Contaminants of Emerging Concern: Occurrence, Fate, and Remediation

CHAPTER 4

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation Tabish Nawaz, Sukalyan Sengupta Civil and Environmental Engineering Department, University of Massachusetts Dartmouth, Dartmouth, MA, United States

4.1 Introduction In the past few decades, scientific achievements have made a significant impact on human activities and lifestyle. These developments, made largely possible by advances in synthetic chemistry, have led to an exponential increase in pharmaceuticals, personal care, and household products. A study shows that between 1930 and 2000, the global annual production of anthropogenic chemicals increased from 1 million to 400 million tons [1]. As a result, a variety of chemical compounds, previously undetected and thereby unregulated, began entering the environment via waste streams. These included chemical compounds from agricultural pesticides, veterinary pharmaceuticals, pharmaceuticals, personal care, and household products. These chemical compounds are collectively termed as contaminants of emerging concern (CECs). They are defined as any anthropogenic or naturally occurring chemicals or microorganisms that are typically not monitored in the environment, but can potentially enter and harm the environment causing adverse ecological and/or human health effects. The two criteria for long-term classification of a substance as an emerging contaminant are (i) its persistence in the environment and/or (ii) potential harmful human and ecotoxicological effects. In some cases, the release may have been going on for a while, and is now widespread, but the substance is not regarded as an emerging contaminant due to the lack of its reported adverse effects. Conversely, new information, for example, as in the case of endocrine disruption by nonylphenol, can result in reclassification of a well-known contaminant as an emerging contaminant. What is “emerging” and particularly “concerning” about CECs is that the improvement in analytical techniques have only recently begun to detect these chemicals in the level as low as ng/L in the wastewater streams, and whose toxicological risks, though reported in laboratory trials, largely remain unknown in the environment, as concentration is in trace levels. The fact that the existing water/wastewater treatment infrastructure is not designed to tackle them, as Advances in Water Purification Techniques. https://doi.org/10.1016/B978-0-12-814790-0.00004-1 Copyright # 2019 Elsevier Inc. All rights reserved.

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evidenced by the persistence of CECs posttreatment at wastewater treatment plants (WWTPs) and drinking water treatment plants (DWTPs) further adds to the concern with respect to their persistence in the environment. Also disconcerting to the scientific community is the lack of clarity about these contaminants’ fate, transformation pathways, and toxicity of their degradation products in the environment. Answering these questions remains a major environmental challenge. Though the repercussions of CECs are still being studied, and conclusive levels that can be harmful under the environmental conditions over a long time are being evaluated, statistics show [2] that between 2002 and 2011, more than 50% of the total chemicals produced belonged to environmentally harmful category, with more than 70% of these chemicals having significant environmental impact. This indicates a likelihood of their toxic effects upon release in the environment over long time and persistent use. However, the pace of discovering these contaminants and our understanding of them in the environment is painfully slower than their production and subsequent release. Therefore, it takes several years for a contaminant to be termed as emerging. Much of the future studies on toxicology need to focus on these CECs, for they have come into widespread use before anything substantial is known about their environmental impacts. Many consumers meet their water demands from sources that are replenished either directly or indirectly from WWTPs. Therefore, waste streams from diverse sources such as households, hospitals, hotels, agriculture, and industries are analyzed and treated at centralized facilities such as WWTPs. Subsequently the CECs typically enter surface water and groundwater, and are carried into the marine environment as far as the Arctic Circle. In general, wastewater influent and effluent streams are analyzed for certain contaminants regulated by agencies such as United States Environmental Protection Agency (US EPA) and European Environment Agency (EEA). The problem of CECs is also the problem of analytical methods and existing regulations. Since CECs (both from the contaminants and their degradation products) are often in the range of few ng/L to mg/L against a high background matrix of waste streams, their detection requires elevated level of sensitivity and selectivity. The availability of analytical methods such as liquid chromatography (LC) along with mass spectrometry (MS)—especially high-resolution mass spectrometry (LC-HRMS) and time-of-flight mass spectrometry (LCTOFMS)—and gas chromatography mass spectrometry (GC-MS) have enabled detection of CECs in the ng/L concentration range. This has also impelled agencies like the EPA to update their current contaminant list regularly. The regulations and analysis, though they are crucial and provide extent and level of contamination and the resulting response, do not preclude the need for developing and incorporating innovative wastewater treatment methodologies in our current treatment practices. So far, an organized, concerted effort in this respect is lacking. However, many studies have been reported for the removal of certain CECs. Another key step for large-scale

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 69 elimination of CECs requires implementation of more decentralized treatment strategies. As different waste streams possess differing level of water quality and kinds of CECs, decentralized treatment schemes, where source separated waste streams are treated with methodologies most suited to them, would address the problem more effectively than centralized schemes currently employed in WWTPs, where one-approach-for-all is applied, resulting in poor removal of CECs. This chapter investigates the problem of CECs from regulatory, occurrence, fate, environmental impact, and treatment perspectives. We begin by examining the present state of regulatory standards for drinking and wastewater in the United States, and the future initiatives proposed by the EPA in dealing with the CECs. We also analyze the occurrence and fate of CECs in the environment, focusing both its spatial and temporal aspects. Treatment schemes have been reviewed and discussed.

4.2 Regulations and Analytical Methods 4.2.1 Currently Regulated Contaminants Currently, the EPA regulates 88 contaminants in drinking water under National Primary Drinking Water Regulations (NPDWRs). According to the EPA, the regulations are legally enforceable primary standards and treatment techniques that apply to public water systems. These primary standards are in place to protect public health by limiting the levels of contaminants in drinking water. There also exists separate National Secondary Drinking Water Regulations (NSDWRs) covering 15 contaminants. These regulations are nonenforceable guidelines regarding contaminants that can potentially cause cosmetic effects (such as skin or teeth decoloration) or aesthetic effects (such as taste, odor, or color) in drinking water. The secondary regulatory standards are merely recommendations to water systems by the EPA and are not legally binding on them. However, it is entirely up to respective states to decide whether to adopt them as enforceable standards. The complete list of 88 contaminants in NPDWR and 15 contaminants in NSDWR with their maximum concentration levels, common sources of potential health effects, and public health goal levels are provided in the tables available on USEPA website (https://www.epa.gov/ground-water-and-drinking-water/national-primarydrinking-water-regulations and https://www.epa.gov/dwstandardsregulations/secondarydrinking-water-standards-guidance-nuisance-chemicals).

4.2.2 Wastewater Treatment and Discharge Regulation: Clean Water Act The EPA regulates the discharge and treatment of wastewater under the Clean Water Act (CWA). An earlier act, Water Pollution Control Act, 1948, is considered the basis for the CWA. In 1972, the earlier act was significantly reorganized and expanded to form the CWA. The act lays down the basic structure for regulating the release of contaminants into the US waters and

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establishing quality criteria for surface waters. Under the act, EPA has set wastewater standards for industry and for all contaminants in surface water. The CWA makes discharge of any pollutant from point sources, such as pipes or man-made ditches, illegal unless a mandatory approval has been obtained from National Pollutant Discharge Elimination System (NPDES). Individual households connected to a municipal system, using septic tanks, or not making any surface discharge are exempted from NPDES permit. The rest of the facilities, including industrial and municipal, must obtain a NPDES permit. The act also develops and updates analytical methods from time to time under CWA Analytical Methods. For CECs, two analytical methods have been developed—Method 1694 for a suite of 74 pharmaceuticals and personal care products by HPLC (high-pressure liquid chromatography)/MS/MS, and Method 1698 for a suite of 27 steroids and hormones by highresolution gas chromatography (HRGC)/HRMS, that may be determined in wastewater effluents and influents, and sewage sludge (biosolids). The list of contaminants covered in the two methods is provided in Appendix A. First published in 2007, these methods have been validated only in a single laboratory and have not undergone multilaboratory validation. Since their promulgation, many laboratories have implemented the methods or their variants. The latest update to these methods were made in August 2017, when EPA added certain methods (methods# 608.3, 624.1, 625.1) reviewed under the Alternate Test Procedures (ATPs) program, and reviewed the procedure for the estimation of the method detection limit. The methods were developed to screen WWTP matrices consisting of influent, effluent and biosolids for a wide array of pharmaceuticals, and personal care products. These methods are not regulatory, and are only used to screen samples from various sources to profile occurrence of CECs. The methods are open to improvement from the scientific community and are thus upgraded accordingly. For instance, in an August 2017 update, method numbers 608, 624, and 625 were replaced by 608.3, 624.1, and 625.1 respectively. The CWA has also developed Method 1614A for brominated diphenyl ethers in water soil, sediment, and tissue by HRGC/ HRMS and Method 1699 for pesticides in water, soil, sediment, biosolids, and tissue by HRGC/ HRMS. Promulgation of newer, more sensitive, and selective methods is necessary for accurate detection of CECs; based on which EPA regularly updates the list of unregulated contaminants in a contaminant candidate list (CCL). The CCL contains a list of drinking water contaminants that are reported or expected to occur in public water systems and are presently unregulated from EPA drinking water standards. Drinking water regulation: Safe Drinking Water Act (SDWA). The Safe Drinking Water Act (SDWA) is the main federal law that gives the EPA the authority to regulate contaminants in drinking water and ensure its quality. This law regulates all waters that are actually or potentially utilized for drinking purposes, irrespective of their sources— above ground or underground. The SDWA was first passed by Congress in 1974, and was

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 71 further amended in 1986 and 1996. The SDWA allows EPA to set national health-based standards for drinking water, protecting consumers against both naturally occurring and manmade contaminants that can potentially be found in drinking water. Under the SDWA, EPA develops NPDWRs for a new contaminant by a three-step procedure. The first step involves identification of the unregulated contaminants, preparing and publishing a list of unregulated contaminants in a CCL, and then based on monitoring data, risk analysis, and other relevant information, prioritization of the contaminants is performed. In the next step, EPA carries out a series of evaluations to determine whether they should initiate the rule making process for developing an NPDWR for a given contaminant. The EPA essentially checks if the contaminant meets three criteria: (1) if the contaminant may cause an adverse health effect on consumers, (2) the contaminant occurs or has substantial likelihood of occurrence in drinking water with a frequency and at a level of public health concern, and (3) based on the discretion of EPA administrator, if the regulation of the contaminant in question leads to meaningful opportunity for reducing health risks to consumers served by public water systems. Following the two steps, EPA publishes a preliminary regulatory determination report, allows the public and experts to comment and discuss with states and other federal agencies, reviews and considers the comments and suggestions, and publishes a final notice in the Federal Register. Subsequently, EPA makes a decision if the contaminant in question necessitates regulation or not. In the case of regulation, EPA begins to frame rules in the NPDWR and reviews it every 6 years to check if changes are required. In the case of decision against regulation, EPA may issue a health advisory or take no further action. The complete list of currently regulated contaminants can be found at the EPA website, under the document titled NPDWRs. The EPA currently lists 88 contaminants under NPDWR, including CECs, heavy metals, microorganisms, and turbidity, which is a surrogate parameter for the concentration of colloidal solid. The NPDWR also contains maximum contaminant level of the contaminants, potential health risks due to their exposure, and the desired level for meeting the public health goal. The SDWA also makes it necessary for the EPA to evaluate unregulated contaminants that may require future regulations by maintaining CCL. The CCL is a list of contaminants that are (1) not currently listed in NPDWR, (2) known or expected to occur at public water systems, and (3) may warrant regulation. The SDWA requires EPA to publish the CCL every 5 years. So far four drafts of CCL have been prepared starting with CCL—1 (50 chemicals and 10 microorganisms) that came in 1998; CCL—2 (43 chemicals and 9 microorganisms), and CCL—3 (112 chemicals and 12 microorganisms) came in 2005 and 2009 respectively; the latest being CCL—4 (97 chemicals and 12 microorganisms) (see Appendix B) which was prepared in 2016. After publishing the CCL, EPA is required to determine whether to regulate at least five contaminants using a separate process called Regulatory Determinations. The EPA compiles and evaluates additional data on all the contaminants in CCL, conditional to availability, and determines which contaminants have sufficient information to be tested against the three criteria of the

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SDWA (discussed above). The timely update to CCL shows the dynamic nature of the regulation task with respect to CECs. The EPA utilizes CCL to identify priority contaminants for regulatory decision-making and information collection. As per SDWA, EPA also maintains a National Contaminant Occurrence Database (NCOD), where unregulated contaminant occurrence data, 6-year review data, and ambient and/or source water data are prepared and maintained. The SDWA also lays down rules for EPA to monitor unregulated contaminants under Unregulated Contaminant Monitoring Rule (UCMR) (see Appendix C). The UCMR program was developed to monitor up to 30 contaminants every 5 years, in large systems and sample of small public water systems serving less than 10,000 people. Occurrence data are collected through UCMR for NCOD. The UCMR was developed and works in close coordination with CCL. In developing UCMR, health risks of potential contaminants are extensively studied, based on which contaminants are ranked upon their severity and become candidates for future CCLs. The data developed under UCMR is one of the primary sources of occurrence and exposure information which EPA uses for developing regulatory decisions for emerging contaminants.

4.3 Emerging Contaminants 4.3.1 Occurrence, Fate, and Environmental Impact The CECs, like any other conventional pollutant, result from domestic, commercial, industrial, and agricultural activities. Therefore, waste discharges from households, commercial entities like hospitals and hotels, industrial units, agricultural, and animal/poultry farms have been reported to contain chemical micropollutants in them. Additionally, the CECs, due to their persistence and long-distance mobility, have been detected in various areas where they were never used or discharged. Pharmaceuticals and personal care products from household use release pollutants into the environment. Also, agricultural and animal husbandry practices, for example, spreading of manure/sludge/pesticides on agricultural fields result in leaching of contaminants to surface and groundwater. Additionally, hormones/drugs in animal urine or excrements may wash into wastewater. Certain pesticide metabolites, which are biologically active and toxic, have also been detected in wastewater effluents. The CECs therefore originate from many sources, among which the most significant are WWTPs, seepage from septic tanks, and landfill leachate. These sources can be classified as point sources and nonpoint/diffused sources. Point-source discharge takes place via a discrete spatial location from where its input enters into aquatic systems in generally a more constrained or confined manner. Examples include municipal sewage, industrial wastewater, landfill leachate, and discharge from hospitals. In contrast, nonpoint discharge takes place from ill-defined, diffused sources occurring over large geographical scales, for example, agricultural runoff, stormwater runoff, and groundwater leaking municipal sewage lines. The contaminants from these two kinds of sources finally end up in aquatic resources (surface water and groundwater). Fig. 4.1

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 73

Landfill

Animal waste

Aerial sources

Septic tanks

lagoon Leakage*

Sources and pathways Major source Minor source Major pathway/flux Minor pathway/flux Major removal proc. or pathway Rapid route bypassing natural attenuation MAR Managed artificial recharge

Hospital waste/eff.

Leakage*

Soil zone

Runoff or

MAR

Leachate

Domestic waste/eff.

Sewerage system

Wastewater treatment proc

WWT effluent and biosolids MAR Unsaturated zone

Groundwater

Industrial waste/eff.

Accidental leaks, emergencies, urban runoff

Untreated discharge*

Surface water

Receptors

Fig. 4.1 Schematic of sources and pathways of CECs in the environment. From D.J. Lapworth, N. Baran, M.E. Stuart, R.S. Ward, Emerging organic contaminants in groundwater: a review of sources, fate and occurrence, Environ. Pollut. 163 (2012) 287–303, with permission.

illustrates a schematic view of the pathways by which CECs from various sources meet their receptors. Though, the sources and receptors are known and defined, very little is known about CECs’ pathways connecting them. This is due to poor information available on the physical and chemical properties of the target species, and the complexity of its behavior in the environmental systems, particularly at trace concentrations. Determining an accurate picture of CECs pathways in the environment remains a key area for advanced research. The scheme in Fig. 4.1 shows that wastewater streams from various anthropogenic sources are among the major carriers of CECs into the environment. A majority of these streams are treated at WWTPs and subsequently released into the environment receptors such as groundwater, surface water, and to consumers via drinking water supplies. Several studies reported detections of CECs in water samples from WWTPs [3], DWTPs [3a], surface water [4], and groundwater [5]. These CECs belong to distinct categories; in the following sections the appearance, fate, and environmental impact of each category of CECs are discussed. 4.3.1.1 Pharmaceutical-based CECs Prescription and over-the-counter drugs constitute pharmaceuticals. They include antibiotics, birth control pills, antidiabetics, antidepressants, betablockers, lipid regulators, impotence drugs, painkillers, tranquilizers, and other medicines. The global annual per person consumption of pharmaceuticals is 15 g in developing countries and 50–150 g in developed countries [6]. It is estimated that roughly 3000 different substances are used as pharmaceuticals; however, only a handful of these chemicals in the environment (
150) have been investigated so far [7]. The chief sources of pharmaceuticals in the environment are drugs excreted or disposed-off in the domestic sewer systems or leaky landfills [8, 9]; hospital effluents; and

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animal husbandry and aquaculture runoffs [10]. Blood pressure reducers, hormones, psychiatric drugs, antibiotics, and painkillers have been found in WWTPs effluents and their recipient water bodies [11]. The release of pharmaceuticals and their metabolites into water bodies has been observed globally. A study reports detection of 631 human- and veterinary-based pharmaceuticals over their detection limit in 71 countries [12]. The study concludes that the major source of CECs into the environment worldwide is urban wastewater effluent, although contributions from pharmaceutical industry, agricultural activities, and aquaculture have local significance. Many studies also reported the presence of pharmaceuticals and their metabolites in marine water. One study estimates the presence of 196 pharmaceuticals by monitoring 50 marine sites. Mostly antibiotics such as erythromycin, sulfamethoxazole, and trimethoprim; also, carbamazepine, ibuprofen, acetaminophen were detected in maximum concentration in the lower μg/L range [13]. The antidepressant lamotrigine and its conjugate 2-N-glucruonide was determined by Ferrer and Thurman [14] in wastewater, surface water, drinking water, and groundwater. In the surface water, a high average concentration of glucuronide (209 ng/L) was discovered, implying that the conjugate passed through WWTPs without any significant degradation. The point discharge by WWTPs effluents into two streams—Boulder Creek (Colorado) and Fourmile Creek (Iowa)—has been demonstrated to result in ng/L levels of antidepressants [15]. Metcalfe et al. [16] analyzed six antidepressants and five human metabolites in Canadian WWTPs and river water samples. The two metabolites of venlafaxine —O- and N-desmethyl venlafaxine—were in maximum concentrations among the detected CECs with average values at 2.1 and 1.1 μg/L in a WWTP and 0.109 and 0.047 μg/L in rivers, respectively. Prasse et al. [17] studied nine antiviral drugs and their metabolites in surface water and raw and treated wastewater samples; the maximum concentration was observed to be around 190 and 170 ng/L for acyclovir and zidovudine, respectively. Oseltamivir carboxylate, an antiviral drug metabolite, was detected in WWTP samples and rivers in Japan, with levels rising during flu seasons, reaching up to 293 ng/L in WWTPs effluents and 190 ng/L in rivers [18]. Pharmaceuticals classified as glucocorticoids, such as cortisone, cortisol, prednisolone, triamcinolone amide, and prednisone, were detected in hospital wastewaters in concentrations ranging between 13 and 1918 ng/L [19]. Lindberg et al. [20] analyzed six antimycotics in wastewater and sludge samples. The study noted that fluconazole was the only antimycotic observed in the raw wastewater and WWTP effluent samples, with concentration between 90 and 140 ng/L, whereas clotrimazole, ketoconazole, and econazole were detected in all sludge samples but not in the WWTP effluent samples. Svanfelt et al. [21] noted thyroid hormone derivative in WWTP influent and effluent samples at 64 and 22 ng/L respectively. Between 2006 and 2007, 51 CECs were monitored by Benotti et al. [22] in the US drinking water samples; pharmaceuticals (carbamazepine, naproxen, trimethoprim, sulfamethoxazole, atenolol, and gemfibrozil) were most commonly detected with median concentrations <10 ng/L. Boles and Wells [23] reviewed the environmental occurrence of amphetamines and methamphetamines. Opioids (such as oxycodone and methadone) and muscle relaxants (metaxolone) were

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 75 detected at high concentration levels (oxycodone: 1700 μg/L; metaxolone: 3800 μg/L) in WWTP effluents linked to pharmaceutical processing facilities [24]. A seasonal and temporal variation of cocaine was noted in a WWTP in Brussel-Noord, Belgium by van Nuijs et al. [25]. The study reported insignificant seasonal variation, but a constant monthly use with elevated levels during weekends. This clearly reflects how human behavior directly affects the level of CECs in the environment. The seasonal variation of antiepileptic drug carbamazepine and betablockers was studied in a Swedish river-lake system [26]. A total of 73 pharmaceuticals were reported in wastewater and surface water samples by Gros et al. [27]. Loos et al. [28] performed a pan European study across 23 countries to detect polar organic persistent compounds, along with pharmaceuticals, in groundwater samples. The study noted carbamazepine to be the only pharmaceutical that was found above the quality standard for pesticides in groundwater. Reviewing the occurrence of pharmaceuticals and their transformation products (TPs) in drinking water, [29]reported the presence of lipid regulators, bezafibrate, clofibric acid, gemfibrozil, the antiepileptic drug carbamazepine, and ibuprofen. The TPs identified were Gd(III) from diethylenetriamine pentaacetate (DTPA) complex which is used for magnetic resonance imaging (MRI), ozonation products of carbamazepine, chlorination products of sulfamethoxazole, and the popular pain reliever acetaminophen. More than 30 antibiotics have been reported in sewage influent and effluent samples, surface water, groundwater, and potable waters. Most antibiotics are water soluble, which makes their excretion via body waste convenient. [30]reported that
90% of the antibiotics ingested by animals and humans are released as urine, and up to 75% of them via feces. The summary of occurrence of CECs from pharmaceuticals in different media is presented in Table 4.1. Pharmaceuticals are biologically active molecules; therefore, they undergo transformation due to metabolic processes inside the body of the consumer, and after discharge into the waste streams. Biological transformation reactions of CECs during wastewater treatment processes under aerobic conditions have been the focus of several review articles [43–46]. Detoxification processes within the body carried out by bacteria are also potential pathways of transformation. Singhal and Perez-Garcia [47] reported conjugation reaction or conversion to more polar TPs for better release of the cell. Biological wastewater treatment processes are typically carried out into aerobic and anoxic stages, which enable the transformation of organic molecules and organically bound nitrogen. Since the aerobic stage has a higher biological activity, it is surmised that significant transformation of CECs occurs under these conditions. In a first study, [48]reported the conversion of betablocker atenolol to atenolol acid via amide hydrolysis and the hydroxylation of the hypoglycemic substance glibenclamide in activated sludge. Helbling et al. [49] reported demethylation and hydroxylation of antiepileptic diazepam in contact with activated sludge. In a laboratory-scale study of bioreactors spiked with 5 and 500 μg/L of trimethoprim, Jewell et al. [50, 51] reported four and two TPs. At the lower concentration (5 μg/L), the study identified a completely different transformation route. The difference caused due to concentration suggests the importance of determining accurate levels of pharmaceuticals CECs in the environmental systems, which would enable their

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Chapter 4 Table 4.1 Summary of occurrence findings of CECs CECs

Occurrence Media

Carbamazepine Sulfamethoxazole Ibuprofen Diclofenac Clofibric acid Paracetamol Ketoprofen Triclosan Iopamidol Lincomycin Propyphenazone Sulfamethazine N,N-Diethyl-metatoluamide Phenazone Salicylic acid Ciprofloxacin Trimethoprim Naproxen Diclofenac Mefenamic acid Acetaminophen

Groundwater, WWTP effluent, freshwater-rivers, canals, seawater

Concentration Range (ng/L)

References

Pharmaceuticals 1.64–99,194 5–2800 0.6–12,000 2.5–590 4–7300 15–120,000 3–2886 7–2110 130–2400 100–320 15–1250 120–616 454–6500

Lapworth et al. [5], Pal et al. [31], Arpin-Pont et al. [13]

25–3950 <0.3–2098 23–3353 <0.5–7900 <0.3–5100 <0.5–3300 <0.3–554 1.8–777 Lifestyle compounds

Caffeine Cotinine Benzophenone-4 Phenylbenzimidazole sulfonic acid Climbazole Benzophenone-3

Groundwater Raw wastewater, seawater Raw wastewater

WWTP influent, river, seawater, swimming pool

13–110,000 60–400 5100 (max)

Lapworth et al. [5]

3900 (max)

Wick et al. [32], Ramos et al. [33] Wick et al. [32]

1400 (max) 28 (max in river); 127 (max in WWTP)

Wick et al. [32] Pedrouzo et al. [34], Tsui et al. [35]

EDCs and hormones Estriol Progesterone Thyroid 17b-Estradiol Estrone

River, canal, estuaryseawater Hospital wastewater Raw wastewater Groundwater

12

Pal et al. [31]

227 (max) 10.5–84.9 0.79–120 0.1–45

Guedes-Alonso et al. [36] Svanfelt et al. [21] Lapworth [5]

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 77 Table 4.1 CECs

Summary of occurrence findings of CECs—Cont’d Occurrence Media

Concentration Range (ng/L)

References

Flame retardants Tris(chloropropyl) phosphate Tetrabromobisphenol A Marbon

Swimming pools

1180

Teo et al. [37]

Seawater River sediment

2.8 450 ng/g

Gu et al. [38] Guo et al. [39]

63,000 320–510 487 (max) 266 (max) 890 (max) 7400

Kowal et al. [40] Zhou et al. [41] Loos et al. [28]

Pesticides N, N-Dimethylsulfamide DDT Desthylatrazine Desethylterbutylazine N-Chloridazon Desphenyl-chloridazon (DPC)

River Groundwater Groundwater, rivers, wastewater

Buttliglieri et al. [42]

tracking to be more accurate and their remediation more effective. It also highlights the importance of utilizing environmentally relevant concentrations when analyzing transformation pathways, particularly for antibiotics. The transformation pathways suggested by the study were validated by the detection of 2,4-diaminopyrimidine-5-carboxylic acid in an activated sludge reactor effluent at 61 ng/L concentration (accounting
52% of the transformed product). Boix et al. [52] reported the degradation of irbesartan and ibuprofen in an activated sludge sample via hydroxylation of their aromatic moieties. Similarly, the diaminopteridine moiety of methotrexate was observed to be hydroxylated in activated sludge as a minor pathway reaction; also, methotrexate formed a secondary amine via demethylation that can occur with molecules with methyl amine moieties during biological wastewater treatment leading to the formation of primary or secondary amines [53]. Citalopram was also reported to be demethylated in contact with activated sludge. These reactions are a form of N-dealkylation with many CECs. With amine group in biological media, reactions such as N-oxidation and conjugation (e.g., acyl conjugation) have been observed [54]. In the presence of ammonia oxidizing bacteria (AOB), deamination of the aniline moiety and oxidation of nitro group have been observed for sulfamethoxazole [55]. Poirier-Larabie et al. [56] noted decarboxylation of diclofenac resembling the decarboxylation of 4-hydroxycinnamic acid to 4-hydroxystyrene. Jewell et al. [50, 51] reported biotransformation of diclofenac during wastewater treatment. The reactions like the well-known conjugation reactions of pharmaceuticals inside the body [57] have been noticed to occur in activated sludge samples, for example, the sulfate conjugation of triclosan [58] and dextrorphan [59]. Funke et al. [60] investigated the fate of five antiviral drugs—abacavir, emtricitabine, ganciclovir, lamivudine, and zidovudine—during biological wastewater treatment; oxidation of terminal hydroxy— moiety to carboxylate group was observed for each drug. Lamivudine and emtricitabine

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showed oxidation of thioether moiety to sulfoxides. In the influent to WWTPs, antiviral drugs were detected (emtricitabine: 980 ng/L), however, in the effluent, chiefly TPs were noticed with concentration reaching up to 1320 ng/L for carboxy-abacavir. Hirte et al. [61] studied the hydrolysis of amoxicillin and reported that the hydrolysis rate as well as specimen of TPs were strongly dependent on pH. Despite extensive availability of literature on the fate of pharmaceuticals in the environment, results about their environmental effects are limited. This is mainly attributed to their low concentration in the environment. This is even more concerning as nothing can be said reliably about their long-term effects, whereas the human consumption of these substances continues to rise. However, pharmaceuticals in waterbodies are considered as ecotoxicological hazards, as these molecules are bioactive in their original forms and some of them have been reported to harm many aquatic lifeforms. Currently, the estimated levels of pharmaceuticals in the environment are substantially lower than their lowest observed effect concentrations (LOECs), with a few notable exceptions of salicylic acid, diclofenac, propranolol, clofibric acid, carbamazepine, and fluoxetine; in the case of diclofenac, the LOEC for toxicity in fish was in the same range as wastewater concentration, similarly the LOEC for zooplankton and benthic organism with respect to propranolol and fluoxetine was near the maximum values measured in wastewater effluents [62]. The sex hormones, such as estrogens and androgens typically pose the maximum potential concern, followed by cardiovascular drugs, antibiotics, and antineoplastics [63]. Blood lipid lowering drugs (bezafibrate, fenofibrate, and gemfibrozil) showed high estrogenic activity in a breast cancer cells proliferation study [64]. Plankton and algal growth have been reported to suffer in the presence of the antibiotic ciprofloxacin at its environmentally pertinent concentrations [65]. A dramatic drop in vulture population (>95%; roughly 40 million) in Pakistan has been caused by the renal failure induced by the residue of veterinary applied diclofenac [66]. Many studies have reported that antibiotics discharged into the environment may alter the ecosystem. In some case, benthic organisms near aquaculture sites have been adversely affected by the antibiotics. Wild organisms near pisciculture stations have been found to contain in vivo a significant concentration of antibiotics, for example, in Norway wild fish caught near aquaculture areas contained elevated levels of oxytetracycline where cultured fish were treated with drugs. This suggests the possibility of transportation of oxytetracycline into the natural environment from the controlled environment. Additionally, many antibiotic-resistant microbes have been found near these fish farms. Moreover, the microbial surveys in these areas show that roughly 20 groups of microbes found in aquaculture regions are potentially pathogenic to humans. This leads to a dangerous situation for human health when pathogens in aquatic environment develop antibiotic resistance. It also poses a grave threat to other creatures’ growth and populations in the aquatic environment, for example, Vibrio cholerae, which is a serious human pathogen after developing antibiotic resistance in Ecuadorian shrimp farms (where massive amounts of antibiotics were being used), led to luminescent vibriosis in the shrimp population,

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 79 Table 4.2 Summary of the ecotoxicological effects of CECs Contaminants

Ecotoxicological Effects

Sucralose Acesulfame

Toxic effects on gills, muscles and brain of carps [67] Increased oxidative stress on Carassius auratus [68, 69]; Zebra fish embryo toxicity [70–72] Phytoplankton and bacterioplankton significantly reduced for concentration 500 μg/L [73]; toxic to zebra fish embryo [74] and rainbow trout [75] Disruption of cell membranes and oxidative stress [76] Cancer, reproductive and developmental effects, endometriosis, ulcerative colitis, thyroid disease [77] Bladder and colorectal cancer, miscarriage, birth defects [54]; evolution of resistant E. coli [70–72] Acute toxicity to aquatic organisms [54]; bleaching and death of coral reefs [35] Environmentally persistent, transported over long distances [54] Shellfish poisoning, human illness [78]; liver, kidney, gills, muscle, and intestine damage the fish [79] Damage to digestive tract and internal organs of seabirds, seals, sea lions, dolphins, whales, marine reptiles, and zooplankton [54] Hormonal effects in humans leading to breast cancer [80] Stimulates human estrogen receptor, estrogenic to breast cancer cells, rainbow trout estrogen receptor [81] Miscarriage, pregnancy complications [82] Feminization in fish [83]; endocrine disruption Affects gills and kidneys of fish, brown trout [84] Harmful to aquatic organisms, inhibited polyp regeneration and reduced reproduction [85] Growth impairment in human embryonic kidney cells [86]

Silver nanoparticles

Nano zerovalent iron PFASs DBPs UV filters Flame retardants Algal toxins Microplastics

Bisphenol A Butylated hydroxyanisole

Phthalates Estrone, 17-ß estradiol and 17-α ethynylestradiol Diclofenac Ibuprofen

Mixture of atenolol, bezafibrate, carbamazepine, cyclophosphamide, ciprofloxacin, furosemide, hydrochlorothiazide, ibuprofen, licomycin, ofloxacin, ranitidine, salbutamol, and sulfamethoxazole in 10– 1000 ng/L C60 fullerenes Genotoxic response for concentration 2.2 μg/L; DNA damage in humans [87]

resulting in huge economic loss. Table 4.2 summarizes the relevant studies on ecotoxicological effects of pharmaceuticals. 4.3.1.2 Life-style compound-based CECs Several life-style products, such as cosmetics, fragrances, food additives, etc., release CECs over their life cycle. In this section, some of the critical compounds are discussed with respect to their occurrence, fate, and environmental impact. Artificial sweeteners (sucralose,

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saccharine, acesulfame, aspartame, neotame, stevioside, glycyrrhizic acid, neohesperidine dihydrochalcone, and cyclamate) are increasingly used in foods and soft drinks globally. They are mostly stable and enter the environment via treated wastewater after their direct disposal or excretion through urine after consumption. So far, the contamination from artificial sweeteners was mostly considered and studied in Western countries; however, a recent study reports their noteworthy levels in surface waters of Philippines (sucralose, saccharine, acesulfame, and cyclamate), and Myanmar (acesulfame and cyclamate), and in groundwater in Vietnam (acesulfame) [88]. Sucralose and acefulame, due to their high stability and elevated levels in the environment (ppb level), are increasingly being recognized as potential tracers for human inputs into the environment (as a measure of groundwater contamination by wastewater), replacing caffeine [54]. Jmaiff Blackstock et al. [89] reported elevated levels of acesulfame in swimming pools and hot tubs. More than 250 samples from 31 pools and hot tubs showed acesulfame in concentration range 30–7110 ng/L. Toxicity concerns for artificial sweeteners have been reported for aquatic organisms. Recent studies have reported toxicity for their TPs. At environmental levels (0.05 and 155 μg/L), sucralose has been reported to affect the gills, muscle tissues, brain, and liver of Carp. No bioaccumulation of sucralose was observed, but oxidative damage in lipids and protein was noticed [67]. Acesulfame photolysis products showed toxic effect on zebra fish embryo; the photolysis is reported to occur naturally in sunlight, and the work reports six new TPs of acesulfame [70–72]. Carassius auratus experienced oxidative stress in the presence of UV photolysis products of acesulfame; however, acesulfame itself did not cause any stress [68, 69]. Simulated solar radiation (>290 nm) and UV irradiation study was performed on acesulfame, sucralose, saccharin, and cyclamic acid. The UV irradiation degraded all the artificial sweeteners except acesulfame which degraded only with solar radiation; acesulfame degraded at a faster rate (more than three order of magnitude) than the other artificial sweeteners [90]. Acesulfame when exposed to chlorination in a controlled lab study released 10 disinfection byproducts (DBPs), with two new chlorinated products discovered for the first time; five of these DBPs have been reported in chlorinated drinking water samples [91]. Another major class of CECs emanating from life-style use is UV filter. The UV filters are found in many products such as sunscreens, cosmetics, shampoos, and hair dyes for protecting hair and skin against sun radiation. The UV filters are mainly of two types: organic and inorganic. Organic UV filters are predominantly lipophilic species with conjugated aromatic structures. Many UV filters are used in combination in the products. The UV filters have often been detected in environmental waters via swimming activities or indirectly through treated wastewater, releasing in the range of μg/L concentrations [54]. Toxic effects of UV filters have been reported. The bleaching and death of coral reefs have been partly attributed to UV filters; five UV filters benzophenone (BP)-1, BP-3, BP-8, octocrylene (OC), and octyl dimethylp-aminobenzoic acid (ODPABA) were detected in the coral tissues (highest concentrations were reported for BP-3 and BP-8 at 31.8 and 24.7 ng/g of coral tissue, respectively) [35].

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 81 Higher levels were reported during the wet season, implying contribution of UV filters from swimmers. BP-3 exceeded the cutoff mark for larval deformities and death in >20% coral samples; bioaccumulation factors ranged from 2.21 to 3.01 in soft coral tissue. In seawater BP-3 was reported to be 25.8 ng/L. Ekowati et al. [92] reported 14 UV filters in 17 swimming pools across Catalonia, Spain with 1-H-benzotriazole (1HBT) being most frequent and 4-methylbenzylidene camphor (4-MBC) at the highest level (up to 69.3 ng/L). Vila et al. [93] detected 22 UV filters with concentration reaching up to 540 μg/L. A detailed review of organic UV filters detected in WWTPs, and their fate in wastewater and sludge samples has been studied by Ramos et al. [33]. Application of synthetic musk compounds as fragrance additives has become widespread in many consumer products such as perfumes, laundry detergents, lotions, deodorants, and sunscreens. Structurally, they can be nitroaromatic (musk xylene: 1-tert-butyl-3,5-dimethyl2,4,6-trinitrobenzene), musk ketone (4-tert-butyl-3,5-dimethyl-3,5-dinitroacetophenone), or polycyclic [e.g., tonalide, galaxolide, celestolide, cashmeran, phantolide (as trade names)]. They have been increasingly reported in environmental specimens, including humans and wildlife, and consequently they have caused considerable concern with respect to their toxicity. Due to its high lipophilic character, musk tends to accumulate in biota, sludge, and sediments. Levels up to 190 ng/g of lipids have been detected in humans [94]. Galaxolide was the most abundantly detected musk analyzed in wastewater, with concentrations as 2069 and 1432 ng/L in influent and effluent, respectively [94]. The other musk compounds detected were Cashmeran, phantolide, and tonalide in the concentrations up to 94, 26, and 88 ng/L respectively. Despite nitromusks being banned for use in cosmetics in Europe, musk xylenes and musk ketones were still found in wastewater and rivers [94]. Wombacher and Hornbuckle [95] studied eight synthetic musk compounds in influent to a DWTP, and their attenuation during different stages of conventional treatment scheme via lime softening. They reported the presence of galaxolide in 100% of the samples; in posttreatment processes, galaxolide and tonalide were detected at 2.2 and 0.51 ng/L (average concentrations), respectively. Sumner et al. [96] examined inputs and distributions of synthetic musk fragrances in estuarine and coastal environments; high levels of galaxolide (2098 ng/L) and tonalide (159 ng/L) were detected in wastewater treatment effluents which are released into the Tamar and Plym Estuaries (United Kingdom). Polycyclic musks were measured in water, sediment, and fish from Upper Hudson River (New York); bioaccumulation was noticed in several fish species, and 26 and 23 ng/L of galaxolide and tolamide, respectively, were detected in water [97]. Clara et al. [98] determined household to be the major source of polycyclic musk fragrances in the influent to WWTP. In the past 50 years, the use of phthalates has increased considerably in consumer products. Average global production of phthalates has been estimated to be 3 million tons per year [3]. Phthalates are commonly used as plasticizers in plastics, for example, PVC, to improve their flexibility and cold resistance. They are also used as fixative agents in cosmetics. The most

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commonly used phthalate is di-2-ethylhexyl phthalate (DEHP) particularly for fragrances, food containers, catheters, and blood bags [4, 99, 100]. Among phthalates, only six molecules are regularly analyzed in the wastewater. The DEHP is the most commonly used phthalate, with its concentration in the influent and effluent of treatment plants being highest. The removal rate for the most phthalates studied is >90%; diisobutyl phthalate (DiBP) has a much lower removal rate, the reason being scarce data on its occurrence [3]. 4.3.1.3 Endocrine disrupting compounds and hormones Certain substances, both synthetic and natural, by mimicking hormones can interfere or disrupt normal hormonal functions. Due to their ecotoxicological risks, Endocrine disrupting compounds (EDCs) pose grave concerns. These natural as well as man-made substances have been known or predicted to influence the endocrine system. The examples include natural estrogens (17ß-estradiol, estrone), natural androgens (testosterone), phytosteroids (17ß-sitosterol), isoflavenoids (daidzeine), synthetic estrogens (17α-ethinylestradiol), pesticides (atrazine), phthalates, alkylphenol ethoxylate surfactants, dioxins, coplanar polychlorinated biphenyls (PCBs), parabens (hydroxybenzoate derivatives), bisphenol A, and organotins. Since these chemicals belong to different classes, they therefore possess variable action and potencies for endocrine disruption. The EDCs have been detected in biological samples. Markman et al. [101] determined levels of four EDCs (E2, dibutylphthalate, dioctylphthalate, and bisphenol A) in earthworms exposed to sewage effluents. Significantly elevated levels of these EDCs were found in the worms residing in sewage percolating filter beds. This implies that earthworms can be utilized as bioindicators for EDCs. Noguchi et al. [102] detected 2,6-dimethyl-4-nitrophenol and 1-hydroxy-pyrene (known estrogenic compounds) in diesel exhausts. Sarmah et al. [103] studied the occurrence of natural and synthetic estrogenic compounds in sewage and animal effluents in a dairy region of New Zealand, finding high level of E2 (19–1360 ng/L), and its degradation product E3 (41–3123 ng/L) in the dairy effluents. The overall load for the five EDCs measured varied between 60 and >4000 ng/L. In lagoons associated with animal husbandry (pigs, poultry, and cattle rearing) several estrogens were identified [104]. The EDCs have also been discovered in air [105], and in cereals [106]. Loos et al. [107] detected octyl- and nonylphenols, their ethoxylate and carboxylate in textile industry and WWTPs effluents and surface waters. Nonylphenol, nonylphenol ethoxylates, triclosan, and bisphenol A were measured in wastewater and sewage sludge [108]. Nevado et al. [109] detected atrazine, simazine, propazine, ametryn, prometrun, terbutryne, and three chloro-s-triazine in surface waters. Several studies attribute EDCs for the decline in certain species, such as heightened sterility in American alligators, change of sex in fish and shellfish [7]. 4.3.1.4 Flame retardant-based CECs Brominated flame retardants are used in many household products such as children’s sleepwear, foam cushions in sofas and chair, electronics, computers, and plastics. Particularly in polyurethane foams, brominated flame retardants roughly constitute up to 30% by weight.

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 83 Although flame retardants make human life safer, in environment they pose ecotoxicological risks, mainly due to their persistence and transport over longer distance. Polybrominated diphenyl ethers (PBDEs) were once considered a popular flame retardant, but currently they are banned in several countries. However, PBDEs are still manufactured in China, and can be potentially present in many products already sold. Therefore, their release via these products remains a matter of concern [54]. There are new replacements as well for PBDEs available in the market and detected in the environment (e.g., organophosphates and chloroorganophosphates). [110, 111]estimated the environmental fate of 94 halogenated and organophosphate PBDE replacements. The study reported that roughly 60% of the new replacements had persistence and long-distance transport similar to their PBDEs predecessor. Guo et al. [39] detected Marbon, an isomeric form of Dechlorane Plus, in 59 samples from five Lake Michigan tributaries. Teo et al. [37] investigated public swimming pools (15 in number, including spas, indoor, and outdoor pools) for flame retardants; concentrations ranged from 5 to 1180 ng/L. The indoor pools showed higher levels than outdoor pools, and swimsuits were observed to contribute to the presence of flame retardants. Gu et al. [38] analyzed flame retardants in aquaculture; tetrabromobisphenol A (TBBPA) level in the seawater was found to be the highest (up to 2.8 ng/L). Flame retardants have been found in the environmental waters, aquatic organisms, eggs, wastewater, landfill leachate and sediment, and breast milk, with flame retardants detected in distant regions due to long-distance transport. [110, 111]studied photochemical transformation of five brominated flame retardants. Zhang et al. [112] detected two flame retardants [tetrabromobisphenol A and Bis(2-hydroxyethyl)-ether] in the environmental water in levels up to 7.7 μg/L. 4.3.1.5 Pesticides and herbicides Contamination of herbicides and pesticides remains the focus of numerous environmental research. Research has shown that TPs of pesticides formed via hydrolysis, oxidation, biodegradation, or photolysis have been detected at much higher concentrations than their parent pesticides and are as much, and in some cases, even more toxic. Many TPs are currently on the watch list by EPA under CCL. An extensive review was published by Vidal et al. [113] on the detection of TPs (>100 TPs of 49 pesticides) in the environmental, biological, and food samples. Benvenuto et al. [114] determined triazine and its TPs in surface water and wastewater samples. Kowal et al. [40] investigated N,N-dimethylsulfamide, a polar pesticide TP in environmental waters; >600 samples of drinking water, surface water, and groundwater were analyzed in the Rhine and Ruhr region of Germany; roughly 65% of the samples contained quantifiable levels with concentrations up to 63 μg/L. Loos et al. [28], while studying environmental water samples across Europe, reported that the pesticides TPs were the most frequently detected analytes, with many analytes showing the highest concentrations in the groundwater samples; for example, 55 samples (49%) showed concentration levels as high as 487 and 266 ng/L for desethylatrazine and desethylterbutylazine

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respectively. Buttiglieri et al. [42] analyzed >500 groundwater, surface water, and wastewater samples for the occurrence and degradation of N-chloridazon; the pesticide concentration was estimated up to 0.89 μg/L, and the degradation product—desphenyl-chloridazon (DPC) was noticed at much elevated levels, up to 7.4 μg/L. Under aerobic conditions, N-chloridazon was 100% transformed to DPC, which remained stable for 98 days. In a study in Southern France, MCPA [(4-chloro-2-methylphenoxy)acetic acid] transformed by photolysis via following pathways: MCPA ! 4-chloro-2-methylphenol (CMP) ! 4-chloro-2-methyl-6-nitrophenol (CMNP). The CMNP was noted to be more environmentally persistent than its parent compound. Serious concerns have been raised over the genotoxic effects of nitro compounds. The azole fungicides are potentially harmful as their mode of action can disrupt endocrine systems of various organisms. Pelosi et al. [115] demonstrated the toxic effect of indiscriminate use of pesticides on earthworms. The toxic effects of pesticides on spiders [116], carabid beetle [117], pollinators like honey bee, bumble bees, fruit flies [118], and humans [119] are well documented. 4.3.1.6 Per- and polyfluoroalkyl substances-based CECs Per- and polyfluoroalkyl substances (PFASs) most commonly observed in the environment mainly include perfluorooctanoic acid (PFOA), perfluorosulfonate (PFOS); other PFACs are 4-carbon-PFASs [e.g., perfluorobutanoic acid (PFBA), perfluorobutanesulfonate (PFBS)] and GenX (perfluoro-2-propoxypropanoic acid). In the last 2 years, many new classes of PFASs have been discovered, for example, new polyfluorinated carboxylic acids and sulfates, new anionic, cationic and zwitterionic PFASs, and perfluoroalkyl ether carboxylic acids and sulfonic acids. The PFASs make the strongest bond (CdF) in chemistry, and repel both grease and water. They are used in nonstick cooking utensils, food packaging, carpets, fabrics, personal care products, paints, firefighting foams, and adhesives. It is estimated that >3000 PFASs are currently available in the global market [120]. High stability of PFASs and their potential to accumulate in red blood cells raise concerns about their fate in the environment. Almost every person has some measurable quantities of PFOS and PFOA in their blood [54]. The PFOA and PFOS currently feature on CCL-4 and six PFASs on UCMR-3. The EPA has issued a health advisory in November 2016 for PFOA and PFOS combined concentration <70 ng/L in drinking water. The PFASs can potentially bioaccumulate in plants and crops after PFAS-containing biosolids are used as fertilizers. The PFASs are also mobile in groundwater and have been detected in drinking water worldwide [54]. The PFASs have been detected in cats (particularly living indoors) [121], fast food packages (400 samples studied, for example, sandwich wrappers, pastry bags, French fry box, pizza box, paper cups, beverage containers; 33% of the samples showed fluorine in them with a significant number of total fluorinated compounds were unknown polyfluorinated substances) [122], surface water (12 novel perfluoroalkylether

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 85 carboxylic acids and sulfonic acids detected in North Carolina) [123], drinking water [124], landfill leachates [125], snow deposition in Antarctica [126], and oceans [127]. The PFASs raises many health concerns including cancer, reproductive failures, developmental retardation, endometriosis, immunotoxicity, bioaccumulation, thyroid, colitis, and ulcer-based diseases. The PFASs have also been reported to transfer from mother to babies via breast milk and placenta. The PFASs have also been found in elevated levels in seals, polar bears, and killer whales in location much remoter than where these substances are manufactured and used. 4.3.1.7 Disinfection by-products The Disinfection by-products (DBPs) are the unintended result of destroying harmful pathogens in drinking water by the reactions of disinfectants (chlorine, ozone, chloramines, chlorine dioxide, UV) with organic matter, bromide, and iodide. Out of 700 identified DBPs, only 11 are currently regulated by EPA. The precursors to DBPs are mainly anthropogenic substances in addition to the main precursor natural organic matter (NOM). The DBPs have been reported from linear alkyl benzenesulfonate surfactants, UV filter, flame retardant, tamoxifen metabolites, drugs, benzotriazoles, and benzothiazoles [54]. Tang et al. [128] identified 116 peptide-based DBPs in finished drinking water, which were not present in the raw source water. Kimura et al. [129] discovered on five human volunteer subjects that those consuming regular tap water as compared to spring water showed greater total organic chlorine (TOCl) and total organic iodine (TOI) content than the controls. However, total organic bromine (TOBr) content was slightly lower. [70–72]identified > 500 compounds in chlorinated, chloraminated, and ozonated drinking water samples, with 41 being confirmed as DBPs. Hladik et al. [130] discovered a novel class of iodine-based DBPs, due to the use of iodine sanitizers in dairy farms. Iodo-trihalomethane (THMs) were detected within dairy facilities and surface water receiving dairy wastewater across three states. Han et al. [131] discovered new Cl-, Br-, and I-DBPs in water samples treated with chlorine dioxide. Postigo et al. [132] reported two new DBPs (chloroiodomethane and ethyl iodoacetate) in chlorinated and chloraminated water samples. The DBP formation and occurrence in drinking water was comprehensively studied by Richardson [133]; the study listed more than 600 DBPs resulting from various disinfection and disinfectants combinations. The health risks from DBPs include miscarriage, birth defects, and cancer (bladder and colorectal), with bladder cancer appearing most consistently in many studies. Regli et al. [134], after studying 201 DWTPs samples, concluded that increase of bromide concentration over 50 μg/L causes potential risk of cancer in populations being served by 90% of these DWTPs. A study by Wright et al. [135] reported a correlation between cardiac birth defects in babies and specific DBPs in Massachusetts. The study made a case for further examination of brominated acetonitriles for future epidemiological investigations. [70–72] determined development of antibiotic resistance in E. coli at high as well as low concentrations of the two DBPs—iodoacetic acid and chlorite. Wagner and Plewa [136] published the most

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toxicology data on DBPs by studying 103 DBPs for mammalian cell cytotoxicity and genotoxicity. 4.3.1.8 Nanomaterials Nanomaterials have at least one dimension in the length scale of 1–100 nm. Nanomaterials continue to be the focus of research with respect to material development, occurrence, and fate in the environment and toxicology. Nanosilver (nAg) as nanoparticles remains the most dominant nanomaterial, with application ranging in medical bandages, socks, textiles, food containers, baby toys, blankets, and towels. Other nanomaterials include graphene, nZnO, nTiO2, nCeO2, nAu, and nFe(0). nAg has been detected in laundry wash water [137] and biosolids of WWTPs [138]. Markus et al. [139] in a modeling study involving the transport of ZnO, TiO2, and Ag nanoparticles along the Rhine River noted that nanoparticles can be transported over long distances, akin to suspended particulate matter. Canesi and Corsi [140] presented a summary of the effects of nanoparticles on aquatic invertebrates. The study observed the accumulation of nanoparticles in the gills and digestive tracts leading to immune modification, oxidative stress, and embryo poisoning. Accumulation of nanoparticles has been demonstrated in carrots by Ebbs et al. [141]. Low bioaccumulation of carbon nanotubes was noticed in plants, invertebrates, and nonmammalian vertebrates [142]. 4.3.1.9 Microplastics Microplastics have been reported for growing contamination in ocean, lakes, and rivers around the world [54]. They are 5 mm or smaller in size, and result from degradation of bigger plastic materials that enter the environment or from their direct application in products. A new piece of legislation, the Microbead-Free Waters Act, bans all microplastics in cosmetic products or toothpaste produced or sold in the United States. The accumulation of microplastics has been taking place for a long time, with the maximum accumulation observed in ocean gyres, leading to the formation of plastic islands. The intake of microplastics by seabirds, seals, sea lions, whales, dolphins, zooplankton, and marine reptiles are well reported [54]. The concerns arising from its ingestion include obstruction of digestive tracts and as potential vectors for chemical contaminants already on their surfaces. Potential impact on human health has been extensively covered in a review by Wright and Kelley [143]. 4.3.1.10 Dioxane 1,4-Dioxane is used in industries as a solvent stabilizer in the production of numerous products. It is a probable human carcinogen. Some new data have been reported on its prevalence in environmental waters [54]. It is currently listed in CCL-4 by EPA. 1,4-Dioxane is highly soluble and stable in water, therefore, its treatment is vexing.

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 87 Adamson et al. [144] analyzing the data from 4864 DWTPs reported the presence of 1,4-dioxane in 21% of the plants, in 7% of them, the concentrations exceeded the health reference level (HRL) for cancer (0.35 μg/L). Among 28 contaminants listed in UCMR-3, it ranks second among the most detected contaminant in the environmental samples and also surpassed the HRL. The states having highest detection frequency are California, New Jersey, Pennsylvania, Illinois, and Ohio; South Carolina had the highest percentage of plants where 1,4-dioxane was detected (55%) [54]. Sun et al. [145] reported in surface water the concentration range of nondetected contaminant to be 436 μg/L (Cape Fear River watershed, North Carolina) and in the WWTPs effluent the level was as high as 1405 μg/L. In Llobregat River, Spain, 1,4-dioxane level ranged between 0.29 and 241.1 μg/L. 4.3.1.11 Algal toxins Algal toxins result from algal blooms. They include (i) fresh water microcystins, nodularins, anatoxins, cylindrospermopsin, and saxitoxins, and (ii) seawater brevetoxins. Microcystins are the most commonly reported, and they have numerous variants, the most common type contain the amino acids leucine and arginine in its structure [54]. They are currently listed in CCL-4 by the EPA. A comprehensive review on the microcystins in estuarine and marine waters was recently published by Preece et al. [78]. Drobac et al. [79] studied water samples and fish from 13 fish ponds in Serbia during a major algal bloom. The study reported the presence of saxitoxins, microcystins, and nodularin. The fish were diagnosed with liver, kidney, gills, muscles, and intestine damage caused due to the algal toxins. Microcystins-LR (MC-LR) was detected in lettuce and arugula with levels higher than recommended by World Health Organization (WHO) [146]. In a nationwide survey for cyanotoxins in 1161 lakes in United States, microcystins, saxitoxins, and cylindrospermopsin were detected in 4%–32% of the samples, with mean concentrations between 0.0061 and 3.0 μg/L. Parker et al. [147] demonstrated that halogen radicals can facilitate the photodegradation of microcystins in estuaries. Lehman et al. [148] showed that intense drought can promote algal blooms and enhance release of microcystins. In The San Francisco Estuary a severe drought led to increase in total microcystins levels that surpassed the previous dry and wet year by factors of 11 and 65, respectively.

4.3.2 Treatment and Remediation Methodologies The detection of CECs in the environment has led to proposals for upgrading the existing WWTPs and implementation of novel technologies for biodegradation of organic matter [e.g., membrane bioreactor (MBR)] have been mooted [149]. The upstream source-separated treatment (specific to waste streams) can be a potential solution. To implement decentralized schemes, treatment technologies are needed for the removal of specific contaminants. This has led to the publication of numerous studies focused on the development of removal and

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remediation technologies for the contaminants. Currently developed technologies fall into the following major categories: adsorption-based schemes, advanced oxidation processes (AOPs) using hydrogen peroxide or radiation and ozone-based oxidation, UV radiation, gamma radiation, and electrooxidation. The above treatment methodologies have also been used in conjunction with other technologies like oxidation process combined with nanofiltration and reverse osmosis [150]. The choice of the technologies or their combination depends on the quality requirements of the reclaimed water. In this section, we will review each treatment schemes with respect to their performance, and challenges associated with their implementation. 4.3.2.1 Conventional treatment schemes In a typical WWTP, the treatment schemes comprise a primary physicochemical treatment step followed by a secondary step in a biological reactor formed by activated sludge. Deblonde et al. [3]and Pal et al. [31] summarized the data on the removal of CECs in WWTPs. It is evident that the conventional treatment in WWTPs possesses limited capacity to eliminate CECs from wastewaters. It is mainly attributed to the inability of microorganisms to metabolize CECs as a source of carbon and to their inhibition in the presence of CECs [151]. However, this aspect remains to be further investigated. Some CECs that can be removed in primary treatment steps by adsorption include ibuprofen, naproxen, sulfamethoxazole, and iopromide; further biological treatment steps eliminate 30%–75% of antibiotics and antiinflammatories [152]. From the most frequently used analgesics, paracetamol, acetylsalicylic acid, and ibuprofen can be removed by WWTPs, but diclofenac was very poorly eliminated [153]. Daughton and Ternes [11] reported the spontaneous hydrolysis of penicillins in water and precipitation of tetracyclines (TCs) with Ca2+ resulting in its accumulation in sludge treatment plant. The MBRs have been reported to improve the CECs elimination performance before disinfection in a conventional WWTP (96% removal as compared to 85% removal without MBR) [154, 155]. The application of chlorine remains the most widely used conventional treatment method for disinfecting drinking water. Glassmeyer et al. (2017) on surveying the influent and effluent samples from 29 DWTPs concluded that many CECs are incompletely removed during treatment and persist in the distributed water used for potable purposes. Moreover, concerns are also being raised for some of DBPs resulting from chlorination. It is quite clear that the present WWTPs and DWTPs are designed for the treatment and removal of contaminants and nutrients that come under the existing regulations. However, the occurrence of CECs in the waste streams requires advanced treatment. In addition, also the wide variation in properties among CECs requires multitude of treatment methodologies, each capable of exploiting varying properties for effective remediation scenario.

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 89 4.3.2.2 Adsorption-based removal processes Adsorption is a surface phenomenon. Removal of CECs by adsorption involves physicochemical interaction between the adsorbent (e.g., activated carbon) and the adsorbate (CECs). The most popular adsorbent used for eliminating pharmaceutical CECs is activated carbon. Though the exact mechanism of removal by activated carbon remains unknown, experts have suggested the following mechanisms [151]—(i) dispersive interaction between the electrons of graphene plane of activated carbon and the aromatic ring, (ii) electron donoracceptor interaction (π-π) with carbonyl surface group acting as the electron donor and the aromatic ring of CECs as the acceptor, and (iii) electrostatic/dispersive interaction bridged by hydrogen bonding. The uncertainty around the adsorption mechanism of aromatic compounds over activated carbon has fueled the impossibility of ever achieving a general mechanism of the process. However, a few observed rules exist wholly as working principles, if not mechanistically, for understanding and designing the process. Radovic et al. [156] summed them up as (1) the process parameters such as adsorbate solubility, adsorbate and adsorbent hydrophobicity, and the strength of interactions govern the process efficacy, (2) the interaction strength can be modulated by varying the adsorbate or adsorbent aromatic rings’ functional groups, and (3) the pH of the medium has a vital role in the adsorption process. Granular activated carbon has also been reported to effectively remove a large number of CECs (including nonaromatic CECs) in packed-bed mode, with breakthrough curves demonstrating that the CECs with greater hydrophilicity breached activated carbon faster than hydrophobic compounds [157]. The key advantage of the adsorption-based CECs removal is that it does not lead to the formation of toxic or pharmacologically active by-products. The available literature shows that activated carbon has typically shown a high adsorption capacity for eliminating CECs [158, 159]. However, the adsorption by activated carbon suffers from certain deficiencies, for example, slow pollutant uptake (in the order of hours) [160, 161] and poor elimination of many relatively hydrophilic contaminants [162]. Additionally, the regeneration process of the spent activated carbon is energy intensive, needing heating to 500–900°C, and the performance is also not fully restored [163, 164]. Therefore, many other adsorbents have been studied for removing CECs. Rossner et al. [165] studied and compared the performance of four different adsorbents—activated carbon, carbonaceous resin, and two high-silica zeolites—for the removal of a mixture of 28 CECs (at environmentally relevant concentrations,
200–900 ng/L) from lake water. The results indicated that of the four adsorbents, coconutshell-based activated carbon (dose: 10 mg/L) is the most effective in removing the contaminants (98% removal for 24 of the 25 stable CECs in the mixture). The greater ˚ activated carbon performance was related to its larger pore volume with width in the 6–9 A range. In the case of the zeolite, the study found that its uniform pores do not provide an effective broad-spectrum barrier against the variety of contaminants utilized in the study. [158] tested activated carbon, carbon nanofiber, carbon nanotube, and high surface area graphite

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for the removal of nalidixic acid (NAL), 1,8-dichlorooctane (DCO), and methyl-phenoxyethanol (MPET). The study established that the role of adsorbate chemical properties as well as the adsorbent morphology were influencing parameters for determining the adsorption capacity of the adsorbents. The study also reported that for CNT with curved graphene layer, the aromaticity of the adsorbates was a determining factor, whereas for the remaining cases, the functional groups in the aromatic rings influenced the uptake behavior of the adsorbents. Bhadra and Jhung [166] synthesized a series of metal-azolate frameworks and subsequently pyrolyzed them to obtain porous carbonaceous material and used them as adsorbents for the removal of CECs. The study established hydrogen bonding to be the ´ lvarez et al. [167] synthesized carbon xerogels by governing adsorption mechanism. A polycondensation of resorcinol with formaldehyde (molar ratio 1:2) and modified them to remove diclofenac and caffeine. For caffeine removal, the adsorbent (treated with urea) showed an uptake capacity of 182.5 mg/g, whereas in the case of diclofenac, the sulfuric acid-treated xerogels removed 80 mg of diclofenac per g of its weight. An excellent review has been presented on biochar as alternative adsorbent for CECs removal by Mohan et al. [168]. Alsbaiee et al. [169] reported a novel adsorbent prepared from ß-cyclodextrin polymer that showed improved kinetic and equilibrium parameters as compared to activated carbon. ß-cyclodextrin quickly removed a variety of CECs with adsorption rate constants 15–200 times greater than those of activated carbon. Also, the adsorbent was regenerated many times using a mild washing step without any loss in its performance. Some studies have also been reported on the application of silica-based adsorbents [169a], superparamagnetic nanoparticles [170], orange peels [171], premixed polyaluminum chloride (PACL) with carbonaceous matter for removing CECs. The summary of the results of major studies in the adsorption-based CECs remediation is presented in Table 4.3. 4.3.2.3 AOP-based treatment schemes The AOPs are all based on the free radical generation (HO , O2  ,HO 2 ). These highly reactive species successfully attack most organic molecules, with enhanced reaction rate constants ranging between 106 and 109 M1 s1. The AOPs have been successfully used for the oxidation of numerous organic and inorganic molecules and are therefore versatile in application. There are three major variants of AOPs, namely AOPs based on ozone, UV, and gamma radiation. 4.3.2.3.1 Ozone-based AOPs

In recent times, a great deal of research focus has been placed on the application of O3, O3/OH, O3/H2O2, and O3/activated carbon systems for the elimination of CECs. The process is highly reactive with reduced bromate formation in the case of O3/H2O2 oxidation as compared to ozone alone [77]. Snyder et al. [172] in a comprehensive study using ozonation, and O3/H2O2 evaluated the removal efficiency for 16 CECs. Ikehata et al. [173] evaluated the kinetic

Table 4.3 Summary of remediation methods for CECs Process/Material Details

CEC Treated

Performance

Remark

Reference

Adsorption

Activated carbon

Nitroimidazoles

Adsorption capacity: 1.04–2.04 mmol/g

Rivera-Utrilla et al. [159]

Activated carbon, carbon nanofibers, and nanotubes and graphite

NAL, DCO, and MPET

Increase in adsorption rate with decreasing oxygen content and increasing hydrophobicity of the adsorbents The adsorbates’ chemical properties and the adsorbents’ morphology affected the adsorption capacity H-bonding determined to be the main adsorption mechanism

Bhadra and Jhung [166]

The highest adsorption capacity for coconut-shellbased activated carbon was attributed to its larger pore ˚ width volume with 6–9 A Faster rate kinetics obtained as compared commercial activated carbons

Rossner et al. [165]

Faster uptake kinetics for a variety of CECs with rate constants 15–200 times that of activated carbons Adsorption capacity strongly dependent on adsorbents’ morphology, and oxygen functionality

Alsbaiee et al. [169]

Different pyrolyzed metal-azolate frameworks and activated carbon Activated carbon, carbonaceous resin and two high silica zeolites Carbon xerogels prepared by polycondensation of resorcinol and formaldehyde Porous β-cyclodextrin polymer

Multiwall carbon nanotubes

Adsorption capacity for all the adsorbents followed the order: DCO ≫NAL >MPET Salicylic acid, clofibric Adsorption capacity was acid, diclofenac in the range of 160– sodium, bisphenol A 440 mg/g and oxybenzone 98% removal by coconut28 CECs in the shell-based activated concentration range of 200–900 ng/L from carbon for 24 out of 25 lake samples stable CECs Caffeine and diclofenac

Adsorption capacities of 182.5 and 80 mg/g were reported for caffeine and diclofenac respectively

8 CECs including bisphenol A, S, and metolachlor

Adsorption capacity for bisphenol A was 88 mg/g

NAL, DCO and MPET

Highest adsorption capacity reported was 380 mg/g

˜o et al. Patin [158]

´ lvarez et al. A [167]

˜o et al. Patin [158]

(Continued)

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 91

Method

AOPs

Summary of remediation methods for CECs—Cont’d

Process/Material Details

CEC Treated

Performance

Remark

Reference

Activated biochar

Benzophene (BZP), benzotriazole (BZT), bisphenol A (BPA) and 17 ß-estradiol (E2)

Biochar exhibited 5%– 30% greater removal efficiency than commercial activated carbon

Kim et al. [171a]

Premixed poly aluminum chloride (PACL), powdered activated carbon and carbon nanotubes Direct ozonation on 10 secondary wastewater samples Direct oxidation by O3, and O3/H2O2

Salicylic acid, ibuprofen and diclofenac

Adsorption capacity increased 5 times after adding PACL to the adsorbents

Removal efficiency followed the order: E2 > BZP > BPA> BZT with biochar showing the higher adsorption capacity than commercial active carbon for all the contaminants The hydrophobicity of the contaminants played a dominant role in their adsorption

19 CECs and and microorganisms

Ozone with and without H2O2 oxidized the CECs

Clofibric acid, ibuprofen, and diclofenac Nitroimidazole

98% removal efficiency for 5 mg/L O3 and 1.8 mg/L H2O2 Enhanced reduction observed in total organic and toxicity 95% TOC removed in 120 min with complete elimination of toxicity

Ozonation + adsorption on activated carbon Ozonation + adsorption on activated carbon

Ozone can be utilized for standalone oxidation process or in association with H2O2. The major advantage of the process is its nonselectivity

Jiang et al. [171b]

Snyder et al. [172] Ikehata et al. [173]

Nitroimidazole showed high Sanchez-Polo reactivity toward hydroxyl et al. [174] radicals Diclofenac Simple ozonation without Beltran et al. activated carbons achieved [175] only 40% TOC removal in 120 min Ozonation +activated Tetracyclines (TCs) Complete TC In conjunction with activated Gomez-Pacheco carbon+ degradation within carbon, ozonation resulted in et al. [176] H2O2 + biological 10 min higher TC, TOC, and toxicity treatment removal Okamoto et al. Phthalate esters (PEs) Pes degraded to PE-4OH By PE-mediated UV radiation with [177] under UV radiation with photosensitization H2O2 is H2O2 generated from O2 and H+ and and without H2O2 decomposed to hydroxyl

Chapter 4

Method

92

Table 4.3

Table 4.3 Method

Process/Material Details

Summary of remediation methods for CECs—Cont’d CEC Treated

Photodegradation in the presence of TiO2 and natural organic matter (NOM)

Clofirbic acid, carbamazepine and iomeprol

Photodegradation with sulfate radical

Cytarabine antineoplastic

Gamma radiation

Nitroimidazoles

Remark

Reference

Rapid photooxidation observed

radical, thus oxidizing the PE benzene ring. The presence of oxygen doubled the direct photolysis rates of naproxen and propranolol

Lin and Reinhard [178]

39% mineralization in 90 min; the photoFenton system showed 100% oxidation in 60 min Transformation, deactivation and minimization of persistent compounds were achieved Degradation up to 99% was achieved

Total mineralization with 100 min of solar irradiation

The presence of NOM and other organic substances retarded the photocatalysis of a specific persistent substance

Sulfate radical more effective than hydroxyl radical in the photodegradation Higher degradation under Slightly lower removal rate in subterranean and acidic condition than basic or neutral media surface waters than in ultrapure water and was markedly lower in wastewater.

Vogna et al. [179], Ravina et al. [180], Perez-Estrada et al. [181] Doll and Frimmel [182, 183]

Ocampo-Perez et al. [184] Sanchez-Polo et al. [185]

(Continued)

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 93

Gemfibrozil, Direct ibuprofen, photodegradation with a xenon arc lamp ketoprofen, naproxen, and propranolol (765 W/m2; 290 nm estriol, estrone, 17β< λ < 700 nm) estradiol, and 17αethinylestradiol UV/H2O2; photoDiclofenac Fenton system

Performance

94

Table 4.3

Summary of remediation methods for CECs—Cont’d

Process/Material Details

CEC Treated

Performance

Remark

Reference

Electrochemical oxidation

Stirred tank reactor; 60 mA for 480 min; Ti/ SnO2 anode; 0.2 M Na2SO4 as electrolyte; pH 6.2 Stirred tank reactor; 350 mA for 270 min; BDD anode; 0.1 M Na2SO4 as electrolyte; pH 6 Stirred tank reactor; 300 mA for 420 min; Pt BDD anode; 0.05 M Na2SO4 as electrolyte; pH 3 Stirred tank reactor; 300 mA for 360 min; Pt BDD anode; 0.1 M Na2SO4 and 0.1 M NaCl as electrolyte Stirred tank reactor; 80 mA for 300 min; Ti/ RuO2 anode; 0.1 M NaCl as electrolyte Stirred tank reactor; 20 mA for 300 min; BDD anode; 0.1 M NaClO4 and 0.1 M NaCl as electrolyte

17 α-ethinylestradiol

79% TOC removal

Feng et al. [185a]

17 ß-estradiol

94% TOC removal

Yoshihara and Murugananthan [185b]

Clofibric acid

36% TOC removal

Sires et al. [185c]

Diclofenac

46% TOC removal

Brillas et al. [185d]

Paracetamol

28% TOC removal

Boudreau et al. [185e]

Sulfamethoxazole

78% TOC removal

Boudreau et al. [185f]

Chapter 4

Method

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 95 parameters for the oxidation of pharmaceutical in wastewater by O3 and O3/H2O2. High efficiency (98%) was observed for 5 mg/L of O3 and 1.8 mg/L H2O2 in the mineralization of clofibric acid, ibuprofen, and diclofenac; oxidation with O3 yielded only 32% removal of diclofenac. The major disadvantage with O3/H2O2 oxidation systems is their nonselectivity. The organic matter also competes for OH radical and therefore reduces the oxidation efficiency of the target species. Many studies have also focused on the application of ozonation in conjunction with activated carbon system to eliminate organic matter from water, however, very few works have focused exclusively on removing the pharmaceuticals. [174]using ozonation and activated carbon reported increased effectiveness as compared to ozonation or O3/H2O2 alone in removing nitroimidazole; the study also reported enhanced reduction in total organic and toxicity. Another study reported removal of diclofenac from water using ozonation with activated carbon in contrast to direct ozonation where removal was not noticed [175]. Go´mez-Pacheco et al. [176] also compared the effectiveness of ozone and schemes that utilize simultaneously activated carbon, hydrogen peroxide, and biological treatment for the elimination of TCs from water. In conjunction with activated carbon, ozonation resulted in higher TC removal and reduction in TOC and toxicity was observed. In Table 4.3, the key results on ozone-based AOPs are summarized. 4.3.2.3.2 UV radiation-based AOPs (Photolysis)

Hydroxyl radicals can also be generated by photochemical degradation of certain parent compounds. The degradation is a function of radiation wavelength (λ), compounds molar absorption capacity, and quantum yield [186]. In photolysis, the most commonly used radiation has λ in the range of 200–400 nm (UV spectrum region). The molecule degradation depends on its capacity to absorb given radiation and reach an excited state, leading to bond rupture and degradation. If molecule is not directly degraded by photolysis, then indirect degradation by radical formation commences. The major advantages of photolysis over nonphotochemical degradation are (1) some of the contaminants are also degraded by direct photolysis, (2) additional chemical reagents are not needed, (3) it leads to reduction in the quantity of certain oxidants needed in combined systems, and (4) it is less impacted by drastic change in pH. The UV radiation is particularly helpful in removing pharmaceuticals as most of these compounds are photoactive, absorbing luminous radiation. Various studies have shown that many pharmacologically active compounds undergo photodegradation, since they possess aromatic rings, heteroatoms, and certain functional groups that permit absorption of solar radiation or initiate reaction by photosensitive species that cause their photodegradation in water [178, 187].

96

Chapter 4

The photodegradation is further kinetically improved by the addition of H2O2 during photooxidation, as highly reactive hydroxyl radicals reduce the required dose of UV radiation [188]. Okamoto et al. [177] analyzed the removal of phthalate esters under UV light and sunlight irradiation. They also investigated the photodegradation products and their estrogenic activities, concluding that these endocrine disruptors acquire unequivocal estrogenic activity by light exposure. Naproxen undergoes photooxidation [178], however, its TPs are more toxic than its parent compound [189]. Similarly, TPs of triclosan and triclocarban photodegradation are more toxic and persistent in water [190]. Rapid decomposition of diclofenac by photooxidation has been studied by many investigators, implying that this pathway describes its degradation mechanism [191–193]. UV/H2O2 was reported to degrade diclofenac with 39% mineralization in 90 min [179], whereas the photo-Fenton system showed 100% oxidation in 60 min for the same drug and achieved total mineralization with 100 min of solar irradiation [180, 181]. The photodegradation of clofibric acid into many aromatic and aliphatic compounds has been reported to be facilitated by the presence of TiO2 [182, 183]. Carbamazepine has been studied by many research groups for photooxidation [179, 182, 183, 194]; epoxycarbamazepine has been determined as a main subproduct with many other TPs, which have not been completely identified. Ocampo-Perez et al. [184] used another radical species (sulfate radical) and showed it has better efficacy than HO• radicals in removing cytarabine antineoplastic. Table 4.3 summarizes major studies on the application of UV radiation-based AOPs 4.3.2.3.3 Gamma radiation-based AOPs (Radiolysis)

Radiolysis involves production of radicals, highly reactive electrons, ions, and neutral species via exposure of water with high-energy electromagnetic waves/radiation, X-rays, gamma radiation from radioactive sources of 60Co and 137Cs or electron linear accelerators [195–197]. The reactive species formed by ionizing radiations via colliding with water molecules are shown in Eq. (4.1) [151].  + H2 O



> 2:8 HO + 2:7 e aq + 0:6 H + 0:72 H2 O2 + 2:7 H3 O + 0:45 H2

(4.1)

The chemical species represented in Eq. (4.1) as the products are the primary radiolytic products that attack and degrade the contaminants’ molecules [198]. The degradation proceeds in the following stages: (i) active free radicals’ formation (in
1010 s timescale), (ii) diffusion of radicals (in 103 s order of time), and (iii) reaction between primary radiolytic products and contaminants’ molecules. The radiolytic processes based on high-energy electromagnetic radiations are reported to be more economic and effective for large-scale applications as compared to other schemes applied for the removal of CECs [199]. Sometimes AOPs based on gamma radiation are described as Advanced Oxidation/Reduction Processes because of the

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 97 nature of species—e aq and H• are strong reducing agents and HO• acts as a strong oxidizing agent—formed during the radiolysis of water molecule [200]. The established efficacy of the process has spawned research into its application for removing CECs [200–202]. In some countries the technology has also been implemented in treatment plants, however, it has yet to achieve any widespread application; the concerns are mainly inadequate safety conditions and knowledge of its performance [200, 201, 203]. Sanchez-Polo et al. [185] studied the degradation of nitroimidazoles in waste and drinking water using gamma irradiation. The study reported higher degradation under acidic condition than basic or neutral media. The work also established an optimum H2O2 concentration for maximum degradation of the pollutant. The degradation efficiency was also observed in the presence of radical scavengers like t-BuOH (HO• scavenger) and thiourea (e aq , H• and HO• scavenger); both the compounds reduced the degradation rate, implying that the decomposition proceeds via two pathways—oxidation by HO• radicals and reduction by e aq and H• radicals. Ocampo-Perez et al. [204] studied cytarabine degradation in the presence

of Cl, CO3 2 , NO3  , NO2  , and humic acid. The study reported decrease in degradation rate in the presence of each species, mainly due to the competition to cytarabine from these species for HO• radicals in the solution. By adding H2O2 as promoter of HO• radicals, the

degradation rate was enhanced. Velo-Gala et al. [205] and [206]reported similar results for the degradation of diatrizoate and TCs, respectively. In Table 4.3, the summary of major studies is provided. 4.3.2.4 Electrochemical oxidation The most commonly applied electrochemical scheme in wastewater treatment is electrochemical oxidation, also known as anodic oxidation. The process is utilized for oxidizing the pollutants in electrolytic cell and proceeds via direct electron transfer to the anode, and then subsequently formed radicals (physically sorbed OH or chemisorbed “active oxygen”) mediate oxidation of the contaminants [207–209]. Based on the presence of these two species, two approaches have been proposed for contaminant degradation—(i) the electrochemical conversion, where chemisorbed “active oxygen” selectively convert the organics to biodegradable molecules like carboxylates, (ii) the electrochemical combustion, where physically sorbed OH mineralizes the organic molecules [151]. The anode material has a pronounced effect irrespective of whether electrochemical conversion or combustion takes place. Comninellis [207] elucidated on the varying behavior by a model that assumes “active” and “nonactive” anodes [e.g., Pt, IrO2, and RuO2 as active anode; PbO2, SnO2, and boron-doped diamond (BDD) as “nonactive anode”]. In both the anodes, represented as A, water is oxidized and forms physically sorbed hydroxyl radical [A(OH)], and this radical subsequently interacts with the surface and transforms itself to into chemisorbed

98

Chapter 4

active oxygen or superoxide, AO. The AO/A pair then becomes the mediator in oxidizing the organic compounds. Conversely, the nonactive anodic surface interacts weakly with A(OH), which facilitates this radical to directly react with organic molecules to achieve 100% mineralization [151]. In anodic oxidation, apart from physically sorbed OH species (which is the strongest oxidant), different reactive oxygen species are also formed—heterogenous OH, H2O2, and O3. The BDD is considered the most optimum anode material for oxidizing organic matters [209]. In the presence of chloride ions, the generated active chlorine species (Cl2, HClO/ClO, and ClO2  ) also become important, since these species can also attack the organic molecules and compete with radicals [209]. Some research groups have also suggested the existence of ClO3  and even the total transformation of chloride ions to ClO4  [210, 211]. In Table 4.3, the key results of electrochemical oxidation of CECs have been summarized.

4.4 Conclusions and Future Directions Significant developments have taken place toward detection, measurement, environmental fate, and removal of CECs. However, many issues remain unresolved or partially understood. Some of them are listed as follows: (a) Many CECs are yet to be identified in the environment. This is evident from the fact that out of 3000 pharmaceuticals, only
150 have been detected. The continued discovery of PFASs in the environment also supports this assertion. (b) Despite some of the CECs being banned (like a few flame retardants), their new replacements continually enter the market. Therefore, strategies need to be developed that target this “Chemical Conveyor Belt” (to borrow the term from Heather Stapleton of Duke University) [77]. (c) Since many CECs undergo transformation and form a variety of TPs, a robust detection strategy should be developed that identifies the unknown compounds from some empirical formula information of the molecules and their fragment ions. This would constitute an enormous task given a large multitude of contaminants could be found in the environment. Therefore, methods should be developed that prioritize the extensive list of CECs and performs surrogate (group) measurements [77]. (d) There should be focused studies on improving AOPs and other treatment technologies. More decentralized schemes should be incorporated in our waste treatment methodologies. (e) For evolving a holistic approach, toxicology, analytical and remediation studies should be conducted in conjunction with one another.

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 99

Appendix A Method 1694: 74 Pharmaceuticals and Personal Care Products in Water, Soil, Sediment, and Biosolids by HPLC/MS/MS PPCP

CAS #

Classification

Acetaminophen Albuterol Ampicillin Anhydrochlortetracycline (ACTC) Anhydrotetracycline (ATC) Azithromycin Caffeine Carbadox Carbamazepine Cefotaxime Chlortetracycline (CTC) Cimetidine Ciprofloxacin Clarithromycin Clinafloxacin Cloxacillin Codeine Cotinine Dehydronifedipine Demeclocycline Digoxigenin Digoxin Diltiazem 1,7-Dimethylxanthine Diphenhydramine Doxycycline Enrofloxacin 4-Epianhydrochlortetracycline (EACTC) 4-Epianhydrotetracycline (EATC) 4-Epichlortetracycline (ECTC) 4-Epioxytetracycline (EOTC) 4-Epitetracycline (ETC) Erythromycin Erythromycin anhydrate Flumequine Fluoxetine Gemfibrozil Ibuprofen Isochlortetracycline (ICTC) Lincomycin Lomefloxacin Metformin Miconazole

103-90-2 18559-94-9 69-53-4 13803-65-1 4496-85-9 83905-01-5 58-08-2 6804 07 05 298-46-4 63527-52-6 57-62-5 51481-61-9 85721-33-1 81103-11-9 105956-97-6 61-72-3 76-57-3 486-56-6 67035-22-7 127-33-3 1672-46-4 20830-75-5 42399-41-7 611-59-6 58-73-1 564-25-0 93106-60-6 158018-53-2 4465-65-0 14297-93-9 14206-58-7 23313-80-6 114-07-8 59319-72-1 42835-25-6 54910-89-3 25812-30-0 15687-27-1 514-53-4 154-21-2 98079-51-7 657-24-9 22916-47-8

Antipyretic, analgesic Antiasthmatic β-lactam antibiotics Chlorotetracycline degradate Chlorotetracycline degradate Macrolide antibiotic Stimulant Quinoxaline antibiotic Anticonvulsant Cephalosporin antibiotic Tetracycline antibiotic Antiacid reflux Quinoline antibiotic Macrolide antibiotic Quinoline antibiotic β-lactam antibiotics Opiate Nicotine metabolite Nifedipine metabolite Tetracycline antibiotic Immunohistochemical marker steroid Cardiac glycoside Antihypertensive Antispasmodic, caffeine metabolite Antihistamine Tetracycline antibiotic Quinolone antibiotic Chlorotetracycline degradate Chlorotetracycline degradate Chlorotetracycline degradate Oxytetracycline degradate Tetracycline degradate Macrolide antibiotic Macrolide antibiotic Quinolone antibiotic SSRI antidepressant Antilipemic Analgesic Chlorotetracycline degradate Lincosamide antibiotic Quinoline antibiotic Antidiabetes drug Antifungal agent (Continued)

100 Chapter 4 Method 1694: 74 Pharmaceuticals and Personal Care Products in Water, Soil, Sediment, and Biosolids by HPLC/MS/MS—Cont’d PPCP

CAS #

Classification

Minocycline Naproxen Norfloxacin Norgestimate Ofloxacin Ormetoprim Oxacillin Oxolinic acid Oxytetracycline (OTC) Penicillin G Penicillin V Ranitidine Roxithromycin Sarafloxacin Sulfachloropyridazine Sulfadiazine Sulfadimethoxine Sulfamerazine Sulfamethazine Sulfamethizole Sulfamethoxazole Sulfanilamide Sulfathiazole Tetracycline (TC) Thiabendazole Triclocarban Triclosan Trimethoprim Tylosin Virginiamycin Warfarin

10118-91-8 22204-53-1 70458-96-7 35189-28-7 82419-36-1 6981-18-6 66-79-5 14698-29-4 79-57-2 61-33-6 87-08-1 66357-35-5 80214-83-1 98105-99-8 80-32-0 68-35-9 122-11-2 127-79-7 57-68-1 144-82-1 723-46-6 63-74-1 72-14-0 60-54-8 148-79-8 101-20-2 3380-34-5 738-70-5 1401-69-0 11006-76-1 81-81-2

Tetracycline antibiotic Nonsteroidal antiinflammatory drug Quinoline antibiotic Hormonal contraceptives Quinoline antibiotic Macrolide antibiotic β-lactam antibiotics Quinolone antibiotic Tetracycline antibiotic β-lactam antibiotics β-lactam antibiotics Antiacid reflux Macrolide antibiotic Fluoroquinolone antibiotic Sulfonamide antibiotic Sulfonamide antibiotic Sulfonamide antibiotic Sulfonamide antibiotic Sulfonamide antibiotic Sulfonamide antibiotic Sulfonamide antibiotic Sulfonamide antibiotic Sulfonamide antibiotic Tetracycline antibiotic Fungicide and parasiticide Antimicrobial, disinfectant Antimicrobial, disinfectant Pyrimidine antibiotic Macrolide antibiotic Macrolide antibiotic Anticoagulant

Method 1698: 27 Steroids and Hormones in Water, Soil, Sediment, and Biosolids by HRGC/HRMS Steroid/Hormone

CAS #

Classification

Androstenedione Androsterone Equilenin Equilin 17a-Ethynyl estradiol (EE2) Desogestrel Mestranol Norethindrone Norgestrel Campesterol beta-Sitosterol Stigmasterol

63-05-8 53-41-8 517-09-9 474-86-2 57-63-6 54024-22-5 72-33-3 68-22-4 6533-00-2 474-62-4 83-46-5 83-48-7

Anabolic agent Hormone metabolite Hormone replacement Hormone replacement Ovulation inhibitor Ovulation inhibitor Ovulation inhibitor Ovulation inhibitor Ovulation inhibitor Phytosterol (plant sterol) Phytosterol (plant sterol) Phytosterol (plant sterol)

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 101 Method 1698: 27 Steroids and Hormones in Water, Soil, Sediment, and Biosolids by HRGC/HRMS—Cont’d Steroid/Hormone

CAS #

Classification

Beta-stigmastanol 17a-Estradiol 17b-Estradiol (E2) Estriol (E3) Estrone (E1) Progesterone Testosterone 17a-Dihydroequilin Cholestanol Cholesterol Desmosterol Ergosterol b-Estradiol-3-benzoate Coprostanol Epi-coprostanol

83-45-4 57-91-0 50-28-2 50-27-1 53-16-7 57-83-0 58-22-0 651-55-8 80-97-7 57-88-5 313-04-2 57-87-4 50-50-0 360-68-9 516-92-7

Phytosterol (plant sterol) Sex hormone Sex hormone Sex hormone Sex hormone Sex hormone Sex hormone Sterol Sterol Sterol Sterol Sterol Sterol Sterol Sterol

Appendix B Contaminant Candidate List-4 (CCL-4) Chemical Contaminants 1,1-Dichloroethane 1,1,1,2-Tetrachloroethane 1,2,3-Trichloropropane 1,3-Butadiene 1,4-Dioxane 17-α-Estradiol 1-Butanol 2-Methoxyethanol 2-Propen-1-ol 3-Hydroxycarbofuran 4,40 -Methylenedianiline Acephate Acetaldehyde Acetamide Acetochlor Acetochlor ethanesulfonic acid (ESA) Acetochlor oxanilic acid (OA) Acrolein Alachlor ethanesulfonic acid (ESA)

Methanol Methyl bromide (bromomethane) Methyl tert-butyl ether (MTBE) Metolachlor Metolachlor ethanesulfonic acid (ESA) Metolachlor oxanilic acid (OA) Molybdenum Nitrobenzene Nitroglycerin N-Methyl-2-pyrrolidone N-Nitrosodiethylamine (NDEA) N-Nitrosodimethylamine (NDMA) N-Nitroso-di-n-propylamine (NDPA) N-Nitrosodiphenylamine N-Nitrosopyrrolidine (NPYR) Nonylphenol Norethindrone (19-norethisterone) n-Propylbenzene o-Toluidine (Continued)

102 Chapter 4 Contaminant Candidate List-4 (CCL-4)—Cont’d Chemical Contaminants Alachlor oxanilic acid (OA) α-Hexachlorocyclohexane Aniline Bensulide Benzyl chloride Butylated hydroxyanisole Captan Chlorate Chloromethane (methyl chloride) Clethodim Cobalt Cumene hydroperoxide Cyanotoxins Dicrotophos Dimethipin Diuron Equilenin Equilin Erythromycin Estradiol (17-β estradiol) Estriol Estrone Ethinyl estradiol (17-α ethynyl estradiol) Ethoprop Ethylene glycol Ethylene oxide Ethylene thiourea Formaldehyde Germanium HCFC-22 Halon 1011 (bromochloromethane) Hexane Hydrazine Manganese Mestranol Methamidophos

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

Contaminants of Emerging Concern: Occurrence, Fate, and Remediation 103

Appendix C Unregulated Contaminant Monitoring Rule-4 (UCMR-4) Contaminants and Approved Methods by US EPA Contaminant

EPA Method Cyanotoxins

Total microcystins Microcystin-LA Microcystin-LF Microcystin-LR Microcystin-LY Microcystin-RR Microcystin-YR Nodularin Anatoxin-a Cylindrospermopsin

EPA 546 EPA 544 EPA 544 EPA 544 EPA 544 EPA 544 EPA 544 EPA 544 EPA 545 EPA 545 Metals

Germanium Manganese

EPA 200.8, ASTM D5673-10, SM 3125 EPA 200.8, ASTM D5673-10, SM 3125 Pesticides and pesticide manufacturing product

α-Hexachlorocyclohexane Chlorpyrifos Dimethipin Ethoprop Oxyfluorfen Profenofos Tebuconazole Total permethrin (cis- and trans-) Tribufos

EPA 525.3 EPA 525.3 EPA 525.3 EPA 525.3 EPA 525.3 EPA 525.3 EPA 525.3 EPA 525.3 EPA 525.3

Brominated haloacetic acid groups HAA5 HAA6Br HAA9

EPA 552.3 or EPA 557 EPA 552.3 or EPA 557 EPA 552.3 or EPA 557 Alcohols

1-Butanol 2-Methoxyethanol 2-Propen-1-ol

EPA 541 EPA 541 EPA 541 Other semivolatile chemicals

butylated hydroxyanisole o-Toluidine Quinoline

EPA 530 EPA 530 EPA 530 Indicators

Total organic carbon (TOC)

Bromide

SM 5310 B, SM 5310 C, SM 5310 D (21st edition), or SM 5310 B-00, SM 5310 C-00, SM 5310 D-00 (SM online), EPA method 415.3 (rev. 1.1 or 1.2) EPA methods 300.0 (rev. 2.1), 300.1 (rev. 1.0), 317.0 (rev. 2.0), 326.0 (rev. 1.0) or ASTM D 6581-12

HAA5, chloro-, bromo-, dichloro-, dibromo-, and trichloro-acetic acid; HAA6Br, bromo-, bromochloro-, dibromo-, bromodichloro-, chlorodibromo-, tribromo-acetic acid; HAA9, chloro-, bromo-, dichloro-, bromochloro-, dibromo-, trichloro-, bromodichloro-, chlorodibromo-, and tribromo-acetic acid.

104 Chapter 4

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