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Effects of Produced Water and Production Chemical Additives on Marine Environments: A Toxicological Review

By

Jill Schmeichel

Submitted to the Graduate Faculty of North Carolina State University in partial fulfillment of the requirements for the Degree of Master of Environmental Assessment

Raleigh, North Carolina

2017

Approved by advisory committee:

Catherine LePrevost, PhD, Chair Waverly Kallestad, PhD, Co-Chair

April 7, 2017 2

ABSTRACT

Schmeichel, Jill, Masters of Environmental Assessment.

Effects of Produced Water and Production Chemical Additives on Marine Environments: A Toxicological Review

Review of current research suggests that produced water discharges are unlikely to induce widespread acute toxicological effects in marine environments due to the highly dilute concentrations of chemical constituents entering receiving waters, and their fate and transport behaviors post-entry. Biological uptake of chemical constituents known to be toxic has been observed in field and laboratory studies, though generally not at levels inducing acute toxicological impact. Evidence of health effects in biomarker studies suggests that more research is required to understand the impacts of long-term, low-dose exposure to produced water, particularly for marine organisms in early life stages. While there is somewhat limited research in the peer-reviewed literature specific to production chemicals, their impacts appear to be negligible as constituent concentrations have been diluted to the extent that they do not elicit toxic responses, even in cases where they may be intrinsically toxic.

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ACKNOWLEDGEMENTS

I would like to extend a very sincere thank you to my advisors, Dr. Waverly Kallestad and Dr. Catherine LePrevost, for their support and academic assistance over the course of this research. I would also like to thank Dr. Jerry Neff and Greg Durell for their insightful reviews of earlier drafts of this paper. Last, but not least, thank you to my husband Dean for his support and encouragement along the way, and my son Matthew for always reminding me what’s most important in life!

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Table of Contents

ABSTRACT ...... 2 ACKNOWLEDGEMENTS ...... 3 1.0 Introduction ...... 1 2.0 Produced Water ...... 3 2.1 Structure and Physicochemical Properties of Produced Water Constituents ...... 4 2.1.1. Monocyclic Aromatic Hydrocarbons ...... 4 2.1.2. Polycyclic Aromatic Hydrocarbons ...... 5 2.1.3. Phenol and Alkylphenol ...... 7 2.2. Models for Predicting the Dispersion and Ecotoxicity of Chemicals in Produced Water ...... 9 2.3 Eco- Toxicological Impacts ...... 10 2.3.1 Reproductive and Endocrine Impacts ...... 11 2.3.2 CYP1A Induction (EROD Oxidation) ...... 11 2.3.3 Other Impacts ...... 12 2.3.4 Chronic and Sub-Lethal Impacts ...... 13 3.0 Production Chemicals ...... 16 3.1 Physicochemical Properties of Production Chemicals ...... 17 3.2 Production Chemical Use ...... 18 3.3 Eco- Toxicological Impacts ...... 19 4.0 Conclusions ...... 21 5.0 References ...... 22

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1.0 Introduction

The world consumed over 62 quadrillion BTU (British thermal units) of petroleum- and natural gas- based energy resources in 2012 (Bilgen, 2014), and demand continues to rise as the global population grows and per capita gross domestic product (GDP) and energy consumption increase (Siirola, 2014). To meet this growing world energy demand, oil production is occurring in resource-rich regions around the globe, predominantly Russia, Saudi Arabia, United States, Iran, and Canada (International Energy Agency (IEA), 2012), with offshore (i.e. marine) fields producing an increasing portion of global supply (IEA, 2012).

Crude oil and natural gas produced from offshore oilfields usually are extracted from subsea geological formations as part of an oil-gas-water emulsion. The volume of water associated with oil is typically low during early stages of production, but may increase to several times the volume of oil produced toward the end of a well’s lifespan (Henderson, Grigson, Johnson, & Roddie, 1999). Once this water has been separated from the oil and gas, it is known as produced water and becomes a waste material that is most commonly either discharged to sea or injected into the hydrocarbon- bearing formation to enhance oil production or into another subsea formation for disposal (Fakhru’l- Razi, Pendashteh, Abdullah, Biak, Madaeni, & Abidin, 2009). Although treated prior to release, discharged produced water most often retains some hydrocarbon components from the crude oil as well as other natural chemical constituents, such as heavy metals and dissolved solids from the geological formation and synthetic chemicals that are added during the production process to prevent equipment corrosion and bacterial growth and for a myriad of other purposes (Fakhru’l-Razi et al., 2009). Estimates suggest approximately 250 million barrels (~39x109 L) of water are produced daily from world-wide oil and gas fields, with about 40% being discharged into marine environments (International Association of Oil and Gas Producers (OGP), 2005).

The environmental fate and transport of crude oil in marine environments is well-documented, its study triggered to a large extent by historical spills such as the Amoco Cadiz oil spill in French coastal waters in 1978, the Exxon Valdez spill in Alaska, USA in 1989, the BP Deepwater Horizon well blow-out in the Gulf of Mexico in 2010, and others. Volatile components typically evaporate, while other chemical components may be degraded by microorganisms (e.g., biodegradation), broken down through photochemical reactions (e.g., abiotic degradation), or mixed and sorbed to sediment and other particulates and dispersed with ocean currents. Sedimentary association of crude oil components is a particular environmental concern as the chemicals may persist in the environment for years or decades (National Research Council (NRC), 2003).

Understanding of the impacts of crude oil to marine life is limited. Lethality, reduction in fitness through sub-lethal effects, and disruption to the structure and function of marine communities and ecosystems have been observed in laboratory studies and after well-studied oil spills (NRC, 2003), but determination of long-term population effects at environmentally relevant concentrations has posed significant scientific challenges (NRC, 2003). Multiple temporal and spatial variables have made deciphering the eco-toxicological impacts in marine populations and ecosystems difficult (Wiens, 2013; NRC, 2003). 2

Crude oil can form a significant fraction of produced water composition, at documented ranges from 2- 565 mg/L (Fakhru’l-Razi et al., 2009). Based on this relationship between crude oil and produced water, the eco-toxicological impacts associated with crude oil in marine environments have often been extended to produced water. However, the chemical constitution and regional variability of produced water are well studied, and research indicates that the hydrocarbon constituents in crude oil vastly exceed those that are present at measurable concentrations in treated produced water, particularly once the discharge is diluted upon (re)entry to the receiving environment, for example via discharge at sea (Durell, Røe Utvik, Johnsen, Frost, & Neff, 2006; Neff, Lee & DeBlois, 2011a; Neff, Sauer & Hart, 2011b; Røe Utvik, 1999; Veil, Puder, Elcock, & Redweik Jr, 2004). Therefore, the impacts of crude oil in marine environments are likely not directly transferable to the impacts resulting from produced water. On that basis, research into the environmental impacts of produced water itself is expanding. Many of the same challenges relating to the identification of eco- toxicological effects as were observed in the study of crude oil in lighter oiled areas present themselves in produced water research.

In addition to crude oil, a secondary common component of produced water is synthetic chemicals, which are often referred to as production chemicals because they are added to assist in the oil production process. They serve a range of purposes and have a correspondingly wide range of chemical compositions, although their precise formulations are often undisclosed, even to the producers that use them, as they are proprietary to the chemical manufacturers. The concentration of most production chemicals is low in treated produced water (Neff et al., 2011a), as they may be largely “spent” in use prior to entering the produced water stream, or may be removed to varying extents during produced water treatment prior to discharge. Production chemicals are the subject of much industry, regulatory, and environmental interest because of their widespread use and potential toxicity to marine environments, although there is currently somewhat little available research on the effects of production chemicals in marine environments and the extent to which they may contribute toxicity to the produced water mixture.

The anthropogenic contribution of crude oil and production chemicals to marine environments is controlled to some extent by regulation, and produced water management and chemical use regulations are in place in many offshore producing regions around the world. Some of the regulatory approaches to limiting the risk to marine environments from crude oil and production chemicals include legislating parameter limits for produced water discharges (current regulatory limits for total oil and grease in produced water for offshore disposal ranging from 29 to 40 mg/L (Neff et al., 2011a)), chemical substitution programs phasing out the use of recognized marine toxicants in production processes and replacing with other formulations that are less toxic (e.g., OSPAR Commission PLONOR list [“Pose Little or No Risk to the Environment”], Ekins, Vanner, & Firebrace, 2007), mandating environmental testing and analysis of discharges from producers, and encouraging producers to limit releases and improve produced water discharge treatment technologies.

The objectives of this study are to review and summarize findings on the toxicologically and ecologically relevant environmental impacts of produced water, and to assess whether produced water discharge is likely associated with adverse environmental impact. Furthermore, literature regarding the impact of production chemicals will be reviewed to explore if any observed toxicological impacts of produced water may be directly attributable to these additives, understanding that their entry into the environment is primarily via produced water discharge. The scope of this review is 3

focused on peer- reviewed literature, with an acknowledgement that some volume of research is proprietary to industry stakeholders and is therefore not publicly available and/or is present in Industry reports. Additionally, this review will be limited to studies examining environmentally relevant contaminant concentrations that are representative of produced water and actual chemical dosing rates (thereby excluding studies discussing impacts of spills and other unplanned release events), and will be restricted to the study of chemicals used in the oil production process (thereby excluding chemicals used in exploration, drilling, refining, and other phases of oil and gas resource development).

2.0 Produced Water

The term ‘produced water’ refers to the water mixture that is removed from a geological formation during the extraction of oil and gas, and potentially also includes water which was injected into the reservoir to maintain pressure and oil production (Holdway, 2002). Produced water is a complex mixture with many variables influencing its characteristics, including the age and location of the oil field, the geological characteristics of the formation from which the water is originating, the type of hydrocarbon product being produced, the production history of the reservoir, and the operational conditions under which it originates (Fakhru’l-Razi et al., 2009). While the composition of produced water is considered highly variable (Durell et al., 2006) and constituent concentrations can vary between different sources by orders of magnitude (Neff et al., 2011a; Fakhru’l-Razi et al., 2009), it is similar across oil production facilities in terms of its major constituents (Fakhru’l-Razi et al., 2009). Fakhru’l-Razi et al. (2009) summarize the components of produced water to include crude oil, which is a mixture of aliphatic and aromatic hydrocarbons; dissolved formation minerals, including heavy metals and radioactive materials; production chemicals, which are typically synthetic additives; solids such as formation solids, corrosion and scale materials, bacteria, waxes and asphaltenes; and dissolved gases.

Oil is a generic term representing a wide array of compounds, mainly hydrocarbons, which may be present in produced water as dispersed droplets and/or dissolved in the water phase, depending on their solubility and structural properties (OGP, 2005). Aliphatic hydrocarbons are typically found in the dispersed phase, while carboxylic acids are most often found in the dissolved phase. Aromatics can be in either, or sometimes in both, depending on their molecular weight and structural complexity, with lower molecular weight compounds tending to be relatively more water soluble and thus more often present in the water (dissolved) phase (OGP, 2005).

Produced water is generated in large volumes in the production phase of conventional oil wells. Approximately 1.1 m3 is generated for each 1.0 m3 of oil produced worldwide (Neff et al., 2011a), making it definitively the largest waste stream associated with the production process (Arctic Monitoring and Assessment Programme (AMAP), 2010). Produced water is typically treated to remove the dispersed crude oil content (that is, droplets of crude oil, typically ranging from 1 to 10 um in size) (Neff et al., 2011a) before it is either discharged as a waste material into the sea, or is re- injected into a sub-sea formation for disposal (Ekins et al., 2007; Yeung et al., 2015). Environmental regulations in most jurisdictions dictate the allowable water quality parameters for discharged waters and often include maximum oil-in-water concentration limits, ranging between 14 mg/L and 39 mg/L (OGP, 2005). Current treatment methods are not entirely effective, and small suspended oil particles, 4

micro-emulsions, dissolved elements, and organic chemicals are often still present in treated produced water (Fakhru’l-Razi et al., 2009). Similar work has also demonstrated the presence of nonregulated compounds, specifically persistent organic contaminants such as hexachlorobenzene, decachlorobiphenyl, and octachlorodibenzofuran, in produced water (Balaam, Chan-Man, Roberts & Thomas, 2009). The most abundant organic chemicals in most treated produced waters are water- soluble low molecular weight organic acids (primarily mono- and di-carboxylic acids) and monocyclic aromatic hydrocarbons (MAHs) including benzene, ethyl benzene, toluene, and xylenes (Neff et al., 2011a). Produced water components thought to contribute most to the ecological risk in marine environments based on their chemical characteristics are the MAHs, polycyclic aromatic hydrocarbons (PAHs), related heterocyclic aromatic compounds, and sometimes one or more metals such as iron, lead, mercury, and zinc (OGP, 2005).

2.1 Structure and Physicochemical Properties of Produced Water Constituents

The structure and physicochemical properties of produced water compounds are significant in terms of their likelihood to be associated with adverse impacts, largely based on their potential to bioaccumulate and their susceptibility to biodegrade. Similarity in composition and commonalities in production operations allow for generalizations to be made about the characteristics and risk of produced water in marine environments.

2.1.1. Monocyclic Aromatic Hydrocarbons Benzene, toluene, ethylbenzene, and xylene (BTEX) are low molecular weight monocyclic aromatic hydrocarbons (MAHs). They are moderately soluble in seawater, highly volatile, and have a moderate affinity for partitioning into the lipid tissues of aquatic organisms (OGP, 2005; Neff, 2002). BTEX are rarely included when considering the ecotoxicological effects of produced water on marine environments (Bakke, Klungsøyr, & Sanni, 2013). This is largely because they are not accumulated to a large degree in marine organisms (OGP, 2005), and although total concentrations of BTEX may be as high 10,000 µg/L or greater in treated produced water, they dilute, evaporate, and are degraded very rapidly in the receiving water environment following discharge (Neff et al., 2011a; Neff, 2002). A study illustrating the dilution of BTEX showed a 14,900-fold reduction in BTEX concentration twenty meters down-current from a produced water discharge point (concentration in the treated produced water was 6,140 µg/L, versus 0.43 µg/L twenty meters downstream) (Neff, 2002). The main removal mechanisms for BTEX from the water column are evaporation, adsorption to sediment organic matter, biodegradation, and photolysis (Neff, 2002). Because of their high volatility, evaporation accounts for the greatest loss of BTEX from water (Neff, 2002). Under moderately calm open water conditions, the residence time of BTEX in the aqueous phase is roughly two days (Neff, 2002). Under more turbulent conditions, the half-life for BTEX in the water column may be only a few hours due to good vertical mixing (Neff, 2002).

Mechanisms of acute BTEX toxicity to marine organisms are thought to include non-specific mode of action (non-polar narcosis), alterations of cell membrane permeability particularly in the gills (Meyerhof, 1975; Morrow et al., 1975 as cited in OGP 2005), and potentially also developmental defects (Kjorsvik et al, 1982 as cited in OGP, 2005). Toxicity generally increases with increasing molecular weight although the rapid loss of BTEX in seawater limits exposure (OGP, 2005). Table 1 5

shows predicted median lethal concentrations for BTEX compounds. A Gulf of Mexico study demonstrated that most fish, crabs, and bivalve mollusks collected near offshore platforms discharging produced water did not contain detectable concentrations of benzene, toluene, or ethylbenzene (Neff, 2002). Roughly 95 percent of marine animals sampled contained less than 0.003 µg/g dry weight benzene, toluene, and ethylbenzene, with highest concentrations being detected in fish muscle at 0.046 µg/g of toluene (Neff, 2002). (Xylene was not analyzed in this study). As Table 1 illustrates, these concentrations are well below noted threshold effects levels. The absence of comparable studies supports the notion that MAHs are generally not the focus of ecotoxicological study with respect to offshore oil production.

TABLE 1. Common monocyclic aromatic compounds (MAHs) in produced water, their structure, and a comparison of representative produced water concentrations and their acute and chronic toxicities Chemical Structure Concentration in Treated PW Acute Toxicity to Marine Animals Chronic Toxicity to Marine Animals (prior to discharge)

Concentration Reference Organism Lethal Reference Organism Predicted No Reference (µg/L) Concentration Effect (LC50) Concentration (µg/L) (PNEC) or No Observed Effect Concentration (NOEC) (µg/L) benzene 800 - 4,500 Røe Utvik, 1999 freshwater 35,420 3,4 Neff, 2002 (no data) (no data) (no data) and marine animals 1,600 - 1,700 Gabardo et al., 2011 toluene 1,000 - 3,500 Røe Utvik, 1999 freshwater 27320 3,4 Neff, 2002 (no data) (no data) (no data) and marine animals 1,900 - 2,100 Gabardo et al., 2011 ethylbenzene 300 - 600 Røe Utvik, 1999 freshwater 19020 3,4 Neff, 2002 (no data) (no data) (no data) and marine animals 200 Gabardo et al., 2011 xylene (m-, p-, o-) 200 - 700 Røe Utvik, 1999 freshwater 2,900 (m-)3 Neff, 2002 (no data) (no data) (no data) and marine 8,570 (p-)3 animals 10,950 (o-)3 900 Gabardo et al., 2011

1. Value is sourced from the US Environmental Protection Agency’s AQUIRE database (1997) 2. Value is a geometric mean based on 48-hour exposure or greater, as per available data from the AQUIRE database (1997)

2.1.2. Polycyclic Aromatic Hydrocarbons PAHs are represented by hundreds of compounds with a wide variety of structures and properties. They are relatively insoluble in water and their potential to bioaccumulate increases with increasing molecular weight (OGP, 2005). Typically, a suite of 16 priority PAHs (per United States Environmental Protection Agency’s Toxic Pollutants list pursuant to 40 CFR 401.15) and/or a larger suite of 30-50 PAHs, including alkylated homologues for more detailed evaluation, comprise the focus for environmental assessments (Neff et al., 2011a). Their movement in the environment depends on properties such as how easily they dissolve in water and how easily they evaporate in air (Agency for 6

Toxic Substances Disease Registry (ATSDR, 1995), as well as how they are formed, as petrogenic PAHs appear to be more bioavailable than pyrogenic PAHs (Hylland, 2006). Petrogenic PAH are typically found in greater abundance in petroleum products (e.g., oil) while pyrogenic PAH are relatively enriched in products of incomplete combustion. A survey of literature shows produced water concentration ranges for the 2-3 ring (lower molecular weight, generally petrogenic) structures to be at least an order of magnitude greater than the 4-5 ring (higher molecular weight, typically pyrogenic) PAHs (Table 2). PAHs tend to be removed from the water column through volatilization to atmosphere (particularly the lower molecular weight fractions). Alternatively, they may be taken up by aquatic biota or persist in the environment through sorption to sediment or suspended particulates (typically the heavier fractions) (ATSDR, 1995, OGP, 2005;). Generally, PAHs are lipophilic and are taken up by marine organisms either directly from water or through diet but are not biomagnified through food chains as many animals are able to metabolize and eliminate them (Hylland, 2006)

Naphthalenes, specifically parent naphthalene and its C1-C4 alkylated homologues, are characterized by moderate water solubility and have a relatively low potential to bioaccumulate (OGP, 2005) but can be bioaccumulated in lower trophic level organisms that lack the required enzymatic functioning to metabolize them (Neff, 2011a). Because naphthalenes are rapidly degraded in the water column, they are thought to present a low risk of effects in the marine environment (OGP, 2005).

Higher molecular weight PAHs (more than four rings) tend to have very low water solubility and are usually present at very low or undetectable concentrations in produced water (OGP, 2005). They tend to remain associated with oil droplets and sorb tightly to particulates, thus having low availability to marine organisms (OGP, 2005). Their high affinity for particles and organic matter allows them to remain in marine sediment for months to years (Hylland, 2006), making contaminated sediment a particular source of concern for long-term organism exposure.

Higher molecular weight PAHs are the subject of growing environmental interest because of PAH (typically anthropogenic) in contaminated sediment, and their association with various adverse effects in fish, including carcinogenicity and impairment of immunological, histological, genetic, reproductive, and developmental systems (Kane Driscoll, McArdle, Menzie, Reiss, & Steevens, 2010). Uncertainties remain with respect to the mechanisms of PAH toxicity (OGP, 2005), but it has been clearly demonstrated that there are multiple potential modes of action for toxicity (including non- polar narcosis, phototoxicity, and biochemical activation which may result in mutagenicity, carcinogenicity, and teratogenicity), depending on the compound. Other influencing factors for mechanism of toxicity include whether exposure is acute or chronic, the organism, and the environmental compartment in which exposure occurs (OGP, 2005). Toxicity studies on fish are highly variable, with differences in the route and duration of exposure, and the number and type of PAHs to which fish are exposed (Kane Driscoll et al., 2010). Some of these studies relate to crude oil and others to combusted oil products, such as soot and coal ash.

In terms of acute toxicity, LOAELs (lowest observed adverse effect levels) as low as 0.06 and 0.07 mg/kg/day have been measured in fish following exposure to PAH mixtures from crude oil or field- collected sediments, although study authors in those cases noted that co-occurring contaminants may have contributed to the observed toxicity (Kane Driscoll et al., 2010). Total PAH concentrations at 0.5 - 23 µg/L in the water have been associated with yolk sac edema, morphological damage, delayed development, or death in a variety of species of developing fish and embryos, following exposure to 7

PAHs dissolved in crude oil (Carls at al., 2008; Carls & Meador, 2009). Studies have shown that a significant mechanism of toxicity for PAHs is phototoxicity, as ultraviolet light has been shown in some cases to dramatically increase the toxicity of some PAHs in some invertebrates, fish, and bivalve embryos (Hylland, 2006; Lee, 2003). Increased mortality and a variety of sub-lethal effects have been observed following laboratory exposure to a combination of PAHs and sunlight (Finch et al., 2016; Alloy, Boube, Griffitt & Roberts, 2015.; Lee, 2003).

TABLE 2. Common polycyclic aromatic compounds (PAHs) in produced water, their structure, and a comparison of representative produced water concentrations and their acute and chronic toxicities Chemical Structure Concentration in Treated PW Acute Toxicity to Marine Animals Chronic Toxicity to Marine Animals Concentration Reference Organism Lethal Reference Organism Predicted No Reference (µg/L) Concentration Effect (LC50) Concentration (µg/L) (PNEC) or No Observed Effect Concentration (NOEC) (µg/L)

naphthalene 1 80 - 1,060 Røe Utvik, 1999 freshwater 2,140 3 Neff, 2002 mussels 196 (PNEC) Neff et al., and marine 2006 animals

80 - 530 Neff, 2002 mussels 8 4,890 Neff et al., 2006 230 - 1,590 Durell et al., 2006 phenanthrene 60 - 100 Røe Utvik, 1999 freshwater 166 3 Neff, 2002 mussels 14.7(PNEC) Neff et al., and marine 2006 animals

7.4 - 19 Neff, 2002 mussels 8 368 Neff et al., 2006 benzo(a)pyrene not detected - Røe Utvik, 1999 freshwater 5 3 Neff, 2002 mussels 0.30 (PNEC) Neff et al., 0.2 and marine 2006 animals

not detected Neff et al., mussels 8 7.61 Neff et al., 2011a 2006

2-3 ring structures (various) 400 - 6,700 Bakke et al., (no data) (no data) (no data) (no data) (no data) (no data) 2013 10 - 390 Durell et al., 2006 4-5 ring structures (various) 0.2 - 6 Durell et al., (no data) (no data) (no data) (no data) (no data) (no data) 2006

4-6 ring structures (various) 0.4 - 12 Bakke et al., (no data) (no data) (no data) (no data) (no data) (no data) 2013

1. Naphthalene is generally considered to be a PAH although it does not fit the structural definition of having 3-fused rings as commonly defines PAHs 2. Value is sourced from the US Environmental Protection Agency's AQUIRE database (1997)

3. Estimated value based on published log Kow values

2.1.3. Phenol and Alkylphenol In terms of chemical structure, phenol is a hydroxybenzene. The five aromatic carbons not bonded to the hydroxyl group react easily with methyl carbons to form methyl phenols (alkylphenols) (Neff, 2002). Phenols are thought to be derived from the natural biodegradation of oil in the reservoir (Neff et al, 2006), and are also contributed to marine environments anthropogenically as a waste by- product from a variety of industrial processes (Neff, 2002). The phenol composition in produced 8

water is predominantly the less alkylated (C1-C3) alkylphenols, whereas those with higher molecular weight (consisting of branched alkyl chains) are typically only present at trace levels (Beyer et al., 2012; Neff, 2002) (Table 3). Phenol partitions almost entirely into solution (OGP, 2005). Alkylphenols with side chains of greater than five carbons are mainly associated with oil droplets (OGP, 2005). C8 and C9 alkylphenols do not occur naturally in oil (OGP, 2005). It is the long-chained alkylphenols (C7- C9) that are considered to be the most toxic phenols, as they can exhibit strong endocrine disruption, though they are rare in produced water (OGP, 2005, Neff et al., 2011a).

There is relatively little published information about the marine ecotoxicology of phenol as it is not considered to be particularly toxic to aquatic organisms, and is not environmentally persistent (OGP, 2005; Neff, 2002). Toxicity seems to increase with higher taxonomic position (Neff, 2002) and with alkyl chain branching (Neff, 2002), and it is the alkylphenols that are more abundant in marine environments (OGP, 2005). C1 to C3 alkylphenols typically have high water solubility and are degraded rapidly in marine environments (OGP, 2005). C4 and C5 homologues have a higher affinity for lipid phases (OGP, 2005). C6 to C9 homologues are typically hydrophobic and tend not to partition from oil into produced water in detectable amounts (OGP, 2005).

The mechanism of toxicity of phenols is believed to be related to bioaccumulation in target tissue lipids (Neff, 2002). C1 to C3 homologues are thought to behave as non-specific narcotic toxicants, capable of inducing acute effects in aquatic organisms (McCarty et al., 1992; McCarty et al., 1985, as cited in OGP, 2005) via compound accumulation in the target tissue lipids to a concentration that induces disruption (Neff, 2002). Moderately and highly branched alkylphenols, such as nonylphenol, have been associated with estrogenic effects in vertebrates (AMAP, 2010; Hylland et al., 2008; Kovarova, Divisova, & Svobodova, 2013; Neff, 2002; Sundt & Bjorkblom, 2011) and, to a lesser extent, in algae, clams, shrimp, and crustaceans (Hylland et al., 2008). Effects have been observed in species in laboratory studies (Meier et al., 2010; OGP, 2005), but not in wild populations (OGP, 2005), and the compounds are not detectable in environmental samples (OGP, 2005) (Table 3). Observed concentrations of phenol in produced water are typically lower than those associated with ecotoxicological impact (Table 3). Data as to the concentration of phenol downstream from produced water discharge points is lacking, and is generally estimated by models based on the phenol composition at the point of receptor entry as predicted by the modelled fate of the produced water plume (Beyer et al., 2012; Neff, 2002) .

Estimates for the levels at which phenols elicit acute toxicity have varied over time, and have decreased, presumably with advancements in measurement techniques. A 1999 study, for example, reported that nonylphenol and octylphenol are both acutely toxic to fish (17-3000 µg/L), invertebrates (20-3000 µg/L), and algae (27-5000 µg/L), with chronic NOAEC (no observed adverse effect concentration) values as low as 6 mg/L in fish and 3.7 µg/L in invertebrates (Servos, 1999) (Table 3). More recent studies have reported much lower values. The LOAEL for alkylphenols (C10 – C13) has been observed at 4 µg/kg, with NOAEL (no observed adverse effect level) at 0.4 µg/kg, but a conservative LOAEL of C4-C7 alkylphenols could be as low as 0.5 µg/kg if the exposure route is taken into account (Beyer et al., 2012). A 2012 risk assessment using cod, saithe, and haddock showed that in exposure simulations involving produced water and performed using the DREAM (Dose-related Risk and Effect Assessment Model) software model, only when the risk threshold was dropped to 0.02 µg/kg/day did a relatively small number of fish reach a level indicative of risk. This threshold is significantly lower than the previously observed NOAEL level of 0.4 µg/kg, suggesting that exposure to 9

alkylphenols through produced water is not likely to cause adverse effects on fish populations (Beyer et al., 2012). Based on their endocrine disruption properties, which are generally manifested after prolonged periods of exposure, highly alkylated alkylphenols (particularly octyl- and nonylphenol) are considered to have high chronic toxicity (Zha, Wang, Wang, & Ingersoll, 2007; Segner et al., 2003), though not necessarily at environmentally relevant concentrations.

TABLE 3. Common phenols and alkylphenols in produced water, their structure, and a comparison of representative produced water concentrations and their acute and chronic toxicities Chemical Structure Concentration in Treated PW Acute Toxicity to Marine Animals Chronic Toxicity to Marine Animals (prior to discharge) Concentration Reference Organism Lethal Reference Organism Predicted No Reference (µg/L) Concentration Effect (LC50) Concentration (µg/L) (PNEC) or No Observed Effect Concentration (NOEC) (µg/L) phenol 1 410 - 11,450 Røe Utvik, 1999 rainbow trout 13,100 Neff, 2002 (no data) (no data) (no data) 9 - 23,000 Fakhru'l-Razi pink salmon 3,730 Neff, 2002 et al., 2009 367 - 438 Neff et al., kelp shrimp 10,310 Neff, 2002 2011a alkylphenol 9.1 - 636 Farmen et al., (no data) (no data) (no data) (no data) (no data) (no data) (AP) 1 (various) (sum APs) 2010 2.8 - 290 Beyer et al., (APs, C4-C5) 2012 0.08 - 14 Beyer et al., (APs, C6+) 2012 nonylphenol not detected Neff, 2002 fish 17-3,00 Servos, fish 6 -23 (NOEC) Servos, invertebrates 20-3,000 1999 invertebrates 24-202 (NOEC) 1999 algae 2-2,500 algae 10-694 (NOEC)

1. Phenols and APs are often grouped together in the reviewed literature, and chemical species within the groups are often not specified

2.2. Models for Predicting the Dispersion and Ecotoxicity of Chemicals in Produced Water

Toxicity to MAHs, PAHS, and phenols (including alkylphenols), as with any other substance or group of substances, depends on the chemical dose to which a receptor organism is exposed. Other contributing factors to toxicity include the frequency and duration of chemical exposure, the route of exposure, biological factors (e.g., species, age, size, health status), chemical factors (e.g., solubility, polarity), and environmental factors (e.g., temperature, sunlight, and other water and air parameters) (Hodgson, 2010). Dilution of discharge into receiving ocean waters, precipitation, volatilization of low molecular weight hydrocarbons, physical-chemical reactions with other chemical species in seawater, adsorption onto particulate matter, and biodegradation into simpler and potentially less toxic chemicals all impact the volume and concentration of chemicals discharged into marine waters through produced water and therefore their potential to cause adverse impacts (Veil et al., 2004).

There are numerous well-established models that assist with estimating the fate of produced water following discharge into the ocean, the most common and relevant of these being plume dispersion models. Plume dispersion models are used to predict the dispersion, and more specifically the dilution rates and patterns, of discharged produced water. DREAM (Dose-related Risk and Effect 10

Response Model) (also an ecological risk model), CORMIX (Cornell Mixing Zone Export System), and OOC (Offshore Operators Committee) are several often-cited models deployed in this area of research. They may be used alone, but are sometimes used in tandem with monitoring studies to assess the validity of modeled versus empirical data (Niu, Lee, Robinson, Cobanli & Li, 2016; Neff et al., 2011a). Models vary in what they are specifically designed to estimate (e.g., short- versus long- term chemical dispersion, near- mixing zone versus far, steady state versus non-steady state) (Neff et al., 2011a). However, inputs into these models are generally similar and include parameters such as the volume and characteristics of the produced water, the rate and location of discharge, direction and speed of wind and water currents, air and water temperatures, water depth and stratification, and chemical composition differences between the produced water and ambient seawater (Niu et al., 2016; Neff et al., 2011a).

In terms of chemical composition, salinity is an important factor in determining how produced water disperses. The salinity of produced water can range from a few parts per thousand to roughly 300 parts per thousand, compared to a salinity of 32-36 parts per thousand for seawater (Neff et al., 2011a). Literature explaining the basis for differences in the salinity of produced water is lacking, and a recent study (Ozgun et al., 2013) showed seasonal and locational variability in a set of nine Turkish production wells. Highly saline produced water is usually as dense or denser than ocean water and disperses below the ocean surface, diluting rapidly upon discharge into well-mixed receiving waters (Neff et al., 2011a). Low-salinity produced water may form a surface plume and dilute more slowly (Neff et al., 2011a).

While plume dispersion models remain limited in their capacity to predict the fate of the various chemical components of produced water, they have evolved to enable very accurate simulations of near-field mixing processes of the produced water plume (e.g., where the outflow characteristics determine plume behavior, typically less than 500 meters from a discharge point), and reasonably good far-field mixing (e.g., where ambient flow conditions determine plume behavior, typically at least 1000 meters from a discharge point). Plume dispersion models consistently predict a rapid initial dilution of produced water discharge by 30- to 100-fold within the first few tens of meters of the outfall, followed by slower rates of dilution at greater distances (Neff et al., 2011a). These model results are supported by other field and laboratory research which has observed typically low hydrocarbon concentrations even at short distances from a discharge point (Harman, Tollefsen, Bøyum, Thomas, & Grung, 2008; Harman et al., (2009). Regulations in some jurisdictions recognize the dilution effect by allowing platform operators to discharge water exceeding set criteria into specified regulatory ‘mixing zones’. These mixing zones are defined as areas where wastewaters mix with receiving waters for dilution, and where applicable water quality criteria do not apply.

2.3 Eco- Toxicological Impacts

Research on produced water largely began in the 1990s and, to an extent, has advanced understanding of the environmental impacts of produced water discharge into marine environments. Study focus has evolved over this timeframe, with the earlier studies primarily aiming to define the profiles of produced water and how these may differ between regions, oilfields, and platforms. Once an understanding had generally been reached regarding the major constituents of produced water, research focus shifted toward attempting to understand the concentrations of these constituents in actively producing offshore areas, their fate and transport in marine environments, and, more 11

recently, the toxicological impacts to relevant organisms following exposure, with attention to acute responses.

2.3.1 Reproductive and Endocrine Impacts On the basis that crude oil is a component of produced water, albeit at varying concentrations, research has focused on the extent to which the toxic effects of crude oil extend to produced water. Endocrine and reproductive effects are among the more commonly studied endpoints for produced water discharge. It is relatively well established that some components of produced water, particularly alkylphenols, do have the potential to elicit endocrine and reproductive effects in fish, with observed effects on gonadal development, induction of vitellogenin, and inhibition of spermatogenesis and oogenesis (Bakke et al., 2013). However, researchers that have specifically studied exposure to produced water at environmentally relevant concentrations have generally concluded that widespread effects on endocrine and reproductive systems of fish at the population level are not likely (Bakke et al., 2013; Meier et al., 2010) even though some endocrine effects have been observed on individuals in environmental conditions which reflect those in close proximity to discharge points (Meier et al., 2010; Sundt & Bjorkblom, 2011; Tollefsen, Sundt, Beyer, Meier, & Hylland, 2011). Table 4 summarizes available studies that have shown toxicological response following exposure to environmentally relevant concentrations of produced water. In some of these cases, researchers have noted the aforementioned reproductive effects but not actual effects on more general health parameters, such as gonadosomatic, hepatosomatic, or fish condition index, therefore suggesting that endocrine disrupting effects were too low to elicit clear physiological or growth effects (Bakke et al., 2013).

TABLE 4. Reproductive studies showing the toxicological impacts of produced water on marine environments Study Reference Location Study Type Species Exposure Medium Impact Impacted Lifestage Not Impacted Meier et al., North Sea laboratory Atlantic cod real produced abnormal development of larvae, juveniles embryo survival, hatching 2010 water sample, pigmentation, inability to success diluted to 1% 1 feed, significantly elevated biomarkers for plasma and liver effects Tollefsen et al., North Sea laboratory Atlantic cod real produced up-regulation of sex- early adult plasma vitellogenin levels 2, 2011 water sample, steroid binding capacity in condition index 3, diluted to 0.5% blood, indicating hepatomsomatic index 4, interference with blood gonadosomatic index 5 steroid transport Sundt & North Sea laboratory Atlantic cod real produced increased levels of pre-spawning condition index, Bjorkblom, 2011 water sample, vitellogenin, impaired gonadosomatic index, diluted to 0.066% oocyte development, hepatosomatic index reduced estrogen levels, alterations in testicular development

1. Per Neff et al. (2011a), computer modelling has consistently estimated produced water dilution to be 1:30 at 10 meters from discharge point, 1:100 at 100m, and 1:1000 at 1 km 2. Plasma vitellogenin is a reproductive biomarker 3. Fish condition index is a measure of overall health, based on weight relative to a healthy fish of the same type and length 4. Hepatosomatic index is a measure of liver health and is also used to assess the level of environmental contaminants to which a species has been exposed 5. Gonadosomatic index is a measure of sexual maturity, based in ovarian/testicular development

2.3.2 CYP1A Induction (EROD Oxidation) Cytochrome P450 (CYP1A) is a monooxygenase enzyme which plays a notable role in the metabolism of hydrocarbons (particularly PAHs) in some marine organisms (primarily fish) following exposure to crude oil. The presence of PAHs in the cell triggers the induction and expression of CYP1A (Stagg & McIntosh, 1996). Within the cell, the CYP1A enzyme catalyzes a reaction that inserts an oxygen atom 12

into the hydrocarbon molecule, rendering it amenable to conjugation (and ultimately excretion) by Phase II enzymes, but at the same time increases the likelihood for PAHs to interact with other molecules within the cell (Stagg & McIntosh, 1996). Because of their potential to cause DNA damage (Stagg & McIntosh, 1996), and elicit other impacts on reproductive, developmental, genetic, immunological, and histological systems (Kane Driscoll et al, 2010), PAHs present an increased risk to organisms during embryogenesis and early developmental stages (Stagg & McIntosh, 1996). CYP1A enzyme activity in tissues can be measured as oxidation of EROD (ethyoxyresorufin-O-deethylase), which is a commonly studied endpoint as a biomarker of exposure to PAHs.

Results from available representative studies on CYP1A induction and/or EROD activity are displayed in Table 5. There are several limitations in this area of study which have prevented a clear determination as to the association between adverse toxicological impact and produced water discharge. The first of these challenges is the limited amount of research on environmentally relevant concentrations of produced water as a complex mixture rather than impacts from crude oil or the larger scale impact of oil installations. Second is the difficulty in determining the source of hydrocarbons triggering observed impacts in field-based studies, particularly for locations not near oil platforms. Additional sources of hydrocarbon inputs to marine environments may include shipping vessels, atmospheric deposition, coastal discharges and river and land runoff (Stagg & McIntosh, 1996). Finally, the association between EROD induction and deleterious effects is not clear (Stagg & McIntosh, 1996), as EROD induction is a biomarker and not an actual measure of adverse impact.

TABLE 5. CYPA1A Induction studies showing the toxicological impacts of produced water on marine environments Study Reference Location Study Type Species Exposure Medium Impact Impacted Lifestage Not Impacted Stagg & McIntosh, 1996 North Sea field fish (species not 1 ug l-1 1 induction of EROD activity larvae (no other specified) parameters studied) Abrahamson et al., 2008 North Sea research station Atlantic cod lab: real produced water concentration-dependant immature or in early EROD activity in- (laboratory and field) sample, diluted to 0.5% induction of EROD mature state situ and 0.1% 2 activity in laboratory field: studied in-situ study

1. Measured hydrocarbon fluorescence level. This is 5-10 times greater than background hydrocarbon levels in the North Sea 2. 0.1% dilution was used to approximate produced water concentration at 50-100m from a platform

2.3.3 Other Impacts Literature is available on other effects such as the genotoxic and cytotoxic effects of oil compounds on marine organisms, but studies specifically on the effects of produced water discharge are limited. Rybakovas, Baršiene, & Lang (2009) studied the genotoxic and cytotoxic effects in selected offshore areas of the Baltic and North Seas (Table 6). They looked at offshore drilling zones, areas with extensive traffic and/or major river influences, and areas thought to be comparatively far from potential pollution sources. Three fish species were collected from each sampling location using a trawler, their micronuclei and nuclear buds were assessed as genotoxicity endpoints, and their fragmented apoptotic cells were analyzed as cytotoxic endpoints. While no production areas were studied in the Baltic Sea (as there is very little production activity in this area), researchers found increased levels of genotoxic and cytotoxic effects in areas close to oil platforms in the North Sea, in zones with heavy shipping traffic, and in areas that were potentially impacted by contamination from a major river source. Genetic and cellular abnormalities were observed, as were adaptations to heavily polluted environments and genetic differences between studied fish populations in the Baltic and North Seas. These effects were associated with zones of elevated PAHs but were not directly attributable to produced water discharge. Additionally, ambient water temperature was cited as 13

being a factor in some of the cellular impacts as higher water temperatures affect mitotic activity and micronuclei formation in fish (Rybakovas et al., 2009).

Limited research has been done to assess other acute toxicological endpoints, including studies on oxidative stress and protein expression (Table 6), and these have generally resulted in a similar conclusion that adverse impact is concentration dependent, and is minimal at concentrations entering marine environments through produced water discharge. Several studies have also been done on the effects of microbial community structure in areas proximal to production platforms (Table 6), and these showed unique population characteristics between produced water and surrounding seawater (Yeung et al., 2015; Yeung et al., 2011) but, again, no clear evidence of adverse impact.

TABLE 6. Other studies showing the toxicological impacts of produced water on marine environments Study Reference Location Study Type Species Exposure Endpoint Impact Impacted Lifestage Not Impacted Medium Rybakovas et al., 2009 Baltic Sea field flounder, dab, ocean water, Genotoxicity/ highest levels of genotoxic and erythrocytes of adults (no other and North and cod offshore zones Cytotoxicity cytotoxic effects observed in areas (mature) parameters Sea close to production platforms, in studied) shipping zones, and in areas contaminated by large rivers; seasonal differences in nuclear abnormalities; species- and tissue- specific differences Farmen et al., 2010 North Sea laboratory rainbow trout produced Oxidative oxidative stress responses but only in hepatocytes of juveniles effects at in-situ water stress and AP and PAH concentrations higher concentration samples, cytotoxicity than what are representative of areas levels extracted into surrounding oil platforms water soluble and particulate organic fractions

Bohne-Kjersem et al., North Sea laboratory Atlantic cod synthetic Plasma Protein changes in protein expression juvenile (no other 2009 water sample, Expression 1 parameters spiked with AP studied) and PAH to mimic produced water Yeung et al., 2011 Atlantic laboratory bacterial produced Population different bacterial community in not applicable no differences Canada colonies water, Effects produced water and sediment versus between areas of seawater seawater produced water discharge and seawater Yeung et al., 2015 Atlantic laboratory bacterial produced Population unique microbial characteristics in not applicable no effects on Canada colonies water, Effects produced water versus seawater produced water seawater discharge on bacterial community in areas of discharge

1. Plasma protein expression (changes in blood protein) is being used in this study as a biomarker for a variety of effects, including immunological, fibrinolysis, fertility, bone resorption, fatty acid metabolism, oxidative stress, cell mobility, and apoptosis

2.3.4 Chronic and Sub-Lethal Impacts Literature on the chronic impacts of produced water discharge is limited to such an extent that it prevents a broad scientific understanding of the chronic ecological risk to marine environments. Much of the available research on long-term fish health is in the form of biomarker studies which are used rather extensively in North America and the North Sea (George et al, 1995, as cited in Gagnon, 2011). The concept of biomarkers in field-collected organisms is based on the recognition that an adverse effect will become apparent at the subcellular level before the effect appears at higher levels of biological organization such as population and community levels (Stein et al., 1998, as cited in Gagnon, 2011). Additionally, studying population level effects in fish is both expensive and impractical in the open ocean in the absence of major impacts (Gagnon, 2011). Table 7 outlines some of the chronic research currently available, predominantly in the form of biomarker studies. Study results have shown minor effects in gene expression in exposed fish populations (Holth, et al., 14

2008; Holth, Thorsen, Olsvik, & Hylland, 2010), evidence of the presence of some hydrocarbon constituents in tissues of caged mussels located close to discharge points (Brooks et al., 2011a), effects on health status biomarkers in mussels (Brooks et al, 2011a), and various behavioral and reproductive impacts in red abalone larvae, kelp, and sea urchins (Holdway, 2002). There is a noteworthy pattern in some of this research, showing that while impacts appeared to be associated with produced water concentration, there is not always an observable linear response in which effects increase with increasing produced water concentration, as evidenced in Brooks et al. (2011a), Brooks et al. (2011b), and Holdway (2002) and discussed in more detail below.

Sub-lethal effects following chronic exposure to produced water have been reported in fish and invertebrates at exposure concentrations below 10,000 ppm (1% dilution of produced water) (Bakke et al., 2013; Brooks et al., 2011a; Neff et al., 2011; Holdway (2002)). Brooks et al. (2011a) observed in a laboratory study that mussels exposed to produced water concentrations at 0.01% (100 ppm) had a marked health effect as measured by their Integrative Biological Response index (a widely accepted approach to assessing overall health status using biochemical, genotoxicity, and histochemical biomarkers (Brooks et al, 2011a)). The same study found no relationship specifically between the biological effects and contaminant concentrations measured, including at concentrations up to 1% produced water (10,000 ppm). Researchers hypothesized that additive or synergistic effects from chemicals present in very low concentrations or non-polar narcosis had occurred (Smith et al., 1998; Hannam et al., 2009 as cited in Brooks et al, 2011b), or that results were attributable to other, non- measured contaminants such as alkylphenols, organic acids or decalins (Brooks et al., 2011b). A laboratory study of red abalone larvae showed sub-lethal effects on settlement, metamorphosis, viability, and swimming behavior concentrations of 100 ppm of produced water (Holdway, 2002) and reduction in reproductive functions in giant kelp also at 100 ppm, although in the latter case the same results were not shown in field studies (Holdway, 2002). Concentrations of 1 ppm were shown to reduce sea urchin fertilization success in the same study, although fertilization was successful for 50% of test organisms at the highest tested concentration of 10,000 ppm (Holdway, 2002). Holdway (2002) also cites a 1991 study by Rabalais, McKee, Reed, and Means, which shows benthic community structure effects in terms of both species number and individual species abundance. Study results show impacts ranging from no effects at any distance, to effects at an 800m distance from the discharge point.

15

TABLE 7. Chronic impact studies showing the toxicological impact of produced water on marine environments Study Reference Location Study Type Species Exposure Medium Impact Impacted Lifestage Not impacted

Holth et al., 2008 North Sea laboratory zebrafish synthetic produced water down regulation in adult, offpsring of reproduction, mixture, diluted to 0.5% gene transcription, exposed adults recruitment and 0.05% reduction in fish condition factor (males), spinal cord deformations in first generation offspring Holth et al., 2009a North Sea laboratory Atlantic cod synthetic produced water differences in gene juvenile physiological mixture; PAH expression parameters 1., concentrations at 5.4 and condition factors, 0.54 ug L-1 ; AP at 11.4 and time to maturation 1.14 ug L-1 Holth et al., 2009b North Sea laboratory Atlantic cod synthetic produced water hepatic DNA not specified EROD activity mixture; PAH adduct formation concentrations at 5.4 and 0.54 ug L-1 ; AP at 11.4 and 1.14 ug L-1 Bohne-Kjersem et al., North Sea laboratory Atlantic cod surrogate produced water differences in juvenile (no other endpoints 2009 mixture, diluted to 0.01% protein studied) expression 2 Grung et al., 2009 North Sea laboratory Atlantic cod synthetic produced water elevated bile juvenile (no other endpoints mixture, diluted to 0.5% metabolites 3 studied) and 0.05% Brooks et al., 2011 North Sea field mussels, in-situ produced water, elevated bile adult other biomarkers 4 Atlantic cod measured at 200-250m metabolites from discharge point

Gagnon, 2011 Australia field surface fish in-situ produced water at physiological not specified (no other endpoints species 100m from discharge point parameters 5 studied) present in vicinity of platforms (various species; different at each facility) Codi King et al., 2011 Australia field Stripey in-situ produced water, elevated bile randomly collected other biomarkers 6 seaperch measured at 200m and metabolites from in-situ, (tropical fish 1000m from discharge ranged 2-15 years species) point Hamoutene et al., 2011 Canada laboratory Atlantic cod diluted produced water at oxidative junveile to adult weight gain, food concentrations of 0 ppm, metabolism (attained sexual intake, serum 0.01% (100ppm), and maturity during metabolites, resting 0.02% (200ppm) course of immunity, and experiment) stress response

Burridge et al., 2011 Canada laboratory Atlantic cod diluted produced water at embryo growth and juvenile EROD response, concentrations ranging heart rate were growth, plasma from 0ppm to 66% slowed 7 vitellogenin (6600ppm) Mathieu et al., 2011 Canada field American in-situ produced water in no significant adult fish condition, skin plaice the vicinity of the oil impacts and organ lesions, development site and at a EROD activity, reference location 8 heamotology, various histopathological indices in liver and gills

1. Physiological parameters included condition factor, liver somatic index, gonadosomatic index, and hematocrit

2. Protein expression impacts were related to (fibrinolysis, immune system, fertility, bone resorption, fatty acid metabolism, oxidative stress, impaired cell mobility, apoptosis) 3. Bile metabolites are a frequently used biomarker for exposure to PAHs and to a lesser extent, Aps 4. Other studied biomarkers were CYP1A (cytochrome P450 1A2), GST (glutathione s-transferase), ZRP (zona radiata protein), and Vtg (vitellogenin) in fish; LMS (lysosomal membrane stability) and MN (micronuclei formation) in mussels 5. Minor reduction in condition factor at one of three sites; liver somatic index elevated at two sites; elevated EROD activity and DNA damage at one site discharging high volumes of PW; stress proteins HSP70 elevated at all three sites; naphthalene and pyrene biliary 16

metabolites detected at significant levels at all three sites 6. Other studied biomarkers were hepatosomatic index, total cytochrome P450, and cholinesterase (ChE) 7. Mortality and development abnormalities were observed at very high concentrations of produced water (10% or 1000ppm and 66% or 6600 ppm) 8. Precise fish collection locations were not specified in study

It is difficult to establish conclusions as to the broad eco-toxicological impacts of discharges on marine environments due to the number of variables involved, including the range of geographical locations and changing environmental conditions associated with the study area; the variable characteristics of discharge water; and the extensive roster of toxicological endpoints to be considered. Despite the challenges, there is general scientific consensus on the major findings relating to the impacts of produced water in offshore environments. While there is evidence indicating that some chemical compounds in produced water are potentially toxic to marine organisms based on their inherent toxicity and observed impacts in studies on crude oil, sewage discharge, and other environmental outputs that have some chemical similarity to produced water, the current body of literature indicates that the assessed risk of acute toxicity to marine environments from discharged produced water is low (Ekins et al., 2007; Henderson et al., 1999; Røe Utvik, 1999; Yeung et al., 2015), with the main risk factor appearing to be the concentration of constituents rather than the total discharge volume of the potentially hazardous components it contains (Ekins et al., 2007). However, discharge volumes remain an important consideration as produced water effluent volumes generally comprise the largest operational discharge associated with production drilling. Chronic risk and risk to ecosystems and marine populations following exposure to produced water discharges remain understudied, and there is an outstanding need to develop methodologies and analytical techniques which can be used to reliably study and predict chronic impacts (Fakhru’l-Razi et al., 2009). Biomarkers have been widely advocated as a vital addition to the risk assessment procedure (Galloway et al., 2004, as cited in Codi King et al., 2011); however, it is also recognized that they should be applied as part of a suite of techniques rather than as a single assessment tool (Stebbing and Dethlefsen, 1992; and Galloway et al., 2004, as cited in Codi King et al, 2011).

3.0 Production Chemicals

Production chemicals are widely used in oil and gas development and production activities to treat and prevent a range of common problems which may be encountered during normal operations. Chemicals may be added at all stages of the production process, from the water injection phase in which the hydrocarbon resources are being extracted, through the oil-water separation phase, and including the transport phase where the recovered oil is sent to pipelines. Production chemicals are often introduced to the marine environment as components of discharged produced water. While the challenges associated with studying environmental impacts of discharged produced water are notable, they are amplified in trying to account for the contribution of individual components, such as production chemicals, to the overall toxicity of the waste stream. Added to the changing environmental conditions, variable discharge characteristics, and diverse endpoints and receptors in the study of produced water discharge are the further difficulties in understanding the broad range of production chemicals in use, and measuring their concentrations in waste streams and in the environment, and their toxicological effects.

In contrast to production chemicals, the crude oil component of produced water discharges has been 17

studied and is documented as being significant, but concentrations are not sufficiently high to elicit acute impacts, as the acute toxicity of produced water is thought to be low. A similar extrapolation may be extended to production chemical constituents, and that is to assume that their contribution to the acute toxicity of discharged produced water must be minimal based on the current understanding about impacts of produced water. However, because the toxicity of produced water is thought to be related to the concentration of its hazardous components more so than the total volume discharged, an understanding of the inherent toxicities of production chemicals, as well as their concentrations in produced water, is necessary to evaluate the contribution they make to aquatic toxicity.

3.1 Physicochemical Properties of Production Chemicals

Production chemicals are often complex mixtures of various compounds (Veil et al., 2004), some of which are known to producers and operators, while many others remain the proprietary to the supplying chemical manufacturers for business reasons. Some of the common chemical components of production chemicals are identified in Table 8, along with their chemical class and function.

TABLE 8. Common known constituents of production chemicals and their chemical class functions Group Function Common Chemical Components Corrosion inhibitors To prevent corrosion and scaling of pipes, sodium sulfite, zinc carbonate, ammonium equipment, and machinery bisulphite, zinc chromate, diammonium phosphate

Demulsifiers To separate oil-water emulsions oxyalkylated alkylphenols, aromatic hydrocarbons, glycols, cationic and non-ionic surfactants

Biocides To prevent micro-organism development and gluteraldehydes, sodium hypochlorite, microbial degradation quaternary ammonium salts, oxyalkylated phenols, fatty diamines, thiazolines, carbamates, paraformaldehyde, dichlorophenols

Hydrate inhibitors To prevent formation of hydrates which can plug methanol, ethylene glycol, alkyl triazines, amines, pipelines, equipment, and machinery amine phosponates

Defoamers To prevent or minimize foam in order to lower flourosilicones, octyl alcohol, aluminum stearate, interfacial tension various glycols, silicones, sulfonated hydrocarbons

Solvents To reduce freeze points and assist with other isopropanol, isobutanol, butanol, ethylene glycol, stabilizing functions methanol, esters, ethers

Source: Ekins et al., 2007

Known constituents of common production chemicals tend to be polar and hydrophilic compounds, which are not as amenable to routine analysis as nonpolar, hydrophobic compounds (McCormack, Jones, Hetheridge & Rowland, 2001), but are less prone to persist in the environment. Adding to the complexities of studying these mixtures, Henderson et al. (1999) note that in their study that it was not possible to differentiate between the toxic contribution of the production chemical and the crude oil content in the aqueous phase of produced water because the toxicological contribution of the oil could not be separately accounted for. With only general knowledge about the class of chemical constituents in a production chemical and the lack of suitable analytical methods, it is difficult to understand and predict the fate of these chemicals in the environment and therefore assess their potential impacts (McCormack et al., 2001). 18

3.2 Production Chemical Use

Scale and corrosion inhibitors, defoamers, and demulsifiers are generally the most heavily used production chemicals (AMAP, 2010). A sample of production chemical use from the North Sea is presented in Table 9 (Neff et al., 2011a).

TABLE 9. Production chemical use and discharge volumes from North Sea oil and gas platforms Chemical Typical Use Concentration Phase Association of Amount Discharged to North (ppm, v/v) Chemical Sea (t/y)

Scale inhibitor 3-10 Water 1143 Corrosion inhibitor 25-100 Oil 216 5-15 Water 22 H2O/ O2 scavenger Biocide 10-200 Water 81 Emulsion breaker 10-200 Oil 9 Coagulants & flocculants <3 Water 127 Gas treatment chemicals Variable Water 2846 Source: Neff et al., 2011a

A 1999 study of production chemicals indicates that most offshore production platforms use two or three types of chemical in their operations, with very few using none (Henderson et al., 1999). There are some indications that chemical use is significant and increasing (Henderson et al., 1999), (Table 10); however, not all studies support that trend. La Vedrine et al. (2015) cite work by Igunnu and George (2013) who determined that chemical use tends to increase as oil fields age.

TABLE 10. World demand for production chemicals, with projections to year 2021 Region

Million Dollars % Annual Growth

2006 2011 2016 2021 2006-2011 2011-2016

World 10,749 18,245 27,930 40,850 11.2 8.9 North America 5,580 10,240 15,270 21,100 12.9 8.3 United States 4,300 8,350 12,450 17,000 14.2 8.3 Canada & Mexico 1,280 1,890 2,820 4,100 8.1 8.3 Central & South America 650 1,220 2,300 4,050 13.4 13.5 Europe 1,460 1,890 2,560 3,650 5.3 6.3 Africa/Mideast 1,531 2,430 3,850 5,850 9.7 9.6 Asia/Pacific 1,528 2,465 3,950 6,200 10 9.9 Source: Upstream Pumping, 2016

In their recent study on production chemical usage in the United Kingdom (UK), La Vedrine et al. (2015) observed that the total quantity of discharged chemical product remained roughly the same over the study timeframe of 2000 through 2012. According to that study, over one million total tonnes of chemical products were discharged between 2006 and 2012, and these included products ranging from non-hazardous to highly hazardous as per UK rankings, with the greatest discharge amounts from the gas hydrate inhibitor class (155,758 tonnes, or 29% of total chemicals discharged), followed by scale inhibitors (62,473 tonnes, or 12% of total), hydrogen sulfide scavengers (59,599 tonnes, or 11% of total), and corrosion inhibitors (52,185 tonnes, or 10% of total). Demulsifiers and biocides each contributed approximately 5% of the total chemicals discharged (La Vedrine et al., 2015). A 2007 paper on chemical usage in the Artic reported similar findings on the use of chemical 19

product classes, stating that scale and corrosion inhibitors, defoamers, and demulsifiers were the most widely used in that region, with smaller amounts of hydrogen sulfide scavengers, gas hydrate inhibitors, and a range of others (AMAP, 2010).

The discharge of particularly hazardous chemical components (sometimes referred to as substitutable substances) within chemical products was also reviewed as part of the La Vedrine et al. study, and was found to have declined significantly in the 2000 through 2012 study period. The discharge of corrosion inhibitors accounted for the largest contribution of substitutable substance discharges. The corrosion inhibitor class products were discharged in relatively high volumes proportionally to other product functions (10% of total discharges) and their discharges have remained relatively constant despite containing the highest proportion of substitutable substances. La Vedrine et al. report that attempts to replace the more toxic constituents in this product class with ones posing lesser risk have resulted in reduced product efficacy, thus explaining the relatively constant discharge rates in this product class while substitutable substance discharges have declined in total and in other product classes. Scale inhibitors contributed roughly 12% of total discharges and contained a high proportion of substitutable substance chemicals in 2006, although they were used in declining proportions between 2006 and 2012. The hydrogen sulfide scavenger class included a small and declining proportion of chemical constituents considered particularly toxic. Gas hydrate inhibitors have used minimal proportions of substitutable substances and are therefore considered to be low-risk chemicals (La Vedrine et al., 2015).

Veil et al. (2004) identified the production chemicals having the greatest aquatic toxicity risk for adverse effects in marine organisms as being corrosion inhibitors, biocides, and reverse emulsion breakers, which is generally consistent with the La Vedrine et al. findings although the Veil et al. study did not specify the approach used to make these determinations. Veil et al. noted that these higher risk substances may undergo reactions prior to being discharged that reduce or eliminate their toxicity. While it is arguable that chemical composition and toxicity are highly variable for different formulations within the same chemical class, available literature suggests there are enough constituent similarities to make assertions about their relative toxicities in marine environments (La Vedrine et al, 2015; Grigson et al., 2000; Henderson et al, 1999).

3.3 Eco- Toxicological Impacts

Despite the widespread use of production chemicals, but perhaps because of the challenges associated with this area of study, there is very little published literature on the environmental impacts or contributing toxicity of production chemicals to the produced water mixtures that are entering marine environments. While there is potentially more information available in the private records of industry and government, only a handful of studies are publicly available.

Henderson et al. (1999) tested eleven heavily used production chemicals in the North Sea for their acute toxicity to marine bacterium using the Microtox rapid biological screening test, with the selected chemicals representing a range of functions and toxicities as per the UK’s Offshore Chemical Notification Scheme (OCNS) ranking system. The researchers observed that, for most production chemicals tested, there was little change in the toxicity response of test organisms to the aqueous 20

phase of a crude oil/brine mixture when production chemicals were added in their normal field dosage range. They also observed that in all cases where greater toxicological responses were seen, they were elicited by production chemicals of which a large or moderate proportion of their toxic components partitioned into the aqueous phase. Researchers concluded from their study that production chemicals may only make a small contribution to the overall acute toxicity profile of produced water discharges and even chemicals which are classified as highly toxic may not actually present an acute toxicity risk at dosages representing normal operating conditions. The same conclusion may not apply to all endpoints and/or marine organisms, however, as a 2013 study of genotoxic and cytotoxic responses in fish concluded that production chemicals occurring at very low concentrations, even below detection limits, may still act as genotoxicants or cytotoxicants and that their presence can lead to unpredictable genotoxic responses to pollution (Barsiene, Rybakovas, Lang, Andreikenaite, & Michailovas, 2013). This discrepancy highlights the need for continued research in this realm, particularly for sub-acute and chronic toxicity impacts.

Besides acute toxicological response, the Henderson et al. (1999) study also examined the partitioning behavior of eight production chemicals. Researchers found that 100% of the toxic components of biocides partitioned into the aqueous phase in both the biocides tested. Three corrosion inhibitor chemicals were tested and considerable variation was observed in their partitioning behavior, with anywhere from 0.5% to 93% of the chemicals’ toxicity partitioning into the aqueous phase. Of the three demulsifiers examined, two were classified with the highest toxicity category per OCNS ratings; however, none of the three produced a detectable level of toxicity in the aqueous phase at field- representative dosages. Dosing the demulsifiers in concentrations unlikely to be used under normal operating conditions did elicit a measurable response. In a similar study performed by Grigson, Wilkinson, Johnson, Moffat, & McIntosh, 2000), three corrosion inhibitors and three demulsifiers were evaluated for their partitioning behaviors between oil and water phases to determine likely discharge residues. Their analyses revealed that corrosion inhibitors partitioned primarily into the aqueous phase, and demulsifiers partitioned into the oil phase, largely in agreement with the work completed by Henderson et al. Grigson et al. (2000) also investigated levels of oilfield chemicals entering the marine environment via produced water discharge, using mass spectrometry and wet chemical analysis techniques. They observed that scale inhibitors were discharged in the produced water at roughly their dosing concentrations, and corrosion inhibitors could not be detected in the produced water (applying a detection limit of 5 mg/L). Low levels of benzoalkonium quaternary ammonium salts, which are typical corrosion inhibitor chemicals, were detected in sediment within 500m from the production platform.

Table 11 summarizes some acute toxicity study results for various production chemical types (Holdway, 2002). It is important to note that these toxicity values are based on production chemical concentrations while in process lines, and are not reflective of the production chemical concentrations in treated produced water. However, the table does demonstrate the inherent toxicity potentials, and the differences between the listed chemical types. The development of regulations around discharge composition, as well as initiatives such as the chemical substitution program and chemical selection guidelines to reduce the environmental hazards associated with production chemical use, has led to reductions in the use of high toxicity chemicals, suggesting, then, that the toxicity values presented may be higher than would be applicable to current formulations. For example, the data presented in Table 11 were studied prior to the 2015 La Vedrine study, which found significant decreases between 2000 and 2012 in the volume of highly toxic chemical 21

constituents entering marine environments through produced water.

TABLE 11. Acute toxicity of chemicals added in the production process / oil separation pipeline

Chemical Type Acute Toxicity EC50 (mg/L) Algae Brine Shrimp Microtox (minutes exposed) Reference (Skeletonema costatum) (Artemia salina) (Vibrio fisheri) Corrosion inhibitor 0.2 - 2 > 20 - 25 15 - 50 (15) Brendehaug et al., 1992 7.1 - 21.9 (5 - 30) Henderson et al., 1999 Scale inhibitor 60 1000 > 1000 (15) Brendehaug et al., 1992 Demulsifier 20 30 20 (15) Brendehaug et al., 1992 2.1 - 112 (5 - 15) Henderson et al., 1999 Biocide not applicable not applicable 15.2 - 33.7 (15 - 45) Henderson et al., 1999

Source: Holdway, 2002

1. Note that these toxicity values are based on production chemical concentrations while in process lines, and are not reflective of the production chemical concentrations in treated produced water

4.0 Conclusions

Most currently available research indicates that the acute toxicological risk of discharged produced water to marine environments is minimal. The chronic risks associated with long-term, low-level exposure to produced water are less understood. While there is some evidence of potential impact, based largely on biomarker studies which show reduced health in some organisms primarily in early life stages, there is no clear evidence of impact at the population or community structure levels. There is general scientific consensus that the water-soluble fractions of polycyclic aromatic hydrocarbons and alkylphenols are the produced water components that are most significantly associated with marine toxicity. Regulations in regions such as the North Sea and offshore areas in the United States have restricted or eliminated the use of chemical additives identified as being the most hazardous to marine environments from discharged produced water. Similarly, legislation around discharged water quality and chemical use are becoming increasingly rigorous. The contribution of production chemicals to the overall toxicity of produced water discharges is believed to be minimal, although there are recognized challenges in assessing their environmental impact as their full chemical makeup is often unknown, appropriate analytical methods are currently lacking or are insufficient for this purpose, and their impacts are difficult and/or impossible to separate from other constituents of produced water discharges. Regardless of source, it appears that the environmental concentration of potentially toxic chemicals is of greater ecological concern than the volume of those constituents entering the marine environment. However, both the concentration and the volume of potentially toxic chemicals must be considered together when assessing the risks of discharged produced water to marine environments. Knowledge gaps in understanding the chronic and cumulative risks of produced water discharges need yet to be filled, and the study of production chemicals is in its infancy. There is still much that needs to be learned in order that appropriate actions can be taken by producers in terms of chemical product selection and application of effective water treatment processes, and by regulatory agencies in order that they can determine the best way to protect and preserve marine environments in a manner that does not threaten petroleum resource development. 22

5.0 References

Abrahamson, A., Brandt, I., Brunström, B., Sundt, R. C., Jørgensen, E. H. (2008). Monitoring contaminants from oil production at sea by measuring gill EROD activity in Atlantic cod (Gadus morhua). Environmental Pollution, 153(1), 169-175. doi:10.1016/j.envpol.2007.07.025

ATSDR (Agency for Toxic Substances and Disease Registry). (1995). Toxicological profile for Polycyclic Aromatic Hydrocarbons (PAHs). Atlanta, GA: U.S. Department of Health and Human Services, Public Health Service.

Alloy, M. M., Boube, I., Griffitt, R. J., Oris, J. T., & Roberts, A. P. (2015). Photo-induced toxicity of Deepwater Horizon slick oil to blue crab (Callinectes sapidus) larvae: Photo-induced toxicity in blue crab zoea. Environmental Toxicology and Chemistry, 34(9), 2061-2066. doi:10.1002/etc.3026

AMAP (Arctic Monitoring and Assessment Programme). (2010). Assessment 2007: Oil and Gas Activities in the Arctic - Effects and Potential Effects. Volume 2. Arctic Monitoring and Assessment Programme, Oslo, Norway. vii + 277 pp.

Bakke, T., Klungsøyr, J., & Sanni, S. (2013). Environmental impacts of produced water and drilling waste discharges from the Norwegian offshore petroleum industry. Marine Environmental Research 92:154-169. doi:10.1016/j.marenvres.2013.09.012

Balaam, J. L., Chan-Man, Y., Roberts, P. H., & Thomas, K. V. (2009). Identification of nonregulated pollutants in North Sea-produced water discharges. Environmental Toxicology and Chemistry / SETAC, 28(6), 1159. doi:10.1897/08-488.1

Barsiene, J., Rybakovas, A., Lang, T., Andreikenaite, L., & Michailovas, A. (2013). Environmental genotoxicity and cytotoxicity levels in fish from the North Sea offshore region and Atlantic coastal waters. Marine Pollution Bulletin, 68(1-2), 106-116. doi:10.1016/j.marpolbul.2012.12.011

Beyer, J., Myhrevolatil, L. P., Sundt, R. C., Meier, S., Tollefsen, K., Vabø, R., et al. (2012). Environmental risk assessment of alkylphenols from offshore produced water on fish reproduction. Marine Environmental Research, 75, 2-9. doi:10.1016/j.marenvres.2011.11.011

Bilgen, S. (2014). Structure and environmental impact of global energy consumption. Renewable and Sustainable Energy Reviews, 38, 890-902. doi:10.1016/j.rser.2014.07.004

Bohne-Kjersem, A., Skadsheim, A., Goksøyr, A., & Grøsvik, B. E. (2009). Candidate biomarker discovery in plasma of juvenile cod ( Gadus morhua) exposed to crude North Sea oil, alkyl phenols and polycyclic aromatic hydrocarbons (PAHs). Marine Environmental Research, 68(5), 268-277. doi:10.1016/j.marenvres.2009.06.016

Brooks, S. J., Skarphéðinsdóttir, H., Liewenborg, B., Sundt, R. C., Harman, C., Grung, M., et al. (2011a). Water column monitoring of the biological effects of produced water from the Ekofisk offshore 23

oil installation from 2006 to 2009. Journal of Toxicology and Environmental Health, Part A, 74(7), 582-604. doi:10.1080/15287394.2011.550566

Brooks, S., Harman, C., Zaldibar, B., Izagirre, U., Glette, T. & Marigomez, I. (2011b). Integrated biomarker assessment of the effects exerted by treated produced water form an offshore natural gas processing plant in the North Sea on the mussel Mytilus edulis. Marine Pollution Bulletin, 62: 327-339.

Burridge, L., Boudreau, M., Lyons, M., & Courtenay, S. (2011). Effects of Hibernia production water on the survival, growth and biochemistry of juvenile Atlantic cod (Gadus morhua) and northern mummichog (Fundulus heteroclitus macrolepidotus). (pp. 329-344). New York, NY: Springer New York, 329-344. doi:10.1007/978-1-4614-0046-2_17

Carls, M. G., Holland, L., Larsen, M., Collier, T. K., Scholz, N. L., & Incardona, J. P. (2008). Fish embryos are damaged by dissolved PAHs, not oil particles. Aquatic Toxicology, 88(2), 121-127. doi:10.1016/j.aquatox.2008.03.014

Carls, M. G., & Meador, J. P. (2009). A perspective on the toxicity of petrogenic PAHs to developing fish embryos related to environmental chemistry. Human and Ecological Risk Assessment: An International Journal, 15(6), 1084-1098. doi:10.1080/10807030903304708

Code of Federal Regulations (40 CFR) 401.15 – Toxic Pollutants. Available at: https://www.law.cornell.edu/cfr/text/40/401.15 . Accessed on March 7, 2017.

Codi King, S., Conwell, C., Haasch, M., Mondon, J., Müeller, J., Zhu, S., et al. (2011). Field evaluation of a suite of biomarkers in an Australian tropical reef species, stripey seaperch (Lutjanus carponotatus): Assessment of produced water from the Harriet A platform. (pp. 261-294). New York, NY: Springer New York. doi:10.1007/978-1-4614-0046-2_14

Durell, G., Røe Utvik, T., Johnsen, S., Frost, T., & Neff, J. (2006). Oil well produced water discharges to the North Sea. part I: Comparison of deployed mussels ( Mytilus edulis), semi-permeable membrane devices, and the DREAM model predictions to estimate the dispersion of polycyclic aromatic hydrocarbons. Marine Environmental Research, 62(3), 194-223. doi:10.1016/j.marenvres.2006.03.013

Ekins, P., Vanner, R., & Firebrace, J. (2007). Zero emissions of oil in water from offshore oil and gas installations: Economic and environmental implications. Journal of Cleaner Production, 15(13), 1302-1315. doi:10.1016/j.jclepro.2006.07.014

EPA (U.S. Environmental Protection Agency). 1997. AQUIRE: Aquatic Toxicity Information Retrieval Database. Office of Research and Development, National Health and Environmental Effects Research Laboratory, Mid-Continent Ecology Division, U.S. Environmental Protection Agency, Duluth, MN.

Fakhru’l-Razi, A., Pendashteh, A., Abdullah, L. C., Biak, D. R. A., Madaeni, S. S., & Abidin, Z. Z. (2009). Review of technologies for oil and gas produced water treatment. Journal of Hazardous Materials, 170(2), 530-551. doi:10.1016/j.jhazmat.2009.05.044 24

Farmen, E., Harman, C., Hylland, K., & Tollefsen, K. (2010). Produced water extracts from North Sea oil production platforms result in cellular oxidative stress in a rainbow trout in vitro bioassay. Marine Pollution Bulletin, 60(7), 1092-1098. doi:10.1016/j.marpolbul.2010.01.015

Finch, B. E., Stefansson, E. S., Langdon, C. J., Pargee, S. M., Blunt, S. M., Gage, S. J., et al. (2016). Photo-enhanced toxicity of two weathered Macondo crude oils to early life stages of the eastern oyster (Crassostrea virginica). Marine Pollution Bulletin, 113(1-2), 316-323. doi:10.1016/j.marpolbul.2016.10.008

Gabardo, IT, Platte, EB, Araujo, Pugatti, FH. (2011). Evaluation of produced water from Brazilian offshore platforms. Pages 89-214 In: K. Lee and J. Neff (Eds.). Produced Water: Environmental Risks and Advances in Mitigation Technologies. Springer Science Publishers, New York.

Gagnon, M. M. (2011). Evidence of exposure of fish to produced water at three offshore facilities, North West Shelf, Australia. (pp. 295-309). New York, NY: Springer New York. doi:10.1007/978-1- 4614-0046-2_15

Galloway, T. S., Brown, R. J., Browne, M. A., Dissanayake, A., Lowe, D., Jones, M. B., et al. (2004). Ecosystem management bioindicators: The ECOMAN project – A multi-biomarker approach to ecosystem management. Marine Environmental Research, 58(2–5), 233-237. doi:http://dx.doi.org.prox.lib.ncsu.edu/10.1016/j.marenvres.2004.03.064

George, S.G., Wright, J. & Conroy, J. (1995). Archives of Environmental Contamination and Toxicology, 29: 530. doi:10.1007/BF00208384

Grigson, S. J. W., Wilkinson, A., Johnson, P., Moffat, C. F., & McIntosh, A. D. (2000). Measurement of oilfield chemical residues in produced water discharges and marine sediments. Rapid Communications in Mass Spectrometry, 14(23), 2210-2219. doi:10.1002/1097- 0231(20001215)14:23<2210::AID-RCM154>3.0.CO;2-8

Grung, M., Holth, T. F., Jacobsen, M. R., & Hylland, K. (2009). Polycyclic aromatic hydrocarbon (PAH) metabolites in Atlantic cod exposed via water or diet to a synthetic produced water. Journal of Toxicology and Environmental Health, Part A, 72(3), 254-265. doi:10.1080/15287390802539210

Hamoutene, D., Volkoff, H., Parrish, C., Samuelson, S., Mabrouk, G., Mansour, A., et al. (2011). Effect of produced water on innate immunity, feeding and antioxidant metabolism in Atlantic cod (Gadus morhua). (pp. 311-328). New York, NY: Springer New York. doi:10.1007/978-1-4614- 0046-2_16

Hannam, M. L., Bamber, S. D., Sundt, R. C., & Galloway, T. S. (2009). Immune modulation in the blue mussel mytilus edulis exposed to north sea produced water. Environmental Pollution, 157(6), 1939-1944. doi:10.1016/j.envpol.2008.12.031

Harman, C., Tollefsen, K. -., Bøyum, O., Thomas, K., & Grung, M. (2008). Uptake rates of alkylphenols, PAHs and carbazoles in semipermeable membrane devices (SPMDs) and polar organic chemical 25

integrative samplers (POCIS). Chemosphere, 72(10), 1510-1516. doi:10.1016/j.chemosphere.2008.04.091

Harman, C., Brooks, S., Sundt, R. C., Meier, S., & Grung, M. (2011). Field comparison of passive sampling and biological approaches for measuring exposure to PAH and alkylphenols from offshore produced water discharges. Marine Pollution Bulletin, 63(5), 141-148. doi:10.1016/j.marpolbul.2010.12.023

Harman, C., Thomas, K. V., Tollefsen, K. E., Meier, S., Bøyum, O., & Grung, M. (2009). Monitoring the freely dissolved concentrations of polycyclic aromatic hydrocarbons (PAH) and alkylphenols (AP) around a Norwegian oil platform by holistic passive sampling. Marine Pollution Bulletin, 58(11), 1671-1679. doi:10.1016/j.marpolbul.2009.06.022

Henderson, S. B., Grigson, S. J. W., Johnson, P., & Roddie, B. D. (1999). Potential impact of production chemicals on the toxicity of produced water discharges from North Sea oil platforms. Marine Pollution Bulletin, 38(12), 1141-1151. doi:10.1016/S0025-326X(99)00144-7

Hodgson, E (Ed).2010). A Textbook of Modern Toxicology (4th ed.). Hoboken, N.J: John Wiley & Sons.

Holth, TF, Thorsen, A, Olsvik, PA, Hylland, K. (2010). Long-term exposure of Atlantic cod (Gadus morhua) to components of produced water: Condition, gonad maturation, and gene expression. Journal of Fisheries and Aquatic Sciences 67(10):1685-1698.

Holth, T. F., Beylich, B. A., Skarphédinsdóttir, H., Liewenborg, B., Grung, M., Hylland, K., et al. (2009). Genotoxicity of environmentally relevant concentrations of water-soluble oil components in cod (Gadus morhua). Environmental Science & Technology, 43(9), 3329.

Holth, T. F., Nourizadeh-Lillabadi, R., Blaesbjerg, M., Grung, M., Holbech, H., Petersen, G. I., et al. (2008). Differential gene expression and biomarkers in zebrafish ( Danio rerio) following exposure to produced water components. Aquatic Toxicology, 90(4), 277-291. doi:10.1016/j.aquatox.2008.08.020

Holdway, D. A. (2002). The acute and chronic effects of wastes associated with offshore oil and gas production on temperate and tropical marine ecological processes. Marine Pollution Bulletin, 44(3), 185-203. doi:10.1016/S0025-326X(01)00197-7

Hylland, K. (2006). Polycyclic aromatic hydrocarbon (PAH) ecotoxicology in marine ecosystems. Journal of Toxicology and Environmental Health, Part A, 69(1), 109-123. doi:10.1080/15287390500259327

Hylland, K., Pinturier, L., Balk, L., Baršiene, J., Marigòmez, I., Feist, S. W., et al. (2008). Water column monitoring near oil installations in the North Sea 2001–2004. Marine Pollution Bulletin, 56(3), 414-429. doi:10.1016/j.marpolbul.2007.11.004

Igunnu, E. T., & Chen, G. Z. (2014). Produced water treatment technologies. International Journal of Low-Carbon Technologies, 9(3), 157-177. doi:10.1093/ijlct/cts049 26

International Energy Agency (Organization for Economic Cooperation and Development) (2012). UK: 2012 Energy Policy Review: Energy Policies of IEA Countries. Record number: 2012 IIS 2380- S17.91. 182pp.

Johnsen, S., Frost, T.K., Hjelsvold, M., Utvik, T.R. (2000). The environmental impact factor – A proposed tool for produced water impact reduction, management, and regulation. SPE 61178, Presented at the SPE HSE Conference, Stavanger, Norway, Society of Petroleum Engineers, Richardson, TX.

Kane Driscoll, SB, McArdle, ME, Menzie CA, Reiss, M. Steevens, JA. (2010). A framework for using dose as a metric to assess toxicity of fish to PAHs. Ecotoxicology and Environmental Safety 73:486-490.

Kjørsvik, E., Sæthre, L. J., & Lønning, S. (1982). Effects of short-term exposure to xylenes on the early cleavage stages of cod eggs (Gadus morhua L.). Sarsia, 67(4), 299-308. doi:10.1080/00364827.1982.10421344

Kovarova, J, Blahova, J, Divisova, L, Svobodova, Z. (2013). Alkylphenol ethoxylates and alkylphenols - Update information on occurrence, fate and toxicity in aquatic environment. Polish Journal of Veterinary Sciences 16(4): 763-772.

La Vedrine, M. A., Sheahan, D. A., Gioia, R., Rowles, B., Kroeger, S., Phillips, C., et al. (2015). Substitution of hazardous offshore chemicals in UK waters: An evaluation of their use and discharge from 2000 to 2012. Journal of Cleaner Production, 87, 675-682. doi:10.1016/j.jclepro.2014.09.053

Lee, R.F. (2003). Photo-oxidation and photo-toxicity of crude and refined oils. Spill Science & Technology Bulletin 8:157-162.

Mathieu, A., Hanlon, J., Myers, M., Melvin, W., French, B., King, T., et al. (2011). Studies on fish health around the Terra Nova oil development site on the Grand Banks before and after discharge of produced water. (pp. 375-399). New York, NY: Springer New York. doi:10.1007/978-1-4614-0046- 2_20

McCarty, L. S., Dixon, D. G., MacKay, D., Smith, A. D. and Ozburn, G. W. (1992). Residue-based interpretation of toxicity and bioconcentration QSARs from aquatic bioassays: Neutral narcotic organics. Environmental Toxicology and Chemistry, 11: 917–930. doi:10.1002/etc.5620110705

McCarty, L. S., Hodson, P. V., Craig, G. R. and Kaiser, K. L. E. (1985). The use of quantitative structure- activity relationships to predict the acute and chronic toxicities of organic chemicals to fish. Environmental Toxicology and Chemistry, 4: 595–606.

McCormack, P., Jones, P., Hetheridge, M. J., & Rowland, S. J. (2001). Analysis of oilfield produced waters and production chemicals by electrospray ionisation multi-stage mass spectrometry (ESI- MS). Water Research, 35(15), 3567-3578. doi:10.1016/S0043-1354(01)00070-7

Meier, S., Goksøyr, A., Folkvord, A., Klungsøyr, J., Svardal, A., Craig Morton, H., et al. (2010). Development of Atlantic cod ( Gadus morhua) exposed to produced water during early life 27

stages: Effects on embryos, larvae, and juvenile fish. Marine Environmental Research, 70(5), 383- 394. doi:10.1016/j.marenvres.2010.08.002

Meyerhoff, R.D. (1975). Acute toxicity of benzene, a component of crude oil, to juvenile striped bass (Morone saxatilis). Journal of the Fisheries Research Board of Canada, 32, 1864-1866

Morrow, J., Gritz, R., & Kirton, M. (1975). Effects of some components of crude oil on young coho salmon. Copeia, 1975(2), 326-331. doi:10.2307/1442886

Neff, JM. (2002). Bioaccumulation in Marine Organisms. Effects of Contaminants from Oil Well Produced Water. Elsevier Science Publishers, Amsterdam. 452 pp.

Neff, J.M., K. Lee, and E. DeBlois. (2011a). Produced water: overview of composition, fates, and effects. Pages 3-56 In: K. Lee and J. Neff (Eds.). Produced Water: Environmental Risks and Advances in Mitigation Technologies. Springer Science Publishers, New York.

Neff, J., Sauer, T., & Hart, A. (2011b). Bioaccumulation of hydrocarbons from produced water discharged to offshore waters of the US Gulf of Mexico. Pages. 441-477 in: K. Lee and J. Neff (Eds.). Produced Water: Environmental Risks and Advances in Mitigation Technologies. Springer Science Publishers, New York.

Neff, J. M., & Durell, G. S. (2012). Bioaccumulation of petroleum hydrocarbons in arctic amphipods in the oil development area of the Alaskan Beaufort Sea. Integrated Environmental Assessment and Management, 8(2), 301-319. doi:10.1002/ieam.1247

Neff, J. M., Johnsen, S., Frost, T. K., Røe Utvik, T. I., & Durell, G. S. (2006). Oil well produced water discharges to the North Sea. Part II: Comparison of deployed mussels ( Mytilus edulis) and the DREAM model to predict ecological risk. Marine Environmental Research, 62(3), 224-246. doi:10.1016/j.marenvres.2006.03.012

Niu, H., Lee, K., Robinson, B., Cobanli, S., & Li, P. (2016). Monitoring and modeling the dispersion of produced water on the Scotian shelf. Environmental Systems Research, 5(1), 1-15. doi:10.1186/s40068-016-0070-5

OGP (International Association of Oil & Gas Producers). (2005). Fate and Effects of naturally occurring substances in produced water on the marine environment. OGP Report No. 364. OGP, London, England. 42 pp.

Ozgun, H., Ersahin, M. E., Erdem, S., Atay, B., Sayili, S., Eren, E., et al. (2013). Comparative evaluation for characterization of produced water generated from oil, gas, and oil–gas production fields. CLEAN – Soil, Air, Water, 41(12), 1175-1182. doi:10.1002/clen.201200204

Rabalais, N.N., B.A. McKee, D.J. Reed, and J.C. Means. (1991). Fate and effects of nearshore discharges of OCS produced waters. Vol. 1: Executive Summary. Vol. 2: Technical Report. Vol. 3. Appendices. OCS Studies MMS 91-004, MMS 91-005, and MMS 91-006. U.S. Dept. of the Interior, Minerals Management Service, Gulf of Mexico OCS Regional Office, New Orleans, LA. 28

Røe Utvik, T. I. (1999). Chemical characterisation of produced water from four offshore oil production platforms in the North Sea. Chemosphere, 39(15), 2593-2606. doi:10.1016/S0045- 6535(99)00171-X

Rybakovas, A., Baršiene, J., & Lang, T. (2009). Environmental genotoxicity and cytotoxicity in the offshore zones of the baltic and the North Seas. Marine Environmental Research, 68(5), 246-256. doi:10.1016/j.marenvres.2009.06.014

Segner, H., Caroll, K., Fenske, M., Janssen, C. R., Maack, G., Pascoe, D., et al. (2003). Identification of endocrine-disrupting effects in aquatic vertebrates and invertebrates: Report from the European IDEA project. Ecotoxicology and Environmental Safety, 54(3), 302-314. doi:10.1016/S0147- 6513(02)00039-8

Servos, M. (1999). Review of the aquatic toxicity, estrogenic responses and bioaccumulation of alkylphenols and alkylphenol polyethoxylates. Water Quality Research Journal of Canada, 34(1), 123-177.

Siirola, J. J. (2014). Speculations on global energy demand and supply going forward. Current Opinion in Chemical Engineering, 5, 96-100. doi:10.1016/j.coche.2014.07.002

Smith, J. P., Tyler, A. O., & Sabeur, Z. A. (1998). Ecotoxicological assessment of produced waters in Indonesia. Environmental Toxicology and Water Quality, 13(4), 323-336. doi:10.1002/(SICI)1098- 2256(1998)13:4<323::AID-TOX5>3.0.CO;2-9

Stagg, R., & McIntosh, A. (1996). Hydrocarbon concentrations in the northern North Sea and effects on fish larvae. Science of the Total Environment, 186(3), 189-201. doi:10.1016/0048- 9697(96)05111-X

Stebbing, A.R.D., & Dethlefsen, V. (1992). Introduction to the Bremerhaven workshop of biological effects of contaminants. Marine Ecology Progress Series, 91:1-8

Stien, X., Percic, P., Gnassia-Barelli, M., Roméo, M., & Lafaurie, M. (1998). Evaluation of biomarkers in caged fishes and mussels to assess the quality of waters in a bay of the NW Mediterranean Sea, doi:http://dx.doi.org.prox.lib.ncsu.edu/10.1016/S0269-7491(98)00013-X

Sundt, R. C., Ruus, A., Jonsson, H., Skarpheinsdottir, H., Meier, S., Grung, M., et al. (2012). Biomarker responses in Atlantic cod (Gadus morhua) exposed to produced water from a North Sea oil field: Laboratory and field assessments. Marine Pollution Bulletin, 64(1), 144-152. doi:10.1016/j.marpolbul.2011.10.005

Sundt, R.C., & Bjorkblom, C. (2011). Effects of produced water on reproductive parameters in pre- spawning Altantic cod (Gadus morhua). Journal of Toxicology and Environmental Health, Part A, 74: 7-9, 543-554, doi:10.1080/15287394.2011.550563 29

Thomas, K. V., Balaam, J., Hurst, M. R., & Thain, J. E. (2004). Identification of in vitro estrogen and androgen receptor agonists in North Sea offshore produced water discharges. Environmental Toxicology and Chemistry / SETAC, 23(5), 1156. doi:10.1897/03-239

Tollefsen, K. E., Sundt, R. C., Beyer, J., Meier, S., & Hylland, K. (2011). Endocrine modulation in Atlantic cod (Gadus morhua L.) exposed to alkylphenols, polyaromatic hydrocarbons, produced water, and dispersed oil. Journal of Toxicology and Environmental Health, Part A, 74(7), 529-542. doi:10.1080/15287394.2011.550562

Tollefsen, K., Harman, C., Smith, A., & Thomas, K. V. (2007). Estrogen receptor (ER) agonists and androgen receptor (AR) antagonists in effluents from Norwegian North Sea oil production platforms. Marine Pollution Bulletin, 54(3), 277-283. doi:10.1016/j.marpolbul.2006.07.012

Transportation Research Board and National Research Council. (2003). Oil in the Sea III: Inputs, Fates, and Effects. Washington, DC: The National Academies Press. doi:10.17226/10388.

Upstream Pumping [Webpage]. Available at: http://www.upstreampumping.com/article/upstream- market/oilfield-chemicals. Accessed on March 9, 2016.

Veil, J. A., Puder, M. G., Elcock, D., & Redweik Jr, R. J. (2004). A white paper describing produced water from production of crude oil, natural gas, and coal bed methane. Argonne National Laboratory, Technical Report.

Wiens, J.A., (Eds.) (2013). Oil in the environment: Legacies and lessons of the Exxon Valdez spill

Yeung, C. W., Lee, K., Cobanli, S., King, T., Bugden, J., Whyte, L. G., et al. (2015). Characterization of the microbial community structure and the physicochemical properties of produced water and seawater from the Hibernia oil production platform. Environmental Science and Pollution Research, 22(22), 17697-17715. doi:10.1007/s11356-015-4947-z

Yeung, C. W., Law, B. A., Milligan, T. G., Lee, K., Whyte, L. G., & Greer, C. W. (2011). Analysis of bacterial diversity and metals in produced water, seawater and sediments from an offshore oil and gas production platform. Marine Pollution Bulletin, 62(10), 2095-2105. doi:10.1016/j.marpolbul.2011.07.018

Zha, J., Wang, N., Wang, Z., & Ingersoll, C. (2007). Histological alternation and vitellogenin induction in adult rare minnow ( Gobiocypris rarus) after exposure to ethynylestradiol and nonylphenol. Chemosphere, 66(3), 488-495. doi:10.1016/j.chemosphere.2006.05.071

Zhao, L., Chen, Z., & Lee, K. (2008). A risk assessment model for produced water discharge from offshore petroleum platforms-development and validation. Marine Pollution Bulletin, 56(11), 1890-1897. doi:10.1016/j.marpolbul.2008.07.013