Fate of N-Nitrosodimethylamine in recycled water after recharge into anaerobic aquifer

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Fate of N-Nitrosodimethylamine in recycled water after recharge into anaerobic aquifer B.M. Patterson a,b,*, M.M. Pitoi a,b, A.J. Furness a, T.P. Bastow a, A.J. McKinley b a b

CSIRO Land and Water, Private Bag 5, Wembley, WA 6913, Australia School of Biomedical, Biomolecular and Chemical Sciences, University of WA, Crawley, WA 6009, Australia

article info

abstract

Article history:

Laboratory and field experiments were undertaken to assess the fate of N-nitro-

Received 26 October 2011

sodimethylamine (NDMA) in aerobic recycled water that was recharged into a deep

Received in revised form

anaerobic pyritic aquifer, as part of a managed aquifer recharge (MAR) strategy. Laboratory

13 December 2011

studies demonstrated a high mobility of NDMA in the Leederville aquifer system with

Accepted 14 December 2011

a retardation coefficient of 1.1. Anaerobic degradation column and

Available online 23 December 2011

studies showed that anaerobic conditions of the aquifer provided a suitable environment

14

C-NDMA microcosm

for the biodegradation of NDMA with first-order kinetics. At microgram per litre concenKeywords:

trations, inhibition of biodegradation was observed with degradation half-lives (260
20

MAR

days) up to an order of magnitude greater than at nanogram per litre concentrations

NDMA

(25e150 days), which are more typical of environmental concentrations. No threshold

Degradation

effects were observed at the lower ng L1 concentrations with NDMA concentrations

Recycled water

reduced from 560 ng L1 to <6 ng L1 over a 42 day

N-Nitrosamines

experiment.

14

C-NDMA aerobic microcosm

Aerobic 14C-NDMA microcosm studies were also undertaken to assess potential aerobic degradation, likely to occur close to the recharge bore. These microcosm experiments showed a faster degradation rate than anaerobic microcosms, with a degradation half-life of 8
2 days, after a lag period of approximately 10 days. Results from a MAR field trial recharging the Leederville aquifer with aerobic recycled water showed that NDMA concentrations reduced from 2.5
1.0 ng L1 to 1.3
0.4 ng L1 between the recharge bore and a monitoring location 20 m down gradient (an estimated aquifer residence time of 10 days), consistent with data from the aerobic microcosm experiment. Further down gradient, in the anaerobic zone of the aquifer, NDMA degradation could not be assessed, as NDMA concentrations were too close to their analytical detection limit (<1 ng L1). Crown Copyright ª 2011 Published by Elsevier Ltd. All rights reserved.

1.

Introduction

NDMA is a probable human carcinogen (IARC, 1987; USEPA, 1987) and may be formed during wastewater treatment. NDMA also has a low sorption to sediments (Gunnison et al.,

2000; Yang et al., 2005) thus increasing the risk of contaminating large volumes of an aquifer, if NDMA is present in wastewater used for managed aquifer recharge (MAR). Although the evidence of NDMA carcinogenic effects to humans are limited, numerous experiments have shown it to

* Corresponding author. School of Biomedical, Biomolecular and Chemical Sciences, University of WA, Crawley, WA 6009, Australia. Tel.: þ61 8 93336276; fax: þ61 8 9333 6211. E-mail address: [email protected] (B.M. Patterson). 0043-1354/$ e see front matter Crown Copyright ª 2011 Published by Elsevier Ltd. All rights reserved. doi:10.1016/j.watres.2011.12.032

w a t e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 2 6 0 e1 2 7 2

be highly toxic, carcinogenic and mutagenic to animals (Futakuchi et al., 1996; Ishinishi et al., 1988; Liteplo and Meek, 2001; Richardson et al., 2007). Based on a lifetime cancer risk of 105, USEPA recommended a maximum concentration of 7 ng L1 in drinking water (USEPA, 1987). Drinking water maximum guidelines value for NDMA in Australia is 10 ng L1 (EPHC, 2008). NDMA is formed during disinfection of drinking water and wastewater, especially disinfection by chloramination (Choi and Valentine, 2002; Mitch and Sedlak, 2002; Najm and Trussel, 2001; Pehlivanoglu-Mantas et al., 2006). However, chlorination (Krasner et al., 2009; Pehlivanoglu-Mantas and Sedlak, 2006; Pehlivanoglu-Mantas et al., 2006) and ozonation (Zhao et al., 2008) can also form NDMA. During disinfection, NDMA was formed by the reaction of monochloramine with dimethylamine (DMA) (Choi and Valentine, 2002; Mitch and Sedlak, 2002) or other organic precursors (Mitch et al., 2003a; Pehlivanoglu-Mantas and Sedlak, 2006). NDMA is the most abundant and frequently detected nitrosamine in primary and secondary wastewater effluent (Krauss and Hollender, 2008; Krauss et al., 2009; Zhao et al., 2008). It was reported to be found in drinking water and wastewater treatment plants in USA, Canada, Switzerland and UK (Krasner et al., 2009; Krauss and Hollender, 2008; Krauss et al., 2009; Schreiber and Mitch, 2006; Templeton and Chen, 2010; Zhao et al., 2008) with concentrations in wastewater up to 400 ng L1 (Schreiber and Mitch, 2006). The higher NDMA concentrations that are typically observed in wastewater effluent compared to drinking water are likely a result of the higher concentrations of anthropogenic pollutants with dimethylamine functional groups that provide precursors for NDMA formation (Le Roux et al., 2011). The occurrence of NDMA was also reported in the rubber and tyre industry in 1979 (Fajen et al., 1979), although measurements of their concentrations in samples from German rubber industry plants can be traced back to 1965 (de Vocht et al., 2007). Apart from the rubber industry, NDMA was also found in snuff tobacco at trace concentrations (Brunnemann et al., 1982) and in human urine samples at concentrations ranging between 20 and 100 ng L1 (Kakizoe et al., 1979). NDMA may not be completely removed during wastewater treatment. Instead, NDMA concentrations may increase when chloramines are employed to prevent bio-fouling in membrane treatment systems such as microfiltration (Plumlee et al., 2008). Also, NDMA is less degradable in wastewater treatment plants compared to other nitrosamines such as N-nitrosodiethylamine, N-nitroso-di-n-propylamine and N-nitrosopiperidine (Hollender et al., 2009; Krauss et al., 2009) with an average aqueous removal efficiency of 65% in 20 wastewater treatment plants investigated in Switzerland. Reverse osmosis (RO) treatment, which removes many hydrophobic compounds from wastewater (Steinle-Darling et al., 2007), was found to be ineffective for complete NDMA removal because the NDMA molecule is small, polar, and uncharged (Mitch et al., 2003b; Steinle-Darling et al., 2007). Only 24e56% of NDMA was removed during RO from an advanced wastewater treatment plant in Southern California which was consistent with other reported removal rates of 45e65% and 10e70% (Plumlee et al., 2008). Khan and

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McDonald (2010) reported that less than 40% of spiked NDMA was removed during RO. Steinle-Darling et al. (2007) also reported removal rates of 54e70% for three different RO membrane materials tested. Apart from RO, activated sludge, sand filtration and ultraviolet treatment have been reported to reduce NDMA from wastewater. Investigation of 21 wastewater plants in Switzerland showed that activated sludge was able to remove NDMA with an average removal efficiency of 43% while sand filtration removed NDMA with an average removal efficiency of 38% (Krauss et al., 2009). Ultraviolet (UV) treatment reduced NDMA concentrations between 43% and 66%, and combining RO with UV treatment reduced NDMA concentration by up to 77% (Plumlee et al., 2008). When recycled water is used as the recharge water for MAR, biogeochemical changes during aquifer storage or aquifer passage may result in beneficial water quality changes such as the natural attenuation of trace organics (Dillon, 2005). Therefore, for NDMA that is not completely removed by various wastewater treatments, biodegradation of NDMA during aquifer passage (as part of MAR) may provide possible additional attenuation. However, the persistence of NDMA in groundwater aquifers from RO treated wastewater has been responsible for the closure of municipal drinking water wells in the United States (Mitch et al., 2003b; Sharp et al., 2005), suggesting possible limited NDMA biodegradation in aquifers. NDMA biodegradation has been previously investigated (Table 1) in sediments via batch studies (Kaplan and Kaplan, 1985; Gunnison et al., 2000; Bradley et al., 2005; Yang et al., 2005; Arienzo et al., 2006; Gan et al., 2006; Szecsody et al., 2008), column studies (Drewes et al., 2006; Szecsody et al., 2008; Patterson et al., 2010, 2011) or through field observations (Zhou et al., 2009). These investigations (Table 1) showed that NDMA degradation occurred in the majority of experiments with half-lives between 1.4 and 80 days. For experiments conducted with the same sediment under aerobic and anaerobic conditions (Gunnison et al., 2000), no substantial differences in half-lives were observed under these different redox conditions. However, degradation rates varied for the different sediments examined, suggesting NDMA degradation rates were likely site specific. Additionally, NDMA degradation rates were likely concentration dependent with decreased degradation rates at high concentrations for the same sediments (Gunnison et al., 2000; Szecsody et al., 2008; Yang et al., 2005). For the sediment that showed limited degradation (Patterson et al., 2010), further research has now been conducted to determine if this limited degradation was due to inhibitory concentration effects and/or redox conditions. Experiments were conducted to mimic the injection of NDMA-contaminated recycled water into an aquifer under anaerobic and aerobic conditions with the specific purposes to: (i) assess the degradation of NDMA in an aquifer at ng L1 concentrations that are more typical in environmental water samples; (ii) determine concentration effects on degradation rates and assess if higher concentrations (mg L1 concentrations) would provide comparable fate data; and (iii) assess changes in degradation rates for longer-term experiments and relate this data to degradation lag times and bacterial acclimation.

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Table 1 e Comparison of degradation rates of NDMA. Sources

Types of soil

Kaplan and Kaplan (1985)

Topsoil from a garden shop

Gunnison et al. (2000)

Aquifers from several locations in Rocky Mountain Arsenal, Colorado, US NDMA-impacted surface soil Surface soil with no history of NDMA Surface soils

Bradley et al. (2005) Yang et al. (2005) Arienzo et al. (2006) Gan et al. (2006) Drewes et al. (2006) Szecsody et al. (2008)

Zhou et al. (2009) Patterson et al. (2010) Patterson et al. (2011)

Surface soil at several depths Surface soils at different depths Aquifers from Arizona Aquifers from California, Washington, and New Jersey

A site in Los Angeles County, California, USA (Field study) Anaerobic pyritic aquifer from Perth, Western Australia Aerobic sandy aquifer from Perth, Western Australia

Initial concentration in soil or water 1

Aerobic condition Anaerobic condition 10a 35a 51a 51a 42a 11.6e35

0.01 mg g 1 mg g1 100 mg g1 1000 mg g1 10000 mg g1 50 mg L1

47 mg L1 47 mg L1 250 mg kg1 25 mg kg1 100 mg kg1 250 mg kg1 120 ng L1 6 ng L1e2.5 mg L1 influent water 10.2 ng L1e2.5 mg L1 influent water Elevated concentrations during the field study 590 mg L1 590 mg L1

Half-life (days)

26.3e38.5

60a 92a 4.1e22.5 4.0e14.8 3.1e13.1 13.5e79.7 1.42b 14.25 (average) 58 69.5a >100 >50

a calculated from the rate constants (as first-order degradation). b anoxic conditions.

The research undertaken consisted of (i) a 675 day high concentration (590 mg L1) stop-flow column experiment, (ii) a 520 day low concentration (200 ng L1) slow-flow column experiment, and (iii) an anaerobic 38 day, anaerobic 160 day, and aerobic 42 day low concentration (560 ng L1) 14C-NDMA microcosm experiments. Additionally, field data from a MAR site recharging aerobic recycled water into the Leederville aquifer in Perth, Western Australia was interpreted to provide a comparison between laboratory and field data.

2.

Materials and methods

2.1.

Chemicals

NDMA and sodium azide were sourced from SigmaeAldrich (Sydney, Australia). d8-NDMA was sourced from CDN isotopes (Honsby, Australia). N-[methyle14C]NDMA which was sourced from Moravek Biochemicals and Radiochemicals (Brea, California, USA) had a specific activity of 54 mCi mmol1, and a concentration in sterile water of 1.4 mg mL1. Sodium bromide was sourced from Hayashi Pure Chemical Ind. Ltd. (Osaka, Japan).

2.2.

Aquifer sediment

The sediment used in all experiments was anaerobic Leederville aquifer sediment collected from a trial MAR site in Perth, Western Australia. The sediment was collected as core

samples or drill cuttings from the confined aquifer over the proposed MAR injection depth interval between 120 m and 220 m below ground level. To prevent sediment oxidation, collected sediment was immediately stored in sealed air-tight containers and kept at 4  C. The mineralogy of the sediment was described in Patterson et al. (2010). In short, the sediment was predominantly quartz (72%), K-feldspar (24%), and minor quantities of pyrite (2%) and Na-feldspar (based on X-ray diffraction) with 0.32% of sediment organic matter (SOM) (Patterson et al., 2010). Particle size distribution of the sediment was >1 mm (10%); 1 mme500 mm (20%); 500 mme250 mm (50%); 250 mme125 mm (17%) and <125 mm (3%). Based on incubation experiments by Descourvie`res et al. (2010), the sediment was highly reductive with a measured reductive capacity between 29 and 143 mmol O2 g1, and pyrite oxidation was the dominant oxygen consuming reaction.

2.3.

Column experiments

2.3.1.

Column influent water

Recycled water used for the experiments was secondary treated municipal wastewater collected from the Beenyup wastewater treatment plant, Perth, Western Australia, which has undergone further purification through ultrafiltration, RO and UV disinfection. The chemistry of the recycled water is given in Table 2. For the column experiments, the recycled water was spiked with NDMA and bromide prior entering the column. The influent water bromide concentration was 50 mg L1,

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Table 2 e Chemistry of recycled water and Leederville aquifer groundwater. Analytes pH Chloride (mg L1) Sulphate (mg L1) Nitrate-N (mg L1) Sodium (mg L1) Potassium (mg L1) Magnesium (mg L1) Calcium (mg L1) Conductivity (mS/cm) Dissolved oxygen (mg L1) Dissolved organic carbon (DOC) (mg L1) NDMA (ng L1)

Recycled water

Groundwater

6.9 3.0 0.38 <1 5.0 0.55 0.18 1.6 34 8.5 <1

6.8 73 42 <1 66 3.6 11 60 710 <1 1.2

8

<2

while the influent water NDMA concentration was 200 ng L1 and 590 mg L1 for low and high concentration column experiments respectively. Bromide was added to the column influent water as a conservative tracer to determine the retardation coefficient of NDMA in the column experiments. To differentiate between abiotic and biotic processes in the column experiments, two columns were used for each column experiment. One column was operated as a non-sterile experimental column while the other column was operated as a sterile control column. The microbial activity in the sterile control column was suppressed by the addition of the metabolic inhibitor sodium azide (0.65 g L1) to the influent water. The influent water for sterile and non-sterile columns was prepared separately every two weeks and stored in a separate 5 L SKC Flexfoil Grab Bag prior to entering the columns. These samples bags were selected to minimise sorption and photo-oxidation of NDMA during storage. Analysis of influent water of the sterile and non-sterile columns after a two week storage period showed that the NDMA concentration in sample bags remained at >80% of the initial concentration.

2.3.2.

Low NDMA concentration column setup

Two stainless steel columns were used for low NDMA concentration experiments. This column material was chosen to minimise sorption and photo-oxidation of NDMA within the column. One column was used as a sterile control, while the other was used as non-sterile column. The columns were 100 cm in length and 14.5 cm in diameter. To prevent sediment migrating into the influent or effluent tubing, a stainless steel grate with holes (1.0 cm in diameter) and a stainless steel mesh (0.1 cm diameter) was fixed at the bottom and the top of each column. The base of each column, below the stainless steel mesh, provided a 500 mL mixing chamber and to collect influent water prior to entering the column. Each column has 11 sampling ports at 0 (just below the stainless steel mesh), 4, 11, 18, 25, 32, 39, 46, 59, 75, and 92 cm from the bottom of the column. Each sampling port comprised a 0.4 cm i.d. stainless steel tube that protruded 6.0 cm from the wall into the centre of the column. The inner end of the tube was fixed with a stainless mesh (0.1 cm diameter) to avoid sediment migrating into the tube whereas the outer end contained a silicon septum to allow a syringe needle to be inserted for water sample collection.

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The columns were wet packed with homogenised anaerobic Leederville aquifer sediment. To maintain anaerobic conditions during packing, the column gas space was flushed with nitrogen gas while anaerobic groundwater from the Leederville aquifer (Table 2) was added to maintain a water level above the packed sediment to avoid sediment stratification. The Leederville groundwater was continually flushed through the column in a flow rate of 5000 mL d1 for two weeks after the packing in order to confirm the permeability of the sediment-packed column. After the two weeks, the flow was reduced to 190 mL d1 for two months to stabilize the groundwater chemistry in the columns before the groundwater was replaced with recycled water for 2 months. After this time NDMA/bromide spiked recycled water was introduced into the columns. The flow of the influent water into each column was maintained at 190 mL d1 using separate MCP standard drive pumps (ISMATEC), giving a linear velocity of approximately 2.2 cm day1, estimated from the bromide tracer test. This gave a water residence time of 42 days.

2.3.3.

High NDMA concentration column setup

The set up for columns used for high concentration experiments were similar to low concentration experiments with the exception that the columns were 200 cm long and the position of sampling ports were at 0, 4, 11, 18, 25, 32, 39, 46, 59, 75, 92, 109, 125, 142, 158, 175 and 192 cm from the base of each column. Further details of the column configuration is described in Patterson et al. (2010). For the stop-flow experiment, after a w12 months period of continuous injection of NDMA spiked recycled water the influent water flow to high NDMA concentration columns was ceased. Influent water flow was ceased for 675 day. This was undertaken to increase the column water residence time within the columns.

2.3.4.

Column water sample collection

For the low NDMA concentration column experiment, water samples from each sampling port of the columns were collected using 100 mL gas tight syringes (SGE). For NDMA analysis, a 50 mL sample was stored in a Pyrex screw cap test tube (Corning) with a Teflon lined septa cap prior to extraction and analysis. For the high NDMA concentration experiment, a minimum volume of sample was taken from each port (5 mL) to minimise perturbing effect to the columns. A 50 mL subsample was diluted to 50 mL using Milli-Q water and then processed similarly to the samples from low NDMA concentration column experiment. The remaining part of the sample was used for bromide and dimethylamine (DMA) analysis. Samples for bromide analysis were filtered using Bulk Aerodisc 32 mm filters with 0.45 mm Supor membranes (PAL life) and stored in 750 mL polypropylene vials capped with PTFE/ silicone lined caps prior to analysis by ion chromatography.

2.3.5.

Analytical methods

For the laboratory experiments, NDMA was analysed using micro liquideliquid extraction followed by GCeMS analysis as developed by Ranwala (2009). This method has a limit of detection of 2 ng L1 and recovery of 30
5%. In summary, a 50 mL water sample was spiked with d8-NDMA as a surrogate standard and extracted three times with 10 mL of dichloromethane (DCM) for each extraction. The DCM

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containing NDMA was then dried with anhydrous sodium sulfate and evaporated down to 2 mL. A clean-up step was subsequently applied to the 2 mL sample in order to remove interferences. For the clean-up step, the sample (2 mL) was loaded onto a small silica gel column and eluted with 5% of acetone in DCM (2 mL) followed by 10% of acetone in DCM (2 mL). The first 4 mL of the eluent was discharged and the last 2 mL (10% of acetone in DCM) containing NDMA was collected, concentrated to 50 mL, an internal standard added (d8-naphthalene, 2.5 ng) and the sample subjected to GCeMS analysis. GCeMS analysis was carried out with an Agilent 5975 MSD, interfaced with a Agilent 6890 gas chromatograph fitted with fused-silica open tubular columns coated with a AT-5ms stationary phase (Alltech, 60 m  0.25 mm i.d., 0.25 mm film thickness). The GC oven temperature was programmed from  35 C (1 min) to 150  C at 5  C min1 and then to 280  C (5 min) at 30  C min1. Samples for analysis were injected (1.2 mL) into  a vaporising injector (260 C) operated in splitless mode (0.4 min) using an autosampler. Helium was used as carrier gas at a constant flow of 1.2 mL min1. Typical MSD conditions  were: ionisation energy 70 eV; source temperature 230 C; electron multiplier voltage 1700 V. The mass spectrometer was run in selected ion monitoring (SIM) mode. The ions monitored for NDMA were 74 and 80 m/z. The peaks in the chromatograms were integrated using Chemstation D 2.00.275 and Microsoft Excel software. For the field trail, NDMA was analysed by GCeMS analysis at the ChemCentre in Western Australia, based on EPA Method 521 with a detection limit 1 ng L1. For bromide analysis an internal standard was added to the filtered water sample prior to analysis, i.e. 50 mL of 200 mg L1 of LiF was added to 500 mL of a filtered water sample. 10 mL of the mixture was then analysed using a Dionex ICS-3000 RFIC with an AS18 4 mm column. A 33 mM potassium hydroxide eluent was used at a flow rate of 1 mL min1 with a column temperature of 40  C. DMA was analysed using a Dionex ICS3000 RFIC with a CS16 5 mm column. Methanesulfonic acid (30 mM) was used as an eluent at flow of 1.0 mL min1 and a column temperature of 40  C.

2.4. 14

14

C-NDMA microcosm experiments

C-NDMA microcosm experiments were undertaken to provide unambiguous NDMA mineralisation data for low NDMA concentrations. As mass balance 14C-NDMA mineralisation data required destructive sampling of the microcosms, two microcosm experiments were undertaken for anaerobic conditions. A shorter 38 day experiment with more intensive sub-sampling, and a longer 160 day experiment to provide a longer timeframe for transformation of 14C-NDMA to 14 CeCO2. A 42 day aerobic microcosm experiment was also undertaken to assess potential aerobic degradation that may occur close to the recharge bore during injection of aerobic recycled water if the reduction capacity of the near injection sediment became exhausted. Microcosm experiments were undertaken in 250 mL Bellco respirometry flasks (Belco Glass, Vineland, New Jersey). Ten grams of anaerobic Leederville sediment was added to each anaerobic flask, and 10 g of aerobic sediment (anaerobic sediment aerated under saturated conditions for 12 months) was added to each aerobic flask. Accounting for the moisture content of the sediment,

groundwater was added to the flask to give a total water content of 100 mL. Sterile control microcosms were prepared by dosing these flasks with either sodium azide to a final concentration of 50 mg L1 or mercuric chloride to a final concentration of 20 mg L1. The sidearm of the flask was filled with 25 mL of 1 M NaOH to capture 14CeCO2 produced. The sidearm was then sealed with a Teflon coated rubber stopper and a long stainless steel 18-guage needle was passed through the rubber stopper and into the NaOH solution. A Teflon tap was connected to the Luer connection of the needle for periodic collection of NaOH samples (2 mL). Once the NaOH sample was collected, a 2 mL N2 or O2 sample was injected into the microcosm. N-[methy-14C]NDMA (1.4 mg mL1 NDMA; 1.0 mCi mL1) was diluted fifty fold with deionised water to give a concentration of 28 mg mL1. Then 2 mL of the diluted NDMA solution was added to each flask to give a final concentration of 560 ng L1 (0.04 mCi). Immediately prior to the addition of the NDMA solution, anaerobic microcosms were flushed with N2 gas for 60 s. Once the microcosm was flushed, the NDMA solution was immediately added and the microcosms sealed with a Teflon coated rubber stopper. Aerobic microcosms were flushed with O2 before NDMA addition and sealing with a Teflon coated rubber stopper. To determine the radioactivity initially added to the microcosms as 14C-NDMA, a triplicate solution (2 mL) of the diluted NDMA solution in 100 mL of deionised water was prepared. Each solution (2 mL) was then collected and 10 mL of scintillant (Starscint, Packard) added. The radioactivity of each solution was then determined in a 2250CA Tri-carb Packard Liquid Scintillation Analyser. At the end of the experiment, a mass balance of radioactivity was undertaken. The sediment slurry was acidified with 10 mL of 1 M sulphuric acid to convert any labelled carbon present as bicarbonate or carbonate in to CO2. After equilibration of a day, the NaOH trap was again sampled to determine bicarbonate/carbonate formation. Also, the slurry was filtered and 2 ml of the collected supernatant was mixed with 10 mL of scintillant and the radioactivity measured to determine non-metabolised 14C-NDMA and other aqueous nonvolatile radioactive compounds remaining in the slurry. For the anaerobic 160 day experiment, NaOH samples (2 mL) were collected from the sidearm of each microcosm on day 0, 7, 22, 42, 92, 114, 149 and 160. For the anaerobic 38 day experiment, NaOH samples (2 mL) were collected from the sidearm of each microcosm on day 0, 4, 7, 11, 14, 17, 22, 29, 31, 35 and 38. For the aerobic 42 day experiment, NaOH samples (2 mL) were collected from the sidearm of each microcosm on day 0, 4, 7, 11, 14, 21, 26, 33, 35, 39 and 42. For each NaOH sample that was collected, 10 mL of scintillant (Starscint, Packard) was added, and the radioactivity of each solution was then determined in a 2250CA Tri-carb Packard Liquid Scintillation Analyser. All counts were adjusted for background interference and recalculated based on the volume of NaOH in the sidearm at the time of sampling.

2.5.

Field assessment

NDMA degradation was also assessed at field scale during an MAR field trial investigating the groundwater replenishment

w a t e r r e s e a r c h 4 6 ( 2 0 1 2 ) 1 2 6 0 e1 2 7 2

of an anaerobic aquifer (Leederville aquifer) using aerobic recycled water in Perth, Western Australia. The recycled water used was secondary treated municipal wastewater collected from the Beenyup wastewater treatment plant, Perth, Western Australia that had undergone ultrafiltration, RO and UV disinfection at the Water Corporation’s Advanced Water Recycling Plant. Details of the field trial known as the Groundwater Replenishment Trial are given in Water Corporation (2009). NDMA concentration data of the injected recycled water was compared to data from monitoring bores located 20, 60 and 120 m from the recharge bore. At a depth of 129 m below ground, groundwater residence times between the recharge bore and monitoring bores were 10 days, 61 days and 270 days for the 20, 60 and 120 m monitoring locations. For each monitored location, a residence time was estimated from the recharge water breakthrough curve by calculating the time to reach a 50% reduction in electrical conductivity. The average NDMA concentration for each monitoring location was determined from groundwater samples collected once recharge water had reached the monitored location (determined from the reduction in electrical conductivity). Groundwater samples prior to recharge water breakthrough were not included in the interpretation of NDMA degradation.

3.

Results and discussion

3.1.

Retardation coefficient

After NDMA and bromide were introduced into the columns, a retardation coefficient (R) was determined by comparing the migration rate of NDMA along the column to the migration rate of the conservative tracer bromide (Fig. 1). Data from the sterile columns were used in preference to the non-sterile column data to eliminate the potential for biodegradation

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confounding the interpretation of the retardation data. Porosity studies based on bromide breakthrough data of the sterile and non-sterile columns showed that column packing of both columns were similar. NDMA and Br breakthrough profiles after 21 days (for low NDMA concentration column experiment) and 15 days (for high NDMA concentration column experiment) for the sterile columns are shown in Fig. 1. For the low NDMA concentration column experiment (Fig. 1b), the non-zero NDMA concentrations at the effluent end of the column (80e100 cm) was likely a result of relatively high background NDMA concentrations (w50 ng L1) in the recycled water at this time. Improvements in the recycled water treatment system over time have reduced NDMA concentrations in the recycled water. During the laboratory experiments, background NDMA concentrations in the recycled water were on average 8 ng L1, and during the field trial was 2.5 ng L1. R values (assuming linear sorption isotherms) for NDMA were determined by (i) initially fitting the bromide data to the convectionedispersion equation (Parker and Vangenuchten, 1984) with a nonlinear least squares fitting routine based on the LevenbergeMarquardt algorithm (MICROCAL, 1995) using Origin v7 software, then (ii) fitting the data for NDMA to the convectionedispersion equation constrained using the bromide fitted parameters, except the R value which was used as the fitting parameter. For the low NDMA concentration column experiment data (Fig. 1b), the breakthrough data was confounded by the non-zero NDMA present in the recycled water and 0.25 was used as the final concentration fitted parameter. From the experiential data, an R value of 1.1 was estimated for the low NDMA concentration column experiment. This estimated R value was consistent with the reported R value of 1.1 in the same sediment at 3 order of magnitude higher concentration (Patterson et al., 2010). While sorption studies at only two different concentrations in the same

Fig. 1 e NDMA (eBe) and bromide (e-e) breakthrough profiles for A) high NDMA concentration column experiment 15 days after NDMA delivery commenced (modified after Patterson et al., 2010) and B) low NDMA concentration column experiment 21 days after NDMA delivery commenced.

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sediment are not adequate to confirm linearity of sorption behaviour, the similar R values at these two different concentrations would suggest that the sorption isotherm of NDMA in Leederville sediment was near linear and sorption was not substantially concentration dependant over ng L1emg L1 concentration range. Sorption behaviour of NDMA was previously reported to be linear (r2 > 0.95) for surface soils (Yang et al., 2005) and (r2 > 0.90) for subsurface soils (Gunnison et al., 2000). For these soils, NDMA sorption was relatively low with R values between 2.5 and 5.3, based on distribution coefficients between 0.45 and 1.14 L kg1 (Gunnison et al., 2000; Yang et al., 2005) and a porosity of 0.4 and bulk density of 1.5 kg L1. The low NDMA R values, estimated from these column experiments, suggest that NDMA would be highly mobile in the Leederville aquifer. The low octanol-water partition coefficient of NDMA (log Kow ¼ 0.57) would explain its low sorption behaviour observed in the Leederville sediment (organic carbon content of 0.32% w/w). Based on this sorption data, NDMA would travel at a velocity of at least 90% of groundwater velocity if it is present in the Leederville aquifer.

3.2. Degradation e high NDMA concentration stop-flow column experiment An NDMA half-life degradation rate of >100 days was previously reported (Patterson et al., 2010) in Leederville aquifer sediment at a concentration of 590 mg L1. In these high NDMA concentration experiments, with a column residence time of 39 days, there was insufficient NDMA removal within the column to provide a more accurate degradation half-life. To overcome this, the column water residence time was increased by ceasing the flow of influent water into the columns. Therefore, this high concentration column experiment was continued as stop-flow column experiment. During the stop-flow column experiment, water samples were collected from sampling ports at 4, 25, 46, 92, 125, 142, 158, 175, and 192 cm from the base of each column. Water samples from these same ports were also collected prior to ceasing water flow through the column. The average NDMA column residence time prior to ceasing water flow was 21 days. Column water samples were collected again at average column residence time of 158, 361, 507, and 675 days. NDMA concentration in all sampling ports during the stopflow column experiments are shown in Fig. 2. NDMA concentrations in the non-sterile column decreased over time at all sampling ports while NDMA concentrations in the sterile column remained relatively unchanged. The slower NDMA concentration decrease at the influent end of the column was possibly due to (i) diffusion of NDMA into the column from the 500 mL mixing chamber at the base of the column containing relatively high influent water concentrations of NDMA, or (ii) different geochemical conditions at influent end of the column as a result of aerobic water injection over a period of 12 months into a reductive pyritic sediment, or (iii) biodegradation competition with other trace organics from recycled water that were sorbed to sediment at the influent end of the column. Competitive biodegradation has been reported previously to decrease biodegradation of a chemical in the presence of other chemicals (Qiu et al., 2009;

Fig. 2 e NDMA concentrations along the sterile and nonsterile columns during stop-flow column experiment. Results are for the sterile column on day 21 (B), day 158 (,), day 361 (6), day 506 (>) and day 675 (7) and for the non-sterile column on day 21 (eCe), day 158 (e-e), day 361 (e:e), day 506 (eAe), and day 675 (e;e).

Stringfellow and Aitken, 1995). Therefore, due to these potential confounding additional processes at the influent end of the columns, data from ports at distance <50 cm from the influent end of the columns were not used during estimation of NDMA degradation half-life. Average NDMA concentration data for each sampling time (excluding data for ports <50 cm from the influent end of the column) were plotted against residence time (Fig. 3). Degradation rates were then determined by fitting a first-order halflife curve to the experimental data using Origin v7 software. As a first-order degradation profile could be fitted to the concentration data with a greater degree of confidence than a zero order profile, NDMA degradation was assumed to have first-order kinetics. Other authors (Gunnison et al., 2000;

Fig. 3 e Average NDMA concentrations versus column residence time for the sterile (,) and non-sterile (-) stopflow column experiment, along with fitted half-live curves.

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Fig. 4 e NDMA concentrations versus column residence time for the sterile (B) and non-sterile (-) columns, with fitted half-life curves. Data is for the low NDMA concentration column experiment at day 351.

Kaplan and Kaplan, 1985) have also found first-order kinetics better to describe NDMA degradation. First-order half-life curves fitted to the non-sterile and sterile column data gave an estimated half-life of 290
25 days for the non-sterile column, and 4300
3100 days for the sterile column.

3.3. Degradation e low NDMA concentration column experiment To determine the biodegradation rates, NDMA concentration data was plotted against column residence time (e.g. Fig. 4). Column residence time was calculated from the distance of the sampling ports along the column and the linear flow velocity of NDMA/bromide tracer through the columns. Fitting a first-order degradation profile to the non-sterile column data gave estimated half-life degradation rates of between 25 and 150 days of the 520 day experiment. As there was no trend of

decreasing half-lives with time, a biodegradation lag time or an increase in microbial activity over time could not explain the variability observed. The variation in half-lives during the experiment was possible due to a combination of the relatively slow degradation of NDMA with the relatively short column water residence time (51 days). As a result, effluent concentrations of NDMA were only reduced between 20 and 55% of the influent concentration. Increasing the residence time via a stop-flow experiment was not possible at these low NDMA concentrations, due to the amount of water required to be collected from a non-flowing column (50 mL per sampling port to achieve ng L1 detection limits). Minor losses were also observed in the control column. These losses may be attributed to analytical variability or slow degradation of NDMA in the sodium azide sterilized column as some microorganisms were persistent under azide treatment (Lichstein and Soule, 1943). Assuming slow degradation of NDMA, a degradation half-life between 160 and 240 days was estimated for the sterile column.

3.4.

14

C-NDMA microcosm experiments

The final distributions of 14C radioactivity from the anaerobic 160 day and 38 day microcosm experiments and aerobic 42 day microcosm experiment are given in Table 3. For the control microcosms, the percentage of 14C-NDMA remaining within the microcosm solution as either 14C-NDMA and/or soluble degradation products at the end of the experiment ranged between 70% and 90%. For the anaerobic experimental microcosms, 43% of the added 14C-NDMA remained within the microcosm solution as either 14C-NDMA and/or soluble degradation products at the end of the 38 day experiment and only 2% remained after 160 days. For the aerobic experimental microcosm, <1% of the added 14C-NDMA remained within the microcosm solution at the end of the 42 day experiment. Recoveries of 14CO2 and/or volatilized 14C-NDMA were 30% and 51% for the 38 and 160 day anaerobic experiments, and 33% for the aerobic experiment. No 14 C  HCO 3 formation (<1%) was observed in any of the experimental or control microcosms.

Table 3 e Final distribution of 14C radioactivity from the anaerobic microcosms after 38 and 160 days,a and aerobic microcosms after 42 days. Experimental conditions

Experiment duration (days)

Percentage NaOH trap solutionb 14

b

CO2 and/or Volatilised 14C-NDMA Anaerobic experimental Anaerobic control Anaerobic experimental Anaerobic control Aerobic experimental Aerobic control a b c d

38 38 160 160 42 42

14

30
8 5
1 51
3 15
2 33
6 4
1

C recovered Microcosm solution

14

C-NDMA/soluble degradation productsc 43
13 87
7 2
1 70
2 <1 90
4

Recoveries are given as the percentage of radioactivity initially added as 14C-NDMA. Recovered from NaOH trap. Recovered from acidified microcosm solution at the end of the experiment. Difference in counts in NaOH trap before and after acidification of microcosm solution.

CeHCO3d

14

<1 <1 <1 <1 <1 <1

14

Cetotal 73 92 53 85 33 94

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Mass balance recoveries for the sterile control microcosms ranged from 85% to 94%. For the non-sterile microcosms, mass balance recoveries ranged from 33% to73%. The unaccounted radioactivity (27e47 %) in the anaerobic and (67%) aerobic microcosms may have been due to the formation of other degradation compounds that were not monitored. Potential non-monitored degradation compounds could have been sorbed to the sediment in the microcosms, taken up as bacterial biomass or volatilized (e.g. methane) and not collected in the NaOH trap. Bradley et al. (2005) observed up to 41% methane produced from NDMA degradation in anoxic microcosms and suggested direct microbiological demethylation of NDMA or methanogenic biodegradation of NDMA and/or its degradation products. In aerobic microcosms, Szecsody et al. (2008) observed 51% carbon dioxide produced with the remainder as aqueous species (18%), sorbed to sediment (2%), incorporated or sorbed to microbes (6.4%) or unaccounted (20%). To compare 14C-NDMA microcosm experiment data to column experimental data, 14C-NDMA loss was calculated rather than 14C-NDMA mineralisation to 14CeCO2. 14C-NDMA concentration data over time were calculated from 14CeCO2 formation data from the NaOH trap, scaled to the final 14CNDMA concentration in the microcosms at the end of the experiment. Plots of 14C-NDMA concentration versus time for the 160 day and 38 day anaerobic microcosm experiments and 42 day aerobic microcosm experiment are shown in Fig. 5. Also shown are half-life curves fitted to the experimental data. Estimated half-life degradation rates were 29
2 and 46
7 days for the 38 and 160 day anaerobic experiments and 8
2 days for the aerobic experiment. For the control microcosms, half-life degradation rates were between 220 and 260 days. The minor losses in 14C-NDMA observed in the control microcosms were possibly due to slow volatilisation of 14CNDMA from the microcosm solution (Henry’s Law constant of

0.143 atm m3 M1 at 25  C; Suthersan, 1997) and partitioning into the NaOH trap solution over time. For the aerobic non-sterile microcosm, the initial limited NDMA degradation followed by rapid degradation enabled the determination of a microbiological lag time (time before NDMA degradation commenced). For this microcosm, a lag time of approximately 10 days was observed (Fig. 5B). This data would suggest that there were bacteria present within the microcosm at the start of the experiment capable of acclimation to aerobic biodegradation of NDMA. For the anaerobic non-sterile microcosms, a lag time was not apparent, possibly due to the slower degradation rates observed in these microcosms.

3.5.

Mechanism for removal of NDMA

The substantial half-life differences (Table 4) between sterile and non-sterile (i) stop-flow experiments, (ii) low NDMA concentration column experiments, and (ii) 14C-NDMA mineralisation microcosm experiments provides strong evidence for the degradation of NDMA via biodegradation. The Leederville sediment has a high reductive capacity (Descourvie`res et al., 2010) and a rapid oxygen consumption rate of 2.0 mg L1 h1 (Patterson et al., 2010). Therefore, dissolved oxygen from aerobic recycled water would be consumed rapidly and aerobic conditions would only be present at the inlet of the columns for a short period of time (<24 h). As NDMA was present in low NDMA concentration columns for greater than 24 h, the potential mechanism for removal of NDMA in all column experiments would be through anaerobic degradation. Biodegradation of NDMA under anaerobic conditions was reported via a denitrosation pathway in which NDMA was converted to the intermediate dimethylamine (DMA) and nitrite (Rowland and Grasso, 1975). Szecsody et al. (2008) also detected DMA as a transient degradation product of NDMA in

Fig. 5 e 14C-NDMA concentration (calculated from 14CeCO2 formation and scaled to the final 14C-NDMA concentration in the microcosms at the end of the experiment) over time for A) the 38 and 160 day anaerobic microcosm experiments and B) aerobic microcosm experiment. Also included are fitted degradation half-life curves.

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Table 4 e Half-live degradation rates determined from column and Experiment

a

High NDMA Conc. Column High NDMA Conc. Stop-flow column Low NDMA Conc. Column Anaerobic 14C-NDMA Microcosm Anaerobic 14C-NDMA Microcosm Aerobic 14C-NDMA Microcosm

14

C-NDMA microcosm experiments.

NDMA initial concentration (ng L1)

Water residence time (days)

590,000 590,000 200 560 560 560

39 675 51 38 160 42

NDMA degradation half-life (days) Non-sterile

Sterile

>100 290
20 25e150 29
2 46
7 8
2

>100 4300
3100 160e240 220
30 260
40 260
30

a Patterson et al. (2010).

a reduced sediment. Bradley et al. (2005) also showed NDMA degradation under anaerobic soil conditions with stoichiometric formation of carbon dioxide and methane. An alternative reductive pathway has also been reported (Grilli and Prodi, 1975) where N,N-dimethylhydrazine was formed as an intermediate. Therefore, it is likely that the degradation of NDMA in the reductive anaerobic Leederville sediment is through a reductive pathway similar to denitrosation or hydrazine with the potential transient formation of DMA and final products of carbon dioxide and methane. An attempt was made to analyse DMA in the columns of the high NDMA concentration stop-flow column experiment at day 675. However, DMA concentrations in all sampling ports of nonsterile and sterile columns were below the detection limit (5 mg L1). The failure in detecting DMA in the column water may suggest an alternative reductive degradation pathway or the reduction of NDMA into DMA was likely the limiting reaction in the NDMA degradation resulting in no transient accumulation of DMA. Electron donation for this reductive degradation from DOC was unlikely due to the low DOC concentration of (i) the influent recycled water used (<1 mg L1), and (ii) groundwater from the site of sediment collection (1.2 mg L1). This is consistent with Nalinakumari et al. (2010), who suggested that the RO treated water did not provide sufficient primary substrate to support the degradation of NDMA. However, electron donation was possibly from the oxidation of SOM (0.32% w/w) or oxidation of reduced minerals such as pyrite (2% w/w) present in the Leederville sediment. Evidence for pyrite oxidation included the rapid oxygen consumption (8 mg L1 to < 2 mg L1) and increased sulphate concentration (from <2 mg L1 to 10 mg L1) along the first 20 cm of the sterile and non-sterile columns of the low NDMA concentration experiment. Also, as the recycled water was not buffered prior to injection, the mineral dissolution observed (Patterson et al., 2010) may possibly provide trace nutrients to enhance NDMA degradation. Under aerobic conditions, an NDMA hydroxylation degradation pathway was proposed by Harder and Attwood (1975). Szecsody et al. (2008) also showed aerobic NDMA degradation with 51% of converted to carbon dioxide via a monooxygenase enzyme pathway. For the field trial, aerobic degradation may occur close to the recharge bore during injection of aerobic recycled water. Under these conditions, sulphide oxidizing bacteria responsible for pyrite oxidation may also play a role in NDMA biodegradation.

3.6. Comparison of degradation rates at high and low concentration In these experiments, NDMA was degraded at both low (ng L1) and high (mg L1) concentrations. Half-life degradation rates (Table 4) at the ng L1 concentrations investigated (200e560 ng L1), were comparable for the 14C-NDMA anaerobic microcosm experiments (half-life 29e46 days) and the low NDMA concentration column experiment (half-life 25e150 days). There was a of 2e12-fold increase in the degradation half-live (half-life 290
20 days) when the NDMA concentration was increased 3 orders of magnitude, which indicated that degradation rate of NDMA was concentration dependant in the Leederville aquifer over ng L1 to mg L1 concentration range. Concentration inhibitory effects on NDMA degradation have previously been reported (Gunnison et al., 2000; Yang et al., 2005; Szecsody et al., 2008,). NDMA rate constant data measured by Kaplan and Kaplan (1985) indicated 4-fold higher half-life when the initial concentration was increased to 6 order of magnitude from 0.01 mg to 10,000 mg per gram soil (see Table 1). Similarly, rate constants of NDMA mineralisation in experiments undertaken by Gunnison et al. (2000) indicated 3-fold increase in half-life when the initial concentration of NDMA was increased to 3 orders of magnitude (50 mg L1 to 50 mg L1). Literature degradation rates for NDMA (Table 1) were variable for similar concentrations, suggesting that degradation rates were likely to be site specific. The half-life of NDMA in anoxic (mild denitrifying conditions) column experiments at 120 ng L1 was reported to be 34 h (Drewes et al., 2006) while in our experiments, under more reductive anaerobic conditions at 200 ng L1, the half-life was substantially greater (25e150 days). Different rates of NDMA degradation in different types of soil could be due to different soil activity as showed by Yang et al. (2005). Bradley et al. (2005) attributed biodegradation of NDMA to common soil microorganisms since it was degraded in both NDMAimpacted soil and soil with no history of NDMA. However, in the experiment conducted by Drewes et al. (2006), it was found that soil with a long history of organic contamination gave a shorter half-life. They suggested that long periods of contamination may have resulted in the development of a greater diversity of microorganisms which seemed to enhance NDMA degradation. Based on the data from Drewes et al. (2006), NDMA degradation rates in the Leederville aquifer may increase with time, if the naturally occurring

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microorganisms were exposed to NDMA contamination for a greater length of time. The effect of redox conditions on NDMA degradation were based on 14C-NDMA microcosm studies with the same initial NDMA concentration. Aerobic microcosm experiments showed substantially faster degradation rates (half-life 8
2 days) than anaerobic experiments (half-life 29e46 days). These results are in contrast to Gunnison et al. (2000) who observed no substantial differences in half-lives under aerobic and anaerobic conditions (Table 1), further suggesting that degradation rates were likely to be site specific. No threshold effects were observed at the lower ng L1 concentrations with NDMA concentrations reduced from 560 ng L1 to <6 ng L1 during the 42 day 14C-NDMA aerobic microcosm experiment, and from 560 ng L1 to 10 ng L1 during the 160 day 14C-NDMA anaerobic microcosm experiment. These results were consistent with the observations by Kaplan and Kaplan (1985).

3.7.

Field assessment

Average NDMA concentrations of the influent recharge recycled water during the field trial was 2.5
1.0 ng L1 (n ¼ 15). Average groundwater concentrations (post-breakthrough of the recharge water) at 20, 60 and 120 m from the recharge bore were 1.3
0.4 ng L1 (n ¼ 25), 1.2
0.4 ng L1 (n ¼ 24), and 1.3
0.5 ng L1 (n ¼ 12). There was a significant (student t test, P < 0.0001) decrease in NDMA over the first 20 m (between recharge bore and 20 m monitoring location) with NDMA concentrations reducing to approximately half the original recharge concentration. Based on the recycled water breakthrough data, an aquifer residence time of 10 days was estimated between the recharge bore and 20 m monitoring location. This would give an estimated degradation half-life of approximately 10 days. As the influent recharge recycled water was aerobic, the reduction in NDMA over the first 20 m could have been via aerobic biodegradation. Further away from the recharge bore, the aquifer was anaerobic and reduction in NDMA could be via anaerobic biodegradation. However, between 20 and 120 m no substantial reduction in NDMA was observed. But as NDMA concentrations between 20 and 120 m were close to the analytical detection limit (<1 ng L1), assessment of NDMA degradation was problematic. Therefore, anaerobic NDMA degradation could not be confirmed at the field scale.

4.

Conclusion

Results of the laboratory and field investigations indicated that recharging the Leederville aquifer with NDMAcontaminated aerobic recycled water is likely to result in NDMA biodegradation during aquifer passage. Based on aerobic 14C-NDMA microcosm studies and field observations, rapid aerobic biodegradation of NDMA (half-life of 8
2 days) should occur in the aerobic zone close to the recharge bore. Further away from the recharge bore in the anaerobic zone of the aquifer, slower biodegradation rates (half-life of 25e150 days) via reductive degradation or co-metabolism should occur, based on anaerobic 14C-NDMA microcosm studies and

low concentration column studies. However, this could not be confirmed based on field observations due to low concentrations between 20 and 120 m from the recharge bore. Concentration inhibition of anaerobic biodegradation was observed with degradation half-lives up to an order of magnitude greater at mg L1 NDMA concentrations compared to ng L1 concentrations. No threshold effect was observed at the lower ng L1 concentrations with NDMA concentrations reduced from 560 ng L1 to <6 ng L1 during the 42 day 14 C-NDMA aerobic microcosm experiment, and 560 ng L1 to 10 ng L1 during the 160 day 14C-NDMA anaerobic microcosm experiment. Therefore, given a sufficient aquifer residence time or travel distance between recycled water injection and groundwater extraction, NDMA (if present in the recycled water) should be degraded naturally during aquifer passage. Transferability of these finding to predict the potential for other nitrosamines to biodegrade within the Leederville aquifer is problematic, as the mechanism and bacteria responsible for biodegradation may be different. However, Pitoi et al. (2011) demonstrated N-nitrosomorpholine biodegradation under the same anaerobic geochemical conditions. Therefore, it is plausible that other nitrosamines which are capable of biodegradation via a reductive degradation mechanism may also biodegrade within the Leederville aquifer.

Acknowledgements This research was made possible through funding from CSIRO Water for a Healthy Country Flagship Program and the Water Corporation of Western Australia. The field trial was supported by the Australian Government’s Water for the Future initiative through the Water Smart Australia program.

references

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