Sulfamate in environmental waters

Science of the Total Environment 695 (2019) 133734

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Science of the Total Environment journal homepage: www.elsevier.com/locate/scitotenv

Sulfamate in environmental waters D.R. Van Stempvoort a,⁎, J. Spoelstra a,b, S.J. Brown a, W.D. Robertson b, R. Post c, S.A. Smyth d a

Watershed Hydrology and Ecology Research Division, Environment and Climate Change Canada, 867 Lakeshore Road, Burlington, Ontario L7S 1A1, Canada Department of Earth and Environmental Sciences, University of Waterloo, 200 University Avenue West, Waterloo, Ontario N2L 3G1, Canada c Nottawasaga Valley Conservation Authority, 8195 8th Line, Utopia, Ontario L0M 1T0, Canada d Emerging Priorities Division, Environment and Climate Change Canada, 867 Lakeshore Road, Burlington, Ontario L7S 1A1, Canada b

H I G H L I G H T S

G R A P H I C A L

A B S T R A C T

• Information about the occurrence and fate of sulfamate (sulfamic acid) in the environment is lacking. • We addressed this science gap by analyzing sulfamate in samples of precipitation, surface water, groundwater and wastewater. • Sulfamate was detected in most environmental water samples at concentrations ranging up to 128,000 ng/L. • Contextual information obtained in this study suggests that sulfamate detected in environmental waters is derived from various sources.

a r t i c l e

i n f o

Article history: Received 30 May 2019 Received in revised form 31 July 2019 Accepted 1 August 2019 Available online 02 August 2019 Editor: Damia Barcelo Keywords: Sulfamate Sulfamic acid Wastewater tracer Surface water Groundwater Ion chromatography

a b s t r a c t Although sulfamate (the anion of sulfamic acid) has been in use for decades in various industrial and other applications, there is no previously published information about its occurrence and fate in environmental waters. In this study sulfamate was widely detected in environmental waters in Ontario, Canada, ranging up to 128,000 ng/L. It was always detected (N100 ng/L) in bulk precipitation samples and streams, it was usually detected in samples of lake water, and often detected in groundwater. Spatial and temporal variations suggest that both widespread atmospheric deposition and localized land-based anthropogenic sources of sulfamate may be important. Lower concentrations or non-detections of sulfamate in waters that had relatively low dissolved oxygen (e.g. some groundwaters) suggest that sulfamate may be degraded in the environment under suboxic or anoxic conditions. Given our findings of a wide distribution of sulfamate in environmental waters, including precipitation, it is not likely to be very useful as a wastewater tracer. Crown Copyright © 2019 Published by Elsevier B.V. All rights reserved.

1. Introduction

⁎ Corresponding author. E-mail address: [email protected] (D.R. Van Stempvoort).

https://doi.org/10.1016/j.scitotenv.2019.133734 0048-9697/Crown Copyright © 2019 Published by Elsevier B.V. All rights reserved.

Recently, Castronovo et al. (2017) observed sulfamate (reported as “sulfamic acid”, see Fig. 1 and text below) in municipal wastewater effluent, and suggested that it was a promising candidate as a “wastewater tracer”. They proposed that it might be a more appropriate

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Fig. 1. Top row shows sulfamic acid (left), which readily dissociates in water to form sulfamate (right) (pKa = 1.0 ± 0.1: Williams, 2004). Bottom row shows three artificial sweeteners that have structures related to sulfamate.

wastewater tracer than the artificial sweetener acesulfame, of which it is a degradation product (Fig. 1). However, there are other potential sources of sulfamate in wastewater and in various environmental media. For example, since its introduction as an industrial chemical (Cupery, 1938), sulfamic acid and its derivative salts have been widely used for many purposes including as an industrial cleaning agent (Metzger, 2012), a herbicide (ammonium sulfamate: Kamrin and Montgomery, 1999), and as a fireproofing agent (ammonium sulfamate: Lewin et al., 2002). Furthermore, some laboratory studies have suggested that sulfamate and/or sulfamic acid may form in the atmosphere by reaction of ammonia (NH3) and sulfur trioxide (SO3), particularly in polluted air masses (Shen et al., 1990; Lovejoy and Hanson, 1996; Larson and Tao, 2001; Pawlowski et al., 2003; Pszona et al., 2015) including flue gases (Hirota et al., 1996). There are only a few known biosynthesized organic compounds that contain nitrogen‑sulfur bonds, and the majority of these are classified as sulfamates (i.e. sulfamate is the functional group containing these N–S bonds) (Petkowski et al., 2018). However, whether the degradation of these compounds in nature releases sulfamate to the environment is a knowledge gap. Sulfamic acid (H3NSO3), which is sometimes referred to as − amidosulfonic acid, can be described as a zwitterion (NH+ 3 ·SO3 ) that is readily soluble in water (Metzger, 2012), where it dissociates to form the anion sulfamate (NH2SO− 3 ) (Fig. 1). Similarly, sulfamate salts, including ammonium sulfamate, generally have very high solubilities in water. Hence, most references in this paper are to sulfamate rather than sulfamic acid. Given its extensive production and broad use (including sulfamic acid) for more than seventy years, the fact that it probably forms “naturally” in the atmosphere, and its high solubility in water, sulfamate potentially occurs widely in the environment. Castronovo et al. (2017) found similar concentrations of sulfamate in influent and effluent wastewater indicating that it is relatively stable during municipal wastewater treatment and thus expected to be present in receiving environmental waters. However, there is virtually no published information about concentrations of sulfamate in environmental media. The objective of this paper is to address this knowledge gap by reporting what are, to our knowledge, the first published concentrations of sulfamate in environmental waters. The field sites that were included in this study (Fig. 2) are located in the Province of Ontario in Canada at sites where the authors are currently conducting ongoing aquatic chemistry research and thus could obtain samples readily. We also

include analyses of sulfamate in samples of municipal and septic wastewater as contextual information. 1.1. Previous relevant research Based on laboratory experiments, Scheurer et al. (2012) reported that sulfamate (“amidosulfonic acid”) was one of the main products that formed during oxidation of the artificial sweetener cyclamate by ozone, and sulfamate was also detected as an oxidation product during similar “ozonation” experiments with acesulfame. In subsequent experiments in which the artificial sweetener acesulfame was added to activated sludge from a wastewater treatment plant, Castronovo et al. (2017) reported that sulfamate was the “predominant transformation product” of the degradation of acesulfame. Similarly, in experiments with enrichment cultures of microorganisms derived from wastewater, Kahl et al. (2018) also observed formation of sulfamate associated with degradation of acesulfame. Taking into consideration chemical structures, there is at least one more artificial sweetener that is a potential source of sulfamate via degradation: saccharin (Fig. 1). Despite some evidence that sulfamate appears to be persistent during treatment of municipal wastewater (Castronovo et al., 2017), laboratory studies have indicated that some microorganisms are capable of degrading sulfamate (Rein and Cook, 1999; Fulton and Cooper, 2005; Linder, 2012). Bodek and Smith (1980) reported an ion chromatography method for analyzing ammonium sulfamate in air, but to our knowledge no published data for actual air samples have been reported. Pinto et al. (2016) recently reviewed methods to analyze sulfamate as an impurity in pharmaceutical products, and some have measured it as a product of the degradation of the drug topiramate (Li and Rossi, 1995; Pinto et al., 2019). Castronovo et al. (2017) reported an IC-MS/MS method to analyze sulfamate, which they applied to an investigation of wastewater as described above (along with LC/MS/MS for quantitative analysis of acesulfame and IC/QTOF for the identification of the transformation products). 2. Materials and methods 2.1. Field sites The field sites for this study are located in three areas in Ontario, Canada (Fig. 2):

D.R. Van Stempvoort et al. / Science of the Total Environment 695 (2019) 133734

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Fig. 2. Field sampling locations in Ontario, Canada. The red circles on map (b) indicate specific locations where surface water or groundwater samples were collected as mentioned in the text. The locations of lakewide samples from Lake of the Woods are not shown. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

2.1.1. Upper Thames basin Two rural headwater catchments (Cedar Creek, 9515 ha; Nissouri Creek, 3090 ha) in southern Ontario are dominated by agricultural use (Table 1), which includes dairy operations, other livestock, (largely hogs, poultry) and mixed crops (largely corn, soy beans, grain, hay) (Ontario Ministry of Agriculture, Food and Rural Affairs, 2018). The small headwater streams and constructed drains in these catchments do not receive influxes of treated municipal wastewater, but one of the stream subcatchments flows through a small urban area in which the residences have private septic systems to treat wastewater. Detailed information about land use, hydrology and other characteristics of these catchments has been provided by Ontario Ministry of the Environment (2012) and Upper Thames River Conservation Authority (2017). 2.1.2. Lake of the Woods basin This 7,003,000 ha basin consists largely of undeveloped forest, wetlands and lakes. Locally there is some urban development (e.g., city of Kenora), mining and recreational developments (cottages and resorts) and some agricultural land use. Lake sampling occurred at lakewide monitoring locations, and also from research sites that have recently been established at Keewatin (Kenora), Poplar Bay and Sioux Narrows.

A few stream samples were collected in the Rainy River sub-basin, a catchment that includes agricultural land, a recently developed mine, forest and wetlands. Agricultural land uses in the Canadian portion of the Rainy River sub-basin include cropland (hay, grains) and livestock production (beef, sheep) (Ontario Ministry of Agriculture, Food and Rural Affairs, 2018). Detailed information about the hydrology and other characteristics of the basin has been provided by De Sellas et al. (2009) and Pascoe et al. (2014).

2.1.3. Minesing wetlands Designated as a wetland of international importance by the Ramsar Convention on Wetlands, this undeveloped area spans over 6000 ha. The eastern portion of the wetlands, conifer swamps and fen complex is sustained by groundwater discharge and is relatively isolated from the surface water influences that dominate the remainder of the wetland. It is surrounded by farms which are largely cash crops such as corn, soybeans, wheat and hay (Ontario Ministry of Agriculture, Food and Rural Affairs, 2018) with limited livestock operations. Bradford (1999) and Post et al. (2018) have provided further details on the physiography, biology and hydrology of the wetlands.

Table 1 Land use in the catchments of the ten streams that were sampled in the Upper Thames basin. Catchment

Area (km2)

% agriculture and undifferentiated rural

% community/infrastructure

% sand, gravel extraction, mine

% forest, hedges

% wetlands

% open water

CC-SW-Aa CC-SW-Ba CC-SW-Ca CC-SW-Da CC-SW-Ea NC-SW-Ab NC-SW-Bb NC-SW-Cb NC-SW-Db NC-SW-E

83.7 2.7 10.7 1.8 5.3 2.3 8.5 2.7 23.7 1.6

77.33 85.97 79.25 61.05 78.80 83.47 75.74 59.22 81.02 77.58

6.94 1.73 9.49 1.74 2.64 2.64 2.37 1.53 2.51 0.04

1.01 0.00 0.09 0.00 0.00 0.00 0.00 0.00 0.00 0.00

4.71 4.19 7.30 7.44 5.61 8.07 7.39 8.40 9.09 19.41

9.12 7.60 2.69 29.67 12.95 5.83 14.18 30.50 7.12 2.98

0.17 0.00 0.12 0.00 0.00 0.00 0.00 0.00 0.05 0.00

a b

Catchments CC-SW-B, -C, -D, and –E are subcatchments of CC-SW-A; catchment CC-SW-D is a subcatchment of CC-SW-E. Catchments NC-SW-A, -B, and -C are subcatchments of NC-SW-D.

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2.2. Field sampling methods Environmental water samples were collected between May 2017 and January 2019 in the three field areas. Samples of precipitation, surface water and groundwater were collected in polyethylene bottles, immediately placed in coolers for transportation to the laboratory. Most samples were filtered immediately in the field (0.45 μm membranes), were frozen within 48 h of collection, and remained frozen until the day analyzed. Exceptions are indicated below. At two sites in the Upper Thames basin, bulk precipitation samples were obtained at 2 to 4 week intervals in collectors open to the atmosphere (Palmex Rain Sampler RS1, described by Gröning et al., 2012). These samples have likely been impacted by both wet (precipitation) and dry atmospheric deposition. Following their collection, these precipitation samples were filtered in the laboratory (0.45 μm membranes). Rinses indicated minor to negligible contribution of sulfamate from the precipitation collection equipment to the samples (Fig. S2.1 in Supporting Information). Surface water was sampled from ten headwater streams (including constructed drains) in the Upper Thames basin (2 to 4 week intervals), from the Lake of the Woods (excursions in May, June, September and October 2017; May, June, August, September and October 2018), and from a stream in the Lake of the Woods basin (one day event in October 2017). Lakewide monitoring samples were collected from Lake of the Woods at two discrete depths, 1 m below surface and two meters from the bottom, using a Van Dorne or Niskin sampler, as described by Pascoe et al. (2014). The rest of the samples were collected as nearsurface (b 1 m) grab samples using a pole sampler. A few of the Lake of the Woods samples (urban nearshore and Poplar Bay) were collected at depths N1 m below surface using a Niskin sampler. Most of the surface water samples were filtered immediately in the field, lakewide samples collected at Lake of the Woods were filtered within 24 h of collection (0.45 μm membranes). Peristaltic pumps were used to collect groundwater samples from shallow depths (usually b2 m below ground); most were obtained using the drive point methodology described by Roy and Bickerton (2010). In the Upper Thames basin these drive points were installed at eight locations in riparian zones along headwater streams and drains and were maintained in place for the duration of the study for sampling at two week intervals. In the Lake of the Woods basin groundwater was sampled by shallow drive points on an event basis along an urban shoreline (two days in May 2017), along the shore of Poplar Bay (October 2018), and along the shore of Lake of the Woods at Sioux Narrows (October 2018). Other groundwater samples from the Lake of the Woods basin were collected at shallow depths (0.4 to 2.1 m) from monitoring points installed with the use of a soil auger in the vicinity of septic systems on one or two events (October 2017, September 2018). Shallow groundwater samples were also obtained from a network of piezometers in the Minesing Wetlands (May and November 2017); further details on this network are given by Bradford (1999) and Post et al. (2018). Some of the surface water and groundwater samples from Lake of the Woods were stored on ice in coolers for several days (bweek) before transfer to frozen storage in the laboratory. 2.3. Sampling of municipal and septic wastewater Composite samples of both raw (influent) wastewater and final effluent were obtained between April and October 2018 at fourteen municipal wastewater treatment plants across Canada. The sampling method, described in detail by Guerra et al. (2015), involved the use of HACH Sigma 900 refrigerated autosamplers which were programmed to collect 200 mL every 15 min over 24 h to generate each composite sample. For samples collected close to our laboratory, a subsample was filtered (0.2 μm) and transferred to a polyethylene bottle and frozen until analysis. For other samples, on the day they were collected

20 mL subsamples were filtered (sterile SFCA 0.2 μm) and collected in glass vials, mixed with 5.5 g of isopropanol (for preservation), and refrigerated until analysis. Testing indicated no difference in the results of analyses of sulfamate and artificial sweeteners for these two storage methods (results not shown). A peristaltic pump was used to collect septic wastewater samples from septic tanks at five locations (three cottages, one visitor centre, one park campground) near the shores of Lake of the Woods in August and September 2018. The sample bottles, field-filtering, sample handling and storage procedures were as described above for the environmental water samples. 2.4. Analytical methods The instruments used to analyze sulfamate and artificial sweeteners were a Dionex (Sunnyvale, CA, USA) 5000 ion chromatography (IC) system coupled to a QTRAP 5500 (AB Sciex, Concord, ON, CAN) triplequadrupole mass-spectrometer, operated in negative electrospray ionization (ESI) mode. Injection volume was 100 μL onto a Dionex IONPAC® AS20 analytical column (2 × 250 mm) with a KOH eluent gradient. Two multiple reaction monitoring (MRM) transitions were monitored for sulfamate; quantification MRM m/z 96–80 and qualification MRM m/z 96–64. The declustering potential was −60 and the collision energy was set to −26 for the quantification MRM and − 61 for the qualification MRM. The retention time (r.t.) for sulfamate was 4.8 min and the deuterated internal standard acesulfame (m/z 166–86) area response (r.t. 10.8 min) was used to correct for instrument variability and some sample matrix. Sulfamate (r.t. 4.8 min) eluted 0.8 min before chloride (r.t. 5.6 min). The recovery of 1000 ng/L of sulfamate was 110.3% in a matrix of 100 mg/L chloride. At chloride concentrations N100 mg/L the recovery of sulfamate could not be adequately corrected by the acesulfame internal standard. The minimum detection limit was 25 ng/L. A quadratic regression with 1/x2 weighting was applied and the r2 was N0.999. The mean and % standard deviation of a 125 ng/L standard injected 3 times over a 48 h period of a sample run was 119.7 ng/L and 2.7%. Further instrument details and operating parameters are provided in the Supporting Information. The sweetener analysis has been described previously in Van Stempvoort et al. (2011b) and is also included in the Supporting Information. The sweetener detection limits were as follows: acesulfame 2 ng/L; saccharin 2 ng/L and cyclamate 3 ng/L. Sulfamate standards were made by dissolving Sulfamic Acid (reagent grade 99.3%, Sigma Aldrich, St. Louis, MO) in Milli-Q water and diluting with Milli-Q water to the appropriate concentration. To determine sulfamate storage stability by freezing, a set of 30 samples from Lake of the Woods was collected in 2 bottles. One was stored frozen for 12 days before analysis and the other 419 days before analysis. The samples analyzed on day 419 ranged from a low sulfamate recovery of 74.6% to a high recovery of 113%. The median sulfamate recovery was 94.0% after 419 days of frozen storage. In this sample set the median recovery of acesulfame was 117.1%. Cyclamate and saccharin were not present at high enough concentrations to access. Previous frozen storage tests with cyclamate and saccharin from 227 days to 6 years resulted in recoveries ranging from a low of 69% to a high of 118% (Table S1.3 in Supporting Information). For each sampling event, additional samples of surface water and groundwater samples were collected for laboratory analyses of anions by ion chromatography (Dionex 2500 system with a Dionex AS18 4 mm analytical column), major cations and metals by inductively coupled plasma - optical emission spectrometry (HORIBA Jobin Yvon ULTIMA 2), and ammonium (spectrophotometry using a salicylatenitroprusside colorimetric method, absorbance measured at 640 nm). When measured, dissolved oxygen was determined in the field; the majority of these measurements were by an optical dissolved oxygen sensor (YSI ProDSS Multiparameter instrument; YSI Incorporated, Yellow Springs, OH).

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3. Results and discussion

ha·year:

3.1. Upper Thames basin

ψt ¼ cwþd  pa

3.1.1. Precipitation data Over the period August 2017 to January 2019, sulfamate concentrations in bulk precipitation (wet plus dry) in the two rural Upper Thames catchments ranged from 185 to 2583 ng/L (Fig. 3; Table 2). These results indicate that atmospheric deposition plays an important role in transporting sulfamate to surface environments, which may then impact surface water and groundwater. The consistent presence of sulfamate in all precipitation samples suggests that this compound may have formed in the atmosphere, as suggested by some previous studies (e.g., Larson and Tao, 2001; Pawlowski et al., 2003; Pszona et al., 2015). Although seasonal trends of sulfamate concentrations in precipitation were not readily apparent (Fig. 4), two sample t-tests indicated significantly lower concentrations in summer compared to winter and spring (Table S2.1 in Supporting Information). The reason for this is unknown. There was a weak inverse correlation (Pearson correlation = −0.276) between the volume of precipitation collected (per sample) and the sulfamate concentration (Fig. S2.2 in Supporting Information) that was statistically significant (p-value = 0.044), suggesting that washout (precipitation scavenging) played a role (Atlas and Giam, 1988). A weak correlation between sulfamate and saccharin in precipitation samples (Table S2.2 in Supporting Information) suggests that some of the sulfamate is derived from saccharin or from a source common to both (cf. discussion of Minesing Wetlands groundwater).

3.1.2. Estimated atmospheric deposition rate based on precipitation (wet plus dry) data The volume weighted mean concentration of sulfamate in precipitation (wet plus dry) based on all 2017–19 samples from the Upper Thames basin was 727 ng/L. Based on 1981–2010 Climate Normals for the nearest station (Woodstock, Ontario, average annual precipitation of 0.969 m; located 6–17 km from the two precipitation sampling sites), using Eq. (1), the estimated annual total atmospheric deposition (ψt, wet plus dry) of sulfamate in the Woodstock area is 7040 mg/

5

ð1Þ

where cw+d is the volume-weighted sulfamate concentration in the wet plus dry (open) collector (ng/L = μg/m3), and pa is the average annual precipitation (m), and given that 1 μg/m2·year = 10 mg/ha·year. This equation assumes that the volume-weighted mean concentration in samples from each wet plus dry collector was representative for inferring the annual rate of total (wet plus dry) atmospheric deposition of sulfamate at the collection site. As an alternative approach, we used our precipitation volume data for the entire period of collection to calculate the loading on an annual basis: ψt ¼



mtot−pc =acoll  ðdcoll =365Þ

ð2Þ

where mtot-pc is the total mass (mg) of sulfamate measured in precipitation at each site over the entire collection period (using our concentration and precipitation volume data), acoll is the area of the precipitation collection (ha), and dcoll is the number of days for the entire precipitation collection period. Using this approach, we calculated the annual total atmospheric deposition (ψt, wet plus dry) of sulfamate to be 6450 and 10,550 mg/ha·year at the two precipitation sites. 3.1.3. Stream data In the ten headwater streams (including constructed drains) sampled every two weeks in the rural Upper Thames basin catchments from June 2017 through March 2018, the sulfamate concentrations ranged over more than two orders of magnitude from 144 to 68,107 ng/L (Fig. 3, Table 2). The ubiquity and overlap with the precipitation concentrations (Fig. 3) suggests that atmospheric deposition was an important source of the sulfamate found in the streams. In contrast, the highest concentrations in the streams are almost two orders of magnitude higher than those observed in precipitation (Fig. 3), which does not seem to be compatible with precipitation as the sole source of sulfamate in these samples. These samples must have been affected by localized additional sources of sulfamate. As discussed below, the use

Fig. 3. Box and whisker plots of sulfamate concentrations measured in environmental water samples. Also shown for comparison are concentrations in septic wastewater samples (5 septic tanks from shoreline developments at Lake of the Woods) and samples from 14 municipal wastewater treatment plants across Canada (locations not shown). The whiskers indicate the full range of data; boxes indicate the range for the second and third quartiles of the data, medians for each are shown as a solid black line in the boxes. Note method detection limit (mdl = 25 ng/L); for this plot, non-detections were assigned values of 10 ng/L.

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Table 2 Sulfamate concentrations in selected environmental waters, Province of Ontario, Canada. Also, for comparison, sulfamate in septic wastewater (5 systems located near shoreline of Lake of Woods) and in municipal wastewater (14 wastewater treatment plants across Canada, locations not shown). Type of sample Environmental waters Upper Thames watersheds Precipitation Surface water Riparian groundwater Lake of the Woods Basin lake samples Lakewide Urban nearshore Sioux Narrows area Poplar Bay Other Lake of the Woods Basin samples Stream samples Urban storm drain Urban nearshore shallow groundwater Groundwater in vicinity of septic systems Other shallow groundwater Minesing Wetlands Shallow groundwater Wastewater Septic wastewater (tanks) Municipal wastewater Raw influent Final effluent

Number of samples

Mean (ng/L)a

Standard Deviation (ng/L)a

Minimum (ng/L)

Maximum (ng/L)

54 215 192

842 5513 405

547 9306 1254

185 144 n.d.

2583 68,107 12,174

140 40 22 88

606 1485 178 139

298 2558 42 232

n.d. 467 116 n.d.

1277 17,177 232 1109

9 1 10 33 13

23,998 7969 1729 5175 701

46,478 – 2134 4220 736

269 – 85 n.d. 104

128,140 – 6414 15,127 2534

84

1005

7138

n.d.

65,076

10

3890

2668

1273

9536

42 42

87,864 76,641

209,652 133,989

3477 10,656

1,075,826 721,687

“n.d.” = Less than the method detection limit of 25 ng/L. a mean and standard deviation values were calculated assuming that all non-detections (bmdl) had concentrations of 10 ng/L (mdl = 25).

of hydrologic data provides further evidence for non-atmospheric sources of sulfamate in these catchments. Though there is considerable scatter in the stream sulfamate data (Fig. 5), when all samples were grouped by seasons, there were significant seasonal differences, with significantly higher sulfamate concentrations in the fall and winter compared to spring and summer (Fig. 5; Table S2.1 in Supporting Information). This seasonality may be related to similar seasonal trends in concentrations of sulfamate in precipitation (summer low), and in groundwater (Fig. 5; Table S2.1 in Supporting Information) as discussed below. Perhaps the periods of highest sulfamate concentrations in surface water are related to seasonal increases in land-based sources of sulfamate, for example seasonal application of manure to fields. However, manure application in Canada tends to be focused in both fall and spring, with reduced rates in winter (Beaulieu, 2004). For most of

the sites there was no significant correlation between sulfamate concentrations and surface water temperature, though there was a weak negative correlation at three sites (Table S2.3 and Fig. S2.3 in Supporting Information). Similarly there was no significant correlation between the concentrations of sulfamate and dissolved oxygen at most sites (Table S2.3 and Fig. S2.4 in Supporting Information). In contrast, for the two sites with sufficient stream discharge data (collected during this study, represent the two largest catchments: Table 1), there was a strong positive correlation between the sulfamate concentrations and stream discharge (Table S2.3 and Fig. S2.5 in Supporting Information). Collectively, these results suggest that the fluctuations in sulfamate concentrations in the streams are strongly related to both seasonal and episodic variations in hydrological conditions and probably less influenced by seasonal variations in temperature and/or microbial activity. 3.1.4. Export rates (per catchment area) based on stream data The mean sulfamate concentrations in the ten streams in the Upper Thames basin were used to estimate total annual export (Ea-t) of sulfamate per unit area in each catchment: Ea−t ¼ ðcs  roa–c Þ=arc

Fig. 4. Temporal variation of sulfamate in bulk precipitation (wet plus dry) sampled from the Upper Thames basin in the Cedar Creek (blue circles) and Nissouri Creek (red circles) sub-basins. (For interpretation of the references to colour in this figure legend, the reader is referred to the web version of this article.)

ð3Þ

where cs is the mean sulfamate concentration in the stream (ng/L = μg/ m3), roa-rc is the annual runoff in the catchment at the point of sampling (m3/year), and arc is the area of the same catchment (ha) upstream of the point of sampling. Based on Eq. (3), the inferred rates of export per unit area in the ten catchments range from 1604 to 86,035 mg/ha·year (Table 3). The results for the two largest catchments (CC-SW-A, NC-SW-D) were 17,859 to 18,644 mg/ha·year. When compared to the above estimated rates of atmospheric deposition (6450 to 10,550 mg/ha·year), these results suggest that in some catchments, atmospheric deposition may provide the majority of the sulfamate observed in the streams. However, this statement must be qualified by the evidence for some loss (possibly by degradation) of sulfamate within these catchments. For example, the lowest export (1604 mg/ha·year) was for the stream with largest percentage of wetlands in its catchment (NC-SW-C: Tables 1

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Fig. 5. Seasonality of sulfamate in surface water samples and groundwater samples in the Upper Thames basin. Each sampling location shown by a unique symbol. Season breaks are shown by vertical dotted lines.

and 3), suggesting that wetlands may be environments where enhanced degradation of sulfamate occurs (cf. groundwater data below). The fact that some stream catchments export sulfamate at rates (Table 3) that are much higher than the estimated rate of annual atmospheric deposition of this chemical suggests that there are other sources of sulfamate in these catchments. For example, dairy operations could be important sources given that sulfamic acid is commonly used as

cleaning agent in this sector (Metzger, 2012). However, further detailed investigation would be required to identify the specific sources of the additional sulfamate. 3.1.5. Groundwater data Sulfamate concentrations in shallow riparian groundwater sampled every two weeks at eight locations in the Upper Thames basin

Table 3 Inferred total annual export (per unit area) of sulfamate from ten Upper Thames catchments based on stream data. Catchment Mean stream sulfamate concentration (ng/L) Total sulfamate export (mg/ha·year)a

CC-SW-A

CC-SW-B

CC-SW-C

CC-SW-D

CC-SW-E

NC-SW-A

NC-SW-B

NC-SW-C

NC-SW-D

NC-SW-E

4878

3072

3261

23,274

7007

3943

4295

395

4475

1676

18,644

11,448

12,131

86,035

26,059

16,001

17,313

1604

17,859

6689

a Based on Eq. (3) (see text), using the Ontario Flow Assessment Tool (Ontario Ministry of Natural Resources and Forestry, 2018) to determine annual stream flow and upstream area for each catchment.

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catchments ranged from non-detectable (b25 ng/L) to 12,174 ng/L (Figs. 3, 5, Table 2). The mean (and median) concentrations in groundwater were much lower than in either precipitation or surface water (Fig. 3, Table 2). Overall, these concentrations were significantly lower than those observed in the streams in the same catchments (two sample t-test, p b 0.005). These results suggest that sulfamate is attenuated or degraded in the subsurface. Groundwater generally experiences a wide range of redox conditions compared to surface waters and degradation of sulfamate under relatively reducing conditions could be responsible for the lower sulfamate concentrations in groundwater relative to surface water and precipitation (cf. discussion of Lake of the Woods groundwater data below). The average concentration (± standard deviation) of dissolved O2 in the riparian groundwater in the Upper Thames basin was 1.7 ± 1.0 mg/L (111 measurements), whereas the mean concentration of dissolved O2 in the streams was 10.8 ± 2.9 mg/L (177 measurements). It is probable that the shallow groundwater varies in redox conditions, and includes anoxic microsites, as observed in other studies (e.g., Shati et al., 1996). When all groundwater samples were grouped together by season, there were statistically significant seasonal differences in sulfamate concentrations, with the highest in fall and winter (Fig. 5; Table S2.1 in Supporting Information). As discussed above, similar trends observed in surface water concentrations appeared to be strongly related to seasonal differences in stream discharge, which may correspond to seasonal differences in hydrological conditions in the subsurface (e.g., groundwater flow rates). At two of the groundwater sampling sites there was a strong negative correlation between sulfamate concentrations and temperature (Table S2.4 in Supporting Information); perhaps the higher concentrations were related to reduced microbial activities during colder periods in the shallow riparian subsurface. Where data were available there was no significant correlation between dissolved oxygen and sulfamate concentrations in groundwater (Table S2.4 in Supporting Information). 3.1.6. Sulfamate relationship to acesulfame Not surprisingly, only two of the precipitation samples had detections of acesulfame (Fig. 6); this sweetener is not expected to be a significant contaminant in the atmosphere. In contrast, saccharin was detected in all of the precipitation samples (average 133 ng/L, range 8 to 1249 ng/L). However, some of our quality control samples indicated potential contamination of some of these samples by saccharin derived from an unknown source (atmospheric dust?) (Fig. S2.1 in Supporting Information).

Similar to our results for saccharin, in a study conducted in an urban area of China, saccharin was detected in every precipitation sample and had the highest single concentration (1300 ng/L) and the highest mean concentration (284 ng/L) of several artificial sweeteners that were detected (Gan et al., 2013). The authors are not aware of any other study that has analyzed sweeteners in precipitation samples. Though the source(s) of the atmospheric sulfamate that we detected in precipitation samples are not known, it is possible that the origin of at least some of this sulfamate is related to atmospheric saccharin that we also detected in precipitation samples. Based on these observations, we conclude that the sulfamate detected in precipitation was derived from one or more non-wastewater sources, possibly from natural processes in the atmosphere, as discussed in the introduction, and possibly from a source related to saccharin. Many of the stream and groundwater samples from the Upper Thames basin had measurable concentrations of acesulfame (Fig. 6) and saccharin (not shown). The origin of acesulfame detected in the streams and groundwater in this dominantly rural area (Table 1) is probably wastewater derived from septic systems (Spoelstra et al., 2017). The highest concentrations of this sweetener were observed in samples from a stream that flows through a small urban community (that has many septic systems), and also in samples from groundwater along this stream. The origin of saccharin is less certain, it has been used for many decades in various applications, including as an artificial sweetener in food and beverages, but also as a component of animal feed (Buerge et al., 2011; Ma et al., 2017) and is also known to be a breakdown product of sulfonylurea herbicides (Berger and Wolfe, 1996). Although many of the stream and groundwater samples from the Upper Thames basin have co-detections of acesulfame with sulfamate, there is no obvious relationship between the two in these samples (Fig. 6 and Table S2.2 in Supporting Information). Some of the groundwater samples with acesulfame detections had non-detections of sulfamate, and vice-versa (Fig. 6). Overall, sulfamate was very weakly correlated with the sweetener acesulfame in the stream samples, but not with saccharin (Table S2.2 in Supporting Information). As discussed above, many of the sulfamate concentrations in the streams and groundwater were similar to those in precipitation (Figs. 3, 6), suggesting that atmospheric deposition could account for much of the sulfamate detected in the streams and groundwater. Although the ranges of sulfamate concentrations in the streams and groundwater at the Upper Thames sites were similar to the range we observed in septic wastewater (at Lake of the Woods; see Fig. 3 and further discussion

Fig. 6. Concentrations of sulfamate plotted against the sweetener acesulfame in environmental waters sampled in the Upper Thames basin; mdl = method detection limit.

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below), the stream and groundwater samples had much lower acesulfame concentrations compared to the septic wastewater. 3.2. Lake of the Woods 3.2.1. Lake samples In the lakewide samples from Lake of the Woods, sulfamate concentrations ranged from non-detectable to 1277 ng/L (Figs. 3, 7, Fig. S2.6 in Supporting information). Most concentrations were between 100 and 1000 ng/L (Fig. 7). The lowest concentrations were in relatively pristine, peripheral areas of the basin, and highest concentrations were from the south part of the lake, through the middle, and northward towards the city of Kenora (Fig. S2.6), along what is inferred to be the main south-tonorth flowpath through the lake (De Sellas et al., 2009). For the lakewide samples collected in 2017, the mean concentration of sulfamate was higher in June compared to September, and higher in samples collected at depth (2 m above bottom) compared to surface samples (1 m) (Table S2.5 in Supporting Information). However, the only statistically significant differences were between the June samples (both depths) and the September surface samples (1 m) (Table S2.5 in Supporting Information). As with the Upper Thames stream data (above), the combination of nearly all (83 of 84) detections of sulfamate in the lakewide samples, along with observed differences in concentrations across the lake, suggest that sources may include regional scale atmospheric deposition (including the relatively pristine sites), combined with localized additional impact from other sources. Localized impacts could include increased rates of atmospheric deposition (perhaps related to polluted air), or increased rates of land-based releases from industrial/urban sources, such as wastewater effluent. The urban nearshore lake samples, which were collected from along the northern edge of Lake of the Woods (Keewatin neighbourhood of

Fig. 7. Sulfamate concentrations plotted against the concentrations of the artificial sweetener acesulfame in Lake of the Woods (lakewide, nearshore urban, Sioux Narrows, Poplar Bay) and in streams in the Rainy River sub-basin (Rainy River flows into Lake of the Woods); mdl = method detection limit.

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City of Kenora), had relatively high sulfamate concentrations ranging from 467 to 17,177 ng/L (Fig. 7; Table 2). Most concentrations were around 1000 ng/L, which is the high end of the range for the lakewide monitoring stations (Fig. 7). The highest concentration was close to the outfall of a storm sewer. This variation suggests some nearshore impact from localized urban sources. In the lakewide samples from Lake of the Woods, there was no significant correlation of sulfamate and the artificial sweetener acesulfame (Table S2.2 in Supporting Information). Such a relationship might be expected if the sulfamate was derived from wastewater, given that acesulfame has been shown to be a good tracer of municipal and domestic wastewater (Buerge et al., 2009; Van Stempvoort et al., 2011a, 2011b). However, a lack of correlation is not diagnostic because this could result from mixing of wastewaters of different composition or when one or more of these analytes has been extensively degraded. Some of the highest concentrations of sulfamate in the urban nearshore waters were associated with high concentrations of acesulfame (Fig. 7). This observation suggests that some of the sulfamate may be derived from wastewater, an interpretation that is supported by the positive correlation of sulfamate with acesulfame in these samples (Table S2.2 in Supporting Information). Following Castronovo et al. (2017), the elevated sulfamate in the nearshore waters at Kenora could be largely a metabolite derived from the degradation of acesulfame present in the wastewater. Alternatively, it could be a coexisting constituent of wastewater, for example derived from cleaning products (see introduction). The elevated sulfamate concentrations observed in these nearshore urban samples cannot be traced directly to treated wastewater effluent from the City of Kenora, which is discharged into the Winnipeg River downstream of Lake of the Woods. Poplar Bay is a relatively enclosed bay in Lake of the Woods, with extensive cottage development. As anticipated, the samples from Poplar Bay (Fig. 7) had relatively high acesulfame concentrations, indicating significant influx of septic wastewater (cottages) to the bay. In contrast, the sulfamate concentrations were generally low in this bay (Fig. 7; Table 2), and there was no significant correlation between the two (Table S2.2 in Supporting Information). Given the above evidence from the Upper Thames basin for possible enhanced degradation of sulfamate under low O2/anoxic conditions, perhaps the low sulfamate concentrations in Poplar Bay are the result of microbial degradation in the deeper part of this bay, which experiences seasonal low oxygen concentrations (Pascoe et al., 2014), including inferred historic hypoxic conditions (Summers et al., 2012). However, depth profiles of sulfamate in Poplar Bay that were obtained in 2018 show that the shallowest water often had the lowest concentrations (Fig. S2.7 in Supporting Information). Samples from the Sioux Narrows area have relatively low sulfamate (Table 2), reflecting relatively pristine conditions overall in this portion of Lake of the Woods, or enhanced degradation. Some of these samples that were collected in nearshore settings have elevated acesulfame concentrations (Fig. 7), which is likely related to seepage of septic wastewater into the lake. Overall, the paired concentrations of acesulfame and sulfamate in the Lake of the Woods samples (Fig. 7) indicate a complex relationship, and yield no clear evidence about whether a significant portion of the sulfamate in this lake is derived from wastewater. 3.2.2. Stream samples The highest concentration of sulfamate measured in this study (128,000 ng/L) was one of the samples from a stream in the Rainy River sub-basin (Fig. 3). This sample, taken downstream of a recent mine development, also had detectable acesulfame (21 ng/L). There are also farms in the area. The exact source of the elevated sulfamate in this stream is unknown, but it seems likely that it is largely derived from one or more localized, land-based anthropogenic sources. Sulfamate appears to be weakly correlated with the sweetener acesulfame in the Rainy River sub-basin stream samples, but not at a

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statistically significant (0.05) level (Table S2.2 in Supporting Information). 3.2.3. Groundwater samples In shallow groundwater samples collected from an urban shoreline of Lake of the Woods in the city of Kenora, the sulfamate concentrations ranged from 80 to 6414 ng/L (Figs. 3, 8; Table 2). The sulfamate was not significantly correlated with the sweeteners acesulfame or saccharin (Table S2.2 in Supporting Information), suggesting that wastewater is not an important source of the sulfamate in these groundwater samples. The lowest sulfamate concentrations in the urban groundwater at Kenora (lower than adjacent surface water: Table 2) were associated with low dissolved oxygen, low nitrate, and elevated ammonium. The sulfamate and nitrate concentrations were significantly correlated (at the 0.02 level: Table S2.6 in Supporting Information). These relationships suggest that microbial degradation of sulfamate has occurred preferentially under relatively reducing conditions in the groundwater (cf. Upper Thames and Minesing data). Elsewhere within the Lake of the Woods basin, elevated concentrations of sulfamate were detected in groundwater impacted by septic plumes (up to 15,127 ng/L) and in other groundwater (up to 2534 ng/L) (Table 2). In the samples impacted by septic plumes, the concentrations of acesulfame were also high (Fig. 8), consistent with sulfamate being derived from wastewater, either as a degradation product of acesulfame as suggested by Castronovo et al. (2017), or as a coexisting constituent of wastewater. These septic-impacted samples also had elevated concentrations of the sweeteners cyclamate (up to 7996 ng/L) and saccharin (up to 3579 ng/L). It is uncertain how much the degradation of these three sweeteners may have contributed to the sulfamate that is observed in these groundwater samples (see introduction). Statistical analysis indicated no significant correlation between sulfamate and the three sweeteners in these septic plumes, although a weak positive relationship with acesulfame is suggested (Table S2.2 in Supporting Information). The concentrations of acesulfame (and the other sweeteners) were generally much higher in the groundwater at these septic sites compared to the urban groundwater in Kenora and the other groundwater samples collected at Lake of

the Woods, whereas the concentrations of sulfamate were similar in all these groups of samples (Fig. 8). The fact that the concentrations of sulfamate were very similar in the septic wastewater samples and in all of the types of groundwater samples that we collected at Lake of the Woods (Figs. 3, 8) indicates that at least much of the sulfamate that we detected in groundwater samples was not derived from septic wastewater. 3.3. Minesing wetlands 3.3.1. Groundwater samples For groundwater samples collected in the Minesing Wetlands, most of the sulfamate concentrations were lower than those measured in precipitation (wet plus dry) in the Upper Thames basin (Figs. 3, 9, Table 2). Furthermore, wastewater from municipal effluents or septic systems are not present in the wetlands. These facts suggest that the sulfamate in the Minesing Wetlands groundwater may be derived predominantly from atmospheric deposition. In May 2017 sulfamate was generally more abundant in the shallowest groundwater samples of paired shallow (just below surface of the peat) and deep (~1.5 to 3 mbgs) piezometers, and was often undetectable in the deeper groundwater (Fig. 9). This pattern suggests that atmospheric-derived sulfamate is degraded during downward flow of groundwater as conditions in the deeper subsurface of the wetland become more reducing, as evidenced by the general lack of nitrate and increasing iron concentrations in the deeper groundwater samples (data not shown). However, the pattern of concentrations of sulfamate was very different between the spring and late fall samplings. In November 2018, sulfamate was detected in all of the wetland groundwater samples and the higher concentrations tend to be in the deeper piezometers. Conditions in the deeper wetland groundwater in November were less reducing, with higher nitrate and lower iron concentrations (data not shown), supporting the idea that sulfamate would be less prone to degradation and therefore be found at higher concentrations at this time. The cause of this seasonal variation is unknown. We hypothesize that the elevated sulfamate in the deeper November samples was

Fig. 8. Sulfamate and acesulfame concentrations in groundwater from the Lake of the Woods basin; mdl = method detection limit.

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Fig. 9. Plots of saccharin versus sulfamate in the Minesing groundwater samples; mdl = method detection limit.

derived from degradation of sulfamate-containing compounds that were already enriched at these depths. Alternatively, perhaps the November 2018 sulfamate profile was affected by post-summer seasonal recharge, which moved a zone of sulfamate-enriched groundwater from near the water table to the deeper monitoring points below. Further investigation would be required to test these hypotheses. Generally the only sweetener found in the groundwater samples from the Minesing Wetlands was saccharin. If domestic wastewater was present in the wetlands, acesulfame would have also been detected in the groundwater, and therefore the saccharin (and sulfamate) must be from a non-wastewater source. It has previously been shown that the herbicide glyphosate is also present in the wetland and that the source is likely atmospheric deposition from the adjacent uplands (Van Stempvoort et al., 2016). In addition to sulfamate produced by reactions in the atmosphere, it is possible that some of the sulfamate delivered to the Minesing Wetlands by atmospheric deposition originates from use as a herbicide in the surrounding uplands. There was no significant correlation of sulfamate and saccharin for the May 2017 samples (Fig. 9, Table S2.2 in Supporting Information). In the November 2017 wetland groundwater samples, the highest concentration of sulfamate was associated with the highest saccharin, yielding a strong positive correlation between these two analytes (Fig. 9, Table S2.2 in Supporting Information). When this outlier (highest concentrations) was removed, there was no significant correlation (Table S2.2 in Supporting Information). The anomalous sample with the highest concentrations of sulfamate (by two orders of magnitude) and saccharin suggests that the source of this high sulfamate was related to saccharin in some way, perhaps a land-based source (i.e., not atmospheric). What kind of nonatmospheric source could have produced elevated sulfamate and/or saccharin in these wetland groundwater samples is unknown. The fact that all of the other samples yielded no correlation with saccharin (Table S2.2 in Supporting Information) suggests that in most of these samples the sulfamate was derived from atmospheric deposition. As explained above this is reasonable given that the sulfamate concentrations in most of the wetland groundwater samples were lower than those measured elsewhere in precipitation (Fig. 3). 3.4. Contextual municipal wastewater data Our Canadian municipal wastewater data (Fig. S2.8 in Supporting Information) indicate that sulfamate concentrations are similar in both raw influent and final effluent samples, and are not significantly different based on a two sample t-test (p = 0.771). In contrast, as a group the acesulfame concentrations are significantly lower in the effluent

samples (two sample t-test, p b 0.005). These results support earlier evidence that, compared to acesulfame, sulfamate appears to be relatively persistent during municipal wastewater treatment (Castronovo et al., 2017), but do not suggest net conversion of acesulfame to sulfamate (Fig. S2.8 in Supporting Information). Although the samples of municipal wastewater tended to have higher sulfamate concentrations than environmental water samples, there was overlap in their concentration ranges (Fig. 3). Given this observation, plus the fact that the fluxes of treated wastewater that are discharged to the environment are very small fractions compared to the environmental water cycle fluxes at our study sites (for example, the rate of precipitation), it is evident that municipal wastewater could only account, at most, for a small secondary fraction of the sulfamate that we observed in these environmental water samples.

4. Conclusions Sulfamate was detected in almost all of the environmental water samples that were analyzed, ranging in concentration from nondetectable to 128,140 ng/L. Mean concentrations for the various groups of surface water samples discussed in this study ranged from 140 to 24,000 ng/L. Sulfamate was detected in all of the precipitation samples, and was found in N99% of the surface water samples at urban and rural sites, and sites relatively remote from development areas. Lowest concentrations and non-detections were in groundwater samples that had relatively low dissolved oxygen concentrations, and in surface water in a restricted bay that is known to be affected by seasonal hypoxia. These patterns suggest that degradation of sulfamate may be enhanced in environmental waters under relatively reducing conditions. As a follow-up to this study, the potential for loss and/or microbial degradation of sulfamate in such settings should be investigated. The geographic and spatial distribution of concentrations suggested higher amounts of sulfamate in the vicinity of urban and/or industrial developments. Ubiquitous but uneven distribution suggests a mix of atmospheric deposition (all sites) and localized urban / industrial sources at sites with the highest concentrations. Given the relatively high concentrations in some environmental waters, the risk of negative impacts (toxicity) on aquatic organisms should be investigated. Sulfamate was sometimes associated with presence of the artificial sweetener acesulfame, but not consistently. Our evidence indicates that the sulfamate detected in the environmental water samples that we analyzed was probably mostly from non-wastewater sources. Since we demonstrate that the breakdown of acesulfame is clearly not

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the only source of sulfamate in environmental waters, it would be challenging to use sulfamate as a “wastewater tracer”. Declaration of competing interest The authors have no competing interests to declare. Acknowledgements The following people from Environment and Climate Change Canada (ECCC) assisted with collection of environmental water samples: Ross MacKay (all sites), Pam Collins (Upper Thames), Tim Pascoe and Tana McDaniel (Lake of the Woods). Steve Teslic, Korey Broad and Alexandra Auyeung of ECCC collected the municipal wastewater samples. Michael Saunders (Nottawasaga Valley Conservation Authority) assisted with collection of the groundwater samples at the Minesing Wetlands. Pam Collins (ECCC) conducted laboratory analyses of anions, cations and metals, and ammonium. Thanks also goes to Oxford County, the City of Kenora, the municipality of Sioux Narrows and Nestor Falls, and several private landowners for allowing access to sites for collection of precipitation and/or groundwater samples. This study was funded by Environment and Climate Change Canada. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi. org/10.1016/j.scitotenv.2019.133734. References Atlas, E., Giam, C.S., 1988. Ambient concentration and precipitation scavenging of atmospheric organic pollutants. Water Air and Soil Pollution 38 (1–2), 19–36. Beaulieu, M.S., 2004. Manure Management in Canada. Article 2, Farm Environmental Management in Canada, June 2004. Vol. 1, No. 2. Statistics Canada, Catalogue No. 21–021-MIE. Berger, B.M., Wolfe, N.L., 1996. Multiresidue determination of sulfonylurea herbicides by capillary electrophoresis for hydrolysis studies in water and sediments. Fresenius Journal of Analytical Chemistry 356 (8), 508–511. Bodek, I., Smith, R.H., 1980. Determination of ammonium sulfamate in air using ion chromatography. Am. Ind. Hyg. Assoc. J. 41 (8), 603–607. Bradford, A.L., 1999. A Hydrobiological Study of Minesing Swamp, Ontario. Queen's University, Kingston, ON PhD thesis. (408 pp). Buerge, I.J., Buser, H.-R., Kahle, M., Müller, M.D., Poiger, T., 2009. Ubiquitous occurrence of the artificial sweetener acesulfame in the aquatic environment: an ideal chemical marker of domestic wastewater in groundwater. Environmental Science & Technology 43 (12), 4381–4385. Buerge, I.J., Keller, M., Buser, H.-R., Müller, M.D., Poiger, T. 2011. Saccharin and other artificial sweeteners in soils: estimated inputs from agriculture and households, degradation, and leaching to groundwater. Environ. Sci. Technol. 45 (2), 615–621. http:// dx.doi.org/https://doi.org/10.1021/es1031272. Castronovo, S., Wick, A., Scheurer, M., Nödler, K., Schulz, M., Ternes, T.A., 2017. Biodegradation of the artificial sweetener acesulfame in biological wastewater treatment and sandfilters. Water Res. 110, 342–353. Cupery, M., 1938. Sulfamic acid: a new industrial chemical. Industrial & Engineering Chemistry 30 (6), 627–631. De Sellas, A.M., Paterson, A.M., Clark, B.J., Baratono, N.G., Sellers, T.S., 2009. State of the Lake Report for Lake of the Woods and Rainy River Basin. Lake of the Woods Sustainability Foundation, Kenora. Fulton, C.K., Cooper, R.A., 2005. Catabolism of sulfamate by Mycobacterium sp. CF1. Environ. Microbiol. 7 (3), 378–381. Gan, Z., Sun, H., Feng, B., Wang, R., Zhang, Y., 2013. Occurrence of seven artificial sweeteners in the aquatic environment and precipitation of Tianjin, China. Water Res. 47 (14), 4928–4937. Gröning, M., Lutz, H.O., Roller-Lutz, Z., Kralik, M., Gourcy, L., Pöltenstein, L., 2012. A simple rain collector preventing water re-evaporation dedicated for δ18O and δ2H analysis of cumulative precipitation samples. J. Hydrol. 448–449, 195–200. Guerra, P., Kim, M., Teslic, S., Alaee, M., Smyth, S.A., 2015. Bisphenol-A removal in various wastewater treatment processes: operational conditions, mass balance, and optimization. J. Environ. Manag. 152, 192–200. Hirota, K., Mäkelä, J., Tokunaga, O., 1996. Reactions of sulfur dioxide with ammonia:dependence on oxygen and nitric oxide. Industrial Engineering & Chemistry Research 35 (10), 3362–3368. Kahl, S., Kleinsteuber, S., Nivala, J., van Afferden, M., Reemtsma, T., 2018. Emerging biodegradation of the previously persistent artificial sweetener acesulfame in biological wastewater treatment. Environmental Science & Technology 52 (5), 2717–2725. Kamrin, M.A., Montgomery, J.H., 1999. Agrochemical and Pesticide Desk Reference on CDROM. CRC Press LLC, Boca Raton, FL, USA.

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