Final Screening Assessment

Petroleum Sector Stream Approach

Heavy Fuel Oils
[Industry-Restricted]
Chemical Abstracts Service Registry Numbers

64741-75-9
68783-08-4
70592-76-6
70592-77-7
70592-78-8

Environment Canada
Health Canada
July 2013

(PDF Format - 584 KB)

Table of contents

• Synopsis
• Introduction
  • Grouping of Petroleum Substances
• Substance Identity
• Physical and Chemical Properties
• Sources
• Uses
• Releases to the Environment
  • Potential On-site Releases
  • Potential Releases from Transportation
• Environmental Fate
• Persistence and Bioaccumulation Potential
  • Environmental Persistence
  • Potential for Bioaccumulation
• Potential to Cause Ecological Harm
  • Ecological Effects Assessment
  • Ecological Exposure Assessment
  • Characterization of Ecological Risk
  • Uncertainties in Evaluation of Ecological Risk
• Potential to Cause Harm to Human Health
  • Exposure Assessment
  • Health Effects Assissment
  • Characterizaion of Risk to Human Health
  • Uncertainties in Eavluation of Risk to Human Health
• Conclusion
• References
• Appendices

Synopsis

The Ministers of the Environment and of Health have conducted a screening assessment of the following industry-restricted heavy fuel oils (HFOs):

These substances were identified as high priorities for action during the categorization of substances on the Domestic Substances List (DSL), as they were determined to present greatest potential or intermediate potential for exposure of individuals in Canada, and were considered to present a high hazard to human health. These substances met the ecological categorization criteria for persistence or bioaccumulation potential and inherent toxicity to aquatic organisms. These substances were included in the Petroleum Sector Stream Approach (PSSA) because they are related to the petroleum sector and are considered to be of Unknown or Variable composition, Complex reaction products or Biological materials (UVCBs).

These HFOs are a group of complex combinations of petroleum hydrocarbons that serve as blending stocks in final fuel products or as intermediate products of distillation or residue derived from refinery distillation or cracking units. The final fuel products usually consist of a blend of HFOs, as well as higher-quality hydrocarbons as diluents. The HFOs considered in this assessment are complex combinations composed of aromatic, aliphatic and cycloalkane hydrocarbons with carbon ranges spanning C₇–C₅₀ and a typical boiling point range of 121–600°C. In order to predict the overall behaviour of these complex substances for the purposes of assessing the potential for ecological effects, representative structures have been selected from each chemical class in the substances.

The HFOs considered in this screening assessment have been identified as industry restricted (i.e., they are a subset of HFOs that may leave a petroleum sector facility and be transported to other industrial facilities). According to information submitted under section 71 of the Canadian Environmental Protection Act, 1999 (CEPA 1999), and other sources of information, these HFOs are transported in large volumes from refinery or upgrader facilities to other industrial facilities by pipelines, ships, trains and trucks; therefore, exposure of the environment is expected.

Based on results of comparison of levels expected to cause harm to organisms with estimated exposure levels and the relatively low expected frequency of spills to water and soil during loading/unloading and transport operations, these five HFOs have a low risk of harm to aquatic or soil organisms.

Based on the information presented in this screening assessment on the frequency and magnitude of spills, there is low risk of harm to organisms or the broader integrity of the environment from these substances. It is concluded that the industry-restricted HFOs (CAS RNs 64741-75-9, 68783-08-4, 70592-76-6, 70592-77-7 and 70592-78-8) do not meet the criteria under paragraph 64(a) or 64(b) of the Canadian Environmental Protection Act, 1999 (CEPA 1999) 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 effect for the initial categorization of industry-restricted HFO substances was carcinogenicity, based primarily on classifications by international agencies. Several cancer studies conducted in laboratory animals resulted in the development of skin tumours following repeated dermal application of HFO substances. HFOs demonstrated genotoxicity in in vivo and in vitro assays and may also adversely affect reproduction and development in laboratory animals when applied dermally. There are no carcinogenicity studies by the inhalation route to inform the carcinogenic potential of these substances in the general population following inhalation exposure. Information on additional HFO substances in the PSSA that are similar from a processing and physical-chemical perspective was considered for characterization of human health effects.

General population exposure to industry-restricted HFOs results primarily from inhalation of ambient air containing HFO vapours due to evaporative emissions during transportation. Due to the relatively low volatility of the HFO substances, as defined by their physical-chemical properties, evaporative emissions into the air are expected to be minimal. The margins between the upper-bounding estimate of exposure, the maximum air concentration of HFOs (1.28 µg/m³), and the critical inhalation effect levels are considered to be highly conservative and adequately protective to account for uncertainties related to health effects and exposure. The likelihood of inhalation exposure of the general population is considered to be low; thus, the risk to human health is likewise considered to be low. General population exposure to industry-restricted HFOs via the oral and dermal routes is not expected; therefore, risk to human health from the industry-restricted HFOs via these routes is not expected.

Based on the information presented in this screening assessment, it is concluded that the industry-restricted HFOs (CAS RNs 64741-75-9, 68783-08-4, 70592-76-6, 70592-77-7 and 70592-78-8) do not meet the criteria under paragraph 64(c) of the Canadian Environmental Protection Act, 1999 (CEPA 1999) 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 concluded that the five industry-restricted HFOs listed under CAS RNs 64741-75-9, 68783-08-4, 70592-76-6, 70592-77-7 and 70592-78-8 do not meet any of the criteria set out in section 64 of CEPA1999.

Introduction

The Canadian Environmental Protection Act, 1999 (CEPA1999) (Canada 1999) requires the Minister of the Environment and the Minister of Health to conduct screening assessments of substances that have met the categorization criteria set out in the Act to determine whether these substances present or may present a risk to the environment or to human health.

Based on the information obtained through the categorization process, the Ministers identified a number of substances as high priorities for action. These include substances that:

• met all of the ecological categorization criteria, including persistence (P), bioaccumulation potential (B) and inherent toxicity to aquatic organisms (iT), and were believed to be in commerce in Canada; and/or
• met the categorization criteria for greatest potential for exposure (GPE) or intermediate potential for exposure (IPE) and had been identified as posing a high hazard to human health based on classifications by other national or international agencies for carcinogenicity, genotoxicity, developmental toxicity or reproductive toxicity.

A key element of the Government of Canada’s Chemicals Management Plan is the Petroleum Sector Stream Approach (PSSA), which involves the assessment of approximately 160 petroleum substances that are considered high priorities for action. These substances are primarily related to the petroleum sector and are considered to be of Unknown or Variable composition, Complex reaction products or Biological materials (UVCBs).

Screening assessments focus on information critical to determining whether a substance meets the criteria set out in section 64 of CEPA1999. Screening assessments examine scientific information and develop conclusions by incorporating a weight-of-evidence approach and precaution^[1] .

Grouping of Petroleum Substances

The high priority petroleum substances fall into nine groups of substances (Table A1.1 in Appendix 1) based on similarities in production, toxicity and physical-chemical properties. In order to conduct the screening assessments, each high priority petroleum substance was placed into one of five categories (“streams”) depending on its production and uses in Canada:

An analysis of the available data determined that 16 petroleum substances are evaluated under Stream 2, as described above. These occur within five of the nine substance groupings: heavy fuel oils (HFOs), gas oils, petroleum and refinery gases, low boiling point naphthas and crude oils.

This screening assessment addresses five industry-restricted HFO substances described under Chemical Abstracts Service Registry Numbers (CAS RNs) 64741-75-9, 68783-08-4, 70592-76-6, 70592-77-7 and 70592-78-8. These substances were identified as GPE or IPE during the categorization exercise, and were considered to present a high hazard to human health. These substances met the ecological categorization criteria for persistence or bioaccumulation potential and inherent toxicity to aquatic organisms. According to information submitted under section 71 of CEPA 1999 (Environment Canada 2008, 2009), these substances can be consumed on-site or transported from refineries and upgraders to other industrial facilities, but they are not sold directly to consumers. These substances were included in the PSSA because they are related to the petroleum sector and are all complex combinations of petroleum hydrocarbons.

Seven site-restricted HFOs were previously assessed under Stream 1, and an additional nine HFOs are being assessed separately, as they belong to Streams 3 and 4 (as described above). The health effects of the industry-restricted HFOs were assessed using health effects data pooled across all high priority HFOs due to insufficient data specific to the industry-restricted HFOs.

Included in this Stream 2 screening assessment is the consideration of information on chemical properties, uses, exposure and effects, including the additional information submitted under section 71 of CEPA 1999. Data relevant to the screening assessment of these substances were identified in original literature, review and assessment documents, and stakeholder research reports and from recent literature searches, up to March 2010 for the human exposure and environmental sections of the document and up to September 2011 for the health effects section of the document. Key studies were critically evaluated, and modelling results were used to reach conclusions.

Characterizing risk to the environment involves the consideration of data relevant to environmental behaviour, persistence, bioaccumulation and toxicity, combined with an estimation of exposure of potentially affected non-human organisms from the major sources of release to the environment. To predict the overall environmental behaviour and properties of complex substances such as these industry-restricted HFOs, representative structures were selected from each chemical class contained within the substances. Conclusions regarding risk to the environment are based on an estimation of environmental concentrations resulting from releases and the potential for these concentrations to have a negative impact on non-human organisms. As well, other lines of evidence including fate, temporal/spatial presence in the environment, and hazardous properties are taken into account. The ecological portion of the screening assessment summarizes the most pertinent data on environmental behaviour and effects and does not represent an exhaustive or critical review of all available data. Environmental models and comparisons with similar petroleum substances may have assisted in the assessment.

Evaluation of risk to human health involves consideration of data relevant to estimation of exposure (non-occupational) of the general population, as well as information on health effects. Health effects were assessed using toxicological data pooled across high priority HFO substances. Decisions for risk to human health are based on the nature of the critical effect and margins between conservative effect levels and estimates of exposure, taking into account confidence in the completeness of the identified databases on both exposure and effects, within a screening context. The screening assessment does not represent an exhaustive or critical review of all available data. Rather, it presents a summary of the critical information upon which the final conclusion is based.

This screening assessment was prepared by staff in the Existing Substances Programs at Health Canada and Environment Canada and incorporates input from other programs within these departments. The human health and ecological portions of this assessment have undergone external written peer review/consultation. Comments on the technical portions relevant to human health were received from scientific experts selected and directed by Toxicology Excellence for Risk Assessment (TERA), including Dr. Michael Dourson (TERA), Dr. Stephen Embso-Mattingly (NewFields Environmental Forensics Practice, LLC), Dr. Susan Griffin (United States Environmental Protection Agency [U.S. EPA]) and Dr. Donna Vorhees (Science Collaborative). Although external comments were taken into consideration, the final content and outcome of the screening assessment remain the responsibility of Health Canada and Environment Canada.

The critical information and considerations upon which the final screening assessment is based are summarized below.

Substance Identity

These HFOs are a group of complex petroleum combinations of petroleum hydrocarbons that serve as blending constituents in final fuel products or as intermediate products of distillate or residue derived from refinery distillation or cracking units with a typical carbon range of C₂₀–C₅₀ (CONCAWE 1998). The final fuel product usually consists of a blend of HFOs and high-quality hydrocarbons that have been produced in the refinery or upgrader facilities. The HFOs considered in this assessment are complex mixtures composed of aromatic, aliphatic and cycloalkane hydrocarbons with carbon ranges spanning C₇–C₅₀ and boiling point ranges of 121–600°C (Table A2.1 and A2.2 in Appendix 2; API 2004; CONCAWE 1998). The ratio of aliphatic to aromatic hydrocarbons is important for estimating environmental behaviour; however, very few data exist for these five CAS RNs, so a ratio of 50:50 has been assumed. This ratio will not bias results and is within the range of other types of HFOs (50–79% aromatics) (ATSDR 1999; API 2004).

These UVCB substances are complex combinations of hydrocarbon molecules that originate in nature or are the result of chemical reactions and processes that take place during the upgrading and refining process. Given their complex and variable compositions, they could not practicably be formed by simply combining individual constituents.

Physical and Chemical Properties

The composition and physical-chemical properties of HFOs vary depending upon the sources of crude oils or bitumen and the processing steps involved. A summary of experimental data on the physical-chemical properties of industry-restricted HFOs is presented in Table 1.

The theoretical vapour pressures of individual substances comprising HFOs are low to moderate due to their high molecular weights. However, the actual vapour pressures will be influenced by the substance composition of the HFO in which they occur. Water solubilities of all HFOs are low, and octanol–water partition coefficient estimations vary considerably due to the complex nature of these substances.

To predict the environmental behaviour and fate of complex petroleum products such as these HFOs, representative structures were selected from each chemical class contained within the mixture. Forty-seven structures were selected from a database in PETROTOX (2009) based on boiling point ranges for each HFO (Table A2.3 in Appendix 2), the number of data on each structure and the middle of the boiling point range of similar structures. As the compositions of most HFOs are not well defined and are indeed variable, representative structures could not be chosen based on their proportion in the mixture. This resulted in the selection of representative structures for alkanes, isoalkanes, one-ring cycloalkanes, two-ring cycloalkanes, polycycloalkanes, cycloalkane monoaromatics, cycloalkane diaromatics and one-, two-, three-, four-, five- and six-ring aromatics ranging from C₉–C₅₀ (Table A2.4 in Appendix 2). Physical-chemical data for each representative structure were assembled from scientific literature and from the group of environmental models included in the U.S. EPA’s Estimation Programs Interface Suite (EPI Suite 2008) (Table A2.4 in Appendix 2).

Sources

Industry-restricted HFOs are produced in Canadian refineries and upgraders. The CAS RN descriptions (NCI 2006) and typical process flow diagrams (Hopkinson 2008) indicate the origin of these HFOs. Information submitted under section 71 of CEPA 1999 shows that these substances can be intermediate streams consumed within a facility or be transported off-site by pipeline, truck, train and ship for use as a feedstock in other industrial facilities or for disposal (Environment Canada 2008, 2009).

CAS RN 64741-75-9 is a residual fraction from distillation of hydrocracking effluents in both refineries and upgraders.

CAS RN 68783-08-4 is a general description of distillates from atmospheric distillation of crude oil in refineries, primarily ranging from C₇–C₃₅.

CAS RNs 70592-76-6, 70592-77-7 and 70592-78-8 have slight differences in their dominant carbon range, but they all refer to distillates from vacuum fractionation of the residue produced from atmospheric distillation of crude oil.

Uses

According to the information collected through the Notice with respect to certain high priority petroleum substances published under section 71 of CEPA 1999 (Environment Canada 2008) and the Notice with respect to potentially industry-limited high priority petroleum substances (Environment Canada 2009), these industry-restricted HFO substances have been identified as being consumed at the facility or transferred to another industrial facility for use as feedstock or for disposal. Although these substances were identified by multiple use codes established during the development of the Domestic Substances List (DSL), it has been determined from information submitted under section 71 of CEPA 1999, voluntary submissions from industry, an in-depth literature review and a search of material safety data sheets that these industry-restricted HFOs (i.e., the CAS RNs identified in this screening assessment) may leave a refinery or an upgrading facility and be transported to another industrial facility for use as a feedstock, or for disposal, but do not reach the public market in the form originally acquired.

Releases to the Environment

Potential releases of industry-restricted HFOs consist of releases within facilities from activities associated with processing these substances, as well as releases related to transportation of these substances between industrial facilities.

Due to the complex nature of the petroleum industry and transportation industry, as well as the ambiguity in the literature in the use of the terminology that is critical to the understanding of the Stream 2 PSSA assessments, it is important that the definitions specific to the assessment of the industry-restricted petroleum substances are well understood. Table 2 lists the terminology specific to the present assessment.

Table 2. Definitions of terms specific to the PSSA assessments of industry-restricted petroleum substances

Terminology

Definition

Release

A generic term to define a leak, spill, vent, or other release of a gaseous or liquid substance, including controlled release and unintentional release, as defined below, but not including catastrophic events.

Controlled release

Any planned release for safety or maintenance purposes that is considered part of routine operations and occurs under controlled conditions.

Unintentional release

Any unplanned release of a petroleum substance. Causes can include equipment failure, poor maintenance, lack of proper operating practices, adverse weather-related events or other unforeseen factors, but can also be a routine part of normal operations. The following two categories are included under unintentional releases: (1) unintentional leaks or spills that occur from processing, handling and transport of a petroleum substance; such leaks or spills can be reduced or controlled by the industry; and (2) accidental releases that may not be controllable by the industry. Only unintentional leaks or spills (category 1 defined above) are considered in the assessment of the potential of industry-restricted petroleum substances to cause ecological harm.

Fugitive release

A specific type of unintentional release. It refers to an unintentional release, which occurs under normal operating conditions, of a gaseous substance into ambient air and which may occur on a routine basis. Fugitive releases can be reduced but may not be entirely preventable due to the substance’s physical-chemical properties, equipment design, and operating conditions. Evaporative emission during the transportation of petroleum substances is a fugitive release and is considered in the human exposure analysis for purposes of assessing the potential of the substance to cause harm to human health.

Potential On-site Releases

Potential releases of HFO substances from refineries or upgraders can be characterized as either controlled or unintentional releases. Controlled releases are planned releases from pressure relief valves, venting valves and drain systems for safety purposes or maintenance. Unintentional releases are typically characterized as unplanned releases due to spills or leaks from various equipment, valves, piping or flanges. Refinery and upgrader operations are highly regulated, and regulatory requirements are established under various jurisdictions. As well, voluntary non-regulatory measures implemented by the petroleum industry are in place to manage these releases (SENES 2009).

Controlled Releases

The industry-restricted HFOs considered in this screening assessment originate from distillation columns in a refinery or an upgrader, either as a residue (bottom product) or as a distillate. Thus, the potential locations for the controlled release of these HFOs include relief valves, venting valves and drain valves on the piping or vessels where these streams are generated.

Under typical operating conditions, controlled releases of these HFO substances would be captured in a closed system^[3] , according to defined procedures, and returned to the processing facility or to the facility’s wastewater treatment plant. In both cases, exposure of the general population or the environment to these industry-restricted HFOs is not expected.

Unintentional Releases

Unintentional releases (including fugitive releases) occur from equipment (e.g., pumps, storage tanks), seals, valves, piping, flanges, etc. during processing and handling of petroleum substances and can be greater in situations of poor maintenance or operating practices. Regulatory and non-regulatory measures are in place to reduce these events at petroleum refineries and upgraders (Appendix 3) (SENES 2009). Rather than being specific to one substance, these measures are developed in a more generic way in order to reduce unintentional releases of all substances in the petroleum sector.

Conclusion for Potential On-site Releases

Based on the information presented in this screening assessment and in the screening assessment of the Stream 1 (site-restricted) HFOs, exposure of the general population or the environment to the on-site releases (controlled or unintentional) of industry-restricted HFOs is not expected.

Potential Releases from Transportation

As these industry-restricted HFOs can be transported between facilities, releases may also occur during transportation. In general, three operating procedures are involved during the process of transportation: loading, transit and unloading.

The on-site handling of petroleum substances for transportation is often regulated at the federal and provincial/territorial levels with legislation covering loading and unloading (Appendix 3).

Storage of industry-restricted HFOs may be required before they are transported off-site. Releases of HFO vapours from the storage tanks into the air are expected to be small because the HFOs have low volatility. All relevant releases from storage, including leaks, spills and breathing loss (expulsion of vapour due to changes in temperature and pressure), will be similar to the aforementioned potential on-site releases and will be managed under the relevant legislation currently in place.

Tanks or containers for transferring petroleum substances are typically dedicated vessels; thus, washing or cleaning is not required on a routine basis (U.S. EPA 2008a; OECD 2009). As such, exposure of the general population and the environment to the HFOs considered in this screening assessment from tank cleaning is not expected. Cleaning facilities require processing of grey-water to meet local and provincial release standards.

Release Estimation

Information on the transportation quantities and relevant transportation modes was collected under section 71 of CEPA1999 (Environment Canada 2009) with respect to each CAS RN considered in this screening assessment. Four modes of transportation--ships, pipelines, trucks and trains--were identified as being involved in moving industry-restricted HFOs to other industrial facilities. The total transport quantity of the five HFOs considered in this assessment is about 3 million tonnes (2.9 × 109 L) (year 2006). Two types of potential releases occur during transportation and are considered in this screening assessment. These are evaporative emissions and unintentional releases (e.g., spills or leaks) during the handling and transit processes.

Evaporative emissions are similar to breathing loss of organic substances from storage tanks. The quantity lost depends on the volatility of the substances, temperature or pressure changes that occur during transportation, and tightness of transport vessels and settings of valves. Ambient air is the receiving medium for evaporative emissions.

Evaporative emissions to the environment were considered in transportation by ships, trucks and trains and were estimated based on empirical equations from the U.S. EPA (2008a), physical-chemical properties (e.g., vapour pressure, molecular weight and density of vapours) of these HFO substances, and the annual transported quantities. No evaporative emissions are considered for pipeline systems, as typical releases are generated as a result of leaks through seals, flanges and valves and are defined as unintentional releases.

Unintentional releases of the HFO substances due to spills generally enter water or soil, depending on the modes of transportation involved. Due to the relatively low volatility of the HFOs, as defined by their physical-chemical properties, evaporative emissions into the air from spills would occur in a lower proportion compared with the proportions entering water or soil.

Potential releases associated with the transport of these HFOs to marine, freshwater and soil environments were assessed through analysis of historical spill data (2000–2009) from the Environment Canada Spill Line database (Environment Canada 2011). There was no spills category for HFOs; spills of Bunker C fuel oil were therefore used. The releases labelled as Bunker C fuel oil (Fuel Oil No. 6) would also include these industry-restricted HFOs. There were also a small number of releases that were generically labelled as just “Bunker,” and there was no indication as to what specific type of Bunker was released. Thus, all releases labelled as “Bunker” were considered to be Bunker C fuel oil. Bunker C is considered a heavy fuel oil but is not industry-restricted and has a wider distribution. Thus, it is expected that the actual number and volume of industry-restricted HFO spills are considerably lower than those of Bunker C fuel oil spills, but this could not be reliably determined. Of note was the large-volume spill of 734 000 L in 2005, which is known to be a Bunker C fuel oil spill into Lake Wabamun, Alberta; it was not included in the release estimate, as it was known not to be an industry-restricted HFO spill. As well, extremely large spills with no known origin were not included, as these were likely from environmental emergencies training exercises, which are not differentiated from actual events in the Environment Canada Spill Line database (Environment Canada 2011). Spills where collisions, poor road conditions and/or adverse weather-related events were listed as a source, reason or cause of spill were not included in the release estimate.

Many of the individual reports had no estimate of the volume released into the environment. In order to account for the underestimation of the volume released, the estimated total volumes were extrapolated by assuming that the distribution of reported volumes released was representative of all releases (Table A4.1 in Appendix 4). From 2000–2009, the extrapolated total volume of spills of HFOs to all media (soil, salt water and fresh water) was 2.4 million litres from 339 spills (Table A4.1 in Appendix 4).

The historical spill data were also separated into the specific compartment affected, so that the estimated average release quantity per spill to each compartment could be determined. Within each compartment, a similar extrapolation was conducted to account for reported spills with no associated volumes. The estimated average quantities of these HFOs released per spill to fresh water and salt water from ship transport are shown in Table 3. These average spill volumes were based on Bunker C releases from the Spills Line database because HFOs are handled the same as Bunker C fuel oil for loading/unloading and ship transport. However, spills that were specifically known to be Bunker C were not included in the release estimate. There is no distinction in the database as to whether the spills occur during loading, transport or unloading. Thus, the average spill volume will be used for each of the scenarios.

Table 3. Average release quantities per spill of industry-restricted HFOs to various compartments (L/spill) based on historical Bunker C spill data from 2000–2009 from Environment Canada (2011)

The largest fraction of HFO spills documented by Environment Canada from 2000–2009 affected land (130 incidents), followed by 10⁸ releases to sea water and 53 releases to fresh water. For some reported spills, the compartment affected was not documented, whereas for others, multiple compartments were included; thus, this total does not equate to the total reported spills shown in Table A4.1 (Appendix 4). These numbers are considered to be a low estimate of actual releases, as not all provinces were reporting their spills to Environment Canada for all years, and some provinces have minimum reportable spill quantities. Releases to groundwater were not included in the analysis.

The Environment Canada (2011) Spill Line database provides three columns of data (sources, causes and reasons) for many releases of Bunker C fuel oil. The data in these columns were analyzed to determine how and why the majority of HFO releases occur (Tables A4.2a–c in Appendix 4).

The industrial areas where the majority of HFO releases occurred (Table A4.2a in Appendix 4) were other watercraft (25% of the volume), pipelines (20% of the volume) and marine tankers (20% of the volume). Releases at storage depots and facilities accounted for about 2% of the volume, refineries accounted for 2%, tank and transport trucks accounted for 3%, trains accounted for 4% and “other” sources accounted for 9%. The majority of truck releases were in New Brunswick (50%), and the rest were reported in Newfoundland and Labrador, Nova Scotia, Quebec, Prince Edward Island, Ontario and British Columbia.

The Environment Canada Spill Line data were also analyzed for causes of HFO leaks (Table A4.2b in Appendix 4). It was found that pipe leaks accounted for 38% of the volume released, which is consistent with pipelines being a major source of Bunker C releases (see Table A4.2a). Likewise, sinking and grounding of vessels accounted for 13% and 6% of the total volume, respectively, which is also consistent with the high total spill volume by watercraft as a source. Twenty-five percent of the volume spilled was due to unknown causes, and 8% was due to “other” causes.

Analyzing reasons for releases, the data (Table A4.2c in Appendix 4) identified material failure as a major cause of releases, accounting for 16% of the volume released. Unknown reasons accounted for 43% of the volume, human error and negligence accounted for 18%, and fire and explosion accounted for 6% of the volume (from a single spill). The remaining 17% was divided over a wide variety of reasons.

For purposes of assessing the potential exposure of the environment from the transportation of industry-restricted HFOs, the ecological assessment focuses on unintentional releases to water and soil due to spills. Releases to water contributed significantly greater volumes than releases to soil. In comparison, assessment of potential exposure of the general population from transportation of industry-restricted HFOs focuses on evaporative emission, which occurs during regular operation activities. Although spills occur during transit and in loading or unloading operations, such releases are considered to occur on a non-routine or unpredictable basis in distinct locations and are therefore not considered in the assessment of exposure of the general population.

In addition, as relevant legislation and best practices are in place for on-site handling of these industry-restricted HFOs (Appendix 3), non-occupational human exposure as a result of loading and unloading is not expected and is not considered in the human exposure assessment.

This assessment does not include illegal releases of fuel oil at sea in Canadian jurisdictions. Transport Canada has in place a National Aerial Surveillance Program to monitor and deter such releases (Transport Canada 2010).

Environmental Fate

When petroleum substances are released into the environment, four major fate processes will take place: dissolution in water, volatilization, biodegradation and adsorption. These processes will cause changes in the composition of these UVCB substances. In the case of spills on land or water surfaces, another fate process, photodegradation, can also be significant.

The rates of dissolution in water or volatilization of individual petroleum components are retarded by the complex nature of these petroleum mixtures. The solubility and volatility of individual components in mixtures are proportional to the solubility or volatility of the components in its pure state and its concentration in the mixture. Solubility and volatility of a component decrease when the component is present in a mixture (Banerjee 1984; Potter and Simmons 1998).

Each of the fate processes affects hydrocarbon families differently. Aromatics tend to be more water soluble than aliphatics of the same carbon number, whereas aliphatics tend to be more volatile (Gustafson et al. 1997). Thus, when a petroleum mixture is released into the environment, the principal water contaminants are likely to be aromatics while aliphatics will be the principal air contaminants (Potter and Simmons 1998). The trend in volatility by component class is as follows: alkenes ≈ alkanes greater than aromatics ≈ cycloalkanes. The most soluble and volatile components have the lowest molecular weight; thus, there is a general shift to higher molecular weight components in residual materials.

Biodegradation is almost always operative when petroleum mixtures are released into the environment. It has been widely demonstrated that nearly all soils and sediments have populations of bacteria and other organisms capable of degrading petroleum hydrocarbons (Pancirov and Brown 1975). Degradation occurs both in the presence and absence of oxygen. Two key factors that determine degradation rates are oxygen supply and molecular structure. In general, degradation is more rapid under aerobic conditions. Decreasing trends in degradation rates according to structure are (Potter and Simmons 1998): 

1. n-alkanes (especially in the C₁₀–C₂₅ range which are degraded readily);
2. isoalkanes;
3. alkenes;
4. benzene, toluene, ethylbenzene and xylenes (BTEX) (when present in concentrations that are not toxic to the microorganisms);
5. monoaromatics;
6. polynuclear (polycyclic) aromatic hydrocarbons (PAHs); and
7. higher molecular weight cycloalkanes (which may degrade very slowly (Pancirov and Brown 1975)).

These trends typically result in the depletion of the more readily degradable components and the accumulation of the most resistant in residues.

Level III fugacity modelling of representative hydrocarbons contained in the HFO group of substances was performed using EQC (2003) (Table A5.1 in Appendix 5) based on their physical-chemical properties as given in Table A2.4 (Appendix 2).

If released solely to air, all C₉–C₁₅ representative structures will remain in air. With an increase in molecular size, the proportion remaining in air declines. Some of the C₂₀ components will also remain primarily in air, except for alkanes, polycycloalkanes, cycloalkane monoaromatics, and four-, five- and six-ring PAHs. Moderate amounts of the C₃₀ isoalkanes (70%) will also remain in air with the same pattern of decreasing partitioning to air with increasing molecular size (Table A5.1 in Appendix 5). Aside from the C₃₀ isoalkanes, the C₃₀ and C₅₀ representative structures of HFOs will partition almost entirely to soil.

If released solely to water, most C₉ representative structures will remain in water, with the exception of alkanes, which will partition almost equally between sediment and water. The C₁₅ one- to three-ring aromatics will also undergo significant partitioning between sediments and water (12–49% into water), while all other representative structures will partition largely to sediment. Volatilization from water surfaces is not expected to be an important fate process despite the presence of some representative structures with moderate to very high estimated Henry’s Law constants. Thus, if water is a receiving medium, all HFOs are expected to have a large proportion of the mixture partitioning to sediment (Table A5.1 in Appendix 5). It is likely, with a release situation into water where the HFO is not immediately in contact with sediments or suspended matter, that the moderate to high Henry’s Law constants will drive the C₉–C₂₀ representative structures out of the water. The tendencies for evaporation and sorption are competing and the exact nature of the release would dictate how the HFO behaves.

If released to soil, all representative structures of HFOs are expected to have high sorption to soil (i.e., expected to be immobile with greater than  99% remaining in the soil). Competing with this tendency are evaporative forces. Volatilization from moist soil surfaces may be an important fate process based upon estimated Henry’s Law constant values of 5.1 to 1.3 × 10⁶ Pa·m³/mol. Lower molecular weight representative structures of HFOs (alkanes, isoalkanes, cycloalkanes and one-ring aromatics) may slightly to substantially volatilize from dry soil surfaces based upon their moderate vapour pressures (Table A5.1 in Appendix 5).

Fugacity estimations in soil do not take into account situations where large quantities of a hydrocarbon mixture enter the soil compartment. When soil organic matter and other sorption sites in soil are fully saturated, 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 (Arthurs et al. 1995), the NAPL will be immobile; 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). 

Persistence and Bioaccumulation Potential

Environmental Persistence

In water, hydrolysis half-lives could not be predicted for hydrocarbons using the HYDROWIN (2008) model. Alkanes, alkenes, benzenes, biphenyls, PAHs and heterocyclic PAHs are all known to be resistant to hydrolysis (Lyman et al. 1990).

Since no empirical data were available on the degradation of these HFOs as complex mixtures, a QSAR-based weight-of-evidence approach (Environment Canada 2007) was applied using the BioHCWin (2008), BIOWIN 3,4,5,6 (2009), CATABOL (c2004-2008) and TOPKAT (2004) biodegradation models (Table A5.2 in Appendix 5).

Using an extrapolation ratio of 1:1:4 for a water : soil : sediment biodegradation half-life (Boethling et al. 1995), the representative structures that are persistent in water are also persistent in soil (half-life greater than or equal to 182 days) and in sediments (half-life greater than or equal to 365 days).

Using compositional data on Fuel Oil No. 6 and reading across to these HFOs (Tables A5.3 and A5.4 in Appendix 5), the average weight percent of components that are expected to be persistent ranges from approximately 50–60%.

AOPWIN (2008) is a model that calculates atmospheric oxidation half-lives of compounds in contact with hydroxyl radicals in the troposphere under the influence of sunlight. Atmospheric oxidation rates were calculated for all of the representative structures. Although the low vapour pressures of these representative structures indicate that volatilization may not be a very significant fate process, oxidation half-lives of less than 1 day (Table A5.5 in Appendix 5) indicate that this would be a relatively rapid removal process if these substances were introduced into the atmosphere (Atkinson 1990; API 2004).

Persistence Conclusion

Based on results from AOPWIN (2008), there would be a relatively rapid removal process if these HFOs were introduced into the atmosphere, based on oxidation half-lives of less than 1 day. These HFOs thus do not meet the criterion for persistence in air (half-life greater than or equal to 2 days) as defined in the Persistence and Bioaccumulation Regulations (Canada 2000). With regard to the primary and ultimate biodegradation modelling, the C₃₀–C₅₀ isoalkanes, C₃₀–C₅₀ one-ring cycloalkanes, C₁₅–C₅₀ two-ring cycloalkanes, C₁₄–C₂₂ polycycloalkanes, C₃₀–C₅₀ one-ring aromatics, C₁₅–C₂₀ cycloalkane monoaromatics, C₁₅–C₅₀ two-ring aromatics, C₁₂ cycloalkane diaromatics, C₂₀–C₅₀ three-ring aromatics, C₁₆–C₂₀ four-ring aromatics, C₂₀–C₃₀ five-ring aromatics and C₂₂ six-ring aromatics in these HFOs meet the criteria for persistence (half-lives in soil and water greater than or equal to 182 days and half-life in sediment greater than or equal to 365 days). These HFOs are estimated to contain approximately 50–60% of components (C₁₀–C₅₀) by weight that meet the persistence criteria as defined in the Persistence and Bioaccumulation Regulations (Canada 2000).

Potential for Bioaccumulation

Bioconcentration Factors (BCF) and Bioaccumulation Factors (BAF)

Experimental studies

Since no empirical data on the bioaccumulation of HFOs or its components were found, empirical data on the bioaccumulation of components of Fuel Oil No. 6 was used in a read-across approach. A predictive approach using a bioconcentration/bioaccumulation factor (BCF/BAF) model was also applied (Arnot and Gobas 2003, 2004). According to the Persistence and Bioaccumulation Regulations (Canada 2000) a substance is bioaccumulative if its BCF or BAF is greater than or equal to 5000; however, measures of BAF are the preferred metric for assessing the bioaccumulation potential of substances. This is because BCF may not adequately account for the bioaccumulation potential of substances via the diet, which predominates for substances with log K_ow greater than ~4.5 (Arnot and Gobas 2003).

Neff et al. (1976) exposed clams (Rangia cuneata), oysters (Crassostrea virginica) and fish (Fundulus similus) to the water-soluble fraction of Fuel Oil No. 2 (0.41 kg/L [2 ppm] total naphthalenes) for 2 hours, followed by depuration of hydrocarbons for 366 hours. All fish organs examined showed rapid accumulation of naphthalenes within the 2-hour exposure period, with the gallbladder and brain of fish accumulating the highest concentrations. BAFs of naphthalenes in clams ranged from 2.3–26.7 L/kg wet weight (ww) (Table A5.6 in Appendix 5). Release of naphthalenes by fish began immediately following transfer to fresh water, reaching undetectable levels after 366 hours (~15 days).

Peterson and Kristensen (1998) exposed eggs and larvae of zebrafish (Brachydanio rerio) and larvae of cod (Gadus morhua), herring (Clupea harengus), and turbot (Scophthalmus maximus) to 14C-labelled PAHs (naphthalene, phenanthrene, pyrene and benzo[a]pyrene B[a]P. The experiments were performed in a semistatic test system and steady-state was not reached during the embryonic stage except for naphthalene. High BCFs were found in all cases, indicating that bioaccumulation can occur during early life stages as fish larvae have higher lipid contents and lower metabolic capabilities than juvenile or adult fish.

Burkhard and Lukasewycz (2000) compiled data on tissue (lake trout; Salvelinus namaycush), water and sediment concentrations of PAHsfrom three separate published works and used the data to derive BAFs. BAFs for PAHs in these fish were 87, 1550 and 3990 L/kg ww for phenanthrene, fluoranthene and chrysene/triphenylene, respectively (Table A5.6 in Appendix 5). Burkhard and Lukasewycz (2000) note that there is significant uncertainty in the BAFs for phenanthrene and fluoranthene, as both chemicals were present in the tissues at concentrations just greater than the method detection limit.

Hardy et al. (1974) carried out an experiment giving cod (Gadus morhua) single doses of hexadecane (a C₁₆ alkane) in the diet and tracked metabolites. Entirely unchanged hexadecane was found in the liver. Hardy et al. (1974) suggest that such findings do not support high metabolic conversion of hexadecane in the liver of cod, and n-alkanes were preferentially deposited in liver over flesh of cod. However, the liver is the major site of chemical biotransformation, so higher concentrations in liver would be expected. Cravedi and Tulliez (1981) dosed rainbow trout with dodecyl cyclohexane (a C₁₈ alkyl cycloalkane) and studied its elimination and metabolism from the fish. Approximately 75% of the dose was absorbed. A major source of unmodified substance elimination was through the gills. Considerable amounts were also metabolized to a fatty acid and distributed throughout the body and 14% was excreted in urine (Cravedi and Tulliez 1981).

Cravedi and Tulliez (1983) also studied the dietary uptake of 1% C₁₃–C₂₂ n-alkanes in rainbow trout for 7 months. Trout were dosed with 10 000 ppm total alkanes in feed, and showed preferential fixation of C₁₃–C₁₄ n-alkanes in the adipose tissue. The mean accumulated mass of n-alkanes was 958 ppm per fish, so that a calculated BCF (diet) was 0.1. n-Alkanes longer than C₁₆ were well retained (over 60% of accumulated n-alkanes remained after 8 weeks of depuration), while short-chain ( less than C₁₆) n-alkane concentrations decreased more rapidly (only 20–50% remained after 8 weeks of depuration).

Colombo et al. (2007) studied the bioaccumulation dynamics of C₁₂–C₂₅ n-alkanes and aliphatic unresolved complex hydrocarbons (UCM) in a detritivorous fish (Prochilodus lineatus) collected from the sewage-impacted Buenos Aires coastal area. Fish muscles contained large amounts of C₁₂–C₂₅ n-alkanes and aliphatic UCM, reflecting the chronic bioaccumulation of fossil fuels from sewage particulates. The hydrocarbon composition in fish muscles was enriched in C₁₅–C₁₇ n-alkanes relative to fresh crude oil and settling particulates. The bioaccumulation factors (BAFs: 0.4–6.4 dw or 0.07–0.94 lipid-organic carbon) plotted against K_ow showed a parabolic pattern maximizing at C₁₄–C₁₈.

McCain et al. (1978) reported that 1- and 2-methylnaphthalene and 1,2,3,4-tetramethylbenzene were accumulated to a greater extent than other oil components in English sole (Parophrys vetulus) from oil-contaminated sediments. Tissue burdens of hydrocarbons decreased with increasing exposure time, such that after 27 days of exposure, only the liver had a detectable hydrocarbon burden. McCain et al. (1978) suggested that induction of the aryl hydrocarbon hydroxylase enzyme system eventually resulted in hydrocarbon removal.

Weinstein and Oris (1999) found that 4-day-old fathead minnows (Pimephales promelas) bioconcentrated fluoranthene (BCF 9054 L/kg) with only 24 hours exposure and steady-state was reached. They observed that the age of the fish likely impacted the ability to depurate fluoranthene and that older, more mature fish would be unlikely to bioacumulate PAHs. Weinstein and Oris (1999) used a static renewal system which is less preferable to flow-through designs where consistent exposures can be maintained, thus this study was considered to be of low reliability. However, the study does show that bioaccumulation is important for toxicity in the early life stages (Weinstein and Oris 1999). In contrast, De Maagd (1996) found a BCF of 3388 L/kg ww for fluoranthene in adult fathead minnows.

Guppies (Poecilia reticulata) bioconcentrated pyrene, producing BCFs in the range of 4786–11 300 L/kg ww (depending on the type of test) after 48 hours of exposure, while lighter-weight PAHs had lower BCFs (1050–2238 L/kg ww for fluorene and 4550–7244 L/kg ww for anthracene) (De Voogt et al. 1991). The fish were capable of depurating pyrene completely within 160 hours of cessation of exposure. However, anthracene was only 70% and fluorene only 20% depurated within 200 hours. The BCF results for anthracene and pyrene by De Voogt et al. (1991) were not considered reliable in determining the bioconcentration potential of these substances due the lack of evidence that a steady-state had been reached within the 48-hour exposure. Likewise, there was poor recovery of pyrene at the end of the static bioconcentration experiment (62%) which had the BCF of 11 300 L/kg ww.

Jimenez et al. (1987) exposed bluegill sunfish (Lepomis macrochirus) to [14C]B[a]P in a flow-through system for 48 hours to determine the effects of temperature and feeding on B[a]P uptake and elimination. The uptake of B[a]P was twice as fast for fish that were fed compared to those denied food in fish and uptake was slower at lower temperatures. A BCF of 608 L/kg ww was found for B[a]P for fed bluegill sunfish at 23ºC and it appears that steady-state was reached.

Jonsson et al. (2004) used a long-term (36-day) study to determine the bioconcentration of pyrene in sheepshead minnows (Cyprinodon variegatus). Fish reached a steady state after 4–7 days of exposure. The BCFs were 145 and 97 L/kg ww for PAH concentrations of 7.57 and 72.3 µg/L, respectively, which were likely due to biotransformation of PAHs by the fish.

Mollusc studies have typically found high potentials for the bioconcentration of PAHs. This may be caused by the relatively slow rates of depuration when compared to fish studies coupled with fairly rapid uptake. Other works have shown that BCFs for PAHs in molluscs and some crustaceans are considerably higher than in fish (Table A5.7 in Appendix 5). Unlike fish and some crustaceans, molluscs are unable to rapidly metabolize aromatic hydrocarbons. Accumulation can occur in stable tissue compartments with low hydrocarbon turnover and that are not readily exchangeable (Stegeman and Teal 1973; Neff et al. 1976).

The zebra mussel (Dreissena polymorpha) showed fast uptake of B[a]P and pyrene over a 6-hour exposure which led to high BCF values (Bruner et al. 1994; Gossiaux et al. 1996). After 3 days of depuration, body concentrations of B[a]P had dropped to less than 50%, and after 2 weeks, the concentrations had been reduced to 5–20% (Gossiaux et al. 1996). Pyrene elimination was highly temperature dependent, with depuration occurring more rapidly at higher temperatures and occurring very slowly at colder temperatures (Gossiaux et al. 1996). Lipid content was also important to the bioconcentration values, with higher lipid contents accumulating PAHs more readily, whereas body size did not affect the BCF values (Bruner et al. 1994).

McLeese and Burridge (1987) studied the bioaccumulation potential of PAHs by a number of saltwater invertebrates using PAH-seawater solutions or PAH-contaminated sediments. When PAHs were dissolved in water, fluoranthene, pyrene, triphenylene and perylene produced high BCF values in mussels (Mytilus edulis) (5920, 4430, 11 390 and 10 500 L/kg ww, respectively) and clams (Mya arenaria) (4120, 6430, 5540 and 10 000 L/kg ww, respectively) after short (96-hour) exposures. However, when PAHs are present in the sediment, only mussels have a high potential for bioconcentration (BCF of 5950 L/kg ww for fluoranthene, 5000 L/kg ww for pyrene and 9500 L/kg ww for perylene). All of these substances can be depurated from molluscs given time, but heavier PAHs (triphenylene and perylene) depurate more slowly than lighter PAHs (phenanthrene, fluoranthene and pyrene). Shrimps and polychaetes did not readily bioaccumulate PAHs.

Although some crustaceans can readily bioaccumulate higher-weight PAHs, they can also rapidly depurate PAHs. After 6 hours of exposure to B[a]P, the amphipod Pontoporeia hoyi and the freshwater shrimp Mysis relicta showed rapid uptake (Evans and Landrum 1989), but also had rapid depuration over 10–26 days (Evans and Landrum 1989). In Daphnia magna exposed to PAHs for 24 hours, high bioconcentration ( greater than 5000 L/kg) was observed with 11 higher-weight PAHs, ranging from 6100 L/kg ww for chrysene to 50 000 L/kg ww for dibenz[ah]anthracene (Newsted and Giesy 1987). Depuration was not studied.

Other invertebrates have also been shown to bioaccumulate petroleum hydrocarbons. Muijs and Jonker (2010) studied the bioaccumulation of petroleum hydrocarbons (total and divided into three different carbon ranges) over 49 days by the aquatic worm, Lumbriculus variegatus, after exposure to a series of 14 field-contaminated sediments with a known history of oil pollution. A maximum tissue concentration was reached for the C₁₁–C₁₆ fraction after 14 days of exposure but then decreased; other fractions did not show any decrease in tissue concentration once a maximum was achieved. After 28 days of exposure, it was estimated that 70–90% of equilibrium was reached, though it was noted that it may take greater than 90 days for hydrocarbons greater than C₃₄ to reach equilibrium. Characterization of the accumulated hydrocarbons was not determined, however, alkanes from C₁₀–C₃₄ were identified in the aquatic worms. The accumulation of higher molecular weight alkanes may possibly be due to ingestion of organic matter to which the chemicals are sorbed. Depuration was not studied.

Overall, BCF values determined for various PAHs (Table A5.7 in Appendix 5) were highly variable, ranging from 180 to over 28 000 L/kg ww. The majority of BCF studies on PAHs have found that bioconcentration can occur after short exposure times but that the majority of organisms also exhibit rapid depuration once the contaminant is removed. However, some components have been shown to meet the persistence criteria.

Three studies on BAFs of PAHs in aquatic organisms were found. Hence, experimental values of BAFs from the work of Neff et al. (1976), Zhou et al. (1997) and Burkhard and Lukasewyez (2000) were compiled for comparison with modelled data (Arnot and Gobas 2003) (Table A5.6 in Appendix 5). In general, the modelled values approximate the measured (Table A5.8 in Appendix 5) for the selected PAHs. None of the measured and modelled values were shown to be bioaccumulative according to the criteria (BAF  greater than or equal to  5000) in the Persistence and Bioaccumulation Regulations (Canada 2000), with the exception of the substituted PAH isoheptylfluorene and 2-isohexylphenanthrene (see Table A5.8 in Appendix 5).

Characterizing BCF/BAF

In characterizing bioaccumulation, the derivation of a BAF is preferred over a BCF since chemical exposure through the diet is not included in the latter (Barron 1990). BCFs are typically derived under laboratory controlled conditions. According to Arnot and Gobas (2006), the BCF is a poor descriptor of biomagnification in food webs because it is derived from laboratory experiments and does not include dietary exposure. Thus, BCFs based on laboratory studies have been shown to underestimate bioaccumulation potential or biomagnification of chemicals in the food web, as predators consume prey containing lipophilic compounds (U.S. EPA 1995). As hydrophobicity increases, dietary uptake is likely to be more important than absorption from water (Arnot and Gobas 2003). Furthermore, laboratory BCFs have been shown to overestimate bioaccumulation potential when a chemical is bound or tightly sorbed to sediment (i.e., less bioavailable).

Due to the scarcity of measured BAF values (Table A5.6 in Appendix 5), BCFs from various published works were compiled (Table A5.9a in Appendix 5) and used to help verify measured and modelled BAF values. In contrast to the few available experimental BAFs on PAHs, a suite of BCFs for components of HFOs were found, including alkanes, isoalkanes, two-ring cycloalkanes, one-ring aromatics, cycloalkane monoaromatics, cycloalkane diaromatics and polyaromatics (Table A5.9a in Appendix 5). Model estimates of these BCFs were also produced using a kinetic mass-balance model (Arnot and Gobas 2003) to fit the model kinetic elimination constants to agree with the observed BCF data in order to generate BAF predictions that reflect the known elimination rates.

A kinetic mass-balance model is, in principle, considered to provide the most reliable prediction method for determining bioaccumulation potential because it allows for correction of the kinetic rate constants and bioavailability parameters, when possible. BCF and BAF model predictions are considered “in domain” for this hydrocarbon assessment because it is based on first principles. As long as the mechanistic domain (passive diffusion), global parameter domain (range of empirical log K_ow and molecular weight), as well as metabolism domain (corrected metabolic rate [kM]) are satisfied, predictions are considered valid (Arnot and Gobas 2003, 2006). The kinetic mass-balance model developed by Arnot and Gobas (2003, 2004) was employed using metabolic rate constants normalized to both conditions of the study and a representative middle trophic level fish as outlined in Arnot et al. (2008a, b) when the BCF or growth-corrected elimination rate constant is known. Both BCF and biomagnification factor (BMF) empirical data were used to correct default model uptake and elimination parameters, which are summarized in Table A5.9b (Appendix 5).

In Table A5.9b (Appendix 5), some metabolic rate constants calculated from the empirical BCF data were negative, suggesting that the metabolic rate is essentially zero and that other routes of elimination are more important. Accordingly, no metabolic rate correction was used when predicting the BCF and BAF for these structures. Gut and tissue metabolism is generally not regarded as an important elimination process for chemicals with log K_ow less than ~4.5 (Arnot et al. 2008a, b; Arnot and Gobas 2006), but this can depend on the size and lipid content of fish used in testing.

In Table A5.9a (Appendix 5), only the C₁₅ isoalkane (2,6,10-trimethyldodecane), C₈ one-ring cycloalkane (ethylcyclohexane), and C₁₃ two-ring aromatics (2-isopropylnaphthalene) had measured and/or modelled BCFs or BAFs greater than or equal to 5000. However, the measured diaromatic (2-isopropylnaphthalene) that was found to be highly bioaccumulative contains the isopropyl functional group that is considered atypical in petroleum and requires a more thorough appraisal of reasonableness of model predictions based on available experimental information (Lampi et al. 2010). As well, Neff et al. (1976) found that the C₁₂ and C₁₃ diaromatics (alkylated naphthalenes and biphenyls) were not highly bioaccumulative in clams upon exposure to an oil-in-water dispersion of Fuel Oil No. 2. Thus, the combined weight of evidence suggests that these C₁₂ and C₁₃ diaromatics are not likely to be highly bioaccumulative. For the C₈ cyclohexane (ethyl cyclohexane), the predicted BAF (Arnot and Gobas 2004) for the middle trophic level fish is 5495 L/kg ww, which just exceeds the criterion (BAF greater than or equal to 5000), suggesting that it is bioaccumulative when all routes of uptake are considered. This prediction, however, was generated with a metabolic rate equal to zero because of the potential error associated with the estimate of metabolism rates (see Table A5.9b in Appendix 5). Factoring in metabolism, it is expected that the BAF would be lower and likely below 5000. As well, the experimental BCF suggests this C₈ cycloalkane is not highly bioaccumulative (Table A5.9a in Appendix 5). Combining these lines of reasoning, this suggests that this C₈ cycloalkane is also not likely to be bioaccumulative according to the Canadian criteria. For the C₁₅ isoalkane (2,6,10-trimethyldodecane), two predicted BAFs are presented (575 and 47 863 L/kg ww). The latter BAF of 47 863 L/kg ww is preferred, as the depuration rate constant from the study was available to calculate the metabolic rate constant. This higher predicted BAF value is also in agreement with the slow rate of metabolism. Combining these lines of reasoning, this suggests that this C₁₅ isoalkane is likely bioaccumulative according to the Canadian criteria.

Most components greater than C₂₀ have an estimated log K_ow greater than 8 and were excluded from the modelling, as predictions may be highly uncertain due to limitations of the model (Arnot and Gobas 2003). In Arnot and Gobas (2006), at a log K_ow of 8.0, the empirical distribution of “acceptable” fish BCF data shows that there are very few chemicals with fish BCFs exceeding the Canadian criterion of BCF greater than or equal to 5000. Examination of Environment Canada’s empirical BCF/BAF database for DSL and non-DSL chemicals developed by Arnot and Gobas (2003) and further by Arnot (2005, 2006) shows that these are all highly chlorinated substances (i.e., decachlorobiphenyl, nonachlorobiphenyl, heptachlorobiphenyl), which have BCFs in the 105 range, noting that octachloro naphthalene has a measured BCF of less than 1000 L/kg ww, (Fox et al. 1994; Gobas et al. 1989; Oliver and Niimi 1988) and all have log K_ow values less than 8.0. Therefore, the predicted BCF and BAF values with log K_ow greater than 8 were considered out of the parametric domain of the Arnot-Gobas model (2003) and considered highly uncertain and not reliable.

BCF and BAF model estimates were also generated for an additional twenty-six C₉–C₂₂ linear and cyclic representative structures using the modified Arnot-Gobas three trophic level model (2004) (Table A5.8 in Appendix 5), as no empirical bioaccumulation data were identified for these substances. Metabolism and dietary assimilation efficiency kinetics were corrected for these predictions based on analogue BCF and BMF test data. From this analysis, only one C₁₄ polycycloalkane was predicted to have a BCF that suggested a high bioconcentration potential. However, one isoalkane, several polycycloalkanes, one- and two-ring cycloalkanes and one-, two- and three-ring aromatics were found to have high bioaccumulation factors. The log K_ow for these structures suggests that dietary uptake can predominate (up to 87% of total uptake) but will not be the sole route of exposure as some substances are expected to have a 90% bioavailable fraction in the water column. BAF is therefore considered the most appropriate metric for assessing the bioaccumulation potential of these structures and represents a comparison of whole-body burdens compared with concentrations in water. The BCF and BAF predictions for these fractions are within the parametric, mechanistic and metabolic domains of the model and so are considered reliable.

Biomagnification Factors (BMF) and Trophic Magnification Factors (TMFs)

BMF values from ExxonMobil Biomedical Sciences Inc. (EMBSI), used to derive kinetic information for 15 substances, are reported in Table A5.9a (Appendix 5) (Lampi et al. 2010). None of these analogues have BMFs greater than 1, suggesting that these hydrocarbons will not biomagnify when compared to the concentrations expected in food items. A combination of metabolism, low dietary assimilation efficiency and growth dilution appear to limit the biomagnification potential of these compounds (see Tables A5.9a and A5.9b in Appendix 5).

Lampi et al. (2010) also summarized TMFs for PAHs from three field studies. The TMFs for various PAHs are summarized in Table A5.10 (Appendix 5). Field-based TMFs for the PAHs studied are mostly less than 1, except fluorene and acenaphthene, which are approximately 1. A combination of metabolism, low dietary assimilation efficiency and growth dilution appear to limit the trophic magnification potential of these compounds as well. Therefore, it is not likely that the linear, cyclic and aromatic components of HFOs will undergo biomagnification or trophic magnification.

Broman et al. (1990) studied TMFs for 19 PAHs in a marine food chain (seston to mussels (M. edulis) to ducks (Somateria mollissima)), and did not find TMFs greater than 1.

Biota-Sediment Accumulation Factors (BSAFs)

Lampi et al. (2010) also summarized the available BSAF data for several PAHs from a database compiled by the U.S. EPA (2008a). Median field-based fish BSAF values for PAHs expected to be found in HFOs (acenaphthylene, acenaphthene, benzo[a]anthracene, B[a]P, benzo[e]pyrene, benzo[b]fluoranthene, benzo[b+k]fluoranthene, benzo[j+k]fluoranthene, benzo[k]fluoranthene, benzo[ghi]perylene, chrysene, fluoranthene, fluorene, indeno[123-cd]pyrene, dibenz[ah]anthracene, perylene, naphthalene, phenanthrene and pyrene) ranged from 10^-4 to 10-1. Ninetieth percentile BSAF values ranged from 10^-4 to just less than 1, with naphthalene being the only PAH with a BSAF close to but below 1. None of the PAHs have fish BSAFs greater than one. This is expected, given the same rationale for low BMF and TMF values. However, data were not extracted for invertebrate BSAFs from the U.S. EPA database. In the case of invertebrates, these factors can be much greater than one, because invertebrates do not have the same metabolic competency as fish (e.g., B[a]P) (Muijs and Jonker 2010; Stegeman and Teal 1973; Neff et al. 1976).

As previously noted, Muijs and Jonker (2010) studied the bioaccumulation of oil in the aquatic worm, L. variegatus. Resulting BSAFs varied from 0.01–2.3. The wide range is likely related to the differences in oil weathering status. The BSAF values for separate hydrocarbon blocks appeared to be relatively constant up to C₂₂, indicating that L. variegatus proportionally accumulated these fractions from sediment. Beyond C₂₂, BSAFs decreased for all sediments studied, likely due to the reduced bioavailability of the higher boiling point fractions such as PAHs. Likewise, there may be enhanced sorption of PAHs to sediment and in some cases the nonaqueous phase liquid (NAPL). Muijs and Jonker (2010) also suggest that the studied aquatic worm may even avoid NAPLs, which may also limit the bioaccumulation of the very hydrophobic fractions.

Bioaccumulation Conclusion

Non-PAH Components

As noted previously, of the parameters that have prescribed Canadian regulatory criteria, BAF values are preferred over BCF values because they represent the potential accumulation in biota from all exposure sources and thus represent a more complete picture of the total body burden of chemicals. Biomagnification (BMF), trophic or foodweb magnification (TMF) and biota-sediment accumulation factors (BSAF) are also considered very important for understanding the pattern of bioaccumulation and are used in a weight of evidence for the overall bioaccumulation potential of a chemical.

In general, the majority of less than C₁₅ components (alkanes, isoalkanes and cycloalkane monoaromatics) were not found to meet the Persistence and Bioaccumulation Regulations (Canada 2000). This conclusion is based on consistencies found between available BCF and BAF experimental data and BCF and BAF kinetic mass-balance model predictions using the Arnot-Gobas (2004) three trophic level model.

The majority of components greater than or equal to C₂₀ (alkanes, isoalkanes, one-ring cycloalkanes, two-ring cycloalkanes and one-ring aromatics) have estimated log K_ows greater than 8 and were therefore excluded from modelling, as predictions may be highly uncertain due to limitations of the model (Arnot and Gobas 2003). Likewise, for these greater than or equal to C₂₀ components, no experimental measured BCFs were found.

In terms of the polycycloalkanes, the C₁₈ polycycloalkane (hydrochrysene) did not meet the criterion of BCF or BAF greater than or equal to 5000 for its modelled BAF prediction using the modified Arnot-Gobas three trophic level model (2004), whereas the C₁₄ and C₂₂ polycycloalkanes were found to meet this criteria based on the same model (Table A5.8 in Appendix 5) The metabolic rate constant (0.45/day) for hydrochrysene suggests a rapid rate of metabolism in comparison to the lower metabolic rate constants (0.01/day and 0.04/day) for the C₁₄ and C₂₂ polycycloalkanes. Study details from experimental evidence for a similar polycycloalkane could not be obtained to determine predicted BCFs and BAFs, thus the available evidence suggests that the C₁₈ polycycloalkane (hydrochrysene) is not bioaccumulative based on modelled results alone.

The C₁₄ and C₂₂ polycycloalkanes, C₁₅ one-ring aromatics, C₁₅–C₂₀ cycloalkane monoaromatics and C₂₀ cycloalkane diaromatics were found to meet the bioaccumulation criteria based on modelled results from the Arnot-Gobas (2004) three trophic level model. For these particular components, the metabolic rate constants range from 0.01–0.08 (day^-1), suggesting a slow rate of metabolism. In the case of C₁₄ and C₂₂ polycycloalkanes, C₁₅ one-ring aromatic and C₂₀ cycloalkane monoaromatic, only experimental BMFs for comparative analogues were available. The BMFs were all less than 1, suggesting that these components will not biomagnify. In the case of the C₁₅ cycloalkane monoaromatic, only an experimental BCF (3418 L/kg ww) for a similar component (octahydro-phenanthrene) was found. However, considering the slow metabolic rate of 0.197 (day^-1) for octahydrophenanthrene, there is the potential that predicted BCFs and BAFs for the C₁₅ cycloalkane monoaromatic could exceed the Canadian criteria, although this cannot be determined due to the lack of details from the relevant study. Lastly, the only analogue similar to the C₂₀ cycloalkane diaromatic (isoheptylfluorene) is fluorene, which has an experimental BCF of 1030 L/kg ww. However, the presence of an isoheptyl group may affect the bioaccumulation potential of fluorine, and the low kM value suggests a slow rate of metabolism. Overall, the available evidence suggests that these components are likely to bioaccumulate based on available modelled and experimental results.

BMF values for 15 substances comprising some isoalkanes, one- and two-ring cycloalkanes, polycycloalkanes, one-ring aromatics, cycloalkane monoaromatics, cycloalkane diaromatics and three- and four-ring aromatics (see Table A5.9a in Appendix 5) show that no components have BMFs greater than 1. This suggests that these particular hydrocarbons will not biomagnify when compared to concentrations expected in food items. Thus, the available evidence suggests that there is limited biomagnification of petroleum hydrocarbons. It is possible that BSAFs will be greater than 1 for invertebrates (up to 2.3 for total petroleum hydrocarbons in L. variegatus (Muijs and Jonker 2010)) as they do not have the same metabolic competency as fish. However, BSAFs will likely decrease beyond C₂₂ due to reduced bioavailability of the higher boiling point fractions (Muijs and Jonker 2010).

Overall, there is consistent empirical and predicted evidence to suggest that 10 representative structures (C₁₅ isoalkane, C₁₅ one-ring cycloalkanes, C₁₅ two-ring cycloalkane, C₁₄ and C₂₂ polycycloalkane, C₁₅ one-ring aromatic and C₁₅–C₂₀ cycloalkane monoaromatics) meet the bioaccumulation criteria as defined in the Persistence and Bioaccumulation Regulations (Canada 2000). These components are associated with a slow rate of metabolism and are highly lipophilic. Exposures from water and the diet, when combined, suggests that the rate of uptake would exceed that of the total elimination rate. However, these components are not expected to biomagnify in aquatic foodwebs largely because a combination of metabolism, low dietary assimilation efficiency and growth dilution allows the elimination rate to exceed the total uptake rate.

PAH Components

In general, the majority of less than C₁₅ components (cycloalkane diaromatics and three-ring PAHs) were not found to meet the Persistence and Bioaccumulation Regulations (Canada 2000). The majority of components greater than or equal to C₂₀ (two-, three- and five-ring aromatics) have estimated log K_ows greater than 8 and were therefore excluded from modelling, as predictions may be highly uncertain due to limitations of the model (Arnot and Gobas 2003).

Experimental BAFs and BCFs suggest that PAHs, as a whole, have low bioaccumulation potential in fish. This is due in part to the metabolism of PAHs by fish, resulting in low or nondetectable concentrations of the parent PAHs in fish tissues (Varanasi et al. 1989). Regarding BAF, none of the measured or modelled values were shown to meet the bioaccumulation criterion (BAF greater than or equal to 5000) as defined in the Persistence and Bioaccumulation Regulations (Canada 2000) with the exception of modelled BAF values for isoheptylfluorene and 2-isohexylphenanthrene (see Table A5.8 in Appendix 5). Lampi et al. (2010) found that isopropyl functional groups increased the bioaccumulation potential of naphthalene although isopropyl groups are considered atypical in petroleum. Thus highly alkylated PAHs, especially those with iso- groups, likely have a greater potential to bioaccumulate simply from increased partitioning to lipophillic tissues in biota and possibly some hindrance of biotransformation. Lack of experimental or field data for alkyl-PAHs larger than naphthalene prevents drawing a concrete conclusion for these substances, albeit Neff et al. (1976) found that as naphthalene becomes increasingly alkylated, there is an increase in bioaccumulation potential. With regards to the modelled BAF values for isoheptylfluorene and 2-isohexylphenanthrene, the only similar analogues (fluorene and phenanthrene) have experimental BCFs of 1030 L/kg ww and 2944 L/kg ww, respectively, which are both slightly higher than the predicted BCFs using the mass-balance kinetic model (Table A5.8 in Appendix 5). However, there is some uncertainty surrounding the kinetic rate constants used to model BCF and BAF for isoheptylfluorene and 2-isohexylphenanthrene (e.g., the metabolic rate constants were either estimated from QSARs or based on analogue data), as well as the degree of trophic magnification within the foodweb used by the model), suggesting that the BAFs may be overestimated. However, given that the log K_ow of these compounds is between 7.0 and 7.5, the optimal range for high bioaccumulation from the diet and water coupled with a possible slow rate of metabolism and that a TMF for fluorene is approximately 1 (Table A5.10 in Appendix 5), a high bioaccumulation potential may still be likely.

None of the modelled BCF values for representative PAHs were shown to meet the bioconcentration criterion (BCF greater than or equal to 5000) as defined in the Persistence and Bioaccumulation Regulations (Canada 2000) (see Table A5.8 in Appendix 5). This is largely due to the lower contribution of chemical uptake from water from highly hydrophobic substances, but also because PAHs such as naphthalene, phenanthrene and B[a]P are metabolized by fish resulting in very low or nondetectable concentrations of the parent PAHs in fish tissues (Varanasi et al. 1989). However, measured BCFs in fish for some of the PAHs exceed the bioaccumulation criterion, including fluoranthene, anthracene and pyrene (Table A5.8 in Appendix 5). For fluoranthene, Weinstein and Oris (1999) reported a BCF of 9054 L/kg ww in fathead minnows, Burkhard and Lukasewycz (2000) determined a BAF of 1550 L/kg ww in trout, and De Maagd (1996) determined a BCF of 3388 L/kg ww in fathead minnows. As previously mentioned, the Weinstein and Oris (1999) and De Voogt et al. (1991) studies, as well as Peterson and Kristensen (1998), reported high BCF values and contain sufficient levels of uncertainty, or the early life stage results cannot easily be interpreted versus other studies or regulatory criteria for bioaccumulation. The findings of these studies were thus considered equivocal and received a lower weighting for determining bioaccumulation potential according to criteria. The high laboratory BCFs are also not consistent with field measured BAFs in fish for fluoranthene. Consequently there is greater evidence weight and consistency from kinetic data, modelled BCF and BAF values, and laboratory and field evidence for vertebrates (i.e., fish) to suggest that vertebrates possess sufficient metabolic capacities and other elimination processes to mitigate body burdens of PAHs below levels considered by criteria to be high levels of bioaccumulation.

Empirical BCF data for invertebrates, namely molluscs (fluoranthene and pyrene) and Daphnia magna (chrysene, benzo[a]anthracene, benzo[k]fluoranthene, B[a]P, benzo[e]pyrene and benzo[ghi]perylene) have been shown to be high. In the case of D. magna, benzo[a]anthracene, B[a]P, benzo[e]pyrene and benzo[ghi]perylene have BCF values exceeding bioaccumulation criteria at approximately 10 000 L/kg ww (Table A5.7 in Appendix 5). This indicates that there is potential for body burdens to reach toxic levels in these lower trophic level organisms as they lack the metabolic capability to eliminate PAHs in comparison to fish. Thus, high accumulation patterns are found in both the lab and field. There is also potential for these body burdens to exceed the internal narcotic thresholds, assuming PAH exposure is constant and continuous. However, the majority of BCF studies on PAHs have found that bioconcentration by invertebrates can occur quickly but that the majority of organisms also exhibit rapid depuration once the contaminant is removed. Therefore, exposure duration is critical to bioaccumulation and toxicity.

Field-based TMFs for PAHs were mostly less than 1, with the exception of fluorene and acenaphthene which are approximately one (Table A5.10 in Appendix 5). It appears that biomagnification and trophic magnification are mitigated by a combination of metabolism, low dietary assimilation efficiency and growth dilution through the food-chain. Thus, the available evidence suggests that there is limited biomagnification and trophic magnification for PAHs.

As PAHs tend to accumulate in sediments, benthic organisms may be continuously exposed to the contaminants. Because invertebrates do not have the same metabolic competency as fish (Muijs and Jonker 2010; Stegeman and Teal 1973; Neff et al. 1976), the bioaccumulation potential in invertebrates is expected to be higher than in fish. While only BSAFs for fish were found for some PAHs and were below one, it is possible that BSAFs will be greater than 1 for invertebrates as they have lower metabolic competencies than fish, but BSAFs will likely decrease beyond C₂₂ due to reduced bioavailability of the higher boiling point fractions (Muijs and Jonker 2010).

Overall, the evidence indicates that 4 representative polycyclic aromatic hydrocarbon structures (C₂₀ three-ring PAHs, C₁₈ four-ring PAHs, C₂₀ five-ring PAHs and C₂₂ six-ring PAHs) meet the bioaccumulation criteria as defined in the Persistence and Bioaccumulation Regulations (Canada 2000).

Proportion of Bioaccumulative Components in HFOs

Based on the boiling point ranges of each individual CAS RN (Table A2.4 in Appendix 2), the proportions of components that are expected to be bioaccumulative range from approximately 5 to 25% by weight.  These proportions are based on Canadian samples of Fuel Oil No. 6, as the chemical characterization of these industry-restricted HFOs is unknown (Table A5.11 in Appendix 5). A more detailed analysis of how these bioaccumulative proportions were determined is shown in Table A5.11 (Appendix 5).

Thus, up to approximately 25% of components by weight of these industry-restricted HFOs may be bioaccumulative based on criteria in the Persistence and Bioaccumulation Regulations (Canada 2000).

Potential to Cause Ecological Harm

Ecological Effects Assessment

Information relevant to the toxicity of HFOs to various organisms is provided below. As well, PAHs are components of HFOs and have been considered in a previous regulatory assessment. PAHs are on the List of Toxic Substances under Schedule 1 of CEPA 1999 (Environment Canada 2010c).

Evidence from field and laboratory studies using field samples indicates that biota are adversely affected at various Canadian sites contaminated by PAHs of different industrial origins (Canada 1994).

There are potential hazards associated with the metabolism of PAHs such as B[a]P. This process may create metabolites that are potent mutagens. Under laboratory conditions, neoplastic and genotoxic effects have been associated with exposure to PAHs for both terrestrial and aquatic organisms. In field studies, preliminary stages of chemically induced carcinogenesis have been shown (Environment Canada 1994).

Aquatic Compartment

No experimental data were available for the aquatic toxicity of these industry-restricted HFOs; therefore, data from Fuel Oil No. 6 were used in a read-across approach to estimate the potential for aquatic toxicity. Other studies have shown that, with HFOs, variations in aquatic toxicity exist, in part, due to differences in boiling point ranges determining the composition of the HFOs (ECB 2000b).

Table A5.12 (Appendix 5) presents Fuel Oil No. 6 acute toxicity data. Aquatic median lethal concentration (LC₅₀) values range from 0.9– greater than 10 000 mg/L. Oil-in-water dispersions have been shown to be not nearly as hazardous to aquatic organisms as the water-soluble fraction. The lowest marine toxicity value of 0.9 mg/L was determined in a 48-hour acute LC₅₀ test using water-soluble fractions with Mysidopsis almyra (a mysid shrimp) (Neff and Anderson 1981). The same value was determined in a 96-hour LC₅₀ with Capitella capitata (a marine worm) (Rossi et al. 1976). The lowest freshwater value of 4.1 mg/L was determined by MacLean and Doe (1989) in a 48-hour EC₅₀ test withDaphnia magna.

Avian Effects

HFOs can have a wide variety of effects on birds, especially sea birds. Heavy oils, including HFOs, can destroy the insulation provided by feathers, resulting in increased mortality due to exposure. HFOs are also directly toxic to birds through ingestion. The preening of feathers to clean them of oil, and the reduced insulation from oiled feathers increases metabolic requirements to the point where birds may starve to death while trying to keep warm.

As well, nesting birds that come into contact with fuel oils may transfer oil from their feathers and feet to their eggs during incubation. Toxicity to bird eggs via this route has been shown (Environment Canada 2010b; Michigan 2010). Fuel Oil No. 6 is similar to four of the HFOs considered here (64741-75-9, 70592-76-6, 70592-77-7 and 70592-78-8) and can be used in a read-across approach for toxicity. Szaro (1979) found that 5 µL of Fuel Oil No. 6 applied to eggs significantly reduced hatching success to 36% and 6-day survival to 52% in mallard ducks (Anas platyrhyncos).

Fuel Oil No. 2 is similar to the light HFO (CAS RN 68783-08-4) and can be used as a toxicity surrogate. In tests on mallard duck eggs, lowest-observed-effect concentrations (LOECs) were found at 1 μL/egg (20% reduction in hatchability with a 28% reduction in duckling survival post-hatch) (Albers and Szaro 1978; Szaro et al. 1978). Coon et al. (1979) determined that a 5 μL/egg treatment reduced hatchability by 28% compared with controls with eggs of great black-backed gull (Larus marinus). Common eider duck (Somateria mollissima), Louisiana heron (Hydranassa tricolour), laughing gull (Larus atricilla) and sandwich tern (Sterna sandvicensis) eggs experienced from 20–81% mortality at 20 μL/egg (Albers and Szaro 1978; White et al. 1979).

Terrestrial Compartment

The Canada-Wide Standards for Petroleum Hydrocarbons in Soil (CCME 2008) were used as a data source for effects of HFOs on terrestrial ecosystems. These standards were developed based on consideration of four fractions of total petroleum hydrocarbons (TPHs): F1 (C₆–C₁₀), F2 ( greater than C₁₀–C₁₆), F3 ( greater than C₁₆–C₃₄) and F4 ( greater than C₃₄). Fraction 3 (F3) is most like HFOs. Standards were developed for four land-use classes (agricultural, residential, commercial, industrial) and two soil types (coarse grained and fine grained). The land-use and soil type with the lowest standard is typically agricultural coarse-grained soils. The F3 standard for soil contact by non-human organisms for agricultural coarse-grained soils is 300 mg/kg dw (CCME 2008).

Ecological Exposure Assessment

Estimations of releases of these HFOs were made using data included in responses to a notice published under section 71 of CEPA 1999 (Environment Canada 2009), along with estimations of losses to the sea on Canada’s east coast by Risk Management Research Institute (RMRI 2007) and Environment Canada’s Spill Line database (Environment Canada 2011).

Aquatic Compartment

To determine the predicted environmental concentration (PEC) in water, the volume of water predicted to be in contact with spilled oil was provided by a report prepared by the Risk Management Research Institute (RMRI 2007). This work estimated the risk of oil spills in Hazard Zones around the southern coast of Newfoundland and Labrador based on the nature of the water (open or partially constricted), the type of vessels travelling through the zones, and the quantities of oil transported. The estimated volume of water in contact with spilled oil was dependent on the volume of oil spilled during the event and the hazard zone of the spill.

For the ship loading and unloading scenarios, the volume of water in contact with oil is from Hazard Zone 1, as this region includes loading operations at Whiffen Head and Come By Chance refinery in Newfoundland and Labrador (RMRI 2007). For the ship transport scenarios, the estimated volume of water in contact with oil is the average volume of water from Hazard Zone 2 (outer Placentia Bay), as this area is a major ship transportation corridor. The area of a slick created within Hazard Zones around Newfoundland was estimated for specific volume ranges of oil using ocean spill dispersion models, and then the volume of contacted water was estimated by multiplying the area by 10 to represent the top 10 meters of water. This estimate assumes that all of the water is equally contacted by the petroleum product spilled. This work was originally developed for crude oil, but it can be applied to HFOs as they have a similar density.

In the case of marine loading and unloading of HFOs by ship, an estimated 13 646 kg of fuel oil on average could be lost in one event to salt water (Table 3). At an average density of 1.04 kg/L (API 2004) this is equivalent to 83 barrels of fuel oil and is therefore expected to be in contact with 150 × 109 litres of water (Table A5.13 in Appendix 5). This volume is estimated from the enclosed waters found at wharves and loading terminals. The resulting concentration in water would be 0.09 mg/L (1.38 × 1010 mg/150 × 109 litres), which is considered the marine PEC for ship loading and unloading.

The situation is similar for marine transportation of HFOs by ship. In this case, 83 barrels of fuel oil is expected to be in contact with 6250 × 109 litres of water (Table A5.13 in Appendix 5). This volume is estimated from the open ocean of Placentia Bay. The resulting concentration in water would be 0.002 mg/L (1.38 × 1010 mg/6250 × 109 litres), which is considered the marine PEC for ship transport.

In the case of the freshwater loading and unloading of HFOs by ship, an estimated 15 262 kg of fuel oil could be lost in one event to fresh water (Table 3). At an average density of 1.04 kg/L (API 2004) this is equivalent to 92 barrels of fuel oil and is therefore expected to be in contact with 150 × 109 litres of water (Table A5.13 in Appendix 5). This volume is estimated from the enclosed waters found at wharves and loading terminals. The resulting concentration in water would be 0.1 mg/L (1.53 × 1010 mg/150 × 109 litres), which is considered the freshwater PEC for ship loading and unloading.

In the case of the freshwater transportation of HFOs by ship, an estimated 15 262 kg of fuel oil could be lost in one event to fresh water (Table 3). At an average density of 1.04 kg/L (API 2004) this is equivalent to 92 barrels of fuel oil and is therefore expected to be in contact with 6250 × 109 litres of water (Table A5.13 in Appendix 5). This volume is estimated from the open ocean of Placentia Bay. The resulting concentration in water would be 0.002 mg/L (1.53 × 1010 mg/6250 × 109 litres), which is considered the freshwater PEC for ship transport.

Terrestrial Compartment

Less than one release event per year for pipeline transport of HFOs is predicted based on the short distance of transport determined from information submitted under section 71 of CEPA 1999 (Environment Canada 2009) and the average spill rate per length of pipeline (1 spill per 11 100 km of pipeline, as found in NEB 2008). Likewise, only two of the five industry-restricted HFOs are transported by pipeline. From the historical Canadian data from the Spill Line database for Bunker C (Environment Canada 2011), only 13 spills of HFOs from pipelines were reported over 10 years (2000–2009). Thus, less than 1 release event per year is expected for pipeline loading, transport and unloading for industry-restricted HFOs.

It is estimated that there will be less than or equal to 1 release event per year each for train and truck loading, unloading and transport based on historical release information from the Spill Line database (Environment Canada 2011). Spill events are expected to generally occur at industrial facilities for industry-restricted HFOs. It was additionally considered that these infrequent releases would likely occur on a hard surface and not on soil; therefore releases from truck and train are not considered to be of high importance under these circumstances. It is expected that the actual release frequency for these industry-restricted HFOs is lower, as the Spill Line database release information was for Bunker C.

Characterization of Ecological Risk

The approach taken in this ecological screening assessment was to examine available scientific information and develop conclusions based on a weight-of-evidence approach as required under CEPA 1999. For each endpoint organism, an estimate of the potential to cause adverse effects and predicted no-effect concentration (PNEC) was determined. Also, a PEC was determined from the aquatic exposure scenario. The PNEC is the lowest critical toxicity value (CTV) for the organism of interest divided by an appropriate assessment factor. A risk quotient (RQ = PEC/PNEC) was calculated for each endpoint organism and is an important line of evidence in evaluating the potential risk to the environment.

Since a read-across approach can be used with Fuel Oil No. 6, the CTVs for this assessment are selected from empirical data available for Fuel Oil No. 6 (Table A5.12 in Appendix 5). For the marine scenarios, a CTV of 0.9 mg/L is selected based on the 48-hour acute LC₅₀ value for Mysidopsis almyra. For the freshwater exposure scenarios for ship loading/unloading and transport, the selected CTV is the 96-hour acute EC₅₀ (immobilization) of 4.1 mg/L for Daphnia magna (Table A5.12). An assessment factor of 10 is used to account for the extrapolation of modelled data to field effects.

Table 4 is the summary of the risk quotients for the industry-restricted HFOs. Only spills to marine water during the loading/unloading of ships were determined to be potentially harmful to fish, as the RQ greater than or equal to 1. 

For all aquatic spill scenarios, the critical spill volume for HFOs required to obtain an RQ = 1 and the frequency of spills above that threshold was determined from the Environment Canada Spill Line database (Environment Canada 2011) (see Table 5).

For the marine and freshwater scenarios during ship transport, critical spill volumes of 835 000 L and 6 700 000 L of fuel oil, respectively, are needed to obtain an RQ of 1 for aquatic organisms (Table 5) based on toxicity estimations and spill dispersion models of the volume of water affected. None of the reported spills from 2000–2009 were greater than these threshold volumes during ship transport and therefore, the expected number of spills per year above this volume is zero. As well, the whole dataset from the Environment Canada Spill Line database for Bunker C was used as a surrogate for these HFOs. The amount of HFOs released is unknown, but is certainly less than the total volume of Bunker C. Thus, spills to fresh water and salt water during transport are not considered harmful to aquatic organisms.

For the scenarios for ship loading/unloading, spill volumes of 13 000 L and 58 000 L of fuel oil are needed to obtain an RQ of 1 for aquatic organisms in marine and fresh waters, respectively (Table 5). There are some reported spills above these threshold volumes during the loading/unloading of ships in marine (15% of spills) and fresh water (8% of spills). However, these frequencies equate to an expected less than 2 spills per year above the threshold volume in marine waters and less than one spill per year in fresh water. These frequencies are based on the entire dataset from the Environment Canada Spill Line database for spills of Bunker C, which are expected to be more frequent than spills of industry-restricted HFOs. The RQ for marine loading was 1, and for fresh water loading the risk quotient was below 1, based on average spill volumes. Thus, based on the RQs and relatively low expected number of spills per year, spills of these HFOs to water during loading and unloading are considered to be infrequent and pose a low risk of harm to aquatic organisms.

These spill volumes were calculated based on models developed by RMRI (2007) relating the volume spilled and concentration of petroleum substance in the water. These models take into consideration dispersion of the petroleum substance spilled and, therefore, the calculated spill volume relating to a risk quotient of 1 is not for the acute, initial exposure to the spilled material. It is recognized that local, acute effects may occur during the inital phase of a spill before significant dispersion occurs.

Both field reports and experiments have shown that commercial blends of HFOs can be toxic to aquatic birds through ingestion (CONCAWE 1998; Environment Canada 2010b; Michigan 2010), contact with feathers (Environment Canada 2010b) and contact with eggs (Albers and Szaro 1978; Coon et al. 1979; CONCAWE 1998). The negative effects of oil on feathers, however, are not specific to HFOs and are primarily based on Bunker C fuel oil. Indeed, average spills to marine water for these industry-restricted HFOs are based on the Environment Canada Spill Line data for Bunker C fuel oil (Environment Canada 2011). Use of this data overestimates the number of spills of the industry-restricted HFOs considered in this assessment. Thus, there is a low frequency of releases of these industry-restricted HFOs to marine waters, and thus low risk to sea birds through direct toxicity and indirect effects.

Based on the estimated less than 1 HFO release event per year for pipeline transport, HFOs pose a low risk of harm to terrestrial non-human organisms.

With regard to truck and train releases, a risk quotient was not determined. Release frequency and volumes from trains are less certain due to a lack of definitive data. The Spill Line database reports small numbers of Bunker C spills via train (11 spills) and truck (32 spills) from 2000–2009. Considering the cause and reason of spill, it was determined that for each scenario of loading, transport and unloading of trains, less than 1 spill per year is expected. By the same analysis, less than or equal to 1 spill per year each for loading, transport and unloading by truck is expected. Thus, terrestrial impacts from train and truck transport of HFOs are unlikely to cause harm due to their low frequency (less than 1 spill per year for loading/unloading and transport). Likewise, the estimated spills from truck loading and unloading were not considered to be of high importance, as they would likely occur on a hard surface and not on soil.

Based on results from AOPWIN (2008), there would be a relatively rapid removal process if these HFOs are introduced into the atmosphere, based on oxidation half-lives of less than 1 day. With regard to the primary and ultimate biodegradation modelling, the C₃₀–C₅₀ isoalkanes, C₃₀–C₅₀ one-ring cycloalkanes, C₁₅–C₅₀ two-ring cycloalkanes, C₁₄–C₂₂ polycycloalkanes, C₃₀–C₅₀ one-ring aromatics, C₁₀–C₂₀ cycloalkane monoaromatics, C₁₅–C₅₀ two-ring aromatics, C₁₂–C₂₀ cycloalkane diaromatics, C₂₀–C₅₀ three-ring aromatics, C₁₆–C₂₀ four-ring aromatics, C₂₀–C₃₀ five-ring aromatics and C₂₂ six-ring aromatics in these HFOs meet or exceed the criteria for persistence (half-lives in soil and water greater than or equal to 182 days and half-life in sediment greater than or equal to 365 days) defined in the Persistence and Bioaccumulation Regulations (Canada 2000). Based on the available predicted information, these HFOs contain approximately 50–60% by weight of components that may persist sufficiently in soil, water and sediment to meet the regulatory criteria.

Based on the combined evidence of empirical data and predicted analysis of BCFs, BAFs, BMFs, TMFs and BSAFs, the HFOs assessed in this report may contain approximately 25 % by weight of components that meet the criteria for bioaccumulation as defined in the Persistence and Bioaccumulation Regulations (Canada 2000), but are not likely biomagnified in food webs. Both empirical and predicted BCFs and predicted BAFs are greater than or equal to 5000 for isoalkanes, cycloalkanes and some aromatic substances. There is consistent steady-state and kinetic evidence to suggest that these components do not metabolize very quickly and have sufficient dietary assimilation efficiency that, when tissue levels are compared with the bioavailable fraction in water, accumulation factors are expected to be high.

In general, fish can efficiently metabolize aromatic compounds.  Of the aromatic representative structures of HFOs with high bioaccumulation potential, only two (a C₂₀ cycloalkane diaromatic and a C₂₀ three-ring PAH) were bioaccumulative (i.e., BCF or BAF greater than 5000).  Both structures contain isoalkyl functional groups which may hinder biotransformation.  There is some evidence that alkylation increases bioaccumulation of naphthalene (Neff et al. 1976, Lampi et al. 2010) but it is not known if this can be generalized to larger PAHs or if any potential increase in bioaccumulation due to alkylation will be sufficient to exceed the Canadian criteria.

Some lower trophic level organisms (i.e., invertebrates) appear to lack the capacity to efficiently metabolize aromatic compounds, resulting in bioaccumulation that can be above Canadian criteria for some aromatic components of HFOs.  This is the case for the C₁₈ four-ring PAHs, C₂₀ five-ring PAHs, and C₂₂ six-ring PAHs, which were bioconcentrated to high levels (BCF greater than 5000) by invertebrates (e.g., Daphnia, molluscs) but not by fish.  There is potential for such bioaccumulative components to reach toxic levels in organisms if exposure is constant, continuous and of sufficient magnitude; however, this is unlikely in the water column following a spill scenario due to relatively rapid dispersal.

Bioaccumulation of aromatic compounds might be lower in natural environments than what is observed in the laboratory.  PAHs may sorb to organic material suspended in the water column (dissolved humic material) which decreases their overall bioavailability primarily due to an increase in size. This has been observed with fish (Weinstein and Oris 1999) and Daphnia (McCarthy et al. 1985).

As shown in Table A5.14 (Appendix 5), some components may meet both the persistence and bioaccumulation criteria in the Persistence and Bioaccumulation Regulations. The HFOs assessed in this report may contain approximately 15% of these components by weight. These include the C₁₅ dicycloalkanes, C₁₄ and C₂₂ polycycloalkanes, C₁₅–C₂₀ cycloalkane monoaromatics, C₂₀ three-ring aromatics, C₁₈ four-ring aromatics, C₂₀ five-ring aromatics and C₂₂ six-ring aromatics.

Based on the information presented in this screening assessment on the frequency and magnitude of spills, there is low risk of harm to organisms or the broader integrity of the environment from these substances. It is concluded that these industry-restricted HFOs (CAS RNs 64741-75-9, 68783-08-4, 70592-76-6, 70592-77-7 and 70592-78-8) do not meet the criteria under paragraph 64(a) or 64(b) of the Canadian Environmental Protection Act, 1999 (CEPA 1999) as they are not entering the environment in a quantity or concentration or under conditions that may have an immediate or long-term harmful effect on the environment or its biological diversity.

Uncertainties in Evaluation of Ecological Risk

This analysis addresses uncertainty associated with each component of the current assessment, including but not limited to selection of representative structures and quantification, exposure estimation, effects estimation, and risk characterization.

All modelling of the substance’s physical-chemical properties, as well as persistence, bioaccumulation and toxicity characteristics, is based on chemical structures. As these industry-restricted HFO are UVCBs, they cannot be represented by a single, discrete chemical structure. The specific chemical compositions of these HFOs are variable and not well defined. HFO streams under the same CAS RNs can vary significantly in the number, identity and proportion of components, depending on operating conditions, feedstocks and processing units. Therefore, for the purposes of modelling, a suite of representative structures that provide average estimates for the entire range of components likely present was identified. Specifically, these structures were used to assess the fate and hazard properties of HFOs. Given that more than one representative structure may be used for the same carbon range and type of component, it is recognized that structure-related uncertainties exist for these substances. The physical-chemical properties of 48 representative structures were used to estimate the overall behaviour of these HFOs, in order to represent the expected range in physical-chemical characteristics. Given the large number of potential permutations of the type and percentages of the structures in HFOs, there is uncertainty in the results associated with modelling.

Uncertainty arises from the non-uniformity of spill data. The available data on spills generally do not report values for each specific transported substance by CAS RN. For marine transportation, Environment Canada reported spills data for substances similar to these heavy fuel oils, specifically Bunker C fuel oil. Spill data specific to these industry-restricted HFOs are not available for each mode of transportation. The use of a generic loss fraction factor, derived from the available data, introduces uncertainty in the estimation of transportation releases.

Similarly, historical spills data classified as Bunker C fuel oil from the Emergencies Spill Line database (Environment Canada 2011) were used in the ship, truck and train transport release scenarios for these industry-restricted HFOs. The amount of HFOs released is unknown, but is certainly less than the total volume of Bunker C. There is uncertainty in the estimation of the actual HFO loading, transport and unloading releases.

The fate, food chain interactions and toxicity of a number of petroleum hydrocarbons depend to a large extent upon their chemical form. As such, conservative assumptions about chemical form, bioavailability, and absorption through the digestive tract were generally carried forward in the risk assessment. HFO representative structures were assessd with the conservative assumption that all of them were bioavailable.

This assessment involves the prediction of effects on biota using measured inputs and modelled accumulation or exposures. The process typically relies on modelled exposures for organisms at higher trophic levels. However, all models are simplifications of natural systems or processes, and therefore, rely on a number of assumptions. These, in turn, create uncertainties in the outcomes.

The BAF model calculations were derived from a large database of measured BAF values from the Great Lakes for chemicals that are poorly metabolized (e.g., PCBs). With metabolic biotransformation, the BAF model predictions are in general agreement with measured BAFs in fish. There is some uncertainty when estimating the biotransformation used by the model at the first trophic level. Many petroleum hydrocarbons are readily metabolized, somewhat by invertebrates and at much higher levels in fish.

The significance and impact of bioaccumulation is species specific and is dependent on a range of factors such as species, size and the environmental conditions. At present, there are no field data on the study of bioaccumulation of industry-restricted HFOs as a class; therefore, predicting effects is based on modelling their BAFs based on laboratory-acquired partitioning data.

Potential to Cause Harm to Human Health

Exposure Assessment

HFO substances are a group of heavy petroleum streams produced in oil refinery and upgrader facilities. Due to the physical-chemical properties of HFOs, the dermal route is an important route of occupational exposure. In a recent study to quantify such workplace exposures, Yvette et al. (2011) found that dermal exposures were generally low. However, the authors indicated that the presence of HFO components with some degree of carcinogenic potential identified in all of the HFO blends they investigated requires that control measures to maintain low dermal exposure levels should be strictly adhered to, and additional means of reducing HFO exposure even further should continue to be sought.

Due to the relatively low volatility of the industry-restricted HFOs (see Table 1) and relevant regulations in place to limit potential releases during handling of petroleum substances, general population exposure to these substances by ingestion and inhalation during loading and unloading is expected to be negligible and will not be considered further.

Significant concentrations of hydrogen sulfide are known to accumulate in the headspaces of storage tanks that contain HFOs. Heating of such tanks may cause the decomposition of some of the sulfur-containing compounds, which release hydrogen sulfide. There is also evidence that vapours of light hydrocarbons accumulate in the headspaces of HFO tanks (CONCAWE 1998).

The human health assessment of industry-restricted petroleum substances focuses on the fugitive releases that occur when petroleum substances escape into ambient air. These include evaporative emissions from tanks during the various modes of transportation of petroleum substances. The unintentional release (leaks or spills) data used in the ecological assessment are, for the purposes of assessing the potential to cause harm to human health, considered to refer to releases that occur on a non-routine or unpredictable basis in specific geographical locations. These unintentional releases (leaks or spills) typically do not contribute to the potential for exposure of the general population in Canada.

Evaporative emissions of the industry-restricted HFO substances during transit will enter ambient air. As such, inhalation is the primary potential route of bystander exposure, which may occur as the substances are being transported between facilities, and is therefore the focus of the current human health exposure assessment.

Inhalation from Ambient Air

As monitoring data on HFOs in the environment are not available, the HFO vapour level in ambient air was estimated using SCREEN3 (1996), a screening-level Gaussian air dispersion model based on the Industrial Source Complex (ISC) model (for assessing pollutant concentrations from various sources in an industry complex). The driver for air dispersion in the SCREEN3 model is wind. The maximum calculated exposure concentration is selected based on a built-in meteorological data matrix of different combinations of meteorological conditions, including wind speed, turbulence and humidity. This model directly predicts concentrations resulting from point, area and volume source releases. SCREEN3 estimates the maximum concentrations of a chemical at chosen receptor heights and at various distances for a given population in the vicinity of the release source in the direction downwind from the prevalent wind 1 hour after a given release event. During a 24-hour period, for point emission sources, the maximum 1-hour exposure as assessed by the ISC Version 3, is multiplied by a factor of 0.4 to account for variable wind directions. This gives maximum concentration within 24-hour exposure (U.S. EPA 1992). Similarly, for exposure events happening over the span of a year, it can be expected that the direction of the prevalent winds will be even more variable and uncorrelated to the wind direction for a single event; thus, the maximum exposure concentration for one year is determined by multiplying the maximum 1-hour exposure by a factor of 0.08. Such scaling factors are not required for non-point source emissions. However, to prevent overestimation of the exposures, we use a scaling factor of 0.2 to obtain the yearly exposure concentration from the value of the maximum 1-hour exposure determined from SCREEN3 calculations. Detailed input parameters for SCREEN3 are listed in Table A6.1 (Appendix 6). As a conservative estimate, the regular evaporative emission during 1 day of transit is assumed to originate from a defined area rather than a moving line source; as such, actual levels are expected to be lower, considering that the release source is typically moving.

Estimated regular evaporative emission to air during transit of industry-restricted HFOs is presented in Table A6.2 (Appendix 6) as a range to cover the losses from the various transportation modes involved. Formulas for evaporative emissions of HFOs from truck and train transit of HFOs are not available in the AP 42 guidelines (U.S. EPA 2008a). A conservative estimate for these transit losses may be calculated by using stationary storage tank formulas adapted to typical dimensions of truck and train tanks. Even at this level of conservatism, due to the low volatility of the HFOs, the evaporative emissions from truck and train transit are small. The upper value in the range is related to evaporative emission from ship transit. Emission rates in grams per second per square metre (g/s·m²) are derived based on the loss quantity of kilograms per day (Table A6.2 in Appendix 6) and the estimated emission areas for different transportation modes (Table A6.1 in Appendix 6). This emission rate (g/s·m²) was used for determining the concentration of the HFO vapours in ambient air by SCREEN3 (1996).

As evaporative emission quantities are different for various transportation modes, for those industry-restricted HFOs with more than one mode of transportation, the maximum concentrations of ambient HFO vapours during 24 hours are presented as a range in Table A6.3 (Appendix 6). A conservative estimate of exposure was chosen by using the maximum concentrations at 50 m (for transportation by trains), as these were the highest exposure values obtained, compared with those at distances farther from the release sources. The upper-bounding estimate of the maximum concentration in ambient air at 50 m was 1.28 µg/m³.

It should be noted that the estimated air concentrations of HFOs are considered to be conservative, as SCREEN3 is, by design, a conservative screening-level tool used as a rapid approach to estimate the air dispersion of various chemicals. Another consideration is that the releases of the industry-restricted HFO vapours that occur during the transit process occur continuously from a moving source (a line source) rather than from a stationary point source. As such, the actual concentration of the HFO vapours around a moving line source, for any given location, will be considerably lower than that represented by the total daily release quantity from a point release source, as was used in this assessment. Thus, the assumption of total daily evaporative emission within one defined area is considered to be a conservative estimate of the actual substance concentration in ambient air. Placing the receptor at 50 m from the release source is also conservative, as most Canadians do not reside within 50 m of HFO transport.

Health Effects Assessment

Given the limited number of studies available that specifically evaluate the health effects of the industry-restricted HFO substances, an adequately representative toxicological dataset unique to these substances could not be obtained. Therefore, to characterize the health effects of these HFOs, additional HFOs in the PSSA that are similar from both a process and a physical-chemical perspective were also considered. Because both the industry-restricted and the additional HFO substances have similar physical-chemical properties, their toxicological properties are likely similar. The health effects data were therefore pooled and used to construct a toxicological profile to represent all HFOs. Accordingly, the health effects of HFOs are represented as a group, not by individual CAS RNs.

Appendix 7 contains a summary of available health effects information on HFOs in laboratory animals. A summary of key studies is presented below. The HFO category of petroleum mixtures represented in Table A7.1 (Appendix 7) includes both residual fuels from distillation or cracking units and blended products. It consists of aromatic, aliphatic and cycloalkane hydrocarbons. Heavy fuels may also contain hydrogen sulfide, as well as a broad range of chemicals that are tumourigenic (e.g., PAHs), and the quantities present in HFOs can vary (CONCAWE 1998; Yvette et al. 2011).

HFOs have low acute toxicity. Inhalation exposure resulted in an LC₅₀ of greater than  3700 mg/m³ in rats. Oral exposure resulted in a median lethal dose (LD₅₀) of greater than  2000– greater than  25 000 mg/kg-bw in rats. Dermal exposure resulted in an LD₅₀ of greater than  2000– greater than  5350 mg/kg-bw in rabbits and greater than  2000 mg/kg-bw in rats (CONCAWE 1998; ECB 2000a; API 2004; U.S. EPA 2005). Minimal to moderate skin irritation was observed for acute dermal exposure. Available data indicate that HFOs and HFO components are not eye irritants (CONCAWE 1998).

In an acute oral study conducted for CAS RN 64741-62-4, a single dose of 2000 mg/kg-bw or a single dose of 125, 500 or 2000 mg/kg-bw was administered to pregnant Sprague-Dawley rats on one of gestation days 11–15 or on gestation day 12, respectively. Decreased maternal body weight gain and thymus weights were reported, regardless of treatment day, for the gestation day segment of the study. Dose-related decreased maternal body weight gain and thymus weights were reported for the dose-response segment of the study (Feuston et al. 1989; Feuston and Mackerer 1996).

One short-term inhalation study was conducted for CAS RN 64742-90-1. A lowest-observed-adverse-effect concentration (LOAEC) of 540 mg/m³ was observed for decreased body weight (concentration-related) and increased liver weight in Fischer 344 rats following administration of 540 or 2000 mg/m³, 6 hours/day for 9 days (Gordon 1983).

Short-term and subchronic dermal studies conducted over periods of 3 days to 13 weeks are available for HFO substances, including one industry-restricted substance (CAS RN 68783-08-4). Slight to severe skin irritation was observed in several studies; the lowest dose reported for skin irritation was 8 mg/kg-bw per day (Mobil 1994a, b). Selected systemic effects observed in these studies included decreased body weight gain and body weight, decreased thymus weights, increased liver weights and changes in hematological parameters (e.g., platelets, hemoglobin, red blood cells) and serum chemistry (i.e., liver enzymes and other indicators of liver toxicity) (API 1983; Mobil 1988, 1990, 1992, 1994a, b; UBTL 1990, 1994; Feuston et al. 1994, 1997). A lowest-observed-adverse-effect level (LOAEL) of 1 mg/kg-bw per day was reported for maternal toxicity following dermal exposure of pregnant CD rats to CAS RN 64741-62-4 at doses of 0.05, 1, 10, 50 or 250 mg/kg-bw per day from gestation days 0–19. Effects observed at the LOAEL included significantly decreased body weight gain, body weight and feed consumption, as well as decreased gravid uterine weight and the occurrence of red vaginal exudates (Hoberman et al. 1995). For subchronic exposure, a LOAEL of 8 mg/kg-bw per day was established following dermal exposure of male and female rats to CAS RN 64741-62-4 or 64741-81-7 at doses of 8, 30, 125, 500 or 2000 mg/kg-bw per day for 13 weeks. Effects noted at the LOAEL included decreased platelet counts and increased liver weights, as well as dose-related skin irritation (Mobil 1988, 1992, 1994b; Feuston et al. 1994, 1997). Lack of testing at doses lower than 8 mg/kg-bw per day lowers confidence in the LOAEL.

The genotoxicity of HFOs has been evaluated in invivo and in vitro assays. Results from in vivo genotoxicity testing of three HFO substances were mixed (i.e., both positive and negative results were obtained for the same assay and endpoint). Positive results were observed in mice and rats for micronuclei induction, sister chromatid exchange and unscheduled deoxyribonucleic acid (DNA) synthesis when HFOs were administered orally or by intraperitoneal injection (Khan and Goode 1984; API 1985a, b). Negative results were also observed for micronuclei induction, as well as for chromosomal aberrations (API 1985c; Mobil 1987a).

In vitro assays evaluating the genotoxicity of HFOs also exhibited mixed results. Positive results were obtained in the Ames test battery and mouse lymphoma assays, as well as for cell transformation and unscheduled DNA synthesis (Brecher and Goode 1983, 1984; Blackburn et al. 1984, 1986; API 1985c,d, 1986a; Mobil 1985; Feuston et al. 1994). Regarding CAS RN 68553-00-4, negative results were obtained in the Ames and mouse lymphoma assays, as well as for forward mutations and sister chromatid exchange (Farrow et al. 1983; Vandermeulen et al. 1985; Vandermeulen and Lee 1986). Additional negative results were observed only for one cytogenetic assay and one forward mutation assay for CAS RNs 64741-57-7 and 64741-62-4, respectively (API 1985e; Mobil 1987b). Equivocal results were observed in one forward mutation assay and one sister chromatid exchange assay and for cell transformation (Papciak and Goode 1984; API 1985f, 1986b).

The overall genotoxicity database indicates that although the results varied depending on the substance tested and the assay used, HFOs do exhibit genotoxic potential.

The European Commission has classified industry-restricted HFOs as Category 2 carcinogens (R45: may cause cancer) (European Commission 1994; ESIS 2008). The United Nations’ Globally Harmonized System of Classification and Labelling of Chemicals has classified these substances as Category 1B carcinogens (H350: may cause cancer) (European Commission 2008a). The International Agency for Research on Cancer (IARC) has classified residual (heavy) fuel oils as Group 2B carcinogens (possibly carcinogenic to humans) (IARC 1989a).

A number of skin-painting studies were conducted in laboratory animals to investigate the dermal carcinogenicity of HFOs using both chronic and initiation/promotion methodologies. Skin tumours, including both malignant carcinomas and benign papillomas, were frequently observed in mice, rabbits and monkeys (Smith et al. 1951; Shubik and Saffiotti 1955; Shapiro and Getmanets 1962; Saffiotti and Shubik 1963; Getmanets 1967; Weil and Condra 1977; Bingham and Barkley 1979; Sun Petroleum Products Co. 1979; Bingham et al. 1980; Lewis 1983; Blackburn et al. 1984, 1986; API 1989a, b; McKee et al. 1990). Exposure durations for the chronic studies ranged from 25 weeks to lifetime, with reported tumour latency periods ranging from 8 to 113 weeks. In several studies, however, the durations of exposures and latencies were not specified. In a chronic study, male mice were dermally treated with CAS RN 64741-62-4 at doses of 8.4, 16.8, 42.0, 83.8 or 167.6 mg/kg-bw, 3 times per week for a lifetime. Significant skin tumour formation was observed at all doses in a dose-response fashion (McKee et al. 1990). In the one initiation study that was identified, male mice were dermally treated with CAS RN 64741-62-4 at a dose of 16.8 mg/kg-bw once per day for 5 consecutive days. Significant skin tumour formation was observed at this dose. In the corresponding promotion study, no increase in histologically confirmed tumour incidence was observed. A statistically significant increase in the number of mice with gross masses (and shortened latency periods) was observed, however, indicating possible weak promoting activity (API 1989a).

Regarding the tumourigenicity of HFOs, it is recognized that these substances may contain appreciable concentrations of components that are tumourigenic, such as PAHs, and the quantity of this fraction can vary depending on the nature and amount of diluent fractions and whether the residue component is cracked or uncracked. The Government of Canada has previously completed a human health risk assessment of five PAHs, including a critical review of relevant data, under the Priority Substances Program. Based primarily on the results of carcinogenicity bioassays in animal models, these PAHs were classified as probably carcinogenic to humans: substances for which there is believed to be some chance of adverse effects at any level of exposure (Canada 1994). Due to the lack of exposure to HFOs, evaluating the contribution of HFO components to carcinogenic activity is beyond the scope of the current assessment.

HFOs have also been investigated for their reproductive and developmental effects. A LOAEL of 1 mg/kg-bw per day was identified for reproductive toxicity after dermal exposure of pregnant rat dams to CAS RN 64741-62-4 during gestation days 0–19 (the no-observed-adverse-effect level [NOAEL] was 0.05 mg/kg-bw per day). Reproductive effects included decreased number of live fetuses, increased incidences of resorptions and early resorptions and increased percentage of dead or resorbed conceptuses per litter. Fetal developmental variations were also observed in this study but were determined by the authors not to be treatment related (Hoberman et al. 1995). A LOAEL of 8 mg/kg-bw per day for treatment-related developmental toxicity was determined in another study, based on an increased incidence of fetal external abnormalities, including cleft palate, micrognathia (shortened lower jaw) and kinked tail, when catalytically cracked clarified oil was applied dermally to pregnant rats (Mobil 1987c; Feuston et al. 1989). These effects were noted to occur at low incidences. Reproductive toxicity and further developmental effects were observed at 30 mg/kg-bw per day. Reproductive effects included an increased incidence of resorptions and a decreased number of viable fetuses at and above 30 mg/kg-bw per day. At 250 mg/kg-bw per day, no viable offspring were produced (Mobil 1987c; Feuston et al. 1989). In another study, various HFO substances were applied dermally to rats. Substance-dependent LOAELs ranged from 30 to 500 mg/kg-bw per day based on fetal resorption rates ranging from 35.1–78.0% (Feuston et al. 1994).

Only one oral reproductive and developmental study was identified for any HFO substance. A LOAEL of greater than or equal to  125 mg/kg-bw was established based on a dose-related increase in resorptions (concomitant decrease in litter size), decreased fetal body weight and increased incidences of skeletal malformations in this acute study that exposed pregnant Sprague-Dawley rats to CAS RN 64741-62-4 during gestation (Feuston and Mackerer 1996). No reproductive or developmental studies were identified for any HFO substance via the inhalation route of exposure.

Although results varied depending on the substance tested, the overall weight of evidence suggests that HFOs exhibit reproductive and developmental toxicity in laboratory animals. The most sensitive LOAEL is 1 mg/kg-bw per day for reproductive and developmental effects.

Epidemiological data were not available for consideration in the human health effects evaluation of HFO substances.

Characterization of Risk to Human Health

Industry-restricted HFOs were identified as high priorities for action during categorization of the DSL because they were determined to present greatest potential or intermediate potential for exposure of individuals in Canada, and were considered to present a high hazard to human health. A critical effect for the initial categorization of industry-restricted HFO substances was carcinogenicity, based primarily on classifications by international agencies. These substances are classified as Category 2 carcinogens by the European Commission (European Commission 1994; ESIS 2008), Category 1B carcinogens using the Globally Harmonized System (European Commission 2008a) and Group 2B carcinogens by IARC (1989a). Several cancer studies conducted in laboratory animals resulted in the development of skin tumours following repeated dermal application of HFO substances (API 1989a; McKee et al. 1990). Skin carcinomas and papillomas developed in 100% of mice tested after 36 weeks of dermal exposure to an HFO substance at 167.6 mg/kg-bw per day, whereas tumours developed in 18% of the mice exposed to the lowest dose of 8.4 mg/kg-bw per day (McKee et al. 1990). HFOs demonstrated genotoxicity in in vivo and in vitro assays when applied dermally, and a mode of action for the induction of tumours involving direct interaction with genetic material cannot be precluded. There are no carcinogenicity studies by the inhalation route to inform the carcinogenic potential of these substances in the general population following inhalation exposure.

Given that the potential for general population exposure to the industry-restricted HFOs results primarily from inhalation of ambient air containing HFO vapours due to evaporative emissions during transportation and that the estimated maximum air concentration (1.28 µg/m³) is considered to be low, the risk to human health is likewise considered to be low. The conservative nature of the ambient air concentration estimated is highlighted by a bystander being placed at 50 m and the assumption of total daily evaporative emissions occurring within a defined geographic area from a stationary point source (under normal operating conditions, evaporative emissions occur predominantly from a moving source; thus, the releases are diluted across a large geographic area).

General population exposure to industry-restricted HFOs via the dermal and oral routes is not expected; therefore, risk to human health from these routes of exposure is not expected.

With respect to non-cancer effects, decreased body weights and increased liver weights in rats were the primary adverse effects observed following a short-term repeated inhalation exposure of 6 hours/day for 9 days. A critical LOAEC of 540 mg/m³ was reported in the single available inhalation study. Comparison of this critical effect level for inhalation exposure in rats with the estimated maximum daily exposure concentration of 1.28 µg/m³ in ambient air results in an MOE of approximately 420 000. The margin is considered more than adequately protective to account for the uncertainties in the data set for the human health risk assessment for both cancer and non-cancer effects, especially in light of the highly conservative nature of the estimated exposures.

Uncertainties in Evaluation of Human Health Risk

The PSSA screening assessments evaluate substances that are complex combinations of hydrocarbons (UVCBs) composed of a number of substances in various proportions due to the source of the crude oil or bitumen and its subsequent processing. Monitoring information or provincial release limits from petroleum facilities target broad releases, such as releases of oils and grease, to water or air. These widely encompassing release categories do not allow for the detection of individual complex mixtures or production streams. As such, the monitoring of broad releases cannot provide sufficient data to associate a detected release with a specific substance identified by a CAS RN, nor can the proportion of releases attributed to individual CAS RNs be defined.

Uncertainty exists by using empirical equations for the estimation of evaporative emissions. It is noted that the transit evaporative emissions also vary with physical conditions, such as the tightness of transport vessels or the valve settings. The screening estimation of evaporative emissions does not account for these.

There is uncertainty regarding the conservative estimation of human exposure because of the lack of monitoring data of HFOs in the ambient air and the use of modelling. SCREEN3 modelling of the dispersion profile of HFO vapours requires limited input parameters and non-site-specific meteorological data. These assumptions will introduce more uncertainty compared with other complex dispersion models (Tables A6.1 and A6.2 in Appendix 6).

Because the relative differences in absorption of HFOs through the inhalation, dermal and oral routes of exposure are not well documented, a conservative assumption of 100% absorption was made. Thus, the internal (systemic) doses were considered to be equivalent to the external doses that were used for treatment of the laboratory animals.

As the industry-restricted HFOs are UVCBs, their specific compositions are not well defined. HFO streams under the same CAS RN can vary significantly in the number, identity and proportion of components. Consequently, it is difficult to obtain a representative toxicological dataset for these specific HFO CAS RNs. For this reason, all available health effects data on HFO substances were pooled across CAS RNs to develop a comprehensive toxicological profile. More research by the scientific community or the petroleum sector to elucidate the compositions of petroleum substances would allow for better characterization of the potential health risks associated with possible exposure to these substances.

Uncertainty also exists due to the paucity of data available regarding the physical-chemical properties of certain HFOs. The densities of the specific CAS RNs were not provided in the health effects studies; thus, these properties were often obtained from alternative sources. However, because each sample of a particular CAS RN can be slightly different in its composition (as stated previously), these properties may not be entirely representative of a specific sample tested in any one study.

Uncertainty also exists because certain details of the laboratory animals (i.e., sex, strain, body weight and minute volume) were often not stated in the health effects studies and were obtained from laboratory standard data. Thus, these data may not be entirely representative of the physical features of the actual test animals used in the studies.

Conclusion

Based on comparison of levels expected to cause harm to organisms with estimated exposure levels, these HFOs have low risk to cause harm to aquatic life in the confined marine waters around loading wharfs due to the low estimated frequency of -- and, hence, exposure to the environment from -- unintentional spills of these HFOs during ship loading.

Based on the information presented in this screening assessment on the frequency and magnitude of spills, there is low risk of harm to organisms or the broader integrity of the environment from these substances. It is concluded that the five industry-restricted HFOs (CAS RNs 64741-75-9, 68783-08-4, 70592-76-6, 70592-77-7 and 70592-78-8) do not meet the criteria under paragraphs 64(a) or 64(b) of CEPA 1999, 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.

Based on the information presented in this screening assessment, the critical effect for the initial categorization of risk to human health was carcinogenicity. However, because the estimates of exposure indicate that the potential exposure of the general population to industry-restricted HFOs from ambient air is expected to be very low, resulting in an extraordinarily large MOE (approximately 420 000), the likelihood of inhalation exposure of Canadians is considered to be very low. Exposure of the general population to industry-restricted HFOs via the dermal and oral routes is not expected. Therefore, based on the adequacy of the margins between estimated exposure to industry-restricted HFO substances and critical effect levels, it is concluded that the five industry-restricted HFOs (CAS RNs 64741-75-9, 68783-08-4, 70592-76-6, 70592-77-7 and 70592-78-8) do not meet the criteria under paragraph 64(c) of CEPA 1999, 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 concluded that these five industry-restricted HFOs listed under CAS RNs 64741-75-9, 68783-08-4, 70592-76-6, 70592-77-7 and 70592-78-8 do not meet any of the criteria set out in section 64 of CEPA1999.

References

Albers PH, Szaro RC. 1978. Effects of no. 2 fuel oil on common eider eggs. Mar Pollut Bull 9:138–139.

Anderson JW, Neff JM, Cox BA, Tatem HE, Hightower GM. 1974. Characteristics of dispersions and water-soluble extracts of crude and refined oils and their toxicity to estuarine crustaceans and fish. Mar Biol 27:75–88.

[AOPWIN] Atmospheric Oxidation Program for Microsoft Windows [Estimation Model]. 2008. Version 1.92a. Washington (DC): U.S. Environmental Protection Agency, Office of Pollution Prevention and Toxics; Syracuse (NY): Syracuse Research Corporation. Available from: www.epa.gov/oppt/exposure/pubs/episuite.htm

[API] American Petroleum Institute. 1983. A 28-day dermal toxicity study of catalytic cracked clarified oil, API sample 81-15. Study conducted by Borriston Laboratories Inc. Washington (DC): API. API Medical Research Publication No.: 30-32854. [cited in CONCAWE 1998; ECB 2000a, b].

[API] American Petroleum Institute. 1985a. In-vivo sister chromatid exchange (SCE) assay. API sample 81-15 catalytically cracked clarified oil (CAS 64741-62-4). Study conducted by Microbiological Associates, Inc. Washington (DC): API. API Health and Environmental Sciences Department Report No.: 32-32254. [cited in CONCAWE 1998; ECB 2000a; API 2004].

[API] American Petroleum Institute. 1985b. An evaluation of the potential of RO-1, 81-15 and PS8-76D5-SAT to induce unscheduled DNA synthesis (UDS) in the in-vivo/in-vitro hepatocyte DNA repair assay. Study conducted by SRI International. Washington (DC): API. API Medical Research Publication No.: 32-32406. [cited in CONCAWE 1998; ECB 2000a; API 2004].

[API] American Petroleum Institute. 1985c. Mutagenicity evaluation studies of catalytically cracked clarified oil API sample 81-15 in the rat bone marrow cytogenetic assay and in the mouse lymphoma forward mutation assay. Study conducted by Litton Bionetics, Inc. Washington (DC): API. API Medical Research Publication No.: 32-30534. [cited in CONCAWE 1998; ECB 2000a; API 2004].

[API] American Petroleum Institute. 1985d. An evaluation of the potential of RO-1, 81-15 and PS8-76D5-SAT to induce unscheduled DNA synthesis (UDS) in primary rat hepatocyte cultures. Study conducted by SRI International. Washington (DC): API. API Medical Research Publication No.: 32-32407. [cited in CONCAWE 1998; ECB 2000a; API 2004].

[API] American Petroleum Institute. 1985e. CHO/HGPRT (Chinese hamster ovary/hypoxanthine-guanine phosphoribosyl transferase) mammalian cell forward gene mutation assay of API sample 81-15 catalytic cracked clarified oil (CAS 64741-62-4). Study conducted by Pharmakon Research International, Inc. Washington (DC): API. API Medical Research Publication No.: 32-32118. [cited in CONCAWE 1998; ECB 2000a; API 2004].

[API] American Petroleum Institute. 1985f. Sister chromatid exchange (SCE) assay in Chinese hamster ovary (CHO) cells. API sample 81-15 catalytically cracked clarified oil (CAS 64741-62-4). Study conducted by Microbiological Associates, Inc. Washington (DC): API. API Health and Environmental Sciences Department Report No.: 32-32750. [cited in CONCAWE 1998; ECB 2000a; API 2004].

[API] American Petroleum Institute. 1986a. Salmonella/mammalian-microsome plate incorporation mutagenicity assay (Ames test). Study conducted by Microbiological Associates, Inc. Washington (DC): API. API Medical Research Publication No.: 33-30599. [cited in CONCAWE 1998; ECB 2000a; API 2004].

[API] American Petroleum Institute. 1986b. Morphological transformation of BALB/3T3 mouse embryo cells. API sample 81-15 catalytically cracked clarified oil (CAS 64741-62-4). Study conducted by Microbiological Associates, Inc. Washington (DC): API. API Health and Environmental Sciences Department Report No.: 33-32638. [cited in in CONCAWE 1998; ECB 2000a; API 2004].

[API] American Petroleum Institute. 1989a. Short-term dermal tumorigenesis study of selected petroleum hydrocarbons in male CD-1 mice. Initiation and promotion phases. Study conducted by IIT Research Institute. Washington (DC): API. API Health and Environmental Sciences Department Report No.: 36-32643. [cited in CONCAWE 1998].

[API] American Petroleum Institute. 1989b. Lifetime dermal carcinogenesis/chronic toxicity screening bioassay of refinery streams in C3H/HeJ mice (AP-135r). Study conducted by Primate Research Institute, New Mexico State University. Washington (DC): API. API Health and Environmental Sciences Department Report No.: 36-31364. [cited in CONCAWE 1998; ECB 2000a].

[API] American Petroleum Institute. 2004. High Production Volume (HPV) Challenge Program. Robust summary of information on heavy fuel oils. Washington (DC): API. Available from: www.epa.gov/hpv/pubs/summaries/heavyfos/c15368rs.pdf

Arnot JA. 2005. Bioconcentration factor and bioaccumulation factor assessments for organic chemicals on the Canadian domestic substances list: Task 1: Supplementation of Environment Canada’s BCF database. Report to Environment Canada, New Substances Branch.

Arnot JA. 2006. Bioconcentration factor and bioaccumulation factor assessments for organic chemicals on the Canadian domestic substances list: Database update. Report to Environment Canada, New Substances Branch.

Arnot J, Gobas F. 2003. A generic QSAR for assessing the bioaccumulation potential of organic chemicals in aquatic food webs. QSAR Comb Sci [Internet] [cited 2007 Mar 15]; 22(3):337–345. Available from: www3.interscience.wiley.com/journal/104557877/home [restricted access].

Arnot J, Gobas F. 2004. A food web bioaccumulation model for organic chemicals in aquatic ecosystems. Environ Toxicol Chem 23:2343-2355.

Arnot J, Gobas F. 2006. A review of bioconcentration factor (BCF) and bioaccumulation factor (BAF) assessments for organic chemicals in aquatic systems. Environ Rev 14:257–297.

Arnot JA, Mackay D, Bonnell M. 2008a. Estimating metabolic biotransformation rates in fish from laboratory data. Environ Toxicol Chem 27(2):341–351.

Arnot JA, Mackay D, Parkerton TF, Bonnell M. 2008b. A database of fish biotransformation rates for organic chemicals. Environ Toxicol Chem 27(11):2263–2270.

Arthurs P, Stiver WH, Zytner RG. 1995. Passive volatilization of gasoline from soil. J Soil Contam. 4: 1-13.

Atkinson R. 1990. Gas-phase tropospheric chemistry of organic compounds: a review. Atmos Environ 24A:1–41.

[ATSDR] Agency for Toxic Substances and Disease Registry. 1999. Toxicological profile for total petroleum hydrocarbons (TPH). Atlanta (GA): U.S. Department of Health and Human Services, Public Health Service, ATSDR. [cited 2009 Mar]. Available from: www.atsdr.cdc.gov/toxprofiles/tp.asp?id=424&tid=75

Banerjee S. 1984. Solubility of organic mixtures in water.Environ Sci Technol 18:587–591.

Barron MG. 1990. Bioconcentration. Environ Sci Technol 24:1612–1618.

[BCFBAF] Bioconcentration Factor and Bioaccumulation Factor Estimation Program for Windows [Estimation Model]. 2008. Version 4.00. Washington (DC): U.S. Environmental Protection Agency, Office of Pollution Prevention and Toxics; Syracuse (NY): Syracuse Research Corporation. Available from: www.epa.gov/oppt/exposure/pubs/episuite.htm

Bingham E, Barkley W. 1979. Bioassay of complex mixtures derived from fossil fuels. Environ Health Perspect 30:157–163. [cited in IARC 1984].

Bingham E, Trosset RP, Warshawsky D. 1980. Carcinogenic potential of petroleum hydrocarbons: a critical review of the literature. J Environ Pathol Toxicol 3:483–563. [cited in IARC 1989a].

[BioHCWin] Biodegradation of Petroleum Hydrocarbons Estimation Program for Windows [Estimation Model]. 2008. Version 1.01a. Washington (DC): U.S. Environmental Protection Agency, Office of Pollution Prevention and Toxics; Syracuse (NY): Syracuse Research Corporation. Available from: www.epa.gov/oppt/exposure/pubs/episuite.htm

[BIOWIN] Biodegradation Estimation Program for Windows [Estimation Model]. Version 4.10. 2009. Washington (DC): U.S. Environmental Protection Agency, Office of Pollution Prevention and Toxics; Syracuse (NY): Syracuse Research Corporation. Available from:www.epa.gov/oppt/exposure/pubs/episuite.htm

Blackburn GR, Deitch RA, Schreiner CA, Mehlman MA, Mackerer CR. 1984. Estimation of the dermal carcinogenic activity of petroleum fractions using a modified Ames Assay. Cell Biol Toxicol 1:67–80.

Blackburn GR, Deitch RA, Schreiner CA, Mackerer CR. 1986. Predicting carcinogenicity of petroleum distillation fractions using a modified Salmonella mutagenicity assay. Cell Biol Toxicol 2:63–84.

Boethling RS, Howard PH, Beauman JA, Larosche ME. 1995. Factors for intermedia extrapolations in biodegradability assessment. Chemosphere 30(4):741–752.

Brecher S, Goode JW. 1983. BALB/3T3 transformation test. Aromatic pyrolysis oil. Project No. 2084. Prepared for Gulf Oil Chemicals Co., Houston, TX. Pittsburgh (PA): Gulf Life Sciences Center. [cited in U.S. EPA 2005].

Brecher S, Goode JW. 1984. Hepatocyte primary culture/DNA repair test of aromatic pyrolysis oil. Project No. 2083. Prepared for Gulf Oil Chemicals Co., Houston, TX. Pittsburgh (PA): Gulf Life Sciences Center. [cited in U.S. EPA 2005].

Broman DC, Lindberg NC, Zwbuhr Y. 1990. An in-situ study on the distribution, biotransformation and flux of polycyclic aromatic hydrocarbons (PAHs) in an aquatic food chain (seston-Mytilus edulis L.-Somateria mollissima) from the Baltic: an ecotoxicological perspective. Environ Toxicol Chem 9:429–442.

Brost EJ, DeVaull GE. 2000. Non-aqueous phase liquid (NAPL) mobility limits in soil. API Soil and Groundwater Research Bulletin No. 9. Washington (DC): American Petroleum Institute (API).

Bruner KA, Fisher SW, Landrum PF. 1994. The role of the zebra mussel, Dreissena polymorpha, in contaminant cycling: I. The effect of body size and lipid content on the bioconcentration of PCBs and PAHs. J Great Lakes Res 10(4):725–734.

Burkhard L, Lukasewycz M. 2000. Some bioaccumulation factors and biota–sediment accumulation factors for polycyclic aromatic hydrocarbons in lake trout. Environ Toxicol Chem 19(5):1427–1429.

Byrne CJ, Calder JA. 1977. Effect of the water-soluble fractions of crude, refined and waste oils on the embryonic and larval stages of the quahog clam Mercenaria sp. Mar Biol 40:225–231.

Canada. 1994. Polycyclic aromatic hydrocarbons [Internet]. Ottawa (ON): Environment Canada; Health Canada. (Priority substances list assessment report). Available from: www.hc-sc.gc.ca/ewh-semt/alt_formats/hecs-sesc/pdf/pubs/contaminants/psl1-lsp1/hydrocarb_aromat_polycycl/hydrocarbons-hydrocarbures-eng.pdf

Canada. 1999. Canadian Environmental Protection Act, 1999. S.C., 1999, c. 33. Canada Gazette, Part III, vol. 22, no. 3. Available from: www.gazette.gc.ca/archives/p3/1999/g3-02203.pdf

Canada. 2000. Canadian Environmental Protection Act, 1999: Persistence and Bioaccumulation Regulations, P.C. 2000-348, 23 March 2000, SOR/2000-107. Canada Gazette, Part II, vol. 134, no. 7, p. 607−612. Available from: www.gazette.gc.ca/archives/p2/2000/2000-03-29/pdf/g2-13407.pdf

Canada. 2001. Canada Shipping Act, 2001. S.C., 2001, c. 26. Canada Gazette, Part III, vol. 24, no. 4, chapter 26. Available from: www.gazette.gc.ca/archives/p3/2001/g3-02404.pdf

Canada. 2010. Petroleum Refinery Liquid Effluent Regulations (C.R.C., c. 828) to the Fisheries Act. [cited 2010 Feb 19]. Available from: http://laws.justice.gc.ca/eng/C.R.C.-c.828/index.html

[CanLII] Canadian Legal Information Institute [databases on the Internet]. 2001– . Ottawa (ON): CanLII. [cited 2009]. Available from: www.canlii.org/en/index.php

Carlson RM, Oyler AR, Gerhart EH, Caple R, Welch KJ, Kopperman HL, Bodenner D, Swanson D. 1979. Implications to the aquatic environment of polynuclear aromatic hydrocarbons liberated from northern great plains coal. U.S. EPA Environmental Research Laboratory (EPA 600/3-79-093). [cited in Lampi and Parkerton 2009].

[CATABOL] Probabilistic assessment of biodegradability and metabolic pathways [Computer Model]. c2004-2008. Version 5.10.3. Bourgas (BG): Bourgas Prof. Assen Zlatarov University, Laboratory of Mathematical Chemistry. Available from: http://oasis-lmc.org/?section= software&swid=1

[CCME] Canadian Council of Ministers of the Environment. 2008. Canada-wide standard for petroleum hydrocarbons (PHC) in soil. User Guidance. Winnipeg (MB): Canadian Council of Ministers of the Environment. Report number PN-1398. Available from: http://www.ccme.ca/assets/pdf/pn_1398_phc_user_guide_1.1_e.pdf

Charbeneau RJ, Linz DG, Newell CJ, Rixey WG, Ryan JA. 1998. Proceedings from the Pellston workshop on assessing contaminated soils: from soil-contaminant interactions to ecosystem management. Pellston (MI): Pellston Workshop on Assessing Contaminated Soils. p. 23–27.

[CITI] Chemicals Inspection and Testing Institute. 1992. Bioaccumulation and Biodegradation Data on Existing Chemicals Based on the CSCL Japan. Tokyo, Japan.

Colombo J, Cappelletti N, Migoya M, Speranza E. 2007. Bioaccumulation of anthropogenic contaminants by detritivorous fish in the Río de la Plata estuary: 1--Aliphatic hydrocarbons. Chemosphere 68:2128–2135.

[CONCAWE] CONservation of Clean Air and Water in Europe. 1998. Heavy fuel oils. Prepared by CONCAWE’s Petroleum Products and Health Management Groups. Brussels (BE): CONCAWE. Product Dossier No.: 98/109.

Coon NC, Albers PH, Szaro RC. 1979. No. 2 fuel oil decreases embryonic survival of great black-backed gulls. Bull Environ Contam Toxicol 21:152–156.

Cravedi J, Tulliez J. 1981. Distribution and elimination routes of a naphthenic hydrocarbon (dodecylcyclohexane) in rainbow trout (Salmo gairdneri). Bull Environ Contam Toxicol 26:337–344.

Cravedi J, Tulliez J. 1983. Hydrocarbon disposition, lipid content, and fatty acid composition in trout after long-term dietary exposure to n-alkanes. Environ Res 32(2):398–413.

Curry P, Kramer G, Newhook R, Sitwell J, Somers D, Tracy B, Oostdam JV. 1993. Reference values for Canadian populations. Prepared by the Environmental Health Directorate Working Group on Reference Values. Ottawa (ON): Health Canada. [update of 1988 version].

De Maagd PGJ. 1996. Polycyclic aromatic hydrocarbons: fate and effects in the aquatic environment. Ph.D. Thesis. Utrecht University. Utrecht, NL. [cited in European Commission 2008b].

De Voogt P, van Hattum B, Leonards P, Klamer JC, Govers H. 1991. Bioconcentration of polycyclic heteroaromatic hydrocarbons in the guppy (Poecilia reticula). Aquat Toxicol 20:169–194.

[ECB] European Chemicals Bureau. 2000a. IUCLID dataset for clarified oils (petroleum), catalytic cracked (CAS No. 64741-62-4). Available from: http://esis.jrc.ec.europa.eu/doc/IUCLID/data_sheets/64741624.pdf

[ECB] European Chemicals Bureau. 2000b. IUCLID dataset for fuel oil, No. 6. (CAS No. 68553-00-4). Available from: http://esis.jrc.ec.europa.eu/doc/IUCLID/data_sheets/68553004.pdf

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2004a. Fish, dietary bioaccumulation study, study no. 0400947. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi et al. 2010].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2004b. Fish, aqueous bioaccumulation study, study no. 0409544. Annandale (NJ): ExxonMobil Biomedical Sciences Inc.

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2005a. Fish, dietary bioaccumulation study, study no. 0409547. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi et al. 2010].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2005b. Fish, aqueous bioaccumulation test, study no. 0523644. Annandale (NJ): ExxonMobil Biomedical Sciences Inc.

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2005c. Fish, dietary bioaccumulation study, study no. 0523747. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi et al. 2010].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2005d. Fish, dietary bioaccumulation study, study no. 0531047. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi et al. 2010].

[EMBSI Exxon Mobil Biomedical Sciences, Inc. 2006a. Fish, dietary bioaccumulation study, study no. 0674247. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi et al. 2010].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2006b. Fish, dietary bioaccumulation test, study no. 0681647. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi et al. 2010].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2007a. Fish, dietary bioaccumulation study, study no. 0796347T. ExxonMobil Biomedical Sciences Inc. Annandale, NJ. [cited in Lampi and Parkerton 2009].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2007b. Fish, dietary bioaccumulation study, study no. 0796347C. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi and Parkerton 2009].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2008a. Fish, dietary bioaccumulation study, study no. 0711047. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi et al. 2010].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2008b. Fish, dietary bioaccumulation study, study no. 0821047. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi et al. 2010].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2008c. Fish, dietary bioaccumulation study, study no. 0818447. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi et al. 2010].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2009. Fish, dietary bioaccumulation study, study no. 0818447. Annandale (NJ): ExxonMobil Biomedical Sciences Inc. [cited in Lampi et al. 2010].

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2010a. Fish, dietary bioaccumulation study, study no. 0818447. Supplement. Annandale (NJ): ExxonMobil Biomedical Sciences Inc.

[EMBSI] Exxon Mobil Biomedical Sciences, Inc. 2010b. Fish, dietary bioaccumulation study, study no. 0681647. Annandale (NJ): ExxonMobil Biomedical Sciences Inc.

Environment Canada. 1994. Polycyclic aromatic hydrocarbons. ISBN 0-662-22209-1, Cat. No. En40-215/42E. Available from: http://www.hc-sc.gc.ca/ewh-semt/pubs/contaminants/psl1-lsp1/hydrocarb_aromat_polycycl/index-eng.php

Environment Canada. 2007. Guidance for conducting ecological assessments under CEPA 1999. Science Resource Technical Series, Draft Module on QSARs. Draft document prepared by Environment Canada, Existing Substances Division.

Environment Canada. 2008. Data for petroleum sector stream substances collected underthe Canadian Environmental Protection Act, 1999, section71:Notice with respect to certain high priority petroleum substances. Data prepared by: Environment Canada, Oil, Gas, and Alternative Energy Division.

Environment Canada. 2009. Data for petroleum sector stream substances collected underthe Canadian Environmental Protection Act, 1999, section71: Notice with respect to potentially industry-limited high priority petroleum substances. Data prepared by: Environment Canada, Oil, Gas, and Alternative Energy Division.

Environment Canada. 2010a. Oil properties database. Available from: http://www.etc-cte.ec.gc.ca/databases/OilProperties/oil_prop_e.html

Environment Canada. 2010b. Birds oiled at sea. [accessed 2010 Feb 17]. Available from: www.atl.ec.gc.ca/boas/index_e.html

Environment Canada. 2010c. CEPA Environmental Registry. Toxic Substances List – Schedule 1. Available from: www.ec.gc.ca/lcpe-cepa/default.asp?lang=En&n=0DA2924D-1

Environment Canada. 2011. Spill Line data report on petroleum product spills 2000–2009. Gatineau (QC): Environment Canada, Environmental Emergencies Division. Unpublished information.

[EPI Suite] Estimation Programs Interface Suite for Microsoft Windows [Estimation Model]. 2008. Version 3.4. Washington (DC): U.S. Environmental Protection Agency, Office of Pollution Prevention and Toxics; Syracuse (NY): Syracuse Research Corporation. Available from: www.epa.gov/oppt/exposure/pubs/episuite.htm

[EQC] Equilibrium Criterion Model. 2003. Version 2.02. Peterborough (ON): Trent University, Canadian Environmental Modelling Centre. Available from: www.trentu.ca/cemc/

[ESIS] European Chemical Substances Information System [database on the Internet]. 2008. Database developed by the European Chemicals Bureau (ECB). [cited 2011 Aug 29]. Available from: http://esis.jrc.ec.europa.eu/

European Commission. 1994. Commission Directive 94/69/EC of 19 December 1994. Annex II. Official Journal of the European Communities. 31.12.94. L381, vol. 37. European Commission. 21st ATP.

European Commission. 2008a. Details on substances classified in Annex VI to Regulation (EC) No 1272/2008 [database on the Internet]. Database developed by the European Commission. [cited 2011 Aug 29]. Available from: http://esis.jrc.ec.europa.eu/index.php?PGM=cla

European Commission. 2008b. European Union risk assessment report. Coal-tar pitch, high temperature. CAS No.: 65996-93-2. EINECS No.:266-028-2. Risk assessment. Environment. May 2008. Luxembourg (NL): Office for Official Publications of the European Communities. Available from: http://ecb.jrc.ec.europa.eu/DOCUMENTS/Existing-Chemicals/RISK_ASSESSMENT/REPORT/pitchcoaltarreport323.pdf

Evans MS, Landrum PF. 1989. Toxicokinetics of DDE, benzo(a)pyrene, and 2,4,5,2’, 4’,5’-hexachlorobiphenyl in Pontoporeia hoyi and Mysis relicta. J Great Lakes Res 15:589–600.

Farrow MG, McCarroll N, Cortina T, Draus M, Munson A, Steinberg M, Kirwin C, Thomas W. 1983. In vitro mutagenicity and genotoxicity of fuels and paraffinic hydrocarbons in the Ames, sister chromatid exchange, and mouse lymphoma assays. Toxicologist 3:36 (Abstract No. 144). [cited in IARC 1989a].

Feuston MH, Mackerer CR. 1996. Developmental toxicity of clarified slurry oil, syntower bottoms, and distillate aromatic extract administered as a single oral dose to pregnant rats. J Toxicol Environ Health 49(1):45–66.

Feuston MH, Kerstetter SL, Singer EJ, Mehlman MA. 1989. Developmental toxicity of clarified slurry oil applied dermally to rats. Toxicol Ind Health5(3):587–599.

Feuston MH, Low LK, Hamilton CE, Mackerer CR. 1994. Correlation of systemic and developmental toxicities with chemical component classes of refinery streams. Fundam Appl Toxicol 22:622–630.

Feuston MH, Hamilton CE, Mackerer CR. 1997. Systemic and developmental toxicity of dermally applied syntower bottoms in rats. Fundam Appl Toxicol 35:166–176.

Fox K, Zauke G, Butte W. 1994. Kinetics of bioconcentration and clearance of 28 polychlorinated biphenyl congeners in zebrafish (Brachydanio rerio). Ecotoxicol Environ Saf 28(1):99–109.

Fuhr B. 2008. Hydrocarbon composition of fuel products for risk assessment modelling. Final Report. Edmonton (AB): Alberta Research Council, Fuels and Lubricants Laboratory for Environment Canada..

Getmanets IY. 1967. Comparative evaluation of the carcinogenic properties of cracking residues of high- and low-paraffin petroleum. Gig Tr Prof Zabol 11:53–55. [cited in IARC 1989b].

Giddings JM, Parkhurst BR, Gehrs CW, Millemann RE. 1980. Toxicity of a coal liquefaction product to aquatic organisms. Bull Environ Contam Toxicol 25:1–6. [cited in CONCAWE 1998].

Gobas F, Clark K, Shiu W, Mackay D. 1989. Bioconcentration of polybrominated benzenes and biphenyls and related superhydrophobic chemicals in fish: role of bioavailability and elimination into the feces. Environ Toxicol Chem 8:231–245.

Golder Associates Ltd. 2006. Final Draft Report on Lakewater and Aquatics Long-term Monitoring – 2005–2006 Interpretive Report, Wabamun Lake Derailment Site, Canadian National Railway Company. PIN 2401257. Wabamun, Alberta.

Gordon T. 1983. Nine-day repeated dose inhalation toxicity study in rats. Aromatic pyrolysis oil. Project No. 2035. Prepared for Gulf Oil Chemicals Co., Houston, TX. Pittsburgh (PA): Gulf Life Sciences Center. [cited in U.S. EPA 2005].

Gossiaux DC, Landrum PF, Fisher SW. 1996. Effect of temperature on the accumulation kinetics of PAHs and PCBs in the zebra mussel, Dreissena polymorpha. J Great Lakes Res 22(2):379–388.

Gustafson JB, Tell JG, Orem D. 1997. Selection of representative TPH fractions based on fate and transport considerations. Vol. 3. Total Petroleum Hydrocarbon Criteria Working Group Series, Association for Environmental Health and Sciences Foundation. 109 p. Available from: www.aehsfoundation.org/publications.aspx

Hardy R, Mackie P, Whittle K, McIntyre A. 1974. Discrimination in the assimilation of n-alkanes in fish. Nature 252:577–578.

Health Canada. 1998. Exposure factors for assessing total daily intake of priority substances by the general population of Canada. Ottawa (ON): Health Canada, Environmental Health Directorate. Unpublished report.

Hebert GW, Kussat RH. 1972. A laboratory evaluation of the toxicity of certain oils and chemical oil dispersants to juvenile coho salmon and staghorn sculpins. Canada, Department of the Environment Fisheries Service. Technical Report 1972-6.

Hoberman AM, Christian MS, Lovre S, Roth R, Koschier F. 1995. Developmental toxicity study of clarified slurry oil (CSO) in the rat. Fundam Appl Toxicol 28:34–40.

Hollister TA, Ward GS, Parrish PR. 1980. Acute toxicity of a #6 fuel oil to marine organisms. Bull Environ Contam Toxicol 24:656–661.

Hopkinson R. 2008. Priority substances under Environment Canada’s Chemicals Management Plan for the petroleum sector. Richmond (BC): Levelton Consultants Ltd.

[HYDROWIN]. Aqueous Hydrolysis Fate Program for Windows [Estimation Model]. 2008. Version 2.00. Washington (DC): U.S. Environmental Protection Agency, Office of Pollution Prevention and Toxics; Syracuse (NY): Syracuse Research Corporation. Available from: www.epa.gov/oppt/exposure/pubs/episuite.htm

[IARC] IARC Working Group on the Evaluation of Carcinogenic Risks to Humans. 1984. Mineral oils. IARC Monogr Eval Carcinog Risks Hum 33:87–168.

[IARC] IARC Working Group on the Evaluation of Carcinogenic Risks to Humans. 1989a. Fuel oils. IARC Monogr Eval Carcinog Risks Hum 45:239–270.

[IARC] IARC Working Group on the Evaluation of Carcinogenic Risks to Humans. 1989b. Occupational exposures in petroleum refining. IARC Monogr Eval Carcinog Risks Hum 45:39–117.

Jimenez BD, Cirmo CP, McCarthy JF. 1987. Effects of feeding and temperature on uptake, elimination and metabolism of benzo(a)pyrene in the bluegill sunfish (Lepomis macrochirus). Aquat Toxicol 10:41–57.

[JNITE]. 2010. Japanese National Institute of Technology and Evaluation. 2010. Official Bulletin of Economy, Trade and Industry [database on the Internet]. [accessed 2010 Sept]. Available from: http://www.nite.go.jp/index-e.html

Jonsson G, Bechmann RK, Bamber SD, Baussant T. 2004. Bioconcentration, biotransformation, and elimination of polycyclic aromatic hydrocarbons in Sheepshead minnows (Cyprinodon variegates) exposed to contaminated seawater. Environ Toxicol Chem23: 1538–1548. [Cited in Golder Associates Ltd. 2006]

Khan SH, Goode JW. 1984. Micronucleus test: aromatic pyrolysis oil orally for 2 days. Project No. 2082. Prepared for Gulf Oil Chemicals Co., Houston, TX. Pittsburgh (PA): Gulf Life Sciences Center. [cited in U.S. EPA 2005].

Koyama J, Uno Seiichi, Kohno K. 2004. Polycyclic aromatic hydrocarbon contamination and recovery characteristics in some organisms after the Nakhodka oil spill. Mar Poll Bull 49:1054–1061.

Lampi M, Parkerton T. 2009. Bioaccumulation assessment of PAHs. Prepared for CONCAWE. Available from:http://echa.europa.eu/doc/about/organisation/msc/msc_rcoms2009/rcom_af/Concawe_Bioaccumulation_Assessment_PAHs.pdf

Lampi M, Paumen ML, Parkerton T. 2010. An evaluation of the persistence, bioaccumulation and toxicity of petroleum hydrocarbons. Report prepared for CONCAWE, Brussels, Belgium. Annandale (NJ): ExxonMobil Biomedical Sciences, Inc.; Machelen (BE): ExxonMobil Petroleum and Chemical.

Lewis SC. 1983. Crude petroleum and selected fractions. Skin cancer bioassays. Prog Exp Tumor Res 26:68–84. [cited in IARC 1989b].

Lyman W, Reehl W, Rosenblatt D, editors. 1990. Handbook of chemical property estimation methods. Environmental behaviour of organic compounds. Washington (DC): American Chemical Society.

MacLean MM, Doe KG. 1989. The comparative toxicity of crude and refined oils to Daphnia magna and Artemia. Manuscript Report EE-111. Ottawa (ON): Environment Canada. 72 p. [cited in Environment Canada 2010a].

McCain BB, Hodgins HO, Gronlund WD, Hawkes JW, Brown DW, Myers MS. 1978. Bioavailability of crude oil from experimentally oiled sediment to English sole Parophrys vetulus. J Fish Res Board Can 35(5):657–664.

McCarthy JF, Jimenez BD, Barbee T. 1985. Effect of dissolved humic material on accumulation of polycyclic aromatic hydrocarbons: Structure-activity relationships. Aquat Toxicol 7:145–156.

McKee RH, Nicolich MJ, Scala RA, Lewis SC. 1990. Estimation of epidermal carcinogenic potency. Fundam Appl Toxicol 15:320–328.

McLeese DW, Burridge LE. 1987. Comparative accumulation of PAHs in four marine invertebrates. In Capuzzo IM, Kester DR, editors. Oceanic processes in marine pollution. Malabar (FL): Kirger RE. p. 109–118 [cited in European Commission 2008b].

Michigan. 2010. Michigan Department of Natural Resources and Environment. Oil intoxication. [accessed 2010 Feb 17]. Available from: www.michigan.gov/dnr/0,1607,7-153-10370_12150_12220-27243--,00.html

[Mobil] Mobil Oil Corporation. 1985. A modified Ames pre-incubation mutagenesis assay for determination of specific mutagenicity of the DMSO extract of heavy vacuum gas oil. Study No. 52261. Princeton (NJ): Mobil Oil Corporation, Environmental and Health Science Laboratory. [cited in API 2004].

[Mobil] Mobil Oil Corporation. 1987a. Micronucleus assay of bone marrow red blood cells from rats treated via dermal administration of heavy vacuum gas oil. Study No. 61591. Princeton (NJ): Mobil Oil Corporation, Environmental and Health Science Laboratory. [cited in API 2004].

[Mobil] Mobil Oil Corporation. 1987b. Metaphase analysis of Chinese hamster ovary (CHO) cells treated in vitro with a DMSO extract of heavy vacuum gas oil. Study No. 52262. Princeton (NJ): Mobil Oil Corporation, Environmental and Health Science Laboratory. [cited in API 2004].

[Mobil] Mobil Oil Corporation. 1987c. Clarified slurry oil: developmental toxicity study in rats. Study No. 50541. Princeton (NJ): Mobil Oil Corporation, Environmental and Health Science Laboratory. [cited in ECB 2000b].

[Mobil] Mobil Oil Corporation. 1987d. A static 96-hour acute toxicity study of process oil to bluegill sunfish. Pennington (NJ): Mobil Environmental and Health Science Laboratory. [cited in API 2004].

[Mobil] Mobil Oil Corporation. 1987e. A static 48-hour acute toxicity study of process oil to Daphnia magna. Pennington (NJ): Mobil Environmental and Health Science Laboratory. [cited in API 2004].

[Mobil] Mobil Oil Corporation. 1987f. A static 96-hour acute toxicity study of process oil to Selenastrum capricornutum. Pennington (NJ): Mobil Environmental and Health Science Laboratory. [cited in API 2004].

[Mobil] Mobil Oil Corporation. 1988. Thirteen-week dermal administration of syntower bottoms to rats. Study No. 62710. Princeton (NJ): Mobil Oil Corporation, Environmental and Health Science Laboratory. [cited in ECB 2000b].

[Mobil] Mobil Oil Corporation. 1990. Developmental toxicity study in rats exposed dermally to Ferndale syntower bottoms. Study No. 62934. Princeton (NJ): Mobil Oil Corporation, Environmental and Health Science Laboratory. [cited in ECB 2000b].

[Mobil] Mobil Oil Corporation. 1991. Developmental toxicity study in rats exposed dermally to heavy atmospheric gas oil. Study No. 64146. Princeton (NJ): Mobil Oil Corporation, Environmental and Health Science Laboratory. [cited in API 2004].

[Mobil] Mobil Oil Corporation. 1992. Thirteen-week dermal administration of heavy atmospheric gas oil to rats. Study No. 63456. Princeton (NJ): Mobil Oil Corporation, Environmental and Health Science Laboratory. [cited in API 2004].

[Mobil] Mobil Oil Corporation. 1994a. Developmental toxicity study in rats exposed dermally to heavy coker gas oil (HCGO). Study No. 64168. Princeton (NJ): Mobil Oil Corporation, Environmental and Health Science Laboratory. [cited in ECB 2000b].

[Mobil] Mobil Oil Corporation. 1994b. Thirteen-week dermal administration of Joliet heavy coker gas oil to rats. Study No. 64165. Princeton (NJ): Mobil Oil Corporation, Environmental and Health Science Laboratory. [cited in ECB 2000b].

[MSDS] Material Safety Data Sheet. 2007. HAGO [heavy atmospheric gas oil]. Calgary (AB): NOVA Chemicals. Available from: www.novachem.com/appl/prodfinder/docs/chemical/HeavyAtmosphericGasOil_MSDS_EN.pdf

Muijs B, Jonker M. 2010. A closer look at bioaccumulation of petroleum hydrocarbon mixtures in aquatic worms. Environmental Toxicology and Chemistry, V29, 9, pp 1943 – 1949.

[NCI] National Chemical Inventories [database on a CD-ROM]. 2006. Columbus (OH): American Chemical Society, Chemical Abstracts Service. Available from: www.cas.org/products/cd/nci/require.html

[NEB] National Energy Board. 2008. Focus on safety and environment: a comparative analysis of pipeline performance 2000–2006. Available from: www.neb.gc.ca/clf-nsi/rsftyndthnvrnmnt/sfty/sftyprfrmncndctr/fcsnsfty/2008/fctsht0825-eng.html

Neff JM, Anderson JW. 1981. Response of marine animals to petroleum and specific petroleum hydrocarbons. London (UK): Applied Science Publishers Ltd. [cited in CONCAWE 1998].

Neff JM, Cox BA, Dixit D, Anderson JW. 1976. Accumulation and release of petroleum-derived aromatic hydrocarbons by four species of marine animals. Mar Biol 38:279–289.

Newsted JL, Giesy JP. 1987. Predictive models for photoinduced acute toxicity of polycyclic aromatic hydrocarbons to Daphnia magna, Strauss (Cladocera, crustacean). Environ Toxicol Chem 6:445–461.

Nfon E, Cousins IT, Broman D. 2008. Biomagnification of organic pollutants in benthic and pelagic marine food chains from the Baltic Sea. Sci Total Environ 397:190–204. [cited in Lampi et al. 2010].

Niimi A, Palazzo V. 1986. Biological half-lives of eight polycyclic aromatic hydrocarbons (PAHs) in rainbow trout (Salmo gairdneri). Water Res 20(4):503–507.

[OECD] Organisation for Economic Co-operation and Development. 2009. Emission scenario document on transport and storage of chemicals. OECD series on emission scenario documents, No. 24. Available from: http://www.oecd.org/officialdocuments/displaydocumentpdf?cote=env/jm/mono(2009)26&doclanguage=en

Oliver B, Niimi A. 1988. Trophodynamic analysis of polychlorinated biphenyl congeners and other chlorinated hydrocarbons in the Lake Ontario ecosystem. Environ Sci Technol 22:388–397

Pancirov R, Brown R. 1975. Analytical methods for polynuclear aromatic hydrocarbons in crude oils, heating oils, and marine tissues.Proceedings of a Conference on Prevention and Control of Oil Pollution, San Francisco, California. Washington (DC): American Petroleum Institute. p. 103–113. [cited in Potter and Simmons 1998].

Papciak MS, Goode JW. 1984. CHO/HGPRT test using aromatic pyrolysis oil. Project No. 2081. Prepared for Gulf Oil Chemicals Co., Houston, TX. Pittsburgh (PA): Gulf Life Sciences Center. [cited in U.S. EPA 2005].

Peterson GI, Kristensen P. 1998. Bioaccumulation of lipophilic substances in fish early life stages. Environ Toxicol Chem 17:1385-1395.

[PETROTOX] Spreadsheet model to calculate toxicity of petroleum hydrocarbons to aquatic organisms. 2009. Version 3.04. Brussels (BE): CONservation of Clean Air and Water in Europe (CONCAWE). Available from: www.concawe.be/content/default.asp?PageID=778

Potter T, Simmons K. 1998. Composition of petroleum mixtures. Vol 2. Total Petroleum Hydrocarbon Criteria Working Group Series, Association for Environmental Health and Sciences Foundation. 109 p. Available from: www.aehsfoundation.org/publications.aspx

Rhodes SW, Risher JF. 1995. Toxicology profile for fuel oils. Atlanta (GA): U.S. Department of Health and Human Services, Agency for Toxic Substances and Disease Registry, Division of Toxicology/Toxicology Information Branch. 231 p.

[RMRI] Risk Management Research Institute. 2007. Quantitative assessment of oil spill risk for the south coast of Newfoundland and Labrador. St. John’s (NL): RMRI (Canada). Draft report for Transport Canada.

Rossi SS, Anderson JW, Ward GS. 1976. Toxicity of water-soluble fractions of four test oils for the polychaetous annelids, Neanthes arenaceodentata and Capitella capitata. Environ Pollut 10:9–18.

Saffiotti U, Shubik P. 1963. Studies on promoting action in skin carcinogenesis. Natl Cancer Inst Monogr 10:489–507. [cited in IARC 1984].

Salazar-Coria L, Amezxua-Allieri MA, Tenorio-Torres M, Gonzales-Marcias C. 2007. Polyaromatic hydrocarbons (PAHs) and metal evaluation after a diesel spill in Oaxaca, Mexico. Bull Environ Contam Toxicol 79:462–467.

Salem H, Katz SA, editors. 2006. Inhalation toxicology. 2^nd ed. Boca Raton (FL): CRC Press, Taylor & Francis Group.

[SCREEN3] Screening Tool Program for Windows [Screening Model]. 1996. Version 4.10. Research Triangle Park (NC): U.S. Environmental Protection Agency, Office of Air Quality Planning and Standards, Emissions, Monitoring, and Analysis Division. Available from: www.epa.gov/scram001/dispersion_screening.htm

[SENES] SENES Consultants Limited. 2009. Review of current and proposed regulatory and non-regulatory management tools pertaining to selected petroleum substances under the Chemicals Management Plan. Report to Health Canada. Richmond Hill (ON): SENES.

Shapiro DD, Getmanets IY. 1962. Blastomogenic properties of petroleum of different sources. Gig Sanit 27:38–42. [cited in IARC 1989b].

Shubik P, Saffiotti U. 1955. The carcinogenic and promoting action of low boiling catalytically cracked oils. Acta Unio Int Contra Cancrum 11:707–711. [cited in IARC 1984].

Simpson BJ. 2005. Analysis of petroleum hydrocarbon streams on the Health Canada CEPA/DSL Draft Maximal List. Report to the Canadian Petroleum Products Institute.

Smith WE, Sunderland DA, Sugiura K. 1951. Experimental analysis of the carcinogenic activity of certain petroleum products. Arch Ind Hyg Occup Med 4:299–314. [cited in IARC 1984].

[SPARC]. 2009. Sparc performs automated reasoning in chemistry. [Internet]. U.S. Environmental Protection Agency. Ecosystems Research Division. Available from: www.epa.gov/athens/research/projects/sparc/

Sprague JB, Carson WG. 1970. Toxicity tests with oil dispersants in connection with oil spill at Chedabucto Bay, Nova Scotia. Fisheries Research Board of Canada. Technical Report No. 201.

Stegeman JJ, Teal JM. 1973. Accumulation, release and rention of petroleum hydrocarbons by the oyster Crassostrea virginica. Mar Biol 22:37–44.

Strobel CJ, Brenowitz AH. 1981. Effects of bunker C oil on juvenile horseshoe crabs (Limulus polyphemus). Estuarine Res Fed 4(2):157–159.

Sun Petroleum Products Co. 1979. Letter from Sun Petroleum Products Company submitting additional information requested in USEPA letter dated February 6, 1979, with attachments. EPA/OTS No. 8EHQ-0379-0140. NTIS/OTS-0200438. [Abstract]. [cited in TOXLINE 2009].

Szaro RC. 1979. Bunker C fuel oil reduces mallard egg hatchability. Bull Environm Contam Toxicol. 22: 731-732.

Szaro RC, Albers PH, Coon N. 1978. Petroleum: effects on mallard egg hatchability. J Wildl Manag 42(2):404–406.

Tagatz ME. 1961. Reduced oxygen tolerance and toxicity of petroleum products to juvenile American shad. Chesapeake Sci 2:65–71.

Takeuchi I, Miyoshi N, Mizukawa K, Takada H, Ikemoto T, Omori K, Tsuchiya K. 2009. Biomagnification profiles of polycyclic aromatic hydrocarbons, alkylphenols and polychlorinated biphenyls in Tokyo Bay elucidated by δ13C and δ15N isotope ratios as guides to trophic web structure. Mar Pollut Bull 58:663–671.

Tatem HE, Cox BA, Anderson JW.1978. The toxicity of oils and petroleum hydrocarbons to estuarinecrustaceans. Estuarine Coastal Mar Sci 6:365–373.

Tolls J, van Dijk J. 2002. Bioconcentration of n-dodecane and its highly branched isomer 2,2,4,6,6-pentamethylheptane in fathead minnows. Chemosphere 47:1049–1057.

[TOPKAT] Toxicity Prediction by Komputer Assisted Technology [Internet]. 2004. Version 6.1. San Diego (CA): Accelrys Software Inc. Available from: http://accelrys.com/mini/toxicology/predictive-functionality.html

[TOXLINE] Toxicology Literature Online [database on the Internet]. 1974– . Bethesda (MD): National Library of Medicine (US). [revised 2009 Apr 18; cited 2009 Apr 21]. Available from: http://toxnet.nlm.nih.gov/cgi-bin/sis/htmlgen?toxadv.htm

Transport Canada. 2010. National Aerial Surveillance Program. Synopsis available from: http://www.tc.gc.ca/eng/marinesafety/oep-ers-nasp-2195.htm

[UBTL] Utah Biomedical Testing Laboratory Inc. 1990. 28 day dermal toxicity study in rats. Salt Lake City (UT): UBTL. Report No.: ATX-90-0066. [cited in API 2004].

[UBTL] Utah Biomedical Testing Laboratory Inc. 1994. A developmental toxicity screen in female rats administered F-228 dermally during gestation days 0 to 20. Study No. 66479. Salt Lake City (UT): UBTL. Report No.: ATX-91-0267. [cited in API 2004].

[U.S. EPA] United States Environmental Protection Agency. 1992. Screening procedures for estimating the air quality impact of stationary sources. Revised. Research Triangle Park (NC): U.S. EPA, Office of Air and Radiation, Office of Air Quality Planning and Standards. Report No.: EPA-454/R-92-019. Available from: http://cfpub.epa.gov/ols/catalog/catalog_display.cfm?&FIELD1=SUBJECT&INPUT1=Wind%20velocity&TYPE1=EXACT&item_count=27

[U.S. EPA] United States Environmental Protection Agency. 1995. Great Lakes Water Quality Initiative Technical Support Document for the Procedure to Determine Bioaccumulation Factors, Report No. EPA-820-B-95-005.

[U.S. EPA] United States Environmental Protection Agency. 2005. Appendix II. Robust summaries of studies used to characterize the fuel oils category. Washington (DC): U.S. Environmental Protection Agency, High Production Volume Chemical Program. Available from: www.epa.gov/hpv/pubs/summaries/fueloils/c13435rr3.pdf

[U.S. EPA] United States Environmental Protection Agency. 2008a. Transportation and marketing of petroleum liquids. In: AP 42, Compilation of air pollutant emission factors, vol 1. 5^th ed. Washington (DC): U.S. EPA. Available from: www.epa.gov/ttn/chief/ap42/ch05/final/c05s02.pdf

[U.S. EPA] United States Environmental Protection Agency. 2008b. Biota-Sediment Accumulation Factor Data Set. [cited in Lampi et al. 2010]. Available from: www.epa.gov/med/Prods_Pubs/bsaf.zip

Vandermeulen HH, Lee RW. 1986. Lack of mutagenic activity of crude and refined oils in the unicellular alga Chlamydomonas reinhardtii. Bull Environ Contam Toxicol 36:250–253. [cited in IARC 1989a].

Vandermeulen HH, Foda A, Stuttard C. 1985. Toxicity vs mutagenicity of some crude oils, distillates, and their water soluble fractions. Water Res 19:1283–1289. [cited in IARC 1989a].

Varanasi U, Stein J, Nishimoto M. 1989. Biotransformation and disposition of PAH in fish. In Varanasi U, ed, Metabolism of Polycyclic Aromatic Hydrocarbons in the Aquatic Environment, Boca Raton (FL): CRC Press. pp 93–150. [cited in Burkhard and Lukasewycz 2000].

Wan Y, Jin X, Hu J, Jin F. 2007. Trophic dilution of polycyclic aromatic hydrocarbons (PAHs) in a marine food web from Bohai Bay, North China. Environ Sci Technol 41:3109–3114.

Weil CS, Condra NI. 1977. Experimental carcinogenesis of pyrolysis fuel oil. Am Ind Hyg Assoc J 38:730–733. [cited in IARC 1989b].

Weinstein JE, Oris JT. 1999. Humic acids reduce the bioaccumulation and photoinduced toxicity of fluoranthene to fish. Environ Toxicol Chem 18(9):2087–2094.

White DH, Kirke KA, Coon NC. 1979. Effect of No. 2 fuel oil on hatchability of marine and estuarine bird eggs. Bull Environ Contam Toxicol 21:7–10.

Yakata N, Sudo Y, Tadokoro H. 2006. Influence of dispersants on bioconcentration factors of seven organic compounds with different lipophilicities and structures Chemosphere 64:1885–1891. [cited in Lampi et al. 2010].

Yvette C, Van Tongeren M, Urbanus J, Cherrie JW. 2011. An assessment of dermal exposure to heavy fuel oil (HFO) in occupational settings. Ann Occup Hyg 55(3):319–328.

Zhou S, Heras H, Ackman RG. 1997. Role of adipocytes in the muscle tissue of Atlantic salmon (Salmo salar) in the uptake, release and retention of water-soluble fraction of crude oil hydrocarbons. Mar Biol 127:545–553.

Appendices

• Appendix 1: Petroleum Substance Groupings
• Appendix 2: Physical and Chemical Data Tables for Industry-restricted HFOs
• Appendix 3: Measures Designed to Prevent, Minimize or Manage Unintentional Releases
• Appendix 4: Release Estimation of Industry-restricted HFOs During Transportation
• Appendix 5: Modelling Results for Environmental Properties of Industry-restricted HFOs
• Appendix 6: Modelling Results for Human Exposure to Industry-restricted HFOs
• Appendix 7: Summary of health effects information from pooled health effects data for HFO substances

Footnotes

Date modified:

2013-10-01