Emerging Organic Pollutants in Waste Waters and Sludge (Handbook of Environmental Chemistry)

The Handbook of Environmental Chemistry Vol. 5, Part O (2005): 1– 24 DOI 10.1007/b98605 © Springer-Verlag Berlin Heidelb...

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The Handbook of Environmental Chemistry Vol. 5, Part O (2005): 1– 24 DOI 10.1007/b98605 © Springer-Verlag Berlin Heidelberg 2005

Estrogens and Progestogens in Wastewater, Sludge, Sediments, and Soil Marina Kuster · Maria J. López de Alda (✉) · Damià Barceló Department of Environmental Chemistry, IIQAB-CSIC, Jordi Girona 18–26, 08034 Barcelona, Spain [email protected]

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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2 Usage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.1 Human Medicine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2.2 Animal Farming . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

3 4 4

3 Sources and Fate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 3.1 Environmental Distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . .

5 7

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Occurrence . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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5

Toxicity Identification Evaluation (TIE) Approaches . . . . . . . . . . . . . . .

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6 6.1 6.2 6.3

Analytical Methods . . . . Sampling . . . . . . . . . . Sample Pretreatment . . . Analyses . . . . . . . . . .

. . . .

16 19 19 20

7

Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Abstract Estrogens and progestogens are two classes of female steroidal hormones whose presence in the environment has been associated with the appearance of certain alarming reproductive and development effects, such as feminization, decreased fertility, and hermaphroditism, in living organisms exposed to these compounds. Synthetic chemicals resembling these natural hormones are now well established in human medicine (mainly as contraceptives and for treatment of hormonal disorders) and in animal farming practices (usually as growth promoters). They are therefore produced on a large scale every year. Mainly due to unsuccessful removal in wastewater treatment plants, they are continuously released into the aquatic environment.Adverse effects on aquatic wildlife at concentrations as low as ~1 ng L–1 have been reported. Studies have also shown that estrogens and progestogens are easily distributed in the environment and may accumulate in river sediments. However, little is known about their long-term environmental impact. In this chapter, the main sources of estrogens and progestogens, their principal pathways into the aquatic environment, and the primary routes of exposure to these compounds are discussed. This chapter also reviews the methods described so far for the analysis of estrogens and progestogens in wastewater, sludge, sediments, and soils as well as the environmental levels found in these compartments. Keywords Estrogens · Progestogens · Environmental analysis · Occurrence · Fate

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Abbreviations APCI Atmospheric-pressure chemical ionization EC European Community EDC Endocrine disrupting compound EEF Molar-based 17b-estradiol equivalency factor ELISA Enzyme-linked immunosorbent assay ER-CALUX Estrogen receptor-mediated chemically activated luciferase gene expression assay ESI Electrospray ionization EXAMS Exposure assessment modeling system FDA United States Food and Drug Administration GC Gas chromatography GPC Gel-permeation chromatography HPLC High-performance liquid chromatography LC Liquid chromatography LLE Liquid–liquid extraction LOD Limit of detection LOQ Limit of quantitation MCF-7 Cell proliferation (E-screen) MS Mass spectrometry MSTFA N-methyl-N-(trimethylsilyl)trifluoroacetamide NI Negative ion PFPA Pentafluoropropionic anhydride PI Positive ion RAM Restricted access material SIM Selected ion monitoring SPE Solid-phase extraction SRM Selected reaction monitoring STP Sewage treatment plant TIE Toxicity identification and evaluation USEPA United States Environmental Protection Agency WW Wastewater WWTP Wastewater treatment plant YES Yeast-based recombinant estrogen receptor–reporter assay

1 Introduction Chemicals used in a wide range of applications in our modern society are produced on a large scale worldwide. Because of their physical and chemical properties, many of these substances or their metabolites end up in the environment, where they can induce adverse effects on wildlife organisms. The environmental presence of endocrine disrupting compounds has become a hot topic to the point that it competes with other priority health concerns such as the environmental pollution by carcinogenic compounds [1]. Among the various categories of substances with reported endocrine disrupting properties – polychlorinated organic compounds, pesticides, organotins, alkyl phenols and

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alkyl phenol ethoxylates, phthalates, bi-phenolic compounds, fitoestrogens and microestrogens, etc. – the group of female sexual hormones and related synthetic steroids stands out because of their estrogenic potency. Many studies have confirmed the presence of estrogens and progestogens at concentrations of toxicological concern in the aquatic environment. Already at very low concentrations of ~1 ng L–1 endocrine disrupting effects, such as decreased fertility, feminization, and hermaphroditism of aquatic organisms, are assigned to this class of steroidal hormones [2–5]. Due to their strong endocrine disrupting potency, special attention has been given to the natural estrogens estradiol and estrone, as well as to the synthetic estrogen ethynylestradiol [6]. Synthetic chemicals, resembling the action of natural hormones, find wide application in both human and veterinary medicine and in animal farming practices. Both natural and synthetic estrogens and progestogens are eliminated, either as free compounds or in their conjugated form, primarily through the urine but also in the feces. These substances enter the aquatic environment mainly via wastewater treatment plant (WWTP) effluents (after incomplete removal in the plant) and untreated discharges, and through runoff of sewage sludge used in agriculture [7, 8]. Once in the waterways they may undergo a series of processes, such as photolysis, biodegradation, and sorption to bedsediments, where estrogens and progestogens may persist for long periods [9]. At present, the environmental occurrence of these substances is not subjected to regulation. However, there are concrete indications that the presence of the most active estrogens in the aquatic environment will be regulated in the near future. This calls for efficient and reliable analytical methods for routine monitoring and control. Since the consequences linked to the presence of these compounds in the environment were first made public, numerous analytical methods for their quantification in different environmental matrices have been developed. Most of these methods have focused on surface waters, while wastewater (WW), sludge, and principally sediments and soils, have received comparatively less attention, probably due to the complexity of these matrices. This chapter reviews the most advanced methods applied to the analysis of estrogens and progestogens in these complex matrices, together with the environmental levels found in these natural systems. The main sources of the most environmentally relevant estrogens and progestogens, their physicochemical properties, their principal pathways into the aquatic environment, the primary routes of exposure to these compounds, and data regarding their activity as endocrine disruptors are discussed in this chapter.

2 Usage Large quantities of pharmacologically active substances are used annually in human medicine for diagnosis, treatment, and prevention of illness or to avoid unwanted pregnancy. In animal and fish farming, drugs are mostly adminis-

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tered as food additives for preventing illness, as growth promoters or parasiticides [10]. During the last five decades, the consumption of estrogens and progestogens for some of these purposes has experienced a steady growth, and this fact, together with the discovery of their negative ecological effects, has contributed to the current concern about their occurrence in the environment. 2.1 Human Medicine In terms of binding to the human estrogen receptor, estradiol is the principal endogenous phenolic steroid estrogen, which is oxidized in the metabolic processes to estrone and further transformed to estriol. The natural hormones are rapidly metabolized and are therefore orally inactive or active only at very high concentrations. Blocking the oxidation to estrone by, for instance, introducing an ethynyl group in position 17a or 17b of estradiol leads to much more stable products, which remain longer in the body. The consequence of this increased stability is that the so-formed synthetic steroid ethynylestradiol is excreted up to 80% unchanged in its conjugated form [11]. Estradiol also forms the backbone structure used in the engineering of other synthetic estrogens, such as mestranol and estradiol valerate, also utilized in human hormone treatments [11]. One of the main applications of estrogens and progestogens is in contraceptives. The estrogen content in birth control pills is usually in the range of 20 to 50 mg daily [12]. As for the progestogenic content, it varies depending on the type of contraceptive. Thus, in combined oral formulations the progestogenic content is in the range of 0.25 to 2 mg daily, whereas in progestogen-only contraceptives, it is lower (30–500 mg daily). Besides contraception, the uses of estrogens can largely be put into three main groups: the management of the menopausal and postmenopausal syndrome (its widest use); physiological replacement therapy in deficiency states; and the treatment of prostatic cancer and of breast cancer in postmenopausal women. In the same way as estrogens, progestogens are used in the treatment of several other conditions such as infertility, endometriosis, in the management of certain breast and endometrial cancers, and either alone or in combination with estrogens in the treatment of menstrual disorders, among others. The therapeutic doses required in the treatment of many of these diseases are often significantly larger than those employed in contraception. 2.2 Animal Farming Estrogens and progestogens are mainly used as growth promoters in animal farming, and for the development of single-sex populations of fish in aquaculture. Some naturally occurring sexual steroids such as estradiol, progesterone,

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and testosterone, and synthetic chemicals such as zeranol (estrogenic), melengestrol acetate (gestagenic), and trenbolone acetate (androgenic) have growthpromoting effects. Due to the improvement of weight gain and feed efficiency in meat-production animals, administration of sex steroids to cattle has been a common practice for many years in several meat-exporting countries, including the USA. The most widely used substances are estrogens, either in the form of 17b-estradiol, estradiol benzoate, or the synthetic zeranol. Progesterone, testosterone, and the two synthetic hormones trenbolone acetate and melengestrol acetate are generally used in combination with estrogens [13]. In contrast, no hormone applications for use in commercial-level poultry have been United States Food and Drug Administration (FDA)-approved since the agency’s withdrawal of the cancer-causing hormone diethylstilbestrol in the 1950s. In the European Community (EC), the use of hormonal substances for the promotion of animal growth is prohibited (Directive 96/22/EC). The ban was applied without discrimination internally and to imports from third countries as from January 1, 1989. As a result, countries wishing to export bovine meat and meat products to the EC were required either to have an equivalent legislation or to follow a hormone-free cattle program [14]. In aquaculture, steroidal compounds are used to develop single-sex populations of fish to optimize growth. Sex determination in fish is primarily under genetic control but may be influenced by various environmental conditions, such as temperature, social environment, pH, stocking density, and exposure to exogenous hormones or hormone-like chemicals [15]. Thus, all-male [16] and all-female [17] fish stocks may be obtained through exposure to androgens and estrogens, respectively. The potencies of sex steroids to induce sex reversal are different for each steroid. Functional sex reversal from female to male is carried out by using 17a-methyltestosterone, 19-norethynyltestosterone, or methylandrosterone (concentration range: 0.1–100 mg/kg diet). 11-Ketotestosterone and androsterone have also been used but the dosage required is higher than those of synthetic androgens. Phenotypical feminization is induced successfully by using estradiol, although estrone and ethynylestradiol are used as well [18].

3 Sources and Fate Figure 1 shows the principal routes of environmental exposure to estrogens and progestogens. The most relevant ways by which these compounds enter the environment and reach aquatic systems or the food chain are through WWTP effluents, untreated discharges, and runoff of manure and sewage sludge used in agriculture [7, 8, 10, 19–21, 45]. Human excretion is thought to be the principal source of estrogens and progestogens. These compounds are readily adsorbed from the gastrointestinal tract and through the skin or mucous membranes, and are metabolized in the

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Fig. 1 Routes of environmental exposure to estrogens and progestogens

liver with some undergoing enterohepatic recycling. Excreted hormones and their metabolites are found in urine, usually as water-soluble conjugates, and a small amount of “free” estrogens occur in feces [12, 22]. The normal daily estrogen secretion of women is 24–100 mg, depending on the menstrual cycle, and can rise to 30 mg toward the end of pregnancy. The excreted hormones and metabolites collected in the sewer systems end up in WWTPs, where different processes of varying efficiency are applied. Field data suggest that the activated sludge treatment process can consistently remove over 85% of estradiol, estriol, and ethynylestradiol, and a lower, variable percentage of estrone [23]. On the contrary, the concentration of unconjugated steroids in the effluent of WWTPs has occasionally been found to be higher than that in the corresponding influent. Thus, many studies suggest that the conjugated forms (mainly glucuronides and sulfates) are readily converted to the more active free compounds in both the sewer system and WWTPs, as a result of the activity of the b-glucuronidase and arylsulfatase enzymes present in these systems [7, 24–27]. In activated sludge treatment works the principal mechanisms for removal of these compounds are likely to be sorption and biodegradation. Based on the log Kow, sorption to sludge is predicted to play an important role in the removal of hydrophobic compounds (e.g., mestranol) from the aqueous phase. However, this does not seem to be the case for more hydrophilic compounds, such as estriol, estrone, and their glucuronide and sulfate conjugates. With regards to biodegradation, the extent of which depends on factors such as nitrifying bacteria, sludge retention times, aeration, and temperature, some laboratory test studies indicate that estradiol is more readily mineralized than ethynylestradiol

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or estrone and that synthetic estrogens in general exhibit greater recalcitrance in the activated sludge process [23, 27, 28]. More advanced water purification techniques, utilizing UV-irradiation, ozonization, or activated charcoal, may significantly improve the removal of these compounds, but these techniques are not broadly applied due to their high cost. Thus, current European activated sludge treatment plants, with a hydraulic residence time not greater than 14 h, can in most cases not completely eliminate all the estrogens and progestogens from the effluent [23]. As previously mentioned, contamination of water resources by estrogens and progestogens may also occur through runoff from manure and sewage sludge used in agriculture. Most of the drugs used for animals end up in their urine and feces.When this manure, or the sludge from sewage treatment plants, is dispersed onto the field, the unmetabolized drugs present or their metabolites, depending on their mobility in the soil system, may reach the groundwater (as a result of leaching from fields) or the surface water in the vicinity (through runoff) and affect terrestrial and aquatic organisms [29]. Other disposal options for the sewage sludge are landfill, dumping at sea (forbidden in the EU since 1998) [30], and incineration. The most popular for solid waste disposal is landfill. However, many of the disposal sites are open dumps without protective barriers or leachate-collection systems, which represent a potential risk to the quality of the nearby groundwater. Another increasingly important source of estrogens and progestogens in the environment is, as mentioned before, fish farming. Treatment with steroids is usually carried out by feeding, although in species where male sex differentiation is initiated before feeding commences (e.g., salmon), other procedures are used, such as immersion of alevins [18]. Drugs used in aquaculture as feed additives are discharged directly into the water. It has been estimated that around 70% of the drugs administered end up in the environment surrounding the farm, due to overfeeding, loss of appetite by diseased fish, and poor adsorption of the drugs [31]. 3.1 Environmental Distribution The introduction of estrogens and progestogens into the environment is a function of the way several factors are combined. The manufactured quantity and the dosage applied (amount, frequency, and duration) combined with the excretion efficiency of the compound and its metabolites, the capability of adsorption and desorption on soil, and the metabolic decomposition in sewage treatment are examples of necessary factors to assess environmental exposure. In general the fate and effect of a substance in the environment is dependent on the distribution into the different natural systems, such as air, water, and solids (soil, particles, sediment, and biota). Information on the physical and chemical properties (KH, Kd, and Kow vapor pressure) of a compound may help determine whether it is likely to concentrate in the aquatic, terrestrial, or atmospheric

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environment. Table 1 lists the main physicochemical characteristics of the most relevant estrogens and progestogens from the environmental point of view.With regards to the water solubility, it might be worth pointing out that the steroid solubility in WW may be markedly lower than that in distilled water [32]. Once estrogens and progestogens have reached the waterways, a series of processes, such as, photolysis, biodegradation, and sorption to bed-sediments, can contribute to their elimination from the environmental water. Given the relatively low polarity of these compounds, with octanol–water partition coefficients mostly between 103 and 105, sorption to bed-sediments appears to be a likely process. Kd values calculated for estriol, norethindrone, and progesterone in a Spanish river (479, 128, and 204, respectively) as the ratio between the sediment concentration (ng kg–1) and the water concentration (ng L–1) indicate that, in fact, these compounds exhibit a general tendency to accumulate in sediments. Jurgens et al. [33] carried out a series of laboratory experiments to study the behavior of estrogens in the aquatic environment and set up a model to estimate their likely environmental concentrations in the water column and bed-sediments. According to this study, between 13 and 92% of the estrogens entering a river system would end up in the bed-sediment compartment with the majority of sorption occurring within the first 24 h of contact. A similar approach conducted by Lai et al. [9] to investigate the partitioning of estrogens from water to sediments, kinetics of sorption, and the influence of various environmental variables (salinity, total organic carbon, etc.) indicated that sorption takes place rapidly within the first half hour, slows down within the next half hour, and steadily decreases afterward. Furthermore, the synthetic estrogens (mestranol and ethynylestradiol), with their higher Kow values, were shown to partition to the sediment to a greater extent than the natural estrogens. At higher estrogen concentrations, there was a decrease in estrogen removal from the aqueous phase, while higher levels of sediment induced greater removal. The sorption of estrogens to sediments correlated to the total organic carbon content. However, the presence of organic carbon was not a prerequisite for sorption. Tests performed with laboratory saline water resulted in an increase of estrogen removal from the water phase compared to unsalted waters, which is consistent with partitioning experiments using actual field water samples. The addition of estradiol valerate, with a particularly high Kow, suppressed sorption of other estrogens, suggesting that it competed with other compounds for the binding sites. A series of experiments was also conducted by Bowman et al. [34] to ascertain the effects of differing environmental factors on the sediment–water interactions of natural estrogens (estradiol and estrone) under estuarine conditions. Sorption onto sediment particles was in this case relatively slow, with sorption equilibrium being reached in about 10 and 170 h for estrone and estradiol, respectively. On the other hand, true partition coefficients calculated on colloids were found to be around two orders of magnitude greater that those on sediment particles. Hence, it was concluded that under estuarine conditions, and in comparison to other more hydrophobic compounds, both estrone and estradiol

c

b

Values are at 25 °C if not specified. Estimated data. Experimental data.

000050-28-2 000050-27-1 000053-16-7 000057-63-6 000056-53-1 000068-22-4 000057-83-0 000797-63-7

Estradiol Estriol Estrone Ethinyl estradiol Diethylstilbestrol Norethindrone Progesterone D-Norgestrel

a

CAS number

Compound

272.39 288.39 270.37 296.41 268.36 298.43 314.47 312.46

Molecular weight

3.6 (27 °C) 441b 30 11.3 (27 °C) 12 7.04 8.81 2.05b

Water solubility a, c (mg L–1) 4.01 2.45 3.13 3.67 5.07 2.97 3.87 3.48

Log Powc

1.26E-008 1.97E-010 1.42E-007 2.67E-009 1.41E-008 7.31E-009 1.3E-006 3.93E-010

Vapor pressure a, b (mm Hg)

3.64E-011 1.33E-012 3.8E-010 7.94E-012 5.8E-012 5.8E-010 6.49E-008 7.7E-010

Henry’s law constant a, b (atm-m3 mole–1)

Table 1 Physico-chemical Properties of selected Estrogens and Progestogens [(http://esc.syrres.com/interkow/physdemo.htm)]

Estrogens and Progestogens in Wastewater, Sludge, Sediments, and Soil 9

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would be expected to remain mainly in the dissolved phase and to have a strong tendency for bioaccumulation. Nevertheless, it is not clear yet whether sorption or biodegradation processes play a major role. Studies conducted with activated sludge [25, 35] pointed at biodegradation as the mechanism contributing the most to the elimination of estrogens from the aqueous phase, while losses by sorption effects were considered rather unlikely. However, Jurgens et al. [33], based on a designed exposure assessment modeling system (EXAMS) model, postulate that degradation processes in rivers are unimportant under average flow conditions, as they account for only 2–8% of the input loading. This is in agreement with the theory presented by Huang et al. [36], according to which the main removal mechanism for hormones in WWTPs would be sorption onto particles and not biotransformation. Degradation studies carried out in waters from five English rivers indicate that estradiol has a half-life of 3–27 days [33]. Estrone was found to be the first degradation product of estradiol but no investigations of the subsequent byproducts were conducted. The poorest degradation rates were observed in the estuary river water samples, where the high salt content might have inhibited microbial degradation. Furthermore, ethynylestradiol (half-life 46 days) was found to be more stable than 17b-estradiol (half-life 4 days, e.g., in the River Thames). These half-life values might correspond to ideal summer temperatures. However, under winter conditions these compounds could be twice as persistent. In activated sludge, the synthetic estrogens ethynylestradiol and mestranol have been shown to remain stable and intact over 5 days, while progestogens are already up to more than 50% disintegrated after 48 h [32]. Under the anaerobic, dark conditions normally present in the subsurface layers of river sediments, these compounds are expected to undergo a slow photodecomposition and biodegradation. On the other hand, desorption from sediments has been shown to be significantly less important than sorption, with desorption distribution coefficients two to three times lower than those obtained for the sorption process [33]. In an environment like this, river sediments can therefore act as sinks where estrogens and progestogens may persist for long periods, be transported to other areas, and be eventually released by diffusion across the sediment water-column interface or by scouring in storm events [9]. The concentration of estrogens and progestogens in bed-sediments is predicted to increase over time; thus, bed-sediments can be anticipated as environmental reservoirs from where these substances may eventually become bioavailable [37].

4 Occurrence Most research of estrogens and progestogens has been conducted on water samples and less frequently on solid samples. Soils and sediments, in particular, have received very little attention and thus literature data on these matrices are very

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scarce. The same applies to the analytes investigated. Whereas free estrogens, both natural (e.g., estradiol, estrone, estriol) and synthetic (e.g., ethynylestradiol, mestranol, diethylstilbestrol), have often been investigated, estrogen conjugates [38, 39] and progestogens [40] have seldom been studied, probably due to their lower estrogenic potency. Figure 2 shows the chemical structure of the estrogens and progestogens most frequently investigated in environmental samples. Table 2 summarizes the literature data available on the occurrence of these two classes of steroidal compounds in WW, sludge, and sediments. Natural hormones (and their metabolites) have always been present in the environment. The growing use of both natural and synthetic estrogens and progestogens in human medicine and in livestock farming (see Sect. 2, Usage) has led to an increase of their occurrence in natural systems. Due to steady population growth and regional population density, an irregular distribution of these pollutants is found. Particular concern is given to certain areas where high levels were detected, e.g., areas adjacent to agricultural and animal farms.

Table 2 Environmental occurrence of estrogens and progestogens

Matrix (location)

Compounds

Concentration (ng L–1 or ng g–1)

Ref.

Estrogens and progestogens Natural and synthetic estrogens Natural and synthetic estrogens Natural and synthetic estrogens

0.4–188 Æ 0.3–82.1 <0.2–115 Æ <0.2–21.5 <0.5–140 Æ <0.4–47 2–8.6 Æ 3.8–18

[35] [40] [59] [66]

River sediment Germany Spain UK

Natural and synthetic estrogens Estrogens and progestogens Estrone

<0.2–2 0.05–22.8 <0.04–0.388

[45] [46] [42]

Activated and digested sludge Germany

WATER Wastewater STP influent Æ effluent Italy Spain The Netherlands France SOLID SAMPLES

Natural and synthetic estrogens

<2–49

[45]

Activated sludge Israel

Estrogen

19–64

[19]

Soil

–

–

–

Natural and synthetic estrogens

<0.1–2.5 mg/g

[48]

BIOTA Rainbow trout bile Sweden

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Fig. 2 Molecular structure of the environmentally most relevant natural and synthetic estrogens and progestogens

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As previously mentioned, discharged domestic effluents represent the most significant input of estrogens and progestogens to the aquatic environment. According to the studies carried out until now, mostly in densely populated regions, estrogens and progestogens are normally present in domestic sewage in the nanogram per liter range and occasionally in the microgram per liter range (see Table 2). The nature of the compounds present in the sewage system does, however, change as a function of their transport route. D’Ascenzo et al. [39] have recently conducted a very comprehensive study where the presence of the three most common natural estrogens and their conjugated forms was investigated in female urine, in a septic tank collecting domestic WW, and in influents and effluents of six activated sludge WWTPs. A group of 73 women was selected to represent a typical cross section of the female inhabitants of a Roman condominium. On average, the concentrations of conjugated estriol, estradiol, and estrone measured in the urine were 106, 14, and 32 mg/day, respectively.Apart from some estriol found in pregnancy urine, free estrogens were not detected and estrogen sulfates represented 21% of the total conjugated estrogens content. This situation, however, changed markedly in the condominium collecting tank. Here, significant amounts of free estrogens were observed and the estrogen sulfate to estrogen glucuronate ratio rose to 55/45, which was attributed to the ready deconjugation of the glucuronated estrogens by the b-glucuronidase enzyme produced presumably in large amounts by fecal bacteria (Escherichia coli). At the WWTP entrance, free estrogens and sulfated estrogens were the dominant species. Finally, the sewage treatment was found to completely remove residues of estrogen glucuronates, and with good efficiency (84–97%) the other analytes, but not estrone (61%) and estrone-3-sulfate (64%). In STP effluents, total extractable estrogens and conjugates have been detected at levels up to 1 mg/L [9, 11, 26]. Despite the wide variability in terms of removal efficiency reported for different WWTPs, a general trend has been observed with respect to the identity of the compounds most frequently detected in WWTP effluents. Thus, of the various compounds most commonly monitored – namely, estradiol, its metabolites estriol and estrone, and the synthetic estrogen ethynylestradiol – estrone is the most ubiquitous both in WWTP effluents and in environmental waters in general, while the most potent estrogens estradiol and ethynylestradiol have only occasionally been detected [26, 40–42].As for the conjugates, the very few studies that have attempted their determination pointed out estrone sulfates as the most abundant, while glucuronides are most often found below the limit of detection [26, 36, 38, 39]. Downstream of WWTPs, in the receiving river waters, the concentration of estrogens and progestogens is normally considerably lower than that in the corresponding effluent and decreases with distance from the WWTP [35, 36, 39, 42]. In this kind of compartment, the presence of estrogens and progestogens has been reported to occur in the low nanogram per liter range [19, 43, 44]. In activated and digested sewage sludge, the concentration of ethynylestradiol (17 ng g–1), estrone (37 ng g–1), and estradiol (49 ng g–1), found in one of the

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very few studies conducted in this kind of matrix (mestranol was not detected), indicates that estrogens may remain unchanged during sludge digestion [45]. River sediments, likewise sewage sludge, have rarely been investigated [42, 45–47] and to the authors’ knowledge there are no published reports on the occurrence of estrogens and progestogens in either marine sediments or soils. In one of the studies conducted with river sediments (in the UK), only the latter of the three estrogens estradiol, ethynylestradiol, and estrone was detected (>0.04–0.388 ng g–1 wet weight) [42]. Estriol and norethindrone were the compounds most frequently detected in sediments collected in two rivers from the northeast of Spain, where maximum concentrations were obtained for ethynylestradiol (22.8 ng g–1 dry weight) and estrone (11.9 ng g–1) [46]. In both studies, large concentration variations between sites and between the same site sampled on different occasions were observed, which might be explained by the difficulty in obtaining representative samples and the variability of factors such as the gaseous/redox conditions influencing degradation rates. In another study, conducted in Germany, estradiol, estrone, and ethynylestradiol were found at concentrations up to 2 ng g–1 (estrone), whereas mestranol, a prodrug for ethynylestradiol, was not detected [45]. Finally, neither estrogens nor progestogens were detected in river sediments from Portugal [47]. Table 2 includes an example of bioaccumulation, as described by Larsson et al. [48], where estrogen concentrations 4 to 6 orders of magnitude higher than those in water were found in the bile of a rainbow trout caged downstream of WWTPs.

5 Toxicity Identification Evaluation (TIE) Approaches Most of the studies conducted to assess the environmental occurrence and fate of estrogens and progestogens have focused on the determination of specific target compounds. However, by simply following the disappearance of a substance, one cannot conclude that the environmental risk has vanished. The derived degradation products or metabolites may also cause environmental adverse effects. The identification of these products is a difficult task, due to the great number of compounds that can possibly be generated, the high costs, and the lack of analytical standards. At the beginning of the 1990s the United States Environmental Protection Agency (USEPA) developed the so-called TIE procedures. These approaches were originally designed to identify the presence of health and environmentally relevant compounds in WWs [49, 50]. However, since their introduction, they have become established, powerful tools for determining the causative agents of effects (such as acute toxicity, (geno)toxicity, and endocrine disrupting potential) in aqueous and solid environmental samples. The general scheme of TIE for the effect-based analysis is presented in Fig. 3. As can be seen, the main purpose of TIE is to convert complex environmental

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15

Fig. 3 General scheme of TIE procedure

extracts into fractions where the identification of the compounds responsible for the effects observed (by means of suitable bioassays) is feasible using, e.g., mass spectrometric methods. A survey of the TIE procedures used for the identification of endocrine disrupting compounds (EDCs) has been recently published by Petrovic et al. [51]. An example of a TIE approach is that described by Desbrow et al. [7]. In this work, the endocrine disrupting activity detected in effluents of seven UK WWTPs by means of a yeast-based screening assay [52] was mainly attributed to the presence of estradiol, estrone, and ethynylestradiol. However, to assess the estrogenic activity different bioassays may be used, e.g., the yeast-based recombinant estrogen receptor–reporter assay (YES), the MCF-7 cell proliferation (E-screen), and the estrogen receptor-mediated chemically activated

16

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Table 3 Relative estrogenic potencies as determined by different bioassays (expressed as EEF – the molar-based 17b-estradiol equivalency factor) (adapted from [51])

Compound/bioassay

YES

MCF-7

ER-CALUX

17b-estradiol Estriol Ethinyl estradiol Diethylstilbestrol Estrone

1 3.7E-1 1.9E-1–1.2 4.5E-2–1.1 1.9E-2–1E-1

1

1

1.25–1.9 2.5 1E-2

1.2 5.6E-2

luciferase gene expression assay (ER-CALUX). Table 3 lists the relative estrogenic potencies determined for the most important estrogens by different bioassays. As the estrogenic activity of two compounds may vary in different approaches (bioassay classes), it needs to be stressed that the final outcome of the TIE, with regard to the identity of the causative agent, may vary.

6 Analytical Methods The analysis of steroid sexual hormones and related synthetic compounds in WW, soil, sludge, and sediment samples is a challenging task. This is due to both the complex environmental matrices and the requirement of low detection limits. Therefore, the use of complicated, time- and labor-consuming analytical procedures is necessary. Many studies have reported the analysis of estrogens and progestogens in WW and other water samples, but for solid samples, little is found in the literature. This chapter briefly reviews the methods described for WW samples and discuss the very few methods used up to now to analyze solid samples (sediments, sludge, and soil). For more detailed data we refer to the reviews: – Environmental behavior and analysis of veterinary and human drugs in soils, sediments, and sludge [53] – Determination of endocrine disrupters in sewage treatment and receiving waters [54] – Review of analytical methods for the determination of estrogens and progestogens in WWs [55] – Liquid chromatography–(tandem) mass spectrometry of selected emerging pollutants (steroid sex hormones, drugs and alkyl phenolic surfactants) in the aquatic environment [56] Some of these methods are summarized in Table 4 and details concerning each step of the analytical process are described in Sects. 6.1–6.3.

E3, E2, EE, E1

WWTP influent, effluent, and others 1 LOQ 0.08–0.6 2–500

200 5 0.07–0.18 0.003–15

HPLC–MS/MS (NI-ESI) HPLC-DAD–MS (NI-ESI) ELISA GC–MS/MS (E2) HPLC-DAD–MS (NI-ESI) HPLC–MS (NI-ESI) (SIM) HPLC–MS/MS (NI-ESI) HPLC–MS (NI-ESI) HPLC–MS/MS (NI-ESI)

SPE (C18 disk), hydrolysis, HPLC fraction On-line SPE (PLRP-s column) SPE (C18 column)

[61]

[69]

[68]

[60]

[36]

[35] [57]

[43]

[7] [26]

[66]

Ref.

E1, estrone; E1-3G, estrone 3-(b-D-glucuronide); E1-3S, estrone 3-sulfate; 16a-OH-E1, 16a-hydroxyestrone; E2, 17b-estradiol; aE2, 17a-estradiol; E2-3G,17b-estradiol 3-(b-D-glucuronide); E2-3S, 17b-estradiol 3-sulfate; E2-17A, 17b-estradiol 17-acetate;E2-17G, 17b-estradiol 17-(b-D-glucuronide); E2-17 V, 17b-estradiol 17-valerate; E3, estriol; E3-3G, estriol 3-(b-D-glucuronide); E3-3S, estriol 3-sulfate; E3-16G, estriol 16a-(b-D-glucuronide); EE, ethynylestradiol; DES, diethystilbestrol; MES, mestranol; LEV, levonorgestrel; NOR, norethindrone; PROG, progesterone; MSTFA, N-methyl-N(trimethylsilyl)trifluoroaceatamide; PFPA, pentafluoropropionic anhydride.

0.1 0.2–0.4 10–200

0.2 0.1–2.4

LOQ 0.04–0.32

LOD (ng L–1 or ng g–1)

GC–MS/MS

GC–MS GC–MS/MS

GC–MS

Analytical method

Speedisk-C18 and derivatization PFPA SPE (C18), HPLC fraction, LLE SPE (SDB-XC disk, C18 or NH2 column), HPLC fraction) SPE (C18), silica gel, derivatization SPE (Carbograph-4) SPE (C18 column)

Sample preparation

SPE (LiChrolut EN +C18 col.) + immunoaffinity extraction E2, E3, E1, EE, E3-3G, E2-3G, SPE (Carbograph-4) E1-3G, E3-16G, E2-17G, E3-3S, E2-3S, E1-3S

E2, E1

E3, E2, EE, E1, DES, NOR, LEV, PROG E2, E3, E1, EE, DES

E2, E1, MES, aE2, E2-17 V, 16 a-OH-E1, E2-17A E3, E2, EE, E1 E3, E2, EE, E1, DES, NOR, LEV, PROG E2, EE, E2-17G, E2-3S

E2, EE, E1 E2, EE, E1, aE2

Compound

Sample

Table 4 Analytical methods described for the determination of estrogens and progestogens in environmental samples

Estrogens and Progestogens in Wastewater, Sludge, Sediments, and Soil 17

Sludge

Solvent extraction, silica gel column, SPE (RP-C18), HPLC (RP-C18) cleanup, derivatization MSTFA Ultrasonic extraction with MeOH-acetone, SPE (C18) SPE (RAM cartridges (ADS C4))

E2, EE, E1, MES

NOR, LEV, PROG

NOR, LEV, PROG

Solvent extraction, GPC, silica gel column, derivatization with MSTFA

SPE (RAM cartridges (ADS C4))

E3, E2, EE, E1, DES

E2, EE, E1, MES

Ultrasonic extraction with MeOH-acetone, SPE (C18)

E3, E2, EE, E1, DES

Solvent extraction with sonification, HPLC fractioning, derivatization

E2, EE, E1

Ultrasonic extraction with

SPE (Carbograph-4)

E3, E2, EE, E1

E2, EE, E1

SPE (SDB-XC disk, C18 or NH2 column), HPLC fraction)

E2, EE, E1

Natural river sediment

SPE (Envi-Carb col.)

E2, E3, E1, EE

WWTP influent, effluent, and others

Sample preparation

Compound

Sample

Table 4 (continued)

GC–ion-trap MS/MS

HPLC–MS (PI-ESI) (SIM)

HPLC–MS (PI-ESI) (SIM)

GC–MS/MS (EI)

HPLC–MS (NI-ESI) (SIM)

HPLC–MS (NI-ESI) (SIM)

GC–MS/MS DCM, HPLC fractioning, derivatization

GC–MS/MS

HPLC–MS/MS (NI-ESI)

GC–MS/MS

HPLC–MS/MS (PI-APCI)

Analytical method

LOQ 2–4

0.5

0.04

LOQ 0.2–0.4

1–5

0.05–1

0.04–5

0.4–1

0.2–0.5

0.1–1.8

LOQ 0.5–1

LOD (ng L–1 or ng g–1)

[45]

[63]

[46]

[45]

[63]

[46]

[42]

[42]

[59]

[59]

[58]

Ref.

18 M. Kuster et al.

Estrogens and Progestogens in Wastewater, Sludge, Sediments, and Soil

19

6.1 Sampling The choice of the sampling procedure is essential to obtain significant, representative results of the environmental occurrence of these compounds. The place and time of sample collection has to be planned carefully. The sampling procedure and transport should not influence the matrix to be analyzed. In general, exposure to light, oxygen, and high temperatures should be avoided due to the risk of transformation of the analytes or other organic components present in the environmental samples. Wastewater samples have usually been collected in precleaned amber glass containers. Both discrete and composite samples have been used for the analysis of effluents and influents of WWTPs. Unpreserved samples are normally stored at 4 °C for 48 h, or frozen [48]. Other authors add chemical agents such as methanol, sulfuric acid, or mercuric chloride to prevent bacterial activity during storage, and/or store the samples in supports used for extraction [26, 35, 57]. The devices used for sampling of solid samples (sludge, sediment, and soil) are usually grab samplers or corers. Box corers or multicorers can be employed if more detailed information on the spatial distribution of the analytes is needed. The samples are stored in the dark at 4 °C or more commonly at –20 °C, preferably in glass containers [53]. Very often, solid samples are also dried or lyophilized prior to storage. 6.2 Sample Pretreatment An essential step in the analysis of trace pollutants in environmental matrices is the pretreatment procedure. Methods that are more efficient have been developed in the last few years, facilitating subsequent chromatographic analysis. Because of the complexity of the matrices, the sample pretreatment procedure includes both extraction and purification of the target analytes. Filtration is the first step of WW sample preparation because of the high loading of organic material and suspended particles. This step is essential to prevent clogging of the adsorbent bed by suspended solids, if solid-phase extraction (SPE) is performed. In the case of immunochemical assays, previous filtration avoids undesired adsorption onto antibodies. Most of the studies reviewed employed glass-fiber filters with a pore size between 0.22 and 1.2 mm. Investigations showed no significant loss of the analytes after this filtration procedure [7, 36]. However, the filtration system is usually washed with an organic solvent to ensure that no analyte is left adsorbed to the particles [35, 58, 59]. Extraction of estrogens and progestogens is mostly performed by off-line SPE (on-line SPE has been reported by Lopez de Alda et al. [60]), using either disks or, more frequently, cartridges. Octadecyl (C18)-bonded silica in both cartridge [7, 57] and disk [36] format, graphitized carbon black [35, 58, 59], Isolut

20

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ENV+ cartridges [48], and SDB-XC disks [26, 59] are the sorbents more widely used. The combination of C18 and SDB has also been reported, showing good recoveries for the investigated analytes [43]. Previous adjustment of the sample pH is performed sometimes [43]. The occurrence of conjugated estrogens has only been investigated in a few works [26, 36, 48, 61]. The analysis of these compounds by means of immunoassays requires their previous conversion to the corresponding free estrogens by enzymatic hydrolysis [26, 36, 48]. In this case, the concentrations of the conjugated hormones are determined from the differences between the results obtained for hydrolyzed and unhydrolyzed samples. It should, however, be remarked that these enzymes convert estradiol glucuronide quantitatively into estradiol, but convert only ca. 30% of estradiol sulfate into estradiol [36]. Levels of the sulfate-conjugated hormones might therefore be underestimated. By contrast, LC–MS enables the simultaneous determination of the free and conjugated forms without the need for hydrolysis. Previous derivatization of the extract is necessary to improve the stability of the compounds and the sensitivity and precision of subsequent GC–MS analysis. Silyl derivatives formed for example with MSTFA [43], halogenated alkene derivatives produced with heptafluorobutyric anhydride (HFBA) [36] or pentafluoropropionic acid [58] or anhydride (PFPA), as well as acetate derivatives formed using acetic anhydride [48] have been widely employed. Solid samples are in general more difficult to handle than liquid ones. The target analytes are extracted from their solid matrix by sonification, by pressurized-liquid extraction (PLE), or by simple blending or stirring of the sample with polar organic solvent solutions, e.g., methanol/acetone solutions [42, 45–47, 62, 63]. Cleanup of the extracts was performed using SPE (C18 cartridges [45, 46], RAM cartridges (ADS C4) [63], silica gel [45]), gel-permeation chromatography (GPC) [45], and semipreparative HPLC [45]. RAMs are bifunctional sorbents that combine size exclusion and reversed-phase retention mechanisms. 6.3 Analyses The analysis of WW samples has been dominated by the use of immunoassays and GC–MS techniques. However, in recent years, LC–MS and LC–MS/MS have gained in popularity, because the above-mentioned preceding hydrolysis step (needed for immunoassay analysis) and derivatization step (needed for GC–MS analysis) are not necessary. Biological techniques, e.g., immunoassays, are among the most sensitive analytical methods, but are limited by the availability of the specific antisera and are subject to cross-reactivity. Huang et al. [36] employed an enzymelinked immunosorbent assay (ELISA) for determination of estradiol, its conjugates, and ethynylestradiol in wastewaster treatment plant effluents (see Table 4). The reported limit of detection (LOD) of 0.1 ng L–1 reflects the sen-

Estrogens and Progestogens in Wastewater, Sludge, Sediments, and Soil

21

sitivity of this method. Low LODs in the range of pg L–1 to 2 ng L–1 have also been achieved by using other immunoassay [64, 65] and radioimmunoassay [19, 22] protocols. The analysis of estrogens and progestogens by GC–MS has been carried out with a variety of capillary columns using helium as carrier gas [7, 26, 36, 43, 59, 66]. LODs in the range of 0.1–1.8 ng L–1 have been achieved. In terms of sensitivity, GC– and HPLC–tandem mass spectrometry are comparable techniques. However, the derivatization carried out prior to GC separation is time consuming and can be a source of inaccuracy [7]. For the environmental analysis of estrogens and progestogens by HPLC–MS, both electrospray ionization (ESI) in the negative-ion (NI) mode for estrogens

Fig. 4 Reconstructed SRM chromatograms obtained from the LC–ESI-MS/MS analysis of a 100 ng mL–1 standard mixture of estrogens (in the NI mode) and progestogens (in the PI mode). Column: Purospher STAR RP-18e (125¥2 mm, 5 mm, Merck). Mobile phase: gradient acetonitrile/water. Flow rate: 0.2 mL min–1

22

M. Kuster et al.

and in the positive-ion (PI) mode for progestogens and, to a lesser extent, atmospheric-pressure chemical ionization (APCI) in the PI mode have been used. Chromatographic separation has been performed on octadecyl silica stationary phases. According to a recent study carried out to compare the performance of various MS techniques (GC–MS, LC–MS, and LC–MS/MS) [67], LC–ESI-MS/MS is the technique of choice for analysis of these compounds, based on sensitivity and selectivity. The same study also indicates that, although the limits of detection achieved by LC–MS in the selected ion monitoring (SIM) mode and by LC–MS/MS in the selected reaction monitoring (SRM) mode are in general comparable, the higher selectivity of the latter is essential to avoid false positive determinations in the analysis of real environmental samples. Figure 4 shows representative chromatograms obtained from the analysis of estrogens and progestogens by LC–ESI-MS/MS. For analysis of solid samples, GC–MS/MS [45] and more frequently HPLC–MS have been used [46, 63]. Limits of detection vary from 0.04 to 4 ng g–1.

7 Conclusions What is the real danger imposed by the presence of estrogens and progestogens in aquatic systems? The studies carried out up to now indicate that the occurrence of these compounds in surface waters is an issue of concern, and that wastewater treatment plant effluents play a major role in their introduction into the environment. However, more information on their environmental presence, transport, and fate is needed to assess their ultimate ecosystem impacts. Main data gaps are localized in the area of environmental solid samples. Possible future dangers from accumulation in sediments and soil are at present unpredictable and should be investigated thoroughly. Besides environmental levels, toxicological data and bioavailability and degradation studies should be available. In the field of analysis, important progress has been made in terms of sensitivity and selectivity. LC–ESI-MS/MS appears to be the technique of choice for their determination as it provides reliable results at subnanogram per liter or per gram levels. However, sample preparation is identified as the main bottleneck in the analysis of these compounds. Quite tedious and time-consuming procedures are still required, especially in the case of complex matrices such as sewage sludge. To reduce the use of these sexual hormones seems to be an impossible task, as no substitute compounds can be applied instead of estrogens and progestogens to the described medicinal and farming applications. However, since WWTP effluents constitute the most important source of these compounds, remediation actions can be performed at this level by introducing more advanced and efficient treatment processes, especially in plants receiving high inputs of urban discharges from highly populated regions.

Estrogens and Progestogens in Wastewater, Sludge, Sediments, and Soil

23

Acknowledgements This work was supported by the Energy, Environmental and Sustainable Development Program (Project ARTDEMO EVK1-CT2002-00114), and by the Spanish Ministry of Science and Technology (Projects BQU2002-10903-E and PPQ2001-1805-CO3-01). Maria José López de Alda acknowledges her Ramon y Cajal contract from the Spanish Ministry of Science and Technology.

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32. Norpoth K, Nhrkorn A, Kirchner M, Holsen H, Teipel H (1973) Zbl Bakt Hyg I Abt Orig B 156:500 33. Jurgens MD,Williams RJ, Johnson AC (1999) Research and development technical report P161. Environment Agency, Bristol 34. Bowman JC, Zhou JL, Readman JW (2002) Mar Chem 77:263 35. Baronti C, Curini R, D’Ascenzo G, Di Corcia A, Centili A, Samperi R (2000) Environ Sci Technol 34:5059 36. Huang CH, Sedlak DL (2001) Environ Toxicol Chem 20:133 37. Lai KM, Scrimshaw MD, Lester JN (2002) Sci Total Environ 289:159 38. Isobe T, Shiraishi H, Yasuda M, Shinoda A, Suzuki H, Morita M (2003) J Chromatogr A 984:195 39. D’Ascenzo G, Di Corcia A, Gentili A, Mancini R, Mastropasqua R, Nazzari M, Samperi R (2003) Sci Total Environ 302:199 40. Petrovic M, Solé M, López de Alda MJ, Barceló D (2002) Environ Toxicol Chem 21:2146 41. Xiao XY, McCalley DV, McEvoy J (2001) J Chromatogr A 923:195 42. Williams RJ, Johnson AC, Smith JJL, Kanda R (2003) Environ Sci Technol 37:1744 43. Ternes TA, Stumpf M, Mueller J, Haberer K, Wilken RD, Servos M (1999) Sci Total Environ 225:81 44. Kolpin DW, Furlong ET, Meyer MT, Thurman EM, Zaugg SD, Barber LB, Bastón HT (2002) Environ Sci Technol 36:1202 45. Ternes TA, Andersen H, Gilberg D, Bonerz M (2002) Anal Chem 74:3498 46. López de Alda MJ, Gil A, Paz E, Barceló D (2002) Analyst 127:1299 47. Céspedes R, Petrovic M, Raldúa D, Saura U, Piña B, Lacorte S, Viana P, Barceló D (2004) Anal Bioanal Chem 378:697 48. Larsson DGJ, Adolfsson-Erici M, Parkkonen J, Petterson M, Berg AH, Olsson PE, Förlin L (1999) Aquatic Toxicol 45:91 49. Ankley GT, Burkhard LP (1992) Environ Toxicol Chem 11:1235 50. Norberg-King TJ, Durhan EJ, Robert E, Ankley GT (1991) Environ Toxicol Chem 10:891 51. Petrovic M, Eljarrat E, López de Alda MJ, Barceló D (2004) Anal Bioanal Chem 378:549 52. Routledge EJ, Sumpter JP (1996) Environ Toxicol Chem 15:241 53. Díaz-Cruz MS, López de Alda MJ, Barceló D (2003) Trend Anal Chem 22:340 54. Gomes RL, Scrimshaw MD, Lester JN (2003) Trend Anal Chem 22:697 55. López de Alda MJ, Barceló D (2001) Fresenius J Anal Chem 371:437 56. López de Alda MJ, Díaz-Cruz S, Barceló D (2003) J Chromatogr A 1000:503 57. López de Alda MJ, Barceló D (2000) J Chromatogr A 892:391 58. Laganà A, Bacaloni A, Fago G, Marino A (2000) Rapid Commun Mass Spectrom 14:401 59. Johnson AC, Belfroid A, Di Corcia A (2000) Sci Total Environ 256:163 60. López de Alda MJ, Barceló D (2002) J Chromatogr A 911:203 61. Gentili A, Perret D, Marchese S, Mastropasqua R, Curini R, Di Corcia A (2002) Chromatographia 56:25 62. Williams RJ, Johnson AC, Smith JJL, Kanda R (2003) Environ Sci Technol 37:1744 63. Petrovic M, Tavazzi S, Barcelo D (2002) J Chromatogr A 971:37 64. Shishida K, Echigo S, Kosaka K, Tabasaki M, Matsuda T, Takigami H,Yamada H, Shimizu Y, Matsui S (2000) Environ Technol 21:553 65. Aherne GW, Briggs R (1998) J Pharm Pharmacol 41:735 66. Mouatassim-Souali A, Tamisier-Karolak SL, Perdiz D, Cargouet M, Levi Y (2003) J Sep Sci 26:105 67. Díaz-Cruz MS, López de Alda MJ, López R, Barceló D (2003) J Mass Spectrom 38:917 68. Croley TR, Hughes RJ, Koenig BG, Metcalfe CD, March RE (2000) Rapid Commun Mass Spectrom 14:1087 69. Ferguson PL, Iden CR, McElroy AE, Brownawell BJ (2001) Anal Chem 73:3890

The Handbook of Environmental Chemistry Vol. 5, Part O (2005): 25– 51 DOI 10.1007/b98606 © Springer-Verlag Berlin Heidelberg 2005

Organic Compounds in Paper Mill Wastewaters A. Latorre1 · A. Rigol2 · S. Lacorte1 (✉) · D. Barceló1 1

2

Department of Environmental Chemistry, IIQAB-CSIC, Jordi Girona 18–26, 08034 Barcelona, Catalonia, Spain [email protected] Department of Analytical Chemistry, University of Barcelona, Av. Diagonal 647, 08028 Barcelona, Catalonia, Spain

1

Pulp and Paper Mill Wastewaters . . . . . . . . . . . . . . . . . . . . . . . . . .

27

2

Legislation Related to Pulp and Paper Mill Industries

. . . . . . . . . . . . . .

30

3 3.1 3.2 3.3 3.4 3.5

Chemical Characterization of Pulp and Paper Mill Waters Biocides . . . . . . . . . . . . . . . . . . . . . . . . . . . Resin and Fatty Acids . . . . . . . . . . . . . . . . . . . . Surfactants and Plasticizers . . . . . . . . . . . . . . . . . Lignin and Hemicelluloses . . . . . . . . . . . . . . . . . Chlorinated Compounds . . . . . . . . . . . . . . . . . .

. . . . . .

32 34 38 40 41 42

4

Toxicity of the Effluents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

43

5

Ain Emissions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

45

6

Removal Strategies

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

46

7

Conclusions and Future Recommendations . . . . . . . . . . . . . . . . . . . .

48

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

49

References

. . . . . .

. . . . . .

. . . . . .

. . . . . .

. . . . . .

. . . . . .

. . . . . .

. . . . . .

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. . . . . .

Abstract This chapter is focused on the problem caused by the effluent discharges from paper and pulp mills. At present, three aspects should be considered in paper and pulp wastewater management: (1) the toxicity and high BOD5 of whitewaters and effluents; (2) the lack of knowledge on specific compounds responsible for the toxicity of the liquid and solid residue (sludge) and (3) the difficulty of treating whitewaters, which are characterized by the presence of suspended solids, colour odour, a high organic content, and an overall high toxicity. This chapter attempts to give an overview of organic compounds that contribute to the toxicity of paper mill waters and effluents, their levels, toxicological characterization and the methodologies used for their analysis. Families of compounds that are included are natural compounds such as resin and fatty acids, lignins, lignans and carbohydrates, and additives used during paper making such as surfactants, biocides and slimicides. In addition, part of the chapter is devoted to describing the wastewater treatment strategies used to decrease the toxicity and BOD5 of the effluents, which are used to indirectly phase out toxic organic pollutants from paper and pulp whitewaters (Table 1). Keywords Paper mill · Whitewaters · Effluents · Organic compounds · Analytical methods · Toxicity · Treatment

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Table 1 List of acronyms ordered by compounds and techniques

Compounds AOX APEO BOD BPA BSTFA COD DBNPA DCM DHA DS ECF FA Kj-N LAS MBT MFO MIB MTBE NOx NP NPEC OC OP OPEC PCB PCDBT PCDD PCDF PCP PFB RA TCA TCDD TCF TCMTB TOC VSC VOC

Absorbable organic halides Alkylphenol ethoxylate Biochemical oxygen demand Bisphenol A Bis(trimethylsilyl)trifluoroacetamide Chemical oxygen demand 2,2-Dibromo-3-nitrilopropionamide Dichloromethane Dehydroabietic acid Dry solid Elementary chlorine- free Fatty acids Kjeldahl nitrogen Linear alkylbenzene sulfonates Methylene-bis-(thyiocyanate) Mixed-function oxidase Methylisoborneol Methyl tert-butyl ether The sum of nitrogen oxide (NO) and nitrogen dioxide (NO2) expressed as NO2 Nonylphenol Nonylphenol ethoxycarboxylate Organochlorine Octylphenol Octylphenol ethoxycarboxylate Polychloroinated biphenyls Polychlorinated dibenzothiophene Polychlorinated dibenzo-p-dioxins Polychlorinated dibenzo-p-furans Pentachlorophenol Pentafluorobenzyl Resin acids 2,4,6-Trichloroanisole Tetrachloro dibenzo dioxin Totally chlorine- free 2-(Thiocyanomethylthio)-benzothiazole Total organic carbon Volatile sulphur compounds Volatile organic compounds

Organic Compounds in Paper Mill Wastewaters

27

Table 1 (continued)

Techniques APCI CZE ECD EROD ESIP FID GC GPC HRGC HRMS IS LC LLE LOD MEKC MS SPE SPME TU

Atmospheric pressure chemical ionization Capillary zone electrophoresis Electronic-capture detector Ethoxyresorufin-O-deethylase Electrospray ionization Flame ionization detector Gas chromatography Gel-permeation chromatography High-resolution gas chromatography High-resolution mass spectrometry Internal standard Liquid chromatography Liquid–liquid extraction Limit of detection Micellar electrokinetic chromatography Mass spectrometry Solid-phase extraction Solid-phase micro extraction Toxicity units

1 Pulp and Paper Mill Wastewaters The pulp and paper industry is the sixth largest polluter (after the oil, cement, leather, textile and steel industries), discharging a variety of gaseous, liquid and solid wastes into the environment [1]. The main environmental issues are emissions to water and air, sludge build-up and energy consumption. It is the pollution of water bodies, however, which is of major concern because large volumes of wastewater are generated for each metric ton of paper produced, depending on the raw material, finished product and extent of water reuse. Untreated paper mill effluent discharges cause considerable damage to the receiving waters, since they have high biochemical oxygen demand (BOD), chemical oxygen demand (COD), chlorinated compounds (measured as adsorbable organic halides, AOX), suspended solids (mainly fibres), fatty acids, tannins, resin acids, lignin and its derivatives, sulphur and sulphur compounds, etc. [1]. While some of these pollutants are naturally occurring wood extractives (tannins, resin acids, lignin), others are xenobiotic compounds that are formed during the process of pulping and paper making (chlorinated lignins, resin acids and phenols, dioxins and furans) [2, 3]. Some of the pollutants listed above, notably polychlorinated dibenzodioxins (PCDD) and dibenzofurans (PCDF), are recalcitrant to degradation and tend to persist in nature.

28

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The accumulation of organic matter in whitewaters is due to the paper-making process itself. Figure 1 depicts a mass stream overview of a pulp and paper mill. In paper mills, paper is made from wood, pulp or recycled paper by mixing the raw material with water to a fibre suspension which is ground, diluted and evenly distributed on the wire of a paper machine. On the wire, the pulp is dewatered and the paper sheets start to form, with a final dryness of 90–95%. Only 60–70% of the fibres are retained on the wire, and the rest end up in whitewaters that must be recovered and recycled to the paper machine to achieve maximum productivity. Due to this process, whitewater allows build-up of organic matter in the system, which causes a number of troubles in paper production such as neutralization of cationic retention chemicals due to the anionic nature of organic matter, growth of microorganisms due to high concentrations of organic matter (causing depletion of oxygen and production of hydrogen sulphide), formation of biofilms and generation of “stickies” due to lipophilic extractives which are accumulated in the paper [4]. In addition, whitewaters produce corrosion in the paper machine. The type and amount of organic compounds in whitewater and effluents depend on the raw material, paper-making process, additives used and type of energy supply. On average, paper production generates from 10 to 50 m3 of wastewater per ton

Fig. 1 Mass stream overview of a pulp and paper mill. The presence of some substances depends on the raw material, paper-making process, additives used and type of energy supply

Organic Compounds in Paper Mill Wastewaters

29

of paper [5]. Wastewater is either released untreated, treated or recycled. Pulp and paper mill effluents discharged into freshwater, estuarine and marine ecosystems alter aquatic habitats, affect aquatic life and adversely impact human health. Chronic sublethal toxic effects measured as increased liver mixed-function oxidase activity (MFO) and symptoms of altered reproductive capacity in fish and aquatic invertebrates have been detected in the discharge of treated pulp mill effluents [6]. In addition to that, large amounts of solid wastes are generated as by-products of the wastewater treatment plants. Mean emission levels are 0.3–1 kg/ton paper [5]. Primary sludge (solids removed during physical wastewater treatment prior to biological treatment) is rich in wood fibre and volatile solids. Secondary sludge (product of biological treatment) may contain organochlorine compounds (chlorophenols, chlorocatechols, chloroguaiacols, chlorovanillins and chlorosyringaldheydes) as well as trace levels of some dioxin and furan congeners, generated by chlorine bleaching [7]. There is ample evidence of the adverse health and environmental effects linked to organochlorinated compounds, often related to endocrine disruption (growth retardation, thyroid dysfunction, decreased fertility, feminization or masculinization of biota, etc.) [8]. Table 2 reports some common parameters used for the characterization of paper mill effluents. Biodegradable organic carbon, associated with families of non-chlorinated organic materials, is measured by biochemical and chemical oxygen demand (BOD, COD) methods. Some of the compounds found in paper mill effluents and sludge, including chlorinated compounds and several wood extractive constituents found in pulp liquors, are refractory (resistant to rapid biological degradation) and thus not measurable by the BOD5 analytical method.At present, there is a lack of information on the characterization of the Table 2 Average levels of several parameters used for the characterization of whitewaters and sludge

Type of pollutant

Typical example

Levels

Air emissions

Malodorous gases of reduced sulphur compounds, measured as total reduced sulphur Particulate matter Sulphur oxides/nitrogen oxides Volatile organic compounds (VOCs)

0.3–3 kg/t of air-dried pulp 75–150 kg/t 0.5–30/1–3 kg/t 15 kg/t

Liquid effluents

Biochemical oxygen demand (BOD) Chemical oxygen demand (COD) Absorbable organic halides (AOX) Kj-N P total

10–40 kg/t 10–60 kg/t 0–4 kg/t 3–13 mg/L 0.5–1.8 mg/L

Solid wastes

Sludge from primary and secondary treatment

50–550 kg/t

30

A. Latorre et al.

toxic organic fraction of pulp and paper mill effluents to evaluate their potential effects towards the environment, and the specific compounds, groups of compounds or organic fractions that cause harmful effects should be determined. The objectives of the proposed chapter are to describe the methods used to characterize pulp and paper mill effluents chemically and toxicologically, with the ultimate goal of obtaining a deep knowledge on the compounds responsible for toxicity, their concentration in the different types of industries and treatment methods used for their removal.

2 Legislation Related to Pulp and Paper Mill Industries As highlighted by the EU, in order to improve the management and control of industrial effluents, the Council Directive 96/61/EC on integrated pollution prevention and control (IPPC Directive) [5] is implementing the best available technologies to ensure a high level of protection of the environment. The IPPC Directive has started an exchange of information between EU member states and the pulp and paper industries concerning the best available techniques aimed at: (1) improving the quality of water by installation of treatment plants; (2) minimizing water consumption, since the pulp and paper mill industry is the second highest consumer of fresh water in Europe, generating six billion m3 of wastewater annually; (3) resolving the chronic toxicity and ecotoxicity associated with paper and pulp effluents; and (4) reducing the amounts of additives used (or substituting less toxic compounds). As a result, this sector still requires data and investment to improve water quality, which will be performed by revising and establishing emission limits. To assess these necessities, and to support the European IPPC Directives [5], new analytical techniques are emerging to determine those pollutants which may induce toxicity, may negatively influence water treatment or may affect paper production. Historically, the pulp and paper industry throughout the world has been regarded as particularly polluting to aquatic environments. Until the 1950s, it was common for pulp mills and many other industries to discharge untreated, toxic effluents directly into rivers and seas [9]. Nowadays, the development of new technologies directed to the treatment of industrial wastewaters is a European Community priority, which aims to reduce the organic pollution generated by industrial activities, and in the last instance, to reuse effluent waters. Since 1999, member states have been encouraged to comply with Directive 76/464/CEE [10] (and daughter Directives 86/280/CEE, 88/347/CEE and 90/415/CEE), as well as the Water Framework Directive (WFD) [11] related to the monitoring of toxic, persistent compounds with high accumulation potential. The IPPC Directive [5] have the objective of reviewing and implementing strategies and measures to control the sources of pollution and improve the quality of water. Some European countries (e.g. Portugal, France, Germany) have started to analyse the

Bleached Kraft pulp Bleached sulphite pulp

Bleached Kraft pulp

Italy

United Kingdom

Bleached sulphite pulp

Bleached Kraft pulp Bleached sulphite pulp

Bleached sulphite pulp

Bleached Kraft pulp

Bleached sulphite pulp

Bleached Kraft pulp

Ireland

Germany

France

Bleached Kraft pulp

Austria

Bleached sulphite pulp

Type of pulp

Country

No achievable guidance levels proposed No achievable guidance levels proposed

160 mg/L 160 mg/L

No limit No limit

Exist: 40 kg/t New: 25 kg/t Exist: 40 kg/t New: 25 kg/t

Exist: 65 kg/t New: 20 kg/t Exist: 45 kg/t New: 35 kg/t

Exist: 30 kg/t New: 20 kg/t Exist: 40 kg/t New: 25 kg/t

COD

10–50 mg/L

10–50 mg/L

40 mg/L 40 mg/L

90% removal or 50 mg/L 90% removal or 50 mg/L

Exist: 35 mg/L New: 30 mg/L Exist: 35 mg/L New: 30 mg/L

Exist: 3.98 kg/t New: 3 kg/t Exist: 6.5 kg/t New: 5 kg/t

Exist: 3 kg/t New: 2 kg/t Exist: 3 kg/t New: 2 kg/t

BOD5

10–50 mg/L

10–50 mg/L

80 mg/L 80 mg/L

No limit No limit

No values (part of COD) No values (part of COD)

Exist: 6.5 kg/t New: 5 kg/t Exist: 6.5 kg/t New: 5 kg/t

Exist: 5 kg/t New: 2.5 kg/t Exist: 5 kg/t New: 2.5 kg/t

TSS

<1.5 kg/t

<1.5 kg/t

No requirements No requirements

0.1 mg/L 0.1 mg/L

Exist: 2.5 kg/t New: 0.25 kg/t

Exist: 0 kg/t

1 kg/t (yearly average)

1 kg/t (yearly average)

Exist: 2.5 kg/t New: 0.25 kg/t Exist: 0.2 kg/t New: 0.1 kg/t

AOX

Table 3 Current national discharge limits and proposal levels for production of bleached Kraft and bleached sulphite pulp [5]

Organic Compounds in Paper Mill Wastewaters 31

32

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levels of various priority pollutants in surface waters, sediments and biota according to the European policy. Most countries still lack relevant and precise data on the pollution generated by the paper and pulp industries, and identification of the specific compounds responsible for the toxicity of the effluent is a subject still to be covered. Environmental limits or guidelines for the pulp and paper industry vary significantly between European countries, despite efforts to create a more uniform system [5]. As an example the discharge limits for bleached Kraft and bleached sulphite pulp are given in Table 3. In some countries, such as Austria or France, there is different legislation depending on the type of paper mill. Normally, the most restrictive corresponds to the bleached Kraft pulp. In other countries, such as Ireland, there are no requirements or limits for some parameters. In 1997, the US EPA finalized a new set of federal guidelines, the so-called cluster rules [12]. The guidelines contain limits for 12 different types of mills, each of them distinguished by four different technical levels depending on the type of technology applied. Limit values are given as pollutant load expressed as kilograms of pollutant per tonne of product, distinguishing maximum values per day and average monthly values. The regulations apply to any pulp, paper or paperboard mill that discharges process waters. Compared to recent permit requirements in Europe (see Table 3) the limitations for existing mills set in the US cluster rules are lenient. There is special legislation for other countries. For example Canada, one of the most important pulp and paper producing countries, has set several general guidelines, and the actual limits and guidelines can be different in each state [5]. The wastewater limit systems in Canada are mostly based on load kilograms per tonne production. Different limits are applied for different types of mills. Some parameters are measured as concentrations. Toxicity limits are also used.

3 Chemical Characterization of Pulp and Paper Mill Waters The analysis of organic pollutants in whitewaters and effluents requires procedures which can eliminate suspended matter and fibres and still permit the extraction and efficient recovery of target analytes. Whitewaters and effluents, especially from closed-cycle systems, are characterized by a very high total organic content (TOC) (up to 5,000 mg/L), by the presence of particulate matter and by the formation of microfibres which, if not eliminated, may affect the extraction efficiency. To remove suspended matter and particles, sample filtration through 1, 0.7 and 0.45 mm filters is necessary, or otherwise centrifugation at 2,000 rpm for 20 min. Most apolar compounds might be retained in the particulate fraction of the sample and thus, the filter should also be extracted [13]. Due to the large amount of organic matter, the extraction pro-

Organic Compounds in Paper Mill Wastewaters

Fig. 2 Chemical structure of organic compounds identified in paper mill effluents

33

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cedure must be adapted to this matrix and the chemistry of the target compounds should be considered. The chemical structure of the most representative compounds of each family are shown in Fig. 2. Table 4 summarizes the extraction methods for different families of chemicals generally found in paper mill effluents, pointing out the detection limit and the percentage of recovery. 3.1 Biocides Depending on the type of paper produced, it is common to dose biocides for wood preservation and during paper making to decrease the problems related to microbial, fungal and algal growth. This means an undesired handling of toxic chemicals and the risk of negative effects in the receiving water, as biocides are discharged with the wastewater. Furthermore, high concentrations of biocides in whitewater may diminish the efficiency of secondary treatment, available in some paper mills. There are two main classes of biocides [14]. On the one hand, there are oxidizing agents such as chlorine dioxide and hydrogen peroxide. These happen to be the same chemicals that are widely used for pulp bleaching. The oxidizing action either kills the bacteria and fungi outright or it weakens the cell walls so that they are more susceptible to the other main classes of biocides. The other class involves highly toxic organic chemicals, such as thiocyanates, isothiazolins, cyanobutane, dithiocarbamate and bromo compounds, which are used for wood preservation in place of traditional chlorophenols [15].A possible third category consists of materials that have an ability to inhibit biological film formation, e.g. surfactants such as alkyl sulphosuccinates. The size and type of the paper mill (recycle, Kraft, pulp, etc.) and open/ closed circuits are crucial to determine the type of biocides to be used and their doses. The fate of biocides is as follows: a fraction will degrade (chemically or biologically), a fraction will remain in the circulating waters and finally, a fraction will be present in the effluent or remain in the solid matter. An additional problem is that, due to their physicochemical properties, some biocides may be fibre retentive and can accumulate in the final paper product [16]. Little information is available on biocides in pulp and paper mill whitewaters and effluents due to the complexity of the water matrix, but generally LC techniques are employed since paper mill biocides are usually highly soluble and polar compounds. A recent study recommends the use of LC–ESI-MS in the positive-ion mode for the determination of TCMTB and DBNPA in effluent waters [17]. Figure 3 indicates the average level of biocides found in paper mill waters. DBNPA and TCMTB were detected in process waters of a recycling paper mill at concentrations of 8–116 and 2–4 mg/L, respectively [18]. MBT was detected at concentrations of 0.19 and 0.02 mg/L in whitewater and primary effluent, respectively, after SPE and LC-UV detection [17]. The same method was found to be suitable for determining this product in paper, after

Matrix

Paper food packaging

Paper-recycling process waters

Fortified tissue paper Secondary-treated effluent

Surface-treated lumber

Paper mill effluents

Primary effluent

Paper mill effluents and river waters

Effluents from bleaching processes

Whitewater, effluents

Wood resin

Paper-recycling process waters

Compounds

DBNPA

TCMTB DBNPA

MDC

TCMTB

RA and FA

FA esters

RA

RA and FA

RA and FA

RA and FA

RA and FA

LLE with MTBE, followed by derivatization step

Extraction with DCM, followed by derivatization step

LLE with MTBE Direct injection

LLE with MTBE, followed by derivatization step

SPE, followed by derivatization step

SPE LLE with DCM

LLE with DCM, followed by derivatization step

Extraction with ACN

Hot water extraction followed by SPE

SPE

Hot water extraction

Extraction

Method

GC

LC

LC

GC

GC LC

GC

GC

LC

LC

LC

MEKC

Separation

MS

MS

(APCI)-MS

FID

MS Fluoresc.

MS

MS

UV

UV

MS

UV

Detection

Table 4 Current methods for the isolation and analysis of toxic compounds found in papermill waters

0.007–0.2 mg/L

6–12 mg injected

0.3–32 mg/L 0.5–81.3 mg/L

n.r.

n.r. 0.001–0.02

n.r. n.r.

n.r.

n.r.

81–106

n.r.

70–101 75–95

n.r.

75 91–95

n.r. n.r.

n.r.

96.2

51 88

n.r. n.r.

1.5 mg/L 80 mg/L n.r. 0.01 mg/L

53.0

17, 32

29

23

27, 28

26 mg/L

25

21

19

18

17

16

Recoveries Ref. (%)

1700 mg/g

LOD

Organic Compounds in Paper Mill Wastewaters 35

Matrix

Paper-recycling process waters

Wastewater from paper production

Paper mill effluents

Pulp fibres 1%

Paper mill effluents

Paper mill effluents

Sediments from pulp and paper mill

Bleached pulp mill effluents

Compounds

NP1EC, OP1EC NP, OP, LAS

BPA

NPEC

Hemicelluloses

Lignin

Lignin

OC, PCB, CDPE, PAH, PCDD, PCDF

PCDBT

Table 4 (continued)

Soxhlet extraction of filters with toluene

GPC

CuO degradation, followed by a derivatization step

Permanganate degradation, followed by LLE with acetone/DCM

Acid methanolysis with HCl, followed by a silylation step Hydrolysis and methylation reaction

SPE

SPE

SPE

Extraction

Method

HRGC

HRGC

GC

CZE

GC

GC

GC

LC

Separation

HRMS

HRMS

MS

n.r.

n.r.

90–110

n.r.

<90

<10 ng/L

n.r.

n.r.

n.r.

n.r.

n.r.

54

53

48

46

39

38

36

35

17

Recoveries Ref. (%)

0.55–2.9 mg/L n.r. 0.05–0.30 n.r.

n.r.

MS UV MS

n.r.

0.2–2 mg/L

n.r.

10–80 mg/L

FID

MS

MS

ESI-MS

Detection

LOD

36 A. Latorre et al.

LLE with Hexane HS-SPME

Nesting along contaminated rivers

Effluent from bleaching processes

Pulp mill effluents

Contaminated sediment

Sediment

Air, water and sediments

Pulp mill effluents

River

Drinking water

Chlorophenols, chloroguaiacols, chlorovanillin, chlorocatechol

Chlorophenols, chloroguaiacols, chlorosyringol

Chlorophenols, chloroguaiacols, chlorovanillin, chlorocatechol

Chlorophenols, chloroguaiacols, chlorocatechol,

VSC

VSC

MIB, geosmin

HS-SPME LLE with DCM

Cryogenic trap

Extraction with N-hexane, followed by derivatization step

Dean–Stark Soxhlet extraction

SPE

LLE with diethyl ether/acetone, followed by derivatization step

GPC

Extraction

OC, PCB, PCDD, PCDF

Method

Matrix

Compounds

Table 4 (continued)

GC

GC

GC GC

GC

GC

GC

LC

GC

GC HRGC

Separation

MS

MS

MS MS

FID

MS

MS

Amperometric electrode

FID

EDC HRMS

Detection

1.2–3.3 ng/L

0.5 ng/L

0.7–5 ng/L n.r.

10 pg/L

n.r.

n.r.

0.4–6 mg/L

n.r.

n.r. n.r.

LOD

n.r.

97–103

n.r. n.r.

n.r.

n.r.

n.r.

84–100

n.r.

n.r. n.r.

73

71

69 70

68

57

52

37

27

55

Recoveries Ref. (%)

Organic Compounds in Paper Mill Wastewaters 37

38

A. Latorre et al.

Fig. 3 Levels of different families of organic compounds in paper mill effluent waters

extraction with boiling water. The LC-UV detector was also applied to determine TCMTB and chlorophenols [19] from lumber surfaces after extraction of sticks (1¥2 mm) with acetonitrile. Micellar electrokinetic chromatography (MEKC) was optimized for the determination of ten biocides in paper food packaging, with the inherent advantages of rapidity, simplicity and no use of toxic reagents [16]. 3.2 Resin and Fatty Acids Wood extractives include lipophilic (fatty and resin acids, sterols, steryl esters and triglycerides) and hydrophilic (lignans, low-molecular-mass lignins, ligninlike substances and hemicelluloses) compounds that dissolve in whitewaters during paper production. Among them, resin and fatty acids have a high tendency to form pitch deposits and stickies that alter the machine functioning and decrease the paper physical properties (tensile strength, opacity, brightness, etc.) [20]. On the other hand, wood extractives accumulated in whitewaters can end up in the effluent, and are potential toxicants to biota. Figure 3 shows the concentrations of resin and fatty acids in whitewaters. The levels detected de-

Organic Compounds in Paper Mill Wastewaters

39

pend on the paper-making procedure. Concentrations of 20 to 400 mg/L have been detected in water [16, 20, 21] and 1,474 mg/g d.w. in sediment [22]. These compounds are not removed by primary flocculation whereas a 50% decrease or more is observed after biological treatment, showing the effectiveness of this treatment in reducing the concentration of these type of compound [23]. The analysis of resin and fatty acids is fully reviewed elsewhere [24]. The structures of the most representative resin and fatty acids, dehydroabietic and palmitic acids, are shown in Fig. 2. Liquid–liquid extraction (LLE) has been used to extract resin and fatty acids [25]; methyl tert-butyl ether provides an excellent solvent for the extraction of these compounds [26–29]. Whitewater pH is generally between 6 and 8, and extraction can be performed under neutral or alkaline conditions [30], although acidic conditions avoid microbial growth during sample storage [28]. The determination of resin acids in wood extractives has been traditionally performed by GC with a flame ionization detector (GC-FID) on either a DB1 or HP5 analytical column [31]. This procedure permits determination of the different compounds (resin acids and hemicelluloses) at microgram per liter levels. With non-selective detectors, compound identification is performed by retention time comparison against a standard. However, the analysis of complex water matrices leads to interferences and coelutions, which make identification and quantification difficult. Therefore, two columns of different polarity should be used for compound confirmation. This situation changes if an MS detector is used. Spectra with molecular fragment or cluster ions are generated which provide structural information on the ionized compound. With GC, derivatization of the extract is needed either using methylation agents such as diazomethane, which presents severe health hazards, or bis(trimethylsilyl)trifluoroacetamide (BSTFA) combined with trimethylchlorosilane. Another derivative of resin acids (RAs) is the pentafluorobenzyl (PFB) esters. Figure 4a shows a GC–MS chromatogram which permits the complete separation of 15 resin and fatty acids, after derivatization with BSTFA. The main problem is that the derivatized extract has a very short halflife, generally of 12 h, and the sample has to be derivatized once more if the injection sequence fails. In addition, the derivative may affect the long-term performance of the GC–MS system which will necessitate extra cleaning. GC–MS permits identification of resin and fatty acids from whitewaters of different paper mills [32]. Given the poor volatility of resin and fatty acids, liquid chromatography coupled to mass spectrometry in the negative-ion mode can be used without the need for derivatization [31, 34]. APCI and ESI are suitable interfaces, although no fragmentation is observed even at high fragmentor voltages. Figure 4b shows a LC–MS chromatogram. There is coelution of non-aromatic RAs when using either C18 or C8 columns, but this is not a problem since the concentration of total RAs indicates the quality of paper mill water. Depending on the type of the paper mill process the concentration of resin and fatty acids range wideley. For example, downstream of a bleached Kraft mill

40

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a

b Fig. 4a, b (a) GC–MS total ion chromatogram in EI of a standard containing resin and fatty acids at 7 mg/mL, after derivatization with BSTFA. (b) LC–APCI-MS total ion chromatogram of the same standard. Identification numbers: 1=palmitic acid; 2=margaric acid; 3=linoleic acid; 4=oleic acid; 5=stearic acid; 6=pimaric acid; 7=sandarocopimaric acid; 8=isopimaric acid; 9=palustric acid; 10=levopimaric acid; 11=dehydroabietic acid; 12=abietic acid; 13=neoabietic acid; 14=chlorodehydroabietic acid; and 15=dichlorodehydroabietic acid

effluent discharge, the concentration of resin acids found in river sediment is 139 mg/g. [22]. This level is ten times lower compared with the level found in the sludge of this Kraft mill. A similar behaviour was observed for the analysis of resin and fatty acids in effluent and river water closed to the paper mill. The concentration decreases in most cases, due to degradation and dilution. Only for some resin acids, such as dehydroabietic acid, was the concentration observed 11 km downstream similar to the level found in the source, due to their high stability [26]. 3.3 Surfactants and Plasticizers Surfactants, such as linear alkylbenzene sulfonates (LAS) and alkylphenol ethoxylates, are present in whitewaters because of their use as cleaning agents or as additives in antifoamers, deinkers, dispersants, etc. The non-ionic surfactants alkylphenol ethoxylates (APEO) degrade to nonylphenol (NP) or to a

Organic Compounds in Paper Mill Wastewaters

41

lesser extent, to octylphenol (OP), which are considered as persistent environmental pollutants. LAS and APEO have been detected in whitewaters of paper mills at concentrations up to 5,000 mg/L for the former [16] and from 0.3 to 10 mg/L for NP and OP [35]. The total concentration of nonylphenol ethoxycarboxylates (NPEC) in paper mill effluents ranged from below detection to 1,300 mg/L [36]. These compounds can also be analysed by LC–MS, with the additional advantage that long-chain alkylphenol ethoxylates and carboxylates can be simultaneously determined [34]. Due to the complexity of the matrix, the identification and quantification of compounds should be controlled by the addition of an adequate surrogate or internal standard. Heptylphenol can be used for the analysis of alkylphenols. Paper mill whitewaters and effluents are rich in bisphenol A (BPA), which is used in great quantities for the production of epoxy resins and polycarbonate plastics. Its presence in effluents has been reported as a result of its use in the manufacture of thermal paper or due to migration from plastic containers at the high water temperatures of whitewaters [35]. This compound is preferably analysed by GC–MS. The levels encountered in paper mill effluents are between 28 and 72 mg/L [36, 37]. Another study revealed levels up to 226 mg/L [33]. Special in vitro test systems and animal experiments have demonstrated a weak oestrogenicity for BPA. Since aquatic wildlife could be endangered by paper mill waste discharges at the concentration that BPA is found, its survey in paper mill effluents should be taken into consideration. 3.4 Lignin and Hemicelluloses Wood consists mainly of cellulose, hemicelluloses and lignin in various proportions. The amounts and compositions of these component groups depend primarily on the wood species [38]. In chemical pulping, a significant part of the hemicelluloses is dissolved from the fibres into the pulping liquor. The rest remains in the fibre or is adsorbed into it, significantly affecting the properties of the cellulose fibres or paper produced [39]. It has been demonstrated that the presence of soft or hardwood hemicelluloses in the cellulosic pulp can improve some features of paper making. The plasticity and the high superficial area conferred by hemicelluloses result in an increased binding among the fibres and a higher tensile strength in the paper sheet. However, high amounts of hemicelluloses seem to be deleterious to the mechanical properties of the paper due to a decrease in the individual fibre resistance, and to the optical properties due to the low opacity in the paper sheet [40]. Some studies have demonstrated a relationship between the degradation of hemicellulose components, such arabinose and mannose, and wood strength losses. The significant reduction in strength observed during incipient decay of wood by brown rot fungi is therefore likely to be due to hemicellulose decomposition [41]. A basic method for the analysis of hemicelluloses is the determination of their constituent sugar residues obtained by acid hydrolysis

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or methanolysis (hydrochloric acid in anhydrous methanol). The liberated monosaccharides are converted into the corresponding methyl glycosides, and carboxyl groups of uronic acids are esterified with methyl groups. Thereby, methanolysis gives the advantage of a reasonable stability of the released methyl glycosides, and allows simultaneous analysis of acid and neutral sugars by capillary GC–MS or LC after suitable derivatization [42]. On the other hand, lignins are natural polymers in plant cell walls and represent, after cellulose, the most abundant polymer in nature, with a very complex structure. The lignin composition will be different not only among plants of different origin, but also among different tissues of an individual plant [43]. It is formed by removal of water from sugars to create aromatic structures. Lignin resists attack by most microorganisms, and it is a main component of pulp and paper mill effluent waters. Lignin produces coloured waters [44], unlike hemicellulose, and is an undesirable polymer whose removal during pulping requires high amounts of energy and chemicals [45]. Extracted lignins from non-wood fibres are potential raw materials for new industrial applications [43]. As a result, lignin should be monitored. In order to obtain information about the structure of lignin there were some studies based on the oxidative treatment of the molecule by potassium permanganate [46]. This method involves the selective degradation of all aliphatic side chains attached to aromatic groups in lignin, resulting in the formation of a mixture. The identification of these as well as the amount of each individual acid provides information about the substitution pattern in a particular lignin. From the degradation product obtained it is possible to deduce the structure of lignin. Up to now, the mixture of aromatic acids obtained from permanganate oxidation of lignin has been analysed by GC after esterification [47], but Javor et al. developed a new methodology using capillary zone electrophoresis (CZE), which provides rapid results with the avoidance of time-consuming preparation of esters of the resulting aromatic acids. Another methodology, based on the CuO degradation of lignin, is also considered a suitable technique for their analysis [48].After CuO degradation of lignin, nine products corresponding to three lignin units (p-hydroxyphenyl, guaiacyl and syringyl) could be identified. The degradation products can be easily derivatized, separated by GC and identified by MS. 3.5 Chlorinated Compounds Since the late 1970s, much emphasis was put on the role of chlorinated substances formed in the bleach plants. Bleaching effluents from bleached chemical pulp plants are one of the remaining pollution problems of pulp mills due to the large amounts of chlorinated organic matter discharged into the environment [49]. The bleaching of chemical pulp is accomplished in several stages, to some of which chlorine is added in different forms. The chlorine reacts with lignin and other organic matter present in the pulp giving chlorinated compounds. During the last decade, there has been a drastic decrease in the use of

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molecular chlorine as bleaching agent, which has been replaced by chlorine dioxide, molecular oxygen, peroxide and ozone. This has led to a decrease in adsorbable organic halides (AOX), which is the main parameter used by regulatory agencies to determine the discharge of chlorinated organics [50]. Reduction of AOX has also been achieved by the installation of treatment plants. Current trends are directed towards closed-cycle systems using either elementary chlorine-free (ECF) or totally chlorine-free (TCF) bleaching pulp. However, this is only possible if no chlorinated agents have been used within the process. For some time chlorophenols, and especially pentachlorophenol (PCP), were used as wood preservatives, and as a result they have been encountered in the water [51] and sediments [52] of several paper mills. However their use is now restricted and the wood chain is organized so that wood is rapidly consumed and the use of fungicides is minimized. The presence of chlorine and chlorinated compounds is also the source of dioxins and furans during paper making, and these compounds have been detected in sediments in the vicinity of a pulp and paper mill [53] and in effluents, along with polychlorinated dibenzothiophenes [54]. A recent study found high concentrations of PCDD and PCDF along with PCP in nestling tissue (Tachycineta bicolor) collected downstream of paper pulp mills, suggesting that the primary source of contaminants was the use of PCP for timber preservation [55]. In addition, it has been shown that dioxins bioaccumulate in fish downstream of pulp and paper mills [56]. The levels of chlorinated compounds of different families are shown in Fig. 3. The survey of PCP and other chlorinated compounds has been traditionally performed by the measurement of AOX, which gives a measure of the total chlorinated organic compounds [57]. Typical AOX levels are between 0.01–0.1 kg/t. However, to specifically determine the different families of organochlorinated compounds in paper mill whitewaters and effluents, several analytical methods have been developed. Current official methods for the analysis of chlorophenols, e.g. US-EPA 604, 625 and 8041, are based on LLE followed by GC using electron capture detection (ECD) or MS. However, there is a general trend to use SPE and LC to avoid the use of toxic organic solvents and derivatization procedures. A complete review of LC methods for the analysis of chlorophenols is given elsewhere [58]. Levels of chlorinated organic compounds in paper mill waters are between 1 and 100 mg/L, as shown Fig. 3. The analytical protocol for the analysis of dioxins and furans is well established and follows the EPA method 8280A.

4 Toxicity of the Effluents Some effects have been observed in fauna living close to paper mill discharges, such as skin and physiological diseases in fish and a decrease in the number of juveniles, changes in communities and population structure, changes in growth rates, and delayed sexual maturation and reproduction, among others [2, 58, 59].

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In addition, oxygen depletion is common in such effluents, causing anoxia to fish and other aquatic specimens. The toxicity of paper mill whitewaters and effluents can be measured using acute toxicity tests such as Microtox or ToxAlert. These tests measure the bioluminescence inhibition of Vibrio fischeri caused by the presence of different toxicants in the water sample. Toxic substances will cause changes in cell structures and/or metabolic pathways of marine Vibrio fischeri, which are rapidly reflected in a bioluminescence decrease. The LC50 value of several surfactants, resin acids, fatty acids and biocides has been determined by ToxAlert using individual compounds and mixtures, and the combination of chemical analysis and effect studies permitted the toxicity of whitewaters and effluents of several paper mills to be assessed [33]. Figure 5 represents the percentage of bioluminescence inhibition using ToxAlert and the total organic load of sample (sum of resin and fatty acids, surfactants and biocides) of an untreated effluent and whitewaters corresponding to different paper mills which had undergone several treatments. In cases where the concentration of organic compounds was high, a high percentage of bioluminescence inhibition was observed. On the other hand, in four samples the organic load was low, as well as the percentage of inhibition using ToxAlert. However, untreated Kraft and print paper showed a low organic compound load and a high toxicity. This is attributed to the presence of other compounds not considered

Fig. 5 Total organic composition (resin and fatty acids, biocides and surfactants) and percentage of bioluminescence inhibition of several types of waters (recycle, Kraft, print board in open and closed circuit) submitted to primary or biological treatment. Identification letters: A=effluent; B=recycle untreated; C=recycle primary treatment; D=Kraft untreated; E=Kraft biologically treated; F=print paper untreated; G=print paper biologically treated; H=board untreated; I=board biologically treated; J=board closed loop; and K=board biologically treated

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or detected with chemical analysis. Nevertheless it can be concluded that the combination of chemical analysis and bioluminescence inhibition assay permits evaluation of the quality and efficiency of treatments. The measurement of the ethoxyresorufin-O-deethylase (EROD) activity is another sensitive parameter to detect the effects of paper mill industrial effluents on living organisms in the receiving waters. The EROD activity is a measure of the activity of the cytochrome P-450 enzyme system, which plays a central role in the transformation and elimination of xenobiotics. Increased EROD activity has been shown as far as 40 km from pulp mills, and EROD induction in fish caused by pulp mill effluents remains after biological treatment [60]. It is specified that EROD activity and erythrocytic nuclear abnormalities are induced by abietic and dehydroabietic acid [60]. However, it is difficult to identify the chemical compounds that are responsible for these effects. LC50 values have been tested in several fish species and levels below 2 mg/L have been reported for resin acids [61] and below 0.1 mg/L for some biocides used in the paper industry, such as MBT and TCMTB [34]. Wood extractives (resin and fatty acids, sterols, etc.), diterpene alcohols and juvabiones account for 70–100% of the toxicity in various paper mill effluent streams [62]. However, toxicity depends on the treatment [63] and recent papers relate toxic effects towards aquatic biota due to the presence of resin acids in a secondary-treated bleached Kraft pulp mill effluent [64], and due to nonylphenol polyethoxy carboxylate metabolites of non-ionic surfactants in a US paper mill effluent [65]. Moreover, resin acids and, to a smaller extent, unsaturated fatty acids have been reported as major contributors to the toxicity of paper industry effluents to aquatic organisms, causing chronic sublethal toxicity, genotoxicity and potential bioaccumulation in fish tissues [65]. Endocrine disruption is being highlighted in modern toxicology. Relatively little is known about the potential endocrine effects of paper mill effluents on aquatic organisms. Field surveys on fish in proximity to sewage plants show hermaphrodism and laboratory studies also confirm this phenomenon [66].

5 Air Emissions The environmental impact of Kraft (sulphate) pulp mills associated with atmospheric pollution is due to the emissions of volatile reduced sulphur compounds (VSC) [67]. VSC are formed as a result of the anaerobic decomposition of organic matter, such as hydrogen sulphide (H2S), methyl mercaptan (CH3SH), dimethyl mercaptan (CH3SCH3) and dimethyl disulphide (CH3SSCH3). They are formed in water, and due to their volatility can be emitted to air. In general, these compounds have very low olfactive detection levels. This explains their detection by humans even in small quantities and at great distances from the emission sources. At encountered levels the toxicity of these compounds is negligible. However, being a nuisance, they are subjected to particular attention

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as the pulp and paper industry is continuously faced with more stringent air emission limits concerning gaseous pollutants. Most of the techniques for the analysis of VSC in aqueous matrices use the purge-and-trap method with cryogenic trapping of these analytes in glass tubes [68]. The use of solid-phase microextraction (SPME), working in the headspace (HS) mode, seems to be a good alternative to the traditional techniques [69, 70]. Abalos et al. [69] analysed effluents from a recycled paper mill, obtaining levels between 7 and 24 mg/L, with dimethyl sulphide being the main compound detected, which may originate from the sodium hydrosulphite and sodium metabisulphite used as bleaching reagents during the process. Lower levels were found in a bleached Kraft pulp mill effluent, with values around 0.5–2 mg/L [71]. The release of mill wastewater effluents may be a significant contributor to mill odours. One example of this pollution is the presence of two terpenoids, geosmin (trans-1,10-dimethyl-trans-9-decalol) and 2-methylisoborneol (MIB), caused by the presence of actinomycetes (bacteria) and blue-green algae (cyanobacteria). Both of these compounds are associated with water from spring runoff and/or eutrophic systems [72], and are responsible for the majority of the reported taste and odour events in surface waters close to paper mills. Current methods for detection and quantification at low levels require large sample volumes (100–1,000 mL) and intensive sample concentration procedures [71]. Recently, a HS-SPME–GC method was developed that minimized sample manipulation and time consumption [73]. Watson et al. analysed different paper mill wastewater treatment plants, obtaining a wide range of concentrations depending on the sampling point, with levels between 13 ng/L and 127 mg/L. Other compounds, such as 2,4,6-trichloroanisole (TCA), 2-isopropyl-3methoxypyrazine and 2-isobutyl-3-methoxypyrazine, were found downstream of a pulp mill effluent, and were considered as off-flavours. These compounds are by-products of chlorination, or can be produced by actinomycetes or other biota [74].

6 Removal Strategies For both environmental and economic reasons, many paper mills aim at lower water consumption and a decrease of water discharge. These can be achieved by recycling water but unfortunately, closing the water system is far from easy because an increased recycle of whitewaters leads to an accumulation of soluble organic matter and salts in the paper mill. The advantages and disadvantages of closed-cycle systems in paper mills are shown in Table 4 [61]. However, the problems derived from a build-up of organic matter in the whitewater systems has forced many mills that have been trying a closed-cycle approach to open up their systems again and continue to discharge great amounts of wastewater. The first effect of paper mill wastewater discharge is the depletion of oxygen in the receiving waters, caused by oxygen-consuming microbial degradation of

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readily biodegradable organic matter being discharged. This has led to a rapid interest in developing new methods for in-mill treatment of whitewater to remove organic matter.A review of the treatment of pulp and paper mill effluents indicates processes that minimize the discharge of wastewater into the environment [75]. Evaporation techniques and chemical treatment are very costly operations, and membrane filtration often suffers from fouling problems, which decrease the efficiency and increase the operating costs [76]. Biological treatment is undisputedly the most effective and economical way of removing great amounts of organic matter from wastewaters. The possibility of using in-mill biotreatment was proposed in the 1980s, and during the 1990s biological treatment and reuse of recycle fibre mill process water was applied in some mills [77, 78]. The objective of these treatments is to reduce BOD, which is the direct cause of oxygen consumption. However, being rather conventional, biological treatment plants operating under normal biological conditions (<40 °C, pH around neutral) require extensive modifications of the environmental conditions, such as cooling. This is costly and highly undesirable as it causes heat losses and less efficient production in the paper mill, which is optimally operated at higher temperatures. The insight into this has led to a number of studies on the possibility of operating in-mill treatment at higher temperatures [79–81]. It has even been possible to treat acidic whitewater with high efficiency at a pH as low as 3.5 [81]. Effluents from the mill are treated in bioprocesses such as aerated lagoons or activated sludge, whereas whitewaters undergo an anaerobic treatment followed by activated sludge. The efficiency of the treatments is controlled through measurements of generic parameters such as COD and BOD. It is assumed that removing as much of the organic matter as possible will solve the problem. BOD is removed to a great extent, generally more than 95%. Still, several problems related to the reuse of biologically treated whitewaters have been encountered: – Biological treatment removes the bulk of the organic matter, but the fraction remaining, often dominated by lignin, makes biotreatment difficult. This gives a significant increase in the colour of the treated water, and unacceptable colouring of the product for such paper qualities for which the colour is important. – Aerobic biotreatment effectively eliminates odours from organic acids and sulphide. However, in cases where biotreated water has been reused in paper production, the product has suffered from a weak “soily” smell that is unacceptable and has ruled out the continued use of biotreated water. – It is often necessary to dose nutrients into the bioprocess to achieve a good performance. However, this leads to nutrients entering the whitewater system with the reused water. As microbial activities in the whitewater systems are generally nutrient limited, the increased supply of nutrients may lead to a considerably increased growth of microorganisms and increased slime problems, rather than the decrease that is the aim of biotreatment [81].

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– In cases where water is discharged, considerable effects on life in the receiving waters, especially from chemical pulping effluents, have been encountered. Pathological changes in fish have been observed, as well as effects on physiological and biochemical parameters. It is obvious that these serious effects on the ecosystem are not due to readily biodegradable organic matter, but rather to compounds more resistant to biological treatment. These factors have stopped the installation of biological treatment plants at a number of large Kraft mills, which now continue to discharge great volumes of untreated effluent. However, there is a growing tendency to install advanced post-treatment stages to deal with the remaining problems. The combination of biological treatment and membrane filtration has found a special interest. Pauly and Kappen [82] studied the combination of thermophilic anaerobic treatment and ultrafiltration, and found the biotreatment improved the performance of filtration. However, the problems with odour remained. The improved performance of membrane filtration after biological treatment was also observed by Nuortila-Jokinen [76], in which case biotreatment was to be considered more of a pre-treatment and filtration the main treatment. However, combined biotreatment and advanced filtration, such as ultrafiltration and nano-filtration, are expensive solutions and large amounts of reject streams are formed which have to be further tested. Therefore, combined processes are undoubtedly needed, and it is important to identify the most cost-effective solutions that will give a satisfactory result for each type of paper production. Effective technologies should be directed towards (1) elimination of the organic compounds responsible for the toxicity of paper and pulp effluents and related emissions and (2) reduction of the amount of solid waste going to landfills.

7 Conclusions and Future Recommendations The pulp and paper industry is the greatest industrial polluter in terms of wastewater volumes and organic discharge. Compounds encountered in whitewaters are natural wood components such as resin and fatty acids, and additives added in the process such as wood preservatives, biocides and surfactants and plasticizers. Since the introduction of the best available technologies and according to the IPPC Directive, there has been an improvement in the pulp and paper sector such as minimization of the use of chlorine, additives, energy and fresh water which has lead to a reduction of emissions of toxic compounds to water, air and sludge. Generic parameters such as COD, BOD, AOX, total suspended soils, SO2 and NOx are systematically controlled and maximum discharge limits are well satisfied. However, a recent IPPC Reference Document on the pulp and paper industry indicates that there is insufficient information on the organic composition of whitewaters, effluents and sludge from pulp and paper mills, and on the sampling and analytical methods that should be used for their

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characterization. As indicated in this document, water quality among the different pulp and paper mills can only be assessed by measuring legislated parameters (BOD, COD, metals etc.) and the specific organic/inorganic composition, and by the toxicological characterization of whitewaters, effluents and solid waste. This chapter has attempted to give an overview of the organic compounds present in pulp paper mill whitewater, the levels encountered and their toxicological implication. It has also highlighted the treatments performed and the tools which are nowadays used to remove COD, toxicity and organic load of pulp and paper mill whitewater for an environmentally friendly paper production process. Recently, much effort has been devoted to correlating toxicity studies and the chemical characterization of pulp and paper mill effluents. This permits a much stricter control of the treatment that should be performed and of the quality of the water, which still in many cases is discharged to the environment. Acknowledgements This study has been supported by the EU Energy, Environmental and Sustainable Development Program (CLOSEDCYCLE, Contract No. EVK1-2000-00749) and Ministerio de Ciencia y Tecnología (PPQ2000-3007-CE). T. Welander and A. Malmqvist are acknowledged for providing information on pulp and paper mill treatments.

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The Handbook of Environmental Chemistry Vol. 5, Part O (2005): 53– 77 DOI 10.1007/b98607 © Springer-Verlag Berlin Heidelberg 2005

Evaluation of Pesticides in Wastewaters. A Combined (Chemical and Biological) Analytical Approach M. D. Hernando1 · I. Ferrer2 · A. Agüera2 · A. R. Fernandez-Alba2 (✉) 1

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Department of Environmental Chemistry, IIQAB-CSIC, Jordi Girona 18–26, 08034 Barcelona, Spain Department of Analytical Chemistry, University of Almería, 04120 Almería, Spain [email protected]

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Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

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Chemical Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . Sample Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . Liquid–Liquid Extraction . . . . . . . . . . . . . . . . . . . . . . . Solid-Phase Extraction . . . . . . . . . . . . . . . . . . . . . . . . Semipermeable Membrane Devices and Other Membrane Processes Cleanup Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . Methods of Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . Gas Chromatography . . . . . . . . . . . . . . . . . . . . . . . . . Liquid Chromatography–Mass Spectrometry (LC–MS) . . . . . . .

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Abstract The current status of the analysis of pesticides in wastewater by chromatographic techniques and toxicity bioassays is reviewed and evaluated. When using chromatographic techniques, the low concentrations of pesticides present and the complexity of the wastewater matrices require a sample concentration step prior to measurement. Also, cleanup techniques need to be applied for better detection of the analytes and to avoid ion suppression. The most commonly used methods of analysis for the detection of pesticides in wastewater samples involve GC–MS and LC–MS. However, an evaluation only based on chemical analysis may be insufficient without information related to the negative effects generated. Bioassays play an important role in the detection and screening of the toxic effects of pesticides in complex samples such as wastewaters. They provide a response that relates to the overall effects (synergism, antagonism) of the chemicals present in wastewaters and they assess the short- (acute) and long-term (chronic) effects. Therefore, both chemical and biological analytical strategies are relevant to the correct evaluation of pesticides in wastewaters, their behavior during wastewater treatment, and the reuse of water resources. Keywords Pesticides · Wastewater · GC–MS · LC–MS · Toxicity bioassays

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1 Introduction Pesticides in wastewaters come typically from point sources of contamination such as disposal sites and landfills where industrial or agricultural wastes are buried without any consideration, as well as discharges from industrial effluents from pesticide production plants. Furthermore, nonpoint sources derived from regular agricultural activities, especially in intensive agricultural areas, and accidental spills can also be significant. Urban use of pesticides is also possible in large cities where the use of herbicides and insecticides may result in runoff into the sewers. These sewers in turn may expel pesticides into wastewater treatment plants (WWTPs). Due to the partial to complete resistance of many pesticides to biodegradation during the wastewater treatment processes, these compounds can escape elimination in WWTPs and enter into the aquatic environment. As a consequence, their evaluation represents an important objective in the efficiency of WWTPs and water quality. Until the beginning of the 1990s, halogenated, nonpolar pesticides were the focus of interest and a part of intensive water monitoring programs in many developed countries, and subsequently a drastic reduction of emission was achieved after adoption of appropriate measures [1]. Today, in industrialized countries we can consider the presence of these compounds as having less importance. But they are used as effective pesticides (e.g., lindane, DDT) and still represent a big issue in developing countries in terms of the environment and human pollution [2–4]. Awareness of the presence of nonpolar pesticides in wastewaters is achieved mainly through the use of gas chromatography. Conversely, a “new”generation of pesticides considered as “emerging contaminants” with a wide range of structures and typically with high polarity has only been recognized for the last few years. As a consequence of this, high polarity and sometimes thermally labile LC-based methods are generally more suitable for their analysis. Therefore, the interest in evaluating these compounds in wastewater clearly remains, as is shown by the inclusion of an important number of pesticides in the list of 33 priority substances issued in the last EU Water Framework Directive [5]. In general, we can consider the analytical methods for pesticides well documented and evaluated as a consequence of the important routine monitoring programs for food and drinking water [6–13]. Nevertheless, the complex nature of wastewaters is a great limitation to chemical analyses in their ability to totally evaluate pesticide content in the low microgram per liter range (or even below that).A second point of interest is the large number of pesticides, around 800, on the market with a very wide range of structures and physicochemical properties, which makes it very difficult to develop adequate target multiresidue methods that cover enough of them, even without taking into consideration the formation of possible transformation products. Consequently, there is a lack of knowledge concerning this kind of pollution and a need to apply sophisticated and powerful analytical techniques to per-

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form adequate identification and quantification of pesticides in wastewaters. In such situations, chemical analysis based on the concentrations of a limited number of compounds has serious limitations, even if the treated effluents meet the threshold concentration levels for discharge. The major one is the inability to account for the contribution of the negative effects of the target pollutants in the mixture, and to make feasible an effective evaluation treatment process to allow the reuse of wastewater. Furthermore, wastewaters are not polluted by a single chemical, but rather by a mixture of numerous chemicals, and this fact can be the main reason for the toxic impacts of wastewater samples. The mixtures of pesticides and other pollutants may cause toxicity even if each individual chemical is below its threshold concentration because of interactive effects among them. It means that the combined effect of various chemicals can be the result of additive effects of individual chemicals, or they can even produce a greater toxic effect showing synergism [14, 15]. In the light of these limitations, effective additional tools able to assess the biological responses of the pesticides present, as well as their interaction with the other chemicals, have to be introduced to complete the evaluation of wastewaters. Bioassays on water samples provide a direct functional response that can relate to the negative effects of a single pesticide and overall toxic properties of the complex mixture of compounds present in a sample [16]. This study is an overview focused on the application of the main analytical strategies based on chemical analysis and biological toxicity assays for pesticides, to be used as a combined approach for the evaluation of pesticides in wastewaters.

2 Chemical Analysis 2.1 Sample Treatment Due to the predicted and previously detected low concentrations of pesticides in environmental samples (usually around the nanogram per liter level), a preconcentration step of the water samples is necessary prior to measurement. In this way, a preconcentration factor of several orders of magnitude (200–1,000fold) is mandatory to reach the low detection limits necessary for the identification of pesticides, especially in complex wastewater samples.Also, the use of surrogate standards (e.g., triphenyl phosphate) added before the extraction step is a common practice in order to account for possible errors during the extraction process and for quantitative purposes. The commonly used extraction methods for polar compounds from water matrices involve isolation using liquid–liquid extraction (LLE) and solid-phase extraction (SPE), which are commented on below. Other methods such as semipermeable membrane devices (SPMD) are also mentioned.

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2.1.1 Liquid–Liquid Extraction LLE has been used in the past for the extraction of pesticides from environmental water samples [17]. However, its application in the extraction of wastewater samples is scarce due to the low efficiency of extraction, especially for polar analytes. Because of the vast amount of surfactants and natural products present in wastewater samples, emulsions are formed which complicate the process of extraction and lead to low extraction recoveries. However, there have been some useful applications of LLE to wastewater analyses. For example, LLE was found to be effective for the isolation of herbicide and pesticide organic compounds from industrial wastewater samples and also from complex matrices [18]. 2.1.2 Solid-Phase Extraction SPE procedures are used not only to extract traces of organic compounds from environmental samples, but also to remove the interfering components of the complex matrices in order to obtain a cleaner extract containing the analytes of interest. In this sense, it is a good sample treatment method for the analysis of wastewater. In the last few years, there has been a considerable interest in developing new selective and sensitive methods for extracting and isolating components from complex environmental matrices. The selectivity is the degree to which an extraction technique can separate the analyte from interferences in the original sample.Accordingly, the selectivity of stationary phases is an important parameter to be taken into account when compounds are to be extracted from wastewater samples, since the main objective is to remove interferences and facilitate further analysis by conventional analytical methodologies such as gas chromatography (GC) or liquid chromatography (LC). SPE using C18 or polymeric phases has been used widely for the determination of pesticides in water samples [19, 20]. These stationary phases are generally nonselective and can lead to difficulties with interferences coextracted from the wastewater matrices. Most of the polar pesticides cannot be determined owing to their coelution with the matrix peak, which is obtained at the beginning of the chromatogram when wastewater samples are analyzed by chromatographic techniques [21]. This matrix peak is a coeluting interferent caused by humic substances present in natural waters. The chromatographic methodologies used are commonly not selective toward the coextracted compounds present in environmental samples, and consequently it is of primary importance to use a selective sorbent for the preceding step (SPE) in the entire analysis. The main goal of the SPE step is to provide a cleaner extract, free of matrix interferences. This is the first step in the development of a highly selective and sensitive methodology that can be applied to the determination of traces of organic contaminants in complex environmental samples. In other

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words, the more selective the SPE step, the better the sensitivity achieved. For these reasons, efforts have been made to develop new selective sorbent materials for the analysis of wastewater samples. Modified silica with a C18 reversed-phase sorbent has historically been the most popular packing material, owing to its greater capacity compared to other bonded silicas, such as the C8 or CN types [22]. Applications of C18 sorbents include the isolation of hydrophobic species from aqueous solutions. The mechanism of interaction with such sorbents depends on van der Waals forces, and secondary interactions such as hydrogen bonding and dipole–dipole interactions. Nevertheless, the main drawbacks of such sorbents are their limited breakthrough volumes for polar analytes, and their narrow pH stability range. For these reasons, reversed-phase polymeric sorbents are also used frequently in environmental applications for the trace enrichment of soluble molecules that are not isolated by reversed-phase sorbents such as C18. The most widely used polymeric sorbents are the styrene–divinylbenzene copolymers (SDB), which are among the classical reversed-phase sorbents introduced in the 1960s [20]. They are currently produced in purified form and are useful for the isolation of more polar solutes that have low capacities on the C18 reversed-phase sorbents. Their broader pH-stability range increases the flexibility of the method since the pH of the wastewater samples is usually in the high range. Moreover, these kinds of sorbents have a greater surface area per gram, so they can retain the most water-soluble analytes. Another advantage of the aromatic sorbents derives from their selective interaction with aromatic rings in the analytes. Because the styrene–divinylbenzene structures contain aromatic rings, they have the ability to sorb analytes by specific p–p interactions. More recently, many immunosorbents based on antigen–antibody interactions have been developed for the selective isolation of many pesticides in water samples [23]. They have proven to be very suitable for the highly selective preconcentration of organic contaminants from complex environmental samples, such as sediments and sludges. Since such sorbents are tailor-made for specific applications, their cost is high compared to conventional sorbents [24]. However, they are very limited for multiresidue applications and therefore only useful in wastewater analysis for those cases when a conventional sorbent is not suitable. On the other hand, molecularly imprinted polymers have been developed as well and are gaining applicability in some environmental areas, and could be a promisingly useful tool for the trace enrichment of organic contaminants in complex mixtures in forthcoming years [25]. 2.1.3 Semipermeable Membrane Devices and Other Membrane Processes SPMD have gained widespread use for sampling hydrophobic chemicals from water. In these membranes the more hydrophobic compounds are retained and are further recovered with organic solvents. As an example, SPMD have been applied to the analysis of pesticides in wastewater samples [26].

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Other membrane processes such as microfiltration, ultrafiltration, reverse osmosis, and colloid-enhanced ultrafiltration have been applied to the separation of beta-cypermethrin from wastewater samples [27]. In this study, a separation of above 92% was performed by reverse osmosis by the use of composite membranes and above 80% by colloid-enhanced ultrafiltration by the use of nonionic surfactants. 2.2 Cleanup Techniques In environmental analysis, organic compounds are usually present at low concentrations and are often masked by complex patterns of interfering components. Therefore, cleanup steps are necessary for the analysis of wastewater samples, especially before analysis by gas or liquid chromatographic techniques. Florisil and silica phases are the most commonly used cleanup methods for removing organic acids and humic substances from sample extracts when analyzing hydrophobic compounds such as organochlorine pesticides by gas chromatographic techniques. SPE can be easily applied as a cleanup method for this kind of matrix as well. Accordingly, sequential SPE has been applied as a preconcentration and cleanup method in the analysis of some pesticides in wastewater samples before analyses by liquid chromatographic techniques [28]. Other cleanup methods involve the rapid and effective anion-exchange capacity of the anion-exchange phases to remove humic substances, which are present in complex water and soil samples [29]. The high selectivity of the SAX disk for humic substances allows these interferents to be effectively removed from water samples during trace enrichment of herbicides from complex water extracts. The concept of layering disks can be used to first remove the humic impurities on the SAX disk with simultaneous isolation of herbicides on the lower C18 disk. The concept of stacking adsorbents for trace enrichment was first introduced in the early 1980s with XAD adsorbents. Both anion-exchange and reversed-phase methods can then be used to isolate both natural and contaminant organic compounds from water. More recently, SPE cartridges have been introduced with layered adsorbents, which facilitate treatment of the aqueous samples. Total or partial ion suppression is a well-known LC–MS effect, which is induced by coeluting matrix components that can have a dramatic effect on the intensity of the analyte signal.As can be observed in Fig. 1, analyte suppression occurs as a consequence of the different matrix interferences present in wastewater samples, making the identification and/or quantification process difficult or unfeasible. Even when working under selection ion monitoring (SIM) conditions, these matrix effects can cause ion suppression in the detection of some analytes that are present at low levels of concentration, as seen in this figure. Several papers have reported this effect [30–32] and different alternatives to overcome these problems, such as the inclusion of a size-exclusion step [33] or sequential SPE [28], have been applied for the determination of pesticides in

Fig. 1 SPE–LC–ESI-MS analysis (SIM mode) of two wastewater samples, spiked at different levels of concentration. Compounds: (1) ciromazine, (2) oxamil, (3) metomil, (4) carbendazime, (5) thiabendazole, (6) imidacloprid, (7) acetamiprid, (8) thiacloprid

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environmental matrices. The advantage of the fractionation of the extract is the separation of the pollutant mix into different group classes, depending on their chemical properties, which reduces matrix interferences and makes detection easier and more reliable. Solid-phase microextraction (SPME) is also a useful alternative to conventional sample cleanup with LLE or SPE. SPME is based on the enrichment of analytes by a partitioning process between a polymeric phase coated on a fused-silica fiber and its surrounding aqueous solution. SPME combines sample preparation in terms of extraction from a matrix of interfering compounds with an enrichment process in a single step. A method for the determination of metazachlor in wastewater samples is described in the literature [34]. In this study, SPME was shown to be a suitable and simple sample preparation method for the determination of metazachlor in wastewater by GC–AED. 2.3 Methods of Analysis A wide range of analytical techniques have been developed in order to identify the organic contaminants often present at trace levels in complex environmental samples such as wastewaters. These techniques mainly use gas chromatography (GC) and liquid chromatography (LC). Most of the continuously monitored water contaminants are determined via gas chromatography–mass spectrometry (GC–MS). However, an adequate separation of polar compounds via GC typically requires derivatization of the polar moieties (e.g., BSTFA derivatives). In addition to this, as the analyte groups show different properties concerning the number and kind of functional groups, it is quite difficult to develop a universal derivatization procedure suitable for all the target analytes. Furthermore, the presence in wastewater of many other organic compounds requires the use of labeled standards, which can make application of this method unfeasible [35]. 2.3.1 Gas Chromatography GC is coupled with many detectors for the analysis of pesticides in wastewater. At the present time the most popular is GC–MS, which will be discussed in more detail later in this section. The flame ionization detector (FID) is another nonselective detector that identifies compounds containing carbon but does not give specific information on chemical structure (but is often used for quantification because of the linear response and sensitivity). Other detectors are specific and only detect certain species or groups of pesticides. They include electron capture, nitrogen–phosphorus, thermionic specific, and flame photometric detectors. The electron capture detector (ECD) is very sensitive to chlorinated organic pesticides, such as the organochlorine compounds (OCs, DDT, dieldrin, etc.). It has a long history of use in many environmental methods,

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especially those advocated by the U.S. Environmental Protection Agency. The next most common is the nitrogen–phosphorus detector (called the NP detector), which is effective for many herbicides and insecticides because they contain either nitrogen (triazines and acetanilides) or phosphorus (OP insecticides). This is an inexpensive detector that has been widely used in the analysis of pesticides in water. The flame photometric detector (FPD) works on chemiluminescence and detects pesticides that contain sulfur, phosphorus, and some metals, such as manganese [36]. The thermionic specific detector (TSD) and FPD work on pesticides that also contain sulfur, nitrogen, and phosphorus [37]. The combination of these groups of detectors has value for specific compound identification in complex matrices, where GC–MS may have serious interferences. Wastewater is such an example. A specific detector for measuring elemental compositions as a percentage is the atomic emission detector (AED), which can detect all elements, except helium, separately due to its multichannel ability and selectivity, making it more sensitive than the more commonly used detectors cited above. This type of detector has been used for the detection of pesticides in wastewater samples [34]. Finally, gas chromatography coupled to mass spectrometry (GC–MS) is the most universal technique for the analysis of pesticides in water samples [38]. The high sensitivity and selectivity of modern GC–MS instruments enables low limits of detection depending on the matrix and in particular on the chemical structure of the pesticide. With most instruments, full-scan spectra can be evaluated at the low nanogram level, which means 1 or 10 pg analyte injected into the GC–MS system with the sample. Spectral averaging and background subtraction facilities provided by the data system are generally used to remove contributions from the matrix background or partially resolved contaminants. However, with very weak spectra, these data processing procedures may lead to corrected mass spectra of dubious validity. Changing from full spectral scanning to selected ion monitoring using the reduced number of mass channels leads to considerably improved detection limits for the specified target compound ions. The different types of ionization include electron impact (EI) and chemical ionization (CI). One advantage of negative chemical ionization (NCI) is in the analysis of organochlorine insecticides in complex matrices, because the background does not ionize and the pesticides are easily detected (see Fig. 2). The ion suppression and matrix interference effect is clearly shown in this figure when analyzing wastewater samples in the EI mode, even under SIM conditions. As an example, Fig. 3 shows the analysis of a real wastewater sample in the NCI mode where three compounds were identified by the corresponding mass spectra. Another approach in GC is that of using more power in the separation by doing GC¥GC. In this approach, a second column is used with a different type of stationary phase than the primary stationary phase, and fast chromatography using TOF-MS as the detector is carried out [39]. This technique uses only TOF-MS as the detector since it has the most sensitivity for fast-eluting peaks. The method has been applied to complicated matrix analysis.

Fig. 2 GC–MS chromatograms of a spiked wastewater extract with triclosan, endosulfan, and oxifluorfen obtained under EI (full scan) and EI (SIM) conditions at a concentration level of 625 mg L–1 and NCI (full scan) conditions at 250 mg L–1

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Fig. 3 SPE–GC–NCI-MS chromatogram obtained from a real wastewater sample (influent) where triclosan and endosulfan-a, -b and sulfate were detected at 4.9 mg L–1, 160 ng L–1, 128 ng L–1, and 15 ng L–1, respectively, under SIM conditions

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One of the known disadvantages of the use of GC is the need for previous derivatization of some of the most polar pesticides before analysis can be carried out [40]. These derivatization steps might produce low-efficiency results in complex wastewater matrices, which make the analysis rather difficult and cumbersome. However, the reproducibility in retention times when using GC techniques is so precise, that specific identifications of pesticides can be made even in complex environmental samples. Quantification is usually achieved by a standard addition method, use of labeled internal standards, and/or external calibration curves. In order to allow for matrix interferences the most reliable method for a correct quantitation of the analytes is the isotope dilution method, which takes into account intrinsic matrix responses, using a deuterated internal standard or carbon-13-labeled internal standard with the same chemistry as the pesticide being analyzed (i.e., d-5 atrazine for atrazine analysis). Quality analytical parameters are usually achieved by participation in interlaboratory exercises and/or the analysis of certified reference materials [21]. 2.3.2 Liquid Chromatography–Mass Spectrometry (LC–MS) Due to the high amount of interferences present in wastewater samples, UV detection is not possible for the identification of pesticides at low levels of concentration (see Fig. 4). As can be noted in this figure, the humic and fulvic acid peak at the beginning of the chromatogram masks the identification of the most polar pesticides and complicates the identification and quantitation of the analytes. Furthermore, a higher level of confidence (molecular or fragment structural information) is necessary for the correct identification of analytes in such complex matrices. In this sense, liquid chromatography coupled to mass spectrometric detection is the best choice for the analysis of pesticides in wastewater samples. Since polar, nonvolatile, thermally unstable or highmolecular-weight compounds are unsuitable for gas chromatography–mass spectrometry (GC–MS) analysis, the use of LC–MS has become a robust and routinely applicable tool in environmental laboratories [41, 42]. Non-GCamenable compounds include 15–20% of the present-day pesticides, e.g., phenylureas and carbamates, the phenoxyalkanoic acids, and a large majority of all pesticide transformation products [43]. The performance of LC–MS in the analysis of polar and thermally labile pesticides that are not amenable to GC–MS has been well demonstrated in several studies [44, 45]. In the last few years, interfaces based on atmospheric pressure ionization (API) have resulted in an increase in the number of applications in environmental determinations. In this respect, atmospheric pressure chemical ionization (APCI) and electrospray ionization (ESI) have recently become the most universal techniques for environmental analysis due to their high sensitivity, the possibility of detecting a broad range of analytes, and the useful structural information obtained via fragmentation similar to collision-induced dissociation (CID) [46]. Compared

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Fig. 4 SPE–LC-DAD analysis of a wastewater sample. Peak identification number and peak retention times (min): (1) azinphosmethy, (11) parathion-methyl, (4) malathion, (3) fenitrothion, (8) azinphos-ethyl, (6) chlorphenvinphos, (10) parathion-ethyl, (7) diazinon [from ref. 21]

with older mass spectrometric detection techniques such as TSP and PB, API techniques offer both structural confirmation and high sensitivity for target compounds in environmental samples. One of the great advantages of the ESI interface is its high sensitivity for ionic pesticides such as many herbicide metabolites containing a sulfonic or a carboxylic group in the chemical structure [47]. The advent of high-performance liquid chromatography–mass spectrometry (HPLC–MS) using quadrupole instruments has made analysis of polar pesticides in water a common procedure [45, 48]. Many classes of pesticides are easily analyzed by LC–MS and a more challenging task is to identify the degradation products of pesticides. During the past 5 years many papers have been published on the analysis of pesticides and their degradation products by HPLC–quadrupole MS [49]; however, there are several shortcomings yet to be overcome. For example, often polar pesticides give only a protonated or deprotonated molecule or a weak fragment ion, especially when the interface is ESI. The fragmentor or cone voltage is used to enhance CID in the source and transport regions of the electrospray source, and this fragmentation voltage may

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vary substantially among different analytes and sources, which makes fragmentation difficult to predict in an analysis of unknown compounds. Second, there are no universal libraries available for pesticide analysis by HPLC–MS, as in electron impact GC–MS; this problem makes identification of unknown pesticides or their degradates nearly impossible by simple HPLC–quadrupole MS analysis. These shortcomings may be overcome partially by the application of timeof-flight mass spectrometry (TOF-MS) and liquid chromatography–quadrupole ion-trap tandem mass spectrometry (LC–QIT-MS/MS) [50–52]. The LC–QITMS/MS does MS/MS in time rather than in space, which means that ions are retained in a trap through a set time period. If all the ions are ejected, then the result is a full-scan spectrum. If the protonated or deprotonated molecule is retained in the trap and all others are ejected, and this ion is fragmented, the result is MS/MS. This process may be repeated multiple times, which results in MSn. In contrast, triple quadrupole MS/MS does the isolation and fragmentation in space, which means that the fragmentation is continuous in time, but the selected ion travels through the flight tube of the mass spectrometer to the collision chamber where fragmentation occurs, and then on to the third quadrupole for the mass spectrum. Two advantages of the ion trap are that it gives excellent sensitivity while trapping ions in full-scan mode, which then may be selected and fragmented to yield MS/MS spectra, and second is the ability of the ion trap to do MSn [50]. Typically, three or four isolations and fragmentations are possible before the sensitivity is too low to record ions in unknown samples. The ability to do multiple isolation and fragmentation allows one to build a library of spectra using standard compounds, which give both characteristic fragmentations and diagnostic ions that can then be used to identify unknown pesticides or their degradates. TOF-MS is also useful for identification of synthesized standards to verify the analysis of QIT-MS/MS when no commercial standards are available and new standards are synthesized, as well as the identification of degradates in actual groundwater samples [52].

3 Toxicity Biological Assays 3.1 Bioassays Applied to Evaluate the Toxicity of Pesticides The toxic effects of pesticides can be diverse and depend on the sensitivity of organisms to these toxicants, and the pesticide concentration or bioavailability. Typically, the short- and long-term effects of pesticides have been evaluated through acute or chronic toxicity bioassays, respectively, using lethality endpoints and sublethal endpoints (e.g., growth and reproduction), particularly these last in chronic bioassays.

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3.1.1 Acute Toxicity Bioassays Most commonly, bioassays for the evaluation of the acute toxic effects of pesticides are based on single aquatic species selected to be representative of a range of taxonomic and functional groups, i.e., bacteria, algae, invertebrates or fish [53, 54]. Generally, toxicity evaluation using a single species is the alternative of choice rather than the use of multiple species, because extrapolation of effects to an ecosystem is more difficult and can often lead to incorrect conclusions. The selection of suitable single species and protocols is not a trivial task and may be dependent on various factors. Some of these include simplicity, low cost, or modest material and equipment demand. However, a higher sensitivity than other species to toxicants may be decisive in this choice in order to serve as warning systems. Table 1 shows the sensitivity in terms of effective concentration (EC50), which is the toxicity endpoint for the organisms (bacteria, crustaceans, algae, and fish) selected for the toxicity bioassays. These toxicity bioassays are usually classified according to the test species involved. Fish assays have been extensively used for laboratory studies. Among commonly used species are Pimphales promelas or Oncorhynchus mykiss. These species are relatively sensitive and respond to a variety of water constituents and contaminants including pesticides. P. promelas is a widely distributed species in the aquatic environment, and its use for whole effluent toxicity (WET) procedures is also well established [55, 56]. Reported lethal concentrations for pesticides such as chlorotalonil or chlorpyriphos (EC50=22.6 and 381 mg/l, respectively) showed these compounds as “harmful to aquatic organisms” and “not harmful”, respectively, according to toxicity categories [56, 57]. Generally, in addition to the relative sensitivity (Table 1), the use of these bioassays presents some disadvantages such as standardization problems, time consumption or need of specialized equipment [58–60]. Invertebrate species have been widely used in toxicity studies of pesticides [61]. Zooplankton play a key role in the food chain because they occupy a central position. Therefore, their responses to natural and anthropogenic stresses are intimately linked with other food predator organisms. The most widely accepted bioassays employ species such as Ceriodaphnia dubia, Daphnia magna, Artemia salina, or Thamnocephalus platyurus [62–64]. D. magna has been used for many years as a standard aquatic test species and formally endorsed by the major international organizations such as the EEC, OECD, and ASTM [65–67]. Its choice is mainly because it represents the zooplankton community and is a species of worldwide occurrence. In addition, it has a greater sensitivity to toxicants, particularly pesticides, compared with other aquatic species [61, 68] (Table 1). Algae are of vital importance in the primary production of the aquatic ecosystem because they are primary producers of the food chain. Several species of green algae are used in toxicity studies of pesticides, especially herbicides such as Chlorella vulgaris, Chlorella pyrenoidosa, or the standard test microalga

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Table 1 Effective concentration (EC50) values of pesticides for bacteria, algae, crustaceans, and fish

Pesticides

EC50 (mg l–1) Bacteria

Fenamiphos Benomil Pentachlorlophenol Paraquat Deltametrin Dichlorvos Chlorpyriphos Metalaxil Carbendazim Procymidone Zineb Chlorothalonil b-Cypermethrinm Dichlofluanid Permethrin Carbofuran Diuron Isoproturon Atrazine Formetanate Pirimiphos-methyl Malathion Cyromazine

Algae

35.1a

References Crustaceans

Fish

0.005 0.05 b

0.55 14,800 1

<1.0 2, b 0.005g 2.10–4,e 3.21,e

0.08 1, a

21.1b 34.6 b 0.74 b 0.52 b 0.007 1, c

0.03 2, f

0.132,c

1.03,f

3812, h

22.6 3, h 0.02 g 0.2 g

31.2

1, a

0.022,f 0.041, c <1.0 d <1.0 d

7.41, a

8.62,f 0.07 2, f 4.10–4, f 1.8·10–3, f 10.7 f

[15] [71] [77] [77]1, [68] 2 [58] [63] [68]1, [57]2 [71] [71] [71] [71] [14]1, 2, [56] 3 [59] [14]1, 2, 3 [60] [15]1, 2 [14]1, 2 [68] [68] [15]1, 2 [14] [63] [68]

a

Vibrio fischeri (EC50 at 15 min). Chlorella pyrenoidosa (EC50 at 96 h). c Selenastrum capricornutum (EC at 72 h). 50 d Chlorella vulgaris (EC at 96 h). 50 e Artemia salina (EC at 48 h). 50 f Daphnia magna (EC at 48 h). 50 g Poecilia reticulata (EC at 48 h). 50 h Oncorhynchus mykiss (EC at 96 h). 50 b

Selenastrum capricornutum [14, 68, 69]. Herbicides play an important role in agricultural practices and as a consequence, they can affect nontarget organisms, modifying the structure and function of aquatic communities due to the alterations of the specie composition species in algal communities. The effects of new herbicides in agricultural activities have been recently published [68]. Paraquat, diuron, isoproturon or atrazine (see Table 1) are examples of herbicides considered as very toxic according to toxicity categories established in the Directive 93/67/EEC [57, 68, 70, 71], with toxicity endpoint values expressed as effective concentration (EC50) less than 1 mg/l.

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Bioassays using bacteria as indicators constitute at present one of the most widely applied screening techniques for identifying the toxicity of substances and commercial products [72]. The main reason is that, unlike toxicity bioassays using more complex organisms such as fish, bacterial bioassays are much quicker and cheaper.A great number of bioassays based on very different methods and microorganisms have been described, and include studies of the effects of toxicants on different parameters such as growth inhibition and enzymatic activity [73, 74]. Among the bacteria employed, such as Pseudomonas putida or Escherichia coli as indicators, V. fischeri is the most common standardized bacteria specie used in toxicity bioassays [14, 72, 75]. The advantages are its sensitivity, reproducibility, and it is a rapid and simple test (Table 1). For toxicity evaluation of pesticides, there are reported data showing the sensitivity and the utility of this test [14, 75–77]. The different sensitivity of the species indicates that a single bioassay does not satisfy the correct evaluation of the wastewater. Thus, normally, various species are used because toxic substances may produce a specific response in one species but not in another. Therefore, there is practically generalized consensus on the use of a battery of bioassays involving different trophic levels of species. The application of this approach is considered an efficient and essential tool for predicting environmental hazards to the aquatic ecosystems. According to several authors, the most appropriate way to assess ecotoxicity is the use of four different test organisms of increasing levels of biological organization. This system includes the use of bacteria, crustaceans, algae, and fish to assess the toxicity of chemicals such as pesticides in wastewater, and it would be performed sequentially, going to the next level when the sample was found to be nontoxic [14, 76, 78]. However, a general perception is that, for practical and ethical reasons, the use of fish is not frequently included in these studies. 3.1.2 Chronic Toxicity Bioassays Episodic pollution events can adequately be addressed by acute toxicity bioassays, however these are not sufficient to investigate the water quality for delayed toxicity effects of chemicals present. Chronic effects of pesticides can include carcinogenicity, teratogenicity, mutagenicity, neurotoxicity, and reproductive effects (endocrine disruption). Most insecticides, especially the organophosphate group, cause neurotoxicity as their major mode of action. Assessment of the neurotoxicity includes neurochemical endpoints such as cholinesterase (including acetylcholinesterase, which is the major neurotransmitter in vertebrates such as fish, and other enzymes such as butyrylcholinesterase) inhibition and behavioral endpoints such as swimming speed [79]. Studies done in rats show the neurotoxic action of insecticides such as dimethoate, methyl parathion, dichlorvos, ethyl parathion or propoxur after a prolonged exposure [80, 81].

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Chemical carcinogenicity has been the target of a large list of scientific publications, because it is one of the toxicological endpoints that poses the highest concern. The standard bioassays in rodents used to assess the carcinogenic potential of chemicals are extremely long and costly and require the sacrifice of a large number of animals. For these reasons, mutagenicity bioassays are presented as alternatives to evaluate the DNA-damaging activity [82, 83]. The types of genetic lesions expected can be chromosomal deletion, loss or translocation, mitotic recombination or base substitution [82]. Therefore, regular practices to evaluate the possible genetic lesions recommend the use of a battery of bioassays including a bacterial test for gene mutation, either an in vitro test for chromosomal aberrations or a mammalian cell mutagenesis test, and a general test for DNA damage [84, 85]. A great number of studies on the mutagenic activity of pesticides have been published. Examples of these show that the chloroacetanilides, classified as herbicides, have a consistent positive induction for gene mutations [86]. More recently, toxicity studies have shown the importance of noncancer endpoints in chronic toxicity assessment, with increasing emphasis on endpoints such as endocrine disruption. The endocrine system as a target of pesticide toxicity can manifest reproductive consequences, particularly in terms of steroid hormone function, resulting in the manifestation of demasculinization in fish. The gonad histology and serum vitellogenin (VTG) protein levels have been widely used as endpoints for screening and testing of potential endocrine-active compounds and are currently subject to validation by the OECD and associated scientific groups [87–89]. Some reports have demonstrated that the presence of organochlorines, such as dieldrin, heptachlor or aldrin, appears to be closely linked to the induction of VTG synthesis [90, 91]. However, bioassays based on yeast strains are very promising among the test systems available because of their physiological simplicity, easy handling, and low costs [92, 93]. In general, they rely on yeast constructs expressing an estrogen receptor which, upon binding of suitable substrates, acts as a transcriptional enhancer for an estrogen-responsive DNA-element-controlled reporter gene, in most cases bacterial b-galactosidase. The activity of this enzyme can be determined photometrically by using a chromogenic substrate and thus may serve as a measure of the estrogenic potency of the samples under investigation. Several active components such herbicides and insecticides (e.g., endosulfan, dieldrin or toxaphene) have been reported to possess estrogenic activity [94, 95]. 3.2 Toxicity Studies of Wastewater Containing Pesticides While reported data on the acute and chronic toxicity of many pesticides is plentiful, few studies have been published on toxicity bioassays applied to wastewaters containing pesticides. The application of toxicity bioassays to the quality control of wastewaters offers several advantages in addition to being a

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biological measure able to detect toxic effects. Among these, sensitivity, easy handling, speed, simplicity, and low costs are the features of choice for routine purposes. In the last few years, interest in toxicity bioassays for assessing wastewater has been increasing and recent publications are focused on this approach [96–98]. Some of these studies proved that there is no correlation between chemical and ecotoxicological parameters. Control based on global chemical parameters such as biochemical oxygen demand (BOD), chemical oxygen demand (COD) or total organic carbon (TOC) may be insufficient, even if the treated effluents meet the threshold concentration levels for discharge. This case is illustrated in Fig. 5, which shows a monitoring study performed on influent and effluent wastewaters. Samples corresponding to toxic effluents showed permissible TOC levels [96]. Wastewater from agricultural areas that arrives at wastewater treatment plants (WWTPs) is highly variable in nature. Intermittent or accidental episodes of toxic substances can have a damaging effect on the receiving waters, when the

a

b Fig. 5a, b Monitoring study of wastewaters based on chemical and ecotoxicological parameters [from ref. 96]

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influent has not been properly treated. Consequently, rapid methods of wastewater toxicity assessment represent a very useful tool, acting as an early warning system. The capability of detecting toxic responses in a short time allows quick decisions to be made regarding the convenience of effluent discharge. In others words, the capability of detecting toxic effects is one of the best applications of bioassays in the quality control of wastewaters, because it allows detecion of unwanted toxicity, potential problems in the treatment station, and contamination peaks in effluent toxicity before discharging it to receiving waters [76]. The sensitivity of test species is a decisive feature in the choice of bioassays to evaluate the toxicity. Despite the diversity of test species available, in many regulatory schemes the invertebrate species recommended for acute and chronic testing is the cladoceran D. magna [99, 100]. In the U.S., the Food and Drug Administration and Office of Pollution Prevention and Toxics (OPPT) of the Environmental Protection Agency recommend that acute data should be collected with Daphnia species (D. pulex and D. magna) [101]. Presumably, the focus on D. magna results from its high sensitivity to environmental contaminants relative to other species, mainly invertebrate species. The sensitivity of D. magna to pesticides has been demonstrated in recent publications showing its capability of detecting toxic responses at concentration levels as low as nanograms per liter [14]. This means that toxicants at environmentally realistic concentrations can be detected by this bioassay. However, the detection limit of standardized bioassays may be too high to detect toxicity and hence pesticide contamination. Therefore, in these cases, preconcentration of the samples is necessary. Bioassays combined with preconcentration of the wastewater have been proved to be a useful strategy for screening and monitoring in the initial assessment of water pollution by pesticides [102, 103]. Even if bioassays are able to detect toxicity in nonconcentrated samples, this strategy is a useful approach in order to obtain the toxicity endpoint (e.g., effective concentration EC50) from a full concentration–response relationship. This combined methodology was applied in a screening study from an agricultural area where methyl parathion, lambda-cyhalothrin, and endosulfan are the most commonly used pesticide chemicals. Acute toxicity was detected in surface water from agricultural areas using standardized bioassays with the algae S. capricornutum and crustacean D. magna [104]. Whole effluent toxicity (WET) monitoring offers several advantages because this toxicity evaluation has to account for the presence of unknown toxicants, the interactions among multiple toxicants, and the alterations in toxicant bioavailability caused by the effluent matrix. When evaluating the toxicity of the complex samples, the detection and identification of regulated or specific chemicals is a key need for controlling effluent quality. Thus, the identification of toxic compounds in complex samples has been the objective of reported studies using toxicity-based procedures. Combined protocols involving chemical analysis and toxicity evaluation became known collectively as toxicity identification evaluation (TIE), and nowadays they are techniques well

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established and developed by the USEPA [104]. TIE methods have been found to be effective tools for characterizing and identifying toxicants in samples of effluents, sediments, ambient waters, and other complex mixtures [105, 106]. Regarding the identification of pesticides in wastewaters or ambient waters, few studies have been published. The use of TIE methods allows the detection of toxic surface waters and the identification of herbicides (molinate, mefenacet, symetryn or esprocarb) as major compounds in rivers from agricultural areas [106]. Recent applications including TIE studies were conducted on influent and effluent wastewaters from wastewater treatment plants that received wastewaters from agricultural areas. This approach was used to detect a possible cause–effect relationship between the plant discharge and the receiving water quality [107]. The results of this study showed the detection of lindane and pp¢-DDE in fish, and chemical investigations revealed ammonia and micropollutants as factors of WWTP effluent impact on receiving waters [108]. As was mentioned above, the interaction among multiple chemicals is one of the main reasons for the wastewater toxicity. The application of TIE methods using acute toxicity (D. magna) guided chemical analysis was applied for water quality evaluation of agricultural land runoff and 11 pesticides widely applied were used as target compounds. Pesticides such as dymeron, flutolanil, and mefenacet were detected in concentrations ranging from 6.2 to 29.7 mg/l; however, these concentrations appeared to be too low to have toxic effects because their effective toxic concentrations were from 5 to 10 mg/l. Therefore, it was impossible for the authors to conclude in this study that the observed daphnia toxicity resulted from a single highly toxic substance. The toxicity was attributed to the combined effect of the pesticides [109]. The utility of the bioassays to assess the interactions among pesticides (additive effects, synergism or antagonism) have been demonstrated in different studies [14, 15]. It is especially relevant to consider the combined effect of pollutants because several pesticides and other contaminants can occur in ambient waters from agricultural areas [110, 111]. The global effect can have a greater negative impact than the single pollutants. For a predictive assessment of the aquatic toxicity of pesticide mixtures, two concepts, concentration addition and independent action, are used. Concentration addition is generally regarded as a reasonable expectation for the joint toxicity of acting substances [112]. Following this model, the concentration of each toxicant is expressed as a fraction of its EC50 (toxic unit, TU). In this model, the EC50 of a mixture is the sum of the single TU and equals unity. Therefore, when the sum of TU exceeds unity, the combined effect is more than additive and when it is less than additive, the substances act antagonistically. Synergism is a common interactive effect among pesticide mixtures. Experiments on pesticide mixtures showed a synergistic effect for 60% of the studied cases [14]. Table 2 shows the combined effects evaluated by three different toxicity bioassays. Therefore, it is evident that the consideration of single pesticides alone is not sufficient for determining the environmental impact of wastewaters.

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Table 2 Combined toxicity effects of pesticides evaluated by three toxicity bioassays

Binary mixtures

Bioassays V. fischeri (15 min) 1 T.U.M

D. magna (48 h) S T.U.i2

1 T.U.M

S. capricornutum (72 h) S T.U.i2

1 T.U.M

S T.U.i2

Irgarol 1501– Diruon

26 5.8 Synergistic +

550 183 Synergistic +

0.36 0.01 Synergistic ++

Irgarol 1501– Sea nine 211

7.4 333 Antagonistic ++

550 Additive

5 5 Additive

Irgarol 1501– Chlorothalonil

5.4 0.87 Synergistic +

1100 414 Synergistic

0.271 2 Antagonistic +

Irgarol 1501– Dichlofluanid

9.4 15.6 Additive

132 Additive

0.152 0.046 Synergistic +

Irgarol 1501– TCMTB

333 25.6 Synergistic ++

330 151 Synergistic +

581

160

10 1.08 Synergistic ++

1 , experimental toxicity. T.U.M S T.U.i2, theoretical toxicity. +=factor≥3. ++=factor≥10.

Acknowledgements This work has been supported by the Project CICYT PPQ2001-1805C03-03 from the Ministry of Science and Technology.

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The Handbook of Environmental Chemistry Vol. 5, Part O (2005): 79– 118 DOI 10.1007/b98608 © Springer-Verlag Berlin Heidelberg 2005

Fragrance Materials in Wastewater Treatment Staci L. Simonich (✉) Oregon State University, Department of Environmental and Molecular Toxicology and Department of Chemistry, 1141 Agricultural and Life Sciences, Corvallis, OR 97331-7301, USA [email protected]

1 1.1 1.2 1.3

Introduction to Fragrance Materials . . . . . . . . Use and Disposal . . . . . . . . . . . . . . . . . . . Chemical Structures . . . . . . . . . . . . . . . . . Physical-Chemical Properties and Biodegradability

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81 81 81 84

2 2.1 2.2 2.3 2.4 2.5

Analytical Chemistry of Fragrance Materials Laboratory Quality Control . . . . . . . . . . Standards . . . . . . . . . . . . . . . . . . . . Aqueous Matrices . . . . . . . . . . . . . . . Solid Matrices . . . . . . . . . . . . . . . . . Analysis . . . . . . . . . . . . . . . . . . . .

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86 86 86 87 91 91

3 Sampling Wastewater Treatment Plants for Fragrance Materials . . . . . . . . . 3.1 Selection of Wastewater Treatment Plants . . . . . . . . . . . . . . . . . . . . . 3.2 Wastewater Treatment Plant Sampling . . . . . . . . . . . . . . . . . . . . . . .

92 92 94

Mechanisms of Fragrance Material Removal During Wastewater Treatment . . Biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Sorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Volatilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

95 95 97 97

4 4.1 4.2 4.3

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5 Measurement of Fragrance Materials in Wastewater Treatment . . . . . . . . . 98 5.1 Concentrations in Treatment Plants . . . . . . . . . . . . . . . . . . . . . . . . 98 5.2 Removal During Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 109 6 Predicting Fragrance Material Removal During Wastewater Treatment . . . . . 113 6.1 Framework for Aquatic Risk Assessment . . . . . . . . . . . . . . . . . . . . . . 113 6.2 Simple Treat Model . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 115 7

Conclusions

References

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Abstract In recent years, there has been significant interest in understanding the input of fragrance materials (FMs) to aquatic ecosystems, and this has driven a substantial amount of research on the removal of FMs during wastewater treatment. Because FMs are semivolatile and have a wide range of physical-chemical properties and biodegradabilities, understanding their removal during the treatment process is complex. The mechanisms of FM removal from wastewater include biodegradation, sorption, and/or volatilization.A wide array of analytical methods have been developed to measure FMs in wastewater influent, primary effluent, final effluent, and solids. Wastewater studies have been conducted in the U.S. and Europe. Finally, the efficient removal of FMs during wastewater treatment is not only dependent on the biodegradability and physical-chemical properties of the FM, but is also highly dependent on plant operation and design. Keywords Fragrance materials · Wastewater treatment · Polycyclic musks · Nitromusks Abbreviations ADBI Celestolide (4-acetyl-1,1-dimethyl-6-tert-butylindene) AHDI Phantolide (6-acetyl-1,1,2,3,3,5-hexamethyldihydroindene) AHTN Tonalid (7-acetyl-1,1,3,4,4,6,-hexamethyl-1,2,3,4-tetrahydronaphthalene) ASE Accelerated solvent extraction ATII Traseolide (5-acetyl-1,1,2,6-tetramethyl-3-isopropylindene) BOD Biochemical oxygen demand CAS Chemical Abstracts Service DPMI Cashmeran (6,7,-dihydro-1,1,2,3,3-pentamethyl-4(5H)-indanone) ECD Electron-capture detector FM Fragrance material GC Gas chromatography GC–MS Gas chromatography–mass spectrometry GC–MS/MS Gas chromatography–tandem mass spectrometry GPC Gel-permeation chromatography HHCB Galaxolide (1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-gamma-2benzopyran) HRT Hydraulic retention time Kd Sorption coefficient to activated sludge Kow Octanol–water partition coefficient MA Musk ambrette (1-tert-butyl-2,4-dimethyl-6-methoxy-3,5-dinitrobenzene) MDL Method detection limit MK Musk ketone (3,5-dinitro-2,6-dimethyl-4-tert-butylacetophenone) MM Musk moskene (1,1,3,3,5-pentamethyl-4,6-dinitroindane) MT Musk tibetene (1-tert-butyl-3,4,5-trimethyl-2,6-dinitrobenzene) MX Musk xylene [1-(1,1-dimethylethyl)-3,5-dimethyl-2,4,6-trinitrobenzene] NM Nitromusk NPD Nitrogen–phosphorus detector OTNE 1-(1,2,3,4,5,6,7,8-Octahydro-2,3,8,8-tetramethyl-2-naphthalenyl)ethanone PCM Polycyclic musk SOC Semivolatile organic compound SPE Solid-phase extraction SPME Solid-phase microextraction SRT Solids retention time TSS Total suspended solids

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1 Introduction to Fragrance Materials 1.1 Use and Disposal Over 2,000 distinct chemicals are currently available globally for formulation into fragrances [1]. Although the consumer is most aware of the use of these chemicals in fine fragrances, by far the largest volume of these chemicals is used in laundry detergents, fabric softeners, household cleaning products, and air fresheners [1]. Of these consumer products, fabric softeners and laundry detergents represent the largest volume uses of fragrances and the largest release to the environment through down-the-drain disposal by consumers following product use [1]. Fragrance materials (FMs) are added to consumer products to mask malodors and to deliver consumer-preferred odors [2]. Although these chemicals are used in low concentrations in consumer products, the volume of laundry detergents and fabric softeners sold throughout the globe can result in significant volumes of FMs being released into the environment. Based on a 1995–1996 survey, approximately 90% of these compounds are used globally at less than 10 metric tons per year [1], with less than 1% being used in volumes approaching 4,000 metric tons per year [2]. Because the majority of the FM volume enters the environment through down-the-drain disposal of consumer products, it is important to understand the removal and fate of these chemicals during municipal wastewater treatment. These semivolatile organic compounds (SOCs) may undergo a complex combination of biodegradation, sorption, and/or volatilization during wastewater treatment. In addition, few SOCs have been studied in wastewater treatment because few of the conventional SOCs (such as pesticides and products of incomplete combustion) enter the environment through down-the-drain disposal and wastewater treatment. The objective of this chapter is to review the state of the science in understanding the removal of FMs during municipal wastewater treatment. Others have reviewed the general environmental fate of FMs, in particular the polycyclic musks (PCMs) and the nitromusks (NMs) [3–6]. 1.2 Chemical Structures The chemical structures of the majority of FMs that have been studied in wastewater treatment are given in Figs. 1–3. Figure 1 shows a variety of FM structures that include alcohols, aldehydes, and ketones, including: benzyl acetate (phenylmethyl ester acetic acid), methyl salicylate (2-hydroxy-methyl ester benzoic acid), methyl dihydrojasmonate (3-oxo-2-pentyl-methyl ester cyclopentaneacetic acid), terpineol (4-trimethyl-3-cyclohexene-1-methanol), benzyl salicylate (2-hydroxy-phenylmethyl ester benzoic acid), isobornyl acetate

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Fig. 1 Fragrance materials studied in wastewater treatment [2, 11]

(1,7,7-trimethyl acetate bicyclo[2.2.1]heptan-2-ol), g-methyl ionone [3-methyl4-(2,6,6-trimethyl-2-cyclohexen-1-yl)-3-buten-2-one], p-t-bucinal [4-(1,1dimethylethyl)-a-methyl-benzenepropanal], hexylcinnamaldehyde [2-(phenylmethylene)-octanal], hexyl salicylate (2-hydroxy-hexyl ester benzoic acid), OTNE [1-(1,2,3,4,5,6,7,8-octahydro-2,3,8,8-tetramethyl-2-naphthalenyl)ethanone], and acetyl cedrene [3R-(3a,3ab,7b,8aa))-1-(2,3,4,7,8,8a-hexahydro3,6,8,8-tetramethyl-1H-3a,7-methanoazulen-5-yl)ethan-1-one]. Figure 2 shows the structures of the FMs categorized as nitromusks (nitroaromatic compounds), including: musk ketone or MK (3,5-dinitro-2,6-dimethyl-4-tert-butylacetophenone), musk xylene or MX [1-(1,1-dimethylethyl)-3,5-dimethyl-2,4,6-trinitrobenzene], musk ambrette or MA (1-tert-butyl-2,4-dimethyl-6-methoxy-3,5dinitrobenzene), musk tibetene or MT (1-tert-butyl-3,4,5-trimethyl-2,6-dinitrobenzene), and musk moskene or MM (1,1,3,3,5-pentamethyl-4,6-dinitroindane). Finally, Fig. 3 shows the structures of FMs categorized as polycyclic

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Fig. 2 Nitromusks studied in wastewater treatment

Fig. 3 Polycyclic musks studied in wastewater treatment

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musks (PCMs), including AHTN or Tonalid (7-acetyl-1,1,3,4,4,6,-hexamethyl1,2,3,4-tetrahydronaphthalene), HHCB or Galaxolide (1,3,4,6,7,8-hexahydro4,6,6,7,8,8-hexamethylcyclopenta-gamma-2-benzopyran),ADBI or Celestolide (4-acetyl-1,1-dimethyl-6-tert-butylindene), AHDI or Phantolide (6-acetyl1,1,2,3,3,5-hexamethyldihydroindene), ATII or Traseolide (5-acetyl-1,1,2,6tetramethyl-3-isopropylindene), and DPMI or Cashmeran (6,7,-dihydro1,1,2,3,3-pentamethyl-4(5H)-indanone). The nitromusks (NMs) (shown in Fig. 2) and the polycyclic musks (PCMs) (shown in Fig. 3) have been studied in wastewater treatment by a variety of research groups because they have been detected in the aquatic environment (see for example [7–10]), indicating that they escape wastewater treatment to some degree. The FMs shown in Fig. 1 were studied in wastewater treatment by Simonich et al. [2, 11] because of their relatively large volumes and wide range of physical-chemical properties and biodegradability (see below). It is clear from Figs. 1–3 that FMs have an interesting and wide range of chemical structures. This results in a wide array of perceived odors, including musk, wood, fruit, and flower-like odors. The FMs in Figs. 1–3 will be the focus of this chapter. 1.3 Physical-Chemical Properties and Biodegradability Table 1 lists the CAS numbers, molecular weight, physical-chemical properties (including the log of the octanol–water partition coefficient, sorption coefficient for activated sludge, water solubility, vapor pressure, and Henry’s law constant), and biodegradability of selected FMs [2, 11]. It is clear that the wide range of FM chemical structures results in a wide range of physical-chemical properties (some properties ranging over six orders of magnitude). From the vapor pressures given in Table 1, it is also clear that most FMs can be classified as SOCs (having vapor pressures less than 1 Pa) [2], and have the potential to partition into a variety of environmental compartments once released into the environment. The properties listed in Table 1 are of interest because they govern FM removal during wastewater treatment and their fate in the environment. The log octanol–water partition coefficients (log Kow) for the selected FMs range from 2.1 to 5.9. The more hydrophobic FMs (with high octanol–water partition coefficients) are desirable in consumer products because they tend to be more substantive on fabrics and provide a residual fabric odor. Unfortunately, this hydrophobicity may also result in bioaccumulation in organisms in the environment. The sorption coefficient for activated sludge (Kd) is an important property to consider for removal of FMs due to sorption onto solids during wastewater treatment. The Kd value ranges from 132 to 15,400 L kg–1 for the selected FMs. The water solubility and vapor pressure of the FMs listed in Table 1 also vary greatly: from 0.49 to 1,687 mg L–1 for water solubility and from 0.00003 to 21.9 Pa for vapor pressure. Finally, the Henry’s law constant (a mea-

CAS number

140-11-4 119-36-8 24851-98-7 98-55-5 118-58-1 125-12-2 127-51-5 80-54-6 81-14-1 81-15-2 101-86-0 6259-76-3 54464-57-2 32388-55-9 1506-02-1 1222-05-5

Fragrance material

Benzyl acetate Methyl salicylate Methyl dihydrojasmonate Terpineol Benzyl salicylate Isobornyl acetate g-Methyl ionone p-t-Bucinal Musk ketone Musk xylene Hexylcinnamaldehyde Hexyl salicylate OTNE Acetyl cedrene AHTN HHCB

150.2 152.2 226.3 154.3 228.3 196.3 192.3 204.3 294.3 297.3 216.3 222.3 234.4 246.4 258.4 258.4

2.1 2.6 3.0 3.3 4.3 4.3 4.6 4.2 4.3 4.9 4.9 5.5 5.7 5.6-5.9 5.7 5.9

MW log Kow (g mol–1)

Water solubility (mg L–1)

132 1,265.0 247 1,687.0 408 91.72 595 335.7 2,081 24.59 2,081 23.23 3,030 9.0 1,836 33.0 2,081 1.9 4,412 0.49 4,412 2.75 9,355 6.08 12,020 2.68 12,020 1.28 12,020 (10,040) 1.25 15,400 (12,780) 1.75

Kd (L kg–1)

21.9 0.750 0.00549 4.09 0.000449 10.0 1.30 0.477 0.00004 0.00003 0.027 0.00325 0.203 0.058 0.0608 0.0727

Vapor pressure (Pa) 2.04 0.0607 0.135 0.939 0.00416 84.4 89.4 12.4 0.0061 0.018 5.00 0.118 31.8 14.7 12.5 11.3

Ready Ready Ready Ready Inherent Ready Inherent Ready Not biodegradable Not biodegradable Inherent Ready Not biodegradable Inherent Not biodegradable Not biodegradable

Henry’s law Biodegradability constant (Pa m3 mol–1)

Table 1 Summary of selected FM physical-chemical properties and biodegradability relevant to wastewater treatment [2, 11]

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sure of air–water partitioning) for the selected FMs in Table 1 ranges from 0.00416 to 89.4 Pa m3 mol–1, indicating that some FMs may undergo volatilization during wastewater treatment. FMs also have a wide range of biodegradabilities (see Table 1). Some FMs pass the OECD ready biodegradability test criteria, including the 10-day window [11]. Other FMs pass the OECD inherent biodegradability test or produce CO2 in the OECD ready biodegradability test, but do not meet the 10-day window [11]. Still other FMs do not biodegrade in standard OECD biodegradation tests but undergo biotransformation in more realistic tests [11–13].

2 Analytical Chemistry of Fragrance Materials In order to understand the removal of FMs during wastewater treatment, it is necessary to measure these compounds throughout the wastewater treatment process. Because of the complex nature of wastewater matrices and the low concentration of FMs (0.001–60 mg/L) [11] throughout the treatment plant, accurate and sensitive analytical methods have been developed by a number of researchers. Fortunately, the analytical techniques developed to measure traditional SOCs, such as solvent extraction, extract concentration, and analysis by gas chromatography–mass spectrometry, in general also apply to FMs. 2.1 Laboratory Quality Control Because of the ubiquitous nature of FMs in consumer products, it is critical that any analytical chemistry laboratory measuring these compounds takes extra precautions to avoid laboratory contamination of samples. Several researchers [2, 11, 14–17] have pointed out that likely sources of FM contamination in the modern-day laboratory include the use of consumer products and fine fragrances by laboratory workers, fragrances in soaps used to clean glassware and the laboratory, and laboratory supplies such as gloves. Before beginning the analysis of FMs at low concentrations, the laboratory should analyze several laboratory blank samples to assess the degree to which the laboratory is contaminated. With every set of samples analyzed, the laboratory should also analyze a laboratory and field blank sample. Laboratory workers should be advised to be aware of their personal use of fragrance-enhanced consumer products and the potential for laboratory contamination. 2.2 Standards As interest in measuring FMs in the environment has increased, researchers have used a variety of means to obtain FM standards for use in analytical chemistry.

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The sources of these standards include FM manufacturers, synthesis by researchers (especially in the case of NM metabolites), and specialty chemical catalogs. When possible, it is preferable to obtain these standards directly from the FM manufacturers in order to use the authentic material being discharged to wastewater treatment. In all cases, the purchased, synthesized, or obtained standards must be extensively analyzed to confirm the chemical structure and purity. Also of importance is the appropriate use of surrogates and internal standards for quantification of FMs in wastewater and environmental matrices. Ideally, stable isotope-labeled analogs (such as stable perdeuterated analogs) of the FMs are used for this purpose if gas chromatography–mass spectrometry (GC–MS) is the analysis technique. Simonich et al. [2, 11] synthesized eight perdeuterated FMs, including d3-benzyl acetate, d3-g-methyl ionone, d3-methyl dihydrojasmonate, d3-OTNE, d4-acetyl cedrene, d6-musk xylene, d3-AHTN, and d7-musk ketone, through base-catalyzed exchange of protons with deuterium and used these as surrogates to quantify 16 FMs in wastewater treatment matrices. d3-Terpineol [2, 11, 18] and d6-HHCB [2, 8, 11] have also been used by several researchers. Others have used d34-hexadecane [19], 2,4,5-trichlorotoluene [20], pentachloronitrobenzene [14], 2,2¢-dinitrobiphenyl [14], and perdeuterated polycyclic aromatic hydrocarbons [14, 18, 19, 21, 22] as surrogates or internal standards to measure FMs in wastewater treatment. Finally, the amino metabolites of the NMs have been synthesized by researchers and used as standards. These synthesis methods include reduction of NMs with hydrogen in the presence of Pd/charcoal to form the amino metabolites [15, 16, 23] or reaction of NMs with hydrazine hydrate and Raney nickel [14, 23]. A metabolite of HHCB, HHCB-lactone, has also been synthesized and used as a standard [17]. 2.3 Aqueous Matrices Table 2 lists a variety of analytical methods used to measure FMs in wastewater treatment. Only three studies [2, 11, 24] have attempted to examine fragrance materials throughout the wastewater treatment process, including the analysis of FMs in influent, primary and/or secondary settling, sludge, and/or final effluent. Others have measured FMs in wastewater treatment influent and/or effluent [8, 14, 19, 20, 22]. Still others have focused solely on measuring FMs in sewage sludge and digested sewage sludge [15–18, 21]. Because wastewater treatment processes consist of solid and aqueous phases, analytical methods have been developed to measure FMs in both of these matrices. Although some researchers have chosen to extract aqueous wastewater matrices with traditional methods, such as liquid–liquid extraction [5, 22], other researchers have used solid-phase extraction (SPE) to exhaustively extract FMs from these matrices [2, 8, 11, 14, 19, 20]. Simonich et al. [2, 11] used C18 Bakerbond Speedisks to extract 16 FMs, with a wide range of polarities, from 0.5-L influent and primary effluent and 1.0-L final effluent samples.Verbruggen et al.

Fragrance materials

16 FMs, including those in Table 1, PCMs (AHTN and HHCB), and NMs (MX and MK)

PCMs (AHTN and HHCB) – freely dissolved and total concentrations

PCMs and NMs

NMs and monoamino metabolites

Researchers

Simonich et al. [2, 11]

ArtolaGaricano et al. [24]

Kanda et al. [22]

Rimkus et al. [23]

Influent, effluent

Influent, effluent

Influent, primary settler, aeration tank, secondary effluent, primary sludge, and waste sludge

Influent, primary effluent, activated sludge solids, and final effluent

WWTP matrix

Germany

United Kingdom

The Netherlands

U.S., United Kingdom, and The Netherlands

Location

Table 2 Summary of analytical methods used to measure FMs in wastewater treatment

– Liquid–liquid extraction with hexane – Silica gel and alumina chromatography – GC–MS/MS, GC–MS, GC-ECD, and GC-PND – Limits of quantification=1 ng/L

– Liquid–liquid extraction with dichloromethane – GC–MS – Recovery=69–95%; limit of detection =3.7–8.5 ng/L

Freely dissolved – SPME with GC–MS – Limit of quantification=0.1 mg/L Total concentration – Liquid–liquid extraction with cyclohexane – Silica gel chromatography – GC–MS – Recovery=85–106%; limit of quantification =0.1 mg/L

– Extraction of 0.5–1 L influent, primary effluent, and final effluent with C18 SPE – Extraction of sludge by accelerated solvent extraction with dichloromethane – Silica gel chromatography – Analysis by stable isotope dilution GC–MS using nine perdeuterated FMs – Recovery=97–115%; limit of quantification =0.5–35 ng/L

Analytical method

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Fragrance materials

PCMs and NMs

PCMs (AHTN and HHCB)

NMs, PCMs, and nitromusk metabolites

PCMs (AHTN and HHCB)

NMs and amino metabolites

Researchers

Ricking et al. [19]

Verbruggen et al. [20]

Osemwengie et al. [14]

Buerge et al. [8]

Berset et al. [16]

Table 2 (continued)

Sewage sludge

Wastewater effluent

Wastewater effluent

Wastewater effluent

Wastewater effluent

WWTP matrix

Switzerland

Switzerland

U.S.

The Netherlands

Canada and Sweden

Location

– Extraction with hexane by agitation – Gel-permeation chromatography and silica gel chromatography – GC–MS/MS, GC–MS (EI, NCI, and PCI), 1H and 13C NMR – Recovery=51–116%; detection limit=50 ng/L

– Macroporous polystyrene–divinylbenzene adsorbent – Silica gel chromatography – GC–MS – Recovery=81–141%; LOD=10 ng/L

– On-site 60-L extraction with NEXUS sorbent – Silica gel and gel-permeation chromatography – GC–MS – Recovery=80–97%; method detection limit =0.02–0.3 ng/L

– Biomimetic and exhaustive extraction – C18 Empore disks – GC–MS – Recovery >95%; detection limit=0.1 ng/L

– SPE and filtration and extraction with n-pentane and dichloromethane – Silica gel chromatography – GC–MS – Detection limit=0.5 ng/L

Analytical method

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Fragrance materials

NMs, PCMs, and amino metabolites

PCMs and HHCB-lactone

NMs and PCMs

22 FMs; including PCMs and NMs

Researchers

Herren et al. [15]

Kupper et al. [17]

Stevens et al. [21]

Difrancesco et al. [18]

Table 2 (continued)

Digested sewage sludge

Digested sewage sludge

Sewage sludge

Sewage sludge

WWTP matrix

U.S.

United Kingdom

Switzerland

Switzerland

Location

– Accelerated solvent extraction with dichloromethane – Silica gel chromatography – GC–MS

– Soxhlet extraction with dichloromethane – Silica gel and gel-permeation chromatography – GC–MS

– Extraction with hexane and stirring – GC–MS – Recovery=79–108%; LOD=15–30 umg/kg d.m.; LOQ=45–90 umg/kg d.m.

– Extraction with hexane by agitation – Gel-permeation chromatography and silica gel chromatography – GC–MS/MS, GC–MS (EI, NCI, and PCI) – Recovery=50–118%; detection limit=100 ng/L

Analytical method

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[20] used C18 3 M Empore Disks to exhaustively and biomimetically extract PCMs (AHTN and HHCB) from 10-L effluent samples. Osemwengie et al. [14] used 6 g of Nexus sorbent [polystyrene cross-linked with 50% divinylbenzene and poly(methyl methacrylate)] to extract 60 L of effluent for a variety of NMs, PCMs, and nitromusk metabolites. Buerge et al. [8] used 10 mL of Bio-Beads SM-2 (a macroporous polystyrene–divinylbenzene adsorbent) to extract HHCB and AHTN from 200-mL effluent samples. Finally, Artola-Garicano et al. [24] used 1-cm lengths of a 100-mm polydimethylsiloxane solid-phase microextraction (SPME) fiber to measure free concentrations of AHTN and HHCB throughout wastewater treatment. Many of the aqueous methods listed in Table 2 utilize adsorption chromatography, such as silica or alumina chromatography, to purify extracts prior to analysis. 2.4 Solid Matrices Researchers have chosen to extract sewage sludge for FMs in its wet form using liquid–liquid extraction with solvent [15–17, 24], or to centrifuge the wet sludge, decant the water phase, and extract the sludge by Soxhlet extraction with solvent [21] or at elevated temperature and pressure using accelerated solvent extraction [2, 18]. Herren and Berset [15] and Berset et al. [16] extracted 1 L of wet sewage sludge for NMs and their metabolites with 600 mL hexane for 2 h with vigorous agitation. Artola-Garicano et al. [24] measured the free concentration of AHTN and HHCB in 10 mL wet sludge by negligible depletion SPME and the total concentration by extracting with 6 mL cyclohexane during 2 h of shaking. Stevens et al. [21] extracted NMs and PCMs in 2.5 g dried, centrifuged, and digested sludge by Soxhlet extraction for 18 h with 280 mL dichloromethane. Finally, Simonich et al. [2] and Difrancesco et al. [18] used accelerated solvent extraction (at 60 °C and 2,000 PSI) with dichloromethane to extract a wide range of FMs from centrifuged activated sludge solids and digested and dewatered sludges. In the methods used to extract centrifuged sludges, Na2SO4 was used to remove water from the sample prior to solvent extraction. Many of the sludge methods listed in Table 2 utilize gel-permeation chromatography (GPC) to remove high molecular weight interferences and/or adsorption chromatography, such as silica or alumina chromatography, to purify extracts prior to analysis. 2.5 Analysis Because FMs are semivolatile, they are amenable to analysis by gas chromatography (GC) and gas chromatography–mass spectrometry (GC–MS) without derivitization. Table 2 shows that all of the analytical methods developed to measure FMs in wastewater treatment to date utilize GC or GC–MS.

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In general, FMs can be chromatographically resolved using a 30-m nonpolar GC column, such as a DB-5 [2, 11, 24] or DB-1701 [18] column. Rimkus et al. [23] used GC with an electron-capture detector (ECD) and nitrogen–phosphorus detector (NPD) to analyze for NMs and their metabolites. However, the majority of studies use GC–MS with electron impact ionization to detect and quantify the wide array of FM structures [2, 5, 8, 11, 14, 18–21, 24].An ion-trap mass spectrometer has been used to analyze for FMs by GC–MS/MS and negative chemical ionization has been used to improve sensitivity for the NMs [15, 16]. Because of the volatile nature of FMs, care must be taken when evaporating the extraction solvent to low volumes prior to analysis, and volatile solvents should be used in order to minimize the loss of FMs during this step [2].

3 Sampling Wastewater Treatment Plants for Fragrance Materials 3.1 Selection of Wastewater Treatment Plants Because FMs are used in consumer products, it is important that investigations of their removal during wastewater treatment be conducted at municipal wastewater treatment plants primarily receiving domestic wastewater (>80% domestic wastewater) [2]. In addition, the operation of these plants should be well documented and reported, including: plant design, wastewater flow, hydraulic (HRT) and solids retention times (SRT), and biochemical oxygen demand (BOD) and total suspended solids (TSS) removal. Specific wastewater treatment plants should be selected to represent a range of dry and wet climates, geographic locations (such as the U.S. and Europe), and plant designs (including primary treatment only, activated sludge, carousel, oxidation ditch, trickling filter, rotating biological contactor, lagoon, etc.) in order to obtain a comprehensive understanding of FM removal during wastewater treatment. Plant designs should be selected so that the most prevalent types of wastewater treatment plant design for that geography are sampled, based on both total wastewater flow and number of wastewater treatment plants [11]. For example, in the U.S. approximately 81% of the wastewater flow is treated by activated sludge plants, 7% by trickling filter, 6% by lagoon, 3% by oxidation ditch, 2% by rotating biological contactor, and 1% by primary treatment [11]. In Europe, activated sludge, carousel, trickling filter, and oxidation ditch are among the most common types of wastewater treatment [11]. In a study of the removal of 16 FMs from wastewater, Simonich et al. [11] conducted their studies at 17 wastewater treatment plants with plant flows of 1.4¥106–1.0¥108 L day–1. Twelve of the plants were located in different states and regions of the U.S. and five plants were located in Europe. In addition, five of the 17 plants were activated sludge plants, two were carousel plants, two were oxidation ditch plants, five were trickling filter plants, one was a rotating bio-

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logical contactor plant, and two were lagoons. Artola-Garicano et al. [24] studied the removal of freely dissolved and total concentrations of AHTN and HHCB in four wastewater treatment plants located in The Netherlands. These plants had flow rates ranging from 800 to 4,200 m3/h, HRTs ranging from 4.8 to 8.9 h, and SRTs ranging from 8 to 22 days [24]. Kanda et al. [22] studied PCMs and NMs in influent and effluent from six wastewater treatment plants in the U.K. The flow rates of these plants ranged from 103 to 3,198 m3/day and included rotating biological contactor with reed beds, submerged aerated filter, oxidation ditch, biological filter bed, activated sludge, and trickling filter plant designs. Finally, Buerge et al. [8] measured HHCB and AHTN in effluents from five wastewater treatment plants in Switzerland with flow rates ranging from 3,177 to 14,250 m3/day–1. Daily BOD and TSS removal at the plant should be measured and evaluated in order to judge how well the plant operates during low and high flow conditions. Measurements of BOD and TSS removal should be done on the days in which FM monitoring takes place at the plant. This is important, because Simonich et al. [11] showed that the overall removal of biodegradable, nonsorptive FMs from most plant designs is positively correlated with plant BOD removal and that the overall removal of nonbiodegradable, sorptive FMs is positively correlated with plant TSS removal (see Fig. 4). This was true for all plant designs except for lagoons, which had poor BOD and TSS removal (due to aquatic vegetation growing in lagoons) but had good removal of FMs due to long HRTs and SRTs. BOD and TSS removal is governed by both plant design and daily operation.

Fig. 4 Correlation of the measured overall removal of terpineol with plant 5-day BOD removal and the measured overall removal of HHCB with plant TSS removal. Regressions include all wastewater treatment plants studied by Simonich et al. [11] except for the two lagoons (see text), and are significant at the 99.9% level; n=15. Dashed lines are the 95% confidence intervals of the regressions

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3.2 Wastewater Treatment Plant Sampling When planning which wastewater compartments to measure and at what frequency, it is important to consider the potential for FM concentrations to vary throughout the day, within the plant. Simonich et al. [2] measured 16 FMs in the influent of a U.S. wastewater treatment plant every 2 h over a 24-h period. Their data indicate that the total FM concentration in influent varies greatly throughout a 24-h period, with a relative standard deviation in total FM concentration of 38.9% (see Fig. 5). This variation in influent concentrations has been observed for other consumer product chemicals, such as surfactants [2], and is a function of consumer use and disposal of these chemicals and water discharge volume changing throughout the course of the day. In general, Simonich et al. [2] measured low FM influent concentrations from 11:30 p.m. to 7:30 a.m. and high influent concentrations from 9:30 a.m. to 9:30 p.m. These time periods are consistent with consumer use of these chemicals, considering that sewer transport times to wastewater treatment plants can be on the order of 0.5–2 h in some locations. Simonich et al. [2] also observed that the total FM concentration in final effluent varied significantly less than the influent concentrations, with a relative standard deviation of 8.0% (see Fig. 5). This is most likely due to the capacity of the treatment process to treat fluctuations in influent concentration due to the long residence times within the plant.

Fig. 5 Diurnal fluctuations in total FM concentration in influent and final effluent collected from an activated sludge wastewater treatment plant [2]. Hourly samples were combined to represent a 2-h period

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Artola-Garicano et al. [24] measured the free and total concentrations of AHTN and HHCB in the influent of a wastewater treatment plant in The Netherlands every 2 h over a 24-h period. Their data indicate that the variation in total concentration of AHTN and HHCB in influent was 19%, while the variation in free concentration was less than 10% over the 24-h period. These authors suggested that fluctuations in water volume cause fluctuations in total concentrations; however, for hydrophobic FMs such as AHTN and HHCB, the solids act as a reservoir and stabilize the free concentrations. If the objective of measuring FMs, or other consumer product chemicals for that matter, in wastewater treatment is to understand FM removal and mechanisms of removal across wastewater treatment processes, then it is important to collect samples at least every 2 h and composite these samples into a single, flow-based 24-h sample. Otherwise, the results may be significantly over- or underestimated depending on the time of the day the sample was collected. However, if the objective is to monitor FMs in only final effluent or sludge, representative grab sampling may be sufficient.

4 Mechanisms of Fragrance Material Removal During Wastewater Treatment Because of their wide range of physical-chemical properties and biodegradabilities (see Table 1), FMs have the potential to biodegrade, sorb to solids, and/or volatilize during wastewater treatment. The relative importance of these removal mechanisms will depend on the specific FM, the plant design, and the kinetics of each of these processes within the plant. It is also important to acknowledge that FMs bound to solids are not available for biodegradation or volatilization, and it is thought that only FMs freely dissolved in the aqueous phase are available for these processes [24]. 4.1 Biodegradation As shown in Table 1, many FMs meet the biodegradation criteria of a ready or inherent test. If a FM meets the criteria of a ready test, with or without acclimation, a first-order biodegradation rate of 3 h–1 in activated sludge can be assumed [1]. For FMs that show extensive biodegradation but fail the ready test criteria, a first-order rate of 0.3 h–1 can be assumed for activated sludge treatment [1]. Table 1 also indicates that some FMs, including the PCMs and NMs, do not pass ready or inherent biodegradation tests. However, this does not mean that these FMs do not undergo biotransformation to polar metabolites under realistic conditions. These realistic biodegradation tests may be conducted in vitro, in bench-top die-away studies, or as continuous activated sludge and porous pot tests. Ideally, the conditions should include: (1) realistic FM concentrations

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Table 3 AHTN and HHCB biotransformation rates in activated sludge measured by Federle et al. [12] and Artola-Garicano et al. [13]

Federle et al. [12]

Artola-Garicano et al. [13]

PCM

Total concentration (mg/L)

kbiodeg (h–1)

AHTN

5 50 5 25

0.015±0.001 0.008±0.001 0.010±0.002 0.021±0.003

HHCB

Total concentration (mg/L)

kbiodeg (h–1)

5.25

0.023

10.33

0.071

(often achieved through the use of radioisotope-labeled FMs), (2) realistic acclimated sludge concentrations, and (3) realistic exposure times. Several researchers have studied the biotransformation of HHCB and AHTN under realistic activated sludge conditions. Federle et al. [12] studied the biotransformation of 14C-HHCB in activated sludge die-away tests using realistic HHCB concentrations (5 and 25 mg/L) and acclimated sludge concentrations (approximately 2,500 mg/L). The polar biotransformation products of HHCB, including the lactone and hydroxy acid of HHCB, were identified and their corresponding octanol–water partition coefficients estimated by HPLC. The firstorder rate constant for parent HHCB biotransformation in activated sludge was determined to be 0.010–0.021 h–1, depending on HHCB concentration, in these tests (see Table 3). Federle et al. [12] also studied the biotransformation of 14C-AHTN in similar activated sludge die-away tests and in a continuous activated sludge test. Polar biotransformation products of AHTN were identified and their octanol–water partition coefficients were estimated from HPLC data. The first-order rate constant for parent AHTN biotransformation in activated sludge was determined to be 0.008–0.015 h–1, depending on AHTN concentration in the activated sludge die-away test (see Table 3) [12]. The overall removal of AHTN (due to biotransformation, sorption, volatilization) in the continuous activated sludge test was 86.4% and the removal of parent AHTN due to biotransformation was estimated to be 37.4%. Artola-Garicano et al. [13] studied the biodegradation of AHTN and HHCB in activated sludge by measuring the free concentration, using negligible depletion SPME, and total concentration over time. The first-order rate constant for parent AHTN biotransformation was determined to be 0.023 h–1, while the first-order rate constant for parent HHCB biotransformation was 0.071 h–1 (see Table 3) [13]. These authors also determined that microbial biodegradation activity was the rate-limiting step in biotransformation of these compounds and not desorption from the activated sludge. Table 3 shows that the first-order rate constants for parent AHTN and HHCB biotransformation, determined by Federle et al. [12] and Artola-Garicano et al.

Fragrance Materials in Wastewater Treatment

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[13], are similar even though the techniques used to determine these rate constants were quite different and the sources of activated sludge were different (U.S. and Europe). Finally, there is empirical evidence of the biotransformation of NMs and PCMs during wastewater treatment from measurements of the amino metabolites of the NMs and the lactone of HHCB in sewage sludge [15–17]. 4.2 Sorption For the hydrophobic, nonbiodegradable FMs listed in Table 1, such as the NMs and PCMs, removal due to sorption on sewage solids is a significant removal mechanism. Evidence of the significance of this removal mechanism is the measurement of PCMs and NMs in sewage sludge throughout the world [15–18, 21]. Of the FMs listed, Difrancesco et al. [18] measured acetyl cedrene, hexyl salicylate, hexylcinnamic aldehyde, AHTN, HHCB, g-methyl ionone, musk ketone, musk xylene, and OTNE in digested and dewatered sludge samples. The results indicate that FMs with activated sludge sorption coefficients (Kd) as low as 2,000 L kg–1 have the potential to be removed to a significant degree due to sorption to sewage sludge. FM sorption coefficients for sewage sludge have most often been estimated using the octanol–water partition coefficient of the FM rather than measured directly. There is general agreement between the Kd values measured for AHTN and HHCB and the estimates based on their log Kow values (see Table 1) [11]. Artola-Garicano et al. [13] determined the organic carbon normalized sorption coefficient for activated sludge to be 6,681 L kg–1 for HHCB and 7,018 L kg–1 for AHTN. Finally, Federle et al. [12] estimated the removal of 14C-AHTN in a continuous activated sludge test to be 44.7% based on sorption to activated sludge alone. 4.3 Volatilization FMs have the potential to volatilize and enter the atmosphere during manufacturing and consumer use and disposal. The PCMs and NMs have been detected in ambient air [9, 25]. However, most FMs have atmospheric lifetimes sufficiently short that they are unlikely to undergo atmospheric long-range transport [26]. Because some FMs have large Henry’s law constants (Table 1) and some wastewater treatment plant designs have active aeration and large surface areas that are exposed to the atmosphere, it is likely that some FMs volatilize during wastewater treatment. However, the FMs with large Henry’s law constants also have large Kd values (see Table 1), so that in portions of the treatment process with active aeration and high solids concentrations (such as activated sludge) it is not entirely clear whether volatilization or sorption to solids will be the dominant loss mechanism. Although there are limited experimental data on FM volatilization during wastewater treatment, volatilization appears to ac-

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S. L. Simonich

count for <5% of the total FM removal during wastewater treatment. Federle et al. [12] estimated the removal of 14C-AHTN in a continuous activated sludge test to be 3.4% based on volatilization alone.

5 Measurement of Fragrance Materials in Wastewater Treatment 5.1 Concentrations in Treatment Plants Given the variety of analytical methods used to measure FMs in wastewater treatment and the number of researchers measuring these compounds in wastewater collected from the U.S. and Europe (see Table 2), it is important to compare and contrast the concentrations of FMs measured throughout wastewater treatment by different researchers, in different geographies. In addition, one might expect that there would be plant-to-plant variation in the FM concentration in wastewater treatment because of differences in the volume of FMs used (per capita FM use) and differences in the per capita water use for a given geography. These measurements and differences are important to understand because, ultimately, the concentration of FMs in the final effluent and sewage sludge are used to develop aquatic and terrestrial environmental risk assessments for these compounds [1]. Simonich et al. [11] compared the concentration of 16 FMs in wastewater influent collected from 12 U.S. treatment plants and five treatment plants from the U.K. and The Netherlands (see Table 4). It is important to characterize FMs in influent because it is a good representation of the relative amounts of FMs being used by consumers and there is minimal opportunity for biodegradation, sorption, and/or volatilization in transit to the wastewater treatment plant. In addition, because the same analytical methods were used by the same laboratory to measure U.S. and European influent in the Simonich et al. [11] study, we can directly compare concentrations between the U.S. and Europe without questioning differences in laboratory procedures or analytical methodology. Simonich et al. found that the influent concentrations of 14 of the 16 FMs measured were statistically similar in both U.S. and European treatment plants [11]. Only isobornyl acetate and OTNE influent concentrations were statistically different, with higher concentrations of these FMs in European influent. The influent concentrations of the 16 FMs are given in Table 4. Figure 6A shows the normalized relative profile of FMs measured in U.S. and European influent during the [11] study. In both sets of influent, terpineol dominated the FM profile and had the highest concentrations, while the nitromusks (MX and MK) had the lowest concentrations. The large standard deviations in the influent concentrations (Table 4 and Fig. 6A) indicate that there is significant variability in influent concentrations within both U.S. and European treatment plants.

Location and number

U.S.; n=12 plants

U.K. and The Netherlands; n=5 plants

Researchers

Simonich et al. [2, 11]

Simonich et al. [11]

1999–2000

1997–1999

Years sampled

AHTN=5.97±3.88; HHCB=9.71±5.09

AHTN=12.5±7.35; HHCB=16.6±10.4

PCM concentration (mg/L) – mean ±standard deviation and/or range

Table 4 Summary of FM influent concentrations to wastewater treatment

MK=0.996±0.741; MX=0.248±0.136

MK=0.640±0.395; MX=0.386±0.299

NM concentration (mg/L) – mean ±standard deviation and/or range

Benzyl acetate=9.85±10.2; methyl salicylate=11.3±13.0; methyl dihydrojasmonate=11.9±5.31; terpineol=56.3±33.9; benzyl salicylate=10.2±4.51; isobornyl acetate=37.1±28.4; g-methyl ionone=3.63±1.90; p-t-bucinal=2.56±1.96; hexylcinnamaldehyde=12.8±7.27; hexyl salicylate=6.89±3.63; OTNE=9.00±3.77; acetyl cedrene=7.15±4.32

Benzyl acetate=3.74±3.46; methyl salicylate=10.2±9.69; methyl dihydrojasmonate =7.21±4.19; terpineol=63.7±36.4; benzyl salicylate=19.5±10.8; isobornyl acetate=6.47±8.53; g-methyl ionone=3.37±2.56; p-t-bucinal=1.61±0.731; hexylcinnamaldehyde=15.3±12.1; hexyl salicylate=5.48±3.56; OTNE=3.55±1.93; acetyl cedrene=4.97±2.27

Other FM concentration (mg/L)– mean±standard deviation and/or range

Fragrance Materials in Wastewater Treatment 99

UK; n=6 plants

Germany; n=1 plant

Kanda et al. [22]

Rimkus et al. [23]

1996

2001

The Netherlands; 2001 n=4 plants

Artola-Garicano et al. [24] – total concentrations

Years sampled

Location and number

Researchers

Table 4 (continued)

Not measured

AHTN=2.2–8.1; HHCB=8.4–19.2; DPMI=<0.01–0.4; ADBI=<0.01–0.44; AHDI=<0.01–0.1; ATII=<0.01–2.9

AHTN=0.54±0.05 –1.76±0.09; HHCB=1.42±0.12 –4.30±0.23

PCM concentration (mg/L) – mean ±standard deviation and/or range

MX=0.150; 4-NH2-MX=<0.01; 2-NH2-MX=<0.01; MK=0.55; 2-NH2-MX=<0.01

MA=<0.01; MX=<0.01–4.7; MM=<0.01; MT=<0.01; MK=<0.01–2.9

Not measured

NM concentration (mg/L) – mean ±standard deviation and/or range

Not measured

Not measured

Not measured

Other FM concentration (mg/L)– mean±standard deviation and/or range

100 S. L. Simonich

Fragrance Materials in Wastewater Treatment

101

A

B Fig. 6A, B Average relative profile and standard deviation of FMs in A influent and B primary effluent in the U.S. and Europe [11]. The highest concentration FM was normalized to 1. The highest concentration FM (in mg/L) is in parentheses. The error bars represent the normalized standard deviation of the mean

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When we compare the PCM and NM influent concentrations measured by Simonich et al. to those of other researchers (Table 4), we can ascertain some general trends. AHTN and HHCB are the highest concentration PCMs in both U.S. and European influent, with concentrations ranging from 1 to 20 mg/L, and HHCB concentrations being greater than AHTN concentrations. The ratio of HHCB to AHTN concentration in U.S. and European influent is in the range of 1.3–3.8. In general, the concentration of other PCMs in influent are either below the limit of quantitation or less than 1 mg/L [22]. The NM concentrations in U.S. and European influent are significantly less than the PCM concentrations and are in the range of 0.2–5 mg/L. MX and MK are the highest concentration NMs in influent [22]. In general, the amino metabolites of the nitromusks are not detected in influent because there is limited opportunity for biotransformation in transit to the wastewater treatment plants. A larger number of researchers have measured FMs in final effluent from wastewater treatment (Table 5) because this is the most important wastewater parameter for accessing FM discharge to aquatic ecosystems. The concentration of FMs in final effluent is a function of the FM concentration in influent and the efficiency of FM removal across the plant (including plant design and operation). In the Simonich et al. [11] study, the concentration of 16 FMs in final effluent ranged from 0.001–7.6 mg/L in the U.S. to 0.01–4.6 mg/L in Europe (see Table 5). In addition, terpineol no longer dominated the relative FM profile in final effluent, except for final effluent collected from two European trickling filter plants (see Fig. 7) [11]. As the figure indicates, nonbiodegradable and inherently biodegradable, sorptive FMs dominated the relative FM profile in final effluent (including HHCB, AHTN, OTNE, and acetyl cedrene). Of the PCMs, AHTN and HHCB have the highest concentrations in final effluent (see Table 5) [14, 19, 22]. In the U.S., the concentration of AHTN in final effluent ranged from 0.024 to 1.7 mg/L, while the concentration of HHCB ranged from 0.032 to 2.2 mg/L (see Table 5) [11, 14]. In Europe, the final effluent concentrations of AHTN ranged from 0.11 to 2.7 mg/L and the concentrations of HHCB ranged from 0.21 to 6.4 mg/L (see Table 5). MX and MK are the most prevalent NMs in final effluent and are in the concentration range of <1–710 ng/L in European effluents and <MDL–112 ng/L in U.S. effluents (see Table 5). The amino metabolites of the NMs have been detected in U.S. and European effluent in the concentration range of <MDL–250 ng/L (see Table 5). It is also important to consider the concentration of FMs in sewage sludge because this is another route in which FMs may be removed during wastewater treatment and be released to the environment through land application of sludge solids. Only one study has attempted to measure a large number of FMs in digested sludge. Difrancesco et al. [18] detected AHTN, HHCB, MK, g-methyl ionone, hexylcinnamaldehyde, hexyl salicylate, and acetyl cedrene in digested sludge from U.S. wastewater treatment plants (see Table 6). The other FMs they studied were not detected.As Table 6 indicates, the concentrations of these FMs

Location and number

U.S.; n=12 plants

U.K. and The Netherlands; n=5 plants

Researchers

Simonich et al. [2, 11]

Simonich et al. [11]

1999–2000

1997–1999

Years sampled

AHTN=620–2,670; HHCB=980–4,620

AHTN=24–1,710; HHCB=32–2,210

PCM concentration (mg/L) – mean ±standard deviation and/or range

MK=40–770; MX=10–170

MK=10–67; MX=1–112

NM concentration (mg/L) – mean ±standard deviation and/or range

Table 5 Summary of FM final effluent concentrations from wastewater treatment

Benzyl acetate=60–260; methyl salicylate=40–220; methyl dihydrojasmonate=26–1,920; terpineol=80–15,100; benzyl salicylate=20–1,960; isobornyl acetate=10–290; g-methyl ionone=30–730; p-t-bucinal=40–180; hexylcinnamaldehyde=20–910; hexyl salicylate=10–910; OTNE=490–3,190; acetyl cedrene=70–1,430

Benzyl acetate=2–252; methyl salicylate=13–693; methyl dihydrojasmonate=3–456; terpineol=11–1,079; benzyl salicylate=5–1,025; isobornyl acetate=7–112; g-methyl ionone=7–214; p-t-bucinal=13–258; hexylcinnamaldehyde=10–77; hexyl salicylate=1–243; OTNE=25–615; acetyl cedrene=12–1,359

Other FM concentration (mg/L)– mean±standard deviation and/or range

Fragrance Materials in Wastewater Treatment 103

UK; n=6 plants

Germany; n=3 plants

Canada; n=3 plants

Kanda et al. [22]

Rimkus et al. [23]

Ricking et al. [19]

2002

1996

2001

The Netherlands; 2001 n=4 plants

Artola-Garicano et al. [24] – total concentrations

Years sampled

Location and number

Researchers

Table 5 (continued)

AHTN=42–104; HHCB=157–423; DPMI=<1; ADBI=2–8; AHDI=2–5; ATII=<1

Not measured

AHTN=310–2,700; HHCB=1,100–6,400; DPMI=<10–160; ADBI=<10–91; AHDI=<10–48; ATII=<10–790

AHTN=420±60 –1,200±180; HHCB=1,250±20 –2,220±90

PCM concentration (mg/L) – mean ±standard deviation and/or range

MX=<1; MK=<1

MX=<3–10; 4-NH2-MX=7–34; 2-NH2-MX=<1–10; MK=6–94; 2-NH2-MK=15–250

MA=<10; MX=<10–650; MM=<10; MT=<10; MK=<10–710

Not measured

NM concentration (mg/L) – mean ±standard deviation and/or range

Not measured

Not measured

Not measured

Not measured

Other FM concentration (mg/L)– mean±standard deviation and/or range

104 S. L. Simonich

Location and number

Sweden; n=5 plants

U.S.; n=2 plants

Switzerland; n=5

Researchers

Ricking et al. [19]

Osemwengie et al. [14]

Buerge et al. [8]

Table 5 (continued)

2001

Unknown

1999

Years sampled

AHTN=310–760; HHCB=720–1,950

AHTN=26.6–92.2; HHCB=35.0–152; DPMI=<MDL; ADBI=0.3–2.1; AHDI=2.4–5; ATII=<MDL–126

AHTN=110–520; HHCB=205–1,300; DPMI=<1; ADBI=4–19; AHDI=2–6; ATII=<1

PCM concentration (mg/L) – mean ±standard deviation and/or range

Not measured

MX=<MDL–1.3; MK=<MDL–27.5; MA=<MDL; MM=<MDL; MT=<MDL; 4-NH2-MX= <MDL–31.5; 2-NH2-MX= <MDL–0.9; NH2-MK=<MDL

MX=<1; MK=<1

NM concentration (mg/L) – mean ±standard deviation and/or range

Not measured

Not measured

Not measured

Other FM concentration (mg/L)– mean±standard deviation and/or range

Fragrance Materials in Wastewater Treatment 105

Fig. 7 Average relative profile and standard deviation of FMs in final effluent in U.S. and European plants [11]. The highest concentration FM was normalized to 1. The highest concentration FM (in mg/L) is in parentheses. The error bars represent the normalized standard deviation of the mean

Location and number

The Netherlands; n=4 plants

Switzerland; n=10 plants

Switzerland; n=12 plants

Researchers, sample type, and concentration units

Artola-Garicano et al. [24] – total concentration waste sludge in mg/L

Berset et al. [16] – concentration in mg/kg dry weight

Herren et al. [15] – concentration in mg/kg dry weight

Unknown

Unknown

2001

Years sampled

Table 6 Summary of FM concentrations in wastewater sludge

AHTN=741–4,161; HHCB=2,293–12,157; DPMI=38.4–332; ADBI=41–330; AHDI=64.9–843; ATII=n.q.

Not measured

AHTN=12.38±0.19 –82.67±43.09; HHCB=29.47±10.84 –234.60±111.86

PCM concentration – mean±standard deviation and/or range

MA=nd; MX=nd–32.5; MM=nd; MT=nd; MK=nd–7.0, amino metabolites= nd–36.2

MA=nd; MX=nd–32.5; MM=nd; MT=nd; MK=nd–7.1, amino metabolites= nd–49.1

Not measured

NM concentration – mean±standard deviation and/or range

Not measured

Not measured

Not measured

Other FM concentration – mean±standard deviation and/or range

Fragrance Materials in Wastewater Treatment 107

Location and number

Switzerland; n=16 plants

U.K.; n=14 plants

U.S.; n=2 plants

Researchers, sample type, and concentration units

Kupper et al. [17] – concentration in mg/kg dry weight

Stevens et al. [21] – digested sludge concentration in mg/kg dry weight

Difrancesco et al. [18] – digested sludge concentration in mg/kg dry weight

Table 6 (continued)

2000 and 2002

Unknown

2001

Years sampled

AHTN=8,100–51,000; HHCB=21,800–86,000

AHTN=120–16,000; HHCB=1,900–81,000; ADBI=10–260; AHDI=32–1,100; ATII=44–1,100; DPMI – not detected

AHTN=2,500–11,200; HHCB=7,400–3,600; ADBI=100–1,100; AHDI=200–1,800; ATII=200–1,000; HHCB lactone =600–3,500

PCM concentration – mean±standard deviation and/or range

MK=1,300; MX not detected

MA, MX, MM, and MT not detected

Not measured

NM concentration – mean±standard deviation and/or range

g-Methyl ionone=1,100–3,800; hexylcinnamaldehyde=4,100; hexyl salicylate=1,500; OTNE=7,300–30,700; acetyl cedrene=900–31,300; remaining FMs not detected

Not measured

Not measured

Other FM concentration – mean±standard deviation and/or range

108 S. L. Simonich

Fragrance Materials in Wastewater Treatment

109

ranged from 900 to 86,000 mg/kg dry weight, with AHTN and HHCB having the highest concentrations (8,100–86,000 mg/kg dry weight) [18]. The remaining studies on sewage sludge were conducted at European wastewater treatment plants (see Table 6). Of the PCMs, AHTN and HHCB had the highest concentrations on European sludge (120–81,000 mg/kg dry weight), however the other PCMs were also detected. In one study, the lactone of HHCB was also detected on sewage sludge [17]. The NMs and their amino metabolites have been detected on sewage sludge, however their concentrations (nd– 1,300 mg/kg dry weight) are much lower than the PCMs. 5.2 Removal During Treatment Removal of FMs during wastewater treatment is a function of the tendency for FMs to biodegrade, sorb, and/or volatilize during treatment, plant design, and plant operation. FMs may be removed from wastewater during primary and secondary treatment. Simonich et al. [11] showed that FMs undergo significant removal during primary treatment, ranging from 14 to 50% removal (Table 7). Because both sorptive and nonsorptive FMs are removed, these authors suggested that both sorption and biodegradation play a role in the removal of FMs during primary treatment. Further evidence of this is in the comparison of Figs. 6A and 6B, in which the relative profile of FMs does not change significantly from influent (Fig. 6A) to primary effluent (Fig. 6B), although the FM concentrations decrease. The lack of enhancement of biodegradable, nonsorptive FMs in the primary effluent relative profile suggests that removal due to biodegradation, as well as sorption, occurs during primary treatment. The large standard deviations in the primary treatment removals measured by Simonich et al. [11] suggest that there is significant plant-to-plant variability in primary removal of FMs. Simonich et al. [2, 11], Artola-Garicano et al. [24], Kanda et al. [22], and Rimkus et al. [5] studied the removal of FMs following primary and secondary treatment (see Table 7). Simonich et al. [11] showed that FM removal is dependent on plant design and operation and ranges from 50 to 99.9%. The overall removal (primary + secondary treatment) of 16 FMs ranged from 87.8 to 99.9% for activated sludge plants, 58.6 to 99.8% for carousel plants, 88.9 to 99.9% for oxidation ditch plants, 71.3 to 98.6% for trickling filter plants, 80.8 to 99.9% for a rotating biological contactor plant, and 96.7 to 99.9% for lagoons. Lagoons resulted in the most effective removal of FMs and, in general, the lowest final effluent concentrations due to long retention times. The relative FM profiles in final effluent (Fig. 7), broken down by treatment type, gives some perspective on which plant designs are best at removing biodegradable chemicals (significant enhancement of nonbiodegradable, sorptive FMs as in the case of activated sludge, rotating biological contactor, and oxidation ditch plants) and those which are not (enhancement of both nonbiodegradable, sorp-

Location and number

U.S., UK, The Netherlands; n=17 plants

Researchers

Simonich et al. [2, 11]

1997 – 2000

Years sampled Primary removal only: MX=41.2±21.9; MK=26.6±21.5 Primary+ secondary removal: MX=87.6–99.9; MK=85.2–96.7

Primary+ secondary removal: AHTN=50.6–99.9; HHCB=63.5–99.7

Percent NM removal

Primary removal only: AHTN=28.9±20.1; HHCB=29.9±23.4

Percent PCM removal

Table 7 Summary of percent removal of FMs during wastewater treatment

Primary+secondary removal: benzyl acetate=86.4–99.9; methyl salicylate=92.0–99.9; methyl dihydrojasmonate=81.9–99.9; terpineol=95.4–99.9; benzyl salicylate=90.3–99.9; isobornyl acetate=84.5–99.9; g-methyl ionone=83.1–99.8; p-t-bucinal=84.8–99.3; hexylcinnamaldehyde=95.3–99.9; hexyl salicylate=96.4–99.9; OTNE=51.4–99.4; acetyl cedrene=71.3–99.9

Primary removal only: benzyl acetate=28.2±27.5; methyl salicylate=40.4±32.2; methyl dihydrojasmonate=14.6±19.4; terpineol=15.5±12.0; benzyl salicylate=41.8±25.0; isobornyl acetate=29.0±29.5; g-methyl ionone=20.8±19.6; p-t-bucinal=50.6±23.4; hexylcinnamaldehyde=47.1±19.4; hexyl salicylate=37.3±21.0; OTNE=28.8±22.7; acetyl cedrene=31.6±20.3

Percent other FM removal

110 S. L. Simonich

Location and number

The Netherlands; n=4 plants

UK; n=6 plants

Germany; n=1

Researchers

Artola-Garicano et al. [24] – total concentrations

Kanda et al. [22]

Rimkus et al. [23]

Table 7 (continued)

1996

2001

2001

Years sampled

Not measured

Primary+ secondary removal: AHTN=40.0–96.17; HHCB=39.05–93.49; others not given

Primary+ secondary removal: AHTN=14.3–56.3; HHCB=12.0–59.8

Percent PCM removal

Primary+ secondary removal: MX=93.3; MK=98.9

Primary+ secondary removal: MX=80.3–86.2; MK=53.6–64.5

Not measured

Percent NM removal

Not measured

Not measured

Not measured

Percent other FM removal

Fragrance Materials in Wastewater Treatment 111

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S. L. Simonich

tive FMs and some biodegradable, nonsorptive FMs as in the case of trickling filter). As mentioned previously, Simonich et al. [11] showed that plant operation, including the efficiency of removal of TSS and BOD, affects the removal of FMs to a significant degree (see Fig. 4 and earlier discussion) and is likely a more important variable in insuring efficient removal of FMs during wastewater treatment than is plant design alone. The overall removal (primary + secondary treatment) of AHTN and HHCB has been measured by Simonich et al. [11], Artola-Garicano et al. [24], and Kanda et al. [22] (see Table 7). Simonich et al. measured the overall removal of AHTN to be 50.6–99.9% and the removal of HHCB to be 63.5–99.7% in the U.S. and Europe (UK and The Netherlands), depending on treatment type and TSS removal. Artola-Garicano et al. measured the overall removal of AHTN to be 14.3–56.3% and the removal of HHCB to be 12.0–59.8% at four treatment plants in The Netherlands, based on total concentrations. Finally, Kanda et al. measured the overall removal of AHTN to be 40.0–96.2% and the removal of HHCB to be 39.1–93.5% at six different treatment plants in the U.K. The overall removal of AHTN and HHCB measured by Artola-Garicano et al. [24] appears to be significantly lower than the overall removals measured by Simonich et al. [11] and Kanda et al. [22]. This may be due, in part, to the collection of grab samples by Artola-Garicano et al. and the collection of flowbased composite samples by Simonich et al. and Kanda et al., to the relatively short hydraulic retention times in several of the treatment plants monitored by Artola-Garicano et al., and/or excessive levels of TSS in effluent from the plants monitored by Artola-Garicano et al. As previously mentioned, the FM influent concentrations vary significantly throughout the day (Fig. 5). Artola-Garicano et al. [24] confirmed that the total AHTN and HHCB concentration in influent varied throughout the day, while the free FM concentration showed less variability. Simonich et al. [11] measured relatively low overall removals of AHTN and HHCB at two carousel treatment plants in The Netherlands (58.6 and 63.5%, respectively); however, these overall removals were not as low as those reported by Artola-Garicano et al. [24] for other treatment plants in The Netherlands. Finally, the range of AHTN and HHCB overall removal measured by Simonich et al. [11] and Kanda et al. [22] is comparable. Both studies collected flow-composite samples and sampled a variety of different plant designs. Simonich et al. [11], Kanda et al. [22], and Rimkus et al. [23] measured the removal of MX and MK during wastewater treatment (see Table 7). Simonich et al. and Rimkus et al. measured the overall removals of MX and MK in the range of 85–99.9%. Kanda et al. measured the removal of MX in the range of 80–86% and the removal of MK in the range of 53–65%, however the removal of these NMs was not reported for all six treatment plants.

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6 Predicting Fragrance Material Removal During Wastewater Treatment 6.1 Framework for Aquatic Risk Assessment Because of the large number of FMs in commerce, a framework document was developed to prioritize FMs for aquatic risk assessment [1]. An integral part of this risk assessment is an accurate prediction of FM concentration in the aquatic environment. In order to do this, the framework document outlines some simple calculations to estimate the concentration and removal of FMs during wastewater treatment. This is done by first using the annual FM volume of use and the per capita water use in the geographic area to predict an average influent concentration for the U.S. and Europe. Next, the FM removal during primary treatment (sorption and settling only) is predicted using the octanol–water partition coefficient. Finally, the FM removal during secondary treatment and final effluent concentrations are predicted in the first tier of the framework using the octanol–water partition coefficient to estimate sorption. Biodegradation rates are added to sorption in the second tier of the framework if a more refined exposure assessment is needed. Readily biodegradable, inherently biodegradable, and nonbiodegradable FMs are assumed to have biodegradation rates of 3, 0.3, and 0 h–1, respectively [1]. This simple model does not account for volatilization and the assumptions and equations in the framework document are directly applicable to all primary treatment, but only to activated sludge secondary treatment. The largest single source of error in the model is likely the estimation of annual FM volume use because of uses outside the fragrance industry and, in some cases, natural sources [1]. Because of the large number of FMs being evaluated and the wide range of physical-chemical properties and biodegradabilities they represent, it is important to determine if calculations outlined in the framework document are conservative for predicting FM concentration and removal during wastewater treatment. The Simonich et al. [11] dataset for 16 FMs measured in 17 wastewater treatment plants was used to evaluate the assumptions made in the framework document. Figure 8A shows the regression of the measured percent primary removal (Table 7) with the percent primary removal predicted by the framework. All of the plants with primary treatment studied by Simonich et al. are included in this regression. Because of the large variation in measured primary removal (Table 7), the correlation is not statistically significant. Also, the framework prediction does not account for potential biodegradation during primary treatment, and Simonich et al. measured significant removal of biodegradable, nonsorptive FMs during primary treatment (Table 7 and Fig. 6B) [11]. Figure 8B shows the regression of the measured percent overall removal from activated sludge treatment plants only measured by Simonich et al. with

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A

B

Fig. 8A, B Correlation of A measured primary removal with predicted primary removal and B measured overall removal with predicted overall removal for activated sludge plants, using the second tier of the framework model and accounting for sorption and biodegradation [11]. The error bars represent the standard deviation of the mean. The regression for overall removal (B) is significant at the 98% level, while the regression for primary removal (A) is not statistically significant; n=16. Dashed lines are the 95% confidence intervals for the regressions

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the percent overall removal predicted by the second tier of the framework (accounting for sorption and biodegradation) for activated sludge plants. This correlation is significant at the 98% level, however the framework significantly underpredicts overall removal (slope=0.118). This is particularly true for MK, MX, OTNE, AHTN, and HHCB, which are assumed in the framework to be nonbiodegradable (Table 1) and to have biodegradation rates of 0 h–1. This suggests that sorption alone does not account for the removals Simonich et al. measured for MK, MX, OTNE,AHTN, and HHCB and that biotransformation and/or volatilization may be playing a role in the removal of these FMs from activated sludge. Finally, these data show that the calculations outlined in the framework document for estimating FM concentrations in and removal from activated sludge wastewater treatment are predictive. Finally, Fig. 9 shows that the assumptions made in the framework document [1] are conservative for predicting final effluent concentrations, regardless of treatment type and geography (US and Europe). 6.2 Simple Treat Model Artola-Garicano et al. [27] compared their measured removals of AHTN and HHCB [24] to the predicted removal of these compounds by the wastewater treatment plant model Simple Treat 3.0. Simple Treat is a fugacity-based, ninebox model that breaks the treatment plant process into influent, primary settler, primary sludge, aeration tank, solid/liquid separator, effluent, and waste sludge and is a steady-state, nonequilibrium model [27]. The model inputs include information on the emission scenario of the FM, FM physical-chemical properties, and FM biodegradation rate in activated sludge. In general, the Simple Treat model predicted the overall removal of total AHTN and HHCB within a factor of 4 of Artola-Garicano et al.’s measured removals for three wastewater treatment plants located in The Netherlands. However, the free AHTN and HHCB concentrations predicted by Simple Treat were inversely related to the measured free concentrations of these compounds [27]. As previously mentioned, the overall removal of AHTN and HHCB measured by Artola-Garicano et al. (14–60%) was significantly less than the overall removal of these compounds measured by Simonich et al. [11] and Kanda et al. [22] (39–99.9%).

7 Conclusions In recent years, there has been significant interest in understanding the input of FMs to aquatic ecosystems and this has driven the substantial amount of research that has been conducted on the removal of FMs in wastewater treatment. Because FMs are semivolatile and have a wide range of physical-chemical properties and biodegradabilities, understanding their removal during the treatment process is

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complex. The mechanisms of FM removal from wastewater include biodegradation, sorption, and/or volatilization. A wide array of analytical methods have been developed to measure FMs in wastewater influent, primary effluent, final effluent, and solids, and wastewater studies have been conducted in the U.S. and Europe. Finally, the efficient removal of FMs during wastewater treatment is not only dependent on the biodegradability and physical-chemical properties of the FM, but is also highly dependent on plant operation and design.

A

B Fig. 9A–D Comparison of measured FM concentrations in all plants to predicted FM concentrations in the U.S. and Europe [1] for A influent, B primary effluent, C final effluent using the first tier (sorption only) of the framework, and D final effluent using the second tier (sorption and biodegradation) of the framework [11]. Points above the line have measured values greater than predicted by the framework model [1], while those below the line have measured values less than predicted. The filled circles represent concentrations in the U.S. and the open circles represent concentrations in Europe

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C

D Fig. 9C, D (continued)

References 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11.

Salvito DT, Senna RJ, Federle TW (2002) Environ Toxicol Chem 21:1301 Simonich SL, Begley WM, Debaere G, Eckhoff WS (2000) Environ Sci Technol 34:959 Balk F, Ford RA (1999) Toxicol Lett 111:57 Tas JW, Balk F, Ford RA, van de Plassche EJ (1997) Chemosphere 35:2973 Rimkus GG (1999) Toxicol Lett 111:37 Heberer T (2002) Acta Hydrochim Hydrobiol 30:227 Oros DR, Jarman WM, Lowe T, David N, Lowe S, Davis JA (2003) Mar Pollut Bull 46:1102 Buerge IJ, Buser HR, Muller MD, Poiger T (2003) Environ Sci Technol 37:5636 Peck AM, Hornbuckle KC (2004) Environ Sci Technol 38:367 Standley LJ, Kaplan LA, Smith D (2000) Environ Sci Technol 34:3124 Simonich SL, Federle TW, Eckhoff WS, Rottiers A, Webb S, Sabaliunas D, De Wolf W (2002) Environ Sci Technol 36:2839

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12. Federle TW (2004) 13. Artola-Garicano E, Borkent I, Damen K, Jager T, Vaes WHJ (2003) Environ Sci Technol 37:116 14. Osemwengie LI, Steinberg S (2001) J Chromatogr A 932:107 15. Herren D, Berset JD (2000) Chemosphere 40:565 16. Berset JD, Bigler P, Herren D (2000) Anal Chem 72:2124 17. Kupper T, Berset JD, Etter-Holzer R, Furrer R, Tarradellas J (2004) Chemosphere 54:1111 18. DiFrancesco AM, Chiu PC, Standley LJ, Allen HE, Salvito D (2004) Environ Sci Technol 38:194 19. Ricking M, Schwarzbauer J, Hellou J, Svenson A, Zitko V (2003) Mar Pollut Bull 46:410 20. Verbruggen EMJ, Van Loon W, Tonkes M, Van Duijn P, Seinen W, Hermens JLM (1999) Environ Sci Technol 33:801 21. Stevens JL, Northcott GL, Stern GA, Tomy GT, Jones KC (2003) Environ Sci Technol 37:462 22. Kanda R, Griffin P, James HA, Fothergill J (2003) J Environ Monit 5:823 23. Rimkus GG, Gatermann R, Huhnerfuss H (1999) Toxicol Lett 111:5 24. Artola-Garicano E, Borkent I, Hermens JLM, Vaes WHJ (2003) Environ Sci Technol 37:3111 25. Kallenborn R, Gatermann R, Planting S, Rimkus GG, Lund M, Schlabach M, Burkow IC (1999) J Chromatogr A 846:295 26. Aschmann SM, Arey J, Atkinson R, Simonich SL (2001) Environ Sci Technol 35:3595 27. Artola-Garicano E, Hermens JLM, Vaes WHJ (2003) Water Res 37:4377

The Handbook of Environmental Chemistry Vol. 5, Part O (2005): 119– 180 DOI 10.1007/b98609 © Springer-Verlag Berlin Heidelberg 2005

Immunochemical Determination of Industrial Emerging Pollutants M.-Carmen Estévez · Héctor Font · Mikaela Nichkova · J.-Pablo Salvador · Begoña Varela · Francisco Sánchez-Baeza · M.-Pilar Marco (✉) Department of Biological Organic Chemistry, IIQAB-CSIC, Jordi Girona 18–26, 08034 Barcelona, Spain [email protected]

1

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 122

2 2.1 2.1.1 2.1.1.1 2.1.1.2 2.1.1.3 2.1.2 2.1.3 2.1.4

Immunochemical Techniques . . . . . . . . . . . . . Antibody-Based Analytical Methods . . . . . . . . . Immunoassays . . . . . . . . . . . . . . . . . . . . . Enzyme-Linked Immunosorbent Assay (ELISA) . . . Enzyme-Multiplied Immunoassay Technique (EMIT) Polarization Fluoroimmunoassay (PFIA) . . . . . . . Flow-Injection Immunoassay (FIIA) . . . . . . . . . Immunosensors . . . . . . . . . . . . . . . . . . . . Immunoaffinity Chromatography (IAC) . . . . . . .

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Immunochemical Methods for Surfactants . . Anionic Surfactants . . . . . . . . . . . . . . . Nonionic Surfactants . . . . . . . . . . . . . . Cationic Surfactants . . . . . . . . . . . . . . .

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Immunochemical Methods for Polychlorinated and Polybrominated Compounds . . . . . . . . PCBs . . . . . . . . . . . . . . . . . . . . . . . PCDDs and PCDFs . . . . . . . . . . . . . . . . Chlorophenols . . . . . . . . . . . . . . . . . .

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Other Industrial Residues . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167 Bisphenol A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 167 Phthalate Esters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 170

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General Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 171

References

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Abstract A significant number of immunochemical methods have been described for the determination of the most important emerging pollutants. The present chapter is a compilation of the information available today regarding immunochemical determination of industrial residues with a high potential risk of causing negative effects in the environment, wildlife, and public health. Homogeneous immunoassays, ELISAs, FIIAs, immunosensors, and selective immunoaffinity sample treatment methods have been reported for the analysis of an important number of these substances. The bases of these methods are briefly presented.

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Immunochemical methods for anionic (LAS), nonionic (APEs and APs), and cationic surfactants (BDD12AC and DDAC) are extensively reviewed and the features of these assays discussed, particularly if examples of their application to environmental samples have been described. Similarly, a great amount of information has been collected regarding immunochemical determination of organochlorinated substances such as PCBs, PCDDs, PCDFs, and chlorophenols. On the contrary, immunochemical analysis of organobrominated substances, such as the BFR agents, seems to be still a goal. Immunochemical methods have also been reported for bisphenol A and phthalates showing excellent features. The commercial availability of some of these methods is also presented. Keywords Immunochemical techniques · Surfactants · Polyhalogenated compounds · Bisphenol A · Phthalate esters Abbreviations Ab Antibody ADMBAC Alkyldimethylbenzylammonium compounds AES Alkyl ether sulfates Ag Antigen AP Alkylphenol APEC Alkylphenol ethoxy carboxylate APE Alkylphenol polyethoxylate AS Alcohol sulfates ATMAC Alkyltrimethylammonium compounds BBP Butylbenzyl phthalate BDD12AC Benzyldimethyldodecylammonium chloride BFR Brominated flame retardants BMP-IA Bacterial magnetic particle-based immmunoassay BP Bromophenol BPA Bisphenol A BSA Bovine serum albumin CIA Capillary immunoassay CLIA Chemiluminescence immunoassay CP Chlorophenol CR Cross-reactivity DADMAC Dialkyldimethylammonium compounds DBP Dibutyl phthalate DCP Dichlorophenol DDAC Didecyldimethylammonium chloride DDT Dichlorodiphenyltrichloroethane DEHP Diethylhexyl phthalate DEQ Diesterquat DMSO Dimethylsulfoxide EDC Endocrine disrupter chemical EIA Enzyme immunoassay ELISA Enzyme-linked immunosorbent assay EMIT Enzyme-multiplied immunoassay technique EO Ethoxylene unit EPA Environmental Protection Agency EQ Esterquat FA Fatty acids FIA Fluorescence immunoassay

Immunochemical Determination of Industrial Emerging Pollutants FIIA FOH GC GC–MS HPLC HRP HTS IA IAC IC50 IgG LAS LC LC–MS LDS LIA LIC LIF LOD MAb NP NPE OP OPE PAb PBB PBDD PBDE PBDF PCB PCDD PCDE PCDF PCP PFIA QAC QFIA RAb RIA SAS SDS SPC STP TBBPA TBP TCDD TCP I-TEF I-TEQ TtCP WWTP

Flow-injection immunoassay Fatty alcohols Gas chromatography Gas chromatography–mass spectrometry High-performance liquid chromatography Horseradish peroxidase High-throughput screening Immunoassay Immunoaffinity chromatography Concentration at 50% of signal inhibition Immunoglobulin G Linear alkylbenzenesulfonates Liquid chromatography Liquid chromatography–mass spectrometry Linear 4-dodecylbenzenesulfonic acid sodium salt Liposome immunoaggregation Liposome immunocompetition Laser-induced fluorescence Limit of detection Monoclonal antibody Nonylphenol Nonylphenol polyethoxylate Octylphenol Octylphenol polyethoxylate Polyclonal antibody Polybrominated biphenyl Polybrominated dibenzo-p-dioxin Polybrominated diphenyl ethers Polybrominated dibenzofurans Polychlorinated biphenyl Polychlorinated dibenzo-p-dioxin Polychlorinated diphenyl ethers Polychlorinated dibenzofurans Pentachlorophenol Polarization fluoroimmunoassay Quaternary ammonium compounds Quenching fluorescence immunoassay Recombinant antibody Radioimmunoassay Secondary alkyl sulfonates Sodium dodecyl sulfate Sulfophenyl carboxylate Sewage treatment plant Tetrabromobisphenol A Tribromophenol Tetrachlorodibenzo-p-dioxin Trichlorophenol Toxic equivalent factor Toxic equivalent quotient Tetrachlorophenol Wastewater treatment plant

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1 Introduction Chemicals and secondary by-products from industry, household chemicals, personal care products, and pharmaceuticals such as drugs, antibiotics, and hormones, are some of the substances that have been grouped under the expression emerging pollutants, making reference to their recent increased use and release into the environment and to the fact that most of them were not previously considered as contaminants (see Table 1 for the most important groups of emerging pollutants). The adverse effects derived from a continuing exposure to these substances are often unknown, and regulations are still not well established [1, 2]. Much interest has been focused on these compounds not only because of their possible adverse effects, but also for the great amounts that are produced worldwide. Most of the substances considered as emerging contaminants are widespread in everyday life and applied in differTable 1 Some of the most important emerging pollutants divided into chemicals with an industrial origin and pharmaceuticals

Industrial chemicals

Pharmaceuticals

Surfactants and their metabolites Nonionic surfactants Anionic surfactants Cationic surfactants

Antibiotics Fluoroquinolones Sulfamides Penicillins Tetracyclines Macrolides

Organochlorinated substances Polychlorinated biphenyls Chlorophenols Dioxins Organobrominated substances Polybrominated biphenyl ethers Bromophenols Dioxin-like compounds Tetrabromobisphenol A Industrial additives and others Phthalate esters Bisphenol A

Steroid hormones Estrogens Androgens Gestagens Corticosteroids Analgesics Paracetamol Aspirin Ibuprofen Diclofenac Tranquilizers (psychiatric drugs) Diazepam Cytostatic agents Methotrexate Cyclophosphamide Ifosfamide

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ent fields such as pharmaceuticals (for both animal and human use), drugs, hormones, and surfactants. An important number of these substances have an industrial origin. Some of them, like the pesticides, arrive intentionally in the environment and their use and release should be theoretically controlled. However, many of them have not been purposely produced as bioactive substances but more as components or additives of certain materials. Their significant growth in the chemical industry has not only been produced as a consequence of the discovery of new active principles in the pharmaceutical or pesticide area, but also because of the expansion of new technologies (electronics, containers, textiles, plastics, resins, foams, etc.), that require the development of new materials and substances with particular features. Most of these substances enter or are discharged to water and air sources without regulated controls. Wastewater treatment plants (WWTPs) are often not yet adapted to completely remove them, and therefore these new compounds can be found to some extent in wastewater effluents as well as in soil and sludge. The release into the environment of this large amount of chemicals has become an increasing concern for the authorities and for the scientific community. Although there is positive pressure by governmental bodies and agencies to push chemical industries toward the development of more environmentally friendly compounds that are not persistent and are easily biodegradable, the final metabolites readily formed can be more ubiquitously distributed and/or present more toxicological effects than the parent compounds. That is the case for alkylphenol polyethoxylates, a major group of nonionic surfactants used worldwide, whose major breakdown products are the alkylphenols, which are considered relevant endocrine disrupters, especially nonylphenol. The same effects have been reported in the case of certain plastic and polymer additives such as bisphenol A or some phthalate esters (see BKH report [3]). Other kinds of substances are not produced intentionally but are generated as by-products of industrial processes such as combustion or waste incineration. This is the case for the dioxins or the PCBs. Other polyhalogenated compounds are also of concern because of their persistence in the environment. This is the case for the chlorophenols, used for many years as insecticides, and wood and textile preservatives. Their use is today restricted in most of the developed countries but residues can still be detected in many environmental compartments. Organobrominated substances, used as flame retardant additives, have emerged as a new generation of polyhalogenated substances whose environmental and toxicological impacts are still not completely determined, although some evidence suggests an endocrine disrupter action. Some of the substances considered in this chapter do not have toxicity by themselves, but may affect the permeability or solubility of other pollutants present in the environment. This is the case for the anionic surfactants such as LAS, whose presence in the environment is also a worry due to their high production and use. The continuous development of more specific and sensitive analytical techniques has allowed the detection of traces of these substances in many com-

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partments of the environment. Recently, the US Geological Survey reported the development of five new analytical methods for the detection of 95 organic wastewater contaminants, many of them considered as emerging pollutants, such as several antibiotics, drugs (both prescription and nonprescription), steroid hormones, phthalates, and nonionic surfactants [4]. As fast as the risk assessment is being carried out and a regulation is trying to be established for these substances, more and more rapid, sensitive, and specific methodologies capable of detecting this huge amount of compounds are continuously demanded. Immunochemical techniques can fulfill all these requirements because of their specificity, selectivity, and demonstrated high sample throughput capabilities. The objective of this chapter has been to collect and to provide information on the immunochemical techniques available today for the determination of an important number of the emerging pollutants of industrial origin. The emerging pollutants considered here have been selected by attending to their toxicological risk, environmental relevance, or their regular use or production (their generic structure and some reported data regarding their environmental levels as well as their more probable biodegradation pathways are briefly detailed in Table 2). For many of these substances there are immunochemical methods available that have been applied with significant success to the analysis of environmental samples. For some of these compounds, different antibodies (both monoclonal and polyclonal) have been produced and manufactured by different companies either as immunochemical reagents or as immunochemical assay kits. The availability of antibodies, as the key reagents of any immunochemical technique, opens the possibility of developing a wide variety of methods, such as those described before, depending on the necessities. However, in most cases the methods commercially available are ELISA kits in a variety of formats and supports such as magnetic particles, microtiter plates, test strips, tubes, etc. (see Table 3 for examples of commercially available immunochemical techniques).

2 Immunochemical Techniques The key component of immunochemical techniques is the antibody (Ab). Antibodies are globular proteins generated by the immune system as a defense against foreign agents (antigens,Ag). Their structure varies depending on their isotypes. There are five different families of immunoglobulins (IgG, IgM, IgA, IgD, and IgE) differing in their charge, size, amino acid sequence, and carbohydrates attached. The most abundant class in mammal serum and the most used in immunochemical applications is the IgG subclass. The IgGs (MW 150 kDa) are composed of four polypeptide chains: two identical “heavy” (H) chains that carry covalently attached oligosaccharide groups, and two identical, nonglycosylated, “light” (L) chains. The heavy chains are interlinked by disulfide bonds, and each light chain is joined by a disulfide bond to a heavy

Chemical Structure

R=C8, C9

R=C8, C9 n=1–40 Major component in the production of APEs and breakdown product in their degradation. Plasticizers and stabilizers in plastics

Household and commercial detergents Emulsifiers Textile and leather industry Pharmaceutical and personal care products (PPCPs)

Uses/Origin

– Sewage effluents: 0.025–330 mg L–1 (NP) and 0.0022–73 mg L–1 (OP). Usually below 10 mg L–1 [9] – Found in air [10] and sediments (up to 14000 mg Kg–1) [5, 11] – SW: Usually NP levels below 1 mg L–1. High levels in Spain (0.15–644 mg L–1) [12] and England rivers (0.2–180 mg L–1) [13]. – OP levels below 0.1 mg L–1 [5]

– STPs effluents in Switzerland: NP1EO (30–65 mg L–1), NP2EO (47–77 mg L–1) and in Japan NP1EO (0.21–2.96 mg L–1) [5] – SW: usually below 1 mg L–1 and maximum peak values around 20 mg L–1[5]

Environmental Occurrence

More persistent than APE. Half-life (river waters): 30–58 days (NP) and 7–50 days (OP) [5]

Readily biodegradable in WWTPs under aerobic and anaerobic conditions [6] to short chain alkylphenol ethoxylates and alkylphenols Half-life=1 to 4 weeks [7, 8]

Environmental Fate

DW: drinking water; GW: ground water; SW: surface water; WW: waste water; STP: sewage treatment plant; WWTP: wastewater treatment plant.

NP (Nonylphenol) OP (Octylphenol)

Alkylphenols (APs)

NPE (nonylphenol ethoxylates) OPE (octylphenol ethoxyates)

Alkylphenol Ethoxylates (APEs)

Nonionic Surfactant

Substance

Table 2 Summarized table with the more important compounds included in the surfactants, polyhalogenated compounds, and other industrial residues. Their generic chemical structure and the use or origin are shown. Some reported data regarding their environmental occurrence and the more probably environmental fate are also given

Immunochemical Determination of Industrial Emerging Pollutants 125

Chemical Structure

R=C12–C18

Fatty Amides Polyethoxylates

R=C12–C18

Fatty Esters Polyethoxylates

R=C9–C15 n=4–14

Fatty Alcohol Ethoxylates (AE)

Substance

Table 2 (continued)

Foam stabilization Emulsifier Solubilizer, antistatic, wetting agent in PPCPs Hair shampoo, liquid soaps, shaving creams and other PPCPs

Emulsifiers Textile and leather industries Component in PPCPs

Household and laundry. Pulp and paper manufacturing Textile dyeing Emulsifiers, spray adjuvants

Uses/Origin

No data found

No data found

– In GW levels: (61–189 ng L–1 for the different AEOs C12EO3–9 – Total conc. of 710 ng L–1 [14] – Soil interstitial water: (48–73 ng L–1 for C12EO3–5 (total conc. in the deeper layers at 194 ng L–1) [14] – Treated sewage 6.5 mg L–1 [15], 12.5–300 mg L–1 [16] – Sewage sludge 10–190 mg Kg–1 [17]

Environmental Occurrence

Contradictory data related to they biodegradability [19]. Readily biodegradable [20, 21]

Easily biodegradable (slower when it has more than 50 units of ethylene oxide units [20]

Readily biodegradable (>80% in 28 days for linear AE and 40% for branched AE) [18]. Slower degradation (AE>20 ethoxylene units) [19]

Environmental Fate

126 M.-C. Estévez et al.

Chemical Structure

m+n=C3–C11

Sulfophenyl Carboxylates (SPCs)

m+n=C10–C14

Anionic Surfactants Linear Alkylbenzene Sulfonates (LAS)

Substance

Table 2 (continued)

Major metabolite in LAS biodegradation

Household and commercial detergents

Uses/Origin

– SW and sewage effluents: 0.5–3.2 mg L–1 [25] – Seawater: [26] – Drinking water: 1.6–3.7 mg L–1 [23] – Raw river waters: 1.8–5 mg L–1 [27] – Up and downstream of WWTPs: <1–101 mg L–1 [28]

– STPs effluents: 19–71 mg L–1 [15] – WW effluents levels: 0.09–0.9 mg L–1 [18] – WW sludge: <3 mg g–1 [18] – SW in North Sea (<0.05–9.4 mg L–1) [22] – SW in Brazil (14–155 mg L–1 [23] – SW in Philippines (2.2–102 mg L–1) [24]

Environmental Occurrence

Found mainly in aquatic compartments. Shorter alkyl chain SPCs (≤C5) are expected to be found as the distance from the discharge point of LAS increases

Readily degradable with a half-life of 1–87 days. 10–35% adsorbed in the particulate are matter [18]

Environmental Fate

Immunochemical Determination of Industrial Emerging Pollutants 127

Alkyl Ether Sulfates (AES)

Alkyl Sulfates (AS)

Secondary Alkyl Sulfonates (SAS)

Primary Alkyl sulfonates

Alkyl Sulfonates

Substance

Table 2 (continued)

R=C10–C14 m=1–4

R=C12–C18

R + R1=C12–C18

R=C11–C17

Chemical Structure

Liquid bath soaps, hair shampoos, and mechanical dishwashing agents. Ingredient in industrial cleaning agents

Laundry detergents Wool-washing agents, soap bars and liquid bath soaps, hair shampoos, and toothpastes

Liquid detergents (dish washing agents, cleaning agents, and hair shampoos). Commercial products are almost exclusively composed of SAS

Uses/Origin

– Effluents of seven representative STPs: C12–15 AES: 3 and 12 mg L–1 [15]

– STPs effluents: C12–15 AS between 1.2 and 12 mg L–1 [15]

No data found

Environmental Occurrence

Readily biodegradable in WWTPs under both aerobic and anaerobic conditions [18, 29]

Fast biodegradation under aerobic and anaerobic conditions. Effective removal in WWTPs [18]

Can be adsorbed onto sludge but under aerobic conditions are readily biodegradable (primary degradation in WWTP <90% in 3 days [18]). Not degraded in anoxic conditions

Environmental Fate

128 M.-C. Estévez et al.

Chemical Structure

R¢=Methyl or benzyl X=Cl or Br (or Methyl sulfate)

Fabric softener Emulsifying agents Biocides, disinfectants

Uses/Origin

m+n =1–10

Polychlorinated biphenyls (PCBs)

Organochlorinated Substances

R=C16–C18

Insulating fluid in electrical equipment and as hydraulic fluids

Fabric softener

Quaternary Carboxyalkyl Ammonium Compounds (Esterquats)

ATMAC: Alkyltrimethyl-ammonium compounds DADMAC: Dialkyldimethyl-ammonium compounds ADMBAC: Alkyldimethylbenzylammonium compounds

Quaternary Ammonium Compounds (QACs)

Cationic Surfactants

Substance

Table 2 (continued)

– SW (up to 100–500 ng L–1) [33, 34]

No data found

Values found for ditallowdimethylammonium chloride (DTDMAC): – STPs-influents: 375–4300 mg L–1 – In SW of rivers: 2–34 mg L–1 – Effluent of STPs: 11–55 mg L–1 (reviewed in [30]) – WWTPs influents: 340–480 mg L–1 (US); ≈ 1000 mg L–1 [31]

Environmental Occurrence

Very persistent in the environment. Low levels in water and air. High levels in soils, sediments and animals

Easily biodegraded under aerobic conditions. It’s also assumed their degradation in anoxic conditions [19, 32]

Highly adsorbed onto particulate matter [18]. Short half-life under aerobic conditions. Poorly anaerobically biodegraded [19]

Environmental Fate

Immunochemical Determination of Industrial Emerging Pollutants 129

Chlorophenols

Polychlorinated dibenzofuran (PCDF)

Polychlorinated dibenzop-dioxin (PCDD)

Polychlorinated diphenylethers (PCDE)

Dioxin Like Substances

Substance

Table 2 (continued)

m=1–5

m+n=1–8

m+n=1–8

m+n =1–10

Chemical Structure

Preservatives agents, pesticides

Are not intentionally produced. They are formed as byproducts in several processes

Uses/Origin

– GW near sawmills or waste sites from (0.03 to 91.3 mg L–1) [41] – Drinking water: 0.03–0.7 mg L–1 [42–44] – SW (several countries): 1.6–26.6 mg L–1 [45, 46] – Sediments (Canada) 25–10,000 mg Kg–1 [47]

– Sediments from polluted areas: >20 pg I-TEQ g–1 dw [37] – Effluents from waste incinerator plants (Japan): 3.3–120,000 pg L–1. (73.5 TEQ) (reviewed in [38]) – Soil: 0.1–1080 pg I-TEQ g–1 and sediments: 0.42–8 pg I-TEQ g–1 (Spain) [39, 40]

– Marine sediments (Finland): <1 pg I-TEQ g–1 dw [35, 36]

Environmental Occurrence

Most of them go into water. They can be transformed by photolysis into dioxins [48, 49]

Very persistent in the environment. Very low levels are found in water and air. High levels in soils, sediments and animals. Occurrence in the environment related to chlorination degree obtained by photolysis: >chlorination degree> persistence

Environmental Fate

130 M.-C. Estévez et al.

Chemical Structure

Polybrominated diphenylethers (PBDE)

Dioxin Like Substances

m+n =1–10

m+n=1–10

Organobrominated Substances Polybrominated biphenyls (PBB)

Substance

Table 2 (continued)

Flame retardant

Flame retardant

Uses/Origin

– Sewage sludge: 11–28 ng g–1 (Br10) [55] – Water (Japan) in 1988: <0.1 mg L–1 [38] and sediments (Japan) in 1996 <25–580 ng g–1 (Br10) [38] – SW: 0.158 ng L–1; sewage sludge: 2290–4890 ng g–1; sediments: 132 ng g–1 (reviewed in [56]) – Soil (US) <0.1–31.6 ng g–1 [57]

– Water Pine river (USA): 0.01–3.2 mg L–1 [50, 51] – Sediments: 0.33–0.84 (dw) mg Kg–1 [52] – Water: <0.05 mg L–1 and sediments: <8 ng g–1 (Japan) in 1989 [38]

Environmental Occurrence

Compounds highly brominated attach strongly to sediments. They are slowly degraded in the environment

Compounds highly brominated attach strongly to sediments. They are slowly degraded in the environment [53, 54]

Environmental Fate

Immunochemical Determination of Industrial Emerging Pollutants 131

m=1–5

m+n=1+8

m+n=1–8

Chemical Structure

Tetrabromobisphenol A (TBBPA)

Bromophenols

Polybrominated dibenzofuran (PBDF)

Polybrominated dibenzo-p-dioxin (PBDD)

Dioxin Like Substances

Substance

Table 2 (continued)

Flame Retardant

Flame retardant Wood preservative

Are not intentionally produced. They are formed as byproducts in several processes

Uses/Origin

– Sewage sludge (2.9–76 ng g–1 dw) [55] – Sediments (Japan) 1988: 2–108 ng g–1 [38]

– Raw and treated water from potable water treatment plants (0.6 and 1.3 ng L–1) [61] – Sediments (Japan)1986: <0.5–4 ng g–1 [38]

– Sewage sludge from municipal WWTPs (Germany): 0.29–3.05 mg Kg–1 (Br1–Br5) [58] – Sediments (Japan): 0.03–0.37 ng g–1 (Br4–Br6) [38] – River and marine sediments: 0.03–0.37 mg Kg–1 (Br4–Br6) [60] – Sediments 1.98–17.4 ng g–1 [38]

Environmental Occurrence

TBBPA can be metabolized by microbes and broken down by photolysis [64]

They can be dehalogenated in marine and river sediments [62, 63]

Very persistent in the environment Faster degradation than PCDDs [59]. Very low levels are found in water and air. High levels in soils, sediments and animals. Occurrence in the environment related to bromination degree obtained by photolysis: >bromination degree> persistence

Environmental Fate

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Production of resins (polycarbonate and epoxy resins). Component in flame retardant production Antioxidant, preservative

Bisphenol A

Uses/Origin

Plasticizers in PVC production. Component in the manufacture of cosmetics, inks, and adhesives

Chemical Structure

Di-n-butyl phthalate (DBP) Diethyl phthalate (DEP) Butyl benzyl phthalate (BBP) Di(2-ethylhexyl) phthalate (DEHP)

Phthalate Esters

Others

Substance

Table 2 (continued)

– River water mean values: 0.016 mg L–1 (Europe) and 0.5 mg L–1 (US) [66]. – SW: <0.001–1 mg L–1 [9] – WW effluents mean values: 1.5 mg L–1 [67]

– SW levels are near to 10 mg L–1; in rivers, between 0.5–1 mg L–1 and in sea water between 0.005–0.7 mg L–1 – US streams: 2.5 mg L–1 (DEHP) and 0.25 mg L–1 (DEP) [4]

Environmental Occurrence

Not persistent in surface water. Rapidly biodegraded in aquatic environments [68] and removed in WWTP. Half-life: 1–4 days [69] in water. Accumulated in anoxic sediments [9]

Fast biodegradation under aerobic conditions. Half-life in water: 1–15 days Half-life in soils: 7 days – several months [65]

Environmental Fate

Immunochemical Determination of Industrial Emerging Pollutants 133

Immunochemical kit

ELISA kit ELISA kit ELISA kit

ELISA kit

PCBs

PCB RISc soil test kit PCB RISC liquid waste test system PCB immunoassay kit EnviroGard PCB in soil PCB in soil (tube assay) D TECH® PCB test kit DELFIA PCB (soil and food) PCB RaPID Assay® PCBs ELISA kit, 100T (magnetic particle)

Organochlorinated substances

LAS

Anionic Surfactants

APEs AE AP

Nonionic Surfactants

Analyte

EnSys, Inc. (Strategic Diagnostic Inc.) EnSys, Inc. (Strategic Diagnostic Inc.) Hach Millipore Inc. (Strategic Diagnostic Inc.) EnviroLogix Strategic Diagnostic Inc. Hybrizyme Ohmicron Corp. (Strategic Diagnostics Inc.) Takeda Chemical Industries L-EC

Japan EnviroChemicals, Ltd.

Japan EnviroChemicals, Ltd. Takeda Chemical Industries L-EC Japan EnviroChemicals, Ltd.

Supplier/manufacturer

http://www.sdix.com/ http://www.sdix.com/ http://www.hach.com/ http://www.sdix.com/ http://www.envirologix.com/ http://www.sdix.com/ http://www.hybrizyme.com/ http://www.sdix.com/ http://www.takeda.co.jp/index-e.html

http://www.jechem.co.jp/eco/

http://www.jechem.co.jp/eco/ http://www.takeda.co.jp/index-e.html http://www.jechem.co.jp/eco/

Contact

Table 3 Some representative commercial immunochemical assay kits for the most important emerging pollutants with an industrial origin. The supplier and the contact web page are also listed

134 M.-C. Estévez et al.

Immunochemical kit

Bisphenol A

ELISA kit ELISA kit

EnSys Dioxin High-performance dioxin/furan immunoassay kit insert (IN-DF1) DELFIA TCDD test kit

Dioxins

Others

PENTA RISc® EnviroGardTMin soil D TECH® PCP Test PCP RaPID Assay®

PCP

Organochlorinated substances

Analyte

Table 3 (continued)

http://www.jechem.co.jp/eco/ http://www.jechem.co.jp/eco/

http://www.hybrizyme.com/

Hybrizyme Japan EnviroChemicals, Ltd. Japan EnviroChemicals, Ltd.

http://www.sdix.com/ http://www.cape-tech.com/

http://www.sdix.com/ http://www.sdix.com/ http://www.sdix.com/ http://www.sdix.com/

Contact

EnSys Inc. (Strategic Diagnostics Inc.) CAPE Technologies

EnSys Inc. (Strategic Diagnostics Inc.) Millipore (Strategic Diagnostics Inc) Strategic Diagnostics Inc. Ohmicron Corp. (Strategic Diagnostics Inc.)

Supplier/manufacturer

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chain. All four chains contain defined constant and variable, locally hypervariable, regions within their amino acid sequence (see Fig. 1). The Fc fragment is the constant (crystallized) region and it is involved in the immune regulation, whereas the Fab (antibody binding fraction) fragment is the region that contains the variable fraction (Fv) with the specific binding sites that allow interaction with the Ag. Since the publication of the first assay based on the use of antibodies to determine insulin [70], development in this area has undergone rapid growth. The application of immunochemical methods to detect substances not only in the clinical field, but also in food and environmental areas has led to the necessity to obtain more and more specific antibodies in considerable amounts and at low cost. The important progress made during recent years in the molecular biology and genetic engineering fields has favored this fact. Currently, we can speak of three different techniques to obtain antibodies, yielding what we know as polyclonal (PAb), monoclonal (MAb), and recombinant (RAb) antibodies. In principle it is possible to obtain antibodies for any kind of substance. In the case of small molecules (i.e., molecular weight less than 1,000 Da), the design and synthesis of an appropriate immunizing hapten followed by its covalent attachment to a carrier molecule [71] has been until now unavoidable; however, knowledge obtained while engineering new antibody molecules may reduce the effort necessary in this aspect. Polyclonal antibodies are obtained directly from the serum of the immunized animals (sometimes a purification step is carried out before their use). A family of clones is obtained that recognize the global structure of the hapten immunized, exhibiting each clone to a specific binding to different epitopes in the molecule. Therefore, the affinity of a PAb will be a combination of the

Fig. 1 Scheme showing the basic H2L2 structure of the immunoglobulins of type G (IgG). It is formed by two pairs of polypeptide chains interlinked by disulfide bonds. The Fc fragment is the constant (crystallized) region and it is involved in the immune regulation, whereas the Fab (antibody binding fraction) fragment is the region that contains the variable fraction (Fv) with the specific binding sites that allow the interaction with the Ag. Fragments obtained after papain or pepsin digestion are also shown

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activity of each clone for the target analyte. The host animal is usually a rabbit, but when great amounts of serum are required the use of goats, pigs, or sheep has been described. The process to finally obtain PAbs is simple once you have the immunizing hapten but because of the animal variability, a lack of reproducibility can be found from animal to animal. This fact can be a problem when a constant supply of identical antisera is required. Monoclonal antibodies are produced by the fusion of antibody-producing spleen cells from an immunized animal (mouse) with mutant tumor cells derived from myelomas [72]. In contrast with PAbs, a unique IgG molecule is obtained from a single cell clone and theoretically this technology provides an unlimited source of the antibody with identical affinity for the antigen, as long as the hybridoma line is stable. However, the screening process to isolate the desired clone is usually tedious and time-consuming, the cost of production of MAbs is higher than for PAbs, and sometimes they have lower affinities to small molecules than PAbs. Whereas in both PAbs and MAbs the specificity and affinity of the final antibody will be a consequence of the immunizing hapten chosen and the immunization protocol, in recombinant antibody phage display technology these problems can be in part solved by the generation of a variety of Ab fragments mimicking the immune response in vitro. The whole process involves the following steps: (a) the preparation of Ab encoding libraries derived from different methods (usually by isolation of mRNA from hybridoma, spleen cells, or lymphocytes of immunized mice); (b) cloning of the genes in a bacterial plasmid vector; (c) expression in bacteria (E. coli) and coinfection with helper bacteriophage virus, displaying Ab fragments on its surface as a fusion with normally occurring coat protein; and (d) screening for antigen specificity and antigen-driven selection [73, 74]. The ability of this methodology to design the antibody polypeptidic structure, as well as to modify the existing fragments, can allow one to improve the affinity of the antibodies or even to change their selectivity. Although this technology was developed initially for therapeutic purposes, RAb fragments have also been used in environmental analysis. Recombinant antibodies have been produced for insecticides like parathion [75], for dioxins [76], and for pesticides like triazines [77, 78], although at present they have not achieved the affinity levels of the corresponding MAbs or PAbs. 2.1 Antibody-Based Analytical Methods Immunochemical techniques are based on the immunological reaction derived from the binding of the antibody to the corresponding antigen. This reaction is reversible and is stabilized by electrostatic forces, hydrogen bonds, and Van der Waals interactions. The formed complex has an affinity constant (ka) that can achieve values around the order of 1010 M–1. This great affinity and specificity between the specific antibody and the antigen (or the analyte) have turned these techniques into powerful analytical tools to detect and quantify

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substances at low concentrations and trace levels. During recent decades many efforts have been made and important advances have been achieved in this field. The advances made in microelectronics combined with the availability of antibodies for a great variety of substances such as proteins, macromolecules, or low molecular weight molecules such as drugs, metabolites, or environmental, agricultural and food pollutants, have been crucial for this progress. Despite the robustness and good detectability achieved nowadays by the chromatographic analytical methods when coupled to sensitive detectors, they often require expert personnel and expensive equipment. Moreover, preconcentration and cleanup procedures to remove potential interferences prior to the analysis are usually mandatory. All these factors lead to an increase of the final cost of these methodologies and the analysis time. Alternatively, immunochemical techniques are simple, fast, and very specific and sensitive. Nowadays automation and the possibility of development of high-throughput screening (HTS) have been demonstrated. Overall we can say that they constitute excellent tools to be exploited in monitoring programs where a great number of samples need to be analyzed. One of their drawbacks is often the fact that matrix effects should be carefully evaluated beforehand since, in contrast to other analytical techniques, specific and nonspecific signals are not so easy to distinguish leading to overestimation or to false positives. Contrariwise, false negatives are very seldom seen in these techniques. Thus, as effective screening techniques, immunochemical methods are complementary to the standard analytical techniques. Several immunochemical techniques have been developed as analytical tools or in sample treatment methods to separate an analyte from complex matrices. Some of the most important or more frequently used are described below. 2.1.1 Immunoassays Nowadays, immunoassay (IA) is the most extensive immunochemical methodology.A great number of IAs have been developed for the detection of pollutants at trace levels [71, 79–81], such as pesticides and other kinds of industrial residues, not only in different environmental compartments (water, soils, sediments, etc.) but also in food and biological matrices. Many of those developed for pesticides are commercially available and the US EPA (US Environmental Protection Agency) has validated and included some of them in the SW-846 method list [82].Wide application has also been found in pharmaceutical, veterinary, and forensic analysis as well as, more recently, in human exposure assessment to a variety of industrial chemicals or contaminants such as polyaromatic hydrocarbons (PAHs), polychlorinated biphenyls, (PCBs), or pesticides [83–85]. For human biomonitoring, where large a number of samples are often analyzed, the possibility of adapting immunoassays to HTS makes them particularly suited to field studies or large-scale monitoring. In addition, the ability to recognize not only the target analyte but also other structurally related compounds (Ab cross-

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reactivity) may allow both parent molecules and their metabolites to be detected simultaneously. In this context, the EPA encourages the development of immunochemical techniques for human exposure monitoring [84, 85]. In immunoassays the reaction Ab–Ag is quantified by means of labels, under competitive conditions. The general procedure involves a competition step between a fixed concentration of a labeled derivative and the free analyte for a limited amount (low concentration) of Ab. The amount of labeled Ag can then be measured and therefore the amount of free Ag. Several kinds of markers can be used as labels. The first immunoassay was based on the use of radioisotopes (radioimmunoassays, RIA) [86], but they have been replaced by more environmentally friendly and less hazardous substances. By using fluorescent (fluorescein, rhodamine, etc.), chemiluminescent (i.e., luminol), or bioluminescent markers, techniques such as fluoroimmunoassay (FIA) and chemiluminescence immunoassay (CLIA) have been developed, although in FIAs, for instance, the sensitivity achieved is in many cases not as good as expected, sometimes because fluorophores are exposed to many interferences that can lead to quenching of the signal. The use of enzyme labels (EIA, enzyme immunoassay) offers the possibility of increasing detectability, due to the amplification produced depending on the enzyme turnover, and the option of using a variety of substrates producing colored, fluorescent, or chemiluminescent products. Nowadays enzymes such as horseradish peroxidase (HRP), alkaline phosphatase (AP), and glucose oxidase (GO) are the most frequently used labels in immunoassay. Immunoassays can be performed in solution (homogeneous format) or by immobilization of one of the immunoreagents on a solid support (heterogeneous format). The solid support can be tubes, nitrocellulose paper, magnetic particles, microspheres, polystyrene plates, etc. The most used supports are microtiter plates where up to 96 (or in some cases up to 384) samples can be processed simultaneously, making use of very small sample volumes. In heterogeneous assays a separation between the bound and the free phases is required, whereas in the homogeneous one, the detection step is carried out in solution, with both fractions (bound and free) in the immunoreagent mixture. Homogeneous assays are faster, simpler, and can be easily adapted to the available automated analyzers often used in clinical chemistry. However, they are often less sensitive and are more exposed to matrix interferences, so washing steps to help remove these interferences must be performed. The common aspect of all these assays when applied to the analysis of small organic molecules is the fact that the assay takes place under competitive configurations, as we will see below. In contrast, for the determination of large substances, this is not always necessary. 2.1.1.1 Enzyme-Linked Immunosorbent Assay (ELISA) Among the heterogeneous assays, ELISA is the most common and frequently used for environmental monitoring. Examples of its wide applicability can be found in recent reviews [71, 79]. The most usual configurations for the analy-

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a

Direct Competitive ELISA

b

Indirect Competitive ELISA

Fig. 2a, b Scheme of the two ELISA formats most frequently used for the analysis of low molecular weight analytes. a Direct competitive ELISA. The Ab is coated on the surface and a competition is established between the analyte and the enzyme tracer. After washing, the addition of a substrate produces a chromogen product that is easily quantified. b Indirect competitive ELISA.A coating antigen is immobilized on the solid support and the specific IgG and the analyte are in solution. After removal of unbound reagents, a secondary IgG labeled with the enzyme (IgG-enzyme), which specifically recognizes the Ab, is added. After another washing step the amount bound is also quantified by the addition of the substrate solution

sis of small molecules are shown in Fig. 2. In the direct format (see Fig. 2a), the Ab is coated onto the solid support (usually a microtiter plate) and an equilibrium is established between the Ab, the free analyte, and the enzyme tracer (both of them in solution).After a washing step, where all the unbound reagents are removed, the amount of label bound to the Ab is measured, the signal being inversely proportional to the amount of analyte in the sample. A direct assay can also be performed by immobilizing an analog of the analyte (coating antigen) on the plate and performing the competition step with the free analyte for the labeled Ab in solution. In the indirect format (see Fig. 2b), the coating antigen is coated on the plate, but in this case the amount of analyte present in the sample is indirectly measured by measuring the bound Ab with a second Ab that is conveniently labeled (AntiIgG-enzyme). Although this format has a step more, it has often proved to be more robust. 2.1.1.2 Enzyme-Multiplied Immunoassay Technique (EMIT) EMIT is one of the most common EIAs working under homogeneous conditions [87]. The principle is the competition for the specific antibody between the analyte and an analog labeled with a particular enzyme (usually glucose6-phosphate dehydrogenase, G6P-DH) such that the enzyme activity decreases upon binding of the labeled antigen to the antibody. In this format the analyte

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concentration is directly proportional to the enzyme activity measured. This kind of assay has been widely applied in the clinical analysis field for the determination of drugs of abuse in several biological matrices such as urine, blood, or tissues [88–92], but some examples related to the analysis of pesticides have also been reported [93]. 2.1.1.3 Polarization Fluoroimmunoassay (PFIA) PFIA also works under homogeneous conditions and makes use of fluorescent labels (usually fluorescein). The principle is the excitation of the sample with plane polarized light. The free labeled antigen rotates rapidly, emitting light in many different planes, resulting in a decrease in the intensity of vertical polarized light. But when it binds to a large molecule like the antibody the rotation is slower, leading to an increase in the emitted light measured. In the absence of the analyte, the light measured is thus very small since a great part of the labeled antigen will be bound to the antibody. The presence of the analyte is thus evidenced in a direct manner by the increase of the light measured. As in the case of EMIT this technique has also been used in the clinical area, sometimes comparing precisely with EMIT in drug analysis, usually as screening methodology [94–96], but also for environmental pollutants such as pesticides [97–99]. 2.1.2 Flow-Injection Immunoassay (FIIA) In recent years the combination of immunochemical methodologies with automated devices has grown in importance. In this context, flow-injection systems coupled to immunoassays (FIIA) [100–102] offer rapid analysis of a great number of samples, providing rapid results and good levels of sensitivity, and allowing continuous monitoring. In these devices, a small immunoreactor continuously receives buffer or reagents through different valves. The technique can work in homogeneous (usually with fluorophores as label) or heterogeneous conditions (usually with the Ab immobilized). The more generic FIIA works as is shown in Fig. 3. By means of a buffer valve all the solutions are continuously flowing through the system and the immunoreagents (Ag and labeled Ag, both together or in sequential steps) are injected in the system. The amount of labeled Ag is detected downstream and quantified. Subsequently, the system can be regenerated by passing buffer solutions. Several FIIAs have been developed for environmental pollutants such as pesticides and also industrial chemicals, as will be discussed in following sections [100, 103–106].

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Fig. 3 Generic FIIA system. A heterogeneous format is shown. The antibodies are immobilized in the immunoreactor. The analyte and the labeled Ag (in this case with an enzyme) are passed through the system and the competition step takes place. The flow of the substrate solution through the system allows the determination of the amount of bound labeled Ag, which is then detected and measured

2.1.3 Immunosensors In recent years many efforts have been made to develop immunochemical techniques integrating the recognition elements and the detection components, in order to obtain small devices with the ability to carry out direct, selective, and continuous measurements of one or several analytes present in the sample. In this context biosensors can fulfill these requirements. Biosensors are analytical devices consisting of a biological component (enzyme, receptor, DNA, cell, Ab, etc.) in intimate contact with a physical transducer that converts the biorecognition process into a measurable signal (electrical or optical) (see Fig. 4). In

Fig. 4 Schematic representation of a generic biosensor with the essential components (biorecognition element, transducer, and electronic part involved in data processing and display)

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immunosensors the Ab or the Ag is immobilized on the surface of the transducer and the formation of the immunocomplex is detected by the transducer. Although there are many approaches for direct immunosensors (the changes generated when the antibody binds the analyte), in the case of small organic molecules the detection usually takes place under competitive configurations with or without the use of labels. Several transducer principles have been described. Electrochemical immunosensors are based on the use of amperometric, potentiometric, conductimetric, or impedimetric transducers. Optical immunosensors make use of optical fibers, the evanescent wave, or the surface plasmon resonance principle, among others. Piezoelectric immunosensors are based on the shift of the resonance frequency of piezoelectric crystals produced after formation of the immunocomplex. Finally some examples can also be found of thermometric immunosensors, where the heat of the reaction produced as a consequence of the immunoreaction, usually coupled to an enzyme amplification system, is detected. When the use of labels is required, the nature of these depends on the transducer principle of the immunosensor. Thus, the use of electroactive, fluorescent, or mass labels among others has been described. Extensive information on the transducer principles, methods, and examples of all these types of immunosensors can be found in recent reviews [107–110].Although the use of biosensors has been mainly reported for clinical purposes [111], nowadays their application has been extended to different areas, for example the detection of microorganisms such as viruses and bacteria [112], the detection of drugs, and also in the control of environmental pollutants [102, 113, 114]. 2.1.4 Immunoaffinity Chromatography (IAC) The application of immunochemical techniques is not only in the development of analytical detection tools, but also the inherent specificity of the immunoreaction can be exploited to develop selective extraction procedures to be used prior to the analysis by any other analytical method (either immunochemical or a conventional one, like chromatography). Immunoaffinity chromatography (IAC) [115–117] for trace analysis of low molecular weight analytes in complex matrices has several advantages over other solid phases. Apart from selectivity, the antibody cross-reactivity allows the extraction of both the analyte and its metabolites or other structurally related compounds. Preconcentration of the analyte may assist in increasing the detectability of certain analytical methods. The essential hydrophilic media necessary when handling biomolecules allow the purification of polar substances, which can be hindered when other conventional phases are used. Figure 5 shows a scheme of a basic IAC procedure. The Abs are covalently bound to a solid support and packed in small columns, which are conditioned and loaded with the sample. After a washing step that allows removal of the nonspecifically retained compounds, the target analyte is eluted under appropriate conditions (usually by changing the buffer composition). Finally, the column can be regenerated and reused several times.

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Fig. 5 Sequential steps involved in an immunoaffinity extraction procedure

Immunoaffinity procedures can be performed either on-line or off-line, and can be coupled to chromatographic systems [118, 119] or even to immunoassays [120]. Many examples can be found in the literature regarding the use of immunoaffinity extraction of drugs and pharmaceuticals from biological matrices, as well as of organic pollutants such as pesticides from environmental samples [115, 121–124].

3 Immunochemical Methods for Surfactants The use of surfactants in detergent formulations was extended worldwide several decades ago. More environmentally friendly compounds with surfactant properties have gradually replaced the natural soaps. These new substances are characterized by the presence in the chemical structure of both hydrophilic (usually charged) and hydrophobic groups (particularly long linear alkyl chains). This fact imparts unique properties to these compounds as surfaceactive agents. Depending on the charge of the hydrophilic part of the molecule four clear groups can be distinguished: anionic, cationic, nonionic, and amphoteric surfactants. The production volume of these compounds has increased considerably since they have been used as substitutes for soap-based detergents. Thus, in 30 years (between 1940 and 1970) the production of synthetic surfactants in the USA went from 4.5¥103 to 4.5¥106 t [18], whereas that of the natural surfactants has decreased but not to the same extent. According to the European Committee of Surfactants and their Organic Intermediates (CESIO), the total production in western Europe in 2000 was 2.5¥106 metric tons, with anionic and nonionic surfactants being the most abundant ones at production volumes near to 1¥106 and 1.2¥106 metric tons, respectively [125]. Cationic surfactants are produced to a lesser extent in western Europe, constituting approximately 8% (2¥105 metric tons) of the total production of surfactants.

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Fig. 6 General structures of the most important surfactants and metabolites: alkylphenol polyethoxylate (APE); alkylphenol (AP); alkyl ether (AE); alkylphenol ethoxy carboxylate (APEC); linear alkylbenzenesulfonates (LAS); alkyltrimethylammonium compounds (ATMAC); dialkyldimethylammonium compounds (DADMAC); alkyldimethylbenzylammonium compounds (ADMBAC); esterquat (EQ); diesterquats (DEQ). X is usually a chlorine or bromine atom. DDAC (didecyldimethylammonium chloride) and BDD12AC (benzyldimethyldodecylammonium) are the two target analytes with a reported immunochemical technique developed for their analysis [153, 154]

Contrary to anionic and nonionic agents, they have poor detergency and are used more in the preparation of germicides, fabric softeners, and emulsifiers. Amphoteric surfactants are produced in much smaller amounts (5¥104 metric tons, near to 2% of the total production) [125]; they are biodegradable and their ecotoxicological importance can be considered low. Their environmental occurrence up to know has been just occasional. Liquid chromatography coupled to mass spectrometry (LC–MS) is the most usual technique applied for the detection of anionic [126–128], nonionic

ELISA (direct) FIIA FIIA CIA-GDH biosensor ELISA Automated BMP-IA CIA-GDH biosensor ELISA

NP10EO

OP10EO

NPE

OPE

Alkylphenol ethoxylates (APEs)

Immunochemical techniquea

Buffer Buffer Buffer Buffer Buffer

378 mg L–1 104 mg L–1 605 mg L–1 42 mg L–1

6.6 mg L–1

Buffer

17.7 mg L–1

0.5 mg L–1

Buffer/water Buffer

24.5 mg L–1

IC50

Matrix

8.9 mg L–1 2.4 mg L–1

LOD

Sensitivity

[146] [146]

[146] [146] [147]

[145]

[144] [144, 145]

References

CIA-GDH biosensor: capillary immunoassay coupled to a glucose dehydrogenase biosensor; ELISA: enzyme-linked immunosorbent assay; BMP-IA: bacterial magnetic particle-based immunoassay; FIIA: flow-injection immunoassay; PFIA: polarization fluoroimmunoassay. b BDD AC: benzyldimethyldodecylammonium chloride; DDAC: didecyldimethylammonium chloride. 12

a

Nonionic surfactants

Analyte

Table 4 Immunochemical techniques developed for the detection of surfactants. The sensitivity of the method and the matrix considered are shown

146 M.-C. Estévez et al.

Cationic surfactants

Anionic surfactants

Nonionic surfactants

Table 4 (continued)

ELISA CIA-GDH biosensor

OP

ELISA ELISA

BDD12AC

DDAC

Quaternary ammonium compounds (QACs)b

ELISA ELISA FIIA PFIA PFIA Automated BMP-based IA

Linear alkylbenzene sulfonates (LAS)

ELISA ELISA ELISA FIIA PFIA PFIA ELISA CIA-GDH biosensor PFIA

Immunochemical techniquea

NP

Alkylphenols (AP)

Analyte

0.66 mg L–1 29 mg L–1

8 mg L–1

9 mg L–1

Buffer

Buffer, (milk)

Buffer Buffer Spiked rain/water Buffer Buffer Buffer

346 mg L–1 1560 mg L–1 40 mg L–1 19.8 mg L–1 387 mg L–1

Buffer Buffer

590 mg L–1 1033 mg L–1 42 mg L–1 42 mg L–1 769 mg L–1 4481 mg L–1 42 mg L–1

Matrix

Buffer Buffer Buffer Buffer Buffer Buffer Buffer Buffer Buffer

IC50

0.043 mg L–1

0.5 mg L–1 35 ng L–1

19.5 mg L–1

20 mg L–1

9 mg L–1

52 mg L–1 7.9 mg L–1 9 mg L–1

10 mg L–1 76 mg L–1

LOD

Sensitivity

[154]

[153]

[151] [144] [144] [144] [152] [147]

[146] [146]

[148, 149] [144] [144] [144, 145] [144] [150] [146] [146] [150]

References

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[129–138], and cationic surfactants [139], achieving good levels of sensitivity. Gas chromatography also coupled to mass spectrometry (GC–MS) has been used more sporadically [140–142]. As is well known, these methodologies can achieve very low limits of detection (LOD), sometimes in the range of micrograms per liter and even nanograms per liter. However, the analytical protocols employed exhibit important drawbacks derived from the high polarity and solubility of these compounds in water. This fact makes the necessary prior extraction/preconcentration step complicated. Solid-phase extraction procedures are often tedious to ensure the efficient recovery of analytes from water samples [16, 143]. Moreover, usually they are not single substances but technical mixtures of several compounds and isomers (see Fig. 6 for chemical structures). Immunochemical techniques offer in this case the advantage that the analyses are generally performed in water and often have sufficient detectability to allow direct analysis of the sample. Therefore, in principle there is no need to use organic solvents or to extract the surfactant from the aqueous sample. Table 4 summarizes most of the immunochemical techniques that have been developed for these compounds and will be briefly commented on in the next few sections. 3.1 Anionic Surfactants Linear alkylbenzenesulfonates (LAS) represent more than 40% of all surfactants used [18] with a production of nearly 8.5¥105 metric tons. Other anionic agents such as alkyl sulfonates (AS), alkyl ether sulfates (AES), or fatty alcohol sulfates (FAS) are also produced, with the same final applications but lower consumption (see Fig. 6 for chemical structures).All of them are easily biodegradable and efficiently removed by the WWTPs (efficiency near 99%) [28], but because of their large consumption they end up reaching rivers and marine environments. Both LAS and their metabolites, the sulfophenyl carboxylates (SPCs), are usually found in surface waters at levels around a few micrograms per liter, although this level increases in areas where sewage effluents are not connected to municipal WWTPs [23–25, 27]. They have also been detected in seawater [26] and drinking water (1.6–3.7 mg L–1) [23]. Although LAS cannot be considered as toxic substances (none of the anionic surfactants mentioned above is included in the list of dangerous substances of the Council Directive 67/548/EEC), the formation of micelles helps to dissolve and transport many nonpolar organic pollutants, preventing their degradation and enhancing the toxic action of other substances. Thus, synergistic effects are also observed with pesticides (i.e., DDT and dieldrin) and heavy metals like Cd or Hg [155]. Although several immunochemical methods have been reported for LAS, few examples of their application to real environmental matrices have appeared. The first immunochemical method for LAS was reported by Fujita et al. [151]. It is a direct ELISA and uses MAbs generated against 5-sulfophenyl valeric acid conjugated to BSA through the carboxylic acid, thus preserving the sulfonic group

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intact. The LOD achieved is around 20 mg L–1 and the working range between 20 and 500 mg L–1. No cross-reactivity was observed for sodium dodecyl sulfate (SDS), other sodium fatty acid salts (such as sodium palmitate, laureate, and stereate), and other surfactants such as nonylphenol polyethoxylates or shortchain SPCs. Since the long alkyl chains of LAS (between 9 and 13 carbon atoms) were well recognized, there was a risk that long-chain SPCs would cross-react in the assay; however, these compounds were not evaluated. The assay was evaluated in terms of precision and accuracy by measuring spiked river water samples. Recoveries between 81 and 100% were obtained. Similarly the assay was also validated with HPLC. Franek et al. [144] also used several sulfophenyl carboxylic acids with different chain lengths as immunizing haptens to produce a large number of polyclonal antibodies. Direct and indirect ELISAs [144], a FIIA [144], and a PFIA [152] have been developed with these antibodies. For FIIA the tracer was a hapten conjugated to b-galactosidase. The sensitivity reported is quite good and in the same range for both ELISA and FIIA methods (near 20 mg L–1, see Table 4). Unfortunately it is referenced to the linear 4-dodecylbenzenesulfonic acid sodium salt (LDS) instead of to the commercial mixture, which does not give a real picture of the detectability of these methods. The flow format allowed the analysis of ten samples per hour. Matrix effects were also studied using spiked surface and rainwater samples. The latter produced nonspecific interferences in the assay whereas surface water could be directly measured without problems. For PFIA [152], several antibodies and fluorescein thiocarbamyl ethylenediamine (EDF) tracers were screened but the sensitivity was worse. The best combination afforded a LOD of 0.5 mg L–1 and a dynamic range between 3 and 85 mg L–1. These values are about 3 orders of magnitude higher than those obtained for the ELISA developed with the same antibody [144]; however, this technique allows higher sample loads since about seven samples can be analyzed in 10 min. SDS only cross-reacts about 5%. Recently Matsunaga et al. [147] developed an automated immunoassay system based on the immobilization of monoclonal antibodies [151] to magnetic particles, synthesized by magnetic bacteria (Ab-BMP). The assay is performed in microtiter plates mounted in the reaction station. This kind of immunoassay allows the suspension and subsequent easy separation of antibody from the rest of the immunoreagents via a magnet coupled to the tips used in the automated pipette. The detection limit obtained in the competitive assays for LAS is very good (35 ng L–1), similar to those obtained by GC–MS or LC–MS, and shows a wide working range (between 35 ng L–1 and 35 mg L–1). The assay showed low cross-reactivity with short alkyl chain alkylbenzenesulfonates and with SDS, but surprisingly NP is recognized with a 34% cross-reactivity. Finally, MAbs and an immunoassay kit for LAS have been commercialized (see Table 3). The working range of the assay is between 20 and 500 mg L–1. The antibodies are highly specific for LAS with alkyl chains between C8 and C12, whereas other anionic and nonionic surfactants tested showed no cross-reactivity.

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3.2 Nonionic Surfactants During recent decades the nonionic surfactants used most worldwide have been the alkylphenol polyethoxylates (APEs). They are widely used in detergent formulations, both for industrial and domestic applications. They are also used as plasticizers and stabilizers in plastics. About 5¥105 tons are produced annually [156], nonylphenol ethoxylate (NPE) (and to a lesser extent octylphenol, OPE) being the most prevalent one (NPEs represent about 80% of APEs used). This great consumption involves high levels of discharges into the environment. In the WWTPs they are readily biodegraded under both aerobic and anaerobic conditions [6, 157] with removal efficiencies of 90–99%, although it can decrease when high loads of APEs are produced [7]. The mechanism involves the loss of ethoxylate units and the oxidation of the phenol group to form alkylphenol ethoxy carboxylates (APECs). The final breakdown products are APEs with one or two ethoxylate units (AP1EO, AP2EO), APECs, and AP (see Fig. 6 for chemical structures). All these metabolites have lost their surfactant properties and are much more persistent in the aquatic environment [5, 18]. Several studies have indicated that whereas the parent compounds seem to present low toxicity in organisms [19, 158], APE metabolites, particularly APs (both NP and OP) and short-chain APEs, are highly toxic and their estrogenic activity is remarkable [3, 159–163]. The most polar APECs are mainly detected in wastewater, effluents, and rivers, whereas APs, which are more lipophilic, tend to be adsorbed onto soil, sediments, and sludge. The distribution and fate of this family of compounds have been widely reported during recent years [4, 5, 9, 164–168]. The levels of these pollutants in the environment may exceed the predicted no effect concentration (PNEC of 0.33 mg L–1) [165] proposed in a risk assessment report of the EU. As a consequence, their use for household and industrial cleaning has been banned in several countries of the EU, being substituted by less toxic and more environmentally friendly nonionic surfactants such as alcohol ethoxylates (AEs) [156]. In the USA these evidences have also led to regulatory actions related to the domestic use of APEs. It seems that AEs are rapidly biodegraded, although it depends on the nature of the alkyl chain, mainly their branching degree (the degradation decreases when the branching of the chain increases) [19]. Immunochemical methods have been reported for both APEs and their metabolites, especially APs.A discussion of the immunochemical methodologies reported to date, the effect of the immunizing haptens employed, and the features of these techniques were recently reviewed [169]. Unfortunately, the detectability achieved is usually far from what is necessary for direct application to environmental samples. Moreover, the selectivity for APs versus APEs is not always satisfactory. Thus, Goda et al. [148] developed a direct ELISA for NP with a LOD of 10 mg L–1 and a working range between 70 and 1,000 mg L–1, but APEs with one to ten ethoxylate units are also well recognized.

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MAbs for AP and APEs have been produced and commercialized by Takeda Chemical Industries [170] (see Table 3) and have opened up the opportunity to develop a great variety of assays. FIIAs and ELISAs for NP and NP10EO have been developed with these antibodies [144]. The detectability achieved is similar for both methods, although NP10EO is better recognized than NP. Thus, while the FIIA dynamic range reported for NP10EO was 10–500 mg L–1, for NP it was 100–5,000 mg L–1. Using the same antibodies an improved FIIA for APEs (both NPEs and OPEs) and NP was thoroughly evaluated and validated by Badea et al., who studied the influence of different factors such as the presence of organic solvents or heavy metals in the assay media, and its performance in several water matrices (tap water, surface water, and wastewater) [145]. The LOD reported for NP was 51 mg L–1 and about 2.5 mg L–1 for NP10EO, although short NPEO were also highly recognized. The method was applied to the determination of these surfactants in the influent and effluent of WWTPs. The same MAbs have been immobilized on magnetic particles and used to develop an automated immunoassay method for APEs (NPEs with 10–20 EO units) [147]. The wide dynamic range reported (between 6.6 mg L–1 and 66 mg L–1) is related to a low slope assay, which has a direct negative influence on the immunoassay precision. NPEs with two ethoxylate units and NP are also recognized in this assay (49 and 31%, respectively). Finally, with the same antibodies a capillary immunoassay (CIA) coupled to an enzyme biosensor (glucose dehydrogenase, GDH, biosensor) has been developed for APs (NP and OP) and APEs (NPE and OPE) [146]. The competitive step takes place off-line in an antibody-coated plastic capillary and once it has been washed, it is integrated in the flow-injection system with the biosensor as detector unit. The detectability is much worse than that for the ELISA format and, as with the other formats, APEs are better recognized than APs. Franek et al. have also produced polyclonal antibodies for NP [144]. Different antibodies were raised using various para-alkyl hydroxyphenyl compounds with different alkyl chain lengths as immunizing haptens. An indirect ELISA and PFIA have been developed with these antibodies. The detectability accomplished is between 1 and 2 orders of magnitude higher for the ELISA format. While the IC50 of the ELISA was 590 mg L–1, the limit of detection of the PFIA was around 8 mg L–1. As happened with the LAS PFIA method, the sample throughput is very high but it is necessary to develop compatible sample concentration methods in order to apply this technique for environmental monitoring purposes. Subsequently, Yakovleva et al. [150] also reported the development of PFIA for NP, obtaining sensitivities in the same order of magnitude (LOD of 9 mg L–1 and a dynamic range between 10 and 177 mg L–1). Although the cross-reactivity studies showed no recognition of other phenolic compounds, no APE or other related surfactants were tested, which would have been interesting in order to determine the capability of the method to discriminate between NP and its parent compounds.

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3.3 Cationic Surfactants The most important cationic surfactants are those derived from quaternary ammonium salts where one or two hydrophobic groups (usually a long alkyl chain, between 12 and 18 carbon atoms) are attached to positively charged nitrogen. The other two groups are short alkyl chains, usually a methyl or a benzyl group. The most used salts in the formulation of commercial products are quaternary ammonium compounds (QACs). Figure 6 shows a generic structure of different QACs and their abbreviated names. Alkylquats can be considered those structures where the hydrophobic alkyl chain is directly linked to the nitrogen atom (e.g., alkyltrimethylammonium compounds (ATMAC), dialkyldimethylammonium compounds (DADMAC) or alkyldimethylbenzylammonium compounds (ADMBAC)), whereas in the esterquats (EQ) (or diesterquats, DEQ) the hydrophobic group is linked through an ester bond.As happens with LAS or APEs, for their preparation several raw substances are obtained from natural oils and therefore the commercial product is often constituted of a mixture of different alkyl chain lengths. The properties of these surfactants include not only their surfactant activity but also their biocide effect. They are active principles in several household products such as fabric softeners or hair conditioners, and are also in disinfectants, biocides, emulsifiers, wetting agents, and processing additives.After use they are discharged to STPs or directly to surface waters.Although it seems they are readily biodegraded under aerobic conditions (a half-life of 2.5 h for octadecyltrimethylammonium chloride in wastewater [171]), little has been reported concerning anaerobic degradation [172–174]. It seems that alkylquats are poorly degraded under anoxic conditions since the presence of O2 is required in the first steps, for the cleavage of the C–N bond, or for the w-oxidation of the alkyl chain. In contrast, esterquats, whose biodegradation mechanism involves in a first step the hydrolysis of the ester bond, can undergo anaerobic biodegradation [32]. The toxicity of these compounds [173, 175] can be relatively high compared to other surfactants, but their poor solubility and their tendency to adsorb to solids or to complex with anionic substances considerably reduce the real risk and adverse effects for the aquatic environment. [30, 31, 176]. The use of alkylquats has been substituted by the more easily biodegradable and less toxic esterquats that are nowadays the cationic surfactants produced in higher volumes. A competitive indirect ELISA was developed [154], using PAbs raised against an immunizing hapten that preserved both long alkyl chains and with one methyl group substituted by the spacer arm. A LOD of 8 mg L–1 was achieved when didecyldimethylammonium chloride (DDAC) was used as analyte (see Fig. 6 for structure). The specificity of the assay was evaluated by testing shortchain QACs (with methyl or ethyl groups) that were poorly or not recognized, demonstrating that the methyl group or the charged nitrogen atoms are not the main epitopes. Fatty acids (FA) or fatty alcohols (FOH) with long alkyl chains

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were also tested, and high levels of cross-reactivity were observed for those with alkyl chain lengths between 10 and 12 carbon units. Shorter and longer chains were less recognized, which indicates that the main epitope is the decyl chain. A better detectability has been reported on another ELISA for benzyldimethyldodecylammonium chloride (BDD12AC) [153], a component of the benzalkonium chloride (BAK) which is a mixture of three alkyldimethylbenzylammonium chlorides with different alkyl chain lengths (C12, C14, and C16). BAK is widely used and there is also public concern due to its toxicity. The PAbs were raised using as immunizing hapten an analog of the analyte where a methyl group was substituted by the spacer arm. The LOD accomplished was 43 mg L–1 for the bromide analog. No recognition was observed for FAs, FOHs, tertiary amines with long alkyl chains and benzyl amines, and QACs with short (methyl) or medium (C6) alkyl chains. Dialkyldimethylammonium compounds (with C10 and C12 alkyl chains) were slightly recognized, while benzyldimethylalkylammonium compounds (BDACs, with C6–C16 alkyl chains) showed crossreactivity values between 42 and 106%. This ELISA has been validated by HPLC using spiked samples and commercial products.

4 Immunochemical Methods for Polychlorinated and Polybrominated Compounds During recent decades much concern has been focused on the adverse health effects of organochlorinated substances. This group involves several families of compounds (chlorophenols, CP; polychlorinated biphenyls, PCB; polychlorinated dibenzodioxins, PCDD; polychlorinated diphenyl ethers, PCDE; polychlorinated dibenzofurans, PCDF etc.) known to be highly persistent and in some cases with a clearly proven toxicity in organisms. Recently, much more attention is also being paid to organobrominated compounds because of their growing use and production. Analog families of the organochlorinated substances can be found (bromophenols, BP; polybrominated biphenyls, PBB; polybrominated dioxins, PBDD; polybrominated diphenyl ethers, PBDE etc.). Each family has a common and generic carbon-based structure and a different degree of halogen substitution (either Cl or Br), resulting in chemical compounds or commercial products containing mixtures [53]. The structures of the most common polybrominated and polychlorinated substances are shown in Fig. 7. Although the use of organochlorinated substances has been abolished in most of the developed countries, their extensive use during the past decades and their persistence in the environment determine their actual widespread distribution. Moreover, some substances are not used or commercialized but are still important synthetic intermediates for the preparation of other substances. Thus, chlorophenols are important intermediates in the production of pesticides or other chemicals. Other substances such as PCDEs or PCDDs are

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Fig. 7 Generic chemical structures of polyhalogenated compounds. X=Cl, Br. (I) Polychlorinated biphenyls (PCBs), polybrominated biphenyls (PBBs); (II) chlorophenols (CPs), bromophenols (BPs); (III) polychlorinated diphenyl ethers (PCDE), polybrominated diphenyl ethers (PBDE); (IV) polychlorinated dibenzo-p-dioxin (PCDD), polybrominated dibenzo-pdioxin (PBDD); (V) polychlorinated dibenzofuran (PCDF), polybrominated dibenzofuran (PBDF); (VI) tetrabromobisphenol A (TBBPA)

not intentionally produced, but are generated as undesired by-products in various industrial activities and all combustion processes [34]. They can be formed by chemical, photochemical, or thermal reactions from precursors [177]. Chlorophenols are also unintentionally formed when water with a high content of organic material is disinfected with chlorine or during wood pulp bleaching processes. PCBs are used as lubricants, fire retardants, immersion oils, and dielectric heat transfer fluids. For the latter, there was an estimated total production of 1.5 million tons in 1992 [178, 179]. The toxicity and the environmental and ecological impact of these substances have been extensively reviewed, although there are still questions and controversies on the effects after long time exposures. Moreover, the toxicity varies greatly between families and congeners (for reviews see [180–184]). Thus, the greatest concern exists around PCDDs, followed by PCDFs, PCDEs, and PCBs. From the PCDD family, tetrachlorodibenzo-p-dioxin (TCCD) has been identified as the most toxic congener [182]. Data regarding human dioxin exposure have been associated with an increased risk of severe skin lesions such as chloracne and hyperpigmentation, altered liver function and lipid metabolism, general weakness associated with drastic weight loss, depression of the immune system,

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and endocrine- and nervous-system abnormalities. TCDD is considered a potent carcinogenic, teratogenic, and fetotoxic chemical in certain animals. Human populations occupationally or accidentally exposed to chemicals contaminated with dioxin have increased incidences of soft-tissue sarcoma and non-Hodgkin’s lymphoma. On the other hand, some DDT, PCDD, PCDF, and PCB congeners have been classified as high-concern potential endocrine disrupting substances in the list of the 66 priority substances according to the BKH report [3]. The use of organobrominated substances has shown extraordinary growth during the last few years because of their properties as flame retardants (BFRs, brominated flame retardants) and preservatives for woods, plastics, textiles, electronic circuitry, and other materials [185]. The most frequently used are tetrabromobisphenol A, PBB, PBDE (penta-, octa-, deca-brominated diphenyl ether (oxide) formulations), and hexabromocyclododecane. The production, application, and potential environmental occurrence of the most important BFRs have recently been reviewed [186]. The global production (Europe, Asia, and USA) of BFRs increased from 106,700 to 203,500 tons from 1989 to 1999, the most spectacular increase being in Asia (from 28,700 to 119,900 tons over the same period) [186, 187]. Their impact on wildlife and the environment has been reviewed by authors in different countries [38, 56, 57, 188–191].Although an important amount of information exists on the potential risk of the organochlorinated substances, more studies have to be done regarding the toxicity of the polybrominated substances. It is necessary to collect more data on the toxicity effects in different species and on the metabolism and fate of the BFRs [192, 193]. Some studies indicate that the toxicity of the chlorinated and brominated analogs could be very similar, but the results are still very much dependent on the assay used. Thus, toxic effects, including teratogenicity, carcinogenicity, and neurotoxicity, have been observed for some BFR congeners, in particular the PBDEs (for recent reviews see [194, 195]). The endocrine-disrupter activity of the BFRs is being investigated. Hence, PBBs have also been classified as of high concern in the BKH report [3]; on the other hand, some evidence has been provided on the disruption of the thyroid hormone system by BFRs, with particular emphasis on the PBDEs [196]. The application of immunoassays to the analysis of organohalogenated compounds has not been as frequent as for more water-soluble species [197]. Since immunoassays are typically aqueous-based systems, the low water solubility of these compounds makes the use of immunochemical techniques more challenging since several facts have to be considered, starting from a careful selection of the immunizing hapten to aspects such as the preparation of sample extracts or handling of the standards to the last stages in the assay optimization. Other problems derive from the tendency of these substances to bind to the lab ware employed, especially to the plastic ware commonly used for immunoassays (plates, pipette tips, cuvettes, etc.). In spite of these problems a variety of immunochemical techniques have been reported for the analysis of organochlorinated emerging pollutants. Conversely, to our knowledge no immunochemical techniques have been reported for the determination of

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organobrominated substances, although it has been reported that antibodies raised against chlorophenols have the ability to recognize the corresponding brominated analogs even more [198, 199]. Moreover, preliminary experiments performed in our group have shown that the application of the 2,4,6-trichlorophenol immunoassay to the analysis of 2,4,6-tribromophenol from wood extracts would be feasible [200]. In the following we will describe briefly the features of some of the immunochemical techniques available today for the detection PCBs, PCDDs/PCDFs, and chlorophenols (see Table 5). 4.1 PCBs The first assays for PCBs date from the 1980s [228, 229] and since then several other attempts have also been carried out (reviewed in [169, 230]). The described assays are usually based on the analysis of PCBs such as Aroclors, not as congeners, and use PAbs produced using an Aroclor or a single PCB as a hapten. RIAs have been developed and used to detect them in biological matrices such as milk or blood [83, 207, 231], whereas other kinds of immunoassays have been reported to detect them in environmental samples such as water, soils, and sediments [201–203, 232, 233]. The detectability reported depends on the hapten used to raise antibodies and also on the congener used to calibrate the assay. Simple extraction methods rendering extracts compatible with the aqueous media of the immunochemical methods are sometimes reported in order to be able to perform assays on-site. Usually a water-miscible solvent such as methanol is present at a certain percentage to improve solubility and to diminish adsorption of the analyte to the glass containers or plastic surfaces [205, 210].

a

b

Fig. 8a, b a Liposome immunocompetition assay (LIC); R1, liposome/PCB competition zone. b Liposome immunoaggregation assay (LIA); R2, liposome/antibody aggregation zone. C1, C2: anti-biotin capture zones. Published with permission of ACS Copyright Office [209]

ELISA RIA

ELISA

RIA

EIA magnetic particle

ELISA FOI

ELISA

ELISA

ELISA

Immunochemical techniqueb

38 mg L–1 0.013 mg kg–1

25 mg L–1

0.2 mg L–1 500 mg Kg–1

1.57 mg (10.5 ng g–1) 5–12.9 mg L–1 10 mg L–1

L–1

1.34 mg L–1 (8.95 ng g–1)

LOD

Sensitivity

212 mg L–1

2 mg L–1 20 mg L–1 217 mg L–1

Buffer Milk

Water Soil Blood Milk Buffer

Soil River water, soil

Soil, sediments

Soil

1,000 mg L–1 25 mg L–1

Soil, sediments

22 mg L–1

IC50

Matrix

[206] [207]

[205] [205] [83] [83] [206]

[203] [204]

[201]

[202]

[201]

References

DCP: dichlorophenol; PCP: pentachlorophenol; 2,3,7,8-TCDD: 2,3,7,8-tetrachlorodibenzo-p-dioxin; TCP: trichlorophenol; TMDD: 2,3,7-trichloro8-methyldibenzo-p-dioxin. b EIA magnetic particle: enzyme immunoassay based on magnetic particles; ELISA: enzyme-linked immunosorbent assay; FOI: fiber optic immunosensor; LIA: liposome immunoaggregation assay; LIC: liposome immunocompetition; LIF-microdroplets-QFIA: laser-induced fluorescence detection in microdroplets with quenching- fluorescence immunoassay; RIA: radio immunoassay.

a

Aroclor 1260

Aroclor 1254

Aroclor 1242

Aroclor 1248

PCBs

Analytea

Table 5 Immunochemical techniques developed for the detection of organochlorinated substances

Immunochemical Determination of Industrial Emerging Pollutants 157

ELISA

3,4,3¢,4¢-tetrachlorobisphenyl

TMDD

2,3,7,8-TCDD

ELISA ELISA ELISA

ELISA ELISA

ELISA ELISA ELISA ELISA

LIA LIC

2-chlorobiphenyl

Dioxins PCDDs/PCDFs

FIIA

Immunochemical techniqueb

Aroclors 1242, 1248, 1254, 1260

PCBs

Analytea

Table 5 (continued)

0.01 mg L–1 0.004 mg L–1

0.1 mg L–1 0.025 mg L–1

0.24 mg L–1 0.036 mg L–1 16 ng mL–1

19.5 pg per tube 10.4 ng mL–1

4 mg L–1 200 pg per well 1 ng per well

Buffer Buffer Buffer

Buffer Buffer Buffer Soil Water Fly ash Buffer

Buffer

0.2 mg L–1

Buffer Buffer Buffer 0.9 mg L–1

IC50

Matrix

0.26 pmol 0.4 nmol

1 mg L–1

LOD

Sensitivity

[216] [217] [218]

[211] [212] [213] [214] [214] [215] [76]

[210]

[209] [209]

[208]

References

158 M.-C. Estévez et al.

ELISA ELISA

ELISA

2,4-DCP

2 mg L–1

0.053 mg 0.07 mg L–1 0.26 mg L–1 0.8 mg L–1

L–1

0.23 mg L 0.32 mg L–1 1.28 mg L–1 6.46 mg L–1

–1

7.8 mg L–1

1.6 mg L–1 0.2 mg L–1

ELISA

Buffer

Buffer Water Urine Serum

Urine Water

Buffer Buffer Buffer Buffer

1.53 mg L–1 1.44 mg L–1 2.76 mg L–1 0.45 mg L–1

0.2 mg L–1 0.2 mg L–1 0.2 mg L–1 0.04 mg L–1

ELISA ELISA ELISA LIF-microdroplets-QFIA

Soil Buffer

500 mg L–1 30–40 mg L–1

Water Buffer

ELISA ELISA

2.9 mg L–1 2.2 mg L–1

IC50

Matrix

0.1 mg L–1 0.06 mg L–1

LOD

Sensitivity

ELISA EIA magnetic particles

Immunochemical techniqueb

2,4,5-TCP

2,4,6-TCP

PCP

Chlorophenols

Analytea

Table 5 (continued)

[148]

[226] [227] [227] [227]

[225]

[198] [199] [223] [224]

[221] [222]

[219] [220]

References

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Two liposome-based immunomigration techniques have been developed for PCBs based on a test-strip format for on-site analysis (see Fig. 8). The liposome immunocompetition (LIC) assay format measures the competitive reaction between analyte-tagged liposomes and the sample analyte for immobilized antibodies and can detect 0.4 nmol of PCB in less than 8 min. The liposome immunoaggregation (LIA) assay is based on the principle of immunoaggregation between anti-PCB antibodies and analyte-tagged liposomes, and detects the inhibition of immunospecific liposome aggregation in solution produced by the presence of the analyte. This last assay can detect 2.6 pmol of PCB in less than 23 min. Both formats utilize capillary action to transport liposome-containing solutions along strips of nitrocellulose. Measurement of color intensity is then carried out visually or with a desktop scanner [209, 234]. A continuous semiautomated FIIA system has been reported [208, 235]. In this device the analyte-containing medium is allowed to flow through a column containing the antibodies immobilized on a support. First, the antibodies are saturated with a fluorescent dye-labeled analog of the analyte. As the analyte passes through the immunosorbent, some dye-labeled antigen is displaced and is then detected in a fluorometer located downstream from the column. The LOD achieved for the developed system is 1 mg L–1. Several ELISA formats have been developed for PCB determination. Thus, MAbs have been developed for coplanar PCBs [210], which are the most toxic congeners. The ELISA developed is highly selective for PCB 77 and 126, showing IC50 values of 0.9 and 1.2 mg L–1, respectively. Noncoplanar PCBs, PCDDs, PCDFs, or single-ring halogenated compounds, including chlorinated benzenes and phenols, do not interfere with this assay. Johnson et al. [201] also produced PAbs for PCBs and an indirect ELISA has been optimized for the detection of several Aroclors, giving LODs of about 1.5 mg L–1 that correspond to 9 ng g–1 in soils and a dynamic range between 50 and 1,333 ng g–1. Potential contaminants usually found in soil samples such as chlorophenols and chlorobenzenes were also tested, exhibiting cross-reactivities lower than 3%. A competitive enzyme immunoassay for the quantification of PCBs in water has been developed using PAbs covalently attached to amine-terminated superparamagnetic particles as solid support [205]. The assay detected various Aroclors (1016, 1232, 1242, 1248, 1254, 1260, 1262, and 1268) with detection limits of 0.2 mg L–1 in water and 500 mg kg–1 in soil using Aroclor 1254 as standard. The assay was validated using the GC-EPA method 8080 in water, demonstrating good performance and excellent precision and accuracy. This assay is available as a commercial kit to be applied to different matrices such as soil, wipes, and water. Kim et al. [206] have used commercial MAbs for development of an immunoassay for the determination of PCBs in insulating oils. A dynamic range between 30 and 1,000 mg L–1 has been achieved for the assay in methanol and allows analysis of diluted oils containing >35 mg mL–1 PCBs (neat). The procedure requires a previous pretreatment of oil samples (either a solid-phase extraction or washing with KOH-EtOH/sulfuric acid to remove interference).

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There are some other firms that also commercialize immunoassay kits for PCBs based on different formats (see Table 3). They have been mainly validated for soil and wipe matrices, although the manufacturer can provide optional protocols for testing sediment or water samples. A fiber optic immunosensor (FOI) has also been reported for detection of PCBs in Aroclors [204]. The quartz fiber surface is coated with PAbs against PCBs and the competitive assay takes place using as fluorescent tracer, an analog of the analyte coupled to 2,4,5-trichlorophenoxybutyrate (TCPB) on the Ab-coated fiber. The LOD achieved is around 10 mg L–1. A high-performance immunoaffinity chromatographic (HPIAC) method has been developed with the aim of improving the analytical methodology of PCBs [236, 237]. The IAC column prepared using PAbs generated against the coplanar toxic PCB congeners reduces the cleanup steps, time of analysis, and the costs because of the ability to selectively retain PCBs. Moreover, this sample treatment method reduces the use of hazardous organic solvents. 4.2 PCDDs and PCDFs Several attempts have been made to set up immunochemical techniques for dioxin analysis (reviewed in [230, 238, 239]). Frequently the detectability and selectivity accomplished have not been considered appropriate for the direct analysis of environmental samples. We should notice that due to the poor solubility of PCDDs and PCDFs in water, the levels of these contaminants in aqueous samples is very low. For this reason analysts usually prefer the use of chromatographic and spectrometric methods that perform using organic solvents. However, the speed and high sample throughput that can be accomplished with the immunochemical methods have prompted several research groups and companies to establish immunochemical methods. The first IA for dioxins was a RIA developed by Albro et al. [240]. It was quite time-consuming and utilized PAbs showing low specificity. MAbs were developed later by Kennel et al. [241], but they also lacked sufficient detectability to analyze dioxins in solution. The first ELISA was developed by Stanker et al. using MAbs (DD3) generated against 2,3,7,8-TCCD [211, 242]. This congener is normally used as analyte because of its recognized higher potential toxicity. The selectivity of the ELISA was very similar to that of the RIA. The optimized assay had an IC50 around 200 pg per well [212]. Langley et al. [213] also reported a PAb-based ELISA with an IC50 value of 1 ng of TCDD per well. Several years later Harrison and Carlson [214] developed a tube test and a microplate test using the DD3 MAb and the two formats displayed detection limits of 100 pg per tube and 25 pg per well. Sanborn et al. [218] reported the production of PAbs using haptens containing an unsaturation in the spacer arm (between the halogenated dibenzo-p-dioxin ring system and the protein to which it is conjugated), which provides a rigid handle structure. The chemical structure of the hapten was similar to that of TCDD (i.e., 2,3,7,8- or 1,2,3,7,8-) but the polar

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groups for hydrogen bonding were lacking. In the haptens synthesized, one substituent is an alkyl chain (between 3 and 5 carbon atoms) containing at least one double bond or aromatic ring with a carboxylic group at the end of the molecule. An ELISA was developed that showed an IC50 of 0.8 ng per well (16 ng mL–1) when 2,3,7-trichloro-8-methyldibenzo-p-dioxin (TMDD) was used as analytical surrogate standard in order to avoid using the toxic congener. It was proved that this analyte responded similarly to 2,3,7,8-TCDD. The same research group was able to improve the assay by introducing several modifications, such as the chemical structures of the competitor haptens used as coating (i.e., trans-3-(7,8-dichlorodibenzo-p-dioxin-2-yl)-cis-2-methylpropenoic acid coupled to BSA) or the content of DMSO in the assay (up to 37%). The ELISA exhibited an IC50 value of 12 pg per well (240 pg mL–1), with a working range from 2 to 240 pg per well (40 to 4,800 pg mL–1) [216]. Subsequently, the use of a new coating antigen for the indirect assay was optimized to finally reach a LOD of 4 ng L–1 [217]. Correlation with GC–MS was carried out, achieving good agreements for soil samples without the necessity for any additional cleanup step prior to the analysis. Intensive work in this field has continued in order to improve detectability and to establish reliable immunochemical protocols, including appropriate sample treatment methods, for the analysis of PCDDs and PCDFs in real samples [214, 238, 243–246]. The development of selective extraction procedures as well as solvent exchange methods to finally achieve immunoassay-compatible extracts with solvents has been critical. Thus, an immunoaffinity extraction/ cleanup protocol was developed by Harrison et al. [243] and then used on an assay employing an improved MAb (DF1). Handling of standards and samples has also been simplified by the use of detergent keeper in the solvent exchange procedure. Based on an assumption of quantitative recovery, a tenfold concentration of the original sample was accomplished before the immunochemical measurement. The final analytical method (sample treatment plus immunochemical method) allowed improvement of the sensitivity up to 100 pg L–1, starting from 2 L of water [238, 244, 246]. Also an attempt has been made to clone and express recombinant Fab antibodies against dioxins [76] using two hybridoma cell lines (DD1 And DD3) [211]. The option of cloning Fab fragments rather than scFv was chosen because they are usually more stable, which is important when environmental analysis of complex matrices is going to be carried out. Using 2,3,7,8-TCDD as standard an indirect ELISA has been developed. As in the aforementioned work, considerations regarding the use of organic solvent (MeOH) have been taken into account and an assay with IC50 values around 10.4–14 ng mL–1 was achieved depending on the Fab used. These sensitivities and the results of cross-reactivity studies are very similar to those obtained with the respective MAbs, showing their usefulness as analytical reagents. Several of these assays have become commercially available (see Table 3). The IA kit in a coated tube method [247], developed by Cape Technologies, has been made available for the analysis of various types of sample extracts (fly ash,

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soil, stack gas, tissue, sediment, water, etc.) after the application of conventional extraction methods, followed by a solvent exchange step to provide hydro-alcoholic (methanol–water) extracts compatible with the immunoassay methods. Also the production of specific Abs for PCDDs/PCDFs has been directed toward the development of immunoaffinity procedures [248, 249]. Shelver et al. reported several works regarding the use of IAC to selectively extract and analyze these compounds from complex matrices such as milk or serum [250–253]. Moreover, a separation of very similar dioxin congeners (i.e., 1,3,7,8-TCDD and 2,3,7,8-TCDD) was also examined [254]. It is also worth mentioning that some authors have tried to demonstrate a correlation between congener immunochemical recognition and the I-TEF (toxicity equivalent factor, relative toxicity related to that of 2,3,7,8-TCDD), which indicates the potential for predicting the I-TEQ (toxic equivalent quotient, total toxicity of a mixture attending to the individual TEF values) [214, 215, 243, 247, 255]. This would provide a method for estimating the I-TEQ value of a sample by multiplying each mass concentration obtained by GC–MS by the corresponding immunoassay cross-reactivity value [245]. Harrison and Carlson [247] used their immunoassay to correlate PCDD/F congener recognition profiles with the congener toxicity in order to estimate the TEQ of real samples [243, 245]. 4.3 Chlorophenols Among all chlorophenols, 2,4,6-trichlorophenol (TCP) and pentachlorophenol (PCP) are listed as priority pollutants by the US Environmental Protection Agency (EPA) (IRIS electronic database) and the EU [256]. In particular, PCP has been classified as a B2 probable carcinogen for humans from animal toxicity studies and human clinical data. There have been various attempts to develop immunoassays for chlorophenols in environmental samples (soil, water). Noguera et al. [219] produced PAbs for PCP. By using pentachlorophenoxypropionic acid as immunizing hapten (so that the five chlorine atoms are kept intact in the molecule) a direct ELISA was developed with a LOD in water of 0.1 mg L–1 and a working range between 0.3 and 30.5 mg L–1. The immunoassay is very specific for PCP since only one of the compounds tested (2,3,5,6-tetrachlorophenol) shows a significant degree of cross-reactivity (21.3%). Other CPs and related phenols are not recognized (CR<0.3%). The assay is applicable to real environmental water with a minimal sample pretreatment for river and lake samples. A competitive heterogeneous immunoassay based on the use of magnetic particles as solid support has also been developed using commercial PAbs for PCP [220]. Good sensitivities have been achieved with this format (the LOD is about 60 mg L–1) and it shows low cross-reactivity with related compounds (only for tetrachlorophenols, TtCPs, is the degree of recognition significant with CR of 54% for 2,3,5,6-TtCP and 15% for 2,3,4,6-TtCP). The assay allows the

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analysis of water samples (e.g., river water, pond water, or groundwater) without sample pretreatment as well as quantification of soil samples including a simple extraction procedure. It provides results in 60 min and can be easily adapted to on-site monitoring. The IA has also been compared with GC–MS and HPLC, obtaining in both cases good correlation levels for spiked water and soil samples. Many studies have been carried out in recent years in order to detect PCP in environmental samples, and most of them are based on the use of commercially available IAs for PCP in water and urine (shown in Table 3). Thus, several assays such as PENTA-RISc, PCP RaPID-Assay, or another IA developed by Westinghouse Bio-Analytic Systems (WBAS) have been evaluated for on-site screening tests of PCP in soils or aqueous matrices such as surface, drinking, or ground water [221, 222, 257]. PCP RaPID-Assay was evaluated using certified wastewater samples, soil samples, and certified reference materials [258, 259].A critical comparison of the ELISA method with an online liquid–solid extraction (LSE) method followed by liquid chromatography (LC-UV or LC-MS) analysis was performed.A good correlation was found between both methods, although for some samples undesirable matrix effects were also observed. PAbs have been developed against 2,4,6- and 2,4,5-TCP after careful study of the chemical structure of these substance using computer-assisted molecular modeling tools and theoretical calculations [223, 226]. For 2,4,6-TCP both direct [223] and indirect [198, 199] ELISA formats have been developed. The direct assay uses a homologous hapten (same chemical structure as the immunizing hapten) coupled to horseradish peroxidase as tracer. The microtiter plate ELISA can be carried out in about 1 h and it has a LOD of 0.2±0.06 mg L–1. The assay tolerates samples having a wide range of ionic strengths (from 4 to 56 mS cm–1) and pH values (between 5.5 and 9.5). The indirect ELISA format has a similar detectability but it has proven to be more robust to matrix interferences. This ELISA uses a heterologous hapten coupled to BSA as coating antigen and it can be performed in 1.5 h. As with other microtiter plate ELISA methods, the advantage is that many samples can be processed simultaneously. Both direct and indirect assays show a similar pattern of selectivity. Thus, the indirect ELISA [198] recognizes to a much lesser extent other chlorinated phenols, such as 2,3,4,6-tetrachlorophenol (2,3,4,6-TtCP, 21%), 2,4,5-TCP (12%), recognized than the corresponding chlorinated analogs (ex. 2,4,6-TBP, 710%; 2,4-dibromophenol, 119%). The indirect ELISA formats have been evaluated for the analysis of several types of water samples [199] and also for urine samples [198] in order to perform human exposure assessment studies. Using the same PAbs an optical biosensor system has been developed for 2,4,6-TCP [224]. The principle is the detection of laser-induced fluorescence (LIF) in single microdroplets by a homogeneous quenching fluorescence immunoassay (QFIA). The competitive immunoassay occurs in microdroplets (d=58.4 mm) produced by a piezoelectric generator system. A continuous Ar ion laser (488 nm) excites the fluorescent tracer and its fluorescence is detected by a spectrometer attached to a cooled, charge-coupled device (CCD) camera

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Fig. 9 Scheme of the instrumental setup. Microdroplets are generated by a vibrating orifice aerosol generator (orifice diameter=10 mm) consisting of a syringe pump and a piezoceramic oscillator. Droplets are illuminated by a laser beam (continuous Ar ion laser, l=488 nm) that is focused to a laser beam waist of ~200 mm at the trajectory of the droplet stream. The beam diameter is chosen to be larger than the droplet diameter to ensure that all fluorophores in the sample are illuminated. A holographic laser band-pass filter eliminates undesirable plasma lines from the laser source and transmits only the laser line at 488 nm. The fluorescence is collected by a microscope objective lens (N.A. of 0.55) and focused onto the entrance slit of the imaging spectrometer. Spectra are recorded with a thermoelectrically cooled camera with a 512¥512 pixel charge coupled device (CCD) detector. A 488-nm holographic Raman notch filter placed in front of the slit of the spectrometer blocks elastically scattered laser radiation. Published with permission of ACS Copyright Office [224]

(see Fig. 9). The fluorescence is quenched by specific binding of the TCP PAbs to the fluorescent tracer (homologous hapten covalently coupled to fluorescein). The quenching effect is diminished by the presence of the analyte. Therefore an increase in the signal is produced in a dose-dependent manner when TCP is present in the sample. The LIF-microdroplet-QFIA method shows a LOD of 0.04 mg L–1 in buffer. The QFIA performed in microtiter plate format using the same immunoreagents achieved a LOD of 0.36 mg L–1 in buffer. With the LIFmicrodroplet-QFIA system, urine samples can be directly analyzed just after buffer dilution reaching a LOD of 1.6 mg L–1, which is sufficient detectability for occupational exposure risk assessment. PAbs against 2,4,5-TCP have been prepared after theoretical and molecular modeling chemical studies [226]. Competitive direct and indirect ELISAs have been developed, but as before the latter format was shown to be more robust. The indirect immunoassay has an excellent LOD near 0.05 mg L–1. The selectivity of the assay is high in relation to other chlorophenols frequently present in real samples, but as with the 2,4,6-TCP assay the brominated analogs may also be recognized. The 2,4,5-TCP immunoassay is stable in media with pH

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values ranging from 6.6 to 10.5 and ionic strength values varying within 20 and 80 mS cm–1. It shows a good accuracy and the coefficients of variation within and between assays are around 12% or lower. The assay has been evaluated for the analysis of water samples and complex biological matrices, such as serum and urine. While the water samples could be analyzed without any sample pretreatment, the serum and urine samples produced important interferences. The investigation of simple sample treatment procedures compatible with the immunochemical method allowed the establishment of reliable analytical protocols for straightforward immunochemical determination of 2,4,5-TCP in natural waters, urine, and serum reaching LODs of 0.07, 0.26, and 0.8 mg L–1, respectively [227]. Whereas the immunoassays for PCP and TCP show a great level of specificity for these analytes, Noguera et al. [225] have recently developed an ELISA for screening the total concentration of chlorophenols in environmental samples. Using 3,5-dichloro-4-hydroxyphenyl propionic acid as immunizing hapten (the structure contains two chlorine atoms in the ortho position to the phenol group, i.e., the pattern of 2,6-dichlorophenol), PAbs were obtained and two immunoassays developed (in both direct and indirect formats). The most sensitive one was the direct format and after the optimization of several parameters such as surfactant concentration, ionic strength, or time of incubation a LOD of 0.2 mg L–1, an IC50 value of 7.8 mg L–1, and a dynamic range between 0.7 and 86.4 mg L–1 were obtained when 2,4,6-TCP was used as standard. The cross-reactivity studies carried out show high levels of recognition of other CPs. Thus, the most recognized one was 2,6-dichlorophenol (2,6-DCP; CR=66.7%) since the pattern of substitution is the same as that in the immunizing hapten, 2,3,5,6-TtCP (CR=34%) and PCP (CR=33%). 2,4,5-TCP, 2,5-DCP, and 2,4-DCP were recognized to a much lesser extent (<7%) and chlorophenols with just one or no chlorine in the molecule or other related compounds without the phenol group didn’t cross-react (CR<0.09%). Owing to the broad degree of recognition of this assay it has been tested as a potential screening tool in order to estimate contamination by chlorophenols. Several preliminary studies have been carried out with spiked water samples by the addition of different amounts of CPs with CR levels higher than 1%, and the results of the assay, expressed as equivalents of 2,4,6-TCP, showed good correlation levels when a regression analysis was performed, with recoveries around 95%. Goda et al. reported the only attempt carried out regarding the production of MAbs against CPs [148]. By using a similar hapten to the aforementioned one (in this case the spacer arm has six carbon atoms instead of three, but the pattern of 2,6-dichlorophenol is also retained), a direct ELISA using 2,4-DCP as standard was developed. A LOD of 2 mg L–1 and a working range between 2 and 60 mg L–1 were achieved. This assay also shows a high degree of recognition of several chlorophenols tested but, unlike the direct assay of Noguera et al [225], the pattern of recognition is different. Thus, high or moderate cross-reactivity values were observed for some mono- or dichlorophenols (e.g., 4-chlorophenol showed a CR=100%), and even for phenol (24%), whereas for more substituted

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CPs the recognition was lower (i.e., 2,4,6-TCP, 13%; 2,3,6-TCP, 3%; 2,3,4,6-TtCP, 9%; and PCP only 0.3%). Thus, this assay can be useful for determination of the overall concentration of CPs instead of the single compound 2,4-DCP, since an overestimation can be observed in the analysis.

5 Other Industrial Residues From the wide variety of emerging pollutants of industrial origin that could be considered here, bisphenol A (BPA) and phthalate esters (PE) are of especial relevance not only because of the high volumes produced and their widespread use, but also because of their demonstrated toxicity, particularly as endocrine disrupters. Both of them have been included in the final report of the European Commission toward the establishment of a priority list of endocrine disrupter chemicals, EDCs [3], and have been rated as of high risk of exposure for human and wildlife populations. Because of their structural characteristics these compounds cannot be included in any of the groups described above, so they will be described in this section (see Fig. 10). 5.1 Bisphenol A Bisphenol A (2,2-bis(4-hydroxydiphenyl)propane, BPA) is a man-made chemical mainly used in the manufacture of polycarbonate and epoxy resins. These plastics are used in the preparation of containers such as food and drink packaging as well as a great variety of products including compact discs, optical lenses, thermal paper, adhesives, powder paints, or even in dental composite fillings and sealants. To a minor extent (about 10%), BPA is used in PVC production, as an antioxidant and preservative, or as a flame retardant (i.e., tetrabromobisphenol A, see above). This wide range of applications has led to high levels of worldwide production (around 2.6¥106 tonnes in 1999). Besides the discharge into the environment from their output, another important source of

Fig. 10 Chemical structure of phthalate esters and bisphenol A. DMP: dimethyl phthalate; DEP: diethyl phthalate; DBP: di-n-butyl phthalate; DOP: dioctyl phthalate; DEHP: diethylhexyl phthalate; BBP: butylbenzyl phthalate

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release and human and wildlife exposure comes from leaching from these products as a result of incomplete polymerization or hydrolysis of the polymers. Therefore BPA has become not only an environmental pollutant but also a food contaminant. Due to both its industrial and domestic applications it can be expected to be found in sewage, influent and effluent wastewater, and sewage sludge. Their discharge into the environment from industrial emitters and communal wastewater has been monitored [67], with detection of BPA at concentrations between 35 and 50 mg L–1 in several wastewater samples (i.e., hospitals, household areas, and food, chemical, and paper industries). It was found that almost 90% of the total load was removed, with an effluent mean concentration of about 1.5 mg L–1. The levels in surface waters are usually lower [9]. The acute toxicity levels for BPA have been measured for several organisms such as algae, invertebrates, and fish and range from 1 to 20 mg L–1 [69, 260], but of most serious concern is their proven estrogenic activity [261, 262]. Several studies have shown that alterations in reproductive organs in female rats, fish, and mice can be produced [263–267]. BPA is mainly determined using chromatographic techniques such as GC–MS [141, 142, 168, 268], HPLC-UV [269], HPLC–MS [142], or HPLC with fluorescence detection [270]. Immunochemical methods can improve monitoring efficiency because of their known capability to reach low limits of detection and to process many samples. Specific PAbs and MAbs have been produced against BPA to develop ELISAs (see Table 6). Due to the evident risk of human exposure to BPA, further application of these immunoassays has been carried out to analyze biological matrices such as serum samples. An indirect ELISA with PAbs was developed by Zhao et al. [271] with a good LOD of 0.1 mg L–1 and a dynamic range between 1 and 10,000 mg L–1 in water. The PAbs were raised using as immunizing hapten a bisphenolic structure preserving both phenolic groups and replacing one methyl group by the spacer arm. No important matrix effects were found when spiked real water samples were analyzed. Other phenolic compounds did not interfere in this assay. When spiked serum samples were analyzed a dilution factor of 1:10 was required in order to obtain quantitative recoveries, placing the LOD still at a good level (around 2 mg L–1). With these antibodies an immunoaffinity procedure for the selective extraction of BPA from serum samples has been developed providing good recovery levels (about 90%) [272]. Ohkuma et al. [273] also prepared PAbs for BPA but in this case one carboxyalkyl ether as spacer arm was substituted for one of the phenolic groups. With these antibodies a direct competitive ELISA was developed, achieving a LOD of 0.3 mg L–1 and a working range between 0.3 and 100 mg L–1. Accuracy evaluation was carried out by analyzing spiked human serum samples; recovery values between 82 and 97% were obtained. The interferences caused by the matrix were negligible. Correlation studies were also done with a conventional chromatographic technique (GC–MS), obtaining a regression coefficient of 0.990, thus demonstrating the possibility of using this immunochemical method to directly determine BPA in serum. Phenols and other endocrine disrupters such as alkylphenols and phthalates were not recognized in this assay.

b

a

Buffer

0.2 mg L–1

Automated BMP-IA

2.3 ng

L–1

0.3 mg L–1

ELISA

Buffer

Water

Serum

0.1 mg L 2 mg L–1

ELISA

ELISA

Water Serum

–1

5 mg L–1

Buffer

Buffer

0.57 mg L–1

IC50

Matrix

97 ng L–1

LOD

Sensitivity

ELISA

ELISA

DBP

Bisphenol A

TR-FIA

DMP (DEP, DBP, BBP and DOP)

Phthalate estersb

Immunochemical techniquea

[147]

[275]

[273]

[271] [271]

[148]

[148]

[276]

References

ELISA: enzyme-linked immunosorbent assay; BMP-IA: bacterial magnetic particle-based immunoassay; TR-FIA: time-resolved fluoroimmunoassay. DEP: diethyl phthalate; DBP: di-n-butyl phthalate; BBP: butylbenzyl phthalate; DMP: dimethyl phthalate; DOP: dioctyl phthalate.

Others

Analyte

Table 6 Immunochemical methods developed for the detection and quantification of BPA and phthalate esters

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Other attempts have been made to detect BPA at a low concentration range. Thus Kodaira et al. [274] analyzed BPA in urine samples with an assay that showed a working range between 0.5 and 5 mg L–1. The assay was validated by HPLC. DeMeulenaer et al. [275] developed an indirect competitive ELISA using PAbs obtained from chicken egg yolk, but the assay achieved an IC50 value of only 570 mg L–1. The production and use of monoclonal antibodies against BPA have also been reported. Nishii et al. [277] developed an ELISA with MAbs selected for their resistance to organic solvents, achieving a LOD in the order of 1 mg L–1. Goda et al. [148] also prepared specific MAbs for BPA and developed a direct ELISA with a LOD of 5 mg L–1 and a dynamic range between 5 and 500 mg L–1. A high degree of recognition was observed for two bisphenolic compounds, produced and used to a lesser extent, whereas other related compounds and BPA metabolites showed no cross-reactivity. Takeda Chemical Industries, Ltd. have commercialized ELISA kits using these antibodies. The immunoassay shows a better sensitivity (i.e., the dynamic range is between 0.05 and 10 mg L–1) and a high specificity toward other related compounds.An immunoassay based on the immobilization of monoclonal antibodies on the surface of magnetic particles has been developed [147]. The sensitivities achieved were also very good but the assay shows a wide dynamic range (between 2.3 ng L–1 and 2.3 mg L–1), indicating a low slope assay as occurred with the assay developed for APE based on the same principle (see above). 5.2 Phthalate Esters Phthalate esters are used to manufacture cosmetic products, inks, adhesives, and solvents, but they are mainly used as additives in PVC production. This kind of plastic can be found in a wide range of products such as enclosures for food containers, defoaming agents, soft squeeze toys, and teething rings. Among the different phthalates that are used for these purposes, diethylhexyl phthalate (DEHP) represents over 90% of the total phthalate production, near to 54,000 tonnes per year [278, 279], followed by butylbenzyl phthalate (BBP), dibutyl phthalate (DBP), and dioctyl phthalate (DOP). Due to this wide range of applications as well as their high consumption, these compounds can enter into the environment through different routes, increasing the risk of exposure of the human population through food and also by inhalation and dermal contact. For instance, many pipes and bags used in hospitals are made of PVC and high levels of DEHP in blood have been found in patients treated with hemodialysis [280–282]. The origin of these levels can be the leaching of phthalates from the tubes or the bags containing blood. These plasticizers are poorly soluble in water, and therefore it is expected to find them not only in wastewater but also adsorbed onto sewage sludge and soils. The environmental fate of phthalates has been extensively reviewed [65] and it can be concluded that their aerobic degradation occurs rapidly, preventing their accumulation in water and

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even in soil. The biodegradation is slower in anaerobic or cold environments, but several reported experiments indicate that the bioaccumulation of these compounds is limited by biotransformation. Regarding their toxicological data, many studies have been done [283–286]. Their effect as testicular toxicants has been proven [287, 288], and great concern has also been focused on their endocrine disrupting effects, which have been reviewed [289]. Several in vitro studies confirm their estrogenic activity [161, 290], mainly for BBP and DBP, although it is quite weak compared with the natural estrogen 17b-estradiol [291]. Other in vivo experiments also show that DBP and DEHP can produce alterations in male sexual differentiation and in reproduction [292]. The mechanism of action of these compounds is unclear, since some studies indicate that they may act as antiestrogens [293]. A high degree of removal efficiency of DEHP in STPs [294] has been found. Phthalates have been detected in groundwater, rivers, and drinking water [290] as well as in industrial effluents, sewage sludge, and soils [295]. Their usual analysis involves the use of both GC and LC coupled to MS detection, although sometimes a solid-phase extraction is required prior to the analysis [37, 142, 296–298]. To our knowledge, only two attempts have been made to obtain specific antibodies for phthalate esters (see Table 6), although none of them is capable of detecting DEHP, the most used phthalate and also the most persistent one in the environment. The first immunochemical technique is based on the development of a time-resolved fluoroimmunoassay (TR-FIA) [276], a heterogeneous immunoassay that uses europium chelates as labels. Their special fluorescent properties, such as their longer decay time (over hundreds of milliseconds) compared with other more conventional organic molecules (around nanoseconds), determine the good detectability of this method. PAbs were produced against an immunogen derived from dimethyl phthalate (DMP). The assay developed using these antibodies showed a LOD of 97 ng L–1 and a working range between 97 ng L–1 and 388 mg L–1 for DMP. Other phthalates such as diethyl phthalate (DEP), DBP, BBP, and DOP were recognized to a similar extent (between 97 and 110%). MAbs were also produced by Goda et al. [149], and a direct ELISA has been developed. The LOD accomplished for DBP was 0.2 mg L–1 with a dynamic range of 0.2–4 mg L–1. DEHP was not recognized (CR<1%). BBP showed a cross-reactivity value of 155% and dipropyl phthalate (DPrP) and dipentyl phthalate (DPnP) cross-reacted 60 and 51%, respectively.

6 General Summary It has been widely demonstrated that immunochemical techniques offer a good alternative to conventional methodologies in many areas due to the high sensitivity and selectivity achieved for the antibodies toward the target analytes.

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In clinical and environmental analysis, their use has been broadly spread because of their sensitivity, specificity, and high sample processing capabilities. In the case of emerging pollutants of industrial origin, a variety of immunochemical methods and formats have been developed (RIA, ELISA, PFIA, FIIA, immunosensors, immunoaffinity extraction procedures, etc.). Several surfactants can readily be detected by these methods. However, although attempts have been made toward the immunochemical determination of NP, under suspicion because of its potential estrogenic activity, the methodologies reported also recognize its parent compounds. Immunochemical analytical methods for PCBs and PCDDs/PCDFs have been reported, showing good detectability levels. Most of the problems that arise while analyzing these substances are related to their lack of solubility in the aqueous-based systems of the immunochemical methods. However, protocols have been developed using organic solvents and water mixtures. BPA can be efficiently detected with the immunochemical methods available today, not only in water samples, but also in more complex biological matrices such as serum, with detection limits in the order of micrograms per liter or even lower. Regarding phthalates, more efforts must be directed toward the production of specific antibodies for the main phthalate DEHP, since the immunochemical techniques reported up to now do not recognize this congener. Unfortunately, some of the methods described in this chapter are not yet being used as regular screening and analytical methods in environmental control laboratories. A reason for this may lie in the lack of knowledge on the performance of these types of techniques by certain analytical sectors, and also in the lack of validated protocols for a wide range of sample matrices. Immunoassay methods may suffer from undesirable matrix effects that may lead to false positive or negative results. It is wrong to assume that the selectivity of the immunochemical method is sufficiently high to overcome nonspecific interactions of the antibodies with the matrix components. Rigorous evaluation of the performance of these methods on each sample matrix of interest, and the consequent establishment of appropriate sample treatment methods, are required to ensure reliability and to convince control laboratories of the efficiency of these techniques. A close collaboration and interchange of the expertise of analytical chemists and immunochemists are needed to accomplish this goal and benefit from the advantages of these methods, to assess risk and protect public health from the adverse effects of these types of pollutants. Acknowledgements This work has been supported by CICYT (BIO2000-0351-P4-05, AGL2001-5005-E) and by the EC: nanotechnology and nanosciences, knowledge-based multifunctional materials, new production processes and devices (contract number NMP505485-1).

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228. Newsome WH, Shields JB (1981) Int J Environ Anal Chem 10:295 229. Luster MI,Albro PW, Clark G, Chae K, Chaudhary SK, Lawson LD, Corbett JT, McKinney JD (1979) Toxicol Appl Pharmacol 50:147 230. Diaz-Ferrero J, Rodriguez-Larena MC, Comellas L, Jimenez B (1997) TrAC Trends Anal Chem 16:563 231. Niessner R (1993) Anal Method Instrum 1:134 232. Mapes JP, McKenzie K, Stewart TN, L.R. M, Studabaker WB, Manning WB, Friedman SB (1993) Bull Environ Contam Toxicol 50:219 233. Goon DJW, Nagasawa HT, Keyler DE, Ross CA, Pentel PR (1994) Bioconj Chem 5:418 234. Reeves SG, Siebert STA, Roberts MA, Durst RA (1995) TrAC Trends Anal Chem 14:351 235. Kusterbeck AW, Wemhoff GA, Charles PT,Yeager D, Bredehorst R,Vogel C-W, Ligler FS (1990) J Immunol Methods 135:191 236. Concejero MA, Galve R, Abad E, Herradon B, Marco MP, Rivera J, Gonzalez MJ, De Frutos M (2000) Organohalogen Compd. 45:37 237. Concejero MA, Galve R, Herradon B, Gonzalez MJ, M. Df (2001) Anal Chem 73:3119 238. Harrison RO, Eduljee GH (1999) Sci Total Environ 239:1 239. Behnisch PA, Hosoe K, Sakai S (2001) Environ Int 27:413 240. Albro PW, Luster M, Chae K, Chaudhary K, Clark G (1979) Toxicol Appl Pharmacol 50:137 241. Kennel SJ, Jason C, Albro PW, Mason G, Safe SH (1986) Toxicol Appl Pharmacol 82:256 242. Vanderlaan M, Stanker LH, Watkins B (1988) Environ Sci Technol 7:859 243. Harrison RO, Carlson RE (1998) Organohalogen Compd 35:43 244. Harrison RO (2001) CAPE Technologies, South Portland, ME, USA 245. Harrison RO, Carlson RE (1997) Organohalogen Compd 40:31 246. Harrison RO, Carlson RE (1999) Organohalogen Compd 40:31 247. Harrison PTC, Carlson RE (1997) Organohalogen Compd 31:139 248. Shelver WL, Larsen GL, Huwe JK (1998) J Chromatogr B Biomed Sci Appl 705:261 249. Shelver WL, Huwe JK, Stanker LH, Patterson DG, Turner WE (2000) Organohalogen Compd 45:33 250. Shelver WL, Huwe JK (1999) J Liq Chromatogr Relat Technol 22:813 251. Shelver WL, Huwe JK, Stanker LH, Larsen GL (1999) Orgahalogen Compd 40:53 252. Shelver WL, Shan G, Gee SJ, Stanker LH, Hammock BD (2002) Anal Chim Acta 457:199 253. Huwe JK, Shelver WL, Stanker LH, Patterson DG, Turner WE (2001) J Chromatogr B Biomed Sci Appl 757:285 254. Shelver WL, Huwe JK (1998) Organohalogen Compd 35:13 255. Kolic TM, Reiner EJ, MacPherson KA, Harrison PTC (1998) Proceedings of the second biennial international conference on chemical measurement and monitoring of the environment, Ottawa 256. Fielding M, Barcelo D, Helweg S, Galassi L, Tortensson L, Van Zonen P, Wolter R, Angeletti G (1992). Pesticides in ground and drinking water. Water pollution research report 27. Commission of the European Communities, Brussels 257. Gerlach RW, Van Emon JM (1997) Chemosphere 35:2727 258. Oubina A, Puig D, Gascon J, Barcelo D (1997) Anal Chim Acta 346:49 259. Castillo M, Oubina A, Barcelo D (1998) Environ Sci Technol 32:2180 260. Staples CA, Woodburn K, Caspers N, Hall AT, Klecka GM (2002) Hum Ecol Risk Assess 8:1083 261. Sonnenschein C, Soto AM (1998) J Steroid Biochem Mol Biol 65:143 262. Krishnan AV, Stathis P, Permuth SF, Tokes L, Fieldman D (1993) Endocrinology 132:2279 263. Kim JC, Shin HC, Cha SW, Koh WS, Chung MK, Han SS (2001) Life Sci 69:2611 264. Lindholst C, Pedersen KL, Pedersen SN (2000) Aquat Toxicol 48:87

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265. Nagel SC,Vom Saal FS, Thayer CA, Dhar MG, Boechler M,Welshons WV (1997) Environ Health Perspect 105:70 266. Steinmetz R, Michner NA, Grant A, Allen DL, Bigsby RM, Ben-Johnathan N (1998) Endocrinology 139:2741 267. Yokota H, Tsuruda Y, Maeda M, Shima Y, Tadokoro H, Nakazono A, Honjo T, Kobayasji K (2000) Environ Toxicol Chem 19:1925 268. Braun P, Moeder M, Schrader S, Popp P, Kuschk P, Engewald W (2003) J Chromatogr A 988:41 269. Brossa L, Pocurull E, Borrull F, Marce RM (2002) Chromatographia 56:573 270. Tsuda T, Suga K, Kaneda E, Ohsuga M (2000) J Chromatogr B Biomed Sci Appl 746:305 271. Zhao MP, Li YZ, Guo ZQ, Zhang XX, Chang WB (2002) Talanta 57:1205 272. Zhao M, Liu Y, Li Y, Zhang X, Chang W (2003) J Chromatogr B Biomed Sci Appl 783:401 273. Ohkuma H, Abe K, Ito M, Kokado A, Kambegawa A, Maeda M (2002) Analyst 127:93 274. Kodaira T, Kato I, Li J, Mochizuki T, Hoshino M, Usuki Y, Oguri H, Yanaihara N (2000) Biomed Res 21:117 275. DeMeulenaer B, Baert K, Lanckriet H, VanHoed V, Huyghebaert A (2002) J Agric Food Chem 50:5273 276. Ius A, Bacigalupo MA, Meroni G, Pistillo A, Roda A (1993) Fresenius J Anal Chem 345:589 277. Nishii S, Soya Y, Matsui K, Ishibashi T, Kawamura Y (2000) Bunseki Kagaku 49:969 278. Report on carcinogens, 10th edn. US Department of Health and Human Services, Public Health Service, National Toxicology Program (2002) http://ehp.niehs.nih.gov/roc/ 279. Angelidaki I, Mogensen AS, Ahring BK (2000) Biodegradation 11:377 280. Dine T, Luyckx M, Gressier B, Brunet C, Souhait J, Nogarede S, Vanpoucke J, Courbon F, Plusquellec Y, Houin G (2000) Med Eng Phys 22:157 281. Hill SS, Shaw BR, Wu AHB (2001) Clin Chim Acta 304:1 282. Kambia K, Dine T, Azar R, Gressier B, Luyckx M, Brunet C (2001) Int J Pharm 229:139 283. Api AM (2001) Food Chem Toxicol 39:97 284. Kavlock R, Boekelheide K, Chapin R, Cunningham M, Faustman E, Foster P, Golub M, Henderson R, Hinberg I, Little R (2002) Reprod Toxicol 16:453 285. Kavlock R, Boekelheide K, Chapin R, Cunningham M, Faustman E, Foster P, Golub M, Henderson R, Hinberg I, Little R (2002) Reprod Toxicol 16:529 286. Kavlock R, Boekelheide K, Chapin R, Cunningham M, Faustman E, Foster P, Golub M, Henderson R, Hinberg I, Little R (2002) Reprod Toxicol 16:721 287. Wine RN, Li L, Barnes LH, Gulati DK, Chaplin RE (1997) Environ Health Perspect 105:102 288. Li L-H, Jester WF, Orth JM (1998) Toxicol Appl Pharmacol 153:258 289. Moore NP (2000) Reprod Toxicol 14:183 290. Jobling S, Reynolds TRW, Parker M, Sumpter J (1995) Environ Health Perspect 103:582 291. Harris CA, Hentu P, Parker MG, Sumpter JP (1997) Environ Health Perspect 105:802 292. Davis BJ, Maronpot RR, Heindel JJ (1994) Toxicol Appl Pharmacol 128:216 293. Sohoni P, Sumpter JP (1998) J Endocrinol 158:327 294. Marttinen SK, Kettunen RH, Sormunen KM, Rintala JA (2003) Water Res 37:1385 295. Vikelsoe J, Thomsen M, Carlsen L (2002) Sci Total Environ 296:105 296. Kataoka H, Ise M, Narimatsu S (2002) J Sep Sci 25:77 297. Prokupkova G, Holadova K, Poustka J, Hajslova J (2002) Anal Chim Acta 457:211 298. Castillo M, Barcelo D (2001) Anal Chim Acta 426:253

The Handbook of Environmental Chemistry Vol. 5, Part O (2005): 181– 244 DOI 10.1007/b98616 © Springer-Verlag Berlin Heidelberg 2005

Immunochemical Determination of Pharmaceuticals and Personal Care Products as Emerging Pollutants M.-Carmen Estévez · Héctor Font · Mikaela Nichkova · J.-Pablo Salvador · Begoña Varela · Francisco Sánchez-Baeza · M.-Pilar Marco (✉) Department of Biological Organic Chemistry, IIQAB-CSIC, Jordi Girona 18–26, 08034 Barcelona, Spain [email protected] 1

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 183

2 2.1 2.2 2.3 2.4 2.5 2.6

Antibiotics . . . . Penicillins . . . . Chloramphenicol Tetracyclines . . . Sulfonamides . . Fluoroquinolones Macrolides . . . .

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Steroid Hormones Estrogens . . . . Androgens . . . . Gestagens . . . . Corticosteroids .

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4 Other Drugs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 233 4.1 Analgesics and NSAIDs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 234 4.2 Cytostatic Agents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 237 5

General Summary . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 238

References

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Abstract A review on immunochemical methods for the analysis of pharmaceuticals is presented. A broad range of pharmaceutical categories and personal care products may reach the aquatic environment after excretion through industrial, domestic, and hospital wastewater. With few exceptions pharmaceuticals for human medicine are not high-production chemicals and the expected environmental concentrations should be low. However, the use of some of these chemicals in veterinary medicine increases the probability that the concentration values in the aquatic environment might reach higher levels. On the other hand certain drugs with limited use are of concern because of their high pharmacological potency, which creates a risk even at trace levels. Attending to these considerations and to the potential human risks, this review focuses on antibiotics, hormones, analgesics, nonsteroidal anti-inflammatory drugs, and cytostatic agents. Although these procedures have only been applied to the analysis of environmental samples on a few occasions, immunochemical methods for several of these substances exist and some of them are commercially available due to their use in clinical laboratories and forensic medicine.

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Keywords Immunochemical techniques · Antibiotics · Steroid hormones · Analgesics · Cytostatic agents Abbreviations BSA Bovine serum albumin CAP Chloramphenicol CE Capillary electrophoresis COD Codeine CR Cross-reactivity E1 Estrone E2 b-Estradiol E3 Estriol EE2 Ethynylestradiol EIA Enzyme immunoassay ELIFA Enzyme-linked immunofiltration assay ELISA Enzyme-linked immunosorbent assay EMIT Enzyme-multiplied immunoassay technique ETIA Energy transfer immunoassay EW Evanescent wave FPIA Fluorescence polarization immunoassay GC Gas chromatography HPIAC High-performance immunoaffinity chromatography HPLC High-performance liquid chromatography IAC Immunoaffinity chromatography LIF Laser-induced fluorescence LOD Limit of detection MAb Monoclonal antibody MECC Micellar electrokinetic capillary chromatography MIAC Multi-immunoaffinity chromatography MOR Morphine MRL Maximum residue level MS Mass spectrometry MTX Methotrexate NSAID Nonsteroidal anti-inflammatory drug PAb Polyclonal antibody PEC Predicted environmental concentration PFIA Polarization fluoroimmunoassay PNEC Predicted no effect concentration PPCPs Pharmaceutical and personal care products RIA Radioimmunoassay SPIA Sol particle immunoassay SPR Surface plasmon resonance STPs Sewage treatment plants TIRF Total internal reflection fluorescence immunoassay WWTPs Wastewater treatment plants

Immunochemical Determination of Pharmaceuticals and Personal Care Products

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1 Introduction Pharmaceuticals and personal care products (PPCPs) are a diverse group of chemicals used internally or externally in the body of humans, domestic animals, and plants. They include substances such as human and veterinary drugs, diagnostic agents (i.e., X-ray contrast media), nutraceuticals (bioactive food supplements), certain feed and food additives, sunscreen agents, fragrances, cosmetic additives, etc. Their presence in the environment is determined by their worldwide frequent use by multitudes of individuals or by their use as veterinary drugs. Thousands of tons of pharmacologically active substances are used yearly to treat illnesses, to prevent unwanted pregnancy, or to face the stress of modern life. For example, about 50,000 drugs are registered in Germany for human use, 2,700 of which are responsible for 90% of the total consumption and which, in turn, contain about 900 different active substances [1]. In the UK approximately 3,000 active substances are licensed [2]. In animal farming the use of antibiotics, feed additives, and hormones has become a usual practice for certain farmers. Substances such as the UV filters used for sunscreens and cosmetics, designed to remain on the skin, are generally washed off either during bathing or swimming, or are transferred to towels or clothes which will be finally washed. All these are examples of the usual use and release to the environment of PPCPs. Several reviews discuss and present real data on the environmental occurrence of the most relevant PPCPs [1, 3–7]. We must moreover remark that their use is continuing and escalating at the same time as new arrays of more potent chemicals are being introduced onto the market. Pharmaceuticals are inherently biologically active and remarkably potent agents. Often they are resistant to biodegradation, as a certain metabolic stability is necessary for their pharmacological action. PPCPs are released into the environment unaltered or as still active metabolites. Thus, human and veterinary drugs are frequently excreted as glucuronide or sulfate conjugates that can easily be hydrolyzed to release again the active parent compound in the environment. Certain PPCPs or their metabolites are highly soluble in water. This parameter combined with a lack of biodegradability limits the removal of PPCPs in wastewater treatment plants (WWTPs) [4, 8].Additionally, antibiotics and disinfectants are supposed to disturb the wastewater treatment process and the microbial ecology in surface waters. Furthermore, resistant bacteria may be selected in the aeration tanks of STPs (sewage treatment plants) by the antibiotic substances present [1]. In these cases the drugs enter the aquatic environment, contaminating ground and surface waters and eventually also reaching drinking water. Some other PPCPs are more lipophilic, thus showing a tendency to bioaccumulate in organisms that fortuitously will be used for human consumption. The consequence is that PPCPs enter the environment resulting in real exposure of humans and wildlife to these active substances [5, 9–12]. The concern produced within the scientific community, the authorities, and governmental bodies has prompted the establishment of certain regulations

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[13, 14]. Thus, before any new veterinary pharmaceutical product can obtain a marketing authorization, a review must be carried out by national or European Union (EU) authorities to ensure its efficacy, quality, and safety to public health and the environment. The requirements for ecotoxicity testing are regulated by Directive 2001/82/EC. The European Agency for the Evaluation of Medicinal Products (EMEA), which coordinates the evaluation and supervision of medicinal products for both human and veterinary use, has published guidance documents on how to perform environmental risk assessment of these products [15]. The International Cooperation on Harmonization of Technical Requirements for Authorization of Veterinary Medicinal Products (VICH) formed by the EU, USA, and Japan (Australia and New Zealand participate as observers) are trying to establish uniform risk assessment criteria [16]. The general idea is to predict the environmental concentration (PEC) of these substances based on calculations such as number of prescriptions, doses, degradation, environmental models, etc. [12]. The PEC value is then compared to the lowest effective concentration found, for the parent compound or its metabolites, in certain ecotoxicity tests performed in soil and/or water to establish what is known as the predicted no effect concentration (PNEC). The ratio PEC/PNEC should be lower than 1. Otherwise a risk to the environment is assumed and risk mitigation measures should be linked to the authorization of the product [17]. The complete guidance is still being developed, but already some data dealing with these parameters have been published on the most frequently used pharmaceuticals (i.e., [18, 19] and personal care products (i.e., [17]). In a study performed in Denmark, taking only pharmaceuticals prescribed for human medicine into consideration, ibuprofen, acetylsalicylic acid, and paracetamol exceeded the PEC/PNEC reference value [18]. Similar conclusions were drawn for paracetamol, amoxicillin, oxytetracycline, and mefenamic acid in a study on the aquatic environmental assessment of the top 25 English most prescribed pharmaceuticals [19]. Because of their inherent bioactivity, trace levels of these substances can have a negative impact on the environment and public health. This fact, together with the above-mentioned considerations, gives rise to substantial analytical problems. The different formats and benefits that the immunochemical methods can confer on the analysis of trace contaminants in aqueous-based samples has been widely demonstrated [47–56] (see also Chap. 5 in this volume). For this particular case immunochemical techniques offer an additional benefit derived from the possibility of developing single or class-specific methods according to necessity. Environmental monitoring of PPCPs requires efficient methodologies able to detect trace levels of contamination caused by both parent compounds and their metabolites. The present chapter describes some of the immunochemical methods available today for the analysis of the most important PPCPs regarding their potential impact on the environment (see Table 1). The PPCPs treated in this chapter have been selected according to their production and use. Because of the recent concern about their environmental impact, the aim of most of the immunochemical techniques reported for ana-

Human medicine

Ofloxacin

– High levels in hospital WW [4, 5]

– WW effluents in Switzerland 45–120 ng L–1 in [4] – high level in hospital WW [4] – US streams: 0.12 mg L–1 [20]

– WW effluents (Switzerland): 249–405 ng L–1 [4] – Hospital effluents: 3–87 mg L–1 [4, 5] – High level in hospital WW [4] – US streams: 0.2 mg L–1 [20]

Environmental Occurrence

It’s strongly sorbed onto soil. Persistent in the environment [1]

It’s strongly sorbed onto soil. Persistent in the environment [1]

It’s strongly sorbed onto soil. Persistent in the environment [1]

Environmental Fate

DW: drinking water; GW: ground water; SW: surface water; WW: waste water; STP: sewage treatment plant; WWTP: wastewater treatment plant.

Human medicine

Norfloxacin

Uses/Origin

Human medicine

Chemical Structure

Ciprofloxacin

Fluoroquinolones

Antibiotics

Substance

Table 1 Summarized table with the more important PPCPs divided into antibiotics, steroid hormones, and other drugs. Their generic chemical structures and the use or origin are shown. Some reported data regarding their environmental occurrence and the more probable environmental fate are also given

Immunochemical Determination of Pharmaceuticals and Personal Care Products 185

Veterinary medicine Human medicine

Veterinary medicine Aquaculture Human medicine

Veterinary medicine Human medicine

Sulfadimethoxine

Sulfamethazine/ Sulfadimidine

Sulfathiazole

Uses/Origin

Veterinary medicine Aquaculture Human medicine

Chemical Structure

Sulfamethoxazole

Sulfonamides

Substance

Table 1 (continued)

– Denmark landfill leaches: 0.04–6.47 mg L–1 [5]

Non degradable in sewage treatment

Non degradable in sewage treatment

Non degradable in sewage treatment

– US streams: 0.06 mg L–1 [20] – GW of a landfill from waste pharmaceutical production: 5 mg L–1 (total sulfonamides) [23] – High concentrations in landfill leaches in Denmark [4, 5] – US streams: 0.22 mg L–1 [20] – GW in Germany: 10–100 mg L–1 [21]

Non-degradable in sewage treatment

Environmental Fate

– GW: Up to 410 ng L–1 [4, 21, 22] – STP effluents in Germany: 0.40 mg L–1 [21] – River waters: 1 mg L–1 [23] – US streams: 0.15 mg L–1 [20]

Environmental Occurrence

186 M.-C. Estévez et al.

Veterinary medicine Human medicine

Veterinary medicine Human medicine

Veterinary medicine Human medicine

Penicillin V

Oxacillin

Ampicillin

Uses/Origin

Veterinary medicine Human medicine

Chemical Structure

Penicillin G

Penicillins

Substance

Table 1 (continued)

Hydrolysis in water of the b-lactam ring

Hydrolysis in water of the b-lactam ring

– SW in Germany: 0.26 ng L–1 [25]

Hydrolysis in water of the b-lactam ring

– River water: 25 ng L–1 [24] – Potable water 10 ng L–1 [24]

Not detected in the environment

Hydrolysis in water of the b-lactam ring

Environmental Fate

Not detected in the environment

Environmental Occurrence

Immunochemical Determination of Pharmaceuticals and Personal Care Products 187

Aquaculture Veterinary medicine Human medicine

Aquaculture Human medicine Growth promoter in veterinary medicine

Tetracycline

Chlortetracycline

Veterinary medicine Human medicine

Uses/Origin

Aquaculture Veterinary medicine Human medicine

Chemical Structure

Oxytetracycline

Tetracyclines

Cloxacillin

Penicillins

Substance

Table 1 (continued)

Accumulates in sewage sludges or sediments. It forms stable complexes with Ca2+. Persistent in anoxic conditions Accumulates in sewage sludges or sediment Degradation rate of tetracycline in liquid manure is approximately 50% in 5 months Photodescomposition Accumulates in sewage sludges or sediment After 30 days at 33 °C, 44% remained

– River water: 1 mg L–1 [3] – US streams: 0.11 mg L–1 [20]

– SW in US: 0.42 mg L–1 [20] – River water: 1 mg L–1 [3]

Hydrolysis in water of the b-lactam ring

Environmental Fate

– Detected in molluscs and wild fish in Norway 189–285 mg L–1 in sediments in fish farming [26] – River water: 1 mg L–1 [3] – US streams: 0.34 mg L–1 [20]

Not detected in the environment

Environmental Occurrence

188 M.-C. Estévez et al.

Human medicine

Clarithromycin

Uses/Origin

Veterinary medicine Human medicine Aquaculture

Chemical Structure

Erythromycin

Macrolides

Substance

Table 1 (continued)

– STP effluents in Germany: 0.24 mg L–1 [23] – GW in Germany 0.24–0.87 mg L–1 [23] – River Po in Italy: 1.7 ng L–1 [27]

– GW in Germany: up to 200 ng L–1 [25] – SW in Germany: 0.15 mg L–1 [23] – STP effluents in Germany. 2.5 mg L–1 [23] – River water: 1 mg L–1 [3] – River Po in Italy: 3.2 ng L–1 [27]

Environmental Occurrence

More stability than erythromycin in acid conditions

Not degradable in the environment [3]. Labile in acid conditions

Environmental Fate

Immunochemical Determination of Pharmaceuticals and Personal Care Products 189

Chloramphenicol

Trimethoprim

Roxithromycin

Macrolides

Substance

Chemical Structure

Table 1 (continued)

Aquaculture Exceptional cases in human medicine Veterinary use is forbidden

Veterinary medicine Human medicine Mixture with sulfonamides

Human medicine

Uses/Origin

– Sewage and surface level: low at mg L–1 [4] – STP in germany: 0.56 mg L–1 [23]

– STP effluents (Germany): 0.32 mg L–1 [23] – US streams: 0.15 mg L–1 [20]

– US streams: 0.05 mg L–1 [20]

Environmental Occurrence

Hydrolysis of chloramphenicol glucuronide in chloramphenicol

Half-life >1 year

No data found

Environmental Fate

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Metabolite of estradiol

Metabolite of estradiol

Oral contraceptive

Estrone (E1)

Estriol (E3)

Ethynylestradiol (EE2)

Uses/Origin

Endogenous hormone

Chemical Structure

Estradiol (E2)

Estrogens

Hormones

Substance

Table 1 (continued)

Mean Half-life: 17 days. High sorption onto sediments

Average removal efficiency 96%. Degradation not reported

– Sewage effluent: <0.1–42 ng L–1 (different countries) [28] – SW <0.1–3.4 ng L–1

– Sewage effluent: <0.053–62 ng L–1 (different countries) [28, 29] – SW: <0.053–30.8 ng L–1

Mean Half-life: 3.0 days. Main product of degradation of estradiol

Mean Half-life: 2.8 days. Sorption in the sediments

Environmental Fate

– Sewage effluent: <0.1–220 ng L–1 [28] – SW <0.1–17 ng L–1

– Sewage effluent: <0.1–88 ng L–1 [28, 29] – SW <0.05–15 ng L–1

Environmental Occurrence

Immunochemical Determination of Pharmaceuticals and Personal Care Products 191

Growth promoter Anabolic steroid

Growth promoter Anabolic steroid

Methyltestosterone (MT)

Trenbolone (Tr)

Oral contraceptive

Uses/Origin

Endogenous hormone Growth promoter Anabolic steroid

Chemical Structure

Testosterone (T)

Androgens

Mestranol (MeEE2)

Estrogens

Substance

Table 1 (continued)

– After 5.5 months in soil fertilized with liquid manure: 0.16–0.10 mg kg–1

– Pond Water after treatment with MT food: <5 mg L–1 [32]

– Raw sewage in WWTP: 16 to 700 ng mL–1 – GW: 1.0 ng L–1 – Runoff water from manured fields: 215 ng L–1 [13]

– Effluents: <1–8 ng L–1 [30]

Environmental Occurrence

Traceable after 8 days of fertilization, not detectable after 40 days [31]

Phelps et al. [32] found that MT in the water returned to the background levels within one week after hormone administration

T runs off by leaching of aqueous solution from the soil [31]

Degraded in aerobic conditions in sludge (80%) and 7% hydrolyzed to EE2

Environmental Fate

192 M.-C. Estévez et al.

Anti-inflammatory agent

Anti-inflammatory agent

Cortisone

Cortisol

Corticosteroids

Oral Contraceptive

Norethindrone

Uses/Origin

Endogenous hormone

Chemical Structure

Progesterone

Gestagens

Substance

Table 1 (continued)

No data found

No data found

– River samples (UK) 17 ng L–1 [33] – Effluents: 8–20 ng L–1 [30]

– US streams: 0.11 mg L–1 [20]

Environmental Occurrence

Very stable [34]

Practically insoluble in water [34]

28% Biodegradation in 6 h after the plant treatment and completely in 24 h [33]

No data found

Environmental Fate

Immunochemical Determination of Pharmaceuticals and Personal Care Products 193

No data found

Anti-inflammatory agent Growth promoter

Dexamethasone

Environmental Occurrence

No data found

Uses/Origin

Anti-inflammatory agent Growth promoter

Chemical Structure

Betamethasone

Corticosteroids

Substance

Table 1 (continued)

Practically insoluble in water [34]

Practically insoluble in water. Very stable [34]

Environmental Fate

194 M.-C. Estévez et al.

Pain killer Antithrombotic agent

Metabolite of acetylsalicylic acid

Metabolite of acetylsalicylic acid

Aspirin (acetylsalicylic acid)

Salicylic acid

Gentisic acid

Uses/Origin

Mild analgesic Antiphlogistic

Chemical Structure

Paracetamol (acetominophen)

Analgesics

Others

Substance

Table 1 (continued)

Readily biodegradable

No data found

No data found

– Sewage effluents: 13 mg L–1 [36] – STP effluent, max: 0.14 mg L–1 and river and streams (Germany): 4.1 mg L–1 [35] – STP effluent (Berlin): 0.04 mg L–1 [37] – GW (Berlin) max: 1225 ng L–1 [37] – STP effluents: max: 0.59 mg L–1 and rivers and streams (Germany): 1.2 mg L–1 [35] – GW Berlin max: 540 ng L–1 [37]

Readily degradable after acclimatization

Environmental Fate

– Sewage effluent (England): 1 ng ml–1 [24] – STP effluent, max: 1.5 mg L–1 and river and streams water (Germany): 0.34 mg L–1 [35]

– US streams, max: 10 mg L–1 [20] – STP effluent, max: 6 mg L–1 (Germany) [35]

Environmental Occurrence

Immunochemical Determination of Pharmaceuticals and Personal Care Products 195

Antiphlogistic

Diclofenac

Uses/Origin

Anti-inflammatory agent Pain killer Antiphlogistic Anti-rheumatic

Chemical Structure

Ibuprofen

Analgesics

Substance

Table 1 (continued)

– Inherently biodegradable (<95% removed in WWTPs) [39]

Readily or inherently biodegradable; photolytic degradation [4, 42]

– Rivers: 15–500 ng mL–1 [38] – Effluent: up to 2 ng mL–1 [35, 38] – STP influents: 12–560 ng L–1; STP effluents (Greece): 10–365 ng L–1 [41] – River Aabach (Switzerland): 11–310 ng L–1 [42] – Lakes (Switzerland): <1–12 ng L–1 [42] – WWTP effluent (Switzerland): 0.99 mg L–1 [43]

Environmental Fate

– Different German rivers: 17–139 ng L–1 [38] – STP effluent max: 3.4 mg L–1 and river and streams water (Germany): 0.53 mg L–1 [35] – STP effluent (Berlin): 0.1 mg L–1 [37] – GW (Berlin): max: 200 ng L–1 [37] – STP influents up to 3 mg L–1 and STP effluent: 2 ng L–1, rivers and lakes (Germany): up to 8 ng L–1 [39] – Italian rivers: 90.6–92.4 ng L–1 [40] – Sewage effluent: 1.5, 0.87 and 85 mg L–1; SW: 2.7 mg L–1 [36]

Environmental Occurrence

196 M.-C. Estévez et al.

Chemical Structure

Cancer therapy treatment (chemotherapy)

Cancer therapy treatment (chemotherapy)

Cancer therapy treatment (chemotherapy)

Cyclophosphamide

Ifosfamide

Analgesic

Uses/Origin

Methotrexate

Cytostatic (antineoplastic) agents

Codeine

Analgesics

Substance

Table 1 (continued)

– Treated hospital effluent: 24 ng L–1 [44] – Oncologic hospital effluent: mean: 109 ng L–1 [46] – STP effluent (Germany) max: 2.9 mg L–1 [35]

– Treated hospital effluent from STP: 146 ng L–1 [44, 45] – STP effluent (Germany) max: 20 ng L–1 [35]

– River and potable water: <6.25 ng L–1 [33]

– US streams max: 0.019 mg L–1 median value: 0.012 mg L–1 [20]

Environmental Occurrence

No data found

Not degradable

Persistent

Not data found

Environmental Fate

Immunochemical Determination of Pharmaceuticals and Personal Care Products 197

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lyzing PPCPs has been the analysis of tissues or body fluids. In this case the detectability should be in accordance with the MRL (maximum residue level) established in the legislation (EC Regulation 2377/90). However, considering the complexity of the biological samples, their application to environmental water samples should be straightforward. The use of those formats using radioactive labels, initially developed for biochemical studies, should be avoided whenever possible due to the problems derived from handling and producing radioactive waste.

2 Antibiotics Antibiotics are chemical substances that are able to suppress or kill the growth of bacteria. They are extensively used in human and veterinary medicine as well as in aquaculture. They are administered in veterinary medicine for the treatment and control of infectious diseases such as mastitis, enteritis, peritonitis, and pneumonia. Moreover, certain antibiotic substances have also been used as growth promoters in food producing animals [57, 58].Antibiotic therapy began with the clinical use of sulfonamides in 1936 and was followed by the development of penicillins (1944), chloramphenicol (1947), tetracyclines (1948), and fluoroquinolones (1980) (see Fig. 1 for the chemical structures). Since 1950, and in parallel with the use of antibiotics in human medicine, veterinary use has provided control of diseases in animal farms. This was followed by their use as growth promoters in many countries, although at present this practice is becoming increasingly controversial. In Europe about 10,000 tons of antibiotics are consumed each year (FEDESA, the European Animal Health Association 1998) [59] (see Table 2).According to these data, 5,000 tons are due to veterinary purposes (3,500 tons prophylaxis and therapy, and growth promotion about 1,500 tons). The other half of production is used in medicine.Among the antibiotics used in veterinary practice, Table 2 Veterinary consumption of therapeutic antibiotics in Europe

Product group

% Share

Penicillins Tetracyclines Macrolides Aminoglycosides Fluoroquinolones Trimethoprim/sulfonamides Others Totala

9 66 12 4 1 2 6 100

a

Total consumption: 2,494 tonnes of active ingredient at 100% purity.

Immunochemical Determination of Pharmaceuticals and Personal Care Products

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Fig. 1 Chemical structures of some of the most important antibiotics used nowadays divided into the most representative families: fluoroquinolones, sulfonamides, penicillins, macrolides, and tetracyclines. Another important antibiotic, chloramphenicol, is also shown

200

M.-C. Estévez et al.

penicillins and tetracyclines (also applied in aquaculture) and macrolides are the most frequently administered, whereas in humans fluoroquinolones, macrolides, and aminoglycosides are the most frequently used. The amount of antimicrobial agents used in food animals (cattle, chickens, pigs, and turkeys) in the United States is unknown.At least 17 classes of antimicrobial agents are approved for growth promotion and feed efficiency in the United States, including tetracyclines, penicillins, macrolides, lincomycin (analog of clindamycin), and virginiamycin (analog of quinupristin/dalfopristin). To understand the human health consequences of the use of antimicrobial agents in food animals, it is important to evaluate the quantity of antimicrobial agents used in food animals in the USA. Unfortunately, although reporting systems have recently been implemented in several European countries, no reporting system exists for the quantity of antimicrobial agents used in food animals in the United States. The Animal Health Institute, which reportedly represents 80% of the companies that produce antimicrobial agents for animals in the USA, has estimated that their member companies produced 18 million pounds of antimicrobial agents for therapeutic and nontherapeutic (growth promotion and disease prevention) use in food animals in the USA in 1999 [60]. An alternative report, provided by the Union of Concerned Scientists in 2001, estimated that 29 million pounds of antimicrobial agents are used in food animals annually in the USA, of which 25 million pounds are used for nontherapeutic purposes [61]. Though more precise data on the quantity of antimicrobial agents used in food animals are needed, these initial estimates provide some perspective on the quantity of antimicrobial agents used in food animals in the USA. The most important impact of the misuse of antibiotics is related to the development of resistance mechanisms. Antimicrobial resistance may be viewed as the ability of microorganisms of a certain species to survive or even to grow in the presence of a concentration of an antimicrobial that is usually sufficient to inhibit or kill bacteria of the same species. In the presence of an antimicrobial, organisms with inherent or acquired resistance to the agent will be selected. The bacterial population then comes to consist largely or entirely of resistant bacteria, causing failure of the traditional treatments. It has been reported that more than 70% of bacteria are insensitive against at least one antibiotic. This situation is causing a serious threat for public health, as more and more infections can no longer be treated with the presently known antidotes [62–68]. The World Health Organization has recommended that, unless a risk-based evaluation demonstrates their safety, the growth promotion use in food animals of antimicrobial agents that belong to the same classes of antimicrobial agents used in humans should be terminated [69]. Similar recommendations to discontinue the use of human antimicrobial agents as growth promoters in food animals have been made by several independent organizations in the United States, including the Alliance of Prudent Use of Antibiotics in 2002 [70] and the distinguished Institute of Medicine of the National Academies in 2003 [71]. For this reason, the EU has established the principle of using different antibiotics for

Immunochemical Determination of Pharmaceuticals and Personal Care Products

201

humans and animals. Since 1999 the EU has also banned some antibiotics such as tylosin, spiramycin, virginiamycin, and bacitracin, used as growth promoters (Council Regulation (EC) No. 2821/98), due to their structural relatedness to antimicrobial agents used in human medicine. After administration in humans or animals, these substances pass to the environment, mainly to the aqueous compartment resulting in some high local concentrations (e.g., aquaculture, hospital effluents). Several studies have been carried out in the USA, Germany, Switzerland, and Denmark to investigate the occurrence and fate of the antibacterial drugs in STPs or surface waters (see Table 1). Antibiotic resistance causes an important impact on the ecosystem, water, and soil-dwelling organisms. Moreover, some antibiotics can also produce adverse effects in animals and plants. For example, sulfadimethoxine and bacitracin produce loss of weight in roots and leaves in some plants, and oxytetracycline and tetracycline can kill pinto bean plants at a concentration level of 160 mg L–1 [3]. Chloramphenicol can produce pneumonia and sulfamethazine has been evaluated by the WHO/FAO Expert Committee as a suspected carcinogen. Macrolide antibiotics (clarithromycin, dehydroerythromycin, etc.) and sulfonamides (sulfamethoxazole, sulfadimethoxine, sulfamethazine, and sulfathiazole) are the most prevalent antibiotics found in the environment with levels around a few micrograms per liter, whereas fluoroquinolones, tetracyclines, and penicillins have been detected in fewer cases and usually at low concentrations (nanograms per liter) [3, 20, 23, 72]. This result is not surprising, since penicillins are easily hydrolyzed and tetracyclines readily precipitate with cations such as calcium and are accumulated in sewage sludge or sediments. Several reviews have reported the environmental occurrence of different antibiotics in aquatic and soil compartments. Some of these data are detailed in Table 1. Various techniques based on completely different principles have been used to detect antibiotic residues. Traditionally, most of the tests used take advantage of the antibacterial activity of the antibiotics. These growth inhibition tests have been used in different animal matrices; however, they detect all residue levels of any antibiotic above the MRL and no conclusion may be drawn about the identity of the antibiotic or its concentration [73]. On the other hand, HPLC and GC are highly specific but require extensive sample preparation, sophisticated equipment, and skilled laboratory personnel. Therefore they cannot be used for routine screening of a large number of samples [21, 23, 72, 74]. Immunochemical techniques can be excellent tools to assess contamination of the environment by antibiotics in different matrices due to their high detectability and specificity. Furthermore, immunoassays are excellent tools for screening large numbers of samples in short time periods. Table 3 summarizes some of the immunochemical techniques reported for the detection of several families of antibiotics.As mentioned in the introduction, usually these techniques have been developed to determine antibiotic levels in biological matrices, however their availability opens the door to further applications in analyzing environmental samples.

a

Biacore Q (optical biosensor) Biacore Q (optical biosensor) Biacore 1000 RIA IAC–ELISA ELISA Biacore 1000 system

Biles Water Urine Tissues Porcine bile

0.023 mg mL–1 5 mg L–1 0.17 ng mL–1 10 ng g–1 0.015 mg mL–1

1.39 ng mL–1

Muscle

7.4 ng g–1 0.041 mg mL–1

Serum

Swine liver Milk Honey Buffer

Bovine milk Ovine kidney

Buffer

Serum

10 ng mL–1

2.5 mg L–1

ELISA

6 ng mL–1 35.5 mg L–1 88.0 mg L–1

IC50

Matrix

[83] [84] [85] [86] [87]

[82]

[81]

[78] [79] [79] [80]

[77]

[76]

[75]

References

EIA: enzyme immunoassay; ELISA: enzyme-linked immunosorbent assay; ELIFA: enzyme-linked immunofiltration assay; IAC: immunoaffinity chromatography; SPIA: sol particle immunoassay; SPFIA: solid-phase fluorescence immunoassay; SPR: surface plasmon resonance; RIA: radioimmunoassay.

Sulfamethazine

Sulfathiazole

4.0 g kg–1

7.3 mg

1 ng mL–1

ELISA

Norfloxacin

L–1

10 pg mL–1 c

LOD

Sensitivity

ELISA ELISA

ELISA

Sarafloxacin

Sulfonamides

ELISA

Immunochemical techniquea

Ciprofloxacin

Fluoroquinolones

Analyte

Table 3 Some immunochemical techniques developed for the detection of antibiotics

202 M.-C. Estévez et al.

Biacore Q (optical biosensor) Biacore 1000 system

SPIA

Sulfadiazine

Sulfadimidine

Biacore SPR

Ampicillin

Biacore SPR Biacore SPR RIA SPFIA (Parallux kit)

Biacore SPR

Penicillin G

Penicillin M

ELISA (LacTek kit) qualitative

SPFIA (Parallux kit)

Automated flow-through amperometric IA

Cephalexin

Pencillins

ELISA

Immunochemical techniquea

Sulfadimethoxine

Sulfonamides

Analyte

Table 3 (continued)

Milk Milk Water Bovine and porcine kidney Milk

30 mg L–1

Buffer Milk Bovine and porcine kidney Plasma

5.9 mg L–1 12.5 mg L–1 50 mg L–1

5–7 mg L–1 2.6 mg kg–1 2 mg L–1 50 mg L–1

Raw milk

1 mg mL–1

Urine Milk

Muscle

Liver tissue

Porcine bile

48.8 g L–1 71.6 mg L–1

1.50 mg L–1

10 ng mL–1 20 ng mL–1

g–1

IC50

Matrix

0.028 mg mL–1

5.6 ng

LOD

Sensitivity

[91]

[91] [94] [84] [92]

[93]

[91] [91] [92]

[90]

[89]

[87]

[82]

[88]

References

Immunochemical Determination of Pharmaceuticals and Personal Care Products 203

100 mg kg–1

ELISA (TC Microwell test kit)

RIA ELISA ELISA

ELISA

Tylosin

4 ng mL–1

10 mg 0.4 ng mL–1 0.3 ng mL–1

L–1

8 ng mL–1

0.1 ng mL–1

ELISA

Bovine muscle

Water Bovine muscle Buffer

Buffer

1 mg L–1

0.3 ng mL–1

Milk

20 mg kg–1

EIA ELISA (qualitative) RIA (Charm II RIA test)

ELISA

Bovine and porcine kidney Honey Pork meat Water

300 mg L–1

SPFIA (Parallux kit)

Erythromycin

Macrolides Macrolides

8 ng mL–1

Bovine and porcine kidney Muscle tissue

300 mg L–1

Oxytetracycline

SPFIA (Parallux kit)

RIA SPFIA (Parallux kit)

Tetracyclines

IC50

Matrix

Water Bovine and porcine kidney

LOD

Sensitivity

1 mg L–1 300 mg L–1

Immunochemical techniquea

Chlortetracycline

Tetracyclines

Analyte

Table 3 (continued)

[100]

[84] [100] [99]

[99]

[98]

[96] [95] [97]

[92]

[95]

[92]

[84] [92]

References

204 M.-C. Estévez et al.

Chloramphenicol

Chloramphenicol

Analyte

Table 3 (continued)

ELISA ELISA ELIFA Dipstick EIA ELISA (Le Carte Test) EIA kit (5091CAP1p)

EIA (RIDASCREEN)

RIA (Charm II assay)

Immunochemical techniquea

0.7 ng mL–1 17 ng mL–1 2 mg kg–1 0.1 mg kg–1

5 ng g–1 20 ng mL–1 0.5 ng g–1 0.3 ng mL–1

LOD

Sensitivity

3 ng mL–1 3 ng mL–1

IC50

Tissue Urine Tissue Urine Swine muscle tissue Muscle tissue Milk Milk Meat Shrimp tissue

Matrix

[101] [101] [101] [101] [102] [102] [103] [103] [104] [105]

References

Immunochemical Determination of Pharmaceuticals and Personal Care Products 205

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M.-C. Estévez et al.

2.1 Penicillins Penicillins are one of the most important families of antibiotics used in veterinary and human medicine. But due to their rapid transformation in environmental media (easy hydrolysis of the b-lactam ring), their persistence in environmental samples should be low. Thus, some works aimed at detecting antibiotic residues in water samples point out the absence of penicillin residues in spite of this drug being widely used [24, 25]. Several immunochemical methodologies have been developed for the detection of penicillins in food samples of animal origin [91, 93, 94, 106] (see Table 3). Most of them are based on the use of the commercial surface plasmon resonance (SPR) biosensor Biacore.A SPR is an evanescent electromagnetic field generated at the surface of a metal conductor (usually Ag or Au) when excited by the impact of light of an appropriate wavelength at a particular angle (qp). Surface plasmons are generated by electrons at the metal surfaces that behave differently from those in the bulk of the metal. These electrons are excited by the incident light, producing an oscillation (resonance) of different frequency from that in the bulk of the metal film. The absorption of light energy by the surface plasmons during resonance is observed as a sharp minimum in light reflectance when the varying angle of incidence reaches the critical value. The critical angle depends not only on the wavelength and polarization state of the incident light, but also on the dielectric properties of the medium adjacent to the metal surface and therefore is affected by analytes binding to that surface (see Fig. 2). Thus, when the immunocomplex is formed or dissociated a shift of the SPR angle is observed. The Biacore sensor was applied by Gaudin et al. to detect different penicillins in milk using commercial monoclonal antibodies (MAb) against ampicillin [91]. These MAbs had a higher affinity for the open b-lactam ring compounds than for those with the closed ring. Therefore, the analyses had to be performed by carrying out pretreatment of the samples in order to open the b-lactam ring

Fig. 2 Surface plasmon resonance (SPR) principle. Surface plasmons are excited by the light energy at a critical angle (q) causing an oscillation and the generation of an evanescent wave. Under this condition a decrease in the reflected light intensity is observed. The angle q depends on the dielectric medium close to the metal surface and therefore is strongly affected by molecules directly adsorbed on the metal surface. This principle allows the direct detection of the interaction of the analyte and the antibody

Immunochemical Determination of Pharmaceuticals and Personal Care Products

207

of the penicillins and accomplish acceptable limits of detection (for ampicillin: 5.9 mg L–1 in buffer and 12.5 mg L–1 in milk).With these antibodies a high crossreactivity (CR) was observed with penicillin G, penicillin V, amoxicillin, and cloxacillin. Gustavsson et al. [94] tried to improve the procedure of the Biacore using carboxypeptidase and antibodies against a hydrolyzed peptide generated by an enzymatic reaction. This method had the advantage of detecting only the intact b-lactam structure. Penicillins inhibit this enzyme and therefore the amount of penicillin present in the sample can be measured by a decrease in the concentration of the hydrolyzed peptide. This method could be applied to the analysis of penicillins in milk with limits of detection around few micrograms per liter. Moreover, several ELISA (enzyme-linked immunosorbent assay)-type immunoassays have also been described for analyzing penicillins in different matrices (see Table 3). Thus, the determination of penicillins in plasma samples from cattle by ELISA has been reported [93]. Monoclonal antibodies against ampicillin have been used to develop a direct ELISA and a multi-immunoaffinity chromatography method (MIAC) for penicillins [106]. Moreover, many immunoassay kits have become commercially available (see Table 4). Thus, the LacTek ELISA was established as a qualitative rapid prediction test of amoxicillin and ampicillin residues in tissues and, applying a modified methodology, also in milk. The test is performed in test tubes and takes 7 min to complete one analysis. Analyte and an enzyme tracer compete for an antibody coated on the tube wall. After washing, a color developer is added to visualize the surfacebound complex. The color intensity is measured in a spectrophotometer and compared with a penicillin standard indicating positive or negative results. At the moment there are also LacTek tests available to detect tetracyclines, sulfonamides, and chloramphenicol. The Charm II 6600/7600 is a semiquantitative radioimmunoassay (RIA) developed to analyze b-lactam antibiotics in multiple food matrices (tissue, urine, milk, honey, etc). It is based on the use of 1H- and 14C-tagged drug tracers. After a step of competition between tracer antibiotic and sample residues for the antibody, the bound tracer–antibody complex is separated by centrifugation from the unbound tracer. The complex is analyzed in a scintillation counter for 1 min to obtain a resultant count. Samples with high-count results are considered negative while samples with low count are considered positive. The assay is very fast (about 10 min per sample) and the detection limit varies depending on the antimicrobial drug and the matrix. For example, penicillin is detected in milk at levels around 3.5 mg L–1 whereas, in the case of animal tissues, the sensitivity reach levels of 50 mg L–1. This test has also been applied to the analysis of other kinds of antibiotics such as tetracyclines, chloramphenicol, sulfonamides, and macrolides [84]. The Delvo X-Press and SNAP tests are cost-effective, rapid b-lactam immunoreceptor assays developed for the screening of cow’s milk before milk intake at the laboratories of the receiving stations. The Beta Screen Test employs a fluorescence endpoint and takes about 10 min per assay, showing a high speci-

Antibiotics

5101SUDA1p

Sulfadiazine SNAP test Biacore X LacTek Delvo X-Press Beta Screen Test Parallux Charm II 6600/7600

LacTek 5101SUL1p

Sulfamethazine

Penicillins

Charm II 6600/7600 Biacore X Parallux

Charm ROSA (Strip ELISA) Biacore prototype 5101ERFX1p

IA kit

Sulfonamides

Sulfonamides

Enrofloxacin

Fluoroquinolones

Analyte

IDEXX Laboratories Inc. Biacore IDEXX Laboratories Inc. Gist-Brocades BV Advanced Instruments Inc. IDEXX Laboratories Inc Charm Sciences Inc.

Euro-Diagnostica

IDEEX Laboratories Inc. Euro-Diagnostica

http://www.Idexx.com http://www.biacore.com http://www.Idexx.com http://www.dsm.com/ http://www.aitests.com/ http://www.idexx.com http://www.charm.com

http://www.elisa-tek.com

http://www.Idexx.com http://www.elisa-tek.com

http://www.charm.com http://www.biacore.com http://www.Idexx.com

http://www.biacore.com http://www.elisa-tek.com

Biacore Euro-Diagnostica Charm Sciences Inc. Biacore IDEEX Laboratories Inc.

http://www.charm.com

Contact

Charm Sciences Inc.

Supplier/manufacturer

Table 4 Some representative commercial immunochemical assay kits for the most important PPCPs. The supplier and the contact web page are also listed

208 M.-C. Estévez et al.

Antibiotics

Parallux Parallux Parallux Parallux

Tetracycline/ Penicillins

Sulfamethazine/ Tetracyclines

Sulfonamides/ Tetracyclines/ Penicillins

Charm II 6600/7600 LacTek RIDASCREEN ELISA kit 5091CAP1p Veratox

Charm II 6600/7600

Charm II 6600/7600 RIDASCREEN LacTek TC Microwell

IA kit

Sulfamethazine/ Penicillins

Multianalyte

Chloramphenicol

Other antibiotics

Macrolides

Tetracyclines

Analyte

Table 4 (continued)

IDEXX Laboratories Inc.

IDEXX Laboratories Inc.

IDEXX Laboratories Inc.

IDEXX Laboratories Inc.

Charm Sciences Inc. IDEXX laboratories Inc. Biopharm AG In Vitro Biologics Ltd. Euro-Diagnostica Neogen Corporation

Charm Sciences Inc.

Charm Sciences Inc. Biopharm AG IDEXX Laboratories Inc. Idetek Inc.

Supplier/manufacturer

http://www.idexx.com

http://www.idexx.com

http://www.idexx.com

http://www.idexx.com

http://www.charm.com http://www.idexx.com http://www.r-biopharm.com http://www.invitrobiologics.com http://www.elisa-tek.com http://www.neogen.com

http://www.charm.com

http://www.charm.com http://www.r-biopharm.com http://www.idexx.com

Contact

Immunochemical Determination of Pharmaceuticals and Personal Care Products 209

Hormones

EE2 EIA kit Estrogen ELISA kit ES EIA kit

Biosense Laboratories

EE2 ELISA kit

Ethynylestradiol (EE2)

Estrogens (E1+E2+E3)

Japan EnviroChemicals, Ltd

E1 EIA kit

Estrone (E1)

Japan EnviroChemicals, Ltd Biosense Laboratories

Biosense laboratories

Oxford Biomedical Research Atlas Link

Assay Designs Inc.

Estriol EIA immunoassay kit Estriol ELISA kit Estriol EIA kit

Estriol (E3)

Immunometric Ltd. Japan EnviroChemicals, Ltd. Assay Designs Inc. Biosense Laboratories Atlas Link Oxford Biomedical Research

Supplier/manufacturer

Estradiol EIA kit 17b-Estradiol ELISA kit 17b-Estradiol EIA kit E2 EIA kit Estradiol EIA kit Estradiol ELISA kit

IA kit

Estradiol (E2)

Estrogens

Analyte

Table 4 (continued)

http://www.biosense.com/

http://www.biosense.com/

http://www.biosense.com/

http://www.atlaslink-inc.com

http://www.atlaslink-inc.com http://www.oxfordbiomed.com/

Contact

210 M.-C. Estévez et al.

Hormones Testosterone EIA kit Salivary assay test Colorimetric EIA kit Testosterone ELISA kit Androstenedione ELISA kit Trenbolone ELISA test kit Nortestosterone ELISA test kit Kit for stanozolol, ELISA Stanozolol ELISA test kit

Androstenedione

Trenbolone

Nortestosterone

Stanozolol

IA kit

Testosterone

Androgens

Analyte

Table 4 (continued)

In Vitro Biologics Ltd

Tecna

InVitro Biologics Ltd

InVitro Biologics Ltd

Oxford Biomedical Research

Immunometric Ltd. Salimetric Assay Designs Inc. Oxford Biomedical Research

Supplier/manufacturer

http://www.invitrobiologics.com

http://www.invitrobiologics.com/

http://www.invitrobiologics.com/

http://www.oxfordbiomed.com/

http://www.oxfordbiomed.com/

Contact

Immunochemical Determination of Pharmaceuticals and Personal Care Products 211

Hormones

Corticosteone EIA kit ELISA test kit Bio-X corticosteroids ELISA kit

Corticosterone

Dexamethasone

Corticosteroids (family) Progesterone EIA kit Progesterone ELISA kit Progesterone EIA kit EIA kit

Progesterone

17a-OH Progesterone

Gestagens

Cortisol EIA kit Cortisol EIA test kit Cortisol EIA kit

IA kit

Cortisol

Corticosteroids

Analyte

Table 4 (continued)

Assay Designs Inc.

Assay Designs Inc. Oxford Biomedical Research Atlas Link

Bio-X Diagnostics

InVitro Biologics Ltd

Assay Designs Inc.

Assay Designs Inc. Atlas Link Oxford Biomedical Research

Supplier/manufacturer

http://www.assaydesigns.com

http://www.atlaslink-inc.com

http://www.oxfordbiomed.com

http://www.biox.com/

http://www.invitrobiologics.com

http://www.assaydesigns.com

http://www.atlaslink-inc.com http://www.oxfordbiomed.com

Contact

212 M.-C. Estévez et al.

Others

Methotrexate

Cytostatics FPIA (Adx, TDxFLx, AxSYM Systems) EMIT

FPIA (Adx, TDxFLx, AxSYM Systems)

http://abbott.com http://syva.com

Syva Corporation

http://abbott.com

http://www.cozart.co.uk http://neogen.com

http://syva.com

http://www.diagnostix.ca http://www.cozart.co.uk http://www.stech.com

Contact

Abbott Laboratoires

Abbott Laboratoires

Cozart Bioscience Neogen

Microplate EIA ELISA

Salicylate

Syva Corporation

EMIT

Opiates (morphine, codeine, cocaine, etc.)

Diagnostix, Inc. Cozart Bioscience Ora-Sure/STC Technologies, Inc.

Supplier/manufacturer

Single-step ELISA Microplate EIA Microplate EIA

IA kit

Fenantyl

Analgesics

Analyte

Table 4 (continued)

Immunochemical Determination of Pharmaceuticals and Personal Care Products 213

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ficity for penicillin. Finally, the Parallux, a solid-phase fluorescence immunoassay (SPFIA) intended for use as a rapid detection method (it takes only 4 min) in raw bovine milk, can detect all six major b-lactam antibiotics in one test. This system has been developed as a multianalyte method to detect simultaneously the presence of a variety of antibiotics from different families (see Table 4). The kit is based on the use of antibodies immobilized on four glass capillary tubes presented in the form of a disposable cartridge. Two different kinds of cartridges are included: one consists of four tubes each containing the same range of antibodies (the “cillins” multicartridge), so that four different samples can be screened simultaneously. The other one also includes four tubes but each containing different antibodies (the “individual” cartridge), so that a positive sample can be further identified. The sample is mixed with dried reagents, and antibiotic competitively binds to the coated tube; after the tube is washed and the cartridge centrifuged, the fluorescence signal is measured with a limit of detection (LOD) for penicillin of 50 mg L–1. This test has also been applied to the detection of tetracyclines, sulfonamides, cephapirin, and ceftiofur. To our knowledge only one work has been reported on the use of a commercial immunochemical test to detect penicillins (penicillin G, penicillin V, ampicillin, cloxacillin, and oxacillin) in several environmental compartments. Thus, Campagnolo et al. [84] measured penicillin using the Charm II RIA test in water samples proximal to a US farm. The LOD of the technique was 2 mg L–1. 2.2 Chloramphenicol Chloramphenicol (CAP) is a broad-spectrum antibiotic that was widely used in veterinary medicine. Since 1994 the use of CAP is banned in the EU because of certain toxicological problems (i.e., aplastic anemia and the “grey baby syndrome”) observed in its administration to humans [107] that have prompted the establishment of a zero tolerance for the presence of these residues in meat and animal products.As a consequence, many efforts have been made to develop sensitive methodologies capable of detecting CAP residues or its metabolites. Several qualitative and quantitative immunochemical methods for CAP analysis in biological matrices of animal origin have been described [101, 102, 104, 105] (see Table 3).Van de Water et al. [102] described an ELISA that detected CAP in swine muscle tissue with an IC50 value of 3 ng mL–1. This immunoassay was improved and subsequently optimized incorporating the streptavidin– biotin amplification system. There are also several commercially available test kits (see Table 4). RIDASCREEN is a competitive enzyme immunoassay for the quantitative analysis of CAP residues in milk, eggs, and meat in a microtiter plate. The measurement is made photometrically, obtaining a LOD of 100 ng L–1 in meat and eggs and 150 ng L–1 in milk. The test has been also applied to the analysis of tetracyclines. On the other hand, the 5091CAP1p test is a direct competitive enzyme immunoassay for the quantitative analysis of these residues in milk, eggs, meat,

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urine, tissue, and honey. In this test in a 96-well microtiter plate is coated with antibodies. The LOD is 0.2 ng mL–1 in urine and milk, with a cross-reactivity of 65% with chloramphenicol glucuronide. Veratox is a semiquantitative CAP ELISA test designed to detect the presence of CAP in farm-raised shrimp. The test format provides a 48-well plate operating on the basis of competition between the enzyme conjugate and the antibiotic in the sample for the antibody-coated wells.After the addition of a colored substrate, a semiquantitative determination can be made of the level of drug present. Results are obtained in approximately 90 min showing low cross-reactivity to other antibiotics and a sensitivity of 2 ng g–1. Lynas et al. [101] compared the use of two immunochemical methods, Charm II RIA and the RIDASCREEN EIA (enzyme immunoassay), in tissues and fluids of treated cattle. Charm II reached LODs of 5 ng g–1 and 20 ng mL–1 in tissues and urine, respectively, whereas RIDASCREEN improved these values to 0.5 ng g–1 for tissues and 0.3 ng mL–1 in urine. One of the advantages of these methods is that the CAP metabolites are also detected. On the other hand, Keukens et al. [104] used the Le Carte test kit, a polyclonal antibody (PAb)based assay for the regulatory control of this antibiotic in meat with a LOD of 20 mg kg–1. The assay was also applied to urine samples with a detectability of 5 mg L–1. Moreover, Impens et al. [105] used the 5091CAP1p test for the screening of CAP in shrimp tissues and Van de Water et al. [108] developed a MAbbased cleanup procedure, prior to HPLC, for the analysis of CAP in eggs and milk. At present we have not found examples of the application of any of these immunochemical methods to environmental samples, but as mentioned in the introduction, it can be assumed that application to wastewater samples should produce fewer matrix effects than those produced by the biological samples. 2.3 Tetracyclines Tetracyclines are broad-spectrum antibiotics with activity against Gram-positive and Gram-negative bacteria that have been widely used for the treatment of infectious diseases in veterinary and human medicine, as well as additives in animal foodstuffs. Normally tetracyclines are not found at high levels in the environment because they readily precipitate with cations such as calcium and are accumulated in sewage sludge or sediments [3, 20, 26]. Immunochemical methods have been developed and placed on the market to analyze tetracycline residues (see Table 4). Thus, a qualitative EIA has been developed and used to analyze tetracyclines in honey samples with a detection level of 20 µg/kg–1 [96].A microplate-based indirect ELISA has been developed to analyze tetracyclines using polyclonal antibodies. The assay could measure tetracycline in the range between 0.1 and 6 ng mL–1. Other tetracycline antibiotics such as chlortetracycline, rolitetracycline, or minocycline are also highly recognized in this assay [98]. Several immunoassay kits are commercially available for the analysis of tetracyclines although, to our knowledge, none of them

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has been applied to the analysis of environmental samples (see Table 4). The RIDASCREEN EIA (see above for CAP) recognizes four tetracyclines (tetracycline, chlortetracycline, minocycline, and rolitetracycline). It has been used to analyze milk samples with a LOD around 1.5 mg L–1 for tetracycline and 15.5 mg L–1 for oxytetracycline. This kit can also be applied to the analysis of more complex matrices such as meat and honey. The Parallux system (see above for penicillins) has also become available to detect tetracyclines and sulfonamides (see below). This responds to the presence of tetracycline, chlortetracycline, and oxytetracycline with sensitivities of 100, 125, and 100 mg L–1, respectively. Recently, Okerman et al. [92] described the use of this kit on bovine and porcine kidneys with a sensitivity of 300 mg L–1. Moreover, De Wasch et al. [95] reported the use of the TC Microwell test kit to analyze oxytetracycline in pork meat samples, reaching sensitivity levels of around 100 mg kg–1. The use of the Charm II RIA test to analyze tetracycline antibiotics in water (both surface and groundwater) has been reported [84, 97]. This RIA, which was initially developed to analyze tetracycline in serum, urine, and milk, was subsequently adapted to analyze water samples at concentration levels around 1 mg L–1. Thus, samples from hog lagoons, surface water samples, and groundwater samples were tested using the RIA method and the results confirmed by LC–MS. 2.4 Sulfonamides Sulfonamides are antibiotics widely used in animal husbandry and as feed additives. A large number of immunoassay screening methods for sulfonamides in foods and other related complex matrices have been reported in the literature [78, 81, 85, 88, 89, 96, 109] (see Table 3 for immunochemical methods.). Lee et al. [78] described the development of ELISAs with a series of MAbs that can detect sulfathiazole in animal tissues with IC50 values ranging from 6 to 21 ng mL–1 in swine liver samples. Verheijen et al. [89] described the development of a sol particle immunoassay (SPIA) for the detection of sulfadimidine residues at the qualitative level. This kind of immunoassay is based on the use of dyed colloidal particles as labels (i.e., gold, carbon, or latex). The device here reported is a one-step strip test where the antigen is immobilized on the membrane, and is based on the use of affinity-purified PAbs anti-sulfadimidinelabeled with colloidal gold particles in the mobile phase. The use of colored particles as labels allows their use as a direct detector reagent. This test has been applied to the analysis of urine and milk samples, giving positive results at concentrations above 10 and 20 ng mL–1, respectively. Situ et al. [87] evaluated the performance of a high-throughput SPR to simultaneously analyze eight samples and also to detect sulfamethazine and sulfadiazine in porcine bile in an online system. Most of the antibodies developed (both monoclonal and polyclonal) only detect individual sulfonamides. However, due to the widespread use of sulfon-

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amides and the variety of congeners that can potentially be used, several attempts have been made to produce antibodies showing broader specificity [80, 110–112]. Thus, Spinks et al. [110] carried out molecular modeling studies of the hapten structure in order to accomplish this aim, although no conclusive results were obtained. Korpimäki et al. [80] studied the use of protein engineering to modify MAbs so that they would recognize a wider range of sulfonamides with similar affinities, and have achieved a significant improvement in this aspect when compared to the wild-type antibody. As occurred with the other antibiotics, commercial immunoassay formats, also available as kits for tetracyclines and penicillins such as the Parallux, the LacTek, or the Charm II, have also been placed on the market for the analysis of sulfonamides (see Table 4). Thus, the Parallux detects sulfamethazine and sulfadimethoxine in raw milk with a LOD of 10 mg L–1. The Charm II detects almost all sulfonamides in honey and milk with a LOD in the range from 1 to 10 mg L–1, whereas LacTek is able to detect sulfamethazine. Moreover, the 5101SUL1p and 5101SUDA1p tests reach LOD values for sulfamethazine and sulfadiazine of around 0.2 mg L–1 and they have been applied to the analysis of urine, milk, and plasma. These tests have proved to be efficient as a point of care for “on-site” applications on farms. Moreover, commercially available antibodies can be found from several sources such as Silver Lake Research, US Biological, Cortex Biochem. Inc., Accurate Chemical Scientific, Fitzgerald Industries International Inc., and Biotrend Chemikalien GmbH. We have found only one attempt to use immunoassays to detect sulfonamides in environmental samples.As in the case of penicillins and tetracyclines and also for fluoroquinolones (see below), Campagnolo et al. [84] measured sulfonamides in water samples proximal to a farm in Iowa using a commercial Charm II RIA test, accomplishing a LOD of 5 mg L–1 for sulfamethazine. 2.5 Fluoroquinolones Fluoroquinolones have mainly been used in human medicine but more recently some fluoroquinolones (enrofloxacin, sarafloxacin, ciprofloxacin) are also being employed in veterinary medicine. Conventional methods like spectrofluorometric assays and HPLC have normally been used to detect fluoroquinolones in human samples, although some immunoassays have also been described [75–77, 113] (see Table 3). Thus, the production of MAb and the development of an indirect ELISA against sarafloxacin have been reported [76, 113]. The IC50 ranged from 7.3 to 48 mg L–1 depending on the MAb used. The antibodies obtained for sarafloxacin were also able to recognize enrofloxacin, difloxacin, norfloxacin, and trovafloxacin. With the same MAbs a high-performance immunoaffinity chromatography (HPIAC) was developed that consisted of the extraction of the analyte by using the immunoaffinity capture columns coupled online to a reversed-phase liquid chromatography system. The optimized procedure was applied to the detection of fluoroquinolones in serum [114], chicken liver [115],

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and milk [116]. The relative affinity of these immobilized MAbs toward the different fluoroquinolones allowed the gradual elution, separation, and individual quantification of two fluoroquinolones [117]. Snitkoff et al. [75] reported the development of an EIA for the detection of ciprofloxacin in serum, which was sensitive at picogram per milliliter levels of the antibiotic and no cross-reaction with its metabolites was observed. Gobbo et al. [118] recently described the production of PAb for ciprofloxacin with the aim of detecting fluoroquinolones in Brazilian livestock. On the other hand, Bucknall et al. [77] produced antibodies for quinolones and fluoroquinolones with the aim of developing both generic and specific immunoassays. ELISAs for ciprofloxacin, enrofloxacin, flumequine, and nalidixic acid were developed with sensitivity values around 4 mg kg–1 (on both the generic and specific assays) in bovine milk and ovine kidney. Regarding commercially available immunochemical kits, we could mention the Charm ROSA Enrofloxacin Test that detects ciprofloxacin and enrofloxacin equally (see Table 4) and the 5101ERFX1p test. This last one is a direct competitive ELISA, which uses MAbs and has a LOD of 3 ng g–1 in tissues. Some other companies do have antibodies available as reagents for different applications such as Biodesign International and QED Bioscience Inc. We have only found one example of the application of an immunoassay kit to the analysis of fluoroquinolones in environmental samples [84]. The assay is able to detect enrofloxcin as standard analyte with sensitivity levels of 5 mg L–1. 2.6 Macrolides Macrolide antibiotics are macrocyclic lactones widely used in veterinary medicine to treat diseases and infections and also as feed additives to promote animal growth. Some immunochemical methods have been developed to analyze macrolides, such as a RIA for erythromycin A and its chemical by-products [119] or an ELISA with a broad range of recognition for macrolide antibiotics [99] (see Table 3). As mentioned earlier, the RIA has the disadvantage of handling and producing radioactive residues. The ELISA developed by Yao et al. uses PAbs and has a LOD as low as 0.3 ng mL–1 with an IC50 value of 8 ng mL–1. Macrolides with 12-, 14-, or 16-carbon rings possessing amino-substituted sugar moieties are well recognized by these antibodies, regardless of the presence of neutral sugar residues. In contrast, little or no cross-reactivity was observed with the acrocyclic lactone ring structure (tylactone) or macrolides containing only neutral sugar. Also a Japanese patent [120] reported the production of antibodies against 16-membered-ring macrolide antibiotics such as ricamycin, midecamycin, josamycin, leucomycin A7, and leucomycin V. Regarding sensors, Draisci et al. [100] reported the development of an electrochemical competitive ELISA for the detection of erythromycin and tylosin in bovine muscle. They used MAbs against these two macrolides and the activity

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of an enzyme label was electrochemically measured with an electroactive substrate. The detection limits were 0.4 and 4 ng mL–1 for erythromycin and tylosin, respectively. The specificity of the assay showed only cross-reactivity with roxithromycin, a macrolide derived from erythromycin. The Charm II 6600/7600 system is the only commercial immunoassay test available for the detection of several macrolides in different matrices (see Table 4). The LOD of erythromycin in milk is 40 mg L–1 and 100 mg L–1 in tissues. Several common methods for the detection of these compounds in environmental media have been proposed such as microbiological assay and conventional chromatographic methods. However, only one example of the application of immunochemical methods to the analysis of environmental samples has been reported [84]. In this case, erythromycin could be measured at concentrations higher than 10 mg L–1 using the Charm II 6600/7600 assay in water samples proximal to livestock farms.

3 Steroid Hormones Steroid hormones are a group of biologically active compounds controlling human body functions related to the endocrine system and the immune system. Steroids are synthesized from cholesterol and have in common a cyclopentan-o-perhydrophenanthrene ring. Natural steroids are secreted by the adrenal cortex, testis, ovary, and placenta in humans and animals, and include progestagens, corticoids, androgens, and estrogens [121]. Estrogens (estradiol, estrone, and estriol) are predominantly female hormones which are important for maintaining the health of the reproductive tissues, breast, skin, and brain, are secreted from the ovary, and work to make an ovulatory phase of menstrual cycles. Gestagens (progesterone) act as hormone balancers, particularly of estrogens; they are excreted from the ovary and act during the luteinizing phase of the menstrual cycle and are involved in maturation of the endometrium of the uterus. Androgens (testosterone, dehydroepiandrosterone, and androstenedione) play an important role in tissue regeneration, especially of the skin, bones, and muscles. Glucocorticoids (cortisol) are produced by the adrenal glands in response to stressors such as emotional upheaval, exercise, surgery, illness, and starvation [122]. All the steroid hormones cause their action by passing through the plasma membrane and binding to intracellular receptors. Besides these endogenous hormones, many synthetic steroids have been produced in order to take advantage of their high bioactivity. Table 5 summarizes both natural and the most important synthetic steroids for each group as well as their more usual applications either in veterinary or human medicine; their chemical structures can be seen in Fig. 3. As well as the endogenous steroids, the xenosteroids and their metabolites are excreted by humans and animals in the form of glucuronide or sulfates [31]. These steroids end up in the environment through sewage discharge and ani-

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Table 5 Most important natural and synthetic steroid hormones

Endogenous hormones

Synthetic hormones

Use

Androgens

Testosterone

Boldenone Nandrolone Methyltestosterone Trenbolone Stanozolol

Growth promoter in livestock

Estrogens

Estradiol Estriol Estrone

Ethynylestradiol Mestranol

Contraceptive

Progestagens

Progesterone 17a-Hydroxyprogesterone

Norgestrel Norethindrone Melengestrol Medroxyprogesterone

Contraceptive

Corticosteroids

Cortisone Cortisol

Dexamethasone Betamethasone Deoxymethasone Prednisone Prednisolone Flumethasone

Growth promoter Anti-inflammatory

mal waste disposal. All of these compounds have been detected in effluents of STPs and surface waters. The water solubility of steroid conjugates is higher than that of the free forms and they are also more easily metabolized by degrading bacteria. The octanol/water partition coefficient of steroids is high and therefore adsorption on sediments and suspended soils and accumulation take place. These compounds have a half-life of between 2 and 6 days in the environment depending on the environmental conditions and the climate [28]. The fate and occurrence in the environment of the different steroid hormones are shown in Table 1. The unconjugated compounds are the active forms responsible for a potential environmental impact, especially in the aquatic compartment, at very low concentration levels [123]. Compounds like the estrogens have an endocrine disrupting activity, causing adverse effects such as the induction of vitellogenesis (plasma vitellogenin induction) and feminization of male fish [28]. For instance, concentrations between 4.7 and 7.9 ng L–1 of estradiol led to induction of vitellogenin in juvenile rainbow trout [124]. Therefore, vitellogenin induction in male or juvenile fish has become a useful biomarker for identifying estrogenic contamination of the aquatic environment [125]. Regarding androgens, they can induce a masculinization of the female sexual organs. The administration of androgens could be made in the free form or as esters (mainly acetate) [126]. After administration, the ester compound is hy-

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Fig. 3 General structures of the most important natural and synthetic steroid hormones

drolyzed to the free form under phase I metabolism, leading to the active compound. As a result of the continuous growth of the population and of livestock farming, the level of endogenous hormones excreted into the environment has gradually increased. However, the nonethical human and veterinary practices related to the use of the natural and synthetic sex hormones as anabolic substances and growth promoters are more worrying. For instance, it has been reported that about 33 tons of estrogens, 7.1 tons of androgens, and 322 tons of gestagens are excreted per year by livestock in the EU [31]. These data don’t include the synthetic steroids, such as ethynylestradiol or norethindrone, which are commonly used as oral contraceptives. The use of hormones, both natural and synthetic, to enhance growth and as reproductive aids for synchronization of the ovarian cycle, has been regulated for animal drugs because they alter the structure or function of the animal. For these reasons, the EU has banned the use of these compounds as growth promoters in food-production animals (Directive 85/649/EEC replaced by Directive 88/146/EEC). In order to detect the

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use of legal or illegal natural hormones or xenobiotic drugs, and to prevent the inappropriate use of therapeutic drugs, veterinary and public health control laboratories require efficient screening methods. Directive 96/23/EC and the Directive 2377/90/EC regulate the MRLs and the analytical methods to detect them. Hormone implants are widely used in the USA, Australia, and Canada where their use is allowed. The use of progesterone, testosterone, estradiol, zeranol, and trenbolone acetate for animal food production has been regulated by the US Food and Drug Administration (FDA) and by the Food and Agriculture Organization of the World Health Organization (FAO/WHO). The number of analytical methodologies currently available for determination of steroids in water samples is limited. The methodologies are based on either biological or chromatographic techniques [30, 127, 128], which are usually accurate but time-consuming methodologies. Immunoassays can be extremely sensitive and often can be directly applied to the analysis of water samples. Many examples can be found in the literature regarding immunochemical determination of steroid residues in biological matrices. In contrast, their application to environmental samples, and particularly wastewater, has rarely been reported. However, as mentioned before their application to less complex matrices such as aqueous samples can open the possibility to perform more and more efficient controls of the contamination of the environment by these groups of substances. We will briefly describe the most important immunochemical methods reported for the most relevant steroids that can be detected in the environment (see Table 1). Although usually the antibodies and immunochemical methods developed can recognize different congeners of the same family, few examples of real multianalyte (several families screened simultaneously) procedures have been described. It is only worth mentioning the case of immunoaffinity methods of extraction. Thus, some of these columns have been prepared not only to selectively extract a single family of steroids, but also to extract a larger number of analytes using a multi-immunoaffinity chromatographic column (MIAC). Different antibodies are linked to the solid support, allowing extraction of several steroids at once and therefore multianalyte-screening procedures using immunochemical or conventional analytical methods. The convenience of these procedures for biological and environmental monitoring programs of veterinary drug residues is that often, animal treatments are performed using cocktails of different substances (i.e., anabolic steroids with corticosteroids, or estrogens with gestagens). Dubois et al. [129] used MIAC combined with GC–MS detection for the screening of different anabolic steroids in urine and feces from bovine specimens. Information about 12 different compounds could be obtained in just one run. The MIAC gel columns were prepared by mixing individual gels prepared with different specific antibodies (against methyltestosterone, nortestosterone, fluoxymesterone, zeranol, clostebol, ethynylestradiol, diethylstilbestrol, and trenbolone). The fractions selectively eluted were evaporated to dryness, dissolved in ethanol, derivatized, and injected into the GC–MS system. The MIAC gel could be regenerated with mixtures of methanol/

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water, water, and finally PBS and stored at 4 °C. More examples can be found in the literature regarding the use of multi-immunoaffinity columns, for example against different anabolic steroids [129–131]. 3.1 Estrogens Due to their endocrine disrupting activity estrogens have been the most studied regarding residues in the environment. Table 6 shows some examples of the large number of different immunological screening methods reported for estrogens.A first RIA was described in 1985 for the analysis of ethynylestradiol in water [33], and also a RIA was employed for the detection of 17b-estradiol in wastewater [132]. This method allows rapid, sensitive, and inexpensive screening of a large number of samples. However, as already mentioned, the major disadvantage of RIA is that it requires radioisotopes and scintillation fluids. Huang et al. [159] used an ELISA to quantify estrogenic hormones in wastewater effluents and surface water. The LODs accomplished levels around 0.1 ng L–1 in wastewater effluents and 0.05 ng L–1 in surface waters. Results indicated that the concentrations of the estrogenic hormones 17b-estradiol and 17a-ethinylestradiol discharged by WWTPs were comparable to those that induce vitellogenesis in fish. Some hormones appear to be removed by effluent filtration, of which >95% of estrogenic hormones are removed by reverse osmosis. Compared to GC–MS/MS, the ELISA method had lower detection limits and was less susceptible to matrix interference. This produced a certain discrepancy in the results obtained by both methods that was attributed to the fact that the concentrations measured were near to the LOD of the GC–MS/MS method. Another technique that has been applied to detect 17b-estradiol in wastewater is an electrochemical ELISA [133]. The activity of the label enzyme (horseradish peroxidase) was measured electrochemically using 3,3¢,5,5¢-tetramethylbenzidine (TMB) as electrochemical substrate, accomplishing a LOD of 5 pg mL–1. The interday and intraday precision (RSD) ranged from 1 to 3% and from 3 to 6%, respectively. Analysis of wastewater from three different treatment plants demonstrated the absence of matrix effects if an extraction with diethyl ether–water was performed or the samples were just diluted 1:1 with buffer. Validation of the method was performed by analyzing 36 samples and comparing the results with those obtained by LC–ESI-MS/MS (liquid chromatography–electrospray ionization-tandem mass spectrometry). The results correlated very well with an R2 of 0.960. Commonly the development of these techniques has been directed toward the detection of the estrogen family instead of an individual compound. For instance, Goda et al. [136] developed different ELISAs for several hormone disrupting chemicals (HDCs) and one of them, based on commercial MAbs, is addressed to the detection of total natural estrogens (ES): estrone (E1), estradiol (E2), and estriol (E3). Depending on the MAb used and considering E2 as standard analyte, several assays were developed with working ranges within

ELISA

ELISA

Estrogens

Ethynylestradiol

a

TIRF-IA ETIA

Estrone

Wastewater effluent Surface water Buffer Buffer Water samples

0.1 ng L–1 0.05 ng L–1 0.07 mg L–1 0.01 mg L–1 5 ng L–1

1.07 kg L–1 2.7 mg L–1

Buffer

0.1 mg L–1

Buffer Buffer

0.07 mg L–1 0.5 mg L–1

Wastewater Buffer Buffer Wastewater effluent Surface water Buffer Saliva Buffer

0.51 mg L–1 0.81 mg L–1

1.84 mg L–1 1.2 mg L–1

IC50

Matrix

12 pg per well 10 pg per well

5 pg mL–1 0.16 mg L–1 0.85 mg L–1 0.1 ng L–1 0.05 ng L–1 0.1 mg L–1

LOD

Sensitivity

[135] [135] [134] [134] [33]

[136]

[134] [134]

[137] [138]

[133] [134] [134] [135] [135] [136]

References

EIA: enzyme immunoassay; ELIFA: enzyme-linked immunofiltration assay; ELISA: enzyme-linked immunosorbent assay; ETIA energy transfer immunoassay; RIA: radioimmunoassay;TIRF-IA: total internal reflection fluorescence immunoassay.

TIRF-IA ETIA RIA

ELISA Chemiluminescence ELISA

ELISA

ELISA TIRF-IA ETIA ELISA

Immunochemical techniquea

Estriol

Estradiol

Estrogens

Analyte

Table 6 Immunochemical techniques developed for the detection of steroid hormones

224 M.-C. Estévez et al.

ELISA

ELISA ELISA

ELISA ELISA

Boldenone

Trenbolone

Nandrolone

ELISA ELISA RIA

EIA

Progesterone

Norethindrone

Gestagens

ELISA ELISA ELISA

Immunochemical techniquea

Testosterone

Androgens

Analyte

Table 6 (continued)

Plasma Human serum Water Water

10 ng L–1

Equine urine Bovine bile

1 ng mL–1 >2 ng mL–1 0.5 nmol L–1 3.8 pg per tube 5 ng L–1

Meat samples Urine Muscle tissue

0.1 mg L–1 0.1 ng mL–1 0.02 ng g–1

Human serum Human serum Human serum Urine Faeces

IC50

Matrix

26 pg per well 0.1 ng g–1

3.9 pg mL–1 2.5 pg per well 10 pg per well

LOD

Sensitivity

[33]

[147] [148] [33]

[145] [146]

[143] [144] [144]

[142] [142]

[139] [140] [141]

References

Immunochemical Determination of Pharmaceuticals and Personal Care Products 225

Direct luminescence EIA

ELISA ELISA

ELISA

ELISA

ELISA EIA ELIFA ELISA ELISA ELISA

ELISA

Cortisol

Prednisolone

Betamethasone

Dexamethasone

Flumethasone

Immunochemical techniquea

Cortisone

Corticosteroids

Analyte

Table 6 (continued)

Saliva Human serum Faeces Equine urine

1.4 nmol L–1 2.8 ng mL–1 0.1 ng g–1 12.5 ng mL–1

Equine urine Plasma Equine urine Urine Urine and blood Equine urine Equine urine

3.1 ng 0.51 ng mL–1 390 ng mL–1 4 ng mL–1 0.01 ng mL–1 2 ng mL–1 2.5 ng mL–1

mL–1

Human blood plasma

IC50

Matrix

2 nmol L–1

LOD

Sensitivity

[153]

[153] [154] [155] [156] [157] [158]

[153]

[152]

[150] [151]

[149]

References

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0.1–5 mg L–1. The cross-reactivity measurements were 100% for E2, 87% for E1, and 55% for E3. The authors claim the possibility of obtaining a global value for estrogens in a particular environmental sample. Coille et al. [134] described the use of two fluorescence immunochemical methods to detect different estrogenic compounds in wastewater: TIRF (total internal reflection fluorescence immunoassay) and ETIA (energy transfer immunoassay). The first is an immunosensor technique based on the evanescent wave (EW) phenomenon using fluorescence detection. The antigen is bound to the transducer surface, which usually is an optical fiber inside a flow cell, and interacts with a fluorescent compound-labeled antibody. Light travels by the fiber optic by total internal reflection and the associated EW interacts with the immobilized antigen. As a consequence of the biorecognition reaction with the labeled antibody, a change is produced in the features of the light that is traveling (see Fig. 4). The ETIA works under homogeneous conditions. In this format the antibody is labeled with a donor fluorescent dye, whereas the Ag is labeled with an acceptor dye via a BSA molecule. When they form the complex a quenching of the fluorescence of the labeled antibody by energy transfer is observed. In the presence of the analyte an increase in the fluorescence is produced. The LODs obtained by TIRF for estrone, estradiol, and ethynylestradiol were 0.07, 0.16, and 0.07 mg L–1, respectively, whereas using ETIA the corresponding detectability accomplished was 0.5, 0.85, and 0.01 mg L–1. Immunoaffinity purification procedures have also been described for estrogens and applied to environmental and biological samples [160–163]. Ferguson et al. [160] described a method, based on immunoaffinity extraction coupled to LC–ESI-MS, for the determination of the steroid estrogens b-estradiol (E2), estrone (E1), and a-ethynylestradiol (EE2) in wastewater. The use of highly selective immunosorbents for sample preparation prior to the analysis allowed the removal of interfering sample matrix components present in the wastewater extracts that would otherwise cause severe ionization suppression of the estrogens during the electrospray process. The authors claim that the use of immunoextraction removed much of the isobaric noise from the selected ion monitoring chromatograms, increasing the signal-to-noise ratios and improv-

Fig. 4 Waveguide evanescent wave (EW) principle. Light is propagated through the waveguide (n1) and an electromagnetic field (called EW) is generated in the external medium (n2). The EW interacts with immobilized molecules that absorb energy, leading to attenuation in the reflected light of the waveguide

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ing the detectability of the analytical method (0.18 and 0.07 ng L–1 for E2 and E1, respectively). The optimized method was applied to the analysis of estrogens in two wastewater effluents. Recoveries of E2 and E1 were excellent (>90%), while EE2 was not retained (recovery <2%) from effluent extracts due to its structural differences with the immunizing hapten. The precision of the method was high, with relative standard deviations below 5%. The concentrations of E2 found in wastewater were 0.77–6.4 ng L–1, while levels of E1 were higher (1.6–18 ng L–1). Farjam et al. [163] developed an immunoaffinity precolumn (immuno-precolumn) immobilizing antibodies directed against estrogen steroids on Sepharose. They evaluated different desorbing techniques, suitable for online coupling to an HPLC-UV system. The most effective approach used 95:5 methanol–water mixtures, although the use of cross-reactants to elute the target was also considered. The final system consisted of a column-switching unit allowing preconcentration of the samples on an immunoaffinity sorbent. After elution the analytes were concentrated on a C18-bonded silica precolumn, and then separated on a C18-bonded silica analytical column. The minimum concentration of estrogens detected in urine using this system was around 200 ng L–1 with a repeatability of 6–8%. The total analysis time was 45 min, which gave an estimation of about 30 analyses per day in this automated unattended method [164]. Besides these works, there are a lot of commercially available immunoassays. Their main application is directed toward clinical analysis and food quality. Table 4 shows the most representative kits that can be found nowadays on the market. 3.2 Androgens The main areas of application of immunochemical techniques for androgens is doping control in athletes, forensic chemistry, farm animals for human consumption, and food analysis [165–167]. Immunochemical methods for androgen detection have been applied to a great variety of matrices (see Table 6); however, to our knowledge, their application in the environment field has not yet been recorded. One of the anabolic steroids widely used is 19-nortestosterone. Different ELISAs have been developed for the analysis of this compound in feeds, food from animal origin, and for doping control [145, 146]. Similarly, ELISA methods have also been developed for other anabolic steroids such as boldenone and trenbolone, used as growth promoters [142, 144]. In the case of testosterone, several immunochemical studies have been addressed to establish the real physiological levels of this hormone in different animals.A highly sensitive microplate-based direct ELISA was developed to analyze testosterone levels in human serum [139]. The specificity and accuracy of the assay were established, demonstrating negligible cross-reactivity with other related steroids. Previously other ELISA methods had been reported to analyze testosterone in human plasma [140, 141]. Thus, Rassie et al. [140] developed a PAb-based direct

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immunoassay performed in microplates with a sensitivity of 2.5 pg per well. The assay was very specific for testosterone and did not show any cross-reaction with other related C19 steroids tested. Replacement of immunoassay plates by polypropylene tubes raised the detection limits to 25 pg per tube, but improved the range of testosterone that could be measured up to 10,000 pg. Similarly, Rao et al. [141] developed a direct immunoassay on microplates using PAbs and penicillinase as tracer. The LOD was 10 pg of testosterone with a dynamic range between 15 and 1,000 pg. No interference was produced by other common androgens, estradiol, or progesterone, whereas a low level of cross-reactivity with 5a-dihydrotestosterone (6.2%) and 11b-hydroxytestosterone (1%) was observed. Specific antibodies for androgen compounds have also been used to develop immunoaffinity columns in order to include a purification step prior to the analysis of androgens such as stanozolol [168], methyltestosterone [169], testosterone, trenbolone, and nortestosterone [170]. In the case of methyltestosterone, a comparison made between XAD solid-phase extraction and immunoaffinity procedures showed that immunoaffinity could be more efficient in isolating and concentrating anabolic steroids from complex matrices such as urine and serum [169].An immunoaffinity precolumn packed with Sepharose-immobilized PAbs against 17b-19-nortestosterone (b-19-NT) was used for the selective online pretreatment of raw extracts of urine, bile, and tissue samples followed by HPLC-UV detection (247 nm) [171]. b-19-NT and its metabolite 17a-19nortestosterone (a-19-NT) could be detected in these biological samples with detection limits around 0.05 mg kg–1. Using the same immunosorbent 17b- and 17a-trenbolone were also detected. By percolating high sample volumes it was possible to confirm these results by GC–MS. Farjam et al. [172] also evaluated an immunoaffinity purification procedure coupled to GC. The immunosorbent contained antibodies raised against the synthetic steroid hormone b-19-NT. The online connection between the immunoaffinity precolumn and the capillary GC was performed with an interface that consisted of a 10¥2-mm reversed-phase precolumn and a diphenyltetramethyldisilazane-deactivated GC retention gap. After preconcentration on the immunoaffinity precolumn the analytes were eluted and reconcentrated on the reversed-phase precolumn. Subsequently, this precolumn was desorbed with 75 mL of ethyl acetate, which was directly introduced into the retention gap by using partially concurrent solvent evaporation. The system allowed the automated pretreatment and GC analysis of 19-norsteroids at the nanogram per liter level. A variety of immunoassay test kits for androgens are available from commercial sources. Table 4 indicates some of the most important companies commercializing these assays. 3.3 Gestagens As with estrogens and androgens, several commercial immunoassay kits are nowadays available for the analysis of gestagens [173–175] (see Table 4). Sim-

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ilarly, different research groups have also made efforts to develop assays and to demonstrate the performance of those assays in a variety of sample matrices (see Table 6). Thus, Aherne et al. described different immunoassays for detecting natural and synthetic steroids in water [33]. Norethindrone and progesterone were detected at concentrations above 10 and 5 ng L–1 using an EIA and a RIA, respectively. Results below the level of detection were obtained in all the samples examined (eight river samples and six potable supply samples), except for two river samples that contained norethindrone (17 ng L–1) and one river sample and one potable water sample that were positive for progesterone (6 ng L–1). They concluded that the presence of norethindrone in river water was caused by the low biodegradation of norethindrone in the usual 6-h water treatment processes. Thanks to these studies the authors found that a 24-h treatment of the sludge system was necessary. 3.4 Corticosteroids Corticosteroids are synthetic glucocorticoids that produce a strong anti-inflammatory effect. Corticosteroids such as dexamethasone are commonly used in veterinary practice for the treatment of illnesses such as respiratory and gastrointestinal disorders. However, corticosteroids are also used illegally as growth promoters in animal feed. The misuse of glucocorticoids in livestock production occurs, sometimes in combination with b-adrenergics in mixtures or cocktails to achieve growth promotion of food-producing farm animals. The aim is to reduce meat fat, to increase the appetite of the animals, and to increase the efficiency of the use of b-agonists. For consumer health and safety, the use of these compounds for this purpose is banned within the EU (Council Directive 86/469/EEC). Their therapeutic use is also regulated by the MRLs that have been established for dexamethasone, betamethasone, prednisolone, and methylprednisolone in different tissues (EC Regulation 2377/90). Owing to their high potency they are very effective in low doses, which results in low residue levels in biological matrices. Therefore, there is a requirement for sensitive analytical methods for the quantification and confirmation of these compounds in biological samples. Previously GC–MS methods have been used for the analysis of corticosteroids. However, recent developments in liquid chromatography–mass spectrometry technology have led to reports on the use of LC-based approaches for analysis of these compounds. The use of LC–MS reduces the analysis time due to elimination of the lengthy derivatization or oxidation procedures necessary for GC–MS methods. Efficient screening procedures based on ELISA methods have been described for the most important corticosteroids (see Table 6) [153, 176]. Rodriguez et al. [153] reported an ELISA using PAbs raised against flumethasone that also recognized several synthetic corticosteroids such as dexamethasone (CR=80%) and betamethasone (CR=20%), while endogenous corticosteroids such as cortisol gave very low cross-reactivity (<0.5%). This assay can be directly applied to

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1:10 diluted urine samples without hydrolysis of glucuronide or sulfate conjugates or any other treatment of samples. The PAbs were obtained by immunizing sheep with a flumethasone derivative linked to human serum albumin. Flumethasone, dexamethasone, and betamethasone could be detected at levels around 2.5, 3.1, and 12.5 ng mL–1, respectively. Similar results were obtained by Roberts et al. [158] with antibodies raised against dexamethasone using 21hemisuccinate dexamethasone coupled to BSA as immunizing hapten (Fig. 5). These Abs also recognized flumethasone, betamethasone, and deoxymethasone (see Table 7 for CR values). The immunizing hapten chemical structure strongly determines the specificity of the resulting antibodies. However, considering the structural similarities, the presence of common epitopes determines the high cross-reactivity observed in these assays (see Fig. 5 for chemical structures). A direct enzyme immunoassay for screening of synthetic glucocorticoids in biological samples was also developed by Meyer et al. [154], by raising antibodies against dexamethasone using the same immunizing hapten as Roberts et al. (21-hemisuccinate-BSA) and using prednisolone-21-hemisuccinate horseradish peroxidase as tracer. The system had an analog cross-reactivity pattern with similar CR values: dexamethasone (I) (100%), flumethasone (103%), betamethasone (45%), triamcinolone (18%), and prednisolone (17%). The natural glucocorticoids cortisone and cortisol, the gestagens progesterone and pregnenolone, and the androgen testosterone were not recognized (CR<0.4%). Recently, there

Fig. 5 Chemical structure of the hapten conjugated to BSA used as immunogen by Roberts et al. [158] and Meyer et al. [154] to raise antibodies against dexamethasone. Structures of other corticosteroids subsequently tested by Creaser et al. [180] in order to evaluate the selectivity of the ELISA obtained are also shown

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has also been an increasing interest in the use of saliva to detect drugs. In this context, Anfossi et al. [150] described the use of an ELISA for cortisol that achieves a LOD of 1.4 nmol L–1 in this matrix and recoveries from spiked samples of between 80 and 120%. An ELIFA (enzyme-linked immunofiltration assay) method has also been reported [155] for the rapid detection and semiquantification of dexamethasone in equine urine samples. The assay consists of an indirect competitive ELISA in which dexamethasone in standards or samples competes for the antibody binding with a dexamethasone–protein conjugate immobilized as a spot on the surface of a cellulose nitrate filter. The sheep anti-dexamethasone antibodies are complexed with an alkaline phosphatase-labeled second antibody. The filtration system allows rapid washing and incubation steps, so the signal could be visualized in just 15 min by an insoluble colored dye as a spot on the filter at the site of the immobilized drug–protein conjugate. The assay has a LOD of 390 ng mL–1 for a visual endpoint in which the color intensity of spots developed in the presence of samples is compared with those of standards. Twelve filters can be processed in a single batch consisting of two standards and ten samples. Immunoaffinity procedures have also been developed to selectively extract corticosteroids from different sample matrices. Thus, Seymour et al. demonstrated the higher efficiency of the immunoaffinity methods compared with the conventional extraction procedures using organic solvents [177]. Immunosorbents have also been used for online procedures followed by HLPC-UV [178, 179], HPLC–APCI-MS [179, 180], GC–MS [176, 181], or capillary electrophoresis [182]. Poly(hydroxyethyl methacrylate) (HEMA) was evaluated as a support material for the anti-dexamethasone antibodies used in IAC. The online IAC– HPLC–MS allowed determination of dexamethasone and flumethasone in equine urine with LODs in the range 3–4 ng mL–1 [180]. The cross-reactivity values obtained in the ELISA and the recoveries of an IAC–HPLC procedure are presented in Table 7. Bagnati et al. developed an immunoaffinity extraction Table 7 Cross-reactivities for selected corticosteroids using IAC–HPLC (reproduced from [180] with editor permission)

Corticosteroid

Dexamethasone Flumethasone Betamethasone Deoxymethasone Cortisol Prednisolone Antiserum batch: AD60.

Relative cross-reactivity (%) IAC–HPLC

ELISA

100 96 30 27 0 0

100 96 37 21 1 4

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method for dexamethasone and betamethasone in bovine urine, followed by HPLC fractionation and GC–MS detection [181]. The immunoaffinity cartridge was inserted in an automatic HPLC system for online extraction and purification. The purified collected fractions containing the analytes of interest were derivatized to yield the tetra-trimethylsilyl derivatives of the three corticosteroids, which were analyzed by gas chromatography–selected ion monitoring mass spectrometry. The method allowed a detection limit of 0.1 ng mL–1 for dexamethasone and 0.2 ng mL–1 for betamethasone.

4 Other Drugs This section is dedicated to providing information on the immunochemical methods available today to determine drugs that do not belong to the above groups, but that years of unrestricted emission to the environment require to be considered [183]. From the broad range of pharmaceuticals that can reach the environment, drugs such as analgesics and nonsteroidal anti-inflamatory drugs (NSAIDs) are regularly employed, often even without prescription. On the other hand, cytostatic agents are of concern not because of their production volume but for their high pharmacological potency. In Germany for instance the total quantities of acetylsalicylic acid sold per year have been estimated to be greater than 500 tons, 75 tons for diclofenac, and 180 tons for ibuprofen [35]. The same occurs in other EU countries where common drugs such as paracetamol or aspirin are sold in quantities comparable to high production volume materials – close to or exceeding 1,000 tons per year [184]. Ibuprofen, which is in the top ten list of pharmaceuticals used in Denmark in 1995, is used in yearly amounts of 33 tons and analgesics 28 tons [3]. Psychiatric drugs were used in a yearly amount of 7.4 tons in Denmark in 1995 [35]. Antineoplastics (cytostatic agents) differ from the other groups by the fact that they are mainly utilized in the hospital sector and by their intrinsic mutagenic action.About 13–14 kg of cyclophosphamide is used in hospitals per year [10]. In addition, 5,969 kg is prescribed for sale at private pharmacies. Considering all aspects, sex hormones, antibacterials, and antineoplastic agents were identified by Christensen as the three most relevant groups of chemicals concerning their potential human risk as a consequence of drug exposure via the environment [10]. Immunochemical methods for hormones and antibiotics have already been discussed above. In this section we will describe methods based on the use of antibodies for the analysis of analgesics, NSAIDs, and cytostatic agents. As occurs with other groups, after administration these drugs are excreted into wastewater, enter the aquatic environment, and eventually can reach drinking water if they are not biodegradable or eliminated during sewage treatment. Data on the environmental occurrence of the pharmaceuticals treated in this section are found in Table 1.

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4.1 Analgesics and NSAIDs Acetylsalicylic acid (aspirin) is still the most widely used analgesic, anti-inflammatory, and antipyretic agent followed by paracetamol. Ibuprofen [2-(4isobutylphenyl)propionic acid] and diclofenac (diclofenac-Na) from the NSAID group are used extensively for the treatment of rheumatic disorders, arthritis, pain, and fever. Fentanyl is a very strong opioid with analgesic properties 80 times stronger than those of morphine. The narcotics law therefore regulates its use. It is used in major surgery and in the treatment of pain in tumor patients [185]. Other opioids, including codeine (COD), morphine (MOR), and heroin have been used therapeutically and/or consumed illicitly for many years. Most of these substances have been shown to be readily or inherently biodegradable [11, 18, 39, 42]. Photodegradation was identified as the main elimination process of diclofenac in lake water [4, 42].With a relatively high sorption coefficient to particles, ibuprofen might be eliminated by sedimentation [43]. In contrast to diclofenac, ibuprofen and its metabolites are efficiently degraded (>95%) during treatment in WWTPs [39].Acetylsalicylic acid and its metabolite (salicylic acid) were detected in 22 and 33, respectively, of the 49 STP effluents analyzed by Richarson et al. [24]. The same authors reported that they found these substances in rivers and streams at levels between 0.2 and 0.5 mg L–1. Several other analgesics and NSAIDs such as aminophenazone, fenoprofen, indomethazine, ketoprofen, mefenamic acid, naproxen, and phenazone have also been detected in sewage, and surface and groundwater samples (i.e., [21, 36]). Immunochemical methods have been reported for the determination of these substances in body fluids (see Table 8) in clinical and forensic analyses. In the case of illicit use of opioid drugs, methods have also been reported for the control of drug abuse and assessment of intoxication using body fluids, tissue extracts, post-mortem specimens, and seizure samples. For this reason there are several commercially available immunochemical methods (see Table 4). Some research groups have used commercial immunoreagents (antibodies, tracers, and other conjugates) to develop new immunochemical methods. Thus, capillary zone electrophoresis or micellar electrokinetic capillary chromatography-based immunoassays with laser-induced fluorescence detection have been used for the determination of salicylate and gentisic acid in urine [187]. Similarly, Wey et al. [196] developed two rapid, competitive binding, electrokinetic capillary-based immunoassays recognizing urinary opioids (COD, codeine-6-glucuronide, dihydrocodeine (DHC), dihydrocodeine-6-glucuronide, MOR, morphine-3-glucuronide, and ethylmorphine (EMOR)).Aliquots of urine and immunoreagents of a commercial, broadly cross-reacting polarization fluoroimmunoassay (PFIA) for opiates were combined and analyzed by capillary zone electrophoresis or micellar electrokinetic capillary chromatography with laser-induced fluorescence detection. Assay sensitivities for COD and

a

60 ng L–1

0.0048 ng mL–1

ELISA (chemiluminescence detection) ELISA (spectrophotometric detection) ELISA ELISA

[193] [194]

[192]

[192]

[191]

[189] [190]

[188]

[187]

[187]

[186]

[186]

References

CE-IA-LIF: capillary electrophoresis-based immunoassay with laser-induced fluorescence detection; EIA: enzyme immunoassay; ELISA: enzymelinked immunosorbent assay; EMIT: enzyme-multiplied immunoassay technique; MECC-IA: micellar electrokinetic capillary chromatographybased immunoassay; RIA: radioimmunoassay.

0.25 ng mL–1 0.5 ng mL–1

0.045 ng mL–1

Pure, tap, and surface water; wastewater Umbilical cord, maternal plasma Umbilical cord, maternal plasma Human urine Human urine

6 ng L–1

ELISA

Diclofenac

Buffer Buffer

3.62 ng mL–1 100 pg per assay

ELISA ELISA

Ibuprofen

Urine

Salicylic acid

Urine

10 mg mL–1 (SA)

Serum

Serum

Plants

ELISA

Gentisic acid (GA)

IC50

Matrix

0.39 mmol L–1

CE-IA-LIF

Salicylic acid (SA)

<20 mg mL–1

LOD

Sensitivity

5 mg mL–1 (GA)

MECC-IA

CE-IA-LIF

Salicylate

MECC-IA

Immunochemical techniquea

Analgesics Paracetamol (acetaminophen)

Analyte

Table 8 Some representative immunochemical techniques developed for the detection of analgesics, NSAIDs, and cytostatic agents

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Methotrexate

Cytostatic (antineoplastic)

EMIT

Morphine

RIA EIA ELISA CE-IA-LIF

ELISA ELISA ELISA

ELISA CE-IA-LIF (commercial reagents)

Immunochemical techniquea

Codeine

Analgesics

Analyte

Table 8 (continued)

6.25 ng L–1 50 pg mL–1 5¥10–12 g mL–1 5 pg

Water samples Serum Buffer Buffer

Blood Bile Tissue Equine blood, urine Buffer Urine

0.020 mg L–1 0.200 mg L–1 0.100 mg L–1 100 pg mL–1 100 pg mL–1

Buffer Human urine

400 pg mL–1

IC50

Matrix

1 ng mL–1 10 ng mL–1

LOD

Sensitivity

[33] [201] [202] [203]

[197] [197] [197] [198] [199] [200]

[195] [196]

References

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MOR were comparable (10 ng mL–1), whereas those for DHC and EMOR were about fourfold lower. Furthermore, glucuronides were shown to react like the corresponding free opioids.Validation with real urine samples was performed with identification of the peaks by capillary electrophoresis–ion-trap mass spectrometry (CE–MS) after solid-phase extraction. A PFIA, commercialized by Abbot, is one of the most common immunochemical methods used in clinical laboratories to analyze salicylic acid (SA) in serum. The assay also recognizes gentisic acid (GA) but is insensitive to salicylamide, salicyluric acid, and conjugates of SA and of its metabolites [187]. With the aim of investigating the mechanisms involved in the hypersensitivity reactions, an enantioselective immunoaffinity extraction method has been developed that specifically isolates peptide fragments that have been modified with optically active ibuprofen. The antibodies were obtained by immunizing rabbits with (S)-ibuprofen coupled to BSA through a b-alanine group. The elicited antibody strongly recognizes the asymmetric center and the isobutylphenyl moiety of (S)-ibuprofen and its conjugates.A 0.5-mL aliquot of the immunosorbent (11.5 mg of IgG per mL gel) prepared by immobilization of the antibody was capable of retaining up to 1 mg of (S)-ibuprofen [190]. The immunochemical methods available today for the analysis of analgesics or NSAIDs should be easily adaptable to the analysis of environmental samples, although few examples have been reported. In this context, an indirect ELISA has been developed and applied to the determination of diclofenac in tap water, surface water, and wastewater samples [191]. The authors used a diclofenac-BSA conjugate as immunogen to produce antisera. The ELISA showed a LOD of 6 ng L–1 in buffer. A greater recognition for the glucuronide conjugates was observed. In order to validate the assay the results obtained were compared with those from GC–MS. 4.2 Cytostatic Agents Cytostatic drugs are frequently used in chemotherapeutic treatments. Residues of these substances should exclusively occur in hospital sewage at low concentrations. Among the cytostatic agents more frequently employed we should consider: methotrexate (MTX) (4-amino-10-methylfolic acid), a folic acid antagonist; the alkylating antineoplastic drug cyclophosphamide is one of the oldest known cytostatics and is one of the most frequently used agents in cancer chemotherapy; and ifosfamide is a widely used antitumor agent. Cytostatic agents fall far below the quantitative importance of other drugs. However, seen from the potential ecotoxicological impact viewpoint, they are an important group of drugs with a high potential risk for humans and wildlife. Although their effects against higher aquatic organisms such as fish or algae have only been seldom investigated [1], their carcinogenic, mutagenic, and embryotoxic effects are clearly demonstrated [204]. Most of the active substances investigated proved to have a low biodegradability. Therefore, the active substances are ex-

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pected to pass unchanged through municipal STP and thus reach surface waters [1] when they are not eliminated by adsorption onto sewage sludge. Steger-Hartmann et al. [44] did not observe a significant reduction in a laboratory-scale STP. In four out of 16 effluent samples from German STPs, cyclophosphamide was detected at maximum concentrations of 20 ng L–1. Ifosfamide was detected in only two samples, but in one of those it reached a value of 2.9 mg L–1 [35]. To our knowledge cytostatics have not been detected in surface waters but they have an estimated PEC of 0.8 ng L–1 [1, 7, 46]. Few examples of immunochemical methods for cytostatic agents have been reported (Table 8).Within the context of work performed by Aherne et al. [33], on the use of immunochemical methods in the analysis of microcontaminants in water samples, was reported the use of a RIA for the detection of methotrexate with a LOD of 6.25 ng mL–1. With the exception of a hospital effluent (concentration of 1 ng mL–1 of methotrexate was found), all samples (river and potable water) were negative. Ferrua et al. [201] developed an EIA with enzyme-labeled Ab and an analog antigen of MTX bound to polystyrene spheres through a methylated bovine albumin carrier. Serum samples of treated patients were analyzed, and good agreements with other methodologies developed to measure MTX were obtained. Recently, the use of commercial antibodies for MTX for the development of an immunoassay by capillary electrophoresis with laser-induced fluorescence detection has also been described [203], achieving good sensitivities. The procedure includes an initial competition step between the immunoreagents and the analyte followed by the separation of the species and detection, both steps carried out simultaneously with CE. The rapid separation allows the reduction of the time per assay to a few minutes.

5 General Summary Immunochemical methods have been developed or are commercially available for the analysis of the most important groups of pharmaceuticals with an increased human risk when in contact with the environment. However, additional work is necessary in order to expand the number of families of compounds that can be detected using these methods, and especially to adapt them to the analysis of environmental samples. Considering the complexity of the biological matrices, there is great promise regarding their potential application to the analysis of water samples. Within the advantages of implementing these methods for environmental monitoring purposes are their simplicity and the high sample processing capabilities. However, we must consider the low concentrations expected in the environment even though, due to their high bioactivity, these levels may be sufficient to cause adverse effects on the ecosystem.

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Acknowledgements This work has been supported by CICYT (BIO2000-0351-P4-05,AGL20015005-E) and by the EC: nanotechnology and nanosciences, knowledge-based multifunctional materials, new production processes and devices (contract number NMP-505485–1).

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The Handbook of Environmental Chemistry Vol. 5, Part O (2005): 245– 272 DOI 10.1007/b98617 © Springer-Verlag Berlin Heidelberg 2005

Wastewater Quality Monitoring: On-Line/On-Site Measurement Olivier Thomas1 (✉) · Marie-Florence Pouet1 1

Environment and Sustainable Development Institute, Université de Sherbrooke, QC, Canada [email protected]; [email protected]

1

Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 246

2 2.1 2.2 2.3 2.4 2.5

Measuring, Why and How? . . . . . . . . . . . . . . . . The Main Practical Objectives of Measuring . . . . . . . Sampling Versus On-Site Measurement . . . . . . . . . Types of Results . . . . . . . . . . . . . . . . . . . . . . Analytical Characteristics and Measurement Objectives Parameters and Substances . . . . . . . . . . . . . . . .

3 3.1 3.2 3.3 3.4

On-Site Measurement: From Needs to Solutions Minimizing the Measurement Error . . . . . . . Taking Account of Variability . . . . . . . . . . . Preventing Sample Aging . . . . . . . . . . . . . On-Site Measurement Constraints and Solutions

4 4.1 4.2 4.2.1 4.2.2 4.3 4.3.1 4.3.2 4.4 4.5

Parameters and Substances Monitored by On-Line Systems . . . . . . . On-Line Monitoring: From Laboratory-Transposed Methods to Software Sensors . . . . . . . On-Line Monitoring Systems . . . . . . . . . . . . . . . . . . . . . . . . Physical and Aggregate Properties . . . . . . . . . . . . . . . . . . . . . Inorganic Constituents . . . . . . . . . . . . . . . . . . . . . . . . . . . Organic Constituents . . . . . . . . . . . . . . . . . . . . . . . . . . . . Aggregate Properties . . . . . . . . . . . . . . . . . . . . . . . . . . . . Specific Organic Constituents . . . . . . . . . . . . . . . . . . . . . . . Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Non-Parametric Measurement for Detection of Accident or Disturbance

5 5.1 5.2 5.2.1 5.2.2 5.2.3

Validation and Developments Systems Validation . . . . . Developments . . . . . . . . Optical Techniques . . . . . Biosensors . . . . . . . . . . Software Sensors . . . . . .

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Abstract Sampling and laboratory analysis are not well adapted to wastewater quality monitoring in a process control or hazards prevention context, for which on-line/on-site measurement is preferable. Before considering the implementation and constraints of on-line systems, the reasons for and ways of monitoring are discussed. The main existing and up-and-coming solutions are then presented, showing that with respect to the number of parameters and substances to be monitored, for regulation purposes for example, only a few of them are measurable with on-line devices. Keywords Wastewater quality · On-site measurement · On-line monitoring · Emerging pollutants Abbreviations BOD Biochemical oxygen demand COD Chemical oxygen demand CSO Combined sewer overflow ORP Oxido-reduction potential PAH Polycyclic aromatic hydrocarbons SAC Spectral absorption coefficient (UV 254) TKN Total Kjeldhal nitrogen TOC Total organic carbon TSS Total suspended solids UV Ultraviolet UV/UV UV degradation + detection VOC Volatile organic compounds WW Wastewater(s) WWTP Wastewater treatment plant

1 Introduction We all agree that the continuous on-line/in situ detection of pollutants in water and wastewater should be the best practice for true quality monitoring. This is particularly relevant for the monitoring of emerging pollutants if we consider that for the other types of pollutants, there already exist some suitable systems. The main topic of this chapter is to show that there are some available devices for the on-line monitoring of emerging pollutants or, at least, some interesting developments. But first, it is indispensable to outline several important points in order to be sure that all potential users or developers of on-line/in situ measurement systems do know the main limits and constraints of the exercise.

2 Measuring, Why and How? First of all, let us explain that the term measurement used in this chapter must be considered in its generally accepted meaning, including all efforts carried out for giving end-users the quantitative or qualitative result that is more often use-

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ful for a decision-making process. This obviously includes all analytical systems but also all test kits, field-portable devices and on-/off-line instruments. 2.1 The Main Practical Objectives of Measuring In the domain of wastewater quality monitoring, the main reason for measuring is related to regulation compliance or contractual needs, in order to check if concentrations of substances or parameter values are below the threshold limits, for example at the outlet of a treatment plant, before discharge. Moreover, wastewater quality monitoring programmes are also planned for some specific sewer parts, such as combined sewage overflow (CSO) or industrial connections. For this application, the monitoring procedure is more often based on the use of the automatic sampling–laboratory analysis scheme. Flow measurement is generally coupled to wastewater quality measurement not only for sampling assistance, but also for daily load calculations, complementary to concentrations or parameter threshold values. The second main reason for wastewater quality monitoring is related to process control, particularly for treatment plants where analysers and sensors are generally used with physico-chemical or biological reactors, including settling tanks. This application is mainly encountered for important wastewater treatment plants, either urban (majority domestic) or industrial, where the storage capacities are rather small with regard to the flow to be treated. Obviously, on-line systems are preferable in this case, but the available instruments often limit the choice. Hazards prevention can also be a reason for wastewater quality monitoring, in order to protect biological treatment plants from toxic shock loads, for example, or to prevent potential toxic effects on the receiving medium. This application is mainly found in industrial contexts where the presence of toxic pollutants may occur. In this case, on-line systems are obviously preferable for real-time warning. The last reason is for improving the scientific knowledge of wastewater quality. This scientific need is of great interest with regard to the scope of this book. A lot of research has been carried out in the domain of water considered as a resource (drinking, surface and ground water) because of health considerations and economic reasons, but somewhat less for wastewater quality itself, due to poor interest. However, it is well known that wastewater constitutes a huge problem, even in developed countries, for environmental protection from chemical (including emerging pollutants) and other sanitary (pathogenic) risks to human health. This is the reason why research on wastewater quality must be encouraged with the development of suitable on-site measurement procedures.

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2.2 Sampling Versus On-Site Measurement As previously presented, and depending on the objective of the measurement, a procedure has to be chosen from two main ones (Fig. 1). The first is the classical procedure recommended and even required in official texts for regulation monitoring, based on sampling and laboratory analysis, including several steps between sampling and analysis: conditioning, storage, transportation and pretreatment. The other procedure, carried out on site, is based on the existence of on-line measurement systems or on the use of field-portable devices or test kits. Actually, the two approaches are often combined, taking into account either the scientific relevance of some practices (e.g. on-site measurement of dissolved gases and temperature), or the availability of systems for on-line monitoring. The constraints and methodology related to this last procedure will be explained in more in detail in the following sections. With respect to the measurement objectives, the two procedures are not equivalent, depending on the purpose (Table 1). If it is obvious that the classical procedure will still be preferred (at least because of the lack of suitable online systems), the relevant control of the treatment process cannot be envisaged without on-line monitoring. In fact, the two procedures might generally be considered as complementary for most applications.

Fig. 1 Measurement procedures for wastewater quality monitoring

Table 1 Choice of measurement procedure with respect to wastewater measurement needs

Sampling Regulation compliance Process control Hazards prevention Scientific knowledge

On-site

✓

✓

✓ ✓ ✓

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2.3 Types of Results Even if quantitative results are more often expected for wastewater quality measurement, qualitative information is of great interest, as is the case for other applications of the analytical sciences (in the health sector, the use of test kits and biodiagnostic systems leads to quick and useful information, often far from a classical analytical result). In fact, quantitative analysis gives the concentration not only of one substance, but also of a group of comparable substances (surfactants, PAH, ...), and even the value of a specific (TOC, TKN, ...) or aggregate (BOD, COD, toxicity, ...) parameter. In this context, total indices are often proposed as parameters complementary to classical analytical results [1]. From semi-quantitative results to non-parametric measurement, there exist several levels of qualitative information: – The first one is a degraded response of quantitative analysis, given by quick tests or kits designed for on-site use. The results obtained are then of a semiquantitative nature, often relying on a number of reagent drops or a colour change. The sensitivity of the kits is very coarse, depending on the scale of responses. This approach is very interesting for on-site studies from grab sampling, as it can provide assistance for focusing on sites of interest (or some period of time), in case of medium variability (pre-measurement). – The most common type of qualitative analysis is related to a binary response, with a “presence–absence”,“yes–no” or “lower than–greater than” answer to a main assumption, as for example: is the concentration lower than a threshold value? This method is interesting for the detection of unknown pollutants and can be easily carried out with screening procedures coupling real qualitative analysis with semi-quantitative responses. – A third mode of qualitative measurement is a non-parametric one [2]. This concept is based on the direct exploitation of an analytical signal (absorbance, intensity, potential, ...) without parameter calculation, leading to the simple characterization of the studied sample (Fig. 2). For example, fingerprinting or image analysis can be considered as non-parametric measurement. With respect to wastewater quality monitoring, quantitative and/or qualitative results can be chosen (Table 2). Table 2 Adequacy between measurement needs and types of results

Quantitative

Regulation compliance Process control Hazards prevention Scientific knowledge

✓ ✓ ✓ ✓

Semiquantitative

Qualitative

Nonparametric

–

– ✓ ✓ ✓

– ✓

? ✓ ✓

? ✓

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Fig. 2 Non-parametric measurement concept

2.4 Analytical Characteristics and Measurement Objectives If we try to refine the adequacy between the measurement procedures and the practical needs for wastewater quality monitoring, different metrological (analytical) characteristics have to be considered, such as detection limit, reliability and robustness (Table 3). Even if it is very difficult to compare the analytical methods carried out in the laboratory with on-site measurements (with on-line or tests kits), this presentation points out the main features of the measurement required for different needs. These characteristics define the quality of the available information [3], which constitutes one of the major problems that Table 3 Measurement needs versus analytical characteristics for wastewater monitoring

Regulation compliance Low detection limit, Sensitivity Rapidity Reliability Simplicity Relevance Robustness Availabilitya a

Process control

Hazards prevention

✓

Scientific knowledge ✓

✓

✓

✓

✓ ✓

✓ ✓

✓

✓

✓

✓

Ratio of number of exploitable results and number of produced results.

✓ ✓

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analysts have to face.A rapid response, even coarse, is sometimes preferable for decision making. In fact, there exists a great difference between the two types of needs in terms of traceability. On the one hand, the regulation compliance need and more often the scientific need require us to be as confident as possible about the trueness of the results (or at least the closeness between the results and true values) and to store the data for further exploitation. On the other hand, process control and hazards prevention are based on the real-time exploitation of the results. 2.5 Parameters and Substances A lot of substances and components are present in wastewaters and can be measured, especially the emerging pollutants. However, in practice, the aggregate parameters (BOD, COD, TSS, ...) and the physico-chemical ones (temperature, pH, dissolved oxygen, conductivity, turbidity, ...) are more often monitored. The only specific compounds generally considered are the N and P forms, and in case of industrial wastewaters, some specific pollutants such as organics (phenolics, hydrocarbons, ...) or metallic compounds. But if we take into account the emerging pollutants and compounds, the choice of which is guided by environmental considerations (mainly risks for health), then surfactants, endocrine disruptors, pesticides, other industrial organics (PAH, aromatic amines, ...) or inorganics (sulphides, arsenic, ...) and microbiological indicators (pathogens) must also be considered.

3 On-Site Measurement: From Needs to Solutions The previous chapters have shown that the classical procedure based on sampling and laboratory analysis is not suitable for the majority of cases, especially in an industrial context, where it is obvious that the efficiency of (treatment) processes must be always guaranteed or at least, most of the time. But there are other reasons for considering on-site (on-line) measurement. Experimental errors related to the numerous steps between sampling and analysis, medium variability with space and time, and sample aging are some good reasons. 3.1 Minimizing the Measurement Error The usual way to get information on wastewater quality is first sampling using an autosampler and then transportation of samples to the laboratory for analysis. Between sampling and analysis, several steps are needed: storage/conditioning, transportation, preparation (filtration, pre-concentration, cleanup, ...).

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Table 4 Comparison between technical steps used for the three main measurement procedures

Sampling Storage Transportation Preparation Analysis Transcription/transmission

On-line/off-line (grab sample)

In situ

Classical

–/✓

✓

✓ ✓

✓? ✓ ✓

✓ ✓ ✓ ✓ ✓ ✓

These steps are sources of error, and thus some recommendations must be carried out in order to minimize the final error in the analytical result [4]: – Sampling, including position of sampler inlet or device in the flow or tank, size of strainer, hose diameter and minimum flow rate for the sampling line, ... [5, 6] – Cooling or chemical preservation in adapted flasks, after sampling [7] – Rapid transportation from sampling sites to laboratory – Recovery tests or use of internal standards for pre-treatment and use of reference materials for analytical calibration and traceability [5]. Depending on the type of measurement chosen, some of the previous steps can obviously be avoided (Table 4). By minimizing the number of technical steps between sampling (or even on-line sensing) and the analytical result, the global error of measurement will be reduced. 3.2 Taking Account of Variability Wastewater is usually a very complex medium as explained further, the composition of which (urban or industrial) is changing with time and space. If time variability is well known for the load variation between night and day for urban wastewater or between weekdays and weekends for industrial wastewater, then space variability has been less studied.We easily understand that for an urban wastewater network, the composition of the medium will change with industrial discharge or with the presence of a hospital for example (these establishments being responsible for emerging pollutant discharges depending on the efficiency of the wastewater treatment, if any). But space variability can easily be studied inside industrial sites [2]. Table 5 shows the evolution of the qualitative variability of a refinery wastewater network, decreasing from upstream units (desalter) to the outlet of the treatment plant (biofilter). This study was based on the use of UV spectrophotometry for the estimation of phenols and sulphides [9] and on the study of hidden isosbestic points in UV spectra

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Table 5 Variability study along a refinery wastewater network [2]

Effluents

Comments

Inlet of the desalter pH 7.6, Ammonium 49.8, Sulphide 12, COD 2000, Phenols 200, TSS 100, Q 50 m3/h

– Mixture of outlets of different strippers. – UV spectra very structured. – lHIP=248 nm – V=70.8%

Outlet of a storage basin ph 8.4, Ammonium 46.4, Sulphide 1.7, COD 480, Phenols 9.1, TSS 59, Q 30 m3/h Outlet of the biofilter ph 7.1, Ammonium 37.8, Sulphide 0.1, COD 83, Phenols 0, TSS 15, Q 200 m3/h

Normalized UV spectra

– UV spectra few structured. – lHIP=224 nm – V=44.4%

– Few structured spectra. – Presence of nitrate. – lHIP=224 nm – V=7.7%

HIP: hidden isosbestic point; Q: flow rate; V: variability calculated from the number of spectra not crossing at the HIP divided by the total number of spectra.

sets [10]. Taking account of time and space variability, on-site measurement is the only suitable procedure (unfortunately limited by the available systems). 3.3 Preventing Sample Aging Another factor to be considered is that wastewater is rarely stable in its composition after sampling. Mainly due to physico-chemical or biological transformation, sample aging is very difficult to prevent even with low-temperature conservation after sampling, which is proposed for biodegradation inhibition. Among the chemical substances involved in physico-chemical aging, surfactants play a major role in aggregation/adsorption phenomena (Fig. 3). This explains why the surfactant concentration in the liquid phase can decrease by up to 30%, and at the same time total suspended solids can increase by up to 30% [8].

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Fig. 3 Aggregation/adsorption phenomena during aging processes [8]

A biologically treated wastewater is more stable than a raw one or a physicochemically treated effluent. Before biological treatment, raw urban wastewater may show strong variation in its composition due to adsorption biodegradation even at low temperature. This problem depends on the nature of the wastewater (a raw urban one being generally less stable than an industrial wastewater, except from the food industry), but seems to be sufficiently important for a lot of substances and parameters possibly to be transferred from dissolved to colloidal or solid phases between sampling and analysis, with a risk of adsorption–complexation–release for metallic compounds, and degradation into by-products for organics. This reinforces the interest in on-site measurement. 3.4 On-Site Measurement Constraints and Solutions On-site measurement thus seems to be the optimal solution for wastewater quality monitoring. However, as seen previously, wastewater composition is very complex and varyies with time and space. The implementation of on-site measurement must take into account some constraints related to the risk of sampling line clogging or sensor fouling (in the case of on-line measurement). In order to prevent this risk, or at least to space maintenance procedures, on-site measurement must be carried out carefully, depending on the type of system used. The principal factor of complexity is obviously related to the origin of the wastewaters and to the presence of fouling and/or clogging materials (Table 6). Even if raw suspended matter seems to be at first responsible for preventing measurement, some other components such as soluble substances (grease, petroleum by-products, ...) or living organisms (from bacteria to mussels) can be the main problem. Except for treated wastewater, almost all raw effluents contain solid compounds or grease and hydrocarbons responsible for fouling/clogging. As seen previously, on-site measurement is thus preferable, in order to place the instru-

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Table 6 Wastewater types and risk of fouling/clogging

Wastewater (WW) type

Suspended solids

Fouling/clogging occurrence

Factor

Raw urban WW Towns Small communities

+ Variable

+ +

Grease, bacteria id

Industrial WW Refinery, chem. plants Pulp and paper Textile Agrofood Metal transforming

– +++ + +++ +

++ +++ + +++ –

Hydrocarbons Fibres (cellulose) Fibres (textile) Biological solids None

Agricultural WW

+

+

Variable

Treated WW

–

–

None

ment as close as possible to the medium to be characterized, and three main possibilities exist (Fig. 4). The first is on-line measurement, with the sensor placed inside the flow to be monitored, and in this case without sampling. The second method is off-line measurement, for which the sensor is placed in a sampling loop with a high flow rate. If an automatic analyser is used for monitoring, a feeding line is connected to the “rapid” sampling loop. The third case is simpler, as measurement is carried out with a field-portable system, more often after grab sampling. On-line and off-line measurements are often grouped and considered as permanent or continuous measurements, when results are given with a more or less regular time interval. Taking into account the complexity of some wastewaters (mainly industrial), it is thus quite impossible to carry out perfect sampling for a complete representation of the medium. But if we assume that emerging pollutants are mainly present in the soluble fraction of wastewater, the only re-

Fig. 4 On-site measurement types

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Table 7 On-line sensor/analysing equipment “properties” (from [11])

“Property”

Main possibilities

Placement of on-line sensor

On-line, off-line, in situ, ...

Principle of sampling

External sampling, no external sampling

Principle of sample pre-treatment

No treatment, filtration, sedimentation, ...

Principle of measuring

Continuous, batch, ...

Principle of chemical/ physical method

Photometric, colorimetric, enzymatic, titrimetric,...

Number of determinants

Single parameter, multiple parameters

Need for supplies

Consumables, no consumables

Service intervals

Long, medium or short intervals

maining question is: does the presence of heterogeneous fractions in wastewaters affect directly or indirectly the representativity of the measurement? Directly means that, in the case of on-site measurement, the response may be affected by clogging (off-line), fouling (on-line) or existing interferences. On-line sensor/analysing equipment is an automatic measuring device giving an output signal linked to the value of one determinant (or more) from a medium, continuously or at regular time intervals. The choice and implementation of an on-line sensor must take into account several items or “properties” related to use constraints, which are presented in Table 7 (derived from [11]). To conclude this short discussion of on-site measurement implementation, Table 8 presents the main problems concerning the representativity of the Table 8 Main on-site measurement problems

Potential problems

Recommendations

On-line

Representativity Fouling Availability

Good positioning of sensor Anti-fouling devices Adapted service Double sensing preferable

Off-line

Representativity Clogging Availability

Good positioning of strainer Design of the strainer Suitable flow rate and hole diameter Adapted service

Representativity

Repeated samples and measurement

In situ

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measurement, the fouling/clogging risk for sensor or sampling line and the availability of the results. In fact, this last criterion integrates the others, as it is the judge of the final efficiency of the measurement chain. The availability is thus the percentage of “good” response of the system (in terms of metrology).

4 Parameters and Substances Monitored by On-Line Systems Before considering some existing solutions for permanent measurement of wastewater quality, let us describe briefly the different types of approaches. 4.1 On-Line Monitoring: From Laboratory-Transposed Methods to Software Sensors The permanent measurement methods can be grouped into three classes, depending on the principle used in terms of closeness to an existing laboratory method, generally considered as a standard (or reference) method: – Transposed laboratory methods – This first group is historically the most important as the first developments were carried out in that way. All colorimetric systems using automatic sampling feeding a fast reaction/detection line (for example, with a flow-injection procedure) have been developed from classical procedures, first to increase the analytical rate in laboratories before being transposed for off-line measurement. – Rapid equivalent methods – Generally based on a principle different from that of the corresponding laboratory method, alternative or surrogate systems are used more and more often for on-line and off-line monitoring. For example, the spectrophotometric methods or biosensors proposed for the measurement of organic compounds or electrochemical techniques for metals must be considered as alternative methods. – Software sensors and related methods – This last group is considered because of the complexity of wastewater composition and of treatment process control. As all relevant parameters are not directly measurable, as will be seen hereafter, the use of more or less complex mathematical models for the calculation (estimation) of some of them is sometimes proposed. Software sensing is thus based on methods that allow calculation of the value of a parameter from the measurement of one or more other parameters, the measurement principle of which is completely different from an existing standard/reference method, or has no direct relation. Statistical correlative methods can also be considered in this group. Some examples will be presented in the following section.

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4.2 On-Line Monitoring Systems In the past 5 years, some recommendations have been made for the development of new on-site sensors/analysing equipment [12, 13]. Two recent reviews present the state of the art of wastewater quality monitoring. The first [14] focuses on existing and innovative technologies for the main parameters used for the measurement of organic load (BOD, COD, TOC) for regulation needs, and the second [15] is a review of on-line monitoring equipment designed for wastewater treatment processes. These studies do not refer to emerging pollutants, and can be completed for the other parameters by the following considerations. 4.2.1 Physical and Aggregate Properties Relatively far from the present topic and well known, the on-line measurement of the physical and aggregate properties of wastewater does not present any problem. Conductivity, temperature, turbidity and oxido-reduction potential (ORP) are easily measured by well-designed sensors, because these parameters are also used for treatment process control. In practice, turbidity is more used for the treatment of natural water, and ORP for the biological treatment of wastewater. However, conductivity and temperature are often monitored at the same time as the other parameters in this section. The measurement of total suspended solids (TSS) in wastewater always constitutes a challenge because of the difficulty of transposing the laboratory procedure (filtration, drying and weighing). Turbidity measurement is sometimes used for TSS estimation by correlation, but generally without good agreement (because of solids heterogeneity and the effect of colloids). However, a non-contact device coupling both scattered light for TSS and fluorescence for organic load (COD) estimation has been proposed [16], giving good results around 100–200 mg L–1 for crude wastewater. Another study shows the interest of considering the UV absorption response of the heterogeneous fraction for TSS study and estimation [17]. 4.2.2 Inorganic Constituents There are a lot of interesting inorganic constituents to be measured in wastewater, either metallic or non-metallic, but few of them can be measured with on-line systems. Table 9 presents a selection of recent works dealing with online/on-site monitoring systems for inorganic analytes. For metallic constituents, some attempts have been made using electrochemical techniques but without real success, because of the existing interferences in wastewater and the high detection limit required with respect to

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Table 9 Selected on-line/on-site methods for inorganic constituent monitoring

Analyte

Principle

Reference

Metals Cu, Pb, Cd Cd Cr(VI)

Electrochemistry Biosensing UV spectrophotometry

[18] [19] [20]

UV spectrophotometry Biosensing Electrochemistry UV/UV

[21, 22] [24] [23] [25, 26] [26]

Colorimetry UV spectrophotometry Biosensing

[29] [27] [28]

Other Sulphide

UV spectrophotometry

[32]

Multi-ions Phosphate, iron, sulphate Chloride, sulphate, phosphate, ...

Continuous-flow analysis Capillary electrophoresis

[30] [31]

Nitrogen compounds Nitrate Nitrite Ammonium Nitrate + ammonium TKN Phosphorus compounds

regulation needs. A recent study [18] proposed the use of stripping voltammetry with thick-film graphite and screen-printed electrodes. The analysis is performed in three steps: sample pre-treatment, accumulation of the analyte on the electrode surface, and measurement. Cu, Pb and Cd can be determined with a detection limit around 5 mg L–1. Unfortunately, this technique has not been sufficiently checked for wastewater. Some commercial systems based on polarography can be envisaged, but they not really efficient for wastewater. Another method is explored with gene expression-based biosensors [19] for Cd measurement. A Cd-responsive promoter from E. coli is fused to a promoterless lacZ gene and monitored with an electrochemical assay of b-galactosidase activity. The expected detection limit is about 0.1 mg L–1 with a response of a few minutes.As for stripping voltammetry, no real tests have been carried out for wastewater. However, biosensors can be considered as a promising technique for wastewater monitoring. Hexavalent chromium is also a toxic compound (like lead, cadmium, mercury) and can be easily detected with UV spectrophotometry [20]. This system works for the quality control of electroplating treated wastewater with a detection limit of 5 mg L–1. For non-metallic constituents, several systems exist especially for nutrients monitoring, considering their importance in the eutrophication phenomenon. The on-line measurement of some nitrogen compounds (nitrate and ammo-

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nium) has been possible for at least 20 years with the use of selective electrodes at the beginning and UV detection more recently. For nitrate measurement, UV spectrophotometry is the best method because the UV-specific signal of nitrate can be simply extracted from the raw spectrum of a wastewater [21, 22]. An intercomparison study was carried out some years ago for NH4+ on-line analysers with several instruments using selective electrodes or UV spectrophotometry [23], showing that the performances were acceptable but installation and maintenance are crucial. For other forms of nitrogen, nitrite and organic constituents, few techniques have been proposed due to the lesser importance of nitrite in wastewater management (low concentration and non-stable), and to the difficulty in being selective for the organic forms. A biosensor for nitrite [24] was recently proposed for monitoring nitrite concentration in activated sludge exposed to oxic/anoxic cycles. The biosensor contains bacteria reducing only NO2– into N2O, which is subsequently monitored by a built-in electrochemical sensor. Up to 90% of the response is obtained in about 1 min, and the detection limit is around 5 mg L–1, a concentration sufficient for treatment process monitoring. Unfortunately, the maximum operational lifetime of the NO2– biosensor is 6 weeks and some problems may occur with time. A new in situ probe [25] was presented for the continuous measurement of ammonium and nitrate in a biological wastewater treatment plant. Based on the use of electrochemical measurement, the sensor can be immersed and requires minimum maintenance. The tests carried out to compare its performance with those of other procedures (including UV for nitrate) showed that the results were rather close, with a detection limit of 0.1 mg L–1 for both analytes. Another principle, based on the use of UV photo-oxidation of reduced forms of nitrogen (ammonium and organic) into nitrate (measured by UV spectrophotometry), allows the selective determination of nitrate, ammonium and organic nitrogen, and thus of TKN [26], with a detection limit of 1 mg L–1. This method is commercially available. On-line phosphate measurement is more often limited to orthophosphates, the total phosphorus measurement needing a mineralization step that is difficult to carry out on site. However, some recent works have been published [27, 28] based on the use of a biosensor or of UV spectrophotometry (after reagent addition). The limit of detection is rather high for this analyte (0.5 mg L–1 or higher). Some coupled systems allow measurement of the main N and P forms (nitrate, ammonia and orthophosphates) [22, 27, 29], among which is a system based on membrane technology in combination with semi-micro continuousflow analysis (mCFA) with classical colorimetry. With the same principle (classical colorimetry), another system [30] proposes the measurement of phosphate, iron and sulphate by flow-injection analysis (FIA). These systems are derived from laboratory procedures, as in a recent work [31] where capillary electrophoresis (CE) was used for the separation of inorganic and organic ions from waters in a pulp and paper process. Chloride, thiosulphate, sulphate, oxalate,

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sulphite, hydrogen sulphide, formiate, carbonate, phosphate and acetate are separated in 5 min after filtration. Except for the development of on-line systems for nutrients monitoring, the measurement of other inorganic non-metallic constituents is rather rare. Some commercial systems based on electrochemical sensing are proposed for the measurement of cyanide. A simple and rapid procedure for sulphide measurement in crude oil refinery wastewater has been developed [32]. Based on the deconvolution of the UV spectrum of a sample, this method has a detection limit of 0.5 mg L–1 and has been validated for crude oil refinery wastewater. Even if few systems are proposed for inorganic compounds (with regard to the number of potential pollutants), instruments or sensors for parameters used for treatment process control are available: UV systems for residual chlorine in deodorization, electrochemical sensors for dissolved oxygen (with nowadays a luminescent dissolved-oxygen probe utilizing a sensor coated with a luminescent material) and a colorimetric technique for residual ozone. In conclusion it must be noted that a lot of developments are still needed in order to increase the possibility of on-line/on-site monitoring of mineral constituents, including the speciation of metallic compounds with regard to health risks. 4.3 Organic Constituents The monitoring of organic constituents in wastewater concerns mainly the measurement of aggregate properties like oxygen demand parameters (BOD and COD) and also the detection of specific compounds, generally expressed as the total sum of the concentrations of their congeners. Table 10 displays a selection of on-line/on-site methods for the monitoring of organic constituents and related parameters. 4.3.1 Aggregate Properties Among organic constituent measurements, that of aggregate properties (BOD and COD) and specific parameters (TOC for example) has been well developed for more than 20 years. Concerning BOD, a recent review on biosensors [33] has been published. BOD biofilm-based sensors as well as respirometric systems, other measuring principles, and the commercial BOD instruments are discussed and compared regarding their performance characteristics like linearity, response time, precision, agreement between BOD values obtained from the biosensors and the conventional 5-day test, as well as toxic resistance to various compounds and operational stability. Some new developments are also proposed such as a system based on the use of electrochemically active bacteria in combination with a microbial fuel cell [34], giving good responses over 60 days, or a biosensor developed for fast

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Table 10 Organic constituents: on-line/on-site systems

Parameter

Principle

Reference

BOD

Sensor array Respirometer-type BOD sensors Biosensor Short-term BOD

[37] [33] [34] [35]

COD, TOC, ...

UV spectrophotometry FTIR

[38, 39] [40]

Hydrocarbons

IR evanescent wave

[41]

Organohalogenated compounds

ATR-FTIR

[42]

Phenols

Amperometric biosensor UV spectrophotometry

[43] [9]

Aromatic amines

UV spectrophotometry SPME–HPLC SPME–GC

[44, 45] [50] [51]

Surfactants

UV spectrophotometry On-line titration

[46] [47]

Explosives

SPME–biosensor

[49]

Screening

SPE (biosorbents)–HPLC–MS SPME–GC

[50] [51]

estimation of short-term biochemical oxygen demand (BODst) [35], leading to generally good agreement with the reference method for process control applications. For on-site measurement from grab sampling, a compact optical device with disposable strips for BOD determination has been developed [36]. The system includes three pairs of light-emitting diodes and photodiodes, and the disposable strips are made of inexpensive, transparent polycarbonate plates, where Pseudomonas fluorescens is immobilized. Using the 2,6-dichlorophenol-indophenol sodium salt as chromophore, a linear relationship was observed between the bioluminescence of the exposed strip and the BOD value. Another way for BOD estimation is the use of sensor arrays [37]. An electronic nose incorporating a non-specific sensor array of 12 conducting polymers was evaluated for its ability to monitor wastewater samples. A statistical approach (canonical correlation analysis) showed a linear relationship between the sensor responses and BOD over 5 months for some subsets of samples, leading to the prediction of BOD values from electronic nose analysis using neural network analysis. In the same way (use of a principle very different from the reference method), UV spectrophotometry is often proposed for BOD and even COD estimation. Among numerous works two main approaches are used for the exploitation of

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UV spectra. The first one is based on the absorbance measurement at 254 nm (the emitting ray of a low-pressure mercury lamp), often corrected by another absorbance measurement (for example at 280 or 350 nm) for compensation of interferences. This method can be advantageously replaced by a semi-deterministic deconvolution of UV spectra, taking account of the interferences for their own estimation (TSS for example). Some industrial applications for the pulp and paper industry or petrochemical plants [38, 39] have shown good estimations of aggregate and specific parameters, with a good availability of results with regard to other instruments (TOC-meter for example). Before considering the on-line measurement of some specific organic constituents, a final work dealing with the use of a Fourier-transform infrared (FT-IR) spectrometer as an on-line sensor can be cited [40]. Although limited to high concentrations, this method based on mid-IR analysis and calibration gives a rapid estimation of chemical oxygen demand (COD), total organic carbon (TOC), volatile fatty acids (VFA) and partial and total alkalinity (PA and TA) in anaerobic digestion processes for the treatment of industrial wastewaters. 4.3.2 Specific Organic Constituents The first group of interesting organic constituents is hydrocarbons. Classically measured by mid-IR spectrometry after solvent extraction, they can also be measured with near-IR devices such as a polymer-coated quartz glass optical fibre and direct spectrophotometric measurement of the extracted species in the polymer through the evanescent wave [41]. The proposed system can be used for the quantitative in situ analysis of organic pollutants like chlorinated hydrocarbons, aromatic hydrocarbons, or fuels in a broad concentration range from around 200 mg L–1 up to a few hundred milligrams per litre.As for the reference method, this instrument provides a signal corresponding to the sum of concentrations of the extracted organic compounds by measuring the integral absorption at the C–H overtone bands in the near-IR spectral range. Always based on the use of IR spectrophotometry, a novel attenuated total reflection–Fourier-transform infrared (ATR–FTIR) sensor [42] was proposed for the on-line monitoring of a dechlorination process. Organohalogenated compounds such as trichloroethylene (TCE), tetrachloroethylene (PCE) and carbon tetrachloride (CT) were detected with a limit of a few milligrams per litre, after extraction on the ATR internal-reflection element coated with a hydrophobic polymer. As for all IR techniques, partial least squares (PLS) calibration models are needed. As previously, this system is promising for bioprocess control and optimization. For phenolic compounds, amperometric biosensors have recently been designed using bacterial cells [43]. For this purpose, Pseudomonas putida immobilized on the surface of thick-film working electrodes made of gold, by using a gelatin membrane cross-linked with glutaraldehyde, was used and the respi-

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ration corresponding to phenol degradation was followed with a commercial oxygen electrode. Phenol detection was performed in synthetic wastewater samples. For refinery wastewaters, a UV spectrophotometry system, based on spectrum deconvolution, has been proposed for the simultaneous determination of sulphides and phenols with a detection limit of 5 mg L–1 [9]. Aromatic amines from the (bio)degradation of azo dyes or nitroaromatic explosives must also be monitored, mainly through the sum of their concentrations. However, taking account of the standard solution used for the calibration of the colorimetric reference method (4-nitroaniline), some attempts are proposed for the on-line specific determination of the most important single compounds [44, 45]. Both urban and industrial wastewater often contains high concentrations of surfactants. Cationic (like alkylbenzene sulphonates) and non-ionic surfactants (like alcohol ethoxylates) are among the most-used surfactants and are discharged into sewers in widely varying concentrations. Two on-line methods have been designed for the monitoring of cationic surfactants with UV spectrophotometry [46] and non-ionic surfactants by on-line titration [47]. The detection limits are around 10 mg L–1. Endocrine disruptors are nowadays considered among the most important emerging pollutants in wastewater, but they are not actually monitored on-line. A recent study [48] described the implementation of a broad-spectrum analytical scheme for the screening of more than 200 compounds (endocrine disruptors, pharmaceutical compounds, ...) in urban wastewater. For other specific organic compounds, a study concerning the improvement of immunoassays with a solid-phase extraction (SPE) membrane was reported for the on-site detection in soils and water of energetic materials (i.e. explosives) [49], but unfortunately it was not really tested for wastewater. Otherwise, there are some on-line procedures for the screening and detection of several specific organic constituents. The first [50], concerning polar pollutants, consists of a selective SPE based on antigen/antibody interactions, followed by liquid chromatography and diode-array or mass spectrometric detection. Class-selective immunosorbents have been developed for polyaromatic hydrocarbons, benzidine and congeners, nitroaniline and aromatic amines. Another procedure [51], using an on-site manual step of solid-phase microextraction (SPME) before gas chromatographic analysis, was designed for the detection of organic compounds in industrial wastewater. 4.4 Toxicity On-line or on-site toxicity evaluation is a great challenge due to the complexity of measuring the different impacts (from trouble from to death of organisms) of several substances, the effect of which is often increased by synergy. Furthermore for wastewater, toxicity monitoring must be implemented in several locations: raw sewer, treatment plant or discharge point. Toxicity can be eval-

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uated through some effects on organisms such as respirometric modification, bioluminescence or electrochemistry biosensing and even death (rarely used for on-line systems). Historically, respirometers have been used for wastewater biodegradability evaluation. More recently [52], a mobile on-line respirometer was proposed and tested for monitoring the activated sludge inhibition due to industrial discharges in a sewer network. A derived portable device called a Baroxymeter [53], based on monitoring the respiration of a bacterial culture by pressure measurements and using respiration inhibition as a toxicity alert, was proposed for the rapid detection of the toxicity effect of some toxic substances. But the most-used toxicity tests are based on bioluminescence inhibition, the responses of which are sometimes difficult to interpret particularly for wastewater quality monitoring. A comparison between a bioluminescence test kit (Microtox) and a respirometry approach for the toxicity study of seven organic and five inorganic toxic compounds was performed [54]. The bioluminescent response proved to have a higher sensitivity to toxicants but was less representative of the effects on activated sludge compared to respirometry, due to the nature of the microorganisms involved in each procedure. Recent studies, including the use of Microtox and ToxAlert test kits [55, 56], were carried out for the determination of the toxicity of some non-ionic surfactants and other compounds (aromatic hydrocarbons, endocrine disruptors) before implementation on raw and treated wastewater, followed by the identification and quantification of polar organic cytotoxic substances for samples with more than 20% inhibition. Furthermore, the study of their contribution to the total toxicity was obtained using sequential solid-phase extraction (SSPE) before liquid chromatography–mass spectrometry (LC–MS) detection. This combined procedure allows one to focus only on samples containing toxic substances. A derived combined approach uses an amperometric biosensor [57] with a whole-cell (E. coli) sensing part, for industrial application (textile and tannery wastewaters) and detection of phenolic compounds, non-ionic surfactants and benzenesulphonate compounds. As in the previous studies, chemical analysis (SSPE followed by LC–MS) revealed the pollutants responsible for the observed toxicity. A portable microbial sensing system [58] was developed for detecting the toxicity of pre-treated wastewater. The signal of the modified electrode containing a bacterial culture renewed every 9 h, within 8 min of contact with toxic solutions or samples, is roughly correlated with toxicity. A novel slow-release biosensor delivery for on-line monitoring instrumentation [59] allows both simple toxicity testing and more complex toxicity fingerprinting of industrial effluents, with the exploitation of kinetic (dose–response) and dynamic (response with time) signals. Furthermore, the slow release of biosensors immobilized in a polyvinyl alcohol (PVA) matrix greatly improved biosensor delivery, did not affect the sensitivity of toxicity testing, and demonstrated great potential for inclusion in on-line monitoring instrumentation.

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In conclusion, a review of the application of whole-cell biosensors for early warning systems [60] showed that electrochemical biosensors are well suited for on-site use in the monitoring of general toxicity as well as hydrocarbons and heavy metals. 4.5 Non-Parametric Measurement for Detection of Accident or Disturbance A final up-and-coming application, based on non-parametric measurement, is used more and more for process control and hazards prevention, for example for shock load prevention or toxic events. This qualitative approach uses integrated information coming from multiple physical signals: – Several existing physico-chemical sensors, with data reduction algorithms and filtering methods [61] (see “software sensing” in last part) – A chemical sensor array (consisting of eight conducting polymer sensors) derived from an electronic nose [62], for the characterization of headspace gas from a sparged liquid sample – Only one instrument giving several responses, such as the absorption spectrum of a UV detector [63].

5 Validation and Developments 5.1 Systems Validation Generally speaking, alternative methods (including on-line, off-line or in situ methods) may be used provided it can be demonstrated that equivalent results with those of reference procedures can be obtained. The experiments are generally carried out with standard solutions and reference materials for the determination of the method characteristics. The equivalence between methods must be statistically verified by plotting the results (Fig. 5) and checking the coordinates of the experimental regression line (comparison of the slope and intercept values, which must be not statistically different from respectively 1 and 0 values of the theoretical line). Once the equivalence between methods is confirmed, the validation procedure results given for on-/off-line instruments (permanent measurement) must be completed, taking into account that the sampling procedure is different for a lab method compared to a permanent one. For example, considering that regulatory constraints require 24 hours of composite sampling before lab analysis, the challenge is to obtain equivalent results with this procedure and with permanent measurement. In this case, the results to be compared are the mean of values for each measurement during the permanent acquisition, and the ref-

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Fig. 5 Comparison of reference and alternative methods

erence value of the corresponding composite sample. Then, the final choice of the measurement system must consider the specifications and performance tests of the selected system [11] and the future availability of the measurement estimated from other end users’ experiences [39]. For on-site measurement, such as colorimetric field kits giving semi-quantitative responses, a simple validation can be carried out [64] taking account of the non-continuous distribution of measurement values and that the measuring steps are often non-uniformly distributed over the measuring range. A recent study dealt with new elements related to metrological analysis in the field of (electrical safety) testing, such as measurement uncertainty and traceability [65]. It is important that the measurement result and its uncertainty are correctly evaluated so that the right conclusion of conformity or nonconformity with specifications is made, as is the case for wastewater quality regulation needs. In conclusion, one must insist on the importance of the main metrological characteristic, the traceabilility (generally of a result), ensuring a clear (unbroken) relationship between the final result and the complete measurement scheme by using appropriate procedures, standards and calibrated equipment. However, for chemical metrology and particularly for on-site measurement, some adaptations are needed for a wider meaning of traceability [66]. 5.2 Developments At the end of this chapter, the main methods of on-site measurement system development are briefly presented. 5.2.1 Optical Techniques Historically used but still in progress are the optical methods, the applications of which in wastewater quality monitoring have recently been reviewed [67, 68].

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Before seeing a (future?) whole integrated system, mixing UV–visible–infrared and fluorimetric methods, the first route is the development of UV-based microsystems, including some relevant spectral exploitation techniques such as the semi-deterministic one [69, 70]. Fluorescence data could be used to quantify oxygen demand values (chemical and biochemical) and total organic carbon values. Furthermore, the fluorescence spectral response can be apportioned to biodegradable (BOD) and non-biodegradable (COD-BOD) dissolved organics [71]. Other studies outline the advantages and drawbacks of the use of fluorescence techniques for wastewater quality monitoring [72, 73]. Less suitable for the purpose are infrared techniques, which are limited by the strong absorption of water. However, they can be envisaged for the monitoring of highly concentrated organic pollutants [74], particularly with the development of mid-infrared transparent optical fibres and waveguides, the surface of which can be chemically modified to enhanced analyte recognition based on tunable properties of enrichment or (bio)chemical recognition layers. The use of attenuated total reflection (ATR) devices with Fourier-transform IR spectroscopy is also proposed for organic compound monitoring [75]. The application of near-IR spectroscopy for real-time monitoring of glucose, lactic acid, acetic acid and biomass in liquid cultures of microorganisms of the genera Lactobacillus and Staphylococcus has been recently published [76]. The NIR spectrum acquired by the optical-fibre probe immersed in the culture is exploited using a partial least squares (PLS) calibration step, a classical method for IR techniques. A final optical application deals with the measurement of intracellular nicotinamide adenine dinucleotide (NADH) by fluorescence [77], giving information about the physiological status of wastewater treatment plant biomass. This indirect method could be envisaged for toxicity estimation. 5.2.2 Biosensors Biosensors are based on the direct spatial coupling of immobilized biologically active compounds with a signal transducer and an electronic amplifier. Due to the reaction between the biorecognition molecule (receptor) and a target analyte, a transducer signal is produced [27]. Biosensors are promising techniques for environmental monitoring applications such as toxicity, bioavailability and multi-pollutant screening, even if some limitations still exist [78]. With regard to the diversity of compounds and the complexity of matrices (particularly wastewaters), further developments have to be made before considering biosensing as an operational solution for on-line monitoring. Currently, a large spectrum of microbial biosensors have been developed that enable the monitoring of pollutants by measuring light, fluorescence, colour or electric current and electrochemical signal [60]. A recent study [19] shows that whole-cell biosensors based on the detection of changes in gene

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expression can be applied to environmental alterations of the cellular response (promoters responding to some physical parameters such as metabolites or environmental stress agents). A biosensor for measurement of inhibitors of nitrification in environmental samples has also been developed. The biosensor consists of a Clark oxygen probe as a transducer and an immobilized mixed nitrifying culture as the microbial component. The measuring principle is based on the direct determination of bacterial metabolic activity by measuring the oxygen consumption rate of the microbial immobilizate [79]. In conclusion, more than 40 years after the first electrode with an immobilized-enzyme membrane was produced, future developments in biosensor design will inevitably focus upon the technology of new materials, especially the new copolymers that promise to solve the biocompatibility problem and offer the prospect of more widespread use of biosensors in clinical (and environmental) monitoring [80]. 5.2.3 Software Sensors As seen previously for some specific applications such as wastewater treatment plants, software sensors can be envisaged to provide on-line estimation of nonmeasurable variables, model parameters or to overcome measurement delays [81–83]. Software sensors have been developed mainly for monitoring bioprocesses because the control system design of bioreactors is not straightforward due to [84] significant model uncertainty, lack of reliable on-line sensors, the non-linear and time-varying nature of the system or slow response of the process. A software sensor combines the theoretical knowledge of a system through a mathematical model, and the practical knowledge of its actual functioning through measurements. If the inputs acting on the system are known (and provided that theoretical conditions are fulfilled) and if, moreover, the model is a sufficient approximation of the real system, then the software sensor estimates the whole state of the system [81]. There are two types of approaches in developing software sensors [85], the first estimating the required parameters on the basis of a deterministic model and the second being a black-box approach depending only on the observed values. In practice, for wastewater treatment applications, the main techniques available are [82]: – The representation of the biological conservation of substrate to cell mass by an overall chemical reaction. The stoichiometric relationships are then used to calculate various rates such as cell mass concentration [83]. – The previous method supposes complete knowledge of the system and depends on the measurement quality of instruments (errors, availability), leading to severe effects on the accuracy of the on-line estimates. Therefore, a good noise filtration algorithm (like the Kalman filter or derivative) should be employed to improve the reliability of the estimated values before their use.

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Table 11 Software sensor main types for wastewater treatment applications

Types

Objectives

References

Estimation through elemental balances

Anaerobic bioreactor parameters estimation

[81]

Observer and filtering techniques

Anaerobic bioreactor control

[86]

Nitrification bioreactor control

[88]

TKN estimation

[85]

Fluorescence spectra exploitation for organic constituents estimation

[89]

Estimation of wastewater parameters (COD, NH4+, ...)

[90]

Artificial neural networks or hybrid ANN

– Another procedure uses artificial neural networks (ANN) derived from artificial intelligence techniques.Among several ANN algorithms, the feed-forward one, made up of interconnected neurone-like elements, can model complex non-linear systems easily, depending on the status of the training data. If there are noise and uncertainty in the training data, a problem of overfitting often arises but can be solved by data pre-processing, using a principal component analysis (PCA) for example [85]. Table 11 gives some examples of different software applied to the monitoring of bioreactors and the estimation of wastewater parameters.

References 1. 2. 3. 4. 5. 6. 7. 8. 9. 10. 11. 12. 13. 14.

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