bioremediation of oil field waste containing elevated concentrations of weathered...

PETROLEUM TECHNOLOGY ALLIANCE CANADA BIOREMEDIATION OF OIL FIELD WASTE CONTAINING ELEVATED CONCENTRATIONS OF WEATHERED PETROLEUM HYDROCARBONS LITERATURE REVIEW

Bioremediation of Oil Field Waste Containing Elevated Concentrations of Weathered Petroleum Hydrocarbons Literature Review

EPR03029

26 January 2007

Environment & Water Resources

Suite 705, 10240 – 124 Street Edmonton, AB T5N 3W6 Canada Telephone: +1 780 496 9055 Toll Free: 1 877 566 3933 Facsimile: +1 780 496 9575 worleyparsons.com © Copyright 2006 WorleyParsons Komex


PROJECT EPR03029 - BIOREMEDIATION OF OIL FIELD WASTE CONTAINING ELEVATED CONCENTRATIONS OF WEATHERED PETROLEUM HYDROCARBONS FILE LOC.: EDMONTON

REV DESCRIPTION ORIG REVIEW WORLEY- PARSONS APPROVAL

DATE CLIENT APPROVAL

DATE

A B.Bennett

S.Murphy

S.Murphy

20-Dec-06

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Disclaimer

The information presented in this document was compiled and interpreted exclusively for the purposes stated in Section 1.2 of the document. WorleyParsons Komex provided this report for Petroleum Technology Alliance Canada solely for the purpose noted above.

WorleyParsons Komex has exercised reasonable skill, care, and diligence to assess the information acquired during the preparation of this report, but makes no guarantees or warranties as to the accuracy or completeness of this information. The information contained in this report is based upon, and limited by, the circumstances and conditions acknowledged herein, and upon information available at the time of its preparation. The information provided by others is believed to be accurate but cannot be guaranteed.

WorleyParsons Komex does not accept any responsibility for the use of this report for any purpose other than that stated in Section 1.2 and does not accept responsibility to any third party for the use in whole or in part of the contents of this report. Any alternative use, including that by a third party, or any reliance on, or decisions based on this document, is the responsibility of the alternative user or third party.

Any questions concerning the information or its interpretation should be directed to Brooke Bennett.

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CONTENTS 1. INTRODUCTION..............................................................................................................1

1.1 Background..........................................................................................................1

1.2 Scope of Review..................................................................................................1

2. PETROLEUM HYDROCARBONS ....................................................................................2

2.1 Petroleum ............................................................................................................2

2.2 Hydrocarbons ......................................................................................................3

2.3 Whole Crude Oil and Petroleum Fraction Definitions ............................................4

2.3.1 Petrogenic, Biogenic, and Pyrogenic Hydrocarbons ...........................................4

2.3.2 Fraction Defined by the Petroleum Industry ........................................................5

2.3.3 Canada-Wide Standards for Petroleum Hydrocarbons in Soil .............................7

3. CHEMICAL BIOAVAILABILITY.......................................................................................12

4. PETROLEUM HYDROCARBONS IN SOIL.....................................................................15

4.1 Biodegradation...................................................................................................16

4.2 Bioremediation...................................................................................................19

4.2.1 Compost Use in Bioremediation Programs .......................................................20

5. PETROLEUM HYDROCARBON TOXICITY....................................................................24

5.1.1 Toxicity Testing................................................................................................25

6. ORGANIC MATTER AND PETROLEUM HYDROCARBON ANALYSES.........................27

7. SUMMARY AND CONCLUSIONS..................................................................................31

8. CLOSURE......................................................................................................................32

9. REFERENCES...............................................................................................................33

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Tables within Text

TABLE A ELEMENTAL COMPOSITION OF NOMINAL PETROLEUM............................... 3

TABLE B EQUIVALENT CARBON NUMBER RANGES OF THE CCME CANADA-WIDE STANDARD PETROLEUM HYDROCARBON FRACTIONS............................... 8

TABLE C COMPOUNDS PRESENT IN BOTH CRUDE OIL AND NATURAL ORGANIC MATTER (AFTER JORGENSON IN WHITE AND IRVINE, 1996) ..................... 28

Figures within Text

FIGURE 1 HYDROCARBON STRUCTURAL RELATIONSHIPS.......................................... 2

FIGURE 2 BENZENE CHEMICAL STRUCTURE ................................................................ 4

FIGURE 3 SARA PETROLEUM HYDROCARBON FRACTIONS (JEWELL ET AL., 1974)... 7

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ACRONYMS

API American Petroleum Institute

BTEX Benzene, Toluene, Ethylbenzene, Total xylenes

C Carbon

CCME Canadian Council of Ministers of the Environment

CWS Canadian-Wide Standards

EC-L Effects concentration low

ECN Equivalent Carbon Number

ECOTAG Ecological Task Advisory Group

ECx Effects-endpoints

ERAC Environmental Research Advisory Council

F3 Fraction 3 of the Canadian Council of Ministers of the Environment regulated petroleum hydrocarbons (hydrocarbon compounds greater than C16 to C34).

GC/FID Gas chromatography with a flame ionizing detector

GC/MS Gas chromatography mass spectrometry

H Hydrogen

LCx Lethal concentration

MOG Mineral oil and grease

NOM Natural organic matter

NOx Nitrogen oxides

NRC National Research Council

O&G Oil and grease

PAH Polycyclic aromatic hydrocarbon

PHC Petroleum hydrocarbon

SARA Saturates-aromatics-resins-asphaltenes

SOM Soil organic matter

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SOx Sulphur oxides

SQG Soil quality guideline

TEC Threshold effects concentration

TPH Total petroleum hydrocarbons

TPHCWG Total Petroleum Hydrocarbon Criteria Working Group

U.S.EPA United States Environmental Protection Agency

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1. INTRODUCTION

1.1 Background

Weathered or partially degraded petroleum hydrocarbons (PHCs) are a source of contamination at thousands of inactive oil and gas sites across Alberta. Generally, the portion of PHCs in soils that remain decades after the initial contamination has occurred consist of as the Canadian Council of Ministers of the Environment (CCME) fraction (F)3 and F(4). F3 PHCs are generally high molecular weight hydrocarbons (compounds with C>16 to C34). Bioremediation techniques are generally successful at removing freshly added petroleum to soils (e.g., from spills); however, the chemical and physical characteristics of weathered PHCs do not make weathered PHCs amenable to traditional bioremediation techniques. Soils contaminated with concentrations of PHC F3 exceeding the applicable guidelines are typically sent to landfills for disposal during site remediation activities. One bioremediation technique that may potentially reduce PHC F3 concentrations in soils is to add organic amendments, such as mature compost and bark mulch, to contaminated soils.

In the last decade, a number of studies have examined microbial degradation of PHCs in soils by the addition of mature compost with varying degrees of success (Al-Daher et al. 2001; Graham, 2005; Heidman, 2005; Kastner et al. 1995; Mahro et al. 1994; Stegmann et al. 1991; Vasudevan and Rajaram, 2001). Few studies have examined the biodegradation of weathered PHCs with the addition of mature compost and bark mulch as bulking agents (Abiola and Olenyk, 1998; Chaw and Stoklas, 2001). Bioremediation of weathered PHCs using mature compost and/or bark mulch could potentially reduce site remediation costs and reduce the volume of contaminated soil sent to landfills for disposal during remediation activities.

1.2 Scope of Review

The purpose of this report is to present a literature review on the topics related to PHCs in soils and the bioremediation of PHCs in soils, including the addition of compost and bark mulch as organic amendments in bioremediation trials. This literature review was prepared to meet the reporting requirements for funds granted from the Environmental Research Advisory Council (ERAC) through the Petroleum Technology Alliance Canada to an associated bioremediation research project (WorleyParsons Komex, 2006).

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2. PETROLEUM HYDROCARBONS

2.1 Petroleum

Petroleum is composed of complex mixtures of hydrocarbons along with other minor constituents, such as nitrogen, sulphur, and heavy metals. In some parts of the oil and gas industry, petroleum refers to both crude oil and natural gas. In the refining industry, petroleum refers only to crude oil (Van Dyke, 1997), which is the definition that will be used throughout this review.

Crude oil is found in naturally occurring geologic deposits formed from organic decomposition products of ancient plants and animals. To form crude oil deposits, plant and animal residues were gradually buried in sediments over geologic time, and subsequently subjected to heat and pressure as the sediments were buried. The residues were slowly reduced under anaerobic conditions to relatively pure hydrocarbons. Crude oil is principally composed of liquid hydrocarbons having four or more carbon atoms (Manning and Thompson, 1991; Van Dyke, 1997). Alkanes, cycloalkanes, and aromatics are the three principle classes of hydrocarbons present in petroleum (Strausz and Lown, 2003; Figure 1). Alkenes and alkynes are found only in trace quantities in petroleum (Potter and Simmons, 1998; Van Dyke, 1997), and will not be discussed further.

HYDROCARBONS

Aliphatics Aromatics

Alkynes Alkanes Alkenes Polycyclic Monoaromatics

aromatic

hydrocarbons Cycloalkanes

Figure 1 Hydrocarbon structural relationships

Since the original buried organisms contained other compounds in their structures in addition to carbon and hydrogen, petroleum deposits also contain lesser amounts of heteroatom compounds and trace



concentrations of metals (Cole, 1994). Heteroatom compounds contain nitrogen, oxygen, sulphur, or metals bound within their molecular structure. The range of elemental composition of petroleum is shown in Table A (Speight, 1998).

Table A Elemental composition of nominal petroleum

Element Range weight (%)

Carbon 83 to 87

Hydrogen 10 to 14

Nitrogen 0.10 to 2.0

Oxygen 0.050 to 1.5

Sulphur 0.050 to 6.0

Nitrogen oxides (NOx) and sulphur oxides (SOx) will form during fuel combustion if concentrations of heteroatoms are not reduced during oil refining and fuel production. The most common metal contaminants in heavy gas oils and residues are nickel and vanadium (Beuther and Flinn, 1963). Catalysts are used in the refining process to remove metals from oils; however, metals may form deposits on the catalysts. Such deposits result in considerable loss of selectivity so that reactions tend to produce more coke (ash) and gas instead of the desired products, gasoline, diesel, etc.

2.2 Hydrocarbons

Hydrocarbons are organic compounds composed of carbon (C) atoms linked to hydrogen (H) atoms through high-energy bonds that can be broken during combustion. Hydrocarbons are divided into aromatic and aliphatic compounds. Aromatic compounds are made up of one or more cyclic hydrocarbon structure(s) with double bonds linking the carbon atoms (Speight, 1998). Aliphatic compounds are organic compounds in which carbon atoms are joined together in straight or branched chains.

The simplest aromatic compound is benzene (C6H6; Figure 2). Aromatics are subdivided into two main classes: monoaromatics and polycyclic aromatic hydrocarbons (PAHs; Figure 1). Monoaromatic compounds have one benzene ring while PAHs have two or more fused benzene rings (Cole, 1994; Van Dyke, 1997).



Figure 2 Benzene chemical structure

Aliphatics include all other hydrocarbons that are not aromatics. Aliphatics are subdivided into three main classes: alkanes (paraffins), alkenes (olefins), and alkynes. The compounds in each class contain different numbers of carbon atoms, but share some common structural features. Alkanes are compounds with single carbon-carbon bonds, alkenes have double bonds, and alkynes have triple bonds. Cycloalkanes (naphthenes) are a group of alkanes in which carbon atoms form cyclic structures with single carbon-carbon bonds (Strausz and Lown, 2003). Figure 1 illustrates hydrocarbon structural relationships (after Potter and Simmons, 1998).

2.3 Whole Crude Oil and Petroleum Fraction Definitions

Whole crude oil and petroleum fractions are described and defined using dozens of terms and definitions, depending on the type of business sector using the terms. In Canada, different operational definitions are used to describe petroleum fractions in the petroleum industry and environmental sectors. Generally, the industrial sector separates PHC fractions based either on boiling point or polarity, while environmental regulations define them based on compound carbon numbers. Some of the historical and operational definitions for both sectors are outlined below. This report will focus on the whole oil definitions in Section 2.3.1 and the PHC fractions as defined by the CCME (Section 2.3.3).

2.3.1 Petrogenic, Biogenic, and Pyrogenic Hydrocarbons

PHC “fingerprinting” is commonly done to characterize spilled oils and to link spilled oils to known sources of oil in order to select appropriate spill responses, allocate responsibility, and to take effective cleanup measures (Wang and Fingas, 2003). To assist in petroleum hydrocarbon fingerprinting, PHCs from a spill may be divided into the following three groups based on the origin of the oil or the transformations the oil has undergone in soil or marine environments:

• petrogenic;

• biogenic; and



• pyrogenic.

Petrogenic hydrocarbons are generated in organic rich source rocks exposed to elevated temperature for long time periods. This category includes crude oil and its refined products. Biogenic hydrocarbons are generated by biologic processes or in the early stages of diagenesis in marine sediments. In this review, the term biogenic hydrocarbons will also be applied to hydrocarbon compounds in compost. Pyrogenic hydrocarbons are generated in combustion processes, such as in engines or flare pits, and are the direct result of anthropogenic activities (Page et al., 1995).

Certain chemical characteristics can be used to distinguish petrogenic, biogenic, and pyrogenic hydrocarbons. Petrogenic hydrocarbons characteristically have families of related PAHs, where the unsubstituted parent PAH for each family is less abundant than its alkylated homologues (i.e., the same compound but with one or more methyl or ethyl side groups). Petroleum with an unsubstituted parent PAH that is less abundant than its alkylated homologues is a characteristic used to distinguish petrogenic hydrocarbons from pyrogenic hydrocarbons. In pyrogenic hydrocarbons, the unsubstituted parent PAH is more abundant than its alkylated homologues. Biogenic hydrocarbons in subtidal sediments are distinguished by the presence of perylene. Perylene is an unsubtitued PAH produced in subtidal sediments during early diagenesis (Page et al., 1995, 1996). The presence of perylene in decomposed soil organic matter (SOM) has not been reported; therefore the use of perylene as a marker compound (biomarker) of biogenic hydrocarbons in SOM is unknown.

2.3.2 Fraction Defined by the Petroleum Industry

In the petroleum industry, crude oil is separated (refined) using various techniques depending on the source of the petroleum. Extracted crude oil is of limited use; however, refining the oil produces fuels and thousands of petrochemical products. During separation, crude oils are classified according to grade: conventional light oil; conventional medium oil; heavy oil; extra heavy oil; and bitumen. Conventional crude oil is petroleum found in liquid form, flows naturally, or is capable of being pumped without further processing or dilution. Heavy oil, extra heavy oil, and bitumen are considered unconventional oils because they cannot be produced, transported, or refined by conventional methods. Crude oils may also be classified based on their specific gravity, American Petroleum Institute (API) gravity, and viscosity (Strausz and Lown, 2003).

Residual Petroleum Hydrocarbons from Petroleum Refining

A residual fraction is the non-distillable fraction of petroleum that remains following petroleum refining. This residual fraction is a useful source of petroleum fuel if further processed, and is used as feedstocks for further processing. During petroleum refining, residual oils are commonly grouped together based on the general characteristics of the compounds (e.g., polars or asphaltenes) rather than the identification of specific compound types (e.g., diaromatics or cycloalkanes) due to the



complexity of the mixtures. Petroleum technologists have commonly defined four classes of components in residual oils: asphaltenes; maltenes; resins; and oils (Jewell et al., 1974).

Heavy oils are high boiling point materials that are included in the residual fraction and are the less mobile (high viscosity) constituents of petroleum. Heavy oils are blended into several grades of heating oil and bunker fuel for ships (e.g., Bunker C fuel). Bitumen is even less mobile than heavy oils and is the soluble form of organic matter found in tar sands deposits. An asphalt is the product of a refinery operation where a residuum is treated to make a product that meets specifications for a variety of road/highway construction and other uses (Speight, 1998).

The SARA (saturates-aromatics-resins-asphaltenes) method is one method used historically to separate and define residual oils (compounds with boiling points greater than 260°C; Jewell et al., 1974). The SARA petroleum compound classes are still commonly used to define PHC fractions in the petroleum industry. The separation of the residual oils is done by elution with various solvents based on the polarity of the various fractions.

The following is a brief description of the SARA method (Jewell et al., 1974), and Figure 3 outlines the analytical separation steps (solvents) and products in the SARA method. First, the asphaltene fraction is separated from the heavy oil. Asphaltenes are the highest molecular-weight fraction of petroleum and are defined as the fraction that is insoluble in n-heptane. The maltene fraction is what remains once the asphaltenes have been removed, and is defined as the fraction of heavy oil that is soluble in n-heptane. The resins are separated from the maltene fraction by adsorption onto a clay (Attapulgus clay, or Fuller’s earth), or eluted from silica gel by polar elluents. When eluted from the silica gel column, the resin fraction is commonly called the polar fraction or polars. The saturates and aromatics are eluted from the adsorbent (Attapulgus clay, silica gel) with n-pentane or benzene. The saturates are then separated from the aromatics on an alumina column; the saturates are eluted with n-pentane and the aromatics with benzene (Speight, 1998; Strausz and Lown, 2003). Other standard analytical test methods separate PHCs into a combination of saturate, aromatic, aliphatic, or polar fractions.



Heavy oil

Figure 3 SARA petroleum hydrocarbon fractions (Jewell et al., 1974)

2.3.3 Canada-Wide Standards for Petroleum Hydrocarbons in Soil

The CCME developed the Canada-Wide Standards (CWS) for PHCs in soil for contaminated site assessment and remediation. The CCME CWS were adopted on May 1, 2001 by the Canadian federal and provincial environment ministers (CCME, 2001a). The CWS were developed to be protective of human and environmental health, are risk-based, and can be applied at any of three successive Tiers. Tier 1 applies national generic numerical values to surface and subsurface soils, Tier 2 values are adjustments to the Tier 1 values to accommodate unique site characteristics, and Tier 3 values are developed from site-specific human health or environmental risk assessments. Tier 1 levels were developed using risk assessment and risk management techniques (CCME, 2000).

Maltenes Asphaltenes

Resins/Polar Oils

Saturates

n-Heptane

Attapulgus clay

n-Pentane/Benzene

SiO2/Al2O3 column

Benzene n-Pentane

Aromatics



Canadian Council of Ministers of the Environment Petroleum Fractions

The CWS subdivides PHCs into fractions according to specified ranges of equivalent carbon numbers (ECN; CCME, 2000). Each fraction is composed of subfractions previously defined by the U.S. Total Petroleum Hydrocarbon Criteria Working Group (TPHCWG; Edwards et al., 1997; Gustafson et al., 1997). The subfractions that form the four CWS fractions (F1 to F4) are described according to relevant physical-chemical properties (e.g., solubility, Henry’s Law constant, etc.) and toxicological characteristics (reference doses/concentrations). Subdividing the fractions using physical-chemical and toxicological properties was done to allow the prediction of chemical fate, exposure, and potential risk of the hydrocarbon fractions. The CWS fractions separated based on ECNs are shown in Table B with the corresponding boiling point ranges (approximate) included for reference purposes.

Table B Equivalent carbon number ranges of the CCME Canada-Wide Standard petroleum hydrocarbon fractions

CCME CWS Fraction

Equivalent Carbon Number (ECN) Range

Boiling Point Ranges*

PHC F1 1nC6 to nC10 50°C to 173°C

PHC F2 2nC>10 to nC16 174°C to 288°C

PHC F3 3nC>16 to nC34 289°C to 481°C

PHC F4 nC>34 greater than 481°C 1 minus benzene, toluene, ethylbenzene, and xylenes (BTEX) 2 minus naphthalene 3 minus PAHs

* Boduszynski, 1987; ESG International Inc., 2003

BTEX and PAHs (including naphthalene) are assessed and managed separately during remediation (CCME, 2000; O’Connor Associates Environmental Inc. and Meridian Environmental Inc., 2001). The environmental fate of PAHs is of potential concern due to their toxic, mutagenic, and carcinogenic properties (Weissenfels et al., 1992). Even though PAHs are managed separately under CCME guidelines, PAHs have been studied thoroughly with respect to sorption, degradation, and bioavailability. References to these studies will be included in this review.



Guideline Development for the Canada-Wide Standards of Petroleum Hydrocarbons in Soils

The PHC CWS in soils were developed for four generic land uses (agricultural, residential/parkland, commercial, and industrial) with consideration of the proximity to surface or subsurface receptors. Generic Tier 1 values were derived through a systematic evaluation of all pathways of exposure of concern for human and ecological receptors. The main exposure pathways identified were: soil contact; soil ingestion; groundwater/surface water; vapour inhalation (humans only); produce, meat and milk produced on-site (humans only); and off-site migration of soil/dust. Not all pathways are relevant for every land use. The recommended human health soil quality level under Tier 1 is based on the single pathway that results in the greatest exposure. For ecological receptors, there is a large range of organisms that may potentially be exposed to PHCs. In Canada, as opposed to other countries or jurisdictions, emphasis is placed on ecological exposure pathways based on direct contact between plant roots or soil invertebrates with contaminated soil. This eco-soil contact pathway is intended to protect key ecological receptors that are indicative of soil quality (CCME, 2000).

The CCME protocol for the derivation of soil quality guidelines (SQGs) based on direct soil contact to invertebrates and plants (eco-soil contact pathway) is outlined in CCME (1996a). Agricultural and residential/parkland land use SQGs were developed by applying the threshold effects concentration (TEC) concept to toxicity data sets. TEC is the concentration of a chemical below which no adverse effect is expected to occur. TEC is calculated as the 25th percentile of the effects and no-effects data distribution (CCME, 2000). Commercial and industrial land use SQGs are developed by applying the effects concentration low (EC-L) concept to toxicity data sets. EC-L is a level of protection determined for commercial and industrial land use above a TEC. EC-L is calculated as the 25th percentile of effects data distribution (CCME, 2000).

Several modifications to the CCME (1996a) protocol were made when deriving the Tier 1 CWS for PHCs in soil based on the eco-soil contact pathway due to the availability of PHC toxicity data for terrestrial receptors (CCME, 2000). Two of these modifications were: (1) only effects-endpoints (ECx) or lethal concentrations (LCx) were used; and (2) toxicity endpoints response levels were standardized at or near the 50% response level for sublethal studies. The SQG for PHCs for agricultural and residential/parkland land use were developed using the effects data set for soil and plant invertebrates, similar to the EC-L protocol for commercial/industrial land uses outlined by CCME (1996a). The 50th percentile of the plant effects (not mortality) data was used to derive a soil quality benchmark for commercial and industrial land uses.

To derive the Tier 1 guidelines, data from PHC toxicity tests and studies for ecological receptors were sorted using a weight of evidence approach. The weight of evidence approach is defined as the critical evaluation and adoption of new numerical protocols where required to facilitate the incorporation of high quality but disparate types of information on the risks of PHCs to ecological receptors (CCME,



2000). The Tier 1 levels for ecological health protection were derived from chronic, sub-chronic, acute, and lethal responses of plant and invertebrate species relevant to the sustainable functioning of soil under the four land uses. Toxicity tests for plants and invertebrates were performed using coarse textured soils (greater than 50% by mass of particles greater than 75 µm mean diameter) with whole crude and individual PHC fractions from a fresh Alberta Federated Crude Oil (ESG International Inc. [ESG], 2003). Therefore, the results are expected to be most applicable to coarse-grained soils subjected to a fresh spill of PHC product (CCME, 2000). The Tier 1 soil contact standards were derived assuming 100% bioavailability of the product as determined by total soil concentration of each PHC fraction measured in soil (Murphy and Charrois, 2003; Stantec Consulting Ltd. [Stantec], 2004). The guideline values for fine-textured soils (greater than 50% by mass of particles less than 75 µm mean diameter) were extrapolated from the ecotoxicity results for exposures in coarse textured soils, based on assumptions about reduced bioavailability. Toxicity studies for plants and invertebrates was not performed on fine textured soils.

ESG (2003) provided the EC20 (likely similar to EC25 endpoints) and EC50 response levels of plants and soil invertebrates to fractionated federated crude oil (F2 to F4) using regression-based statistical techniques. EC50 is the concentration of a toxicant that can be expected to cause a defined non lethal effect in 50% of a given population of organisms under defined conditions. Revisions to the CCME protocols for derivation of SQGs (2006) propose using a standardized effects level of EC25 as opposed to EC50 (Axiom Environmental Inc. [Axiom]/UMA Engineering Ltd. [UMA], 2004). A 50% reduction in an ecologically relevant response factor (e.g., growth, reproduction) may be too high relative to policy-based environmental protection goals. Axiom/UMA (2004) made a comparison between reported EC20 and EC50, and found that on average the EC20 endpoints for chronic toxicity of whole crude to plants were 35% of the EC50 values. Based on the definitive end-points for plants, the F2 EC20 was on average 37% of the EC50 value, and for F3 was 26% of the EC50 value. Based on the definitive end-points for soil invertebrates, the F2 EC20 was on average 59% of the EC50 value, and for F3 was 45% of the EC50 value. From these comparisons, if standardization changes to EC20 from EC50, the benchmark values for the CWS for PHCs would be lowered significantly. More detailed Tier 3 ecological risk assessments are required to conclude if EC20 values are adequate or overly conservative to protect environmental receptors (Axiom/UMA, 2004).

Petroleum Hydrocarbon Fraction 3

CCME PHC F3 encompasses PHC constituents with an effective boiling point range from nC>16 to nC34. F3 includes unbranched saturated alkanes and PAHs, as well as an appreciable portion of alkyl-PAHs, such as alkylated naphthalene. Components of F3, in particular those in the high boiling point range, are not readily degraded (n-alkanes greater than C22 and PAHs greater than 3 rings), and tend to be more resistant to either natural microbial-mediated attenuation or bioremediation techniques (Section 3).



During the adoption of the CCME CWS for PHCs in soil, F3 concentrations were a major issue for site remediation activities and cleanup costs at the majority of upstream oil and gas sites contaminated with hydrocarbons (Axiom/UMA, 2004) due to the resistant nature of F3 components to biological degradation. Soil from historic oil and gas sites contaminated with F3 PHCs is commonly excavated and disposed in landfills during site remediation activities due to limitations with meeting Tier 1 endpoints. Weathered PHCs in soils have different chemical and physical characteristics compared to PHC sources freshly added to soils. Weathered PHCs are less amenable to desorption, volatilization, biodegradation, and extraction compared to PHCs freshly added to soils (Alexander, 1995; Bragg et al., 1994). A large portion of weathered PHCs are biologically unavailable and generally high molecular weight compounds, and thus may be very difficult to degrade. Loehr and Webster (1996) concluded that it is inaccurate to apply results from studies using freshly added chemicals to predict the behaviour of chemicals that have been in contact with soil for extended periods of time. Site-specific information about the degradation, desorption, or mobilisation of chemicals should be evaluated with field soils, not extrapolated from spiked soil data. Evidence from work done by Tindal (2005) and Visser and Reid (2005) shows that the CWS for weathered F3 in fine-textured soils are overly conservative and need to be re-examined, since the values for fine-textured soils were originally extrapolated from the results for coarse-textured soils (CCME, 2000).

During the final stages of development of the CWS, the Ecological Task Advisory Group (EcoTAG) recommended a strategy to develop SQGs from complex mixtures (CCME, 2000). EcoTAG also suggested that perhaps the range of compounds included in the F3 fraction was too broad, and there may be differences in the ecotoxicity of PHCs in the range of nC16 to nC21 (termed fraction “F3a”) and nC>21 to nC34 (termed fraction “F3b”). A soil ecotoxicity study was undertaken to separate the F3 fraction of a fresh Federated Crude Oil from Alberta into F3a and F3b (Axiom/UMA 2004). The study concluded that in most of the longer duration tests, Fraction 3b did not appear to be intrinsically less toxic than Fraction 3a. The results did not support further division of the F3 fraction. The authors concluded that there is a possibility that differences in the relative toxicity of the two subfractions would be evident in more weathered PHC mixtures.



3. CHEMICAL BIOAVAILABILITY

The concept of chemical bioavailability to humans and ecological receptors has recently gained interest in the waste and environmental industries when deciding waste cleanup goals and acceptable treatment endpoints. Bioavailability is the quantity of a chemical in soil or other environmental medium that is actually accessible to an organism for assimilation and generating a potential toxic effect (Alexander, 2000; Alexander and Alexander, 2000). The U.S. National Research Council (NRC) has found that the term bioavailability has been defined differently by different disciplines; therefore, the NRC defines bioavailability processes. Bioavailability processes are the individual physical, chemical, and biological interactions that determine the exposure of organisms to chemicals associated with soils and sediments (Ehlers and Luthy, 2003). As chemicals in soils undergo biological, chemical, and physical alterations, including sorption/desorption, diffusion, and partitioning to soil particles and organic matter (Kastner and Mahro, 1996; Sims et al., 1990; Weissenfels et al., 1992), the total concentration available to ecological receptors decreases, and thus, bioavailability over time decreases.

Contaminant weathering, aging, and sequestration are time-dependent processes that influence the bioavailability of chemicals, including PHCs, in soils (Alexander, 2000; Kelsey and Alexander, 1997). The bioavailability of higher molecular weight PHCs in soil, in particular compounds in the F3 and F4 fractions, decreases with time as hydrocarbons weather, age, and become sequestered in soil. Weathering, as applied to PHCs, is the change in chemical composition and bioavailability with time as related to natural processes including volatilization, differential mobility, biodegradation, and stabilization (CCME, 2000). Rates of PHC weathering are typically greatest in aerobic environments where microbial degradation is facilitated (Section 4.1). As PHCs weather in soils, the chemical composition of the petroleum shifts from a broad range of molecular weights to compounds with generally high molecular weights and more branched chains (i.e., PHCs in the F3 and F4 fraction remain). PHC molecules in the F3 and F4 fractions are generally large and complex and do not readily undergo microbial degradation.

Aging describes the processes of sorption and sequestration of chemicals in soils to form stable solid phases over time (Ehlers and Luthy, 2003). Non-extractable hydrocarbon residues may form as chemicals remain in contact with soil particles and organic matter during aging (Alexander, 1995; Oudot, 1984). Aging of organic chemical in soils is poorly understood; however, for non-ionic compounds, such as PHCs, it may involve diffusion within soil aggregates or into organic matter into spaces that are not readily accessible to microorganisms. The magnitude of reduction in bioavailability resulting from aging is different for a single compound in different soils, for different compounds in the same soil, and for different periods of time that a compound has remained in soil (Alexander, 2000). Studies with different soil types suggest that aging is more marked in soils with high levels of organic matter (Hatzinger and Alexander, 1995). Alexander (2000) compiled a list of compounds shown to



become less available for microbial degradation as a result of aging, including various PAHs (naphthalene, phenanthrene, anthracene, fluoranthene, and pyrene) in different soil types.

Sequestration is the time-dependent movement of contaminant molecules into remote and inaccessible areas of soil particles and/or organic matter (Stantec, 2004). The extent of exposure to chemicals in soil is likely to be considerably less if the compound has been sequestered than if it has been freshly applied to the soil particle surface (Huesemann, 1997). Weissenfels et al. (1992) found that sequestration was evident when phenanthrene, anthracene, fluorene, pyrene, chrysene, and other PAHs remaining in the soil from a closed coking plant were not susceptible to biodegradation, but were metabolized when extracted and added back to the soil. Nam and Alexander (2001) found that sequestration of compounds was negligible when biodegradation was rapid, but it was appreciable when microbial degradation was slow. The authors concluded that sequestration can occur even as a substrate is being biodegraded.

The amount and quality of SOM has been proposed as one of the most significant factors dominating hydrophobic contaminant interactions within soil (Hatzinger and Alexander, 1995). The affinity of SOM for non-polar organic compounds depends on its origin and geologic history. Organic matter in unweathered shales and high-grade coals has been shown to enhance sorption capacities more than an order of magnitude larger than organic matter in recent soils, geologically immature material, or highly weathered soil organic material (Luthy et al., 1997). Thus, the bioavailability of organic contaminants in soils is influenced by the presence of SOM and carbonaceous geosorbants, such as black carbon, coal, and kerogen.

Bioavailability can be measured using physicochemical techniques, such as microscopy, chemical extractions, or bioassays. A combination of these techniques may be used to provide detailed information on contaminant bioavailability. If chemical extraction techniques are used, caution should be exercised when interpreting the data, as current chemical extraction techniques measure total concentrations, not bioavailable concentrations, and the results may overestimate the risk posed by the contaminant (Alexander, 2000; Ehlers and Luthy, 2003). Current analytical methods used for extracting contaminants from soils and sediments involve vigorous methods with strong organic solvents, and do not account for the time-dependent change in contaminant availability. Mild extraction techniques using combinations of methanol, ethanol, butanol, and water have been developed for use as surrogate assays for bioavailability (Hatzinger and Alexander, 1995; Kelsey and Alexander, 1997). The use of mild solvents and less vigorous extraction methods compared to standard methods take into account changes in availability of contaminants with time (Alexander, 2000). Generally, the use of standard chemical extraction methods may be used as screening tools for contaminated soils and sediments, and then modified procedures can be used to measure contaminant bioavailability, such as modified chemical extraction methods developed to determine bioavailable arsenic in soils (Rodriguez et al., 2003; Szakova et al., 2005). Bioassays are another tool used to measure the bioavailability of a contaminant (Ehlers and Luthy, 2003). Bioassays are typically conducted to measure the effects of a



type of contamination on a living organism. Biological responses to contaminants and acute and sublethal toxicity can be measured using bioassays; however, bioassays are often time consuming and expensive (Alexander, 2000).

As contaminants become less available to ecological receptors in soils, the potential toxicity of the contaminant diminishes. The concept of bioavailability influencing the toxicity of compounds in soils was developed within the last decade by Alexander (1995). Alexander (1995) and Alexander and Alexander (2000) proposed that the exposure of an organism to a carcinogen or other toxic compounds in soils is not related to the total concentration of the substance in the soil, but is rather related to the amount that is actually available to receptors. Prior to studies on the influence of bioavailability on toxicity, hazard assessments of toxic chemicals in soils were made without concern that bioavailability of compounds may change over time. Alexander (1995) collected documented evidence from other studies of the changes in the bioavailability of toxic chemicals in soils over time. The author found that patterns of disappearance of persistent or residual chemicals in both field and laboratory studies showed a declining availability to microorganisms with residence time in soil. The decline in bioavailability diminished the toxicity of the compounds to test organisms.

If contaminants in soils and sediments are not bioavailable, then more contaminant mass may potentially be left in place without creating additional risk to the receptors. The use of bioavailability tools and concepts during site remediation has not received widespread regulatory and public support, because it is sometimes perceived as a “do nothing” approach. As of 2003, the NRC found no legal recognition of bioavailability for soil cleanup, although bioavailability concepts are emerging for sediment management and are used for biosolids management (Ehlers and Luthy, 2003).



4. PETROLEUM HYDROCARBONS IN SOIL

Hydrocarbons are a ubiquitous class of natural compounds, and are present in low concentrations in most soils and sediments (Rosenberg et al., 1992). At upstream oil and gas facilities, PHC contamination of surface and subsurface soils can result from accidental spills, equipment failure, and leaks in underground facilities, such as underground storage tanks, pipelines, or earthen pits (flare pits, sumps). At abandoned wellsites, the source of PHCs in soils is generally from spills at well centres, the disposal of drilling wastes in sumps, or flare pits that were used to store and/or burn produced fluids generated at the sites. In Alberta, pipeline corrosion and equipment failures were the leading causes of oil releases to soils in 2002 and 2003. The reported total volume of hydrocarbons released to soils in Alberta in 2003 was 5,268 m3, which was a slight increase from 5,188.8 m3, the volume released in 2002 (Alberta Energy and Utilities Board, 2004).

Crude oil is composed of volatile, semi-volatile, and non-volatile constituents. The CCME regulated volatile constituents include: benzene, toluene, ethylbenzene, total xylenes (BTEX); and CCME PHC F1. The group of semi-volatile constituents generally includes CCME PHC F2 and associated PAHs, while the non-volatile constituents include compounds with carbon numbers greater than 16 (PHC F3 and associated PAHs, PHC F4, and residual components).

When fresh product is released to soil, PHCs undergo biological, physical, and chemical alterations. These alterations do not occur independently of each other, but rather together in varying degrees depending on the surrounding soil conditions. Volatile components are released to the air or air-filled pore spaces. Semi- and non-volatile components sorb onto soil particles or organic matter, migrate to the water table in the dissolved phase or colloidal transport of sorbed contaminants, and/or are degraded by soil microorganisms (Potter and Simmons, 1998).

Microbial degradation of PHCs is the main process involved in removing PHC contamination in soils. Biodegradation is a biologically mediated process that chemically alters the structure of a chemical, the common result being the separation of the chemical into smaller components (CCME, 2000). Microbial degradation of complex mixtures of PHCs in soils results in the biotransformation of the original mixture to various end-products depending on the structure of the PHC molecules. Since biodegradation is the major pathway of PHC degradation and removal from soils (Atlas, 1981; Douglas et al., 1996; Fedorak and Westlake, 1981), bioremediation is a common remediation strategy used for soils contaminated with PHCs. Bioremediation is the process of using biological agents, in particular microorganims, to degrade (mineralize) waste components from the environment (e.g., PHCs) into simple inorganic constituents, such as carbon dioxide, water, and mineral components. Bioremediation of PHCs can be undertaken using various methods, and the process of composting is one method that is gaining recognition as an effective bioremediation strategy.



4.1 Biodegradation

Bacteria (including actinomycetes) and fungi in soils are the organisms that degrade petroleum in soils (Bossert and Bartha, 1984). Most soil bacteria are chemoheterotrophs, and obtain energy by oxidizing organic matter and use the metabolic products as sources of carbon for growth. Fungi are heterotrophs, and must obtain carbon and other nutrients from organic matter, similar to bacteria (Sylvia et al., 1999). Examples of carbon sources in soils available to soil bacteria and fungi as energy and nutrient sources are SOM, leaf litter, decayed material, and PHCs.

Microbial metabolism of PHCs occurs most efficiently in aerobic conditions, as oxygen is required as the terminal electron acceptor during oxidation of PHCs by microbes (Atlas, 1981). Anaerobic degradation of PHCs has gained increasing attention in the past two decades and may be significant under certain environmental conditions, such as in aquifers or marshes. Zang and Bennett (2005) present a review of current advances in anaerobic biodegradation of xenobiotics, including PHCs. Soil properties that influence microbial degradation of PHCs include: temperature; oxygen; moisture; pH; inorganic nutrients; soil texture; and organic matter quantity and quality (Bossert and Bartha, 1984).

Low concentrations of biogenic PHCs are present in most soils and sediments. Thus, hydrocarbon-oxidizing bacteria adapted to using hydrocarbons as a source of carbon are located in most natural areas in varying concentrations (Abiola and Olenyk, 1998; Rosenberg et al., 1992; Sims et al., 1990). As early as 1922, the addition of oil to agricultural soils was found to increase the numbers of aerobic bacteria in the soil. A decrease in species diversity accompanied this increase in microbial activity, as hydrocarbon-degrading bacteria were able to selectively utilize the PHCs as a carbon source (Baldwin in Bossert and Bartha, 1984).

To degrade large, complex organic substances, such as certain types of hydrocarbons, bacteria must excrete hydrolytic enzymes into the soil solution. These extracellular enzymes break down complex organic constituents into smaller molecules that can be absorbed into the cell. Enzymes are generally compound-specific so that many types of enzymes may be required to complete biodegradation of few organic constituents (Bollag and Bollag, 1992; Sims et al., 1990; Sylvia et al., 1999).

The types of hydrocarbons in petroleum mixtures have a greater influence on the degradability of individual hydrocarbon components than soil type, nutrient fertilizer addition, or microbial populations (Atlas, 1981; Huesemann, 1995). Also, molecular structure has a profound effect on the overall biodegradability and the biodegradation rate of a specific compound (Alexander, 1999). During PHC biodegradation, the more labile fraction is preferentially degraded leaving a residual that is much less available and mobile, a phenomenon known as biostabilization (Luthy et al., 1997). The sequence of biodegradation is generally: n-alkanes, simple aromatic components, branched alkanes, cycloalkanes, isoprenoids (e.g., pristane and phytane), and condensed aromatic compounds (Fedorak and Westlake, 1981; Connan, in Heath et al., 1997).



The residual fraction that remains in soils following weathering and aging is described as recalcitrant, or resistant to microbial degradation (Stantec, 2004). Organic compounds that are recalcitrant to biodegradation may persist in weathered soils for years. Recalcitrance is a function of both the chemical and physical characteristics of organic compounds (Abiola and Olenyk, 1998). Generally, the persistence of PHCs in the environment increases with an increase in the compound’s molecular weight and boiling point (Potter and Simmons, 1998). A molecule with a complex structure (e.g., fused rings, or compounds with double or triple bonded carbon atoms) may be too bulky or insoluble for microbial uptake and metabolism, or may interfere with binding to an enzyme’s active surface sites (Huesemann, 1997). Methyl branching generally increases the resistance of hydrocarbons to microbial attack (Atlas, 1981). Condensed PAHs and cycloalkanes, as well as high molecular weight alkanes, are mineralized very slowly. Structures containing four or more condensed rings are not easily degraded; however, ring structures of four to six rings may be co-metabolized or co-oxidized in the presence of more easily degraded compounds (Crawford et al., 1993).

Two conceptual models (Huesemann, 1997) have been used to explain residual hydrocarbon fractions that remain in the soil after bioremediation treatment: 1) the contaminant sequestration model; and 2) the inherent recalcitrance model. The contaminant sequestration model is used to explain the incomplete biodegradation of individual hydrophobic compounds such as pesticides or PAHs in soils (Hatzinger and Alexander, 1995; Weissenfels et al., 1992). The contaminant sequestration model hypothesizes that sorption and diffusion from soil particles and organic matter are the main processes that control the rate of biodegradation in soils and limit compound bioavailability (Huesemann, 1997). The inherent recalcitrance model postulates that the incomplete biodegradation of total petroleum hydrocarbons (TPH) or oil and grease (O&G) is due to the presence of certain hydrocarbons that, as a result of their complex molecular structure, are only slowly degraded or even inherently recalcitrant to biodegradation within the time frame of most bioremediation projects (less than five years).

The contaminant sequestration theory has been used to argue that the bioavailability or exposure to residual hydrocarbons is minimal or nonexistent because desorption rates of sequestered hydrocarbons are extremely low, thus presenting minimal toxicity and risk of exposure to residual hydrocarbons. The argument of residual hydrocarbons from bioremediation treatments posing minimal risk and toxicity has led to the view that residual PHCs may be left in the soil without further action and expense. However, bioremediated soils may exhibit toxicity effects even after TPH (O&G and/or PAH) biodegradation; thus, reasoning based solely on the contaminant sequestration theory is likely to result in a significant underestimation of potential soil toxicity effects for bioremediated soils (Huesemann, 1997).

Typically, degradation kinetics of residual compounds shows an initial rapid decline followed by a phase with little or no decrease in concentration (Huesemann, 1995; Jorgensen et al., 2000; Salanitro et al., 1997; Weissenfels et al., 1992). Degradation kinetics has been explained by the changes in availability of compounds to microorganisms as a result of sequestration that occurs during aging.



Degradation kinetics relies on a time-dependent change in bioavailability, as compared to other types of kinetics, such as zero or first order kinetics. The time required to reach the phase where there is little or no decrease or reduction in concentration and the time to reach the final percentage of available contaminant varies among soils and compounds (Alexander, 2000).

For a complete discussion of microbial metabolic pathways for hydrocarbons, please refer to Atlas (1981).

Walker et al. (1976) reported the first data set on rates of microbial degradation of PHCs. PHC degradation was found to be a dynamic process, not properly characterized by measuring total residue, but instead is characterized by the degradation of individual fractions. The authors found that the presence of methyl branches on the alkyl portion of certain hydrocarbons (e.g., pristane, phytane, etc.) resulted in significantly reduced biodegradation rates when compared to the relatively fast biodegradation kinetics of the non-methylated counterparts. In addition, an increase in the number of saturated rings in alicyclic hydrocarbons (cyclical compounds with at least one double bond, but no benzene ring in the structure) resulted in decreased biodegradation rates. The study indicated that the components of petroleum, although degraded simultaneously, were degraded at different rates.

Oudot (1984) demonstrated the preferential degradation of low molecular weight hydrocarbon compounds during a marine microbial degradation experiment. The author observed a progressive decrease of n-alkanes and iso-alkanes at a rate approximately proportional to their molecular weight. Setti (Heath et al., 1997) found that straight chain n-alkanes with greater than nine and with fewer than 28 carbons (C>9 to C<28) were rapidly metabolized within a matter of days, while n-alkanes with greater chain lengths (C>28) were either partially or not degraded over a period of two months. As for compound aromaticity in regards to biodegradability, Fedorak and Westlake (1981) found that simple aromatics (naphthalene and methylnaphthalenes) were at least as readily degradable as n-alkanes (C10 to C27) and that degradation rates decreased with the degree of alkylation. The study also found that aromatic and sulphur-aromatic compounds were degraded according to the number of aromatic rings, with degradation rates decreasing with increasing number of rings (mono- > di- > tri- > tetra-aromatic compounds).

There are few PHC degradation studies of hydrocarbons with boiling points above n-C35, such as compounds found in heavy oils and tars (CCME F4), since degradation rates of heavy oils and tars tend to be reduced compared to PHC compounds with lower boiling points. Huesemann and Moore (1993) found that during landfarming up to 85% of PHCs with carbon numbers above 44 (C>44) were biodegraded; likely saturated compounds. Heath et al. (1997) conducted several trials to investigate the biodegradation rate of high molecular weight aliphatic hydrocarbons of a crude oil containing n-alkanes up to n-hexacontane (n-C60). The authors found that at the end of the 136-day study, low molecular weight n-alkanes (n-C<30) were completely or extensively degraded and 14% of oil remained.



The residual fraction was comprised mainly of C>40 compounds. There was no observed decrease in C>45 compounds during the experiment.

4.2 Bioremediation

Bioremediation is the process of using biological agents, specifically microorganims, to degrade (mineralize) waste components from the environment into simple inorganic constituents, such as carbon dioxide, water, and mineral portions. Bioremediation has become one of the most rapidly developing fields of environmental restoration due to its cost effectiveness, minimal inputs or additions, and ability to often meet the goal of achieving permanent remediation at sites with soil contamination. In 2001, the U.S.EPA officially began using bioremediation as a cleanup method at Superfund sites across the United States (U.S.EPA, 2001). During bioremediation of soils with chronic hydrocarbon contamination, the contaminated soil serves as the microbial source of hydrocarbon degraders (Abiola and Olenyk, 1998).

Bioremediation of soil contaminated with organic chemicals, such as PHCs, is accomplished by degradation of specific organic constituents (i.e., “parent” compounds) in the contamination. In the natural environment, a constituent may not be completely degraded, but undergoes biotransformation into intermediate product(s) that may be less, equally, or more toxic than the parent compound, as well as more or less mobile in the environment (Bollag and Bollag, 1992; Dua et al., 2002; Liu and Suflita, 1993).

Microbial degradation of waste is either aerobic or anaerobic, as discussed in Section 4.1, and may be carried out in place (in situ) or following removal (ex situ). In situ remediation refers to the treatment of contaminated materials in place (e.g., solvent vapour extraction, air sparging). Ex situ remediation refers to the treatment of materials once the material has been excavated, and materials are either treated on-site or removed to another site for treatment. Examples of ex situ treatments are: windrows; biopiles; bioreactors (Cacciatore and McNeil, 1995); and co-composting (Section 4.2.1). Ex situ bioremediation has been used extensively to successfully treat gasoline and diesel contaminated soils in Alberta (Abiola and Olenyk, 1998; Chaw and Stoklas, 2001), and also soils contaminated with crude oil with varying degrees of success (Angehrn et al. in Murphy and Charrois, 2003).

Historically, the success of a bioremediation program was evaluated by measuring the total degradation of a product (Song et al., 1990; Wang et al., 1990) or a limited number of individual oil compounds (Douglas et al. in Jensen et al., 2000). Biostabilization of contaminants in soils occurs as fractions undergo degradation, and increases with residence time (Alexander, 1995). The fractions that remain become progressively less available for uptake by organisms. The change in bioavailability through biostabilization lowers, but does not eliminate, the risk and exposure of the contaminant to receptors, even if total concentrations remain elevated above the applicable guidelines. Currently, the importance of the concepts of risk and exposure in remediation programs are gaining consideration as



acceptable treatment endpoints compared to historical objectives, when the success of a program was based on reaching total concentration endpoints for site cleanup goals.

While bioremediation is an effective remediation strategy, several environmental factors may limit bioremediation and must be considered during remediation planning and monitoring. These environmental factors include: temperature; pH; moisture content; type of microbial population; competition among microbial populations; type and availability of electron donors and acceptors (required for aerobic versus anaerobic degradation); concentration and type of hydrocarbons; and presence of co-contaminants (e.g., metals, salts, etc.; Dua et al., 2002). The contaminant may not be bioavailable to the microbes or may be inaccessible to the appropriate degrading enzymes because of binding to high molecular weight organic compounds (e.g., humus, peat, etc.), low water solubility, or possible steric hindrance at the site that is acted on enzymatically (Blackburn and Hafker, 1993; Bollag and Bollag, 1992). Pollutant availability is a major factor in determining the success or failure of a particular bioremediation strategy (Section 3; Alexander, 2000) and in appraising risk to human health and the environment. To overcome environmental limitations, in situ and ex situ treatments are enhanced by a variety of physical/chemical methods, such as fertilization, aeration, soil pH adjustments, and/or moisture control (Sims et al., 1990). Improvements of soil physical properties, such as moisture, organic matter, and nutrient contents are commonly done to optimize the bioremediation process (Brown et al., 1998).

Site-specific remediation programs should be developed and should include thorough contaminant characterization, treatability studies, and design and implementation of the bioremediation plan. Site-specific treatability studies are necessary components of the bioremediation plan and provide site-specific information concerning the potential rate and extent of bioremediation and identify pathways of contaminant migration. Treatability studies should be undertaken prior to full scale remediation activities. Studies can be conducted in laboratory microcosms, at pilot-scale facilities, or in the field. These studies usually provide optimum and homogeneous conditions with respect to mixing and contact of soil particles with waste constituents and microorganims (Cacciatore and McNeil, 1995; Sims et al., 1990).

4.2.1 Compost Use in Bioremediation Programs

Composting is an aerobic process in which microbial populations biodegrade organic compounds under elevated temperatures to produce a humus-like product called compost (Abiola and Olenyk, 1998). Compost is organic material that is commonly used as a soil amendment or as a medium to grow plants. Contaminated soil can be co-composted with traditional composting substrates, such as yardwaste and sewage sludge. Soil co-composting is the process of simultaneously biodegrading organic materials (i.e., composting substrates) and organic contaminants (U.S.EPA, 2001). The U.S.EPA recently included composting as one of four ex situ bioremediation technologies suitable for



use at Superfund sites (U.S.EPA, 2001), and soil co-composting is currently an emerging bioremediation technology (Beaudin et al., 1999; Guerin, 2001).

Mature compost is a stable humic material, created during composting by combining organic wastes (e.g., yard trimmings, food wastes, manures) in proper ratios into piles, rows, or vessels, and adding bulking agents (e.g., bark mulch, wood chips, sawdust) as necessary to accelerate the breakdown of the organic materials. The characteristics of compost differ depending on the initial substrates that form the compost, nutrient contents and release of nutrients, chemical characteristics, and density. The finished material is allowed to fully stabilize and mature through a curing process. Compost maturity is determined using various criteria, such as enzymatic activity (Ayuso et al., 1996), or chemical parameters such as the percentage of humic carbon compared to extractable carbon, or comparison between water-soluble carbon and total nitrogen (Garcia et al., 1993). In Canada, the Guidelines for Compost Quality (CCME, 1996b) outline quality guidelines for compost, including compost maturity. Compost maturity is based on several factors, including: specific C:N ratios; measured oxygen uptake; bioassays; internal temperature; and curing time.

The use of composted materials for soil amelioration promotes soil sustainability and re-use of organic waste products. Composts act as soil ameliorants capable of changing pH, moisture content, soil structure, and acting as a nutrient source, thereby improving the contaminated soil environment for microbial degradation and bioremediation (Namkoong et al., 2002). Many compost microorganisms and fungi are known to degrade chlorinated and non-chlorinated hydrocarbons, explosives and propellant contaminants, and pesticides in sediments and soils (Bennett et al., 1995; Bumpus et al., 1985; Cole et al., 1995; Davis et al., 1993; Michel Jr. et al., 2001; Rosenbrock et al., 1997). Compost has also been used to enhance soil properties for re-vegetation of flare pit soils during remediation and reclamation of contaminated soils (Abiola et al., 1997). The numerous microbial populations in compost provide an ecological environment for the biodegradation of complex molecules (Chaw and Stoklas, 2001; Hoffmann and Chaw, 1999).

One major concern with the use of compost as an amendment for bioremediation is the problem of mixing non-contaminated material (compost) with contaminated soil and producing a greater quantity of contaminated material if bioremediation is unsuccessful. To avoid this problem, treatability studies followed by pilot-scale testing must be carried out (Abiola et al., 1997; Abiola and Olenyk, 1998; Chaw and Stoklas, 2001; Semple et al., 2001). Another problem that is encountered is that compost, as an organic amendment, has a high sorptive capacity for organic compounds in soil. The addition of compost to contaminated soil may lead to sorption of organic contaminants to the organic matrix of the compost (Stegmann et al., 1991; Weissenfels et al., 1992), which may cause contaminants to become less bioavailable to microbes. Organic contaminants may also remain sorbed to compost when soils are extracted for contaminant analysis. In these instances, measured concentrations of the contaminant in the extract that is to be analyzed are not representative of total contaminant concentrations, and biodegradation may be inferred incorrectly (Kastner and Mahro, 1996).



Various studies have examined the degradation of PAHs in soils, soil-compost mixtures, and composted municipal wastes. Past studies focused on individual PAHs or the PAH fraction versus other petroleum fractions because of the potentially carcinogenic risk that PAHs pose to mammals, PAHs tend to bioconcentrate, and PAHs can be separated from whole oil samples using specific lab procedures, such as GC/MS. Sorption of PAHs by SOM can reduce the bioavailability of PAHs, thus reducing the rates of PAH degradation in soils and the rates of PAH bioconcentration in organisms. The strong association between PAHs and organic matter often limits passive remedial methods, such as bioremediation (Mashayekhi et al., 2006).

The addition of mature compost was found to enhance the degradation of PAHs in soils (Kastner et al, 1995; Kastner and Mahro 1996; Mahro et al., 1994; Martens, 1982; Sasek et al., 2003; Semple et al., 2001; Stegmann et al., 1991; Wischmann and Steinhart, 1997). Martens (1982) and Wischmann and Steinhart (1997) attributed PAH degradation to biotic versus abiotic transformations based on their results. Kastner et al. (1995) and Wischmann and Steinhart (1997) noted that PAHs with three or less rings were readily degraded. Stegmann (1991) and Mahro et al. (1994) concluded that nutrient additions did not assist PAH degradation when mature compost was added to soils. All of the above studies attribute PAH degradation to biotic transformation, but do not define the mechanism through which mature compost enhances degradation (e.g., ameliorates soil structure for native microbial populations, influence of compost on sorptive capacity of the soil, addition of microorganisms from the compost, etc.). Kastner and Mahro (1996) proposed that the organic matrix of the compost was essential for enhancing the degradation of PAHs through supporting indigenous soil microbial populations in the compost. The authors also proposed that organic carbon in the compost was necessary for stimulating PAH degradation.

An additional extraction step was performed in the experiment conducted by Kastner and Mahro (1996) and Wischmann and Steinhart (1997) to examine if the measurement of PAH degradation was attributed to sorption to humic materials in compost. Kastner and Mahro found that the recovery of PAHs did not increase with the additional extraction step, demonstrating that PAHs were not sorbed to the humic materials in the compost. Wischmann and Steinhart (1997) found that only the recovery of benzo[a]pyrene (five ring PAH) increased when using the additional extraction step, demonstrating that a portion of benzo[a]pyrene sorbed to the humic materials.

Several studies have examined co-composting to treat soils contaminated with weathered PHCs (Beaudin et al., 1996, 1999; Guerin, 2001; Jorgensen et al., 2000). Beaudin et al. (1996) measured a 73% decrease in weathered mineral oil and grease (MOG) after 278 days of co-composting in a laboratory-scale reactor. Guerin (2001) found that after six months, the average concentration of TPH (C10 to C36) in four windrows decreased from an average of 3,100 mg/kg to below the criterion of 1,000 mg/kg TPH for clean fill (in Australia). At the end of the study the material was suitable for use as fill on the site. Jorgenson et al. (2000) reported that 70% of MOG was degraded, and discussed the implications of the residual fraction when planning bioremediation programs.



Other studies have examined the effectiveness of adding mature compost to enhance degradation of weathered PHCs. Brown et al. (1998) added composted wastewater sludge with high concentrations of nitrogen (6%), compost with low nitrogen (less than 1%), and sawdust as bulking agents to a soil with weathered, structurally complex PHCs. The authors found that the use of high nitrogen sludge (6%) compost accelerated degradation of the PHCs when compared to an untreated control and to soil mixtures with bulking agents lower in total nitrogen (less than 6%). Near the end of the study, hydrocarbon removal declined due to substrate depletion and because the remaining hydrocarbons were structurally complex. Al-Daher et al. (2001) constructed soil piles in Kuwait from weathered contaminated soil after amending with mature compost, fertilizers, wood chips, dried sewage sludge, and hydrocarbon-utilizing bacteria. The authors found there was no significant influence of compost or sewage sludge addition on the rate of hydrocarbon degradation. Vasudevan and Rajaram (2001) added compost as an amendment during the bioremediation of oil sludge. The authors found that the addition of compost instead of inorganic nutrients did not enhance the removal of PHCs compared to the treatment containing organic nutrients, indicating the lack of suitable hydrocarbon-degrading strains in the compost.

In the above studies on PAH degradation with the addition of compost, and bioremediation studies using co-composting and mature compost, substrate degradation was measured using various laboratory procedures. Al-Daher et al. (2001), Beaudin et al. (1996), Brown et al. (1998), Guerin (2001), Jorgenson et al. (2000), Kastner and Mahro (1996), Martens (1982), Stegmann et al. (1991), and Vasudevan and Rajaram (2001) performed chemical extractions on the test materials and analyzed the extracts for total concentrations of PAHs or hydrocarbons, commonly using gas chromatography with a flame ionization detector (GC/FID). Instead of measuring total hydrocarbons at the end of the study, Wischmann and Steinhart (1997) measured PAH degradation products to study the degradation of coal tar.

Past studies using compost as an amendment during bioremediation trials of PHCs in soils have not shown decisively that compost additions are effective in remediating PHCs, in particular weathered or heavy-end (i.e., C>16) hydrocarbons. The utility of compost in bioremediation projects may only apply to fractions of petroleum, such as PAHs. Due to inconclusive results from the literature, treatability studies are recommended to assess the ability of compost to remediate PHC contamination in soils on a site by site basis.



5. PETROLEUM HYDROCARBON TOXICITY

Toxicity is the property or properties of a material that produce(s) a harmful effect upon a biological system (Crosby, 1998). A toxicant is the compound that produces this biological effect. The toxicity of an organic contaminant in soil is a function of the exposure concentration and the exposure duration on the monitored response of the test organism(s). Of critical importance is the dose that determines biological ramifications. The relationship between dose and biological effects is the dose-response relationship, commonly expressed as a dose-response curve. Concentration and dose are not always directly proportional or comparable from species to species. Crude oil and petroleum products are a significant source of environmental toxicity because of their persistence in the environment and their common usage in daily activities (Alexander, 1995; Landis and Yu, 1999; Stantec, 2004).

Little is currently known about the toxicity of hydrocarbon mixtures (such as fuels) or individual hydrocarbon compounds. Due to the diversity of various hydrocarbon molecular structures found in mixtures, attempting to measure the toxicity of all individual hydrocarbon compounds is impractical, time consuming, and not economically feasible. Studies examining the toxicity of PHCs have found that n-alkanes with fewer than nine carbon atoms (C<9), which are mainly liquids at room temperature, are toxic to most bacteria and are only degraded by few microorganisms. Alkanes with less than nine carbon atoms (C<9) are soluble in water, and the toxicity of hydrocarbons (C<9) is linked to their solubility. As the n-alkane chain length increases above C9, rates of bacterial growth generally increase, demonstrating a decrease in toxicity (Heath et al., 1997). Three-, four-, and five-ring PAHs are the aromatic compounds that have been found to be the most toxic to soil microbes (Huesemann, 1997).

The route of contaminant exposure will influence the overall extent of soil toxicity. Potential exposure routes include ingestion (soil or leachate), direct contact (dermal, plant roots), and inhalation. Potential receptors include humans, wildlife, soil invertebrates, and vegetation (Huesemann, 1997).

Aging (Section 3) is toxicologically significant to organisms because the assimilation, and acute (short-term exposure) and chronic (long-term exposure) toxicity of compounds has been observed to decline with residence time in soils and as compounds become sequestered in soils. Aging reduces the concentrations of compounds (toxicants) that desorb from soil particles or SOM; therefore, ageing also indirectly affects human or ecological exposure to toxicants. Although aging reduces exposure and thus toxicity and risk, ageing does not eliminate exposure and risk (Alexander, 2000).

Laboratory tests that measure bioavailability and toxicity have problems associated with data interpretation because it is not yet clear how aging in nature should be simulated in the laboratory, and the possibility that adding pollutants to test soils using solvents may introduce artifacts (Alexander, 2000). Comparisons between field-aged PHC soils and fresh product in laboratory tests have showed



that there is less toxicity of field-aged hydrocarbons to organisms compared to fresh product, suggesting that aged PHCs are less bioavailable than fresh product (Roy and Visser, 2004).

The mobility, microbiological and chemical accessibility, and toxicity of chemicals in soils should be determined to adequately assess the potential risk that such chemicals pose to human and ecological receptors (Loehr and Webster, 1996). Little attempt has been made to correlate site cleanup with a measured decrease in the health hazards associated with biotreated material. A site that was biotreated but contains concentrations of one or more contaminants above the target levels may have been successfully cleaned up based on risk, even if contaminant concentrations are measured above the applicable guidelines. One strategy to deal with cases where residual TPH concentrations remain above regulatory targets is to prove that the residual hydrocarbon fraction does not pose any serious risks to human health or the environment and is deemed to present an acceptable risk. This approach may ultimately lead to the acceptance of less stringent guidelines when acceptable, which may in turn, result in more frequent site closures using cost-effective bioremediation and may significantly reduce clean up expenses for industries (Alexander, 2000; Huesemann, 1997).

5.1.1 Toxicity Testing

Bioassays measure the effect of a chemical on a test species under specified test conditions to quantify the toxicity of a compound (Loehr in Sims et al., 1990). Toxicity assays may utilize single or multiple species of test organisms. The use of a single species as a screening test is a common procedure employed to quickly assess the treatability of contaminated materials. A single species bioassay does not provide a comprehensive evaluation of the toxicity of a chemical in soil, but is generally a cost-effective approach to measuring the potential toxicity and risk of a compound in soils. Long-term bioassays focus on sub-lethal endpoints (e.g., growth and reproduction, seedling emergence, shoot and root lengths, shoot and root wet mass, and shoot and root dry mass) and are the most relevant type of bioassay from an assessment perspective (van Gestel et al., 2001; ESG, 2003). The evaluation of contaminated soils typically incorporates a battery of terrestrial organisms, multiple endpoints, and a range of sensitivities (Stantec, 2004).

Toxicity tests integrate exposure concentrations, bioavailability, and the toxic interactions between and among substances present in the soil and the physical and chemical properties of the soil and soil pore water (Aquaterre Environmental, 1998). Toxicity tests can be either acute or chronic. Acute screening tests can be either instantaneous or over a period of minutes to days. Test organisms are exposed to relatively high concentrations of the product(s) in soil for relatively short periods of time (e.g., less than 14 days). Chronic, or definitive tests entail exposure to relatively low concentrations of a contaminant in soils for durations that are at least one tenth (1/10) of the life span of a species (e.g., typically more than 14 days). Since some plant species are long lived, running a test for 1/10 of the life span of the species may not be possible; therefore, chronic tests using plants are referred to as definitive tests (Environment Canada, 2005). Definitive tests are considered to be more sensitive in assessing the



toxicity of oil contamination in soils compared to acute tests; however, definitive tests are generally time consuming, expensive, and require several concentrations and many replications to determine a reliable exposure concentration-response relationship (Stantec, 2004).

Screening tests are performed primarily to identify the “effective” concentrations that result in a particular response (e.g., mortality) or range of responses in the test organisms. The results of acute toxicity tests are then used to determine the optimal range of exposure for organisms exposed to sub-lethal levels of contaminants for the longer definitive tests (ESG, 2003). Most ecological risk assessments are based on acute toxicity data because of the high cost and long duration of chronic tests (Stantec, 2004).

Earthworms and terrestrial plants are common test species used in soil ecotoxicity testing. Earthworms are a particularly sensitive group of soil organisms because of the nature of interactions with soil. Earthworms typically ingest organic matter and small soil particles and are in constant contact with soil, making earthworms susceptible to the bioconcentration of soil contaminants. When earthworms consume small soil particles, the uptake of a contaminant increases in the organism relative to the bulk concentrations in the soil (Belfroid in Stantec, 2004). Earthworms also serve as a food source for higher trophic levels (e.g., worm-eating birds), which contributes to the potential for the biomagnification of contaminants in the food chain. Terrestrial plants are also commonly used to evaluate the phytotoxicity of compounds, since seeds and roots are in constant contact with soil particles. Typically, the types of plants used in regulatory testing are species of economic importance (e.g., crops or ornamentals). Environment Canada has developed accepted test procedures, conditions, and methods for toxicity testing for earthworms and terrestrial plants in Canada (Environment Canada, 2004, 2005).



6. ORGANIC MATTER AND PETROLEUM HYDROCARBON ANALYSES

Natural organic matter (NOM), such as peat or humic materials in soils may interfere with PHC analysis using chromatographic methods, such as GC/FID (CCME, 2001b). Analytical methodologies intended to quantify PHC contamination, such as the CCME reference method for PHCs in soils (2001b), are not able to resolve hydrocarbons from NOM versus hydrocarbons from PHCs. The inclusion of hydrocarbon concentrations from NOM reported as part of PHC concentrations may falsely elevate contamination concentrations. NOM interference with PHC analysis was observed during the analysis of organic (peat) soils from Alaska and the analysis of peat soils from the central and west-central portions of Alberta (White and Irvine, 1996; White et al., 1998).

Crude oil is principally composed of non-polar compounds with minor polar constituents; thus, solvents used to extract oil from soil are generally a combination of polar and non-polar solvents (e.g., 50:50 hexane:acetone). NOM contains similar hydrocarbon compounds as crude oil, as listed in Table C; thus, solvents used to extract crude oil from soil will also extract a fraction of NOM if present. The extraction solvent not only extracts compounds from NOM with the same molecular configuration as compounds in crude oil, but all compounds with similar degrees of polarity (White and Irvine, 1996).

The compounds in NOM that partition into non-polar solvents are called soil bitumens or natural fats, waxes, and resins. Soil bitumens compose between 1% and 5% of the NOM in mineral soils, and between 10% and 20% of the NOM in organic soils. Soil bitumens are composed of numerous types of compounds including aliphatic hydrocarbons, fatty acids, alcohols, PAHs, and other complex hydrophobic molecules.

Biogenic interference is the portion of NOM in soil that cannot be distinguished from petroleum in a standard analytical test for contamination, such as the CCME (2001b) reference method for PHCs in soils. Biogenic interference is generally a small fraction of total NOM; however, in organic or peat soils, biogenic hydrocarbon (Section 2.3.1) concentrations may exceed PHC guidelines set by regulatory agencies. White and Irvine (1996) found that methylene chloride, which historically was a common solvent commonly used to extract PHC from soil, extracted 7.5% NOM in an organic clay soil (soil with 20% NOM).

Knowledge of the baseline concentrations of biogenic hydrocarbons in a particular soil is required to assist in delineating contaminated areas at spill sites, set site-appropriate cleanup standards, and evaluate remediation progress (White et al., 1998; Garland et al., 2000). Soil bitumens should be excluded from the solvent phase if the integrity of the target analyte (i.e., crude oil) can be maintained during analysis. One standard procedure for removing biogenic interference from extracted soil samples is to perform a silica gel cleanup procedure prior to analysis. The silica gel cleanup procedure removes polar compounds (both NOM compounds that cause biogenic interference and PHCs) from



extracts (CCME, 2001b). Removal of polar PHCs during a silica gel cleanup procedure falsely lowers total reported PHC concentrations (White and Irvine, 1996). A second standard procedure may be followed instead of performing a silica gel cleanup to determine background NOM concentrations. The procedure is to collect samples from a nearby soil of the same type that is known to be uncontaminated (i.e., background sample) for PHC analysis. When the hydrocarbon analysis of a background sample is compared with that of the contaminated sample, biogenic hydrocarbons and petroleum contamination can be quantified separately (CCME, 2001b; White et al., 1998).

Table C Compounds present in both crude oil and natural organic matter (after Jorgenson in White and Irvine, 1996)

acetic acid heptacosane n-hentriacontane octanoic acid

alkanes heptacosanoic acid n-heptadecane o-xylene

benzene hexacosane n-hexadecane pentacosane

1,2-benzofluorene hexadecanoic acid n-nonacosane pentanoic acid

benzoic acid methane n-nonadecane perylene

butanoic acid methanethiol nonanoic acid phenanthrene

carbazole methanol n-octacosane propanoic acid

decanoic acid m-xylene n-octadecane p-xylene

2,6-dimethylundecane naphthalene n-pentadecane tetradecanoic acid

eicosanoic acid n-dotriacontane n-tetracosane toluene

ethanol n-docosane n-tetradecane

ethylbenzene n-eicosane n-triacosane

formic acid n-heneicosane n-triacontane

Attempts to quantify petroleum organics in soils have used silica, alumina, or fluoracil gels to selectively remove biogenic interference. Since biogenic interference consists of polar, semi-polar, and nonpolar molecules, gel cleanups may be successful at removing biogenic interference from one soil compared to another soil depending on the polarity of the biogenic interference (White and Irvine, 1996). None of the cleanup methods will remove all biogenic interference in any soil, and most gels compromise the semi-polar fraction of petroleum (e.g., 10 to 20% of crude oil; Garland et al., 2000).



The CCME reference method for PHC analysis recommends using a silica gel cleanup prior to GC/FID analysis to remove polar nonhydrocarbon compounds (CCME, 2001b). However, weathered or partially degraded PHCs can contain significant quantities of polar compounds, and a silica gel cleanup may remove polar PHC compounds. When attempting to quantify concentrations of weathered PHCs or PHCs in soils containing organic matter, the reference method recommends comparing a contaminated soil with an uncontaminated soil of the same type from a nearby location. Both should be analyzed without silica gel cleanup. In this case, subtraction of the clean soil from the contaminated soil gives an estimate of the concentration of PHCs in the soil.

Gas chromatography/mass spectroscopy (GC/MS) can be used to identify unique compounds, such as sterols, in NOM that are not present in petroleum (Woolard et al., 1999). GC/MS has been used to identify organic constituents in compost (Kafka and Kuras, 1993; Keeling et al., 1994; Nilsson et al., 2000; Zbytniewski and Buszewski, 2005). Comparisons between background (clean) and contaminated soils are generally done to differentiate NOM from PHC using GC/MS. To date, no studies have reported a methodology to easily and positively separate NOM from PHCs in soil samples using GC/MS.

Several studies from Alaska have examined analytical methods for differentiating NOM from PHC during soil chemical analysis (Garland et al., 2000; White and Irvine, 1996; White et al., 1998; White et al., 2004). White and Irvine (1996) assessed various standard test methods including gravimetric, spectroscopic, GC/MS, and infrared (IR) spectroscopy methods for the ability to differentiate NOM from crude oil in a peat soil from Alaska. The peat soil was spiked with crude oil and extracted with the following solvents: benzene:ethanol (2:1); methylene chloride; ethyl ether; carbon disulphide; benzene; and 1,1,2-trichloro-1,2,2,-trifluoroethane (freon-113). Gravimetric and spectroscopic methods were neither able to separate nor differentiate between hydrocarbon compounds from petroleum and NOM. All solvents extracted a sufficient mass of NOM to interfere with the gravimetric quantification of crude oil. Using GC/MS, extractable compounds from NOM were identified as n-alkanes (C19 to C31), alcohols, aldehydes, and ketones. Although GC/MS can be used to distinguish soil bitumens from a simple petroleum product such as gasoline, selectively separating NOM from compounds in crude oil was not achieved in the study since many compounds from each source were identical (e.g., n-C31). IR spectroscopy was used to quantify soil bitumens and crude oil in samples extracted with freon-113. Crude oil and NOM were found to be indistinguishable in the IR spectra.

Analytical pyrolysis is a small-scale thermal degradation method that is used for the chemical characterization of macromolecular materials, such as humic substances, from their pyrolysis products. Pyrolysis-GC produces a “fingerprint” of the pyrolysis products. The fingerprint can be used to characterize a sample by comparing the fingerprint to other known fingerprints using statistical comparisons (White et al., 2004). Garland et al. (2000) developed a test using pyrolysis-GC/FID to quantify biogenic interference in peat soil samples from northern Alaska. The soil samples had no known history of contamination; thus, all measured “petroleum” was derived from biogenic interference.



A total of 12 biogenic indicators (not listed in the published article) that corresponded to compounds with retention times ranging from approximately 46 to 59 minutes were used to build regression models for biogenic interference. The pyrolysis test was found to be successful at predicting biogenic interference in soil samples more accurately than other standard soil tests (e.g., C:N ratio, pH, percent organic carbon, extractable carbon, humic acids, fulvic acids, low molecular weight acids, hydrophobic neutrals, and hydrophilic neutrals).

White et al. (1998) identified a suite of compounds present in NOM, but not found in petroleum. The suite of compounds found in NOM, referred to as biogenic interference, was identified to infer biogenic interference. Three midrange alkenes were chosen as biogenic indicators because the alkenes adequately characterized the biogenic interference. The indicators that were selected were not positively identified by MS in the study; however, the compounds were identified as the alkenes n-C21, n-C22, and n-C23. Alkenes are commonly observed in the pyrolysate of NOM and are considered pyrolysis products of lipids. A linear regression curve between the biogenic interference and the biogenic indicators was used by the author to predict the amount of biogenic interference in uncontaminated and contaminated samples. Using a linear regression curve to predict biogenic interference is limited by sample variability (White et al., 1998; Woolard et al., 1999).



7. SUMMARY AND CONCLUSIONS

The following summary and conclusions are drawn from and based on the information gathered and presented in the literature review:

• The terminology used to describe and classify PHCs varies between industries (e.g., refining versus environmental), and an understanding of the terms facilitates communication regarding petroleum fractions and applicable guidelines.

• The majority of PHC chemical and physical characteristics related to crude oil refining are well understood and documented. Studies regarding PHC chemical and physical properties in soils relating to the time-dependent processes of weathering, ageing, and sequestration are ongoing.

• Two studies have reported the use of mature compost as an organic amendment during PHC bioremediation programs. Additional studies using mature compost would increase the knowledge base regarding the potential of using mature compost in bioremediation programs.

• The time-dependent processes of weathering, ageing, and sequestration influence the bioavailability of PHCs in soils. The bioavailability of PHCs directly influences the toxicity and associated risk of PHCs in soils during remediation programs. Toxicological properties of PHCs in soils are not well known as data from extensive long-term studies are slowly becoming available, and collecting toxicity data is time-consuming and expensive.

• NOM interferes with PHC analysis since similar hydrocarbon compounds are found in both NOM and PHCs. NOM interference with PHC analysis is soils may be reduced by using gel cleanup methods, such as silica or alumina gels prior to PHC analysis. Gel cleanups remove polar material (NOM and PHCs) from the sample extract; however, cleanup methods may artificially reduce PHC concentrations in soil samples contaminated with weathered PHCs as the majority of weathered PHCs in samples are polar. Another acceptable procedure to separate NOM from PHCs is to compare and subtract PHC concentrations in background soil samples from contaminated soil samples. GC/MS may be used to identify NOM and PHCs; however, an analytical method or technique has not yet been developed using GC/MS or another method to separate NOM from PHCs during routine PHC analysis.



8. CLOSURE

We trust that this report satisfies your current requirements and provides suitable documentation for your records. If you have any questions or require further details, please contact the undersigned at any time.

Report Prepared by WorleyParsons Komex

Brooke Bennett, M.Sc., A.Ag. Environmental Scientist

Senior Review by

Sean Murphy M.Sc., P.Ag. Senior Scientist Vice President Operations



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