Draft assessment - Low boiling point naphthas group

Official title: Draft Assessment - Low Boiling Point Naphthas Group

Environment and Climate Change Canada

Health Canada

March 2024

Synopsis

Pursuant to section 68 of the Canadian Environmental Protection Act, 1999 (CEPA), the Minister of the Environment and the Minister of Health have conducted an assessment of 27 substances referred to collectively under the Chemicals Management Plan as the Low Boiling Point Naphthas (LBPNs) Group. The Chemical Abstracts Service Registry Numbers (CAS RN^{Footnote 1}), their Domestic Substances List (DSL) names, and their subgroupings are listed in the table below.

^a This CAS RN is a UVCB (unknown or variable composition, complex reaction products, or biological materials).

For the purposes of the human health assessment, the LBPN substances in this assessment were divided into 4 subgroups: C₉–C₁₄ hydrocarbon solvents (subgroup 1), C₉ aromatic hydrocarbon solvents (subgroup 2), the C₆–C₉ aliphatic hydrocarbon solvents (subgroup 3), and the LBPNs with no identified uses in products available to consumers (subgroup 4). The primary basis for the subgroups was the differences in health effects, carbon ranges for the hydrocarbon components, aromatic content, and potential uses in products available to consumers. Within each subgroup, the CAS RNs are assumed to be interchangeable with respect to their use in categories of products available to consumers.

The 27 LBPNs in this assessment are complex and highly variable combinations of hydrocarbons produced by the distillation of crude oil, followed by the fractionation of the resultant hydrocarbon streams by boiling range. They contain straight and branched chain alkanes (that is, paraffins and isoparaffins), cycloalkanes (also known as cycloparaffins), alkenes (also known as olefins), and aromatic hydrocarbons, predominantly in the carbon range of C₄ to C₁₄.

The 27 LBPN substances in this assessment may be used at the refinery where they are produced; blended into substances leaving the refinery under different CAS RNs; or transported by truck or train to other petroleum or non-petroleum sector facilities for use as feedstocks or to be blended with other feedstocks, resulting in a new substance with a new CAS RN. The exposures and risks associated with petroleum industry uses of the 27 LPBN substances are considered to be similar to those of the site-restricted and industry-restricted LBPNs for which risk assessment activities are considered to have already taken place under CEPA and are not considered further in this assessment. Several of the 27 LBPN substances are also identified as being used in non-petroleum industries as solvents and are present in products available to consumers, including in cosmetics, automotive products, paints and coatings, adhesives and sealants, household cleaning products, and as formulants in pest control products. Some may also be used in the manufacture of food packaging materials and in incidental additives used in food processing establishments.

Since the aromatic content of the LBPNs in many commercial or industrial uses is unknown, aromatic contents ranging from 0% to 100 % by weight were considered in the ecological assessment. Empirical aquatic toxicity data for LBPNs were comparable with modelled toxicity values. However, the predicted no-effect concentrations were determined on the basis of modelled aquatic toxicity data, which made it possible to estimate toxicity in the most sensitive species for LBPNs with a range of aromatic contents, while taking into consideration the compositional changes after wastewater treatment.

Predicted environmental concentrations of LBPNs were estimated for the 3 exposure scenarios with the highest potential for release to the environment, including a consumer release scenario resulting from the use of products available to consumers such as paints and coatings, adhesives and sealants, personal care and cosmetic products, household cleaners, and automotive care products; a generic formulation scenario for products available to consumers and industrial applications; and a pulp and paper scenario for the use of processing aids by pulp and paper mills. Environmental concentrations in the aquatic environment following wastewater treatment associated with releases from these uses were estimated and compared to modelled predicted no effect concentrations on the basis of the predicted composition of LBPNs in the effluent. In addition, the concentration of LBPNs in soils following the application of biosolids from wastewater treatment facilities to soil were compared to predicted no effect concentrations for soil organisms. On the basis of these comparisons, LBPNs are considered unlikely to be causing ecological harm to aquatic and soil organisms. Because of the volatile nature of LBPNs, terrestrial wildlife may be exposed to LBPNs through inhalation. Comparison of a critical toxicity value and estimated annual emission rate of LBPNs to air in this assessment with those in previous assessments on petroleum site-restricted and industry-restricted LBPNs indicates that the LBPNs examined in this assessment are unlikely to be causing ecological harm to terrestrial wildlife.

The potential for cumulative effects was considered in this assessment by examining cumulative exposures from within the group of Petroleum LBPNs. Considering all available lines of evidence presented in this draft assessment, there is a low risk of harm to the environment from the 27 LBPNs in this assessment. It is proposed to conclude that the 27 LBPNs in this assessment do not meet the criteria under paragraph 64(a) or (b) of CEPA as they 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.

A critical health effect for the LBPNs was initially considered to be genotoxicity or carcinogenicity of benzene. However, the LBPNs in this assessment contain less than 0.1% benzene according to available data; therefore, concerns associated with benzene genotoxicity or carcinogenicity are not considered applicable, and other critical effects were taken into consideration.

For subgroup 1, the C₉–C₁₄ hydrocarbon solvents, chronic neurotoxicity effects on humans and developmental neurotoxicity effects on laboratory animals were identified in inhalation studies. These effects were considered applicable to long-term and short-term exposure inhalation scenarios, respectively. With respect to the dermal route, effects on the peripheral nerve system in laboratory animals were considered to be critical effects following short and long-term exposure. A comparison of the critical effects levels with the estimated exposures from certain cosmetics and other products available to consumers in Canada resulted in margins of exposure (MOEs) that are considered to be potentially inadequate to address uncertainties in the health effects and exposure databases. On the basis of the estimated maximum environmental media exposures (air and water) of these substances, the risk to human health from exposure to these subgroup 1 substances was considered to be low.

For subgroup 2, the C₉ aromatic hydrocarbon solvents, reduced body weights, dose-related mortality in maternal animals, reduced fetal body weights, and delayed ossification in their offspring animals were identified in inhalation studies conducted on laboratory animals. These effects were considered applicable for both short-term and long-term exposure scenarios. With respect to the oral route, effects on the liver and kidneys were considered to be critical effects following long-term exposure. A comparison of the estimated exposures from certain cosmetics and other products available to consumers including nail polish, nail adhesive, spray paint, stain, floor polish, and lacquer, to the critical effect levels resulted in MOEs that are considered to be potentially inadequate. On the basis of the estimated maximum environmental media exposures (air and water) of these substances, the risk to human health from exposure to these subgroup 2 substances was considered to be low.

For subgroup 3, the C₆–C₉ aliphatic hydrocarbon solvents, limited health effects data were available for these substances; therefore, toxicological data from representative UVCBs were used to inform the health effects assessment. A reproductive and developmental toxicity endpoint identified in laboratory animals was considered applicable for short-duration exposures, while long-term exposures were compared to a carcinogenicity study in mice. A comparison of the estimated exposures from certain products available to consumers in Canada to the critical effect levels resulted in MOEs that are considered to be potentially inadequate. On the basis of the estimated maximum environmental media exposures (air and water) of these substances, the risk to human health from exposure to these subgroup 3 substances was considered to be low.

The health effects associated with subgroup 4, the LBPNs with no identified uses in products available to consumers, were characterized in a similar fashion to the previous 3 groupings. On the basis of the estimated maximum environmental media exposures (air and water) of these substances, the risk to human health from exposure to these substances was considered to be low.

The human health assessment took into consideration those groups of individuals within the Canadian population who, due to greater susceptibility or greater exposure, may be more vulnerable to experiencing adverse health effects. In addition, the potential for elevated exposure for people living near non-petroleum industrial facilities that may release some of these substances was considered in the assessment. The potential for cumulative effects was considered in this assessment by examining cumulative exposures from within the groups of Petroleum LBPNs. 

Considering all of the information presented in this draft assessment, it is proposed to conclude that the 9 C₉–C₁₄ hydrocarbon solvents (subgroup 1; CAS RNs 8030-30-6, 8032-32-4, 8052-41-3, 64475-85-0, 64741-41-9, 64741-65-7, 64742-48-9, 64742-82-1, and 64742-88-7), one C₉ aromatic solvents (subgroup 2; CAS RN 64742-95-6), and 7 C₆–C₉ aliphatic solvents (subgroup 3; CAS RNs 64741-66-8, 64741-84-0, 64742-49-0, 64742-89-8, 68410-97-9, 68647-60-9, and 426260-76-6), which occur in products available to consumers meet the criteria under paragraph 64(c) of CEPA as they are entering or may enter the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health. Considering all of the information presented in this draft assessment, it is proposed to conclude that the 10 LBPNs in subgroup 4 with no identified uses in products available to consumers (CAS RNs 64741-68-0, 64741-92-0, 64741-98-6, 68333-81-3, 68512-78-7, 68513-03-1, 68553-14-0, 68603-08-7, 68920-06-9, and 70693-06-0) do not meet the criteria under paragraph 64(c) of CEPA as they are not entering the environment in a quantity or concentration under conditions that constitute or may constitute a danger in Canada to human life or health.

It is therefore proposed to conclude that the 17 LBPNs in C₉–C₁₄ hydrocarbon solvents (subgroup 1; CAS RNs 8030-30-6, 8032-32-4, 8052-41-3, 64475-85-0, 64741-41-9, 64741-65-7, 64742-48-9, 64742-82-1, and 64742-88-7), C₉ aromatic solvents (subgroup 2; CAS RN 64742-95-6) and C₆–C₉ aliphatic solvents (subgroup 3; CAS RNs 64741-66-8, 64741-84-0, 64742-49-0, 64742-89-8, 68410-97-9, 68647-60-9 and 426260-76-6)  meet one or more of the criteria set out in section 64 of CEPA. It is therefore proposed to conclude that the 10 LBPNs in subgroup 4 with no identified uses in products available to consumers (CAS RNs 64741-68-0, 64741-92-0, 64741-98-6, 68333-81-3, 68512-78-7, 68513-03-1, 68553-14-0, 68603-08-7, 68920-06-9, and 70693-06-0) do not meet any of the criteria set out in section 64 of CEPA.

1. Introduction

Pursuant to section 68 of the Canadian Environmental Protection Act, 1999 (CEPA) (Canada 1999), the Minister of the Environment and the Minister of Health have conducted an assessment of 27 substances referred to collectively under the Chemicals Management Plan as the Low Boiling Point Naphthas (LBPNs) Group to determine whether these substances present or may present a risk to the environment or to human health. These substances were identified as priorities for assessment as they met categorization criteria or were considered a priority through other mechanisms (ECCC, HC [modified 2017]). The Chemical Abstracts Service Registry Numbers (CAS RNs) and Domestic Substances List (DSL) names of the 27 LBPNs examined in this assessment are listed in Table A.1 of Appendix A; descriptions of each CAS RN as reported on the DSL or cited in the published literature are provided in ECCC (2022).

The human health assessment focuses on 4 subgroups of LBPNs: C₉–C₁₄ hydrocarbon solvents (subgroup 1, 9 substances), C₉ aromatic hydrocarbon solvents (subgroup 2, 1 substance), the C₆–C₉ aliphatic hydrocarbon solvents (subgroup 3, 7 substances), and the LBPNs with no identified uses in products available to consumers (subgroup 4, 10 substances). The 27 individual LBPN substances are identified in Appendix A. The primary basis for the subgroups was the differences in health effects, presence in products available to consumers, hydrocarbon range, and aromatic content. Subgroup 1 substances consist primarily of hydrocarbons with carbon ranges from C₉ to C₁₄, which contain variable ranges of aromatic content where some are <2% and others range from 2% to 25%. Benzene and sulphur contents in this group are low, with benzene levels typically being <3 ppm. These substances are found in products available to consumers. Subgroup 2 is represented by a single substance, which contains an aromatic content greater than 90%. These substances are found in products available to consumers. Subgroup 3 substances are the products of refining processes. These substances consist primarily of hydrocarbons with carbon ranges from C₆ to C₉, which typically contain less than 1% total aromatics (toluene and xylene). These substances are found in products available to consumers. The health effects of the subgroup 4 substances are classified according to those of the former 3 groups, which are similar. These substances are not found in products available to consumers and are used in the petroleum and non-petroleum chemical industry. Within each subgroup, the CAS RNs are assumed to be interchangeable with respect to their use in categories of products available to consumers.

Petroleum industry site-restricted and industry-restricted LBPNs have previously been assessed (EC, HC 2011, 2013). These assessments evaluated exposures of the general population in Canada to LBPNs with uses restricted to the petroleum industry (refineries and upgraders). These petroleum industry uses include on-site production and use of these substances, or their transportation from production facilities to other petroleum facilities where they are used. It was determined that the site-restricted and industry-restricted LBPNs do not pose a risk to the environment or human health. In addition to petroleum industries, the 27 LBPN substances in this assessment are used in other chemical and consumer industries not related to refineries and upgraders. Air and water releases from non-petroleum industrial uses of these substances are considered in this assessment as the nature and amount of these releases may differ from those of the petroleum industry uses previously assessed.    

Of the 27 substances in this assessment, the 17 substances in subgroups 1, 2, and 3 may be used in food packaging materials, incidental additives, and/or products available to consumers and have potential non-petroleum industry releases where they are widely used as solvents. The remaining 10 substances from subgroup 4 have not been identified as being used in food packaging materials, incidental additives, or products available to consumers in Canada and only have relatively limited petroleum and other chemical industry uses. On the basis of the differences in exposure to the general population, the subgroup 4 substances are therefore assessed as a separate subgroup in this assessment.

This draft assessment includes consideration of information on chemical properties, environmental fate, hazards, uses and exposures, including additional information submitted by stakeholders. Relevant data were identified and targeted literature searches were conducted up to September 2020. Empirical data from key studies as well as results from models were used to reach proposed conclusions. When available and relevant, information presented in assessments from other jurisdictions was considered.

This draft assessment was prepared by staff in the CEPA Risk Assessment Program at Health Canada and Environment and Climate Change Canada and incorporates input from other programs within these departments. The ecological and human health portions of this assessment have undergone external review and/or consultation. Comments on the technical portions relevant to the environment were received from Geoff Granville (GCGranville Consulting Corp) and Dr. Connie Gaudet. Comments on the technical portions relevant to human health were received from Theresa Lopez, Jennifer Flippin, and Joan Garey (Tetra Tech). While external comments were taken into consideration, the final content and outcome of the assessment remain the responsibility of Health Canada and Environment and Climate Change Canada.

This draft assessment focuses on information critical to determining whether substances meet the criteria as set out in section 64 of CEPA by considering scientific information, including information, if available, on subpopulations who may have greater susceptibility or greater exposure, vulnerable environments and cumulative effects^{Footnote 2}, and by incorporating a weight of evidence approach and precaution.^{Footnote 3} This draft  assessment presents the critical information and considerations on which the proposed conclusions are based.

2. Identity of substances

LBPNs are generally complex, liquid combinations of hydrocarbons containing volatile components and are produced by refining or upgrading crude oil or bitumen, or are produced in natural gas processing plants. These petroleum fractions are commonly used as gasoline blending streams and solvents as well as other industrial / consumer products. They are mixtures of varying properties, with a typical boiling point range of -20°C to 230°C, and predominantly fall within the C₄ to C₁₄ carbon range (CONCAWE 2005). LBPNs are the lightest and most volatile fraction of the liquid hydrocarbons in petroleum and are composed of normal and branched alkanes (paraffins), isoalkanes, cycloalkanes (naphthenes), aromatics and, if subject to a cracking process, alkenes (olefins). Some refining processes create LBPNs that predominantly contain 1 or 2 of these chemical classes, while other refinery processes do not significantly influence the chemical composition of the LBPN (API 2008). For example, LBPNs derived from catalytic reforming typically contain high concentrations of aromatics, while those derived from alkylation typically contain no aromatics. Processes such as primary distillation and desulfurization (sweetening) do not significantly influence LBPN composition (API 2008). 4 CAS RNs (64742-82-1, 68553-14-0, 68603-08-7, and 70693-06-0) have boiling ranges above the 230°C LBPN upper limit. The boiling point ranges for 3 CAS RNs included in this assessment (68333-81-3, 68647-60-9 and 68920-06-9) are unknown; however, all have carbon ranges that overlap with those of the LBPNs.

The boiling point and carbon ranges given for LBPNs are useful for defining aspects related to their properties and fate as well as the nature of their primary components. However, these ranges are not definitive values, and there may be overlap with other petroleum classes. For example, 2 CAS RNs (64741-98-6 and 64742-88-7) included in this assessment of LBPNs are classified as kerosenes by the European industry organization, CONservation of Clean Air and Water in Europe (CONCAWE). Substances in the kerosene group have boiling point ranges of around 90 to 290°C and contain components with carbons predominantly in the C₇ to C₁₆ range (CONCAWE 2001a), ranges which overlap those of the LBPNs. This highlights the difficulty encountered in classifying petroleum substances, which contain many components, into discrete categories. All CAS RNs examined in this assessment are considered to have compositions and characteristics that make them suitable for inclusion in the LBPNs Group. 

As these substances are comprised of various combinations of hydrocarbons, they are considered to be Unknown or Variable composition, Complex reaction products, or Biological material (UVCB). These materials are derived from natural sources or complex reactions. A UVCB is not an intentional mixture of discrete substances and is considered a single substance. The complexity and variability of their compositions can make them difficult to fully and consistently characterize. Furthermore, variable initial feedstock composition, changes in refining processes, product specifications, and customer requirements across jurisdictions further complicate this characterization. The major components and approximate aromatic to aliphatic ratios of LBPNs examined in this assessment for risks to the environment are summarized in ECCC (2022).

LBPNs that are primary products of petroleum refining may be processed further into hydrocarbon solvents (see section 4, Sources and uses) and these have compositional characteristics that differ from the original refinery product. Solvents are increasingly being marketed as “narrow cut” products, and their boiling point ranges and the carbon range of their primary components are more restricted than for the corresponding refinery stream substance (CONCAWE 2001a).

For the human health risk assessment, the 27 substances in the LBPNs Group are separated into 4 subgroups, 3 of which are LBPN substances with consumer use and are differentiated on the basis of differences in health effects and aromatic content. The fourth subgroup includes LBPN substances not used in products available to consumers. These 4 subgroups are the C₉–C₁₄ hydrocarbon solvents (subgroup 1), C₉ aromatic hydrocarbon solvents (subgroup 2), the C₆–C₉ aliphatic hydrocarbon solvents (subgroup 3), and the LBPNs with no identified uses in products available to consumers (subgroup 4). Preliminary assessment of the health effects data associated with the LBPNs group revealed differences in the toxicity profiles in the 4 subgroups on the basis of hydrocarbon chain length and aromatic content (see, hazard sections of the Human Health Section, section 8). Subgroup 1, the C₉–C₁₄ hydrocarbon solvents, are predominantly aliphatic in content but contain variable ranges of aromatic content where some are <2% and others range between 2% and 25%. Subgroup 2 contains a single unique substance (that is, CAS RN 64742-95-6) which contains an aromatic content of greater than 90%. Subgroup 3, the C₆–C₉ aliphatic hydrocarbon solvents, are characterized by a total aromatic content of less than 1%. Subgroup 4 consists of substances that are chemically similar to the previous 3 subgroups but only have industrial uses and no identified uses in products available to consumers. The nature of the exposure of the general population to the subgroup 4 substances is different from the other 3 subgroups.

The CAS RN, DSL names, and compositional information for the individual substances in the 4 subgroups are presented in Table 2‑1 to Table 2‑4, respectively. A summary of known compositional information for substances of the LBPN subgroups 2 and 3 are presented in Appendix C.

Benzene and sulphur contents in the subgroup 1 C₉–C₁₄ hydrocarbon solvents group listed in Table 2‑1 are low with benzene levels typically being <3 ppm (OECD 2012c). Total aromatics in this group may be as high as 25%.

Abbreviations: aliph, aliphatic; NA, not available; CAS RN, Chemical Abstracts Service Registry Number; DSL, Domestic Substances List

Registration, Evaluation, Authorization and Restriction of Chemicals (REACH) has introduced requirements for renaming solvents on the basis of their degree of refinement. Within this scheme, hydrocarbon solvent substances currently assigned the same CAS RN will be given different names on the basis of whether they are petroleum raw materials or more highly processed hydrocarbon solvents. The Hydrocarbon Solvents Producers Association (HSPA) has introduced a number of guidance documents related to standardizing the nomenclature of the hydrocarbon substances (including “white spirits”), which may impact the future naming of some of the subgroup 1 substances (ESIG 2011; REACH 2017)

The substance identity of subgroup 2, the C₉ aromatic solvents, is given in Table 2‑2. This UVCB substance is primarily composed of bi- or tri-substituted benzenes such as ethyltoluene or trimethylbenzene.

Abbreviations: arom, aromatic; CAS RN: Chemical Abstracts Service Registry Number; DSL: Domestic Substances List

The substances in subgroup 3, C₆–C₉ aliphatic hydrocarbon solvents, listed in Table 2‑3, are differentiated from other refinery substances such as gasoline and diesel fuel by additional processing steps that result in finished substances with a low odour, narrow boiling range, removal of sulphur- and nitrogen-containing compounds, aromatic content control and low colour (OECD 2010a). These additional refining steps are necessary in order to make substances that have qualities suitable for consumer applications. These substances typically contain less than 1% total aromatics (toluene and xylene). Benzene is intentionally removed to obtain levels of less than 0.01%, and sulphur and nitrogen compounds are removed by the refining process.

The carbon range distribution of subgroup 3, C₆–C₉ aliphatic solvents, is given in Table C-2 of Appendix C. For the substances with reported carbon ranges, C₆ hydrocarbons are reported at a maximum of 8%. This 8% can be comprised of up to 5 C₆ saturated straight and branched alkanes, 2 possible cyclic C₆ aliphatic substances, and a number of unsaturated hexenes. The percentage range of n-hexane in the C₆–C₉ aliphatic solvents, specifically, is unknown.

Abbreviations: aliph, aliphatic; CAS RN: Chemical Abstracts Service Registry Number; DSL: Domestic Substances List

The subgroup 4 LBPNs are listed in Table 2‑4. On the basis of their chemical composition and health hazard, these substances are categorized according to the 3 subgroups 1, 2, and 3 for the purposes of human health hazard evaluation.

Abbreviations: aliph, aliphatic; arom., aromatic; CAS RN, Chemical Abstracts Service Registry Number; DSL, Domestic Substances List

3. Physical and chemical properties

The composition and physical-chemical properties of LBPNs vary depending on the sources of crude oil, bitumen, and natural gas and the processing steps involved. Chemical property data for some of the LBPNs examined in this assessment are summarized in Table B-1 of Appendix B. All LBPNs listed in Table B-1 are liquids at ambient temperature (15°C to 25°C), with melting points in the range of -127°C to 112°C and density values below that of water and ranging from 0.66 g/mL to 0.89 g/mL at 15°C to 25°C. Vapour pressure varies widely between the LBPNs, with some LBPNs having values in the range of 0.3 Pa to 830 Pa at 20°C to 25°C and others exhibiting substantially higher values of 1300 Pa to 5500 Pa at 20°C. Considered together, the data indicate moderate to high volatility. LBPNs have a typical boiling point range of −20°C to 230°C (CONCAWE 2005). See Tables 3-1 to 3-3 for a more detailed analysis of composition ranges by hydrocarbon subgrouping.  

Water solubilities range from very low for the longest-chain alkanes to high for the simplest mono-aromatic components, in the range of 0.01 mg/L to 94.3 mg/L at 25°C, indicating low to moderate solubility in water (Table B-1 of Appendix B). Octanol-water partition coefficients (log K_ow) measured for the LBPNs range from 3.0 to 7.2, with most values at above 3.5, indicating moderate to high hydrophobicity. In general, the aromatic compounds are more soluble than the same-sized alkanes, isoalkanes and cycloalkanes. This indicates that the components likely to remain in water are the one- and two-ring aromatics (C₆–C₁₂). The C₉–C₁₂ alkanes, isoalkanes, and one- and two-ring cycloalkanes are likely to be attracted to sediments on the basis of their low water solubilities and moderate to high log K_ow and log organic carbon-water partition coefficient (K_oc) values.

To predict the physical/chemical properties and ecological fate of complex petroleum substances such as LBPNs, representative structures were chosen from each chemical class found in the substance. As the composition of LBPNs is variable and not well defined, representative structures for alkanes, isoalkanes, n-alkenes, one- and two-ring cycloalkanes, cycloalkane monoaromatics, and one- and two-ring aromatics ranging from C₄ to C₁₂ were selected on the basis of carbon numbers and boiling points for each hydrocarbon class. Physical and chemical data were assembled from scientific literature and from the EPI Suite (c2000-2012) group of environmental models. A summary of empirical and modelled physical and chemical property data for the representative hydrocarbon structures of LBPNs is provided in ECCC (2022). 

In general, lower carbon (C₄ to C₆) aliphatic LBPN representative structures are highly volatile (vapour pressure 1.3×10⁴ Pa to 3.5×10⁵ Pa at 25°C; Henry’s law constant 1.5×10⁴ Pa to 1.8×10⁵ Pa·m³/mol at 25°C), with low to moderate water solubility (9.5 mg/L to 221 mg/L at 25°C) and low to moderate partition coefficients (log K_ow 2.4 to 3.9 and log K_oc 2.1 to 3.4). As carbon number in the aliphatic structures increases to C₉ and C₁₂, vapour pressure and volatility remain high (15 Pa to 1440 Pa at 25°C; Henry’s law constant 9.2×10³ Pa to 2.0×10⁶ Pa·m³/mol at 25°C), while water solubility decreases (3.7×10^-3 mg/L to 25 mg/L at 25°C) and partition coefficients (that is, sorption potential) increase (log K_ow 3.7 to 6.1 and log K_oc 3.2 to 5.3).

Similar trends in volatility, water solubility, and sorption potential are seen with aromatic representative structures. While vapour pressure and volatility are high for all LBPN representative substances, vapour pressure and Henry’s law constant values for the aromatic compounds are consistently lower than for their aliphatic counterparts of the same carbon number, while water solubility values tend to be higher. For example, vapour pressure and Henry’s law constant values for the C₆ aliphatic substance, hexane (CAS RN 110-54-3), are 2.0×10⁴ Pa and 1.8×10⁵ Pa·m³/mol, respectively, while for comparison, those for the C₆ aromatic, benzene (CAS RN 71-43-2), are 1.3×10⁴ Pa and 562 Pa·m³/mol, respectively. The water solubility of hexane is 9.5 mg/L at 25°C, while for comparison, the solubility of the aromatic benzene is 1790 mg/L. Partition coefficients are generally lower for aromatics when compared with aliphatic structures of equivalent carbon number (that is, log K_ow of hexane is 3.9 while that of benzene is 2.1), although trends of increasing sorption potential with increased carbon number remain consistent for both groups (ECCC 2022).

It should be noted that the physical and chemical behaviour of the representative structures may differ if these representative structures are present in a complex substance, such as LBPNs. The vapour pressures of components in a mixture will be lower than their individual vapour pressures according to Raoult’s Law (that is, the total vapour pressure of an ideal mixture is proportional to the sum of the vapour pressures of the mole fractions of each individual component). Similarly to Raoult’s Law, the water solubilities of components in a mixture are lower than when they are present individually (Banerjee 1984; Di Toro et al. 2007). Concurrently, however, when an individual petroleum hydrocarbon chemical that is normally solid under environmental conditions is part of a petroleum mixture (or UVCB), it can be found in a liquid state due to the lowering of its melting point when it is in a mixture (Di Toro et al. 2007). This results in an increase in the vapour pressure and water solubility of the hydrocarbon that is normally solid, as determined by the subcooled vapour pressure (Staikova et al. 2005) and subcooled solubility (Di Toro et al. 2007). 

In general, the C₉–C₁₄ hydrocarbon LBPNs in subgroup 1 have low to moderate water solubilities, moderate to high Log K_ow (log octanol–water partition coefficient), moderate vapour pressure, and a very high Henry’s Law constant; see Table 3‑1.

Abbreviations: (E), Experimental, (M), Modelled, UP, unspecified whether data is experimental or modelled.
^a References: OECD 2010a; OECD 2012c

The C₆–C₉ aliphatic and C₉ aromatic LBPNs in subgroups 2 and 3 have low to moderate water solubilities, moderate Log K_ow (log octanol–water partition coefficient), and very high vapour pressure and Henry’s Law constant; see Table 3‑2.

Abbreviations: (E), Experimental; (M), Modelling; (E)*, Experimental but with restrictions because there is limited information on test procedure; (A), value derived from an analogue(s)
^a References OECD 2011; McKee et al. 2015
^b References OECD 2013a. Phys/chem data is an average for 4 CAS RNs: 64741-84-0, 64742-49-0, 64742-89-8, and 68410-97-9. All 4 substances are liquids at 25°C.
^c References OECD 2010a. Phys/chem data is an average for 3 CAS RNs: 64741-84-0, 64742-49-0, and 64742-89-8. All 3 substances are liquids at 25°C.
^d References OECD 2010a, 2010g, 2013b, 2013c, and 2013d; NICNAS 2019

Subgroup 4 (the no consumer use identified LBPNs) consists of complex mixtures with components that fall predominantly into the C₄–C₁₂ carbon range: alkanes, cycloalkanes, aromatics and, if they are subject to a cracking process, alkenes as well (CONCAWE 2005). Depending on the specific refining and distillation processes involved in their preparation, the chemical composition of these CAS RNs may have a limited carbon range.

Abbreviations: NA, not available
^a ECHA 2019a
^b US EPA 2011b
^c ECHA 2019b
^d US EPA 2009
^e ECHA 2022b
^f ECHA 2022c
^g ECHA 2022a

Experimental and modelled vapour pressures of representative structures for the subgroup 4 substances (no identified uses in products available to consumers) are moderate to very high and decrease with increasing molecular size; see Table 3‑3. This suggests that losses from soil and water will likely be high and that air will be the ultimate receiving environment for most of the components of LBPNs.

4. Sources and uses

Summaries of the manufacture, import, use, and export of these LBPN substances in Canada were prepared according to information reported in response to various CEPA section 71 surveys (Environment Canada 2008, 2009, 2011a, 2012a, 2015) as well as a voluntary data gathering initiative in 2015 (Table 4‑1).

Abbreviations: NA, not available

In total, 25 of the 27 substances were surveyed to obtain information on the manufacture, transport, and uses of LBPNs in Canada. On the basis of initial triaging, 3 substances (CAS RNs 68553-14-0, 68603-08-7 and 68603-08-7) in subgroup 4 were not included in any surveys issued pursuant to a CEPA section 71 notice. The CEPA section 71 surveys for the reporting year 2006 applied only to petroleum refining or upgrading facilities (Environment Canada 2008, 2009) and requested information on the quantities of substances used and their industrial fate (Environment Canada 2008, 2011b), as well as information on the transportation of the substances from refineries to other commercial or industrial facilities (Environment Canada 2009, 2011c). The 2012 and 2015 surveys, however, were applicable to all sectors that met the reporting requirements and they collected information on the manufacture, import, and use of the petroleum LBPNs regarding their use in products available to consumers and their potential for general population exposure. The 2015 survey gathered qualitative information about the import and blending/formulation activities of remaining priority petroleum substances.

Quantity data were collected for the 13 CAS RNs surveyed in the 2010 calendar year (Environment Canada 2011a, 2012b) and for 1 CAS RN surveyed in 2011 (Environment Canada 2012a, 2013), whereas subsequent surveys and data gathering initiatives were used to collect information on uses and did not collect quantity data. The quantity data available for the initial 14 LBPNs surveyed were considered sufficient for the assessment as they provided an estimated order of magnitude of the quantities of LBPNs that could be used within a given application.

According to information submitted in response to CEPA section 71 surveys (Environment Canada 2011a, 2012a, 2012b, 2013), the total quantity surveyed for manufacture and import in Canada was greater than 5 billion kg for 13 CAS RNs in 2010 and greater than 100 million kg for 1 CAS RN in 2011.

According to OECD classification (OECD 2004) of the substances in this assessment, 23 substances (CAS RNs 64741-49-1, 64741-65-7, 64742-48-9, 64742-82-1, 64742-88-7, 8030-30-6, 8032-32-4, 8052-41-3, 64742-95-6, 64741-66-8, 64741-84-0, 64742-49-0, 64742-89-8, 68410-97-9, 68647-60-9, 647421-68-0, 64742-92-0, 64741-98-6, 68513-03-1, 68553-14-0, 68603-08-7, 68920-06-9, and 70693-06-0) are on the 2004 list of High Production Volume (HPV) Chemicals, which are produced globally at levels greater than 1000 tonnes per year.

LBPNs are used in a variety of industrial and commercial processes and products. According to information submitted in response to CEPA section 71 surveys and a voluntary data gathering initiative (Environment Canada 2012a, 2013, 2015; ECCC 2016a,b), in Canada, LBPNs are mainly used as solvents that become part of the formulation or mixture to which they are added, as viscosity adjustors, and as fuels or fuel additives. Products containing LBPNs are used as processing aids in petroleum production and other industrial processes, as lubricants and lubricant additives, in paint and coating additives, as solvents for cleaning and degreasing, in adhesives and sealants, as corrosion inhibitors and anti-scaling agents, and as demulsifiers (that is, specialty chemicals used to separate emulsions) in paraffin control and dewaxing. LBPNs may also be used in automotive care products, plastics and rubber, printing inks, pulp and paper processing aids, household cleaners, personal care products, and as formulants in pest control products. Some LBPNs have specialty applications in products such as in anti-adhesive agents, plating and surface treatment agents, and some pigments.

The majority of the substances in the LBPNs Group are listed on the EU Cosmetics Regulation 1223/2009 Annex II—List of substances prohibited in cosmetic products (Galleria Chemica). In addition, the majority of the substances are restricted under Annex XVII of the REACH Regulations (Galleria Chemica). These substances cannot be used in products and preparations placed on the market for sale to the general public in individual concentrations ³ 0.1 % (European Parliament and Council 1999; 2006; 2008). The use of these LBPN substances in products available to consumers is subject to some restrictions in Australia (IMAP 2016).

Table 4‑2 below presents a summary of the information collected from responses to CEPA section 71 surveys and a voluntary data gathering initiative in relation to the major consumer uses of 25 of 27 of the LBPN substances. These include uses of these LBPN substances or products containing these LBPN substances for which reporting companies have identified as being intended for sale to the general public or intended for use in consumer activities (Environment Canada 2011a, 2012a, 2012b, 2013, 2015; ECCC 2016a,b).

Abbreviations: CBI, confidential business information; NR, none reported above the reporting threshold of 100kg.
^a This value describes the total quantity, reported in response to surveys conducted under section 71 of CEPA,  of the substance imported, manufactured, or used in products that are intended for sale to the general public (Environment Canada 2012a, 2013, 2015). See survey for specific inclusions and exclusions (schedules 2 and 3).
^b Non-confidential uses reported in response to the surveys conducted under section 71 of CEPA, for products intended for sale to the general public (Environment Canada 2012b), other uses were reported for some substances but were considered confidential business information. See survey for specific inclusions and exclusions (schedules 2 and 3).
^c Quantities and uses presented here reflect those from companies that reported the substance as being used in mixtures, products or manufactured items intended for use in consumer activities. It is unclear whether products associated with these uses are intended for sale to the general public.
^d Exact quantities or ranges of quantities of LBPN substances were not surveyed in the 2014 reporting year (Environment Canada 2015). The quantity above pertains only to customers in Canada to whom the notifier sold more than 100 kg of the CAS RN.

According to information from Health Canada’s Pharmaceutical Drugs Directorate (PDD) and Natural and Non-prescription Health Products Directorate (NNHPD), none of the substances in the LBPNs Group were reported to be used as medicinal or non-medicinal ingredients in drugs including in natural health products in Canada (personal communication, email from Health Canada (HC) PDD and NNHPD to the Existing Substances Risk Assessment Bureau (ESRAB) of HC, February 12, 2018; unreferenced).

The substances in the LBPNs Group were not reported to be used as active ingredients in pest control products in Canada. However, CAS RNs 8052-41-3, 64741-65-1, 64741-66-8, 64742-48-9, 64742-82-1, 64742-88-7, 64742-95-6 and 70693-06-0 are present as formulants in currently registered pest control products in Canada (personal communication, email from the Pest Management Regulatory Agency to the ESRAB of Health Canada, September 2017 and February 12, 2018; unreferenced).

The substances in the LBPNs Group are not permitted food additives in Canada. However, some of the LBPN substances (CAS RNs 64741-65-7, 64742-48-9, 64742-82-1, 64742-88-7, 8032-32-4, 8052-41-3, 64742-95-6, 64741-66-8, 64742-49-0, 64742-89-8, and 426260-76-6) may be used as components in the manufacture of food packaging materials (FPM). Specifically, they may be used as components in the formulation of inks, coatings and adhesives whose applications have no direct contact with food. They may also be used as components in the manufacture of coatings and plastics- and paper/paperboard-based food packaging materials, which are in direct contact with food. They may also be used as components of incidental additives^{Footnote 4} (for example, lubricants and/or cleaners) used in food processing establishments with negligible or no food contact (personal communication, emails from the HC Food Directorate to the HC ESRAB, August 2017, November 2017, February 2018, and July 2019; unreferenced).  

Eight LBPN substances in this assessment from subgroups 1, 2, and 3 (CAS RNs 64475-85-0, 64741-66-8, 64742-48-9, 64742-49-0, 64742-95-6, 8030-30-6, 8032-32-4, and 8052-41-3) were identified as ingredients in a number of cosmetics in Canada. Types of products that were reported to contain these substances include nail polishes, nail adhesives, waterproof mascaras, non-permanent eye makeup, hair adhesive removers, depilatory lotions, hair wax removers, hair colours, and cleansers (personal communication, email from the HC Consumer and Hazardous Products Safety Directorate (CHPSD) to the HC ESRAB, December 10, 2019 to May 14, 2020; unreferenced).

Through searches of publicly available databases, online literature, and retailer websites, all of the substances in subgroups 1, 2, and 3 in this assessment were identified as being used as ingredients in many products that are available to consumers in Canada. The majority of products identified were considered to be do-it-yourself (DIY) products. Tables 4‑3 to 4‑5 summarize the types of products available to consumers that were identified to contain the substances in the LBPN subgroups 1, 2, and 3.

No products available to consumers containing any of the 10 substances in subgroup 4. Industrial uses for subgroup 4 were identified. Industrial uses for subgroup 4 are presented in Table 4‑6.

Abbreviations: NA, not available

5. Releases to the environment

LBPNs may be released into the environment from industrial activities associated with their production, processing, transportation, storage, and disposal, as well as during industrial, commercial and consumer uses of products that contain them.

Industry information indicates that LBPNs in Canada are used mainly in industrial processes where they are consumed on the premises of the facilities or are blended into other products or formulations (Environment Canada 2012a, 2013, 2015; ECCC 2016a,b). Releases to the environment from these applications can be estimated by developing an appropriate exposure scenario for each application type.

LBPNs used in commercial and consumer products have the potential to be released into the environment. The high volatility of LBPNs suggests that environmental releases will be primarily to air. Releases to water and land may occur via sources such as wastewater treatment systems (WWTS^{Footnote 5}) and unintentional releases such as spills. No direct releases to soil are anticipated from WWTS; however, indirect releases may result from the application of biosolids to land from WWTS receiving wastewater containing LBPN component substances.

LBPNs used as solvents commonly require rapid evaporation as part of their intended function, for example, in the proper curing of paints and coatings as well as in printing inks, adhesives, and sealants (Cheminfo Services Inc. 2008). The evaporation of solvent LBPNs results in emissions to air unless volatile components are captured, recycled, and/or destroyed in control devices such as thermal oxidizers (Cheminfo Services Inc. 2008). In Canada, only a small proportion of evaporated LBPN solvents are captured, while the great majority are emitted to air (Cheminfo Services Inc. 2008).

9 LBPN CAS RNs examined in this assessment (CAS RNs 8030-30-6, 8032-32-4, 8052-41-3, 64475-85-0, 64741-65-7, 64742-48-9, 64742-88-7, 64742-89-8 and 64742-95-6) appear on the National Pollutant Release Inventory (NPRI), an inventory that tracks releases of targeted contaminants to air, water, and land from industrial facilities in Canada. All NPRI-listed LBPN CAS RNs are classified as Part 5 speciated Volatile Organic Compounds (VOCs), with a requirement for detailed reporting above the reporting threshold of ≥1 tonne of air releases for VOCs (ECCC 2017).

According to information submitted in response to a CEPA section 71 survey for the 2010 reporting year (Environment Canada 2011a, 2012b), sources of releases of substances in the LBPNs Group may include air emissions and releases to water from facilities. In addition to data reported in response to section 71 of CEPA, CAS RNs 64742-89-8 and 64742-95-6 were reported under the NPRI for facilities meeting the reporting threshold of these substances. Table 5‑1 below summarizes the range of quantities reported for the substances in each subgroup on the basis of information submitted in response to a CEPA section 71 survey and/or NPRI, per non-petroleum chemical facility. For all substances captured in the table below, the major release was to air.

^a According to information reported in response to a CEPA section 71 survey (Environment Canada 2011b)
^b According to information reported in response to a CEPA section 71 survey (Environment Canada 2012b)
^c NPRI 2017

As VOCs, naphtha solvents can contribute to ground-level ozone formation, and their emissions are managed in Canada under various regulations and voluntary instruments (ChemInfo Services Inc. 2008). These include ECCC and Canadian Council of Ministers of the Environment (CCME) initiatives to reduce the VOC content of architectural and industrial maintenance coatings, automobile refinish coatings, and consumer and commercial products such as cleaners, degreasers, adhesives, and personal care products, as well as industry agreements and codes of practice for printing and automotive assembly (Environment Canada 2011a, 2012a, 2013, 2015; ECCC 2016a). In addition, air permitting requirements and bylaws designed to reduce emissions from all VOCs, including naphtha solvents, are available for a number of provincial and municipal jurisdictions (ChemInfo Services Inc. 2008).

Unintentional releases of LBPNs may occur through spills to the environment during LBPN storage, transport, or use. Releases of LBPNs from spills were evaluated through the analysis of spills information reported to Environment and Climate Change Canada for Alberta and Ontario between 2014 and 2018 (personal communication, email from the Alberta Ministry of Environment and Parks to the Science and Technology Branch of ECCC, dated 2019; unreferenced; personal communication, email from the Ontario Ministry of the Environment, Conservation and Parks to the Science and Technology Branch of ECCC, dated January 2019; unreferenced). Publicly available spill data from Quebec (Government of Quebec 2019), Manitoba (Government of Manitoba 2019), Saskatchewan (Government of Saskatchewan 2019), Alberta (AER 2019), British Columbia (Government of British Columbia 2019) and the Northwest Territories (Government of Northwest Territories 2019) were examined in the analysis. Spills of petroleum substances under the following descriptions were considered: “naphtha”; “solvent”; “light oil”; “light petroleum”; and “petroleum distillate”. Over the five-year period from 2014 to 2018, 159 spill incidents with an estimated total volume of 250 000 L were reported under the general descriptions indicated above. All reported incidents occurred in Alberta (100 incidents) and Ontario (59 incidents), with no spills reported in the remaining provinces and territory. These spills are not specific to the LBPNs examined in this assessment, and they provide a conservative estimate of releases.

6. Environmental fate and behaviour

6.1 Environmental distribution

When petroleum substances are released into the environment, the major fate processes include dissolution in water, volatilization, adsorption, biodegradation, and photodegradation. These processes will cause changes in the composition of these substances.

Biodegradation almost always occurs when petroleum mixtures are released into the environment. Studies have found populations of bacteria and other organisms (for example, fungi and yeasts) that are capable of degrading petroleum hydrocarbons in both fresh and marine waters and sediments, as well as soils (Atlas 1981). Degradation occurs in both in the presence and absence of oxygen. In general, aromatic components tend to be degraded more slowly than aliphatic components, although the degradation of some higher molecular weight cycloalkanes may be very slow (Atlas 1981; Potter and Simmons 1998).

3 weathering processes—dissolution in water, volatilization, and biodegradation—typically result in the depletion of the more readily soluble, volatile, and degradable compounds and the accumulation of those most resistant to these processes in residues. The physical properties of moderate to very high volatility and very low to moderate water solubility (Table B-1, Appendix B) suggest that LBPNs will tend to distribute primarily into air when first released into the environment. However, due to the complex interaction of components within a mixture, which impacts their physical-chemical properties and behaviour, it is difficult to predict the fate of a complex mixture such as LBPNs. Therefore, as a general indication of the fate of LBPNs, the physical-chemical properties of representative structures of LBPNs (ECCC 2022) were examined.

If released into air, most components of LBPNs are expected to remain in this compartment. An exception is the two-ring aromatics, where a small proportion may distribute out of the air and into water and soil. Many LBPN components (C₄–C₆) are extremely volatile, with vapour pressures exceeding 10 000 Pa, and most other components are highly volatile with vapour pressures above 100 Pa (ECCC 2022). The larger (C₁₂) alkanes, cycloalkanes and one- and two-ring aromatics are moderately volatile, with vapour pressures ranging from about 1 Pa to 20 Pa at 25°C. 

If released into water, most components of LBPNs are expected to partition to air because of their high volatility. LBPNs are less dense than water (Table B-1, Appendix B), suggesting that when they are released into this medium, they will tend to rise to the surface of the water and spread out as a slick. This increases the exposure surface, allowing the more volatile components, which comprise the majority of LBPNs, a greater opportunity to volatize. The remaining components can then be expected to occur in the dissolved phase or to be adsorbed to suspended particulates, depending on their individual water solubilities and sorption potential (as expressed by partition coefficients).

A slick is more likely to be present when release of the LBPN occurs at or close to the water surface, which increases the likelihood of volatilization of the LBPN components. If the release occurs deep in the water column, solubility of the component substances is expected to predominate. Many representative components of the LBPNs are sufficiently soluble to remain within the water column. This includes the C₄ alkanes, isoalkanes, and alkenes; the C₆ one-ring cycloalkanes and one-ring aromatics; the C₉ two-ring cycloalkanes and one-ring aromatics; and the C₁₀ cycloalkane monoaromatics and two-ring aromatics. All of these components are moderately to highly soluble in water, with solubility values of the pure substances ranging from 25 mg/L to 1790 mg/L at 25°C (ECCC 2022). On the basis of low or very low water solubility ranging from 13 mg/L to less than 1 mg/L at 25°C, C₆–C₁₂ alkanes and isoalkanes, C₉–C₁₂ alkenes and one-ring cycloalkanes, and C₁₂ one- and two-ring aromatic representative components are not expected to remain in the water column and are expected to partition to particulates suspended in the water column and/or sediment, though the tendency to sorb is greater for the aliphatic structures than for the aromatic ones.

If released to soil, highly volatile components of LBPNs such as the C₄–C₆ alkanes and isoalkanes, can be expected to partition to air (volatilize), while other less volatile components will remain in the soil. Within the soil, the movement of components will be determined by their partitioning between soil particulates, soil pore water, and soil pore air. Larger, more hydrophobic LBPN components such as the C₁₂ aliphatics and aromatics with K_oc values ranging from 4.4 to 5.3 (see Table 2.1, ECCC 2022) are expected to sorb to soil particulates and thus remain relatively immobile. More soluble components with lower sorption potential, such as the C₆ aromatics, may become dissolved in soil pore water or soil pore air, facilitating their movement away from the site of release.

When large quantities of a hydrocarbon mixture enter the soil compartment, soil organic matter and other sorption sites in soil become saturated, and the hydrocarbons will begin to form a separate phase (a non-aqueous phase liquid or NAPL) in the soil. At concentrations below the retention capacity for the hydrocarbon in the soil, the NAPL will be immobile (Arthurs et al. 1995); this is referred to as residual NAPL (Brost and DeVaull 2000). Above the retention capacity, the NAPL becomes mobile and will move within the soil (Arthurs et al. 1995; Brost and DeVaull 2000). LBPNs moving through soil have the potential to be transported into groundwater, where they can contaminate both the groundwater and any surface waters receiving inflow from the groundwater.

6.2 Environmental persistence

Due to the complex nature of LBPNs, their persistence was evaluated on the basis of empirical and/or modelled data for representative petroleum hydrocarbons expected to be similar to those released into the environment. These representative structures do not include all possible individual components present within an LBPN nor do they necessarily provide a complete picture of the full range of persistence potential for any given chemical class (for example, alkanes, one-ring aromatics) or carbon number (for example, C₁₂). Thus, the modelling results do not indicate the persistence of all substances in a specific class and carbon range but instead give a more general indication of these properties.

Some empirical biodegradation data are available for the LBPNs, and these are summarized in Table 3.1 of ECCC (2022). Heavy straight-run petroleum naphtha (CAS RN 64741-41-9) and hydrodesulfurized heavy petroleum naphtha (CAS RN 64742-82-1) were 77% and 75% degraded, respectively, within a 28-day testing period (OECD 2012c; API 2014); therefore, they both met readily biodegradable criteria according to OECD Guideline 301F (manometric respirometry test; OECD 1992). Using the AOPWIN estimation program in EPI Suite, light alkylate petroleum naphtha (CAS RN 64741-66-8) was shown to have an estimated atmospheric half-life of 1.07 to 15.99 days and was 42% degraded in water in a 28-day testing period (API 2014; ECCC 2022). It was considered inherently biodegradable on the basis of a CONCAWE test method for determining the inherent aerobic biodegradability of oil products (CONCAWE 1999). Hydrotreated heavy petroleum naphtha (CAS RN 64742-48-9) and hydrotreated light petroleum naphtha (CAS RN 64742-49-0) were also readily biodegradable on the basis of results from the OECD 301F manometric respirometry test, although the hydrotreated heavy petroleum naphtha product did not meet a 10-day biodegradation window (OECD 2010e).

The persistence of a suite of representative petroleum hydrocarbons for LBPNs was characterized on the basis of empirical and/or modelled data. Model results and the weighing of information are reported in the technical document on petroleum substance persistence and bioaccumulation (Environment Canada 2014) and results are summarized in ECCC (2022).

Reaction with photochemically-produced hydroxyl radicals is the main degradation pathway for LBPN representative structures in the atmosphere, although alkenes may also photodegrade through reaction with ozone and nitrate radicals (AOPWIN 2010). Empirical and modelled atmospheric half-lives for most representative structures of LBPNs are less than 2 days, indicating that most LBPN components are unlikely to persist in air (Environment Canada 2014). However, some C₄–C₆ alkanes, C₄–C₆ isoalkanes, and C₆ one-ring aromatic components have the potential to remain in the atmosphere for longer periods, allowing them to be potentially transported long distances and to remote regions (Environment Canada 2014).

On the basis of their chemical structure, components of the LBPNs are not expected to hydrolyze under environmental conditions (Lyman et al. 1990; Environment Canada 2014).

On the basis of modelled biodegradation results for representative structures in water, soil, and sediment, the following components of LBPNs are expected to have half-lives of greater than 6 months in water and soils and greater than one year in sediments: C₁₂ one-ring aromatics, C₁₀ cycloalkane monoaromatics, and C₁₀ two-ring aromatics (Environment Canada 2014). The C₉ one-ring aromatics have half-lives of greater than one year in sediments.

6.3 Potential for bioaccumulation

The bioaccumulation potential for a suite of representative petroleum hydrocarbons for LBPNs was characterized on the basis of empirical and/or modelled data. Bioaccumulation factors (BAFs) are the preferred metric for assessing the bioaccumulation potential of substances, as the bioconcentration factor (BCF) may not adequately account for the bioaccumulation potential of substances via the diet, which predominates for substances with a log K_ow of greater than approximately 4.5 (Arnot and Gobas 2003).

In addition to fish BCF and BAF data, bioaccumulation data for aquatic invertebrate species were also considered. Biota-sediment/soil accumulation factors (BSAFs), trophic magnification factors (TMFs), and biomagnification factors (BMFs) were also considered in characterizing bioaccumulation potential.

Empirical and modelled bioaccumulation data for petroleum hydrocarbons can be found in Environment Canada (2014), and results for the LBPNs are summarized in ECCC (2022).

Overall, there is consistent empirical and predicted evidence to suggest that the following components have the potential for high bioaccumulation, with BAF/BCF values close to or greater than 5000: C₉ n-alkanes, C₁₂ n-alkenes, and C₁₂ one-ring cycloalkanes. In addition, empirical BMF and TMF values of greater than 1 have been reported for C₁₀ two-ring aromatics, indicating that these substances have the potential to biomagnify in aquatic or terrestrial food webs. Most LBPN representative components are not expected to biomagnify relative to their concentration in the diet, largely because the combination of metabolism, growth dilution, and low dietary assimilation efficiency of these components allows the elimination rate to exceed the uptake rate when exposure occurs through the diet only (Environment Canada 2014).

7. Potential to cause ecological harm

7.1 Ecological effects assessment

7.1.1 Effects on aquatic organisms

Empirical aquatic toxicity data are available for 6 of the CAS RNs in this assessment, and these data were considered in evaluating the potential toxicity of LBPNs (ECCC 2022). However, the LBPNs within this assessment that have experimental data are primarily composed of aliphatic constituents and may not fully represent the potential toxicity of mainly aromatic LBPNs. Therefore, experimental toxicity data for gasoline, a primarily aromatic LBPN, are also considered. Empirical and modelled data are also available for representative structures of LBPNs, and these data are considered as lines of evidence in the evaluation of potential for toxicity.

Empirical aquatic toxicity data for the LBPNs are presented in ECCC (2022) and summarized in Table 7-1 below. Data are available for 3 species of fish, 3 invertebrate species and 1 algal species. Due to the low to moderate water solubility of LBPNs, testing was conducted using water-accommodated fractions (WAFs) of the LBPN. WAFs are laboratory-prepared aqueous media derived from low-energy mixing of a poorly soluble test material such as a petroleum product. WAFs are essentially free of particles of bulk material, containing only the fraction that is dissolved or present as a stable dispersion or emulsion (Singer et al. 2001).

Exposure concentrations are often not measured in WAF test solutions and results are reported in terms of the LBPN loading rate^{Footnote 6} rather than concentrations (for example, a median lethal loading [LL₅₀] rather than a median lethal concentration [LC₅₀]). This approach is common in testing with poorly water soluble UVCBs such as petroleum products. Where effects endpoints are presented in terms of measured concentrations, the identity of the analytes measured in the test solutions must also be reported. For example, endpoints that are based on measured concentrations are provided in API (2014) for aquatic toxicity testing conducted on LBPN CAS RN 64741-66-8 (naphtha, petroleum, light alkylate). Information provided with the toxicity testing results indicates that measured concentrations represented the sum of 7 hydrocarbon components (identities not specified) that were measured in the WAF solutions. These compounds do not represent all of the hydrocarbons present in the dissolved fraction of the test solutions, and test endpoints calculated from these measured values would be expected to be lower than if all dissolved components were included in the measurements (API 2014).

Abbreviations: EL₅₀, loading rate of test substance resulting in a specified effect (for example, immobilization, growth) in 50% of the test species exposed to the WAF; LL₅₀, median lethal loading rate, loading rate of the test substance that results in 50% mortality in a population of test organisms exposed to the WAF; NOELR, no observed effect loading rate; WAF, water-accommodated fraction, aqueous medium containing only the fraction of the petroleum substance that is dissolved or present as a stable dispersion or emulsion.

Median acute effect levels based on WAF loading rates (LL₅₀ or median effective loading rate, EL₅₀) ranged from 3.6 mg/L to 51 mg/L in fish and 1.4 mg/L to 43 mg/L in invertebrates (Table 7.1). Lowest chronic median effect (EL₅₀) values were 5.2, 1.2, and 2.5 mg/L for fish, daphnids, and algae, respectively, while a 5-day no-observed effect level on the basis of a substance loading rate (NOELR) of 2.1 mg/L was reported for domestic sewage sludge microorganisms. The results summarized in Table 7‑1 indicate that the WAFs of the LBPNs tested were not highly hazardous to aquatic organisms.

Toxicity values are comparable between species (that is, fish, aquatic invertebrates, and algae) and do not differ significantly for acute vs. chronic testing. However, as indicated, the empirical data were derived for predominantly aliphatic LBPNs and may not represent the hazard potential of LBPNs that are comprised mainly of aromatic compounds. For read-across purposes, empirical toxicity data for unleaded gasoline, as an aromatic LBPN, were also considered. The measured acute toxicity values for gasoline-blend WAFs in closed test systems were all within 2 orders of magnitude. In these studies, algae were the most sensitive species, with a 96-h median inhibition loading rate or IL₅₀ of 1.4 mg/L, followed by rainbow trout (Oncorhynchus mykiss) with a 96-h LL₅₀ of 11 mg/L, and Daphnia with a 48-h EL₅₀ of 12 mg/L (CONCAWE 1996a). In comparison, lower toxicity values (0.3, 1.2, and 3 mg/L for invertebrates, and 2.7 mg/L for fish) were obtained in flow-through closed test systems using a water-soluble fraction (WSF), as cited by CONCAWE (1992). These values are similar to those obtained for aliphatic LBPNs, as summarized in Table 7.1.

It should be noted that the toxicity data are for individual CAS RNs, and no information is provided on the composition of the specific samples tested. As discussed in section 2 (Identity of substances), the CAS RN of an LBPN is insufficient to determine the degree and severity of processing, and therefore the final composition, of any specific sample of the CAS RN.

CONCAWE has developed an aquatic toxicity model specifically for petroleum hydrocarbon mixtures called PETROTOX (2012). This model is on the basis of chemical action via narcosis and accounts for additive effects using the toxic unit approach. PETROTOX can model petroleum hydrocarbon toxicity for C₄–C₄₁ compounds dissolved in the water fraction.^{Footnote 7} Substances smaller than C₄ are considered too volatile to impart significant aquatic toxicity, while those larger than C₄₁ are considered too hydrophobic and immobile to impart significant aquatic toxicity. PETROTOX generates toxicity estimates in terms of loading rates rather than concentrations, thereby accounting for the poor solubility of petroleum substances in water. The model can also estimate a chronic no-observed-effect loading rate (NOELR) by utilizing an average acute-to-chronic ratio (ACR).

The moderate to high volatility of LBPNs suggests that air will be an important receiving medium for these substances. However, some release into water is also expected to occur, for example, during use of LBPNs at industrial facilities or in products available to consumers. Release of LBPNs to the aquatic environment will primarily occur following secondary wastewater treatment (see section 7.2, Ecological exposure assessment). As LBPNs are UVCBs consisting of various individual components, each with its own physical and chemical properties that impact its removal during wastewater treatment, such treatment will result in the differential removal of components of LBPNs. Thus, the relative proportion of individual components in the LBPN released following wastewater treatment differs from that of the LBPN at the time when it originally entered the treatment system. In order to determine the toxicity of the modified LBPN released in effluent following wastewater treatment, the removal of hydrocarbons during wastewater treatment and, thus, the composition of the post-wastewater treatment LBPN, was estimated. The estimation of the removal of hydrocarbons utilizes the library of hydrocarbon representative structures, their physical-chemical properties, and the mapping scheme of hydrocarbons to certain hydrocarbon blocks found within the PETROTOX v3.06 model (PETROTOX 2012). The percent removal of hydrocarbon blocks during wastewater treatment is estimated on the basis of the removal of individual hydrocarbon representative structures using the UVCB Modifier, which incorporates information from the PETROTOX library and estimates from the SimpleTreat version 3.1 model (SimpleTreat 2003). SimpleTreat is a wastewater treatment model that estimates removal of substances via sorption, volatilization and degradation but does not provide information on degradation products. From this analysis, the new relative proportion of components in the LBPN following wastewater treatment was estimated on the basis of the hydrocarbon blocks, and PETROTOX was used to estimate the acute toxicity and chronic NOELR for 4 aquatic species utilizing the post-wastewater treatment composition of the released LBPN. PETROTOX v3.06 uses an ACR of 3.83 to determine the NOELR; however, more recent analysis has adjusted the average ACR to 5.22 (McGrath et al. 2018). Therefore, the ACR in PETROTOX v3.06 was manually adjusted to 5.22 to account for this new analysis when calculating chronic values.

Acute LL₅₀ and chronic NOELRs were generated for 4 species using LBPNs ranging from 100% aliphatic (low aromatic) to 100% aromatic (high aromatic) content prior to wastewater treatment. The results are summarized in Tables 7.2 to 7.5. PETROTOX was run using the low-resolution mode and 4 hydrocarbon blocks, with boiling point ranges of 0.1°C to 69.61°C, 69.62°C to 128.64°C, 128.65°C to 179.24°C, and 179.25 °C to 221.59°C. Boiling point ranges were selected on the basis of the boiling point for the C₄ and C₆, >C₆ and C₈, >C₈ and C₁₀, and >C₁₀ and C₁₂ n-alkanes. The lowest NOELR for a given composition was used as the critical toxicity value (CTV). A default headspace of 10% was selected for the model runs.

Petroleum hydrocarbons, such as those found in the LBPNs, are expected to have similar toxicities to freshwater and marine species because their primary mode of toxic action is non-polar narcosis. As neutral narcotic substances, petroleum hydrocarbons are not affected by the dissolved salts present in greater quantities in sea water. For this reason, both freshwater and marine aquatic toxicity data were considered together in determining the aquatic CTV. 

Model results for a 100% aliphatic LBPN, an 80% aliphatic: 20% aromatic LBPN, a 50% aliphatic: 50% aromatic LBPN, and a 100% aromatic LBPN are presented in Tables 7‑2 to 7‑5, respectively. The lowest toxicity endpoints obtained using PETROTOX are comparable with those derived for the same organism in laboratory testing, although modelled values are generally lower. This may be due to the default headspace value used for toxicity estimation in the PETROTOX model (10%) compared with that used in laboratory tests, where headspace may be greater than 10% even in closed test systems. For example, the lowest empirical chronic endpoint for green algae was 2.5 mg/L (Table 7‑1), while the corresponding chronic modelled values were 0.1 mg/L to 0.4 mg/L (Table 7‑2 to Table 7‑5). It should be noted that the chronic PETROTOX value is an NOELR (no-effect level), while the empirical chronic value is an EL₅₀ (median effect level).

Some patterns were evident within the PETROTOX modelling. For all species and both acute and chronic endpoints, lowest endpoint values were determined for the 100% aromatic LBPN (Table 7‑5) and highest values for the 100% aliphatic LBPN (Table 7‑2), with intermediate values obtained for the 80% aliphatic: 20% aromatic and 50% aliphatic: 50% aromatic mixed LBPN (Table 7‑3 and Table 7‑4). In addition, chronic endpoint values were consistently lower than acute values and post-wastewater treatment chronic values were most commonly lower than pre-wastewater treatment values. Wastewater treatment increased the proportion of aromatic components of LBPNs; the composition of LBPNs changed from 80% aliphatic: 20% aromatic to 70% aliphatic: 30% aromatic, and from 50% aliphatic: 50% aromatic to 36.6% aliphatic: 63.4% aromatic post wastewater treatment. The predicted increase in relative toxicity following wastewater treatment may result from a loss of the more volatile and degradable LBPN components, resulting in a higher proportion of soluble substances with greater exposure potential for aquatic organisms. However, it should be noted that the overall release of LBPNs to effluent is greatly reduced during wastewater treatment, resulting in much lower exposure concentrations and absolute toxicity in the treated effluent. In summary, the overall lowest values, and therefore the highest toxicity, were associated with 100% aromatic LBPNs and chronic endpoints (that is, NOELRs) post wastewater treatment. The marine amphipod, Rhepoxynius abronius, was the most sensitive test species in PETROTOX.

Abbreviations: LL₅₀, median lethal loading; NOELR, no-observed-effect loading rate

Abbreviations: LL₅₀, median lethal loading; NOELR, no-observed-effect loading rate
^a Composition became 70% aliphatic: 30% aromatic after wastewater treatment

Abbreviations: LL₅₀, median lethal loading; NOELR, no-observed-effect loading rate
^a Composition became 36.6% aliphatic and 63.4% aromatic after wastewater treatment

Abbreviations: LL₅₀, median lethal loading; NOELR, no-observed-effect loading rate

On the basis of the PETROTOX modelling results, CTVs of 0.09, 0.07, 0.06 and 0.05 mg/L were determined using the chronic NOELR values for Rhepoxynius abronius of post-wastewater treatment LBPNs initially composed of 100% aliphatic components, 80% aliphatic: 20% aromatic components, 50% aliphatic: 50% aromatic components, and 100% aromatic components, respectively. As these values represent chronic no‑effect values for the most sensitive species, no assessment factor was applied to convert the CTVs to predicted no-effect concentrations (PNECs). Experimental toxicity data for sediment-dwelling organisms are not available; however, it is expected that their sensitivity to LBPNs would be similar to that of aquatic organisms.

7.1.2 Effects on soil-dwelling organisms

Few empirical terrestrial toxicity data were found for the LBPNs. Data derived for gasoline (CAS RNs 86290-81-5 and 8006-61-9) in addition to some older data for Stoddard solvent (CAS RN 8052-41-3) were used in order to estimate toxicity for the group.

ESG International (2000) investigated the effects of additive-free motor gasoline on 2 soil invertebrate species (earthworm, Eisenia fetida, and springtails, Onychiurus folsomi) and 4 plant species (alfalfa, Medicago sativa; barley, Hordeum vulgare; corn, Zea mays; and red fescue, Festuca rubra). Lowest 7-day median lethal (LC₅₀) concentrations of 630 mg/kg and 710 mg/kg soil dry weight (dw) were reported for earthworms in closed-air and open-air test systems, respectively. Lowest acute median inhibition (IC₅₀) concentrations for growth in plant species were 1770 mg/kg and 2700 mg/kg soil dw for barley and corn, respectively. The results indicate that gasoline has low to moderate hazard potential for the species tested. Although chronic testing was also conducted, loss of the test substance due to high volatility made the maintenance of stable exposure concentrations difficult and the results are therefore not presented here.

ECHA (c2007-2018) reported on the estimation of potential long-term effects of gasoline on terrestrial species through direct soil contact by characterizing chronic terrestrial toxicity endpoints with the use of the petroleum hydrocarbon toxicity model, PETROTOX (2012). Generic PNECs of between 0.4 mg/kg and 20.8 mg/kg soil dw were calculated from the model, with the no-effect range applicable to terrestrial invertebrates, plants and microorganisms (ECHA c2007-2018). These values are substantially lower than the lowest 7-day L/IC₅₀ values determined by ESG International (2000) for soil invertebrates (630 and 710 mg/kg soil dw) and terrestrial plants (1770 and 2700 mg/kg soil dw). ESG International (2000) reported difficulties in maintaining stable test concentrations during the longer 9-day to14-day chronic testing due to the high volatility of the test substance; because of this, the chronic test results are not provided here. However, it is possible that the lowest chronic effect concentrations would be lower than those reported for the shorter duration studies had longer continuous exposure been possible. While the ECHA (c2007-2018) results are deemed acceptable for inclusion as a line of evidence in the assessment, it should be noted that the PETROTOX model is designed to estimate hazard to aquatic species and may not fully represent the hazard potential for soil organisms. ECHA (c2007-2018) assigns a reliability of 2 (reliable with restrictions) to the model output values.

2 studies examined the effects of Stoddard solvent (CAS RN 8052-41-3) on terrestrial organisms. Voigt (1953) reported that spraying Stoddard solvent onto the root tips of tree seedlings resulted in an increased oxygen uptake of 38.5%, 7.6%, 18.8%, and 19% for Jack pine (Pinus banksiana), red pine (P. resinosa), white pine (P. strobus), and black locust (Robinia pseudoacacia), respectively. The observed increase in oxygen uptake was attributed to acute temporary injury of the root tips following application of the solvent (Voigt 1953). Stoddard solvent applied to coarse sandy soil at a rate of 100 gallons per acre decreased the resident soil micro-population to 221 colonies from 518 colonies in the untreated control (Persidsky and Wilde 1955). The growth and weight of the soil fungus, Aspergillus niger, mycelia were also adversely affected, with an approximate 30% reduction in both relative to fungi in untreated soil (Persidsky and Wilde 1955).

The Canada-Wide Standards for Petroleum Hydrocarbons (PHC) in Soil (CCME 2008) provides soil standards for petroleum products on the basis of toxicity to a variety of terrestrial organisms (invertebrates, plants). These standards are on the basis of 4 fractions of total petroleum hydrocarbons (TPHs): Fraction 1 (F1) (C₆ to C₁₀), F2 (greater than C₁₀ to C₁₆), F3 (greater than C₁₆ to C₃₄) and F4 (greater than C₃₄), and they assume a ratio of 80% aliphatics to 20% aromatics. LBPNs contain hydrocarbons that are predominantly in the C₄ to C₁₂ carbon range (CONCAWE 2001a, b), indicating that they occur mainly in the F1 fraction and, to a lesser extent, the F2 fraction as outlined in the Standards. The standards are also divided into 4 land-use classes (agricultural, residential, commercial, and industrial) and 2 soil types (coarse-grained and fine-grained soils) for the determination of remedial standards. The most sensitive land-use and soil type is typically agricultural coarse-grained soils. The standards for soil contact by non-human organisms for F1 and F2 are 210 and 150 mg/kg dry weight (dw) of soil, respectively (CCME 2008). The lower value of 150 mg/kg dw of soil is used as a conservative terrestrial CTV. As the Canada-Wide Standards were developed to protect key ecological receptors in the soil (CCME 2008), and as the CTV chosen for use in this assessment is the most protective value (that is, lowest relevant Standard for coarse-grained agricultural soils), no assessment factor was applied and the CTV of 150 mg/kg dw of soil also serves as the PNEC.

7.1.3 Effects on wildlife

Selected endpoints (mortality and reproduction) from studies on small mammals used to evaluate human health effects were also used to evaluate terrestrial wildlife toxicity. For Stoddard solvent (CAS 8052-41-3), a lowest-observed-adverse-effect-concentration (LOAEC) for short-term or subchronic exposure on the basis of non-neoplastic effects was determined to be 214 mg/m³ on the basis of an inflammatory response of the respiratory tract in mice exposed for 4 days (Riley et al 1984). Studies for previously assessed industry- and site-restricted LBPNs are also considered (EC, HC 2011, 2013). For example, CAS RN 64741-55-5 delivered via inhalation at 9041 mg/m³ was considered to be a no-observed-adverse-effect-concentration (NOAEC) for systemic toxicity in rats using a reproductive/developmental toxicity testing protocol (Schreiner et al. 1999; API 2008).

Rats exhibited a median lethal dose (LD₅₀) of 3500 mg/kg bw when orally dosed with CAS RN 68955-35-1 (API 2008). An oral no-observed-adverse-effect-level (NOAEL) of 2000 mg/kg bw/day was determined for CAS RN 64741-55-5 for reproductive and developmental toxicity in rats (Stonybrook Laboratories 1995), and an oral NOAEL of 50 mg/kg bw/day was determined for CAS RN 68513-74-8 for reproductive and developmental toxicity in rabbits (this was the highest dose tested) (Miller and Schardein 1981). These values indicate that these substances are not highly hazardous to terrestrial mammals for these particular endpoints and exposure routes.

Mammalian inhalation toxicity tests for unleaded gasoline were considered as a surrogate for estimating the environmental toxicity of these LBPNs. Adult rats were tested for 6 hours/day, 5 days/week for 13 weeks at up to 6570 mg/m³ (1552 ppm), without any treatment-related mortality (Kuna and Ulrich 1984). MacFarland et al. (1984) found no treatment-related mortalities in rats exposed for 113 weeks to an unleaded gasoline vapour at concentrations of up to 6170 mg/m³. The value of 6570 mg/m³ is the highest no-observed-effect concentration (NOEC) for mortality from studies on animals identified.    

7.2 Ecological exposure assessment

LBPNs may be released to the environment from their various uses. The uses considered for exposure analysis are those that are expected to have the highest potential for release to the environment. Uses as fuels or fuel additives and reaction intermediates are excluded from this analysis as they are expected to have negligible residual amounts after combustion or chemical reactions. For non-fuel, non-intermediate applications, LBPNs were used as additives incorporated into the various products identified on the basis of information submitted in response to CEPA section 71 surveys (Environment Canada 2011a, 2012a, 2012b, 2013, 2015; ECCC 2016a) and a voluntary data gathering initiative (ECCC 2016b).

3 exposure scenarios from the uses of LBPNs in non-fuel and non-intermediate applications were investigated: consumer release scenario resulting from the use of personal care and cosmetic products, paints and coating, adhesives and sealants, household cleaners, and automotive care products; generic formulation scenario for products available to consumers and industrial applications; and pulp and paper scenario for the use of processing aids by pulp and paper mills. Exposure to LBPNs in soils was also estimated because indirect releases may result from the application of biosolids to land from WWTS receiving wastewater containing LBPN component substances.

For LBPN substances considered in the consumer release scenario, 17 LBPNs are identified as being present in products available to consumers. Information on the types of uses and quantity data for 14 of these substances were submitted in response to 2 CEPA section 71 surveys (Environment Canada 2011a, 2012a, 2012b, 2013). No additional substances in the LBPNs Group were identified as being used in products available to consumers according to submissions in response to a later CEPA section 71 survey (Environment Canada 2015; ECCC 2016a) and a voluntary data gathering initiative (ECCC 2016b).

For LBPN substances considered in the generic formulation scenario, all 27 substances in the LBPNs Group are identified as being used in the formulation of both products available to consumers and those for industrial applications. Information on the types of uses and quantity data for 14 of these substances was submitted in response to 2 CEPA section 71 surveys (Environment Canada 2011a, 2012a, 2012b, 2013). The remaining 13 substances were limited to types of uses without quantity data according to information requested in a CEPA section 71 survey (Environment Canada 2015; ECCC 2016a) and a voluntary data gathering initiative (ECCC 2016b).

For LBPN substances considered in the pulp and paper scenario, 4 substances in the LBPNs Group are identified as being in products used in pulp and paper processing aids. Information on the types of uses and quantity data for these substances was submitted in response to a CEPA section 71 survey (Environment Canada 2011a, 2012b). No additional substances in the LBPNs Group were identified as being used in pulp and paper processing aids according to information submitted in response to 2 later CEPA section 71 surveys (Environment Canada 2012a, 2013, 2015; ECCC 2016a) and a voluntary data gathering initiative (ECCC 2016b). The 4 substances do not include those used in printing inks, some of which are expected to end up in pulp and paper mills involved in the deinking of printed paper products.

7.2.1 Calculation of Predicted Environmental Concentrations (PECs) and general assumptions

7 substances were selected as representatives for the calculation of PECs. Each of them accounted for more than 10% of the combined quantity of the substances identified under each scenario, on the basis of the data submitted in response to 2 CEPA section 71 surveys (Environment Canada 2011a, 2012a, 2012b, 2013). These LBPN substances represented more than 80% of all the reported quantities combined under each scenario. Table 7‑6 summarizes the boiling point range and composition used for exposure calculations under each scenario, on the basis of those representative substances. The highest boiling point range was selected for each scenario when multiple substances were used. This resulted in conservative estimates for wastewater treatment removal by ignoring the volatilization of certain light hydrocarbons. The percentage of aliphatics or aromatics was calculated by proportionating each substance used under a scenario. 

The overall wastewater treatment removal for LBPNs and post-treatment compositions were estimated under 3 scenarios: consumers’ release from products available to consumers, formulation of products available to consumers, and pulp and paper (Table 7‑7). As mentioned earlier, a hydrocarbon block method developed by CONCAWE (1996b) was used for the estimation. This method calculates the overall removal on the basis of the removal and proportion of each individual component of a hydrocarbon mixture. The method was programmed into an Excel spreadsheet, called UVCB Modifier, for automation by Environment and Climate Change Canada, and the removal of a hydrocarbon component was pre-estimated and incorporated into UVCB Modifier. SimpleTreat 3.1 (2003) was used to obtain removal estimates of individual components for primary and secondary wastewater treatment systems. For lagoon systems, the removal of a component was assumed to be the same as that for secondary systems, considering a similar treatment level (sedimentation followed by biodegradation) between the 2 types of systems. The boiling point ranges of LBPNs and their pre-treatment aliphatic/aromatic breakdowns, given in Table 7.6, are the basis for the determination of post-treatment compositions. All 3 treatment systems (primary, secondary, and lagoon) are applicable to the consumers’ release and formulation scenarios. The treatment level for the pulp and paper scenario is either secondary or lagoon.

7.2.2 Exposure scenario 1: Consumer release from products available to consumers

Components of LBPNs are expected to enter sewers, and subsequently, the aquatic environment via wastewater treatment systems from the use of many consumer/commercial products (for example, paints and coatings, adhesives and sealants, cleaners, automotive care, cosmetics, and personal care products). The level of the aquatic exposure is directly proportional to use quantity but is also affected by other parameters such as wastewater treatment removal and receiving water dilution.

A model called the Consumer Release Aquatic Model, or CRAM, was developed by Environment and Climate Change Canada to perform release and exposure calculations from the use of consumer and commercial products. The model assumes uniform per-capita use quantities throughout Canada. 3 types of wastewater treatment used in Canada (primary, secondary and lagoon systems) were considered. The level of exposure was estimated near discharge points. User inputs and model built-in defaults are summarized in Table 7‑8. Major user inputs included use quantity, release fraction to sewer, and wastewater treatment removal. Model built-in data included effluent flow and receiving water dilution. The effluent flow was derived from water usage. The dilution was determined as the ratio of the 10th percentile river flow to the effluent flow. The ratio was capped at 10 to account for limited dilution near discharge points in the case of large rivers. For lakes, the dilution was assumed to be ten-fold near discharge points. To conclude, the model generated a distribution of PECs in receiving water for thousands of wastewater treatment systems in Canada (Table 7‑9).

The total annual quantity of LBPNs used in consumer/commercial products was estimated on the basis of an average annual quantity per CAS RN. The average was determined to be 1 200 000 kg/yr per CAS RN from the combined quantity (16 800 000 kg/yr) of the 14 CAS RNs reported for quantities. For the remaining 13 CAS RNs in the LBPNs Group, the average was assumed. Therefore, the combined annual quantity of the 13 CAS RNs that did not have quantity data was estimated to be 15 600 000 kg/yr. The total annual quantity of LBPNs used under the consumers’ release scenario was estimated to be 32 400 000 kg/yr, which is the sum of the combined quantity of the 14 CAS RNs with reported quantities and the combined assumed quantity of the 13 CAS RNs without reported quantities.

CRAM results indicate that the aquatic PECs of LBPNs range from a minimum of 0.2 µg/L to a maximum of 26 µg/L. The median value of 1.2 µg/L implies that half of sites (discharge points of wastewater treatment systems) in Canada show PECs below 1.2 µg/L and the other half above 1.2 µg/L.

The total use quantity of all CAS RNs in the LBPNs Group is unknown and an upper bound was estimated. The use of the upper bound quantity in CRAM modelling resulted in conservative aquatic PECs. These conservative estimates are appropriate for initial screening of the potential ecological exposure concern of these substances.

The fraction lost to sewer is 0.04 for paints and coatings (European Chemicals Bureau 2003). As paints and coatings represent the predominant use of LBPNs (see Table 4-3), their loss fraction (0.04) is expected to be reflective of the typical extent of releases from consumer products and is therefore used in the calculations.

7.2.3 Exposure scenario 2: Generic formulation of products available to consumers

LBPNs are used in the formulation of a wide range of non-fuel, non-intermediate products. These include paints and coatings, processing aids, lubricants and greases, adhesives and sealants, inks and toners, automotive care products, and personal care products, according to information submitted in response to 3 CEPA section 71 surveys (Environment Canada 2011a, 2012a, 2012b, 2013, 2015; ECCC 2016a) and a voluntary data gathering initiative (ECCC 2016b).

A formulation scenario was developed to estimate the aquatic PEC of LBPNs resulting from formulation operations at a given site. Under this scenario, LBPNs are expected to be released to local wastewater treatment systems from equipment cleaning. Any amount remaining in the effluent after wastewater treatment enters the aquatic environment. The aquatic PEC is estimated by:

$PEC=\frac{{10}^9\times a\times Q\times E\times (1-R)}{N\times V}$

where

PEC: predicted environmental concentration in receiving water at a site, µg/L

a: scale-up factor from known use quantities to total use quantities at a site, unitless

Q: total known annual use quantity of LBPNs at a site, kg/yr

E: emission factor to wastewater, fraction

R: overall wastewater treatment removal, fraction

N: annual number of release days, d/yr

V: daily dilution water volume at a site near discharge point, L/d

10⁹: conversion factor to convert kg to µg, µg/kg

One PEC value was calculated per site from the inputs summarized in Table 7-10. A large number of sites (between 100 and 200) were identified from information submitted in response to 2 CEPA section 71 surveys (Environment Canada 2011a, 2012a, 2012b, 2013). Each site corresponded to one wastewater treatment system and encompassed one or more industrial facilities, with each facility involved in the use of one or more CAS RNs. The annual use quantities of these CAS RNs were combined together per facility. These combined quantities were then further combined across different facilities at a site. This yielded a site-combined use quantity (Q) from surveyed quantities. The use quantity was specific to a site and varied in a wide range from 1000 kg/yr to 10 000 000 kg/yr. The daily dilution water volume (V) was also specific to a site and ranged from 1 million L/d to 23 000 million L/d. This parameter is interpreted as the daily volume of water available to dilute a substance released from a wastewater treatment system near the discharge point. It was equal to the effluent flow multiplied by the dilution of the receiving water. The dilution was dependent upon the size of the receiving water but was capped at 10-fold due to limited mixing near the discharge point.

With the above input values, aquatic PECs were estimated for all sites (Table 7-11) and were in the range of 0.2 µg/L to 36 µg/L when the 5^th and 95^th percentiles were considered as the statistically realistic minimum and maximum. PEC values below the 5^th or above the 95^th percentiles were expected to be unreliable as were sensitive to compounding uncertainties associated with the related sites.

7.2.4 Exposure scenario 3: Pulp and Paper

Products containing LBPN CAS RNs were reported to have been sold to pulp and paper mills and were used as process aids in papermaking (Environment Canada 2012b; ECCC 2020). Papermaking is a water-intensive operation. Water is continuously removed from pulp as the pulp is transformed into paper through a paper machine. The water removed is recycled and reused to prepare pulp slurry, forming a water circulation loop. Chemicals used during papermaking are collected into this circulation loop, dissolved in water, and sorbed to pulp fibre. Fresh water is constantly added to prevent chemical buildup in the loop. Meanwhile, excess water is removed and discharged to on-site wastewater treatment. Residual chemicals are released to the aquatic environment via wastewater treatment effluent.

The aquatic exposure or PEC was estimated near the discharge point of a mill’s effluent. The area near the discharge point is part of the ecosystem. It is a sensitive and vulnerable zone for aquatic life because of the presence of harmful chemicals at the highest concentrations. These chemicals are dispersed as they move away from the discharge point. They can also be eliminated over time by natural processes like biodegradation and photodegradation. The area near the discharge point was selected for exposure analysis to ensure the protection of a sensitive and vulnerable zone.

A mass balance method is used for PEC estimates. The principle of the method is described in the European Chemicals Agency’s guidance for environmental exposure assessment (ECHA 2016). An internal database was developed and is regularly updated by ECCC for the Canadian pulp and paper sector. The database contains a range of operational data in support of exposure assessment. In Canada, mills operate continuously throughout the year. Most mills are equipped with on-site wastewater treatment systems and discharge effluent directly to receiving water. A limited number of mills discharge wastewater to off-site treatment systems before it is released to the aquatic environment. The nature of a continuous operation requires that process aids be used continuously. This leads to continuous releases and exposure via effluent. The resulting PEC from a mill can be calculated from the amount of a substance present in effluent and the dilution of the effluent in receiving water. The concepts of emission factor and wastewater treatment removal are used to estimate the amount present in effluent.

The PECs in receiving water from the mills involved with LBPNs were estimated using the following equation: 

$PEC=\frac{{10}^9\times Q\times E\times (1-R)}{N\times F\times D}$

where

PEC: predicted environmental concentration in receiving water near discharge point, µg/L

Q: annual use quantity of LBPNs at a mill, kg/yr

E: emission factor to wastewater treatment, fraction

R: wastewater treatment removal, fraction

N: number of annual operation days, d/yr

F: daily wastewater treatment effluent flow, L/d

D: receiving water dilution factor near discharge point, unitless

10⁹: conversion factor from kg to µg, µg/kg

PEC results and the inputs used are summarized in Table 7-12.

Figure 7-1 shows how the emission factor was estimated. In papermaking, a process water stream recirculated around the paper machine, called white water, is recycled and mixed with pulp feed. To prevent chemical buildup, excess white water is taken out of the recycle loop and discharged to on-site wastewater treatment. Fresh water is added to maintain the balance of white water. It was assumed that the entire amount of LBPN substances used during papermaking entered the white water stream. It was further assumed that this amount entered the pulp feed via the white water recycle.

By definition, the emission factor is the ratio of the aquatic loss quantity to the use quantity (Figure 7-1). The ratio is equivalent to the concentration ratio of LBPNs between excess white water and pulp feed. The concentration consists of 2 phases, dissolved and sorbed. The sorbed concentration can be correlated with the dissolved concentration, assuming equilibrium sorption. Thus, the emission factor is given as

$E=\frac{C_{excess}}{C_{feed}}=\frac{C_{aq}+C_{aq}{K}_d{S}_{excess}}{C_{aq}+C_{aq}{K}_d{S}_{feed}}=\frac{1+K_d{S}_{excess}}{1+K_d{S}_{feed}}$

where

C_excess: concentration of LBPNs in excess white water, mg/L

C_feed: concentration of LBPNs in pulp feed, mg/L

C_aq: dissolved concentration in excess white water or pulp feed, mg/L

K_d: solids-water partition coefficient, L/kg

S_excess: solids content of excess white water, kg/L

S_feed: solids content of pulp feed, kg/L

The dissolved concentration C_aq is the same for excess white water and pulp feed. This is because the aqueous phase is a common phase due to the constant recycling of white water. The equation is simplified by cancelling C_aq.

The solids-water partition coefficient K_d can be estimated from the octanol-water partition coefficient K_ow. An empirical correlation derived for wastewater solids (Dobbs et al. 1989) was used. Wastewater solids and pulp fibre are both organic in nature. They are equivalent to each other in terms of sorptive ability for neutral organic chemicals like LBPNs. The correlation is deemed appropriate.

Log K_d = 0.58logK_ow + 1.14 + logf_om

where

K_ow: octanol-water partition coefficient, unitless

f_om: organic matter fraction of pulp feed solids, unitless

Values for log K_ow range from 3.0 to 7.2 for LBPNs. The lowest value of 3.0 was selected for an upper end estimate for the emission factor. This value represents moderate hydrophobicity. The pulp feed solids consist of fibre (organic matter) and added chemicals added. The total added chemicals are less than 0.5 (OECD 2009). The organic matter fraction f_om is normally greater than 0.5 but was conservatively assumed to be 0.5. This assumption only considered half of the solids for sorption and intentionally overestimates aquatic losses. The combination of low values for both log K_ow and f_om yielded a lower-end estimate of 2.58 for log K_d. This estimate indicates moderate sorption. The emission factor was thus calculated to be 0.23 at a moderate log K_d of 2.58. This result indicates that only a small fraction of LBPNs are lost to excess white water or enters wastewater treatment. The majority is lost to evaporation during the drying of paper sheet. In the calculation, solids in pulp feed S_feed = 0.01 kg/L (Gavrilescue et al. 2008) and solids in excess white water S_excess = 0.0003 kg/L (Vurdiah 2015) were used. Both pulp feed and excess white water are highly diluted. Their density was approximated at 1 kg/L. Although small, the emission factor (E) calculated provides an upper end value for aquatic releases.

As pulp and paper mills normally operate continuously throughout the year, the number of annual operation days (N) was set as 350 d/yr (ECCC internal database).

The mills involved with LBPNs are equipped with secondary treatment. A wastewater removal rate (R) of 93% was previously estimated for secondary treatment (see section 7.2.2).

Lastly, a dilution factor (D) of 10 was chosen for this scenario. The area near the effluent discharge point was selected for PEC calculations. For Canadian mills, receiving water bodies are large and mixing is a long process. The dilution near the discharge point is limited and effluent may need to travel a great distance downstream before being fully diluted. A default value of 10-fold dilution was therefore used to account for this dilution near the point of discharge.

According to information submitted in response to a CEPA section 71 survey (Environment Canada 2012b), 5 mills were identified that used substances from the LBPNs Group as process aids in papermaking. The annual use quantity of all CAS RNs combined at a mill was assumed to be equivalent to the surveyed quantity of sales to the mill. The quantity showed a great degree of variability, ranging from 1000 kg/yr to 100 000 kg/yr. The effluent flow of a mill was also highly variable, ranging from 1 million L/d to 100 million L/d. When calculating a PEC, the mill-specific annual use quantity (Q) and the mill-specific effluent flow (F) were used. Values for the other parameters (E, N, R, and D) were constant and were used consistently across all PECs, as detailed above.  

The calculated PECs for the 5 mills were in the range of 1 µg/L to 36 µg/L. These are considered to be conservative values because a conservative estimate was used for the emission factor. Differences between PECs were a result of differences in the annual use quantity and effluent flow.  

7.2.5 Exposure scenario 4: Soil Exposure

An approach described by the European Chemicals Agency (ECHA 2016) is used to estimate soil exposure to LBPNs resulting from the land application of biosolids. In this approach, the amount of biosolids accumulated within the upper 20 cm layer of soil over 10 consecutive years serves as the basis for the estimation. The loss via degradation, volatilization, leaching, or soil run-off is assumed to be nil, and this assumption yields conservative estimates for soil exposure.

The predicted environmental concentration of LBPNs in soil is estimated using the following equation:

${PEC}_{soil}=\frac{C_s\times A\times N}{d\times \rho }$

where

PEC_soil: predicted environmental concentration in soil, mg/kg

C_s: concentration of LBPNs in biosolids, mg/kg

A: annual biosolids land application rate, kg/m²-yr

N: number of years for biosolids land application, yr

d: soil mixing depth, m

r: dry soil density, kg/m3

The concentration of LBPNs in biosolids (C_s) is estimated on the basis of 2 major sources of releases to wastewater treatment systems. One is the industrial release from formulation facilities under the formulation scenario, while the other is the consumer release in municipalities where the formulation facilities are located. The 2 sources enter the same wastewater treatment systems and a fraction of their combined quantity ends up in biosolids.

The combined quantity of LBPNs is calculated as the sum of the 2 sources.

The total annual industrial release quantity of LBPNs under the generic formulation scenario was obtained by adding up release quantities from all formulation facilities, which is 233 000 kg/yr. The daily release quantity is then 638 kg/d assuming 365 days of release per year.

The total consumer release quantity of LBPNs is estimated on the basis of the population served by the wastewater treatment systems that also receive wastewater from formulation facilities. The population is estimated from the total effluent flow of these treatment systems using an average per capita daily wastewater generation rate, 600 L/person-day (CWWA 2001). The total effluent flow was 8 468 212 400 L/d, according to calculations in generic formulation scenario. The population served is estimated to be 8 468 212 400 L/d / 600 L/d-person = 14 114 000 persons.

As discussed in the consumer release scenario, the total use quantity of LBPNs was 32 400 000 kg/yr for the entire Canadian population of 36 000 000, and the fraction lost to sewer was 0.04. By prorating by population, the total consumer release quantity of LBPNs from formulation-related municipalities is estimated as:

total daily consumer release quantity of LBPNs from formulation related municipalities = 32 400 000 kg/yr × 0.04 × 14 114 000/(36 000 000 × 365 d/yr) = 1392 kg/d

The total combined release quantity from the 2 sources is determined as:

total combined release quantity of LBPNs from formulation facilities and formulation related municipalities = 638 kg/d + 1392 kg/d = 2030 kg/d

The total biosolids quantity is estimated from the total effluent flow using an average biosolids generation rate of 104 mg per litre of effluent (Kim et al. 2013).

Total biosolids quantity from formulation related municipalities = 8 468 212 400 L/d × 104 mg/L = 880 694 000 000 mg/d = 880 694 kg/d

Assuming that the total combined release quantity of LBPNs ends up in biosolids, the value for C_s is estimated by dividing this combined quantity by the total biosolids quantity.

C_s = 2030 kg/d / 880 694 kg/d = 0.0023 kg/kg = 2300 mg/kg

In Canada, the maximum land application rate of biosolids varies and is regulated by provinces/territories. The highest rate occurs in Alberta and is used in the PEC calculations (Alberta Environment 2001).

A = 8.3 tonne/ha-yr = 0.83 kg/m²-yr

A period of 10 consecutive years and a soil mixing depth of 20 cm were suggested by the European Chemicals Agency (ECHA 2016) for the calculation of exposure in biosolids-applied land.

N = 10 yr

d = 0.2 m

The soil density (dry) was reported by Williams (1999) as

r = 1200 kg/m³

The soil PEC was estimated to be 80 mg/kg on a dry basis. This is a conservative estimate because the total combined release quantity of LBPNs is assumed to sorb entirely to biosolids, while only a fraction of the release quantity of LBPNs undergoes this pathway. Although the quantity of LBPNs entering soil was conservatively estimated, the resulting soil PEC is not expected to be high enough to cause the formation of NAPL (free hydrocarbons). LBPNs are predominantly in the C₄ to C₁₂ carbon range (CONCAWE 2001a). This range falls approximately under 2 groups of hydrocarbons (C₆ to C₁₀ and C₁₀ to C₁₆) for which soil standards are available in the Canada-Wide Standards for Petroleum Hydrocarbons in Soil (CCME 2008). These standards (the lowest being 150 mg/kg dry weight) are considered to be below the point for the formation of NAPL, while the soil PEC obtained here is, at 80 mg/kg, even lower. Thus, the form of LBPNs present in soil under this scenario is in a sorbed phase.

7.3 Characterization of ecological risk

The approach taken in this ecological assessment was to examine assessment information and develop proposed conclusions using a weight-of-evidence approach and precaution. Evidence was gathered to determine the potential for LBPNs examined in this assessment to cause harm in the Canadian environment. Lines of evidence considered include those evaluated in this assessment that support the characterization of ecological risk in the Canadian environment. Reliable secondary or indirect lines of evidence were considered when available.

The potential for cumulative effects was considered in this assessment by examining cumulative exposures from within the group of Petroleum LBPNs.

7.3.1 Risk quotient analysis

Risk quotient analyses were performed by comparing conservative estimates of exposure (PECs; see the Ecological Exposure Assessment section) with ecotoxicity information (PNECs; see the Ecological Effects Assessment section) to determine whether there is potential for ecological harm in Canada. Risk quotients (RQs) were calculated by dividing the PEC by the PNEC for relevant environmental compartments and associated exposure scenarios. Aquatic PNECs were the lowest chronic NOELRs for the most sensitive species, Rhepoxynius abronius, as determined by PETROTOX (Table 7‑2 to Table 7‑5). PNEC values were 0.09, 0.07, and 0.05 mg/L for post-wastewater treatment of LBPNs composed initially of 100% aliphatic LBPNs, 80% aliphatic: 20% aromatic LBPNs (this changed to 70% aliphatic: 30% aromatic after wastewater treatment), and 100% aromatic components, respectively. PNEC values were compared to the PECs estimated for the 3 primary release scenarios, which are consumer release, generic formulations, and pulp and paper processing. For terrestrial scenarios, the lowest of the 2 applicable Canada-Wide Standards for eco-soil contact in coarse-grained agricultural soil (that is, 150 mg/kg dw for Fraction 2; CCME 2008) was used as the PNEC. Table 7‑13 presents RQs for the LBPNs.

^a Initially 100% aromatic LBPN post wastewater treatment
^b Initially 80% aliphatic: 20% aromatic LBPN post wastewater treatment but becomes 70% aliphatic: 30% aromatic
^c Initially 100% aliphatic LBPN post wastewater treatment
^d Canada-Wide Standard for Petroleum Hydrocarbons in Soil, F2 Standard for Coarse-Grained Agricultural Soil (CCME 2008)

For the consumer release, formulation, and pulp and paper scenarios, RQs greater than 1 were not determined for any locations regardless of the aromatic content of the LBPNs. For exposure of LBPNs in soils, there was less than 1 on the basis of conservative assumptions.

7.3.2 Consideration of the lines of evidence

To characterize the ecological risk of LBPNs, technical information for various lines of evidence was considered (as discussed in the relevant sections of this report) and qualitatively weighted. The key lines of evidence supporting the assessment conclusion are presented in Table 7‑14. Weighted lines of key evidence considered in order to determine the potential for LBPNs to cause harm in the Canadian environment, with an overall discussion of the weight of evidence, are provided in section 7.3.3. The level of confidence refers to the combined influence of data quality and variability, data gaps, causality, plausibility, and any extrapolation required within the line of evidence. The relevance refers to the impact the line of evidence has when determining the potential to cause harm in the Canadian environment. Qualifiers used in the analysis ranged from low to high, with the assigned weight having 5 possible outcomes, that is, high, moderate to high, moderate, low to moderate, and low.

Abbreviations: PEC, predicted exposure concentration; PHC, petroleum hydrocarbon; PNEC, predicted no-effect concentration; RQ, risk quotient.
^a Level of confidence is determined according to data quality, data variability, data gaps (that is, are the data fit for purpose).
^b Relevance refers to the impact of the evidence in the assessment.
^c Weight is assigned to each line of evidence according to the overall combined weights for level of confidence and relevance in the assessment.

7.3.3 Weight of evidence for determining potential to cause harm to the Canadian environment

LBPNs are the lightest and most volatile fraction of the liquid petroleum hydrocarbons. If released into the air, most components of LBPNs are expected to remain in the air because of their volatility. If released into the water, the volatile components of LBPNs will tend to rise to the water surface while more soluble components will remain in the water, and those with higher sorption potential will adsorb to solid media. If released into the soil, more volatile LBPN components can be expected to distribute into the air, while the less volatile components will remain within the soil.

When released into the environment, LBPNs are expected to preferentially distribute into the air, where the majority of their component substances are predicted to degrade through reaction with photochemically produced hydroxyl radicals. Some LBPN components may also photodegrade through reaction with ozone and nitrate radicals. Most LBPN components are not expected to persist in air, other than the short-chain (C₄ to C₆) alkanes and isoalkanes and one-ring aromatic components, which may remain in air for longer periods of time. Most LBPN components are expected to biodegrade; however, larger (C₉ to C₁₂) one- and two-ring aromatics and cycloalkane monoaromatics, as well as C₅ alkenes, have the potential to persist in water, soil, and/or sediment for long periods of time. Persistency in air, water, soil and sediment increases the duration of exposure to organisms.

LBPNs contain some components that have a high potential to bioaccumulate, specifically larger (that is, C₉ and C₁₂) n-alkanes, n-alkenes, and cycloalkanes. While most LBPN components are not expected to biomagnify, empirical BMF and TMF values of greater than 1 have been reported for some two-ring aromatics, indicating that these substances may biomagnify in aquatic or terrestrial food webs.

Predicted no-effect concentrations (PNECs) for aquatic organisms were determined using the PETROTOX model, which allows for the estimation of LBPN toxicity following compositional changes that occur during wastewater treatment. Lowest chronic modelled no-observed-effect loading rates (NOELRs) of 0.05, 0.07 and 0.09 mg/L were obtained for the most sensitive test species, Rhepoxynius abronius, following wastewater treatment of LBPNs initially composed of 100% aromatic components, 80% aliphatic: 20% aromatic components, and 100% aliphatic components, respectively. These values indicate that LBPNs with a higher aromatic proportion are more hazardous than mainly aliphatic LBPNs. These values were selected as PNECs in the risk quotient analysis and indicate that LBPNs ranging in composition from 0% to 100% aromatic components have the potential to be hazardous to aquatic organisms.

Information submitted in response to 3 CEPA section 71 surveys as well as a voluntary data gathering initiative were used to identify 3 exposure scenarios associated with the highest potential for release to the environment from uses of substances in the LBPNs Group in non-fuel and non-intermediate applications. The following scenarios were identified: a consumer release scenario resulting from the use of products available to consumers such as paints and coatings, adhesives and sealants, personal care and cosmetic products, household cleaners, and automotive care products; a generic formulation scenario for products available to consumers in which substances from the LBPNs Group are used in the products’ formulation; and a pulp and paper scenario in which substances in the LBPNs group are present in processing aids used in pulp and paper mills. Predicted exposure concentrations (PECs) developed for each of these scenarios were compared with the PNECs determined using the PETROTOX model.

The resulting risk quotients (RQs) did not exceed 1 in any of the 3 scenarios considered for LBPNs ranging from 0% to 100% aromatic content. This indicates that LBPNs used in all 3 scenarios are unlikely to be causing ecological harm to the aquatic environment at current levels of exposure in Canada.

Few empirical terrestrial toxicity data are available; however, on the basis of results obtained for gasoline and Stoddard solvent, LBPNs are considered to have low to moderate acute hazard potential to soil species. WHO (1996) noted that, given the volatility of the substance and possible lowered bioavailability due to its sorption to soil, results reported for laboratory testing with Stoddard solvent and tree seedlings were likely to overestimate effects in the field. OECD (2012dc) reported a log K_ow range of 3.5 to 7.2 for Stoddard solvent (Table B-1), suggesting that sorption may indeed contribute to reducing the toxic potential in the environment.

The volatility of LBPNs suggests that inhalation could be an important exposure route for air-breathing wildlife. A previous assessment of industry-restricted LBPNs (EC, HC 2013) determined that margins of exposure derived from comparing a CTV of 14.7 mg/m³ with an upper-bounding exposure of 0.73 µg/m³ (estimated from an annual emission rate of 30 000 kg/year) were adequate to account for uncertainties in the dataset for non-carcinogenic inhalation effects in humans. Although the LBPNs considered in the present assessment have a higher estimated emission rate of 21 000 kg/year to 751 000 kg/year, they also have a higher CTV of 214 mg/m³. Therefore, the substances in the LBPNs Group are unlikely to be causing ecological harm to air-breathing wildlife at current exposure levels in Canada, including terrestrial wildlife.

7.3.4 Sensitivity of conclusion to key uncertainties

LBPNs are particularly complex UVCBs because many of them are created to meet certain product specifications. Coupled with ongoing and continuing improvements to analytical capability and a preference to reduce aromatic content due to the added hazard posed by aromatics as compared to aliphatics, the composition of specific LBPNs has changed over time and in different jurisdictions. Therefore, some uncertainties exist in terms of the composition, physical-chemical properties and toxicities of LBPNs when using that spans a period of over 40 years. However, where applicable, this assessment recognizes the changing nature of these substances.

Compositional data were incomplete for many of the CAS RNs considered in this assessment. However, as PETROTOX modelling covered the full compositional range of LBPN aliphatic and aromatic content, it is not likely that the absence of full compositional information had an impact on the assessment outcome.

There are uncertainties associated with the estimated PECs for the consumer release and formulation scenarios for products available to consumers. For example, the combined annual use quantity of all 27 CAS RNs in LBPNs is uncertain as the total use quantity is unknown and was approximated. For the consumer release scenario, the total quantity was estimated by prorating for the 27 substances from the known use quantities of 14 substances. For the formulation scenario, use quantities were available for 14 of the 27 substances in the LBPNs Group known to be used in the formulation of products available to consumers. The quantity of the 13 CAS RNs without quantity data was estimated by applying a scaling factor on the basis of the total count of CAS RNs. The unknown use quantities of the 13 CAS RNs are assumed to be comparable to the known use quantities of the 14 CAS RNs. The total quantities calculated for all 27 CAS RNs under the consumer release and formulation scenarios can be considered as approximate and may be underestimates or overestimates.

Limited data are available on the quantities of LBPNs manufactured, imported and used in Canada. While 25 of the 27 CAS RNs examined in this assessment were surveyed through formal data gathering initiatives, quantitative data are available for only 14 CAS RNs. The available information indicates that the exposure analysis presented in this assessment captures the major non-fuel and non-intermediate applications in Canada. However, lesser applications have not been quantitatively considered in the assessment and may be contributing to the total loading of LBPNs into the environment.

There is uncertainty related to the emission factor used in the calculations for the formulation scenario. The value chosen (0.5%) is on the basis of formulation operations for paints and coatings (OECD 2009). Water was assumed to be used for cleaning equipment and discharged to sewer afterwards along with residual chemicals. However, the assumption may not be true for all sites. Organic solvents are often used for equipment cleaning and the spent solvents are disposed of as hazardous wastes instead of entering sewer. Overestimates of releases of LBPNs are therefore possible due to this assumption. Another uncertainty relates to on-site wastewater treatment. Many facilities are equipped with such treatment and discharge treated wastewater to sewer. However, these facilities could not be identified and their related treatment details were unknown. To overcome this uncertainty, all facilities were assumed to have no on-site wastewater treatment. This resulted in overestimates for LBPN releases. Overall, the 2 assumptions for the emission factor and on-site wastewater treatment resulted in conservative PECs. For the formulation scenario, confidence is high when the PEC range of 0.2 µg/L to 36 µg/L is used to indicate a low level of concern. The range is conservative and its upper end (36 µg/L or 95^th percentile) should be interpreted as the maximum estimate for aquatic exposure. Any value above the 95^th percentile is not statistically reliable due to the compounding conservatism in the assumptions made with respect to the emission factor and on-site wastewater treatment.

There are uncertainties regarding the spill release estimate of LBPNs, because spills were not specifically reported for LBPNs; instead, spills information on more general terms for petroleum substances were used in spill analysis. This would likely result in conservative estimates of spill frequency and volume. However, the overall spill releases of these LBPNs are minor compared with release from industrial and commercial activities.

No ecological toxicity data for sediment-dwelling organisms are available. A risk quotient on the basis of exposure in sediment may be calculated on the basis of the PEC and PNEC values for water and used for sediment risk characterization. In the calculation, the bottom sediment, the suspended sediment, and the aqueous phase are assumed to be in equilibrium. The benthic and pelagic organisms are assumed to have similar sensitivities to LBPNs. Therefore, this equilibrium approach would result in risk quotients (PEC/PNEC) for sediment that are similar to those for water. The conservative RQs for water, and thus sediments, are all below 1, indicating a low risk of harm to sediment-dwelling organisms from LBPNs in Canada. These assumptions tend to underestimate the risk as they do not take into consideration the ingestion of LBPNs in sediment. However, physical-chemical properties such as low assimilation efficiency, growth dilution, and metabolism (in some organisms) are expected to lower the accumulation potential of LBPNs. In addition, LBPNs are volatile and very soluble in water. When released to the environment, they are expected to partition to air because of their high volatility or remain in water due to their high solubility.

8. Potential to cause harm to human health

On the basis of the general population exposure and human health effects assessments presented for each case, for the purpose of human health risk characterization, the LBPNs are split into 4 subgroups:

Subgroup 1: C₉–C₁₄ hydrocarbon solvents;

Subgroup 2: C₉ aromatic solvents;

Subgroup 3: C₆–C₉ aliphatic solvents; and

Subgroup 4:  No consumer use identified solvents.

Each subgroup is assessed separately in the subsections below.  

A critical health effect for the initial categorization of the LBPNs was genotoxicity or carcinogenicity, according primarily to classifications by international agencies. However, the genotoxicity or carcinogenicity classification is only associated with the presence of benzene. Since the classified LBPNs in this assessment contain less than 0.1% benzene according to available data, and are not associated with genotoxicity or carcinogenicity concerns, the classifications are not considered applicable and other critical effects were taken into consideration (EC 2001).

The human health assessment took into consideration those groups of individuals within the Canadian population who, due to greater susceptibility or greater exposure, may be more vulnerable to experiencing adverse health effects. In addition, the potential for elevated exposure for people living near non-petroleum industrial facilities that may release some of these substances was considered in the assessment. The potential for cumulative effects was considered in this assessment, where applicable, by examining cumulative exposures within the subgroups of petroleum LBPNs.

8.1 Subgroup 1 (C₉–C₁₄ Hydrocarbon Solvents)

8.1.1 Exposure assessment

8.1.1.1 Environmental media and food

No Canadian or international data on the levels of C₉–C₁₄ hydrocarbon solvents in indoor or outdoor air, water, dust, or soil, were identified through searches of available literature. While variable in composition, these substances have a high vapour pressure (up to 830 Pa) and high Henry’s Law constant values (-0.5 to +0.5 [log₁₀ atm·m³/mol]), which suggests that they are likely to enter ambient air if released and have the potential to rapidly volatilize from surface waters (OECD 2010c, 2011, 2012b, 2012c).

Information received in response to a CEPA section 71 survey (Environment Canada 2012b) and data from the NPRI (2020) indicate that certain C₉–C₁₄ hydrocarbon solvents are released to air from facilities that report using these substances (Table 4‑2). Therefore, there may be potential for inhalation exposure to populations residing near these facilities. SCREEN3, a tier-one air dispersion model developed by the US EPA (SCREEN3 1996), was used to estimate the potential outdoor ambient air concentration of the C₉–C₁₄ hydrocarbon solvent substances from releasing facilities.

The largest quantity of a substance released to air according to information received in response to a CEPA section 71 survey by single facility for all of subgroup 1 substances was 100 000 kg (Environment Canada 2012b) and the largest reported quantity of release to air for the subgroup 1 substances reported to the NPRI was 141 000 kg for CAS RN 64742-88-7 in the 2017 reporting year (NPRI 2019). Using the maximum quantity (141 000 kg) and the SCREEN3 model with exposure factors given in Table D-1 of Appendix D, the worst-case maximum yearly ambient air concentration for the subgroup 1 substances in the C₉–C₁₄ hydrocarbon solvents category resulting from facility releases to air was estimated to be 0.71 mg/m³. This concentration represents a systemic dose of 0.14 (adults 19+) to 0.51 (infant 1 yr) mg/kg bw/day, assuming complete inhalation absorption. The resultant air concentration is expected to capture estimated air concentrations of substances in the C₉–C₁₄ hydrocarbon solvents group and is considered conservative as this concentration was used to estimate intake from both indoor and outdoor air for the determination of risk.

Limited data were available on releases to water for certain substances in subgroup 1 on the basis of information received in response to a CEPA section 71 survey (Environment Canada 2012b). The largest quantity reported by a single facility that reported releasing any of the LBPN substances to water was 10 000 kg. A conservative estimate of concentrations (the predicted environmental concentrations or PECs) for LBPNs in surface water was estimated using the 95^th percentile concentrations for LBPNs in surface water from releases to wastewater generated by industrial and down-the-drain sources and calculated using the estimates presented in section 7.2.2. These values are considered conservative as they are estimates of high-end concentrations at the point of release into water. The 95^th percentile concentrations from industrial releases and down-the-drain sources are 36 μg/L and 12 μg/L, respectively. Estimates of potential drinking water intakes were generated using these calculated concentrations, the Health Canada water consumption tables (Health Canada 2017, 2018), and information on predicted wastewater treatment system removal rates (see Table 7‑7 of section 7.2.1). The predicted drinking water intakes for the LBPNs resulting from potential industrial releases and down-the-drain sources ranged from 0.00077 mg/kg bw/day (adults) to 0.0012 mg/kg bw/day (infant 1 yr). These values, which assume that the concentrations in treated wastewater released to receiving water bodies are similar to that of drinking water are conservative and are expected to represent potential worst-case releases of LBPN substances in all subgroups (Environment Canada 2012b).

Some of the LBPN subgroup 1 substances (CAS RNs 64741-65-7, 64742-48-9, 64742-82-1, 64742-88-7 and 8052-41-3) may be used as components in the manufacture of food packaging materials that come into direct contact with food; however, dietary exposures resulting from these uses are considered negligible (personal communication, email from the HC Food Directorate to the HC ESRAB, dated November 2017, February 2018 and February 2020; unreferenced). Other uses of LBPN subgroup 1 substances in the manufacture of food packaging materials do not involve direct food contact; therefore, dietary exposure from these uses is not expected. Some of the LBPN subgroup 1 substances (CAS RNs 64741-65-7, 64742-48-9, 64742-82-1, 64742-88-7, 8032-32-4, and 8052-41-3) may also be used as components in incidental additives (for example, lubricants and/or cleaners) used in food processing establishments; however, dietary exposure, if any, would be considered to be negligible (personal communication, email from the HC Food Directorate to the HC ESRAB, dated February 2018 and February 2020; unreferenced).

8.1.1.2 Products available to consumers

Substances in subgroup 1 were identified as ingredients in a large number of products available to consumers in Canada. With the exception of some cosmetic products, the majority identified were DIY products (see section 4 on Sources and Uses). Since no products that may result in oral exposure to substances in this group were identified, the exposure assessment from products available to consumers focuses on the dermal and inhalation routes.

To evaluate the potential for exposure to substances in subgroup 1, “sentinel” scenarios resulting in the highest level of potential exposure to substances in this LBPN subgroup via the dermal and inhalation routes were selected, taking into consideration frequencies of use, reported concentrations, and availability to Canadians from home use of DIY projects. The vulnerable subpopulation of children and women of reproductive age (that is, teens and adult women) and the sub-population living near facilities were considered in the exposure calculations, health effects assessment, and risk characterization of subgroup 1 substances. The exposure values were selected on the basis of either the age group considered most relevant to the scenario or the relevant sensitive subpopulation and/or were on the basis of the age group that had the highest exposure as determined in Appendix D.

Table 8-1 describes the range of estimated dermal and inhalation exposures to subgroup 1 substances from cosmetics and products available to consumers. With the exception of the first 4 cosmetic products that may have daily uses, exposures are expected to be of short duration and to occur infrequently. As there may be numerous products with different subgroup 1 substances in each category, and because in Table 8‑1 a specific CAS RN is not assigned to each product category, the percentage of LPBN listed in this table is the highest for the subgroup 1 substances in each category. The exposure factors used for these calculations, along with links to the specific Safety Data Sheets used to extract LBPN percent in the whole product are given in Tables D-2 to D-5 of Appendix D. For the dermal exposures in Table 8‑1, direct exposure of the product to skin was calculated as this exposure is compared with a dermal point of departure (POD), adjustment for dermal absorption of the substance is not required.

^a  The exposures identified are for adults unless otherwise noted
^b Internal dose (mg/kg bw/day) = mean concentration on day of exposure [mg/m³] × Inhalation rate [m³/day] / body weight [kg]
^c For ventilation rate of 2.5 hr^-1.
^d For ventilation rate of 0.5 hr^-1.

In the exposure calculations for cosmetic products not meant to be directly applied to the skin, retention factors (typically 0.1) were employed to relate the total amount of product used to the amount of product residing on the skin. In these cases, the amount residing on the skin is from product contact with both hands and the head / scalp.

8.1.2 Health effects assessment

8.1.2.1 International assessments

C₉–C₁₄ aliphatic hydrocarbons solvents have multiple synonyms such as white spirits, mineral spirits, mineral turpentine, petroleum spirit, solvent naphtha, and Stoddard solvent (WHO 1996). White spirit comprises different mixtures of hydrocarbons, mainly C₈–C₁₂ aliphatic, alicyclic, and aromatic hydrocarbons, with an aromatics content ranging from ≤1% to ~25% w/w (WHO 1996). However, the benzene and toluene contents are low (~0.001% w/w) (WHO 1996). The boiling range varies from 130°C to 220°C, for example Stoddard solvent, which has a range of 149°C to 208°C (WHO 1996).

The Scientific Committee on Occupational Exposure Limits (SCOEL), Committee for Risk Assessment of the European Chemicals Agency (ECHA), the Danish Environmental Protection Agency (Danish EPA), and the World Health Organization/Internationally Peer Reviewed Chemical Safety Information/International Programme on Chemical Safety (WHO/INCHEM/IPCS) concluded that there is a strong association between exposure to white spirit containing a high level of aromatics (15% to 20%) and long-term adverse effects on the central nervous system in male workers (WHO 1996; SCOEL 2007; ECHA 2011a; Danish EPA 2014). On the basis of available animal and human studies, SCOEL recommended an eight-hour time weighted average Occupational Exposure Level (OEL) of 116 mg/m³ under specific protective conditions (that is presence of ventilation or wearing a mask) and by applying an uncertainty factor of 2 on the NOAEL of 232 mg/m³ determined by Mikkelsen et al. (1988) (WHO 1996; SCOEL 2007; ECHA 2011a; Danish EPA 2014). The OEL covers white spirit with different aromatic contents, dearomatized white spirit, and white spirit with various aliphatics (SCOEL 2007). SCOEL determined that no major differences were found between the neurotoxic patterns associated with aromatized and dearomatized white spirits in the animal studies. However, there was only limited information on the effects of dearomatized white spirits on humans (SCOEL 2007). ECHA mentioned that the effects observed in humans are mainly related to neurobehavioral changes and are difficult to detect in laboratory animals (ECHA 2011a). SCOEL and ECHA acknowledged the importance of considering major functional changes in the central or peripheral nervous system (SCOEL 2007; ECHA 2011a). ECHA, SCOEL, the Danish EPA and WHO concluded that long-term exposure to white spirit may lead to the impairment of brain function and can therefore be associated with a high risk for the development of a chronic toxic encephalopathy (WHO 1996; SCOEL 2007; ECHA 2011a; Danish EPA 2014). White spirits and other hydrocarbons (toluene, xylene, styrene, and pentane) were listed as causative agents for encephalopathies because of organic solvent exposure (ECHA 2011a).

Indoor air guidelines for hydrocarbon solvents / white spirits (CAS RNs 64741-65-7, 64742-48-9, and 64742-88-7) have been established by the German Federal Environmental Protection Agency using the developmental neurotoxicity study from Hass et al. (2001) as the key study (Sagunski and Mangelsdorf 2005). A limit value of 0.2 mg/m³ was determined to prevent developmental neurotoxic effects (Sagunski and Mangelsdorf 2005).

8.1.2.2 Toxicokinetics

The accumulation of white spirits in adipose tissue was demonstrated by several studies in humans (WHO 1996; SCOEL 2007; ECHA 2011b). The redistribution phase of dearomatized white spirit in adipose tissue was estimated to be approximately 20 hours and the half-life of white spirit in adipose tissue was calculated to be between 46 and 48 hours. After 5 consecutive days of exposure to 600 mg/m³ dearomatized white spirit for 6 hours per day, the maximum concentration was approximately 55 mg/kg fat and the minimum was approximately 35 mg/kg fat in 8 male volunteers (ECHA 2011b). Because the brain is composed of nearly 60% fat, fat tissue and brain tissue are the first tissues to be reached by white spirits (SCOEL 2007; ECHA 2011b). The distribution and accumulation of white spirits in the brain is considered relevant for the adverse effects on the central nervous system.

Since white spirits are mixtures, analysis of the toxicokinetic properties is generally conducted on the numerous different hydrocarbons contained within them. A study showed that the aliphatic content in blood increased alongside increasing molecular size from n-octane to n-dodecane but the concentration in brain only increased from n-octane to n-decane, thereafter declining from n-decane to n-dodecane in rats exposed to 9 different C₈–C12 hydrocarbons at 100 ppm, 12 hours a day for 3 days (WHO 1996). When the aliphatic, alicyclic and aromatic hydrocarbons were compared, it was noted that aromatics produce the highest concentration in blood but were found to have the lowest concentration in brain. For alicyclic and aliphatic hydrocarbons, lower values in blood and higher values in brain were detected, particularly for the alicyclic hydrocarbons (WHO 1996; ECHA 2011b). For the n-alkanes, accumulation in fat occured during the three-day exposure period while for the aromatics, the content in fat peaked on day 1 and decreased remarkably following the next 2 days of exposure. Overall, the alicyclics are most extensively distributed from blood to other tissues (WHO 1996; ECHA 2011b).

8.1.2.3 Neurotoxicity

Aromatics (2% to 20%)

Male Wistar rats (8/dose) were exposed to 0, 1200, 2400, or 4800 mg/m³ (equivalent to 0, 351.2, 702.5, or 1405 mg/kg bw/day) aromatized white spirit (18% aromatics) vapours for 8h per day, 5 days per week for 17 weeks, while another group was exposed for 26 weeks (ECHA  2020a). For the 17-week exposure duration, authors of the study observed clinical signs and body weight changes but did not provide any details. Behavioural tests designed to measure changes in activity coordination, grip strength, and discrimination performance did not reveal significant differences compared to control rats. However, an increase in response time to a light stimulus, and a decrease in tail nerve conduction velocity were observed at the highest dose. For the 26-week exposure, rats had a 10 h recovery after the last exposure. In addition, no exposure-related changes in histopathological analysis in brain, spinal cord, or sciatic nerve were observed.  The authors of the study did not observe any differences after the recovery period in the same tests between controls and treated animals. A NOAEC of 4800 mg/m³ (equivalent to 1405 mg/kg bw/day), the highest dose tested, was determined by the authors of the study on the basis of the absence of persistent changes in neurobehavioral functions. No further details are provided (ECHA 2020a).

Young (3-month old) or old (15-month old) Male Wistar rats (12-4/dose/age) were exposed to 0, 2290, or 4580  mg/m³ (equivalent to 0, 472.4or 946 mg/kg bw/day) of white spirit/Stoddard solvent (containing 20% aromatics) vapours for 6 h/day, 5 days per week for 6 months with a 2-month recovery period to assess neurobehavioral tests and a 4-month recovery prior to the neurochemical analysis and brain histopathology (Ostergaard et al. 1993). Body weight was significantly lower in young rats exposed to the high dose, but there was higher consumption of food and water. Old rats in the high dose group showed only an increase in water consumption. No changes (in body weights or food consumption) were observed in any of the treated old rats. Plasma urea and creatine were significantly increased while alanine aminotransferase (ALAT) activity was significantly reduced in both recovery periods and dose-dependently in all exposed rats (young and old). However, concentrations of noradrenaline, dopamine, and 5-hydroxytryptamine in various parts of the brain were increased in exposed rats (young and old) and significant in the highest dose for both recovery periods. No changes were observed in behavioural tests such as motor activity, functional observation battery, passive avoidance, radial arm maze, or Morris maze. No macro- or microscopic changes were observed in brains. However, the authors of the study observed permanent changes in the neurochemistry of the exposed brains (Ostergaard et al. 1993). In the present assessment, a LOAEC of 2290 mg/m³ was determined on the basis of neurotoxicity effects.

In a similar study, male Wistar rats (7-10/dose/time of exposure) were exposed to 0,  2290, or 4580 mg/m³ (equivalent to 0, 502.7, or 1005.4 mg/kg bw/day) of white spirit/Stoddard solvent (containing 20% aromatics) for 6 h/day, 5 days per week for 3 weeks or for 6 months, with a 3-month recovery period (Lam et al. 1995). In the 3-week exposure, animals showed narcotic effects in all doses. In the 6-month exposure, treated animals had lower body weight and a higher consumption of food and water, suggesting that food efficiency was reduced. No changes in brain weight were observed in all exposure groups. The authors of the study observed a reduction in the level of synaptosomal proteins by a Lowry protein assay, while an increase of noradrenaline, dopamine, 5-hydroxytryptamine, and cholinesterase activities in the brain without the cerebellum was observed in all exposed rats (3-week and 6-month exposures). The authors of the study concluded that exposure to white spirit with aromatics induced functional changes in the brain after 3 weeks of exposure and that these changes may contribute to symptoms of chronic toxic encephalopathy (Lam et al. 1995).  In this assessment, a LOAEC of 2290 mg/m³ (502.7 mg/kg bw/day) was determined on the basis of neurotoxic effects.

Male Wistar rats were exposed to 0, 575, 2875, or 5750 mg/m³ (equivalent to 0, 126.2, 631.1, or 1262.3 mg/kg bw/day) white spirit vapour (containing 11.7% aromatics) by inhalation 6h per day, 5 days per week from 4 to 17 weeks (Savolainen and Pfäffli 1982). No clinical signs of toxicity or changes in body weight were observed. A significant dose-dependant decrease was observed in the cerebellar succinate dehydrogenase activity in all treated rats after 8 weeks of exposure while creatine kinase activity increased after 12 weeks in cerebral hemispheres. A decrease in cerebellar glutathione level was seen at the highest dose after 8-weeks of exposure. The authors of the study concluded that exposure to white spirit may have caused chemical effects in the brain and determined a NOAEC of 575 mg/m³ (equivalent to 126.2 mg/kg bw/day) on the basis of the chemical changes in muscles at the next dose following 17 weeks of exposure (Savolainen and Pfäffli 1982).

Dermal (occluded) exposure (on tail) of male Wistar rats (5/dose) to 855 mg/kg bw/day of dearomatized white spirit (0.3% aromatics) or 691 mg/kg bw/day of aromatized white spirit (11.7% aromatics) for 3 h per day, 5 days per week for 6 weeks did not cause changes in body weight gain in treated rats (Verkkala et al. 1984). Systemic doses were calculated using the average quantity of formulation applied on the skin per day and the rat body weight. The skin of the treated area showed keratolysis. A histopathological analysis of the tail showed that axons teased from all rats exposed to both white spirits had prenodal swellings with widening of the nodes of Ranvier. Electric stimulation of the tail showed an increased duration of the motor response, which became polyphasic after 6 weeks in the dearomatized white spirit group. Aromatized white spirits produced similar responses as dearomatized white spirits, but they were less intense. The authors of the study concluded that the physiological changes in the 2 white spirits were comparable to each other despite the difference in the doses, and the removal of aromatic constituents did not supress the neurotoxic effects (Verkkala et al. 1984). In the present assessment, a lowest observed adverse effect level (LOAEL) of 691 mg/kg bw/day was determined for the dermal route.

Dearomatics (≤2%)

Short-term:

In a short -term study, adult male Wistar rats (9-10/dose) were exposed to 0, 2339, or 4679 mg/m³ (equivalent to 0, 513.5, or 1027.2 mg/kg bw/day) dearomatized white spirit (CAS No. 64742-48-9) by inhalation for 6h per day, 7 days per week for 3 weeks (Lam et al. 1995). In the subcellular fractions of hemispheres and in hippocampus, a significant level of cellular glutathione was observed in both dose groups. In the hippocampus, a significant increase of reactive oxygen species (ROS) was observed in the high dose group. In the liver, a significant increase in ROS and a decrease in glutamine synthetase activity was seen at the highest dose. The study authors concluded that cumulative oxidative damage may occur due to exposure to dearomatized white spirit (Lam et al. 1995). Neurons are among the cell types that are most vulnerable to oxidative insults, largely due to their high metabolic rate and low regenerative potential (Cobb and Cole 2015). In the present assessment, a lowest observed effect concentration (LOEC) of 2339 mg/m³ (equivalent to 513.5 mg/kg bw/day) was determined on the basis of biochemical changes in the brain.  

In an in vitro assay, an increase of ROS and reactive nitrogen species (RNS) was significant in a dose-dependent manner in rat brain synaptosome fraction after an exposure of n-nonane (Myrhe and Fonnum 2001).

In another study, male Wistar rats (10/dose/substance) were exposed to 0, 2339, or 4679 mg/m³ (equivalent to 0, 513.5, or 1027.2 mg/kg bw/day) of dearomatized white spirit (0.4% aromatics) or aromatized white spirit (20% aromatics) by inhalation for 6h per day, 7 days per week for 3 weeks (Lam et al. 2001). No changes were observed in animal body weight. A decrease of 5-hydroxytryptamine transporter (5-HTT) and 5-HT_2A receptor in the forebrain, exposed to both white spirits at the highest dose, was observed as well as a decrease of 5-HT4 at the highest dose of aromatic white spirit only. The authors of the study concluded that inhalation exposure to white spirit caused neurotoxicity at 2339 and 4679 mg/m³, especially the aromatics (Lam et al. 2001).

In a short-term study, male Wistar rats (10/dose/substance) were exposed to 0, 2339, or 4679 mg/m³ (equivalent to 0, 513.5, or 1027.2 mg/kg bw/day) dearomatized white spirit (0.4% aromatics) or aromatized white spirit (20% aromatics) by inhalation for 6h per day, 7 days per week for 4 weeks (Lam et al. 2000). The glial fibrillary acidic protein (GFAP, known to be released following neural injury) was measured after 7-day or 4-weeks of exposure. No clinical signs of toxicity were observed following exposure. With dearomatized white spirit, the expression of GFAP increased in the cerebellum during the first week but not after 4 weeks. In contrast, an increase of GFAP was observed in cerebellum, and medulla oblongata in rats exposed to both doses of aromatized white spirit and in the cerebral cortex and thalamus for the highest dose. The authors of the study concluded that the data are indicative of an aromatic white spirit-induced gliosis in several regions of the rat central nervous system and suggested that short-term exposure to this solvent may be associated with underlying neural damage (Lam et al. 2000). A change in the expression of GFAP may be a more sensitive indicator of neurotoxicity than changes detected by standard histopathology (US EPA 1998). Based the weight of evidence from these results and those from previous studies, in the present assessment, a LOAEC of 2339 mg/m³ (equivalent to 513.5 mg/kg bw/day) was determined for dearomatized white spirit on the basis of adverse modifications in neurochemical changes in the rat brain.

Long-term:

Male Wistar rats (9-10/dose) were exposed to 0, 2339, or 4679 mg/m³ (equivalent to 0, 513.5, or 1027.2 mg/kg bw/day adjusted for 5 days/week) of dearomatized white spirit (CAS No. 64742-48-9) by inhalation for 6h per day, 5 days per week for 6 months (Lund et al. 1996). After a recovery period of 3 months, some rats from each dose were analyzed for neurophysiological, neurobehavioral, and histopathologic changes. An electrode was implanted in the skull of the rats to record visual flash potentials, the somatosensory evoked potentials, and the auditory brain stem response. The authors observed a dose-dependent increase in response in all of these sensory tests. A decrease in motor activity assessed by measuring ambulation in the cage, was observed. However, no effects were observed in learning and memory functions measured by functional observation battery, passive avoidance test, Morris water maze, and radial arm maze test 3 months after the cessation of exposure (Lund et al. 1996). No other clinical effects were observed. The authors concluded that all doses directly affected the central nervous system of exposed rats (Lund et al. 1996). In the present assessment, a LOAEC of 2339 mg/m3 (equivalent to 513.5 mg/kg bw/day) was estimated.

8.1.2.4 Developmental neurotoxicity

Due to the lack of available data on reproductive toxicity, developmental immunotoxicity, developmental neurotoxicity, and endocrine disruption (evidence on hormonal changes), ECHA has recently made proposals for current testing to assess those endpoints for several hydrocarbons, C₉–C₁₄ aromatics (ECHA 2021a).

In an in utero study, pregnant Wistar rats (6/dose) were exposed to 0, 2339, or 4679 mg/m³ (equivalent to 0, 513.5, or 1027.2 mg/kg bw/day) dearomatized white spirit (CAS No. 64742-48-9) for 6h per day from gestational day 7 to 20 by inhalation (Edelfors et al. 1999). Only female offspring were analyzed. No change was observed in exposed offspring body weight at postnatal 35 days. No information was given by authors about dams. At postnatal 35 days, an analysis of the brain from the female offspring showed that cytosolic calcium was increased in synaptosomes, the terminal part of the neuron, in both exposed groups. The increase of intracellular calcium is a second messenger system to induce “programmed cell death” (WHO 2001, Marambaud et al. 2009). The increase of intraneuronal calcium is known as a key signal for neurodegeneration (Marambaud et al. 2009) and it acts as a neurochemical endpoint to classify a chemical as neurotoxic (US EPA 1998).  

In a similar in utero study, pregnant Wistar rats (6/dose) were exposed to 0 or 4679 mg/m³ (equivalent to 1115.7 mg/kg bw/day) dearomatized white spirit (CAS No. 64742-48-9) for 6h per day on gestation days 7 to 20 by inhalation (Hass et al. 2001).  Maternal body weight (-8%) and body weight gain (-26%) were lower in exposed dams but the offspring weight (+7%) was increased. The gestation length, the frequency of post-implantation loss, sex distribution in the litters, the frequency of neonatal death, and offspring body weights were similar in the 2 groups. No changes were recorded in motor function or in the activity in the Open Field. At the age of 2 months, exposed male offspring showed a significant reduction in memory and exposed females showed cognitive function impairment measured by Morris water maze test. At 5 months of age, learning and memory deficits were observed in both exposed sexes. The authors of the study concluded that a developmental exposure of 4679 mg/ m³ (1115.7 mg/kg bw/day) dearomatized white spirit may induce long-term learning and memory deficits in male and female rats (Hass et al. 2001). In the present assessment, a LOAEC of 4679 mg/m³ (equivalent to 1115.7 mg/kg bw/day) was determined on the basis of developmental neurotoxicity.   

8.1.2.5 Epidemiological data

Short-term:

A group of volunteers (unknown gender; 6/dose) between 25 and 59 years of age were exposed to 0, 132, 825, or 2200 mg/m³ (equivalent to 0, 13.9, 86.9, or 231.6 mg/kg bw/day) of Stoddard Solvent for 15 min (Carpenter et al. 1975a,b). No effect was noticed at 132 mg/m³, except for olfactory fatigue. At 825 mg/m³ one volunteer displayed eye irritation and 6 volunteers also reported olfactory fatigue. At 2200 mg/m³, most of the volunteers experienced olfactory fatigue, throat and eye irritation, tears, slight injection of sclera, dizziness, and tasting the solvent due to its strong odor. All volunteers reported that effects disappeared 15 min after leaving the exposure chamber. The authors concluded that a dose of 2200 mg/m³ was not acceptable for workplace exposure (meaning 8 h of exposure per day) (Carpenter et al. 1975a).

In a short-term study, 8 healthy volunteers (4 women and 4 men) were exposed twice in an exposure chamber to increasing concentrations of 0.5 mg/m³ to 600 mg/m³ (equivalent to approximately to 0.051 mg/kg bw/day to 63.2 mg/kg bw/day) of white spirit containing 15% to 20% aromatics or dearomatized white spirit in eight 10-min steps, with a break of one week between 2 sessions (Ernstgard et al. 2008a, b). Volunteers then rated their symptoms related to each exposure. The authors of the study could not determine a NOAEC or LOAEC because their rating system did not show a dose-response. The study authors acknowledged that their results showed no effect at near 50 mg/m³ and below for both white spirits.

In a short-term study, volunteers (6 women and 6 men between 20 and 38 years of age) were exposed in a pseudo-randomized within-subject crossover design to 0, 100, or 300 mg/m³ (equivalent to 0, 10.5, or 31.6 mg/kg bw/day) of dearomatized white spirit (containing less than 0.002% aromatics) or aromatized white spirit (containing 19% of aromatics) for 4 h per day, one day per week, with one week between each dose (Juran et al. 2014). The authors observed that effects were difficult to interpret due to a large variability in the volunteer responses in the neurobehavioral tests and due to a lack of dose-response relationship (Juran et al. 2014). Since all volunteers were exposed to all substances doses, the results may reflect the effects of exposure to more than one substance or may not reflect one acute exposure but rather the results of multiple exposures across time.  

In another study, 12 healthy male volunteers were exposed for 4 h to either 0, 57, or 570 mg/m³ (equivalent approximately to 0, 6, or 60 mg/kg bw/day) aromatized white spirit (21.3% aromatics, CAS 64742-82-1) in 2 sessions spaced 7 days apart (Lammers et al. 2007). Neurobiological tests such as psychomotor skills, or reaction time test showed a significant negative trend effect during exposure and after exposure at 570 mg/m³. The authors of the study compared the results on volunteers with the results from a 3 day-exposure, with one day of recovery, of 8 h per day to 0, 600, 2400, or 4800 mg/m³ (equivalent to 0, 183.9, 735.2, or 1474.6 mg/kg bw/day) of white spirit (14% to 20% aromatics) on male and female rats performed in parallel with the epidemiological study (Lammers et al. 2007; ECHA 2020a). The authors did not observe any effect at 600 mg/m³ in rats. There was a significant increase in response time to stimuli at 2400 and 4800 mg/m³ in rats, but animals recovered after a day without exposure. The authors of the study determined a no-observed effect level (NOEL) of 600 mg/m³ (12.7mg/kg bw/day) for the short term on the basis of the neurotoxic effects in rats at 2400 mg/m³ (735.2 mg/kg bw/day) and above (Lammers et al. 2007; ECHA 2020a). However, because they saw similar significant neurotoxic effects in humans at 570 mg/m³ and in rats at 2400 mg/m³, the authors of the study suggested that humans may be slightly more sensitive than rats to the effects of hydrocarbon solvents on the central nervous system or that more sensitive tests were used with human volunteers. 

A study using Physiologically Based Pharmacokinetic (PBPK) models concluded that an exposure of white spirit would produce brain effects similar to those seen in the literature described in this assessment for humans and rats. The authors of the study confirmed a NOEL of 600 mg/m³ for short exposure of white spirit (Hissink et al. 2007).

Using data from the Institute of Occupational Health of Finland, an epidemiological study on occupational pregnant women evaluated the impact of exposure to different solvents such as white spirits (described as hydrocarbons aliphatic containing 0% to 15% aromatics) on spontaneous abortion during the first term of pregnancy (Lindbohm et al. 1990). The authors used a case-control design where women with spontaneous abortion were defined as cases, and women who gave birth were the controls. The controls were matched with the case for age at the time of conception within 2.5 years, using the nearest available matching. Women were classified by their work tasks including painting, lacquering, gluing, printing, rubber work, plastics work, degreasing, lamination, or dry cleaning. Data on the women’s pregnancy and work histories, chronic and acute diseases, and smoking and use of alcohol during the first trimester of pregnancy were also sought. Exposure was defined as high if the worked handled the white spirits daily or 1-4 days a week and the level of exposure, measured by industrial hygiene, was high. Exposure was defined as low if the worker handled solvents 1 to 4 days a week, and the level of exposure was low or if the workers handled solvents less than once a week. The odds ratios for exposure were estimated using the logistic regression model for individually matched data, on the basis of the conditional likelihood function. A significantly higher odds ratio of spontaneous abortion for high levels of aliphatic hydrocarbon exposure was observed in comparison to the low level group and other solvents such as toluene. The odds ratio of spontaneous abortion was significantly higher among graphic workers (7 women exposed and 3 controls) and painting and lacquering workers (3 women exposed and 3 controls) but not for workers with other work tasks (3 women exposed and 7 controls). However, the odds ratio was significantly higher among graphic workers than among painting and lacquering workers. Because graphic workers used white spirits for cleaning printing machines and as a diluent for printing ink, they were exposed to high levels (212 ppm to 1,167 ppm) for a shorter period than painting and lacquering workers were. The control graphic workers, who were exposed to white spirits at a lower level (4 ppm to 68 ppm) for the same amount of time, did not have spontaneous abortions. The authors of the study indicated that any combined solvent effects cannot be excluded given the multiple exposures to different solvents and that the statistical power of the study was fairly low due to the small size of the study population and the low prevalence of exposure.

In a similar epidemiological study, the association between solvent exposure and spontaneous abortion was examined in an interview study with 1926 women, approximately one-third of whom had experienced spontaneous abortion during the first 20 weeks of gestation (first term of pregnancy) (Windham et al. 1991). The women were questioned about the different types of solvents to which they were exposed, the duration of exposure during pregnancy and the exposure intensity. Women using paint thinners at work had a significantly higher level of spontaneous abortion in comparison with those exposed to other products such as paint strippers or xylene. A significant association with paint thinners and paints containing aliphatic hydrocarbons was observed by the authors of the study. However, no relationship with the duration of exposure (less than 10h per week or more than 10h per week) was found by the authors of the study. The authors determined that the aliphatic hydrocarbons were generally the most used solvent at home (non-occupational use) among the women who experienced spontaneous abortion. The authors concluded that the duration of exposure to aliphatic hydrocarbons did not seem to have an impact on the association with spontaneous abortions during the first term of pregnancy.     

Long-term exposure:

In a cross-sectional study, 236 male painters and 128 male non-exposed joiners were examined using a questionnaire, and 44 male painters and 44 male non-exposed joiners underwent neurological examinations (nerve conduction measurements on the right hand) and neuropsychological tests (Cherry et al. 1985). The average levels of white spirit were determined to be 125 or 578 mg/m³ (equivalent approximately to 9.4 or 43.5 mg/kg bw/day). However, painters were also exposed to other solvents such as trichloroethylene, dichloromethane, methyl n-butyl ketone, or n-butanol during their work. The authors of the study did not remove or incorporate the interaction with the other solvents. The mean exposure period was 11.7 years for painting. Painters significantly and more often reported tingling in the hands and feet without evidence of impaired nerve conduction in the hands, depression, difficulties in concentration, and increased irritability (Cherry et al. 1985).

In a cross-sectional study, 28 male painters exposed to organic solvents reported a reduction of vibration perception thresholds during a mean exposure period of 30 years in comparison with 20 male non-exposed boilermakers (control group) (Demers et al. 1991).  

In a cross-sectional study, the effects of a chronic exposure to white spirit were studied with a random sample of 85 male painters and 85 male bricklayers as a non-exposed group (Mikkelsen et al. 1988). Possible confounders such as age, alcohol intake, or education were identified and taken into account in the analyses (Mikkelsen et al. 1988). The solvents used in the paints were mainly white spirit containing approximately 15% to 20% aromatic hydrocarbons, and 80% to 85% aliphatic hydrocarbons (Mikkelsen et al. 1988). The median years of occupation as a painter was 31 years, and the median of the solvent exposure was 25 years. The authors of the study stated that the risk of developing dementia was associated with long-term solvent exposure. The estimated odds ratio for painters with medium solvent exposure (15 to 30 years) was 3.6. For painters with high solvent exposure (>30 years) to the same dose of solvent, the estimated odds ration was 5.0. The prevalence for painters with a low exposure level (<15 years) was the same as for bricklayers. In psychometric tests, painters with high and medium solvent exposure performed more poorly than painters with low solvent exposure and more poorly than bricklayers in almost all of the tests. The authors of the study noticed a significant increase in abnormal movement coordination in painters with medium and high exposure. Results from the questionnaire and the Computed Axial Tomography of the brain showed an increase in dementia alongside an increase in solvent exposure. On the basis of the previous results, painters exposed to low levels of solvent did not differ significantly from bricklayers; therefore they have little to no risk of dementia in comparison with the medium and high solvent exposure groups (Mikkelsen et al. 1988). Basing their conclusion on the results, the authors of the study determined an average of 232 mg/m³ (equivalent approximately to 17.45mg/kg bw/day) white spirit containing between 15% to 20% of aromatics as a NOAEL for 13 years of exposure, which was confirmed by SCOEL and ECHA (SCOEL 2008; ECHA 2011b).  

In another cohort study, neuropsychiatric effects of white spirits were studied in 135 male painters and 71 male carpenters (control group) for at least 10 years (Lundberg et al. 1995). The solvent exposure time was evaluated using a questionnaire. Neuropsychiatric symptoms compatible with Chronic Toxic Encephalopathy (CTE) were more common among the painters than among the carpenters, and these symptoms became increasingly prevalent as the duration of solvent exposure increased. In the majority of psychometric tests, all of the painters showed worse results than the carpenters. No differences between painters and carpenters in psychiatric evaluations were observed. The authors of the study determined that an exposure of 540 mg/m³ (approximately 40.6 mg/kg bw/day) was related to an elevated risk of symptoms associated with chronic toxic encephalopathy (Lundberg et al. 1995). From an evaluation of the paper, SCOEL and ECHA determined a LOAEL for long-term effects to be approximately 540 mg/m³ white spirit (SCOEL 2008; ECHA 2011b).

An epidemiological study examined the association between prenatal exposure to organic solvents and a decrease in neurobehavioral performance in children between 3 and 7 years old (Till et al. 2001). The study contained 33 children born to mothers who were occupationally exposed to organic solvents during pregnancy and 28 control children born to non-exposed women. The various organic solvents such as aromatic and aliphatic hydrocarbons, halogenated compounds, or polyaromatic hydrocarbons were determined by a questionnaire and MSDS. Children in the solvent-exposed group showed deficits in selective cognitive abilities testing when performing tasks of Graphomotor ability (that is NEPSY Design Copy and Visuo-motor Precision tasks). Receptive and Expressive language abilities testing by NEPSY Body Part Naming or Speeded Naming were also affected in the exposed group, whereas there was no variability in Attention, Fine-motor ability, and Visuo-spatial ability between groups. The authors of the study suggested that increasing exposure intensity during pregnancy is associated with moderate impairment in selective cognitive tasks. The limitation of this study is that the level of chemical inhaled is unknown. It was therefore impossible to isolate the various organic solvents and examine the effects of each separately. The authors of the study concluded that on the basis of the increase in impairments in the neurodevelopment of prenatal exposed children associated with organic solvents, exposure should be closely monitored and women should be advised to minimize their exposure during pregnancy, especially during early pregnancy to prevent effects during critical periods of organogenesis.

8.1.2.6 Repeated doses toxicity

Aromatics (2% to 20%)

Short-term (2-week) inhalation exposure to 0, 138, 275, 550, 1100 or 2200 mg/m³ (6 h/day, 5 days/week) of Stoddard solvent did not cause clinical signs of toxicity in adult male and female rats and mice (n = 5/sex/dose). However, liver appeared as a target organ in both species since a significant increase was observed in absolute and relative liver weight in male rats in 550 mg/m³ (equivalent to 150.4 mg/kg bw/day) or higher and in female rats in 275 mg/m³ (equivalent to 78.1mg/kg bw/day) or higher dose groups. In addition, an increase in liver weight was seen in male or female mice in 275 mg/m³ (equivalent to 100.1 or 102.7 mg/kg bw/day for males and females, respectively) or higher dose groups (NTP 2004).

Likewise, subchronic (3-month) inhalation exposure to 0, 138, 275, 550, 1100 or 2200 mg/m³ (6 h/day, 5 days/week) of Stoddard solvent did not produce clinical signs of toxicity in (n=10/sex/dose) male or female rats and mice. In rats, an increase in relative liver and kidney weight and absolute and relative testis weight accompanied by a decrease in sperm motility was seen in male rats exposed to 550 mg/m³ (equivalent to 134.22 mg/kg bw/day) or higher. There was an increase in the incidence of goblet cell hypertrophy of the nasal respiratory epithelium in male and female rats exposed to 2200 mg/m³ (equivalent to 542.5 mg/kg bw/day for males and589.7mg/kg bw/day for females) of Stoddard solvent. Male mice exposed to the same doses showed an increase in liver weight and decrease in sperm motility at 2200 mg/m³ (equivalent to 770.2 mg/kg bw/day) dose but without any effect on fertility (NTP 2004).

Wistar rats (5/sex/dose) were exposed by inhalation to 0, 2000, 4000, or 7500 mg/m³ (equivalent approximately to 0, 418.5, 838.7, or 1583 mg/kg bw/day for males and 0, 462.2, 926, or 1745.8 mg/kg bw/day for females) aromatized white spirit (19% aromatics) for 6 h/day, 5 days/week for 13 weeks (Carrillo et al. 2014).  Body weights were significantly lower in males and females at the highest dose. An increase of relative spleen weight in males at 4000 mg/m3 and above was observed. In males, a decrease in red blood cells (from -5% to -8%) and packed cell volume (from -3% to -5%) in a dose response manner and an increase in mean cell volume (+3%) and mean corpuscular hemoglobin (+6%) at all doses were observed. Females showed a significant increase in white blood cells values in the mid- and high-dose groups and in mean cell volume only in the highest dose group. The authors of the study concluded that these effects are not considered relevant because they are within historical limits (Carrillo et al. 2014). Spleen observations showed an increase in erythropoietic activity and hemosiderin deposition at mid and high doses in males. In females, there was an increase in deposition of hemosiderin and extra-medullary hematopoiesis at the highest dose. The authors of the study concluded that this was consistent with compensatory processes. On the basis of the body weight reduction at the highest dose, the authors of the study determined a NOAEC of 4000 mg/m³ (926 mg/kg bw/day for females) (Carrillo et al. 2014). Without a recovery period, it is difficult to confirm whether or not effects in blood and spleen are adaptive. On the basis of the modification in blood parameters and spleen effects at 4000 mg/m³ in males, a NOAEC of 2000 mg/m³ (462.2 mg/kg bw/day) was determined for this assessment.

Male rats were exposed to 0, 480, 1100, or 1900 mg/m³ (equivalent approximately to 0, 103.1, 233.9, or 407.3 mg/kg bw/day) Stoddard solvent (14.7% aromatics) by inhalation for 6h per day, 5 days per week for 13 weeks (Carpenter et al. 1975). No clinical sign of toxicity or adverse effects were observed in treated animals. The neurologic system was not tested. On the basis of these results, a NOAEC of 1900 mg/m³ (407.3 mg/kg bw/day), the highest dose tested was determined in this assessment.

Male Wistar rats were exposed to 0, 49, 100, or 230 mg/m³ (equivalent approximately to 0, 10.5, 21.5, or 49 mg/kg bw/day) Stoddard solvent (14.7% aromatics) by inhalation for 6h per day, 5 days per week for 13 weeks (ECHA 2020a). No clinical sign of toxicity or adverse effects were observed. The neurologic system was not tested. A NOAEC of 230 mg/m³ (49 mg/kg bw/day), the highest dose tested, was determined by the authors of the study (ECHA 2020a).

Sprague Dawley rats (5/sex/dose) were administered 0, 116, 347, or 1056 mg/kg bw/day of hydrocarbons, C11-C₁₄, n-alkanes, isoalkanes, cyclics, aromatics (2% to 25%) (CAS 64742-81-0) by gavage for 28 days (ECHA 2020a). No mortality or clinical sign of toxicity or adverse effects were observed. The neurologic system was not tested. The NOAEL was determined to be 1056 mg/kg bw/day, the highest dose tested by the authors of the study (ECHA 2020a).

New Zealand rabbits (10/sex/dose) were administered 0, 200, 1000, or 2000 mg/kg bw/day Stoddard solvent (CAS 8052-41-3) by occluded dermal application for 6 h per day, 3 days per week for 4 weeks (ECHA 2020a). At 2000 mg/kg bw/day, 3 female rabbits developed liver lesions characterized as white streaks or foci with granular surface. These observations were not explained by the authors. Severe and moderate dermal irritation was observed in all treated animals. At the highest dose, there was a significant reduction in weight gain in both sexes, whereas only female body weight gain was reduced at 1000 mg/kg bw/day (ECHA 2020a). The authors of the study observed changes in hematological parameters as well as in albumin and globulin at 2000 mg/kg bw/day; however, they judged them not to be treatment-related, attributing them instead to acute dermal inflammation (ECHA 2020a). The majority of lesions noted upon gross examination at necropsy were seen in the skin and were associated with dermal irritation. Microscopic examination of treated animals revealed lesions in the skin at the application site including thickening and down-growth of the epidermis, hyperkeratosis, and dermal fibrosis. The incidence and severity of the observed lesions were significantly greater in high dose animals compared with controls and animals in the lower dose groups. The authors of the study noticed a small number of lesions on other tissues such as heart, trachea, pancreas, testes, and spleen in all treated groups but did not consider these to be related to the treatment, without giving further details. The neurologic system was not tested. Since the observed effected were not related to treatment, the authors of the study determined a NOAEL of 2000 mg/kg bw/day which was the highest dose (ECHA 2020a). Since the liver effects in females were not explained by the authors and could be related to the treatment, a NOAEL of 1000 mg/kg bw/day is estimated in this assessment.

In a developmental study, female rats (25-26/dose) were exposed to 0, 600, or 2400 mg/m³ (equivalent to 0, 131.7, or 526.9 mg/kg bw/day) white spirit (Stoddard solvent 24% aromatics) by inhalation for 6h per day, on days 6 to 15 of gestation (ECHA 2020a). No maternal or developmental effects were observed. The authors of the study did not test for developmental neurotoxicity. A NOAEC for developmental toxicity of 2400 mg/m³ (equivalent to 526.9 mg/kg bw/day), the highest dose tested, was determined by the authors of the study (ECHA 2020a).

Dearomatics (≤2%)

Sprague Dawley male and female rats (10-15/dose/sex) were exposed to 0, 197, or 561 mg/m³ (equivalent to 0,43.2, or 123.2 mg/kg bw/day for females and 0, 43.7, or 125.3 mg/kg bw/day for males) dearomatized white spirit (containing <0.5% aromatics) by inhalation for 6 h per day, 5 days per week for 12 weeks (Phillips and Egan 1984). No clinical adverse effects were observed at levels tested by the authors of the study (Phillips and Egan 1984). The neurologic system was not tested. In this assessment, a NOAEC of 561 mg/m³ (125.3mg/kg bw/day), the highest dose tested, is estimated.

Male and female rats were administered 0, 500, 2500, or 5000 mg/kg bw/day C10–C13 mixed alkanes material containing 0.5% aromatics by gavage for 13 weeks (EMBSI 1991 as cited in Amoruso et al. 2008). Some animals in the high dose group had a recovery period of 28 days. The neurologic system was not tested. No treatment-related mortalities or adverse effects were observed. (EMBSI 1991 as cited in Amoruso et al. 2008).

8.1.2.7 Genotoxicity and carcinogenicity

WHO, SCOEL, and OECD reported that in vitro and in vivo data on white spirits did not indicate further human risk concern for genotoxic and carcinogenic effects (WHO 1996; SCOEL 2007; OECD 2012b; McKee 2018).

The genotoxicity potential of Stoddard solvents was negative in routine in vivo or in vitro genotoxicity assays such as Ames assay, chromosomal abnormalities in rat bone marrow cells, sister-chromatid exchange (SCE) in human lymphocytes, dominant lethal tests in mice and rats, and micronucleus assay in BALB/c mice (Gochet et al. 1984; WHO 1996).

Chronic (2-year) inhalation exposure to 0 or 138 mg/m³ (male rats only) or to 550, 1100 or 2200 mg/m³ (female rats only) of Stoddard solvent for 6 h/day, 5 days/week (n= 50/sex/dose) caused mortality in the highest dose group but did not cause any changes in the body weight of these animals. There was a significant increase in the incidence of benign and benign or malignant pheochromocytoma (combined) of the adrenal medulla in males in the 550 or 1100 mg/m³ (equivalent to 117.8, or 236.5 mg/kg bw/day) dosing groups. On the basis of those results, the authors of the study concluded that under the 2-year exposure, Stoddard solvent showed evidence of carcinogenic activity in male rats. Furthermore, a slight increase in renal tubule adenoma was seen in the 1100 mg/m³ group. Analysis of cell proliferation in the kidney showed the presence of α 2u-globulin, which is a male rat-specific lesion and is not considered relevant to humans. In the same study, inhalation exposure to 0, 550, 1100, or 2200 mg/m³ (6 h/ day, 5 days/wk) in male or female mice caused a slight increase in the incidence of hepatocellular adenoma in female mice exposed to 2200 mg/m³ (equivalent to 734.8 mg/kg bw/day); however, the authors considered this as no evidence of carcinogenicity (NTP 2004). On the basis of this carcinogenic study, a NOAEC of 138 mg/m³ (equivalent to 29.7 mg/kg bw/day) which is based on the increased incidence of benign and benign or malignant pheochromocytoma of the adrenal medulla in males at 550 mg/m³ (equivalent to 117.8 mg/kg bw/day) and above, is determined in this assessment.

8.1.3 Characterization of risk to human health

Exposure of the general population to substances in subgroup 1 may occur primarily from the use of products available to consumers.

The accumulation of white spirit, especially aliphatic hydrocarbon components, in fat tissue including the brain are considered relevant for the adverse effects on the central nervous system. Taking all available health effects information and toxicokinetic literature into account, there are no major differences in neurotoxicity when comparing aromatized and dearomatized C₉–C₁₄. An identical point of departure is chosen for aromatized and dearomatized C₉–C₁₄ in this assessment.  

For long- and short- duration exposures, a developmental neurotoxicological LOAEC of 4679 mg/m³ (equivalent to 1115.7 mg/kg bw/day), lowest dose tested, was determined on the basis of permanent impairment in cognitive functions and a memory deficit in prenatally exposed offspring rats (Hass et al. 2001). These results are supported by the neurological effects determined for long term exposure, and by chemical and neuronal changes at lower doses in rats. At 2339 mg/m³ (equivalent to 513.5 mg/kg bw/day; cognitive functions and memory were not tested at this dose) and 4679 mg/m³ (equivalent to 1027.2 mg/kg bw/day), an increase of cytosolic calcium in synaptosomes at postnatal day 35 in female offspring in the absence of maternal toxicity was observed and confirmed by in vitro testing (Edelfors et al. 1999).  

To support the risk of developmental neurotoxic effects, all inhalation studies of dearomatized white spirit exposure on adult male rats showed other permanent changes. Changes in neuronal functions such as an increase in GFAP (Lam et al. 2000) and glutathione level (Lam et al. 1994) or a decrease in 5-hydroxytryptamine transporter (5-HTT) and 5-HT_2A receptor (Lam et al. 2001) were seen after a short-term exposure at 2339 mg/m³ (lowest dose tested).  Certain central nervous system changes such as a decrease in motor activity (Lund et al. 1996) were observed after a long-term exposure at the same dose, 2339 mg/m³ (lowest dose tested).  Similar results were observed with aromatized white spirits at an identical dose, 2400 mg/m³. Male adult rats showed a reduction in the level of synaptosomal proteins and an increase in noradrenaline, dopamine, 5-hydroxytryptamine, and cholinesterase activities after a short- and long-term exposure (Ostergaard et al. 1993; Lam et al. 1995).

2 epidemiological studies linked an exposure to white spirits at concentrations greater than 200 ppm (~1300 mg/m^3, assuming average molecular weight of 160 g/mol) to spontaneous abortion regardless of the duration during the first term of pregnancy (Lindbohm et al. 1990, Windham et al. 1991).

Based on the available literature on neurotoxicity in men and laboratory animals, the critical effect levels were converted to an internal dose and a LOAEC of 4679 mg/m³ (equivalent to 1115.7 mg/kg bw/day) is determined as a POD for inhalation and oral routes. This LOAEC is considered to be protective of the neurotoxic effects in humans for short-term exposure (up to 3 hours) and for long-term exposure.  

The dermal study conducted on male Wistar rats was selected for the characterization of the risk in short- and long-term dermal exposure following dermal exposure to aromatized or dearomatized white spirit for 3 h per day, 5 days per week for 6 weeks (Verkkala et al. 1984) in section 8.1.2.3. A LOAEL of 691 mg/kg bw/day was identified as the POD on the basis of adverse effects on the peripheral nerve system (axons with prenodal swellings with widening of the nodes of Ranvier and no motor response following an electric stimulation). After a short- or long-term exposure of white spirits by inhalation, the nervous system was also affected and showed a decrease in motor activity in rats (Lund et al. 1996; ECHA 2011b; Lammers et al. 2007). In Lund et al. (1996), effects are considered irreversible because they are observed after a recovery period (Lund et al. 1996). In 2 epidemiological studies, tingling in hands and feet, and a reduction of sensation in hands and feet after a stimulation were observed in painters exposed to white spirit solvents (Cherry et al. 1985; Demers et al. 1991). As a weight of evidence, the LOAEL as the POD for the dermal route is considered to be protective of the adverse effects observed on the peripheral nerve system for all ages and sexes.   

8.1.3.1 Exposure from environmental media and food

The estimated annual inhalation dose for the highest release source of C₉–C₁₄ hydrocarbon solvents for the general population living in the vicinity of a non-petroleum facility was 0.14 mg/kg bw/day (adults 19+) to 0.51 mg/kg bw/day (1 yr), assuming complete inhalation absorption.  

As described in section 7.2.2, there is the potential for C₉–C₁₄ hydrocarbon solvents to be released to water bodies via industrial releases and down-the-drain scenarios. There is a potential for exposure if these substances are released into water bodies that become a source of drinking water. The 95^th percentile concentrations from industrial releases and down-the-drain sources are 36 μg/L and 12 μg/L, respectively. The predicted drinking water intakes for the LBPNs resulting from potential industrial releases ranged from 0.00077 mg/kg bw/day (adults) to 0.0012 mg/kg bw/day (infant 1 yr). These values, which assume that the concentrations in treated receiving water bodies are similar to those of drinking water are conservative and are expected to represent potential worst-case releases of LBPN substances in all subgroups.

The combined exposure from ambient air and drinking water is 0.14 mg/kg bw/day (adults 19+) to 0.51 mg/kg bw/day (1 yr). Compared to the LOAEL of 1115.7 mg/kg bw/day for neurotoxicological effects from long-term exposures, these exposures give MOEs of 2200 to 8000. Given that they are on the basis of epidemiological data, these MOEs are considered adequate and, therefore, the risk to the general population in Canada from environmental media exposure to C₉–C₁₄ hydrocarbon solvent LBPN substances from non-petroleum industrial facilities is considered to be low.

Dietary exposure, if any, from use of some of the subgroup 1 substances as a component in the manufacture of food packaging materials or as a component of incidental additives is considered to be negligible.

8.1.3.2 Exposure from products available to consumers

Inhalation and dermal exposures to the C₉–C₁₄ hydrocarbon solvents from cosmetics and products available to consumers, the corresponding critical hazard point of departure, and MOE are presented in Table 8‑2. The vulnerable subpopulation of children and women of reproductive age (that is, teens and adult women) and the sub-population living near facilities are considered in the exposure calculations, health effects assessment, and risk characterization of subgroup 1 substances.

^a The exposures identified are for adults unless otherwise noted.
^b Retention factors used for dermal exposures are given in Appendix D.
^c Using dermal LOAEL = 691 mg/kg bw/day on the basis of adverse effects on the peripheral nervous system as the critical endpoint.
^d Using inhalation LOAEC = 4679 mg/m³ leading to an adjusted LOAEL = 1115.7mg/kg bw/day on the basis of adverse effects on developmental neurotoxicity (permanent impairment in cognitive functions and a memory deficit in prenatally exposed offspring)
^e  For ventilation rate of 2.5 h^-1.
^f  For ventilation rate of 0.5 h^-1.

The resulting margins of exposure (MOEs) for dermal exposures to a number of cosmetic products and products available to consumers listed in Table 8-2 are below 300. MOEs for inhalation exposures from a number of products available to consumers with expected short duration infrequent exposures are below 1000. Both of these are considered to be inadequate to address the uncertainty in both the available hazard and exposure data.

The vulnerable subpopulation of children and women of reproductive age (that is teens and adult women) and the sub-population living near facilities are considered in the exposure calculations, health effects assessment, and risk characterization of subgroup 1 substances.

8.2 Subgroup 2 (C₉ Aromatic Hydrocarbon [CAS RN 64742-95-6])

8.2.1 Exposure assessment

8.2.1.1 Environmental media and food

No Canadian or recent international data on the levels of CAS RN 64742-95-6 in indoor or outdoor air, water, dust, or soil, were identified through searches of available literature. While variable in composition, this substance has a high vapour pressure (405 Pa) and high Henry’s Law constant values (-2 [log₁₀ atm·m³/mol]), suggesting that it is likely to enter ambient air if released and has the potential to rapidly volatilize from surface waters (OECD 2012a).

Information received in response to a CEPA section 71 survey (Environment Canada 2012b) and data from the NPRI (2019) indicate that CAS RN 64742-95-6 is released to air from facilities that report use of the substance. Therefore, there may be a potential for inhalation exposure to populations that may reside near these facilities. SCREEN3, a tier-one air dispersion model developed by the US EPA (SCREEN3 1996), was used to estimate the potential ambient air concentration of CAS RN 64742-95-6 from releasing facilities.

The largest quantity released to air reported by a single facility was reported to be 101 883 kg (CAS RN 64742-95-6) in the NPRI 2017 reporting year. However, the quantity reported by another facility associated with a smaller release quantity of 89 000 kg from the NPRI 2017 reporting year (NPRI 2020) resulted in the largest inhalation exposure potential because of the closer distances between residences and the reporting facility. The estimated exposure is based on a combination of the quantity released per year of the substance, number of release days per year, and distance to nearby receptors. Using this quantity (89 000 kg) and the SCREEN3 model with exposure factors given in Table E-1 of Appendix E, the yearly ambient air concentration for CAS RN 64742-95-6 resulting from facility releases to air was estimated to be 0.155 mg/m³. This concentration represents a systemic dose of 0.032 mg/kg bw/day (adults 19+) to 0.113 mg/kg bw/day (infant 1 yr), assuming complete inhalation absorption.    

Limited data were available on releases to water for certain substances in this subgroup, according to information received in response to a CEPA section 71 survey (Environment Canada 2012b). The largest quantity reported to be released to water by a single facility for any of the LBPN substances was 10 000 kg. As described previously (see section 8.1.1.1), a conservative estimate of drinking water intakes was calculated using the 95% percentile for LBPNs from releases to surface water from wastewater from industrial releases and down-the-drain sources. The predicted drinking water intakes for the LBPNs resulting from potential industrial releases ranged from 0.00077 mg/kg bw/day (adults) to 0.00311 mg/kg bw/day (formula fed infants, 0 to 5 months). 

CAS RN 64742-95-6 may be used as a component in the manufacture of certain food packaging materials that have the potential to come into direct contact with food. Assuming a worst-case scenario, that is, that 100% of the substance migrates into the food, the estimated cumulative probable daily intake of CAS RN 64742-95-6 from food packaging materials is 314 ng/kg bw/day for the general population (personal communication, email from the HC Food Directorate to the HC ESRAB, dated February 2020; unreferenced). Other uses of this substance as a component in the manufacture of food packaging materials in incidental additives would not involve contact with food; therefore, dietary exposure form these uses is not expected (personal communication, email from the HC Food Directorate to the HC ESRAB, dated February 2018 and February 2020; unreferenced).

8.2.1.2 Products available to consumers

CAS RN 64742-95-6 was identified as an ingredient in products available to consumers in Canada. With the exception of a limited number of cosmetic products, a majority of the products identified were DIY products; see section 4, Sources and uses. The predominant routes of exposure are expected to be dermal and inhalation as no products that may result in oral exposures were identified.

To evaluate the potential for exposure to CAS RN 64742-95-6, “sentinel” scenarios were selected, taking into consideration frequencies of use, reported concentrations, and availability in Canada. The vulnerable subpopulation of children and women of reproductive age (that is, teens and adult women) and the subpopulation living near facilities are considered in the exposure calculations, health effects assessment, and risk characterization of the subgroup 2 substance. The exposure values were selected based on either the age group considered most relevant to the scenario or relevant sensitive subpopulation and/or were based on the age group that had the highest exposure as determined in Appendix E.

Table 8-3 describes the range of estimates of exposures to the aromatic hydrocarbon solvent CAS RN 64742-95-6 from products available to consumers, calculated using ConsExpo (RIVM 2016). The exposure factors used in determining the exposures are given in Tables E-2 to E-5 of Appendix E. The substance was also identified as an ingredient in an automotive fuel cleaner; however, the potential exposures from this use are expected to be limited. Given the types of products, only adult exposures were considered, except in the case of nail polish and nail adhesive. 

To estimate systemic exposures from the dermal route, a conservative estimate of dermal absorption of 10% was used on the basis of 2 dermal absorption studies conducted with possible discrete components in CAS RN 64742-95-6. The dermal absorption of a C₉ aromatic hydrocarbon, 1,2,4-trimethylbenzene (which may be a component of CAS RN 64742-95-6), was studied by Korinth et al. (Korinth et al. 2003; ECHA 2020b) from a neat liquid and from 50% ethanol solutions on untreated, occluded human skin in a static diffusion cell (in vitro). The dermal absorption of 1,2,4-trimethylbenzene penetrating the skin from these studies was determined to be 0.1% to 0.15% after 8 hours of exposure (ECHA 2020b). The C₈ aromatic hydrocarbon ethylbenzene, which is within the carbon range of CAS RN 64742-95-6, from an in vivo study of albino hairless mice was determined to have a dermal absorption of 3.6% (Susten 1990; ECHA 2020c].

To prevent double counting of exposure from dermal absorption in the inhalation exposure amount, the total amount of systemically available substance from dermal absorption was subtracted from the total available for inhalation.

^a The exposures identified are for adults unless otherwise noted.
^b Inhalation exposure converted to internal dose, assuming complete inhalation absorption and an inhalation rate of 15.1 m³/day (adults 19+), 15.9 m³/day (14–18 years), 13.9 m³/day (9–13 years), 11.1 m³/day (4–8 years), and 9.2 m³/day (2–3 years) (US EPA 2011a). Corrected for the amount of substance that becomes systemically available through dermal absorption, with a dermal adsorption factor of 10%. 
^c For ventilation rate of 2.5 h-1
^d For ventilation rate of 0.5 h-1

8.2.2 Health effects assessment

This subgroup consists of only one substance, high flash aromatic naphtha (CAS RN 64742-95-6).

8.2.2.1 Neurotoxicity

In a study examining neurotoxicity, Sprague-Dawley rats (n=20 males/group) were exposed by inhalation (whole-body) to 0, 496 2120, or 6479 mg/m³ (equivalent to approximately 0,104.7, 451, or 1409.3 mg/kg bw/day) of high flash aromatic naphtha for 6 h per day, 5 days per week for 90 days (Douglas et al. 1993). At the highest concentration, the animals exhibited significantly lower body weights (12%). No treatment-related, neurotoxic effects were observed up to the highest tested concentration of 6479 mg/m³ (approximately 1409.3 mg/kg bw/day).

8.2.2.2 Repeated dose toxicity

The subchronic effects of high flash aromatic naphtha through the oral route were investigated in a 13-week study, in which Sprague-Dawley rats (n=10/sex/group) were gavaged with 0, 30, 125, 500, or 1250 mg/kg bw/day, 5 days per week (TSCATS OTS0556721 as cited in US EPA 2011b). At 500 mg/kg bw/day, clinical signs such as salivation, pale reddish brown oral discharge, and perineal staining were observed. At 500 and 1250 mg/kg bw/day, there were significantly reduced body weights and body weight gains in both sexes, while males exhibited increases in alanine aminotransferase and protein levels. The same dose levels were associated with clinical chemistry changes (increased total bilirubin, albumin, albumin/globulin ratio, alanine aminotransferase, alkaline phosphatase, creatinine, total protein levels, decreased blood urea nitrogen levels) in both sexes. There were also organ weight increases (adrenals, liver) in both sexes. Histopathological examinations revealed liver cell hypertrophy in both sexes. In the report, a NOAEL of 125 mg/kg bw/day was identified by the US EPA on the basis of effects on the liver (increased liver weights, alanine aminotransferase, liver cell hypertrophy) observed at the next dose level of 500 mg/kg bw/day.

To examine the subchronic effects of high flash aromatic naphtha through the inhalation route, rats were exposed to 1800, 3700, or 7400 mg/m³ for 13 weeks (Clark et al. 1989). Exposure to high flash aromatic naphtha resulted in liver and kidney weight increases in female rats at the mid- and high-concentration levels. A low-grade anemia was observed in the female animals at all exposure levels (no further details provided).

The chronic effects of high flash aromatic naphtha through the inhalation route was examined in a study in which Wistar rats (n=50/sex/group) were exposed via inhalation (whole-body) to 0, 470, 970, or 1830 mg/m³ (equivalent to approximately 0, 103.2, 212.9, or 401.7 mg/kg bw/day), 6 h per day, 5 days per week for 12 months (Clark et al. 1989). A subset of the animals (n=15/sex/group) was also subjected to a recovery period of 4 months prior to sacrifice. A possible increase in “aggression” was observed in the males at the highest concentration, which persisted into the recovery period. At the same concentration, there were significantly increased liver (11%) and kidney (10%) weights. In the female animals at the highest concentration, significantly elevated sodium (1%) and albumin (8%) levels were observed. No treatment-related histopathological changes or tumour incidence were observed. The authors concluded that chronic exposure to high flash aromatic naphtha did not result in systemic toxicity up to the highest concentration tested, 1830 mg/m³ (approximately 401.7 mg/kg bw/day when converted to an internal dose). In addition, signs of aggression are consistent with a neurotoxic effect. In this assessment, a NOAEC of 970 mg/m³ (equivalent to 212.9 mg/kg bw/day) is estimated based on these results.

8.2.2.3 Carcinogenicity and genotoxicity

With respect to genotoxicity, there was no evidence that high flash aromatic naphtha would result in genetic mutations or chromosomal aberrations. It was negative for genotoxicity in bacterial mutagenicity assays, in a mammalian cell gene mutation assay, in an in vitro chromosome aberration assay, and in a sister chromatid exchange assay (Schreiner et al. 1989). Negative findings were also observed in an in vivo chromosomal aberration assay (Schreiner et al. 1989).

Studies examining carcinogenicity were not identified. However, the 12-month, inhalation study conducted by Clark et al. (1989) did not reveal any treatment-related differences in tumour at concentrations of up to 1800 mg/m³ (approximately 401.7 mg/kg bw/day). Therefore, CAS RN 64742-95-6 is not expected to result in carcinogenicity.  

8.2.2.4 Developmental and reproductive toxicity

In a developmental toxicity study, CD-1 mice (n=30 females/group) were exposed by inhalation (whole-body) to high flash aromatic naphtha (CAS RN 64742-95-6) at concentration levels of 0, 501, 2458, and 7442 mg/m³ (equivalent to approximately 0, 175.8, 857.3, or 2620.3 mg/kg bw/day), 6 hours per day from GD 6 to 15 (McKee et al. 1990). At 501 mg/m³, there was a decrease in maternal body weight gain (17%) between GD 6 to 18 as well as in the number of live fetuses per litter. At 2458 mg/m³, 2 pregnant mice died (one due to injury, one due to an unknown cause), and maternal body weight gain was significantly reduced (17%) between GD 6 to 18. Fetal body weights were also significantly reduced (7%) and ossification was delayed compared to controls. At the highest concentration, 44% of the pregnant animals died (2 died on the first day of exposure on GD 6, and 12 died between GD 8 to 16). Pregnant mice at this concentration level were also associated with significantly decreased body weight (15%), body weight gains (39%), clinical signs (for example, abnormal gait, laboured breathing, hunched posture, weakness, inadequate grooming, circling, ataxia), and hematological findings (decreased hematocrit, mean corpuscular volume). With respect to the fetuses, the number of live fetuses per litter and fetal body weight were significantly reduced at the middle and highest doses (26% and 34%, respectively). There was also an increase in post-implantation loss at the highest concentration as well as evidence of delayed ossification and an elevated incidence of cleft palate. In this assessment, a LOAEC of 501 mg/m³ (175.8 mg/kg bw/day) was identified as the POD on the basis of maternal toxicity (reduced body weight gain) and developmental toxicity (fewer fetuses per litter), and dams dose-related mortality and more developmental toxicity (reduced fetal body weight and delayed ossification) at the next concentration of 2458 mg/m³.

In a 3-generation reproductive toxicity study, COBS CS rats (n=30/sex/group) were exposed via inhalation (whole-body) to 0, 506, 2433, 7275 mg/m³ (equivalent to approximately 0, 111.1, 534.1, or 1597.1 mg/kg bw/day) of high flash aromatic naphtha (McKee et al. 1990). The animals (F0 generation) were exposed to treatment for 10 weeks prior to mating and during the mating period of 2 weeks, 6 h per day, 5 days per week. Once mating was confirmed, the male animals were sacrificed while the dams continued treatment from GD 0 to 20. The dams were then placed in nesting boxes to deliver pups (F1 generation), and the treatment was stopped. The treatment was re-initiated on postnatal day (PND) 5 and continued until weaning (PND 21). A week after the weaning period, animals in the F1 generation (n=30/sex/group) were exposed for 10 weeks, and then mated to generate the F2 generation. Immediately after weaning on PND 22, treatment was initiated for the animals in the F2 generation (n=40/sex/group). Since the majority of the F2 animals in the high dose group died during the first week of exposure, all survivors were mated to generate the F3 generation.

In the F0 parents, exposure to 2433 mg/m³ resulted in significantly reduced body weight gains (5% to 7%). These effects were more pronounced at the highest concentration of 7264 mg/m³ (14% to 16%) and was accompanied by increased mortality (7 females) and lung lesions (foci of pulmonary macrophages). At the highest concentration, the F1 pups exhibited significantly reduced body weights (12% to 24%) beginning on PND 7 through to the weaning period, 2 days after dams re-initiated treatment at the highest concentration. In addition, following weaning, many of the F1 parents showed clinical signs (for example, ataxia, reduced motor activity) and lung lesions, and some of the female animals died. There was also significantly reduced male fertility (28%) and litter size at birth (30%) compared to controls. The F2 pups also exhibited significantly reduced birth weights and poor survival throughout the lactation period. Exposure of the F2 pups to the highest concentration immediately after weaning (PND 22) resulted in many of the animals (90% males, 85% females) dying within the first week, while the surviving animals exhibited mean body weights that were 21% to 40% below controls. The pups in the F3 generation of the mid- and high-concentration groups had significantly lower body weights compared to controls. A NOAEC of 506 mg/m³ (equivalent to 111.1 mg/kg bw/day) was considered on the basis of parental toxicity (for example, adverse reductions in body weight, lung lesions, mortality) and reproductive and developmental toxicity (for example, reduced fertility, litter size, fetal weight, poor survival) observed at the next concentration of 2433 and 7264 mg/m³.     

8.2.3 Characterization of risk to human health

Subgroup 2 is comprised of a single substance, high flash aromatic naphtha (CAS RN 64742-95-6). On the basis of the data available, it is not likely to result in carcinogenicity or genotoxicity.

The developmental toxicity study conducted on CD-1 mice was selected for the characterization of risk following short-term inhalation and dermal exposures to high flash aromatic naphtha. A LOAEC of 501 mg/m³ (111.1 mg/kg bw/day) was identified as the POD on the basis of maternal toxicity (reduced body weight gain, dose-related mortality) and developmental toxicity (fewer fetuses per litter), and dams dose-related mortality and more developmental toxicity (reduced fetal body weight and delayed ossification) at the next concentration of 2433 mg/m³ (equivalent to 534.1 mg/kg bw/day). Although only one mouse died of an unknown cause and another of injury at 2433 mg/m³, this finding was considered to be toxicologically relevant since more deaths (44%) occurred at higher concentrations (7475 mg/m³) starting from the first day of exposure (GD 6). At these concentrations, the high mortality rates were accompanied by signs of overt toxicity (abnormal gait, laboured breathing, hunched posture, weakness, inadequate grooming, circling, ataxia), and hematological findings (decreased hematocrit, mean corpuscular volume). This point of departure was also considered to be protective of the effects observed in studies of longer duration (for example, low-grade anemia, liver and kidney effects from subchronic and chronic studies) and was therefore considered to be applicable for long-term exposure scenarios. Taking these systemic effects into consideration, the critical effect level was converted to an internal dose, and a LOAEL of 111.1 mg/kg bw/day was derived.

With respect to the oral route, a 13-week gavage study conducted on rats was selected for the characterization of risk following oral exposures. A NOAEL of 125 mg/kg bw/day, which was based on effects on the liver (increased liver weights, alanine aminotransferase, liver cell hypertrophy) and kidneys (increased kidney weights, total protein levels) observed at the next dose level of 500 mg/kg bw/day, was considered to be the most relevant endpoint for characterization of risk to human health from oral exposure.  

8.2.3.1 Exposure from environmental media and food

Potential risks to the general population from possible exposure to LBPNs from their production, use, and transport between petroleum facilities have been previously addressed (EC, HC 2011, 2013). The estimated annual inhalation dose for the highest release source of C₉ aromatic solvents for the general population living in the vicinity of a non-petroleum facility ranged from 0.032 mg/kg bw/day (adults 19+) to 0.113 mg/kg bw/day (infant 1 yr), assuming complete inhalation absorption. Compared to the LOAEL of 111.1 mg/kg bw/day for maternal and developmental toxicological effects, these exposures give an MOE of 1500 to 5300. The MOEs are considered adequate; therefore, the risk to the general population in Canada from inhalation exposure to the subgroup 2 LBPN substance from non-petroleum industrial facilities is considered to be low. 

As described in section 7.2.2, there is the potential for C₉ aromatic solvents to be released to water bodies via industrial releases and down-the-drain scenarios. There is a potential for exposure if these substances are released into water bodies that become a source of drinking water. These 95th percentile concentrations from industrial releases and down-the-drain sources are 36 μg/L and 12 μg/L, respectively. The predicted drinking water intakes for the LBPNs resulting from potential industrial releases ranged from 0.00077 mg/kg bw/day (adults) to 0.00311 mg/kg bw/day (formula-fed infants, 0 to 5 months). Compared to the NOAEL of 125 mg/kg bw/day for liver and kidney effects from oral exposure, these exposures give an MOE of 40 000 to 162 000. The MOEs are considered adequate; therefore, the risk to the general population in Canada from oral exposure to the subgroup 2 LBPN substance from drinking water sources is considered to be low.

A comparison of the estimated cumulative dietary exposure (314 ng/kg bw/day for the general population), under a worst-case scenario from the use of the single LBPN subgroup 2 substance (CAS RN 64742-95-6) in the manufacture of certain food packaging materials with a NOAEL of 125 mg/kg bw/day for liver and kidney effects, results in an MOE of 400 000. This MOE is considered to be adequate; therefore, the risk to the general Canadian population from dietary exposure to this LBPN substance resulting from its use in food packaging materials is considered to be low.

8.2.3.2 Exposure from products available to consumers

Table 8‑4 provides relevant exposure and hazard values for the subgroup 2 substance, as well as resultant margins of exposure, for determination of risk. Due to the short duration of these exposures, the LOAEL of 111.1 mg/kg bw/day for maternal toxicity was considered to be the most relevant point of departure for determining MOEs.

^a The exposures identified are for adults unless otherwise noted.
^b The critical health effect endpoint of LOAEL = 111.1 mg/kg bw/day from a LOAEC of 501 mg/m³ on the basis of maternal and developmental toxicity.
^b For ventilation rate of 0.5 h^-1
c For ventilation rate of 2.5 h^-1

The resulting margins of exposure (MOEs) for nail polish, nail adhesive, spray paint, stain, coating, and lacquer products available to consumers listed in Table 8-4 are below 300 and considered to be inadequate to address the uncertainty in both the available hazard and exposure data for some of the products.

The vulnerable subpopulation of children and women of reproductive age (that is, teens and adult women) and subpopulations living within the vicinity of facilities are considered in the exposure calculations, health effects assessment, and risk characterization of the subgroup substance.

8.3 Subgroup 3 (C₆–C₉ Aliphatic Hydrocarbons)

8.3.1 Exposure assessment

8.3.1.1 Environmental media and food

No data on the substances in LBPN subgroup 3 were identified in indoor or outdoor air, water, dust, or soil, through searches of available literature. Substances in this group, while variable in composition, have high vapour pressures (590 Pa to 31 000 Pa) and high to very high Henry’s Law constant values (-2 to 0.47 [log₁₀ atm•m³/mol]) which suggests that they are likely to enter ambient air if released and have the potential to rapidly volatilize from surface waters (OECD 2010, 2011, 2013).

With the exception of CAS RNs 68647-60-9 and 426260-76-6, information received in response to a CEPA section 71 survey (Environment Canada 2011a,b,c, 2013) and data from the NPRI (2019) indicate that substances in this group are released to air from facilities associated with these substances. Therefore, there may be a potential for inhalation exposure to populations that may reside near these facilities. SCREEN3, a tier-one air dispersion model developed by the US EPA (SCREEN3 1996), was used to estimate the potential outdoors ambient air concentration of substances in this subgroup.

The largest quantity reported by a single facility for releases to air according to information received in response a CEPA section 71 survey (Environment Canada 2011b) for any LBPN substance in subgroup 3 was reported to be 100 000 kg, as shown in Table 4-2 (for CAS RNs 64741-66-8 and 64742-89-8). The quantity reported by a single facility to the NPRI for the 2017 reporting year was 53 000 kg for CAS RN 64742-89-8 (NPRI 2019). Using the quantity of 100 000 kg and basing itself on the upper bounding quantity released per year, number of release days per year, and distance to nearby receptors for the general population (200 m), the SCREEN3 model gives 0.30 mg/m³ as the yearly ambient air concentration resulting from facility releases to air of these substances. See Table F-1 of Appendix F for a complete list of settings for the SCREEN3 calculation. This concentration represents a potential exposure dose that ranges from 0.062 mg/kg bw/day (adults 19+) to 0.22 mg/kg bw/day (infant 1 yr). The potential exposure via releases to air from the other substances in subgroup 3 is expected to be lower than this.

Limited data were available on releases to water for certain substances in this subgroup according to information received in response to a CEPA section 71 survey (Environment Canada 2011a, 2012b). The largest quantity reported by a single facility reporting releases of any of the LBPN substances to water was 10 000 kg. As described previously (see Section 8.1.1.1), a conservative estimate of drinking water intakes was calculated using the 95% percentile for LBPNs from releases to wastewater from industrial releases and down-the-drain sources. The predicted drinking water intakes for the LBPNs resulting from potential industrial releases ranged from 0.00077 mg/kg bw/day (adults) to 0.0012 mg/kg bw/day (infant 1 yr). These values are expected to be representative of potential releases of LBPN substances in all subgroups (Environment Canada 2011a, 2012a, 2012b, 2013). These values, which assume that the concentrations in treated wastewater are similar to those of drinking water, are conservative.

Using the information on potential concentrations of substances in LBPN subgroup 3 from releases to the environment, the highest reported LBPN subgroup 3 intakes from both release to air and water were estimated to be 0.063 mg/kg bw/day (adults 19+) to 0.153 mg/kg bw/day (infant 1 yr), (US EPA 2011b; Health Canada 2017).  

2 of the LBPN subgroup 3 substances (CAS RNs 64742-49-0 and 64742-89-8) may be used as components in the manufacture of food packaging materials that come into direct contact with food; however, dietary exposures from these uses are considered to be negligible. Other uses of LBPN subgroup 3 substances in the manufacture of food packaging materials are such that there would be no contact with food; therefore, dietary exposure from these uses is not expected. Some of the LBPN subgroup 3 substances (CAS RNs 64741-66-8, 64742-49-0, 64742-89-8, and 426260-76-6) may also be used as components in incidental additives (for example, lubricants and/or cleaners) used in food processing establishments; however, dietary exposure, if any, would be considered to be negligible (personal communication, email from the HC Food Directorate to HC ESRAB, dated February 2018 and February 2020; unreferenced).

8.3.1.2 Products available to consumers

Substances in LBPN subgroup 3 were identified as ingredients in a large number of products available to consumers in Canada. With the exception of a limited number of cosmetic products, the majority of products identified were DIY products (see Section 4). No products that may result in oral exposure to substances in this subgroup were identified.

To evaluate the potential for exposure to substances in the C₆–C₉ aliphatic hydrocarbon subgroup, “sentinel” scenarios resulting in the highest level of potential exposure to substances via the inhalation and dermal routes were selected, taking into consideration frequencies of use and reported concentrations. For substances in subgroup 3, estimated concentrations are presented as one-day average exposure concentration values, given that these substances may be associated with short-term health effects (see Section 8.3.2). The vulnerable subpopulation of children and women of reproductive age (that is, teens and adult women) and the sub-population living near facilities were considered in the exposure calculations, health effects assessment, and risk characterization of subgroup 3 substances. The exposure values were selected on the basis of either the age group considered most relevant to the scenario or the relevant sensitive subpopulation and/or were on the basis of the age group that had the highest exposure, as determined in Appendix F.   

Table 8‑5 below describes the range of estimated inhalation exposures to the C₆–C₉ aliphatic hydrocarbons subgroup from products available to consumers. As there may be numerous products with different subgroup 3 substances in each category, no specific CAS RN was assigned to each product category. To prevent double counting of exposure from both the dermal and inhalation routes, the total amount of systemically available substance from dermal exposure was subtracted from the total available for inhalation. The exposure factors for these calculations are given in Table F-2 of Appendix F.

To determine systemic exposures from the dermal route, estimates of the dermal absorption of the substances in this subgroup are required. Direct measurements of dermal absorption have not been identified for the substances of subgroup 3. Given the high vapour pressures of the substances in subgroup 3 (1300 Pa to 31 000 Pa), it is expected that substances in this subgroup will evaporate after a short time in contact with skin. Based on the short duration of dermal exposure due to evaporation and the range of log Kow for these substances (3.2 to 5.7), a dermal absorption of 25% is assumed. This is supported by available maximum flux data (J_max) for 2 individual substances: n-hexane, a C₆ aliphatic substance with a resorption flux of 0.83 μg/cm²/h for 30 min exposure (from in vitro measurement of human skin with closed cell) (Lodén 1986), and n-heptane, a C₇ aliphatic substance with a resorption flux of 113 μg/cm²/h for 10 min exposure (from in vitro measurement of human skin with closed cell) (ECHA 2020d). Exposures calculated for typical scenarios using the J_max of these individual substances resulted in dermal exposures lower than those calculated in Table 8-6 using a 25% dermal absorption factor (Table F-4 of Appendix F).

^a The exposures identified are for adults unless otherwise noted.
^b Internal dose (mg/kg bw/day) = mean concentration on day of exposure [mg/m³]×inhalation rate (m³/day) / body weight [kg]).
^c Assumed use of 2 coats
^d For ventilation rate of 2.5 h^-1.
^e For ventilation rate of 0.5 h^-1.

Table 8‑6 below describes the range of estimated dermal exposures to CAS RNs in the C₆–C₉ aliphatic hydrocarbons subgroup from products available to consumers. The exposure factors used in these calculations are given in Table F-3 of Appendix F.

^a The exposures identified are for adults unless otherwise noted.
^b Dermal exposures include the dermal absorption factor and represent internal doses.

8.3.2 Health effects assessment

The health effects datasets for each individual substance in the C₆–C₉ aliphatic hydrocarbons solvents subgroup were considered to be limited for neurotoxicity, clinical toxicology and genotoxicity.

In a short study examining the neurobehavioral effects of acute exposure to C₆/C₇ cycloparaffinic solvent (CAS RN 64742-89-8), male Wistar rats were exposed through inhalation (whole-body) to 0, 1405, 4255, and 13 955 mg/m³ (equivalent to approximately 0, 411.2, 1245.4, or 4084.6 mg/kg bw/day), 8 hours per day for 3 days (McKee et al. 2011). The animals were evaluated for the functional observational battery (FOB) (Ross 2000), including viability, measures of well-being (for example, body weight), functional observations, motor activity, and visual discrimination performance. At the highest concentration, there was a statistically significant decrease in body temperature (0.5°C) during the 3-day period. There were no significant changes in any of the FOB measurements, and there were no treatment-related effects on motor activity. With respect to measures of visual discrimination, the animals exposed to the highest concentration exhibited significantly increased response latency times (2.11 seconds vs. 1.7 seconds in the control), which did not persist after the 3-day period. The authors indicated that, overall, C₆/C₇ cycloparaffinic solvent (CAS RN 64742-89-8) reversibly affected the response speed component of visual discrimination performance at an exposure level of 13955 mg/m³. A NOAEC of 4255 mg/m³ (1245.4 mg/kg bw/day) was identified as the POD on the basis of the neurological effects observed at the next concentration of 13955 mg/m³.

A neurotoxicity study was conducted using the distillate of light alkylate naphtha (CAS RN 64741-66-8), whereby Sprague-Dawley CD rats (n=12/sex/dose) were exposed via inhalation (whole-body) to 0, 2400, 8100, and 24 300 mg/m³ (equivalent to approximately 0, 510.4, 1724.3, or 5190.9 mg/kg bw/day for males and 0, 562.4, 1892.7, or 5693.1 mg/kg bw/day for females) of light alkylate naphtha distillate, 6 hours per day, 5 days per week for 13 weeks, An additional 12 animals in the control and highest dose groups were exposed as a satellite group for 13 weeks and subsequently maintained for a 28-day recovery period (Schreiner et al. 1998; 1999). Neurobehavioural evaluations of motor activity and a FOB were performed. No treatment-related effects on any endpoint evaluated within the FOB were observed, and no indications of neurotoxicity were observed up to the highest concentration of 24300 mg/m³ (approximately 5693.1 mg/kg bw/day). It should be noted that the test substance used in this study was a distillate generated from CAS RN 64741-66-8 and comprised of a wider carbon range of C4–C₁₀ than the unaltered substance (predominantly C₈–C₁₀). The authors stated that this distillate “may more accurately reflect the entire light vapour fraction to which humans could be exposed.”

In a subchronic toxicology study, Sprague-Dawley CD rats (n=12/sex/dose) were exposed via inhalation (whole-body) to 0, 2400, 8100, and 24300 mg/m³ (equivalent to approximately 0, 510.4, 1724.3, or 5190.9 mg/kg bw/day for males and 0, 562.4, 1892.7, or 5693.1 mg/kg bw/day for females) to the distillate of light alkylate naphtha (CAS RN 64741-66-8), 6 hours per day, 5 days per week for 13 weeks. An additional 12 animals in the control and highest dose groups were exposed as a satellite group for 13 weeks and subsequently maintained for a 28-day recovery period (Schreiner et al. 1998). No adverse effects were observed by the authors.  Exposure to the highest concentration was associated with an increased incidence of red facial staining, and a significant increase in hemoglobin, hematocrit, and erythrocyte levels (≤7% for each parameter). However, with the exception of hemoglobin, other parameters were similar to the control group after the recovery period. Significant increases in absolute and relative liver weights were also observed at the highest concentration in female animals, but these were not present after the recovery period and were not accompanied by histopathological findings. A NOAEC of 24 300 mg/m³ (approximately 5691.1 mg/kg bw/day) was considered in this study by the authors, representing the highest concentration tested (Schreiner et al. 1998).   

For substances in the C₆–C₉ aliphatic hydrocarbons solvents subgroup that have been tested for genotoxicity, no positive findings have been observed. In bacterial assays, naphtha (petroleum), hydrotreated light (CAS RN 64742-49-0) was not mutagenic in Salmonella typhimurium strains TA98, TA100, TA1535, TA1537, or TA1538, nor was it mutagenic in Escherichia coli WP2 with and without metabolic activation. Furthermore, it was not clastogenic in a chromosome aberration test on rat hepatocytes and did not induce mitotic gene conversion in yeast (Shell Research Limited 1983 as cited in OECD 2010a; Meyer 1983 as cited in OECD 2010a; Brooks et al. 1988). In an in vivo micronucleus test on murine bone marrow, no treatment-related increases in the formation of micronuclei were observed up to a limit dose of 2000 mg/kg bw/day (Fox 2003 as cited in OECD 2010a). Light alkylate naphtha (CAS RN 64741-66-8) was neiher mutagenic in an in vitro mouse lymphoma assay nor clastogenic in an in vivo bone marrow cytogenetics assay in rats (American Petroleum Institute 1985 as cited in Bui et al. 1998).   

Studies examining the effects of C₆–C₉ aliphatic hydrocarbon solvents on long-term exposure, reproductive and developmental toxicology, or carcinogenicity, were not identified.

8.3.2.1 Selection of representative UVCB mixtures

As the health effects datasets for the C₆–C₉ aliphatic hydrocarbons solvents subgroup were considered to be limited, toxicological data from similar UVCB substances have been used to inform the human health assessment, where appropriate. A list of the representative UVCB mixtures used to inform this assessment is presented in Table 8‑7.

These substances were identified on the basis of similar composition, aromatic content, functional groups, reactivity, physical-chemical properties, manufacturing processes, functional uses (that is, as solvents), and whether they had relevant empirical data that could be used to read across to substances with limited empirical data.

Similarity of composition, functional groups, and reactivity

The representative UVCBs all represent complex mixtures of hydrocarbons (linear, branched, or cyclic alkanes, mainly saturated) with aromatic contents comparable to the substances in the C₆–C₉ aliphatic hydrocarbon solvents subgroup (<1%). In addition, their predominant carbon distribution ranges from C₆ to C₉. The only functional groups associated with the representative UVCBs are alkyl side chains, similarly to the substances in the C₆–C₉ aliphatic hydrocarbon solvents subgroup.

Similarity of physical-chemical properties

Substances in the C₆–C₉ aliphatic hydrocarbon solvents subgroup, as well as the representative UVCBs, share common boiling point ranges, are liquids at room temperature, are highly volatile, and have similar representative structures and molecular weight ranges.

Similar production methods and use

Hydrocarbon solvents are separated from crude oil by atmospheric distillation and further fractionated by boiling range. Additional refining processes are performed to reduce components such as sulfur, nitrogen, benzenes, and aromatics.

Substances in the C₆–C₉ aliphatic hydrocarbon solvents subgroup, as well as the representative UVCBs, are all used as solvents that have well-defined rates of evaporation and limited aromatic content.

In addition to these considerations, the representative UVCBs represent the upper and lower boundaries of the carbon range spectrum (Appendix C). For example, the representative UVCB commercial hexane contains predominantly lighter hydrocarbons (approximately 99% C₆ isomers), while the other representative UVCBs (C₈-C₉ isoparaffinic solvent, C₉–C₁₁ naphthenic solvent) contain predominantly heavier hydrocarbons (mainly C₈, C₉, and C₁₀ isomers). Substances in the C₆–C₉ aliphatic hydrocarbon solvents subgroup fall within these 2 extremes, providing justification for the use of health effects data from the representative UVCBs to inform the assessment of the C₆–C₉ aliphatic hydrocarbon solvents subgroup.

8.3.2.2 Toxicokinetics

Studies examining the absorption, distribution, metabolism, and elimination of the substances in the C₆–C₉ aliphatic hydrocarbon solvents subgroup were not identified. With respect to representative UVCBs, information was available for commercial hexane (US EPA 1988, as cited in OECD 2013a). In an inhalation rat study, commercial hexane was extensively metabolized (equal to or greater than 95%) and mainly eliminated through exhalation and urinary excretion. The half-life following acute (6 hours) and repeated (8 days) exposures was approximately 8 to 10 hours. Tissues with the highest concentrations upon sacrifice included thymus, adipose tissue, muscle, and skin. Due to the high metabolism and elimination of commercial hexane, the bioaccumulation in fatty tissues is expected to be low.

8.3.2.3 Neurotoxicity

With respect to the health effects data on representative UVCB mixtures, studies were found for commercial hexane. No treatment-related differences in learned behaviour responses were observed in a 6-hour study exposing Sprague-Dawley rats (n=6/sex/group) to up to 31 740 mg/m³ via inhalation (American Petroleum Institute 1990, as cited in OECD OECD 2013a). This concentration represents 75% of the Lower Explosive Limit (LEL).^{Footnote 8} Furthermore, no treatment-related effects on FOB parameters, motor activity, or neuropathology were observed when Sprague-Dawley rats (n=12/sex/group) were exposed up to the highest concentration of 31 740 mg/m³, 6 hours per day, 5 days per week, for 13 weeks (Soiefer et al. 1991).

Epidemiological neurotoxicity data associated with “hydrocarbon solvent” exposure have been identified (Abbritti et al. 1976; Wang et al. 1986, as cited in OECD 2013a; Governa et al. 1987; Mutti et al. 1982; Lewis et al. 2003; Drummond et al. 2006; Rango et al. 2006; Talibov et al. 2014; Hadkhale et al. 2017). However, none of these studies identified specific aliphatic hydrocarbon solvents subgroups. In general, these studies contained insufficient detail on substance identity and measurements to establish robust exposure estimates. Further research identifying the particular substances resulting in health effects of concern, along with robust measures of exposure, is needed.

8.3.2.4 Repeated dose toxicity

With respect to the representative UVCB mixtures, health effects data were available for C₈-C₉ isoparaffin solvent (CAS RN 90622-56-3), C₉–C₁₁ naphthenic solvent (CAS RN 90622-57-4), and commercial hexane. The studies associated with C₈-C₉ isoparaffin solvent and C₉–C₁₁ naphthenic solvent tested similar concentrations in rats and generally revealed effects following exposure at the highest concentrations (that is, 5501 and 5950 mg/m³). Effects included yellow staining of fur, reduced food intake, and increased water intake (OECD 2010d, Carrillo et al. 2018). However, since these effects were completely reversible following exposure, they were not considered for risk characterization purposes.

The studies associated with commercial hexane tested higher concentrations compared to the previous studies. In one study, F344 rats (n=10/sex/group) were exposed to 0, 3182, 10 504, or 31 652 mg/m³ (equivalent to approximately 0, 698.5, 2305.9, or 6948.4 mg/kg bw/day) of commercial hexane vapour (whole-body) for 6 hours per day, 5 days per week, for 13 weeks (American Petroleum Institute 1990 as cited in OECD 2013a). Exposure to the highest concentration resulted in increased adrenal weights, increased levels of platelets (8%), creatinine (20%), protein (6%), albumin (5%), hemorrhage/inflammation in the liver (20% of animals), kidney inflammation (up to 90% of animals), and increased relative liver (11% to 19%) and kidney weights (8%). A NOAEC of 10 503 mg/m³ (2305.9 mg/kg bw/day) was identified in the report on the basis of liver, kidney, and adrenal effects observed at the highest concentration of 31652 mg/m³.

In a similar study, B6C3F1 mice (n=10/sex/group) were exposed to 0, 3181, 9993, and 32 500 mg/m³ (0, 1156, 3631.6, or 11810.8 mg/kg bw/day) of the same substance for 6 hours per day, 5 days per week for 13 weeks (American Petroleum Institute 1990 as cited in OECD 2013a). At the highest concentration, there were significantly increased liver weights relative to body weight and increased mean corpuscular volume (2%). The report indicated that the LOAEC in mice was 32500 mg/m³, implying a NOAEC of 9993 mg/m³ (3631.6 mg/kg bw/day) on the basis of the liver effects observed at the next dose level. 

8.3.2.5 Genotoxicity

Genotoxicity was also not detected for the representative UVCB commercial hexane. It was not mutagenic in Salmonella typhimurium TA98, TA100, TA1535, TA1537, or TA1538 (American Petroleum Institute 1989 as cited in OECD 2013a), not mutagenic in an in vitro mammalian cell gene mutation test in the presence and absence of metabolic activation up to cytotoxic doses (American Petroleum Institute 1990 as cited in OECD 2013a), and not clastogenic in an in vitro chromosome aberration assay up to cytotoxic doses (Daughtrey et al. 1994 as cited in OECD 2013a). With respect to in vivo genotoxicity, commercial hexane did not result in chromosome aberrations in a micronucleus test when rats were exposed to concentrations up to 31 680 mg/m³. Overall, the available information suggests that substances in the C₆–C₉ hydrocarbons solvents subgroup are not likely to be genotoxic.

8.3.2.6 Long-term toxicity and carcinogenicity

Studies examining the effects of C₆–C₉ aliphatic hydrocarbon solvents following chronic inhalation exposure were not identified. However, chronic studies were identified for the representative UVCB mixture, commercial hexane. In a chronic inhalation study, F344 rats and B6C3F1 mice (n=50/sex/group) were exposed via inhalation (whole-body) to 0, 3172, 10 573, and 31 775 mg/m³ (equivalent to approximately 0, 731.7, 2456.6, or 7396.6 mg/kg bw/day for male rats and 0, 778.6, 2608.3, or 7941.3 mg/kg bw/day for female rats) of commercial hexane for 6h/day, 5 days/week for 2 years (Daughtrey et al. 1999). In the rats, the mid- and high-exposure groups were associated with significant decreases in body weight (up to 5% and 11%, respectively) beginning after 15 weeks of exposure. Treatment-related histopathological findings were limited to hyperplasia and inflammation of the epithelium in the nasal turbinates and larynges (indications of respiratory irritation). The authors did not assess pituitary glands in rats. This response was judged to be indicative of upper respiratory tract tissue irritation. These effects were observed in all exposed groups and a LOEC of 3172 mg/m³ (731.7 mg/kg bw/day) was considered for the local effects.

In mice, no treatment-related effects with respect to body weight gains, survival, and hematological parameters were observed up to the highest tested concentration of      31 775 mg/m³ (approximately11224.6 mg/kg bw/day for males and 11461.2 mg/kg bw/day for females) (Daughtrey et al. 1999). With respect to carcinogenicity, a significantly increased incidence of pituitary adenomas was observed in female mice at all dose levels.  At the highest concentration level of 31775 mg/m³, there was also a statistically significant increase in total hepatocellular neoplasms (benign and malignant combined, 32%) observed in female mice only compared with female control animals (14%). However, when adenomas and carcinomas were examined separately, no treatment-related increases in incidence were observed, and a higher proportion of benign tumours was observed compared with malignant carcinomas. In addition, it is unlikely that these tumours resulted from a genotoxic mode-of-action (see Section 8.3.2.5) and uterine tissue from high-exposure females showed a significant decrease in the severity of cystic endometrial hyperplasia along with a decrease in uterine cysts observed at the highest dose. The authors of the study acknowledged the fact that these adenomas and carcinomas may be related to a disruption of the hormonal balance between estrogen and progesterone. In this assessment, a LOAEC of 3172 mg/m³ (equivalent to 1135.9 mg/kg bw/day), the lowest dose tested, is estimated on the basis of the pituitary adenomas at all doses in female mice for chronic inhalation. 

8.3.2.7 Reproductive and developmental toxicity

Health effects data were also available for the representative UVCB commercial hexane. In a developmental toxicity study, pregnant Sprague-Dawley rats (n=25 dams/group) were exposed via inhalation (whole-body) to 0, 3172, 10 573, or 31 719 mg/m³ (equivalent to approximately 0, 714.4, 2381.9, or 7184.1 mg/kg bw/day) for 6 h/day from GD 6 through GD 15 (American Petroleum Institute 1989). At 10 573 mg/m³, the maternal animals exhibited a significant reduction in body weight from GD 9 to 12 but a significant increase from GD 18 to 21. At the highest concentration, there was a significant reduction in body weight gain (19%) and food consumption (13%) from GD 6 to 15. However, these changes did not persist during the post-treatment period from GD 15 to 21. No changes were observed in fetuses. The authors identified a no-observed-effect-level (NOEL) of 3172 mg/m³ on the basis of maternal toxicity observed at the next dose level of 10 573 mg/m³. In this assessment, a NOAEL of 3172 mg/m³ (equivalent to 714.4 mg/kg bw/day) is estimated on the basis of maternal body weight changes at the next dose. 

Under the same exposure conditions and at the same concentrations, a similar study was conducted on pregnant CD-1 mice (n=30/group) with commercial hexane from GD 6 through 15 (American Petroleum Institute 1989). Maternal animals were sacrificed on GD18, earlier than rats were in the previous study. No treatment-related effects on maternal weights or gestational parameters were observed. However, comparison with the previous rat study is difficult because the analyses occurred before the end of the pregnancy. Complete resorption of litters was observed in the control (3%, one litter), low- (7%, 2 litters), and high (10%, 3 litters) concentration groups. No resorption of litters was reported for the middle dose. With respect to the fetuses, there were significantly increased incidences of skeletal variations at the highest concentration (for example, unossified phalanges) in the absence of maternal toxicity, but no treatment-related effects on fetal weights, sex ratio, or internal/external/skeletal malformations were identified. The authors identified a NOEL of 10573 mg/m³ (approximately 3571.3 mg/kg bw/day) for developmental toxicity on the basis of skeletal variations observed at the next concentration level. However, based on the increase in complete litter loss, a LOAEC of 3172 mg/m³ (equivalent to 1073.1mg/kg bw/day), the lowest dose tested, is estimated in this assessment.

In a combined reproductive/developmental toxicity screening test, Sprague-Dawley male and female rats were exposed to the C₇–C₉ isoparaffin solvent via inhalation at 0, 232, 770, or 2315 mg/m³ (equivalent to 0, 50.9, 169, or 508.2 mg/kg bw/day) (OECD 2010d). Female rats were exposed for 6 h/day, 7 days/week, for approximately 6 weeks from 2 weeks prior to mating and during mating through to GD 19. Male rats were exposed beginning 2 weeks prior to mating for approximately 7 consecutive weeks (OECD 2010d). No apparent treatment-related adverse clinical effects were observed, although lower body weights were observed for males in the high dose group (< 10% compared to controls).  Reproductive performance was unaffected by treatment. No statistically significant differences in the number of implantation sites or live births were observed. However, an increase in post-implantation loss was observed in the middle and high dose groups. The highest dose group had more dead pups at day 4 of lactation. No further details were given. In this assessment, the NOAEC for parental effects (males) was 770 mg/m³ (169 mg/kg bw/day) on the basis of male body weight effects; for maternal and developmental toxicity, the NOAEC was 232 mg/m³ (50.9 mg/kg bw/day, lowest dose tested) on the basis of the increase in post implantation loss.  

No adverse effects on maternal Sprague-Dawley rats) were observed after an in utero exposure (GD 6 to 15) up to the highest concentration (5595 mg/m³, approximately 1228.2 mg/kg bw/day) of C₇–C₉ isoparaffin. An increase in the incidence of ossification variations in fetuses was observed at the highest concentration in absence of maternal toxicity, but they were not considered to be adverse because rodents easily overcome this incidence during the lactation period (OECD 2010d; DeSesso and Scialli 2018). No further details were given.

In terms of reproductive toxicity, a two-generation study was conducted on Sprague-Dawley rats (n=25/sex/group) exposed via inhalation (whole-body) to commercial hexane vapours at 0, 3144, 10 555, and 31 786 mg/m³ (equivalent to approximately 0, 686.5, 2304.5, or 6860.5 mg/kg bw/day for males F0 and 0, 752.9, 2566.9, or 7723.0 mg/kg bw/day for females F0) for 6 hours per day, 5 to 7 days per week (American Petroleum Institute 1991). The animals (F0 generation) were exposed 10 weeks prior to mating, for 3 weeks of mating, and during gestation. Parental male animals were sacrificed after parturition, while dams continued to be exposed to commercial hexane through lactation and weaning of the pups (F1). Weanlings from the F1 generation (n = 25/sex/group) were randomly selected, underwent 8 weeks of exposure (pre-mating period), and mated to produce the F2 generation. Overall, no reproductive effects were observed in the two-generation study up to the highest concentration tested of 31 786 mg/m³. The number of deaths during the lactation period increased in the F1 litter at low (12%), middle (19%), and high dose (6%) in comparison with the control (3%). In F2 litters, the number of deaths during the lactation period was higher at the highest dose (8%) in comparison with the control (5%), low dose (4.5%), and middle dose (5%). The highest concentration was also associated with significantly reduced body weights (10%) in the F0 males beginning at week 13, but these effects were no longer observed at week 14. The F1 generation (pups and adults) and the F2 pups exhibited persistent reduced body weights (8% to 10%) at the highest dose. A LOAEC of 3144 mg/m³ (752.9 mg/kg bw/day) is estimated in this assessment based on the increase in F1 pup deaths during the lactation period.

N-hexane as a potential reproductive toxicant

N-hexane may be present in the C₆–C₉ aliphatic hydrocarbon solvents at varying concentrations.

Hexane (n-hexane) was assessed by Health Canada (HC) and Environment Canada (EC) and by ANSES. A LOAEC of 705 mg/m³, the lowest concentration tested for developmental toxicity, was estimated based on an increase in total intrauterine death (early and late resorptions) in female Swiss CD-1 mice exposed to 705, 3524, and 17622 mg/m³ by whole body inhalation (EC, HC 2009; ANSES 2014). 

N-hexane has been determined, in a Harmonised Classification and Labelling, to be a reproductive toxicant category 2 (CMR) by ECHA (ECHA 2021b) and the European harmonised classification and labelling has determined that the reproductive toxicant category 2 must apply if a substance contains more than 3% w/v n-hexane (CONCAWE 2021). The French Agency for Food, Environmental and Occupational Health and Safety (ANSES) also classified n-hexane as a reproductive toxicant and/or identified it as being a potential endocrine disruptor (ANSES 2014). In addition, the California Office of Environmental Health Hazard Assessment (OEHHA) listed n-hexane under Proposition 65 as being known to cause reproductive toxicity (OEHHA 2017).

Animal and human studies have shown that n-hexane has adverse effects on the female reproductive system.

Female mice (10/dose) were exposed to 0, 10.57, 53.22, or 267.16 mg/ m³ n-hexane (equivalent to 0, 2.6, 12.9, or 65.9 mg/kg bw/day) 4h per day, 7 days per week for 5 weeks (Liu et al. 2012). Mice in the highest dose group exhibited decreased activity and appetite; depilation; and rhabdomyolysis (muscle tissue breakdown with release of intracellular contents into circulation) and ulcers on the abdominal area. At the highest dose, one mouse died, the body weight of all animals decreased, and the diestrus phase decreased in mice. An increase in primordial follicle, atretic follicle, and corpus luteum at the highest dose of n-hexane was observed, along with a decrease of mature follicle in middle and high doses. A significant decrease of progesterone in blood was measured in all exposed groups. A non-significant dose decrease in estradiol and an increase in follicle-stimulating hormone (FSH) was observed. A significant increase of apoptotic granulosa cells in ovaries was also observed in middle and high doses. The authors of the study concluded that n-hexane may alter reproductive hormone secretion and promote apoptosis of granulosa cells, which may induce female reproductive impairment (Liu et al. 2012). In this assessment, a LOAEC of 10.57 mg/m³ (2.6 mg/kg bw/day), the lowest dose tested, was estimated on the basis of decreased progesterone at this dose and at 53.22 mg/m³ (12.9 mg/kg bw/day) as well as on an increase of apoptotic granulosa cells and a decrease in mature follicles in ovaries in female mice.

In a prenatal and postnatal study, pregnant Wistar rats (5/dose) were exposed to 0, 360, 1800, 9000, or 45 000 mg/m³ n-hexane (equivalent to 0, 52.7, 263.4, 1317.2, or 6585.8 mg/kg bw/day) by whole body inhalation from GD 1 to 20 for 4h per day (Li et al. 2014 and 2015). No impairment in the dam reproductive tract was observed. Pregnant F0 rats showed irritability and attack tendency at the highest dose only. The number of live pups per litter was significantly lower at the highest dose. The oestrus cycle of F1 females was measured from postnatal day (PND) 36 to 56. At PND 56, the concentrations of estrogen and progesterone in the granulosa cells (ovaries) were measured in the F1 females. A significantly longer duration of the different reproductive cycle phases was observed at different doses in F1 females: the proestrus stage at 360 and 1800 mg/m³, the oestrus at 1800 and 9000 mg/m³, and the diestrus at 45 000 mg/m³. A significant increase of progesterone in blood was observed at 360 and 1800 mg/m³ in F1 females while a significant decrease was seen at 45 000 mg/m³. A significant increase in estrogen was measured only at 9000 and 45 000 mg/m³. The study authors concluded that a prenatal exposure of n-hexane may have induced a longer oestrus cycle via an endocrine dysfunction in F1 female rats even without exposure after birth (Li et al. 2014, 2015).

In a cross-sectional study, hydrocarbons were assessed for potential reproductive endocrine effects in female U.S. Air Force personnel with fuel (primarily JP-8 jet fuel) and solvent exposures (n=63) (Reutman et al. 2002). The study authors compared women exposed to low aliphatic (mean 70.1 ppb of C₆-C₁₆) and high aliphatic (mean 279.6 ppb of C₆-C₁₆) exposure by measuring the concentration of each aliphatic in exhaled breath (Reutman et al. 2002). Among the total aliphatics in both groups, hexane (C₆), heptane (C₇), and decane (C₁₀) had the highest representation.  Hormones were measured in urine and reproductive information was obtained through a questionnaire (Reutman et al. 2002). A regression analysis showed that pre-ovulatory luteinizing hormone (LH) levels were lower in woman in both exposure groups with a total aliphatic concentration superior to the median (Reutman et al. 2002). The authors suggested that aliphatic hydrocarbons may act as reproductive endocrine disruptors, especially on the hypothalamic-pituitary-ovary axis.

In a recent cross-sectional study, the reproductive endocrine effects of n-hexane as a solvent were measured in a group of women exposed to n-hexane in an occupational setting in a leather shoe factory (without personal protection equipment, n=34) and in a control group of women without exposure to solvent (n=32), aged 18 to 37 years (Ruiz-Garcia et al. 2020). For both groups, women presenting any antecedent or current signs of thyroid disorders, polycystic, ovary syndrome, endocrine disease history in the previous 6 months, hysterectomy, oophorectomy, pregnancy, breast feeding at time of current study; and clinical evidence of chronic disease or current infectious disease, treatment with anxiolytics, antidepressants, β-blockers, Ca⁺⁺ channel blockers, hypnotic medication, or hormone replacement therapy as well as use of hormonal contraceptives were not included in this study. Daily exposure of n-hexane was measured in the urine before and after the work shift in both groups. A mean of 49.7 mg/m³ n-hexane, which is the compound with the highest exposure level, was estimated in the workers in a leather shoe factory with a cumulative average of 7.8 years of exposure. In a questionnaire, more women in the n-hexane group reported having oligomenorrhea as well as experiencing a longer time (more than 6 months) getting pregnant. A positive correlation was found between a longer menstrual cycle and “time for getting pregnant” in the n-hexane group. A negative association between FSH and the n-hexane metabolite serum level was observed in participants presenting oligomenorrhea in the exposed group. The study authors concluded that an occupational exposure to n-hexane may induce menstrual disorders and subfertility by acting as an endocrine disruptor on the hypothalamic-pituitary-ovary axis (Ruiz-Garcia et al. 2020).

8.3.3 Characterization of risk to human health

On the basis of the available health effects data, substances in the C₆–C₉ aliphatic hydrocarbon solvents may impair female reproductive health.

With respect to the inhalation route, health effects were observed in studies examining reproductive and developmental toxicity in commercial hexane. These included effects on total litter resorption (American Petroleum Institute 1989), post-implantation loss (American Petroleum Institute 1989, OECD 2010d), mortality during the lactation period in 2 generations and reduced body weight (American Petroleum Institute 1991; OECD 2010d), fetal development (for example, skeletal variations, reduced fetal/pup weights) (American Petroleum Institute 1989;  OECD 2010d), and general toxicity (for example, body weight changes, histopathological findings in the liver) (American Petroleum Institute 1990 as cited in OECD 2013a).

The developmental toxicity study conducted in CD-1 mice was selected for the characterization of risk following short- and long-term inhalation and oral and dermal exposures to C₆–C₉ aliphatic hydrocarbon (incorporating a dermal absorption factor of 25%). A LOAEC of 3172 mg/m³ (equivalent to 1073.1 mg/kg bw/day), lowest dose tested, is estimated on the basis of an increase in complete litter loss.

For environmental media and chronic exposures, the LOAEC of 3172 mg/m³ (equivalent to 1135 mg/kg bw/day), lowest dose tested, from the 2-year carcinogenicity study was also selected as the POD on the basis of the pituitary adenomas at all doses in female mice for chronic inhalation.

N-hexane is classified as a reproductive toxicant category 2 (CMR) and may be present in the C₆–C₉ aliphatic hydrocarbons grouping. On the basis of this classification, the reproductive toxicant category 2 needs to apply if one of the subgroup 3 substances (C₆–C₉ aliphatic hydrocarbons) contains more than 3% w/v n-hexane (CONCAWE 2021).

8.3.3.1 Exposure from environmental media and food

The estimated annual inhalation dose for the highest release source of C₆–C₉ aliphatic solvents for the general population living in the vicinity of a non-petroleum facility was 0.062 mg/kg bw/day (adults 19+) to 0.22 mg/kg bw/day (infant 1 yr), assuming complete inhalation absorption 

As described in Section 7.2.2, there is the potential for C₆–C₉ aliphatic solvents to be released to water bodies via industrial releases and down-the-drain scenarios. There is a potential for exposure if these substances are released into water bodies that become a source of drinking water. The 95^th percentile concentrations in surface water from industrial releases and down-the-drain sources are 36 μg/L and 12 μg/L, respectively. The predicted drinking water intakes for the LBPNs resulting from potential industrial releases ranged from 0.00077 mg/kg bw/day (adults) to 0.0012 mg/kg bw/day (infant 1 yr).

The combined exposure from ambient air and drinking water is 0.063 mg/kg bw/day (adults 19+) to 0.22 mg/kg bw/day (infant 1 yr). Compared to the LOAEL of 1135 mg/kg bw/day for non-genotoxic carcinogenic effects, these exposures give MOEs of 2300 to 7900. These MOEs are considered adequate and therefore the risk to the general population in Canada from environmental media exposure to C₆–C₉ aliphatic solvent LBPN substances from non-petroleum industrial facilities is considered to be low.

Dietary exposure, if any, from use of some of the subgroup 3 substances as a component in the manufacture of food packaging materials or as a component of incidental additives is considered to be negligible.

8.3.3.2 Exposure from products available to consumers

The combined inhalation and dermal exposures (adjusted for 25% dermal absorption) to the C₆–C₉ hydrocarbon solvents from cosmetics and products available to consumers and MOEs, based on the critical point of departure LOAEL = 1073.1 mg/kg bw/day for developmental toxicity are presented in Table 8‑8.

Abbreviations: LOAEL, lowest observed adverse effect level; MOE, margins of exposure.
^a The exposures identified are for adults unless otherwise noted.
^b For ventilation rate of 0.5 h^-1.
^c For ventilation rate of 2.5 h^-1.

The resulting margins of exposure (MOEs) for some products available to consumers as listed in Table 8‑8 are below 300 and considered to be inadequate to address the uncertainty in both the available hazard and exposure data for some of the products.

The vulnerable subpopulation of children and women of reproductive age (that is, teens and adult women) and the sub-population living within the vicinity of facilities are considered in the exposure calculations, health effects assessment, and risk characterization of the subgroup 3 substances.

8.4 Subgroup 4 (no identified uses in products available to consumers)

The 10 substances in this subgroup were originally characterized with other LBPNs in this assessment on the basis of indications that they were used in products available to consumers. However, extensive searches of available databases did not identify uses in products available to consumers, food packaging materials, or incidental additives in Canada. Therefore, only industrial uses of these substances are considered. Environmental media are the only source of general population exposure to these substances.

On the basis of their physical-chemical properties and the accompanying hazard profiles, subgroup 4 LBPN substances were placed in the existing categories of LBPNs discussed in previous sections. In particular, according to the information provided in Table 8‑9, these substances were given hazard profiles categorized in subgroups 1, 2, and 3 which were described previously.

8.4.1 Exposure assessment

8.4.1.1 Environmental media and food

The primary route of general population exposure to the substances in subgroup 4 is inhalation from environmental media releases during the industrial use of these substances. Exposure from potential industrial releases of these substances to ground water is also possible. In addition, there is a potential for exposure to the general population from uses of subgroup 4 substances in other chemical industries. Information regarding releases of chemicals from industry in Canada is collected through CEPA Section 71 surveys and the NPRI. As none of the 10 subgroup 4 substances were subject to a CEPA Section 71 survey or NPRI reporting requirements, information on volumes of these substances released to air/water are not available. Only industrial uses outlined in Table 4-7 are expected to lead to any releases to air/water. As a conservative (upper bounding) estimate of the air releases of subgroup 4 substances during non-petroleum refinery or upgrader uses in industry, the air releases of high-use subgroup 1 (C₉–C₁₄ aliphatic solvents) and subgroup 2 (C₉ aromatic solvents) LBPNs included in the NPRI and summarized in Table 4-2 are used. Since subgroups 1 and 2 have significantly higher usage categories and volumes compared with subgroup 4 substances from similar sectors, the use of air and water release values from these 2 subgroups as representative of releases of subgroup 4 substances is expected to be protective of the general population. For non-petroleum refinery industrial uses of high-use subgroup 1 and 2 LBPNs, the most recent NPRI data from 2017 indicate that air release amounts are in the range of about 140 000 kg/year or less. Some facilities that release high-use LBPNs are close to residential areas. Considering the facility locations and air release information for the subgroup 1, 2, and 3 LBPNs reported in the NPRI, exposures of the general population in the vicinity are determined using the screening-level air dispersion model SCREEN3. The distribution of locations of non-petroleum facilities that use subgroup 4 LBPNs with respect to residential areas is similar to that of the other 3 LBPN subgroups, and it is a conservative estimation to use the SCREEN3 exposure estimate with similar exposure factors for the subgroups. Dermal exposures of the subgroup 4 substances to the general population are not expected. 

The exposure factors for non-petroleum industrial releases of subgroup 4 substances that result in the highest exposure concentration are summarized in Table G-1 of Appendix G. These factors are chosen as the most conservative values given that the facility in question has a large release mass for the high-use LBPN and is located close to residential areas. The maximum yearly concentrations of the LBPN from this facility, at a distance of 140 m from the fence line, are 378 μg/m³. This results in inhalation exposure doses that range from 0.078 mg/kg bw/day to 0.38 mg/kg bw/day for the different subpopulations (adults and infant 1 yr).

Limited data were available on releases to water for certain substances in subgroup 4 from information submitted to a CEPA section 71 survey (Environment Canada 2012b). The largest quantity reported by a single facility reporting releases of any of these substances to water was 10 000 kg. As described previously (see Section 8.1.1.1), a conservative estimate of drinking water intakes was calculated using the 95% percentile for LBPNs in surface water from releases to wastewater from industrial releases and down-the-drain sources. The predicted drinking water intakes for the LBPNs resulting from potential industrial releases ranged from 0.00077 mg/kg bw/day (adults) to 0.00311 mg/kg bw/day (formula-fed infants, 0 to 5 months). These values are expected to be representative of potential releases of LBPN substances in all subgroups (Environment Canada 2012b).

There are no food packaging materials, incidental additives, or products available to consumers that contain any of the subgroup 4 substances; therefore, exposures to the general population from this source are not expected.

8.4.2 Health effects assessment

On the basis of their chemical characterization as given in Tables 2-4 and 8-11 and their physical-chemical properties, the health effect assessment for the 10 substances in subgroup 4 can be related to the C₉–C₁₄ hydrocarbon solvents (subgroup 1), the C₉ aromatic solvents (subgroup 2), and the C₆–C₉ aliphatic solvents (subgroup 3).

8.4.3 Characterization of risk to human health

Exposure of the general population to substances in subgroup 4, if any, would occur via environmental media as there are no food packaging materials, incidental additives or products available to consumers that contain these substances.

The estimated annual inhalation dose at a residential area resulting from the facility release of any of the LBPNs identified in this assessment is estimated to be 378 μg/m³. This leads to inhalation exposure doses in the range of 0.078 mg/kg bw/day to 0.38 mg/kg bw/day for the different sub populations (adults and infant 1 yr).

Estimated exposures to the most highly exposed sub population (infants 1yr) were compared with hazard endpoints for subgroups 1 to 3 and are summarized in Table 8‑10. Comparing these exposures with the NOAEL of 1115.7 mg/kg bw/day for neurotoxic effects from chronic exposure for the subgroup 1 results in an MOE of 2900, which is considered to be adequate to address the uncertainty in both the available hazard and exposure data. Comparison with the NOAEL of 111.1 mg/kg bw/day for maternal and developmental effects in mice for subgroup 2 and 1073 mg/kg bw/day for developmental effects in mice for subgroup 3 results in MOEs of 1050 and 9900, respectively, both of which are also considered to be adequate to address the uncertainty in both the available hazard and exposure data.

As described previously (see Section 8.1.1.1), a conservative estimate of drinking water intakes was calculated using the 95^th percentile PECs from industrial and down-the-drain releases. The predicted drinking water intakes for the LBPNs resulting from potential industrial releases ranged from 0.00077 mg/kg bw/day (adults) to 0.00311 mg/kg bw/day (formula fed infants, 0 to 5 months).

Estimated exposures to the most highly exposed sub population (formula-fed infants, 0 to 5 months) were compared with hazard endpoints for subgroups 1 to 3 and are summarized in Table 8‑10. Comparison with the LOAEL of 513.5 mg/kg bw/day for neurotoxicological effects from subgroup 1, the NOAEL of 111.1 mg/kg bw/day for maternal and developmental effects in mice for subgroup 2, and the NOAEL of 1073.1 mg/kg bw/day for developmental effects in mice for subgroup 3 results in MOEs of 166 000, 36 000, and 340 000, respectively. These MOEs are all considered to be adequate to address the uncertainty in both the available hazard and exposure data.

As actual exposures to the subgroup 4 substances are expected to be less than or equal to the highest exposures identified for the sungroup 1 to 3 substances, the risk to the general population in Canada from inhalation exposure to subgroup 4 LBPN substances from non-petroleum industrial facilities is considered to be low.

The vulnerable subpopulation of children and women of reproductive age (that is, teens and adult women), and the sub-population living within the vicinity of the facilities are considered in the exposure calculations, health effects assessment, and risk characterization of the subgroup 4 substances.

8.5 Uncertainties in evaluation of risk to human health

The key sources of uncertainty are presented in Table 8‑11 below.

+ = uncertainty with potential to cause over-estimation of exposure/risk; - = uncertainty with potential to cause under-estimation of exposure risk; +/- = unknown potential to cause over or under estimation of risk.

9. Conclusion

Considering all available lines of evidence presented in this draft assessment, there is a low risk of harm to the environment from the 27 LBPNs in this assessment. It is proposed to conclude that the 27 substances in the LBPNs Group do not meet the criteria under paragraphs 64(a) or 64(b) of CEPA as they 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.

Considering all the information presented in this draft assessment, it is proposed to conclude that the 9 C₉–C₁₄ hydrocarbon solvents (subgroup 1; CAS RNs 8030-30-6, 8032-32-4, 8052-41-3, 64475-85-0, 64741-41-9, 64741-65-7, 64742-48-9, 64742-82-1, and 64742-88-7), 1 C₉ aromatic solvents (subgroup 2; CAS RN 64742-95-6) and 7 C₆–C₉ aliphatic solvents (subgroup 3; CAS RN 64741-66-8, 64741-84-0, 64742-49-0, 64742-89-8, 68410-97-9, 68647-60-9, and 426260-76-6), which occur in products available to consumers, meet the criteria under paragraph 64(c) of CEPA as they are entering or may enter the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health. Considering all the information presented in this draft assessment, it is proposed to conclude that the 10 LBPNs in subgroup 4 with no consumer uses identified (CAS RNs 64741-68-0, 64741-92-0, 64741-98-6, 68333-81-3, 68512-78-7, 68513-03-1, 68553-14-0, 68603-08-7, 68920-06-9, and 70693-06-0) do not meet the criteria under paragraph 64(c) of CEPA as they 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 to conclude that the 17 LBPNs in C₉–C₁₄ hydrocarbon solvents (subgroup 1; CAS RNs 8030-30-6, 8032-32-4, 8052-41-3, 64475-85-0, 64741-41-9, 64741-65-7, 64742-48-9, 64742-82-1, and 64742-88-7), C₉ aromatic solvents (subgroup 2; CAS RN 64742-95-6), and C₆–C₉ aliphatic solvents (subgroup 3; CAS RN 64741-66-8, 64741-84-0, 64742-49-0, 64742-89-8, 68410-97-9, 68647-60-9, and 426260-76-6) meet one or more of the criteria set out in section 64 of CEPA. It is therefore proposed to conclude that the 10 LBPNs in subgroup 4 with no consumer uses identified (CAS RNs 64741-68-0, 64741-92-0, 64741-98-6, 68333-81-3, 68512-78-7, 68513-03-1, 68553-14-0, 68603-08-7, 68920-06-9, and 70693-06-0) do not meet any of the criteria set out in section 64 of CEPA.

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Appendices

Appendix A. Substance identity information for the LBPNs group

Appendix B. Physical and chemical properties of substances in the LBPNs group

Abbreviations: CAS RN, Chemical Abstracts Service Registry Number; DSL, Domestic Substances List; log K_ow, octanol-water partition coefficient; NA, not applicable.
Note: All LBPNs are liquids at ambient temperature (15°C to 25°C).
^a Range applies to CAS RNs 8032-32-4, 64741-84-0, 64742-48-9, 64742-49-0, and 64742-89-8 (C₇–C₉ aliphatic hydrocarbon solvents category). LBPNs in the OECD hydrocarbon solvents category are based on a narrower distillation range than those of the original refinery product with the same CAS RN (that is, are “narrow cut” products), and physical-chemical property values for these products do not represent the full range of properties for the CAS RN.
^b Range applies to CAS RNs 8052-41-3 and 64742-82-1 (C_9–14 aliphatic [2% to 25% aromatics] hydrocarbon solvents category). LBPNs in the OECD hydrocarbon solvents category are based on a narrower distillation range than those of the original refinery product with the same CAS RN (that is, are “narrow cut” products), and physical-chemical property values for these products do not represent the full range of properties for the CAS RN.
^c Boiling range 88°C to 192°C.
^d Range applies to all substances in the C_9–14 aliphatic [≤2% aromatic] hydrocarbon solvents category, which includes CAS RNs 64741-65-7 and 64742-48-9. LBPNs in the OECD hydrocarbon solvents category are based on a narrower distillation range than those of the original refinery product with the same CAS RN (that is, are “narrow cut” products), and physical-chemical property values for these products do not represent the full range of properties for the CAS RN.
^e Range applies to all substances in the C₉ aromatic hydrocarbon solvents category, which includes CAS RN 64742-95-6. LBPNs in the OECD hydrocarbon solvents category are based on a narrower distillation range than those of the original refinery product with the same CAS RN (that is, are “narrow cut” products), and physical-chemical property values for these products do not represent the full range of properties for the CAS RN.
^f ECHA c2007-2018.
^g Range applies to all substances in the C_10–13 aromatic hydrocarbon solvents category, which includes CAS RN 70693-06-0. LBPNs in the OECD hydrocarbon solvents category are based on a narrower distillation range than those of the original refinery product with the same CAS RN (that is, are “narrow cut” products), and physical-chemical property values for these products do not represent the full range of properties for the CAS RN.

Appendix C. Substance identity

N/A: Not available.
* Representative UVCB.

N/A: Not available.
* Representative UVCB.

Appendix D. Exposure parameters used to estimate exposure to subgroup 1 LBPN substances

^a Professional judgement on the basis of aerial photo analysis
^b Estimated on the basis of the highest quantity of tetrahydrofuran released during the 2017 reporting year by a single submitter and an assumed continuous release (NPRI 2019)
^c Curry et al. (1993)
^d Regulatory default from the SCREEN3 model

N/A = Not applicable
^a As cited in the Canadian exposure factors used in human health risk assessment (Health Canada 2019).

Appendix E. Exposure parameters used to estimate exposure to subgroup 2 (C₉ Aromatic hydrocarbon solvent)

Sentinel exposure scenarios were used to estimate the potential human exposure to the single C₉ aromatic hydrocarbon solvent; scenario assumptions are summarized in Tables D-1 to D-4. Exposures were estimated using ConsExpo Web (RIVM 2016) or algorithms from the model, unless otherwise indicated. Inhalation estimates were converted to internal doses using inhalation rates from US EPA (2011a) and body weights from Health Canada (2015b) for various age groups, where relevant.

Input parameters for the SCREEN3 model used to estimate ambient air concentrations of the C₉ aromatic hydrocarbon substance from the largest emitter reported in 2017 are outlined in Table D-1 below.

^a Professional judgment based on aerial photo analysis
^b Estimation based on the highest quantity of tetrahydrofuran released during the 2017 reporting year by a single submitter and an assumed continuous release (NPRI 2019)
^c Curry et al. (1993)
^d Regulatory default from the SCREEN3 model

^a See Table C-2 of Appendix C for Inhalation rate and body weight information (Health Canada 2019).
^b On the basis of Ficheux et al. (2014, 2016).
^c Mean concentration on day of exposure (that is, average air concentration over the day and accounts for the number of events on one day) was modelled using ConsExpo exposure to vapour values, evaporation model, and the following parameters: exposure duration of 35 min, molecular weight matrix of 124 g/mol, room volume of 1 m³, ventilation rate of 1/h, vapour pressure of 405 Pa, molecular weight of 120.19 g/mol, mass transfer coefficient of 10 m/h, release area of 26.2 cm² for 19+ and 14–18 years; 8.8 cm² for 9–13 years, 4–8 years, and 2–3 years, emission duration of 35 min, and application temperature of 20°C.
^d Internal dose (mg/kg bw/day) = (Mean concentration on day of exposure [mg/m³] / body weight [kg]) × Room volume [m³].

^a See Table C-2 of Appendix C for Inhalation rate and body weight information (Health Canada 2019).
^a See Table C-2 of Appendix C for Inhalation rate and body weight information (Health Canada 2019).
^b Based on Ficheux et al. (2014)
^c Mean concentration on day of exposure (that is, average air concentration over the day and accounts for the number of events on one day) was modelled using ConsExpo exposure to vapour values, evaporation model, and the following parameters: exposure duration of 35 min, molecular weight matrix of 124 g/mol, room volume of 1 m³, ventilation rate of 1/h, vapour pressure of 405 Pa, molecular weight of 120.19 g/mol, mass transfer coefficient of 10 m/h, release area of 26.2 cm² for 19+ and 14–18 years; 8.8 cm² for 9–13 years, emission duration of 35 min, and application temperature of 20°C.
^d Internal dose (mg/kg bw/day) = (Mean concentration on day of exposure [mg/m³] / body weight [kg]) × Room volume [m³].

Appendix F. Exposure parameters used to estimate exposure to subgroup 3 (C₆–C₉ aliphatic hydrocarbon solvent)

Sentinel exposure scenarios were used to estimate the potential human exposure to the single C₆–C₉ aliphatic solvents; scenario assumptions are summarized in Tables E-1 to E-3. Exposures were estimated using ConsExpo Web (RIVM 2016) or algorithms from the model, unless otherwise indicated. Inhalation estimates were converted to internal doses using inhalation rates from US EPA (2011a) and body weights from Health Canada (2015b) for various age groups, where relevant.    

Input parameters for the SCREEN3 model used to estimate ambient air concentrations of the subgroup 3 C₆–C₉ aliphatic solvents substance from the largest emitter reported in 2017 are outlined in Table F-1 below.

^a Professional judgment based on aerial photo analysis.
^b Estimate based on the highest quantity of tetrahydrofuran released during the 2017 reporting year by a single submitter and an assumed continuous release (NPRI 2019).
^c Curry et al. (1993)
^d Regulatory default from the SCREEN3 model.