Water Analysis: Emerging Contaminants and ... - ACS Publications

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Water Analysis: Emerging Contaminants and Current Issues Susan D. Richardson*,† and Thomas A. Ternes‡ †

Department of Chemistry and Biochemistry, University of South Carolina, Columbia, South Carolina 29208, United States Federal Institute of Hydrology, Koblenz, D-56068 Germany

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CONTENTS

Background Major Analysis Trends Sampling and Extraction Trends Chromatography Trends Emerging Contaminant Trends New Emerging Contaminant General Reviews New Regulations/Regulatory Methods New Proposed Regulation for Perchlorate in U.S. Drinking Water Group Approach The Third Unregulated Contaminant Monitoring Rule (UCMR-3) New Regulatory Methods for Drinking Water EPA Method 525.3: Semivolatile Organic Chemicals EPA Method 540: Selected Organic Chemicals EPA Method 524.4: Purgeable Organic Chemicals EPA Method 1615: Enterovirus and Norovirus EPA Method 1623.1: Cryptosporidium and Giardia Sucralose and Other Artificial Sweeteners Nanomaterials PFOA, PFOS, and Other Perfluorinated Compounds Pharmaceuticals and Hormones Environmental Impacts of Pharmaceuticals Biological Transformation Products Elimination/Reaction During Oxidative Water Treatment Photodegradation Opiates and Other Drugs of Abuse Antidiabetic Drugs Cytostatic Pharmaceuticals Multiresidue Methods Sample Stability and Impact of Silanized Glassware Enantiomers Microextraction Method Solid Contact Potentiometric Sensors for Pharmaceuticals Bioassays Drinking Water and Swimming Pool Disinfection Byproducts Drinking Water DBPs Books and Reviews Combining Chemistry with Toxicology and Epidemiology Discovery and Measurement of New DBPs N-DBP Occurrence © 2014 American Chemical Society

Iodo-DBPs New Methods New Swimming Pool Research Formation Studies DBPs of Pollutants Sunscreens/UV Filters Brominated and Emerging Flame Retardants Benzotriazoles and Benzothiazoles Siloxanes Naphthenic Acids Musks Pesticide Transformation Products Antimony Perchlorate Algal Toxins Microorganisms Contaminants on the Horizon: Ionic Liquids and Prions Prions Author Information Corresponding Author Notes Biographies Acknowledgments References

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BACKGROUND This biennial review covers developments in water analysis for emerging environmental contaminants over the period of 2011−2013. Analytical Chemistry’s policy is to limit reviews to a maximum of 250 significant references and to mainly focus on new trends. As a result, only a small fraction of the quality research publications could be discussed. I am excited to again have Thomas Ternes join me this year to cover the section on Pharmaceuticals and Hormones. Thomas coauthored the previous 2011 Review on Water Analysis,1 and as before, this Review is so much better with his contribution. We welcome any comments you have on this Review. 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, and Table 2 provides some useful Web sites.

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Received: February 5, 2014 Published: February 6, 2014 2813

dx.doi.org/10.1021/ac500508t | Anal. Chem. 2014, 86, 2813−2848

Analytical Chemistry

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as a complementary analytical technique to confirm tentative structures proposed by LC/HRMS and LC/MS/MS. Because NMR is not as sensitive as MS, preparative LC coupled to a fraction collector is often used to collect enough material. Examples in this Review include the identification of TPs from pharmaceuticals and ultraviolet (UV) filters. Atmospheric pressure photoionization (APPI) (with LC/MS) has also experienced significant growth in use because it provides improved ionization for more nonpolar compounds, such as nanomaterials, perfluorinated compounds (PFCs), brominated flame retardants, and musks discussed in this Review. Also the number of applications of large volume, direct aqueous injection-LC/MS in environmental analysis is increasing. With this technique, complex water samples can be injected directly onto an LC column, saving the time and cost of using solid phase extraction (SPE) and other preconcentration steps. Examples in this Review include the analysis of iodo-acid disinfection byproducts (DBPs) and pesticide metabolites. Sampling and Extraction Trends. SPE remains the most popular means of extraction and concentration, and new SPE sorbents are available, including ion exchange resins such as Oasis MCX and hypercrosslinked polymer resin (HXLPP), that are being used to capture a broader range of analytes within a single extraction. Solventless extraction techniques, such as solid phase microextraction (SPME), also continue to be used in many applications, with examples involving DBP and brominated flame retardant extraction in this Review. One new extraction trend is the application of ionic liquids for extraction. Ionic liquids are organic salts with a low melting point (1100 review articles on nanomaterials. There is even a monthly journal called ACS Nano (created in 2008). Special focus issues of journals (including two in Environmental Science 2819



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Pycke et al. published a review on strategies for identifying oxidation products of fullerene, together with the parent fullerenes in aquatic and biological samples.40 Most methods developed for water-insoluble nC60 will not work for watersoluble polyoxygenated and polyhydroxylated fullerenes. Proposed strategies for measuring the parent fullerenes with the transformation products include defunctionalization (converting the modified fullerenes back to nC60), functionalization (hydroxylating the parent fullerenes), differential extraction, and filtration. Environmental fate studies continue to be published for nanomaterials. In a particularly interesting one, Alpatova et al. reports the first evidence that nC60 can form chlorinated disinfection byproducts.41 In ozonated-chlorinated aqueous samples, the presence of Cl atoms covalently bound to carbon atoms was confirmed by X-ray photoelectron spectroscopy. In ozonated samples, the morphology of the nC60 spheres changed to irregularly shaped aggregates, and the concentration of atomic carbon substantially decreased in these reaction products. A few studies followed the fate of TiO2 nanoparticles. NanoTiO2 is the highest produced of all nanomaterials and is used in textiles as a UV filter and as a pigment. Windler et al. examined the release of nano-TiO2 from textiles during washing.42 Five textiles released low amounts of TiO2 (0.01−0.06 wt %) in one wash cycle, but another with antimicrobial functionality released much higher amounts (5 mg/L, corresponding to 3.4 wt %). Particle sizes released varied from 60 to 350 nm. Westerhoff et al. examined the occurrence and removal of TiO2 in wastewater treatment.43 Sewage influents ranged from 181 to 1233 μg/L, and WWTPs removed >96% of the influent Ti, with effluents from 10 different plants containing 91% were achieved in environmental waters. Gold NPs were subsequently detected in wastewater influent (but not effluent) at 17 ng/L from a wastewater plant in Munich. The source of the nanogold is not yet clear. Baalousha and Lead investigated the variability in the measured size of nanomaterials by commonly applied techniques and provide an explanation for the varying numbers reported in the literature.36 The authors proposed a validated sample preparation procedure for size evaluation by AFM, along with a quantitative explanation of the variability of the measured sizes by FIFFF, AFM, and DLS. New methods have also been developed for carbon-based nanomaterials. Chen and Ding reported a new LC/APPI-MS/ MS method using ultrasound-assisted dispersive liquid−liquid microextraction for measuring fullerenes in aqueous samples.37 Detection limits of 8, 60, and 150 ng/L were achieved for nC70, nC 60 , and [6,6]-phenyl-C 61 butyric acid methyl ester, respectively. Using this method, nC70 was detected in municipal influent (37 ng/L), nC70 and nC60 were detected in industrial effluent (25 and 130 ng/L, respectively), and nC60 was detected in surface waters (98 ng/L) in Taiwan. Carbon nanotubes were the focus of other new methods including the most sensitive and selective technique available to-date for single-walled nanotubes (SWNTs) developed by Schierz et al.38 This method utilized near-infrared fluorescence spectroscopy to characterize and quantify SWNTs in the aquatic environment. Detection limits of 1.0 μg/L were achieved for water and 62 ng/g for sediment. Plata et al. used thermogravimetry-MS to isolate and quantify SWNTs in environmental samples.39 This method showed promise for distinguishing between incidental (e.g., soot) and engineered (e.g., SWNT) nanoparticles, which was not possible with other techniques.

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PFOA, PFOS, AND OTHER PERFLUORINATED COMPOUNDS Perfluorinated compounds (PFCs) have been manufactured for more than 50 years and have been used to make stain repellents (such as polytetrafluoroethylene and Teflon) that are widely applied to fabrics and carpets. They are also used in the manufacture of paints, adhesives, waxes, polishes, metals, electronics, fire-fighting foams, and caulks, as well as greaseproof coatings for food packaging (e.g., microwave popcorn bags, French fry boxes, hamburger wrappers, etc.). PFCs are unusual chemically, in that they are both hydrophobic (repel 2820



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fluoro-acids via metabolism and biodegradation. Recent studies support this hypothesis. In addition, there is evidence that PFOA itself is volatile (can sublime in its solid form) and that it can also partition from water to air.2 Even the ammonium salt form of PFOA can sublime into air. In addition to providing another unexpected mechanism for atmospheric transport for PFOA, these results also suggest that extra care should be taken in manufacturing facilities that make or use PFOA in formulations in order to minimize workplace exposure from inhalation. PFOS, PFOA, and other PFCs are included in the National Health and Nutrition Examination Survey (NHANES) conducted by the Centers for Disease Control and Prevention (CDC) to provide a better assessment of the distribution of these chemicals in adults and children in the United States (www.cdc.gov/nchs/nhanes.htm). This survey is carried out on a continual basis, with blood and urine collected from thousands of participants in the United States. The most recent report (September 2013) that includes population serum levels of PFCs can be found at http://www.cdc.gov/ exposurereport. The National Toxicology Program is also studying PFOA and several other perfluorocarboxylic acids (PFCAs) and perfluorosulfonates (PFSAs) to better understand their toxicity and persistence in human blood (http:// www.niehs.nih.gov/health/materials/perflourinated_ chemicals_508.pdf). Unlike other contaminants that accumulate in humans (e.g., dioxins, polychlorinated biphenyls), PFCs do not accumulate in fatty tissues but bind to serum proteins and accumulate instead in blood. As such, they have unique profiles of distribution in the body, owing to their unique chemical properties. While PFOS and PFOA were the first fluorinated surfactants to receive considerable attention, significant research has been carried out on PFCAs and PFSAs with shorter and longer chain lengths as well as a recent explosion of research in precursors and newer PFC classes that have been discovered, including perfluoroalkyl-sulfonamides, -ester phosphates, -phosphonates, -ethoxylates, -acrylates, -amino acids, -sulfonamide phosphates, -thioacids, and thioamidosulfonates as well as a new cyclic PFC that was just reported. In addition, there is increased focus on shorter-chain forms, e.g., perfluorobutanoic acid (PFBA) and perfluorobutane sulfonate (PFBS), as manufacturers are shifting to lower molecular weight PFCs that are not bioaccumulative. Also, there are significant, new efforts on PFCs recently identified in aqueous film forming foam (AFFF) formulations used to extinguish fires. Drinking water fate and exposure is an increasingly hot topic for PFCs, with many new papers published over the last 2 years. Post et al. published a nice review on PFOA as an emerging drinking water contaminant, covering fate and transport relevant to drinking water, human exposure and serum levels, toxicokinetics, and potential human health effects.46 The literature suggests that continued human exposure to even low concentrations of PFOA in drinking water can result in elevated body burdens that may increase the risk of health effects. The authors highlight infants as a potentially sensitive subpopulation for PFOA’s developmental effects, with infant exposure higher than adults using the same drinking water source. Drinking water was also the focus of an extensive PFC national screening study in France.47 In this study conducted by Boiteux et al., 331 source water and 110 finished drinking water samples were collected from several regions in France, representing 20% of the national water supply. Of the 10

water) and lipophobic (repel lipids/grease), and they contain one of the strongest chemical bonds (C−F) known. Because of these properties, they are highly stable in the environment (and in biological samples) and have unique profiles of distribution in the body. During 2000−2002, an estimated 5 million kg/yr was produced worldwide, with 40% of this in North America. Two of these PFCs, perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA), have received the most attention. PFOS was once used to make the popular Scotchgard fabric and carpet protector, and since 2002, it is no longer manufactured in the U.S., due to concerns about widespread global distribution in the blood of the general population and in wildlife, including remote locations in the Arctic and North Pacific Oceans. Like PFOS, PFOA is ubiquitous at low levels in humans, even in those living far from any obvious sources.1 In January 2005, the U.S. EPA issued a draft risk assessment of the potential human health effects associated with exposure to PFOA (www.epa.gov/oppt/pfoa/pubs/pfoarisk.html), and in January 2006, the U.S. EPA invited PFC manufacturers to participate in a global stewardship program on PFOA and related chemicals (www.epa.gov/oppt/pfoa/pubs/ stewardship). Participating companies agreed to commit to reducing PFOA from emissions and product content by 95% by 2010 and to work toward eliminating PFOA in emissions and products by 2015. The U.S. EPA has listed PFOA and PFOS on the CCL-3, a priority list for consideration for future regulation in drinking water (http://water.epa.gov/scitech/drinkingwater/ dws/ccl/ccl3.cfm). Six PFCs are also now included on the U.S. EPA’s UCMR-3: PFOA, PFOS, perfluorononanoic acid (PFNA), perfluorohexane sulfonic acid (PFHxA), perfluoroheptanoic acid (PFHpA), and perfluorobutane sulfonic acid (PFBS) (http://water.epa.gov/lawsregs/rulesregs/sdwa/ucmr/ ucmr3). As mentioned earlier in the New Regulations/ Regulatory Methods section, the UCMR requires drinking water utilities in the United States to measure these contaminants in finished drinking water. In Europe, the European Food Safety Authority has established tolerable daily intakes for PFOA and PFOS (www.efsa.europa.eu), and there are new restrictions on the use of PFOS as part of the European Union’s REACH program (http://ec.europa.eu/enterprise/sectors/chemicals/files/reach/ restr_inventory_list_pfos_en.pdf). The European Water Framework Directive (WFD) commits EU Member States to attain a good chemical status of inland water bodies (including marine waters up to one nautical mile from shore). Annex X of the WFD originally contained 33 priority substances/groups. Directive 2013/39/EU of August 12, 2013 added 12 new substances, including PFOS, to Annex X of the Water Framework Directive 2000/60/EC. The environmental quality standards (EQS) for annual average of monthly measurements (AA-EQS) for PFOS are reported for inland waters to be 0.65 ng/L and for coastal waters to be 0.13 ng/L. Potential health concerns include cancer, reproductive and developmental effects, immunotoxicity, ulcerative colitis, and bioaccumulation. Toxicity, bioaccumulation, and human epidemiologic studies continue to be conducted. Research questions include understanding the sources of PFOA and other PFCs, their environmental fate and transport, pathways for human exposure and uptake, and potential health effects. It is hypothesized that the widespread occurrence of PFOA and other fluoro-acids is partly due to the atmospheric or oceanic transport of the more volatile fluorinated telomer alcohols (FTOHs) and subsequent transformation into PFOA and other 2821



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suppressed baseline noise, and gave better detector response. Using this method, drinking water from six European countries was sampled and perfluorooctyl phosphonate (PFOPA) was found at 0.095 ng/L in two samples from Amsterdam. The highest PFC levels found were for PFBS (18.8 ng/L), PFOA (8.6 ng/L), and PFOS (8.8 ng/L). As mentioned earlier, a lot of new research is aimed at PFCs recently identified in aqueous film forming foam (AFFF) formulations. A great example of this includes a new analytical method and occurrence study by Backe et al. to measure zwitterionic, cationic, and anionic PFCs in groundwater and AFFF formulations.51 Large-volume injection (LVI)-LC/MS/ MS was used to measure 26 newly identified AFFF PFCs and 21 legacy PFCs (PFCAs, PFSAs, and fluorotelomer sulfonates) in groundwater from five U.S. military bases and in 12 AFFF formulations. Method detection limits ranged from 0.71 to 67 ng/L. Results revealed the presence of 8 of the 26 newly identified PFCs at levels up to 6900 ng/L. These included the finding of fluorotelomer thioamidosulfonates and the newly identified 3M classes. However, legacy PFCs, including PFOS and PFOA, were found at even higher concentrations, far exceeding the U.S. EPA provisional health advisory levels. Interestingly, profiles of PFCs detected in groundwater differed from those in the AFFF formulations, which may indicate some transformation. Their presence in groundwater results from the use of AFFFs in fire-training activities at these military bases. Several extensive, multicountry/multiocean occurrence studies have also been recently reported. For example, Benskin et al. measured the spatial distribution of C4, C6, and C8 PFSAs, C6− C14 PFCAs, and perfluorooctane sulfonamide in the Atlantic and Arctic Oceans, including previously unstudied coastal waters of North and South America and the Canadian Arctic Archipelago.52 PFOA and PFOS were the predominant PFCs detected in both oceans. In the Northwest Atlantic Ocean, total PFC levels reached a high of 5800 pg/L near Rhode Island, and in the Northeast Atlantic Ocean, total PFC levels reached a high of 980 ng/L near the Canary Islands. Southern Atlantic Ocean levels were generally much lower, except near Rio de la Plata (Argentina/Uruguay), where total levels reached 540 pg/ L. Canadian Arctic Ocean levels reached 250 pg/L. Another extensive occurrence study by Cai et al. followed PFCs from the North Pacific Ocean to the Arctic Ocean.53 Seawater, sea ice core, and snow samples were collected along the eastern coast of Asia and the western and northern coast of Alaska. A total of 14 different PFCs were measured. Average concentrations of total PFCs in surface waters were 560 pg/L for the Northwest Pacific Ocean, 500 pg/L for the Arctic Ocean, and 340 pg/L for the Bering Sea. PFCAs were the dominant PFC class detected, and spatial patterns varied across the regions sampled. The presence of PFCs in snow and ice cores indicated atmospheric deposition of PFCs in the Arctic, and elevated levels of PFCs in the Arctic Ocean indicated that ice melting had an impact on PFC levels and distribution. Interestingly, as with other recent occurrence studies, shorterchain PFCAs, PFBA and PFPeA, were the dominant PFCs in the Arctic Ocean. This study also provided the first detections of several PFCs in the Pacific Ocean (e.g., C4 and C10 PFSA; C4−C7, C10, C11, and C14 PFCAs; perfluorooctane sulfonamide (FOSA); and N-methylperfluorooctane sulfonamidoethanol (MeFOSE)). The Great Lakes also continue to be an intense area of sampling for PFCs. In a recent paper from De Silva et al., the first detection of a cyclic PFC, perfluoroethylcyclohexanesulfo-

PFCs measured using LC/MS/MS, PFOS, perfluorohexane sulfonate (PFHxS), PFOA, and perfluorohexanoic acid (PFHxA) predominated in the source waters (detected in 27%, 13%, 11%, and 7% of the samples, respectively). In finished drinking water, short-chain PFCAs predominated, suggesting a relative effectiveness of certain water treatments in removing PFSAs but also the potential for degradation of PFCA precursors by water treatment processes. A particularly interesting discovery was that 8 of these drinking water treatment plants actually had higher levels of some PFCs, PFBA, perfluoropentanoic acid (PFPeA), PFHxA, and perfluoroheptanoic acid (PFHpA), in the finished water vs the raw source waters. Normally, levels would be expected to be lower in the finished water vs the source water, due to some partial removal, dilution, or degradation. In total, 7 of these 8 plants used activated carbon to treat raw water and results suggest release of PFCs from saturated activated carbon or degradation of precursors during the treatment process. PFHxA was found at the highest levels in finished drinking water, up to 125 ng/L. And as expected, areas with higher population densities showed higher levels of PFCs in their finished drinking water. Another interesting extensive drinking water study was conducted by Llorca et al., who followed 21 PFCs along the whole water cycle (wastewater, river water, tap water, and mineral bottled water) in 32 cities in Germany and Spain.48 A new online LC/MS/MS method was used for their measurement, with 0.83−10 ng/L detection limits. Of the PFCs, PFCAs were found most often in drinking water, with 54% of the tap water samples containing PFBA at levels ranging from 2.4 to 27 ng/L. PFHpA, PFOA, and PFOS ranged from 0.23 to 53 ng/L, 0.16−35 ng/L, and 0.04−258 ng/L, respectively. Overall, PFC levels were higher in surface waters and drinking water from Spain (vs Germany), but German mineral waters sampled contained PFHxA, PFHpA, and PFOS, which were not detected in mineral waters from Spain. It was hypothesized that these PFCs in the mineral water could have leached from the plastic in the bottles. While more than 50% of the tap water samples contained PFCs, only 6 had concentrations of PFOS that exceeded the U.S. EPA’s Provisional Health Advisory level of 200 ng/L. The fate of PFOA and PFOS during different stages of drinking water treatment was the focus of a new study by Flores et al.49 A drinking water treatment plant had just been upgraded from conventional treatment to include advanced treatment processes (ultrafiltration and reverse osmosis) in a parallel treatment line. Results showed that neither preoxidation (with chlorine or ClO2), sand filtration, nor ozonation removed these PFCs. Granular activated carbon was able to remove 64 and 45% of PFOS and PFOA, respectively, in drinking water treatment, and reverse osmosis was the most effective, removing ≥99% of both compounds. With the advanced treatment, levels of PFOA and PFOS were significantly reduced from levels measured in previous years. In a new study from Sweden, Ullah et al. reported the first perfluoroalkyl phosphonate (PFPA) detection in drinking water.50 PFPAs, which contain two acidic protons in their structures, are a new class of PFC recently discovered in surface waters. In this study, the authors developed an LC/MS/MS method to simultaneously measure PFPAs, PFCAs, and PFSAs in drinking water, with method detection limits ranging from 0.014 to 0.17 ng/L. A key to low detection limits for PFPAs was the use of 1-methyl piperidine in the mobile phase, which provided better chromatographic resolution and sharper peaks, 2822



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nate (PFECHS), was made in surface waters and fish in the Great Lakes.54 PFECHS is used as an erosion inhibitor in aircraft hydraulic fluids. A mean log bioaccumulation factor of 2.8 was estimated based on fish: water ratios, which fell between PFOA (2.1) and PFOS (4.5). Landfill leachates were the focus of a new occurrence study by Benskin et al., who measured PFCs and assessed the role of PFSA precursor degradation over time.55 PFPeA and PFHxA were the dominant PFCs found, except during a 2-month period when PFOS, PFOA, and several PFSA precursors (Nalkyl perfluorooctane sulfonamides and fluorotelomer carboxylic acids) increased by a factor of 2−10. Precursors accounted for 71% of the total PFCs on a molar basis, and total PFCs of 8.5−25 kg/year were estimated to leave the municipal landfill via leachate, which is then sent to a wastewater treatment plant. Urban stormwater runoff was the focus of another study by Xiao et al., who used LC/MS to measure 6 PFCs in stormwater runoff from 7 storm events at several different areas corresponding to different watershed land uses.56 The mass flux of PFCs from stormwater runoff in the Minneapolis-St. Paul metropolitan area was estimated at 7.86 kg/year. Interestingly, despite the announced phase-out of PFOSbased fluorochemicals, high levels of PFOS were measured on particles and debris resulting from stormwater runoff in industrial and commercial areas, including high levels near a suspected industrial source where PFOS had been used in past decades. Two other studies focused on hot-spot locations near fluorochemical industries. In the first by Lindstrom et al., 10 PFCs were measured in surface and groundwater samples contaminated by the use of biosolids as a soil amendment on agricultural fields.57 These biosolids were obtained from a local municipal WWTP in Decatur, Alabama, that had received waste from local fluorochemical facilities. Elevated PFC levels were found in surface and well water in the vicinity of these agricultural fields, with 22% exceeding the U.S. EPA’s provisional health advisory level for PFOA in drinking water of 400 ng/L. In the other study, Lasier et al. measured PFCs in surface water and sediments from Northwest Georgia, in close proximity to a major carpet manufacturing center, which uses PFCs in carpet production.58 Bioaccumulation in aquatic oligochaetes (worms) was also investigated. PFCAs with 8 or fewer carbons were the most prominent in surface waters, and those with 10 or more carbons predominated in sediments and worm tissues. PFC patterns in the worm tissues were similar to those in the sediments. Concentrations of PFOS in surface waters from two locations exceeded the U.S. EPA provisional health advisory of 200 ng/L. Airports can also be a source of PFC contamination, as evidenced by a major contamination event several years ago in a creek near Pearson International Airport in Toronto caused by a large release of AFFFs. De Solla et al. carried out a study near Munro International Airport in Hamilton, Ontario, because unexpectedly high levels of PFCs, particularly PFOS, had been found in a lake nearby.59 A combination of LC with negative ion-APPI-MS/MS (for neutral PFCs) and negative ion-ESIMS/MS (for acidic PFCs) was used to measure several PFCs of different compound classes in surface waters, amphipods, and shrimp. PFOS dominated the PFCs detected, with mean PFOS levels of 518 ng/g and 130 ng/L in amphipods and water downstream of the airport. Concentrations declined with distance downstream from the airport. As a result, it was likely that the airport is a major source of PFC contamination, even

though there were no known spill events or publicly reported uses of AFFFs with a fire event. Following the recent discovery of ski waxes as a source of PFC contamination, several environmental studies have investigated this source. Plassman and Berger published such a study from Sweden, where PFCAs (C6 to C22) were measured in snow and soil samples from a ski area.60 LC/MS/MS measurements revealed detection of all of these PFCs in snow and soil samples following a skiing competition and after snowmelt. This study also represented the first report of PFCAs with up to 22 carbons in an environmental sample. Concentrations ranged up to 0.8 μg/L in the snow, with levels decreasing from the start to the finish of the ski trail. Distinct differences were seen in the PFCA patterns in the snow vs the soil, and a larger fraction of higher chain PFCAs was present at the start of the ski trail. The calculated PFCA input from the ski competition was smaller than what was already present in the soil, and results revealed evidence for long-term accumulation. PFCs also continue to be sampled in developing countries. Kim et al. reported the first PFC measurements in Vietnam, where 17 PFCs were measured in environmental waters near waste recycling (e-waste and battery recycling) and disposal sites.61 Highest levels were found in a leachate sample from municipal dumping (360 ng/L), and PFC concentrations from the e-waste and battery recycling locations were higher than those in the rural control samples taken. PFOA, PFNA, and perfluoroundecanoic acid (PFUnDA) were the dominant PFCs overall, and PFHpA and PFHxA were almost as abundant as PFOA in the municipal leachate. New methods continue to be developed for PFCs, including a new, highly sensitive total fluorine method, which can be used to determine how much of the total environmental fluorine contamination is accounted for by measured PFCs (“fluoronomics”, as coined by the authors).62 This method developed by Qin et al. uses LC with continuum source-molecular absorption spectrometry (CS-MAS) and can be used to detect untargeted fluorinated organic compounds in environmental and biological samples.62 Fluorine-specific detection limits of 4 pg of F and 5.26 nM for model fluorine compounds was obtained. This method was demonstrated on fluorinecontaining groundwater. Two new methods involved the use of creative extraction technologies. Surfactant-coated Fe3O4 magnetic nanoparticles were used with LC/ESI-MS/MS in another new method by Zhao et al. to analyze seven PFCs in water.63 The magnetic nanoparticles served as an adsorbent and allowed for convenient isolation by a magnetic field. This method enabled concentration factors of 1600 and extremely low detection limits ranging from 0.022 to 0.31 ng/L. Kaserzon et al. developed a new passive sampler to quantify PFCAs and PFSAs in water.64 The sampler was based on a modified polar organic chemical integrative sampler (POCIS) with a weak anion exchange sorbent (Strata XAW) as the receiving phase enclosed between two polyethersulfone diffusion-limiting membranes, and it was used in conjunction with LC/MS/MS. Trace levels of 0.1−12 ng/L could be detected, and quantitative results for PFCs measured in Sydney Harbor (Australia) compared favorably to traditional SPE-LC/MS/MS techniques. New fate studies of PFCs include one by Murakami et al., who investigated the formation of perfluorinated surfactants from biodegradation of precursors.65 Incubation tests demonstrated that perfluorooctane sulfonamide (FOSA) could be degraded by indigenous microorganisms to form PFOS. The 2823



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addition of nutrients promoted this reaction. A 42-day incubation test resulted in significant formation of PFOS, PFOA, and PFNA from precursors in street runoff but not in rainwater or wastewater effluent. Photolysis was investigated in another study by Taniyasu, who exposed aqueous PFC solutions to natural sunlight at high altitudes in Mt. Mauna Kea, Hawaii, and Mt. Tateyama, Japan.66 PFOS was shown to photolyze, and longer chain PFCs were successively degraded to shorter chain compounds, such as PFBA and PFBS, which were more resistant to photodegradation.

Still, LC/tandem-MS is the method of choice for the determination of all classes of pharmaceuticals in aqueous samples. Mainly ESI and sometimes atmospheric pressure chemical ionization (APCI) are the ionization methods used. Major innovations have been made in modern hybrid mass spectrometry systems (e.g., linear ion trap/Fourier transform (FT)-MS, Q-TOF-MS) coupled to liquid chromatography, providing accurate masses of the analytes and information for mass fragments, which can be used to identify the chemical structures. Buchberger wrote an interesting review summarizing a whole set of analytical techniques utilized for the analysis of pharmaceuticals and personal care products in environmental samples.68 The author highlights sampling and sample preparation techniques, as well as detection methodologies comprising different kinds of LC/hybrid-MS systems, GC/MS, capillary electrophoresis (CE) procedures, and immunochemical methods published so far. Environmental Impacts of Pharmaceuticals. While many pharmaceuticals can have an acute or chronic effect on aquatic or other organisms, most of the lowest observed effect concentrations (LOECs) are substantially above the environmental concentrations typically observed (ng/L to low μg/L). However, there are a few notable exceptions, where chronic toxicity LOECs approach levels observed in wastewater effluents. For chronic toxicity, these include salicylic acid, diclofenac, propranolol, clofibric acid, carbamazepine, and fluoxetine. For example, for diclofenac, the LOEC for fish toxicity was in the range of wastewater concentrations and the LOEC of propranolol and fluoxetine for zooplankton and benthic organisms was near the maximum measured in wastewater effluents. The antibiotic ciprofloxacin has also been shown to have effects on plankton and algae growth at environmentally relevant concentrations.1 Estrogenic effects on wildlife are quite possible with the contraceptive EE2, as it can induce estrogenic effects in fish at extremely low concentrations (low and sub ng/L). Effects include alteration of sex ratios and sexual characteristics and decreased egg fertilization in fish.1 An article in Science (Brodin et al.251) highlighted ecological effects of benzodiazepine anxiolytic oxazepam. Environmentally relevant concentration of 1.8 μg/L oxazepam altered behavior and feeding rate of wild European perch (Perca f luviatilis). Individuals exhibited increased actitivity and reduced sociality when exposed to oxazepam. Bioaccumulation factors (BAF) up to 9.7 were determined. That fresh water fish are up-taking anti-inflammatory drugs (diclofenac, naproxen, ibuprofen) was underlined by the study of Brozinki et al.67 The authors developed an analytical method including an enzymatic hydrolysis of glucuronide conjugates using β-glucuronidase/aryl-sulfase prior to a SPE (Oasis HLB) and with detection via LC/ESI-QqQ-MS. They detected diclofenac, naproxen, and ibuprofen in the bile of wild bream and roach with concentrations up to 148, 102, and 34 ng/mL, respectively. Zuloaga et al. published a review about the extraction, cleanup, and detection techniques for the determination of organic pollutants in sewage sludge.69 The review provided a very comprehensive overview about the determination of lipophilic (more classical pollutants), such as polychlorinated biphenyls (PCBs) or polyaromatic hydrocarbons (PAHs), as well as more polar contaminants, such as pharmaceuticals and personal care product ingredients. Biological Transformation Products. Even though TPs have gained increasing interest as water contaminants, still a

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PHARMACEUTICALS AND HORMONES Pharmaceuticals and hormones have become important emerging contaminants, due to their presence in environmental waters (following incomplete removal in wastewater treatment or diffuse-source contamination), threat to drinking water, and concern about possible estrogenic and other adverse effects, both to wildlife and humans. A major concern for pharmaceuticals also includes the development of bacterial resistance (creation of “Super Bugs”) from the release of antibiotics to the environment, and there are also concerns that antibiotics will decrease biodegradation of leaf and other plant materials, which serves as the primary food source for aquatic life in rivers and streams. It is estimated that approximately 3000 different substances are used as pharmaceutical ingredients, including painkillers, antibiotics, antidiabetics, betablockers, contraceptives, lipid regulators, antidepressants, and impotence drugs. However, only a small subset of these compounds has been investigated in environmental studies so far. Pharmaceuticals are introduced not only by humans but also through veterinary use for livestock, poultry, and fish farming. Various drugs are commonly given to farm animals to prevent illness and disease and to increase the size of the animals. One lingering question is whether the relatively low environmental concentration levels of pharmaceuticals (generally the ng/L range) will cause adverse effects in humans or wildlife. As a consequence, a number of pharmaceuticals and hormones are included on the U.S. EPA’s final CCL-3 (http:// water.epa.gov/scitech/drinkingwater/dws/ccl/ccl3.cfm). These include erythromycin, nitroglycerin, 17α-ethinylestradiol (EE2), 17α-estradiol, estrone (E1), 17β-estradiol (E2), estriol (E3), equilenin, equilin, mestranol, and norethindrone. Article 8b of Directive 2013/39/EU states that “the Commission shall establish a watch list of substances for which Union-wide monitoring data are to be gathered for the purpose of supporting future prioritisation exercises”. Diclofenac as well as the two hormones EE2 and E2 are suggested to be included in the first watch list. There are also increasing “source-to-tap” studies considering the fate of pharmaceuticals from wastewater to river water to source water, and to finished drinking water, such that the complete cycle of pharmaceutical fate is being considered. Innovative analytical instrumentation, such as hybrid mass spectrometry enables the identification and quantification of pharmaceuticals and hormones down to the pg/L and pg/kg range in environmental matrixes and drinking water. Microbial transformation products (TPs) of pharmaceuticals and hormones can be formed during biological wastewater treatment, from contact with sediment and soil, and during bank filtration. Further, TPs can be formed by UV irradiation in surface waters and during oxidative treatment processes, such as ozonation and chlorine disinfection. 2824



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containing nitrite. Under conditions present at WWTPs (neutral pH, low nitrite concentrations, no sample freezing), the formation of appreciable concentrations of nitro- and nitroso-phenolic TPs is unlikely, since this reaction is rather slow at neutral pH and the small amount of nitrite is not sufficient. However, phenolic compounds are transformed in aerobic sludge systems by hydroxylation as well as through the formation of sulfate conjugates, which can be formed with the psychoactive drug dextrorphan, bisphenol A, bisphenylol, as well as with their nitrophenolic TPs. Nödler et al. reported the abiotic transformation of the sulfonamide antibiotic sulfamethoxazole (SMX) under denitrifying conditions in water-sediment systems.73 They identified two TPs, nitro-SMX and desamino-SMX, and confirmed their formation by synthesized standards. In the absence of nitrate under anaerobic conditions, nitro-SMX is reconverted into SMX. Both TPs were detected in spring samples from a ́ German karstic groundwater area. Garcia-Galá n et al. reported that in an aerated fixed-bed reactor, N4-acetylsulfapyridine was completely converted to sulfapyridine until 32 days, while N4acetylsulfamethazine was stable until 90 days.74 The transformation of the benzodiazepines diazepam, oxazepam, and bromazepam in flow-through pilot bioreactors was reported by Kosjek et al.75 Diazepam was converted in 4 biotic TPs, including oxazepam, temazepam, nordazepam (all three known from human metabolism), and a proposed hydroxylated TP. All are formed by N-dealkylation, oxidation, or hydroxylation. For oxazepam, one biotic TP was found, which can be formed by reductive reactions. Applying UV/ H2O2 treatment caused the formation of several further hydroxylated TPs. Mineralization was not observed in any of the processes. Horvat et al. published an excellent review about the analysis, fate, and effects of antihelmintics in the environment.76 Because of the elevated polarity and the molecule size of certain compounds (e.g., macrocyclic lactones), LC/QqQ-MS is the method of choice for detection. However, the number of studies dealing with the occurrence of anthelmintics in environmental matrixes is very scarce. Several benzimidazoles, such as albendazole (ABZ) and fenbendazole (FBZ), have a sulfide moiety, which can be oxidized to albendazole-sulfoxide (ABZ-SO) and fenbendazole-sulfoxide (FBZ-SO). The sulfoxide derivatives can be further oxidized to the sulfone derivatives (ABZ-SO2, FBZ-SO2). The oxidation of the sulfur increases significantly the polarity of the compounds. The ketone group of mebendazole (MBZ) and flubendazole (FLU) can be reduced to a hydroxyl group, leading to enantiomeric TPs due to the formation of a chiral C-atom. The reduced MBZ and FLU TPs have been identified as the principle TPs. Several of the anthelmintics are photosensitive, and thus, a combination of photodegradation and biological transformation is likely. Sulfoxide TPs were also identified by Luft et al.252 for the biocides irgarol and terbutryn. In contact with activated sludge, both were transformed into irgarol sulfoxide and terbutryn sulfoxide, the only transformation product (TP) in significant quantites. Both TPs were tentatively identified by HR-MS, and finally confirmed by NMR as well as by spectra of authentic synthesized reference standards. Irgarol sulfoxide and terbutryn sulfoxide were detected in the effluent of 4 WWTPs up to 22 ng/L and 65 ng/L as well as in German rivers and streams up to 14 ng/L and 34 ng/L, respectively. Luminescent bacteria inhibition test with Vibrio f ischeri indicated that both

limited number of studies have investigated the formation and fate of biological TPs of pharmaceuticals in contact with biologically active matrixes, such as activated sludge or sediments. One reason is the challenge of structural elucidation of TPs present at low concentrations in natural matrixes. Sophisticated analytical techniques are needed, such as hybrid high-resolution mass spectrometry and NMR. Although the target compound is known, with a few exceptions of very simple reactions (e.g., hydrolysis of amides and esters), quadrupole-MS and even high resolution-MS (e.g., LC/ Orbitrap-MS) are often not sufficient to obtain or confirm chemical structures of TPs. The TP structure suggestions based on exact masses and mass fragments have to be confirmed by complementary analytical methods or chemical derivatization reactions specifically altering the new functional moieties formed. Possibilities of analytical methods include a wide variety of currently available NMR techniques or, to a much lesser extent, infrared (IR) spectroscopy. However, a drawback of both techniques is the elevated quantity and the high purity needed for isolated standards. In those cases where no authentic standard is available and only MS spectra of TPs have been obtained, we might better define the suggestions of the chemical TP structures as “tentative identifications” unless further plausibility criteria are fulfilled, confirming the proposed chemical structures. So far, several enzyme-catalyzed reactions are quite commonly observed in the transformation of pharmaceuticals: mono- and dihydroxylation, oxidation of alcohol, aldehyde, and sulfide (thioether) moieties, reduction of ketone groups, ester and amide hydrolysis, N-dealkylation, N-deacetylation, and oxidative decarboxylation. Hernandez et al. reported a study in which retrospective LC/ QTOF-MS analysis was performed to identify TPs.70 WWTP effluents (100 mL) from the Spanish Mediterranean region were preconcentrated via Oasis HLB cartridges and detected via LC/QTOF-MS. They used the MSE mode (simultaneous recording of two acquisition modes at low and high collision energy) to compare the fragmentation patterns of the potential TPs with those of the target compounds. The identification was supported by databases containing high-resolution mass spectra of a large number of compounds as well as software predicting the MS fragmentation. They identified five TPs, such as Ndesmethyl-clarithromycin, 14-hydroxy-clarithromycin, clopidogrel-carboxylic acid, 4-hydroxy-omeprazol sulfide, and 4desmethoxy-omeprazole. A similar approach was used by Gomez-Ramos et al.71 They included the use of an accurate mass database for the systematic identification of TPs in the LC/QTOF-MS spectra. They used an automatic screening procedure including an accurate mass database of compounds and fragments. A further confirmation was suggested via MS fragmentation. However, the authors concluded that an unequivocal confirmation of the chemical structures is frequently not feasible without available authentic standards. Jewell et al. reported that in batch experiments with activated sludge under acidic conditions (pH < 5), phenolic compounds are transformed into nitro- and nitroso-phenolic TPs when nitrite is present.72 The authors mentioned that these abiotic chemical reactions were also observed when water samples were only frozen, since during freezing an unexpected pH drop promotes the reaction of nitrite with phenolic compounds. Hence, the presence of nitrite leads to an artifactual formation of TPs when water samples are frozen. Therefore, special care has to be taken for the storage of wastewater samples 2825



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GC/MS.84 Complex degradation pathways were suggested by Cai et al. with more than 15 intermediates and stable OPs proposed.84 The oxidation kinetics were influenced by the concentrations of bromide and chloride. El Najjar et al. investigated the conversion of the fluoroquinolone antibiotic levofloxacin via chlorination using LC/APCI-QqQ-MS.85 The authors propose the replacement of a carboxylic moiety by a chlorine atom (halodecarboxylation) and a piperazine fragmentation (N-dealkylation). Also for amphetamine-type stimulants, two chlorinated OPs have been identified by Huerta-Fontela et al. with LC/QqQtrap.86 3Chlorobenzo-1,3-dioxole was formed by chlorination of water containing 3,4-methylenedioxyamphetamine (MDA) or 3,4methylenedioxyethylamphetamine (MDEA), while methylenedioxymethamphetamine (MDMA) was converted into 3chlorocatechol. The latter was found in drinking water up to 5.8 ng/L, while 3-chlorobenzo-1,3-dioxole was eliminated after ozonation and GAC treatment. The antibiotic trimethoprim was converted during ClO2 disinfection into monochlorinated and dichlorinated and hydroxylated OPs.87 Even the formation of a nitro-moiety was postulated. Similar to treatment with other oxidants (e.g., ferrate, ozone), its antimicrobial activity should be lost. Li et al. reported the formation of OPs by the chlorination of water containing oxcarbamazepine using LC/IT-MS.88 In addition to monochloro- and dichloro-oxcarbamazepine, the main OP was the unchlorinated 1-(2-benzaldehyde)-(1H, 3H)-quinazoline2,4-dione (BQD), which is already known to be the main OP from ozonation of carbamazepine.89 The ozonation of oxcarbamazepine leads predominately to 1-(2-benzylic acid)(1H, 3H)-quinazoline-2,4-dione (BaQD), which has been also identified as minor oxidation product of BQD via ozonation of carbamazepine. The main OPs of this study were found in the lower ng/L range in drinking water due to the transformation of oxcarbamazepine or carbamazepine. After ozonation, 21 OPs have been observed for trimethoprim by Kuang et al. using LC/ ESI-QqQ-MS and LC/ESI-Q-TOF-MS. Typical reactions were hydroxylation, demethylation, cleavage of the methylene group, and deamination.90 The oxidation of acyclovir and its main biotransformation product (carboxy-acyclovir) via ozone was investigated by Prasse et al.91 At pH 8, a single oxidation product was formed, which was identified via LC/LTQ-Orbitrap-MS and NMR as N-(4-carbamoyl-2-imino-5-oxoimidazolidine)formamido-N-methoxyacetic acid (COFA). Using Vibrio f ischeri, an acute bacterial toxicity was found for COFA, while carboxy-acyclovir showed no toxic effects. Ozonation experiments with guanine and guanosine led to the formation of the respective 2-imino-5oxoimidazolidines, confirming that guanine derivatives, such as carboxy-acyclovir are undergoing the same reactions during ozonation. Furthermore, COFA was even detected in finished drinking water after ozonation and subsequent activated carbon treatment. Electrochemical oxidation with boron doped diamond (BDD) electrodes led to an efficient transformation of sulfamethoxazole.92 In total, 9 OPs were identified by de Vidales et al.,92 including hydroxylation, dihydroxylation, cleavage and hydrolysis of sulfonamide moieties, dealkylation of sulfonamide structures, or oxidation of amines leading to nitro derivatives. Luetke Eversloh et al. investigated the electrochemical transformation of the X-ray contrast medium iopromide.93 Anodic oxidation with BDD and cathodic reduction using a

TPs feature a similar bacterial toxicity than the parent compounds. Elimination/Reaction During Oxidative Water Treatment. Several studies confirmed the efficiency of oxidation processes, such as ozonation, potassium permanganate, advanced oxidation, or ferrate (Fe(VI)), for the transformation/removal of micropollutants. However, it is rather likely that oxidation leads to persistent abiotic oxidation products (OPs) which are of toxicological concern. This might be even more relevant for chlorination, since chlorinated products frequently possess an elevated toxicological potential. It is therefore crucial to identify the oxidation products formed. This is only possible when using advanced high-resolution mass spectrometry, such as LC/Q-TOF-MS, LC/Orbitrap-MS, or LC/Qq-linear ion trap-MS, in combination with NMR techniques. Hu et al. showed that potassium permanganate (Mn(VII)) treatment led to a significant transformation of selected antibiotics (ciprofloxacin, lincomycin, trimethoprim).77 In total, 26 OPs were observed via LC/ESI-QqQ-MS, and complex oxidation pathways could be proposed, but the portion of mineralization was assumed to be negligible. The proposed chemical structures indicate in most cases an increased polarity due to the introduction or release of polar moieties, such as OH-, carboxy-, ketone-, or amide-moieties. However, in all cases, the antibacterial potency was mostly lost, indicating the success of treatment with Mn (VII). Wilde et al. reported an efficient oxidation of three betablockers (atenolol, metoprolol, propranolol) by ferrate (Fe(VI)).78 More than 30 OPs were suggested, based on LC/ESIion trap-MSn measurements. Comparable OPs were reported as previously described by Benner and Ternes using the oxidant ozone (O3).79,80 Zimmermann et al. proposed a mechanism for the transformation of the opiate tramadol using ferrate (Fe(VI)) and ozone as oxidants.81 In total, 6 OPs were identified via LC/ ESI-LTQ-Orbitrap-MS and LC/ESI-QqLIT-MS, which were further oxidized with excess of the oxidants. However, ferrate led mainly to N-dealkylated, formamide, and aldehyde OPs, while ozonation favored ∼90% formation of the tramadol Noxide. As described by Hu et al.77 and Wilde et al.,78 the oxidants were attacking the functional groups with elevated electron densities, such as amino groups, but reactions did not lead to complete mineralization when reasonable oxidant quantities were applied. Therefore, for an overall process evaluation, the (eco)toxicity of the OP mixtures will be most crucial but is very challenging, as it is known for mixture toxicity. Shen and Andrews investigated whether prechlorination of river water or lake water influences the formation of Nnitrosodimethylamine (NDMA) if amine-based pharmaceuticals are present.82 They concluded that this strongly depends on the water source and the natural organic matter (NOM). In general, chlorination and chloramination reduced the NDMA formation in the presence of ranitidine or sumatriptan, but this strongly depends on the compounds present, the chlorination conditions (dose and contact time), and NOM having a major influence on the NDMA formation. The chlorination of water containing phenazone and propyphenazone leads to an elevated number of halogenated, hydroxylated, N-dealkylated, and dephenylated derivatives of both analgesics.83,84 The proposed chemical structures were elucidated via LC/ESI-Q-TOF83 and LC/ESI-QqQtrap and 2826



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phototransformation products were only observed once, at 34.2, 3.2, and 11.9 ng/L. Direct and indirect photolysis (in the presence of bicarbonate and nitrate) caused the rapid transformation of 5-fluorouracil, according to Lin et al.100 Mineralization only occurred to a very minor extent. The phototransformation product identified was suggested to be 3-amino-2-fluoro-acrylic acid. A similar study was reported by Wang and Lin, who investigated the phototransformation of cephalosporin antibiotics in aqueous media.101 Cefazolin and cephapirin underwent mainly a direct photolysis (e.g., by decarboxylation), while cephalexin and cephradine were mainly transformed by indirect photolysis in a bicarbonate-nitrate containing solution. Mineralization was not observed. In total, the chemical structures of 12 TPs were proposed. Captopril, a sulfhydrylcontaining angiotensin-converting enzyme (ACE) inhibitor, was phototransformed mainly to captopril sulfonic acid, which seemed stable to microbial degradation.102 Radjenovic et al. investigated the influence of tramadol transformation by UV−C and UV−C/H2O2 on NDMA formation potential.103 The authors observed the formation of 24 TPs, which were tentatively identified using LC/ESIQqLIT-MS combined with online hydrogen/deuterium (H/D) exchange. Reactions, such as oxidative cleavage of methoxy and methoxybenzene moieties, O-desmethylation, hydroxylation, and ring-opening were proposed. The majority of the TPs were likely to have a higher NDMA formation potential than tramadol. Opiates and Other Drugs of Abuse. A new book on illicit drugs in the environment was published by Castiglioni et al.104 This book features informative chapters authored by many researchers from the U.S. and Europe who are active in this area. Chapters cover environmental occurrence (in wastewater, river water, groundwater, drinking water, air, and particulate matter) in several countries, analytical methods for measuring illicit drugs, implications for ecotoxicology, and the use of mass spectrometry as a tool to track drug use patterns. Pal et al. published an excellent review on illicit drugs in the environment, discussing the potential for distribution of illicit drugs and their metabolites in water, summarizing their occurrence in wastewater, surface water, groundwater, drinking water, sludge, and air, reviewing their potential ecotoxicity, and reviewing analytical methods for their measurement.105 This review also contains helpful tables that summarize the occurrence data, with locations and concentrations listed. Vazquez-Roig et al. published another nice review covering extraction methods (e.g., offline and online SPE and solventless methods), detection methods, and future trends for the analysis of legal and illegal drugs in the aquatic environment.106 The authors discuss advantages and pitfalls of multiclass methods, with an emphasis on new strategies for sample preparation and recent technical developments. France was the focus of a new, extensive occurrence study, in which 17 illicit drugs and metabolites were measured in influents and effluents of 25 WWTPs across the country.107 A new LC/MS/MS method was developed, and the results were used to evaluate drug consumption and the efficiency of wastewater treatment in removing them. Influents ranged in concentration between 6 ng/L for 2-ethylidene-1,5-dimethyl3,3-diphenylpyrrolidine (EDDP, a primary metabolite of methadone) and 3050 ng/L for benzoylecgonine (cocaine metabolite). Consumption maps drawn for cocaine, MDMA, opiates, cannabis, and amphetamine-like drugs revealed that

platinum electrode were used. Mineralization up to 96% was achieved after about 7.5 h. At shorter treatment times, several oxidatively and reductively formed TPs were detected via MSnfragmentation with LC/ESI-Orbitrap-MS, whereas deiodinated iopromide represented the major fraction. During electrolysis, the iodine released from iopromide was found to be 90% as iodide and 10% as iodate, even in the open cell experiments, limiting the potential formation of toxic iodo-DBPs. Chlorinated TPs were not found. Anodic oxidation with mixed metal oxides represents another alternative for electrochemical degradation of emerging contaminants. Radjenovic et al. described the treatment of several pharmaceuticals and pesticides in reverse osmosis concentrates using RuO2/IrO2 electrodes.94 In a similar manner, the same group investigated the electrochemical transformation of metoprolol, resulting in various halogenated byproducts, giving rise to an increased toxicity of the reaction mixture.95 Photodegradation. Calza et al. reported the conversion of carbamazepine and clarithromycin in purified water using heterogeneous photocatalysis with TiO2.96 For carbamazepine, 28 TPs were identified via MS fragmentation using LC/LTQOrbitrap-MS, formed by mono-, di-, trihydroxylation, oxidation, and ring contractions, leading to derivatives of 9-N-aminocarbonylacridine-9-carboxaldehydes. Similar reactions (hydroxylations, oxidations) and the cleavage of the sugar moiety cladinose led to the formation of 29 TPs when the macrolide antibiotic clarithromycin was present. In another study, Calza et al. highlighted the formation of 13 major TPs when a TiO2 solution was irradiated with a 40 W lamp (emission maximum: 360 nm).97 The TPs were formed by reactions, such as mono-, di-, trihydroxylations, oxidation, dephenylation, demethylation, and ring-opening. Until 1 h of irradiation, no TPs could be detected at all. The phototransformation of the thyroid hormone levothyroxine led to the formation of at least 5 TPs according to Svanfelt et al. via LC/Q-TOF-MS.98 The chemical structure of two TPs could be confirmed either by NMR or by an authentic standard, and for three further TPs, tentative chemical structures were proposed. The main TP exhibited a loss of two iodines and a C-atom, while two hydrogens and one oxygen were added. The confirmed chemical structures can only be formed by a ring-opening of the benzene ring, leading to a carboxylic moiety and two double bonds. This is an excellent study exhibiting the strength when HR-MS and NMR are combined to elucidate the chemical structures of TPs formed by complex reactions. Furthermore, the formation of 3,5-diiodotyrosine was confirmed due to the availability of an authentic standard. For the remaining three TPs, a contraction of the 6-membered benzene ring to a cyclopentadiene ring was proposed, accompanied by deiodination, hydroxylation, and oxidation. The identification of phototransformation products of testosterone was accomplished using NMR techniques and LC/ESI-QqQ-MS (Vulliet et al.).99 The chemical structure of three TPs were elucidated, which are formed by complex contraction and rearrangements of the α, β-saturated cyclohexanone ring (ring A). An analytical method consisting of solid phase extraction (Oasis HLB and StrataX) and detection via LC/ESI-QqQ-MS achieved detection limits down to a few ng/L. In four surface waters monitored, testosterone was present in a range of 0.3−15.3 ng/L, while the three 2827



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sludge particles due to their elevated polarities. Only in a very few cases, TPs of cytostatic agents have been identified so far. Uracil arabinoside was formed via hydrolytic deamination of cytarabin, or 7-hydroxymethotrexate was formed by hydroxylation of methorexate. Analytical methods consist of SPE for enrichment and mainly LC/ESI-QqQ-MS or LC/fluorescence detection. Only in a few cases, GC/MS was used after derivatization. Gomez-Canela et al. developed an analytical method with detection via LC/Orbitrap-MS to determine cyclophosphamide and epirubicin in aqueous samples down to 0.35 ng/L with SPE (Oasis HLB) or alternatively, with direct injection down to 3 ng/L.113 A simultaneous determination of 14 cytostatic pharmaceuticals in raw and treated wastewater and surface water with LC/QqQ-MS after SPE (Oasis HLB) was reported by Martin et al.114 The authors observed severe matrix effects due to ion suppression (up to 60%) and ion enhancement (up to 20%) during ionization. A total of 6 of 14 cytostatic pharmaceuticals were found in WWTP influents and WWTP effluents in a range of 3.5−15 ng/L and 1.2−14 ng/L, respectively. Seira et al. developed an analytical method for the determination of ifosfamide and cyclophosphamide in sewage sludge with limits of quantification (LOQs) down to 8.8 and 6.1 μg/kg, respectively.115 The method consists of pressurized solvent extraction of lyophilized sludge, clean up via SPE, and detection via UPLC/ESI-QqLIT-MS. Maximum concentrations were detected up to 42.5 μg/L. Kosjek et al. investigated the biodegradation and removal via UV of 5-fluorouracil and its product capecitabine by using UPLC/ESI-QqTOF-MS.116 Both compounds were biodegraded in aerobic batch experiments with activated sludge. However, TPs could not be observed in these biotic systems, while after photodegradation with a low pressure mercury lamp (254 nm), 5 TPs with 5-fluorouracil and 10 TPs with capecitabine were observed and proposals for their chemical structures were provided. Multiresidue Methods. A multiresidue method was developed by Baker and Kasprzyk-Hordern for the determination of 65 pharmaceuticals and illicit drugs in aqueous matrixes from raw and treated wastewater to river water.117 Detection limits down to 0.5 ng/L were obtained for compounds such as cocaine and benzoylecgonine and up to 154 ng/L for caffeine. After SPE at pH 2 with Oasis MCX, the sample extracts were measured with UPLC/ESI-QqQ-MS. A total of 46 of the selected compounds were detected in raw wastewater and 36 compounds in U.K. rivers and streams. Elevated concentrations were found for amphetamine and tramadol, with maximum concentrations in WWTP influents of 2.3 and 6.2 μg/L, respectively. In addition, the tramadol metabolite nortramadol was present at 7.1 μg/L. Ferrer and Thurman developed an analytical method for the analysis of 100 pharmaceuticals and their transformation products in surface water using LC/ESI-Q-TOF-MS after SPE extraction with Oasis HLB (200 mg, 6 mL), attaining LODs from 1 to 1000 ng/L.118 Chitescu et al. reported a multiresidue screening method to determine 43 pharmaceuticals and fungicides in surface water and groundwater using SPE with Strata X (200 mg, 6 mL) and UPLC/ESI-ExactiveOrbitrap-MS within a concentration range of 10−100 ng/L.119 Gros et al. were able to detect 81 pharmaceuticals in drinking water, surface water, and treated and untreated wastewater

drug use inside a country is not homogeneous. Finally, a few illicit drugs were not well removed in wastewater treatment, leading to higher levels in the receiving waters. Chlorination, hydrolysis, and photolysis products of cocaine and benzoylecgonine were the focus of another very interesting paper by Bijlsma et al.108 UPLC/QTOF-MS was used to identify the 16 TPs formed in the controlled laboratory reactions. Eight chlorination DBPs were formed by the reaction of cocaine with chlorine, 7 TPs were formed by photolysis, and 1 was formed by hydrolysis. Three of these were well-known cocaine metabolites, and their identities could be confirmed with authentic standards (benzoylecgonine, norbenzoylecgonine, and norcocaine). In total, 10 TPs of benzoylecgonine were identified, including 3 chlorination products and 7 photolysis products. TPs resulted from chlorination, dealkylation, hydroxylation, and nitration processes. Following TP identification in controlled lab studies, they were measured in real wastewater influent and effluent, where benzoylecgonine, norcocaine, and norbenzoylecgonine were found as well as 4 previously unreported TPs. Antidiabetic Drugs. Trautwein and Kümmerer investigated the biotransformation of the antidiabetic drug metformin in the Zahn-Wellens test (OECD 303B).109 Using LC/Q-ion trap-MS up to MS3, they were able to identify guanylurea as the main TP, which was formed from metformin by a 2-fold dealkylation and an oxidative deamination. In a WWTP, the concentration of metformin decreased from 56.8 μg/L (influent) to 1.86 μg/L (effluent), while the concentrations of guanylurea increased from 0.40 μg/L (influent) to 1.86 μg/L (eluent). Thus, they concluded that metformin is biotransformed in WWTPs into guanylurea. Scheurer et al. performed a comprehensive study including 5 German WWTPs and several large German rivers and creeks analyzed after SPE (Strata XCW) by LC/ESI-QqQ-MS.110 They concluded that both metformin and guanylurea are permanently discharged in appreciable concentrations in the lower μg/L range in rivers and streams. The concentrations in rivers were directly correlated with their proportion in treated wastewater. Riverbank filtration and artificial groundwater recharge were efficient methods to completely remove metformin and guanylurea, whereas flocculation and activated carbon filtration were ineffective. An analytical method was developed for six antidiabetic drugs with SPE (Oasis HLB) and LC/ESI-Q-TOF-MS detection by Martin et al.111 Metformin was present, with the highest concentrations up to 0.25 μg/L and 0.10 μg/L in Austrian WWTP effluents and in Austrian surface water, respectively. Two other antidiabetic drugs, sitagliptin and vildagliptin, were found with concentrations up to 117 ng/L in WWTP effluents, while glibenclamide and pioglitazone were not found at all. Cytostatic Pharmaceuticals. Kosjek and Heath wrote a comprehensive review summarizing the current literature regarding analytical methods, occurrence, and fate of cytostatic pharmaceuticals.112 In hospital wastewater, concentrations up to 4.5 μg/L (cyclophosfamide) were detected, while in WWTP effluents and in surface water, the concentrations are mainly in the lower ng/L range or even in the pg/L range. Thus, the accurate quantification of cytostatic pharmaceuticals is very challenging, especially in highly TOC loaded matrixes, such as wastewater. The authors concluded that cytostatics are unlikely to be appreciably removed by WWTPs, since (a) their biodegradation potential is very limited based on lab studies and (b) they have only a very low tendency to sorb to activated 2828



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publications are considering the environmental fate of pharmaceutical enantiomers. In a recent study by Barclay et al., an analytical method for enantionselective quantification of fluoxetine and norfluoxetine with chiral LC/ESI-QqQ-MS detection was developed.127 The enantiomeric fractions (EFs) (0.68−0.71) were found to be similar in raw and treated water. The concentrations of the Sisomers of fluoxetine and norfluoxetine are higher than the Risomers. The detected concentrations of fluoxetine and norfluoxetine were always below 20 ng/L. Bagnall et al. reported two chiral LC/ESI-QTOF-MS multianalyte methodologies for the enantiomeric detection of amphetamine, methamphetamine, MDA, MDMA, fluoxetine, and venlaflaxine, as well as the beta-blockers atenolol, propranolol, and metoprolol in WWTP effluent and river water.128 Microextraction Method. Pintado-Herrera et al. developed a multianalyte method for the simultaneous determination of a wide range of pharmaceuticals and ingredients of personal care products in aqueous and solid matrixes.129 Pressurized hot water extraction (a modified version of pressurized liquid extraction) was combined with stir bar sorptive extraction, followed by derivatization using N-(tertbutyldimethylsilyl)-N-methylfluoroacetamide (MTBSTFA) prior to detection by GC/MS. A limit of detection down to 1 ng/L and 0.1 ng/g could be achieved. Solid Contact Potentiometric Sensors for Pharmaceuticals. The development of solid contact sensors for the rapid screening of the sulfonamide antibiotics sulfadiazine and sulfamethoxazole were reported by Almeida et al.130,131 Solidcontact graphite-based electrodes were developed for a selective determination of selected sulfamides in media with elevated concentration levels, such as biological fluids and aquacultures. The template molecule was imprinted in sol−gel and used as the detecting element. This was achieved by employing it either as a sensing layer or as an ionophore of PVC-based membranes and subsequent potentiometric transduction. A cyclodextrin potentiometric sensor was developed for the detection of ibuprofen.132 Detectable concentrations ranges were in the high μg/L range. Bioassays. Nicolardi et al. applied an enzyme-linked immunosorbent assay (ELISA) for the detection of caffeine and cotinine, with detection limits of 0.14 and 0.047 μg/L, respectively.133 They obtained comparable results for rivers and streams as measured by LC/QqQ-MS. The caffeine ELISA displayed cross-reactivities for paraxanthin (68%), theophylline (5.2%), and theobromine (2.8%); the cotinine ELISA showed cross-reactivities for 3-hydroxycotinine (32%) and nicotine (1.1%).

using UPLC/Qq-linear ion trap-MS after SPE with Oasis HLB (60 mg, 3 mL).120 Liu et al. developed an analytical method for the detection of 28 steroids (4 estrogens, 5 glucocorticoids, 14 androgens, 5 progestagens) by LC/ESI-QqQ MS in wastewater, surface water, and sludge samples.121 The authors achieved detection limits down to 0.01 ng/L (surface water) and 0.08 ng/g (sludge). In total, 10 steroids (cortisone, estrone, testosterone, progesterone, epi-androsterone, 4-andrestone-3,17-dione, 17βboldenone, norgestel, androsta-1,4-dien-3,17-dione, and 5αdihydroxytestosterone) were found in surface water, with a maximum concentration of 55.3 ng/L for 5α-dihydroxytestosterone. In sludge samples, 12 steroids were found, with concentrations ranging from 1.6 ng/g (E1) to 372 ng/g (epiandrosterone). Kormos et al. developed a LC/QqQ-MS method for the simultaneous determination of four iodinated X-ray contrast media (ICM) and 46 ICM TPs in raw and treated wastewater, surface water, groundwater, and drinking water.122 LOQs varied between 1 ng/L and 3 ng/L for surface water, groundwater, and drinking water and between 10 ng/L and 30 ng/L for wastewater. In total, 26 TPs were detected above their LOQ in WWTP effluents. Most of the polar ICM TPs, such as iohexol TP599, iomeprol TP643, iopromide TP701A, and iopromide TP643, were not removed or were only partially removed during drinking water treatment. As a consequence, several ICM TPs were detected in drinking water, at concentrations exceeding 100 ng/L, with a maximum of 500 ng/L for iomeprol TP687. A multianalyte method was developed for the detection of 35 human pharmaceuticals, illicit drugs, and bactericides in sediment and sewage sludge.123 After pressurized solvent extraction of freeze-dried samples, the extracts were measure via UPLC/ESI-QqQ-MS. The limits of detection varied from 1 to 50 ng/g. In Scottish sludge, the pharmaceuticals carbamazepine, atenolol, and citalopram ranged from 63 to 317 ng/g and the bactericides triclosan and triclocarban ranged from 19 to 5940 ng/g. The illicit drugs cocaine, its metabolite benzoylecgonine, amphetamine, and methamphetamine were not detected in any of the samples. Sample Stability and Impact of Silanized Glassware. Sample stability is an important aspect to consider for environmental samples because, for example, elevated microbial activity can lead to lowered concentrations in aqueous matrixes. Baker et al. found that 22 of 65 pharmaceuticals and illicit drugs analyzed exhibited a change of more than 15% after 12 h at a temperature of 2 °C.124 This highlights that knowledge of analyte stability in the monitored matrix is crucial prior to designing sampling strategies. Furthermore, the authors found a positive effect of silanized glassware on the recovery of many analytes considered. Carlson et al. reported an average loss of 9% (maximum 19%) of 20 pharmaceuticals (e.g., beta-blockers, sulfonamide antibiotics, carbamazepine) and three pesticides loaded on Oasis HLB stored at −20 °C for 20 months.125 Under the same conditions, losses on polar organic chemical integrative samplers (POCIS) were found to average 11%. Enantiomers. In a comprehensive review, Riberio et al. summarized the analytical methods, fate, and ecotoxicity of chiral pharmaceuticals.126 The authors noted that the availability of methods enabling the detection of enantiomers in environmental matrixes is very limited. Only a few

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DRINKING WATER AND SWIMMING POOL DISINFECTION BYPRODUCTS Drinking Water DBPs. Drinking water DBPs are formed by the reaction of disinfectants (chlorine, chloramines, ozone, chlorine dioxide, etc.) with natural organic matter (NOM) and bromide or iodide in source waters. They can also be formed by the reaction of disinfectants with other organic contaminants, and there is an increasing amount of research in this area. One particularly important discovery in this regard was the formation of high levels of N-nitrosodimethylamine (NDMA) in drinking water that resulted from the reaction of ozone with a fungicide (tolylfluanide) used in Europe.1 New areas in drinking water DBP research include the study of highly 2829



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from several countries. A new epidemiology study by Cantor et al. found an enhanced risk for bladder cancer (odds ratio 5.9) for people with a particular genotype, which can be found in approximately 25% of the U.S. population.1 Dermal and inhalation exposure from showering, bathing, and swimming was a significant risk factor. The findings strengthen the hypothesis that DBPs cause bladder cancer and suggest possible mechanisms as well as classes likely to be implicated. Books and Reviews. Richardson published a new review on DBP formation and occurrence in drinking water.134 This review provides a comprehensive listing of >600 DBPs identified from different disinfectants and disinfectant combinations (updating a 1998 encyclopedia article containing these original comprehensive lists) and includes discussion of formation and occurrence, issues with alternative disinfectants, route of exposure, and formation of “pollutant” DBPs. An excellent new book was written on the chemistry of ozone in water and wastewater treatment by von Sonntag and von Gunten.136 This book details the physical and chemical properties of ozone, ozone kinetics in drinking water and wastewater, inactivation of microorganisms, and toxicological assessment of ozone-induced products of micropollutants, and use of ozonation for drinking water and wastewater treatment, as well as several chapters devoted to specific types of reactions that ozone can participate in. These include reactions with olefins, aromatic compounds, nitrogen-containing compounds, sulfur-containing compounds, compounds with C−H groups that are reactive with ozone, inorganic anions, metal ions, and free radicals. Nice, detailed ozone mechanisms and reaction pathways are given. This book is a must-read, particularly for readers interested in the use of ozone for micropollutant degradation (a hot topic now) but also for anyone desiring a comprehensive look at ozone and OH-radical chemistry. Krasner et al. published an excellent review on the formation, precursors, control, and occurrence of nitrosamines in drinking water.137 This review gives a nice overview of nitrosamine formation mechanisms, through chloramination, ozonation, chlorination, activated carbon, and UV/sunlight reactions. Nitrosamine precursors and model compound studies are also discussed, as well as methods that have been investigated for their removal. These include coagulation and polymer optimization, physical adsorption of precursors, riverbank filtration of precursors, preoxidation of precursors, modification of chloramination process, and destruction of nitrosamines with UV treatment. Occurrence data in the United States, Canada, the U.K., Wales, Scotland, Australia, China, Japan, and Singapore were discussed. The review concluded with research needs, including improved occurrence surveys (with spatial/ temporal coverage and more water quality/operation parameters included), a better understanding of nitrosamine precursors, an understanding of the importance of materials in the distribution systems or plumbing as nitrosamine precursors, refinement and implementation of a total nitrosamine assay, and validation of treatment options at pilot- and full-scale. Bond et al. published a review on precursors of N-DBPs in drinking water.138 Precursors of haloacetonitriles, haloacetamides, halonitromethanes, nitrosamines, and cyanogen halides were discussed, with an emphasis on model compounds that have been studied to-date, including nitrogen-containing precursors such as amino acids, amino sugars, amines, nitrophenols, nitromethane, and nucleic acids, and other small molecules, such as formaldehyde.

genotoxic or carcinogenic DBPs that have been recently identified (e.g., iodoacetic acid, NDMA, and halobenzoquinones), issues with increased formation of many of these with the use of alternative disinfectants (e.g., chloramines and ozone), and routes of exposure besides ingestion. In this regard, there have been several recent studies of DBPs in swimming pools. Other trends include the development of UPLC/MS/ MS methods, increased use of precursor ion scan-MS and HRMS to identify new DBPs, and the combination of analytical chemistry with toxicology to account for toxicological effects with DBPs measured. In addition, near real-time methods are being developed, which could give drinking water utilities a better understanding and control over DBP levels received by consumers and improve exposure characterizations for epidemiologic studies. Toxicologically important DBPs include brominated, iodinated, and nitrogen-containing DBPs (“N-DBPs”). Brominated DBPs are generally more carcinogenic than their chlorinated analogues, and new research is indicating that iodinated compounds are more toxic than their brominated analogues.1 Iodoacetic acid was recently discovered as a chloramination DBP and is the most genotoxic of all DBPs studied to-date.134 Also, just this year, Wei et al. published an important study showing that iodoacetic acid is also tumorigenic in mice.135 Brominated and iodinated DBPs form due to the reaction of the disinfectant (such as chlorine or chloramines) with natural bromide or iodide present in source waters. Coastal cities, where groundwaters and surface waters can be impacted by salt water intrusion, and some inland locations, whose surface waters can be impacted by natural salt deposits from ancient seas or oil-field brines, are examples of locations that can have high bromide and iodide levels. A significant proportion of the U.S. population and several other countries now live in coastal regions that are impacted by bromide and iodide; therefore, exposures to brominated and iodinated DBPs are of growing interest. In 2011, another source of iodine was discovered, Xray contrast media, which contributes to the formation of iodoDBPs. This new discovery is detailed in the section on DBPs of Pollutants. Early evidence in epidemiologic studies indicates that brominated DBPs may be associated with reproductive and developmental effects as well as cancer. Brominated and iodinated DBPs of interest include iodo-acids, bromonitromethanes, iodo-trihalomethanes (iodo-THMs), brominated forms of MX (3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)furanone), haloaldehydes, and haloamides. Iodinated DBPs are increased in formation with chloramination, and bromonitromethanes are increased with the use of preozonation. Besides haloamides, other N-DBPs of interest include NDMA and other nitrosamines, which can form with either chloramination or chlorination (if nitrogen-containing coagulants are used in treatment). Five nitrosamines (NDMA, N-nitrosodiethylamine, N-nitrosodipropylamine, N-nitrosodiphenylamine, and N-nitrosopyrrolidine) as well as formaldehyde (which is a DBP from treatment with ozone, chlorine dioxide, or chlorine), are currently listed on the U.S. EPA’s CCL-3 (http://water.epa. gov/scitech/drinkingwater/dws/ccl/ccl3.cfm). Chloramination has become a popular alternative to chlorination for plants that have difficulty meeting the regulations with chlorine, and its use has increased with the new Stage 2 Disinfectants (D)/DBP Rule (www.epa.gov/safewater/disinfection/stage2). Potential health risks from DBPs include cancer and reproductive/developmental effects, with bladder cancer showing the most consistency in human epidemiologic studies 2830



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Combining Chemistry with Toxicology and Epidemiology. More studies are combining DBP identification/ measurement efforts with toxicology to understand their potential health effects. Also, this year, for the first time, chemistry and toxicology and epidemiology were combined for a multicountry epidemiology study of drinking water and adverse birth outcomes in Europe. In this study called “HiWATE” (Health Impacts of Long-Term Exposure to Disinfection Byproducts in Drinking Water), Jeong et al. published an integrated study of chemistry (target DBP quantification and comprehensive identification) with toxicology (in vitro mammalian cell toxicity of complex drinking water samples) and epidemiology findings.139 In total, 11 drinking water samples were collected from five European countries, each corresponding to a separate epidemiology study for the HiWATE program. More than 90 DBPs were identified, and THMs and HAAs were quantified. The range and type of DBPs reflected the diverse collection sites, the different disinfection processes, and the different characteristics of the source waters. Results revealed that the mammalian cell cytotoxicity correlated with the number of DBPs identified and the several DBP chemical classes. DBP occurrence and cytotoxicity also correlated with epidemiology results (low birth weight and small for gestational age) from two countries. Neale et al. presented an excellent integrated chemistrytoxicology study, in which the formation of DBPs through a full-scale water treatment plant in Australia was assessed using in vitro bioanalytical tools as well as through quantification of halogen-specific adsorbable organic halogen (AOX), characterization of organic matter, and quantification of selected regulated and emerging DBPs.140 Toxicity and formation were evaluated through the different treatment steps comprised of coagulation, sand filtration, chlorination, addition of lime and fluoride, storage, and chloramination. Nonspecific toxicity, using the Microtox assay, Caco-2 and AREc32 cell cytotoxicity assay, and combined algae test, peaked midway through the treatment train, following chlorination and storage. Chlorinated and brominated DBPs caused reactive toxicity to increase after chlorination. Genotoxicity and induction of oxidative stress showed the same pattern as nonspecific toxicity, peaking after chlorination. Quantified chlorinated, brominated, and iodinated DBPs comprised 30−55%, 5−55%, and 9−18% of the adsorbable organic chlorine, bromine, and iodine, respectively. Discovery and Measurement of New DBPs. Halobenzoquinones (HBQs) are a new class of DBP recently identified in Li’s group at the University of Alberta. In a follow-up occurrence and formation study, Zhao et al. reported the measurement of 8 HBQs in 9 drinking water plants in the U.S. and Canada that use chlorination, chloramination, chlorination with chloramination, and ozonation with chloramination.141 SPE with LC/ESI-MS/MS was used for their measurement. 2,6-Dichlorobenzoquinone, 2,6-dibromobenzoquinone, 2,6-dichloro-3-methylbenzoquinone, and 2,3,6-trichlorobenzoquinone were detected in several of the samples, ranging from 4.5 to 275 ng/L. Separate controlled laboratory studies using phenol as a precursor demonstrated that chlorination produced the highest levels of 2,6-dichlorobenzoquinone, while preozonation increased the formation of 2,6-dibromobenzoquinone in the presence of bromide. UV-induced transformation of HBQs was the focus of another study by Qian et al., who reported 90% transformation and subsequent formation of 4 hydroxylated TPs from UV treatment of 4 HBQs (at 1000 mJ/cm2).142 3-Hydroxyl-2,6-dichloro-1,4-benzoquinone, 5-hydroxyl-2,6-di-

chloro-3-methyl-1,4-benzoquinone, 5-hydroxyl-2,3,6-trichloro1,4,-benzoquinone, and 3-hydroxyl-2,6-dibromo-1,4-benzoquinone were observed as products, and these could be further degraded to monohalogenated benzoquinones when the UV dose was higher than 200 mJ/cm2. In another study by Pan and Zhang, precursor ion scan with UPLC/MS/MS was used to identify 11 new aromatic halogenated DBPs in an investigation of the effect of different bromide levels on DBP formation in chlorinated drinking water.143 The new DBPs were classified into four groups: dihalo-4-hydroxybenzaldehydes, dihalo-4-hydroxybenzoic acids, dihalo-salicylic acids, and trihalo-phenols. As has been seen with other DBP classes, increasing bromide levels shifted DBP species to more fully brominated ones but the reactivity differences between HOBr and HOCl forming these four groups of new DBPs were larger than those in reactions forming regulated THMs and HAAs. These aromatic DBPs were suggested as possibly being important intermediate DBPs during chlorination. Ultrahigh resolution with FTICR-MS was used in a couple of new studies to investigate DBPs formed in drinking water treatment. For example, Lavonen et al. used FTICR-MS to identify molecular formulas of new chlorinated DBPs after chlorination and chloramination in Swedish drinking water treatment plants.144 In all, 499 DBPs were detected, of which 230 have not been reported earlier. As a group, the DBPs had significantly lower hydrogen-to-carbon and higher average carbon oxidation states, double bond equivalents per carbon, and oxygen-to-carbon ratios compared to Cl-containing components present before disinfection and CHO formulas in samples taken both before and after disinfection. Type of disinfectant, dose, and predisinfection treatment resulted in distinct patterns in DBP formulas. Zhang et al. also used FTICR-MS to investigate molecular composition of NOM in a source water from China and DBPs formed upon chlorination in a controlled laboratory study.145 In the source water, >4000 NOM components were resolved and 659 one-chlorine containing DBPs and 348 two-chlorine containing products were found in the chlorinated water. Of the 1007 chlorinecontaining DBPs found, only 7 molecular formulas were reported in previous studies. N-DBP Occurrence. Several new studies have investigated the occurrence of emerging N-DBPs. The most extensive occurrence of nitrosamines to-date was reported by Boyd et al., who used a recently developed LC/ESI-MS/MS method to measure 9 nitrosamines in 38 drinking water treatment plants in the United States and Canada.146 Most of the systems were selected based on their use of chloramination, but a few using chlorine or ozone were also included. NDMA, NDPhA, and NMor were detected, with NDMA the most common, being found in finished water from chloramination and chlorination treatment plants and in the distribution system of a plant that used ozone followed by chlorine. Significantly higher levels of NDMA (5−6×) were found in the distribution systems than in the treatment plants, with a maximum level of 130 ng/L. NDPhA was the second-most detected nitrosamine in these drinking waters, and because it is thermally labile, it is seldom measured because most nitrosamine measurements are done with GC/MS (including EPA Method 521), which decomposes it. NDPhA is generally considered less toxicologically important than other nitrosamines because it is a very weak mutagen, but it can induce bladder cancer in rats. The authors also confirmed 2831



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abundant, ranging up to 20.5 and 16.3 ng/L, respectively, in the finished water. Controlled lab experiments revealed that dimethylamine, diethylamine, morpholine, and di-n-butylamine can form the corresponding nitrosamines, with diethylamine and morpholine showing significantly higher yields than dimethylamine. Another occurrence study in China was conducted by Wang et al., who also used a UPLC/MS/MS method to measure nitrosamines in finished drinking water and in distribution systems.152 NDMA was found at all locations sampled, at a maximum of 25.7 ng/L, with higher levels in distribution systems. Iodo-DBPs. As mentioned earlier, iodo-DBPs are among the most genotoxic DBPs measured to-date, and IAA was recently shown to be tumorigenic. As such, research is increasing on this important DBP class. Pan and Zhang published a new method to measure total organic iodine (TOI) using UPLC/ESI-MS for off-line iodide separation and detection.153 This method was proposed as an alternative to the pyrolysis-IC method that has been used to measure TOCl, TOBr, and TOI. One obstacle to overcome with this method was the series of adducts with the same m/z value as iodide (127) formed in the mobile phase and ESI sample cone. By acidifying the absorption solution with formic acid and optimizing instrumental parameters, this was accomplished, and it was more sensitive, accurate, and rapid than the previous pyrolysis-IC method, with 2.5 μg/L quantification limits for a 80 mL water sample. Jones et al. investigated the formation and speciation of iodoTHMs in natural waters treated with preformed NH2Cl vs prechlorination followed by ammonia addition.154 A representative bromide−iodide ratio of 10:1 was used, along with four bromide−iodide levels. Results showed iodo-THM formation was generally lower for prechlorination vs treatment with preformed NH2Cl, due to the oxidation of iodide to iodate by chlorine. However, prechlorination sometimes formed higher iodo-THM levels due to increased brominated I-THMs. The formation of iodoform, IAA, and triiodoacetic acid (TIAA) from chlorine dioxide was investigated by another study by Ye et al.155 Iodoform was the major species observed during ClO2 treatment of NOM in the presence of iodide, and I-DBPs were greater at pH 8. In source waters spiked with iodide, iodoform and IAA formation were strongly correlated to ClO2 dose and water quality. Controlled lab reactions with 18 model compounds revealed considerable formation of I-DBPs, especially for propanoic acid, butanoic acid, resorcinol, hydroquinone, alanine, glutamic acid, phenylalanine, and serine. Finally, in a study of iodide- and bromide-containing waters from Australia, Allard et al. demonstrated that ozone pretreatment at lower pH could be used to selectively oxidize iodide to iodate and minimize I-DBP and bromate formation.156 Ozone was also able to oxidize I-THMs, such that it could be effective in removing any I-THMs already present in the water. New Methods. New low-level detection methods continue to be developed for emerging DBPs. McDonald et al. developed a new isotope dilution-GC/electron ionization (EI)-MS/MS method for measuring eight nitrosamines in drinking water.157 This method utilized direct isotope analogues for all analytes, which ensured accurate quantification, accounting for analytical variability that can occur during sample processing, extraction, and instrumental analysis. Method detection limits of 0.4−4 ng/L were achieved, and the analysis could be performed in 14 min. Ripolles et al. created a new LC/APCI-MS/MS for measuring eight nitrosamines in water.158 This method could achieve detection limits

diphenylamine as a precursor to the formation of NDPhA during chloramination. Jurado-Sanchez et al. published a new study of the occurrence of aromatic amines and nitrosamines in different steps of a drinking water treatment plant.147 In total, 24 amines (including aniline, 2-nitroaniline, and several mono- and dichloroanilines) and 6 nitrosamines were measured using a SPE-GC/MS method. Their concentrations significantly increased (10×) after chloramination, with the formation of 9 amines (4 aromatic and 5 nitrosamines). Interestingly, amines were higher in winter due to low water temperatures, and aromatic amines were detected at their highest concentrations in treated water after rainfall events. Formation and speciation of nine haloacetamides was the focus of another study by Chu et al. SPE (with Oasis HLB), and LC/MS/MS was used for their measurement in source waters that were treated with chlorine or chloramines in a controlled laboratory study.148 These haloamides were only very recently reported in a U.S. Nationwide Occurrence Study, so there has been very little data on their occurrence and formation. Results revealed the highest formation with source waters that had highest dissolved organic nitrogen (DON) from anthropogenic inputs. In addition, higher levels formed during chloramination in low-specific UV absorbance (SUVA) waters with no bromide relative to highSUVA waters with bromide. Dihalo species dominated. Huang et al. published a fascinating study revealing that dichloroacetonitrile (DCAN) and dichloroacetamide (DCAcAm) formed independently during chlorination and chloramination of drinking water.149 Previously, it had been assumed that haloacetamides would be formed by hydrolysis of the corresponding haloacetonitriles, but this study clearly shows that haloacetamides can also form by other pathways. DCAN and DCAcAm formation was compared across a range of NOM isolates and model precursors, and chloramination nearly always formed more DCAcAm than DCAN. In addition, experiments with asparagine as a model precursor suggested DCAcAm formation without a DCAN intermediate, and 15Nlabeled NH2Cl indicated initial rapid formation of both DCAN and DCAcAm by pathways where the nitrogen originated from organic nitrogen precursors. While wastewater effluents and algal substances were more potent precursors for DCAN during chlorination, humic materials were more potent precursors for DCAcAm during chlorination and for both DCAN and DCAcAm during chloramination. Halonitromethanes and haloacetonitriles were the subject of another investigation by Yang et al., who investigated precursors and nitrogen origin of trichloronitromethane (TCNM) and DCAN.150 In this study 31 nitrogen-containing compounds (including amino acids, amines, dipeptides, purines, pyrimidones, and pyrroles) were reacted with chlorine or monochloramine. In general, TCNM and DCAN formation was higher by chlorination than chloramination, and all nitrogen precursors generated TCNM with either chlorine or chloramines. This study was also the first to confirm NH2Cl being a source of nitrogen for TCNM formation, through the use of 15N-labeled NH2Cl. Wang et al. investigated secondary amines as nitrosamine precursors and measured the occurrence of nitrosamines in drinking water from China.151 UPLC/ESI-MS/MS was used to measure the formation of 9 nitrosamines in 12 drinking water treatment plants. Method detection limits ranged from 0.1 to 0.7 ng/L in finished drinking water. All 9 nitrosamines except NDPA were detected, with NDMA and NDEA the most 2832



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used to determine 8 HBQs in 10 swimming pools (detection limits of 0.03−1.2 ng/L).165 2,6-Dichlorobenzoquinone was found in all pools at concentrations ranging from 19 to 299 ng/ L, which is as much as 100× the levels found previously in drinking water. 2,3,6-Trichloro-1,4-benzoquinone, 2,3-dibromo-5,6-dimethyl-1,4-benzoquinone, and 2,6-dibromo-1,4-benzoquinone were also found in some pools, up to 11.2 ng/L, and were notably not found in the input tap water. In addition, the authors had an important finding that ingredients found in personal care products like lotions and sunscreens contain precursors that can form some of these HBQs. This served to explain why some of these HBQs were found in swimming pools but not in tap water. Xiao et al. used UPLC/MS/MS with precursor ion scan-ESI to identify new halogenated DBPs in swimming pool water and investigate their permeability across skin.166 The new pool DBPs were mainly halophenols and halonitrophenols, resulting from chlorination of human body substances, such as urine, in the presence of bromide. Pure standards were available for 2,4dibromophenol, 2,4-dichlorophenol, and 2-bromophenol, such that their permeability values across human skin could be measured. These DBPs were shown to permeate the skin but with less permeability compared to chloroform. However, something very important noted by the authors is that the permeated DBPs increased in an accelerated way with exposure time. In addition, the effect of chlorine on human skin was also investigated, where many DBPs were found to be generated by the interaction of chlorine with the epidermis. It was recommended that swimmers take showers frequently to remove DBPs superficially adsorbed onto the skin and prevent them from deeper penetration. Another fascinating pool study was conducted by Hansen et al., who investigated the formation of DBPs from particles in swimming pool filters.167 The particles (composed mainly of hair and skin cells) were collected from a hot tub with a drum microfilter and were treated with chlorine in the laboratory at different pHs. GC/MS was used to measure THMs, HAAs, HANs, trichloronitromethane, dichloropropanone, and trichloropropanone. Trichloramine was also measured. The formation of DBPs from these human particles was found to be higher than previously reported for a body fluid analogue and filling water. Further, genotoxicity and cytotoxicity measurements showed that toxicity increased with decreasing pH and that HANs were responsible for the majority of the toxicity from measured DBPs. In an extensive study, Simard et al. investigated the variability of chlorination DBP occurrence in 54 indoor and outdoor swimming pools.168 Occurrence of THMs, HAAs, and inorganic chloramines was measured over a period of 1 year (monthly/biweekly sampling). Results revealed much higher levels of DBPs in pool water than drinking water, especially for the HAAs, which reached a maximum of 2224 μg/L in outdoor pools and 1195 μg/L in indoor pools. Lower water renewal in pools was suggested to promote the accumulation of nonvolatile HAAs over time. DBP levels were higher in outdoor pools, and levels increased for heated outdoor pools. Finally, Parinet et al. reported the first measurement of brominated DBPs in seawater pools treated with chlorine.169 In this study, 8 chlorinated pools in France fed with seawater were sampled for THMs and HAAs using GC/MS. Bromoform and dibromoacetic acid were the dominant DBPs detected, with levels up to 18-fold greater than MCLs of 60 and 80 μg/L in drinking water. Statistical analyses showed that brominated

of 1−8 ng/L, and two multiple reaction monitoring (MRM) transitions per compound were used to provide enhanced confidence in the assignments. The idea behind using APCI instead of ESI that that matrix effects tend to be lower than with ESI. Halonitriles were the focus of another new method by Kristiana et al., who used headspace SPME and GC/MS to quantify a series of halonitriles, including halopropionitriles and halobutyronitriles, for which previous methods did not exist.159 Detection limits of 0.9−80 ng/L were achieved, with a linear range over 3 orders of magnitude. This method enabled the first study of their levels in drinking water and revealed that while both chlorine and chloramines could form these halonitriles, 2,2-dichloropropionitrile was formed only in chloraminated drinking water. Allard et al. developed a new SPME-GC/MS method with a programmed temperature vaporizer inlet to measure 10 THMs (chloro-, bromo-, and iodo-THMs) at ng/L in water.160 Limits of detection ranged from 1 to 20 ng/L for iodoform and chloroform, respectively. The suitability of this method was demonstrated for treated groundwater, surface water, seawater, and secondary wastewater, and it was used to show the presence of I-THMs at ng/L at various process stages of an advanced water recycling plant in Australia. Chu et al. created a new method for quantifying 13 haloacetamides in drinking water using SPE extraction with Oasis HLB and LC/APCI-MS/MS.161 Detection limits ranged from 7.6 to 19.7 ng/L, and this method was used to provide the first report of tribromoacetamide and chloroiodoacetamide as DBPs in drinking water. LC/ESI-MS/MS with large volume direct aqueous injection was used by Li et al. in a new method to measure iodo-acid DBPs in drinking water.162 The use of a divert valve technique for the mobile phase solvent delay along with isotopically labeled analogues as internal standards reduced and compensated for the ionization suppression caused by coeluting inorganic anions. Very low detection limits of 0.5−1.9 ng/L were achieved. Li et al. developed another method using GC/ MS/MS to measure 10 HAAs (nine chloro-bromo HAAs and IAA) in drinking water and wastewater.163 Limits of detection ranged from 0.012 to 0.079 μg/L. Electron ionization at 60 eV (rather than the standard 70 eV) was used with 2 MRM transitions for MS/MS. Finally, an important new advance was reported by Li et al. in the traditional pyrolysis-IC TOBr method.164 Previous studies have shown that polar brominated DBPs can penetrate activated carbon (AC) that is used to extract the DBPs from water prior to pyrolysis and quantification by IC. This occurred by reduction by the AC of some of the TOBr to bromide. The authors demonstrated that treating the AC column with ozone (by passing an aqueous ozone solution through it) improved the absorptive capacity of the AC, inhibiting the reductive property of the AC and minimizing penetration of polar brominated DBPs during TOBr analysis. New Swimming Pool Research. Swimming pools are being recognized as an important source of exposure to DBPs. Health concerns include increased risk of bladder cancer from exposure to indoor pools and increased risk of asthma for indoor and outdoor pools.2 Swimming pools continue to be an intense area of study. The recent discovery of HBQs in drinking water was mentioned earlier in this review, and now Li’s group has investigated their formation in swimming pools. In this recent study by Wang et al., a new LC/MS/MS method was 2833



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bromoform, along with high levels of bromate, both of which could be reduced by activated carbon treatment. UV only occasionally produced DBPs, including THMs, HAAs, and nitrite. Delacroix investigated DBP formation and ecotoxicity of ballast water after treatment with chlorine.174 Acute and chronic effects on algae were observed in 25% and >50% of the treated waters, respectively. In total, 22 DBPs were identified, and 4 of these were found at levels that may pose a risk to the local aquatic ecosystem. Bromoform, dibromoacetic acid, dibromochloromethane, dibromomethane, bromodichloromethane, bromochloroacetic acid, bromate, and chlorate were the DBPs found most often, with a maximum concentration of 670 μg/L for bromoform. DBPs of Pollutants. Studies of DBPs have gone beyond the “classic” DBPs formed by the reaction of NOM with disinfectants, such that reactions of environmental pollutants with disinfectants are increasingly being studied. Contaminant DBPs have been recently reported from pesticides, pharmaceuticals, antibacterials, personal care products, and musks. Some of this research has been conducted in order to find ways to degrade and remove these contaminants from wastewater effluents and drinking water sources, but some of this research is being conducted to determine the fate of these contaminants in drinking water treatment. It is not surprising that DBPs can form from these contaminants, as many of them have activated aromatic rings or other structural groups that can readily react with oxidants like chlorine and ozone. However, until recently, these types of DBPs were not investigated. Duirk et al. discovered the formation of highly toxic iodoDBPs (iodo-acids and iodo-THMs) when iodinated X-ray contrast media are treated with chlorine or chloramines.175 This discovery came about from a curiosity in a previous occurrence study, which showed the presence of ppb levels of iodo-DBPs in finished water but nondetectable or barely detectable iodide in the source waters. As a result, other sources of iodine were considered and X-ray contrast media were subsequently found at high levels in these drinking water reservoirs, up to 2700 ng/ L. LC/MS/MS was used to measure them for the first time in drinking water source waters. Of the X-ray contrast media, iopamidol was found the most often and at the highest levels. In subsequent controlled laboratory experiments, it was also the most reactive with chlorine and chloramines, forming the highest levels of iodo-DBPs. NOM also appears to play a role in their formation, with much higher iodo-DBP levels when NOM was present in the treated waters. In addition, toxicity experiments revealed that the reaction mixtures containing iopamidol were much more genotoxic than those without, supporting the formation of these highly genotoxic and cytotoxic DBPs. The role of iodide was also investigated in a study by Vikesland et al. in the halogenation of bisphenol A (BPA), triclosan, and phenols in chlorinated water.176 Over the pH range of 5.5−10, the transformation kinetics of 2,4-dichlorophenol in the presence of 10 μM iodide were 2−15× faster than reactions with free chlorine alone, and for triclosan and BPA, reactions were 3−20× and 230−660× faster. For all test compounds, I-DBPs were rapidly formed. 2,4-Dichlorophenol formed 2,4-dichloro-6-iodophenol; triclosan formed 5-chloro-4iodo-2-(2,4-dichlorophenoxy)phenol, 5-chloro-6-iodo-2-(2,4dichlorophenoxy)phenol, 5-chloro-4,6-diiodo-5−2-(2,4dichlorophenoxy)phenol, and either 4,5-dichloro-6-iodo-2(2,4-dichlorophenoxy)phenol or 5,6-dichloro-4-iodo-2-(2,4-

DBPs resulted mainly from TOC and nitrogen brought by bathers. The seawater pool with the highest DBP levels showed a mean value of 931 μg/L for bromoform and 1089 μg/L for dibromoacetic acid. Formation Studies. Several other interesting formation studies are worthy of note. Wang et al. explored the formation of DBPs from bacteria and from biofilms that can be present in pipes in drinking water distribution systems.170 THMs, HANs, chloral hydrate, chloropicrin, and 1,1,1-trichloropropanone were detected when E. coli was chlorinated or chloraminated. Levels of HANs formed for chlorine and chloramine were similar, but chlorine formed substantially higher levels of THMs (HAAs were not measured in this study). Results indicated that breaking down of bacterial cells released precursors for DBP formation. DBPs also formed when Pseudomonas aeruginosa was cultured with pipe materials (simulating natural biofilms present in pipes), and pipe material played a role, in that biofilms grown on galvanized zinc pipe showed higher DBP formation potential than polyvinyl chloride (PVC) pipe. Several interesting studies investigated DBP formation in salinated waters. Ding et al. explored brominated DBP formation during chlorination of saline sewage effluents.171 In Hong Kong, seawater has been used for decades for toilet flushing, and there is the potential to form toxic brominated DBPs when this salinated wastewater is treated with chlorine, posing potential adverse effects on the marine ecosystem when this water is discharged. The authors used UPLC with precursor ion scan-ESI-MS/MS to identify Br-DBPs formed and also a TOBr/TOCl assay to account for the total brominated and chlorinated organic byproducts formed. A total of 54 major polar Br-DBPs were detected in the chlorinated saline effluents, and 6 were newly reported wastewater DBPs: bromomaleic acid, 5-bromosalicylic acid, 3,5-dibromo-4-hydroxybenzaldehyde, 3,5-dibromo-4-hydroxybenzoic acid, 2,6-dibromo-4-nitrophenol, and 2,4,6-tribromphenol. In the secondary wastewater effluent, polar Br-DBPs formed and reached their maximum levels at different chlorine doses, whereas for the primary effluent, the formation of polar Br-DBPs kept increasing with chlorine dose. However, the primary effluent generated fewer and less Br-DBPs and rarely generated nitrogenous Br-DBPs. Seawater and saltwater marine aquaria were the focus for a study by Shi et al., who used IC/ICPMS and GC/MS to quantify DBPs formed by chlorine and ozone.172 Results revealed high levels (>100 μg/L) of some HAAs and halonitromethanes with brominated species dominating in higher bromide waters. Interestingly, concentrations of bromate and iodate were strongly impacted by factors other than bromide and iodide, including whether the system was open or closed. Ship ballast water treatment was the focus of new studies. These studies were triggered by new international regulations that will soon require reductions in the number of organisms that ships discharge in ballast water. This regulation is intended to reduce the number of invasive species released by ship ballast water. Werschkun et al. reported DBP data from three different types of ballast treatment: chlorine, ozone, and UV.173 Chlorination systems generated THMs, HAAs, and bromate in much larger quantities than for other applications (e.g., drinking water). Interestingly, levels were higher from brackish waters than seawaters. More toxic brominated species predominated in all treated waters. Ozonation produced 2834



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monochloramine, ozone, and chlorine dioxide.180 Results showed that these pesticides can serve as nitrosamine precursors by release of secondary amines through hydrolysis or through reactions with the oxidants. All four oxidants formed nitrosamines, with the highest levels formed by monochloramine and ozone. Kinetic results showed that the dithiocarbamates were not direct precursors, but nitrosamines were formed primarily through reaction of the oxidant with an amine that can be generated through hydrolysis or oxidation. Finally, Kuhlich et al. identified DBPs formed by the chlorination of musks.181 The musk, 6-acetyl-1,1,2,4,4,7hexamethyltetraline (AHTN), produced a carboxylic acid product (AHTN-COOH), which led to two new chlorinated DBPs, including a monochlorinated carboxylic acid and a monochlorinated decarboxylated product. 1,3,4,6,7,8-Hexhydro-4,6,6,7,8,8-hexamethylcyclopenta-γ-2-benzopyran (HHCB) produced HHCB-lactone. The reaction products and intermediates were synthesized and isolated, and reaction mechanisms were proposed.

dichlorophenoxy)phenol; and BPA formed 2-iodo-BPA and either 2-chloro-2′-iodo-BPA or 2-chloro-6-iodo-BPA. Bulloch et al. reported a beautiful study of gemfibrozil wastewater chlorination products using a combination of chemistry and toxicology.177 LC/MS/MS and NMR were used together to identify the DBPs formed, and standards of these DBPs, chlorogemfibrozil and bromogemfibrozil were synthesized, isolated, and characterized. Mass spectrometry was used to follow the in situ halogenation reactions of gemfibrozil and to measure levels of chlorogemfibrozil and bromogemfibrozil formed in chlorinated wastewater effluent. Toxicology results revealed that halogenated DBPs of gemfibrozil enhanced the antiandrogenicity in medaka fish, demonstrating that chlorination may increase the toxicity of pharmaceutically active compounds in surface water. As mentioned earlier in the section on Pharmaceuticals and Hormones, Huerta-Fontela et al. identified new chlorinated DBPs from the illicit drugs 3,4-methylenedioxyamphetamine (MDA) and 3,4-methylenedioxyethylmethamphetamine (MDMA).86 3-Chlorobenzo-2,3-dioxole was found for both MDA and MDEA, and 3-chlorocatechol was formed by MDMA. The formation of these DBPs was followed through the course of simulated and real drinking water treatment. Following its initial formation with chlorine, 3-chlorobenzo-2,3dioxole was eliminated with subsequent ozone and GAC treatment, but the other MDEA DBP was stable upon further treatment and was found in finished drinking water at 0.5−5.8 ng/L. Roberts’ group at Johns Hopkins University recently reported the discovery that Cl2O, a generally overlooked oxidant species, can be important in reactions of chlorine with contaminants like dimethenamid that are less nucleophilic than other aromatic compounds like phenols.1 In a follow-up study involving bromination reactions, Sivey et al. reported the impact of corresponding brominated oxidants (Br2O, BrCl, and Br2) in reactions with dimethenamid.178 In most previous work, HOBr has been assumed to be the active brominating species, but these new findings confirmed the importance of these other brominating species. In particular, bromination rates increased with increasing Cl−, free available chlorine (at constant [HOBr], and excess bromide. Intrinsic brominating reactivity was HOBr ≪ Br2O < BrOCl ≈ Br2 < BrCl for reactions of dimethenamid with chlorine in the presence of bromide. Reactions resulted in the formation of bromodimethenamid as a DBP. Chusaksri et al. investigated chlorine reactions with phenylurea herbicides.179 Chlorination products were monitored by LC/MS/MS, and results revealed reactions taking place at both N atoms and at the ortho- and para-carbons of the aromatic ring. The main chlorinating species were different at different pH. The formation of NDMA from pollutants has been the focus of several studies, including one previously mentioned in the Pharmaceuticals and Hormones section in which Shen and Andrews investigated NDMA formation from eight aminebased pharmaceuticals.82 The presence of NOM enhanced the formation of NDMA from most pharmaceuticals, but prolonged prechlorination inhibited its formation, which was hypothesized to be due to the transformation of NOM into smaller products that could form new bonds with pharmaceuticals. Padhye et al. investigated NDMA and NDEA formation from the oxidation of dithiocarbamate pesticides with chlorine,

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SUNSCREENS/UV FILTERS UV filters used in sunscreens, cosmetics, and other personal care products have increased in interest due to their presence in environmental waters and the potential for endocrine disruption and developmental toxicity. A few UV filters have estrogenic effects similar to E2 (a natural estrogen) as well as the potential for developmental toxicity. Environmental levels of UV filters are not far below the doses that cause toxic effects in animals. There are two types of UV filters, organic UV filters, which work by absorbing UV light, and inorganic UV filters (e.g., TiO2, ZnO), which work by reflecting and scattering UV light. Organic UV filters are increasingly used in personal care products, such as sunscreens, cosmetics, beauty creams, skin lotions, lipsticks, hair sprays, hair dyes, and shampoos. Examples include benzophenone-3 (BP-3), octyl-dimethyl-paminobenzoic acid (ODPABA), 4-methylbenzylidene camphor (4-MBC), ethylhexyl methoxycinnamate (EHMC), octocrylene (OC), iso-amylmethoxycinnamate (IAMC), and phenylbenzimidazole sulfonic acid (PBSA). The majority of these are lipophilic compounds (low water solubility) with conjugated aromatic systems that absorb UV light in the wavelength range of 280−315 nm (UVB) and/or 315−400 nm (UVA). Most sunscreen products contain several UV filters, often in combination with inorganic micropigments. Because of their use in a wide variety of personal care products, these compounds can enter the aquatic environment indirectly from bathing or washing clothes, via wastewater treatment plants, and directly from recreational activities, such as swimming and sunbathing in lakes and rivers. A nice review article was recently published by Gago-Ferrero et al., which summarized LC/MS/MS methods for organic UV filters in the environment.182 UV filters included benzophenones, camphor derivatives, cinnamates, crylenes, benzimidazole derivatives, p-aminobenzoic acid and derivatives, dibenzoyl methane derivatives, salicylates, and triazines, as well as their primary transformation products. In addition, their occurrence in river water, seawater, raw water, reclaimed water, sludge, river sediment, and biota is summarized. New methods have been reported, including ones based on LC−UV, CE/MS, and LC/MS/MS. As mentioned earlier, a recent trend is the use of ionic liquids in analytical methods. For example, Ge and Lee published a new ionic liquid-hollow fiber-liquid phase microextraction-LC−UV method for measur2835



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ing 4 UV filters (BP), 3-(4-methylbenzylidene)-camphor, 2hydroxy-4-methoxybenzophenone, and 2,4-dihydroxybenzophenone in water.183 Of the ionic liquids investigated, 1hexyl-3-methylimidazolium t ris(pentafluoroet hyl)trifluorophosphate [HMIM][FAP] was the best extraction solvent and this method allowed detection limits of 0.3−0.5 μg/L. This procedure involved the injection of a small amount of ionic liquid into a hollow fiber, followed by immersal of the hollow fiber into the sample for extraction, and injection of the extracted solution directly into the LC−UV instrument. Maijo et al. reported a new SPE-inline-CE/ESI-MS method for measuring 4 UV filters (BP-3, 2,2-dihydroxy-4-methoxybenzophenone, 2,4-dihydroxybenzophenone, and 2-phenylbenzimidazole-5-sulfonic acid) in water.184 Oasis MCX was the best SPE sorbent for these analytes, and detection limits of 0.01− 0.05 μg/L were achieved, which are much lower than previous CE/ESI-MS values without inline SPE. This method was successfully demonstrated with the analysis of river waters. Bratkovics and Sapozhnikova reported a new LC/MS/MS method to measure seven common UV filters in fresh water and seawater.185 Oasis HLB was used for SPE, and method reporting limits of 0.5−25 ng/L were achieved. This method was then used for the first measurement of sunscreen compounds in coastal waters from the United States. In these measurements conducted off the coast of South Carolina, concentrations of 10−2013 ng/L were found, with oxybenzone and octocrylene found at the highest levels. Fate studies continue to be conducted for UV filters. For example, Magi et al. reported the study of six UV filters in wastewater treatment in Italy.186 Stir bar sorptive extraction with liquid desorption and LC/MS/MS was used for their measurement, along with triggered MRM, a data-dependent acquisition approach in which the secondary MRM transitions are triggered when the primary transitions (target and qualifier ions) exceed a user-defined threshold. The triggered MRM enabled the generation of product ion spectra used to create a reference MS library. Results showed that four of the UV filters were removed by wastewater treatment; only BP-3 and OC were detected in effluent samples, at lower levels than in the influents. Transformation products have been recently identified for UV filters, including 9 TPs identified from the biotransformation of sulisobenzone (BP-4) in activated sludge. In this study by Beel et al., accurate mass measurements using LTQOrbitrap-MS was used to identify the TPs, along with MSn experiments, and NMR spectroscopy.187 The main TP was synthesized and confirmed. Initial reactions involved reduction of the ketone group to a benzhydrol group, and further reactions resulting in the formation of highly polar TPs. A biodegradation pathway was proposed, and toxicity experiments revealed that the TPs had higher toxicity in Vibrio f ischeri than the parent compound, BP-4. Photolysis of the UV filter octyl methoxycinnamate (OMC) was investigated by MacManus-Spencer et al., who discovered new, previously unreported TPs, including 4-methoxybenzaldehyde, 2-ethylhexanol, cyclodimers, and a dimer hydrolysis product.188 Reactions in simulated and natural sunlight included photoisomerization and subsequent transformation, with estimated half-lives of 1 h for both cis- and trans-OMC.

Review

BROMINATED AND EMERGING FLAME RETARDANTS Brominated flame retardants have been used for many years in a variety of commercial products including children’s sleepwear, foam cushions in chairs, computers, plastics, and electronics. Brominated flame retardants work by releasing bromine free radicals when heated, and these free radicals scavenge other free radicals that are part of the flame propagation process. The use of these flame retardants is believed to have successfully reduced fire-related deaths, injuries, and property damage. However, there is concern because of their widespread presence in the environment and in human and wildlife samples as well as their presence in locations far from where they were produced or used. Polybrominated diphenyl ethers (PBDEs) have been a popular ingredient in flame retardants since the polybrominated biphenyls were banned about 30 years ago. They are environmentally persistent and lipophilic and bioaccumulate in animals and humans. PBDEs are made up of 209 possible congeners containing between 1 and 10 bromine atoms and, of these, 23 congeners are of environmental significance. In recent years, PBDE levels have been increasing significantly. The greatest health concern comes from recent reports of developmental neurotoxicity in mice, but there are also concerns regarding the potential for hormonal disruption and, in some cases, cancer. In 2004, the European Union banned the use of the penta- and octa-BDEs and later, in 2008, banned deca-BDEs. In 2004, the major U.S. manufacturer of PBDE-based flame retardants (Great Lakes Chemical) voluntarily stopped producing the penta- and octa-BDEs. Earlier studies had suggested that deca-BDE was too large to bioaccumulate and would not be a risk to humans. However, research now shows that it can accumulate in animal tissues (including people) and that it can debrominate in the environment and metabolically to form the lower brominated species (including the octa- and penta-BDEs). Several U.S. states banned the penta- and octaBDEs in 2006, and in December 2009, two U.S. producers of deca-BDE agreed to voluntarily phase it out in the United States (www.epa.gov/oppt/existingchemicals/pubs/ actionplans/deccadbe.html). In addition, the U.S. EPA issued a new rule in 2006 to complement the phase-out of the octaand penta-BDEs, ensuring that no new manufacture or importation of these chemicals can occur without first being subject to U.S. EPA evaluation (www.epa.gov/EPA-TOX/ 2006/June/Day-13/t9207.htm). However, despite the halt in manufacture of most of these PBDEs in North America and Europe, they are still present in many consumer products sold previously and can be released into the environment during use and disposal. In addition, there is still the possibility of importing products that contain them. Four of the PBDEs (2,2′,4,4′-tetra-BDE (BDE-47), 2,2′,4,4′,5-penta-BDE (BDE99), 2,2′,4,4′,5,5′-hexa-BDE (BDE-153), and 2,2′,4,4′,6-pentaBDE (BDE-100)) and another brominated flame retardant (2,2′,4,4′,5,5′-hexabromobiphenyl (HBB)) were on the UCMR-2 in the U.S., and national occurrence data are available (http://water.epa.gov/lawsregs/rulesregs/sdwa/ucmr/data. cfm#ucmr2010). Along with the widely measured PBDEs, new “emerging” flame retardants are now being found in the environment. This is partly due to shifting of manufacturers and industries to production and use of different flame retardants (including bromophenols, bromophenyl ethers, brominated phenyl esters, 2836



Review

Several interesting new studies have been published for brominated and emerging flame retardants. Qu et al. used bioassay-directed fractionation and LC/APCI-MS/MS to identify TBBPA diallyl ether as an emerging neurotoxicant in environmental samples from China.193 Primary cultured cerebellar granule neurons viability was used to evaluate the neurotoxicity of separated fractions of river water, sediment, and soil. Results revealed TBBPA diallyl ether as the key neurotoxicant, and a manufacturing plant was identified as the source. Effect-directed analysis with a multiassay screening approach was used by Metcalfe et al. to identify PBDE congeners (BDE47 and BDE-99), a PBDE metabolite, and triclosan as the causative EDCs in treated wastewater.194 A positive correlation was found between the degree of response in the T4-hTTR assay and amounts of these brominated flame retardants, BFR metabolite, and triclosan. LC/MS/MS and GC/MS were used in combination with in vitro testing to identify these causative agents. In another study, Salamova and Hites investigated the occurrence of discontinued and alternative BFRs in the atmosphere and precipitation from the Great Lakes Basin.195 Of the BFRs measured (PBDEs, decabromodiphenylethane [DBDPE], hexabromobenzene [HBB], pentabromoethylbenzene [PBEB], and 1,2-bis(2,4,6-tribromophenoxy)ethane [BTBPE]), highest levels were generally found in urban sites (e.g., Chicago and Cleveland), with a few exceptions. In precipitation, PBDEs were found at higher levels than other BFRs, with a maximum of 42 ng/L in Chicago. Overall halfing times for the BFRs were determined, which demonstrated that PBDEs, HBB, and BTBPE have declined over time. However, levels of PBEB and DBDPE have not changed between 2005 and 2009. Nyholm et al. analyzed emerging BFRs and PBDEs in seepage water, sewage wastewater, sewage sludge, and sediments near suspected sources in Norway.196 TBECH, BTBPE, DBDPE, ethylene bis(tetrabromophthalimide) [EBTP], TBBPA allyl ether (AE), and TBBPA dipropylether (DBPE). An LC/APPI-MS/MS method was developed for the analysis of several of these BFRs. PBDE levels were generally found at higher levels than the emerging BFRs, but several emerging ones were found, with TBBPA DBPE found at the highest levels (81 ng/L) in seepage water from a combined metal recycling and car dismantling site. New methods developed for flame retardants include one by Feo et al., who used LC/ESI-MS/MS to measure 11 hydroxylated PBDEs.197 Hydroxylated PBDEs are formed by metabolism of PBDEs and also from natural sources in marine ecosystems. Some are more neurotoxic than their parent compounds. For this method, two MRM transitions were used and detection limits ranged from 0.17 to 0.72 pg injected. Chen et al. developed a sensitive method using etched stainless steel wire with SPME-GC/negative chemical ionization (NCI)-MS to measure PBDEs in environmental waters.198 The PBDEs exhibited high adsorption affinity toward the etched stainless steel fiber due to electron donation and back-donation between the metal and PBDEs. Low detection limits of 0.2−0.6 ng/L were achieved. Finally, Sanchez-Prado et al. reported a first-time laboratory investigation into the photochemical degradation of 2,2′,4,4′,6pentabromodiphenyl ether (BDE-100) in solid ice samples.199 Eight photolysis products were identified, which indicated stepwise reductive debromination and intramolecular elimi-

other bromoaromatic compounds, and brominated and chlorinated cyclic aliphatic compounds). Global production of brominated flame retardants is extremely high, estimated at 100 000 to 180 000 tons per year.189 Some of these “emerging” flame retardants have been used for many years but were only recently discovered in the environment. For example, Dechlorane Plus, which is used in products worldwide, has been manufactured for more than 40 years but was only recently found in the environment and in biota. Many reviews have been published on brominated and “emerging” flame retardants. These include an excellent review by Papachlimitzou et al., who summarized analytical methods for determining novel, emerging brominated flame retardants in environmental samples.189 The authors provide chemical structures for all 27 of these novel flame retardants and conclude with a list of recommendations for improving methods. Guerra reviewed LC/MS techniques for measuring selected halogenated flame retardants, including PBDEs, tetrabromobisphenol A (TBBPA), hexabromocyclodecane (HBCD), tetrabromoethylcyclohexane (TBECH), and Dechlorane Plus (DP), as well as their metabolites and TPs.190 Van der Veen and de Boer published a review on phosphorus flame retardants (PFRs), providing an overview of their properties, production, environmental occurrence, toxicity, and analytical methods.191 PFRs are often proposed as alternatives to brominated flame retardants and are classified into three main groups: inorganic PFRs, organic PFRs (e.g., organophosphate esters, phosphonates, and phosphinates), and halogenated PFRs (e.g., tris(chloropropyl)phosphate [TCPP] and tris(2-chloroethyl)phosphate [TCEP]). PFR concentrations reported are often higher than for PBDEs, and human exposure through indoor air appears to be higher. Of the PFRs, only the chlorinated ones are carcinogenic. One particularly groundbreaking study was a new discovery of coleaching of decabromodiphenylether (BDE-209) with antimony from PET plastic bottles into bottled water.192 BDE209 is used as a flame retardant in the preparation of polyethylene terephthalate (PET) and polycarbonate (PC) plastics, and Andra et al. just published the first report of PBDEs in bottled water. In this study, bottled water was sampled from Boston, MA grocery stores. The 31 bottled water samples included noncarbonated (e.g., spring, mineral water), carbonated (which contained CO2 and often lime and lemon), and noncarbonated-enriched water (e.g., enriched with various flavors, vitamins, nutrients, minerals, and/or electrolytes). Plastic bottles included PET, PC, high-density polyethylene (HDPE), and polystyrene (PS). Storage conditions were evaluated over time in the light and dark, up to 60 days. ICPMS was used to measure bromine and antimony, and GC/ MS confirmed the presence of BDE-209 in PET bottled water. At the time of purchase (Day 1), total soluble Br concentrations averaged 12, 14, and 36 μg/L Br for noncarbonated, carbonated, and noncarbonated-enriched water samples, respectively. Upon storage up to 60 days, levels of Br and antimony substantially increased in PET bottled water, with increases of Br noted in all types of bottles (PET, HDPE, PS, and PC) but especially for PET bottles. Leaching of antimony and Br were greatest for carbonated and noncarbonated-enriched water samples, suggesting that the presence of dissolved CO2 and the additions of flavor and color may enhance leaching. Further, a highly significant correlation was found between soluble antimony and Br concentrations, suggesting similar leaching behavior. 2837



Review

benzotriazoles were found in all wastewater samples and two seawater samples from the coast of Spain. Loi et al. developed a SPE-LC/MS/MS method for measuring seven benzotriazoles and seven benzothiazoles in wastewater.204 This method was used for the first measurements of these compounds in recycled water from Australia. Detection limits ranged from 2 to 322 ng/ L in secondary wastewater. Results showed that some compounds were incompletely removed by reverse osmosis (RO) treatment, such that 974 and 416 ng/L concentrations were detected for benzotriazoles and benzothiazoles in the post-RO water samples.

nation of HBr. Taking advantage of the high preconcentration factor obtained with SPME at low temperatures, a SPME fiber cooled to 0 °C was used as the photoreaction support for BDE100, allowing identification of a greater number of photolysis products. While this study was conducted in the laboratory, it has implications for potential transformation of PBDEs in cold environments.

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BENZOTRIAZOLES AND BENZOTHIAZOLES Benzotriazoles are complexing agents that are widely used as anticorrosives (e.g., in engine coolants, aircraft deicers, or antifreeze liquids) and for silver protection in dishwashing liquids. The two common forms, benzotriazole (1H-benzotriazole) and tolyltriazole (a mixture of 4- and 5-methyl-1Hbenzotriazole), are soluble in water, resistant to biodegradation, and only partially removed in wastewater treatment. There is new evidence for estrogenic effects in vitro but, so far, not in vivo, in recent fish studies. There is also some evidence that benzotriazole may be a human carcinogen, and Australia now has a drinking water guideline limit of 7 ng/L for tolyltriazole.2 Benzothiazoles are used as anticorrosives and in the manufacture of rubber and other products. Benzothiazoles can be found in rubber materials, herbicides, slimicides, algaecides, fungicides, photosensitizers, azo dyes, drugs, deicing/anti-icing fluids, and food flavors.200 Both benzotriazoles and benzothiazoles are high volume production chemicals. They were recently measured in human urine from a multicountry study in the United States, Greece, Vietnam, Korea, Japan, China, and India.199 Because of their water solubility, LC/MS and LC/MS/MS methods have been recently developed for their measurement in environmental waters. While reports of benzotriazoles are fairly recent (∼last 10 years), studies indicate that they are ubiquitous environmental contaminants. Janna et al. reported an interesting study entitled, “From dishwasher to tap? Xenobiotic substances benzotriazole and tolyltriazole in the environment”.201 This study demonstrated their presence in U.K. wastewaters, rivers, and drinking water and suggested that their use as silver polishing agents in dishwasher tablets and powders may account for a significant proportion of inputs to wastewaters. Benzotriazole and tolyltriazole ranged from 840 to 3605 ng/L and 2685 to 5700 ng/L, respectively, in sewage effluents and from 0.6 to 79.4 ng/L and 80% could be removed in wastewater treatment, with sorption to sludge, biodegradation, and volatilization losses important factors in their fate. The mean total mass of siloxanes entering the WWTP was 15.1 kg/day, and the mean total mass released into the environment from the effluent was 2.67 kg/day. L11, a linear polydimethylsiloxane, along with L10 and D5 were the dominant ones identified in wastewater.

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NAPHTHENIC ACIDS Naphthenic acids (NAs) are a complex mixture of alkylsubstituted acyclic and cycloaliphatic carboxylic acids that dissolve in water at neutral or alkaline pH and have surfactant-like properties. They occur naturally in crude oil deposits across the world (up to 4% by weight) and have also been recently discovered in coal, which could be a source of contamination for groundwater. NAs are toxic to aquatic organisms, including phytoplankton, daphnia, fish, and mammals and are also endocrine disrupting. With decreased conventional crude oil resources, it has become economically feasible to extract heavier 2838



Review

McMurray Formation groundwater vs surface water and shallow groundwater samples, such that this method could be used for source apportionment. The authors cited a need to find suitable standard materials that more accurately reflect the range of compounds found in real NA samples.

oils from oil-sand deposits. Two of the world’s largest accumulation of oil-sands occurs in North and South America. Venezuelan oil-sand deposits contain the largest known petroleum deposits in the world, and the Athabasca oil-sands in Alberta, Canada, are a close second. The Athabasca oil-sands represent more than 25% of Canada’s annual oil production, and most research on NAs has been conducted in this region. Caustic hot water is used in the extraction of crude oil from oilsands, which results in a residual tailing water (0.1−0.2 m3 of tailings per ton of oil-sands processed) that contains high levels of NAs (80−120 mg/L levels are common) and is very toxic. The total volume of tailing ponds is projected to exceed 109 m by the year 2020. Headley et al. published a review of analytical methods for chemical fingerprinting of NAs, which resulted from a workshop held on this subject.207 GC/MS, GC × GC/MS, LC × LC/MS, LC/MS/MS, APPI- and ESI-FT-ICR-MS, and Orbitrap-MS methods were reviewed, as well as newer emerging techniques, including synchronous fluorescence spectroscopy (SFS), X-ray absorption near-edge spectroscopy (XANES), ion-mobility-MS, and isotope ratio-MS. Accurate mass-MS techniques continue to be developed and used to aid in the characterization of these complex NA mixtures. For example, Hindle et al. developed the first ISO17025 accredited method using LC/TOF-MS to measure NAs in technical mixtures and environmental waters.208 LC facilitated a 5-fold reduction in ion suppression as compared to traditional flow injection, and a detection limit of 1 mg/L for total oxy-NAs was achieved. Nyakas used direct infusion-FTICR-MS with and without offline UPLC prefractionation to comprehensively analyze NAs in oil-sands processed water.209 Offline fractionation into eight fractions led to approximately twice as many detected peaks and identified compounds (973 peaks vs 2231 peaks of which 856 and 1734 peaks, respectively could be assigned to chemical formulas based on accurate mass measurements). Naphthenic and oxy-naphthenic acids represented the largest group of molecules with assigned formulas, followed by sulfur-containing compounds and nitrogencontaining compounds. Pereira et al. used LC/Orbitrap-MS to characterize oil-sands process waters, which resulted in the identification of >100 new O2 species with dihydroxy, diketo, or ketohydroxy functional groups.210 LC allowed for separation of various isomers and interference-free MSn experiments. A new derivatization with LC/MS/MS was the basis of another method by Woudneh et al.211 Derivatization with N(3-dimethylaminopropyl)-N′-ethylcarbodiimide allowed a common product ion by positive ion-ESI-MS/MS, which facilitated the quantitative characterization of NAs in environmental waters. A common product ion was formed, regardless of the structure of the NA, resulting in approximately constant relative response factors for various isomers in a given NA chromatographic peak, which allowed quantification using a single standard. This method could be used to distinguish between straight chain fatty acids and NAs and provided up to 3 orders of magnitude greater sensitivity compared to direct NA analysis by negative ion-ESI-MS/MS. Ahad et al. reported a new approach to extract, separate, and characterize high molecular weight organic acids using preparative capillary GC followed by thermal conversion/ elemental analysis-isotope ratio-MS.212 The intramolecular carbon isotope signature was significantly different in deep

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MUSKS Synthetic musk compounds are widely used as fragrance additives in many consumer products, including perfumes, lotions, sunscreens, deodorants, and laundry detergents. They can have nitroaromatic structures, as in the case of musk xylene (1-tert-butyl-3,5-dimethyl-2,4,6-trinitrobenzene) or musk ketone (4-tert-butyl-2,6-dimethyl-3,5-dinitroacetophenone), or polycyclic structures, as in the case of 7-acetyl-1,1,3,4,4,6hexamethyl-1,2,3,4-tetrahydronaphthalene (AHTN; trade name, tonalide) or 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-(g)-2-benzopyran (HHCB; trade name, galaxolide). Because they are widely present in environmental samples, including wildlife and humans, there is growing concern. Musks are highly lipophilic, so they tend to accumulate in sediment, sludge, and biota. Up to 190 ng/g lipid has been reported in humans.2 Several new methods have been recently developed for measuring musks. Lung and Liu reported a new UPLC/APPIMS/MS method for measuring six musks in environmental samples.213 Chromatographic separation could be achieved in 7 min (positive ion mode) and 5.1 min (negative ion mode). Detection limits were below 6 pg, with a linear range of 5−500 ppb. Wang et al. reported a new enantiomeric method for polycyclic musks using GC/MS/MS.214 This method combines a chiral column with a nonchiral HP-5MS column and can effectively resolve all five chiral musks for galaxolide, tonalide, phantolide, traseolide, and cashmeran. Detection limits of 1.01−2.39 ng/L can be achieved in drinking water and surface water. Ionic liquids were used as the stationary phase in GC columns in a method by Sanchez-Prado et al. for musks.215 These ionic liquid GC columns were superior in most cases to the selectivity of a HP5 column. Vallecillos et al. created a new fully automated liquid-based headspace single drop microextraction-GC/MS/MS method for measuring musks in environmental waters.216 Six polycyclic musks, three nitro musks, and one polycyclic musk degradation product could be measured in wastewater at detection limits between 0.01 and 0.03 ng/L. Highest enrichment factors were found with 1 μL of 1-octyl-3-methylimidazolium hexafluorophosphate ionic liquid exposed in the headspace of 10 mL water samples containing NaCl. A molecularly imprinted sorbent was used in another new method by Lopez-Nogueroles et al. to selectively extract nitro musks in environmental waters.217 High enrichment factors between 580 and 827 could be achieved. Several new studies have been conducted for musks, including one by Chase et al. who measured the environmental fate, transport, and transformation of six polycyclic musks and two nitro musks in wastewater, surface waters and their sediments, groundwater, soil cores, and plants from a treated wastewater land application site.218 Musks were detected at effluent-impacted environments at ng/L and ng/g concentrations, which decreased during transport throughout wastewater treatment and land applications. Musks were detected in plant samples at trace levels and unexpectedly at lakes not receiving treatment effluent. 2839



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PESTICIDE TRANSFORMATION PRODUCTS

Review

ANTIMONY

Antimony, which can have both acute and chronic toxicity effects, is regulated in drinking water in the United States, Canada, Europe, and Japan at action levels ranging from 2 to 6 μg/L. Antimony contamination can result from copper or lead smelting, petroleum refineries, or the manufacture of plastics and flame retardants. Antimony can also be present in air particulate matter from automobile emissions (engines and brake linings). Recent studies, however, have shown that it can also leach from polyethylene terephthalate (PET) plastic water bottles, producing the highest levels of human exposure to antimony, up to ∼10 μg/L.1 Antimony trioxide is used as a catalyst in the manufacture of PET plastics, which can contain >100 mg/kg of antimony. This is a concern because of the growing popularity of bottled water. PET bottles have been used for the last 4 decades and have gradually replaced PVC and glass bottles. In Europe, bottled water sales accounted for 44% of the market volume for nonalcoholic drinks in 2009, with an average individual consumption of 105 L per year.225 In the United States, consumption of bottled water is also increasing. New reports from 2012 show an average individual consumption of 30.8 gallons (∼117 L) per year (www. bottledwater.org). Compared to PET bottles, low density polyethylene bottles contain much lower levels (∼1%) of antimony. Bach et al. published a very nice, critical review of chemicals found in PET bottled water and their toxicological assessment.225 Antimony, formaldehyde, and acetaldehyde were noted as clearly migrating from the PET plastic into the water. Formaldehyde and acetaldehyde are generated by thermo-mechanical and thermo-oxidative degradation of PET in the PET melt process in the creation of the bottles. The possibility of other chemicals, such as metals, plasticizers, antioxidants, alkylphenols, UV stabilizers, lubricants, and bisphenol A, were addressed as well as genotoxicity and endocrine disruptor assays that have been used to assess the potential toxicity of bottled water. The authors note that because migration of chemicals from the plastic bottles to the water depends on storage conditions, it is difficult to compare data from all the different studies reported. In addition, the groundbreaking study of coleaching of antimony and decabromodiphenylether (BDE-209) was mentioned earlier.192 This represented the first report of a PBDE leaching from plastic bottles (PET) into bottled water, but it also represents new data on the statistically significant correlation of antimony and PBDE leaching into bottled water. In this study, total leaching of antimony in bottled waters increased from day 1 to day 60 of storage from 100 to 120 ng/ L, with a greater leaching into carbonated water (∼200−300 ng/L after 60 days of storage). Highest overall levels of antimony were found in the noncarbonated-enriched water in PET bottles, with a maximum of ∼1200 ng/L. The lowest levels of antimony were found in noncarbonated waters. It was interesting to note that noncarbonated water is often packaged in various kinds of plastic bottles (PET, PC, HDPE, and PS), whereas the noncarbonated-enriched water is primarily bottled in PET.

Herbicides and pesticides continue to be the focus of much environmental research. Recent studies have focused more on their transformation products because their hydrolysis, oxidation, biodegradation, or photolysis transformation products can be present at greater levels in the environment than the parent pesticide and can be as toxic or more toxic. Several pesticide degradation products are on the U.S. EPA’s new CCL-3: alachlor ethanesulfonic acid (ESA), alachlor oxanilic acid (OA), acetochlor ESA, acetochlor OA, metolachlor ESA, metolachlor OA, 3-hydroxycarbofuran, and terbufos sulfone (http://water.epa.gov/scitech/drinkingwater/dws/ccl/ccl3. cfm), and were previously on the UCMR-2 (alachlor ESA and OA, acetochlor ESA and OA, and metolachlor ESA and OA). The UCMR-2 drinking water data for these contaminants can be found at http://water.epa.gov/lawsregs/rulesregs/sdwa/ ucmr/data.cfm#ucmr2010. LC/MS and LC/MS/MS are now common-place for measuring pesticide degradates, which are generally more polar than the parent pesticides, making LC/ MS ideal for their detection. In addition, researchers are increasingly using UPLC to enable simultaneous analysis of larger groups of pesticides and their degradation products, and TOF-MS and Q-TOF-MS are being used to identify new pesticide degradates, along with MSE scanning (simultaneous recording of two acquisition modes, at low and high collision energy). Fenner et al. reviewed pesticide transformation in the terrestrial and aquatic environment, addressing major challenges in extrapolating from laboratory to field conditions.219 Several new methods have been developed, including one by Reemtsma et al. who created a multianalyte method for measuring 150 pesticide metabolites in groundwater and surface water using direct aqueous injection-LC/ESI-MS/ MS.220 Quantification limits of 0.1 μg/L were achieved for 142 analytes and 0.01 μg/L for 113 analytes. Masia et al. developed a LC/triple quadrupole-MS and LC/Q-TOF-MS screening method with MSE analysis to analyze 43 pesticides and degradation products in surface water and wastewater.221 This strategy allowed the detection of target and nontarget compounds. Photolysis of chloroacetamide pesticides was the focus of a fate study by Souissi et al., who used LC/ESI-MS and GC/MS with EI and chemical ionization (CI) to characterize the photolysis products.222 A total of 15 major degradation products were characterized by these techniques, and in vitro bioassays showed that the products had significant estrogenic activity. Fate during rain events was the focus of two new studies. Olsson et al. investigated the fate of pesticides and their TPs during first-flush events in a semiarid catchment in Israel.223 Results showed a heavy release of most substances to rivers. Levels fell below international guideline values (for drinking water), but levels of chloropyrifos and chloropyrifos oxon were above the acute toxicity level for freshwater invertebrates and fish. The authors stressed the importance of measuring pesticides and their TPs during such events. Petersen et al. studied herbicides and their TPs during flood events.224 Elevated levels were recorded during the floods as opposed to typical grab sampling under nonflood conditions, with pulses of recently applied herbicides being the most prominent.

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PERCHLORATE Perchlorate became an important environmental contaminant following its discovery in a number of water supplies in the western United States. It has since been found in environ2840



Review

perchlorate following fireworks displays from 2008 to 2010 was examined using LC/MS/MS. Following fireworks displays, perchlorate levels increased by 30 to 1 480× in these reflecting ponds, up to a maximum level of 519 μg/L in 2008. The deposition rate of perchlorate from fireworks displays was 4−5 orders of magnitude higher than the annual wet deposition reported for North America, such that it is clear that fireworks are an important source of environmental contamination by perchlorate. Perchlorate levels decreased via first-order kinetics, with an average half-life of 29 days. Human exposure to perchlorate through ingestion of water droplets in aerosols near these reflecting ponds was also assessed. Perchlorate contamination by fireworks could account for 6−32% of the reference dose (RfD) for infants, children, and adults. A fireworks and safety match manufacturing area (Sivakasi, India) was investigated by Isobe et al. as a source of perchlorate contamination.228 The town of Sivakasi is a huge manufacturing area, producing ∼90% of the global production of fireworks, as well as high production of safety matches. Large-volume injection-IC/MS/MS results revealed perchlorate in groundwater, surface water, and drinking water up to 7690, 30, and 0.39 μg/L, respectively. Levels in groundwater were significantly higher in the fireworks/match factory area than in other locations, indicating that these industries were a principal source of perchlorate contamination. In total, 17 out of 57 groundwater samples from Sivakasi exceeded the drinking water guideline value (15 μg/L) proposed by the U.S. EPA. The formation of perchlorate from photodecomposition of aqueous chlorine was investigated in another interesting study by Rao et al.229 Results showed that UV exposure of aqueous chlorine resulted in the formation of low levels of perchlorate, with the amount depending on chlorine and chlorate levels, UV wavelength, and pH. A reaction mechanism involving the reaction of chlorine radical reactions with chlorate was proposed. Measured perchlorate levels for UV-B and UV-C experiments fit the proposed model, but results for UV-A did not, indicating possible involvement of other pathways for UVA exposures. Finally, this update on perchlorate measurement, occurrence, and fate would not be complete without a fascinating and fun study by Schuttlefield et al., who also investigated the role of UV photooxidation in the formation of perchlorate, but on Planet Mars.230 Previous chemical analysis by the Phoenix Mars Lander had provided evidence for high levels of perchlorate in the soil of Mars (0.4−0.6 wt %).

mental waters across the United States and in other parts of the world at μg/L levels, as well as in fresh produce, foods, wines, and beverages from many countries, including those in Europe and the Far East. Perchlorate has also been found in biological samples, and it can be transported by pregnant mothers to their developing babies across the placental barrier. Perchlorate is increasingly being found in environmental waters following fireworks displays, and it is now recognized as a worldwide environmental issue, rather than only being limited to the United States. Perchlorate has been used in solid propellants used for rockets, missiles, and fireworks as well as in highway flares, safety matches, and airbag inflation systems. There is also potential contamination from fertilizers (e.g., Chilean nitrate, where perchlorate co-occurs naturally), and recent work has revealed other natural sources of perchlorate. In addition, perchlorate can be a contaminant in sodium hypochlorite (liquid bleach) that is used in drinking water treatment. Perchlorate is an anion that is very water-soluble and environmentally stable. It can accumulate in plants (including lettuce, wheat, and alfalfa), which can contribute to exposure in humans and animals. In addition, perchlorate is not removed by conventional water treatment processes, so human exposure can also occur through drinking water. Health concerns arise from perchlorate’s ability to displace iodide in the thyroid gland, which can affect metabolism, growth, and development. Because of these concerns and due to the proportion of the U.S. population exposed to it, the U.S. EPA has now decided to regulate perchlorate under the Safe Drinking Water Act (http://water.epa.gov/drink/contaminants/unregulated/ perchlorate.cfm). The regulation is currently being developed, and there is not a proposed MCL as of yet. Perchlorate was previously on the U.S. EPA’s earlier CCL lists (CCL-1 and CCL-2) and is now on the CCL-3. (http://water.epa.gov/ scitech/drinkingwater/dws/ccl/ccl3.cfm). Perchlorate was also included in the first UCMR (http://water.epa.gov/lawsregs/ rulesregs/sdwa/ucmr/data.cfm). The U.S. EPA established a reference dose of 0.0007 mg/kg/day, which translates to a drinking water equivalent level (DWEL) of 24.5 μg/L. Prior to this decision to regulate on a national basis, California had already issued a state regulation of 6 μg/L (in 2007) (http:// www.dtsc.ca.gov/hazardouswaste/perchlorate) and several states had issued advisory levels, ranging from 1 to 18 μg/L (http://www.astswmo.org/Files/Policies_and_Publications/ Federal_Facilities/2011.04_FINAL_Perchlorate_Policy_ Update.pdf). There are several EPA Methods for measuring perchlorate in water, including EPA Method 314.2 (2dimensional ion chromatography (IC) with suppressed conductivity detection), EPA Method 331 (LC/ESI-MS/MS), and EPA Method 332 (IC/ESI-MS/MS) (http://www.epa. gov/ogwdw/methods/pdfs/methods/met331_0.pdf; http:// www.epa.gov/nerlcwww/documents/m_332_0.pdf). Sturchio et al. reviewed the isotopic tracing of perchlorate in the environment.226 Isotopic data are now available for stable isotopes of oxygen and chlorine, as well as 36Cl, in perchlorate samples from natural and synthetic sources, and these are being used to distinguish sources of perchlorate in the environment and for understanding the origin of natural perchlorate. Use of isotopic data for distinguishing sources of perchlorate contamination in drinking water sources is also discussed. Wu et al. published a fascinating study of perchlorate fate in man-made reflecting ponds in Albany, New York.227 The fate of

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ALGAL TOXINS Algal toxins (mostly cyanobacterial toxins produced from bluegreen algae) are of increasing interest in the United States and in other countries around the world. Increased discharges of nutrients (from agricultural runoff and wastewater discharges) have led to increased algal blooms and an accompanying increased incidence of shellfish poisoning, large fish kills, and deaths of livestock and wildlife as well as illness and death in humans. Toxins produced by these algae have been implicated in the adverse effects. The most commonly occurring algal toxins are microcystins, nodularins, anatoxins, cylindrospermopsin, and saxitoxins. “Red tide” toxins are also often found in coastal waters. Microcystins and nodularins are hepatotoxic high molecular weight, cyclic peptide structures. Anatoxins, cylindrospermopsin, and saxitoxins are heterocyclic alkaloids, anatoxins and saxitoxins are neurotoxic, and cylindrospermopsin is hepato2841



Review

used. Low detection limits of 0.03 μg/L could be achieved. Lotierzo et al. introduced new methods using a membranebased ELISA and electrochemical immunosensor for measuring microcystin-LR in water.236 Sub μg/L detection limits were achieved, and these methods were demonstrated on river water and tap water samples. Raman spectroscopy was the focus of another method developed by Halvorson and Vikesland for microcystin-LR identification and quantification.237 SPE was used for extraction of 2 μL water samples, after which a drop coating deposition Raman method was used to enable low μg/ L detection of microcystin-LR. Zamyadi et al. published an interesting study of cyanobacterial and microcystin breakthrough and accumulation in a drinking water treatment plant.238 Their fate was followed: after the addition of coagulant and powdered activated carbon, postclarification, within the clarifier sludge bed, after filtration, and final chlorination. Elevated cyanobacterial cell numbers and total microcystin concentrations (up to 10 mg/L) were found in the clarifiers, with breakthrough of cells and toxins observed in the filtered water. Following chlorination, total microcystins reached 2.47 μg/L in the finished drinking water. Results showed that cyanobacterial cells and their toxins in real environmental blooms were more resistant to chlorination than for laboratory-cultured cells and dissolved standard toxins measured in controlled laboratory studies.

toxic. “Red tide” toxins include brevetoxins, which have heterocyclic polyether structures and are neurotoxic. Microcystins (of which, more than 70 different variants have been isolated and characterized) are the most frequently reported of the algal toxins. The most common microcystins are cyclic heptapeptides that contain the amino acids leucine and arginine in their structures. Nearly every part of the world that uses surface water as a drinking water source has encountered problems with cyanobacteria and their toxins. Algal toxins were on the U.S. EPA’s previous CCLs (CCL-1 and CCL-2) in a general way, “cyanobacteria (blue-green algae, other freshwater algae, and their toxins)”, and now, the CCL-3 has specifically named three cyanobacterial toxins: anatoxin-a, microcystin-LR, and cylindrospermopsin for the new list (http://water.epa.gov/ scitech/drinkingwater/dws/ccl/ccl3.cfm). Several countries, including Australia, Brazil, Canada, France, Italy, Poland, and New Zealand, have guideline values for microcystins, anatoxina, and cylindrospermopsin (ranging from 1.0 to 1.5 μg/L). Many of these toxins have relatively high molecular weights and are highly polar. Several good reviews have been published on algal toxins the last 2 years. For example, Merel et al. reviewed the occurrence and management of harmful cyanobacterial blooms and their toxins in surface water and drinking water.231 The authors discuss formation and monitoring of cyanobacterial blooms, origin of toxicity, occurrence and properties of cyanotoxins (including microcystins, nodularins, cylindrospermopsin, anatoxin-a, anatoxin-a(s), saxitoxins, β-N-methylamino-L-alanine, and other neurotoxins and dermatoxins), detection and quantification of cyanotoxins (including LC/UV, LC/MS, immunoassays, biochemical assays, and mouse assays), and drinking water treatment to remove them. Their economic impact was also discussed as well as research needs. Kaushik et al. summarized methods and approaches for detecting cyanotoxins in environmental samples.232 Conventional biological and analytical methods (including MS) are compared to newer molecular and biosensor approaches. Research needs identified include an integrated analysis system or biosensors that could be used for rapid sampling, concentration, detection, and quantification of all cyanotoxins in field samples. Pantelic et al. summarized existing data on characteristics of cyanotoxins, their production in the environment, and effective treatment processes to remove them from drinking water.233 Treatments discussed include conventional treatment with chlorination, coagulation, ferrate oxidationcoagulation, dissolved air floatation, rapid filtration and slow sand filtration, membrane processes (microfiltration, ultrafiltration, nanofiltration, and RO), and activated carbon; advanced oxidation with permanganate, ozonation, photochemical degradation, Fenton and photo-Fenton processes, and ultrasound degradation; and biological degradation. Several new methods have been developed for algal toxins. Lemoine reported a new method for anatoxin-a using laser diode thermal desorption-APCI-MS/MS.234 Analysis could be made in 15 s, and it enabled the removal of an isobaric interference from phenylalanine. This method avoids the timeconsuming steps of derivatization, chromatographic separation, and solid-phase extraction. Quantification limits of 3 μg/L were achieved as well as good linearity over 2 orders of magnitude. Yu et al. developed a competitive fluorescence immunoassay to measure microcystin-LR in drinking water.235 For this method, an antibody-conjugated quantum dot detection probe and an antigen-immobilized magnetic bead competitor was

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MICROORGANISMS Outbreaks of waterborne illness in the United States and other parts of the world have necessitated improved analytical methods for detecting and identifying microorganisms in water and other environmental samples. Several microorganisms are included on the new CCL-3 (http://water.epa.gov/ scitech/drinkingwater/dws/ccl/ccl3.cfm), and two microorganisms, enteroviruses and noroviruses, are now included on the third Unregulated Contaminant Monitoring Rule (UCMR-3). The U.S. EPA’s National Exposure Research Laboratory in Cincinnati has developed several methods for measuring microorganisms in water (http://www.epa.gov/nerl/topics/ water.html). These include the new EPA Method 1615 discussed earlier in the New Regulations and Regulatory Methods section for enteroviruses and noroviruses, as well as previously created methods for Cryptosporidium, Giardia, E. coli, Aeromonas, coliphages, viruses, total coliforms, and enterococci. Traditional biological methods are often used for detection of microorganisms, including cell culture, immunological methods, quantitative polymerase chain reaction (qPCR), and microscopic identification, but ESI and matrix-assisted laser desorption (MALDI)-MS methods are also often used. Several new reviews summarize methods developed for microorganisms. Aw and Rose reviewed molecular tools for rapid, high-throughput, sensitive, and specific detection of a wide spectrum of pathogens.239 Included are discussions of microarray-based detection, qPCR based detection, and pyrosequencing. A handy table listing qPCR methods and detection limits for CCL-3 contaminants is given, as well as a table of recent metagenomic detection of pathogens in environmental samples with pyrosequencing. The authors note that qPCR is the most sensitive, reliable, and costeffective of the molecular tools available. Detection of E. coli O:157 was the focus of another review by Quilliam et al., who critically assessed current methods, outlined environmental pathways for the movement of E. coli 2842



Review

O:157, and provided future directions for detecting human pathogens in the environment.240 Methods included membrane filtration, culture-based methods, immunological methods, and nucleic acid-based methods. Future directions include the promising technologies of combining microfluidics technology and analytical chemistry with molecular and immunological methods, which have led to the development of biosensors that can be used for real-time detection and in situ monitoring in the field. Methods for detecting and quantifying Legionella was the focus of a review by Buse et al., who also summarized current knowledge of engineered water system characteristics that favor and promote Legionella growth, as well as quantitative microbial risk assessment for Legionella exposures.241 Methods discussed include culture, DNA microarray, flow cytometry, fluorescent in situ hybridization, immunomagnetic separation (with and without culture or qPCR), peptide nucleic acid probe, and qPCR. Ikner et al. published a comprehensive review on concentration and recovery methods of viruses from water.242 Filtration techniques include electronegative and electropositive filtration media and ultracentrifugation. The authors note that recovery from filters can be highly variable, and there is not a way to elute all pathogenic viruses from the filters currently in use. The biological variability of viruses, pertaining to their surface composition, net charge, and tendency to form aggregates, contributes significantly to the different recoveries observed. Levantesi et al. reviewed the occurrence of Salmonella in environmental waters and drinking water and the relevance of these sources for transmission. Control strategies to prevent Salmonella drinking water-related transmission were also discussed, as well as the role of Salmonella-contaminated waters in foodborne outbreaks. Salmonella has been detected in rivers, lakes, ponds, groundwater, and drinking water, with sources linked to humans, domestic animals, and wildlife. Positive correlations between Salmonella frequency and rainfall events indicate that surface runoff plays a major role in its load in the aquatic environment. While Salmonella is a significant health risk in developing countries, it is reported less frequently in developed countries than other disease causing pathogens, such as Cryptosporidium, Giardia, E. coli O:157:H7, norovirus, Shigella, and Campylobacter. Mycobacteria occurrence in water, soil, plants, and air was the focus of another review by Hruska and Kaevska.244 Mycobacteria can cause tuberculosis and leprosy and can participate in inflammatory pathways involved in several other important human diseases like Crohn’s disease, asthma, type 1 diabetes mellitus, psoriasis, arthrosis, and others. Occurrence of mycobacteria in drinking water, bottled water, river and lake water, coastal waters, swimming pools, and hot tubs was summarized. Methods continue to be developed for microorganisms, and a few are highlighted here. Brinkman et al. evaluated new methods for concentrating norovirus, adenovirus, and enterovirus from wastewater.245 Celite (diatomaceous earth) was evaluated, along with large-volume nucleic acid extraction and measurement using qPCR. This procedure yielded 47−98% recovery and was effective for extracting and measuring all three groups of viruses within a single method. Angus et al. created a new field-deployable, near real-time optical microfluidic biosensor method for measuring Cryptosporidium in environmental waters.246 The total assay time took only 10 min and

enabled subsingle oocyst detection. Compared to EPA Method 1623, this method was much faster and simpler. Finally, Maheux et al. reported a new concentration and recovery procedure combined with DNA enrichment and real-time PCR for measuring E. coli and Shigella in drinking water.247 This method was capable of detecting as little as one E. coli or Shigella cell present in a 100 mL drinking water sample and was comparable to EPA Method 1604 in terms of analytical specificity and detection limits but was significantly faster.

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CONTAMINANTS ON THE HORIZON: IONIC LIQUIDS AND PRIONS Ionic liquids are organic salts with a low melting point (

Water Analysis: Emerging Contaminants and ... - ACS Publications

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