Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review

Science of the Total Environment xxx (xxxx) xxx

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Review

Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review Angela C. Scott a,⇑, Warren Zubot b, Craig W. Davis c, John Brogly d a

Unaffiliated private contractor, correspondence c/o Canada’s Oil Sands Innovation Alliance (COSIA), 520 5th Avenue SW, Suite 1700, Calgary, AB T2P3R7, Canada Syncrude Canada Ltd., Edmonton Research Centre, 9421 17 Avenue, Edmonton, AB T6N1H4, Canada c ExxonMobil Biomedical Sciences, Inc., 1545 US Highway 22 East, Annandale, NJ 08801, United States d Canada’s Oil Sands Innovation Alliance (COSIA), 520 5th Avenue SW, Suite 1700, Calgary, AB T2P3R7, Canada b

h i g h l i g h t s

g r a p h i c a l a b s t r a c t

 Partitioning of OSPW-IDO to

synthetic biological storage lipids or membranes is low.  BCFs for OSPW-IDO in fish exposure studies are low, ranging from 0.6 to 53 L/kg-ww.  A biomimetic method used with QSAR modelling has been validated for predicting BCFs.

a r t i c l e

i n f o

Article history: Received 30 May 2019 Received in revised form 17 September 2019 Accepted 17 September 2019 Available online xxxx Editor: Damia Barcelo Keywords: Alberta oil sands OSPW Naphthenic acids Bioaccumulation Bioconcentration factor Reclamation

a b s t r a c t Bitumen recovery via mining in Alberta’s Athabasca region generates large quantities of oil sands process water (OSPW). Aquatic toxicity of OSPW has been well-studied and the class of organic compounds referred to as naphthenic acids (NAs) are consistently implicated as the primary driver. Proposed lease closure options include treated produced waters in reclaimed landscapes such as pit lakes and wetlands. Consequently, it is crucial to understand the bioaccumulation potential of NAs and other OSPW dissolved organics in these environments. Early studies were focussed only on NAs due to analytical limitations, however, later studies investigated additional classes of dissolved organics in OSPW. Reported bioconcentration factors (BCFs) for NAs in fish and amphibians range from 0.24 to 53 L/kg wet-weight. Most quantitative assessments of NAs bioaccumulation potential evaluated commercial NAs mixtures as a surrogate for OSPW and used using single-ion monitoring for measuring NAs concentrations. The resulting BCF values are based on the NA isomers that conform to the formula, C13H22O2. More recently, an advanced analytical technique capable of determining the profile of different isomer classes in OSPW showed that NAs and other OSPW ionizable dissolved organics (OSPW-IDO) have low partitioning to simulated biological storage lipids, suggesting low bioaccumulation potential. Using the same analytical technique to assess in vivo fish exposures, a subsequent study reported a range of BCFs for OSPW NAs between 0.7 and 53 L/kg wet-weight and heteroatomic isomer classes containing S or N heteroatoms had BCFs between 0.6 and 28 L/kg wet-weight. Reported BCFs for all isomer classes of the OSPW-IDO fraction were less than the Canadian standard for bioaccumulative designation (i.e., BCF  5000). Ó 2019 The Author(s). Published by Elsevier B.V. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).

⇑ Corresponding author. E-mail addresses: [email protected] (A.C. Scott), [email protected] (W. Zubot), [email protected] (C.W. Davis), [email protected] (J. Brogly). https://doi.org/10.1016/j.scitotenv.2019.134558 0048-9697/Ó 2019 The Author(s). Published by Elsevier B.V. This is an open access article under the CC BY license (http://creativecommons.org/licenses/by/4.0/).

Please cite this article as: A. C. Scott, W. Zubot, C. W. Davis et al., Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134558

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Contents 1. 2. 3.

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Quantification methods used in bioaccumulation studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bioaccumulation studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1. Investigations of the potential for bioaccumulation of NAs in fish . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.2. Development and evaluation of surrogate techniques to assess bioaccumulation potential of OSPW-IDO . . . . . . . . . . . . . . . . . . . . . . . . . 3.3. Investigation of NAs bioaccumulation potential in frogs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4. Discussion and conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . CRediT authorship contribution statement . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Declaration of Competing Interest . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Acknowledgments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1. Introduction Alberta’s Athabasca oil sands region contains bitumen reserves equivalent to approximately 170 billion barrels of crude oil (NEB, 2006). Surface mining of oil sands ore, followed by hot water extraction, is used to recover the bitumen for subsequent upgrading and refining (Masilyah et al., 2004, Masilyah et al., 2011). Oil sands process water (OSPW) is the term used to describe waters that are used in the alkaline hot water extraction process to recover bitumen from the oil sands (NRCAN – CanmetENERGY, 2010). In accordance with past and current practices, all produced OSPW is retained on-site for recycle and reuse in large settling basins commonly referred to as tailings ponds. The most recent data available shows the total liquid inventory of tailings ponds in the Athabasca region at 1.21x109 m3 as of 2016 (Government of Alberta, 2019). Oil sands mines are required to be reclaimed to boreal ecosystems and will include both terrestrial and aquatic features in the closure landscape. For this to happen, appropriately treated/remediated OSPW must be returned to the Athabasca river watershed in a manner that is protective of human and ecological health. Therefore, it is important to understand OSPW quality, including the chemical composition of the organic and inorganic fractions, bioaccumulation potential, and toxicity. This information can be used to assess risks of releasing treated OSPW and to inform engineering designs suitable for treatment. There can be compositional variability between OSPWs due to factors such as ore composition (e.g., salinity of connate waters), extraction method, tailings pond age, and the hydraulic retention time defined as the total volumetric OSPW inventory divided by the facility’s annual water demand (reviewed by Mahaffey and Dubé, 2017). Organic compounds originating from bitumen are the primary constituents of OSPW that contribute to its toxicity (Hughes et al., 2017b; reviewed by Mahaffey and Dubé, 2017; Li et al., 2017). These compounds include hydrocarbons (i.e., dissolved F1-F4 petroleum hydrocarbon fractions as defined by CCME (2001), benzene, toluene, ethylbenzene, and xylenes), polycyclic aromatic hydrocarbons (PAH), acid extractable organics (AOEs) including naphthenic acids (NAs), and phenolic acids) (Zubot, 2010). Extraction and tailings process aids such as sodium hydroxide and gypsum are water quality modifiers that also contribute to the water chemistry profile. OSPW used for bitumen production (i.e., operational purposes) is, in general, toxic to aquatic life. This constitutes a treatment requirement to support the safe release of OSPW back into the natural environment. NAs were identified early on as the main cause of OSPW toxicity (MacKinnon and Boerger, 1986; Clemente and Fedorak, 2005; Morandi et al., 2015; Hughes et al., 2017b; reviewed by Li et al., 2017). Consequently, numerous studies have attempted to establish accurate methods to characterize NAs and their

00 00 00 00 00 00 00 00 00 00 00

associated toxicity (reviewed by NRCAN – CanmetENERGY, 2010; Kannel and Gan, 2012; Zubot et al., 2012; Pramanik 2016; Mahaffey and Dubé, 2017; Li et al., 2017). NAs have classically been described as cyclic or alicyclic carboxylic acids having the formula CnH2n+zO2, where n represents the carbon number and Z indicates double-bond equivalents (i.e., the degree of unsaturation and/or number of rings) (Brient et al., 1995). Prior to the development of more sophisticated analytical methods, it was believed that the acid extractable organic (AEO) fraction of OSPW was primarily comprised of classical NAs. In fact, the AEO fraction contains many other compounds, including oxy-NAs (i.e., having more than two oxygen atoms, thus not simply carboxylic acids) and other species containing carbon, hydrogen, and various combinations of oxygen, sulfur or nitrogen heteroatoms (Grewer et al., 2010; Headley et al. 2013a, 2013b). Application of high performance liquid chromatographyultrahigh resolution mass spectrometry (HPLC-UHRMS) significantly enhanced characterization of OSPW dissolved organics, providing further evidence for the complexity of this mixture and the presence of many heteroatomic isomer classes in addition to classical NAs (Pereira et al., 2013; Pereira and Martin, 2015). Recent toxicity identification studies based on this method confirmed that the NAs isomer class is the most toxic but showed that other isomer classes contribute to the overall toxicity of OSPW (Morandi et al., 2015). To support progressive aquatic and terrestrial reclamation activities, OSPW must be treated such that it can be safely integrated within the hydrologic cycle. Development of release criteria would enable treated OSPW to be returned to the environment during active mining and reclamation phases. Consequently, systematic investigation of the environmental implications is currently underway (Allen, 2008; Kannel and Gan, 2013; Swigert et al., 2015; McQueen et al., 2017; Hughes et al., 2017b; Sun et al., 2017; Huang et al., 2018; Ajaero et al., 2018). In terms of treatment options for OSPW, it has been shown that OSPW toxicity decreases in scenarios where OSPW is left to age in tailings ponds (with or without stimulation) or in laboratory biodegradation studies (Clemente and Fedorak, 2005; Lo et al., 2006; Han et al., 2009; Toor et al., 2013; Yue et al., 2016; reviewed by Kannel and Gan, 2012). More aggressive treatment methods such as advanced oxidation and adsorption have shown promise, as have sophisticated bioremediation techniques (Zubot et al., 2012; Mahdavi et al., 2015; reviewed by Kannel and Gan, 2012; Xue et al., 2018). Significant questions remain about the economic viability of these options at the scale required for the oil sands operations and implementation of any form of commercial-scale treatment technology would require water release criteria to be in place in order to establish endpoints, defining water treatment objectives. The present review, however, considers another key environmental factor: the bioaccumulation and bioconcentration potential

Please cite this article as: A. C. Scott, W. Zubot, C. W. Davis et al., Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134558

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of NAs and other ionizable components of OSPW. The following terms will be used in this review and are defined based on what is generally accepted in the published literature (Wang, 2016) and by the Government of Canada (CEPA, 1999): Bioconcentration: the intake and retention of a substance in an organism entirely by respiration from water in aquatic ecosystems or from air in terrestrial ecosystems. Bioaccumulation: the intake of a chemical and its concentration in the organism by any mode of entry, including contact, respiration, and ingestion. Biomagnification: passage of a chemical up the food chain, reaching greater, possibly harmful, concentrations at high trophic levels among top predators. Bioconcentration factor (BCF): the ratio of a chemical’s concentration in an organism’s tissue (expressed on a wet-weight mass basis [mg/kg-ww]) relative to the chemical’s concentration in the exposure water (expressed on a volumetric mass basis [mg/L]. Typical units of BCF are L/kg-ww and normalized to a 5% lipid content. 1-octanol water partition coefficient (KOW): the ratio of a chemical’s concentration in 1-octanol to its concentration in water at equilibrium; represents a chemical’s hydrophobicity and is often used as a surrogate for a chemical’s potential to bioconcentrate in lipid and/or organism tissue; often presented in logarithmic form, i.e., log KOW. In Canada, a substance is considered bioaccumulative if it has a BCF  5000 or log KOW  5 (CEPA, 1999). Because bioaccumulation is a phenomenon, not an effect, bioaccumulation per se is not necessarily definitive in concluding an ecological concern (Mackay and Fraser, 2000). Bioaccumulation or bioconcentration potential are typically used as indications of the potential for substances to biomagnify within the food web, however these experimental systems cannot simulate the complex predator–prey dynamics and the relative uptake and elimination processes which ultimately control the ability of a substance to biomagnify (i.e., increasing the burden of the chemical - often expressed as a wet-weight basis in the organism - in higher trophic organisms) (Gobas et al., 2008). A low BCF (i.e., <2000) is often considered strong evidence that the substance has a very limited potential for biomagnification. However, a BCF > 2000 is not always indicative of biomagnification because it does not reflect metabolism and gut-transfer/ assimilation efficiency limitations throughout the aquatic food web (Gobas et al., 1993, 1999). Simple carboxylic acids (Z = 0) are not predicted to be bioaccumulative based on Quantitative Structure-Activity Relationship (QSAR) modelling (Koleva, 2012). However, the mixture of NAs and other dissolved organics in OSPW is extremely complex, including many high molecular weight multi-ring compounds that may be more difficult for organisms to metabolize or eliminate (Arnot et al., 2008a; 2008b). Thus, studying these substances directly is critical in assessing the bioaccumulation of OSPW dissolved organics more broadly. Previous work by CONCAWE has summarized the extensive body of literature regarding the bioaccumulation potential of neutral dissolved fractions of petroleum hydrocarbons (i.e., aliphatics, aromatics, naphthenes) (CITI, 1992; CONCAWE, 2001; Tolls and v Dijk, 2002; Camenzuli et al., 2019). Additionally, biomagnification factors (BMF) (Arnot and Quinn, 2015) and trophic magnification factors (TMF) (Wan et al., 2007; Takeuchi et al., 2009; Khairy et al., 2014) have been collected for a series of PAH compounds. Generally, observed BCFs for dissolved hydrocarbons are <5000 L/ kg-ww, with only one laboratory study1 (out of >100 experimental BCF values) within the CONCAWE data set demonstrating BCF values in excess of the bioaccumulative criteria (i.e., a single di-aromatic 1 Two previously reported studies which indicated BCFs > 5000 for 2 PAH compounds by de Maagd (1996) were later determined to be unreliable studies (Klimisch scores of 3) citing several key deficiencies in experimental design and chemical analysis protocols (Parkerton et al., 2008).

3

mono-naphthenic compound - triisopropylnaphthalene)2. This is consistent with the Arnot – Gobas experimental BCF database (Arnot and Gobas, 2006), which contains 35 hydrocarbon substances, only 2 of which have experimental values that exceed the bioaccumulative criteria (i.e., n-hexadecane and 2,2,4,4,6,8,8-heptamethylno nane). It should be noted that for several of these compounds (n = 3, and 4, respectively), additional experimental BCF studies indicate values below the threshold. This is consistent with BMF and TMF data for these substances which demonstrate a low potential to biomagnify within the food web (Wan et al., 2007; Takeuchi et al., 2009; Khairy et al., 2014; Arnot and Quinn, 2015; Lo et al., 2016). The weight of evidence present in the literature suggests that by and large, the bioaccumulation (and biomagnification) potential of the neutral dissolved fractions of petroleum hydrocarbons is limited, particularly for PAHs. Hereafter, we review studies that examined NAs and other ionizable components of OSPW, which are not included in the 2001 CONCAWE report or related work.

2. Quantification methods used in bioaccumulation studies Reliable quantification of NAs and other ionizable components of OSPW in aqueous systems has been a significant challenge (Grewer et al., 2010; Headley et al. 2013a,b; Hughes et al., 2017a, b); the commercially available Fourier-transform infrared (FTIR) spectroscopy method (Ripmeester and Duford, 2019) and a morerecently developed HPLC-UHRMS method (Pereira et al., 2013; Pereira and Martin, 2015) remain in use today, however, availability of the latter technique is limited. Considering that OSPW contains what has been termed a ‘‘supercomplex mixture” of watersoluble organics from bitumen (Jones et al. 2012), quantifying these compounds in a background of structurally similar fatty acids (i.e., from the tissues of fish or other aquatic organisms) further complicates the development of adequate methodology to determine BCFs. A summary of the evolution of quantification methods for application in bioaccumulation studies is included in Table 1. Initial efforts focussed on designing and validating methods for measuring classical NAs in fish tissues or fluids, specifically. Young et al., (2007) developed a qualitative method to detect the presence of NAs in rainbow trout (Oncorhynchus mykiss). Their method involves a laborious extraction process followed by chemical derivatization and gas chromatography–mass spectrometry (GC– MS) analysis to detect ions with nominal mass-to-charge ratio (m/z) of 267, corresponding to NAs with the formula C13H22O2 (n = 13, Z = -4). This GC–MS technique was previously shown to have good specificity for detecting NAs in various natural and industrial water samples (Merlin et al., 2007). To validate the method, store-bought rainbow trout fillets were spiked with either commercial NAs (refined Merichem) or NAs extracted from OSPW (Syncrude West In Pit, WIP) and then analyzed along with control fish. NAs were only detected in the fish that were spiked and thus, this method was subsequently utilized in preliminary assessments of bioaccumulation including fish exposure studies (Young et al, 2008, 2011; van den Heuvel et al., 2014) and one frog exposure study (Smits et al., 2012). However, it was later discovered that GC–MS consistently over-estimates the concentration of naturally-occurring NAs compared to higher resolution techniques or the industry standard FTIR analysis (Headley et al., 2013a; Hughes et al., 2017a). As a result, GC–MS is generally no longer used for these applications. GC–MS over-estimation or ‘‘inflation” of NA concentrations has the potential to occur in biological matrices as well, which may result in the cancellation of errors when steady-state BCFs are calculated as the ratio of tissue and aqueous 2 The test compound is a synthetically-derived isomer and it is unknown whether it is truly representative of structural isomers present in natural petroleum substances.

Please cite this article as: A. C. Scott, W. Zubot, C. W. Davis et al., Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134558

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Table 1 Summary of analytical methods used to quantify NAs and other OSPW-IDO in biological tissues or fluids. Analysis method

Quantification basis

References

gas-chromatography mass spectrometry (GC–MS)

signal intensity of C13H22O2 isomer group (a representative NA) relative to internal standard

liquid chromatography –high resolution mass spectrometry (LC-HRMS)

average of signal from seven representative NAs ions (C13H22O2, C13H24O2, C15H22O2, C15H24O2, C15H26O2, C15H28O2, C17H26O2) relative to internal standard fluorescence intensity converted to phenanthrene equivalent concentration summed intensity of all detectable ions in isomer classes defined by heteroatomic composition and ionization mode, relative to internal standards

Young et al., 2008, 2011Smits et al., 2012van den Heuvel et al., 2014 van den Heuvel et al., 2014

high performance liquid chromatography (HPLC) high performance liquid chromatography (HPLC) – ultra-high resolution mass spectrometry (UHRMS)

concentrations. Nevertheless, the GC–MS method considers only a single ion representing one empirical class of NAs. A liquid chromatography-high resolution mass spectrometry (LC-HRMS) method and a HPLC-fluorescence technique have also been used to evaluate bioaccumulation of classical NAs in fish bile (van den Heuvel et al., 2014). The LC-HRMS method based quantification on the average relative abundance of seven ions (refer to Table 1) corresponding to saturated (Z = -2 or 8) carboxylic acids with an odd number of carbons. These ions were reportedly chosen because they were deemed least likely to include the background fatty acids found in fish bile. Monitoring several ions offered improved specificity compared to the earlier GC–MS method and the use of high resolution analysis mitigated the over-estimation issue observed with GC–MS. The HPLC fluorescence analysis technique is not selective for NAs but generates an estimate based on parameters used for phenanthrene quantification and, thus, provided ‘‘phenanthrene equivalent NAs concentrations” (PEC-NAs). Results obtained using these two bile NAs analysis methods showed similar trends (i.e., NAs concentrations were generally greater in the bile of fish kept in waters with higher NAs content compared to levels in the bile of fish from reference lakes with <1 mg NAs/L) but produced conflicting absolute quantitative data making it impossible to establish BCFs. The measurement discrepancies may be attributed to the use of small sample sizes combined with the inherent challenges of measuring NAs in a complex background (i.e., fatty acids). Aside from the bioaccumulation work by van den Heuvel et al. (2014), literature searches yielded no other studies employing either of these methods, possibly due to subsequent advancement of a superior approach. The most sophisticated quantification method developed to date is an HPLC-UHRMS method (Pereira et al., 2013; Pereira and Martin, 2015). This method can be used with or without acid extraction of water samples3 and can be conducted using either negative ionization (i.e., for detection of organic acids) or positive ionization (i.e., for detection of organic bases and neutral polar compounds) techniques. Thus, the complex mixture compounds represented by the HPLC-UHRMS data are referred to hereafter as OSPW ionizable dissolved organics (OSPW-IDO). OSPW-IDO are described by the empirical formula CxHyOzSaNb, where subscripts represent the number of atoms, and can be sorted into isomer classes based on the number of sulfur, nitrogen and oxygen atoms in the detected species and the mode of ionization. For example, the notation ‘‘SO-2” describes an isomer class of dissolved organics containing one sulfur and two oxygen atoms plus variable carbon and hydrogen atoms, detected using negative ionization (Pereira et al., 2013). Each OSPW-IDO isomer class consists of multiple chemical species with different structures that can not be distinguished using the HPLCUHRMS method or any other technique available to date. However, this method has been used to detect and quantify thousands of 3 Tissue samples still require homogenization and extraction procedures. An internal standard is used to determine and compensate for recovery efficiency.

van den Heuvel et al., 2014 Zhang et al., 2016

OSPW-IDO including NAs, permitting more comprehensive investigation of the bioaccumulation potential of this fraction of OSPW (Zhang et al. 2015; 2016). The main drawback of this technique is the limited availability of the required analytical instrumentation. The reliability and interpretation of bioaccumulation results is directly correlated with the accuracy/precision and analytical specificity of the quantification techniques used in such investigations. Thus, analytical methodology is an important consideration when evaluating reported results for the purposes of reclamation planning and risk assessment. The adequacy and limitations of methods for characterizing OSPW dissolved organics, particularly the AEO fraction, have been reviewed in detail previously (Headley et al 2013a; Mahaffey and Dubé, 2017), but we have included above a summary of techniques pertinent to bioaccumulation studies. Although initial approaches provided estimates of BCFs based on accepted methods of the time, the ability of the most recent HPLC-UHRMS method to provide quantitative data for groups of chemical species representing the entire profile of OSPW-IDO is a vast improvement over early techniques that relied on single-ion monitoring, low resolution mass spectrometry, or other proxy methods for estimating NAs. 3. Bioaccumulation studies Published in vivo experiments that assessed the bioaccumulation potential of OSPW-IDO are summarized in Table 2. Earlier studies focussed on classical NAs with some studies using the commercially-available Refined Merichem NAs preparation as a surrogate for OSPW NAs. For reference, Fig. 1 shows the profile of NAs in Refined Merichem and in a sample of OSPW from an unspecified tailings pond. The range of carbon numbers and the relative distribution of Z-values (i.e., double-bond equivalents) are similar. However, current analytical methods are not able to completely resolve individual structures and there is evidence that NAs from different sources can have different properties in terms of their toxicity and susceptibility to biodegradation (reviewed by Kindzierski et al., 2012). A summary of the analytical methods used to measure NAs concentrations in extracts of tissues or fluids from organisms used in exposure experiments is given in Table 1 and discussed in the previous section. Laboratory simulations combined with QSAR modelling have also been used to assess bioaccumulation potential of NAs and other OSPW-IDO. No studies have evaluated the potential for biomagnification of OSPW-IDO. 3.1. Investigations of the potential for bioaccumulation of NAs in fish As noted in section 2, initial bioaccumulation studies targeted NAs due to the limitations of analytical methodology as well as established concerns regarding NAs toxicity. Three different exposure scenarios were tested using rainbow trout in 96-h static renewal experiments: (1) spiking aquaria water with NAs of com-

Please cite this article as: A. C. Scott, W. Zubot, C. W. Davis et al., Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134558

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Table 2 Summary of in vivo bioaccumulation studies that were conducted as laboratory static renewal tests (References 1–4, and 6) or in situ by introducing fish into a test pond with OSPW (Reference 5). Reference

Test organism

Source of dissolved organics

Sample type

Analysis basis

Bioaccumulation resultsa

% Relative Standard Deviation

1) Young et al., 2007

Rainbow trout (Oncorhynchus mykiss) Rainbow trout (O. mykiss) Rainbow trout (O. mykiss)

OSPW from Syncrude Pond 9, collected May 1, 2006 Refined Merichem NAs Refined Merichem NAs

whole fish

Single ion monitoring of C13H22O2 isomer class by gas chromatographymass spectrometry (GC–MS) Single ion monitoring of C13H22O2 isomer class by GC–MS Single ion monitoring of C13H22O2 isomer class by GC–MS

Positive detection of NAs in fish tissues but no quantitative results BCF = 2

n/ab

Northern leopard frog (Lithabates pipiens) Yellow perch (Perca flavescens)

Refined Merichem NAs

BCF BCF BCF BCF

18 73 61 8

2) Young et al., 2008 3) Young et al., 2011

4) Smits et al., 2012

5) van den Heuvel et al., 2014

6) Zhang et al., 2016

a b

Japanese medaka (Oryzias latipes)

whole fish Muscle gills liver muscle

Single ion monitoring of C13H22O2 isomer class by GC–MS

OSPW in (1) Syncrude Demonstration Pond, and (2) Syncrude South Bison Pond

whole fish

Single ion monitoring of C13H22O2 isomer class by GC–MS

bile

OSPW from Syncrude Base Mine Lake, collected July 15, 2014

whole fish

Single ion monitoring of C13H22O2 isomer class by GC–MS and liquid chromatography-high resolution mass spectrometry monitoring of seven ions and High performance liquid chromatography (HPLC) fluorescence HPLC-ultrahigh resolution mass spectrometry monitoring all ionizable dissolved organics (dominant species: R-O-2, R-SO+, R-NO+)

4 20 31 0.24

Positive detection of NAs in tissues from fish exposed to source (1); not detected in fish exposed to source (2) Positive detection of NAs in bile from all exposed fish, suggesting biliary excretion of NAs

n/a

BCF range = 0.6 to 53

<20%

BCF of C13H22O2 isomer group = 4.4

not given

n/a

BCF units are L/kg wet weight; a substance is considered bioaccumulative if it has a BCF  5000 (CEPA, 1999). n/a = not applicable

mercial origin (Merichem, 3 mg NAs/L), (2) using undiluted OSPW directly (15 mg NAs/L), and (3) feeding fish food spiked with Merichem NAs (1.5 mg NAs/g of food) (Young et al., 2007). NAs were present in the tissues of all fish analyzed after exposure to NAs regardless of the source or the mode of exposure. The detection method was not quantitative, therefore, BCFs could not be determined from the data. However, the findings indicate that NAs uptake by fish occurs relatively rapidly and that NAs remain in fish tissue under short-term exposure conditions. Analysis of 23 wild fish (four species) captured from aquatic locations near the Athabasca oil sands 4showed NAs concentrations in the range of 0.2 to 2.8 mg NAs/kg fish tissue in five of the fish (22% of samples), with the remainder of fish having concentrations below the detection limit of the method (0.1 mg NAs/kg fish tissue) (Young et al., 2008). Positive detection of NAs was not species-dependent. These results show that wild fish exposed to regional waters containing naturally-occurring NAs (Sun et al., 2017) can take up and store NAs in their tissues. Unfortunately, the study by Young et al. (2008) did not report the NAs concentrations in the waters where these wild fish were collected so it is impossible to know whether all the fish were actually exposed to NAs or if only the five fish that showed NAs in their tissues contacted water with NAs. The absence of quantitative data for NAs in the waters also precludes calculation of BCFs. In the same study by Young et al. (2008), fish exposure experiments were conducted using rainbow trout (O. mykiss) in 96-h static renewal tests with the commercially-available Merichem preparation (i.e., real OSPW was not tested). Aquaria waters were maintained at pH 8.2 (approximate pH of the Athabasca river; Athabasca Watershed Council, 2018). Controlling pH is critical because it governs water solubility and potential partitioning behavior for ionizable compounds (Franco and Trapp, 2000). Results showed that fish from aquaria with 3 mg NAs/L had a mean 4

= = = =

34

Sites included the Athabasca River, Lake Athabasca, and the Clearwater River.

concentration of 5.6 ± 1.9 mg NAs/kg fish tissue between days 2 and 9 of the experiment. This corresponds to a BCF of approximately 2 L/kg wet weight, which is not considered bioaccumulative by CEPA standards (CEPA, 1999). The authors made comparison to literature values of BCFs for resin acids in rainbow trout, noting that the reported range of < 25 to 182 was much higher than their measured BCF of ~ 2 for NAs, despite the structural similarity of resin acids to NAs (the former are monocarboxylic acids but have C–C double bonds). Nonetheless, these results show that NAs can be detected in fish tissue following exposure to aqueous solutions of NAs. On day 4 of the exposure tests, some fish were transferred to aquaria with clean water (no NAs) where extensive depuration was observed. Specifically, after 2 days the average tissue concentration had dropped significantly (p < 0.05) to 0.1 ± 0.2 mg NAs/kg, corresponding to a 95% decrease. These results indicate that the uptake and incorporation of NAs in fish tissues is not permanent, provided exposure to NAs ceases for some time. It should be noted that metabolism was not investigated so it is not possible to tell whether excretion/depuration were exclusively responsible for the observed decreases in NAs. In a subsequent laboratory study, the distribution of NAs in various fish tissues was assessed (Young et al., 2011). The pH in these experiments was 7.6; no reason was given for operating at a lower aquaria pH than in their previous work but pKa’s of NAs are approximately 5, so the expected impact on BCFs would not be significant. Gills and livers of fish exposed to Merichem NAs had statistically higher concentrations of NAs than muscles (p < 0.05), whereas livers had higher concentrations than hearts. NAs were also found in the eggs of fish used in the exposure study. Based on average NAs concentration in fish muscles and aquaria water, the computed BCF is approximately 4. BCFs for fish gills and fish liver were 20 and 31, respectively. However, the accuracy of these two BCF values is questionable considering the very high percent relative standard deviation (%RSD) of the data for mean NAs con-

Please cite this article as: A. C. Scott, W. Zubot, C. W. Davis et al., Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134558

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Fig. 1. Example profiles of NAs in Refined Merichem (a) and an OSPW sample (b) where Intensity represents the peak area of each isomer class corresponding to an R-O-2 species as measured by ultra high performance liquid chromatography ion mobility time-of-flight mass spectrometry. (Reprinted with permission from Huang et al. (2015).)

centrations in these tissues (%RSD = 73 for gills and 61 for livers). A larger sample size could improve precision but BCFs are still far below the bioaccumulative designation threshold BCF value of 5000 (CEPA, 1999). Exposure studies have also been conducted using yellow perch (Perca flavescens), another fish species that is native to the Athabasca region (van den Heuvel et al., 2014). Test fish were introduced to two experimental ponds containing OSPW with NAs

concentrations of 4.6 and 13 mg/L, respectively, and one freshwater reservoir with a NAs concentration of 1.0 mg/L, as measured by high resolution mass spectrometry. The three sites were located on the Syncrude Canada Ltd. lease with the reservoir holding Athabasca River water for plant use. After four months, test fish were collected along with fish from two reference lakes located nearby. Using similar extraction and analysis methods to those of Young et al. (2008), it was shown that NAs from OSPW did not concen-

Please cite this article as: A. C. Scott, W. Zubot, C. W. Davis et al., Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134558

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trate in the tissues of the test fish and were in fact only detected in fish exposed to the OSPW with the highest concentration of NAs (i.e., 13 mg/L). Interestingly, NAs were detected in approximately a third of the fish samples from the reference lakes despite NAs being undetectable in samples of the lake waters. This result was not addressed by the authors, although they did express surprise that NAs were not detected more frequently in fish samples from the actual test ponds considering the results of a previous study where fish were exposed to OSPW (Young et al., 2008). Additionally, the study by van den Heuvel et al. (2014) examined NAs accumulation in bile of the test fish and reference fish. Conflicting results were obtained depending on the analysis technique employed (refer to section 2 of this review), but it was evident that NAs excretion by fish can occur via a biliary route. Compared to reference fish, bile NAs concentrations were significantly higher (p < 0.05) in exposed fish samples despite large variability in the data within each location (i.e., %RSDs ranged between 29% and 189%). Despite analytical measurement challenges, the study findings indicate NAs are susceptible to biochemical removal processes which mitigate bioaccumulation potential within organisms. 3.2. Development and evaluation of surrogate techniques to assess bioaccumulation potential of OSPW-IDO An alternative approach for studying bioaccumulation potential in vivo is to measure the equilibrium partitioning of a chemical between water and a hydrophobic material such as 1-octanol or polydimethylsiloxane (PDMS), which simulates the storage lipids of organisms. A distribution coefficient (DOW or DPDMS, respectively) can then be calculated and its logarithm (log DOW or log DPDMS) interpreted similarly to values for BCF obtained from in vivo exposure studies5. Employing this so-called ‘‘biomimetic” approach, filtered OSPW samples (Syncrude WIP) were incubated with a solid support coated with PDMS (Zhang et al., 2015). Samples were then analyzed by the recently developed HPLC-UHRMS method capable of characterizing OSPW-IDO in terms of several isomer classes6 including NAs (i.e., O–2 species), mono- and polyoxygenated species (Ox), nitrogen- (N) and sulfur- (S) containing species, and a variety of mixed species (i.e., NOx, SOx, NOxS)7 (Pereira et al., 2013; Pereira and Martin, 2015). Using negative mode, it was found that > 95% of OSPW-IDO remained in the OSPW rather than partitioning to the PDMS. The log DOW values estimated from measured DPDMS values ranged from 4.2 to 2.0 for all species detected in negative mode (Table 3). For NAs (i.e., O–2 species), log DOWs were generally quite low, but did range up to 2.0 in some cases, tending to increase with the number of carbons and decrease with double-bond equivalents (i.e., higher Z values). The same trend was observed for OSPW-IDO detected in positive mode, but log DOW ranged up to 5.3 (Table 3) and nearly one quarter of all detected species were found at higher concentration in the PDMS than in the OSPW at equilibrium. Specifically, species belonging to the nonacidic NO+ and SO+ heteroatomic isomer classes had mean log DOW values of 2.5 ± 1.5 and 1.0 ± 1.7, respectively, with maximum log DOW of about 5. 5 A log DOW or log DPDMS > 1.0 indicates a preference for the hydrophobic material; higher values indicate greater potential to bioaccumulate. 6 The method does not permit elucidation of individual chemical structures but allows identification of empirical formula classes. The present review follows the notation used in the original publication wherein only the functional group of each empirical formula class is used to refer to it but note that there is a hydrocarbon portion to each of these compounds. The hydrocarbon group is denoted with an ‘‘R” in Tables 3 and 4. 7 Each species contains a mixture of structural isomers that can not be resolved into pure compounds using this method or any other technique currently available. At this time, the extent of overlap between species that are detected in both positive and negative ionization modes is not known.

7

Simulating bioaccumulation of OSPW-IDO circumvents the challenges of selectively extracting and measuring an extremely complex mixture in the presence of confounding background matrix effects. However, the authors recognized that their study represented only an estimate of bioconcentration potentials and that values they obtained for the DPDMS of organic acids such as NAs were lower than would be expected in vivo because these compounds could more-readily partition to polar materials such as the phospholipid bilayer of cells (PDMS is a non-polar surrogate for biological storage lipids), especially under the predicted environmental conditions. Further, biomimetic approaches only reflect passive diffusion whereas the additional processes of metabolism and active uptake/elimination can impact bioconcentration in vivo. A subsequent study by Zhang et al. (2016) reported another biomimetic assessment of the bioaccumulation potential for OSPWIDO and compared the results to those from in vivo fish exposure experiments conducted in parallel. For the biomimetic experiments, silica beads coated with synthetic phospholipid bilayer (phosphatidylcholine) were incubated with aqueous solutions prepared from extracts of an OSPW samples (Syncrude Base Mine Lake) to estimate the membrane-water distribution ratio (i.e., log DMW) for OSPW-IDO. Analysis by HPLC-UHRMS was performed in both negative and positive mode as per previous work. The results of this study are summarized in Table 3. In negative mode, about 14% of all organic compounds detected had log DMW > 1.0 and were identified as belonging to either the O–2 or SO-2 heteroatomic classes8. The highest measured log DMW value for the O–2 species (NAs) was 4.1. Comparison of this result to those from their previous study (Zhang et al., 2015) clearly showed that OSPW-IDO partition more readily to polar material than to PDMS, suggesting that the PDMS biomimetic method underestimates bioconcentration potential because it does not represent the full suite of molecular interactions that occur at the lipid-water interface. Electrostatic interactions between OSPW-IDO and charged membrane constituents were thought to be responsible for the higher partitioning observed with the simulated membrane as opposed to the simulated storage lipids. For the SO-2 heteroatomic class, log DMW ranged up to 3.4. In positive ionization mode, only 9.1% of all chemical species detected had log DMW > 1.0. These included the polar neutral isomer classes of O+, O+2, NO+, and SO+ with maximum log DMW values of 3.0, 2.5, 4.1, and 4.1, respectively. For all isomer classes, partitioning increased with higher carbon numbers and decreased with double-bond equivalents, reinforcing the general trend observed previously (Zhang et al., 2015). In the next phase of this study, BCFs for OSPW-IDO were computed and then compared to results from in vivo exposure experiments (Zhang et al., 2016). A QSAR model for BCF prediction was adapted for calculation of log BCFs using log DMW and log DOW data from both the current and previous study (i.e., Zhang et al. 2015). From these values, the predicted BCF ranges were calculated (Table 4). Only the polar neutral SO+ and NO+ isomer classes were predicted to be bioaccumulative (i.e., BCF  5000). The predicted BCF range of the acidic O–2 isomer class was 0.8 to 126. Fish exposure studies were conducted using juvenile Japanese medaka (Oryzias latipes) in aquaria containing a 10-fold dilution of filtered OSPW from Syncrude’s Base Mine Lake. Aquaria pH was approximately 8.2, which is comparable to the Athabasca River system (Athabasca Watershed Council, 2018). A static renewal approach similar to that of previous fish exposure studies was used. Samples of fish and water were taken at several timepoints throughout the 12-d exposure and newly developed extraction and analysis methods were employed for quantification of contaminant compounds in the fish tissue (note: water samples were not extracted but were 8 A log DOW or log DMW > 1.0 indicates a preference for the more hydrophobic material (i.e., membrane lipid); higher values indicate greater potential to bioaccumulate.

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Table 3 Summary of bioaccumulation simulations conducted by Zhang et al. 2015 (A) and 2016 (B). Bioaccumulation assessment method A

B

Partitioning to polydimethylsiloxane (PDMS)

Partitioning to phosphatidylcholine

Mass spectrometer ionization mode

Heteroatomic classes with log Da > 1.0

Maximum log Db

Negative Positive

R-O-2

2.0 5.1 5.3 4.1 3.4 3.0 2.5 4.1 4.1

Negative Positive

a b

R-NO+ R-SO+ R-O-2 R-SO-2 R-O+ R-O+2 R-NO+ R-SO+

D = distribution coefficient. for (A), values represent log DOW; for (B), values represent log DMW.

Table 4 BCF ranges predicted by QSAR modelling and BCF ranges measured by fish exposure experiments using Japanese medaka. Heteroatomic isomer class

Predicted BCF range (L/ kg wet weight)

Measured in vivo BCF range (L/kg wet weight)

R-O-2 R-SO-2 R-O+ R-SO+ R-O+2 R-NO+

0.8 0.8 0.8 0.8 0.8 0.8

0.7 to 53 not detecteda not detecteda 0.6 to 28 not detecteda 0.6 to 28

to to to to to to

126 25 126 63,096 1259 15,849

a signal for these compounds was below the detection limit of the HPLC-UHRMS method.

analyzed by the same method as the fish tissue extracts). Only the SO+, NO+ and O–2 isomer classes of OSPW-IDO were detected in the exposed fish. Uptake reached equilibrium after 72 h of exposure and qualitative interpretation of chromatograms obtained by HPLC Orbitrap suggested that some isomers within a given isomer class were more accumulative than others. The range of BCFs for the O–2 isomer class (i.e., classical NAs) was 0.7 to 53 L/kg, which correlated well with the predicted values (Table 4). BCF of ions with the empirical formula C13H22O2 was 4.4 L/kg, which is comparable to BCF values previously reported for this isomer group (Young et al., 2008, 2011). The maximum measured BCF for the SO+ and NO+ isomer classes was 28 L/kg, which is 2- to 100-fold lower than the maximum values predicted by the model (Table 4). Likely, biotransformation processes are involved in eliminating these groups of hydrophobic dissolved organics. None of the species detected in fish tissue had a BCF  5000, and as such, not are considered bioaccumulative by CEPA standards (CEPA, 1999). For the last phase of the study, Zhang et al. (2016) assessed depuration by transferring fish to clean water. Approximately 80% of the O–2 species (NAs) were rapidly depurated over the first 6 h, however, some isomers belonging to this class remained in the fish tissue after 144 h. Biological half-lives for the O–2 isomer class (NAs) ranged from 3.2 to 6.8 h based on data from the first day of depuration. The SO+ and NO+ species were also rapidly depurated over the first 6 h and decreased to below the detection limit before the end of the experiment. Estimated biological half-lives for SO+ and NO+ isomer class were < 3.3 h. When the logarithms of all measured BCFs were plotted against the logarithms of the predicted values, good correlation was observed for species detected in positive mode (p < 0.001, R2 = 0.835) and in negative mode (p < 0.001, R2 = 0.74). This confirms the validity of estimating BCFs from DOW and DMW measurements, suggesting future work could use this method or a comparable model to evaluate bioaccumulation potential for ionizable dissolved organics in OSPW of different origins/age, rather than employing more time-consuming and challenging in vivo approaches.

3.3. Investigation of NAs bioaccumulation potential in frogs In addition to fish, dissolved organics in OSPW also have the potential to impact amphibians, especially during reproductive and early life stages. Amphibians are often considered ‘‘sentinel species” and valuable indicators of ecological and ecosystem health. This is due to their position at the interface of the aquatic and terrestrial/riparian food chains as well as the unique physiology of amphibians (i.e., their highly permeable/absorptive skin) (Cooke, 1981; Storrs and Kiesecker, 2004). Several previous studies have evaluated the toxicological effects and mechanisms of OSPW on amphibians (particularly, NA species) (Pollet and BendellYoung, 2000; Gupta, 2009,). Consequently, deducing the potential of these species to bioaccumulate OSPW-IDO is informative for both aquatic and terrestrial food web systems. The Northern leopard frog (Lithabates pipiens), wood frog (Lithobates sylvatica), Boreal toad (Anaxyrus boreas), and Canadian toad (Anaxyrus hemiophrys) are amphibians native to the Athabasca region of Alberta (Alberta Conservation Association, 2018). A frog exposure study was conducted using Northern leopard frog in aquaria containing waters spiked with Merichem NAs under static renewal conditions (Smits et al., 2012). Employing the extraction and analytical techniques of Young et al. (2011), they were able to detect NAs in the frog muscle tissues albeit at very low concentrations. The reported BCF for the single C13H22O2 isomer group was 0.24 L/kg, indicating a low potential for bioaccumulation. However, the authors cautioned that their work may not represent the bioaccumulation potential of OSPW NAs because commercial preparations of NAs (e.g. Merichem) and OSPW NAs are known to have different chemical profiles. Furthermore, only one of the four amphibian species native to northeastern Alberta was included in this study with only adult frogs tested9. Nonetheless, these findings were consistent with previous work in aquatic organisms that suggested low bioaccumulation potential for the NA chemical class. 4. Discussion and conclusions Determining the potential environmental impacts of OSPW constituents is a high priority. Most of the published work to date has focussed on characterizing the chemical composition of OSPW as well as evaluating and/or mitigating its toxicity. A complex group of dissolved organic compounds, which include NAs, has consistently featured as the main target of these efforts. In addition to the attention given to toxicity of OSPW-IDO, some studies have examined the bioaccumulation potential of the various isomer 9 In a previous study (Hersikorn and Smits, 2011), wood frogs were raised in reclaimed wetlands containing OSPW and solid tailings 7 years old. Bioaccumulation was not evaluated but they did find that thyroid hormone disruption resulted in delayed or complete failure to initiate metamorphosis. Frogs exposed to reclaimed wetlands with older oil sands process materials did not present with adverse physiological or developmental effects.

Please cite this article as: A. C. Scott, W. Zubot, C. W. Davis et al., Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134558

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classes that comprise this fraction of OSPW. To facilitate method development and quantification efforts, most studies have used commercially available preparations as a surrogate for OSPW NAs, a major component of the OSPW-IDO fraction. Commercial preparations offer the convenience of being a pure source of NAs, but their composition and properties differ from OSPW. Recent advances in analytical chemistry, specifically the HPLC-UHRMS method, enables researchers to determine the fate of isomer classes representing all OSPW-IDO, rather than tracking single ions or representative substitutes. The studies reviewed herein indicate that BCFs for NAs and other OSPW-IDO including polar neutral substances are far less than the CEPA threshold for designation as bioaccumulative. This is consistent with physical–chemical properties of NAs and other acidic organic compounds, which typically exist in their ionized form at environmentally relevant pH values, leading to relatively low hydrophobicity (log Kow) and, consequently, low potential for bioaccumulation. In general, the propensity for chemical bioaccumulation increases as hydrophobicity (log Kow) increases (modulated by the compound’s ability to be metabolized (Arnot and Gobas, 2006; Arnot et al. 2008a; 2008b)), as has been welldemonstrated in BCF modeling literature (Neely et al., 1974; Veith et al., 1979; Mackay, 1982; Meylan et al., 1999; Arnot and Gobas, 2003). Nevertheless, the most hydrophobic substances in OSPW (i.e., the NO+ and SO+ isomer classes) were not found to be bioaccumulative in fish exposure studies. This is likely related to both their ability to be metabolised as well as bioavailability limitations for extremely hydrophobic compounds (i.e., log Kow > 10) (Arnot and Gobas, 2006). It should be noted that bioaccumulation can also correlate with a decrease in (sub-cooled liquid) water solubility but this is not always the case as active transport and/or bioavailabilty can modulate the uptake or elimination of the chemical within the organism (Mackay and Fraser, 2000). Previous studies have also identified that NAs are susceptible to physiological processes which result in the rapid depuration of compounds from the organism and the same phenomenon may also explain the observed discrepancy between predicted and observed BCFs for other OSPW-IDO. In any case, research activities to date indicate OSPW-IDO are not bioaccumulative and have yielded a biomimetic method (Zhang et al. 2016) that could be used to estimate bioaccumulation potential of OSPW-IDO from a variety of sources. CRediT authorship contribution statement Angela C. Scott: Visualization, Writing - original draft, Writing review & editing, Project administration. Warren Zubot: Conceptualization, Supervision, Writing - review & editing. Craig W. Davis: Writing - review & editing. John Brogly: Supervision, Writing - review & editing, Funding acquisition. Declaration of Competing Interest The authors declare that they have no known competing financial interests or personal relationships that could have appeared to influence the work reported in this paper. Acknowledgments This work was supported by funding from Canada’s Oil Sands Innovation Alliance (COSIA). Claire Jackson and Stephen Herman of Alberta WaterSMART are acknowledged for their insightful comments and suggestions during the early preparation stages of this review. The authors also thank Dr. Louise Camenzuli (ExxonMobil Petroleum & Chemical) for her helpful discussions.

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Please cite this article as: A. C. Scott, W. Zubot, C. W. Davis et al., Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134558

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Please cite this article as: A. C. Scott, W. Zubot, C. W. Davis et al., Bioaccumulation potential of naphthenic acids and other ionizable dissolved organics in oil sands process water (OSPW) – A review, Science of the Total Environment, https://doi.org/10.1016/j.scitotenv.2019.134558