. 1997; Swegert et al. 1999; Silva and Goncalves 2003; Kohila et al. 2004; Silva et al. 2005), induced apoptosis in the brain (Huh et al. 2005), effects on cholinergic enzyme activities (Bilkei-Gorzo 1993; Zheng and Liang 1998; Dave et al. 2002; Zatta et al. 2002; Kohila et al. 2004), increased cytokine levels (Becaria et al. 2006), increased catalytic efficiency of monoamine oxidases A and B (Huh et al. 2005), increased caspase 3 and 12 (Gong et al. 2005; Huh et al. 2005), increased staining for amyloid precursor protein levels (Gong et al. 2005) and amyloid beta (Ab) levels (Pratico et al. 2002), decrease in long-term potentiation in hippocampal slices (Shi-Lei et al. 2005), and alterations in phospholipid and cholesterol levels in the myelin membrane, synaptosomes or the brain (Sarin et al. 1997; Swegert et al. 1999; Pandya et al. 2001; Silva et al. 2002; Pandya et al. 2004). The lowest administered dose associated with significant effects on biochemical endpoints was 1 mg Al/kg bw/d as administered as aluminum maltolate in drinking water for one year (Huh et al. 2005)¹⁷ .

Other biochemical and biophysical effects observed in the brains of aluminum-exposed rodents included alterations in trace metal (Cu, Zn and Mn) metabolism in the brain (Sanchez et al. 1997; Yang and Wong 2001; Jia et al. 2001a; Fattoretti et al. 2003; Fattoretti et al. 2004), altered synapses in the hippocampus and frontal cortex (Jing et al. 2004), increase in area occupied by mossy fibres in the hippocampal CA3 subfield (Fattoretti et al. 2003; Fattoretti et al. 2004), increase (Flora et al. 2003) and decrease (Jia et al. 2001a) in glutathione peroxidase activity, and increase in catalase activity (Flora et al. 2003). Increased lipid peroxidation was reported by Lal et al. (1993), Gupta and Shukla (1995), Sarin et al. (1997), Pratico et al. (2002), Flora et al. (2003) and Kaneko et al. (2004). Jia (2001a), Gupta and Shukla (1995) and Abd-Elghaffar (2005) reported decreased levels of superoxide dismutase, and Jia et al. (2001a) observed increased levels in malondialdehyde. Johnson et al. (1992) observed decreased levels of cytoskeletal proteins (microtubule associated protein-2, spectrin) in the hippocampus and brain stem.

Inhalation exposure

The toxicological literature for subchronic and chronic inhalation exposure studies is limited. ATSDR (2006) and Krewski et al. (2007) report on several studies of durations of six months (six hours a day, five days a week). The most sensitive and best documented endpoints concerned the respiratory system. The observed effects are those commonly associated with particle inhalation exposure (> 600µg/m³), including a thickening of the alveolar walls, and an increase in alveolar macrophages, granulomatous lesions and relative lung weight (ATSDR 2006).

2.4.2.4 Reproductive and developmental toxicity

Oral exposure

The results of 49 studies investigating gestational, lactational and/or post-weaning exposure of rats, mice and guinea pigs to aluminum salts through diet, through drinking water or by gavage are summarized below. The aluminum salts administered in these studies included chloride, nitrate, sulphate, lactate and hydroxide. In a few studies citrate or ascorbic acid was added to enhance absorption of aluminum.

As discussed in sections 2.4.2.2 and 2.4.2.3, the lack of information on base diet for some studies is a major source of uncertainty with respect to the potential combined dose, particularly when the administered dose was low in comparison to the possible base diet dose (e.g., Clayton et al. 1992; Ravi et al. 2000). There is also uncertainty associated with reported LOELs that are of the same magnitude as the reported LD₅₀ for the administered salt (Johnson et al. 1992; Misawa and Shigeta 1993; Poulos et al. 1996; Llansola et al. 1999).

The most commonly observed neurobehavioural effects in developmental studies included decreased grip strength (Golub et al. 1992b; Golub et al. 1995; Colomina et al. 2005), reduced temperature sensitivity (Donald et al. 1989; Golub et al. 1992b), reduced or delayed auditory startle responsiveness (Misawa and Shigeta 1993; Golub et al. 1994), and impaired negative geotaxis response (Bernuzzi et al. 1986; Bernuzzi et al. 1989a; Muller et al. 1990; Golub et al. 1992b). Decreased activity levels (Cherroret et al. 1992; Misawa and Shigeta 1993), locomotor coordination (Golub et al. 1987; Bernuzzi et al. 1989a; Bernuzzi et al. 1989b; Muller et al. 1990; Golub and Germann 2001b) as well as impaired righting reflex (Bernuzzi et al. 1986; Bernuzzi et al. 1989b) were also observed, although not consistently -- refer to Thorne et al. (1987), Golub et al. (1992b), and Misawa and Shigeta (1993). The lowest administered dose at which effects on these endpoints were observed was 100 mg Al/kg bw/d, observed in Wistar rats administered aluminum lactate in the maternal diet during gestation (Bernuzzi et al. 1989b) as well as in Swiss Webster mice administered aluminum lactate in the maternal diet during gestation, lactation and then in the diet of offspring throughout the lifespan (Golub et al. 2000).

The observations on the effects on learning and memory of developmental exposure to aluminum salts also varied considerably. For example, in some studies improved performance in the maze tasks was observed (Golub et al. 2000; Golub and Germann 2001a; Colomina et al. 2005) while in others impaired performance (Golub and Germann 2001b; Jing et al. 2004) or no change (Thorne et al. 1987) was found. Golub and Germann (2001b) observed diminished maze learning in Swiss Webster mice pups when dams were exposed to aluminum lactate in the diet at a combined dose of 50 mg Al/kg bw/d, but not at 10 mg Al/kg bw/d, during gestation and lactation, and pups were exposed via diet for two weeks following weaning. In this experiment, animals (controls and aluminum-exposed) were fed a sub-optimal diet, designed to simulate the usual diet of U.S. women with regard to recommended dietary amounts of trace elements.

The observations of Roig et al. (2006) suggested a biphasic effect on learning in rats exposed to aluminum nitrate during gestation, lactation and post-weaning; in a two-dose study, the low-dose group (50 mg Al/kg bw/d of aluminum nitrate plus citrate in drinking water) performed significantly better in the water maze test than the high-dose group (100 mg Al/kg bw/d), but there was no significant difference between the high-dose group and the controls. With respect to passive avoidance tests, the same group of researchers also reported improved performance in aluminum exposed animals at an administered dose of 100 mg Al/kg bw/d (Colomina et al. 2005).

Developmental exposure of mice and rats to aluminum salts also produced some evidence of disturbances in brain biochemistry, such as alterations in brain lipid contents and increased lipid peroxidation (Verstraeten et al. 1998; Verstraeten et al. 2002; Nehru and Anand 2005; Sharma and Mishra 2006) or decreased lipid peroxidation (Golub and Germann 2000), decreased levels in superoxide dismutase (Nehru and Anand 2005), delayed expression of a phosphorylated neurofilament protein (Poulos et al. 1996), differential effects on choline acetyltransferase activity in various brain regions (Clayton et al. 1992; Rajasekaran 2000; Ravi et al. 2000), decreased serotonin and noradrenaline levels in specific brain regions (Ravi et al. 2000), decreased concentrations of manganese in brain (Golub et al. 1992b; Golub et al. 1993), alterations to signal transduction pathways associated with glutamate receptors and decreased expression of proteins of the neuronal glutamate-nitric oxide-cGMP pathway (Llansola et al. 1999; Kim 2003), and alterations in secondary messenger systems (Johnson et al. 1992). With respect to biochemical endpoints, the lowest administered dose at which effects were measured was approximately 20 mg Al/kg bw/d, observed by Kim (2003) in which male and female Fisher rats received this dose of aluminum chloride in drinking water for 12 weeks prior to mating, after which treatment at this dose continued in dams during gestation and lactation.

Chen et al. (2002), Wang et al. (2002a) and Wang et al. (2002b) reported impairment of synaptic plasticity, as measured by field potentials in the dentate gyrus of the hippocampus. Johnson et al. (1992) reported decreased levels of microtubule associated protein-2 in the brains of rat pups exposed eight weeks following weaning, although no changes in other cytoskeletal proteins were observed. A significant decrease in myelin sheath width was observed in mice pups exposed during gestation, lactation and then through the diet following weaning (Golub and Tarara 1999), and in guinea pig pups exposed prenatally from GD30 to birth (Golub et al. 2002). These effects were observed at administered doses above 85 mg Al/kg bw/d as aluminum chloride in drinking water of Wistar rat dams (Wang et al. 2002a; Wang et al. 2002b; Chen et al. 2002) and 100 mg Al/kg bw/d in the diet of Swiss Webster mice dams (Golub and Tarara 1999).

Although the focus of the majority of the investigations of prenatal exposure was neurodevelopmental toxicity, effects on some reproductive endpoints were reported as well. Golub et al. (1987), Bernuzzi et al. (1989b), Gomez et al. (1991), Colomina et al. (1992), Belles et al. (1999), Sharma and Mishra (2006) and Paternain et al. (1988) reported reduced maternal weight gain, although no change in this parameter was observed by Donald et al. (1989), Golub et al. (1993), Golub et al. (1995) and Golub et al. (1996a), nor was it reported in the other studies. In regard to pup body weight, Sharma and Mishra (2006), Wang et al. (2002a), Llansola et al. (1999), Cherroret et al. (1995), Misawa and Shigeta (1993), Gomez et al. (1991), Paternain et al. (1988), Domingo et al. (1987), Thorne et al. (1987), Golub and Germann (2001a), Colomina et al. (1992), and Bernuzzi et al. (1989a), Bernuzzi et al. (1989b) reported decreases in aluminum-exposed groups, while other studies reported no effects (Donald et al. 1989; Clayton et al. 1992; Golub et al. 1992b; Golub et al. 1993; Golub et al. 1995; Golub et al. 1996a; Colomina et al. 1994; Verstraeten et al. 1998). The lowest administered dose at which effects on reproductive parameters, including fetal growth, were observed was 13 mg Al/kg bw/d (Paternain et al. 1988; Domingo et al. 1987a), in which Sprague-Dawley rat dams received this dose via gavage as aluminum nitrate.

Cherroret et al. (1995) reported decreased plasma concentrations of total proteins and albumin and increased plasma a1 globulins, which the authors attributed to an inflammation process in young rats exposed postnatally by gavage at doses of 100 to 200 mg Al/kg bw/d. The same research group also observed effects on duodenal enterocytes, with a decrease in microvilli width and significant variation in K, Ca, S and Fe concentrations (Durand et al. 1993).

Other observed reproductive/developmental effects included a decrease in the number of corpora lutea and number of implantation sites (Sharma and Mishra 2006) as well as skeletal malformations (Paternain et al. 1988; Colomina et al. 1992; Sharma and Mishra 2006). Colomina et al. (2005) reported a delay in sexual maturation in both males and females, although this effect was produced at different dose levels in the two sexes (at 50 mg Al/kg bw/d in females and at 100 mg Al/kg bw/d in males). Misawa and Shigeta (1993) observed delayed pinna detachment and eye opening in female pups.

No significant maternal or developmental toxicity, as measured by fetal weight gain, reproductive parameters or fetal malformations, was observed by McCormack et al. (1979) at a combined dietary dose of aluminum chloride of 50 mg Al/kg bw/d, nor by Gomez et al. (1990) where 265 mg Al/kg bw/d of aluminum hydroxide was administered to dams via gavage during gestation.

Inhalation and dermal exposure

No studies were identified concerning the reproductive effects of inhalation or dermal exposure to aluminum salts.

2.4.2.5 Carcinogenicity

The literature concerning oral exposure bioassays is very limited. An increase in gross tumours was reported in male rats and female mice in a one-dose study but few study details were reported (Schroeder and Mitchener 1975a, 1975b, as reported in ATSDR 2006). Two other studies reported no increased incidence of tumours in rats and mice exposed orally to aluminum compounds (Hackenberg 1972; Oneda et al. 1994).

No increased tumour incidence was observed in rats following inhalation of alumina fibres at concentrations of up to 2.45 mg/m³ (Krewski et al. 2007).

The International Agency for Research on Cancer did not classify specific aluminum compounds for carcinogenicity, but classified the exposure circumstances of aluminum production as carcinogenic to humans (Group 1) (IARC 1987).

2.4.2.6 Genotoxicity

The genotoxicity of various aluminum compounds is described in detail by Krewski et al. (2007) and ATSDR (2006). Briefly, aluminum compounds have produced negative results in most short-term in vitro mutagenic assays, including the Rec-assay using Bacillus subtilis, in Salmonella typhimurium TA92, TA 98, TA102, TA104 and TA1000 strains (with and without S9 metabolic activation), and in Escherichia coli (see Krewski et al. 2007).

In vitro studies of rat ascites hepatoma cells reported that aluminum chloride could serve as a stimulator for the crosslinking of chromosomal proteins (Wedrychowski et al 1986a, 1986b, as reported in Krewski et al. 2007, ATSDR 2006). Studies on human blood lymphocytes showed that aluminum chloride could induce positive responses for both micronuclei formation and sister chromatid exchange (see Krewski et al. 2007).

More recently Lima et al. (2007) investigated the genotoxic effects of aluminum chloride in cultured human lymphocytes. Comet assay and chromosome aberrations analysis were used to evaluate DNA-damaging and clastogenic effects of aluminum chloride at different phases of the cell cycle. All tested concentrations (5 to 25 µM aluminum chloride) were cytotoxic, reduced the mitotic index, induced DNA damage and were clastogenic in all phases.

Roy et al. (1991) administered doses of aluminum sulphate and potassium aluminum sulphate in drinking water to male rats at doses ranging from 17 to 171 mg Al/kg bw/d for up to 21 days. The frequency of abnormal cells increased in direct proportion to both the dose and the duration of exposure to the aluminum salts. Most aberrations were chromatid breaks, with translocations recorded at higher doses.

In a recent review of the safety of aluminum from dietary intake, EFSA (2008) summarized indirect mechanisms that might explain the genotoxic effects observed in experimental systems. The proposed mechanisms included cross-linking of DNA with chromosomal proteins, interaction with microtubule assembly and mitotic spindle functioning, induction of oxidative damage, and damage of lysosomal membranes with liberation of DNAase to explain the induction of structural chromosomal aberrations, sister chromatid exchanges, chromosome loss and formation of oxidized bases in experimental systems. EFSA (2008) suggested that these indirect mechanisms of genotoxicity, occurring at relatively high levels of exposure, would not likely be of relevance for humans exposed to aluminum via the diet.

2.4.3 Human studies

In this section, information on the potential human health effects associated with aluminum exposure is briefly summarized with the goal of describing the range of potential effects. As such, various exposure routes are considered in order to identify the possible target organs. This information includes data from case studies, epidemiological investigations into the potential health effects of exposure to aluminum in drinking water, occupational investigations of exposure to aluminum dust and welding fumes, and exposure to aluminum via vaccines and of dermal application of aluminum-containing antiperspirants.

In section 3.2.2 an evaluation of these health effects is presented, in order to: (a) identify critical effects; and (b) determine which, if any, of the human studies may be used to estimate the dose-response relationship. The latter determination is based on the strength of the available evidence and the relevance of the studies to environmental exposure in the general Canadian population.

2.4.3.1 Human case studies of exposure to aluminum

Human cases studies of aluminum toxicity have been well documented for specific medical conditions, most frequently in patients with renal impairment undergoing dialysis with aluminum-contaminated dialysate or receiving medications with elevated aluminum concentration. A small number of case studies or investigations have focused on children and pre-term infants receiving parenteral nutrition. Although the effects in particular sub-groups of susceptible individuals are not representative of exposure conditions for the general population, they are presented in order to identify the target organs of aluminum exposure. A more detailed discussion of these human case studies is presented in the comprehensive reviews InVS-Afssa-Afssaps (2003) and Krewski et al. (2007). As well, a case study is described below in which exposure to aluminum was associated with the accidental discharge of aluminum into the municipal water supply.

Aluminum toxicity in patients with renal impairment

Historically, patients undergoing dialysis treatment were exposed to aluminum through the water used to prepare dialysis solutions and from aluminum compounds prescribed as phosphate binders (Krewski et al. 2007). Today, this exposure is strictly controlled¹⁸. However, in the past, many cases of aluminum-induced encephalopathy, resulting in alterations in behaviour and memory, speech disorders, convulsions and muscle-twitching occurred in dialysis patients (Foley et al. 1981; Alfrey 1993). In cases of intoxication, the aluminum was introduced into the systemic circulation through the dialyzing membrane (in hemodialysis) or abdomen (in peritoneal dialysis) thus bypassing the gastrointestinal barrier, and was therefore completely available at the cellular level. The effects of elevated aluminum exposure in dialysis patients has provided clear evidence for the neurotoxicity of aluminum in humans.

Researchers have also identified cases of individuals with impaired renal function who, because of their reduced capacity to eliminate aluminum and chronic high exposure to aluminum-containing medications, also developed encephalopathy, even though they were not undergoing dialysis (Foley et al. 1981; Sedman et al. 1984; Sherrard et al. 1988; Moreno et al. 1991). A fatal case of aluminum-induced encephalopathy occurred in a patient with chronic renal failure who did not have dialysis treatment, but who consumed large doses of aluminum-containing antacids (Zatta et al. 2004).

Other toxic effects of aluminum observed in dialysis-exposed patients include haematological effects such as anaemia (Bia et al. 1989; Yuan et al. 1989; Shah et al. 1990; Caramelo et al. 1995) and skeletal toxicity (osteomalacia and osteitis fibrosis) (Mathias et al. 1993; Jeffery et al. 1996; Ng et al. 2004).

Aluminum exposure via intravenous nutritional support

Klein (2005) reviewed the human evidence regarding the effects of aluminum exposure via solutions used for intravenous nutritional support with regard to effects on bone (osteomalacia) and the central nervous system. With respect to parenteral nutrition, infants may be a particularly sensitive sub-group because of the immaturity of the blood-brain barrier and renal excretory mechanisms. Bishop et al. (1997) investigated cognitive impairment in pre-term infants in relation to parenteral nutrition. In a randomized trial the researchers found that performance in neurodevelopmental testing conducted at 18 months was significantly better in 92 pre-term infants who had received a low-aluminum nutritional solution as compared to 90 pre-term infants receiving a standard solution with higher aluminum content. No follow-up testing that evaluated cognitive performance in the children of this cohort as they aged was identified.

Investigation of aluminum exposure associated with contamination event in Camelford, UK

Exley and Esiri (2006) reported an unusual case of fatal dementing illness in a 58-year-old woman, resident of Camelford, Cornwall, in the United Kingdom. Fifteen years earlier, at the age of 44 years, this person was exposed to high concentrations of aluminum sulphate in drinking water, which had been accidentally discharged in the drinking water supply of the region. During this event, up to 20,000 people were exposed to aluminum concentrations in drinking water varying from 100 to 600 mg/L. At the autopsy of the woman, a rare form of sporadic early-onset b-amyloid angiopathy in the cerebral cortical and leptomeningeal vessels, and in leptomeningeal vessels over the cerebellum was identified. Coincident high concentrations of aluminum were also found in the severely affected regions of the cortex. To date, this remains the only documented case. Exley and Esiri (2006), who reported this case, state that the role of aluminum is uncertain but may be clarified through future research in similarly exposed and unexposed populations (controls).

2.4.3.2 Epidemiological studies of aluminum exposure via drinking water

By the end of the 1980s, four epidemiological studies with an ecological design (i.e., using group rates of exposure and disease) had reported positive associations between the concentration of aluminum in drinking water and the occurrence of Alzheimer's disease (AD) or of dementia (Vogt 1986; Martyn et al. 1989; Flaten 1990; Frecker 1991). These observations resulted in further research into the relationship of aluminum in drinking water and various dementia syndromes, particularly AD.

Epidemiological studies based on observations of individuals were conducted in the 1990s with the aim of investigating the association between AD or other cognitive dysfunctions and exposure to aluminum in drinking water. Health Canada published a comprehensive review of epidemiological studies in Guidelines for Canadian Drinking Water Quality - Technical Documents: Aluminum (Health Canada 1998b) and in the SOS report of 2000. The discussion presented below summarizes the information presented in the previous reviews and presents more recently published findings of the eight-year follow-up analysis of a large cohort in southwestern France (Rondeau et al. 2000; Rondeau et al. 2001). The study designs and findings of the relevant epidemiological studies are presented in Table B1 (Appendix B). These data have also been described in detail in Krewski et al. (2007) and InVS-Afssa-Afssaps (2003). Analysis of the epidemiological database and its applicability in a quantitative risk assessment is presented in the Hazard Characterization of this assessment (section 3.2.2.1).

Twelve studies are presented in Table B1, based on case-control, cross-sectional, or longitudinal designs. The observations from two Ontario case-control studies are drawn from the same study population -- the Ontario Longitudinal study of Aging (LSA) -- and all the French studies were based on observations from the "Principal lifetime occupation and cognitive impairment in a French elderly cohort" or PAQUID cohort. However, the LSA and PAQUID study populations differ with respect to the case definition and the manner of diagnosis of disease. In the PAQUID investigations, the earlier studies used a case-control design whereas the more recent studies by Rondeau et al. (2000) and Rondeau et al. (2001) used a cohort incidence analysis.

Positive findings for an association between aluminum exposure and AD or other neurological dysfunctions were found to be statistically significant (p < 0.05) in seven of the twelve studies, although the strength and significance of these associations depended on how the data were analysed (see Appendix B) These seven studies were carried out in Ontario (Neri and Hewitt 1991; Forbes et al. 1992; Forbes et al. 1994; Forbes et al. 1995a; Forbes et al. 1995b; Neri et al. 1992; Forbes and Agwani 1994; Forbes and McLachlan 1996; McLachlan et al. 1996), in Quebec (Gauthier et al. 2000) and in France (Michel et al. 1991; Jacqmin et al. 1994; Jacqmin-Gadda et al. 1996; Rondeau et al. 2000; Rondeau et al. 2001).

In Ontario, a series of analyses was conducted on the LSA cohort to investigate the relationship between the concentration of aluminum in drinking water and cognitive impairment, as established by interviews and questionnaires (Forbes et al. 1992; Forbes et al. 1994; Forbes et al. 1995a; Forbes and Agwani 1994). These authors observed statistically significant associations only when they controlled their analyses according to certain physical-chemical parameters of water, such as fluoride, pH, and silica. Since the methods of interviews and questionnaires for characterizing cognitive functions were deemed to be insufficiently specific for accurately detecting neurological impairments, Forbes et al. (1995b) and Forbes and McLachlan (1996) consulted death certificates from individuals on the LSA cohort and examined the association between aluminum in drinking water and AD or presenile dementia as categorized by the corresponding ICD¹⁹ codes. Positive relationships between aluminum and AD and presenile dementia were reported with and without adjustments with different water quality parameters. For instance, some of the highest risks for AD were observed when high concentrations of aluminum (≥ 336 µg Al/L) were combined with high pH (≥ 7.95), low levels of fluoride (< 300 µg/L) or low levels of silica (< 1.5 mg/L).

Neri and Hewitt (1991) and Neri et al. (1992) reported a significant dose-response relationship between AD or presenile dementia and aluminum using hospital discharge records from Ontario, and by matching cases and controls according to age and sex. Another study from Ontario was a case-control analysis from the Canadian Brain Tissue Bank cohort in which AD was confirmed by histopathological criteria (McLachlan et al. 1996).

Although all studies from Ontario assessed the exposure of aluminum based on the data of the water quality surveillance program of the Ontario Ministry of the Environment, only McLachlan et al. (1996) evaluated the past exposure to aluminum²⁰. However, in the McLachlan et al. (1996) study, the analysis was not controlled for potential confounders and modifying factors (e.g., age, sex, education and occupation), and the significant positive associations were not adjusted for other chemical or physical parameters in water.

The single study from Quebec was a case-control analysis of AD and exposure to various aluminum species in residential drinking water (Gauthier et al. 2000). The diagnosis of AD was based on a three-step procedure to discriminate between AD and other neurological disorders. In addition to controlling for a number of confounding factors as well as the aluminum speciation, these authors took into account historical exposure to aluminum in drinking water. Gauthier et al. (2000) reported 16 odds ratios (OR) but observed only one significant positive association (i.e., OR > 1), which was related to the concentration of monomeric organic aluminum in drinking water. This significant association was found, however, when only current exposure was considered, and not for long-term exposure, which would be expected to be more biologically-relevant.

The three studies conducted on populations from the United Kingdom showed no significant association between aluminum concentration in drinking water and neurological dysfunction, following adjustment for sex and age (Wood et al. 1988; Forster et al. 1995; Martyn et al. 1997), but none of these authors adjusted their statistical tests according to the physical-chemical properties of the drinking water. The health outcome in the two case-control studies was AD, diagnosed by a three-step procedure for including cases of presenile dementia (Forster et al. 1995) or by a clinical diagnosis using unspecified criteria (Martyn et al. 1997). This latter study, which took into account past exposure, also did not observe differences between cases and controls when the analyses were restricted to subjects exposed to low levels of silica in drinking water (< 6 mg/L). The cross-sectional study of Wood et al. (1988) was based on data collected from patients from northern England with hip fractures, for whom dementia was evaluated (no information about the diagnostic tests).

The study from Switzerland (Wettstein et al. 1991), which was a cross-sectional examination of mnestic skills in octogenarians from Zurich and aluminum in drinking water, also reported no significant associations when controlling for socio-economic status, age, and education. It should be noted that the high-exposure district in this study had drinking water with a mean aluminum concentration of 98 µg/L. Thus the analysis was carried out for a drinking water supply that was generally lower in aluminum than the drinking water supplies considered in the other investigations.

All the studies from France were based on the PAQUID cohort. The studies of Michel et al. (1991) and Rondeau et al. (2001), reported significant positive associations between the exposure to aluminum in drinking water and the occurrence of AD or dementia diagnosed by a two-step procedure, whereas the positive associations reported by Jacqmin et al. (1994) and by Jacqmin-Gadda et al. (1996) were based on the scores of the Mini-Mental State Examination (MMSE). The results of Michel et al. (1991) have been discounted, however, because of a reliance on potentially unreliable historical information on drinking water concentrations (Jacqmin et al. 1994; Smith 1995; WHO 1997).

Jacqmin et al. (1994) and Jacqmin-Gadda et al. (1996) analysed the same database collected from the PAQUID cohort in different ways, with inconclusive results. The first study included an initial report of the effect of pH on the association between aluminum and cognitive impairment (Jacqmin et al. 1994). Without considering the effect of the pH-aluminum interaction, these authors reported a positive association between aluminum and cognitive impairment, whereas consideration of this interaction resulted in a negative association. These results remained statistically significant only if occupation was included in the logistic regressions. Jacqmin-Gadda et al. (1996) expanded their analyses to include the levels of silica in drinking water. While their results indicate a protective effect of aluminum against cognitive impairment with high level of silica (≥ 10.4 mg/L) and high pH (≥ 7.5), the consideration of the interaction of aluminum and silica in their logistic regression suggests an adverse effect of aluminum on neurological functions.

Rondeau et al. (2000) retained the unimpaired subjects in the studies of Jacqmin et al. (1994) and Jacqmin-Gadda et al. (1996), and evaluated the incidence of dementia and AD one, three, five and eight years after the initial MMSE. This follow-up analysis reported a positive association between aluminum and AD or dementia, after adjustment for age, sex, education and place of residence as well as for consumption of wine and bottled mineral water. This study addressed some of the limitations of previous epidemiological investigations by adjusting for the potential confounders, and while exposure levels were not weighted according to residential history, residential history was considered. At baseline, 91% of the subjects had lived more than ten years in the same parish, with a mean length of residence of 41 years. A total of 3,401 participants were included in the study at baseline, although only 2.6% of the subjects were exposed to an aluminum concentration greater than 100 µg/L. Nonetheless, the associations between aluminum in drinking water and dementia, and aluminum in drinking water and AD, were highly significant. Only two exposure groups (< 100 µg/L or > 100 µg/L) were defined in the principal analysis and no dose-response relationship was found when exposure categories were more finely divided.

Many of the epidemiological studies investigating the association between aluminum in drinking water and the development of cognitive impairment or AD did not control for important potential confounders or modifying factors, or did not adequately characterize past exposure. The Rondeau (2000) study addressed some of these limitations. However, the subjects in the cohort were not generally exposed to high levels of aluminum (97% of subjects exposed to less than 100 µg/L), and within the limited exposure range, no dose-response relationship was observed.

2.4.3.3 Epidemiological investigations of exposure to aluminum in antacids, antiperspirants or food

Only very weak or no associations have been found between repeated exposures to aluminum in antacids and AD in a number of analytical epidemiological studies (Heyman et al. 1984; Graves et al. 1990; Flaten et al. 1991; CSHA 1994; Forster et al. 1995; Lindsay et al. 2002) Positive associations between AD and the use of aluminum containing antiperspirants were reported in two case-control studies, but the interpretation of the results is difficult due to methodological limitations of the studies (e.g., missing data, and misclassification due to varying brands and subtypes of antiperspirant with varying aluminum contents) (Graves et al. 1990; CSHA 1994). This positive observation, however, was not supported by a follow-up study on the CSHA²¹ cohort (Lindsay et al. 2002); the results show that regular use of antiperspirant did not increase the risk of AD.

Rogers and Simon (1999) conducted a pilot study to examine dietary differences in individuals with AD and matched controls (n = 46: 23 subjects, 23 controls). The exposure assessment was based on questionnaires to determine past dietary habits. According to the authors, there may be an association between AD and the consumption of foods containing high levels of aluminum food additives. However, the sample size was very small and the association was statistically significant only for one category of food (pancake, waffle and biscuit).

2.4.3.4 Epidemiological investigations of exposure to aluminum in vaccines

Aluminum adjuvants are included in some vaccines to enhance and extend the immune response of some antigens. Aluminum hydroxide and phosphate salts as well as aluminum sulphate can be used as an adjuvant (Eickhoff and Myers 2002).

Possible associations between AD and historical exposure to vaccines have been investigated in the CSHA cohort (Verreault et al. 2001). Exposure to conventional vaccines appears to lower the risk of developing AD. After adjustments for age, sex and education, the ORs were 0.41 (95% CI 0.27-0.62) for the diphtheria or tetanus vaccines, 0.60 (95% CI 0.37-0.99) for the poliomyelitis vaccines and 0.75 (95% CI 0.54-1.04) for the influenza vaccine. Except for the influenza vaccine, all others contain aluminum-adjuvants (Eickhoff and Myers 2002).

The possible links between the hepatitis B vaccine, which contains aluminum-adjuvants, and the risk of demyelinating diseases such as multiple sclerosis (MS) have been investigated in France (Touze et al. 2000; Touze et al. 2002), England (Sturkenboom et al. 2000; Hernan et al. 2004), the U.S. (Zipp et al. 1999; Ascherio et al. 2001), Canada (Sadovnick and Scheifele 2000) and Europe (Confavreux et al. 2001). Only the study of Hernan et al. (2004) observed a significant positive association between MS and the hepatitis B vaccine, but no association between MS and the tetanus or influenza vaccines, which also contain aluminum adjuvants.

2.4.3.5 Epidemiological investigations of occupational exposure to aluminum

Subclinical neurological effects have been observed in a number of studies of workers chronically exposed to aluminum (aluminum potroom and foundry workers, welders, and miners). Many of these studies involved small numbers of workers and involved the assessment of exposure based on occupation rather than measured airborne aluminum concentrations, and most involved mixed exposures to various dusts and chemicals. Endpoints examined in different studies varied and for those that were similar, results were not always consistent. The types of neurological effects observed included impaired motor function (Hosovski et al. 1990; Sjogren et al. 1996; Kilburn 1998), decreased performance on cognitive tests (attention, memory, visuospatial function) (Hosovski et al. 1990; Rifat et al. 1990; Bast-Pettersen et al. 1994; Kilburn 1998; Akila et al. 1999), subjective neuropsychiatric symptoms (Sjogren et al. 1990; White et al. 1992; Sim et al. 1997) and quantitative electroencephalographic changes (Hanninen et al. 1994).

In one case-control study from England (Salib and Hillier 1996) and two from the U.S. (Gun et al. 1997; Graves et al. 1998), the relationship between the occurrence of AD and occupational exposure to aluminum was investigated. In each study, disease status was defined by standard criteria (e.g., NINCDS-ADRDA and/or DSM)²², and exposure to airborne aluminum (e.g., welding fumes, dusts and flakes) was assessed through occupational history questionnaires administered to informants. In none of these studies was there a significant association between occupational exposure to airborne aluminum and AD.

A four-year longitudinal study investigated neurobehavioural performance in 47 aluminum welders in the train and truck construction industry, with a control group drawn from assembly workers in the same industry (Kieswetter et al. 2007). Exposure to aluminum in dust was assessed through total dust collected on filter samples attached to the welders' helmets as well as through biomonitoring (aluminum in plasma and urine) at the time of neurobehavioural testing (start of investigation, after two years, after four years). The battery of neurobehavioural tests included an evaluation of cognitive abilities, psychomotor performance, attention and memory. This study used a small number of participants to explore the potential use of different biomonitoring measures, dust levels and exposure duration to predict performance in neurobehavioural tests. The study was not designed to find a relationship if one existed, but rather to explore the use of different exposure measures. Although exposure to aluminum among the welders was considered to be high in comparison to other occupational studies of aluminum (88 to 140 µg Al/g creatinine in urine, or approximately 103 to 164 mg Al/L)²³, no association between exposure and neurobehavioural performances was found.

A meta-analysis was conducted for nine investigations of occupational aluminum exposure and neurobehavioural performance, with a total of 449 exposed subjects with mean urinary aluminum concentrations of 13 to 133 µg Al/L (Meyer-Baron et al. 2007). Even if almost all effect sizes indicated an inferior neurobehavioural performance of the exposed group to aluminum, only one out of ten performance variables (the digit symbol test) was statistically significant. However, the statistical significance of the digit symbol results relationship to aluminum exposure was reduced when one study, in which the biomonitoring measure was estimated on the basis of an uncertain conversion factor, was excluded from the analysis. The authors concluded that with respect to occupational exposure, as indicated by urinary concentrations of less than 135 µg Al/L, there is concurring evidence of an impact on cognitive performance and acknowledge that international standardization for exposure is needed.

2.4.4 Mode of action of toxic effects of aluminum

Information related to possible modes of action by which aluminum affects the nervous system, as explored in animal and human studies, has been discussed in a number of recent reviews (Strong et al. 1996; Savory 2000; Kawahara 2005; ATSDR 2006; Savory et al. 2006; Krewski et al. 2007; Shcherbatykh and Carpenter 2007; Goncalves and Silva 2007). In addition, Jeffery et al. (1996) and Krewski et al. (2007) consider the mode of action in relation to bone and hematopoietic tissue.

The mechanism of aluminum neurotoxicity is an area of active research, with multiple lines of investigation. The purpose of the present discussion is to briefly summarize the areas of investigation relating to mode of action of aluminum toxicity, as mostly tested in laboratory rodents or in vitro studies, and present the range of views regarding the relevance of these data to human neurodegeneration, and particularly the development of AD.

Neurotoxic effects

There is evidence from studies in both laboratory animals and humans that absorbed aluminum is distributed to the brain, particularly the cerebral cortex and hippocampus. For example, the accumulation of aluminum in the brains of adult mice, rats and monkeys from the exposed groups was reported in 23 studies of neurological effects of orally administered aluminum (described in section 2.4.2)²⁴. Increased aluminum in the brains of pups exposed only during pregnancy was observed by Sharma and Mishra (2006), but not by others (Colomina et al. 2005; Golub et al. 1992b). Other studies of prenatal exposure in which exposure continued through lactation also reported increased aluminum in the brain (Wang et al. 2002a; Chen et al 2002; Golub et al. 1993). In contrast, Golub et al. (2000) observed decreased aluminum levels in the brains of mice exposed during gestation, lactation and through their lifespan.

Other research documenting the distribution of aluminum in the brain is described in section 2.3.3.2.

The research on aluminum neurotoxicity in laboratory animals has generally focused on the following interrelated categories of biochemical and cellular effects:

• peroxidation of membrane lipids and other sources of oxidative stress;
• increased inflammatory response;
• alterations in the lipid/phospholipid composition of myelin, with consequent effects on neurotransmission and synaptic function;
• impaired glucose metabolism;
• effects on neurotransmission, including cholinergic and glutamatergic systems;
• alterations to second messenger systems (e.g., inositol triphosphate, cAMP and Ca²⁺);
• accumulation of intracellular calcium;
• accumulation of mitochondrial Ca²⁺, resulting in release of cytochrome c and subsequent apoptosis;
• perturbation in the distribution and homeostasis of essential metals with potential adverse metabolic effects;
• alteration of phosphorylation level of neurofilaments, including phosphorylation of tau-protein, and resulting neurofibrillary tangle formation;
• inhibition of axonal transport;
• accumulation of amyloidβ peptide;
• alterations in gene expression and binding to DNA;
• alterations to the permeability of the blood-brain barrier.

There has been some effort to integrate the evidence for the above biochemical effects into a common mechanism, or at least a group of mechanisms of action for the neurotoxicity of aluminum (for example, see Kawahara (2005) and Shcherbatykh and Carpenter (2007)). Strong et al. (1996) argued that "a single unifying mechanism of aluminum neurotoxicity that will encompass all the potential means by which aluminum acts at the cellular level probably does not exist." These authors did, however, propose the following general categories by which aluminum neurotoxicity may be characterized, as a means for focusing future research on mechanisms of action:

• the induction of cytoskeletal pathology in the form of neurofilamentous aggregates: the mechanisms of this induction include those at the level of gene expression to altered post-translational processing (phosphorylation or proteolysis) of neurofilaments;
• alterations in cognition and behaviour in the absence of cytoskeletal pathology but with significant neurochemical and neurophysiological modifications: these include effects on cholinergic activity, signal transduction pathways and glucose metabolism;
• developmental neurotoxicity: research into this lifestage could focus on whether the mechanisms of action of aluminum that have led to neurobehavioural alterations in the developing fetus are similar to those responsible for toxicity in the adult as well as the nature of these alterations (permanent versus transient).

The relationship between the mechanism of aluminum neurotoxicity in animals and to the potential mechanism in AD remains an important topic of discussion. This is a complex debate as the basic cellular mechanism for AD is not clear. The presence of senile plaques composed of Aβ peptides in the brains of individuals with AD is well-documented, but the means by which these peptides produce neurotoxicity is not known (Marchesi 2005). Superimposed on the debate on the mechanisms for AD is the controversy as to whether environmental exposure to aluminum could contribute to the development of AD. The recent literature includes arguments across the spectrum, from the view that no compelling evidence for the "aluminum hypothesis" exists today (Becking and Priest 1997; Wisniewski and Lidsky 1997) to the view that the different animal and epidemiological evidence suggest that environmental aluminum may indeed be an important contributing factor for AD and that it is important not to prematurely reject this hypothesis (Yokel 2000; Gupta et al. 2005; Kawahara 2005; Exley 2006; Miu and Benga 2006; Savory et al. 2006). The proponents of further investigation into the role of aluminum in the development of AD cite, among others, the following lines of evidence, in addition to the epidemiological evidence (described in section 2.4.3.2), for which counter arguments have also been put forward:

• Increased aluminum in the whole brains of AD individuals at autopsy, as compared to age-matched non-AD brains, has been observed in some studies, although not in others. Investigations focusing on the measurement of aluminum in the senile plaques and neurofibrillary tangles of AD brains have also produced variable results, possibly as a result of difficulties and differences of the analytical methods used (Environment Canada and Health Canada 2000; Yokel 2000).
• Aluminum injected into the brain or spinal cord of certain species (e.g., rabbit, cat, guinea pig and ferret) produce effects that have some similarities to AD pathology, although there are significant differences as well.

For example, abnormally phosphorylated tau is the principal protein of the paired helical filaments that make up the neurofibrillary tangles that are diagnostic of AD. Aluminum-induced phosphorylation of tau protein has been demonstrated in somein vitro and in vivo studies (Yokel 2000; Savory et al. 2006). Yet, although aluminum induces neurofilament aggregates in model species, these differ structurally from neurofibrillary tangles that are diagnostic of AD in humans.

The deposition of senile plaques, also a hallmark of AD, is not observed in animal models, but increased immunoreactivity to Aβ and its parent molecule, amyloid precursor protein, via aluminum has been demonstrated in both in vitro and in vivo studies (Environment Canada and Health Canada 2000).

• Dialysis encephalopathy (see section 2.4.3.1 for discussion) is clearly recognized as resulting from aluminum intoxication. This condition has provided clear evidence for the neurotoxicity of aluminum in humans. Nonetheless, the very different clinical symptoms and progression of the two diseases as well as their differing pathologies have been cited as evidence of a lack of causal relationship between aluminum and AD (Wisniewski and Lidsky 1997).

Bone toxicity

In the case of osteomalacia associated with aluminum exposure, two distinct mechanisms of actions are recognized (ATSDR 2006). Firstly, the oral exposure to high levels of aluminum can produce a complex with dietary phosphorus, impairing gastrointestinal absorption of this element necessary for bone mineralization. Secondly, the osteomalacia associated with increased bone concentrations of aluminum, principally located at the mineralization front, is associated with increased mineralization lag time, increased osteoid surface area, low parathyroid hormone levels, and elevated serum calcium levels (ATSDR 2006).

Hematopoetic tissue

Among patients with chronic renal failure who receive dialysis treatment, some individuals will develop a hypochromic microcytic anemia, the severity of which correlates with the plasma and red blood cell aluminum levels and can be reversed by terminating exposure to aluminum or by aluminum chelation with desferrioxamine (Jeffery et al. 1996). While the mechanism for this effect in dialysis patients is not known, Jeffery et al. (1996) suggest that it may be aluminum interference with iron metabolism, possibly through disruption in cellular transfer of iron to ferritin to heme.

¹⁴ The LOELs and NOELs reported in this section may correspond to the doses reported by the researchers, or may be calculated based on reported concentrations in food or drinking water, assuming default values for animal body weight, and food and drinking water consumption rates drawn from Health Canada (1994).
¹⁵ The methodological limitations and uncertainties associated with this study are discussed in section 3.2.3.
¹⁶ The methodological limitations and uncertainties of the Varner et al. (1993) and Varner et al. (1998) studies are discussed in section 3.2.3.
¹⁷ The methodological limitations and uncertainties associated with the study by Huh et al. (2005) are discussed in section 3.2.3.
¹⁸ Cases of elevated aluminum exposure in dyalisis patients are rare, but are still occasionally reported.
See www.cdc.gov/mmwr/preview/mmwrhtml/mm5725a4.htm for a recent example.
¹⁹ International Classification of Disease (World Health Organization).
²⁰ Present exposure (i.e., exposure based on residence at the time of the study or at the time of diagnosis) may poorly characterize the exposure relevant to development of the disease, if the subject has moved frequently in the past, or in the case of a historical change in the water supply (i.e., change in water supply or treatment process).
²¹ Canadian Study of Health and Aging.
²² NINCDS is the National Institute of Neurological and Communicative Disorders and Stroke; ADRDA is the Alzheimer's Disease and Related Disorders Association; DSM is the Diagnostic and Statistical Manual of Mental Disorders (published by American Psychiatric Association).
²³ Meyer-Baron et al. (2007) propose a conversion factor of 1.17 to obtain µg Al/L from µg Al/g creatinine, determined as the mean of reported conversion factors between 0.71 and 1.61.
²⁴ Studies showing accumulation of aluminum in brain regions include Flora et al. (1991, 2003), Golub et al. (1992a), Lal et al. (1993), Varner et al. (1993, 1994, 1998), Florence et al. (1994), Gupta and Shukla (1995), Domingo et al. (1996), Sarin et al. (1997), Somova et al. (1997), Zheng and Liang (1998), Colomina et al. (1999), Kumar (1999), Swegert et al. (1999), Jia et al. (2001a), Baydar et al. (2003), Fattoretti et al. (2004), Jing et al. (2004), Abd-Elghaffar et al. (2005), Huh et al. (2005), Kaur et al. (2006) and Roig et al. (2006).

• 3.1 CEPA 1999 64(a) and 64(b) Environment
• 3.1.1 Environmental risk characterization
• 3.1.1.1 Aquatic organisms
• 3.1.1.1.1 Pelagic
• 3.1.1.1.2 Benthic
• 3.1.1.2 Terrestrial organisms
• 3.1.2 Other lines of evidence relating to aluminum salts
• 3.1.3 Sources of uncertainty
• 3.2 CEPA 1999 64(c): Human health
• 3.2.1 Estimated population exposure
• 3.2.1.1 Air
• 3.2.1.1.1 Estimated average daily intake of total aluminum in outdoor air
• 3.2.1.1.2 Estimated average daily intake of total aluminum in indoor air
• 3.2.1.2 Water
• 3.2.1.3 Soil
• 3.2.1.4 Foods
• 3.2.1.5 Overall estimate of exposure in the Canadian population
• 3.2.2 Hazard characterization
• 3.2.2.1 Effects in humans
• 3.2.2.2 Effects in experimental animals
• 3.2.3 Exposure-response analysis
• 3.2.3.1 Studies pertaining to other life stages
• 3.2.1.2 Identification of the level of concern and associated uncertainties
• 3.2.4 Human health risk characterization for total aluminum
• 3.2.5 Relative contribution of aluminum sulphate, aluminum chloride, and aluminum nitrate to overall exposure to aluminum
• 3.2.6 Uncertainties and degree of confidence in human health risk characterization
• 3.2.7 Recommendations for research
• 3.2.7.1 Exposure assessment
• 3.2.7.2 Exposure-response assessment
• 3.3 Conclusion

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 (i.e., 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 this 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 (e.g., OH^-, SO₄^2-) or organic (e.g., 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 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 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 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 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 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 (e.g., 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 (i.e., 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 (e.g., 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 (e.g., 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 (e.g., 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 (i.e., aluminum chloride, aluminum nitrate, aluminum sulphate, PAC, PASS, 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 polyaluminum chloride (PAC), aluminum chlorohydrate (ACH) and polyaluminum silicate sulphate (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.

3.2.1 Estimated Population Exposure

The average daily intake of aluminum in six age groups in Canada is estimated on the basis of concentrations measured in: (a) indoor and outdoor air (section 2.3.2.1); (b) drinking water (section 2.3.2.2.2); (c) soil (section 2.3.2.4); and (d) food (section 2.3.2.6). Table 3.1 shows the overall estimate of average daily intakes by age group and different environmental media (water, indoor air, outdoor (ambient) air, soil, and food and beverages) for total aluminum.

The average daily intake values were derived using a deterministic exposure assessment, which provides a single point estimate of intake (in this case and estimate of the mean). Probabilistic exposure assessments, on the other hand, provide information on the full range of possible intakes in the study population, and may, as well, give a more accurate estimate of mean exposure. The potential influence of a probabilistic analysis on the current assessment, with regard to the daily total aluminum intake in food, is discussed in more detail in section 3.2.1.4.

Consideration of the environmental media -- drinking water, air, soil and food -- in the derivation of the average daily intake is consistent with other assessments of priority substances. Daily intake of other sources of aluminum (e.g., antacids, vaccines and cosmetics) is difficult to quantify for the general Canadian population, both because of the limited data on exposure and absorption, and the variability in usage within the population. Therefore, these sources were not included in the estimation of the average daily intake. Similarly, occupational exposures to aluminum, such as inhalation exposure by aluminum welders, were not considered in the estimate of average daily intake. All of these additional sources may however, constitute non-negligible exposures to aluminum, and should be considered in the qualitative evaluation of uncertainty associated with the estimate of the average daily intake.

3.2.1.1 Air

3.2.1.1.1 Estimated average daily intake of total aluminum in outdoor air

The estimated average daily intake of total aluminum in airborne particles in outdoor air was determined using more than 10,000 measurements taken over the past ten years at some 50 sites in Canada. The average provincial/territorital total aluminum concentration of 0.17 µg/m³ in PM10 in Canada was used in the daily intake estimate (section 2.3.2.1.1). By age group, average daily intakes for PM10 were very low, ranging from 0.03 µg/kg bw/d for seniors to 0.1 µg/kg bw/d for young children aged six months to four years old.

3.2.1.1.2 Estimated Average Daily Intake of Total Aluminum in Indoor Air

In the case of indoor air, only measurements conducted on PM10 samples were evaluated to estimate intake since the concentration of aluminum in PM2.5 was often below the detection limit. The concentration based on the average daytime and nighttime concentrations of total aluminum is estimated to be 1.49 µg/m³ (section 2.3.2.1.2). The estimated average daily intake from indoor air is therefore higher than that from outdoor air, ranging from 0.3 µg/kg bw/d in adults and seniors to 0.8 µg/kg bw/d in young children aged six months to four years old.

3.2.1.2 Water

On the basis of data provided by municipal drinking water treatment plants from across Canada (section 2.3.2.2.2), the mean total aluminum concentration was estimated to be 101 µg/L. This estimate applies to plants that use coagulant/flocculents containing aluminum salts and secure their water supply from surface water sources. The average daily intake for each age group ranged from 2.0 µg/kg bw/d for adolescents and adults to 10.8 µg/kg bw/d for non-breastfed infants.

3.2.1.3 Soil

The mean total aluminum concentration in soil of approximately 41,000 mg/kg (section 2.3.2.4) was used to estimate the exposure of the Canadian population via soil. The average daily intake of aluminum from soil among infants was 166 µg/kg bw/d, and significantly higher in young children aged six months to four years old, who were found to have an estimated average daily intake of 268 µg/kg bw/d. For the other groups, the average daily intakes of total aluminum are progressively lower from 87 µg/kg bw/d for children aged 5 to 11 years old to 17 µg/kg bw/d for seniors.

3.2.1.4 Foods

For each age group defined in the Canadian population, the estimated mean dietary intake of total aluminum was derived using the fifth Total Diet Study completed in 2000-2002 (Dabeka 2007). Daily intakes of aluminum from food and beverages are presented in Table 3.1. For breastfed infants aged zero to six months old, the exposure to aluminum from human milk was approximately 12 µg/kg bw/d, whereas an intake of 85 µg/kg bw/d was calculated in non-breastfed infants. Among young children aged six months to four years old, the estimated mean daily intake from food was approximately 268 µg/kg bw/d. In the other groups, the mean daily intake of total aluminum ranged from 341 µg/kg bw/d in children aged 5 to 11 years old to 113 µg/kg bw/d in adults over 60 years old.

The above mean intake values of total aluminum in food were derived using a deterministic exposure assessment, which provides a single point estimate of intake but does not provide information about the full range of possible exposures within a population. The deterministic approach in this case is expected to overestimate mean estimates of exposure, in part because the aggregation of food categories inflates the contribution of less frequently consumed foods having higher levels of contamination. Further, the deterministic assessment does not take into account the day-to-day variability in the types of foods consumed by individuals.

Probabilistic exposure assessments estimate the probability of a given exposure in a population. The distribution of intakes that is generated provides more information about the full range of possible intakes in that population. Such statistical modelling can also account for intra- and interindividual variability in eating behaviours. As such, probabilistic exposure assessments, when the datasets are available to allow such assessments, are considered to provide a more accurate picture of exposure than deterministic exposure assessments.

Given that food is recognized to be the predominant source of aluminum exposure in humans, a more thorough evaluation of dietary exposure to aluminum using probabilistic techniques would be warranted (see section 3.2.7 on recommendations for research).

3.2.1.5 Overall Estimate of Exposure in the Canadian Population

The estimated mean daily intake of total aluminum was lower in breastfed than in non -breastfed infants, with levels of 179 and 262 µg/kg bw/d, respectively. The highest EDI of total aluminum was found in young children aged six months to four years old with 541 µg/kg bw/d, whereas for other age groups this intake decreased significantly to 432 µg/kg bw/d in children aged 5 to 11 years old, 293 µg/kg bw/d in adolescents, 163 µg/kg bw/d in adults aged 20 to 59 years old and finally 133 µg/kg bw/d in adults over 60 years old.

The contribution from various environmental media was evaluated for each of the age groups (Table 3.2). In young children aged six months to four years, approximately 50% of the aluminum intake was from food, 50% from ingestion of soil, and less than 1% from the ingestion of drinking water and inhaled particles. The contribution from the ingestion of food increased in the other age groups to 80% or more, whereas the contribution from soil decreased with age to 20% in children aged 5 to 11 years old and approximately 10% in the older age groups. The contribution from the ingestion of drinking water and inhaled particles is very low, at less than 2% or 0.2%, respectively for all age groups other than infants.

In infants, for the exclusively breastfed group, more than 90% of the total aluminum intake was found to be from the ingestion of soil and approximately 7% from the ingestion of human milk. For those infants who consumed infant formula and different food groups and beverages, approximately 30% of total aluminum intake was from the ingestion of food and about 63% from the ingestion of soil²⁵.

With respect to the three salts -- aluminum chloride, aluminum nitrate, and aluminum sulphate -- the only media in which the mean concentration is significantly affected by these the use of these salts is drinking water, in which aluminum sulphate or aluminum chloride may be added during the treatment process. While aluminum sulphate is permitted as an additive in some food products, this use is infrequent and would be expected to have a very minor influence on the total aluminum intake from food. The question of the relative contribution of the three salts to overall exposure to aluminum is discussed in more detail in section 3.2.5.

For those who regularly use aluminum-containing over-the-counter oral therapeutic products (e.g., pharmaceuticals such as antacids), these products represent the major source of daily aluminum intake. Based on the manufacturers' maximum recommended daily doses, EDIs of aluminum from these products may reach  approximately 31,000 µg/kg bw/d.

¹ Assumed to weigh 7.5 kg, to breathe 2.1 m³ of air per day, to drink 0.8 L of water per day (non breastfed) or 0 L of water per day (breastfed), and to ingest 30 mg of soil per day (Health Canada 1998a).
² Assumed to weigh 15.5 kg, to breathe 9.3 m³ of air per day, to drink 0.7 L of water per day and to ingest 100 mg of soil per day (Health Canada 1998a).
³ Assumed to weigh 31.0 kg, to breathe 14.5 m³ of air per day, to drink 1.1 L of water per day and to ingest 65 mg of soil per day (Health Canada 1998a).
⁴ Assumed to weigh 59.4 kg, to breathe 15.8 m³ of air per day, to drink 1.2 L of water per day and to ingest 30 mg of soil per day (Health Canada 1998a).
⁵ Assumed to weigh 70.9 kg, to breathe 16.2 m³ of air per day, to drink 1.5 L of water per day and to ingest 30 mg of soil per day (Health Canada 1998a).
⁶ Assumed to weigh 72.0 kg, to breathe 14.3 m³ of air per day, to drink 1.6 L of water per day and to ingest 30 mg of soil per day (Health Canada 1998a).
⁷ Based on the mean total aluminum concentration from all the drinking water treatment plants in Canada, estimated to be 101.16 µg/L (see section 2.3.2.2.2).
⁸ Based on dietary intake data from the fifth partial Total Diet Study in Canada (Dabeka 2007; see section 2.3.2.6). Data were adjusted for age categories from Health Canada (1998a). For breastfed infants, mean breast milk aluminum concentration of 0.11 mg/kg (section 2.3.2.6) was used, with a human milk density of 1.03 kg/L and an ingestion rate of 0.8 L/d.
⁹ Based on the mean concentration of total aluminum for all Canadian data in ambient air between 1986 and 2006, which is 0.17 µg/m³ in PM₁₀ (see section 2.3.2.1.1).
¹⁰ Based on average daytime and nighttime concentrations of all Canadian data in indoor air for total aluminum, which is about 1.49 µg/m³ (see section 2.3.2.1.2).
¹¹ Based on the mean concentration of total aluminum of 41,475 mg/kg measured in soils and sediments on the entire Canadian territory (see section 2.3.2.4).

3.2.2 Hazard Characterization

The discussion in this section focuses on the broad characterization of the types of effects of concern for the human health risk assessment of aluminum, on the basis of both human and experimental animal data. The suitability of the different sources of data for the exposure-response analysis, presented in section 3.2.3, is evaluated as well.

3.2.2.1 Effects in Humans

Aluminum has been shown to produce neurotoxic effects in humans as well as bone and blood toxicity, during medical treatment in which the gastrointestinal barrier is bypassed (e.g., aluminum-induced encephalopathy through dialysis treatment in patients with renal failure). There is also some epidemiological evidence for long-term cognitive impairment, in pre-term infants receiving aluminum-containing nutritional solution intravenously, and associated with occupational exposures, as discussed in section 2.4.3.1. These exposure conditions are not applicable to the general population, particularly as the exposure to aluminum generally does not occur via ingestion, and therefore human studies have not been used as a basis for characterizing the dose-response relationship for environmental exposures (see section 3.2.3). However, this evidence does support the identification of neurotoxicity and developmental neurotoxicity as endpoints of concern in the human health risk assessment for aluminum.

With respect to the conditions of exposure in the general population, the most relevant available information is provided by the epidemiological investigations into the association between exposure to aluminum through drinking water and AD and other forms of dementia (see section 2.4.3.2). The use of these findings for first identifying an endpoint of concern (i.e., hazard identification), and then for evaluating the exposure-response relationship is discussed below.

The hypothesis of aluminum in drinking water as a risk factor for AD or impaired cognitive function in the elderly is controversial in the scientific community, and has important implications for public health. Hence, it is important to evaluate in detail the weight of evidence for the observed associations, in the context of traditional criteria for causality. This evaluation, for studies published prior to 1998 is presented in the Guidelines for Canadian Drinking Water Quality - Technical Documents: Aluminum (Health Canada 1998b) and in the SOS report (Environment Canada and Health Canada 2000). In the SOS report the criteria of consistency and specificity, strength, dose-response, temporality, biological plausibility, and coherence of the observed association were evaluated, and the conclusion was as follows:

"Overall ... the weight of evidence for causality for the observed associations between aluminum and Alzheimer's disease is weak, at best. There is only limited consistency in the results of the analytical epidemiological studies. While the criteria for diagnosis were generally more stringent in the studies in which there was a positive outcome, there was more consistent control of potential confounding factors in the studies in which no associations were reported. Moreover, while there is some evidence of exposure-response in the individual available studies for the reported association between aluminum and Alzheimer's disease, there is little consistency in results among the different investigations in this respect, at least based on the limited extent of comparison permitted by the available data. There are also limited data to serve as a basis of the extent to which the observed association between aluminum and Alzheimer's disease meets the criterion of temporality. Most limiting, however, in the assessment of the weight of evidence for causality of the observed association is the lack of relevant data on biological plausibility; indeed, there is no hypothesized plausible pathway from exposure to effect with measurable key events, for which sufficient investigation has been conducted to assess weight of evidence against traditional criteria of causality, such as consistency, strength, specificity, dose-response, temporal patterns, biological plausibility and coherence."

Since the publication of the SOS report, a significant positive association between AD and aluminum in drinking water has been observed in the additional analysis of the data from the PAQUID cohort in southwestern France (Rondeau et al. 2000; Rondeau et al. 2001, as described in section 2.4.3.2). While the exposure assessment in this cohort study is improved in relation to previous case-control studies, it is still limited by two factors: the quantification of the aluminum exposure of individuals from other dietary sources and the relatively narrow range of aluminum exposure in the population studied.

Recent reviews of the epidemiological literature have reiterated the limitations of the epidemiological data base, in its entirety, in regard to the causality of the occurrence of aluminum in the environment and AD, while also maintaining that the hypothesis cannot be rejected at this time (InVS-Afssa-Afssaps 2003; ATSDR 2006; JECFA 2006; Krewski et al. 2007). As a result of these limitations, JECFA (2006) and ATSDR (2006) chose not to base their regulatory values for aluminum intake on epidemiological studies.

The epidemiological data on aluminum exposure in drinking water were not used in this assessment for developing the dose-response relationship (see section 3.2.3), because of the lack of evidence for a causal relationship between aluminum in drinking water and AD, and the lack of data on total exposure to aluminum, for which food is the predominant contributor. Nonetheless, the observed associations in some studies between aluminum in drinking water and the development of AD do support further consideration of neurotoxicity as an endpoint of concern in the human health risk assessment for aluminum.

3.2.2.2 Effects in experimental animals

The scientific community has primarily focused its investigations of aluminum toxicity on the endpoints of neurotoxicity and reproductive/developmental toxicity, principally because of the evidence from human case studies and epidemiological studies indicating that these effects may be of concern. A total of 138 toxicological studies, published from 1979 to 2007, reporting on neurotoxicity and reproductive/developmental effects of oral aluminum exposure in rodents, monkeys and dogs, have been evaluated for the present assessment.

The observations of the toxic effects of aluminum may be influenced by dose, aluminum salt, dosing regimen and exposure media as well as animal species and strain, age, sex, and health status. Considering the database evaluated for this assessment, the different studies vary with respect to all of these factors, and with respect to the specific endpoints investigated. Moreover, the majority of studies compare animals exposed at a single dose to a control group. In these single-dose studies, the dose corresponding to a lowest observed effect level (LOEL) or to a no observed effect level (NOEL) is strongly influenced by the researcher's choice of administered dose.

In 2000, in its SOS report, Health Canada summarized the experimental database on aluminum toxicity as follows (Environment Canada and Health Canada 2000):

"Altered performance in a variety of neurobehavioural tests and pathological and biochemical changes to the brain have been observed in studies of the oral administration (i.e., drinking water, diet, gavage) of aluminum salts to mice, rats and monkeys for varying periods of time as adults or during gestation, weaning and/or post-weaning. Interpretation of the results of a number of these studies is limited by designs that focus on testing specific hypotheses rather than examination of a range of neurotoxicity endpoints, the administration of single doses or a lack of an observed dose-response, lack of information on concentrations of aluminum or bioavailability from basal diets, the use of specific ligands to enhance accumulation of aluminum and small group sizes. Indeed, there have been no studies in which a broad range of neurological endpoints (biochemical, behavioural and histopathological) have been investigated in a protocol including multiple dose groups."

Since 2000 the database for neurological and reproductive/developmental endpoints has been considerably expanded. Yet the same limitations apply, most notably in regard to an emphasis on testing specific hypotheses rather than examining a range of neurotoxicity endpoints, testing of single doses or lack of an observed dose-response relationship, and small group sizes. There is no single study that has investigated multiple dose groups for a broad range of neurological endpoints²⁶.

The database does, however, provide a broad range of studies carried out by researchers from many different laboratories. Considered in its entirety, it gives evidence for neurological, neurodevelopmental and reproductive toxicity in experimental animals, including motor (e.g., rotarod test and grip strength), sensory (e.g., auditory startle) and cognitive effects (e.g., maze learning and passive avoidance tests) as well as neuropathological (e.g., neuronal degeneration), and biochemical changes (e.g., alterations in energy metabolism, trace element tissue concentrations and neurotransmission systems).

While no single or limited number of studies provides an adequate basis for characterizing the dose-response relationship, consideration of the database, as a whole, does provide a basis for approximately determining the lower range of doses at which researchers have repeatedly observed statistically significant changes in neurological, neurodevelopmental and/or reproductive endpoints in experimental animals orally exposed to aluminum salts.

3.2.3 Exposure-response analysis

The objective of the exposure-response analysis was to identify the lower range of doses for which oral exposure to aluminum has been shown to produce toxicologically significant effects in multiple studies. This lower range of doses was then considered as the level of concern for the human health risk characterization presented in section 3.2.4.

In order to characterize the lower range of doses at which oral exposure to aluminum produces effects in experimental animals, two subsets of the studies, based primarily on exposure period, were evaluated: (a) neurotoxic effects in adults following subchronic or chronic exposure (greater than 90 days); and (b) neurodevelopmental and reproductive effects in prenatal/lactation exposure studies. The studies included in these subsets are briefly described in Tables C1 and C2 (Appendix C). These two exposure periods were considered to be of greatest relevance to the evaluation of risks from long-term exposure to aluminum. Studies pertaining to other age categories (juvenile or older animals) are discussed separately in section 3.2.3.1.

These subsets include studies with highly diverse experimental conditions, notably with respect to the animal species and strain, type of aluminum salt administered, exposure vehicle as well as other aspects of the experimental methodology²⁷. There is also variability in the reporting of doses. Some researchers adjust the concentration in drinking water for a constant dose in mg Al/kg bw/d and report this value (e.g., Colomina et al. 2005; Colomina et al. 2002; Roig et al. 2006), while others estimate doses in terms of mg Al/kg bw/d based on measures of animal body weight and food and water intake, but keep the same concentration in the diet throughout the experiment (e.g., Golub and Germann 2001b; Golub et al. 2000). In other cases, the dose is reported only as a concentration administered via diet, drinking water or gavage, and the intake in mg Al/kg bw/d has been estimated using Health Canada (1994) reference values for animal body weight and intake.

In the case of the developmental studies, the LOELs are reported as the maternal dose at the beginning of gestation. In the studies where the concentration in drinking water or the diet remained constant, this dose would generally be lower than the received dose, due to increased food and water intake during gestation and lactation. For the purpose of human health risk assessment, however, the maternal dose at the beginning of pregnancy was considered, as this provided a common point of comparison between studies.

One condition that was applied to both subsets of studies was that the experimental administered dose constitutes the principal contribution to total aluminum. As previously discussed, the concentration of aluminum in standard laboratory rodent chow may be significant, contributing approximately 10 mg Al/kg bw/d in rats and 30 mg Al/kg bw/d in mice for a typical concentration of 250 ppm²⁸. In the majority of studies, this base diet concentration is not measured. Base diet concentration would considerably impact the exposure-response analysis if: (a) the bioavailability of the aluminum contained in the chow was of a similar magnitude to the bioavailability of the administered aluminum; and (b) the lab chow were to contribute a large percentage of the total aluminum exposure. While it could be hypothesised that the aluminum in the lab chow, associated with ligands in the food matrix, would be less soluble and therefore less bioavailable than added aluminum, no experimental data were identified to assess the relative bioavailabilities of aluminum in lab chow and added aluminum salts. Therefore, with regard to those studies where base diet was not quantified, studies were included in the two subsets only if the administered dose (D_a) likely exceeded the base diet dose (i.e., D_a > 10 mg Al/kg bw/d for rats and D_a > 30 mg Al/kg bw/d for mice). This approach limits the influence of the unknown base diet aluminum concentration on the exposure-response analysis, but does introduce a bias against inclusion of low dose studies in the exposure-response analysis²⁹. This issue is considered further in the discussion of uncertainties (section 3.2.3.2).

Other conditions applied in the compilation of these subsets were that the doses and other experimental conditions be reported unambiguously. In addition, in the subset of adult studies, studies of juvenile and older animals were not included. Studies based on these other exposure periods are discussed in section 3.2.3.1.

The LOELs of the studies meeting the conditions described above are presented graphically in Figure 3.1. In the four studies in which a LOEL for a specific endpoint is also associated with a NOEL, this is so indicated. Six other studies listed in Tables C1 and C2 found no effects for any endpoints measured (von Linstow Roloff et al. 2002; Domingo et al. 1996; Roig et al. 2006; McCormack et al. 1979; Colomina et al. 1994 and Katz et al. 1984). Consideration of these studies is important in assessing the consistency of the database and are included in the evaluation presented below. However, the studies are not included in Figure 3.1 as no corresponding LOELs for the endpoints were observed.

Considering the studies of Tables C1 and C2 collectively, the following observations concerning the exposure-response relationship for aluminum may be made:

• There is a wide variation in reported LOELs (from 1 to 663 mg Al/kg bw/d). As previously discussed, this variation would be expected, considering the diverse experimental conditions (species, strains, aluminum salt, dosing regimes, dosing vehicle, statistical power and endpoints measured).
• There is a predominance of single dose studies or studies where the LOEL was observed at the lowest dose. Thus, the LOELs in Figure 3.1 may be elevated with respect to the effect levels that might be observed in multiple dose studies.
• For the 16 subchronic and chronic exposure studies for neurotoxicity in adults, the LOELs range between 1 and 500 mg Al/kg bw/d (administered and combined doses -- D_a and D_c -- considered together). Among these studies the neurobehavioural endpoints examined included Morris water maze performance and impaired learning in the shuttle box as well as effects on reflex and motor activity. Biochemical endpoints included alterations in neurotransmission systems, increased apoptosis in the brain, alterations in synaptosomal membrane fluidity and increased lipid peroxidation in the brain.
• For the 22 studies of exposure during gestation and lactation, the LOELs (D_a and D_c) vary between 29 and 663 mg Al/kg bw/d. Neurobehavioural endpoints included grip strength, auditory startle, negative geotaxis and other reflexes, maze learning, thermal sensitivity, and motor development. The observed reproductive/developmental effects included a decrease in the number of corpora lutea and the number of implantation sites, a decrease in placental and fetal weight or reduced pup body weight, an increase in skeletal malformations, and an increase in the number of days to sexual maturity. In addition, alterations in essential element metabolism, deficits in synaptic plasticity in the hippocampus, a decrease in myelin sheath width as well as increased lipid peroxidation and a decrease in superoxide dismutase and catalase activity in the cerebrum and cerebellum were reported in developmental studies.

In order to estimate the lower range of doses at which oral exposure to aluminum produces toxicologically significant neurological or reproductive/developmental effects, the individual studies presented in Tables C1 and C2 were critically reviewed. The limitations of the collective database previously described -- including the use of a single exposure dose, examination of a limited number of endpoints, lack of information on base diet aluminum concentration and small group sizes -- often apply to these studies as well. Nonetheless, some of the studies provided stronger evidence than others for establishing the dose range at which neurological and reproductive/developmental effects may occur. The following discussion focuses particularly on studies documenting LOELs at the lowest doses, and evaluates the findings in relation to three issues: (a) use of a low administered dose; (b) toxicological significance of different endpoints; and (c) methodological strengths and limitations and consistency of study findings.

(a) Use of a low administered dose:

Of the studies included in Figure 3.1 the lowest LOEL was observed by Huh et al. (2005). This study reported apoptosis as well as the activation of the catalytic activity of monoamine oxidases A and B in the brains of Sprague-Dawley rats at a reported combined dose of 1 mg Al/kg bw/d. The aluminum-exposed group received aluminum maltolate in drinking water over a period of 12 months.

This study reported an aluminum concentration of 11.5 ppm in the base diet. Although this is a relatively low value for laboratory chow, it does constitute an aluminum dose (0.6 mg Al/kg bw/d) of nearly twice that of the administered dose (0.38 mg Al/kg bw/d). The use of an administered dose less than the base diet dose raises the question of exposure misclassification of individual animals, as the normal variability in intake between animals may create overlap between the two groups with respect to the dose received. This is considered to be a major limitation of this study.

In spite of the extremely low administered dose, the animals receiving aluminum maltolate were found, after one year, to have approximately four times the amount of aluminum in the brain (462 ng/g) as compared to the controls (110 ng/g)³⁰. This finding suggested a comparable increase in both the fraction of aluminum absorbed into the bloodstream and/or the amount of aluminum distributed to the brain when the aluminum is administered as the maltolate salt. Recently, Zhou et al. (2008) found differences in aluminum oral bioavailability, which were not statistically significant, between the citrate, maltolate and fluoride salts in drinking water. The measured bioavailabilities of all the salts were low (estimated means of 0.5%, 0.61% and 0.35% for maltolate, citrate and fluoride, respectively) and approximately twice the estimated bioavailability of aluminum in food (0.1% to 0.3%, as presented in Table 2.7) as measured with the same experimental protocol. These findings suggest that while aluminum maltolate may be more bioavailable, the increase would not be sufficient to explain the results of Huh et al (2006).

In light of the uncertainty associated with the reported increased brain concentrations in the Huh et al. (2005) study, in addition to the methodological limitation of testing an administered dose that is less than the base diet dose, the study by Huh et al. (2005) was not retained for the purpose of estimating the lower range of aluminum doses at which neurological effects may be expected to occur.

Other investigations with relatively low doses over periods of 12 weeks or longer have also reported neurotoxic effects. These studies were not considered in the exposure-response analysis as the aluminum content in the laboratory chow was not reported, and thus, unlike the study by Huh et al. (2005), the relative contribution of the aluminum in the base diet could not be evaluated. However, it should be noted that LOELs ranging from 0.07 to 22 mg Al/kg bw/d (administered dose) have been associated with a significant increase in brain aluminum levels as well as significant increases in neurobehavioural or histopathological effects (refer to Kaur and Gill 2006; Kaur et al. 2006; Varner et al. 1993; Varner et al. 1994; Varner et al. 1998; Somonova et al. 1997; Fleming and Joshi 1987; Kaneko et al. 2004; and Abd-Elgahaffar et al. 2005). These results were found for different species and for different aluminum salts, administered either in drinking water or by gavage. Thus, the possibility of toxicologically significant neurological effects in this low dose range cannot be discounted. However, the difficulty of interpreting the results of these studies underlines the importance of: (a) quantifying the aluminum content in the base diet and drinking water; and (b) using a purified low-aluminum diet in studies in which the administered dose is also very low.

Among the investigations mentioned above, the study findings with respect to aluminum fluoride are of particular concern, because of the presence of both of these ions in drinking water, either naturally or through addition during the treatment process. Varner et al. (1993), Varner et al. (1994) and Varner et al. (1998), in observing increased aluminum levels in the brain associated with a low administered aluminum fluoride dose, suggested that fluoride may enhance the uptake of aluminum by the brain. At present, the scientific database is very limited with respect to the toxicokinetics and health effects specific to aluminum fluoride.

(b) Toxicological significance of different endpoints:

Considering the 16 subchronic and chronic adult exposure studies, the LOELs range between 19 and 500 mg Al/kg bw/d (administered and combined doses -- D_a and D_c -- considered together, and excluding the Huh et al. (2005) study). For neurobehavioural endpoints (Morris water maze performance, impaired learning in the shuttle box and motor activity), the LOELs of the seven relevant studies vary between 40 to 500 mg Al/kg bw/d (D_a and D_c), with four studies having LOELs at D_as of 40 to 70 mg Al/kg bw/d (Commissaris et al. 1982; Lal et al. 1993; Gong et al. 2005; Mameli et al. 2006). The neurobehavioural endpoints examined constitute standard elements of neurobehavioural testing and impaired performance is considered to be toxicologically significant in the experimental animal.

The biochemical effects observed in the remaining studies included alterations in neurotransmission systems, alterations in synaptosomal membrane fluidity and increased lipid peroxidation in the brain, and were associated with LOELs varying from 19 to 420 mg Al/kg bw/d. These observations provide supportive evidence for neurotoxicity observed via other endpoints as well as information on mechanisms of action, but are more difficult to evaluate with respect to toxicological significance. For this reason, studies with these endpoints were given less weight in the exposure-response evaluation, in comparison to studies that include neurobehavioural endpoints.

Considering the 22 studies of exposure during gestation and lactation, the LOELs (D_a and D_c) varied between 29 and 663 mg Al/kg bw/d. For neurobehavioural endpoints (grip strength, auditory startle, negative geotaxis and other reflexes, maze learning and thermal sensitivity, and motor development), the LOELs (administered doses) ranged from 50 to 155 mg Al/kg bw/d, with the LOELs of two studies falling in the range of 50 to 60 mg Al/kg bw/d (Colomina et al. 2005; Golub and Germann 2001b).

With respect to reproductive parameters, the lowest LOEL was reported by Belles et al. (1999), where aluminum nitrate was administered to pregnant mice via gavage at a dose of 29 mg Al/kg bw/d and observed an increase in the number of early deliveries and reduced fetal body weight. Reduced birth or fetal weight was also observed by Colomina et al. (1992) and Sharma and Mishra (2006) at LOELs ranging between 50 and 70 mg Al/kg bw/d. Morphological effects in offspring were also observed in the latter two studies.

The motor, reflex and learning endpoints examined in the developmental studies as well as the reproductive parameters of fetal growth and morphological variations are all standard endpoints included in neurodevelopmental testing procedures, and considered to be toxicologically significant.

(c) Evaluation of methodology and consistency of results in studies with LOELs of less than 70 mg Al/kg bw/d:

The methodologies and findings of the abovementioned studies with LOELs of less than 70 mg Al/kg bw/d for neurobehavioural or reproductive/developmental endpoints were compared in order to characterize the strength of evidence for the effects observed at these dose levels. With respect to the neurobehavioural effects in adults at exposures greater than 90 days, four studies were evaluated: Mameli et al. (2006), Gong et al. (2005), Lal et al. (1993) and Commissaris et al. (1982). The reproductive/developmental studies included Sharma and Mishra (2006), Belles et al. (1999), Colomina et al. (1992), Colomina et al. (2005) and Golub and Germann (2001b). In addition, investigations in which NOELs were observed for these same endpoints are discussed.

Neurobehavioural effects in adults

Of the four neurobehavioural studies in adults, all were carried out in rats using aluminum chloride, in drinking water (Gong et al. 2005; Mameli et al. 2006; Lal et al. 1993), or in the diet (Commissaris et al. 1982), for periods varying between 90 days and 11 months.

Several weaknesses were identified in the investigations of Commissaris et al. (1982) and Gong et al. (2005). First, exposure information in these two reports was expressed as concentrations in the food or drinking water, and no information was included on intake rates or body weight of the animals. Thus the administered doses (50 and 60 mg Al/kg bw/d, respectively) were calculated on the basis of default intake and body weight values (refer to Health Canada 1994), and are therefore associated with greater uncertainty than had the doses been reported by the researchers on the basis of experimental observations. Moreover, the concentration of aluminum in the base diet was not reported in the two studies, and so the combined dose could not be calculated.

The investigations of Commissaris et al. (1982) and Gong et al. (2005) were also limited by the use of a single aluminum dose and the absence of a group receiving sodium chloride. Thus, a dose-response relationship could not be examined, and the observed effects could not be definitively attributed to the aluminum ion. It should be added that these two investigations were carried out with the primary objective of examining the influence of other test substances on aluminum toxicity -- parathyroid homone and Ginkgo biloba leaf extract, respectively -- and not for the purpose of evaluating aluminum toxicity at different dose levels for different endpoints.

In the study of Lal et al. (1993), adult male Druckrey albino rats were exposed to an administered dose of 52 mg Al/kg bw/d for 180 days in drinking water. Although this dose was not reported directly in this form, information on daily water consumption and average body weight was provided, allowing for calculation of the dose based on experimental data. The investigation included a range of behavioural, biochemical and histopathological endpoints. The researchers observed reduced spontaneous motor activity and impaired learning in the shuttle box and maze tests, in addition to increased lipid peroxidation and decreased Mg²⁺ and Na⁺K⁺-ATPase activities in the brain. The aluminum concentration in different brain regions was significantly increased in the aluminum-exposed animals, but no pathological alterations were observed.

In the context of evaluating the exposure-response relationship, the study by Lal et al. (1993) is more informative than the Commissaris et al. (1982) and Gong et al. (2005) studies, in that the dose is more accurately reported, brain aluminum content was measured and a range of endpoints were examined, with generally consistent findings reported for the different endpoints. Its limitations include the use of a single dose, the absence of a group exposed to sodium chloride and the lack of information on the aluminum concentration in the base diet. Assuming a concentration of 250 ppm of aluminum in the laboratory chow (ATSDR 2006), the corresponding approximate aluminum dose would be 13 mg Al/kg bw/d, leading to an estimated combined dose for the Lal et al (1993) study of 65 mg Al/kg bw/d.

It should be noted that NOELs for impaired learning in the maze and shuttle box tests in aluminum-exposed adults have been observed at doses of 100 and 140 mg Al/kg bw/d, respectively by Domingo et al. (1996) and VonLinstow Roloff et al. (2002). In the study by Domingo et al. (1996) the aluminum was administered to rats as Al nitrate, with added citrate, in drinking water for a period of 6.5 months. Von Linstow Roloff (2002) administered Al sulphate in drinking water to rats for a period of seven months.

Of these four studies, only Mameli et al. (2006) included more than one dose group, and were thereby able to establish a LOEL of 43 mg Al/kg bw/d and a NOEL of 22 mg Al/kg bw/d. At this administered dose the researchers found impairment of the vestibulo-ocular reflex in male rats of different ages (3, 10 and 24 months old) exposed to aluminum chloride in drinking water. Significant increases of aluminum were observed in brain regions (brainstem-cerebellum and cerebrum). This study, which used 20 animals per dose per age group, also included an exposure group for the salt, in this case sodium chloride, such that the observed effects could be more clearly attributed to the aluminum and not the chloride ion. It should be noted, however, that evidence from other studies supporting the effects of aluminum on the vestibulo-ocular reflex is not available, as this endpoint has not been evaluated by other researchers.

In the study by Mameli et al. (2006), the base diet aluminum concentration was measured but not clearly reported, nor was food intake measured. The LOEL of 43 mg Al/kg bw/d is thus the administered dose. The combined dose may be estimated at approximately 50 mg Al/kg bw/d, based on default values for rat dietary intake.

Considering the observations of LOELs and NOELs associated with neurobehavioural effects in adults as well as the probable combined doses, alterations in learning and reflexes may be observed at approximately 50 to 65 mg Al/kg bw/d, based on the LOELs of Mameli et al. (2006) and Lal et al. (1993) expressed as estimated combined dose.

Reproductive effects

With respect to reproductive effects, the lowest LOEL presented in Figure 3.1 is associated with the study of Belles et al. (1999). In this investigation, mice were exposed to aluminum nitrate via gavage from gestational day 6 to 15 at a dose of 29 mg Al/kg bw/d. In addition to the control group, one group received sodium nitrate at a similar nitrate dose. A high mortality (52%) in the aluminum-exposed pregnant mice was observed in this study, which was not observed in other developmental studies in which aluminum nitrate or other aluminum salts were administered at similar or greater doses. Other observations included reduced body weight gain in the dams during gestation and reduced fetal body weight. The number of early deliveries was also increased in the aluminum-exposed animals as compared to the control group, but there was no significant difference in this regard when compared to the sodium nitrate-exposed group.

This study is limited to a single dose, and the aluminum content in the base diet was not measured. The lack of information on base diet is particularly important in studies with mice because of their small body weight. A laboratory chow containing 250 ppm of aluminum would be equivalent to a dose of approximately 33 mg Al/kg bw/d, which is higher than the administered dose in this investigation.

Reduced maternal body weight gain and reduced fetal weight in aluminum-exposed animals were also observed at the LOELs associated with the Sharma and Mishra (2006) and Colomina et al. (1992) studies. A significant reduction in pup weight was also observed at the higher doses tested in the studies of Golub and Germann (2001b) and Colomina et al. (2005), at approximately 100 mg Al/kg bw/d.

In the study by Sharma and Mishra (2006), rats received 70 mg Al/kg bw/d as aluminum chloride via gavage during gestation and lactation. In addition to the effects on fetal weight, the authors observed an increase in skeletal malformations and in oxidative stress in the brains of mothers, fetuses and sucklings. The dose level in this study is based on the measured maternal weights. However, no information on base diet was included. The combined dose, based on a concentration of 250 ppm of aluminum in a typical lab chow and default values of Health Canada (1994), is estimated at approximately 83 mg Al/kg bw/d.

Colomina et al. (1992) administered aluminum lactate to mice through gavage. A LOEL of 57.5 mg Al/kg bw/d (administered dose) was observed for an increased incidence of morphological effects (cleft palate, delayed ossification of parietals), in addition to reduced fetal weight. This study did not report the aluminum content in the base diet. Considering the reported concentration in the laboratory chow used by this research group in other experiments of 42 ppm of aluminum, the estimated base diet dose would be approximately 5.5 mg Al/kg bw/d, based on Health Canada (1994) default values for body weight and food intake in mice. The combined dose would then be estimated at 63 mg Al/kg bw/d.

In contrast to the findings mentioned above, in the study of McCormack et al. (1979), rats were fed aluminum chloride in the diet at maternal dose levels of 25 and 50 mg Al/kg bw/d during gestation, and no differences in fetal growth or skeletal anomalies were observed. Colomina et al. (1994) found no differences in dam body weight, fetal growth or morphological variations in mice exposed via gavage to 104 mg Al/kg bw/d of aluminum hydroxide, during gestation. The latter finding may have resulted from the lower solubility and therefore the lower bioavailability of the hydroxide salt.

Considering the observations of LOELs and NOELs associated with reproductive effects, and the probable combined doses, reductions in fetal and pup body weight may be observed beginning at approximately 60 mg Al/kg bw/d (e.g., Colomina et al. (1992)). The study of Belles (1999), in which a LOEL of 29 mg Al/kg bw/d was observed for reduced fetal weight, is given less weight in this evaluation, in light of the uncertainty associated with the high maternal mortality rate observed in the exposed animals, and the elevated contribution of the base diet to aluminum exposure as compared to the administered dose.

Neurodevelopmental effects

With respect to neurodevelopmental effects, the lowest LOELs presented in Figure 3.1 are associated with the investigations of Colomina et al. (2005) and Golub and Germann (2001b). Both of these studies included exposure through gestation and lactation. The experimental conditions of the two studies, however, differed in many other respects, and these are described briefly below.

Colomina, Roig et al. (2005) exposed female Sprague-Dawley rats to 0, 50, or 100 mg Al/kg bw/d as aluminum nitrate in drinking water with citric acids, in combination with a base diet dose of approximately 3 mg Al/kg bw/d. Aluminum exposure was maintained through gestation, lactation and the life of the dams.

The maternal effects of aluminum administration included decreased food intake (with reduced body weight) during gestation and lactation and decreased water intake during lactation in the 100 mg Al/kg bw/d dose group. No effects were observed with respect to the length of gestation, the number of litters or the number of fetuses per litter. With respect to the pups, there was a significant increase in the number of days until sexual maturation in males in the 100 mg Al/kg bw/d dose group and in females at both 50 and 100 mg Al/kg bw/d. A significant reduction in forelimb grip strength in males was observed in the 100 mg Al/kg bw/d dose group on PND 11 compared controls.

In the water maze task, assessing spatial learning, the performance of aluminum treated rats (50 mg Al/kg bw/d) was significantly improved in comparison to the control group. The pups in the 100 mg Al/kg bw/d dose group were not tested in the water maze test, because of altered maternal food and water intakes in this group. No differences in aluminum-exposed animals were observed with respect to surface righting, negative geotaxis or activity in an open field. The authors also measured aluminum concentration in brain regions but did not find increased levels in any regions in the aluminum-exposed animals.

The study of Golub and Germann (2001b) investigated the long-term consequences of prenatal exposures to aluminum in Swiss Webster mice, in conjunction with a suboptimal base diet. The base diet was designed to simulate the usual diet of young women in the U.S., with respect to estimated phosphate, calcium, iron, magnesium, and zinc intakes. Following breeding, dams were exposed to aluminum in the diet as aluminum lactate. The doses were equivalent to approximately < 1, 10, 50 and 100 mg Al/kg bw/d, as estimated at the beginning of gestation.

The dams were exposed throughout gestation and lactation. Following weaning at 21 days, the pups were fed the same diet as the dams for two weeks (although the per kg dose levels were higher). No effects were observed in the number of dams completing pregnancy, gestation length, weight gain of the dams (GD0 to GD15), litter size or birth weight. By weaning, both males and females in the two highest dose groups weighed significantly less than the controls, although by PND35 only the highest dose group showed this effect.

The female offspring of the highest dose group (maternal exposure of 100 mg Al/kg bw/d) were found to be slower in maze learning at three months old, as indicated by longer latencies during the first three sessions of the four-session learning series. All aluminum treated groups were similar to controls by the fourth session. Differences in aluminum exposed groups were also observed in the cue relocation trials, in which average trial latency was significantly increased at the two highest dose levels (50 and 100 mg Al/kg bw/d) as compared to the control group.

In the motor testing of male offspring at five months old, males in the highest dose group (maternal exposure of 100 mg Al/kg bw/d) had significantly lower hindlimb grip strength and greater number of rotations in the rotarod test (animal losing footing). When body weight was taken into account, only the findings for the rotarod test remained significant.

The investigations by Colomina et al. (2005) and Golub and Germann (2001b) are methodologically superior in many respects to the majority of the studies described in Tables C1 and C2. Both include two dose levels in addition to the control group, quantify the aluminum dose associated with the base diet, and examine a range of reproductive and neurodevelopmental endpoints. The Colomina et al. (2005) study includes measurement of aluminum concentration in different brain regions. The Golub and Germann (2001b) study, however, used an experimental protocol designed to test the influence of a suboptimal diet, which limits comparisons of the findings with other investigations of aluminum toxicity, particularly as no groups were included with equivalent aluminum dose levels and a standard diet.

Interpretation of cognitive and motor test findings in the studies investigating the effects of aluminum exposure is also complicated by a possible biphasic dose-response relationship. For example, in the study by Roig et al. (2006), rats received aluminum nitrate in drinking water during gestation and lactation at administered doses of 50 and 100 mg Al/kg bw/d. No difference in the motor activity of aluminum-exposed pups and controls was found. However, the animals exposed to 50 mg Al/kg bw/d showed an improved performance in maze learning. The performance of animals exposed to 100 mg Al/kg bw/d was significantly reduced as compared to the animals exposed to 50 mg Al/kg bw/d, but not significantly different from controls. Colomina et al. (2005) also observed improved maze performance in aluminum-exposed animals, although the highest exposure group in that study was not tested for this endpoint.

Considering the neurodevelopmental studies described above, diminished performance in learning or motor tests may be observed in animals exposed prenatally or through lactation at maternal combined doses beginning at approximately 50 mg Al/kg bw/d. There is, however, considerable variability in various study results with respect to these endpoints, which also suggest a possible biphasic dose-response relationship in relation to maze learning.

3.2.3.1 Studies Pertaining to Other Life Stages

Some experimental animal studies have focused on life stages not included in the subsets discussed above. These are described below.

Golub and Keen (1999) investigated the effects of aluminum lactate administered in the diet to pubertal mice for four- or eight-week periods at doses of 17, 78, 122 and 152 mg Al/kg bw/d. A significant association between aluminum intake and reduced brain weight was observed in the four-week cohort at 152 mg Al/kg bw/d, but not in the eight-week cohort, suggesting that effects in young animals are reversible, even as exposure continues. There were no consistent effects, however, on startle response or grip strength.

Rajasekaran (2000) administered 53 mg Al/kg bw/d of aluminum chloride via gavage to male pubertal Wistar rats for 30 days. Testing at the end of the exposure period showed a decrease in spontaneous motor activity in the exposed rats, but no effect on motor coordination. Acetyl cholinesterase activity was decreased in the cerebrum but not the cerebellum or brain stem.

Fattoretti et al. (2004) administered aluminum chloride in drinking water to 22-month-old rats, at a dose of 31 mg Al/kg bw/d for six months. They observed an increase in trace elements and aluminum in brain regions, and an increase in the area occupied by the mossy fibres in the hippocampal CA3 zone. No neurobehavioural endpoints were examined in this study.

Colomina et al. (2002) administered aluminum nitrate in drinking water (with citric acid) for 114 days to rats who were 18 months old at the start of the experiment. The weighted dose over the four months was 94 mg Al/kg bw/d. They found a decrease in mean body weight in aluminum-exposed older rats but no differences in brain aluminum concentration. No effects were observed in the passive avoidance test or in open-field activity. However, the percentage of perforated synapses in the brain increased with age and aluminum exposure.

A recent study by Walton (2007a, 2007b) of rats exposed from 12 months to the end of life investigated neurotoxicity endpoints at combined doses of 0.4 and 1.6 mg Al/kg bw/d, doses simulating current estimated low-end and high-end human exposures. Two of the six rats in the high exposure group developed significant impairment in memory tests in old age, and the brains of these rats were examined with respect to aluminum loading and inhibition of PPP2 activity (a major phosphate-removing enzyme active against tau hyperphosphorylation)³¹. The study, limited by the small group size, did not report on differences between the two aluminum-exposure groups, and thus does not provide a basis for conclusions in regard to the relationship between observed biochemical and behavioural effects and aluminum exposure.

3.2.3.2 Identification of the Level of Concern and Associated Uncertainties

On the basis of the 43 studies presented in Tables C1 and C2, and considering additional studies on other age groups, it is recommended that a dose of 50 mg Al/kg bw/d, expressed as a combined dose of total aluminum, be considered as the level at which neurological and reproductive/developmental effects begin to be repeatedly observed in animal studies.

While the dose of 50 mg Al/kg bw/d is an estimation of the lower end of a broad range of LOELs observed under different experimental conditions, it is not considered to be an overly conservative estimate of the effect level of concern. As previously discussed, there are two sources of bias against consideration of lower values of LOEL in the above characterization: (a) low-dose studies were not considered if the administered dose was less than the probable base diet dose; and (b) LOELs from single-dose studies may be overestimates of the actual effect levels. The dose of 50 mg Al/kg bw/d has, however, produced neurotoxic, reproductive and developmental effects in laboratory animals more consistently under a wide range of experimental conditions, as compared to lower doses. This exposure level is therefore retained for the purpose of the characterization of human health risks as the level of concern for neurotoxic, neurodevelopmental and reproductive effects.

Figure 3.1: Compilation of the LOEL values from the two major subsets of studies (Adult exposure > 90 days and Reproductive/developmental) considered in the exposure-response analysis.

Click on image to enlarge.

The numbers represent the 38 studies in which LOELs were observed, as summarized in Tables C1 and C2, and listed below. Where the base diet aluminum level is quantified, the LOEL is expressed as combined dose. NOELs associated with LOELs are indicated when observed.

Study references and endpoints

Reproductive and developmental studies:

1. Bernuzzi et al. 1986: Reduced body weight of pups, impaired negative geotaxis.
2. Golub et al. 1987: Reduced birthweight, decreased body weight gain in pups.
3. Bernuzzi et al. 1989:
   1. Impaired locomotor coordination;
   2. Impaired righting reflex;
   3. Impaired grasping reflex.
4. Muller et al. 1990: Impaired negative geotaxis, impaired performance in suspension and locomotor coordination tests.
5. Gomez et al. 1991: Reduced fetal body weight, increase in skeletal variations.
6. Colomina et al. 1992: Maternal toxicity, reduced fetal body weight (aluminum lactate), increased incidence of morphological effects (aluminum lactate).
7. Misawa and Shigeta 1993: Maternal toxicity, decreased pup weight, delay in pinna detachment and eye opening in females, delayed development of auditory startle in males.
8. Golub et al. 1993: Effects on Mn metabolism.
9. Golub et al. 1994: Reduced auditory startle response.
10. Poulos et al. 1996: Delayed expression of phosphorylated high molecular weight neurofilament protein in tracts in diencephalon, maternal toxicity.
11. Golub et al. 1996: Lower retention of both Mn and Fe.
12. Verstraeten et al. 1998: Increased phospholipid and galactolipid contents in brain myelin, increased lipid peroxidation.
13. Llansola et al. 1999: Decrease in pup body weight, decreased number of cells in cerebellum, disaggregation of microtubules and neuronal death in cerebellar neuron cultures.
14. Belles et al. 1999: Increased mortality of dams and increased early deliveries, reduced fetal body weight.
15. Golub and Tarara 1999: Decreased myelin sheath width.
16. Golub et al. 2000: Reduced forelimb and hindlimb grip strength, decreased thermal sensitivity.
17. Golub and Germann (2001b):
    1. Impaired performance in rotarod test (males);
    2. Decreased weight gain in pups, impaired learning of maze with respect to cue utilization (females).
18. Wang et al. 2002a: Reduced body weight, deficits in synaptic plasticity in dentate gyrus of hippocampus.
19. Chen et al. 2002: Deficits in synaptic plasticity in dentate gyrus of hippocampus.
20. Nehru and Anand 2005: Increased lipid peroxidation, decreased superoxide dismutase and catalase activity in cerebrum and cerebellum.
21. Colomina et al. 2005:
    1. Reduced forelimb strength in males;
    2. Increased number of days to sexual maturation.
22. Sharma and Mishra 2006: Decreased number of corpora lutea, number of implantation sites, placental and fetal weight, increased skeletal malformations, increased oxidative stress in brains of mothers/fetuses and sucklings.

> 90 days exposure studies in adults:

1. Commissaris et al. 1982: Reduced motor activity, impaired learning (shuttle box).
2. Johnson et al. 1992: Decreased levels of microtubule associated protein-2 and spectrin in hippocampus.
3. Golub et al. 1992: Decreased motor activity, hindlimb grip strength and auditory and air puff startle responsiveness.
4. Lal et al. 1993: Reduced spontaneous motor activity; impaired learning (shuttle box, maze), increased brain lipid peroxidation, reduced Mg²⁺- and Na+K+-ATPase activities.
5. Florence et al. 1994: Cytoplasmic vacuolization in astrocytes and neurons.
6. Gupta and Shukla 1995: Increased lipid peroxidation in brain.
7. Zatta et al. 2002: Increased acetylcholinesterase activity.
8. Silva et al. 2002: Increased synaptosomal membrane fluidity, decreased cholesterol/phospholipid ratio in synaptosomes.
9. Flora et al. 2003: Evidence of increased lipid peroxidation in brain.
10. Jing et al. 2004: Impaired performance in Morris water maze, altered synapses in hippocampus and frontal cortex.
11. Gong et al. 2005: Impaired performance in Morris water maze.
12. Shi-Lei et al. 2005: Impaired performance in Morris water maze, decrease in long-term potentiation in hippocampal slices.
13. Silva et al. 2005: Decreased Na+/K+-ATPase activity in brain cortex synaptosomes.
14. Huh et al. 2005: Induced apoptosis in brain, increased efficiency of monoamine oxidases and increased level of caspase 3 and 12 in brain.
15. Rodella et al. 2006: Decreased nitrergic neurons in the somatosensory cortex.
16. Mameli et al. 2006: Impaired vestibulo-ocular reflex.

3.2.4 Human Health Risk Characterization for Total Aluminum

As discussed in the of this assessment, the human health risk characterization consists of a two-stage evaluation. In the first stage, exposure of the general Canadian population to total aluminum in air, drinking water, diet and soil is evaluated and compared with the estimated exposure level of concern. This provides a characterization of the human health risk in relation to total aluminum exposure from the four media.

In the second stage of the risk characterization, the relative contribution of each of the three listed salts (aluminum chloride, aluminum nitrate and aluminum sulphate) to this total aluminum exposure is evaluated.

The human health risk characterization for total aluminum exposure from water, air, soil and food is based on the comparison of the exposure level of concern, identified in the exposure-response analysis of section 3.2.3, and the average EDIs of different age groups in the Canadian population. The ratio of these two levels is generally referred to as the margin of exposure (MOE).

In Table 3.3, the exposure level of concern of 50 mg Al/kg bw/d (see section 3.2.3) is divided by the average EDI for each age group, to calculate the corresponding MOE. The MOEs ranged from 92 to 376 for the various age groups, with toddlers (0.5 to 4 years old) and children (5 to 11 years old) having the lowest MOEs.

¹ Values drawn from Table 3.1

The determination of the MOE for aluminum exposure in the Canadian population that would be considered adequately protective requires consideration of a range of issues, including the severity of the observed effects, the adequacy of the database, the uncertainty and variability associated with the toxicokinetics and toxicodynamics of aluminum as well as the presence of potentially sensitive sub-populations.

The adequacy of the collective database for the neurotoxicity and reproductive/developmental toxicity of orally-administered aluminum was reviewed in section 3.2.2.2. As discussed, there is a clear need for further investigation in experimental animals, in which studies are designed to provide a basis for determining a critical dose for risk assessment. The existing database is nonetheless extensive, providing a basis for the determination of the lower range of LOELs observed in the different studies, carried out under different experimental conditions and for an array of aluminum salts. The neurobehavioural and neurodevelopmental effects most frequently associated with the range of LOELs may be characterized as small but statistically significant changes in performance in motor activity and learning tests. However, as previously discussed, there are also sources of bias against consideration of lower LOELs in the above characterization.

Variability and uncertainty associated with the toxicokinetics and toxicodynamics of aluminum can be considered in the framework of the four chemical-specific adjustment factors generally evaluated in human health risk assessments of chemical substances (IPCS 2001; Meek et al. 2003). When possible, these factors are evaluated quantitatively on the basis of experimental data specific to the substance. When insufficient data are available, default values are used for the purpose of estimating either uncertainty factors in the development of guidelines, or of estimating adequately protective margins of exposure, as in the present assessment. In the case of aluminum neurotoxicity, there is little consensus as to the mode of action, and multiple mechanisms are likely involved, therefore the delineation of chemical-specific adjustment factors is not possible here.

To account for toxicokinetic and toxicodynamic variability and uncertainty, a factor of at least 100 within the MOE would be considered appropriate. An additional uncertainty factor of 10 could have been considered appropriate to use when a LOEL instead of a NOEL is employed; however, in this case, since effects at the lower-bound were generally small changes in performance in motor activity and learning tests, an additional uncertainty factor was not applied.

Other issues, beyond the uncertainty associated with the exposure level of concern of 50 mg Al/kg bw/d, are considered as well in the interpretation of the MOEs presented in Table 3.3:

• Since the EDIs correspond to average intakes, an MOE equal to 100 would imply that approximately half of the population is inadequately protected. Thus, any MOE value close to 100 raises potential concerns. In these cases a probabilistic analysis of intake would allow for a better definition of the level of risk for different segments of the population. In such analyses, the mean EDI is more accurately defined (and may in fact be lower than the mean EDI of a deterministic analysis), and the full distribution of intakes within the population can be evaluated with respect to the exposure level of concern.
• The level of the MOE judged to be acceptable could be reduced³¹ if there were evidence that the overall bioavailability of aluminum from the sources of exposure contributing to the EDI was significantly lower than the bioavailability of aluminum administered in the experimental animal studies (i.e., a relative bioavailability of less than one). This issue was discussed in section 2.3.3.1.6, where it was concluded that the available data do not provide sufficient evidence for relative bioavailabilities significantly different from one, either with respect to comparisons of humans and experimental animals, or with respect to comparisons of different environmental media.
• Sources of exposure (over-the-counter medications, antiperspirants, cosmetics and vaccines), described in sections 2.3.2.6 and 2.3.2.7, are not included in the calculation of the EDI since the estimates of intake would be either highly uncertain (as in the case of antiperspirants) or highly variable (as in the case of antacids) in the general population. For frequent users of antacids or other oral pharmaceuticals containing aluminum, total aluminum intake would likely be considerably higher than the estimated EDI, and consequently their actual margin of exposure could be significantly less than the MOE values presented in Table 3.3.

The MOEs presented in Table 3.3, in particular the MOE of 92 associated with toddlers (0.5 to 4 years old) and the MOE of 116 for children (5 to 11 years old), suggest that a proportion of children may be inadequately protective. Since the MOEs are calculated on the basis of average rather than maximum EDIs, there are segments of all age groups with potentially higher exposure levels. This would particularly include children whose diets include significant quantities of processed foods with high levels of aluminum-containing food additives.

3.2.5 Relative contribution of aluminum sulphate, aluminum chloride, and aluminum nitrate to overall exposure to aluminum

As noted in the Introduction (section 1) three aluminum salts are specifically named for assessment on the PSL2: chloride, nitrate and sulphate. Based on the uses of these three salts described in section 2.2.1, aluminum sulphate and chloride are principally used in water and wastewater treatment, although other minor uses were identified as well. Aluminum nitrate, used in far less quantities than the sulphate and chloride salts, is used in fertilizers and as a chemical reagent in various industries.

Although the data available for the assessment did not allow for quantification of exposure associated with specific salts, it is possible to qualitatively estimate their relative contribution to different environmental media (see Table 3.2). Generally, food is the principal source of exposure to total aluminum followed by soil, while exposure via drinking water and air combined is less than 2 % of total aluminum intake.

Since the major use of sulphate and chloride salts is in water treatment, exposure to these particular salts would be expected via drinking water. Aluminum nitrate has diverse uses, but its use is limited in comparison to the sulphate and chloride salts and is not expected to contribute significantly to aluminum in food and soil, the principle routes of aluminum exposure.

Two of the three salts specifically named on the PSL2 (chloride and nitrate) are not used as food additives. Aluminum sulphate (including its potassium and sodium salts) may be used as a food additive, but other aluminum-containing additives (basic and acidic sodium aluminum phosphate, sodium aluminosilicate) are much more widely used³². This was confirmed through recent information gathered by Health Canada's Food Directorate from those members of the food industry who manufacture products in which aluminum-based food additives are permitted. This information indicates that aluminum sulphate (and its salts) are used as food additives in a limited number of food items, such as muffins, pizza, tortilla, burritos, egg products and some dry bakery mixes, and in quantities less than 0.5% of the final product weight.

These considerations lead to the conclusion that the three aluminum salts on the PSL2 are not significant contributors to the principle media of total aluminum exposure, and hence are not significant contributors to the overall exposure in the general population.

3.2.6 Uncertainties and Degree of Confidence in Human Health Risk Characterization

There is a moderately high degree of confidence in the exposure assessment for aluminum, as it relates to the average external dose associated with food, drinking water, soil and air, and with respect to the intake of different age classes in the Canadian population. There is more uncertainty with respect to the maximum or high-end exposures in the population for the different media.

Food is the predominant source of exposure to aluminum for the general population. The estimates of aluminum concentrations in the diet are based on the most recent data available (2000-2002) and reflect the contribution of different food groups in the diet of Canadians across the country. The estimate of mean concentration of total aluminum in drinking water is based on over 10,000 samples of water from systems across Canada that use aluminum salts in their treatment process.

The data set for the concentrations of total aluminum in soil is comprehensive, including over 40 studies from different environments, primarily from the surface soil. It should be noted, however, that the variability of natural background levels of aluminum in soil is high, and so the variability in intake would also be high, particularly for the toddler age group.

There is reasonable confidence in the estimated concentration of aluminum in ambient air, although data for indoor air for Canadian homes was not available (U.S. data used). However, the contribution of both indoor and outdoor air to total aluminum exposure is negligible. Data gaps with respect to air concentrations would therefore have little influence on the overall exposure estimate.

The exposure estimates, expressed as average EDIs, do not provide a quantitative measure of interindividual variability in the population. Major sources of interindividual variability would be different dietary choices (e.g., processed foods containing aluminum salts as food additives), and consumption of aluminum-containing non-prescription medications. Probabilistic exposure analyses could reduce this uncertainty, as well as providing a more accurate estimate of the mean exposure.

The greatest uncertainty with respect to the exposure assessment is the uncertainty and variability relating to the extent to which different aluminum salts are absorbed from the different media. Although some experimental bioavailability data are available for food and water, collectively the limited aluminum bioavailability data do not indicate that the relative bioavailabilities of aluminum in drinking water, soil and different types of food are significantly different. However, further research in this area, particularly in regard to soil, could provide evidence for significant differences that would in turn influence the human health risk characterization.

With respect to the exposure-response analysis, there is a moderate degree of confidence in the determination of the lowest exposure level associated with neurobehavioural and reproductive/developmental effects under diverse experimental conditions, using different salts, different animal species and different exposure regimes. There is however, considerable uncertainty associated with the interpretation of these results in relation to human exposure, in part because of the potential differences in the bioavailability of aluminum in the human diet as compared to the bioavailability of the aluminum administered in experimental animal studies. On the other hand, the data from epidemiological studies, which would be more applicable to human exposure, were found to be inadequate for establishing the dose-response relationship for environmental exposures in the general population.

In conclusion, the degree of confidence in the average aluminum exposure estimates for different age groups of the Canadian population is moderately high with respect to the concentration of aluminum in different environmental media and the contribution of the different media to cumulative aluminum exposure. Overall, however, the degree of confidence associated with this assessment is more modest, in light of the significant uncertainties in regard to the bioavailability of different aluminum species in different media and with respect to the exposure-response evaluation as it applies to human exposure conditions.

3.2.7 Recommendations for Research

Areas for further research are described briefly below, in order to identify the main avenues for reducing the uncertainties associated with the human health risk assessment for aluminum.

3.2.7.1 Exposure Assessment

Given the importance of food in the total exposure to aluminum, further quantitative analysis of food items with elevated aluminum levels and their associated consumption patterns in different subpopulations in Canada is warranted. A probabilistic analysis of the intake data would provide more information about the full range of possible intakes in the Canadian population. In addition, it would be important to distinguish aluminum food additives from natural aluminum sources in foods, in order to provide a more informed basis for developing a risk management strategy.

Consideration of bioavailability is important to the characterization of human health risks of aluminum if relative bioavailabilities for different exposure media and different species (i.e., humans and experimental animals) differ from unity. This hypothesis could be explored through the determination of bioaccessibilities of aluminum in aluminum-treated drinking water, different soil and dust samples, in selected food items (e.g., processed cheese and packaged bakery items), and in laboratory animal chow, followed by the comparison of these in vitro bioaccessibilities with the in vivobioavailability of aluminum determined in experimental studies for a given media.

In light of the wide use of aluminum-containing products applied to the skin, the dermal absorption of aluminum in humans should be more adequately quantified.

3.2.7.2 Exposure-Response Assessment

Further epidemiological study of aluminum exposure in the Canadian population is called for, to the extent that such research addresses the limitations of previous studies, including the characterization of aluminum exposure by dietary and other sources.

Additional experimental animal studies on toxicokinetics of different salts, including aluminum fluoride as well as the neurological and neurodevelopmental effects of aluminum, should be undertaken, in order to provide information for characterizing the exposure-response relationship. Following OECD guidelines for neurotoxicity and neurodevelopmental toxicity, these studies would include adequate numbers of animals, multiple doses, and examination of a standard array of neurological and neurodevelopmental endpoints. Note that one such study is currently underway in Canada, sponsored by a consortium of aluminum salt producers.

________________________________________

²⁵ Soil would most likely be in the form of household dust for this age group.
²⁶ A good laboratory practice (GLP) study generally following OECD and U.S. EPA Developmental Neurotoxicity guidelines, commissioned by a consortium of aluminum salt producers, is currently underway. The results, however, will not be available before mid-2009.
²⁷ Further categorization of the studies, based on salt administered, animal species, exposure vehicle and a more precisely defined exposure period, was considered but found to be not feasible. Narrowly defined subgroups did not provide an adequate number of studies with common endpoints and dose ranges. On the other hand, the comparison of pooled studies (e.g., drinking water studies vs. dietary administration studies), in order to determine the relative importance of different experimental variables, is limited by the confounding between these variables. Researchers tend to chose similar sets of experimental conditions from one experiment to another. Thus differences in the LOELs observed in a series of studies might be attributed to a particular factor (e.g., drinking water vs. diet) but could also be the result of the researchers' choices to repeatedly use the same single dose of the same salt, in the same exposure vehicle (diet or drinking water). Likewise, evaluation of pools of single-dose studies can mask the influence of an experimental condition, as reported LOELs may be poor estimates of real effect levels.
²⁸ See discussion in section 2.4.4 on typical levels of aluminum in lab chow.
²⁹ The low-dose studies for adult exposure, in which base diet aluminum concentration is not reported, include findings of altered levels of neurotransmitters (Silva and Goncalves 2003; Dave et al. 2002; Bilkei-Gorzo 1993), of changes in the phospholipid content of synaptic plasma membrane (Pandya et al. 2001) or of increased lipid peroxidation in the brain (Kaneko et al. 2004, Pratico et al. 2002, Abd-Elghaffar et al. 2005). Some low-dose studies also documented increased neuronal damage (Varner et al. 1998, 1993; Somova et al. 1997; Abd-Elghaffar et al. 2005) and neuromotor and coordination effects (Bilkei-Gorzo 1993; Sahin et al. 1995). The low-dose prenatal/lactation exposure studies included findings of alterations in neurotransmission (Kim 2003; Ravi et al. 2000) and effects on fetal growth (Paternain et al. 1988; Domingo et al. 1987a).
³⁰ In contrast, Colomina et al. (2002) administered aluminum nitrate, enhanced with citrate, in drinking water, at an average dose of 94 mg Al/kg bw/d, to groups of male rats aged 21 days and 18 months old. The increase in whole brain aluminum concentration in the aluminum-exposed group was not statistically significant. Roig et al (2006) observed an increase of aluminum in brain regions of rats exposed to 100 mg Al/kg bw/d of aluminum nitrate with citrate in drinking water for one year. Observations were made in two-year-old rats, and increases were on the order of three- to ten-fold, depending on the brain region, and with a 22-fold increase in the striatum.
³¹ Neurofibrillary tangles in AD brains are formed from the hyperphosphorylation of tau protein.

CEPA 1999 64(a) and 64 (b): Based on the available data, it is proposed that the three aluminum salts, aluminum chloride, aluminum nitrate and aluminum sulphate, are not entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity or that constitute or may constitute a danger to the environment on which life depends.

CEPA 1999 64(c): Based on available data concerning the exposure of the Canadian population to aluminum chloride, aluminum nitrate and aluminum sulphate, and in consideration of the health effects observed in humans and in experimental animals, it is proposed that these aluminum salts are not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health.

It is therefore proposed that the three aluminum salts, aluminum chloride, aluminum nitrate and aluminum sulphate, do not meet the definition of "toxic" as set out in section 64 of CEPA 1999.

Notwithstanding the above proposed conclusion, the human health risk assessment carried out on the basis of total aluminum exposure suggests thata proportion of individuals, particularly in the younger age groups, may be exposed to levels of total aluminum that, when compared to the exposure level of concern (derived from studies in experimental animals measuring neurotoxic effects), results in margins that may not be adequate.  For most age groups, food is estimated to be the primary source of exposure to aluminum.  In addition, sub-populations exposed to aluminum through the use of aluminum-containing over-the-counter oral therapeutic products would likely have higher exposure to aluminum.

As the three aluminum salts on the Priority Substances List are not significant contributors to Canadians' exposure via the principal media of total aluminum exposure, they are not considered to be significant contributors to the overall health risk.

Food additives in Canada are regulated under the Food and Drugs Act. The Food Directorate of Health Canada is currently re-evaluating dietary exposure to aluminum and reviewing current uses of aluminum-containing food additives, following a decision by the Joint Expert Committee on Food Additives (JECFA) of the World Health Organization to reduce the Provisional Tolerable Weekly Intake (PTWI) for aluminum to 1.0 mg/kg bw. As part of this project, the Food Directorate is conducting a probabilistic assessment of aluminum exposure via foods using the latest data from the Canadian Total Diet Study and from the Canadian Community Health Survey - Cycle 2.2 on Nutrition. The probabilistic exposure assessment will provide a more accurate estimate of the full range of dietary intakes of aluminum, including the proportion of the population that may be exceeding the PTWI. The results of this project will serve as the basis for updating the regulatory provisions (including the maximum levels of use) for aluminum-containing food additives. Further information can be found via Health Canada's Food Additives webpage: http://www.hc-sc.gc.ca/fn-an/securit/addit/index-eng.php.

Abd el-Fattah AA, al-Yousef HM, al-Bekairi AM, al-Sawaf HA. 1998. Vitamin E protects the brain against oxidative injury stimulated by excessive aluminum intake. Biochem Mol Biol Int 46(6): 1175-1180.

Abd-Elghaffar S, El-Sokkary GH, Sharkawy AA. 2005. Aluminum-induced neurotoxicity and oxidative damage in rabbits: protective effect of melatonin. Neuro Endocrinol Lett 26(5): 609-616.

Abercrombie DE, Fowler RC. 1997. Possible aluminum content of canned drinks. Toxicol Ind Health 13(5): 649-654.

Adams WJ, Conard B, Ethier G, Brix KV, Paquin PR, DiToro D. 2000. The challenges of hazard identification and classification of insoluble metals and metal substances for the aquatic environment. HERA 6: 1019-1038.

Adler AJ, Berlyne GM. 1985. Duodenal aluminum absorption in the rat: effect of vitamin D. Am J Physiol 249(2 Pt 1): G209-213.

[ATSDR] Agency for Toxic Substances and Disease Registry. 2006. Draft toxicological profile for aluminum. September 2006 draft for public comments. Washington (DC): U.S. Department of Health and Human Services, Public Health Service. Available from: http://www.atsdr.cdc.gov/toxprofiles/tp22.html

Akila R, Stollery BT, Riihimaki V. 1999. Decrements in cognitive performance in metal inert gas welders exposed to aluminium. Occup Environ Med 56(9): 632-639.

Al Moutaery K, Al Deeb S, Biary N, Morais C, Ahmad Khan H, Tariq M. 2000. Effect of aluminum on neurological recovery in rats following spinal cord injury. J Neurosurg 93(2 Suppl): 276-282.

[AEC] Alberta Environmental Centre. 1984. Alum sludge treatment and disposal. Edmonton (AB): Underwood McLellan Ltd. for Alberta Environment.

[AEC] Alberta Environmental Centre. 1987. Binding, uptake and toxicity of alum sludge. Edmonton (AB): Alberta Environment, Standards and Approvals Division. 117 p.

Alexopoulos E, McCrohan CR, Powell JJ, Jugdaohsingh R, White, KN. 2003. Bioavailability and toxicity of freshly neutralized aluminium to the freshwater crayfish Pacifastacus leniusculus. Arch Environ Contam Toxicol 45: 509-514.

Alfrey AC. 1993. Aluminum and renal disease. Contrib Nephrol 102: 110-124.

Allen DD, Orvig C, Yokel RA. 1995. Evidence for energy-dependent transport of aluminum out of brain extracellular fluid. Toxicology 98: 31-39.

Allen DD, Yokel RA. 1992. Dissimilar aluminum and gallium permeation of the blood-barrier demonstrated by in vivo microdialysis. J Neurochem 58: 903-908.

Almer B, Dickson W, Ekstrom C, Homstrom E, Miller, U. 1974. Effects of acidification on Swedish lakes. Ambio 3: 30-36 [cited in Driscoll CT, Schecher WD. 1990].

Alstad NEW, Kjelsberg BM, Vøllestad LA, Lydersen E, Poléo ABS. 2005. The significance of water ionic strength on aluminium toxicity in brown trout (Salmo trutta L.). Environ Pollut 133: 333-342.

Alva AK, Edwards DG, Blamey FPC. 1986. Relationships between root length of soybean and calculated activities of aluminium monomers in nutrient solution. Soil Sci Soc Am J 50: 959-962.

Alvarez-Hernandez X, Adigosky SR, Stewart B, Glass J. 1994. Iron status affects aluminum uptake and transport by caco-2 Cells. J Nutr 124: 1574-1580.

Amador FC, Santos MS, Oliveira CR. 1999. Lipid peroxidation facilitates aluminum accumulation in rat brain synaptosomes. J Environ Health A 58: 427-435.

Appelberg M. 1985. Changes in haemolymph ion concentration of Astacus astacus L. and Pacifastacus leniusculus (DANA) after exposure to low pH and aluminium. Hydrobiologia 121: 19-25.

Aremu DA, Meshitsuka S. 2005. Accumulation of aluminum by primary culturel astrocytes from aluminum amino acid complex and its apoptotic effect. Brain Res 1031(2): 284-296.

Ares J. 1986. Identification of aluminum species in acid rain forest soil solutions on the basis of Al:F reaction kinetics. II. An example at the Solling area. Soil Sci 142: 13.

Ascherio A, Zhang S, Hernan M, Olek M, Coplan P, Brodovicz K. 2001. Hepatitis B vaccination and the risk of multiple sclerosis: case-control studies. Gastroenterol Clin Biol 25(10): 927-929.

Baker JP. 1982. Effects on fish of metals associated with acidification. In: Johnston J, editor. Acid rain/fisheries. Proceedings of an International Symposium on Acidic Precipitation and Fisheries Impact. American Fisheries Society, Bethesda, Maryland.

Baker JP, Bernard DP, Christensen SW, Sale MJ, Freda J, Heltcher K, Marmorek D, Rowe L, Scanlon P, Suter G, Warren-Hicks W, Welbourn P. 1990. Biological effects of changes in surface water acid-base chemistry. NAPAP Report 13. In: National Acid Precipitation Assessment Program, Acidic deposition: state of science and technology. Vol. II. Oak Ridge (TN): Oak Ridge National Laboratory, Environmental Sciences Division. p. 13.47-13.61 (Publication No. 3609).

Barcelo J, Guevara P, Poschenrieder C. 1993. Silicon amelioration of aluminum toxicity in teosinte (Zea mays L. ssp. mexicana). Plant Soil 154: 249-255.

Barnes LM. 1985. Some characteristics of primary sludges derived from physico-chemically treated sewage. Water Pollut 84: 502-514.

Bast-Pettersen R, Drablos PA, Goffeng LO, Thomassen Y, Torres CG. 1994. Neuropsychological deficit among elderly workers in aluminum production. Am J Ind Med 25(5): 649-662.

Basu S, Das Gupta R, Chaudhuri AN. 2000. Aluminium related changes in brain histology: protection by calcium and nifedipine. Indian J Exp Biol 38(9): 948-950.

Bataineh H, Al-Hamood M, Elbetieha A. 1998. Assessment of aggression, sexual behavior and fertility in adult male rat following long-term ingestion of four industrial metal salts. Hum Exp Toxicol 17: 570-576.

Baydar T, Nagymajtenyi L, Isimer A, Sahin G. 2005. Effect of folic acid supplementation on aluminum accumulation in rats. Nutrition 21(3): 406-410.

Baydar T, Papp A, Aydin A, Nagymajtenyi L, Schulz H, Isimer A, Sahin G. 2003. Accumulation of aluminum in rat brain: does it lead to behavioral and electrophysiological changes? Biol Trace Elem Res 92(3): 231-244.

Becaria A, Lahiri DK, Bondy SC, Chen D, Hamadeh A, Li H, Taylor R, Campbell A. 2006. Aluminum and copper in drinking water enhance inflammatory or oxidative events specifically in the brain. J Neuroimmunol 176(1): 16-23.

Becking G, Priest N, D. 1997. Is aluminium in drinking water a neurotoxic risk to humans? In: Press MU. Managing Health int Aluminium Industry. London. p. 300-307.

Beisinger KE, Christensen GM. 1972. Effects of various metals on survival, growth, reproduction and metabolism of Daphnia magna. J Fish Res Bd Canada 29: 1691-1700.

Bélanger N, Fyles H, Hendershot W. 1999. Chemistry, bioaccumulation and toxicity in the terrestrial environment -- PSL2 assessment of aluminum salts. Montreal (QC): Environment Canada. Unpublished report.

Belles M, Albina ML, Sanchez DJ, Domingo JL. 1999. Lack of protective effects of dietary silicon on aluminium-induced maternal and developmental toxicity in mice. Pharmacol Toxicol 85(1): 1-6.

Bellia JP, Birchall JD, Roberts NB. 1996. The role of silicic acid in the renal excretion of aluminium. Ann Clin Lab Sci 26(3).

Bendell-Young L, Chouinard J, Pick FR. 1994. Metal concentrations in chironomids in relation to peatland geochemistry. Arch Environ Contam Toxicol 27: 186-194.

Bendell-Young L, Pick FR. 1995. Contrasting the geochemistry of aluminum among peatlands. Water Air Soil Pollut 81: 219-240.

Benson WH, Alberts JJ, Allen HE, Hunt CD, Newman MC. 1994. Synopsis of discussion Session on the bioavailability of inorganic contaminants. In: Hamelink JL, Landrum PF, Bergman HL, Benson WH, editors, Bioavailability: Physical, chemical and biological interactions. Proceedings of the 13th Pellston workshop, Michigan, August 17-22, 1992, Boca Raton (FL): Lewis Publishers. p. 63-72.

Bergerioux C, Boisvert J. 1979. Rapid neutron activation method for the determination of minerals in milk. Int J Nucl Med Biol 6(2): 128-131.

Bergman JJ, Boots BF. 1997. Alum sludge management at Buffalo Pound WTP. Report prepared by Buffalo Pound Water Treatment Plant for Environment Canada. Gatineau (QC): Environment Canada. Unpublished report.

Berkowitz J, Anderson MA, Graham RC. 2005. Laboratory investigation of aluminum solubility and solid-phase properties following alum treatment of lake waters. Water Res 39: 3918-3928.

Bernuzzi V, Desor D, Lehr PR. 1986. Effects of prenatal aluminum exposure on neuromotor maturation in the rat. Neurobehav Toxicol Teratol 8(2): 115-119.

Bernuzzi V, Desor D, Lehr PR. 1989a. Effects of postnatal aluminum lactate exposure on neuromotor maturation in the rat. Bull Environ Contam Toxicol 42(3): 451-455.

Bernuzzi V, Desor D, Lehr PR. 1989b. Developmental alternations in offspring of female rats orally intoxicated by aluminum chloride or lactate during gestation. Teratology 40(1): 21-27.

Bertrand A, Robitaille G, Boutin R, Nadeau P. 1995. Growth and ABA responses of maple seedlings to aluminum. Tree Physiol 15: 775-782.

Bertsch PM. 1990. The hydrolytic products of aluminum and their biological significance. Environ Geochem Health 12: 7-14.

Bertsch PM, Parker DR.  1996. Aqueous polynuclear aluminum species. In: Sposito G, editor, The environmental chemistry of aluminum. 2nd edition. Boca Raton, Florida: CRC Press p. 117-168.

Besser JM, Brumbaugh WG, May TW, Church SE, Kimball BA. 2001. Bioavailability of metals in stream food webs and hazards to brook trout (Salvelinus fontinalis) in the Upper Animas River watershed, Colorado. Arch Environ Contam Toxicol 40: 48-59.

Bia MJ, Cooper K, Schnall S, Duffy T, Hendler E, Malluche H, Solomon L. 1989. Aluminum induced anemia: pathogenesis and treatment in patients on chronic hemodialysis. Kidney Int 36(5): 852-858.

Bilkei-Gorzo A. 1993. Neurotoxic effect of enteral aluminium. Food Chem Toxicol 31(5): 357-361.

Birchall JD, Bellia JP, Roberts NB. 1996. On the mechanisms underlying the essentiality of silicon- interactions with aluminium and copper. Coord. Chem. Rev. 149: 231-240.

Birkeland PW. 1984. Soils and geomorphology. New York (NY): Oxford University Press. 372 p.

Bishop NJ, Morley R, Day JP, Lucas A. 1997. Aluminum neurotoxicity in preterm infants receiving intravenous-feeding solutions. N Engl J Med 336(22): 1557-1561.

Booth CE, McDonald DG, Simons BP, Wood CM. 1988. Effects of aluminum and low pH on net ion fluxes and ion balance in brook trout (Salvelinus fontinalis). Can. J. Fish. Aquat. Sci. 45: 1563-1574.

Bowdler NC, Beasley DS, Fritze EC, Goulette AM, Hatton JD, Hession J, Ostman DL, Rugg DJ, Schmittdiel CJ. 1979. Behavioral effects of aluminum ingestion on animal and human subjects. Pharmacol Biochem Behav 10(4): 505-512.

Boyce BF, Eider HY, Fell SG, Nicholson WA, Smith GD, Dempster DW, Gray CC, Boyle IT. 1981. Quantitation and localisation of aluminum in human cancellous bone in renal osteodystrophy. Scan Electron Microsc 3: 329-337.

Braul L, Viraraghavan T, Corkal D. 2001. Cold water effects on enhanced coagulation of high DOC, low turbidity water. Water Quality Research Journal of Canada 36(4): 701-717.

Brezonik PL, Mach CE, Downing G, Richardson N, Brigham M. 1990. Effects of acidification on minor and trace metal chemistry in Little Rock Lake, Wisconsin. Environ Toxicol Chem 9: 871-885.

Brown DJA. 1981. The effects of various cations on the survival of brown trout, Salmo trutta, at low pHs. J. Fish Biol 18: 31-40.

Brown BA, Driscoll CT.  1992. Soluble aluminum silicates: stoichiometry, stability and implications for environmental geochemistry. Science 256: 1667-1670.

Budaveri S, O'Neil MJ, Smith A, Heckelman PE, editors. 1989. The Merck index: An encyclopedia of chemicals, drugs, and biologicals. Rahway (NJ): Merck & Co.

Cameron RS, Ritchie GSP, Robson AD. 1986. Relative toxicities of inorganic aluminium complexes to barley. Soil Sci Soc Am J 50: 1231-1236.

Campbell A, Becaria A, Lahiri DK, Sharman K, Bondy SC. 2004. Chronic exposure to aluminum in drinking water increases inflammatory parameters selectively in the brain. J Neurosci Res 75(4): 565-572.

Campbell PGC. 1995. Interactions between trace metals and organisms: a critique of the free-ion activity model. In: Tessier A, Turner DR, editors, Metal speciation and bioavailability. Chichester: J. Wiley & Sons. p. 45-102.

Campbell PGC, Hansen HJ, Dubreuil B, Nelson WO. 1992. Geochemistry of Quebec North Shore salmon rivers during snowmelt: organic acid pulse and aluminum mobilization. Can. J. Fish. Aquat. Sci. 49: 1938-1952.

Campbell PGC, Lewis AG, Chapman PM, Crowder AA, Fletcher WK, Imber B, Luoma SN, Stokes PM, Winfrey M. 1988. Biologically available metals in sediments. NRCC 27694. Ottawa: National Research Council of Canada. 298 p.

Campbell PGC, Stokes PM. 1985. Acidification and toxicity of metals to aquatic biota. Can. J. Fish. Aquat. Sci. 42: 2034-2049.

Canada. 2000. Canadian Environmental Protection Act: Persistence and Bioaccumulation Regulations, P.C. 2000-348, 23 March, 2000, SOR/2000-107, Canada Gazette. Part II, vol. 134, no. 7, p. 607-612.

[CCME] Canadian Council of Ministers of the Environment. 1998. Protocol for the derivation of Canadian tissue residue guidelines for the protection of wildlife that consume aquatic biota. Winnipeg (MB): Canadian Council of Ministers of the Environment.

[CCME] Canadian Council of Ministers of the Environment. 2003. Canadian water quality guidelines for the protection of aquatic life. Aluminium. In: Canadian environmental quality guidelines. Winnipeg (MB): Canadian Council of Ministers of the Environment.

[CCME] Canadian Council of Ministers of the Environment. 2008. Canada-wide strategy for the management of municipal wastewater effluent. Available at: www.ccme.ca/ourwork/water.html?category_id=81.

[CSHA] Canadian Study of Health and Aging. 1994. The Canadian Study of Health and Aging: Risk factors for Alzheimer's disease in Canada. Neurology 44: 2073-2080.

Caramelo CA, Cannata JB, Rodeles MR, Fernandez Martin JL, Mosquera JR, Monzu B, Outeirino J, Blum G, Andrea C, Lopez Farre AJ and others. 1995. Mechanisms of aluminum-induced microcytosis: lessons from accidental aluminum intoxication. Kidney Int 47(1): 164-168.

Cheminfo Services Inc. 2008. Characterization and analysis of aluminum salts and releases to the environment in Canada. Final report. Vancouver (BC): Prepared for Environment Canada, Environmental Stewardship Branch.

Chen J, Wang M, Ruan D, She J. 2002. Early chronic aluminium exposure impairs long-term potentiation and depression to the rat dentate gyrus in vivo. Neuroscience 112(4): 879-887.

Cherroret G, Bernuzzi V, Desor D, Hutin MF, Burnel D, Lehr PR. 1992. Effects of postnatal aluminum exposure on choline acetyltransferase activity and learning abilities in the rat. Neurotoxicol Teratol 14(4): 259-264.

Cherroret G, Capolaghi B, Hutin M-F, Burnel D, Desor D, Lehr PR. 1995. Effects of postnatal aluminum exposure on biological parameters in the rat plasma. Toxicol Lett 78: 119-125.

Chiba J, Kusumoto M, Shirai S, Ikawa K, Sakamoto S. 2002. The influence of fluoride ingestion on urinary aluminum excretion in humans. Tohoku J Exp Med 196(3): 139-19.

Choinière J, Beaumier M. 1997. Bruits de fond géochimiques pour différents environnements géologiques au Québec.

Church SE, Kimball BA, Frey DL, Ferderer DA, Yager TJ, Vaughn RB. 1997. Source, transport, and partitioning of metals between water, colloids, and bed sediments of the Animas River, Colorado. U.S. Geological Survey Open-File Report 97-151 [cited in Farag et al. 2007].

City of Ottawa. 2002.  City Facilities Environmental Effects Monitoring Project, Summary Report: Study of Ottawa's Water Purification and Wastewater Treatment Plants.  Prepared by the Water Environment Protection Program, Utilities Services, Transportation and Public Works.

Clark KL, Hall RJ. 1985. The effects of elevated hydrogen ion and aluminum concentrations on the survival of amphibian embryos and larvae. Can J Zool. 63: 116-123.

Clayton CA, Perritt RL, Pellizzari ED, Thomas KW, Whitmore RW, Wallace LA, Ozkaynak H, Spengler JD. 1993. Particle Total Exposure Assessment Methodology (PTEAM) study: distributions of aerosol and elemental concentrations in personal, indoor, and outdoor air samples in a southern California community. J Expo Anal Environ Epidemiol 3(2): 227-250.

Clayton RM, Sedowofia SK, Rankin JM, Manning A. 1992. Long-term effects of aluminium on the fetal mouse brain. Life Sci 51(25): 1921-1928.

Cochran M, Chawtur V, Phillips J, Dilena B. 1994. Effect of citrate infusion on urinary aluminium excretion in the rat. Clin Sci 86: 223-226.

Colomina MT, Gomez M, Domingo JL, Corbella J. 1994. Lack of maternal and developmental toxicity in mice given high doses of aluminium hydroxide and ascorbic acid during gestation. Pharmacol Toxicol 74(4-5): 236-239.

Colomina MT, Gomez M, Domingo JL, Llobet JM, Corbella J. 1992. Concurrent ingestion of lactate and aluminum can result in developmental toxicity in mice. Res Commun Chem Pathol Pharmacol 77(1): 95-106.

Colomina MT, Roig JL, Sanchez DJ, Domingo JL. 2002. Influence of age on aluminum-induced neurobehavioral effects and morphological changes in rat brain. Neurotoxicology 23(6): 775-781.

Colomina MT, Roig JL, Torrente M, Vicens P, Domingo JL. 2005. Concurrent exposure to aluminum and stress during pregnancy in rats: Effects on postnatal development and behavior of the offspring. Neurotoxicol Teratol 27(4): 565-574.

Colomina MT, Sanchez DJ, Sanchez-Turet M, Domingo JL. 1999. Behavioral effects of aluminum in mice: influence of restraint stress. Neuropsychobiology 40(3): 142-149.

Commissaris RL, Cordon JJ, Sprague S, Keiser J, Mayor GH, Rech RH. 1982. Behavioral changes in rats after chronic aluminum and parathyroid hormone administration. Neurobehav Toxicol Teratol 4(3): 403-410.

Confavreux C, Suissa S, Saddier P, Bourdes V, Vukusic S. 2001. Vaccinations and the risk of relapse in multiple sclerosis. Vaccines in Multiple Sclerosis Study Group. N Engl J Med 344(5): 319-326.

Connor DJ, Harrell LE, Jope RS. 1989. Reversal of an aluminum-induced behavioral deficit by administration of deferoxamine. Behav Neurosci 103(4): 779-783.

Connor DJ, Jope RS, Harrell LE. 1988. Chronic, oral aluminum administration to rats: cognition and cholinergic parameters. Pharmacol Biochem Behav 31(2): 467-474.

Connor JN, Martin MR. 1989. An assessment of sediment phosphorus inactivation, Kezar Lake, New Hampshire. Water Resources Bulletin 25(4): 845-853.

Cornwell DA, Burmaster JW, Francis JL, Friedline JCJ, Houck C, King PH, Knocke WR, Novak JT, Rolan AT, San Giacomo R. 1987. Committee report: research needs for alum sludge discharge. Sludge Disposal Committee. J Am Water Works Assoc 79: 99-104.

Corrales I, Poschenrieder C, Barcelo J. 1997. Influence of silicon pretreatment on aluminium toxicity in maize roots. Plant Soil 190: 203-209.

Courchesne F, Hendershot WH. 1997. La genèse des podzols. Géographie physique et quaternaire 51: 235-250.

Cournot-Witmer G, Zinngraff J, Plachot JJ, Escaig F, Lefevre R, Boumati P, Bourdeau A, Garadedian M, Galle P, Bourdon R and others. 1981. Aluminium localization in bone from hemodialyzed patients: Relationship to matrix mineralization. Kidney Int 20: 375-378.

Cox AE, Camberato JJ, Smith BR. 1997. Phosphate availability and inorganic transformation in an alum sludge-affected soil. J Environ Qual 26: 1393-1398.

Crane M, Whitehouse P, Comber S, Ellis J, Wilby R. 2005. Climate change influences on environmental and human health chemical standards. Human and Ecological Risk Assessment 11: 289-318.

Cranmer JM, Wilkins JD, Cannon DJ, Smith L. 1986. Fetal-placental-maternal uptake of aluminum in mice following gestational exposure: effect of dose and route of administration. Neurotoxicology 7(2): 601-608.

Cronan CS, Driscoll CT, Newton RM, Kelly JM, Schofield CL, Bartlett RJ, April R. 1990. A comparative analysis of aluminum biogeochemistry in a northeastern and a southeastern forested watershed. Water Resour Res 26: 1413-1430.

Cronan CS, Schofield CL. 1990. Relationships between aqueous aluminum and acidic deposition in forested watersheds of North America and Europe. Environ Sci Technol 24: 1100-1105.

Cronan CS, Walker WJ, Bloom PR. 1986. Predicting aqueous aluminum concentrations in natural waters. Nature (London) 324: 140-143.

Cumming JR, Weinstein LH. 1990. Al mycorrhizal interactions in physiology of pitch pine seedling. Plant Soil 125: 7-18.

Cunat L, Lanhers MC, Joyeux M, Burnel D. 2000. Bioavailability and intestinal absorption of aluminum in rats: effects of aluminum compounds and some dietary constituents. Biol Trace Elem Res 76(1): 31-55.

Dabeka B. 2007. Draft - Résultats préliminaires des données brutes de l'étude de la diète totale - 2000 à 2002. Ottawa.

Dann T. 2007. Output for PM2.5 and PM10 Al for 1986-2006. Personnal communication with Tom Dann. In: Analysis and Air Quality, Environment Canada. Ottawa.

Dave G. 1985. The influence of pH on the toxicity of aluminum, cadmium, and iron to eggs and larvae of the zebrafish, Brachydanio rerio. Ecotoxicol Environ Saf 10: 253-267.

Dave G. 1992. Sediment toxicity and heavy metals in 11 lime reference lakes of Sweden. Water Air Soil Pollut 63: 187-200.

Dave KR, Syal AR, Katyare SS. 2002. Effect of long-term aluminum feeding on kinetics attributes of tissue cholinesterases. Brain Res Bull 58(2): 225-233.

David MB, Driscoll CT. 1984. Aluminum speciation and equilibria in soil solutions of a haplorthod in the Adirondack Mountains (New York, U.S.A.). Geoderma 33: 297-318.

Day JP, Barker J, King SJ, Miller RV, Templar J. 1994. Biological chemistry of aluminium studied using ²⁶Al and accelerator mass spectrometry. Nucl Instrum Methods Phys Res 92: 463-468.

DeWald LE, Sucoff E, Ohno T, Buschena C. 1990. Response of northern red oak (Quercus rubra) seedlings to soil solution aluminium. Can J For Res. 20: 331-336.

Dillon PJ, Evans HE, Scholer PJ. 1988. The effects of acidification on metal budgets of lakes and catchments. Biogeochemistry 3: 201-220.

Divine KK, Lewis JL, Grant PG, Bench G. 1999. Quantitative particle-induced X-ray emission imaging of rat olfactory epithelium applied to the permeability of rat epithelium to inhaled aluminum. Chem Res Toxicol (7): 575-581.

Do P. 1999. Personal communication with Environment Canada. Engineering and Environmental Services, City of Calgary, Alberta.

Domingo JL, Gomez M, Sanchez DJ, Llobet JM, Corbella J. 1993. Effect of various dietary constituents on gastrointestinal absorption of aluminum from drinking water and diet. Res Commun Chem Pathol Pharmacol 79(3): 377-380.

Domingo JL, Llorens J, Sanchez DJ, Gomez M, Llobet JM, Corbella J. 1996. Age-related effects of aluminum ingestion on brain aluminum accumulation and behavior in rats. Life Sci 58(17): 1387-1395.

Domingo JL, Paternain JL, Llobet JM, Corbella J. 1987a. Effects of oral aluminum administration on perinatal and postnatal development in rats. Res Commun Chem Pathol Pharmacol 57(1): 129-132.

Donald JM, Golub MS, Gershwin ME, Keen CL. 1989. Neurobehavioral effects in offspring of mice given excess aluminum in diet during gestation and lactation. Neurotoxicol Teratol 11(4): 345-351.

Driscoll CT, Baker JP, Bisogni JJ, Schofield CL. 1980. Effect of aluminum speciation on fish in dilute acidified waters. Nature 284: 161-164.

Driscoll CT, Bisogni JJ. 1984. Weak acid/base systems in dilute acidified lakes and streams in the Adirondack region of New York State. In: Schnoor JL, editor, Modeling of total acid precipitation impacts. Boston: Butterworth. p. 53-72.

Driscoll CT, Postek KM. 1996. The chemistry of aluminum in surface waters. In: Sposito G, editor, The environmental chemistry of aluminum. 2nd edition. Boca Raton (FL): CRC Press. p. 363-418.

Driscoll CT, Schecher WD. 1988. Aluminum in the environment. In: Sigel H, Sigel A, editors, Metal ions in biological systems. Aluminum and its role in biology. Vol. 24. New York (NY): Marcel Dekker. p. 59-122.

Driscoll CT, Schecher WD. 1990. The chemistry of aluminum in the environment. J. Environ Perspect Health 12: 28-49.

Driscoll CT, Otton JK, Inverfeldt A. 1994. Trace metals speciation and cycling. In: Moldan B, Cerny J, editors, Biogeochemistry of small catchments: a tool for environmental research. New York (NY): Wiley and Sons. pp. 299-322.

Driscoll CT, van Breemen N, Mulder J. 1985. Aluminum chemistry in a forested spodosol. Soil Sci Soc Am J. 49: 437-444.

Drueke TB, Jouhanneau P, Banide H, Lacour B, Yiou F, Raisbeck G. 1997. Effects of silicon, citrate and the fasting state on the intestinal absorption of aluminium in rats. Clin Sci (Lond) 92(1): 63-67.

Duan J, Wang J, Graham N, Wilson F. 2002. Coagulation of humic acid by aluminium sulphate in saline water conditions. Desalination 150: 1-14.

Dudka S, Ponce-Hernandez R, Hutchinson TC. 1995. Current level of total element concentrations in the surface layer of SUdbury's soils. Sci Total Environ 162: 161-171.

Duis K, Oberemm A. 2001. Aluminium and calcium - key factors determining the survival of vendace embryos and larvae in post-mining lakes? Limnologica 31: 3-10.

Dunn CE. 1990. Lithogeochemistry study of the Cretaceous in Central Saskatchewan - Preliminary Report. Saskatchewan Geological Survey 90-4: 193-197.

Durand I, Keller JM, Cherroret G, Colins S, Dauca M, Lehr PR. 1993. Impact d'une intoxication aluminique postnatale précoce sur l'épithélium dodénal de rat: Etudes en microscopie électronique à transmission et en microanalyse de rayons X. Bulletin de l'Académie et de la Société lorraines des sciences 65(1): 3-19.

Dussault EB, Playle RC, Dixon DG, McKinley RS. 2001. Effects of sublethal, acidic aluminum exposure on blood ions and metabolites, cardiac output, heart rate, and stroke volume of rainbow trout, Oncorhynchus mykiss. Fish Physiol Biochem 25: 347-357.

DuVal G, Brubb B, Bentley P. 1986. Aluminum accumulation in the crystalline lens of humans and domestic animals. Trace Elem Med 3: 100-104.

Edwardson JA, Moore PB, Ferrier IN, Lilley JS, Newton GW, Barker J, Templar J, Day JP. 1993. Effect of silicon on gastrointestinal absorption of aluminium. Lancet 342(8865): 211-212.

Eickhoff TC, Myers M. 2002. Workshop summary. Aluminum in vaccines. Vaccine 20 Suppl 3: S1-4. El-Demerdash FM. 2004. Antioxidant effect of vitamin E and selenium on lipid peroxidation, enzyme activities and biochemical parameters in rats exposed to aluminium. J Trace Elem Med Biol 18(1): 113-121.

El-Rahman SS. 2003. Neuropathology of aluminum toxicity in rats (glutamate and GABA impairment). Pharmacol Res 47(3): 189-194.

Environment Canada. 1995. National Analysis of Trends in Emergencies System (NATES) database output.

Environment Canada. 1996. Ecological risk assessments of Priority substances under the Canadian Environmental Protection Act. Resource document. Gatineau (QC): Environment Canada, Commerical Chemicals Branch. Available on request.

Environment Canada. 1997. Notice respecting the second Priority Substances List and di(2-ethylhexyl) phthalate. Canada Gazette, Part I, February 15, 1997. p. 366-368.

Environment Canada. 2008a. Summary of Provincial and Territorial Management Activities Pertaining to Drinking Water Treatment Plant Waste Discharge to Surface Waters in Canada.  Unpublished table.  Existing Substances Division, September 2008. Available on request.

Environment Canada. 2008b. National Enforcement Management Information System and Intelligence System (NEMISIS) database output.

Environment Canada 2008c. Personal communication with Dixon Weir, Director of Water and Wastewater Services, City of Ottawa.

Environment Canada and Health Canada. 2000. State of the science report for aluminum chloride, aluminum nitrate and aluminum sulphate. Ottawa (ON): Environment Canada, Health Canada.

Environmental & OHP. 2005. Étude sur la santé dans la région de Belledune. Area Health study - Appendix A - Human Health RIsk assessment. Department of Health and Wellness, Government of New Brunswick.

European Commission. 2000a. IUCLID (International Uniform Chemical Information Database) Dataset. Existing Chemical Substance ID: 7446-70-0. Aluminium chloride (PDF Format, 232KB). European Chemicals Bureau. Accessed at http://ecb.jrc.ec.europa.eu/iuclid-datasheet/7446700.pdf on July 26, 2007.

European Commission. 2000b. IUCLID Dataset. Existing Chemical Substance ID: 10043-01-3. Aluminium sulphate (PDF Format, 117KB). European Chemicals Bureau. Accessed at http://ecb.jrc.ec.europa.eu/iuclid-datasheet/10043013.pdf on July 26, 2007.

[EFSA] European Food Safety Authority. 2008. Safety of aluminium from dietary intake. Scientific Opinion of the Panel on Food Additives, Flavourings, Processing Aids and Food Contact Materials (AFC). The EFSA Journal 754: 1-4.

Exley C. 2006. Aluminium-adsorbed vaccines. Lancet Infect Dis 6(4): 189.

Exley C, Burgess E, Day JP, Jeffery EH, Melethil S, Yokel RA. 1996. Aluminum toxicokinetics. J Toxicol Environ Health 48(6): 569-584.

Farag AM, Nimick DA, Kimball BA, Church SE, Harper DD, Brumbaugh WG. 2007. Concentrations of metals in water, sediment, biofilm, benthic macroinvertebrates, and fish in the Boulder River watershed, Montana, and the role of colloids in metal uptake. Arch Environ Contam Toxicol 52: 397-409.

Farag AM, Woodward DF, Little EE, Steadman B, Vertucci FA. 1993. The effects of low pH and elevated aluminum on Yellowstone cutthroat trout (Oncorhynchus clarki bouvieri). Environ Toxicol Chem 12: 719-731.

Fattoretti P, Bertoni-Freddari C, Balietti M, Giorgetti B, Solazzi M, Zatta P. 2004. Chronic aluminum administration to old rats results in increased levels of brain metal ions and enlarged hippocampal mossy fibers. Ann N Y Acad Sci 1019: 44-47.

Fattoretti P, Bertoni-Freddari C, Balietti M, Mocchegiani E, Scancar J, Zambenedetti P, Zatta P. 2003. The effect of chronic aluminum(III) administration on the nervous system of aged rats: clues to understand its suggested role in Alzheimer's disease. J Alzheimers Dis 5(6): 437-444.

Fiejka MA, Fiejka E, Dlugaszek M. 1996. Effect of aluminum hydroxide administration on normal mice: tissue distribution and ultrastructural localization of aluminum in liver. Pharmacol Toxicol 78: 123-128.

Findlow JA, Duffield JR, Williamns DR. 1990. The chemical speciation of aluminum in milk. Chem Spec Bioavail 2: 3-32.

Flarend R, Bin T, Elmore D, Hem SL. 2001. A preliminary study of the dermal absorption of aluminium from antiperspirants using aluminium-26. Food Chem Toxicol 39(2): 163-168.

Flaten TP. 1990. Geographical associations between aluminium in drinking water and death rates with dementia (including Alzheimer's disease), Parkinson's disease and amyotrophic lateral sclerosis in Norway. Aluminium and disease. Norway. p 152-167.

Flaten TP, Glattre E, Viste A, Sooreide O. 1991. Mortality from dementia among gastroduodenal ulcer patients. J Epidemiol Community Health 45(3): 203-206.

Fleming J, Joshi JG. 1987. Ferritin: isolation of aluminum-ferritin complex from brain. Proc Natl Acad Sci U S A 84(22): 7866-7870.

Flora SJ, Dhawan M, Tandon SK. 1991. Effects of combined exposure to aluminium and ethanol on aluminium body burden and some neuronal, hepatic and haematopoietic biochemical variables in the rat. Hum Exp Toxicol 10(1): 45-48.

Flora SJ, Mehta A, Satsangi K, Kannan GM, Gupta M. 2003. Aluminum-induced oxidative stress in rat brain: response to combined administration of citric acid and HEDTA. Comp Biochem Physiol C Toxicol Pharmacol 134(3): 319-328.

Florence AL, Gauthier A, Ponsar C, Van den Bosch de Aguilar P, Crichton RR. 1994. An experimental animal model of aluminium overload. Neurodegeneration 3(4): 315-323.

Foley CM, Polinsky MS, Gruskin AB, Baluarte HJ, Grover WD. 1981. Encephalopathy in infants and children with chronic renal disease. Arc Neurol 38: 656-658.

Forbes WF, Agwani N. 1994. Geochemical Risk factors for mental functioning, based on the Ontario Longitudinal study of aging (LSA) III. The effects of different aluminum-containing compounds. La revue canadienne du vieillissement. Can J Aging 13(4): 488-498.

Forbes WF, Agwani N, Lachmaniuk P. 1995a. Geochemical Risk factors for mental functioning, based on the Ontario Longitudinal study of aging (LSA) IV. The role of silicone-containing compounds. La revue canadienne du vieillissement. Can J Aging 14(4): 630-641.

Forbes WF, Gentleman JF, Lessard S. 1995b. Geochemical Risk factors for mental functioning, based on the Ontario Longitudinal study of aging (LSA) V. Comparisons of the results, relevant to aluminum water concentrations, obtained from the LSA and from Death Certificates mentioning Dementia. La revue canadienne du vieillissement. Can J Aging 14(4): 642-656.

Forbes WF, Hayward LM, Agwani N. 1992. Geochemical Risk factors for mental functioning, based on the Ontario Longitudinal study of aging (LSA) I. Results from a preliminary investigation. Can J Aging II(3): 269-280.

Forbes WF, Hayward LM, Agwani N, McAiney CA. 1994. Geochemical Risk factors for mental functioning, based on the Ontario Longitudinal study of aging (LSA) II. The role of pH. Can J Aging 13(3): 249-267.

Forbes WF, McLachlan DR. 1996. Further thoughts on the aluminum-Alzheimer's disease link. J Epidemiol Community Health 50(4): 401-403.

Forster DP, Newens AJ, Kay DW, Edwardson JA. 1995. Risk factors in clinically diagnosed presenile dementia of the Alzheimer type: a case-control study in northern England. J Epidemiol Community Health 49(3): 253-258.

Fortin C, Campbell PGC. 1999. Calculs de spéciation pour l'aluminium rejeté en eaux courantes. Montréal (QC) : Environment Canada, Environmental Protection Service.

Fraga CG, Oteiza PI, Golub MS, Gershwin ME, Keen CL. 1990. Effects of aluminum on brain lipid peroxidation. Toxicol Lett 51(2): 213-219.

France RL, Stokes PM. 1987. Influence of Mn, Ca and Al on hydrogen ion toxicity to the amphipod Hyalella azteca. Can J Zool. 65: 3071-3078.

Frecker MF. 1991. Dementia in Newfoundland: identification of a geographical isolate? J Epidemiol Community Health 45(4): 307-311.

Freda J. 1991. The effects of aluminum and other metals on amphibians. Environ Pollut 71: 305-328.

Freeman RA, Everhart WH. 1971. Toxicity of aluminum hydroxide complexes in neutral and basic media to rainbow trout. Trans. Am. Fish. Soc. 4: 644-658.

Frick KG, Herrmann J. 1990. Aluminum accumulation in a lotic mayfly at low pH -- a laboratory study. Ecotoxicol Environ Saf 19: 81-88.

Froment DH, Buddington B, Miller NL, Alfrey AC. 1989. Effect of solubility on the gastrointestinal absorption of aluminum from various aluminum compounds in the rat. J Lab Clin Med 114(3): 237-242.

Gagnon C, Turcotte P. 2007. Role of colloids in the physical speciation of metals in the dispersion plume of a major municipal effluent. J Water Sci 20(3): 275-285.

Ganrot PO. 1986. Metabolism and possible health effects of aluminum. Environ Health Perspect 65: 363-441.

Garbossa G, Galvez G, Castro ME, Nesse A. 1998. Oral aluminum administration to rats wih normal renal function. 1. Impairment of erythropoiesis. Hum Exp Toxicol 17(6): 312-317.

Garrels RM, Mackenzie FT, Hunt C. 1975. Chemical cycles and the global environment. Los Altos (CA): William Kaufmann Inc.206 p. [cited in Driscoll and Postek 1996].

Garrett RG. 1998. Aluminium levels in Canadian soils and glacial tills from the Prairies and Southern Ontario. Geological Survey of Canada. p 5.

Gauthier E, Fortier I, Courchesne F, Pepin P, Mortimer J, Gauvreau D. 2000. Aluminum forms in drinking water and risk of Alzheimer's disease. Environ Res 84(3): 234-246.

Gensemer RW, Playle RC. 1999. The bioavailability and toxicity of aluminum in aquatic environments. Critical Reviews in Environmental Science and Technology 29(4): 315-450.

George DB, Berk SG, Adams VD, Morgan EL, Roberts RO, Holloway CA, Lott RC, Holt LK, Ting RS, Welch AW. 1991. Alum sludge in the aquatic environment. Denver (CO): American Water Works Association Research Foundation. 224 p.

George DB, Berk SG, Adams VD, Ting RS, Roberts RO, Parks LH, Lott RC. 1995. Toxicity of alum sludge extracts to a freshwater alga, protozoan, fish, and marine bacterium. Arch Environ Contam Toxicol 29: 149-158.

Germain A, Gagnon C, Lind CB. 2000. Entry and exposure characterization for aluminum chloride, aluminum nitrate and aluminum sulfate. Supporting document for Canadian Environmental Protection Act Priority Substances List Assessment Program.Unpublished report. Montreal (PQ): Environment Canada. 111 p.

Giroux M, Rompré M, Carrier D, Audesse P, Lemieux M. 1992. Caractérisation de la teneur en métaux lourds totaux et disponibles des sols du Québec. agrosol V (2): 46-55.

Golub MS, Donald JM, Gershwin ME, Keen CL. 1989. Effects of aluminum ingestion on spontaneous motor activity of mice. Neurotoxicol Teratol 11(3): 231-235.

Golub MS, Germann SL. 1998. Aluminum effects on operant performance and food motivation of mice. Neurotoxicol Teratol 20(4): 421-427.

Golub MS, Germann SL. 2000. Long-term consequences of developmental aluminum (Al) in mice. Neurotoxicol Teratol 22(3): 460.

Golub MS, Germann SL. 2001a. Spatial learning strategies in aged mice exposed to aluminum (Al). Neurotoxicol Teratol 23(3): 293.

Golub MS, Germann SL. 2001b. Long-term consequences of developmental exposure to aluminum in a suboptimal diet for growth and behavior of Swiss Webster mice. Neurotoxicol Teratol 23(4): 365-372.

Golub MS, Germann SL, Han B, Keen CL. 2000. Lifelong feeding of a high aluminum diet to mice. Toxicology 150(1-3): 107-117.

Golub MS, Germann SL, Torrente M. 2002. In utero aluminum exposure affects myelination in the guinea pig. Neurotoxicol Teratol 24(3): 423.

Golub MS, Gershwin ME, Donald JM, Negri S, Keen CL. 1987. Maternal and developmental toxicity of chronic aluminum exposure in mice. Fundam Appl Toxicol 8(3): 346-357.

Golub MS, Han B, Keen CL. 1996a. Iron and manganese uptake by offspring of lactating mice fed a high aluminum diet. Toxicology 109: 111-118.

Golub MS, Han B, Keen CL. 1996b. Developmental patterns of aluminum and five essential mineral elements in the central nervous system of the fetal and infant guinea pig. Biol Trace Elem Res 55(3): 241-251.

Golub MS, Han B, Keen CL, Gershwin ME. 1992a. Effects of dietary aluminum excess and manganese deficiency on neurobehavioral endpoints in adult mice. Toxicol Appl Pharmacol 112(1): 154-160.

Golub MS, Han B, Keen CL, Gershwin ME. 1993. Developmental patterns of aluminum in mouse brain and effects of dietary aluminum excess on manganese deficiency. Toxicology 81(1): 33-47.

Golub MS, Han B, Keen CL, Gershwin ME. 1994. Auditory startle in Swiss Webster mice fed excess aluminum in diet. Neurotoxicol Teratol 16(4): 423-425.

Golub MS, Han B, Keen CL, Gershwin ME, Tarara RP. 1995. Behavioral performance of Swiss Webster mice exposed to excess dietary aluminum during development or during development and as adults. Toxicol Appl Pharmacol 133(1): 64-72.

Golub MS, Keen CL, Gershwin ME. 1992b. Neurodevelopmental effect of aluminum in mice: fostering studies. Neurotoxicol Teratol 14(3): 177-82.

Golub MS, Tarara RP. 1999. Morphometric studies of myelination in the spinal cord of mice exposed developmentally to aluminum. Neurotoxicology 20(6): 953-959.

Gomez M, Bosque MA, Domingo JL, Llobet JM, Corbella J. 1990. Evaluation of the maternal and developmental toxicity of aluminum from high doses of aluminum hydroxide in rats. Vet Hum Toxicol 32(6): 545-548.

Gomez M, Domingo JL, Llobet JM. 1991. Developmental toxicity evaluation of oral aluminum in rats: influence of citrate. Neurotoxicol Teratol 13(3): 323-328.

Goncalves PP, Silva VS. 2007. Does neurotransmission impairment accompany aluminum neurotoxicity? J Inorg Biochem (101): 1291-1338.

Gong QH, Wu Q, Huang XN, Sun AS, Nie J, Shi JS. 2006. Protective effect of Ginkgo biloba leaf extract on learning and memory deficit induced by aluminum in model rats. Chin J Integr Med 12(1): 37-41.

Gong QH, Wu Q, Huang XN, Sun AS, Shi JS. 2005. Protective effects of Ginkgo biloba leaf extract on aluminum-induced brain dysfunction in rats. Life Sci 77(2): 140-148.

Gonzalez-Munoz MJ, Pena A, Meseguer I. 2008. Role of beer as a possible protective factor in preventing Alzheimer's disease. Food Chem Toxicol 46: 49-56.

Gopalakrishnan S, Thilagam H, Raja PV. 2007. Toxicity of heavy metals on embryogenesis and larvae of the marine sedentary polychaete Hydroides elegans. Arch Environ Contam Toxicol 52: 171-178.

Gourrier-Fréry C, Fréry N. 2004. Aluminium. EMC- Toxicologie Pathologie 1: 79-95.

Gramiccioni L, Ingrao G, Milana MR, Santaroni P, Tomassi G. 1996. Aluminium levels in Italian diets and in selected foods from aluminium utensils. Food Addit Contam 13(7): 767-774.

Graves AB, Rosner D, Echeverria D, Mortimer JA, Larson EB. 1998. Occupational exposures to solvents and aluminium and estimated risk of Alzheimer's disease. Occup Environ Med 55(9): 627-633.

Graves AB, White E, K_oepsell TD, Reifler BV, Van Belle G, Larson EB. 1990. The association between aluminum-containing products and Alzheimer's disease. J Clin Epidemiol 43 (1): 35-44.

[GVRD] Greater Vancouver Regional District. 2006. Wastewater. The Greater Vancouver Sewerage & Drainage District quality control annual report 2006. 131 p + Appendices. Available at: http://www.gvrd.bc.ca/sewerage/pdf/Quality
ControlAnnualReport2006forGVS&DD.pdf.

Greger JL, Baier MJ. 1983. Excretion and retention of low or moderate levels of aluminium by human subjects. Food Chem Toxicol 21(4): 473-477.

Greger JL, Chang MM, MacNeil GG. 1994. Tissue turnover of aluminum and Ga-67: effect of iron status. Proceedings of the Society for Experimental Biology and Medicine 207(89-96).

Greger JL, Radzanowski GM. 1995. Tissue aluminium distribution in growing, mature and ageing rats: relationship to changes in gut, kidney and bone metabolism. Food Chem Toxicol 33(10): 867-875.

Guillard O, Fauconneau B, Olichon D, Dedieu G, Deloncle R. 2004. Hyperaluminemia in a woman using an aluminum-containing antiperspirant for 4 years. Am J Med 117(12): 956-959.

Gun RT, Korten AE, Jorm AF, Henderson AS, Broe GA, Creasey H. 1997. Occupational risk factors for alzheimer disease: A case-control Study. Alzheimer Dis Assoc Disord 11(1): 21-27.

Gundersen DT, Bustaman S, Seim WK, Curtis LR. 1994. pH, hardness, and humic acid influence aluminum toxicity to rainbow trout (Oncorhynchus mykiss) in weakly alkaline waters. Can. J. Fish. Aquat. Sci. 51: 1345-1355.

Guo T, Chen Y, Zhang Y, Jin Y. 2006. Alleviation of Al toxicity in barley by addition of calcium. Agric. Sci. China 5(11): 828-833.

Gupta A, Shukla GS. 1995. Effect of chronic aluminum exposure on the levels of conjugated dienes and enzymatic antioxidants in hippocampus and whole brain of rat. Bull Environ Contam Toxicol 55(5): 716-722.

Gupta VB, Anitha S, Hegde ML, Zecca L, Garruto RM, Ravid R, Shankar SK, Stein R, Shanmugavelu P, Jagannatha Rao KS. 2005. Aluminium in Alzheimer's disease: are we still at a crossroad? Cell Mol Life Sci. 62(2): 143-158.

Guyton AC. 1991. Textbook of Medical Physiology. Philadelphia: WB Saunders Co. 1014 p.

Guzyn W. 2003. Phytotoxicology 2001 and 2002. Investigations: Algoma Ore Division, Twp. of Michipicoten (Wawa). Ministry of Environment. Ontario.

Hackenberg U. 1972. Chronic ingestion by rats of standard diet treated with aluminum phosphide. Toxicol Appl Pharmacol 23(1): 147-158.

Hall WS, Hall LW, Jr. 1989. Toxicity of alum sludge to Ceriodaphnia dubia and Pimephales promelas. Bull Environ Contam Toxicol 42: 791-798.

Hall RJ, Driscoll CT, Likens GE, Pratt JM. 1985. Physical, chemical, and biological consequences of episodic aluminum additions to a stream. Limnol Oceanogr 30: 212-220.

Haluschak PW, Mills GF, Eilers RG, Grift S. 1998. Status of selected Trace elements in Agricultural Soils of Southern Manitoba. Agriculture and Agri-food Canada. Soil Resource section. Soils and Crops Branch, Manitoba Agriculture. 46 p.

Hammond KE, Evans DE, Hodson MJ. 1995. Aluminium-silicon interactions in barley (Hordeum vulgare L.) seedlings. Plant Soil 173: 89-95.

Hanninen H, Matikainen E, Kovala T, Valkonen S, Riihimaki V. 1994. Internal load of aluminum and the central nervous system function of aluminum welders. Scand J Work Environ Health 20(4): 279-285.

Haram EM, Weberg R, Berstad A. 1987. Urinary excretion of aluminium after ingestion of sucralfate and an aluminium-containing antacid in man. Scand J Gastroenterol 22(5): 615-618.

Harris WR, Berthon G, Day JP, Exley C, Flaten TP, Forbes WF, Kiss T, Orvig C, Zatta PF. 1996. Speciation of aluminum in biological systems. J Toxicol Environ Health 48(6): 543-568.

Havas M. 1985. Aluminum bioaccumulation and toxicity to Daphnia magna in soft water at low pH. Can. J. Fish. Aquat. Sci. 42: 1741-1748.

Havas M. 1986. A hematoxylin staining technique to locate sites of aluminum binding in aquatic plants and animals. Water, Air, Soil Pollut. 30: 735-741.

Havas M, Likens GE. 1985. Changes in 22Na influx and outflux in Daphnia magna (Straus) as a function of elevated Al concentrations in soft water at low pH. Proc Natl Acad Sci U S A 82: 7345-7349 [cited in Rosseland et al. 1990].

Havens KE. 1990. Aluminum binding to ion exchange sites in acid-sensitive versus acid-tolerant cladocerans. Environ Pollut 64: 133-141.

Health Canada. 1994. Human Health Risk Assessment for Priority Substances.

Health Canada. 1998a. Exposure Factors for Assessing total daily intake of Priority substances by the General Population of Canada - Draft report. Ottawa (ON): Health Canada, Environmental Health Directorate, Priority Substances Section.

Health Canada. 1998b. Guidelines for Canadian Drinking Water Quality - Technical Documents: Aluminum. Ottawa. 22 p.

Health Canada. 1999. CEPA Supporting Documentation - Human Exposure Assessment of aluminium chloride, aluminium nitrate, and aluminium sulfate. Canada: Health Canada. 1-21 p.

Health Canada. 2003. Trace Metal Analysis - Infant Formula [Internet].

Health Canada. 2004. Consolidation of the Food and Drugs Act and the Food and Drug Regulations. Division 16. Food Additives and tables, Part B. July 2004, Ottawa. Available at: http://www.hc-sc.gc.ca/fn-an/legislation/acts-lois/fda-lad/index-eng.php.

Health Canada. 2007a. Guidelines for Canadian Drinking Water Quality Summary Table. Prepared by the Federal-Provincial-Territorial Committee on Drinking Water of the Federal-Provincial-Territorial Committee on Health and the Environment. March 2007.

Health Canada. 2007b. Total water systems in Canada. Personal communication with members of the Federal/Provincial/Territorial Committee on Drinking Water and personnal communications with some municipalities in Quebec and Alberta. Ottawa.

Health Canada. 2007c. Personal communication with Michelle Turner of the McLaughlin Centre of Population Health Risk Assessment, Institute of Population Health, University of Ottawa. Ottawa.

Health Canada. 2008a. Supporting Document for the Second Priority Substances List Draft Assessment for Aluminum Salts.

Health Canada. 2008b. Personal communication with Elena Emelianova of the Food Packaging Materials and Incidental Additives Section, Chemical Health Hazard Assessment Division, Food Directorate, Health Canada.

Hem JD, Robertson CE. 1967. Form and stability of aluminum hydroxide complexes in dilute solution. U.S. Geological Survey. 55 p. Water Supply Paper No. 1827-A.

Hendershot WH, Courchesne F. 1991. Comparison of soil solution chemistry in zero tension and ceramic-cup tension lysimeters. European Journal of Soil Science 42: 577-583.

Hendershot WH, Courchesne F, Jeffries DS. 1995. Aluminum geochemistry at the catchment scale in watersheds influenced by acidic precipitation. In:Sposito G, editor, The environmental chemistry of aluminum. 2nd edition. Boca Raton (FL): CRC Press. p. 419-449.

Henry DA, Goodman WG, Nudelman RK, DiDomenico NC, Alfrey AC, Slatopolsky E, Stanley TM, Coburn JW. 1984. Parenteral aluminum administration in the dog: I. Plasma kinetics, tissue levels, calcium metabolism, and parathyroid hormone. Kidney Int 25(2): 362-369.

Hernan MA, Jick SS, Olek MJ, Jick H. 2004. Recombinant hepatitis B vaccine and the risk of multiple sclerosis: a prospective study. Neurology 63(5): 838-842.

Herrmann J. 1987. Aluminium impact on freshwater invertebrates at low pH: A review. In: Landner, L, editor, Speciation of metals in water, sediments and soil systems. Lecture Notes in Earth Sciences. Volume 11: 157-175 [cited in Rosseland et al.1990].

Heyman A, Wilkinson WE, Stafford JA, Helms MJ, Sigmon AH, Weinberg T. 1984. Alzheimer's disease: a study of epidemiological aspects. Ann Neurol 15(4): 335-341.

Höhr D, Abel J, Wilhelm M. 1989. Renal clearance of aluminium: studies in the isolated perfused rat kidney. Toxicol Lett 45: 165-174.

Horst WJ, Klotz F, Szulkiewicz P. 1990. Mechanical impedance increases        aluminium tolerance of soybean (Glycine max L.). Plant Soil 124: 227-231.

Hosovski E, Mastelica Z, Sunderic D, Radulovic D. 1990. Mental abilities of workers exposed to aluminium. Med Lav 81(2): 119-123.

Hossain MD, Bache DH. 1991. Composition of alum flocs derived from a coloured, low-turbidity water. Aqua 40(5): 298-303.

Hu H, Yang YJ, Li XP, Chen GH. 2005. [Effect of aluminum chloride on motor activity and species-typical behaviors in mice]. Zhonghua Lao Dong Wei Sheng Zhi Ye Bing Za Zhi 23(2): 132-135.

Huh JW, Choi MM, Lee JH, Yang SJ, Kim MJ, Choi J, Lee KH, Lee JE, Cho SW. 2005. Activation of monoamine oxidase isotypes by prolonged intake of aluminum in rat brain. J Inorg Biochem 99(10): 2088-2091.

Hutchinson NJ, Holtze KE, Munro JR, Pawson TW. 1987. Lethal responses of salmonid early life stages to H+ and Al in dilute waters. Belg. J. Zool. 117 (Suppl. 1): 201- 217.

Hutchinson NJ, Sprague JB. 1987. Reduced lethality of Al, Zn and Cu mixtures to American flagfish by complexation with humic substances in acidified soft waters. Environ Toxicol Chem 6: 755-765.

Hutchinson TC, Bozic L, Munoz-Vega G. 1986. Responses of five species of conifer seedlings to aluminium stress. Water Air Soil Pollut 31: 283-294.

Ilvesniemi H. 1992. The combined effect of mineral nutrition and soluble aluminium on Pinus sylvestris and Picea abies seedling. For Ecol Manage. 51: 227-238.

[InVS-Afssa-Afssaps] Institut de veille sanitaire-Agence française de sécurité sanitaire des aliments-Agence française de sécurité sanitaire des produits de santé. 2003. Évaluation des risques sanitaires liés à l'exposition de la population française à l'aluminium: eaux, aliments, produits de santé. 1992 p.

[IARC] International Agency for Research on Cancer. 1987. Aluminum production. IARC monograph. 34 (Suppl 7).

[ICMM] International Council on Mining and Metals. 2007. MERAG: Metals environmental risk assessment guidance. London (UK): ICMM. Available from: http://www.icmm.com/document/15.

Ittel T, Gerish P, Nolte E, Sieberth H. 1993. Fractional absosption of Al is dose-dependent: A ²⁶Al tracer study. Nephrol Dial Transplant 8: 993.

Ittel TH, Kluge R, Sieberth HG. 1988. Enhanced gastrointestinal absorption of aluminium in uraemia: time course and effect of vitamin D. Nephrol Dial Transplant 3(5): 617-623.

Jacqmin H, Commenges D, Letenneur L, Barberger-Gateau P, Dartigues JF. 1994. Components of drinking water and risk of cognitive impairment in the elderly. Am J Epidemiol 139(1): 48-57.

Jacqmin-Gadda H, Commenges D, Letenneur L, Dartigues JF. 1996. Silica and aluminum in drinking water and cognitive impairment in the elderly. Epidemiology 7(3): 281-285.

Jeffery EH, Abreo K, Burgess E, Cannata J, Greger JL. 1996. Systemic aluminum toxicity: effects on bone, hematopoietic tissue, and kidney. J Toxicol Environ Health 48(6): 649-665.

Jia Y, Zhong C, Wang Y. 2001b. Effects of aluminum on amino acid neurotransmitters in hippocampus of rats. Zhonghua Yu Fang Yi Xue Za Zhi 35(6): 397-400.

Jia Y, Zhong C, Wang Y, Zhao R. 2001a. Effects of aluminum intake on the content of aluminum, iron, zinc and lipid peroxidation in the hippocampus of rats. Wei Sheng Yan Jiu 30(3): 132-134.

Jing Y, Wang Z, Song Y. 2004. Quantitative study of aluminum-induced changes in synaptic ultrastructure in rats. Synapse 52(4): 292-298.

Johannessen M. 1980. Aluminium, a buffer in acidic waters? In: Drablos D, Tollan A, editors, Ecological impact of acid precipitation. Proceedings of an international conference, March 11-14, Sandefjord, Norway. p. 222-223.

Johnson GV, Jope RS. 1987. Aluminum alters cyclic AMP and cyclic GMP levels but not presynaptic cholinergic markers in rat brain in vivo. Brain Res 403(1): 1-6.

Johnson GV, Watson AL, Jr., Lartius R, Uemura E, Jope RS. 1992. Dietary aluminum selectively decreases MAP-2 in brains of developing and adult rats. Neurotoxicology 13(2): 463-474.

[JECFA] Joint FAO/WHO Expert Committee on Food Additives. 2007. 67th Meeting of the Joint FAO/WHO Expert Committee on Food Additives and Contaminants, 20-29 June. Rome: Food and Agricultural Organization of the United Nations, World Health Organization.

Jonasson B. 1996. Phosphorus transformations in alum sludge amended soils. Swedish Journal of. Agricultural Research 26: 69-79.

Jones G, Henderson V. 2006. Metal concentrations in soils and produce from gardens in Flin Flon, Manitoba, 2002. Flin FLon: Winnipeg. Section de l'aménagement et de l'habitat et de la surveillance des écosystèmes. Direction de la protection de la faune et des écosystèmes. Conservation Manitoba. Report nr 2006-01. 81 p.

Jones G, Philips F. 2003. Metal concentrations in surface soils of Thompson, September 2001. Thompson: Winnipeg. Section de l'aménagement et de l'habitat et de la surveillance des écosystèmes. Direction de la protection de la faune et des écosystèmes. Conservation Manitoba. Report nr 2003-01. 21 p.

Jones JH. 1938. The metabolism of calcium and phosphorus as influenced by the addition to the diet of salts of metals which form insoluble phosphates. Amer J Physiol 124: 230-237.

Jouhanneau P, Raisbeck GM, Yiou F, Lacour B, Banide H, Drueke TB. 1997. Gastrointestinal absorption, tissue retention, and urinary excretion of dietary aluminum in rats determined by using ²⁶Al. Clin Chem 43(6 Pt 1): 1023-1028.

Jyoti A, Sharma D. 2006. Neuroprotective role of Bacopa monniera extract against aluminium-induced oxidative stress in the hippocampus of rat brain. Neurotoxicology 27(4): 451-457.

Kádár E, Salánki J, Powell J, White KN, McCrohan CR. 2002. Effect of sub-lethal concentrations of aluminium on the filtration activity of the freshwater mussel Anodonta cygnea L. at neutral pH. Acta Biol Hung 53(4): 485-493.

Kaizer RR, Correa MC, Spanevello RM, Morsch VM, Mazzanti CM, Goncalves JF, Schetinger MR. 2005. Acetylcholinesterase activation and enhanced lipid peroxidation after long-term exposure to low levels of aluminum on different mouse brain regions. J Inorg Biochem 99(9): 1865-1870.

Kaneko N, Yasui H, Takada J, Suzuki K, Sakurai H. 2004. Orally administrated aluminum-maltolate complex enhances oxidative stress in the organs of mice. J Inorg Biochem 98(12): 2022-2031.

Karlsson-Norrgren L, Dickson W, Ljungberg O, Runn P. 1986. Acid water and aluminium exposure: gill lesions and aluminium accumulation in farmed brown trout, Salmo trutta L. J Fish Dis 9: 1-9 [cited in Dussault et al., 2001].

Katyal R, Desigan B, Sodhi CP, Ojha S. 1997. Oral aluminum administration and oxidative injury. Biol Trace Elem Res 57(2): 125-130.

Kaur A, Gill KD. 2005. Disruption of neuronal calcium homeostasis after chronic aluminium toxicity in rats. Basic Clin Pharmacol Toxicol 96(2): 118-122.

Kaur A, Gill KD. 2006. Possible peripheral markers for chronic aluminium toxicity in Wistar rats. Toxicol Ind Health 22(1): 39-46.

Kaur A, Joshi K, Minz RW, Gill KD. 2006. Neurofilament phosphorylation and disruption: a possible mechanism of chronic aluminium toxicity in Wistar rats. Toxicology 219(1-3): 1-10.

Kawahara M. 2005. Effects of aluminum on the nervous system and its possible link with neurodegenerative diseases. J Alzheimers Dis 8(2): 171-182; discussion 209-215.

Kiesswetter E, Sch, auml, per M, Buchta M, Schaller KH, Rossbach B, Scherhag H, Zschiesche W, Letzel S. 2007. Longitudinal study on potential neurotoxic effects of aluminium: I. Assessment of exposure and neurobehavioural performance of Al welders in the train and truck construction industry over 4 years. Int Arch Occup Environ Health. 81(1): 41-67.

Kilburn KH. 1998. Neurobehavioral impairment and symptoms associated with aluminum remelting. Arch Environ Health 53(5): 329-335.

Kim K. 2003. Perinatal exposure to aluminum alters neuronal nitric oxide synthase expression in the frontal cortex of rat offspring. Brain Res Bull 61(4): 437-441.

Kimball BA, Callender E, Axtmann EV. 1995. Effects of colloids on metal transport in a river receiving acid mine drainage, upper Arkansas River, Colorado, U.S.A. Applied Geochem 10: 285-306 [cited in Farag et al., 2007].

Kinraide TB. 1997. Reconsidering the rhizotoxicity of hydroxyl, sulphate, and fluoride complexes of aluminium. J Exp Bot 48: 1115-1124.

Kirkwood DE, Nesbitt HW. 1991. Formation and evolution of soils from an acidified watershed: Plastic Lake, Ontario, Canada. Geochim Cosmochim Acta 55: 1295--1308.

Klein GL. 2005. Aluminum: new recognition of an old problem. Curr Opin Pharmacol 5(6): 637-640.

Kobayashi K, Yumoto S, Nagai H, Hosoyama Y, Imamura M, Masuzawa S, Koizumi Y, Ymashita H. 1990. ²⁶Al tracer experiment by accelerator mass spectrometry and its application to the studies for amyotrophic lateral slerosis and Alzheimer's disease. Proceedings of the Japan Academy. Series B, Physical and biological sciences B66: 189-192.

Kohila T, Parkkonen E, Tahti H. 2004. Evaluation of the effects of aluminium, ethanol and their combination on rat brain synaptosomal integral proteins in vitro and after 90-day oral exposure. Arch Toxicol 78(5): 276-282.

Koo WW, Kaplan LA, Krug-Wispe SK. 1988. Aluminum contamination of infant formulas. JPEN J Parenter Enteral Nutr 12(2): 170-173.

Kopácek J, Ulrich K-U, Hejzlar J, Borovec J, Stuchlík E. 2001. Natural inactivation of phosphorus by aluminum in atmospherically acidified water bodies. Water Res 35(16): 3783-3790.

Kram P, Hruska J, Driscoll C, Johnson CE. 1995. Biogeochemistry of aluminum in a forest catchment in the Czech Republic impacted by atmospheric inputs of strong acids. Water Air Soil Pollut 85: 1831--1836.

Krantzberg G. 1989. Metal accumulation by chironomid larvae: the effects of age and body weight on metal body burdens. Hydrobiologia 188/189: 497--506.

Krantzberg G, Stokes PM. 1988. The importance of surface adsorption and pH in metal accumulation by chironomids. Environ Toxicol Chem 7: 653--670.

Krasovskii GN, Vasukovich LY, Chariev OG. 1979. Experimental study of biological effects of leads and aluminum following oral administration. Environ Health Perspect 30: 47-51.

Krewski D, Yokel RA, Nieboer E, Borchelt D, Cohen J, Harry J, Kacew S, Lindsay J, Mahfouz AM, Rondeau V. 2007. Human health risk assessment for aluminium, aluminium oxide, and aluminium hydroxide. J Toxicol Environ Health B Crit Rev. 2007; 10 Suppl 1: 1-269.

Kuja A, Jones R. 2000. Soil contamination in Port Colborne Woodlots: 2000. Port Colborne: Ministry of Environment. Ontario. 30 p.

Kullberg A, Bishop KH, Hargeby A, Jansson M, Peterson RC. 1993. The ecological significance of dissolved organic carbon in acidified waters. Ambio 22(5): 331--337.

Kumar S. 1998. Biphasic effect of aluminium on cholinergic enzyme of rat brain. Neurosci Lett 248(2): 121-123.

Kumar S. 1999. Aluminium-induced biphasic effect. Med Hypotheses 52(6): 557-559.

Kumar S. 2002. Aluminium-induced changes in the rat brain serotonin system. Food Chem Toxicol 40(12): 1875-1880.

Kundert K, Meilke L, Elford T, Maksymetz B, Pernitsky DJ. 2004. Evaluating pH adjustment to investigate seasonal changes in aluminum residuals at a large conventional water treatment plant. In: Proceedings of the 56th Annual Western Canada Water and Wastewater Association Conference and Tradeshow, Calgary, Alberta, October 17-20, 2004. 13 p.

Lacroix GL. 1992. Mitigation of low stream pH and its effects on salmonids. Environ Pollut 78: 157--164.

Lacroix GL, Peterson RH, Belfry CS, Martin-Robichaud DJ.  1993. Aluminum dynamics on gills of Atlantic salmon (Salmo salar) fry in the presence of citrate and effects on integrity of gill structures. Aquat Toxicol 27: 373--402.

Lal B, Gupta A, Gupta A, Murthy RC, Ali MM, Chandra SV. 1993. Aluminum ingestion alters behaviour and some neurochemicals in rats. Indian J Exp Biol 31(1): 30-35.

Lamb DS, Bailey GC. 1981. Acute and chronic effects of alum to midge larva (Diptera:Chironomidae). Bull Environ Contam Toxicol 27: 59--67.

Landry B, Mercier M. 1992. Notions de géologie. 3rd edition. Mont-Royal (PQ): Modulo. p. 565.

LaZerte BD, van Loon G, Anderson B. 1997. Aluminum in water. In: Yokel R, Golub MS, editors, Research Issues in Aluminum Toxicity. Washington (DC): Taylor and Francis. p. 17-46.

Lee W, Westerhoff P. 2006. Dissolved organic nitrogen removal during water treatment by aluminum sulfate and cationic polymer coagulation. Water Res 40: 3767-3774.

Lee YH, Hultberg H, Sverdrup H, Borg GC. 1995. Are ion exchange processes important in controlling the cation chemistry of soil and runoff water. Water Air Soil Pollut 85: 819--1824.

Lévesque L, Mizzen CA, McLachlan DR, Fraser PE. 2000. Ligand specific effects on aluminum incorporation and toxicity in neurons and astrocytes. Brain Res 877(2): 191-202.

Lewandowski J, Schauser I, Hupfer M. 2003. Long term effects of phosphorus precipitations with alum in hypereutrophic Lake Süsser See (Germany). Water Res 37: 3194-3204.

Lewis RJ. 1992. Sax's dangerous properties of industrial materials. Vol. 3. New York (NY): Van Nostrand Reinhold.

Li XP, Yang YJ, Hu H, Wang QN. 2006. Effect of aluminum trichloride on dissociated Ca²⁺ in Hippocampus neuron cell as well as learning and memory. Zhonghua Lao Dong Wei Sheng Zhi Ye Bing Za Zhi 24(3): 161-163.

Liao YH, Yu HS, Ho CK, Wu MT, Yang CY, Chen JR, Chang CC. 2004. Biological monitoring of exposures to aluminium, gallium, indium, arsenic, and antimony in optoelectronic industry workers. J Occup Environ Med 46(9): 931-936.

Likens GE, Bormann FH, Pierce RS, Eaton JS, Johnson NM. 1977. Biogeochemistry of a forested ecosystem. New York (NY): Springer-Verlag.

Lima PD, Leite DS, Vasconcellos MC, Cavalcanti BC, Santos RA, Costa-Lotufo LV, Pessoa C, Moraes MO, Burbano RR. 2007. Genotoxic effects of aluminum chloride in cultured human lymphocytes treated in different phases of cell cycle. Food Chem Toxicol 45: 1154-1159.

Lin S, Evans RL, Schnepper D, Hill T. 1984. Evaluation of wastes from the East St. Louis water treatment plant and their impact on the Mississippi River. Springfield (IL): Department of Energy and Natural Resources, (ISWS/CIR-160/84:1-89, 1984).

Lin SD. 1989. No adverse environmental impacts from water plant discharges. Water Engineering & Management 136: 40--41.

Lindsay J, Laurin D, Verreault R, Hébert R, Helliwell B, Hill GB, McDowell I. 2002. Risk factors for Alzheimer's disease: a prospective analysis from the Canadian study of health and aging. Am J Epidemiol 156(5): 445-453.

Lindsay WL, Vlek PLG, Chien SH. 1989. Phosphate minerals. In: Dixon JB, Weed SB, editors, Minerals in soil environments. Madison (WI): Soil Science Society of America. p. 1089--1130.

Liukkonen-Lilja H, Piepponen S. 1992. Leaching of aluminium from aluminium dishes and packages. Food Addit Contam 9(3): 213-223.

Llansola M, Minana MD, Montoliu C, Saez R, Corbalan R, Manzo L, Felipo V. 1999. Prenatal exposure to aluminum reduces expression of neuronal nitric oxide synthase and of soluble guanylate cyclase and impairs glutamatergic neurotransmission in rat cerebellum. J Neurochem 73(2): 712-718.

Loganathan P, Hedley MJ, Grace ND, Lee J, Cronin SJ, Bolan NS, Zanders JM. 2003. Fertiliser contaminants in New Zealand grazed pasture with special reference to cadmium and fluorine: a review. Australian Journal of Soil Research 41: 501-532 [cited in Manoharan et al. 2007].

Long JF, Nagode LA, Steinmeyer CL, Renkes G. 1994. Comparative effects of calcitriol and parathyroid hormone on serum aluminum in vitamin D-depleted rabbits fed an aluminum-supplemented diet. Res Commun Chem Pathol Pharmacol 83(1): 3--14.

Long JF, Renkes G, Steinmeyer CL, Nagode LA. 1991. Effect of calcitriol infusions on serum aluminum in vitamin D-depleted rabbits fed an aluminum-supplemented ration. Res Commun Chem Pathol Pharmacol 74(1): 89-104.

Lopez FE, Cabrera C, Lorenzo ML, Lopez MC. 2002. Aluminum levels in convenience and fast foods: in vitro study of the absorbable fraction. Sci Total Environ 300(1--3): 69-79.

Lydersen E, Poléo ABS, Muniz IP, Salbu B, Bjornstad HE. 1990a. The effects of naturally occurring high and low molecular weight inorganic and organic species on the yolk-sac larvae of Atlantic salmon (Salmo salar L.) exposed to acidic aluminium-rich lake water. Aquat Toxicol 18: 219--229.

Lydersen E, Salbu B, Poléo ABS, Muniz IP. 1990b. The influences of temperature on aqueous aluminium chemistry. Water Air Soil Pollut 51: 203-215.

Mackie GL, Kilgour BW. 1995. Efficacy and role of alum in removal of zebra mussel veliger larvae from raw water supplies. Water Res 29: 731--744.

MacLean DC, Hansen KS, Schneider RE. 1992. Amelioration of aluminium toxicity in wheat by fluoride. New Phytol 121: 81-88 [cited in Manoharan et al. 2007].

Maitani T, Kubota H, Hori N, Yoshihira K, Takeda M. 1994. Distribution and urinary excretion of aluminum injected with several organic acids into mice: Relationship with chemical state in serum studied by the HPLC-ICP method. J Appl Toxicol 14: 257-261.

Malley DF, Chang PSS. 1985. Effects of aluminium and acid on calcium uptake by the crayfish Oronectes virilis. Arch Environ Contam Toxicol 14: 739-747 [cited in Rosseland et al. 1990].

Malley DF, Findlay DL, Chang PSS. 1982. Ecological effects of acid precipitation on zooplankton. In: D'Itri FN, editor, Acid precipitation: effects on ecological systems. Ann Arbor (MI): Ann Arbor Science, p. 297-327.

Malley DF, Heubner JD, Donkersloot K. 1988. Effects on ionic composition of blood and tissues of Anodonta grandis grandis (Bivalvia) of an addition of aluminum and acid to a lake. Arch Environ Contam Toxicol 17: 479--491.

Mameli O, Caria MA, Melsi P, Zambenedetti P, Ramila M, Zatta P.  2006.  Effect of aluminum consumption on the vestibule-ocular reflex.  Metab Brain Dis 21(2-3): 89-107.

Manoharan V, Loganathan P, Tillman RW, Parfitt RL. 2007. Interactive effects of soil acidity and fluoride on soil solution aluminium chemistry and barley (Hordeum vulgare L.) root growth. Environ Poll 145: 778-786.

Marchesi VT. 2005. An alternative interpretation of the amyloid Ab hypothesis with regard to the pathogenesis of Alzheimer's disease. PNAS 102(26): 9093-9098.

Martin RB. 1996. Ternary complexes of Al+3 and F with a third ligand Coord. Chem. Rev. 149: 23-32.

Martyn CN, Coggon DN, Inskip H, Lacey RF, Young WF. 1997. Aluminum concentrations in drinking water and risk of Alzheimer's disease. Epidemiology 8(3): 281-286.

Mathias R, Salusky I, Harman W, Paredes A, Emans J, Segre G, Goodman W. 1993. Renal bone disease in pediatric and young adult patients on hemodialysis in a children's hospital. J Am Soc Nephrol 3(12): 1938-1946.

Matsumoto H. 2000. Cell biology of aluminium toxicity and tolerance in higher plants. Int Rev Cytol 200: 1-46 [cited in Manoharan et al. 2007].

Mattson MP, Lovell MA, Ehmann WD, Markesbery WR. 1993. Comparison of the effects of elevated intracellular aluminum and calcium levels on neuronal survival and tau immunoreactivity. Brain res 602: 21-31.

McCanny SJ, Hendershot WH, Lechowicz MJ, Shipley B. 1995. The effects of aluminum on Picea rubens: factorial experiments using sand culture. Can J For Res 25: 8--17.

McCormack KM, Ottosen LD, Sanger VL, Sprague S, Mayor GH, Hook JB. 1979. Effects of prenatal administration of aluminum and parathyroid hormone on fetal development in the rat. Soc Exp Biol Med 161: 74--77.

McCormick LH, Steiner KC. 1978. Variation in aluminum tolerance among six genera of trees. Forest Science 24: 565--568.

McDonald DG, Reader JP, Dalziel TRK. 1989. The combined effects of pH and trace metals on fish ionoregulation. In: Morris R, Taylor EW, Brown DJA, Brown JA, editors, Acid toxicity and aquatic animals. Cambridge (MA): Cambridge University Press. p. 221--242.

McGreer JC, Brix KV, Skeaff JM, DeForest DK, Brigham SI, Adams WJ, Green A. 2003. Inverse relationship between bioconcentration factor and exposure concentration for metals: implications for hazard assessment of metals in the aquatic environment. Environ Toxicol Chem 22(5): 1017-1037.

McLachlan DR, Bergeron C, Smith JE, Boomer D, Rifat SL. 1996. Risk for neuropathologically confirmed Alzheimer's disease and residual aluminum in municipal drinking water employing weighted residential histories. Neurology 46(2): 401-405.

Meek B, Renwick A, Sonich-Mulloch C. 2003. Practical application of kinetic data in risk assessment-an IPCS initiative. Toxicol Lett(138): 151-160.

Merck Index. 2006. An encyclopedia of chemicals, drugs, and biologicals. 14th edition. O'Neil M, Smith A, Heckelman PE, eds. Whitehouse Station (NJ): Merck & Co., Inc.

Meyer-Baron M, Sch, auml, per M, Knapp G, van Thriel C. 2007. Occupational aluminum exposure: Evidence in support of its neurobehavioral impact. Neurotoxicology. 28(6): 1068-1078.

Michel P, Commenges D, Dartigues JF, Gagnon M. 1991. Study of the relationship between aluminium concentration in drinking water and risk of Alzheimer's Disease. In: Iqbal K, McLachlan DRC, Winblad B, Wisniewski HM, editors. Alzheimer's disease: Basic mechanism, diagnosis and therapeutic strategies. New York: John Wiley. p 387-391.

Ministère de l'Environnement du Québec and Environment Canada. 2001. Toxic potential assessment of municipal wastewater treatment plant effluents in Quebec. Montreal (QC): Ministère de l'Environnement du Québec and Environment Canada, St. Lawrence Vision 2000, Phase III -- Industrial and Urban component. 136 pp. + appendices.

[MEF] Ministère de l'Environnement et de la faune du Québec and Environnement Canada. 1998. Évaluation de la toxicité des effluents des stations d'épuration municipales du Québec. Rapport d'étape. Campagne de caractérisation d'hiver. Montréal (QC) : Ministère de l'environnement et de la faune du Québec and Environnement Canada, Direction de la protection de l'environnement, Région du Québec, Groupe d'intervention et restauration, 89 pp. + appendices.

Ministers' Expert Advisory Panel. 1995. Report of the Ministers' Expert Advisory Panel on the second Priority Substances List under the Canadian EnvironmentalProtection Act (CEPA). Ottawa (ON): Government of Canada. 26 p.

[ME and MAFRA] Ministry of the Environment and Ministry of Agriculture, Food and Rural Affairs. 1996. Guidelines for the utilization of biosolids and other wastes on agricultural land. Ministry of the Environment and Ministry of Agriculture, Food and  Rural Affairs. March 1996. Available at: http://www.ene.gov.on.ca/envision/gp/3425e.pdf.

Misawa T, Shigeta S. 1993. Effects of prenatal aluminum treatment on development and behavior in the rat. J Toxicol Sci 18(1): 43-48.

Miu AC, Benga O. 2006. Aluminum and Alzheimer's disease: a new look. J Alzheimers Dis 10(2-3): 179-201.

Moreno A, Dominguez P, Dominguez C, Ballabriga A. 1991. High serum aluminum levels and acute reversible encephalopathy in a 4-year old boy with acute renal failure. Eur J Pediatr 150: 513-514.

Morrissey CA, Bendell-Young LI, Elliott JE. 2005. Assessing trace-metal exposure to American dippers in mountain streams of southwestern British Columbia, Canada. Environ Toxicol Chem 24(4): 836-845.

Mortula M, Gibbons MK, Lake CB, Gagnon GA. 2007. The reuse of alum residuals for wastewater treatment: effect on aluminum leachability. In: Proceedings of the 4th Canadian Organic Residuals and Biosolids Management Conference, Moncton, New Brunswick, June 24-27. Canadian Association on Water Quality, Burlington, Ontario.

Mount DR, Ingersoll CG, Gulley DD, Fernandez JD, LaPoint TW, Bergman HL. 1988. Effect of long-term exposure to acid, aluminum, and low calcium on adult brook trout (Salvelinus fontinalis). 1. Survival, growth, fecundity, and progeny survival. Can. J. Fish. Aquat. Sci. 45: 1633--1642.

Muller G, Bernuzzi V, Desor D, Hutin MF, Burnel D, Lehr PR. 1990. Developmental alterations in offspring of female rats orally intoxicated by aluminum lactate at different gestation periods. Teratology 42(3): 253-261.

Muller G, Hutin MF, Burnel D, Lehr PR. 1992. Aluminum transfer through milk in female rats intoxicated by aluminum chloride. Biol Trace Elem Res 34: 79-87.

Muller JP, Steinegger A, Schlatter C. 1993. Contribution of aluminum from packaging materials and cooking utensils to the daily aluminum intake. Z Lebensm Unters Forsch 197(4): 332-341.

Muniz IP, Leivestad H. 1980. Toxic effects of aluminum on the brown trout, Salmo trutta. In: Drablos D, Tollen A, editors, Proceedings of an International Conference on the Ecological Impact of Acid Precipitation. SNSF Project, Oslo.

Narf RP. 1990. Interactions of Chironomidae and Chaoboridae (Diptera) with aluminum sulfate treated lake sediments. Lake and Reservoir Management 6(1): 33--42.

[NRC] National Research Council. Committee on Fluoride in Drinking Water. 2006. Fluoride in drinking water. A scientific review of EPA's standards. Washington, DC: Academies Press. 507 p.

Nehru B, Anand P. 2005. Oxidative damage following chronic aluminium exposure in adult and pup rat brains. J Trace Elem Med Biol 19(2-3): 203-208.

Nelson WO, Campbell PGC. 1991. Review of the effects of acidification on the geochemistry of Al, Cd, Pb and Hg in freshwater environments. Environ Pollut 71: 91--130.

Neri LC, Hewitt D. 1991. Aluminium, Alzheimer's disease, and drinking water. Lancet 338(8763): 390.

Neri LC, Hewitt D, Rifat SL. 1992. Aluminium in drinking water and risk for diagnoses of presenile alzheimer's type dementia (Abstract 453). Neurobiology of aging. p S115.

Neville CM. 1985. Physiological responses of juvenile rainbow trout, Salmo gairdneri, to acid and aluminum -- prediction of field responses from laboratory data. Can. J. Fish. Aquat. Sci. 42: 2004--2019.

Neville CM, LaZerte BD, Ralston JG. 1988. Aluminum. Scientific criteria document for development of provincial water quality objectives and guidelines. Toronto (ON): Ontario Ministry of the Environment, Water Quality Management Working Group I, Water Resources Branch.

Newman MC, Jagoe CH. 1994. Ligands and the bioavailability of metals in aquatic environments. In: Hamelink JL, Landrum PF, Bergman HL, Benson WH, editors, Bioavailability: Physical, chemical and biological Interactions. Proceedings of the 13th Pellston workshop, Michigan, August 17-22, 1992, Boca Raton (FL): Lewis Publishers. p. 39-61.

Ng AH, Hercz G, Kandel R, Grynpas MD. 2004. Association between fluoride, magnesium, aluminum and bone quality in renal osteodystrophy. Bone 34(1): 216-224.

Nieboer E, Gibson BL, Oxman AD, Kramer JR. 1995. Health effects of aluminum: a critical review with emphasis on aluminum in drinking water. Environ. Rev. 3: 29-81.

Nilsson R. 1988. Aluminium in natural waters -- toxicity to fish and man. In: Astruc M, Lester JN, editors, Heavy metals in the hydrological cycle. London (England): Selper Ltd. p. 11--18.

Nilsson SI, Bergkvist B. 1983. Aluminum chemistry and acidification processes in a spodzol on the Swedish west coast. Water Air Soil Pollut 20: 311--329.

Noble AD, Fey MV, Sumner ME. 1988a. Calcium-aluminum balance and the growth of soybean roots in nutrient solutions. Soil Sci Soc Am J 52: 1651-1656 [cited in Manoharan et al. 2007].

Noble AD, Sumner ME, Alva AK. 1988b. Comparison of aluminum and 8-hydroxyquinoline methods in the presence of fluoride for assaying phytotoxic aluminum. Soil Sci. Soc. Am. J. 52: 1059--1063.

Nolte E, Beck E, Winklhofer C, Steinhausen C. 2001. Compartmental model for aluminium biokinetics. Hum Exp Toxicol 20(2): 111-117.

Nordstrom DK, May HM. 1995. The chemistry of aluminum in surface waters. In: Sposito G, editor, The environmental chemistry of aluminum. 2nd edition. Boca Raton (FL): CRC Press. p. 39--80.

Novak JT, Knocke WR, Geertsema W, Dove D, Taylor A,  Mutter R. 1995. An assessment of cropland application of water treatment residuals. Denver (CO): American Water Works Association Research Foundation, 71 p.

Nozaki Y. 1997. A fresh look at element distribution in the North Pacific. Eos, Transactions, American Geophysical Union 78(21): 221-227. Available from: http://www.agu.org/eos_elec/07025e.html.

Nriagu JO, Wong HTK. 1986. What fraction of the total metal flux into lakes is retained in the sediments? Water Air Soil Pollut 31: 999--1006.

Oneda S, Takasaki T, Kuriwaki K, Ohi Y, Umekita Y, Hatanaka S, Fujiyoshi T, Yoshida A, Yoshida H. 1994. Chronic toxicity and tumorigenicity study of aluminum potassium sulfate in B6C3F1 mice. In Vivo 8(3): 271-278.

Orians KJ, Bruland KW. 1985. Dissolved aluminium in the central North Pacific. Nature 316: 427-429.

Orr PL, Craig GR, Nutt SG, Stephenson J. 1992. Evaluation of acute and chronic toxicity of Ontario sewage treatment plant effluents. Toronto (ON): Ontario Ministry of the Environment, Municipal Section, Municipal and Industrial Strategic Abatement. p. 1-314.

Oteiza PI, Keen CL, Han B, Golub MS. 1993. Aluminum accumulation and neurotoxicity in Swiss-Webster mice after long-term dietary exposure to aluminum and citrate. Metabolism 42(10): 1296-300.

Ott SM, Maloney NA, Klein GL, Alfrey AC, Sherrard DJ. 1982. The prevalence of bone aluminum deposition in renal osteodystrophy and its relation to the reponse to calcitriol therapy. N Engl J Med 307: 709-713.

Otto C, Svensson BS. 1983. Properties of acid brown water streams in south Sweden.Arch Hydrobiol 99: 15-36 [cited in Wren and Stephenson 1991].

Owen LMW, Crews HM, Bishop Nj, Massey RC. 1994. Aluminium uptake from some foods by guinea pigs and the characterizatrion of aluminium in in vivo intestinal digesta by SEC-ICP-MS. Food Chem Toxicol 32(8): 697-705.

Pagenkopf GK. 1983. Gill surface interaction model for trace-metal toxicity to fishes: role of complexation, pH and water hardness. Environ Sci Technol 17: 342-347.

Pai SM, Melethil S. 1989. Kinetics of aluminum in rats I: Dose-dependent elimination from blood after intravenous administration. J Pharm Sci 78(3): 200-202.

Pandya JD, Dave KR, Katyare SS. 2001. Effect of long-term aluminum feeding on lipid/phospholipid profiles of rat brain synaptic plasma membranes and microsomes. J Alzheimers Dis 3(6): 531-539.

Pandya JD, Dave KR, Katyare SS. 2004. Effect of long-term aluminum feeding on lipid/phospholipid profiles of rat brain myelin. Lipids Health Dis 3: 13.

Parametrix. 1995. Persistence, bioaccumulation and toxicity of metals and metal compounds. London (UK): International Council on Metals in the Environment [cited in ICMM 2007].

Parent L, Campbell PGC. 1994. Aluminum bioavailability to the green alga Chlorella pyrenoidosa in acidified synthetic soft water. Environ Toxicol Chem 13: 587-598.

Parent L, Twiss MR, Campbell PGC. 1996. Influences of natural dissolved organic matter on the interaction of aluminum with the microalga Chlorella: a test of the free-ion model of trace metal toxicity. Environ Sci Technol 30: 1713-1720.

Parker DR, Bertsch PM. 1992a. Identification and quantification of the "Al13" tridecameric polycation using ferron. Environ Sci Technol 26: 908-914.

Parker DR, Bertsch PM. 1992b. Formation of the "Al13" tridecameric polycation under diverse synthesis conditions. Environ Sci Technol 26: 914-921.

Parker DR, Zelazny LW, Kinraide TB. 1989. Chemical speciation and plant toxicity of aqueous aluminum. In:Lewis TE, editor, Environmental chemistry and toxicology of aluminum. Chelsea (MI): Lewis Publishers. p. 117-145.

Parkhurst BR, Bergman HL, Fernandez J, Gulley DD, Hocket JR, Sanchez DA. 1990. Inorganic monomeric aluminum and pH as predictors of acidic water toxicity to brook trout (Salvelinus fontinalis). Can. J. Fish. Aquat. Sci. 47: 1631-1640.

Paternain JL, Domingo JL, Llobet JM, Corbella J. 1988. Embryotoxic and teratogenic effects of aluminum nitrate in rats upon oral administration. Teratology 38: 253-257.

Pellerin BA, Fernandez IJ, Norton SA, Kahl JS. 2002. Soil aluminum distribution in the near-stream zone at the Bear Brook watershed in Maine. Water Air Soil Pollut 134: 189-204.

Pennington JA. 1988. Aluminium content of foods and diets. Food Addit Contam 5(2): 161-232.

Perl DP, Good PF. 1987. Uptake of aluminium into central nervous system along nasal-olfactory pathways. Lancet 1: 1028.

Perry RH, Green DW, editors. 1984. Perry's chemical engineers' handbook. 6th edition. New York (NY): McGraw-Hill Book Company.

Peterson RH, Bourbonniere RA, Lacroix GL, Martin-Robichaud DJ, Takats P, Brun G. 1989. Responses of Atlantic salmon (Salmo salar) alevins to dissolved organic carbon and dissolved aluminum at low pH. Water Air Soil Pollut 46: 399-413.

Pettersen JC, Hackett DS, Zwicker GM, Sprague GL. 1990. Twenty-six week toxicity study with Kasal (basic sodium aluminum phosphate) in beagle dogs. Environ Geochem Health 12(1-2): 121-123.

Peuranen S, Keinänen M, Tigerstedt C, Kokko J, Vuorinen PJ. 2002. The effects of Fe and Al exposure with or without humic acid at two pH levels on the gills, oxygen consumption and blood and plasma parameters of juvenile grayling (Thymallus thymallus). Arch. Hydrobiol. Supplement 141(3-4): 241-261.

Pichard A. 2005. Aluminium et ses dérivés. Fiche de données toxicologiques et environnementales des substances chimiques. INERIS. Available from : http://www.ineris.fr/index.php?
module=doc&action=getDoc&id_doc_object=134.

Pidwirny M, Gow T. 2002. The changing atmosphere. In: Pidwirny M, editor, Land use and environmental change in the Thompson-Okanagan. Victoria (BC): Royal British Columbia Museum. Available from: http://www.livinglandscapes.bc.ca/thomp-ok/
env-changes/index.html

Pierre F, Baruthio F, Diebold F, Biette P. 1995. Effect of different exposure compounds on urinary kinetics of aluminium and fluoride in industrially exposed workers. Occup Environ Med 52(6): 396-403.

Pilgrim W, Schroeder B. 1997. Multi-Media concentrations of heavy metals and major ions from urban and rural sites in New Brunswick, Canada. Environ Monit Assess 47: 89-108.

Pintro J, Barloy J, Fallavier P. 1996. Aluminum effects on the growth and mineral composition of corn plants cultivated in nutrient solution at low aluminum activity. J Plant Nutr 19: 729-741.

Poléo ABS. 1995. Aluminium polymerization - a mechanism of acute toxicity of aqueous aluminium to fish. Aquat Toxicol 31: 347-356.

Poléo ABS, Bjerkely F. 2000. Effect of unstable aluminium chemistry on Arctic char (Salvelinus alpinus). Can. J. Fish. Aquat. Sci. 57: 1423-1433.

Poléo ABS, Hytterød S. 2003. The effect of aluminium in Atlantic salmon (Salmo salar) with special emphasis on alkaline water. J Inorg Biochem 97: 89-96.

Poléo ABS, Lydersen E, Rosseland BO, Kroglund F, Salbu B, Vogt RD, Kvellestad A. 1994. Increased mortality of fish due to changing Al chemistry of mixing zones between limed streams and acidic tributaries. Water Air Soil Pollut 75: 339-351.

Poléo ABS, Øxnevad SA, Østbye K, Andersen RA, Oughton DH, Vøllestad LA. 1995. Survival of curcian carp, Carassius carassius, exposed to a high low-molecular weight inorganic aluminium challenge. Aquat Sci 57: 350-359 [cited in Alstad et al. 2005].

Polizzi S, Ferrara M, Bugiani M, Barbero D, Baccolo T. 2007. Aluminium and iron air pollution near an iron casting and aluminium foundry in Turin district (Italy). J Inorg Biochem. p 1339-43.

Poulos BK, Perazzolo M, Lee VM, Rudelli R, Wisniewski HM, Soifer D. 1996. Oral aluminum administration during pregnancy and lactation produces gastric and renal lesions in rat mothers and delay in CNS development of their pups. Mol Chem Neuropathol 29(1): 15-26.

Powell J, Thompson R. 1993. The chemistry of aluminium in the gastrointestinal lumen and its uptake and absorption. Proc Nutr Soc 52: 241-253.

Pratico D, Uryu K, Sung S, Tang S, Trojanowski JQ, Lee VM. 2002. Aluminum modulates brain amyloidosis through oxidative stress in APP transgenic mice. FASEB J 16(9): 1138-1140.

Priest ND. 1993. The bioavailability and metabolism of aluminium compounds in man. Proc Nutr Soc 52(1): 231-240.

Priest ND. 2004. The biological behaviour and bioavailability of aluminium in man, with special reference to studies employing aluminium-26 as a tracer: review and study update. J Environ Monit 6(5): 375-403.

Priest ND, Newton D, Day JP, Talbot RJ, Warner AJ. 1995. Human metabolism of aluminium-26 and gallium-67 injected as citrates. Hum Exp Toxicol 14(3): 287-293.

Priest ND, Talbot RJ, Austin JG, Day JP, King SJ, Fifield K, Cresswell RG. 1996. The bioavailability of ²⁶Al-labelled aluminium citrate and aluminium hydroxide in volunteers. Biometals 9(3): 221-228.

Priest ND, Talbot RJ, Newton D, Day JP, King SJ, Fifield LK. 1998. Uptake by man of aluminium in a public water supply. Hum Exp Toxicol 17(6): 296-301.

Quartin VL, Azinheira HG, Nunes MA. 2001. Phosphorus deficiency is responsible for biomass reduction of triticale in nutrient solution with aluminum. J Plant Nutr 24(12): 1901-1911.

Quartley B, Esselmont G, Taylor A, Dobrota M. 1993. Effect of oral aluminium citrate on short-term tissue distribution of aluminium. Food Chem Toxicol 31(8): 543-548.

Radunovic A, Ueda F, Raja KB, Simpson RJ, Templar J, King SJ, Lilley JS, Day JP, Bradbury MW. 1997. Uptake of 26-Al and 67-Ga into brain and other tissues of normal and hypotransferrinaemic mice. Biometals 10(3): 185-191.

Rahnema S, Jennings F. 1999. Accumulation and depletion of aluminum from various tissues of rats following aluminum citrate ingestion. OHIO J Sci 99(5): 98-101.

Rajasekaran K. 2000. Effects of combined exposure to aluminium and ethanol on food intake, motor behaviour and a few biochemical parameters in pubertal rats. Environ Toxicol Pharmacol 9(1-2): 25-30.

Ramamoorthy S. 1988. Effects of pH on speciation and toxicity of aluminum in rainbow trout (Salmo gardneri). Can.J. Fish. Aquat. Sci. 45: 634-642.

Rasmussen PE, Subramanian KS, Jessiman BJ. 2001. A multi-element profile of housedust in relation to exterior dust and soils in the city of Ottawa, Canada. Sci Total Environ 267(1-3): 125-140.

Ravi SM, Prabhu BM, Raju TR, Bindu PN. 2000. Long-term effects of postnatal aluminium exposure on acetylcholinesterase activity and biogenic amine neurotransmitters in rat brain. Indian J Physiol Pharmacol 44(4): 473-478.

Razniewska G, Trzcinka-Ochocka M. 2003. ET-AAS as a method for determination of aluminum in blood serum and urine. Chemia Analityczna 48: 107-113.

Reiber S, Kukull W, Standish-Lee P. 1995. Drinking water aluminium and bioavailability. Journal AWWA: 86-100.

Reimann C, Garrett RG. 2005. Geochemical background--concept and reality. Sci Total Environ 350(1-3): 12-27.

Rifat SL, Eastwood MR, McLachlan DR, Corey PN. 1990. Effect of exposure of miners to aluminium powder. Lancet 336(8724): 1162-1165.

Ritchie GSP. 1995. Soluble aluminium in acidic soils: principles and practicalities. Plant Soil 171: 17-27.

[RMOC] Region of Ottawa-Carleton. 2000.  Regional Facilities Environmental Effects Monitoring Project.  Preliminary Study of Ottawa-Carleton's Water Purification and Wastewater Treatment Plants.  Prepared by the Surface Water Quality Branch, Water Environment Protection Division, Environment and Transportation Department, Region of Ottawa Carleton.

Roberts MH, Diaz RD. 1985. Assessing the effects of alum sludge discharges into coastal streams. Recent advances in sludge treatment and disposal. Denver (CO): American Water Works Association [cited in Cornwell et al. 1987].

Rodella L, Rezzani R, Lanzi R, Bianchi R. 2001. Chronic exposure to aluminium decreases NADPH-diaphorase positive neurons in the rat cerebral cortex. Brain Res 889(1-2): 229-233.

Rodella LF, Ricci F, Borsani E, Rezzani R, Stacchiotti A, Mariani C, Bianchi R. 2006. Exposure to aluminium changes the NADPH-diaphorase/NPY pattern in the rat cerebral cortex. Arch Histol Cytol 69(1): 13-21.

Rogers MA, Simon DG. 1999. A preliminary study of dietary aluminium intake and risk of Alzheimer's disease. Age Ageing 28(2): 205-209.

Roig JL, Fuentes S, Teresa Colomina M, Vicens P, Domingo JL. 2006. Aluminum, restraint stress and aging: behavioral effects in rats after 1 and 2 years of aluminum exposure. Toxicology 218(2-3): 112-124.

Romano LS. 1971. Windsor nutrient removal studies. Sewage treatment. City of Windsor, Ontario, September. 4 p.

Rondeau V, Commenges D, Jacqmin-Gadda H, Dartigues JF. 2000. Relation between aluminum concentrations in drinking water and Alzheimer's disease: an 8-year follow-up study. Am J Epidemiol 152(1): 59-66.

Rondeau V, Jacqmin-Gadda H, Commenges D, Dartigues JF. 2001. Re: aluminum in drinking water and cognitive decline in elderly subjects: the Paquid cohort. Am J Epidemiol 154(3): 288-290.

Roskams AJ, Connor JR. 1990. Aluminum access to the brain: a role for transferrin and its receptor. Proc. Natl. Acad. Sci. USA 87: 9024-9027.

Rosseland BO, Blakar IA, Bulgar A, Krogland F, Kvellstad A, Lydersen E, Oughton DH, Salbu B, Staurnes M, Vogt R. 1992. The mixing zone between limed and acidic river waters: complex aluminium chemistry and extreme toxicity for salmonids.Environ Pollut 78: 3-8.

Rosseland BO, Eldhuset TD, Staurnes M. 1990. Environmental effects of aluminium. Environ Geochem Health 12(1-2): 17-27.

Roy AK, Sharma A, Talukder G. 1991. Effects of aluminum salts on bone marrow chromosomes in rats in vivo. Cytobios 66 : 105-111.

Roy AK, Talukder G, Sharma A. 1991. Similar effects in vivo of two aluminum salts on the liver, kidney, bone, and brain of Rattus norvegicus. Bull Environ Contam Toxicol 47(2): 288-295.

Roy R. 1999a. The chemistry, bioaccumulation and toxicity of aluminum in the aquatic environment for the PSL2 assessment of aluminum salts. Report prepared by Fisheries and Oceans Canada for Environment Canada. Montreal (QC): Environment Canada. 110 pp. Unpublished report.

Roy R. 1999b. A preliminary environmental assessment of aluminum salts in the aquatic environment. Report prepared by Fisheries and Oceans Canada for Environment Canada. Montreal (QC): Environment Canada. 18 p. Unpublished report.

Roy R, Campbell PGC. 1995. Survival time modeling of exposure of juvenile Atlantic salmon (Salmo salar) to mixtures of Al and Zn in soft water at low pH. Aquat Toxicol 33: 155-176.

Roy R, Campbell PGC. 1997. Decreased toxicity of aluminum to juvenile Atlantic salmon (Salmo salar) in acidic soft water containing natural organic matter: a test of the free-ion model. Environ Toxicol Chem 16: 1962-1969.

Ruby MV, Schoof R, Brattin W, Goldade M, Post G, Harnois M, Mosby DE, Casteel SW, Berti W, Carpenter M and others. 1999. Advances in evaluating the oral bioavailability of inorganics in soil for use in human health risk assessment. Environ Sci Technol 33(21): 3697-3705.

Rundgren S, Nilsson P. 1997. Sublethal effects of aluminium on earthworms in acid soil - the usefulness of Dendrodrilus rubidus (SAV.) in a laboratory test system. Pedobiologia 41: 417-436.

Rustad LE, Cronan CS. 1995. Biogeochemical controls on aluminum chemistry in the O horizon of a red spruce (Picea rubens Sarg.) stand in central Maine, USA. Biogeochemistry 29: 107-129.

Sadler K, Lynam S. 1985. The mineral content of some freshwater invertebrates in relation to stream pH and calcium concentration. Central Electric Research Laboratory, Leatherland, Surrey.

Sadler K, Lynam S. 1988. The influence of calcium on aluminium-induced changes in the growth rate and mortality of brown trout, Salmo trutta L. J Fish Biol 33: 171-179.

Sadovnick AD, Scheifele DW. 2000. School-based hepatitis B vaccination programme and adolescent multiple sclerosis. Lancet 355(9203): 549-550.

Sahin G, Taskin T, Benli K, Duru S. 1995. Impairment of motor coordination in mice after ingestion of aluminum chloride. Biol Trace Elem Res 50(1): 79-85.

Salib E, Hillier V. 1996. A case-control study of Alzheimer's disease and aluminium occupation. Br J Psychiatry 168(2): 244-249.

Sanchez DJ, Gomez M, Llobet JM, Corbella J, Domingo JL. 1997. Effects of aluminium on the mineral metabolism of rats in relation to age. Pharmacol Toxicol 80(1): 11-17.

Santschi PH, Nyffeler VP, Anderson RF, Schiff SL, O'Hara P, Hesslein RH. 1986. Response of radioactive trace metals to acid-base titrations in controlled ecosystems: evaluation of transport parameters for applications to whole-lake radio-tracer experiments. Can. J. Fish. Aquat. Sci. 43: 60-77.

Sarin S, Julka D, K.D. G. 1997. Regional alterations in calcium homeostasis in the primate brain following chronic aluminum exposure. Mol Cell Boiochem 168: 95-100.

Savory J. 2000. Aluminium - Mechanism report. Charlottesville, VA: Department of Pathology and Biochemistry and Molecular Genetics. University of Virginia, Health System. 1-57 p.

Savory J, Herman MM, Ghribi O. 2006. Mechanisms of aluminum-induced neurodegeneration in animals: Implications for Alzheimer's disease. J Alzheimers Dis: 135-144.

Schecher WD, McAvoy D. 1994. "MINEQL+: A chemical equilibrium program for personal computers". Version 3.01. Hallowell, Maine: Environmental Research Software; 107 p.

Schemel LE, Kimball BA, Bencala KE. 2000. Colloid formation and metal transport through two mixing zones affected by acid mine drainage near Silverton, Colorado. Appl Geochem 15: 1000-1018 [cited in Farag et al. 2007].

Schier GA. 1996. Effect of aluminum on growth of newly germinated and 1-year-old red spruce (Picea rubens) seedlings. Can J For Res 26: 1781-1787.

Schindler DW. 1988. Effect of acid rain on freshwater ecosystems. Science 239: 149-157.

Schindler DW, Hesslein RH, Wageman R, Broecker WS. 1980. Effects of acidification on mobilization of heavy metals and radionuclides from the sediments of a freshwater lake. Can. J. Fish. Aquat. Sci. 37: 373-377.

Schmidt PF, Zumkley H, Bertram H, Lison A, Winterberg B, Barckhaus R. 1984. Localization of aluminum in bone of patients with dialysis osteomalacia. In: Schramel PBaP, editor. Trace Elements-Analytical Chemistry in Medicine and Biology. New York: Walter de Gruyter & Co. p 475-482.

Schonholzer KW, Sutton RA, Walker VR, Sossi V, Schulzer M, Orvig C, Venczel E, Johnson RR, Vetterli D, Dittrich-Hannen B and others. 1997. Intestinal absorption of trace amounts of aluminium in rats studied with 26aluminium and accelerator mass spectrometry. Clin Sci (Lond) 92(4): 379-383.

Sedman AB, Wilkening GN, Warady BA. 1984. Clinical and laboratory observations. Encephalopathy in childhood secondary to aluminum toxicity. J Pediatr 105: 836-838.

Seip HM, Andersen S, Henriksen A. 1990. Geochemical control of aluminum concentrations in acidified surface waters. J Contam Hydrol 116: 299-305.

Servos MR, Rooke JB, Mackie GL. 1985. Reproduction of selected mollusca in some low alkalinity lakes in south-central Ontario. Can J Zool 63: 511-515.

Shah NR, Oberkircher OR, Lobel JS. 1990. Aluminum-induced microcytosis in a child with moderate renal insufficiency. Am J Pediatr Hematol Oncol 12(1): 77-79.

Shakoor A, Gupta PK, Kataria M. 2003. Influence of aluminium on neurotoxicity of lead in adult male albino rats. Indian J Exp Biol 41(6): 587-591.

Sharma P, Mishra KP. 2006. Aluminum-induced maternal and developmental toxicity and oxidative stress in rat brain: response to combined administration of Tiron and glutathione. Reprod Toxicol 21(3): 313-321.

Shcherbatykh I, Carpenter DO. 2007. The role of metals in the etiology of Alzheimer's disease. J Alzheimers Dis: 191-205.

Sherrard DJ, Walker JV, Boykin JL, Dis. AJK. 1988. Precipitation of dialysis dementia by deferoxamine treatment of aluminum-related bone disease. Am J Kidney Dis 12: 126-130.

Shi-Lei S, Guang-Yu MA, Bachelor LH, Bachelor ZY, Dong HM, Xu XH. 2005. Effect of naloxone on aluminum-induced learning and memory impairment in rats. Neurol India 53(1): 79-82.

Shock SS, Bessinger BA, Lowney YW, Clark JL. 2007. Assessment of the solubility and bioaccessibility of barium and aluminum in soils affected by mine dust deposition. Environ Sci Technol 41(13): 4813-4820.

Silva VS, Cordeiro JM, Matos MJ, Oliveira CR, Goncalves PP. 2002. Aluminum accumulation and membrane fluidity alteration in synaptosomes isolated from rat brain cortex following aluminum ingestion: effect of cholesterol. Neurosci Res 44(2): 181-193.

Silva VS, Duarte AI, Rego AC, Oliveira CR, Goncalves PP. 2005. Effect of chronic exposure to aluminium on isoform expression and activity of rat (Na+/K+)ATPase. Toxicol Sci 88(2): 485-494.

Silva VS, Goncalves PP. 2003. The inhibitory effect of aluminium on the (Na+/K+)ATPase activity of rat brain cortex synaptosomes. J Inorg Biochem 97(1): 143-150.

Sim M, Dick R, Russo J, Bernard B, Grubb P, Krieg E, Jr., Mueller C, McCammon C. 1997. Are aluminium potroom workers at increased risk of neurological disorders? Occup Environ Med 54(4): 229-235.

Sjogren B, Elinder CG, V. L, G. C. 1988. Uptake and urinary excretion of aluminum among welders. Int Arch Occup Environ Health 60: 77-79.

Sjogren B, Gustavsson P, Hogstedt C. 1990. Neuropsychiatric symptoms among welders exposed to neurotoxic metals. Br J Ind Med 47(10): 704-707.

Sjogren B, Iregren A, Frech W, Hagman M, Johansson L, Tesarz M, Wennberg A. 1996. Effects on the nervous system among welders exposed to aluminium and manganese. Occup Environ Med 53(1): 32-40.

Sjogren B, Lidums V, Håkansson M, Hedström L. 1985. Exposure and urinary excretion of aluminum during welding. Scand J Work Environ Health 11: 39-13.

Skinner BJ, Porter SC. 1989. The dynamic earth. Toronto (ON): John Wiley & Sons. 541 p.

Smeltzer E. 1990. A successful alum/aluminate treatment of Lake Morey, Vermont. Lake and Reservoir Management 6: 9-19.

Smith LF. 1995. Public health role, aluminium and alzheimer's disease. Environmetrics 6: 277-286.

Somova LI, Missankov A, Khan MS. 1997. Chronic aluminum intoxication in rats: dose-dependent morphological changes. Methods Find Exp Clin Pharmacol 19(9): 599-604.

Soucek DJ. 2006. Effects of freshly neutralized aluminum on oxygen consumption by freshwater invertebrates. Arch Environ Contam Toxicol 50: 353-360.

Sparks DL, Friedland R, Petanceska S, Schreurs BG, Shi J, Perry G, Smith MA, Sharma A, Derosa S, Ziolkowski C and others. 2006. Trace Copper Levels in the Drinking Water, but not Zinc or Aluminum Influence CNS Alzheimer-Like Pathology. J Nutr Health Aging 10(4): 247-254.

Sparling DW, Lowe TP. 1996. Environmental hazards of aluminum to plants, invertebrates, fish, and wildlife. Rev Environ Contam Toxicol 145: 1-127.

Sposito G, editor. 1996. The environmental chemistry of aluminum. 2nd edition. Boca Raton (FL): CRC Press.

Spry DJ, Wiener JG. 1991. Metal bioavailability and toxicity to fish in low-alkalinity lakes -- a critical review. Environ Pollut 71: 243-304.

Srinivasan PT, Viraghavan T, Kardash B, Bergman J. 1998. Aluminum speciation during drinking water treatment. Water Quality Research Journal of Canada 33: 377-388.

Stauber JL, Florence TM, Davies CM, Adams MS, Buchanan SJ. 1999. Bioavailability of Al in alum-treated drinking water. AWWA 91(11): 84-93.

Steiner KC, Barbour JR, McCormick LH. 1984. Response of Populus hybrids to aluminum toxicity. Forest Science 30: 404-410.

Steinhausen C, Kislinger G, Winklhofer C, Beck E, Hohl C, Nolte E, Ittel TH, Alvarez-Bruckmann MJ. 2004. Investigation of the aluminium biokinetics in humans: a ²⁶Al tracer study. Food Chem Toxicol 42(3): 363-371.

Stevens DP, McLaughlin MJ, Alston AM. 1997. Phytotoxicity of aluminium-fluoride complexes and their uptake from solution culture by Avena sativa and Lycopersicon esculentum. Plant Soil 192: 81-93 [cited in Manoharan et al. 2007].

Strong MJ, Garruto RM, Joshi JG, Mundy WR, Shafer TJ. 1996. Can the mechanisms of aluminum neurotoxicity be integrated into a unified scheme? J Toxicol Environ Health 48(6): 599-613.

Stumm W, Morgan JJ. 1981. Aquatic chemistry. An introduction emphasizing chemical equilibria in natural waters. New York (NY): John Wiley and Sons.

Sturkenboom M, Wolfson C, Rollet E, Heinzlef O, Abenhaim L. 2000. Demyelination multiple sclerosis and hepatitis B vaccination: a population based study in the UK. Neurology 54(A166).

Su C, Evans LJ. 1996. Soil solution chemistry and alfalfa response to CaCO3 and MgCO3 on an acidic gleysol. Canadian Journal of Soil Science 76: 41-47.

Suarez-Fernandez MB, Soldado AB, Sanz-Medel A, Vega JA, Novelli A, Fernandez-Sanchez MT. 1999. Aluminum-induced degeneration of astrocytes occurs via apoptosis and results in neuronal death. Brain Res 835(2): 125-136.

Swegert CV, Dave KR, Katyare SS. 1999. Effect of aluminium-induced Alzheimer like condition on oxidative energy metabolism in rat liver, brain and heart mitochondria. Mech Ageing Dev 112(1): 27-42.

Szerdahelyi P, Kasa P. 1988. Intraventricular administration of the cholinotoxin AF64A increases the accumulation of aluminum in the rat parietal cortex and hippocampus, but not in the frontal cortex. Brain res 444: 356-360.

Takita E, Koyama H, Hara T. 1999. Organic acid metabolism in aluminium-phosphate utilizing cells of carrot (Daucus carota L.). Plant Cell Physiol 40: 489-495 [cited in Manoharan et al. 2007].

Talbot RJ, Newton D, Priest ND, Austin JG, Day JP. 1995. Inter-subject variability in the metabolism of aluminium following intravenous injection as citrate. Hum Exp Toxicol 14(7): 595-599.

Tanaka A, Tadano T, Yamamoto K, Kananura N. 1987. Comparison of toxicity to plants among Al³⁺, AlSO₄⁺ and AlF complex ions. Soil Sci. Plant Nutr. 33: 43-55 [cited in Manoharan et al. 2007].

Taugbol G, Seip HM. 1994. Study of interaction of DOC with aluminium and hydrogen ion in soil and surface water using a simple equilibrium model. Environ Int 20: 353-361.

Taylor GA, Ferrier IN, McLoughlin IJ, Fairbairn AF, McKeith IG, Lett D, Edwardson JA. 1992. Gastrointestinal absorption of aluminium in Alzheimer's disease: response to aluminium citrate. Age Ageing 21(2): 81-90.

Tessier A, Campbell PGC. 1990. Partitioning of trace metals in sediments and its relationship to their accumulation in benthic organisms. In: Broekaert JAC, Guter S, Adams FB, editors, Metal speciation in the environment. Berlin (Germany): Springer-Verlag. p. 545-569.

Thomas KW, Pellizzari ED, Clayton CA, Whitaker DA, Shores RC, Spengler J, Ozkaynak H, Froehlich SE, Wallace LA. 1993. Particle Total Exposure Assessment Methodology (PTEAM) 1990 study: method performance and data quality for personal, indoor, and outdoor monitoring. J Expo Anal Environ Epidemiol 3(2): 203-226.

Thorne BM, Cook A, Donohoe T, Lyon S, Medeiros DM, Moutzoukis C. 1987. Aluminum toxicity and behavior in the weanling Long-Evans rat. Bull Psychon Soc 25(2): 129-132.

Thorne BM, Donohoe T, Lin KN, Lyon S, Medeiros DM, Weaver ML. 1986. Aluminum ingestion and behavior in the Long-Evans rat. Physiol Behav 36(1): 63-67.

Thornton FC, Schaedle M, Raynal DJ. 1986a. Effects of aluminum on the growth, development, and nutrient composition of honeylocust (Gleditsia triacanthos L.) seedlings. Tree Physiol 2: 307-316.

Thornton FC, Schaedle M, Raynal DJ, Zipperer C. 1986b. Effects of aluminium on honeylocust (Gleditsia triacanthos L.) seedlings in solution culture. J Exp Bot 37: 775-785.

Tietge JE, Johnson RD, Bergman HL. 1988. Morphometric changes in gill secondary lamellae of brook trout (Salvelinus fontinalis) after long-term exposure to acid and aluminum. Can. J. Fish. Aquat. Sci. 45: 1643-1648.

Tipping E. 1994. WHAM [Windermere Humic Acid Model] - A chemical equilibrium model and computer code for waters, sediments, and soils incorporating a discrete site electrostatic model of ion-binding by humic substances. Comput Geosci 20: 73-1023.

Touze E, Fourrier A, Rue-Fenouche C, Ronde-Oustau V, Jeantaud I, Begaud B, Alperovitch A. 2002. Hepatitis B vaccination and first central nervous system demyelinating event: a case-control study. Neuroepidemiology 21(4): 180-186.

Touze E, Gout O, Verdier-Taillefer MH, Lyon-Caen O, Alperovitch A. 2000. The first episode of central nervous system demyelinization and hepatitis B virus vaccination. Rev Neurol 156(3): 242-246.

Troutman DE, Peters NE. 1982. Deposition and transport of heavy metals in three lake basins affected by acidic precipitation in the Adirondack Mountains, New York. In: Keith LH editor, Energy and environmental chemistry. Vol. 2. Ann Arbor (MI): Ann Arbor Science, p. 33-61.

Tsunoda M, Sharma RP. 1999a. Modulation of tumor necrosis factor alpha expression in mouse brain after exposure to aluminum in drinking water. Arch Toxicol 73(8-9): 419-426.

Tsunoda M, Sharma RP. 1999b. Altered dopamine turnover in murine hypothalamus after low-dose continuous oral administration of aluminum. J Trace Elem Med Biol 13(4): 224-231.

Turmel MC, Courchesne F. 2007. Influence of microbial activity and trace metal speciation in the rhizosphere on metal uptake by wheat. Unpublished manuscript. Département de géographie, Université de Montréal, Montréal, Québec.

Urban NR, Gorham E, Underwood JK, Martin FB, Ogden JG. 1990. Geochemical processes controlling concentrations of Al, Fe, and Mn in Nova Scotia lakes. Limnol Oceanogr 35: 1516-1534.

Valkonen S, Aitio A. 1997. Analysis of aluminium in serum and urine for the biomonitoring of occupational exposure. Sci Total Environ 199(1-2): 103-110.

van Coillie R, Thellen C, Campbell PGC, Vigneault Y. 1983. Effets toxiques del'aluminium chez les salmonidés en relation avec des conditions physico-chimiques acides. Can. Tech. Rep. Fish. Aquat. Sci. No. 1237. 88 p.

Van Gestel CAM, Hoogerwerf G. 2001. Influence of soil pH on the toxicity of aluminium for Eisenia Andrei (Oligochaeta: Lumbricidae) in an artificial soil substrate. Pedobiologia 45: 385-395.

Van Ginkel MF, van der Voet GB, D'Haese PC, De Broe ME, Wolff FA. 1993. Effect of citric acid and maltol on the accumulation of aluminum in rat brain and bone. J Lab Clin Med 121: 453-460.

Vance GF, Stevenson FJ, Sikora FJ. 1996. Environmental chemistry of aluminum-organic complexes. In: Sposito, G. (ed.), The environmental chemistry of aluminum. Boca Raton (FL): CRC Press. p. 169-220.

Varner JA, Horvath WJ, Huie CW, Naslund HR, Isaacson RL. 1994. Chronic aluminum fluoride administration. I. Behavioral observations. Behav Neural Biol 61(3): 233-241.

Varner JA, Huie CW, Horvath W, Jensen KF, Isaacson RL. 1993. Chronic ALF, Administration: II Selected histological observations. Neurosci Res Commun 13(2): 99-103.

Varner JA, Jensen KF, Horvath W, Isaacson RL. 1998. Chronic administration of aluminum-fluoride or sodium-fluoride to rats in drinking water: alterations in neuronal and cerebrovascular integrity. Brain Res 784(1-2): 284-298.

Vasiloff GN. 1991. Phytotoxicology Section investigation in the vicinity of Welland Chemical, Sarnia on August 23, 1989. Toronto (ON): Ontario Ministry of the Environment, Report PIBS 1434.

Vasiloff GN. 1992. Phytotoxicology surveillance investigation in the vicinity of Welland Chemical, Sarnia, 1991. Toronto (ON): Ontario Ministry of the Environment, Report PIBS 2072E.

Verbost PM, Berntssen MHG, Kroglund F, Lydersen E, Witters HE, Rosseland BO, Salbu B, Wendelaar Bonga SE. 1995. The toxic mixing zone of neutral and acidic river water: Acute aluminium toxicity in brown trout (Salmo trutta L.). Water Air Soil Pollut 85(2): 341-346.

Verreault R, Laurin D, Lindsay J, De Serres G. 2001. Past exposure to vaccines and subsequent risk of Alzheimer's disease. Can Med Assoc J 165(11): 1495-1498.

Verstraeten SV, Erlejman AG, Zago MP, Oteiza PI. 2002. Aluminum affects membrane physical properties in human neuroblastoma (IMR-32) cells both before and after differentiation. Arch Biochem Biophys 399(2): 167-73.

Verstraeten SV, Keen CL, Golub MS, Oteiza PI. 1998. Membrane composition can influence the rate of Al³⁺ -mediated lipid oxidation: effect of galoactolipids. Biochem journal 333: 833-838.

Vogt T. 1986. Water quality and health. A story of a possible relationship between aluminium in drinking water and age-related dementia. Oslo, Norway. 145 p.

von Linstow Roloff E, Platt B, Riedel G. 2002. Long-term study of chronic oral aluminum exposure and spatial working memory in rats. Behav Neurosci 116(2): 351-356.

Walker VR, Sutton RA, Meirav O, Sossi V, Johson R, Klein J, Fink D, Middleton R. 1994. Tissue disposition of 26aluminum i rats measured by accelerator mass spectrometry. Clin Invest Med 17: 420-425.

Walton J, Hams G, Wilcox D. 1994. Bioavailability of aluminium from drinking water: co-exposure with foods and beverage. Water Services Association of Australia. Report nr WSAA No.83.

Wang M, Chen JT, Ruan DY, Xu YZ. 2002a. The influence of developmental period of aluminum exposure on synaptic plasticity in the adult rat dentate gyrus in vivo. Neuroscience 113(2): 411-419.

Wang M, Ruan DY, Chen JT, Xu YZ. 2002b. Lack of effects of vitamin E on aluminium-induced deficit of synaptic plasticity in rat dentate gyrus in vivo. Food Chem Toxicol 40(4): 471-478.

Weatherley NS, Rutt GP, Thomas SP, Ormerod SJ. 1991. Liming acid streams - aluminium toxicity to fish in mixing zones. Water Air Soil Pollut 55: 345-353.

Weberg R, Berstad A. 1986. Gastrointestinal absorption of aluminium from single doses of aluminium containing antacids in man. Eur J Clin Invest 16(5): 428-432.

Wettstein A, Aeppli J, Gautschi K, Peters M. 1991. Failure to find a relationship between mnestic skills of octogenarians and aluminum in drinking water. Int Arch Occup Environ Health 63(2): 97-103.

Wheeler DM, Dodd MB. 1995. Effect of aluminium on yield and plant chemical concentrations of some temperate legumes. Plant Soil 173: 133-145.

Wheeler DM, Edmeades DC, Christie RA. 1992. Effect of aluminum on relative yield and plant chemical concentrations for cereals grown in solution culture at low ionic strength. J Plant Nutr 15: 403-418.

White DM, Longstreth WT, Jr., Rosenstock L, Claypoole KH, Brodkin CA, Townes BD. 1992. Neurologic syndrome in 25 workers from an aluminum smelting plant. Arch Intern Med 152(7): 1443-1448.

Wier, Dixon. 2008.  Director of Water and Wastewater Services, City of Ottawa. Personal communication with John Pasternak, Environment Canada, 6 August 2008.

Wilkinson KJ, Campbell PGC. 1993. Aluminum bioconcentration at the gill surface of juvenile Atlantic salmon in acidic media. Environ Toxicol Chem 12: 2083-2095.

William M, Dave M. 2000. Phytotoxicology Soil investigation- School Yards and Beaches Port Colborne (2000). Port Colborne: Ministry of Environmen,. Ontario. 30 p.

Wisniewski HM, Lidsky TI. 1997. The role of aluminium in alzheimer's disease. In: Press MU, editor. Managing Health in the Aluminium Industry. London. p 263-273.

Witters H, Vangenechten JHD, Van Puymbroeck S, Vanderborght OLJ. 1984. Interference of aluminium and pH on the Na-influx in an aquatic insect Corixa punctata (Illig.). Bull Environ Contam Toxicol 32: 575-579.

Witters HE, Van Puymbroeck S, Stouthart AJHX, Bonga SEW. 1996. Physicochemical changes of aluminium in mixing zones: mortality and physiological disturbances in brown trout (Salmo trutta L.). Environ Toxicol Chem 15: 986-996.

Witters HE, Van Puymbroeck S, Vangenechten JHD, Vanderborght OLJ. 1990. The effect of humic substances on the toxicity of aluminium to adult rainbow trout, Oncorhynchus mykiss (Walbaum). J Fish Biol 37: 43-53.

Wobma P, Kjartanson K, Bellamy B, Pernitsky D. 2001. Effects of cold water temperatures on water treatment unit processes. In: New Horizons in Drinking Water, Proceedings of the Annual Conference of the American Water Works Association, Washington, D.C., June 17-21, 2001.

Wold LA, Moore BC, Dasgupta N. 2005. Life-history responses of Daphnia pulex with exposure to aluminum sulfate. Lake and Reservoir Management 21(4): 383-390.

Wong HKT, Nriagu JO, McCabe KJ. 1989. Aluminum species in porewaters of Kejimkujik and Mountain lakes, Nova Scotia. Water Air Soil Pollut 46: 155-164.

Wood CM, McDonald DG, Booth CE, Simons BP, Ingersoll CG, Bergman HL. 1988a. Physiological evidence of acclimation to acid/aluminum stress in adult brook trout (Salvelinus fontinalis). 1. Blood composition and net sodium fluxes. Can. J. Fish. Aquat. Sci. 45: 1587-1596.

Wood CM, Simons BP, Mount DR, Bergman HL. 1988b. Physiological evidence of acclimation to acid/aluminum stress in adult brook trout (Salvelinus fontinalis). 2. Blood parameters by cannulation. Can. J. Fish. Aquat. Sci. 45: 1597-1605.

Wood DJ, Cooper C, Stevens J, Edwardson J. 1988. Bone mass and dementia in hip fracture patients from areas with different aluminium concentrations in water supplies. Age Ageing 17(6): 415-419.

[WHO] World Health Organization. 1997. Aluminium. Geneva.

Wren CD, Stephenson GL. 1991. The effect of acidification on the accumulation and toxicity of metals to freshwater invertebrates. Environ Pollut 71: 205-241.

Wright RJ, Baligar VC, Wright SF. 1987. Estimation of phytotoxic aluminium in soil solution using three spectrophotometric methods. Soil Sci 144: 224-233 [cited in Manoharan et al. 2007].

Wu J, Zhou CY, Wong MK, Lee HK, Ong CN. 1997. Urine levels of aluminum after drinking tea. Biol Trace Elem Res 57(3): 271-280.

Xu ZX, Pai SM, Melethil S. 1991. Kinetics of aluminum in rats. II: Dose-dependent urinary and biliary excretion. J Pharm Sci 80(10): 946-951.

Yang MS, Wong MH. 2001. Changes in Ca, Cu, Fe, Mg, and Zn contents in mouse brain tissues after prolonged oral ingestion of brick tea liquor containing a high level of Al. Biol Trace Elem Res 80(1): 67-76.

Yokel R. 2006. Blood-brain barrier flux of aluminum, manganese, iron and other metals suspected to contribute to metal-induced neurodegeneration. J Alzheimers Dis 10: 223-253.

Yokel RA. 1985. Toxicity of gestational aluminum exposure to the maternal rabbit and offspring. Toxicol Appl Pharmacol 79: 121-133.

Yokel RA. 2000. The toxicology of aluminum in the brain: a review. Neurotoxicology 21(5): 813-28.

Yokel RA. 2001. Aluminum toxicokinetics at the blood-barrier. In: Exley C, editor. Aluminium and Alzheimer's Disease. New York: Elsevier.

Yokel RA. 2007. Brain aluminum influx rate determined with the tracer ²⁶Al in the rat. Poster presented at the 46th Annual Meeting of the Society of Toxicology, March 25-29, 2007. Charlotte, NC.

Yokel RA, Florence RL. 2006. Aluminum bioavailability from the approved food additive leavening agent acidic sodium aluminum phosphate, incorporated into a baked good, is lower than from water. Toxicology 227: 86-93.

Yokel RA, Hicks CL, Florence RL. 2008. Aluminum bioavailability from basic sodium aluminum phosphate, an approved food additive emulsifying agent, incorporated in cheese. Food Chem Toxicol 46: 2261-2266.

Yokel RA, McNamara PJ. 1985. Aluminum bioavailability and disposition in adult and immature rabbits. Toxicol Appl Pharmacol 77(2): 344-352.

Yokel RA, McNamara PJ. 1988. Influence of renal impairment, chemical form, and serum protein binding on intravenous and oral aluminum kinetics in the rabbit. Toxicol Appl Pharmacol 95(1): 32-43.

Yokel RA, McNamara PJ. 2000. Aluminum bioavailability (a compilation and critical review). Prepared for the Environmental Substances Division, Health Canada.

Yokel RA, McNamara PJ. 2001. Aluminium toxicokinetics: an updated minireview. Pharmacol Toxicol 88(4): 159-167.

Yokel RA, Rhineheimer SS, Brauer RD, Sharma P, Elmore D, McNamara PJ. 2001a. Aluminum bioavailability from drinking water is very low and is not appreciably influenced by stomach contents or water hardness. Toxicology 161: 93-101.

Yokel RA, Rhineheimer SS, Sharma P, Elmore D, McNamara PJ. 2001b. Entry, half-life, and desferrioxamine-accelerated clearance of brain aluminum after a single (26)Al exposure. Toxicol Sci 64(1): 77-82.

Young LB, Harvey HH. 1991. Metal concentrations in chironomids in relation to the geochemical characteristics of surficial sediments. Arch Environ Contam Toxicol 21: 202-211.

Yuan B, Klein MH, Contiguglia RS, Mishell JL, Seligman PA, Miller NL, Molitoris BA, Alfrey AC, Shapiro JI. 1989. The role of aluminum in the pathogenesis of anemia in an outpatient hemodialysis population. Ren Fail 11: 91-96.

Yuan G, Soma M, Seyama H, Theng BKG, Lavkukich LM, Takamatsu T. 1998. Assessing the surface composition of soil particules from some Podzolic soils by X-Ray photoelectron spectroscopy. Geoderma 86: 169-181.

Yumoto S, Nagai H, Matsuzaki H, Kobayashi T, Tada W, Ohki Y, Kakimi S, Kobayashi K. 2000. Transplacental passage of ²⁶Al from pregnant rats to fetuses and ²⁶Al transfer through maternal milk to suckling rats. Nucl Instrum Methods Phys Res B 172: 925-929.

Yumoto S, Nagai H, Matsuzaki H, Matsumura H, Tada W, Nagatsuma E, Kobayashi K. 2001. Aluminium incorporation into the brain of rat fetuses and sucklings. Brain Res Bull 55(2): 229-234.

Zafar TA, Weaver CM, Martin BR, Flarend R, Elmore D. 1997. Aluminum (26AI) metabolism in rats. Proc Soc Exp Biol Med 216(1): 81-85.

Zatta P, Ibn-Lkhayat-Idrissi M, Zambenedetti P, Kilyen M, Kiss T. 2002. In vivo and in vitro effects of aluminum on the activity of mouse brain acetylcholinesterase. Brain Res Bull 59(1): 41-45.

Zatta P, Zambenedetti P, Reusche E, Stellmacher F, Cester A, Albanese P, Meneghel G, Nordio M. 2004. A fatal case of aluminium encephalopathy in a patient with severe chronic renal failure not on dialysis. Nephrol Dial Transplant 19(11): 2929-2931.

Zatta PF, Favarato M, Nicolini M. 1993. Deposition of aluminium in brain tissues of rats exposed to inhalation of aluminum acetylacetonate. Neuro Rep 4: 1119-1122.

Zheng YX, Liang YX. 1998. The antagonistic effects of L-do J Inorg Biochem and eserine on Al-induced neurobehavioral deficits in rats. Biomed Environ Sci 11(4): 321-330.

Zhou Y, Harris WR, Yokel RA. 2008. The influence of citrate, maltolate and fluoride on the gastrointestinal absorption of aluminum at a drinking water-relevant concentration: a ²⁶Al and 14C study. J Inorg Biochem 102(4): 798-808.

Zhou Y, Yokel RA. 2005. The chemical species of aluminum influences its paracellular flux across and uptake into Caco-2 cells, a model of gastrointestinal absorption. Toxicol Sci 87(1): 15-26.

Zipp F, Weil JG, Einhaupl KM. 1999. No increase in demyelinating diseases after hepatitis B vaccination. Nat Med 5(9): 964-5.

Search Methodology, PSL2 Draft Assessment, Aluminum Salts

Toxicological and Epidemiological Data

A comprehensive search of the toxicological and epidemiological literature in relation to the health effects of aluminum was carried out in preparation of the SOS report published in 2000 (Environment Canada and Health Canada 2000). Since publication of this report, literature searches were conducted using the databases Toxline, Pubmed, and Current Contents as well as the organizational Web sites from the standard Existing Substances Division literature search list (ATSDR, ECETOC, IPCS, NICNAS, Health Canada, National Toxicology Program, WHO/Air, WHO/Water). The keywords (in truncated forms) included "aluminum" plus "toxicity", "neurotoxicity", "epidemiology", "bioavailability", "mode of action", "reproductive" and "developmental". For databases functioning on the basis of CAS registration numbers, the CAS RN of aluminum chloride (7446-70-0), aluminum nitrate (13473-90-0) and aluminum sulphate (10043-01-3) were used.

The comprehensive literature search was conducted through 2007. Some articles published in 2008 may also be included.

In addition to evaluating original study reports in the peer-reviewed literature, Health Canada consulted four recent comprehensive reviews of the literature on the toxic effects of aluminum: ATSDR (2006); InVS-Afssa-Afssaps (2003); JECFA (2006); and Krewski et al. (2007). These reviews were used primarily to supplement the literature search and are also cited as sources for some toxicological and exposure information, where appropriate. However, for all issues central to Health Canada's evaluation of human health risks, the original articles were consulted and cited.

Exposure Data

(see text for sources of data for exposure estimates).

Environmental evaluation

Data relevant to the risk characterization of aluminum chloride, aluminum nitrate and aluminum sulphate to the environment were identified from existing review documents, published reference texts and online searches of the following databases: Aqualine, ASFA (Aquatic Sciences and Fisheries Abstracts, Cambridge Scientific Abstracts; 1996), BIOSIS (Biosciences Information Services; 1990-1996), CAB (Commonwealth Agricultural Bureaux), CESARS (Chemical Evaluation Search and Retrieval System, Ontario Ministry of the Environment and Michigan Department of Natural Resources; 1996), Chemical Abstracts (Chemical Abstracts Service, Columbus, Ohio), CHRIS (Chemical Hazard Release Information System; 1964-1985), Current Contents (Institute for Scientific Information; 1993-1996), ELIAS (Environmental Library Integrated Automated System, Environment Canada library; January 1996), Enviroline (R.R. Bowker Publishing Co.; November 1995-June 1996), Environmental Abstracts (1975-February 1996), Environmental Bibliography (Environmental Studies Institute, International Academy at Santa Barbara; 1990-1996), GEOREF (Geo Reference Information System, American Geological Institute; 1990-1996), HSDB (Hazardous Substances Data Bank, U.S. National Library of Medicine; 1990-1996), Life Sciences (Cambridge Scientific Abstracts; 1990-1996), NTIS (National Technical Information Service, U.S. Department of Commerce), Pollution Abstracts (Cambridge Scientific Abstracts, U.S. National Library of Medicine), POLTOX (Cambridge Scientific Abstracts, U.S. National Library of Medicine; 1990-1995), RTECS (Registry of Toxic Effects of Chemical Substances, U.S. National Institute for Occupational Safety and Health; 1996), Toxline (U.S. National Library of Medicine; 1990-1996), TRI93 (Toxic Chemical Release Inventory, U.S. Environmental Protection Agency, Office of Toxic Substances; 1993), USEPA-ASTER (Assessment Tools for the Evaluation of Risk, U.S. Environmental Protection Agency; up to December 21, 1994), WASTEINFO (Waste Management Information Bureau of the American Energy Agency; 1973-September 1995) and Water Resources Abstracts (U.S. Geological Survey, U.S. Department of the Interior; 1990-1996). A further search of the scientific literature was conducted in 2007 using SciFinder, an electronic interface that allows access to six databases: CA Plus (Literature from journals, patents, books, conferences, etc.), Registry (substances), Chemlist (regulatory listing), ChemCats (commercial chemical suppliers), CASReact (reaction database) and Medline.

As well as retrieving references from literature database searches, direct contacts were made with researchers, academics and other government agencies. In addition, a survey of Canadian industry was carried out under authority of section 16 of CEPA (Environment Canada 1997b), and a second review aimed at identifying changes in use trends and quantities was conducted in 2007 (Cheminfo Services Inc. 2008). Companies were required to provide information on uses, releases, environmental concentrations, effects or other data that were available to them and related to aluminum salts. Ongoing scans were conducted of the open literature, conference proceedings and the Internet for relevant information. Data obtained to August 2008 were considered in this assessment report.

Notes:
AD = Alzheimer's disease
Al = Aluminium
F = Fluoride
LSA = Ontario Longitudinal Study of Aging
NFT = neurofibrillary tangles
OR = Odds ratio
PAQUID = Principle Lifetime Occupation and Cognitive Impairment in a French Elderly Cohort study (≥65 years old)
RR = Relative risk
Si = Silicon
Criteria for Alzheimers or dementia diseases:
ADRDA = Alzheimer's Disease and Related Disorders Associations
DSM = Diagnostic and Statistical Manual of Mental Disorders
ICD = International Classification of Diseases (World Health Organization)
MMS = mini-mental state examination
NINCDS = National Institute of Neurological and Communicative Disorders and Stroke

Tables

^* Dose calculated with Health Canada's reference values for body weights and intakes (Health Canada 1994).
^** Dose calculated with author's reported body weights and intakes.

^*Dose calculated with Health Canada's reference values for body weights and intakes (Health Canada 1994).
^**Dose calculated with author's reported body weights and intakes.