Water Analysis: Emerging Contaminants and Current Issues

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Review

Water Analysis: Emerging Contaminants and Current Issues Susan D. Richardson, and Thomas A. Ternes Anal. Chem., Just Accepted Manuscript • DOI: 10.1021/acs.analchem.7b04577 • Publication Date (Web): 07 Nov 2017 Downloaded from http://pubs.acs.org on November 8, 2017

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Analytical Chemistry

Water Analysis: Emerging Contaminants and Current Issues Susan D. Richardson1 and Thomas A. Ternes2 1 2

Department of Chemistry and Biochemistry, University of South Carolina, Columbia, SC 29205 Federal Institute of Hydrology, Koblenz, D-56068 Germany

BACKGROUND This biennial review covers developments in water analysis for emerging environmental contaminants over the period of October 2015-October 2017. Analytical Chemistry’s policy is to limit reviews to a maximum of ~250 significant references and to mainly focus on new trends. Therefore, only a small fraction of the quality research publications are discussed. The previous Water Analysis review (with Susana Kimura) was published in 2016.1 This year, Thomas Ternes has joined back again to cover the section on Pharmaceuticals and Hormones. We welcome any comments you have on this Review ([email protected]).

Numerous abstracts were consulted before choosing the best representative ones to present here. Abstract searches were carried out using Web of Science, and in many cases, full articles were obtained. A table of acronyms is provided (Table 1) as a quick reference to the acronyms of analytical techniques and other terms discussed in this Review. Table 2 provides some useful websites.

Major Analysis Trends. Non-target, unknown analysis continues to be a hot trend, and due to the immense amount of chemical features that can be found using mass spectrometry, detailed workflows are becoming popular to handle all the data, as it can be too time-consuming to manually interpret thousands of unknown compounds and spectral features. High resolution (HR)-mass spectrometry (MS) continues to be a cornerstone of unknown identification, as does MS/MS, library database creation and searching, in silico methods, and user-created software. While electron ionization (EI)-MS libraries (such as the NIST and Wiley databases) can have >760,000 compounds, electrospray ionization (ESI)-MS

1 ACS Paragon Plus Environment


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libraries are lagging behind, due in large part to the lack of standardization for MS conditions used to generate mass spectra and to the differences in spectra produced on different brands and types of mass spectrometers. As a result, many separate databases exist, whether created by mass spectrometry companies, the NIST, or by individual users. Some research groups are posting these for free online use (see Table 2).

Open or semi-open mass spectrometry databases include Metlin, mzCloud, MassBank, European MassBank, MassBank of North America, the Global Natural Products Social Molecular Networking (GNPS), the Human Metabolome Database (HMDB), Stoff-Ident, and ChemSpider. Two commercial MS/MS libraries (NIST and Wiley) are also now available. The NIST2017 library database contains MS/MS spectra for >15,000 compounds, and Wiley’s MSforID database contains MS/MS spectra for >1200 compounds. Workflow tools include Metfrag (a metabolomics MS/MS fragmentation predictor) and Competitive Fragmentation Modeling for Metabolite Identification (CFM-ID), developed by Wishart’s lab at the University of Alberta, which predicts MS/MS spectra for compounds (for input structures) and can rank candidate structures for how well they match the MS/MS spectra. Wishart’s group also recently published a new tool (called CFM-EI) to predict EI mass spectra. This was published in 2016 in Analytical Chemistry2, and is also freely available at http://cfmid.wishartlab.com. CFM-EI incorporates an artificial neural network and was created using a subset of the NIST library as a training set. It can handle odd-electron ions and isotopes and has shown good prediction capability in crossvalidation tests against the NIST database. As such, it can be useful for narrowing the potential chemical structures for unknown identification.

Two new analytical techniques also stand out this year. First, ion mobility-MS was recently developed and is beginning to be used for some environmental applications. Ion mobility-MS offers an additional dimension in separation (by cross-section of the molecules), which can aid in identifying


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compounds in complex environmental mixtures. Examples included in the review this year include the identification of photolysis transformation products (TPs) from the artificial sweetener acesulfame and the identification of naphthenic acid ozonation products. Another technique recently introduced is a software analysis tool called precursor ion exclusion (PIE). This tool helps to overcome a weakness of data dependent acquisition (DDA), a common method for triggering MS/MS scans for unknowns when doing non-target analysis using liquid chromatography (LC)-MS/MS. DDA typically generates MS/MS data for only the most abundant chromatographic peaks observed in a LC-MS analysis, such that the smaller peaks typically go unidentified. Using the PIE approach, Li’s group at the University of Alberta was able to exclude those highly abundant peaks in a second analysis, which allowed a more comprehensive determination of unknown peptides and chlorinated peptides in treated drinking water (see Drinking Water Disinfection Byproduct (DBP) section).

Sampling and Extraction Trends. Fabric phase sorptive extraction (FPSE) is a new type of solid phase extraction (SPE) recently developed by Furton’s group at Florida International University, and several papers were published the last two years using it. This technique uses small squares of cellulosic (or other) fabric that are coated with an ultra-thin sol-gel, which are immersed directly in aqueous environmental water samples to sorb and extract analytes. The fabric piece is then submersed into a small volume of extraction solvent to elute the analytes, after which, they are injected onto a chromatography system. FPSE offers advantages of minimal solvent use, short extraction times, small sample volumes, and high preconcentration factors. Further, the coated fabric pieces can be reused after minimal cleaning without cross-contamination. Results have been demonstrated for drinking water, surface water, wastewater, and biological samples (e.g., urine). Examples of this new technology in this Review include extraction of pharmaceuticals and personal care products, hormones, benzotriazoles, and benzothiazoles, followed by analysis with ultraperformance liquid chromatography (UPLC)-MS/MS.



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SPE, online SPE, and solid phase microextraction (SPME) continue to be popular extraction techniques. Multiple SPE columns with different sorbents are becoming important for comprehensive, non-target analysis, as researchers try to recover more material from water that cannot be recovered using a single phase SPE. For example, Li’s group from the University of Alberta used three complementary SPEs—Oasis HLB, Bond Elut C18, and Bond Elut ENV—to more effectively extract analytes from drinking water (see Drinking Water DBP section). A new multi-method for measuring 302 compounds including many pharmaceuticals in surface water also utilized multiple complementary SPEs: Oasis HLB, Isolute ENV+, Strata XCW cation exchanger, and Strata anion exchanger (see Pharmaceutical and Hormone section). In addition, a new polarity rapid assessment method (PRAM) to characterize nitrosamine precursors uses complementary SPE cartridges in parallel (see Drinking Water DBP section).

Also continuing from the last Water Analysis review, ionic liquids (ILs) are being used for extraction of emerging contaminants (ECs). An example in this Review includes the use of magnetic ionic liquids with stir bar dispersive liquid microextraction for extracting UV filters.

Chromatography Trends. There is nothing particularly new in the chromatographic arena for ECs, as LC, UPLC, gas chromatography (GC), and GCxGC continue to be popular. However, there is a trend in coupling multiple LC and GC columns of different phases (orthogonal) to better separate complex environmental mixtures, especially for comprehensive, non-target analysis. An example includes use of both C18 and hydrophilic interaction liquid chromatography (HILIC) LC columns, which dramatically increased the number of compounds in drinking water that can be detected and resolved by mass spectrometry (see Drinking Water DBP section).

Emerging Contaminant Trends. Transformation products (TPs) continue to be an important subject of many EC studies, with increasing numbers of photolysis and catalytic photolysis experiments published. An important aspect of this topic is that these TPs can often be more toxic than the parent


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compounds, and many studies continue to combine chemistry and toxicology for this reason, as well as using effects-directed analysis (EDA) to find the active compounds in a complex environmental mixture. As with the last Water Analysis Review published in 2016, elaborate transformation pathways continue to be proposed for ECs, with mechanisms determined using LC-MS/MS and LC-HR-MS, and sometimes confirmed by nuclear magnetic resonance (NMR) spectroscopy.

There also continues to be an explosion of research in the per- and polyfluoroalkyl substance (PFAS) area. Progressively more research using HR-MS is being used to more comprehensively identify PFASs in aqueous film forming foams (AFFFs) and in other products, including precursors to the more commonly observed perfluoroalkyl acids and perfluoroalkyl sulfonate end-products. There are also new PFASs being created as old ones are being phased out, creating a “moving target” for research. A review article entitled, “A never-ending story of per- and polyfluoroalkyl substances (PFASs)?” is an apt description of this research area, as there are hundreds, possibly thousands of fluorinated compounds that still remain unknown. New surrogate methods have been recently developed that can quantify total fluorine, which allows researchers to determine what percent the quantified substances represent of the whole mixture. Flame retardants are also a “moving target”, with new chlorinated ones being put on the market as others are discontinued.

New Emerging Contaminant. This year, halomethane sulfonic acids are added as a new emerging contaminant class on the horizon. Non-target screening approaches have recently uncovered them, and results indicate that they may be widespread. Fluoro-, chloro-, bromo-, and iodo-methane sulfonic acids have been tentatively identified, including trifluoromethanesulfonic acid, which is a high production chemical and is regulated in the European Union (EU) under the Registration, Evaluation, and Authorization of Chemicals (REACH) program.



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GENERAL REVIEWS, LARGE OCCURRENCE STUDIES, AND EMERGING AREAS OF FOCUS This section includes general reviews relating, as well as large occurrence studies involving ECs and new emerging areas of focus. Reviews that relate to specific areas (e.g., PFASs, pharmaceuticals, DBPs) can be found in those specific sections. Several reviews and special issues of journals have been published over the last 2 years that relate to water analysis, and several focus specifically on emerging contaminants. Due to reference number limitations, only a few could be cited here.

Emerging Contaminant Reviews. Perez-Fernandez al et al. published an excellent review on recent advancements and future trends in the analysis of ECs, including discussions on sample preparation, LC separations, and mass spectrometry analysis.3 Strengths and weaknesses of miniaturized extraction techniques are also covered, as well as new materials for sample extraction (e.g., nanosized sorbents and ionic liquids) and HR-MS applications. Noguera-Oviedo and Aga wrote a thoughtprovoking review on ‘lessons learned’ from more than two decades of research on ECs in the environment.4 Lessons learned include: (1) ECs are a worldwide phenomenon, with sources mainly deriving from human activities; (2) Treatment does not necessarily mean complete removal of ECs; (3) Metabolites and TPs of ECs matter because some can be more toxic or biologically active than the parent ECs; (4) Unconventional testing of effects and toxicity are needed for ECs (e.g., to account for ecological toxicity when present in the environment as mixtures); and (5) Even the most advanced tools can miss the target (e.g., due to the lack of standards and searchable LC-MS libraries).

A review entitled, Emerging environmental contaminants: Challenges facing our next generation and potential engineering solutions, was the focus of another review by Richardson and Kimura, who discussed sources, occurrence, and issues with ECs, transformation of ECs, effect of climate change on ECs, along with proposed engineering solutions to minimize them in the environment.5 Engineering


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solutions included various oxidation and advanced oxidation processes (AOPs), adsorption on granular activated carbon (GAC) or powdered activated carbon (PAC), and rejection by various membrane technologies. A summary of typical treatments with their advantages, disadvantages, and their costs were summarized and critically compared.

Tiedeken et al. published a review covering the last 20 years on monitoring, sources, receptors, and control measures for three EU Watch List substances: diclofenac, estradiol (E2), and ethinylestradiol (EE2).6 It was noted that diclofenac, E2, and EE2 can often be present at levels higher than the proposed EU annual average environmental quality standards, but that these compounds are not always sufficiently monitored for in many EU member countries. In addition, more sensitive methods are needed for E2 and EE2 to meet the new proposed standards of 0.04 and 0.035 ng/L in water bodies.

Effect-directed analysis (EDA) was the focus of an extensive review by Brack et al., who discussed the selection of bioassays, practical considerations in the design and application of bioassays, sampling strategies and extraction for water, sediment, and biota, fractionation of samples, sample workflow and automation, and toxicant identification.7 Comprehensive examples of in vitro and in vivo bioassays are outlined, along with their applications, advantages, and disadvantages. This is a must-read for anyone seeking more information on EDA. The sensitivity of in vitro bioassays for androgenic, progestagenic, glucocorticoid, thyroid, and estrogenic activity was the focus of another review by Leusch et al. 8 It was suggested that there is currently sufficient sensitivity (with typical sample enrichment) for measuring androgenic, estrogenic, progestagenic, and glucocorticoid activity in environmental waters and drinking water, but that there are currently major knowledge gaps concerning thyroid and antithyroid activity.



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Large Occurrence Studies of ECs. An extensive study of ECs in natural mineral water and spring water sold in France was the subject of another very interesting article by Le Coadou et al.9 Approximately 70% of French-sold bottled waters, encompassing 40 brands, were assessed in this major study that measured 118 pesticides and their TPs, 172 pharmaceuticals, 11 hormones, 8 alkylphenols, 11 phthalates, and 10 PFASs in 14,000 analyses. Fortunately, 99.7% of the measurements were below the limit of quantification (LOQ), most of which were below 10 ng/L. Nineteen compounds were found in 11 samples, including pesticides and their metabolites, PFASs, and alkylphenols. Levels of pesticides were 2x lower than the quality standards for bottled water in France, and their presence indicated that they had been previously used in the recharge region of the aquifers. The presence of low levels of PFASs suggested long-range atmospheric transport and deposition.

New Emerging Area of Focus: Hydraulic Fracturing. Two years ago, hydraulic fracturing (HF, also called hydrofracking or fracking) was added to this review as a new emerging area of focus. The past two years has seen a continued growth in the number of environmental studies on HF, with many addressing potential risks to the environment. HF is accomplished by high-pressure injection of millions of gallons of water and addition of surfactants, sand, and chemicals (including biocides) deep into the ground to fracture shales and extract natural gas into horizontally drilled wells. Most shale gas development is still occurring in the U.S., with many different shale regions being exploited and more than 7000 shale gas wells in Pennsylvania in the Marcellus Shale. HF is now quickly expanding to other countries, to develop their own energy sources. Costa et al. published a review covering five years of shale gas development, with studies on environmental impacts.10 Impacts on water, air, land, seismic activity, occupational and public health, and safety are discussed.

In September 2017, Environmental Science & Technology (ES&T) and ES&T Letters organized a virtual issue containing 28 recent papers (2015-2017) on unconventional oil and gas. Vengosh et al.


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wrote the introduction to that virtual issue and provided an insightful summary of knowledge gaps for unconventional energy development.11 These knowledge gaps include the need for air pollutant emission studies for regions beyond the Barnett and Marcellus Shale (where most studies have been conducted todate); overall and long-term impact on global warming and potential measures for mitigation; characterization of organic and inorganic contaminants in HF wastewater and their interactions; ecological effects of unconventional oil and gas activities; long-term effects of aquifer contamination; whether HF fluid additives or geogenic chemicals and byproducts are drivers of toxicity; technologies to treat and reuse HF wastewater; and the magnitude and duration of human exposure to hazardous air and water pollutants emitted by unconventional oil and gas activities.

In a new study by Luek et al., Fourier transform ion cyclotron resonance mass spectrometry (FTICR-MS) and Orbitrap MS/MS were used to identify molecular formulas for a large number of new compounds in HF wastewaters that have not been reported previously.12 Numerous chlorinated, brominated, and iodinated compounds were observed, with more than 800 formulas containing iodine. Possible origins of these compounds include chemical additives, shale, and biotic or abiotic reactions occurring within the fluids. The higher abundance of the halogenated compounds in flowback water compared to older HF wastewater suggested that these compounds may be associated with HF additives and subsurface reactions (e.g., reactions with chloride, bromide, and iodide and organic matter from the shale with strong oxidant additives). It should be noted that none of these matched known chemical additives.

In another study by Hoelzer et al., GC-MS and GCxGC-MS were used to comprehensively identify compounds in HF wastewater from the Fayetteville Shale.13 Some of the identified compounds were geogenic (e.g., hydrocarbons and hopanes), while others appeared to be undisclosed additives and compounds, including phthalates, radical initiators, halogenated hydrocarbons, and chloromethyl



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alkanoates. It was proposed that the halogenated methanes and acetones were formed as unintended byproducts.

NEW REGULATIONS/REGULATORY METHODS Recent U.S. Rules and Regulations include the U.S. Environmental Protection Agency’s (EPA’s) new Contaminant Candidate List-4 (CCL-4) and the fourth Unregulated Contaminant Monitoring Rule (UCMR-4).

The Fourth Unregulated Contaminant Monitoring Rule (UCMR-4). The new UCMR-4 was finalized in December 2016 and requires drinking water utilities to monitor for 30 contaminants; it will provide national occurrence data for priority unregulated contaminants for future regulatory consideration (www.epa.gov/dwucmr/fourth-unregulated-contaminant-monitoring-rule). This Rule supports the Safe Drinking Water Act and Amendments, which requires every 5 years, a list of no more than 30 unregulated contaminants to be monitored nationally. For the new UCMR-4, ten cyanotoxins (algal toxins) will be monitored over a 4-consecutive month period from March 2018 to November 2020; additional contaminants will be monitored over a 12-month period from January 2018 to December 2020. Cyanotoxins include total microcystins (measured by an enzyme-linked immunoassay [ELISA] method, EPA Method 546), six microcystins, nodularin, anatoxin-a, and cylindrospermopsin, which are measured using EPA Method 544 and 545. Other contaminants include two metals, eight pesticides, one pesticide manufacturing byproduct, three brominated haloacetic acids groups, three alcohols, and three other semivolatile chemicals, as well as total organic carbon (TOC) and bromide, measured as indicator chemicals for DBPs. Table 3 lists the UCMR-4 contaminants and the approved methods for measuring them.


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The New Contaminant Candidate List-4 (CCL-4). In November 2016, the U.S. EPA finalized the CCL-4, the new drinking water priority contaminant list for regulatory decision making and information collection. The contaminants listed either occur or are anticipated to occur in drinking water systems and will be considered for potential regulation. The CCL-4 contains 97 chemicals or chemical groups and 12 microbial contaminants (Table 4), including chemicals used in commerce, pesticides, biological toxins, DBPs, pharmaceuticals, and waterborne pathogens. Many of the chemicals and all of the pathogens on this list are the same as what was previously on the CCL-3. The U.S. EPA made a few changes to the original proposed Draft CCl-4 based on available occurrence and health effects data, which included the addition of two nominated contaminants (manganese and nonylphenol), removal of perchlorate and strontium (which are scheduled for regulation), and removal of 4 contaminants that the U.S. EPA decided not to regulate (1,3-dinitrobenzene, dimethoate, terbufos, and terbufos sulfone). Additional information on the CCL-4 and the process used to create this new list can be found at: www.epa.gov/ccl/contaminant-candidate-list-4-ccl-4-0.

New Regulatory Methods for Drinking Water. Since the last Analytical Chemistry review, one new method, EPA Method 546, has been published for measuring total microcystins in drinking water and ambient water. It is listed along with five other recent methods developed by the U.S. EPA in Table 5. These methods are mostly directed toward the measurement of CCL, UCMR, or regulated chemicals in drinking water. The U.S. EPA will also soon be publishing two ambient water (freshwater) methods: one for microcystins and nodularin, and one for cylindrospermopsin and antoxin-a. These methods should be publically available by the end of 2017. The U.S. EPA’s National Exposure Research Laboratory and the U.S. EPA’s Office of Ground Water and Drinking Water are responsible for most new EPA methods, which can be found at: www.epa.gov/dwanalyticalmethods/analytical-methods-developedepa-analysis-unregulated-contaminants.



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EPA Method 546: Total Microcystins and Nodularins. In September 2016, a new EPA method was created for measuring total microcystins and nodularins in drinking water: EPA Method 546, ‘Determination of Total Microcystins and Nodularins in Drinking Water and Ambient Water by Adda Enzyme-Linked Immunosorbent Assay’ (www.epa.gov/dwanalyticalmethods/analytical-methodsdeveloped-epa-analysis-unregulated-contaminants). This method uses an enzyme-linked immunosorbent assay (ELISA) to measure total microcystins and nodularins in water, and is based on detection of a characteristic feature (Adda amino acid side chain) common to microcystin and nodularin congeners. The ELISA is calibrated against one congener, microcystin-LR, and has a method detection limit (MDL) of 0.26 µg/L. This method was developed in support of the UCMR-4.

SUCRALOSE AND OTHER ARTIFICIAL SWEETENERS Artificial sweeteners are popular in foods and soft drinks throughout the world. Many are very stable and primarily enter the environment through treated wastewater, following direct disposal of products or through urine after consumption of products. While most studies have been carried out in Western countries, a new study published in 2016 showed significant levels of artificial sweeteners in surface waters from the Philippines (acesulfame, sucralose, cyclamate, and saccharin), in ground water from Vietnam (acesulfame), and surface waters from Myanmar (acesulfame and cyclamate).14 Most studies have focused on sucralose and acesulfame, but saccharin, cyclamate, aspartame, neotame, stevioside, glycyrrhizic acid, and neohesperidine dihydrochalcone have also been measured. Due to their stability and high levels found in the environment (often at ppb levels), sucralose and acesulfame are recognized as potential tracers of anthropogenic inputs into environmental waters and are starting to replace caffeine for that purpose. Through their measurement, the percent wastewater contamination in groundwaters can be estimated. A new study in 2017 by Blackstock et al. also reported high levels of acesulfame in swimming pools and hot tubs, illustrating its potential as a urinary marker for the control of


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pool water quality.15 In that study, >250 samples were collected from 31 pools and hot tubs in two Canadian cities; all samples contained acesulfame, with levels ranging from 30-7110 ng/L, which were up to 570x greater than the incoming tap water to the pool facilities. LC-MS/MS was used for measurement. The authors were also able to use the data to estimate urine contribution in two pools that were sampled over a three week time period. The input of urine to each was estimated at 30 and 75 L, respectively.

While these artificial sweeteners are approved for human use, concerns mostly surround their potential for adverse ecological effects. Bioaccumulation is not expected to be an issue, but there can be impacts on the behavior of aquatic organisms, and new research is revealing toxicity for sweetener transformation products (TPs). The last two years has seen an increase in the number of studies showing toxic effects for these artificial sweeteners and their TPs. For example, Saucedo-Vence et al. found effects on the gills, muscle, brain, and liver of carp when exposed to environmentally relevant levels (0.05 and 155 µg/L) of sucralose.16 Sucralose was not found to bioaccumulate in the fish, but induced oxidative damage in lipids and proteins. Another study by Li et al. reported zebra fish embryo toxicity for acesulfame photolysis products, with low g/L levels producing adverse effects in tail detachment, heart rate, hatching rate, and survival rate during embryonic development.17 This photolysis can occur naturally in sunlight, and the authors discovered six new TPs that have not been previously reported. LCion mobility (IM)-quadrupole (Q)-time-of-flight (TOF)-MS coupled to a LC ion trap mass spectrometer was used to identify their structures, and a transformation pathway was proposed.

In another fish toxicity study, Ren et al. reported increased oxidative stress in the liver of Carassius auratus, which were exposed to acesulfame UV photolysis products for 7 days.18 Concentrations of 0.1 and 10 mg/L UV-irradiated acesulfame was used for these experiments. Acesulfame itself did not induce oxidative stress. The authors identified eight TPs using UPLC-MS/MS.



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Photochemical reactions were also the focus of another study by Perkola et al., who carried out experiments on acesulfame, sucralose, saccharin, and cyclamic acid using simulated solar radiation (>290 nm) and UV irradiation.19 UV irradiation degraded all artificial sweeteners, but only acesulfame was found to degrade with solar radiation. Further, acesulfame degraded more than three orders of magnitude faster than the other artificial sweeteners. Three TPs were identified: iso-acesulfame, hydroxylated acesulfame, and hydroxypropanyl sulfate.

Chlorination of acesulfame was carried out in a controlled laboratory study by Li et al., who identified ten DBPs, including two new chlorinated products reported for the first time.20 Interestingly, five of these DBPs—chlorite, chlorate, dichloroacetic acid, dichloroacetamide, and trichloroacetamide— have been reported in chlorinated drinking water. UPLC-QTOF-MS with high resolution was used to generate exact masses of the unknown DBPs, and a reaction pathway was proposed. The authors also demonstrated that the chlorination byproducts were up to 1.8x more toxic than the parent compound (to Vibrio fischeri in the Microtox assay), with a dynamic trend over time, indicating continued formation of DBPs.

Ozonation of sucralose was the focus of another paper by Hu et al., who investigated the kinetics, removal efficiency, and influence of pH, humic acid, and carbonate on sucralose degradation.21 Ozonation was found to initiate by the formation of OH radicals, and rates were significantly higher at pH 7 than 4. Several TPs were identified using a combination of UPLC-high resolution-MS/MS and GC-MS, and a reaction pathway was proposed. TPs included aldehydes, carboxylic acids, and chlorine-containing products (sucralose has three chlorines in its parent structure).

Castronovo et al. reported the biodegradation and formation of TPs from acesulfame in wastewater treatment and sand filters.22 Varied removal at 13 wastewater treatment plants (WWTPs) was


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observed, which was not associated with nitrification. Ion chromatography (IC)-HR-MS was used to identify a prominent TP: sulfamic acid, which was suggested as potentially being a better tracer for wastewater than acesulfame, due to its chemical and microbiological persistence, negligible sorption, and elevated concentrations in WWTP effluents.

Robertson et al. investigated the degradation of sucralose in Canadian groundwater associated with septic systems over a six year period.23 Sucralose was found in all septic tank effluents and nearby groundwater, up to 98 and 77 µg/L, respectively. An apparent slow rate of degradation in the groundwater, suggested that sucralose could be used as an indicator of recent wastewater contamination, and identifying groundwater recharged after 2000 (its earliest use in Canada). Finally, Zirlewagen et al. examined the use of cyclamate and acesulfame for quantifying wastewater contributions in a karst spring.24. Two different scenarios were investigated: (1) a rain-on-snow event in the winter when no sewage overflow occurred, and (2) an intense rainfall event in summer that triggered a combined sewer overflow. Acesulfame was found in all karst spring samples, but cyclamate, which is more degradable, was only detected in the second scenario with the sewage overflow. Cyclamate also correlated with the breakthrough of fecal bacteria, indicating it could be a good marker of sewer overflows. On the other hand, acesulfame is much more stable and could be used as a marker for older contamination.

NANOMATERIALS Nanomaterial (NM) research continues to be a hot topic, with studies investigating their occurrence, fate, and toxicity in the environment, as well as new NM development for new purposes. Nano-silver (nAg) is still the dominant one, with uses in medical bandages, gym socks, t-shirts, food containers, baby blankets, towels, and children’s toys. Graphene research has increased, and studies continue on fullerenes, nanotubes, nZnO, nCeO2, nAu, and nFe(0).



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New toxicity and uptake studies have been published the last two years. First, Canesi and Corsi summarized the effects of nanoparticles (NPs) on marine invertebrates, which were recognized as significant targets for NPs.25 Accumulation of NPs is mainly observed in the digestive tract and gills, and most data is available for bivalves, polychaetes, and echinoderms. The primary biological responses include immuno-modulation, oxidative stress, and embryo toxicity. Interactions of NPs with biomolecules (e.g., polysaccharides, proteins, and colloids) can affect their aggregation and bioavailability. When NPs are inside of an organism, they can interact with plasma proteins, forming a protein corona that can affect uptake and toxicity.

Ebbs et al. investigated the accumulation of zinc, copper, and cerium ions and NPs in carrots.26 The NPs and metal ions were separately mixed in the sand that the carrots were grown in. Interestingly, the NPs were no more toxic than the metal ions and showed lower accumulation in the edible part of the carrot. Another study by Bjorkland et al. found low bioaccumulation of carbon nanotubes (CNTs) in plants, invertebrates, and non-mammalian vertebrates.27 None of the invertebrate and non-mammalian vertebrates demonstrated significant absorption of the NPs from the gut to other tissues. Impacts of silver NPs on a natural estuarine plankton community was the subject of another study by Baptista et al., who found that phytoplankton and bacterioplankton population growth rates were significantly reduced by silver NPs at concentrations ≥500 µg/L.28 Grazing rates also declined, and the authors concluded that the silver NPs did not cause a cascade of effects through the food web, but impacted a specific trophic level.

Potential impacts of in situ remediation by nano zero valent iron (nFe(0)) was reviewed by Lefevre et al.29 nFe(0) has been used for remediation of soil and groundwater, and there is concern that it could interact with and impact microbial communities. Toxicity studies indicate that nFe(0) can disrupt cell membranes and cause oxidative stress in microorganisms through the generation of Fe2+ and reactive oxygen species (ROS), and also that microbes have adaptive responses to this toxicity. In addition,


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evidence shows that nFe(0) can alter the taxonomic and functional composition of indigenous microbial communities.

Life cycle assessment (LCA) has become a popular tool to quantify the overall burden of contaminants to the environment. Pourzahedi et al. carried out a LCA of silver NPs (AgNPs) that extends beyond their potential releases and ecotoxic effects to potential impacts from their production and disposal.30 Fifteen different consumer products containing AgNPs were investigated from cradle to grave. Interestingly, releases of AgNPs from products with highest leaching rates showed a much lower environmental impact than from the production of these products. The contribution of AgNP synthesis ranged from 1-99%. Regarding release of AgNPs from materials, solid polymeric samples lost more Ag during washing compared to fibrous materials. In fact, highest silver releases to the aquatic environment were from food containers, up to 7.92 mg. The authors also recommended that LCA and other assessments should consider environmental burdens and risks associated with products overall and not only focus on marginal impacts from the incorporation of nanomaterials. Mitrano et al. also took a LCA approach for studying AgNP release from textiles.31 Larger releases from textiles occurred during the use phase of the life cycle (through laundering) rather than the disposal phase (in the landfill).

Selck et al. published a EU-U.S. perspective on the status of ecotoxicity testing, research priorities, and future challenges.32 Key future research topics included NP characterization in environmental and biological matrices, NP transformation in the environment, consequences for bioavailability and toxicity, alternative methods to assess exposure, influence of exposure scenarios on bioavailability and toxicity, development of more environmentally realistic bioassays, and uptake, internal distribution, and depuration of NPs.



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Li et al. addressed the extent that WWTP effluents influence the occurrence of AgNPs in surface waters.33 Two WWTPs in southern Germany were studied, along with upstream and downstream measurements in the River Isar. The authors found that >96.4% of AgNPs from wastewater influent were removed by the WWTPs. At the same time, these WWTP effluents were the source of significant increases of AgNPs to the river, but levels decreased substantially at locations 1.5 km downstream of each discharge, such that levels became similar to reference lakes with no industrial sources or WWTPs. The reference lakes were suggested to have natural AgNPs.

Meier et al. investigated the transformation of AgNPs in incinerated sewage sludge from WWTPs.34 For this study, AgNPs were added to a pilot WWTP with anaerobic digestion, and the sludge was incinerated in a bench-scale fluidized bed reactor. X-ray absorption spectroscopy and electron microscopy-energy dispersive X-ray analysis revealed that the AgNPs transformed into Ag2S-NPs during wastewater treatment, but then Ag(0) NPs were reformed during incineration.

The transport of AgNPs by runoff and erosion was the topic of another study by Mahdi et al.35 A flume and rainfall simulator were used for these experiments, which demonstrated that AgNPs can be transported by both overland flow and by sediment due to erosion. Concentrations of AgNPs increased with applied rain events, and increasing the slope of the flume increased total transport with runoff sediments. Yu et al. synthesized isotopically labelled AgNPs (107Ag and 109Ag) to investigate the transformation kinetics of AgNPs and Ag ions in the aquatic environment.36 Transformation between AgNPs and Ag+ was complex and greatly depended on the external conditions. Oxidation of AgNPs predominated in simple water solutions containing only these two species; sunlight initially accelerated the dissolution of AgNPs into Ag ions, but prolonged sunlight exposures induced aggregation of AgNPs and reduced the reaction rate. Dissolved organic carbon also had a pronounced effect, with reduction of Ag+ playing a major role.


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Markus et al. modeled the transport of ZnO, TiO2, and Ag NPs along the Rhine River.37 Six scenarios were examined, including the following: NPs only released via wastewater; NPs released from land through runoff of sludge applied as a fertilizer; and other scenarios based on total emissions from the second scenario. Results indicate that NPs can be transported over long distances, similar to suspended particulate matter. Nearly 1/3 of free NPs are expected to attach to suspended particulate matter due to aggregation. Finally, Adeleye and Keller investigated the interactions between extracellular polymeric substances from freshwater and marine algae and commercial TiO2 NPs (nTiO2) in the aqueous environment.38 Naturally occurring extracellular polymeric substances modified the surface properties and fate of nTiO2 in environmental waters; interactions were driven by electrostatic interactions and chemical bonding between the COO- group of the extracellular polymeric substances and the nTiO2.

PER- AND POLYFLUOROALKYL SUBSTANCES Perfluorinated compounds have been recently renamed as per- and polyfluoroalkyl substances (PFASs) to distinguish them from other simpler perfluorinated compounds that contain only carbon and fluorine atoms (and are not toxic). The most common PFASs found in the environment are perfluorooctanoic acid (PFOA) and perfluorosulfonate (PFOS), but there are many other PFASs, including new 4-carbon PFASs (e.g., perfluorobutanoic acid [PFBA] and perfluorobutane sulfonate [PFBS]) and GenX (perfluoro-2-propoxypropanoic acid) that are in replacements for PFOS and PFOA. New classes of PFASs continue to be discovered, including perfluoroalkyl ether carboxylic acids and sulfonic acids, new polyfluorinated carboxylic acids and sulfates, and new anionic, zwitterionic, and cationic PFASs reported the last two years.



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PFASs have unusual chemical properties, repelling both grease and water, and contain one of the strongest chemical bonds (C-F) known in chemistry. PFASs are used in food packaging, fabrics (e.g., ski jackets), carpets, nonstick cooking pans, paints, adhesives, electronics, personal care products, and firefighting foams. More than 3000 PFASs have been or are currently on the global market.39 Building materials, including composite woods, were also recently identified as another source of PFASs.40 PFASs are of concern because they are highly stable in the environment and accumulate in red blood cells. Nearly every person measured has detectable levels of PFOS and PFOA in their blood.

PFASs are globally transported, particularly atmospherically, and PFAS precursors (‘PreFAAs’), such as dipolyfluoroalkyl phosphates, fluorotelomer alcohols, perfluorooctyl sulfonamides, and sulfonamidoethanols, can biotransform into PFOA, PFOS, and other PFASs. PFASs can also accumulate in plants and crops when PFAS-containing biosolids are applied as fertilizer, and PFASs are mobile in groundwater. 41 PFASs have also been measured in drinking water throughout the world.41 Health concerns for PFASs include cancer, reproductive and developmental effects, endometriosis, bioaccumulation, immunotoxicity, ulcerative colitis, and thyroid disease.1 PFASs can also be transferred from mothers to babies from breast milk and across the placenta. In addition to humans, PFASs have been reported at high levels in polar bears, seals, and killer whales in locations far away from the manufacture and use of these substances.

North America phased out PFOS in 2002, and the U.S phased out PFOA in 2015. In November 2016, the U.S. EPA issued a health advisory for PFOA and PFOS in drinking water at a concentration of 70 ng/L (for the sum of PFOA and PFOS) (www.epa.gov/ground-water-and-drinking-water/drinkingwater-health-advisories-pfoa-and-pfos). PFOA and PFOS are currently on the final CCL-4 (www.epa.gov/ccl/contaminant-candidate-list-4-ccl-4-0), and six PFASs were on the UCMR-3: PFOA, PFOS, perfluorononanoic acid (PFNA), perfluorohexane sulfonic acid (PFHxA), perfluoroheptanoic acid


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(PFHpA), and perfluorobutane sulfonic acid (PFBS) (http://water.epa.gov/lawsregs/rulesregs/sdwa/ucmr/ucmr3), such that there are national occurrence data in drinking water. Europe restricts the use of PFOS as part of the European Union’s REACH program (http://ec.europa.eu/environment/chemicals/reach/reach_en.htm), and the Stockholm Convention lists PFOS as a persistent organic pollutant (POP) (http://chm.pops.int/TheConvention/ThePOPs/ListingofPOPs/tabid/2509/Default.aspx) and has recommended that PFOA should also be included. A recent study of >600 American Red Cross blood donors demonstrates that levels of perfluorohexane sulfonate (PFHxS), PFOS, PFOA, perfluorononanoic acid (PFNA), and perfluorodecanoic acid (PFDA) are declining from 2000 to 2015, with 61%, 88%, 77%, 33%, and 50% reductions, respectively.42

In addition to individually measuring specific PFAS chemical species, four methods are now being used to quantify total fluorine. These methods include adsorbable organic fluorine (AOF, using combustion ion chromatography), total adsorbable fluorine, the total oxidizable precursor (TOP) assay, and particle-induced gamma ray emission (PIGE) spectroscopy. The TOP assay works by oxidizing and converting perfluoroalkylacid (PFAA) precursors into PFAAs, which can easily be measured by LCMS/MS, providing a sum of the PFASs that can be converted to PFAAs in the environment. PIGE is a non-destructive, direct measurement technique that has been applied to paper and textiles and directly quantifies elemental fluorine on surfaces.

Wang et al. wrote an excellent review entitled, “A never-ending story of per- and polyfluoroalkyl substances (PFASs)?”, which provides a “family tree” of the different PFASs that are known, along with an explanation for why PFASs are an intractable management issue, and recommendations for future research.39 Issues discussed include the properties of PFASs (including the high solubility and proteinbinding characteristics of ionic PFAAs that challenge conventional bioaccumulation assessments based



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on octanol-water partitioning coefficients), the large number of PFASs present (and the vast amount of unknown structures), and the lack of effective control measures. Recommendations include targeted research efforts, with a first step of establishing an inventory of legacy and currently-used PFASs, identifying relevant processes for fate and transport, understanding the relationships between molecular structures and their properties and behavior, understanding PFASs as a group and identifying the drivers of the mixture toxicity, and developing viable remediation technologies and safe alternatives.

Bost et al. published a fascinating study on the use of domestic cats as sentinels for PFASs.43 In this study, the serum of 72 pet and feral cats was sampled for nine PFASs. Interestingly, PFAS means in the cats were very similar to humans, as published in the U.S. National Health and Nutrition Examination Survey (NHANES). Highest PFAS levels were found in indoor cats, with unusually high PFHxS levels observed. Total PFAS levels were positively associated with living indoors and with a higher body weight. It was suggested that indoor cats could be sentinels for assessing primary PFAS exposure routes, especially indoor sources relevant to children.

Fluorinated compounds in fast food packaging was the subject of a large study by Schaider et al., who collected approximately 400 samples of food contact papers (e.g., sandwich wrappers and pastry bags), non-contact paper (e.g. outer bags), food contact paperboard (e.g., French fry boxes and pizza boxes), paper cups (for hot and cold drinks), and beverage containers (e.g., milk and juice containers) from fast food restaurants in the U.S.41 PIGE was used to measure total fluorine, and LC-HR-MS was used to identify specific PFASs in a subset of 20 samples. Thirty-three percent of the samples had detectable fluorine concentrations, ranging from 16-800 nmol of F/cm2, most arising from grease-proof products (e.g., food contact papers). Moreover, some of the samples contained PFOA, even though it has been phased out in the U.S. HR-MS indicated a homologous series of unknown polyfluorinated


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compounds, with molecular weights differing by 49.9968 (CF2 group). A significant proportion of the total fluorinated compounds in the samples were unknown polyfluorinated compounds.

Strynar et al. used LC-HR-TOF-MS to discover 12 novel perfluoroalkylether carboxylic acids and sulfonic acids in surface waters in NC.44 Homologous series were found with repeating units of CF2 (m/z 49.9968) or CF2O (65.9917). Negative mass defects (characteristic of halogenated compounds) and differences of m/z 21.9819 (from protonated and sodiated dimers) aided with the structural assignments. Barzen-Hanson also used LC-HR-MS (with a QTOF mass spectrometer) to discover 40 classes of PFASs in aqueous film-forming foams (AFFFs) and in AFFF-impacted ground water.45 New homologues within 17 known classes of PFASs were also identified for the first time in AFFF-impacted groundwater. New classes included anionic, zwitterionic, and cationic PFASs, 34 of which have similar base structures and derive from electrochemical fluorination processes.

Novel PFASs were also reported in surface waters and sediments downstream of manufacturing facilities in Decatur, Alabama, a previously reported hotspot of PFASs.46 Newton et al. used HR-MS to identify nine new polyfluorinated carboxylic acids (PFCAs) that differed by CF2CH2 units. These new PFCAs are believed to be byproducts of a manufacturing process using 1,1-difluoroethene. Two other compounds had structures similar to PFBS and PFHpA, but with a hydrogen atom substituted for a fluorine in the structure. A new polyfluoroalkyl sulfate was also reported.

Dauchy et al. identified 13 emerging PFASs in AFFF concentrates, including 6:1 fluorotelomer thioether amido sulfonate (FTSAS), 4:2 FTSAS, 5:2 and 8:2 FTSAS sulfoxide, 6:2 FTS-C2H4-COOH (hydrolysis product of the FTSAS amide group), 6:2, 4:2, and 10:2 fluorotelomer thio hydroxyl ammonium, and 7:3 fluorotelomer betaine.47 In addition, 154 PFASs were also quantified in AFFFs and water samples located near AFFF-impacted sites in France, and the TOP method was used as a surrogate



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measure of the total fluorinated compounds. The sites included an oil storage depot (where AFFFs had been used to put out a large fire), an international airport, a military airport, and a training center for firefighters. PFAS concentrations at downstream locations were the highest recorded in France and resulted in the closure of drinking water sources.

Drinking water associated with industrial sites, military fire training areas, and WWTPs was the focus of another study by Hu et al., who examined data from the U.S. EPA’s UCMR-3.48 The authors found that the number of such sites was a significant predictor of PFAS detection frequencies and concentrations in drinking water. PFOS and PFOA levels exceeded the U.S. EPA’s lifetime health advisory concentration of 70 ng/L for approximately 6 million U.S. residents.

Drinking water, river water, and sediment located near and far from a major PFAS manufacturing facility was the subject of another study by Boiteux et al., who investigated the concentrations, mass flows, and fate of dozens of PFASs.49 Among the PFASs measured in the river, 6:2 FTSA and 6:2 FTAB were the most predominant and were linked to industrial WWTP discharges. For drinking water, total PFAS concentrations were all below 60 ng/L. Interestingly, Four PFASs were detected as far as 62 km downstream from a manufacturing facility. Numerous unknown PFASs were indicated by the TOP assay.

Landfill leachates were the subject of an Australia-wide assessment of PFASs published by Gallen et al.50 In this study, 27 landfills were sampled. Five PFASs were ubiquitous, with PFHxA the most predominant, up to 25,000 ng/L. Significant associations were observed between PFAS concentrations and landfill age. Younger landfills had a higher burden of wastes containing PFASs and their precursors. In another study, Lang et al. provided a national estimate of PFASs in municipal landfill leachates in the U.S.51 Seventy PFASs were measured in 95 leachate samples. The total volume of


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leachate generated in the U.S. was estimated at 61.1 million m3, with 79% coming from landfills located in wet climates. Further, in most of the leachates, the 5:3 fluorotelomer carboxylic acids were dominant.

Snow deposition was investigated as a source of PFASs in coastal Antarctica by Casal et al.52 Concentrations of 14 ionizable PFASs were one order of magnitude higher in freshly deposited snow than in background surface snow. However, while snowfall was an important source of PFCAs and PFSAs in Maritime Antarctica, it was not a major source of PFOS in seawater.

Several studies investigated the fate of PFASs. Anumol et al. investigated the fate of 6:2 and 8:2 homologues of fluorotelomer unsaturated carboxylic acids in advanced oxidation processes related to wastewater reuse in potable drinking water.53 Following treatment with ozone, ozone/H2O2, and UV/H2O2, these acids transformed almost completely to PFHxA and PFOA, respectively, with almost no other products observed. Thus, it was clear that fluorotelomer compounds can readily oxidize to PFCAs in water using doses relevant to water treatment. Further, ozone/H2O2 treatment showed much greater degradation than UV/H2O2 in the doses tested. Washington and Jenkins investigated the abiotic hydrolysis of fluorotelomer-based polymers (FTPs) as a source of PFCAs.54 This study was done in response to a discrepancy in the literature for previous studies which reported either very long half-lives (1200-1700 years) or much shorter half-lives (10-17 years). This new study reports half-life estimates of 55-89 years, due to hydrolysis at neutral pH. At higher pH (10-12), half-lives can be even shorter (0.7 years). Results also indicate that FTPs could increase PFCAs 4-8x beyond current oceanic loads.

A very interesting study by Yamazaki et al. investigated the impact of the Great East Japan Earthquake of 2011 on the emission, dynamics and transport of PFASs to the ocean.55 This earthquake destroyed buildings and other structures within minutes, releasing PFASs to the environment, and the following tsunami swept them away to coastal waters in hours. Relatively high levels of PFASs were



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measured one month following the earthquake, but one year later, after significant transport by water currents into the ocean, concentrations along the coast returned to background levels. The authors estimated that between 0.8-1.0 tons of PFOS and 4.8-5.1 tons of PFOA were discharged into coastal waters from this event. Simulations using measured PFASs revealed the dynamics of their transport and also estimated the total amount of PFASs stored and used on land before the earthquake.

Sun et al. measured 10 legacy PFASs and seven recently discovered perfluoroalkyl ether carboxylic acids (PFECAs) in the Cape Fear River watershed (North Carolina), as well as their fate in drinking water treatment.56 In raw source waters located in the headwater region of the Cape Fear River basin, PFECAs were not detected, but legacy PFASs were high, with 57 of the 127 sampling days exceeding the U.S. EPA’s lifetime health advisory level of 70 ng/L for PFOA and PFOS. Perfluoro-2propoxypropanoic acid (PFPrOPrA, also known as “GenX”), which is a PFOA replacement chemical, was detected at extremely high levels (up to 4500 ng/L) in a drinking water plant’s raw source waters located downstream of a PFAS manufacturing plant. Six other PFECAs were also detected, three of which had peak areas up to 15x that of GenX. Drinking water treatment with coagulation, ozonation, biofiltration, and disinfection did little to remove them. Powdered activated carbon (PAC) was effective for removing longer-chain PFASs, but was not effective for removing GenX or other perfluorinated ethers. As a result, the authors suggest that GenX presents a greater drinking water challenge than PFOA.

Wang et al. investigated levels, isomer profiles, and estimated discharges of 12 PFCAs and two emerging fluorinated alternatives (6:2 FTS and F-53B) to Chinese rivers.57 Levels varied, with totals up to 1240 ng/L. Branched isomer profiles indicated release during production of PFOA by electrochemical fluorination or use in fluoropolymer manufacture. Results revealed that point source emissions from China are a likely major source of global emissions of several legacy PFCAs and fluorinated alternatives (e.g., F-53B).

Mejia-Avendano et al. studied the generation of PFCAs from aerobic biotransformation


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or quaternary ammonium polyfluoroalkyl surfactants in soil microcosms.58 Perfluorooctaneamido quaternary ammonium salt showed significant degradation and formation of PFOA (30 mol % in 180 days), however perfluorooctane sulfonamide quaternary ammonium salt degradation was very slow, only generating a PFOA yield of 0.3 mol %. This is believed to be due to its stronger sorption to soil because of its longer perfluoroalkyl chain and bulkier sulfonyl group. HR-MS and tandem-MS were used to identify six novel TPs. Biotransformation pathways were also proposed.

Zhang et al. used principal component analysis (PCA), hierarchical clustering, and geospatial analysis to understand source contributions to 14 PFASs measured in surface waters from the northeastern U.S.59 PFOA and PFHxS were found at the highest levels (up to 56 and 43 ng/L, respectively), and PFOS levels have decreased from earlier measurements. Point sources, including airports, textile mills, WWTPs, and the metal smelting industry, were identified in this analysis.

The release of PFASs from carpet and clothing was investigated by Lang et al. using model anaerobic landfill reactors.60 Carpet and clothing samples were obtained in a form that was representative of materials to be disposed in landfills. Carpet samples (used and new scraps) were obtained from a local carpet installer, and clothing (>20 garments, which did not appear to include water-proofed items) were obtained from a second-hand clothing store. Results showed that carpet and clothing are likely sources of PFASs in landfill leachate, but significant releases did not occur until after 100 days. Carpet samples showed greater releases in biotic vs. abiotic reactors, with 5:3 FTCA and PFHxA dominant. Clothing releases were dominated by PFOA, with lower concentrations of the other 69 PFASs measured.

Several new methods were developed for PFASs the last two years. Willach et al. developed a simplified AOF method to study the contribution of PFASs to the total adsorbable organic fluorine in German rivers and a contaminated groundwater.61 Simplification of the method included shortening the



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combustion time to 14 min and accumulation of combustion gases in a single absorption solution, followed by analysis of 1 mL with ion chromatography. Limits of quantification were 0.77 µg/L. For the groundwater, a significant percentage (32-51%) of the AOF could be explained by the 17 individual PFASs measured, but the surface waters had more than 95% unknown AOF. Lin et al. developed a LCOrbitrap-MS method for simultaneously quantifying and identifying perfluoroalkyl sulfonates, fluorotelomer sulfonates, and the emerging 6:2 chlorinated polyfluoroalkyl ether sulfonate in river water.62 Detection limits ranged from 7.1-62 pg/L, and quantification limits from 12-54 pg/L, which are surprisingly comparable to limits from triple quadrupole mass spectrometer methods. Good linearity (r2 > 0.999) and recoveries (63-103%) were also achieved. A benefit of this method is the ability to identify unknown PFASs while providing quantitative data for these fluoroalkyl sulfonates.

Naile et al. developed a new GC-negative chemical ionization (NCI)-MS method to measure isomers and enantiomers of PFCAs in environmental samples.63 Two GC columns were used in tandem (a DB-5MS and a BGB-172 Analytik column), and samples were derivatized with diazomethane prior to analysis. Using this method, eight PFOA isomers could be detected, including four enantiomers. Bach et al. developed a new headspace-SPME-GC-MS method for measuring two perfluoroalkyl iodides, two perfluoroalkyl sulfonamides, three fluorotelomer alcohols, three fluorotelomer iodides, and four fluorotelomer acrylates and methacrylates in water and sediments.64 While LC-MS/MS is mostly used to measure PFASs in water, the simultaneous analysis of non-ionic and ionic PFASs is hampered by ionization suppression of fluorinated telomer alcohols (FTOHs) caused by buffers used in the mobile phases. Also, perfluoroalkyl iodides and fluorotelomer iodides cannot form protonated or deprotonated molecules with electrospray ionization. Thus, this GC-MS method helps to overcome those issues. Limits of quantification of 20-100 ng/L were achieved for tap water samples, along with 76-126% recoveries. This method was subsequently applied to samples near an industrial facility.


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Dasu et al. created a new method for analyzing 14 PFCAs in drinking water using column switching LC-MS/MS.65 This method uses off-line SPE followed by on-line pre-concentration onto a WAX column before analysis on another analytical column with MS/MS. A major advantage of this method is the low sample volume needed (10 mL) vs. other methods that generally require 100-1000 mL volumes. Limits of quantification were 0.59-3.4 ng/L; recoveries ranged from 73-128%. Finally, Gremmel et al. created two new SPE-LC-MS/MS methods for determining 52 PFASs in environmental waters.66 The first method was for acidic PFCAs and PFSAs, as well as their precursors. The second method was for FTOHs and perfluorooctane sulfonamidoethanols, which have very different properties than the PFASs. Quantification limits ranged from 0.3-199 ng/L.

PHARMACEUTICALS AND HORMONES The discharge of pharmaceuticals and their metabolites into the aquatic environment occurs worldwide, as reported by van der Beek et al.67 The authors performed a very comprehensive literature review highlighting that pharmaceuticals have been detected in 71 countries. In total, 631 different human and veterinary pharmaceuticals have been quantified above the limit of detection. It was concluded that globally, the major source for contamination is the discharge of urban wastewater, although emissions from the pharmaceutical industry, agriculture, and aquaculture can be very important locally.

An increasing number of publications is indicating that pharmaceuticals and their metabolites are also present seawater. Arpin-Pont et al. summarized the results of monitoring studies at 50 marine sites, measuring 196 pharmaceuticals and 37 personal care products.68 Predominantly, antibiotics such as erythromycin, sulfamethoxazole, and trimethoprim, as well carbamazepine, ibuprofen, or acetaminophen, were found, with maximum concentrations up to the lower µg/L range.



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The fate and removal of pharmaceuticals in WWTPs and drinking water treatment plants is of high relevance. Yang et al. summarized the removal of a large number of pharmaceuticals in water treatment processes of WWTPs and drinking water treatment plants located in Europe, North America, and Asia.69 They concluded that conventional wastewater treatment and sand filtration are not sufficient to totally remove the selected pharmaceuticals, while ozonation, AOPs, activated carbon, and reverse osmosis can be more efficient. They conclude that several processes, such as membrane filtration, adsorption on activated carbon, and ozonation can be adopted for the removal of ECs. Nevertheless, process stabilities, treatment optimization, and microbial community structures of the biological processes have to be evaluated and improved. Similar results were reported by Petrie et al., indicating that biological wastewater treatment is not sufficient to mineralize pharmaceuticals, which were in most cases only converted into stable transformation products (TPs).70

The number of publications dealing with the detection of veterinary pharmaceuticals is increasing. Two reviews of Snow et al.71 and Kaczala and Blum72 indicated that mainly antibiotics, hormones, and anthelmintic drugs are found in environmental samples. Although those contaminations can be locally very important, the major loads of pharmaceuticals detected in the aquatic environment are caused by pharmaceuticals used for human therapy.

Innovative analytical instrumentation, such as hybrid mass spectrometry, enables the identification and quantification of organic pollutants, such as pharmaceuticals and hormones, down to the picogram per liter and nanogram per kg range in environmental matrices, as well as wastewater and drinking water. While most organic contaminants enter wastewater without being metabolized, pharmaceuticals are frequently transformed in the body, and a combination of non-altered pharmaceuticals and their metabolites are excreted by humans.73,74 Biotic and abiotic transformation of pharmaceuticals and hormones can occur under anaerobic and aerobic conditions, already in the sewers,


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during wastewater treatment, as well in contact with sediments and soil, and during bank filtration. Furthermore, transformation is induced by UV-light in surface waters, as well as during oxidative treatment processes, such as ozonation, advanced oxidation, and chlorine disinfection.

Still, LC-tandem-MS is the method of choice for the quantification of all classes of pharmaceuticals, metabolites, and transformation products in aqueous matrices. ESI and atmospheric pressure chemical ionization (APCI) are the most commonly used LC interfaces. The major innovation can be seen in modern hybrid mass spectrometry systems (e.g. ion trap/FT-MS, QTOF-MS) coupled to liquid chromatography, providing accurate masses of the analytes and mass fragments, which are very useful to identify the chemical structures of TPs. Noguera-Oviedo and Aga wrote a review about studies that are using MS to quantify and to elucidate TPs of pharmaceuticals during biological or oxidative wastewater water treatment.75 A comprehensive overview of the analysis of veterinary pharmaceuticals was published by Obmiakinde et al., providing detailed SPE parameters, as well as the LC-MS/MS detection methods.76 Further innovations are needed in rapid on-line extractions, as well as in on-line derivatization techniques to improve the applications of GC-MS(/MS) detection.

Biological transformation products. Several reviews have been written on the subject of biological transformation reactions of micropollutants in wastewater treatment processes under aerobic conditions.77-80 Transformation reactions of micropollutants can be described as being metabolic or cometabolic.81 Detoxification processes by bacteria are also possible routes of transformation, for example, in the formation of conjugation reactions or conversion to more polar TPs for improved transport out of the cell.82 Typical biological wastewater treatment consists of aerobic and anoxic stages, enabling the transformation of organic compounds and organically bound nitrogen. Due to the higher biological activity in the aerobic stage, it is expected that most transformations of micropollutants including pharmaceuticals will occur under these conditions.83 Currently, high resolution mass spectrometry (e.g.,



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LC-linear ion trap [LTQ]-Orbitrap MS or LC-QTOF-MS) are applied to elucidate the chemical structure of TPs. Frequently, the suggestions of transformation product structures are based on exact masses and mass fragment spectra which need to be confirmed by: i) a thorough interpretation of the MS fragments formed, ii) authentic standards (if available), iii) alternative analytical methods, such as NMR or IR spectroscopy, or iv) chemical reactions, specifically altering the functional moieties present in the molecules.

In the study of Jewell et al., laboratory-scaled bioreactors spiked with 5 µg/L and 500 µg/L trimethoprim were applied to identify the transformations products of the antibiotic trimethoprim using a LTQ Orbitrap Velos mass spectrometer.84 At high initial concentration of 500 µg/L trimethoprim, only two transformations products were found, whereas at 5 µg/L, a completely different transformation pathway was identified, leading to four further TPs. These differences emphasize the importance of using environmentally relevant initial analyte concentrations when elucidating biotransformation pathways, especially for antibiotics. At the low concentration, trimethoprim was demethylated, forming 4desmethyl-trimethoprim, which was then quickly hydroxylated, oxidized, and cleaved, forming 2,4diaminopyrimidine-5-carboxylic acid via TP292 and TP290 formation. (TP292 and TP290 terms indicate transformation products observed at m/z 292 and 290, respectively, for (M+H)+ ions). The relevance of the transformation pathway was confirmed by the detection of 2,4-diaminopyrimidine-5-carboxylic acid in the effluent of an activated sludge reactor system at 61 ng/L, which accounted for 52% of the transformed trimethoprim.

The hydroxylation of aromatic moieties is the first step towards formation of a ring cleavage substrate. Hydroxylation is commonly observed in studies of aromatic micropollutant degradation in activated sludge or in soil environments. Examples are reported by Boix et al. for the hydroxylation of the aromatic moieties of irbesartan and ibuprofen in activated sludge, which were identified via MS2


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spectra using a LTQ Orbitrap XL mass spectrometer, as well as by the formation of 5-hydroxy-diclofenac during incubation of WWTP effluent.86 Heterocyclic aromatics are also known to undergo hydroxylation. For example, the diamino-pteridine moiety of methotrexate was found to be hydroxylated in activated sludge as a minor transformation reaction within a larger transformation pathway.87

Micropollutants with methyl amine moieties can be demethylated during biological wastewater treatment, leading to the formation of primary or secondary amines. For example, the N-demethylation of methotrexate forms a secondary amine87. An N-demethylation of the dimethylamine moiety of citalopram was also found via UPLC-QTOF-MS in contact with activated sludge.88 These reactions are a type of Ndealkylation that are observed with a large variety of micropollutants, also with larger N-alkyl substituents, identified via MS2 spectra using an Orbitrap Q-Exactive mass spectrometer.89 Other reactions of amines commonly found in biological media include N-oxidation and also conjugation, e.g., acyl conjugation. Deamination of the aniline moiety of sulfamethoxazole has also been observed in enriched cultures of ammonia-oxidizing bacteria (AOB), as well as oxidation to a nitro group.90

Aldehyde moieties, which are frequently formed during oxidative transformation, e.g., after the cleavage of amines, are in many cases oxidized to carboxylic acids.89 The decarboxylation of diclofenac measured via LTQ-Orbitrap-XL-MS by Poirier-Larabie et al. resembles the decarboxylation of 4hydroxycinnamic acid to 4-hydroxystyrene.86

Conjugation reactions of pharmaceuticals (or other chemicals) in the human body are well known.91 These reactions have been recently observed in activated sludge, e.g., the sulfate conjugation of triclosan92 and other phenolic compounds, such as dextrorphan,93 as well as N-acylation, N-formylation, and N-succinylation of amines.89



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The biotransformation of diclofenac during wastewater treatment was reported by Jewell et al.94 Laboratory-scale incubation experiments were performed with diclofenac and carriers from a moving bed biofilm reactor of the WWTP in Bad Ragaz, Switzerland, and LC-HR-QTOF-MS was implemented to monitor the biotransformation. Chemical structures were proposed for 20 TPs formed by reactions, such as hydroxylation, decarboxylation, oxidation, amide formation, ring-opening, and reductive dechlorination.

The fate of five antiviral drugs (abacavir, emtricitabine, ganciclovir, lamivudine, and zidovudine) was investigated in biological wastewater treatment. For all five compounds, an oxidation of the terminal hydroxyl-moiety to the carboxylic acid was observed.95 In addition, the oxidation of thioether moieties to sulfoxides was found for emtricitabine and lamivudine. Antiviral drugs were detected in influents of municipal WWTPs with concentrations up to 980 ng L-1 (emtricitabine), while in WWTP effluents, mainly the TPs were found with concentrations up to 1320 ng L-1 (carboxy-abacavir). The concentrations of the TPs ranged from 16 ng L-1 for carboxy-lamivudine to 750 ng L-1 for carboxy-acyclovir.

Although anaerobic conditions prevail in several environmental compartments, up until now, biodegradation studies with ECs such as pharmaceuticals and personal care products have mainly been investigated under aerobic conditions. The review of Ghattas et al. provides a comprehensive survey of anaerobic biotransformation reactions for a variety of well-studied, structurally simple contaminants bearing one or only a few different functional groups/structural moieties.96 Furthermore, this review summarizes anaerobic degradation studies in water, soil, and sediment of more complex contaminants with several functional groups. Abiotic transformation products. Recently, Hirte et al. described the profound study about the hydrolysis of amoxicillin (AMX), a beta-lactam antibiotic.97 The authors confirmed that not only the hydrolysis rate, but also the pattern of TPs was strongly pH-dependent. Three primary TPs (AMX


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penicilloic acid, AMX 2´,5´-diketopiperazine, and AMX penilloic acid) formed by cleavage of the beta lactam ring. Via a subsequent decarboxylation, these primary TPs were even further converted into 23 secondary/tertiary TPs. Thus, a multitude of subsequent reactions follow the first attack of AMX by a water molecule.

Suspect screening. Bade et al. used UPLC-QTOF MS and LC-LTQ-Orbitrap-MS for suspect screening of 107 pharmaceuticals and illicit drugs in WWTP influents and effluents, as well as surface water.98 The aqueous samples were extracted by SPE using Oasis HLB. In total, 28 compounds were detected in Castellon (Eastern Spain) and Cremona (Northern Italy). Irbesartan, valsartan, benzoylecgonine, and caffeine were the most commonly found compounds. The screening results obtained by UPLC-QTOF-MS and LC-LTQ-Orbitrap-MS were mostly comparable, while Orbitrap-MS was able to identify a few more compounds.

Pochodylo and Helbling developed a suspect screening workflow to characterize the occurrence of ECs in water samples from the urban water system of New York State.99 The workflow for the HRMS data (Orbitrap, Q Exactive) contained peak picking, suspect database matching, isotopic pattern scoring, replication filter, blank subtraction, and clustering of suspect hits. The suspect database consisted of 1113 organic chemicals, including pharmaceuticals, personal care products, and pesticides. As a result, the authors were able to identify the presence of 112 ECs in at least one of the 18 samples.

Another suspect screening method was reported for the analysis of pharmaceuticals in WWTPs.100 The authors used an automated approach based on LC-HR-MS, combined with a modelbased prioritization using consumption data, and readily predicted fate properties and a generic mass balance model for activated sludge treatment. In total, 27 new pharmaceuticals were identified, which had not been monitored before. The authors concluded that analytical suspect screening is more selective



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than model-based prioritization, but the modelling results are complementary to analytical screening.

Psychoactive substances. An analytical suspect method for psychoactive substances in influents from two WWTPs, in combination with target analysis using deuterated internal standards, was reported by Baz-Lomba et al.101 The authors concluded that suspect screening can help to identify new psychoactive substances and to include them afterwards into the target method for quantification. A similar procedure was suggested by Andres-Costa et al.102 Solid phase extraction and subsequent UPLCQTOF-MS detection were applied for rapid suspect screening, as well as for the determination of 42 illicit drugs and metabolites, in water. Several ECs were identified using a database that contains more than 2000 compounds that ionize in positive mode. This method enabled the identification of psychoactive compounds in different water matrices. Yao et al. described the determination of 10 illicit drugs and metabolites in aqueous samples by online SPE coupled to LC-MS/MS.103 Detection limits down to 1 ng/L were achieved. The method was applied to surface water samples of Shanghai and four Chinese WWTPs. Pseudoephedrin was detected in concentration up a few ng/L, while 3,4-methylenedioxymethamphetamine (MDMA), 3,4-methylenedioxy-amphetamine(MDA), benzoylecgonine and methadone were not detected. Bade et al. reported a method for the analysis of 10 new psychoactive drugs of cathinon and phenethylamine classes in WWTP influents.104 The method was based on SPE with Oasis MCX prior to UPLC-MS/MS detection. Concentrations in the lower ng/L range were found.

Elimination during oxidative water treatment. Rayaroth et al. reviewed the degradation of pharmaceuticals by ultrasound-based AOPs.105 The authors considered the following processes: sonocatalysis, sono-ozonolysis, sonoelectrochemical, sonophotocatalysis, sono-Fenton, and sonophotoFenton. Maximum degradation of organic compounds was observed at 100-1000 Hz, which represents the high frequency, medium power ultrasound. In general, hydroxyl radicals are generated by pyrolytic cleavage of water molecules. Even though almost all the experiments achieved more than 90% removal


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of the target compounds, the efficiency of the degradation varied with water matrices and pH. Major reactions included hydroxylation, dehalogenation, demethylation, decarboxylation, and deamination.

Trawinski and Skibinski reviewed studies on photodegradation of psychotropic drugs.106 Photocatalysis was concluded as the most promising and more effective process in comparison to regular photolysis. Predominant analytical methods to identify and to determine TPs were LC-MS/MS or LCOrbitrap-MS, but GC-MS was also used after derivatization, as well as NMR, if sufficient quantities were available.

Borowska et al. reported the transformation of the antihistamine drugs cetirizine and fexofenadine, as well as the diuretic drug hydrochlorothiazide, by ozonation.107 At pH 7, all three compounds were very reactive with ozone. The main TPs of cetirizine and fexofenadine were N-oxides, whereas hydrochlorothiazide was converted into chlorothiazide. However, at least seven TPs were identified for each pharmaceutical using the LC-Orbitrap Q Exactive mass spectrometer.

Chon et al. studied the changes of UV absorbance at 254 nm (UVA254) and electron donating capacity as surrogate indicators for assessing: i) the removal of micropollutants, and ii) bromate formation during ozonation of wastewater.108 The electron donating capacity was measured by size .

exclusion chromatography and a post column reaction with ABTS+ (reaction of chlorine with ABTS = 2,2´-azino-bis(3-ethylbenzothiazoline-6-sulfonic acid), which was calibrated against an electrochemical method. The study results indicated that a combination of UVA254 and the electron donating capacity reflects better the intrinsic reactivity of dissolved organic matter (DOM) associated with EC removal and bromate formation than the single use of UVA254 as a surrogate indicator.

Ma et al. described the chlorination of mefenamic acid, tolfenamic acid, and clofenamic acid.109



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The TPs were identified via SPE with C18 material and a UPLC-LTQ-Orbitrap XL mass spectrometer, as well as by ion pair liquid-liquid extraction in ethylacetate using tetrabutylammonium-hydrogensulfate (TBA-HSO4) as an ion pair reagent for anionic TPs and subsequent GC-MS detection. In the hot GC-MS injector, the TP-TBA ion pairs decomposed into TP-butyl derivatives, which were finally detected. TPs were identified, which formed by chlorination and/or hydroxylation of the aromatic rings. Up to two chlorines and two hydroxyl groups were inserted.

Moreira et al. reported the removal of amoxicillin and diclofenac using ozonation and AOPs, such as photolysis, photolytic ozonation, photocatalysis, and photocatalytic ozonation.110 Diclofenac was quickly degraded by direct photolysis (medium pressure vapor arc, >> 300 nm), while amoxicillin was stable. In the presence of ozone, both were rapidly degraded, but not mineralized. However, using photocatalytic ozonation, both compounds were quickly mineralized.

Duan et al. described the degradation of the iodinated pharmaceuticals thyroxine and diatrizoate.111 Direct photolysis led to roughly 40% degradation of both pharmaceuticals, while UV/S2O82- significantly enhanced the degradation. A degradation pathway was proposed involving deiodination and hydroxylation.

Hormones Several studies tried to analyze estrogens down to the pg/L range in order to control the suggested enviromental quality standards for estradiol (400 pg/L) and 17α-ethinyl estradiol (35 pg/L), which are on the Watch List for the European Water Framework Directive. However, neither the disk-based solid phase extraction nor the other methods were able to reach such low concentrations at the pg/L level in natural surface water.112 Even promising passive sampling using Empore SDB-RPS disks were not able to reach such low concentration levels.113


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In the study of Guedes Alonso et al., an extraction method based on sorptive fabric phase (FPSE) coupled to UPLC-MS/MS was developed for the analyis of four progesterones and six androgens in environmental and biological samples.114 Detection limits down to 1.7 ng/L were achieved. Maximum concentrations ranged from 227 ng/L (for progesterone in hospital wastewater) to 3.5 µg/L (for androstenedione in urine). The transformation of progesteron was investigated in a lab-scale study with surface water by Ojoghoro et al.115 In total, five TPs were identifed using LC-QTOF-MS. which were formed by hydroxylation, hydrogenation, dehydrogenation, side chain cleavage, and ring cleavage. Cytostatics. Cesen et al. developed an analytical method to determine the cytostatics ifosfamide (IF) and cyclophosphamide (CP) and their metabolites in hospital wastewater, as well as in influents and effluents of WWTPs.116 The method consisted of SPE using Oasis HLB (keto-CP, CP, IF) and Isolute ENV+ (carboxy CP and N-deCl-CP). Prior to GC-MS detection, the analytes were derivatized by silylation using n-(tert-butyldimethylsilyl)-N-methyltrifluoro-acetamide (MTBSTFA) (for metabolites and TPs) and acetylation using trifluoroacetic anhydride and heptafluorobutyric anhydride (for IF, CP). In hospital wastewater, the analytes were detected at elevated concentrations (e.g., CP: 2.7 µg/L; carboxyCP: 13.2 µg/L), while all analytes were below LOQs (in the lower ng/L range) in WWTP influents and effluents. Speed disc extraction and GC/detection. Drinking water analysis was conducted by speed extraction followed by a simultaneous derivatization with dimethyl(3,3,3-triflouropropyl)silyl diethylamine (DIMETRIS) for beta blockers, hormones, and antiphlogistics, down to a few ng/L-range.117 Onsite large volume solid phase extraction (LVSPE). Onsite large volume solid phase extraction (LVSPE) was reported by Schulze et al.118 The design allows the collection and extraction of 50 L water samples. Extracts of LVSPE were measured using a LC-ESI-LTQ Orbitrap XL mass



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spectrometer. The authors used Chromabond HR-X, HR-XAW, and HR-XCW as sorbents. In total, 251 analytes were extracted, with log D ranging from –3.6 to 9.4. Proton pump inhibitors. A review summarized the analysis, occurrence, fate, and risks of proton pump inhibitors, their metabolites, and TPs in the aquatic environment.119 Since proton pump inhibitors are extensively metabolized in the human body, in general they are not detected in the aquatic environment or only in the lower ng/L range. However, profound studies to determine their metabolites and TPs are still mainly missing. Multi-methods. The detection of a relatively high number of ECs at environmental concentrations is feasible due to materials enabling the simultaneous solid phase extraction of compounds with a broad range of polarities and universal LC-MS/MS detections that does not require derivatization. The development of multi-methods to determine ECs including pharmaceuticals and hormones is facilitated due to the recent develepments in LC-MS/MS techniques.70,120 Aminot et al. reported a multimethod to determine 53 pharmaceuticals in aqueous samples after SPE using LC-MS/MS.121 Furthermore, they were able to determine these analytes also in sediments and suspended matter after microwave-assisted extraction. The occurrence of 67 pharmaceuticals and antifungal residues was reported by Chitescu et al. using SPE and detection via LC-Q Exactive Orbitrap-MS.122 For most analytes, detection limits lower than 5 ng/L could be achieved. In total, 23 analytes were detected up to 166 ng/L (diclofenac), due to the high sensitivity of the analytical method. Valls-Cantenys et al. reported a multi-residue method for the determination of 35 ECs, including pharmaceuticals and iodinated X-ray contrast media.123 The use of 20 isotopically labeled surrogate standards allowed for compensation of matrix effects. ESI enabled a better sensitivity than APCI, although it is known to be more sensitive for matrix effects. Ruff et al. described the monitoring of 302 compounds, including many pharmaceuticals, along the Rhine River using SPE containing a mixture of Oasis HLB, Isolute ENV+, Strata XCW cation exchanger, and Strata anion exchanger.124 For detection, an Orbitrap LTQ mass spectrometer was used.


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For Lake Constance, 83 compounds were quantified at low concentrations; at the German/Dutch border, 143 ECs were detected. The determination of pharmaceuticals in estuarine and coastal systems was described by BaenaNogueras et al.125 The authors were able to reach limits of detection below 1 ng/L using solid phase extraction with Oasis HLB (500 mg) at pH 2.5 and detection with LC-MS/MS. Forty-nine of 83 target compounds were found in the coastal systems of Spain, with concentrations up to 195 ng/L. The importance of appropriate quality assurance/quality control (QA/QC) measures when using multi-methods for the determination of ECs in water was highlighted in a new U.S. nationwide survey.126 The authors reported the analysis of 247 ECs in source and treated drinking water samples from 25 drinking water treatment plants. Six multiple methods based mainly on LC-MS/MS and one on GC-MS was used to analyze 174 pharmaceuticals, personal care products, and pesticides. The authors conclude that for the majority of analytes, the predetermined method performance criteria were met. However, for 16% of the original compounds, improvements are still needed. Overall measurement uncertainties and the implementation of QA/QC measures were discussed for a case study at the Seine River.127 The outcome demonstrated that the comparability of the results from different laboratories was significantly influenced by uncertainties caused by different sampling procedures. New SPE materials/procedures. A universal analytical method based on stir bar sorptive extraction (SBSE) was developed for the detection of ECs, including pharmaceutical, hormones, and UV filter substances, using LC-MS/MS.128 The stir bars were coated with ethylene glycol-modified silicon (EG-silicone) allowing a simultaneous extraction of both polar and non-polar pollutants. The method was successful for the determination of 48 compounds in surface water and drinking water.



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Dynamic fabric phase sorptive extraction (DFPSE) with solgel Carbowax 20M material combined with LS-MS/MS detection was developed to determine pharmaceuticals and personal care products in influents and effluents of WWTPs, as well surface waters.129 Another option is bar adsorptive micro extraction (BAµE) coated with N-vinylpyrrolidone polymers, combined with micro-liquid desorption, followed by LC-diode array detection (DAD).130 Coupling MS detectors with BAµE would be most promising to achieve low LOQs. Becerra-Herrera et al. reported a method to quantify inflammatory drugs by using an improved rotating-disk sorptive extraction (RDSE) procedure and ultra ESI-TOF-MS detection.131 RDSE was able to combine high recoveries with a reduction of the extraction time. Molecular imprinted polymers (MIPs). Ansari and Karimi discussed new developments and all kind of uses of molecular imprinted polymers (MIPs) as innovative sorbents and fibers in SPE and SPME and as sensitive sensors for the analysis of more than 20 different pharmaceuticals.132 However, MIPs are always designed for individual pharmaceuticals. Although the lowest LOQs reported are in the higher ng/L or µg/L range (e.g., diclofenac: 0.74 µg/L; ketoprofen: 0.3 mg/L), the potential of MIPs to isolate individual pharmaceuticals from complex matrices is very promising, as further clean up steps can be avoided. Thus, if individual pharmaceuticals have to be determined in very complex matrices, MIPS might be an option, but currently they are not able compete with multi-component LC-MS/MS analysis. Surface enhanced Raman spectroscopy (SERS): Patze et al. reported that surface enhanced Raman spectroscopy (SERS) can be used to detect the antibiotic sulfamethoxazole down to 2.2×10-9 mol/L.133 Enantiomers. Brienza and Chiron developed an analytical method to determine the enantiomers of climbazole in wastewater and sludge with limits of quantification down to 1 ng/L.134 The authors used a Lux amylose-2 chiral column under isocratic conditions for separation. The detection was done with an


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Exactive Orbitrap mass spectrometer using an ESI interface. Camachos-Munoz et al. developed a multimethod for the enantioselective analysis of 13 pharmaceuticals in influents and effluents of wastewater treatment plants.135 The detection method focused on a supercritical fluid-based chromatography coupled to a triple quadrupole mass spectrometer. Two coated, modified 2.5 µm-polysaccharide-based chiral columns were applied. Detection limits down to the lower ng/L range were achieved. Camachos-Munoz and Kasprzyk-Holdern created another multi-method for the simultaneous enantioselective analysis of chloramphenicol, ibuprofen, ifosfamide, indoprofen, ketoprofen, naproxen, and praziquantel.136 The authors used macrocyclic glycopeptide-based columns with antibiotic teicoplanin as chiral selector.

DRINKING WATER AND SWIMMING POOL DISINFECTION BYPRODUCTS Drinking Water DBPs. DBPs are formed by the reaction of disinfectants (chlorine, chloramines, ozone, chlorine dioxide, UV) with organic matter, bromide, and iodide. They are an unintended consequence of killing harmful pathogens in drinking water to make it microbially safe to drink. While approximately 700 DBP have been identified, only 11 are regulated by the U.S. EPA. In addition to natural organic matter (NOM) that is the main precursor, many anthropogenic pollutants can also serve as precursors. For example, in this Review, DBPs are reported from tamoxifen metabolites, benzodiazepine drugs, benzotriazole, benzothiazole, linear alkylbenzene sulfonate surfactants, a flame retardant, a UV filter, and the algal toxin, microcystin-LR. Health risks include cancer (bladder and colorectal), miscarriage, and birth defects, with bladder cancer showing the most consistency in human epidemiologic studies. An important issue regarding bladder cancer is whether increased bromide concentrations in source waters may increase the bladder cancer risks. This is because bromide reacts with NOM and disinfectants to form brominated DBPs, some of which are suspected to contribute to bladder cancer. As a result, Regli et al. published a paper estimating the potential bladder cancer risk due to increased bromide concentrations in source waters.137 Results based on data from 201 drinking water plants indicated that a bromide increase of 50 µg/L could cause a potential increase between 10-3 and 10-4 excess



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lifetime bladder cancer risk in populations served by approximately 90% of these plants. Those increases are significant (DBP regulations are built around a risk of 1 in a million, 10-6 risk) and are quite possible with increased bromide currently occurring in many cities due to hydraulic fracturing, coal-fired power plants, other industrial discharges, or drought.

A new epidemiologic study was published the last two years that is highly significant and worthy of mention. Wright et al. discovered an association of cardiac birth defects with specific DBPs for babies born in Massachusetts.138 Bromoform, dichloroacetic acid, trichloroacetic acid, and the five regulated haloacetic acids (HAA5: chloro-, bromo-, dichloro-, dibromo- and trichloroacetic acid) were all associated with different types of cardiac birth defects; bromoform showed consistent results with every cardiac birth defect examined. This study lends further concern for brominated DBPs, and the authors suggested that brominated acetonitriles should be examined in further epidemiologic studies, as bromochloroacetonitrile has been reported to cause cardiac defects in rats.

Antibiotic resistance was the focus of a highly significant toxicology study by Li et al., who found that two DBPs—iodoacetic acid and chlorite—had antibiotic-like effects that led to the evolution of resistant E. coli at both high and low exposure concentrations.139 The authors suggested that DBPs like these may be important contributors to the spread of antibacterial resistance globally.

Michael Plewa and Elizabeth Wagner at the University of Illinois are responsible for generating the most toxicology data for DBPs to-date. In 2017, they compiled their vast data for mammalian cell cytotoxicity and genotoxicity for 103 DBPs, providing a detailed presentation of the methodology for the quantitative, comparative analyses on the induction of cytotoxicity and genotoxicity.140 This paper was published in a special issue of Journal of Environmental Sciences (2017) entitled, ‘Disinfection Byproducts in Drinking Water, Recycled Water and Wastewater: Formation, Detection, Toxicity, and


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Health Effects”. Many researchers from the international DBP community, including chemists, toxicologists, engineers, and epidemiologists, published a paper in this special issue—it is a must-read.

Plewa et al. also published a new approach called “TIC-Tox” that can be used to identify forcing agents of toxicity in drinking water.141 This approach combines analytical chemistry data with biological data, multiplying either the total ion current (TIC) observed for DBPs (by GC-MS) or the quantitative concentrations of DBPs by the cytotoxicity or genotoxicity indexes (Tox) for these compounds. The authors used this approach to examine data from a European drinking water epidemiologic study and found that while regulated trihalomethanes (THMs) and HAAs make up a large portion of the overall DBPs quantified, they are a very small component of the overall cytotoxicity. On the other hand, unregulated nitrogen-containing DBPs (N-DBPs), such as haloacetonitriles and haloacetamides, appear to be the major forcing agents of toxicity in these samples.

DBP groups are included on the new UCMR-4: HAA5 (the five regulated HAAs), HAA6Br (the five brominated HAAs), and HAA9 (the chloro- and bromo-HAAs) (www.epa.gov/dwucmr/fourthunregulated-contaminant-monitoring-rule). Total organic carbon (TOC) and bromide are also included on the UCMR-4 as indicator compounds (higher TOC values are typically indicative of higher DBP levels, and higher bromide levels usually translate into higher levels of more toxic brominated DBPs formed). Several other DBPs are included on the new CCL-4, including 5 nitrosamines: N-nitrosodimethylamine (NDMA), N-nitrosodiethylamine (NDEA), N-nitrosodipropylamine (NDPA), N-nitrosodiphenylamine (NDPhA), and N-nitrosopyrrolidine (NPYR); two aldehydes (formaldehyde and acetaldehyde); and chlorate (www.epa.gov/ccl/contaminant-candidate-list-4-ccl-4-0).

New Methods. Several new methods were published for DBPs the last two years. First, Teng et al. developed a new method for non-target identification of peptides and peptide DBPs in drinking



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water.142 This method involved integrated multiple SPE, LC-HR-MS/MS using precursor ion exclusion (PIE), with two complementary LC columns (C18 and HILIC), and searching against the Human Metabolome Database. Using this strategy, 617 peptides were detected in finished drinking water, with 116 identified as DBPs (present in the finished drinking water and not in the raw source water). Twentyfive compounds were confirmed with authentic standards and reactions pathways were proposed. In a follow-up study from the same group, Huang et al. reported the first identification of N-chlorinated dipeptides using complementary HR-TOF-MS and quadrupole ion trap-MS techniques.143 This was done by first reacting three model dipeptides with chlorine, followed by the complementary use of HR-MS and MS/MS to identify the DBPs. A sensitive LC-MS/MS method was then created to analyze for these Nchloro-dipeptide DBPs in real drinking water, where N-chloro-Tyr-Gly, N,N-dichloro-Tyr-Gly, N-chloroPhe-Gly, N-chloro-Tyr-Ala, and N,N-di-chloro-Tyr-Ala were identified. Sayess and Reckhow reported an improved method for total organic iodine (TOI) in drinking water.144 This method used a low pH (1) adsorption of organics from drinking water onto granular activated carbon (GAC), followed by combustion of the GAC, trapping of combustion products in a 2% (v/v) tetramethyl ammonium hydroxide solution (TMAH), and use of a 0.1% (v/v) TMAH wash solution with ICP-MS detection. The primary difference in this method and previous similar ones is the use of ICP-MS for more sensitive detection of the I- ions produced from combustion. Kimura et al. reported a total organic halogen (TOX) method for human urine, using a similar GAC-combustion procedure, where total organic chlorine (TOCl), total organic bromine (TOBr), and total TOI are converted to chloride, bromide, and iodide, respectively, which are then measured using either ion chromatography or ICP-MS. 145 The goal of this method was to be able to use urine as a noninvasive way to assess human exposure to DBPs and other halogenated compounds (such as brominated flame retardants). The final procedure consisted of 1:50 dilution of a 50 mL urine sample with water, adjustment to pH 2, passing this through three consecutive activated carbon (AC) columns, followed by a


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rinsing step with 10 mL of KNO3 solution. The three ACs were then pyrolyzed individually in a quick furnace, and gases were collected in an aqueous solution containing 0.01 mM phosphate and 0.003% hydrogen peroxide. The solution was then analyzed for chloride and bromide with IC and iodide with ICP-MS. Method detection limits were 1.0, 1.0, and 0.14 µg/L for chloride, bromide, and iodide, respectively. Recoveries ranged from 78-99%, and this method was used to measure TOCl, TOBr, and TOI in five human volunteers who consumed chloraminated tap water. Separate controls were done with volunteers consuming spring water without DBPs. TOCl and TOI in urine were higher than the controls, however, TOBr was slightly lower in urine than the controls, indicating other source of exposure to brominated compounds than drinking water. Finally, Li et al. developed a new GCxGC-quadrupole mass spectrometry method to identify unknown DBPs in drinking water.146 More than 500 compounds were tentatively identified in chlorinated, chloraminated, and ozonated water samples, with 41 of the DBPs confirmed using authentic standards. In addition, a free online quantitative structure reactivity (QSAR) model was used to predict their genotoxicity and carcinogenicity. Brominated and iodinated DBPs. Due to their increased toxicity relative to chlorine-containing DBPs, Br- and I-DBPs have been a major focus of research. Hladik et al. published a particularly creative study that investigated a new source of I-DBPs: use of iodine sanitizers in the dairy industry.147 Dairies from three states were investigated, and iodo-THMs were found within dairy processing facilities and in surface waters that received the dairy wastewater. Iodo-THMs comprised 15-29% of the total THMs in surface waters located near the WWTP effluents. In another creative study, Pan et al. reported the first iodo-DBPs from cooking with simulated chlorinated and chloraminated tap water, iodized salt, and wheat flour.148 UPLC-MS/MS using m/z 127 precursor ion scan (for iodine) was used to identify the DBPs, including 3-iodo-4-hydroxybenzaldehyde, 3-iodo-4-hydroxybenzoic acid, 3-iodo-4-hydroxy-5methylbenzoic acid, diiodoacetic acid, 3,5-diiodo-4-hydroxybenzaldehyde, 3,5-diiodo-4-hydroxybenzoic



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acid, 2,6-diiodo-4-nitrophenol, 2,4-diiodo-6-nitrophenol, and 2,4,6-triiodophenol, which were confirmed with authentic standards. A developmental toxicity assay indicated that the phenolic I-DBPs were approximately 50-200x more toxic than aliphatic I-DBPs. Han et al. reported the identification of new Cl-, Br-, and I-DBPs in chlorine dioxide treated water, including the report of a new class of DBPs: trihalomethanols.149 Liquid-liquid extraction (LLE) was used for extraction, and a four-time LLE enabled much more material to be extracted as compared to a single LLE (which is often done). UPLC-MS/MS was used for detection. Two of the trihalomethanols—bromodichloromethanol and trichloromethanol—were confirmed with authentic standards. Interestingly, these trihalomethanols were found to be TPs from THMs formed during chlorine dioxide disinfection. Postigo et al. used the new GC-Orbitrap mass spectrometer to identify I-DBPs in chlorinated and chloraminated water.150 HR-electron ionization (EI)-MS/MS provided accurate masses that could be assigned for both molecular ions and fragments. Two new DBPs: chloroiodomethane and ethyl iodoacetate were identified in chloraminated water for the first time. Zhang et al. reported a significant finding regarding another new source of iodine in the formation of I-DBPs: iodate (IO3-).151 Iodate is regarded as a final oxidation product of iodide, but this study found that UV treatment of iodate at doses of from 50-500 mJ/cm2 can form iodide and subsequently, I-THMs, in simulated drinking water treated with chloramines. Under realistic treatment conditions, with a UV dose of 50 mJ/cm2, DOC of 5.0 mg/L, iodide of 12.7 µg/L, and NH2Cl dose of 5 mg/L, dichloroiodomethane was detected at 0.17 µg/L using a SPME-GC-MS method. Another interesting study by Hao et al. explored the photobromination and photoiodination of DOM in freshwater and seawater.152 Results using negative ion ESI-FT-ICR-MS revealed the formation of organobromine and organoiodine compounds just with sunlight in water (no chlorine or other


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disinfectant). Artificial seawater produced more organobromine and organoiodine compounds than freshwater. High resolution data indicated that the brominated species were mostly hydrocarbons with low oxygen content, along with unsaturated aliphatic compounds, saturated fatty acids, and carbohydrates. In contrast, iodinated species were carboxylic-rich alicyclic molecules, composed of esterified phenolic, carboxylated, and fused alicyclic structures. In a groundbreaking paper, Zhu and Zhang developed a model for the formation of TOCl, TOBr, and TOI in chlorinated and chloraminated drinking water, which can predict their concentrations and total chlorine residual.153 From the modeling results, 57-74% of the total I-DBPs in chlorination were converted to their chlorinated and brominated analogues by substitution with HOCl and HOBr. During chloramination, chloroiodamine was predicted to be an active intermediate species that was responsible for 41-50% of the total I-DBPs, of which, 52-53% underwent deiodination via base-catalyzed hydrolysis. This study provides insights into the kinetic reactions that occur during chlorination and chloramination of drinking water. GAC treatment was the focus of another study by Krasner et al., who investigated the formation of regulated and emerging DBPs when GAC treatment is used with chlorination under controlled laboratory conditions.154 Increased formation of Br-DBPs, particularly dibromoacetonitrile, was observed with the use of GAC due to higher bromide-to-TOC ratios following the use of GAC. Using the TIC-Tox approach cited earlier, the N-DBPs, haloacetonitriles and halonitromethanes, were found to be the driving agents for predicted genotoxicity. Flint. In 2016, Flint, Michigan became the center of attention due to high levels of lead in drinking water, which was uncovered by Marc Edwards. The city of Flint had switched from a fairly pristine source water (Lake Huron) to an impaired and corrosive source water (the Flint River) and also stopped adding the corrosion inhibitor during this switch. As a result, lead leached out of the pipes, and lead levels often exceeded the U.S. EPA’s action levels. During this crisis, residents of Flint were also



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complaining about skin rashes while taking showers, and DBPs, including dichlorobenzene, were suspected as a possible cause. As a result, a study was done to comprehensively assess DBPs in hot and cold Flint tap water. Allen et al. published this study in 2017, detailing the quantification of 61 regulated and emerging DBPs and 50 volatile organic compounds (including dichlorobenzene), along with comprehensive, broadscreen unknown identification with GC-MS.155 Because the city of Flint was back on Lake Huron water during this time, the city of Detroit was also sampled (it also uses Lake Huron water and chlorinates), along with two other cities in Georgia that also use chlorination. While a number of Cland Br-DBPs were detected in Flint hot and cold tap water (and some I-DBPs were identified in one of the Georgia tap waters), there was not anything unusual that was found in Flint waters that stood out as something that could be responsible for the skin rashes. In addition, dichlorobenzene was not detected. As a result, it was suggested that possibly an inorganic chemical or microbial contaminant may be responsible. Point-of-use filters. An important study by Stalter et al. investigated whether point-of-use filters (such as a Brita filter) can effectively remove DBPs and toxicity from chlorinated and chloraminated tap water.156 Absorbable organic halogen (AOX—which is also known as TOX) was used as a surrogate measure for known and unknown DBPs, and three toxicity assays were used to assess cytotoxicity, oxidative stress response, and genotoxicity. Reverse osmosis (RO) filtration and one of the activated carbon filters were the most effective for removing AOX (>94% removal). Seven out of the 11 filters tested reduced cytotoxicity, oxidative stress response, and genotoxicity by >60%. It was suggested that activated carbon filters could provide an important short-term health benefit through the removal of DBPs and toxicity, but regular filter changing is important because bacterial counts increase after filtration, due to biofilm growth on the filter matrix. Formation mechanisms. Le Roux et al. investigated the role of aromatic precursors in the formation of haloacetamides (HAMs) with chloramination.157 Previous research has shown that HAMs


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can form by hydrolysis of haloacetonitriles, but also from other independent pathways directly from NOM. This new research used 15N-labeled NH2Cl to trace the origin of the nitrogen in HAM structures and revealed a two-step profile, including rapid formation (with second order kinetics), incorporating the N from amine groups in the NOM molecules, followed by a slower (and linear) increase incorporating the N from NH2Cl. Yu and Reckhow investigated the formation of a new N-chloro acetamide: N-chloro-2,2dichloroacetamide (N-Cl DCAM) in chlorinated drinking water.158 N-chloro DBPs have been largely overlooked, and this work builds on the earlier discovery by Marinas’ group at the University of Illinois, who first reported evidence for N,2-dichloroacetamide and discovered issues with quenching. In this new study, the authors confirmed that dichloroacetonitrile (DCAN) can degrade to form N-Cl-DCAM, and it was suggested that N-Cl-DCAM has been incorrectly identified as DCAM due to the reduction of the Nbound chlorine by common quenching agents used to quench the free chlorine before measurement. Further, N-Cl-DCAM was predicted to be stable in most drinking water distributions with free chlorine. The authors also developed a new UPLC-ESI-QTOF-MS method to measure seven N-Cl-HAMs, and reported two new brominated N-Cl-HAMs, N-chloro-2,2-bromochloroacetamide, and N-chloro-2,2dibromoacetamide, as well as N-Cl-DCAM, in real drinking water samples. How et al. published an excellent critical review on the formation and stability of Nchloroamides, N-chloramines, N-chloramino acids, and N-chloraldimines (considered together as organic chloramines), along with their occurrence, toxicity, and analytical methods used for their determination.159 The authors also addressed challenges for their analysis in drinking water, including the lack of standards, lack of analytical methods with low detection limits in real drinking waters, short shelf-life for chloramine standards, and issues with degradation due to quenching agents. While most researchers assume that the active chlorinating species in free chlorine systems is HOCl, Lau et al. further examined the role of Cl2 and Cl2O in the formation of DBPs.160 In reactions of



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chlorine with phenol, the presence of chloride (Cl-) was found to be a catalyst and increase chlorination rates at pH < 6, which was attributed to formation of Cl2. At pH 6, under typical chlorine doses used in drinking water treatment, Cl2 was shown to contribute >80% to the overall chlorination rate, whereas Cl2O was a minor player in these reactions with phenols. Wastewater reuse. Due to increasing populations and water scarcity, treated wastewater is being increasingly considered for potable drinking water. Both Singapore and the U.S. have been doing indirect wastewater reuse for some time, and the number of direct potable reuse plants is increasing with a large plant currently being built in El Paso, TX. Wastewater is treated to a much higher standard in these cases, and generally involves microfiltration, RO, and UV-hydrogen peroxide (called “Full Advanced Treatment trains”, following conventional secondary or tertiary wastewater treatment. Then, further disinfection for drinking water treatment (using chlorine or another disinfectant) is generally done. Due to the high energy consumption (electricity) required for RO, potable reuse facilities are considering RO-free alternatives. Chuang and Mitch evaluated one of these alternatives—ozone and biological activated carbon (BAC), followed by chlorination or chloramination—and investigated the formation of 35 regulated and unregulated halogenated DBPs, along with 8 nitrosamines, and bromate.161 BAC treatment resulted in much lower DBP formation at 15 min empty bed contact time (EBCT), and not much further reduction at higher EBCTs, which was attributed to lower dissolved oxygen concentrations inhibiting biological activity. THMs, HAAs and bromate levels were below regulated levels for ozone-BAC treatment, but NDMA exceeded the California notification limit. Using the TIC-Tox approach to calculate DBP-associated toxicity, ozone-BAC treated wastewater disinfected with chloramines showed calculated toxicity levels comparable or slightly higher than traditional microfiltration-RO-UV-H2O2 treatment. Unregulated DBPs, including haloacetonitriles (HANs), HAMs, and haloaldehydes (HALs), were the drivers of the overall calculated toxicity.


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Another wastewater reuse study by McCurry et al. focused on an issue encountered by several water utilities: re-formation of NDMA after advanced oxidation processes (AOPs).162 The authors found that low pH in the RO permeate, coupled with rejection of ammonium ion, promoted the conversion of residual NH2Cl to dichloramine, which is the main reactive disinfectant species to form NDMA. Further, subsequent addition of lime used by utilities to raise the pH for control of corrosion in the distribution system converts amine-based NDMA precursors to their more reactive, neutral forms. To minimize the re-formation of NDMA, the authors recommended reducing the time period between RO treatment and final pH adjustment. The occurrence of nitrosamines and a suite of regulated and unregulated DBPs was assessed in five potable reuse plants in the U.S. by Zeng et al.163 Full Advanced Treatment was used at each of these plants. Low µg/L levels of THMs, HAAs, dichloroacetonitrile, and dichloroacetamide were seen in the initial secondary and tertiary wastewater effluents, with more variability in nitrosamines and up to 320 ng/L NDMA. The use of ozone resulted in higher formation of nitrosamines, HALs, and HAMs, but BAC could then reduce these levels. When chloramine was used before microfiltration (to prevent fouling of these membranes), the highest levels of DBPs were observed. UV-based AOPs effectively removed nitrosamines, but only partially removed other halogenated DBPs, including HAAs, HALs, haloketones (HKs), and trichloronitromethane. The driver of the calculated toxicity (using TICTox) was HANs, which were not effectively removed by RO and AOPs. Low molecular weight DOM was monitored in another wastewater reuse study by Phungsai et al., which used an Orbitrap mass spectrometer to examine transformation and compositional changes following further treatment with biofiltration, ozonation, and chlorination.164 The most intense ions were in the m/z 100-450 range, and compounds containing carbon, hydrogen, oxygen, and sulfur were the most common. Biofiltration was found to preferentially remove CHO-only compounds with high H/C ratios. Ozonation mostly reduced CHOS compounds and CHO compounds with unsaturated structures, producing more saturated and oxidized products. Chlorination produced 168 chlorine-containing DBPs.



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Nitrosamine methods. Two new nitrosamine-related methods are worthy of mention. First, Breider and von Gunten created a new total nitrosamine method using UV photolysis and chemiluminescence detection of nitric oxide.165 UV photolysis was used to cleave the nitroso group to form nitric oxide, which could be sensitively detected at nM levels. Detection limits for nitrosamines ranged between 0.07-0.13 µM, and an r2 of 0.98-0.99 was achieved. This method provided some advantages over the widely celebrated total nitrosamine assay (TONO) created by Bill Mitch’s group, including the lack of chemicals required for denitrosation, increased reproducibility, and the availability of a commercial photoreactor and NO analyzer. In another study, Liao et al. investigated the use of polarity rapid assessment method (PRAM) to characterize nitrosamine precursors and understand their removal in drinking water treatment processes.166 The PRAM method uses SPE cartridges in parallel to isolate and fractionate nitrosamine precursors at ambient pH. Compared to classical resin fractionation, PRAM was found to have a higher selectivity for nitrosamine precursors, such that 50x higher cationic material could be isolated. Nitrosamine formation, occurrence, and fate. Studies continue to probe nitrosamine formation occurrence, and fate. Natural attenuation of NDMA precursors was the focus of a new study by Woods and Dickenson.167 The urban, wastewater-dominated wash examined showed significant nitrosamine precursor reduction along its flow to a spot nine hours downstream. Photolysis and biological degradation were indicated as the removal mechanisms, and not loss due to sorption to sediments. West et al. investigated nitrosamine formation in simulated drinking water treatment with three disinfectants— chlorine, chloramine, and peracetic acid—in the presence of amine precursors.168 This study represents the first time the experimental disinfectant, peracetic acid, has been examined for formation of nitrosamines. Eight nitrosamines were measured, with highest levels observed with chloramination (as observed by others), much lower levels with chlorine, and no nitrosamines observed with peracetic acid. Finally, Bei et al. reported nitrosamine occurrence from a large-scale study of drinking water in China.169


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Forty-four cities from 23 provinces and 155 sampling points were included. NDMA was found in 33% of the finished water from the drinking water plants and in 41% of the drinking water collected at the tap. Increased NDMA levels can often be at the tap because NDMA can continue to form in the distribution system from the treatment plant to the home. Overall, average NDMA levels of 13 ng/L were observed in tap waters, with highest levels coming from the Yangtze River Delta in East China, which had average levels of 28.5 ng/L and formation potential of 204 ng/L. Pollutant DBPs. While NOM is a major precursor in the formation of DBPs, many environmental pollutants/contaminants can also form them when chlorine or other disinfectants are used in wastewater treatment. When not well removed from wastewater treatment, these contaminants also reach downstream drinking water sources and can form DBPs in drinking water when they do. Often, the DBPs of these contaminants are more toxic than their parent compounds, and several examples of this are cited below. Negreira et al. published a very interesting study on DBPs formed by the breast cancer drug tamoxifen and its metabolites.170 Tamoxifen is a pro-drug, such that its metabolites are the active therapeutic species. While tamoxifen was found to be relatively unreactive with chlorine, its metabolites quickly reacted, forming 13 chlorinated DBPs that were tentatively identified using UPLC-quadrupoleOrbitrap mass spectrometry. Time-course profiles revealed their formation in real wastewater treated with chlorine, and QSAR models predicted the DBPs are up to 110x more toxic than their parent compounds. Benzodiazepine drugs, which are the most commonly prescribed psychoactive drugs for anxiety, depression, confusion, insomnia, or panic disorder, were the focus of another study by Carpinteiro et al. 171 Chlorination of the four drugs studied—diazepam, oxazepam, nordazepam, and temazepam—produced 10 DBPs that were tentatively identified using LC-QTOF-MS. Reaction mechanisms involved attack of chlorine at the 1,4-diazepine seven-membered ring, followed by ring opening or formation of a six-membered pyrimidine ring. Benzophenone and quinazoline derivatives were observed. QSAR models suggested that some DBPs were more toxic and/or mutagenic than the parent molecules.



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The flame retardant, tetrabromobisphenol A, was the focus of another DBP study by Gao et al., who found that chlorine produced products by electron transfer in addition to electrophilic substitution.172 Eight DBPs were identified, including a quinone-like product, two dimers, and several halogenated phenols. LC-MS/MS with a precursor ion scan approach was used to identify the products. Proposed mechanisms include the initial oxidation of tetrabromobisphenol A by chlorine to form a phenoxy radical, followed by beta-scission reactions, substitution, dimerization, and oxidation. Benzothiazoles and benzotriazoles were the focus of another DBP study by Nika et al., who used LC-HR-MS/MS to identify the DBPs formed.173 Four benzotriazoles and three benzothiazoles were reacted with chlorine, and 10 DBPs were identified in the controlled laboratory reactions, as well as in reactions involving real wastewaters. Calculated toxicity in fish and algae using the ECOSAR model suggest that these DBPs are more toxic than their parent compounds. Linear alkylbenzene sulfonate (LAS) surfactants were investigated by Gong et al. for DBP formation with chlorination.174 DBPs were identified using UPLC-ESI-MS/MS with precursor ion scan, and total organic halogen (TOX) was also measured as a surrogate of all halogenated products formed. Major products included 2,6-dichloro-3,5-dihydroxy-4-dodecylbenzenesulfonic acid, and 2,6-dibromo3,5-dihydroxy-4-dodecylbenzenesulfonic acid (when bromide was present). TOX results revealed that DBP formation overall was very small. Toxicity of the reaction products was evaluated in marine polychaete embryos, and results showed the following order of toxicity: Chlorinated samples with bromide > chlorinated samples without bromide > non-disinfected LAS surfactants. The popular UV filter used in sunscreens, avobenzone, was examined for DBP formation by Trebse et al., who used GC-MS to identify the products.175 Twenty-five DBPs were identified when avobenzone was treated with UV and chlorine. Two of the primary DBPs—2-chloro-1-(4-tertbutylphenyl)-3-(4-methoxyphenol)-1,3-propanedione and 2,2-dichloro-1-(4-tert-butylphenol)-3-(4methoxyphenyl)-1,3-propanedione—were confirmed by synthesis and GC-MS and NMR analysis. When


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UV and chlorine were applied at the same time, two unique DBP classes were formed—chloroanhydrides and chlorophenols—that were not formed in separate reactions with UV and chlorine. Finally, Chu et al. examined DBPs from the chlorination of the common algal toxin, microcystinLR.176 THMs, HALs, and HANs were formed, and interestingly, higher amounts of chloroform formed with microcystin-LR than with corresponding free amino acids. Microcystin LR contains five amino acids in its structure, but the Adda group, which contains a conjugated diene, was more reactive. This is consistent with previous studies that have reported DBPs formed by the attack of chlorine on the two C=C bonds in the Adda group. Maximum yields of chloroform, chloral hydrate, and dichloroacetonitrile were 43 %, 3.2 %, and 1.5 %, respectively. Swimming pool DBPs. DBPs also form in swimming pools, due to the reaction of the swimming pool disinfectant with remaining NOM in the incoming tap water or with urine, sweat, personal care products, and other inputs from swimmers in the pool. The pool disinfectant can be the same as used for the drinking water coming into the pool facility or it can be a different disinfectant, which can produce a different array of DBPs. Urine is believed to be a major contributor to many pool DBPs, particularly N-DBPs, and Blackstock et al. were able to use the artificial sweetener acesulfame (which enters pools via urine) to estimate urine contributions to pools.15 Daiber et al. published an extensive study on the progressive increase in DBPs and mutagenicity from the source water to tap to swimming pool and spa water.177 Twenty-eight water samples from seven sites, including private pools/spas and pools and spas from three large university aquatic centers) were examined. GC-MS with low and high resolution was used to comprehensively identify >100 DBPs in this study, and membrane introduction mass spectrometry (MIMS) and GC-electron capture detection were used to quantify many DBPs. Total organic chlorine, bromine, and iodine (TOCl, TOBr, and TOI) were also measured. Three new classes of DBPs—bromoimidazoles, bromoanilines, and bromomethanesulfenic acid esters—were identified for the first time. 4,5-Dibromo-1-methyl-1H-imidazole (identified in brominated spa samples)



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and 2,4,5-tribromo-1-methylimidazole (identified in brominated pool and spa waters) were confirmed with authentic standards. Potential precursors of the new bromoimidazole DBPs were suggested to include the amino acid histidine (introduced via urine) and pharmaceuticals and skin-care products containing an imidazole group. Organic extracts of brominated pool/spa water were 1.8x more mutagenic than those from chlorinated pools/spas, and spa waters were 1.7x more mutagenic than pools. Pool and spa waters also showed considerable increases in mutagenicity (2.4 and 4.1x) compared to corresponding tap waters used to fill these pools and spas, providing evidence that human precursors can increase mutagenic potencies and that this increase is associated with increased DBP formation.

Seawater swimming pools were also included in a study with a freshwater pool in a new paper by Manasfi et al. investigating the DBPs and genotoxicity.178 The three indoor seawater pools were dominated by brominated species (e.g., dibromoacetic acid, bromoform, and dibromoacetonitrile), due to the presence of bromide in the seawater, while the outdoor freshwater pool was dominated by chlorinated species (e.g., trichloroacetic acid, chloral hydrate, dichloroacetonitrile, 1,1,1-trichloropropanone, and chloroform). However, the seawater pools were not as genotoxic as the freshwater chlorinated pool, attributed to lower DBP levels (208 µg/L vs. 948 µg/L for seawater and freshwater pools, respectively) and much lower numbers of people frequenting the seawater pools vs the freshwater pool.

Forty-one public indoor swimming pools were investigated in Quebec in a study by Tardif et al., who measured DBPs in the air and in the water.179 While HAA levels were consistently higher than other DBPs in pool water (mean of 294.8 µg/L and maximum of 886.2 µg/L for total HAAs), overall, DBP levels were variable from one pool to another. NDMA was identified in a subset of pool samples, ranging from 2.8 to 105 ng/L. While chloroform was the most dominant DBP found in air (mean of 119.4 µg/m3), brominated THMs were significant, present at 25% the total THM levels at half of the facilities.


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The effect of medium-pressure UV followed by chlorination on DBPs in chlorinated seawater swimming pools was examined by Cheema et al.180 The authors found that UV followed by chlorine caused dibromo- and dichloroacetic acid to decrease significantly (due to photodegradation), but BrTHMs and Br-HANs to increase. HANs were suggested as the main contributor to the estimated toxicity. Public indoor swimming pools in China were the subject of another study by Hang et al., who measured THMs, HAAs, HANs, HKs, and trichloronitromethane in 13 pools.181 HAAs were dominant, followed by THMs, HANs, HKs, and trichloronitromethane. DBP concentrations in an ozone-chlorine treated pool were lower than those treated with chlorine. Formation of DBPs in warmer temperature environments from trichloroisocyanuric acid (a popular form of chlorine ‘shock’) and bromochlorodimethylhydantoin (BCDMH, a bromine-based disinfectant) were examined by Yang et al.182 Simulated pool water was generated using tap water fortified with body fluid analogue (BFA) and treated with these two disinfectants at elevated temperatures. Higher temperatures increased DBP formation, due to the temperature dependence of reaction rates. Among the BFA ingredients, uric acid, citric acid, and hippuric acid were found to be the primary precursors for HAA formation. Finally, Yue et al. investigated the effect of chloride on the formation of volatile DBPs in chlorinated pools.183 Trichloramine, chloroform, and dichloroacetonitrile linearly correlated with chloride concentrations in both controlled laboratory experiments and actual swimming pool water. It was suggested that the higher chloride results in higher levels of Cl2, which is more reactive than hypochlorous acid (HOCl) and can form higher DBP levels. This is consistent with a drinking water study cited above by Lao et al., who found that chloride can boost chlorination rates at certain pHs.160

SUNSCREENS/UV FILTERS UV filters are used in many products, including sunscreens, cosmetics, shampoos and hair dyes to protect against sun damage to skin or hair. Organic UV filters absorb UV light, and inorganic UV filters reflect and scatter UV light. Organic UV filters are mostly lipophilic compounds with conjugated



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aromatic structures. Several UV filters are usually present together in these products. UV filters are often found in environmental waters, introduced directly from activities such as swimming, or indirectly, by treated wastewater, which can release µg/L concentrations. UV filters can be estrogenic, developmentally toxic, or acutely toxic to aquatic organisms. In fact, UV filters have been suggested to be partly responsible for the bleaching and death of coral reefs. Tsui et al. published an important new study investigating the occurrence, distribution, and fate of UV filters in five coral species collected from the Pearl River Estuary in the South China Sea.184 Seawater and sediment was also sampled. Five UV filters—benzophenone (BP)-1, BP-3, and BP-8, octocrylene (OC), and octyl dimethyl-p-aminobenzoic acid (ODPABA) were detected in the coral tissues, with detection frequencies >65% and highest levels for BP-3 and BP-8 (concentrations up to 31.8 and 24.7 ng/g wet weight, respectively).184 Significantly higher levels were found in the wet season, indicating contribution of sunscreen ingredients from swimmers. Bioaccumulation factors in soft coral tissues ranged from 2.21 to 3.01, and BP-3 exceeded threshold values for larval deformities and mortality in more than 20% of the coral samples collected. Corresponding seawater samples ranged up to 25.8 ng/L (BP-3) for UV filters.

Ramos et al. published an excellent review on organic UV filters in WWTPs, detailing their occurrence and fate in wastewater and sludge, along with treatments that could be used to successfully remove them.185 BP-3 and BP-4, which are among the most hydrophilic, are found at the highest concentrations in wastewater influent and effluent, up to the mg/L range, whereas the more hydrophobic ethylhexyl triazone (EHT), 4-methylbenzylidene camphor (4-MBC), and OC tend to sorb to sludge, with levels up to the µg/g (dry weight). The authors indicated that the highest levels reported in the literature could pose a threat to aquatic ecosystems, and seasonal variation in usage and release may endanger sensitive species, particularly during breeding periods. Of all the treatments potentially available, RO is recognized as the best, but due to its high energy cost, it is not generally used for wastewater treatment.


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Finally, the authors pointed out that most data on UV filters in wastewater treatment come from Europe and Asia, with almost nothing from the U.S. Several good fate papers were published the last two years on UV filters. Li et al. used UPLC with HR-MS to discover 11 photolysis TPs of ethyl-p-aminobenzoate (Et-PABA), including 7 new ones not reported previously.186 Interestingly, the photolysis products were less toxic than the parent compound, as evaluated in the Microtox luminescent bacteria assay. Photolysis reaction pathways mainly involved a series of hydroxylation reactions. Avobenzone was the subject of another fate study (mentioned earlier in the DBP section) by Trebse et al., who investigated its fate under chlorination and UV irradiation.175 Twenty-five DBPs were identified, with two unique DBP classes found when UV and chlorine were used together.

UV filter occurrence in swimming pools and spas was the focus of another paper by Ekowati et al., who measured 14 UV filters in 17 pools from sport centers and hotels from Catalonia, Spain.187 BP-1, BP-2, BP-3, BP-8, trihydroxybenzophenone (THB), 4-dihydroxybenzophenone (4-DHB), 4-MBC, 2ethylhexyl-4-dimethylaminobenzoate (OD-PABA), 1-H-benzotriazole (1HBT), 5-methyl-1Hbenzotriazole (MeBT), and 5,6-dimethyl-1H-benzotriazole monohydrate (DMeBT) were all found in the pool and spa samples, with 1HBT observed the most often and 4-MBC at the highest level (up to 69.3 ng/L). Pools that applied treatment with coagulation, sand filtration, UV irradiation, and salt electrolysis had the lowest occurrence of UV filters, overall.

New methods for UV filters include one by Vila et al., who used headspace-SPME and GCMS/MS to measure 14 UV filters in water.188 Only 10 mL of water sample was required, and limits of detection were very low, ranging from 0.068-12 ng/L. This method was then applied to seawater, river water, spa water, swimming pool water, and aquapark water, where 10 of the 14 UV filters were found, up to 692 µg/L. A novel Fe3O4-graphitized carbon black magnetic material was used in a new dispersive



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SPE-UPLC-MS/MS method created by Polvesana et al., who measured 10 UV filters down to 1-5 ng/L detection limits.189 Magnetic ionic liquids (ILs) were used along with stir bar dispersive liquid microextraction in another method by Chisvert et al. for measuring lipophilic UV filters.190 Compounds sorbed to the magnetic IL-coated stir bar and were thermally desorbed directly into a GC-mass spectrometer. Low ng/L detection limits were achieved, with 1% by weight, with tris(1,3-dichloroiosopropyl)phosphate (TDCIPP) the most common one detected. Results indicated that TDCIPP use declined recently, with increased use of tris(chloropropyl)phosphate (TCIPP) and a nonhalogenated aryl phosphate mixture known as TBPP. GC-NCI-MS and UPLC-MS/MS were used for detection.

Novel brominated compounds were tentatively identified in tetrabromobisphenol A (TBBPA)containing plastics in another study by Ballesteros-Gomez et al.193 Fourteen compounds were identified in plastic casings of electric and electronic devices (e.g., an electric adapter, television, decorative item, and router) that were purchased from stores in the Netherlands. These brominated compounds were likely impurities, byproducts or degradation products of TBBPA. Samples were preselected using an initial screening with direct probe ambient mass spectrometry, after which small plastic samples were cut and extracted and analyzed using HR-QTOF-mass spectrometry. Four of the brominated compounds have not been reported previously. Matsukami et al. used refractive index detection and gel permeation chromatography with atmospheric pressure photoionization (APPI)-QTOF-MS to characterize compounds present in commercial organophosphorus flame retardants that are now being used as alternatives for PBDEs.194 Triphenyl phosphate, tris(dimethylphenyl)phosphate, tris(2-chloroisopropyl)phosphate, and new impurities were identified as byproducts. In another study by Zhang et al., the physical-chemical properties and environmental fate were estimated for 94 halogenated and organophosphate PBDE replacements.195 Using three popular models (EPI Suite, SPARC, and Absolv), properties were predicted



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within 102-103 for compounds of low molecular weight ( 25% of Canada’s oil), but research on NAs is now extending to other locations in the world. NAs are found at high concentrations (up to 120 mg/L) in tailing waters that result from the application of caustic hot water to extract the crude oil. While most research focuses on the characterization of NAs in tailings ponds or nearby groundwater or river water, a drinking water contamination incident was the focus of a study by Wang et al., who attributed the event to an oil spill, and used NAs discovered in these samples as markers.206 UPLC-QTOF-MS was used to identify them. Data pointed to the source as a leaking oil pipeline near the drinking water treatment plant.

Barrow et al. used APPI-FT-ICR-MS to expand beyond what had been detected previously in the Alberta Oil Sands process waters.207 In particular, APPI allowed twice as many compounds to be identified in the process water, river water, and groundwater sampled, including low-polarity sulfur-


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containing compounds and hydrocarbons, which do not ionize effectively with more common ESI methods. Duncan et al. created a new method for NAs using condensed-phase membrane introduction mass spectrometry (CP-MIMS) with ESI, MS/MS, and barium ion chemistry.208 The use of barium ions allowed the formation of barium adducts, (M-H+Ba)+, which improved selectivity in complex samples, eliminating isobaric interferences from hydroxylated species. Huang et al. reported another new method using Ag+-SPE to better separate classical, aromatic, oxidized, and heteroatomic NAs, which could be eluted into separate SPE fractions.209 UPLC-ion mobility (IM)-HR-TOF-MS was used for analysis, with a mass tolerance of +/- 1.5 mDa. In an earlier paper by the same group, their UPLC-IM-TOF-MS method was described and used to measure NAs in Oil Sands process water and ozone-treated process waters.210 Ion mobility provided an second dimension of separation. O2-containing NAs were found in the raw process waters; ozonation further oxidized the NAs, with 2-5 oxygens incorporated into the TPs. NA species with CH2CH2-S groups were also found. Effect-directed analysis (EDA) was used in two interesting studies to determine which chemical classes in the oil sands-impacted water were the most acutely toxic. In the first, Morandi et al. subjected water samples in an end-pit lake to a three-stage fractionation using LLE at different pHs, alkaline water washing, and preparative LC, and tested for acute toxicity in the 96 h fathead minnow embryo assay, as well as in the Microtox assay. 211 The most toxic fractions included the final fraction containing mostly NAs, which contained O2+, O+, OS+, and NO+ species. UPLC-MS/MS was used for identification of compounds in the different fractions collected. In the second study, Yue et al. subjected oil sands process-affected water to sequential SPEs and used the Microtox assay for assessment of acute toxicity.212 Orbitrap-MS was used to determine the chemicals in the most toxic fractions, which included tricyclic NAs containing C15-C18 and O2, as well as bicyclic NAs containing C14-C17 and O2. A few O2 containing compounds negatively correlated with toxicity.



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Oxidation of oil sands process water with ferrate, UV/H2O2, and ozone was evaluated by Wang et al. in a study that used FT-ICR-MS and UPLC-TOF-MS to characterize the products.213 Reaction with these three oxidants resulted in transformation of molecules containing Ox, OxSy, and OxN2, with OxSy showing the most sensitive responses. As a result, the authors suggest that OxSy distribution profiles can be used to assess the effectiveness of the oxidation. Spatial and temporal variability of acid-extractable organics (including NAs) in oil sands process-affected waters were assessed in another study by Frank et al.214 The compounds were analyzed using synchronous fluorescence spectroscopy, ESI-MS, HR-MS, GCxGC-TOF-MS, GC-QTOF-MS, and LC-QTOF-MS. Principle component analysis (PCA) could distinguish water samples arising from different ponds, and GCxGC-TOF-MS and GC-QTOF-MS could distinguish samples temporally and spatially. Spatial samples were much more variable than temporal ones, and results suggest that multiple samples in a tailings pond should be used to more adequately characterize the contamination.

BENZOTRIAZOLES AND BENZOTHIAZOLES Benzotriazoles are widely used as corrosion inhibitors and as UV stabilizers in plastics and polymers. They are highly water soluble, resistant to biological degradation, and are not well removed in wastewater treatment, which contributes to their high levels in the environment (up to mg/L). Benzotriazole is estrogenic and is a suspect human carcinogen. A drinking water guideline limit of 7 ng/L was recently established in Australia for tolyltriazole. Benzothiazoles are also used as corrosion inhibitors and in the manufacture of rubber and other products. Both benzotriazoles and benzothiazoles are high volume production chemicals. A new study by LeFevre et al. demonstrates that benzotriazole can be uptaken in lettuce and strawberries that are irrigated with recycled water.215 Median levels of 13.1 and 67.8 ng/g were observed for lettuce and strawberries, respectively. Additionally, strawberries were also found to accumulate benzotriazole plant metabolites.


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Tap water was the focus of another study by Wang et al., who measured benzotriazoles and benzothiazoles in water from 51 major cities in China.216 Benzotriazoles reached a maximum of 310 ng/L, and benzothiazoles reached a maximum of 1310 ng/L. UPLC-ESI-MS/MS was used for their measurement. A chlorinated benzotriazole, 5-chloro-benzotriazole, which may have been formed as a DBP from chlorination, was also found in several tap water samples. Benzothiazoles measured in tap water were indicated to originate from rubber materials used in drinking water distribution systems, whereas benzotriazoles seemed to originate from contaminated raw water sources.

Washing clothes was investigated as a source of benzotriazoles and benzothiazoles in a recent study by Luongo et al.217 These compounds had been previously found to be widespread in textiles used in clothing, and in this study, 27 textile samples purchased from several retail stores in Sweden were laundered in a commercial washing machine. LC-MS/MS was used for analysis. The authors found that benzotriazoles and benzothiazoles do leach during washing, with benzothiazole found in 85% of the samples at average concentrations of 0.53 µg/g. Interestingly, an orange t-shirt made of polyester was found to contain the highest levels of benzothiazole (mean of 1590 ng/g). The average household wastewater emission for one washing was determined to be 0.5 g for benzothiazole. Benzotriazole was detected less frequently and at lower levels.

Snowmelt from engineered snow storage facilities in Alaska was investigated in another study by Alvey et al. as a source of benzotriazole.218 Benzotriazole elution from the snow melt reached a high of 7.4 µg/L, with similar levels observed in nearby creeks.

Several good fate studies have been published the last two years for benzotriazoles and benzothiazoles. For example, photolysis of three benzotriazoles (1H-benzotriazole, 4-methyl-1Hbenzotriazole, and 5-methyl-1H-benzotriazole) was examined by Weidauer et al.219 Thirty-six TPs were



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identified using LC-HR-MS. However, most were highly reactive and did not appear to be persistent in the environment. Borowska et al. investigated UV photolysis and UV/H2O2 oxidation of benzotriazole and benzothiazole in pure water and in synthetic wastewater. Polychromatic UV could efficiently degrade both compounds, and LC-HR-MS/MS was used to identify 11 TPs. Results using the Microtox assay indicates that these TPs are not toxic. Finally, in another study cited earlier in the DBP section, Nika et al. identified 10 DBPs from the reaction of four benzotriazoles and three benzothiazoles with chlorine.173 Rather than detoxifying the benzotriazoles and benzothiazoles (as with UV), the chlorinated DBPs were calculated to be more toxic to fish and algae than the parent compounds.

New methods for benzotriazoles and benzothiazoles include one by Garcia-Guerra, who used FPSE with UPLC-MS/MS to measure benzotriazole UV stabilizers in seawater.221 Enrichment factors of 25 were achieved, along with LOQs of 3.54-29.9 ng/L. This method was then used to measure seawater from the coast of the Gran Canaria Island in Spain, where levels up to 544.9 ng/L were found. Another new method by Ye et al. used ultrathin polydopamine film-coated gold nanoparticles as a substrate for detection of benzotriazole with Raman spectroscopy.222 Limits of detection were 0.119 µg/L.

ALGAL TOXINS Interest in algal toxins continues to grow, as algal blooms appear to be increasing throughout the world. Nutrient runoff and wastewater discharges are largely responsible. Algal toxins can cause fish kills, shellfish poisoning, livestock and wildlife death, and illness in humans. They can also threaten drinking water supplies, as happened in 2014 in Toledo, Ohio, when a large algal bloom occurred over its source water intake on Lake Erie1. The most common algal toxins are fresh water microcystins, nodularins, anatoxins, cylindrospermopsin, and saxitoxins, and seawater brevetoxins. Microcystins are the most frequently reported, and they have many variants, including a common type containing the amino acids leucine and arginine in their structures. Algal toxins have been listed on all of the U.S. EPA’s Contaminant Candidate Lists, and they are currently listed on the final CCL-4. Australia, Brazil, Canada,


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the Czech Republic, France, Italy, Japan, Korea, New Zealand, Norway, Poland, and Spain have guideline limits for microcystins, anatoxin a, and cylindrospermopsin (levels from 1.0 to 1.5 µg/L).

Preece et al. published an excellent review on microcystin measurements in estuarine and marine waters from around the world.223 In addition, factors affecting harmful algal blooms (HABs) that produce microcystins were discussed, as well as factors that affect microcystins in receiving waters. Human health issues from fish and shellfish are also discussed. Cyanobacteria and cyanotoxins in fish ponds and their effects on fish was the focus of an interesting study by Drobac et al.224 Water samples and fish (carp) from 13 fish ponds in Serbia were sampled during a major algal bloom; they contained saxitoxins, microcystins, and nodularin. The algal toxins caused damage in the liver, kidney, gills, intestines, and muscle of the fish. Bioaccumulation and depuration kinetics of microcystin-LR (MC-LR) in lettuce and arugula were determined in another study by Cordiro-Araujo et al.225 The lettuce and arugula were irrigated with MCLR-spiked water (5 and 10 µg/L) for 7 days, then allowed to depurate in pure water for 7 days. LCMS/MS was used for measurement. MC-LR only accumulated in the lettuce, and after the depuration period, 25% remained in the leafy tissue for lettuce spiked at the higher level (10 µg/L). From half-lives calculated, it was estimated that 29-37 days would be required to completely depurate the toxin from the lettuce. Total daily intake of MC-LR exceeded the World Health Organization (WHO) guideline limit of 0.04 µg/kg. Loftin et al. carried out a nationwide survey of cyanotoxins in 1161 lakes in the U.S., as part of the U.S. EPA National Lakes Assessment.226 Microcystins, saxitoxins, and cylindrospermopsin were found in 4-32% of the samples, with mean concentrations ranging from 0.061 to 3.0 µg/L. Cyanobacteria were present in 98% of the samples. Anatoxin-a and nodularin-R were detected in 15 and 3.7% of the samples, respectively. ELISA methods were used to measure the microcystins, saxitoxins, and



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cylindrospermopsin, and LC-MS/MS was used to measure anatoxin-a and nodularin-R. WHO moderate and high risk thresholds were exceeded in many samples. Parker et al. published an interesting fate study showing that halogen radicals can promote photodegradation of microcystins in estuaries.227 Experiments demonstrated that halide salts from seawater increased the quantum yield for photolysis by a factor of 3-6 and that halogen radicals were responsible. This reactive halogen pathway was found to account for a substantial amount of photodegradation along a freshwater-estuarine transect. Zhang et al. investigated the release and degradation of MC-LR from Microcystis aeruginosa for waters treated with chlorine and permanganate.228

Rapid release of MC-LR was observed, even with

very low chlorine and permanganate doses (0.5 mg/L). When the kinetics were investigated, the rate of release was found to be very fast initially, followed by a decrease over time. The presence of ammonia also played an important role in the release of MC-LR. In a study mentioned earlier in the DBP section, Chu et al. found that THMs, haloacetaldehydes, and haloacetonitriles formed from the reaction of chlorine with MC-LR.176 Potential release in sediments was examined in a study by Kieber et al., who investigated whether sunlight is capable of mobilizing MC-LR from Microcystis-containing sediment into the water column.229 In controlled laboratory reactions, a net photorelease was observed, with levels ranging from 0.4-192 µg/L per gram. Levels increased linearly with increasing time of exposure. The authors suggested that sunlight photolysis may be responsible for as much as 100% of the average stock of MC-LR in a freshwater pond in North Carolina. Lehman et al. demonstrated that a severe drought can cause increased algal blooms and increased release of microcystins.230 The San Francisco Estuary was the subject of investigation, where the severe drought resulted in total microcystin concentrations that exceeded previous dry and wet years by a factor


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of 11 and 65, respectively. During this major algal bloom, extremely high levels of nutrients were present, including a 20-year high for phosphorus. Water temperature was also a major contributor. Detailed degradation pathways for the ozonation of cylindrospermopsin was reported in another study by Yan et al.231 Thirty-six degradation products were identified using UPLC-MS/MS. Reactions mostly involved attack on the C=C bond of the uracil ring, along with reaction of the nitrogen in the tricyclic guanidine group. Interestingly, the TPs did not demonstrate cytotoxicity in a human cell assay, so the authors suggest the use of ozone for removing cylindrospermopsin in water treatment. ELISA and LC-MS/MS methods were compared in another paper by Guo et al. for the measurement of microcystins in drinking water.232 The ELISA method was shown to be highly variable and to have false-positives, suggesting that ELISA results should be interpreted with caution. On the other hand, while LC-MS/MS results agreed with spike concentrations for all microcystins, matrix suppression was sometimes observed. New methods created the last two years include one by Zervou et al., who reported a new SPELC-MS/MS method for simultaneously measuring 17 multi-class algal toxins.233 These include cylindrospermopsin, anatoxin-a, nodularin, 12 microcystins, okadaic acid, and domoic acid. A dual SPE cartridge assembly was used, and limits of detection from 1-10 ng/L were achieved. This method was then used to measure these algal toxins in lakes in Greece, with some reported for the first time. Another new method by Li et al. involved the fabrication of a colorimetric aptamer-based sensor for MC-LR detection in water.234 Limits of detection of 0.37 nM were achieved, with a high selectivity of the aptamer for MC-LR.

MICROPLASTICS Microplastics (MPs) have received a great deal of attention, especially the last few years, due to growing contamination in oceans, lakes, and rivers worldwide. MPs, defined as 5 mm or less in size, can



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be present due to degradation of larger plastic materials that enter the environment or from their direct use in products. A new law, the Microbead-Free Waters Act, now bans all microbeads in cosmetic products and toothpaste that is manufactured or distributed in the United States (www.congress.gov/bill/114thcongress/house-bill/1321/text). Of all the types of MPs, fibers are the most predominant, and recent studies have shown that they are easily released from clothes during washing (particularly from polyester fleece clothing).

Accumulation of MPs has been occurring for four decades, with largest accumulations occurring in the ocean gyres, where “plastic islands” are formed. Ingestion of MPs by animals, including seabirds, seals, sea lions, dolphins, whales, marine reptiles, and zooplankton is well documented, and concerns arise from the potential for the MPs (and other plastics) to cause obstruction in the digestive tract and internal organs, as well as serve as vectors for chemical contaminants sorbed on their surfaces.

Wright and Kelly wrote an excellent review evaluating potential human health impacts of MPs, and recommended areas for future research.235 Dietary pathways, such as known sources in seafood, honey, sugar, salt, and beer, are discussed. One study the authors cite discusses measurements of MPs in beer, where 109 MP fragments were found in 1 L of beer. Suggested sources were atmospheric deposition, contamination of water, and materials used in the beer-making process. Another potential pathway of human exposure was inhalation, which could involve airborne particles from sea aerosols, wind-driven transport of MPs from dry soils that had sludge applied as fertilizer, and atmospheric fallout. Potential toxicological pathways are also discussed, including inflammation, genotoxicity, oxidative stress, apoptosis, and necrosis. Seabirds were the focus of another study by Herzke et al., who found that MPs acted more as a passive sampler, reflecting persistent organic pollutant profiles in the gastrointestinal tract, rather than acting as a vector for bioaccumulation of pollutants.236 Specifically, concentrations of polychlorinated


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biphenyls (PCBs), dichloro-diphenyl-trichloroethane (DDT), and PBDEs in the MPs did not differ significantly from what was found in the birds’ liver and muscle tissue. Koelmans et al. also had a similar finding.237 In their evaluation of data from the scientific literature, they concluded that the bioaccumulation from natural prey was far greater than contaminants consumed from MPs. With fibers being a major category of MPs, a new study by Hartline et al. on microfibers released from washing polyester jackets was particularly appropriate.238 Detergent-free washings were done in front-loading and top-loading washing machines, and microfibers were recovered using filters. Levels of microfibers ranged from 0-2 g per wash cycle, with top-loading washing machines leaching approximately 7x the fibers as front-loading machines. A 24-hr continuous wash cycle was also examined, which resulted in further release of fibers. It was estimated a population of 100,000 people would contribute approximately 1.02 kg of fibers each day at a WWTP and that washing synthetic jackets could account for 71-428% of the fibers observed in WWTP influents. While MPs are well documented in skin care products, Hernandez et al. investigated whether nanoplastics could also be present in these products.239 In this study, three facial scrubs were analyzed using sequential filtration, scanning electron microscopy, X-ray photoelectron spectroscopy, and FTinfrared (IR) spectroscopy. The presence of nanoparticles was confirmed, and they were identified specifically as polyethylene nanoplastics, ranging from 24-52 nm in size. Baldwin et al. conducted a massive study investigating plastics in 29 tributaries of the Great Lakes, which had different land covers, wastewater effluent contributions, population densities, and hydrologic conditions.240 MPs made up 98% of all plastic particles found, and they were found in all samples, up to 32 particles per m3. Fibers were the predominant MP found, but curiously, their levels did not correlate with WWTP effluents. It was suggested that atmospheric deposition may play a major role in their presence in streams. Runoff from sludge applied in agriculture did not appear to be a major source.



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Another study by Kooi et al. reported the impacts of biofouling on vertical transport (depth) of MPs in the ocean.241 Their model predicted that microplastic particles can either float, sink to the bottom, or oscillate vertically, depending on the particle’s size and density. Interestingly, a relatively low abundance was predicted at the ocean surface, with highest levels predicted at intermediate depths, and very few reaching the ocean floor. The model demonstrated that particles < 10 µm sink so slowly that they can be dispersed anywhere in the water column. Biofilms were also a topic in a paper by Debroas et al., who discovered that most bacteria sorbed to MPs is of non-marine origin.242 The bacteria were suggested to be “hitchhikers” and included Streptomycetales and cyanobacteria as dominant species. Highest abundances to-date were reported by van der Hal et al., who measured MPs in the Mediterranean Sea off the coast of Israel.243 Of the 17 sites investigated, one contained 324 particles per m3, which was 1-2 orders of magnitude higher than previously reported in other parts of the world. Wastewater treatment is a continued focus for MP research. Mintenig et al. used attenuated total reflection FT-IR spectroscopy and focal plane array-based transmission micro-FT-IR imaging to investigate sludge as a sink for MPs in 12 WWTPs in Germany.244 MPs were made up of 14 different polymer types, with polyethylene predominant. Synthetic fibers were more prevalent than other MPs, and sludge was found to be a major sink. One plant sampled utilized tertiary treatment and resulted in a 97% reduction in MP discharge. In another WWTP study by Murphy et al., MPs were found to be removed predominantly in an early grease removal stage, rather than in sludge.245 This study investigated all stages of wastewater treatment in a city of 650,000 people (Glasgow, Scotland). The wastewater influent contained an average of 15.70 MPs per liter, while the effluent contained only 0.25 MPs per liter. Finally, a new method was reported by Majewsky et al., to determine polyethylene and polypropylene MPs in the environment.246 This method utilized thermogravimetry coupled to differential scanning calorimetry (TGA-DSC), which could clearly identify these types of plastics in wastewater and


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directly quantify polyethylene and polypropylene as total weight concentration. Calibration curves ranged from 240-1540 mg/m3.

IONIC LIQUIDS Ionic liquids (ILs) are organic salts with low melting points (100 ton/year) and is covered in Europe under the REACH regulation. The other chloro-bromo-MSAs tentatively identified were believed to be formed by chlorine disinfection. These halo-MSAs were widespread in the samples collected, suggesting that this is an important new class of emerging contaminant to watch for.

BIOGRAPHIES Susan D. Richardson is the Arthur S. Williams Professor of Chemistry at the University of South Carolina, and was formerly a research chemist at the U.S. Environmental Protection Agency for many years. She received her B.S. degree in Chemistry and Mathematics from Georgia College in 1984 and her Ph.D. degree in Chemistry from Emory University in 1989. Susan’s research has focused on the identification, characterization, and quantification of new toxicologically important disinfection byproducts (DBPs), with special emphasis on alternative disinfectants and polar byproducts. She is particularly interested in promoting new health effects research so that the risks of DBPs can be determined and minimized. She is also interested in studying and minimizing emerging contaminants in water reuse applications. Among her achievements, Susan was recently recognized as an ACS Fellow in 2016.

Thomas A. Ternes graduated with an undergraduate degree in Chemistry from the University of Mainz (Germany) in 1989. In 1993, he completed his Ph.D. at the University of Mainz in Analytical Chemistry. In January 2001, he completed his habilitation and became an official lecturer at the University of Mainz. Since 1995, his research has focused on the analysis 79 ACS Paragon Plus Environment


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and fate of organic pollutants, such as pharmaceuticals and personal care products (PPCPs), in various kinds of environmental matrices. Dr. Ternes was the coordinator of the Pharmacluster project POSEIDON (www.eu-poseidon.com) dealing with the removal of PPCPs in wastewater and drinking water treatment, soil aquifer treatment, and environmental risk assessment. Since May 2003, he has been at the Federal Institute of Hydrology (BfG) in Koblenz, Germany, where he is the head of the water chemistry department, and he is a lecturer at University of KoblenzLandau. Recently, he finished together with Adriano Joss from Eawag the prestigious advanced ERC grant ATHENE which was mainly focused on the biological degradation of emerging contaminants.

ACKNOWLEDGMENTS SR would like to thank Jody Shoemaker of the U.S. EPA for information on new EPA methods (she and others at the U.S. EPA have created many great methods that are used in the United States and throughout the world) and Stig Regli of the U.S. EPA’s Office of Water for helpful information on new regulations. I would also like to thank Emma Schymanski (formerly of EAWAG, now at the University of Luxembourg), Juliane Hollender (EAWAG), and David Wishart (University of Alberta) for the amazing new workflow/database information presented at the Monte Verita Conference in 2016 that was included in this review. And, as always, I would like to thank David Humphries (retired from Alberta Innovates) for daily inspiration, as well as the encouragement of University of South Carolina researchers and scholars who have created a very welcoming and scientifically stimulating atmosphere for this next phase of my career and my fabulous new graduate students, undergrads, and postdocs.


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Table 1. List of Acronyms AFFFs aqueous film forming foams AOPs advanced oxidation processes APCI atmospheric pressure chemical ionization APPI atmospheric pressure photoionization BAC biological activated carbon BP benzophenone CCL Contaminant Candidate List DAD diode array detection DBPs disinfection byproducts DOC dissolved organic carbon DOM dissolved organic matter E2 17β-estradiol ECs emerging contaminants EDA effects-directed analysis EE2 17α-ethinylestradiol ELISA enzyme-linked immunosorbent assay EPA Environmental Protection Agency ESI electrospray ionization EU European Union FPSE fabric phase sorptive extraction FT Fourier-transform FTOHs fluorinated telomer alcohols FTP fluorotelomer-based polymer GAC granular activated carbon GC gas chromatography HAAs haloacetic acids HALs haloaldehydes HAMs haloacetamides HANs haloacetonitriles HF hydraulic fracturing HILIC hydrophilic interaction liquid chromatography HR high resolution IC ion chromatography ICP inductively coupled plasma ICR ion cyclotron resonance ILs ionic liquids IM ion mobility IR infrared LC liquid chromatography LLE liquid-liquid extraction



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LOQ LTQ 4-MBC MC-LR MIMS MPs MS NAs N-DBPs NDMA NMR NMs NOM NPs OC PBDEs PFAA PFASs PFBA PFCAs PFHxA PFNA PFOA PFOS PFSAs PIE POPS QSAR QTOF REACH RO SPE SPME TBBPA TCIPP TDCIPP THMs TOC TOF TOX TPs UCMR UPLC WHO WWTP

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limit of quantification Linear ion trap 4-methylbenzylidene camphor microcystin LR membrane introduction mass spectrometry microplastics mass spectrometry naphthenic acids nitrogen-containing DBPs N-nitrosodimethylamine nuclear magnetic resonance nanomaterials natural organic matter nanoparticles octocrylene polybrominated diphenyl ethers perfluoroalkyl acid per- and polyfluoroalkyl substances perfluorobutanoic acid perfluorocarboxylic acids perfluorohexanoic acid perfluorononanoic acid perfluorooctanoic acid perfluorooctane sulfonate perfluorosulfonic acids precursor ion exclusion persistent organic pollutants quantitative structure-activity relationship quadrupole-time-of-flight Registration, Evaluation, and Authorization of Chemicals reverse osmosis solid phase extraction solid phase microextraction tetrabromobisphenol A tris(chloropropyl)phosphate tris(1,3-dichloroiosopropyl)phosphate trihalomethanes total organic carbon time-of-flight total organic halogen transformation products Unregulated Contaminant Monitoring Rule ultraperformance liquid chromatography World Health Organization wastewater treatment plant


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Table 2. Useful Websites

website www.epa.gov http://www2.epa.gov/dwanalyticalmethods

comments U.S. EPA’s website U.S. EPA approved methods for drinking water compliance monitoring http://www2.epa.gov/water-research/epa-drinking-water-research-methods drinking water methods developed by U.S. EPA’s Office of Water www.standardmethods.org link to Standard Methods Online http://ec.europa.eu/environment/chemicals/reach/reach_en.htm REACH website http://chm.pops.int/TheConvention/ThePOPs/ListingofPOPs/tabid/2509 Stockholm Convention persistent organic pollutants (POPs) https://metlin.scripps.edu METLIN database http://cfmid.wishartlab.com CFM-ID database and tool http://www.hmdb.ca Human Metabolome Database (HMDB) https://omictools.com/metfrag-tool Metfrag www.mzcloud.org mzCloud https://comptox.epa.gov/dashboard/chemical_lists/STOFFIDENT Stoff-Ident database http://www.massbank.jp MassBank http://mona.fiehnlab.ucdavis.edu MassBank of North America https://massbank.eu/MassBank European MassBank http://gnps.ucsd.edu/ProteoSAFe/libraries.jsp Global Natural Products Social Molecular Networking (GNPS)



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Table 3. Unregulated Contaminant Monitoring Rule-4 (UCMR-4) Contaminants and Approved Methods Contaminant

EPA Method

Cyanotoxins Total microcystin

EPA 546

Microcystin-LA

EPA 544

Microcystin-LF

EPA 544

Microcystin-LR

EPA 544

Microcystin-LY

EPA 544

Microcystin-RR

EPA 544

Microcystin-YR

EPA 544

Nodularin

EPA 544

Anatoxin-a

EPA 545

Cylindrospermopsin

EPA 545

Metals EPA 200.8, ASTM D5673-10, SM 3125 Germanium Manganese

EPA 200.8, ASTM D5673-10, SM 3125

Pesticides and pesticide manufacturing product alpha-Hexachlorocyclohexane

EPA 525.3

Chlorpyrifos

EPA 525.3

Dimethipin

EPA 525.3

Ethoprop

EPA 525.3

Oxyfluorfen

EPA 525.3

Profenofos

EPA 525.3

Tebuconazole Total permethrin (cis- & trans-)

EPA 525.3 EPA 525.3

Tribufos

EPA 525.3


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Brominated haloacetic acid groups HAA5

EPA 552.3 or EPA 557

HAA6Br


HAA9


Alcohols 1-Butanol

EPA 541

2-Methoxyethanol

EPA 541

2-Propen-1-ol

EPA 541

Other semivolatile chemicals Butylated hydroxyanisole

EPA 530

o-Toluidine

EPA 530

Quinoline

EPA 530

Indicators Total organic carbon (TOC)

Bromide

SM 5310 B, SM 5310 C, SM 5310 D (21st edition), or SM 5310 B-00, SM 5310 C-00, SM 5310 D-00 (SM Online), EPA Method 415.3 (Rev. 1.1 or 1.2) EPA Methods 300.0 (Rev. 2.1), 300.1 (Rev. 1.0), 317.0 (Rev. 2.0), 326.0 (Rev. 1.0) or ASTM D 6581-12



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Table 4. Contaminant Candidate List-4 (CCL-4) Chemical Contaminants 1,1-Dichloroethane 1,1,1,2-Tetrachloroethane 1,2,3-Trichloropropane 1,3-Butadiene 1,4-Dioxane 17alpha-estradiol 1-Butanol 2-Methoxyethanol 2-Propen-1-ol 3-Hydroxycarbofuran 4,4'-Methylenedianiline Acephate Acetaldehyde Acetamide Acetochlor Acetochlor ethanesulfonic acid (ESA) Acetochlor oxanilic acid (OA) Acrolein Alachlor ethanesulfonic acid (ESA) Alachlor oxanilic acid (OA) alpha-Hexachlorocyclohexane Aniline Bensulide Benzyl chloride Butylated hydroxyanisole Captan Chlorate Chloromethane (Methyl chloride) Clethodim Cobalt Cumene hydroperoxide Cyanotoxins Dicrotophos Dimethipin Diuron Equilenin Equilin Erythromycin


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Estradiol (17-beta estradiol) Estriol Estrone Ethinyl estradiol (17-alpha ethynyl estradiol) Ethoprop Ethylene glycol Ethylene oxide Ethylene thiourea Formaldehyde Germanium HCFC-22 Halon 1011 (bromochloromethane) Hexane Hydrazine Manganese Mestranol Methamidophos Methanol Methyl bromide (bromomethane) Methyl tert-butyl ether (MTBE) Metolachlor Metolachlor ethanesulfonic acid (ESA) Metolachlor oxanilic acid (OA) Molybdenum Nitrobenzene Nitroglycerin N-Methyl-2-pyrrolidone N-Nitrosodiethylamine (NDEA) N-Nitrosodimethylamine (NDMA) N-Nitroso-di-n-propylamine (NDPA) N-Nitrosodiphenylamine N-Nitrosopyrrolidine (NPYR) Nonylphenol Norethindrone (19-Norethisterone) n-Propylbenzene o-Toluidine Oxirane, methyl Oxydemeton-methyl Oxyfluorfen Perfluorooctanesulfonic acid (PFOS)



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Perfluorooctanoic acid (PFOA) Permethrin Profenofos Quinoline RDX (Hexahydro-1,3,5-trinitro-1,3,5-triazine) sec-Butylbenzene Tebuconazole Tebufenozide Tellurium Thiodicarb Thiophanate-methyl Toluene diisocyanate Tribufos Triethylamine Triphenyltin hydroxide (TPTH) Urethane Vanadium Vinclozolin Ziram Microbial Contaminants Adenovirus Caliciviruses Campylobacter jejuni Enterovirus Escherichia coli (0157) Helicobacter pylori Hepatitis A virus Legionella pneumophila Mycobacterium avium Naegleria fowleri Salmonella enterica Shigella sonnei


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