Treatment of Polycyclic Aromatic Hydrocarbons in Oil Sands Process-Affected Water with a Surface Flow Treatment Wetland
Treatment of Polycyclic Aromatic Hydrocarbons in Oil Sands Process-Affected Water with a Surface...
Cancelli, Alexander M.;Gobas, Frank A. P. C.
2020-08-23 00:00:00
environments Article Treatment of Polycyclic Aromatic Hydrocarbons in Oil Sands Process-Aected Water with a Surface Flow Treatment Wetland Alexander M. Cancelli and Frank A. P. C. Gobas * The School of Resource and Environmental Management, Simon Fraser University, 8888 University Drive, Burnaby, BC V5A 1S6, Canada; alexander_cancelli@sfu.ca * Correspondence: gobas@sfu.ca Received: 26 June 2020; Accepted: 20 August 2020; Published: 23 August 2020 Abstract: This study applied a passive sampling approach using low-density polyethylene passive samplers to determine the treatment eciency of the Kearl surface flow treatment wetland for polycyclic aromatic hydrocarbons (PAHs) in Oil Sands Process-aected Waters (OSPW). Treatment eciency was measured as concentration-reduction and mass-removal from the OSPW. The results show that the wetland’s ability to remove individual PAHs from the influent varied substantially among the PAHs investigated. Treatment eciencies of individual PAHs ranged between essentially 0% for certain methylated PAHs (e.g., 2,6-dimethylnaphthalene) to 95% for fluoranthene. Treatment in the Kearl wetland reduced the combined total mass of all detected PAHs by 54 to 83%. This corresponded to a reduction in the concentration of total PAHs in OSPW of 56 to 82% with inflow concentrations of total PAHs ranging from 7.5 to 19.4 ng/L. The concentration of pyrene in water fell below water quality targets in the Muskeg River Interim Management Framework as a result of wetland treatment. The application of the passive samplers for toxicity assessment showed that in this study PAHs in both the influent and euent were not expected to cause acute toxicity. Passive sampling appeared to be a useful and cost-eective method for monitoring contaminants and for determining the treatment eciency of contaminants in the treatment wetland. Keywords: treatment wetlands; oil sands process-aected water; polycyclic aromatic hydrocarbons; passive sampling; toxicity 1. Introduction As demand for freshwater conservation grows, there is a need for sustainable solutions to manage and reuse process-aected waters. In Canada, a considerable volume of Oil Sand Process-aected Waters (OSPW) has been and continues to be generated during bitumen extraction, which contains an array of dierent organic and inorganic contaminants [1]. Polycyclic aromatic Hydrocarbons (PAHs) are commonly associated with OSPW and have been shown to be a potential source of OSPW toxicity [2–4]. Since OSPW is currently subject to a ‘zero discharge’ policy and few treatment options are available, OSPW is either recycled for further use in the extraction process or stored in euent tailings ponds. These euent tailings ponds are susceptible to leaching and erosion, and present adverse risks to migratory birds and wildlife that confuse these areas for safe ecological havens [5–8]. While eorts to develop feasible solutions for OSPW treatment are ongoing, few have been realized to date. Treatment wetlands have emerged as a potentially feasible option to treat OSPW [9–15]. Treatment wetlands are constructed, artificial ecosystems that harness the biogeochemistry of natural systems to reclaim and remediate contaminated land and water. Environments 2020, 7, 64; doi:10.3390/environments7090064 www.mdpi.com/journal/environments Environments 2020, 7, 64 2 of 16 The biogeochemical mechanisms for contaminant removal within a wetland include microbial and plant-mediated biotransformation, chemical transformations, UV degradation, evapotranspiration, and sorption to sediments [12]. The capability of wetlands to harness these biogeochemical processes for wastewater treatment has been demonstrated for a variety of wastewaters including municipal and domestic wastewaters, agricultural runo, pulp and paper wastewater, and waters that contain surfactants, solvents, or pesticides (e.g., [11,16–22]). While treatment has been demonstrated for a variety of wastewaters, many studies have reported treatment wetland performance based on general metrics for water quality such as Biochemical Oxygen Demand (BOD ), Total Nitrogen, Total Phosphorus, or Total Petroleum Hydrocarbons. Although useful as measures of overall water quality, these general metrics do not detail the removal of specific contaminants of concern from wastewaters. Quantifying the removal of individual contaminants by wetland treatment is critical to evaluate the toxicological risk associated with the influent OSPW and euent water and to identify which contaminants are more easily removed from wetlands and which contaminants are not, i.e., high vs. low treatment eciency. Information on the treatment eciency of engineered treatment wetlands is necessary for assessing the feasibility of treatment wetlands for their specific wastewater challenges. The objective of this study is to investigate the capacity of wetlands to treat PAHs in OSPW. Specifically, we investigate the treatment eciency of PAHs in terms of reductions in concentrations, mass loadings, and associated toxicity of PAHs in OSPW in the Kearl Treatment (KT) wetland. 2. Materials and Methods 2.1. Site Description The KT wetland is a free water surface-flow constructed wetland at the Kearl Oil Sands site 0 00 0 00 (approximately 75 km NNE of Fort McMurray, AB, Canada; 57 26 00 N, 111 8 31 W) managed by Imperial Oil Resources Ltd. The KT wetland operates in warmer summer months, typically from May to September. It was designed as a pilot-scale wetland to investigate the treatment of on-site Oil Sands Process-aected Water (OSPW). Water is pumped into the wetland at 5 L/s (430 m /day), resulting in a hydraulic retention time of approximately 14 days. OSPW refers to any water that comes in contact with oil sands or used in an oil sands processing facility. In 2017, OSPW was collected as runo from an overburden disposal area. This overburden disposal area contains stockpiles of excavated overburden from the mined areas at the Kearl Oil Sands site and contains residual amounts of organic contaminants. Since this OSPW is not sourced directly from a tailings pond, these residual contaminants largely consist of PAHs. Concentrations of naphthenic acids and dissolved solids in this OSPW were low, hence the focus on PAHs in this study. The runo is first directed to, and detained in, the north overburden disposal pond located next to the KT wetland which acts as an initial settling basin for suspended solids. Due to the ‘zero-discharge’ policy for all OSPW, the system operates as a closed-circuit and therefore water is recycled from the KT wetland back into the north overburden disposal pond. The outflow was controlled by a submerged pump that was triggered on when water depth reached 1.7 m, and o when water depth receded to 1.0 m in the final deep pool (Figure 1). In 2018, OSPW for the KT wetland was sourced from a drainage pond situated next to an euent tailings area at the Kearl Oil Sands. The OSPW was pumped to the wetland during a single pump event, where it was fully recycled within the wetland (i.e., no external detention pond was used) for the duration of the 2018 study. The wetland consists of six cells in series (3 deep pools, 3 shallow areas) with a longitudinal slope of 0.014%. Water percolates over shallow interior berms that separate adjacent cells. Shallow berms parallel to water flow in the middle of the shallow pool sections were constructed to improve directional water flow, and provide access for wetland monitoring (e.g., vegetation, erosion, water quality). The KT wetland operates at a total volume of water of approximately 6000 m . The deep pools (forebay, deep pool 1, and final deep pool) operate at a depth of 1.7 m, and are dominated by submerged vegetation, but contain a band of emergent vegetation (approximately 5% of area) along the perimeter of the cells Environments 2020, 7, 64 3 of 16 Environments 2020, 7, x FOR PEER REVIEW 3 of 16 where water depth is shallower. The shallow area cells are densely vegetated with a variety of dierent with a variety of different species dominated by common cattails (Typha latifolia) and water sedge species dominated by common cattails (Typha latifolia) and water sedge (Carex aquatilis). (Carex aquatilis). The rooting medium consists of 200 mm of compacted peat soil underneath 300 mm of The rooting medium consists of 200 mm of compacted peat soil underneath 300 mm of non- non-compacted peat soil from the displaced muskeg originally found on location. The peat was placed compacted peat soil from the displaced muskeg originally found on location. The peat was placed over a geosynthetic clay liner that was blanketed with a non-woven geotextile (Figure 1). over a geosynthetic clay liner that was blanketed with a non-woven geotextile (Figure 1). (a) 1.7 m 0.4 m Buoy 300mm non-compacted peat Polyethylene passive Geosynthetic sampling devices clay liner 200mm compacted peat Anchor (b) Figure 1. Schematic diagram of the Kearl Treatment Wetland showing (a) planar view, and (b) cross- Figure 1. Schematic diagram of the Kearl Treatment Wetland showing (a) planar view, and (b) sectional view with the passive sampling devices and rooting medium. Photo shows polyethylene cross-sectional view with the passive sampling devices and rooting medium. Photo shows polyethylene passive samplers in aluminium mesh casings connected to buoys. passive samplers in aluminium mesh casings connected to buoys. 2.2. Water Quality Monitoring 2.2. Water Quality Monitoring Biochemical Oxygen Demand (BOD ), conductivity, Dissolved Inorganic Carbon (DIC), Biochemical Oxygen Demand (BOD5), conductivity, Dissolved Inorganic Carbon (DIC), Dissolved Oxygen (DO), Dissolved Organic Carbon (DOC), pH, Total Dissolved Solids (TDS), Total Dissolved Oxygen (DO), Dissolved Organic Carbon (DOC), pH, Total Dissolved Solids (TDS), Total Suspended Solids (TSS), turbidity, and water temperature (T ) were collected for the OSPW in the water Suspended Solids (TSS), turbidity, and water temperature (Twater) were collected for the OSPW in the wetland to evaluate changes in the quality of OSPW as it passes through the KT wetland. In the 2017 wetland to evaluate changes in the quality of OSPW as it passes through the KT wetland. In the 2017 study, samples were obtained from the forebay and final deep pool on 13 July and 17 August 2017. study, samples were obtained from the forebay and final deep pool on 13 July and 17 August 2017. In the 2018 campaign, OSPW was fully recycled through the wetland, therefore samples were obtained In the 2018 campaign, OSPW was fully recycled through the wetland, therefore samples were from the forebay on both 26 August and 19 September 2018. Water samples were analysed by Maxxam obtained from the forebay on both 26 August and 19 September 2018. Water samples were analysed Analytics (Calgary, AB, Canada) for BOD , Dissolved Inorganic Carbon (DIC), Dissolved Organic by Maxxam Analytics (Calgary, AB, Canada) for BOD5, Dissolved Inorganic Carbon (DIC), Dissolved Carbon (DOC), Total Dissolved Solids (TDS), Total Suspended Solids (TSS). Field measurements were Organic Carbon (DOC), Total Dissolved Solids (TDS), Total Suspended Solids (TSS). Field collected by WorleyParsons Ltd. (Calgary, AB, Canada) for conductivity, Dissolved Oxygen (DO), pH, measurements were collected by WorleyParsons Ltd. (Calgary, AB, Canada) for conductivity, and water temperature using a YSI Professional Plus Multiparameter instrument, and turbidity was Dissolved Oxygen (DO), pH, and water temperature using a YSI Professional Plus Multiparameter ® TM measured using an Orbeco-Hellige TB200 Turbidimeter. ® TM instrument, and turbidity was measured using an Orbeco-Hellige TB200 Turbidimeter. 2.3. Passive Sampling 2.3. Passive Sampling Low-density polyethylene (PE) passive samplers were deployed in triplicate (n = 3) in the Low-density polyethylene (PE) passive samplers were deployed in triplicate (n = 3) in the forebay and final deep pool of the KT wetland. The deeper cells were chosen to ensure samplers forebay and final deep pool of the KT wetland. The deeper cells were chosen to ensure samplers were were deployed within the water column to measure dissolved contaminants in the water. The PE deployed within the water column to measure dissolved contaminants in the water. The PE strips strips (12.70 cm 15.24 cm, 25 m thickness, 0.5 g) were deployed in stainless steel mesh casings, (12.70 cm × 15.24 cm, 25 µ m thickness, 0.5 g) were deployed in stainless steel mesh casings, and and attached to an anchor-buoy system to allow for deployment at the centre of the deep cells, and to attached to an anchor-buoy system to allow for deployment at the centre of the deep cells, and to ensure the PE strips were submerged at approximately 0.3–0.5 m depths below the water surface. ensure the PE strips were submerged at approximately 0.3–0.5 m depths below the water surface. Three deployments of passive samplers occurred, beginning on (1) 21 July 2017 (final deep pool) and Environments 2020, 7, 64 4 of 16 Three deployments of passive samplers occurred, beginning on (1) 21 July 2017 (final deep pool) and 22 July 2017 (forebay) for a 37 day and 36 day deployment, respectively, (2) 28 August 2017 (forebay and final deep pool) for a 31 day deployment, and (3) 25 August 2018 (forebay) and 8 September 2018 (final deep pool) for 14 day deployments each. Since water was fully recycled through the KT wetland in 2018, the deployments in the forebay and final deep pool were done consecutively, i.e., samplers in the forebay were deployed from days 0–14, and samplers in the final deep pool were deployed from days 14–28. The PE samplers were prepared and analysed by SGS Axys Analytical Services Ltd. (Sidney, BC, Canada). Samplers were stored and shipped in aluminium foil, sealed in a plastic bag, and shipped in a cooler with ice packs to maintain a temperature at <4 C. The method of analysis of PE sampler devices follows USEPA Methods 1625B and 8270C/D. Instrumental analysis was performed by low-resolution mass spectrometry (LRMS) using an RTX-5 capillary GC column, which operates at a unit mass resolution in the electron impacts (EI) ionisation mode using multiple ion detection (MID), acquiring at least one characteristic ion for each target analyte and surrogate standard. Quantification of target analytes was performed using the isotope dilution method, and calculations were carried out using HP EnviroQuant and Prolab MS-Extended for targeting and quantification. Sample detection limits are available in the Supplementary Information (Tables S6–S8). The reported concentrations provided by the laboratory were issued as units of mass of chemical per gram of polyethylene passive sampler (i.e., C , ng/g). The partitioning behaviour of each chemical PE between the OSPW and the PE sampler was estimated using the calibration equation for PAHs reported by Lohmann [23]: Log K = 1.22 (0.046 SE)log K 1.22 (0.24 SE) (R = 0.92; SE 0.27) (1) PE-W OW This relationship correlates the polyethylene–water partitioning coecient (K ), with the PE-W octanol–water partition coecient (K ) for each test chemical. The log K for each PAH was OW OW obtained from EPISuite v4.11 [24]. Two field blanks per deployment consisting of clean polyethylene sheets were exposed to ambient air during sampler deployment and collection. Field blanks were wrapped in aluminium foil, sealed in a plastic freezer bag, and refrigerated at <4 C between use. The average concentration of PAHs in the field blanks (C ) were subtracted from the concentrations of PAHs measured in the F.Blank,i deployed PE samplers (C ) to account for background exposure, i.e., C* = C C [25]. PE,i PE,i PE,i F.Blank,i Concentrations were assumed to be negligible (i.e., C* = 0) if the mean concentration of the analyte PE,i found in the field blanks exceeded the concentration of that analyte measured in the deployed passive samplers (C > C ). F.Blank,i PE,i Two performance reference compounds (PRCs) were impregnated into the polyethylene passive 10 14 samplers during lab preparation: fluoranthene-d , and dibenzo(a,h)anthracene-d . These deuterated PRCs were used to evaluate the state of equilibrium between the water and PE by comparing day zero concentrations (C ) to final concentrations (C ) of the PRCs in the samplers after the deployment PRC,0 PRC,t period (t, days). The mass transfer coecients (k , d ) of the two PRCs were determined using: PRC,0 1 k = ln (2) C t PRC,t To relate the depletion rate constant of the performance reference chemicals to the time to reach 95% equilibrium (i.e., 3/k ) between the water and the passive sampler for the target analytes, a linear relationship was developed between log k and log K of the performance reference chemicals. e OW This relationship was then used to determine k from the log K of each target chemical i [26]. e,i OW Environments 2020, 7, 64 5 of 16 This k was then used to account for the lack of achieving equilibrium within the sampling duration e,i by calculating the dissolved concentration (C ) of each target chemical i in water as: WD PE,i C = (3) WD,i k t e,i K 1 e PE W,i Standard errors (SE) of C were derived through error propagation as: WD ! ! 2 2 C C C WD WD WD SE = SE + SE + SE (4) C K k C e WD PE W PE,i C K k PE W e PE where the SE of C* was determined for each chemical from the measured concentrations of the PE,i target analytes in multiple passive samplers; the SE of K was estimated by applying the delta PE-W method to the log-transformed linear regression equation (Equation (1)); the SE of k was determined e,i from measurements of PRC concentrations in the PE samplers. 2.4. Data Analysis Analytes were included in the data analysis if concentrations in the passive samplers exceeded the method detection limit (DL) in at least two of three replicates. Concentrations of PAHs in water below the DL were assigned a concentration equal to one-half of the chemical’s DL (i.e., C = DL/2). This was applied to all final deep pool concentration measurements for 7-methylbenzo[a]pyrene in deployment one, one forebay concentration measurement for 2,6-dimethylnaphthalene in deployment two, and two final deep pool concentration measurements for acenaphthene in deployment three. The great majority of measured concentrations exceeded the DL. For these three substances, a range of average concentrations is provided to reflect the lower estimate (i.e., assuming concentration is zero) and upper estimate (i.e., assuming concentration is equal to the DL). A two-sample t-test assuming unequal variances was performed in JMP , Version 13.1.0 [27] to detect statistical dierences between mean dissolved aqueous concentrations measured in the forebay and final deep pool ( = 0.05). Unequal variances in concentration measurements in the forebay and final deep pool were detected with a two-sided F-test (p < 0.001). 2.5. Wetland Treatment Performance Evaluation 2.5.1. Concentration-Reduction Changes to the concentration of test chemicals in the water (E ) were derived from the C,i concentration of each test chemical (i) in the passive samplers deployed in the influent wastewater eq.in eq.out (forebay; C ) and treated euent (final deep pool; C ) as: PE,i PE, i 0 1 eq.out B C C B C PE, i B C B C E = 1 100 (5) B C C,i B C eq.in @ A PE,i E was estimated directly from the measured concentration of the target analyte i in the passive C,i sampler (C ) to reduce error associated with converting concentrations in passive samplers to those PE,i in water. To correct for the dierent deployment durations of the samplers in forebay and final deep pool, equilibrium concentrations in the PE samplers were estimated using measured k values with: e,i eq. PE,i C = (6) PE,i k t e,i 1 e Environments 2020, 7, 64 6 of 16 2.5.2. Mass Removal Mass removal of PAHs from the wetland was expressed in terms of a mass-loading removal eciency (E ) for each chemical i. E was determined from the concentrations of freely dissolved L,i L,i in out analyte i in the influent (C ) and euent (C ) water, and the corresponding volumetric flow WD,i WD,i rates (L/day) of water entering the forebay (Q ) and leaving from the final deep pool (Q ) of the in out KT wetland: 0 1 out B Q C C out B C WD, i B C B C E = 1 100 (7) L,i B C @ in A Q C in WD, i Q in the KT wetland was controlled and maintained at 432,000 L/day (5 L/s) for the duration in of the study. Q was estimated from the water budget: Q = Q + Q Q , where Q is the out out in P ET P precipitation rate (L/day) and Q is the evapotranspiration rate (L/day) in the KT wetland. The mass ET removal eciency E expresses the removal of dissolved contaminant mass from OSPW. E diers L,i L,i from E in that it accounts for the eects of changes in the volume of water in the wetland due to C,i precipitation and evapotranspiration on concentration of the target chemical i in water. However, it should be stressed that because passive samplers measure only the concentration of dissolved contaminants in the influent and euent, E may not account for all mass of PAHs removed from the L,i wetland, which includes both dissolved and undissolved (sorbed) PAHs. Precipitation, temperature, and relative humidity data were obtained from historical records of the Fort McMurray, AB, Canada weather station available from Alberta Climate Information Services [28]. Total precipitation (P) was 56.0 mm, 38.9 mm, and 50.3 mm during each of the three deployment periods, respectively. Temperature and relative humidity were used to estimate daily evapotranspiration rates from the KT wetland using the Penman–Monteith equation. Evapotranspiration (ET) at the KT wetland was estimated to be 197 mm, 91.4 mm, and 52.9 mm during each of the three deployment periods, respectively. The volumetric rate of precipitation (Q ) and evapotranspiration (Q ) were calculated P ET using the total catchment area and surface area of wetland cells, respectively (i.e., Q = SA P; P catchment Q = S(SA )ET). The total catchment area (SA = 15,264 m ) included everything within ET cells catchment the external berms of the KT wetland. All precipitation within this area was assumed to enter the wetland as runo. The total surface area of all wetland cells (S(SA ) = 7894.6 m ) was estimated at cells the operating water levels, i.e., 1.7 m for deep pools and 0.4 m for shallow cells. 2.5.3. Toxicity The change in OSPW toxicity was estimated using the chemical activity approach [29–31]. Chemical activity (a; unitless) is a thermodynamic quantity related to fugacity and chemical potential which for dilute solutions can be expressed as the ratio of the chemical’s concentration (C; e.g., mol/m ) to the chemical’s solubility in the same media (S; e.g., mol/m ), i.e., a = C/S. The application of chemical activity to assess toxicity for neutral hydrophobic organic chemicals has merit for two main reasons. First, when equilibrated, concentrations of chemicals in dierent media (e.g., the passive sampler, water and organisms in the water) exhibit similar chemical activities. Hence, when at equilibrium, contaminant concentrations in the passive samplers reveal the chemical activity of contaminants in the water and biota that reside in the wetland, or that may be exposed to the influent or euent of the wetland. Unless the contaminants are biomagnified in the food-web, the concentration of the contaminants in the organisms exposed to the wetland water will at most approach the chemical activities in the water and the passive samplers. Biotransformation in the organisms can reduce the chemical activities of the parent compound below the chemical activity in the water. Secondly, studies have shown that a combined chemical activity of PAHs between 0.01 and 0.1 causes acute toxic eects through a mode of toxic action referred to as non-polar narcosis [32–34]. Hence, by converting concentrations of contaminants into chemical activities of contaminants, it is possible to assess whether acute toxic eects can be expected. This makes the chemical activity of PAHs in the water and passive samplers a useful metric for toxicity assessment of individual and mixtures of PAHs. Environments 2020, 7, 64 7 of 16 Chemical activity (a ) of each of the detected PAHs in water entering and leaving the KT wetland was estimated from the concentration of PAHs in the polyethylene samplers (C ) and the solubility of PE,i eq. each PAH in the polyethylene sheets (S ) as a = C /S . S was determined as K S PE,i i PE,i PE,i PE-W,i water,i PE,i where S is the solubility of chemical i in water at 25 C (reported in [20]). The summation of water,i chemical activities for each individual PAH produces a total chemical activity of all analytes present in in out the OSPW influent (S a ) and euent (S a ) of the KT wetland. By comparing the total chemical PE,i PE,i activities of PAH mixture in the water entering and leaving the KT wetland to the chemical activity threshold value for baseline toxicity (a = 0.01), it is possible to assess whether the influent or euent has the potential to be toxic to aquatic biota, and to what degree toxicity or toxicological risk has been reduced through wetland treatment. This approach can also be used for substances that exhibit a greater toxicity (or “excess toxicity”) than the baseline toxicity as long as the degree of excess toxicity is known from empirical toxicity studies or other methods. For such substances, a should be compared to 0.01/
, where
is the excess toxicity defined as toxicity greater than baseline toxicity (i.e.,
> 1). For determining the chemical activities of the PAHs in the PE samplers, changes in temperature during the two deployments were ignored and it was assumed that the mean temperature is adequate for determining the chemical activity of the PAHs. 3. Results and Discussion 3.1. Water Quality Table 1 shows that BOD , Dissolved Inorganic Carbon (DIC), Dissolved Oxygen (DO), Dissolved Organic Carbon (DOC), pH, Total Dissolved Solids (TDS), Total Suspended Solids (TSS), turbidity, and water temperature of OSPW in the wetland were below the upper limits of the water quality targets (WQTs) listed in [35]. Table 1. Biochemical Oxygen Demand (BOD ), conductivity, Dissolved Inorganic Carbon (DIC), Dissolved Oxygen (DO), Dissolved Organic Carbon (DOC), pH, Total Dissolved Solids (TDS), Total Suspended Solids (TSS), turbidity, and temperature (T ) of OSPW in the Kearl Treatment Wetland water during each deployment period. Deployment 1 Deployment 2 Deployment 3 2017A 2017B 2018 13-Jul 17-Aug 26-Aug 19-Sep Parameter Units Forebay FDP Forebay FDP Forebay Forebay WQT BOD mg/L <2.0 <2.0 <2.0 <2.0 — 2 2.4 S/cm Conductivity 811 795 955 919 1700 1700 799 (0.5%) DIC mg/L 74 75 74 75 46 52 mg/L d m DO 4.57 4.46 5.09 6.39 5.04 9.99 1.44 (0.2) DOC mg/L 18 18 18 18 20 20 63.1 pH pH 7.92 7.77 7.97 7.49 7.40 8.24 6.0–10.8 (0.2) TDS mg/L 680 680 680 680 860 1200 TSS mg/L <1.0 <1.0 <1.0 <1.0 12 1.6 82.2 NTU Turbidity 0.84 1.37 0.84 1.11 45 1.5 77 (2%) T C (SD) 20.7 (2.1) 20.6 (2.1) 10.7 (2.2) 25.3 water + m d Water Quality Target (peak target), [35]. <—below the reported detection limit. —minimum value. —data collected on 22-Sep-18. ——parameter not analysed. FDP—Final Deep Pool. ()—Instrument accuracy. DO in water remained constant throughout the wetland during deployment one (DO = 4.57 mg/L; DO = 4.46 mg/L) and increased upon passage through the wetland during forebay FDP deployments two (DO = 5.09 mg/L; DO = 6.39 mg/L) and three (DO = 5.04 mg/L; forebay FDP forebay Environments 2020, 7, 64 8 of 16 DO = 9.99 mg/L). The DO concentrations indicate that reoxygenation of the water occurs as it FDP passes over the internal berms between wetland cells. The average recorded water temperature in the KT wetland was 20.7 (SD 2.1) during deployment one, 20.6 (SD 2.1) during deployment two, and 10.7 (SD 2.2) during deployment three. In a study using a surface flow mesocosm wetland, temperature changes from 18.4 C to 11.3 C showed no significant eects on overall heterotrophic activity [36], possibly due to the diversity in microbial communities contributing to contaminant removal at dierent temperatures [37]. However, the eect on water temperature on heterotrophic activity in wetlands remains an area requiring further investigation. Warmer air temperatures were recorded during deployment one (T = 17.9 C; SD 2.7) compared avg. to deployments two (T = 12.2 C; SD 4.5) and three (T = 5.8 C; SD 4.3). Modelling work using avg. avg. a modified Penman–Monteith equation demonstrated that even a 3 C change in air temperature can induce a 14% change in potential evapotranspiration rates [38]. Therefore, dierences in air temperatures among the deployments are expected to have significant eects on evapotranspiration of water from the wetland. Estimated rates of evapotranspiration range between 2.5 to 9.8, 0.6 to 5.8, 1.0 to 3.1 mm/day and total evapotranspiration was 53.7, 24.9, and 14.4 m /day in deployments one, two, and three, respectively (Table S4). The conductivity of OSPW ranged from 795 to 811 S/cm in deployment one, 919 to 955 S/cm in deployment two and was 1700 S/cm in deployment three and appeared to remain constant throughout the wetland in all three deployments. Water quality targets for conductivity (WQT = 799 S/cm; [35]) were exceeded in all deployments. Elevated levels of conductivity are common in OSPW due to naturally high levels of sodium, bicarbonate, chloride, and sulphate found in waters around the Athabasca oil sands. Since conductivity showed no significant eects on three floating wetland species at values up to 3000 S/cm [39], significant eects on treatment performance as a result of the measured conductivities are not expected in the KT wetland. However, at a conductivity of 4000 S/cm, which is much greater than that in the KT wetland, a 44.4–67.9% reduction in BOD in a continuous flow constructed wetland system with cattails (Typha angustifolia) and Asian crabgrass (Digitaria bicornis) was observed [40]. Total dissolved solids (TDS) ranged between 680 mg/L to 1200 mg/L and were lower than those typically observed in OSPW (i.e., 2000–2500 mg/L; [41]) and within the wide range of concentrations of dissolved solids in surface waters (i.e., 240 mg/L–279,000 mg/L) of the Athabasca oil sands region [42]. No WQT has been issued for TDS. Removal eciencies of organic contaminants in treatment wetlands tend to decrease as salinity increases due to eects on plants, microorganisms, and substrates [43]. The pH of OSPW ranged between 7.77 and 7.92 in deployment one, 7.49 and 7.99 in deployment two, and 7.40 and 8.24 in deployment three. Alkaline pH levels ranging from 8.0–8.4 have been measured in OSPW from other facilities in the Athabasca oil sands region [41] and are common when alkaline reagents such as sodium hydroxide are added to the process water to improve hydrocarbon extraction eciency. TSS concentrations in the OSPW were below detection limits in deployments one and two in 2017. This was due to sedimentation in the overburden disposal pond before water was introduced into the KT wetland. During deployment three in 2018, water was introduced into the wetland during a single rapid pumping event (<24 h) at rates of 4.0–4.7 m /min, which caused turbulence and resuspension of sediment in the source water pond. Since no sedimentation pond was used as a preliminary settling basin, OSPW entering the KT wetland in 2018 contained higher concentrations of TSS than in the 2017 deployments. Within the wetland, TSS reduced by 87% after 24 days of recycling demonstrating the capacity for particulate settling in the KT wetland. 3.2. Polycyclic Aromatic Hydrocarbons in Influent During deployments one and two, freely dissolved concentrations of PAHs in water were highest for chrysene (1.28 (SE 0.35) and 0.37 ng/L (SE 0.039)), fluoranthene (0.92 (SE 0.033) and 0.54 ng/L (SE 0.011)), phenanthrene (0.78 (SE 0.075) and 0.88 ng/L (SE 0.047)), and pyrene (5.41 (SE 0.12) and Environments 2020, 7, 64 9 of 16 3.52 ng/L (SE 0.066)), respectively (Figures 2 and 3). During deployment three freely dissolved PAH concentrations entering the wetland were highest for acenaphthene (2.88 (SE 0.082) ng/L), fluorene (1.34 (SE 0.027) ng/L), phenanthrene (3.83 (SE 0.085) ng/L), pyrene (2.24 (SE 0.13) ng/L), and retene (2.03 (SE 1.4) ng/L) (Figure 4). With the exception of pyrene in the influent of deployment one, concentrations of PAHs in the influent were below the criteria for the protection of aquatic life in Canada (i.e., [44]) and the water quality targets for the Muskeg River (i.e., [35]). Concentrations of PAHs in the influent were similar Environments 2020, 7, x FOR PEER REVIEW 9 of 16 to concentrations measured by Environment and Climate Change Canada (ECCC) in the Athabasca, Peace, and Slave rivers [45] (Figure S1). The concentration of pyrene in the influent exceeded water develop remediation strategies to reduce contaminant concentrations to levels that meet quality targets in the Muskeg River Interim Management Framework [35] and points to the need to environmental quality guidelines and protect wildlife in or visiting the wetland. Environments 2020, 7, x FOR PEER REVIEW 9 of 16 develop remediation strategies to reduce contaminant concentrations to levels that meet environmental quality guidelines and protect wildlife in or visiting the wetland. Forebay Final Deep Pool E EC, i E EL ,i C,i L,i 10000 100% PA H s M ethylated-PA H s 80% 60% 100 40% 20% 0% 1 -20% 1,7-DMPh = 1,7-Dimethylphenanthrene; 1,8-DMPh = 1,8-Dimethylphenanthrene; 1-MeChry = 1-Methylchrysene; 1-MePh = 1-Methylphenanthrene; 2,6- DMPh = 2,6-Dimethylphenanthrene; 2-MePh = 2-Methylphenanthrene; 5,9-DMChry = 5,9-Dimethylchrysene; 7-MeB[a]P = 7-Methylbenzo[a]pyrene; B[b]F = Benzo[b]fluoranthene; B[e]P = Benzo[e]pyrene; Chry = Chrysene; Fla = Fluoranthene; Ph = Phenanthrene; Pyr = Pyrene; Re = Retene Figure 2. Dissolved aqueous concentrations and removal efficiencies of PAHs in the Kearl Treatment Figure 2. Dissolved aqueous concentrations and removal eciencies of PAHs in the Kearl Treatment Wetland during deployment one (2017A). Error bars represents standard errors of the mean. Figure 2. Dissolved aqueous concentrations and removal efficiencies of PAHs in the Kearl Treatment Wetland during deployment one (2017A). Error bars represents standard errors of the mean. Wetland during deployment one (2017A). Error bars represents standard errors of the mean. Figure 3. Dissolved aqueous concentrations and removal eciencies of PAHs in the Kearl Treatment Figure 3. Dissolved aqueous concentrations and removal efficiencies of PAHs in the Kearl Treatment Wetland during deployment two (2017B). Error bars represents standard errors of the mean. Wetland during deployment two (2017B). Error bars represents standard errors of the mean. Figure 3. Dissolved aqueous concentrations and removal efficiencies of PAHs in the Kearl Treatment Wetland during deployment two (2017B). Error bars represents standard errors of the mean. 1,7-DMPh 1,8-DMPh 1-MeChry 1-MePh 2,6-DMPh 2-MePh 5,9-DMChry 7-MeB[a]P B[b]F B[e]P Chry Fl a Ph Pyr Re Concentration (pg/L) Removal Efficiency Envir Envir onments onments 2020 2020 , 7 , ,764 , x FOR PEER REVIEW 10 of 10 16 of 16 Forebay Final Deep Pool E Ec E EL C,i L,i 10000 100% M ethylated-PA H s PA H s 80% 60% 100 40% 20% 0% 1 -20% 1,2,6-TMPh = 1,2,6-Trimethylphenanthrene; 1,7-DMPh = 1,7-Dimethylphenanthrene; 1,8-DMPh = 1,8-Dimethylphenanthrene; 1-MeChry = 1- Methylchrysene; 1-MePh = 1-Methylphenanthrene; 2,3,5-TMN = 2,3,5-Trimethylnaphthalene; 2,3,6-TMN = 2,3,6-Trimethylnaphthalene; 2,4-DMDBt = 2,4-Dimethyldibenzothiophene; 2,6-DMN = 2,6-Dimethylnaphthalene; 2,6-DMPh = 2,6-Dimethylphenanthrene; 2-MePh = 2-Methylphenanthrene; 3- MePh = 3-Methylphenanthrene; 5,9-DMChry = 5,9-Dimethylchrysene; Acen = Acenaphthene; Anth = Anthracene; BiP = Biphenyl; Chry = Chrysene; Fla = Fluoranthene; Fl = Fluorene; Ph = Phenanthrene; Pyr = Pyrene; Re = Retene Figure 4. Dissolved aqueous concentrations and removal eciencies of PAHs in the Kearl Treatment Figure 4. Dissolved aqueous concentrations and removal efficiencies of PAHs in the Kearl Treatment Wetland during deployment three (2018). Error bars represents standard errors of the mean. Wetland during deployment three (2018). Error bars represents standard errors of the mean. 3.3. Wetland Treatment Performance 3.3. Wetland Treatment Performance Treatment in the KT wetland resulted in statistically significant reductions in dissolved aqueous concentration Treatment for in 14 thout e KT of wet 15land PAHs result during ed indeployment statistically sig one, nifi15 cant out red of uctions 20 PAHs in dissolved in deployment aqueous two, concentration for 14 out of 15 PAHs during deployment one, 15 out of 20 PAHs in deployment two, and 19 out of 22 PAHs in deployment three (Figures 2–4). Freely dissolved concentrations of pyrene and 19 out of 22 PAHs in deployment three (Figures 2–4). Freely dissolved concentrations of pyrene in OSPW reduced from 5.41 (SE 0.12) ng/L to 0.46 (SE 0.044) ng/L during deployment one, from 3.52 in OSPW reduced from 5.41 (SE 0.12) ng/L to 0.46 (SE 0.044) ng/L during deployment one, from 3.52 (SE 0.066) ng/L to 0.84 (SE 0.012) ng/L during deployment two, and from 2.24 (SE 0.13) ng/L to 0.58 (SE 0.066) ng/L to 0.84 (SE 0.012) ng/L during deployment two, and from 2.24 (SE 0.13) ng/L to 0.58 (SE 0.0087) ng/L during deployment three. This corresponds to an E for pyrene of 91, 76, and 75% C,i (SE 0.0087) ng/L during deployment three. This corresponds to an EC,i for pyrene of 91, 76, and 75% and an E for pyrene of 92, 76, and 74% for deployments 1–3, respectively (Figures 2–4). Values of L,i and an EL,i for pyrene of 92, 76, and 74% for deployments 1–3, respectively (Figures 2–4). Values of E and E are similar because water inflows and outflows in the wetland were well balanced. C,i L,i EC,i and EL,i are similar because water inflows and outflows in the wetland were well balanced. Pyrene Pyrene measured in deployment one reduced to concentrations that did not exceed the water quality measured in deployment one reduced to concentrations that did not exceed the water quality guideline for pyrene [35]. guideline for pyrene [35]. Both E and E for individual PAHs varied substantially, i.e., from 0% (no statistical dierences C,i L,i Both EC,i and EL,i for individual PAHs varied substantially, i.e., from 0% (no statistical differences in concentration through the wetland) for certain methylated PAHs to 92% (for E ) and 93% (for E ) C,i L,i in concentration through the wetland) for certain methylated PAHs to 92% (for EC,i) and 93% (for EL,i) for fluoranthene during the first deployment. The mean E for all analytes was measured at 72 C,i for fluoranthene during the first deployment. The mean EC,i for all analytes was measured at 72 (SE (SE 4.7)%, 32 (SE 6.9)%, and 50 (SE 5.4)% for each of the three deployments, respectively and closely 4.7)%, 32 (SE 6.9)%, and 50 (SE 5.4)% for each of the three deployments, respectively and closely matched the mean E of 73 (SE 4.4)% for deployment one, 32 (SE 6.9)% for deployment two, and 49 L,i matched the mean EL,i of 73 (SE 4.4)% for deployment one, 32 (SE 6.9)% for deployment two, and 49 (SE 5.5)% for deployment three. The overall reduction in concentration of all PAHs combined (SC ) PAH (SE 5.5)% for deployment three. The overall reduction in concentration of all PAHs combined (ΣCPAH) ranged from 56% (deployment two) to 82% (deployment one). The reduction in total mass of all ranged from 56% (deployment two) to 82% (deployment one). The reduction in total mass of all measur measured ed PAHs PAHs was was 83 83(SE (SE 37)%, 37)%, 5 54 4 (SE (SE 16)% 16)%, , an and d 64 64(SE (SE 19)%, 19)%, inin deploymen deployments ts one, one, two, two, and and three, thr ee, respectively respectivel.y. The The reductions reductions in in mass mass and and concentrations concentrations oof f th the e PA PAHs, Hs, an and d henc hence e EL,i E and and EC,i E in e in ach each L,i C,i deployment were similar because the net change of water in the wetland was close to zero with only deployment were similar because the net change of water in the wetland was close to zero with only small gains and losses of water in the KT wetland. These reductions in concentrations align with small gains and losses of water in the KT wetland. These reductions in concentrations align with those found in other studies. For example, [46] measured an EC for the combined sum of 16 PAHs of those found in other studies. For example, [46] measured an E for the combined sum of 16 PAHs of 56% after 24 h in a pilot-scale surface flow wetland that treated highway runoff in southern Greece. 56% after 24 h in a pilot-scale surface flow wetland that treated highway runo in southern Greece. In a pilot-scale vertical flow constructed wetland system in Munich, Germany [47], a 99% reduction In a pilot-scale vertical flow constructed wetland system in Munich, Germany [47], a 99% reduction in the concentration of phenanthrene in artificial wastewater (8 μg/L) was achieved over 14 days. The in the concentration of phenanthrene in artificial wastewater (8 g/L) was achieved over 14 days. present study measured a reduction in the concentration of phenanthrene in OSPW influent (i.e., 0.77 The present study measured a reduction in the concentration of phenanthrene in OSPW influent ng/L and 3.83 ng/L) in the KT wetland of 36 to 66% over 14 days. (i.e., 0.77 ng/L and 3.83 ng/L) in the KT wetland of 36 to 66% over 14 days. No statistically significant dierences between the influent and euent concentrations were observed for nine PAHs (i.e., benzo[b]fluoranthene (deployment one), 1-methylnaphthalene, 1,2, 6-TMP h , M h 1 7-D P 1,8-DMPh 1-MeChry 1-MePh 2,3, 5-TMN 2,3, 6-TMN 2,4-DMDBt 2,6-DMN 2.6-DMPh 2-MePh 3-MePh 5,9-DMChry Acen Anth BiP Chry Fla Fl Ph Pyr Re Concentration (pg/L) Removal Efficiency Environments 2020, 7, x FOR PEER REVIEW 11 of 16 Environments 2020, 7, 64 11 of 16 No statistically significant differences between the influent and effluent concentrations were observed for nine PAHs (i.e., benzo[b]fluoranthene (deployment one), 1-methylnaphthalene, 2- 2-methylnaphthalene, 2,3,6-trimethylnaphthalene, 2,6-dimethylnaphthalene, fluorene (deployment methylnaphthalene, 2,3,6-trimethylnaphthalene, 2,6-dimethylnaphthalene, fluorene (deployment two), and 2,4-dimethyldibenzothiophene, 2,6-dimethylnaphthalene, and biphenyl (deployment two), and 2,4-dimethyldibenzothiophene, 2,6-dimethylnaphthalene, and biphenyl (deployment three). three). Five of these chemicals are naphthalene-based PAHs with low molecular weights, high Five of these chemicals are naphthalene-based PAHs with low molecular weights, high aqueous aqueous solubilities, and relatively low log KOW (3.86–4.73) when compared to the other test chemicals solubilities, and relatively low log K (3.86–4.73) when compared to the other test chemicals in this OW in this study. High water solubility and low log KOW make substances less susceptible to mechanisms study. High water solubility and low log K make substances less susceptible to mechanisms of OW of chemical removal in rooting media, biofilm, and vegetation because a higher proportion of the chemical removal in rooting media, biofilm, and vegetation because a higher proportion of the chemical chemical remains in the water phase [48]. remains in the water phase [48]. Pyrene, fluoranthene, and chrysene consistently exhibited the highest EL,i in all three Pyrene, fluoranthene, and chrysene consistently exhibited the highest E in all three deployments L,i deployments whereas 2-methylphenanthrene and 2,6-dimethylphenanthrene consistently showed whereas 2-methylphenanthrene and 2,6-dimethylphenanthrene consistently showed the lowest E the lowest EL,i among the three deployments (Figure 5). This consistent pattern of relative treatment L,i among efficienc they thr of ee these deployments PAHs further (Figur indica ete 5s ).th This at chem consistent ical make pattern -up and of pro relative perties pla treatment y an import e ant ciency of role in wetland treatment. Furthermore, since the operation of the wetland was replicated in all three these PAHs further indicates that chemical make-up and properties play an important role in wetland deployments, the considerable variation in EL,i of individual PAHs between deployments shows that treatment. Furthermore, since the operation of the wetland was replicated in all three deployments, environmental conditions and water quality characteristics also play an important role in wetland the considerable variation in E of individual PAHs between deployments shows that environmental L,i treatment. conditions and water quality characteristics also play an important role in wetland treatment. 2017A 2017B 2018 100% 75% 50% 25% 0% Figure 5. Mass-removal eciency of PAHs measured in all three deployments of PES in the Kearl Figure 5. Mass-removal efficiency of PAHs measured in all three deployments of PES in the Kearl treatment wetland. treatment wetland. Toxicity Toxicity Chemical activities of individual PAHs (Table S5) were all far below the baseline toxicity value Chemical activities of individual PAHs (Table S5) were all far below the baseline toxicity value (a ) of 0.01. This indicates that individual PAHs in both the influent and the euent of the wetland (a0) of 0.01. This indicates that individual PAHs in both the influent and the effluent of the wetland were below concentrations that cause acute toxic eects in biota. Retene, chrysene, and pyrene were below concentrations that cause acute toxic effects in biota. Retene, chrysene, and pyrene exhibited the highest chemical activities in the KT wetland throughout the study. The highest exhibited the highest chemical activities in the KT wetland throughout the study. The highest chemical activities for all three of these chemicals were found in the forebay during deployment one, chemical activities for all three of these chemicals were found in the forebay during deployment one, 5 6 5 7 5 −5 −6 −5 −7 −5 wher where e a areten= e = 5.2 5.2 ×10 10 ((SE SE 1.7 1.7× 10 10 ), )a,chr aysene = 4.8 =× 4.8 10 (S 10 E 5.1 (SE × 5.1 10 ), and 10 a), pyren and e = 4.0 a × 10= 4.0 (SE 10 retene pyrene chrysene −7 −5 −5 −6 7 5 5 6 8.6 × 10 ). These chemical activities reduced to 1.2 × 10 (SE 3.1 × 10 ) for retene, 6.1 × 10 (SE (SE 8.6 10 ). These chemical activities reduced to 1.2 10 (SE 3.1 10 ) for retene, 6.1 10 −5 −6 −5 5 6 5 1.6 × 10 ) for chrysene, and 3.4 × 10 (SE 4.0 × 10 ) for pyrene in the final deep pool of the (SE 1.6 10 ) for chrysene, and 3.4 10 (SE 4.0 10 ) for pyrene in the final deep pool of the wetland, corresponding to reductions in chemical activity of 77, 87, and 91%, respectively. The largest wetland, corresponding to reductions in chemical activity of 77, 87, and 91%, respectively. The largest reduction in chemical activity was observed for fluoranthene during deployment one, with a reduction in chemical activity was observed for fluoranthene during deployment one, with a reduction reduction of 92%. of 92%. The total chemical activity (Sa) for all analytes included in the analysis was also below the baseline toxicity value (a ) of 0.01 suggesting concentrations of PAHs in water are below the threshold for acute eects in aquatic biota due to non-polar narcosis. These results are in agreement with the toxicity experiments with the influent OSPW performed by Maxxam Analytics (unpublished data), which showed no acute toxicity of the OSPW to rainbow trout. Sa was highest during deployment three L Environments 2020, 7, x FOR PEER REVIEW 12 of 16 The total chemical activity (Σa) for all analytes included in the analysis was also below the baseline toxicity value (a0) of 0.01 suggesting concentrations of PAHs in water are below the threshold for acute effects in aquatic biota due to non-polar narcosis. These results are in agreement with the Environments 2020, 7, 64 12 of 16 toxicity experiments with the influent OSPW performed by Maxxam Analytics (unpublished data), which showed no acute toxicity of the OSPW to rainbow trout. Σa was highest during deployment 4 −4 6 −6 −4 4 −