emerging organic pollutants in waste waters and sludge

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]

1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

2 Usage . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32.1 Human Medicine . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42.2 Animal Farming . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 4

3 Sources and Fate . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 53.1 Environmental Distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . . 7

4 Occurrence . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

5 Toxicity Identification Evaluation (TIE) Approaches . . . . . . . . . . . . . . . 14

6 Analytical Methods . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 166.1 Sampling . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196.2 Sample Pretreatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 196.3 Analyses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 20

7 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 22

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

Abstract Estrogens and progestogens are two classes of female steroidal hormones whosepresence in the environment has been associated with the appearance of certain alarmingreproductive and development effects, such as feminization, decreased fertility, and her-maphroditism, in living organisms exposed to these compounds. Synthetic chemicals re-sembling these natural hormones are now well established in human medicine (mainly ascontraceptives 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.Mainlydue to unsuccessful removal in wastewater treatment plants, they are continuously releasedinto 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 areeasily 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 sourcesof estrogens and progestogens, their principal pathways into the aquatic environment, andthe primary routes of exposure to these compounds are discussed. This chapter also reviewsthe 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

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

AbbreviationsAPCI Atmospheric-pressure chemical ionizationEC European CommunityEDC Endocrine disrupting compoundEEF Molar-based 17b-estradiol equivalency factorELISA Enzyme-linked immunosorbent assayER-CALUX Estrogen receptor-mediated chemically activated luciferase gene expression

assayESI Electrospray ionizationEXAMS Exposure assessment modeling systemFDA United States Food and Drug AdministrationGC Gas chromatographyGPC Gel-permeation chromatographyHPLC High-performance liquid chromatographyLC Liquid chromatographyLLE Liquid–liquid extractionLOD Limit of detectionLOQ Limit of quantitationMCF-7 Cell proliferation (E-screen)MS Mass spectrometryMSTFA N-methyl-N-(trimethylsilyl)trifluoroacetamideNI Negative ionPFPA Pentafluoropropionic anhydridePI Positive ionRAM Restricted access materialSIM Selected ion monitoringSPE Solid-phase extractionSRM Selected reaction monitoringSTP Sewage treatment plantTIE Toxicity identification and evaluationUSEPA United States Environmental Protection AgencyWW WastewaterWWTP Wastewater treatment plantYES Yeast-based recombinant estrogen receptor–reporter assay

1Introduction

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

2 M. Kuster et al.

alkyl phenol ethoxylates, phthalates, bi-phenolic compounds, fitoestrogens andmicroestrogens, etc. – the group of female sexual hormones and related syn-thetic steroids stands out because of their estrogenic potency. Many studieshave confirmed the presence of estrogens and progestogens at concentrationsof toxicological concern in the aquatic environment. Already at very low con-centrations of ~1 ng L–1 endocrine disrupting effects, such as decreased fertil-ity, feminization, and hermaphroditism of aquatic organisms, are assigned tothis class of steroidal hormones [2–5]. Due to their strong endocrine disruptingpotency, special attention has been given to the natural estrogens estradiol andestrone, as well as to the synthetic estrogen ethynylestradiol [6].

Synthetic chemicals, resembling the action of natural hormones, find wideapplication in both human and veterinary medicine and in animal farmingpractices. Both natural and synthetic estrogens and progestogens are elimi-nated, either as free compounds or in their conjugated form, primarily throughthe urine but also in the feces. These substances enter the aquatic environmentmainly via wastewater treatment plant (WWTP) effluents (after incomplete removal in the plant) and untreated discharges, and through runoff of sewagesludge used in agriculture [7, 8]. Once in the waterways they may undergo a series of processes, such as photolysis, biodegradation, and sorption to bed-sediments, where estrogens and progestogens may persist for long periods [9].At present, the environmental occurrence of these substances is not subjectedto regulation. However, there are concrete indications that the presence of themost active estrogens in the aquatic environment will be regulated in the nearfuture. This calls for efficient and reliable analytical methods for routine mon-itoring and control. Since the consequences linked to the presence of thesecompounds in the environment were first made public, numerous analyticalmethods for their quantification in different environmental matrices have beendeveloped. Most of these methods have focused on surface waters, while waste-water (WW), sludge, and principally sediments and soils, have received com-paratively less attention, probably due to the complexity of these matrices. Thischapter reviews the most advanced methods applied to the analysis of estro-gens and progestogens in these complex matrices, together with the environ-mental levels found in these natural systems. The main sources of the most environmentally relevant estrogens and progestogens, their physicochemicalproperties, their principal pathways into the aquatic environment, the primaryroutes of exposure to these compounds, and data regarding their activity as en-docrine disruptors are discussed in this chapter.

2Usage

Large quantities of pharmacologically active substances are used annually inhuman medicine for diagnosis, treatment, and prevention of illness or to avoidunwanted pregnancy. In animal and fish farming, drugs are mostly adminis-

Estrogens and Progestogens in Wastewater, Sludge, Sediments, and Soil 3

tered as food additives for preventing illness, as growth promoters or parasiti-cides [10]. During the last five decades, the consumption of estrogens andprogestogens for some of these purposes has experienced a steady growth, andthis fact, together with the discovery of their negative ecological effects, hascontributed to the current concern about their occurrence in the environment.

2.1Human Medicine

In terms of binding to the human estrogen receptor, estradiol is the principalendogenous phenolic steroid estrogen, which is oxidized in the metabolicprocesses to estrone and further transformed to estriol. The natural hormonesare rapidly metabolized and are therefore orally inactive or active only at veryhigh concentrations. Blocking the oxidation to estrone by, for instance, intro-ducing an ethynyl group in position 17a or 17b of estradiol leads to much morestable 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 formsthe backbone structure used in the engineering of other synthetic estrogens,such as mestranol and estradiol valerate, also utilized in human hormone treat-ments [11].

One of the main applications of estrogens and progestogens is in contra-ceptives. The estrogen content in birth control pills is usually in the range of20 to 50 mg daily [12]. As for the progestogenic content, it varies depending onthe type of contraceptive. Thus, in combined oral formulations the progesto-genic content is in the range of 0.25 to 2 mg daily, whereas in progestogen-onlycontraceptives, it is lower (30–500 mg daily).

Besides contraception, the uses of estrogens can largely be put into threemain groups: the management of the menopausal and postmenopausal syn-drome (its widest use); physiological replacement therapy in deficiency states;and the treatment of prostatic cancer and of breast cancer in postmenopausalwomen.

In the same way as estrogens, progestogens are used in the treatment ofseveral other conditions such as infertility, endometriosis, in the managementof certain breast and endometrial cancers, and either alone or in combinationwith estrogens in the treatment of menstrual disorders, among others. Thetherapeutic doses required in the treatment of many of these diseases are often significantly larger than those employed in contraception.

2.2Animal Farming

Estrogens and progestogens are mainly used as growth promoters in animalfarming, and for the development of single-sex populations of fish in aquacul-ture. Some naturally occurring sexual steroids such as estradiol, progesterone,

4 M. Kuster et al.

and testosterone, and synthetic chemicals such as zeranol (estrogenic), melen-gestrol acetate (gestagenic), and trenbolone acetate (androgenic) have growth-promoting effects. Due to the improvement of weight gain and feed efficiencyin meat-production animals, administration of sex steroids to cattle has beena common practice for many years in several meat-exporting countries, in-cluding the USA. The most widely used substances are estrogens, either in theform of 17b-estradiol, estradiol benzoate, or the synthetic zeranol. Proges-terone, testosterone, and the two synthetic hormones trenbolone acetate andmelengestrol acetate are generally used in combination with estrogens [13]. Incontrast, no hormone applications for use in commercial-level poultry havebeen United States Food and Drug Administration (FDA)-approved since theagency’s withdrawal of the cancer-causing hormone diethylstilbestrol in the1950s.

In the European Community (EC), the use of hormonal substances for thepromotion of animal growth is prohibited (Directive 96/22/EC). The ban wasapplied without discrimination internally and to imports from third countriesas from January 1, 1989. As a result, countries wishing to export bovine meatand meat products to the EC were required either to have an equivalent legis-lation or to follow a hormone-free cattle program [14].

In aquaculture, steroidal compounds are used to develop single-sex popu-lations of fish to optimize growth. Sex determination in fish is primarily undergenetic control but may be influenced by various environmental conditions,such as temperature, social environment, pH, stocking density, and exposure toexogenous hormones or hormone-like chemicals [15]. Thus, all-male [16] andall-female [17] fish stocks may be obtained through exposure to androgens andestrogens, respectively. The potencies of sex steroids to induce sex reversal aredifferent for each steroid. Functional sex reversal from female to male is carriedout by using 17a-methyltestosterone, 19-norethynyltestosterone, or methylan-drosterone (concentration range: 0.1–100 mg/kg diet). 11-Ketotestosterone andandrosterone have also been used but the dosage required is higher than thoseof synthetic androgens. Phenotypical feminization is induced successfully byusing estradiol, although estrone and ethynylestradiol are used as well [18].

3Sources and Fate

Figure 1 shows the principal routes of environmental exposure to estrogens andprogestogens. The most relevant ways by which these compounds enter the environment and reach aquatic systems or the food chain are through WWTPeffluents, untreated discharges, and runoff of manure and sewage sludge usedin agriculture [7, 8, 10, 19–21, 45].

Human excretion is thought to be the principal source of estrogens andprogestogens. These compounds are readily adsorbed from the gastrointestinaltract and through the skin or mucous membranes, and are metabolized in the


liver with some undergoing enterohepatic recycling. Excreted hormones andtheir metabolites are found in urine, usually as water-soluble conjugates, anda 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 endup in WWTPs, where different processes of varying efficiency are applied. Fielddata suggest that the activated sludge treatment process can consistently removeover 85% of estradiol, estriol, and ethynylestradiol, and a lower, variable per-centage of estrone [23]. On the contrary, the concentration of unconjugatedsteroids in the effluent of WWTPs has occasionally been found to be higherthan that in the corresponding influent. Thus, many studies suggest that theconjugated forms (mainly glucuronides and sulfates) are readily converted tothe more active free compounds in both the sewer system and WWTPs, as a result of the activity of the b-glucuronidase and arylsulfatase enzymes presentin these systems [7, 24–27].

In activated sludge treatment works the principal mechanisms for removalof these compounds are likely to be sorption and biodegradation. Based on thelog Kow, sorption to sludge is predicted to play an important role in the removalof 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 tobiodegradation, the extent of which depends on factors such as nitrifying bac-teria, sludge retention times, aeration, and temperature, some laboratory teststudies indicate that estradiol is more readily mineralized than ethynylestradiol

6 M. Kuster et al.

Fig. 1 Routes of environmental exposure to estrogens and progestogens

or estrone and that synthetic estrogens in general exhibit greater recalcitrancein the activated sludge process [23, 27, 28].

More advanced water purification techniques, utilizing UV-irradiation,ozonization, or activated charcoal, may significantly improve the removal ofthese compounds, but these techniques are not broadly applied due to theirhigh cost. Thus, current European activated sludge treatment plants, with a hydraulic residence time not greater than 14 h, can in most cases not completelyeliminate all the estrogens and progestogens from the effluent [23].

As previously mentioned, contamination of water resources by estrogensand progestogens may also occur through runoff from manure and sewagesludge used in agriculture. Most of the drugs used for animals end up in theirurine 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 (asa result of leaching from fields) or the surface water in the vicinity (throughrunoff) 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 forsolid waste disposal is landfill. However, many of the disposal sites are opendumps without protective barriers or leachate-collection systems, which rep-resent a potential risk to the quality of the nearby groundwater.

Another increasingly important source of estrogens and progestogens in theenvironment is, as mentioned before, fish farming. Treatment with steroids isusually carried out by feeding, although in species where male sex differentia-tion is initiated before feeding commences (e.g., salmon), other procedures areused, such as immersion of alevins [18]. Drugs used in aquaculture as feed ad-ditives are discharged directly into the water. It has been estimated that around70% of the drugs administered end up in the environment surrounding thefarm, due to overfeeding, loss of appetite by diseased fish, and poor adsorptionof the drugs [31].

3.1Environmental Distribution

The introduction of estrogens and progestogens into the environment is a func-tion of the way several factors are combined. The manufactured quantity andthe dosage applied (amount, frequency, and duration) combined with the excretion efficiency of the compound and its metabolites, the capability ofadsorption and desorption on soil, and the metabolic decomposition in sewagetreatment are examples of necessary factors to assess environmental exposure.In general the fate and effect of a substance in the environment is dependent onthe distribution into the different natural systems, such as air, water, and solids(soil, particles, sediment, and biota). Information on the physical and chemicalproperties (KH, Kd, and Kow vapor pressure) of a compound may help determinewhether it is likely to concentrate in the aquatic, terrestrial, or atmospheric


environment. Table 1 lists the main physicochemical characteristics of the mostrelevant estrogens and progestogens from the environmental point of view.Withregards to the water solubility, it might be worth pointing out that the steroidsolubility in WW may be markedly lower than that in distilled water [32].

Once estrogens and progestogens have reached the waterways, a series of pro-cesses, such as, photolysis, biodegradation, and sorption to bed-sediments, cancontribute to their elimination from the environmental water. Given the relativelylow polarity of these compounds, with octanol–water partition coefficientsmostly between 103 and 105, sorption to bed-sediments appears to be a likelyprocess. Kd values calculated for estriol, norethindrone, and progesterone in aSpanish river (479, 128, and 204, respectively) as the ratio between the sedimentconcentration (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 thebehavior of estrogens in the aquatic environment and set up a model to estimatetheir likely environmental concentrations in the water column and bed-sedi-ments.According to this study, between 13 and 92% of the estrogens entering ariver system would end up in the bed-sediment compartment with the major-ity of sorption occurring within the first 24 h of contact.

A similar approach conducted by Lai et al. [9] to investigate the partition-ing of estrogens from water to sediments, kinetics of sorption, and the influenceof various environmental variables (salinity, total organic carbon, etc.) indi-cated that sorption takes place rapidly within the first half hour, slows downwithin the next half hour, and steadily decreases afterward. Furthermore, thesynthetic estrogens (mestranol and ethynylestradiol), with their higher Kow val-ues, were shown to partition to the sediment to a greater extent than the naturalestrogens. At higher estrogen concentrations, there was a decrease in estrogenremoval from the aqueous phase, while higher levels of sediment inducedgreater removal. The sorption of estrogens to sediments correlated to the totalorganic carbon content. However, the presence of organic carbon was not a prerequisite for sorption. Tests performed with laboratory saline water resultedin an increase of estrogen removal from the water phase compared to unsaltedwaters, 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 othercompounds for the binding sites.

A series of experiments was also conducted by Bowman et al. [34] to ascer-tain the effects of differing environmental factors on the sediment–water inter-actions of natural estrogens (estradiol and estrone) under estuarine conditions.Sorption onto sediment particles was in this case relatively slow, with sorptionequilibrium being reached in about 10 and 170 h for estrone and estradiol,respectively. On the other hand, true partition coefficients calculated on colloidswere found to be around two orders of magnitude greater that those on sedi-ment particles. Hence, it was concluded that under estuarine conditions, and incomparison to other more hydrophobic compounds, both estrone and estradiol

8 M. Kuster et al.


Tabl

e1

Phys

ico-

chem

ical

Pro

pert

ies

ofse

lect

ed E

stro

gens

and

Pro

gest

ogen

s [(

http

://es

c.sy

rres

.com

/int

erko

w/p

hysd

emo.

htm

)]

Com

poun

dC

AS

num

ber

Mol

ecul

ar

Wat

er

Log

P ow

cVa

por

Hen

ry’s

law

wei

ght

solu

bilit

ya,

cpr

essu

rea,

bco

nsta

nta,

b

(mg

L–1)

(mm

Hg)

(atm

-m3

mol

e–1)

Estr

adio

l00

0050

-28-

227

2.39

3.6

(27

°C)

4.01

1.26

E-00

83.

64E-

011

Estr

iol

0000

50-2

7-1

288.

3944

1b2.

451.

97E-

010

1.33

E-01

2Es

tron

e00

0053

-16-

727

0.37

303.

131.

42E-

007

3.8E

-010

Ethi

nyl e

stra

diol

0000

57-6

3-6

296.

4111

.3 (2

7°C

)3.

672.

67E-

009

7.94

E-01

2D

ieth

ylst

ilbes

trol

0000

56-5

3-1

268.

3612

5.07

1.41

E-00

85.

8E-0

12N

oret

hind

rone

0000

68-2

2-4

298.

437.

042.

977.

31E-

009

5.8E

-010

Prog

este

rone

0000

57-8

3-0

314.

478.

813.

871.

3E-0

066.

49E-

008

D-N

orge

stre

l00

0797

-63-

731

2.46

2.05

b3.

483.

93E-

010

7.7E

-010

aVa

lues

are

at 2

5°C

ifno

t spe

cifie

d.b

Esti

mat

ed d

ata.

cEx

peri

men

tal d

ata.

would be expected to remain mainly in the dissolved phase and to have a strongtendency for bioaccumulation.

Nevertheless, it is not clear yet whether sorption or biodegradation processesplay a major role. Studies conducted with activated sludge [25, 35] pointed atbiodegradation as the mechanism contributing the most to the eliminationof 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 degra-dation processes in rivers are unimportant under average flow conditions, asthey account for only 2–8% of the input loading. This is in agreement with thetheory presented by Huang et al. [36], according to which the main removalmechanism for hormones in WWTPs would be sorption onto particles and notbiotransformation.

Degradation studies carried out in waters from five English rivers indicatethat estradiol has a half-life of 3–27 days [33]. Estrone was found to be the firstdegradation product of estradiol but no investigations of the subsequent by-products were conducted. The poorest degradation rates were observed in theestuary river water samples,where the high salt content might have inhibited mi-crobial degradation. Furthermore, ethynylestradiol (half-life 46 days) was foundto 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 mestranolhave been shown to remain stable and intact over 5 days, while progestogensare 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 fromsediments has been shown to be significantly less important than sorption, withdesorption distribution coefficients two to three times lower than those obtainedfor the sorption process [33]. In an environment like this, river sediments cantherefore act as sinks where estrogens and progestogens may persist for long periods, be transported to other areas, and be eventually released by diffusionacross the sediment water-column interface or by scouring in storm events [9].The concentration of estrogens and progestogens in bed-sediments is predictedto increase over time; thus, bed-sediments can be anticipated as environmentalreservoirs from where these substances may eventually become bioavailable [37].

4Occurrence

Most research of estrogens and progestogens has been conducted on water sam-ples and less frequently on solid samples. Soils and sediments, in particular, havereceived very little attention and thus literature data on these matrices are very

10 M. Kuster et al.

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 theirlower estrogenic potency. Figure 2 shows the chemical structure of the estrogensand progestogens most frequently investigated in environmental samples.Table 2 summarizes the literature data available on the occurrence of these twoclasses of steroidal compounds in WW, sludge, and sediments.

Natural hormones (and their metabolites) have always been present in theenvironment. The growing use of both natural and synthetic estrogens andprogestogens in human medicine and in livestock farming (see Sect. 2, Usage)has led to an increase of their occurrence in natural systems. Due to steadypopulation growth and regional population density, an irregular distributionof these pollutants is found. Particular concern is given to certain areas wherehigh levels were detected, e.g., areas adjacent to agricultural and animal farms.


Table 2 Environmental occurrence of estrogens and progestogens

Matrix (location) Compounds Concentration Ref.(ng L–1 or ng g–1)

WATER

Wastewater

STP influent Æ effluent

Italy Estrogens and progestogens 0.4–188 Æ 0.3–82.1 [35]Spain Natural and synthetic estrogens <0.2–115 Æ <0.2–21.5 [40]The Netherlands Natural and synthetic estrogens <0.5–140 Æ <0.4–47 [59]France Natural and synthetic estrogens 2–8.6 Æ 3.8–18 [66]

SOLID SAMPLES

River sedimentGermany Natural and synthetic estrogens <0.2–2 [45]Spain Estrogens and progestogens 0.05–22.8 [46]UK Estrone <0.04–0.388 [42]

Activated and digested sludge

Germany Natural and synthetic estrogens <2–49 [45]

Activated sludgeIsrael Estrogen 19–64 [19]

Soil – – –

BIOTA

Rainbow trout bileSweden Natural and synthetic estrogens <0.1–2.5 mg/g [48]

12 M. Kuster et al.

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

As previously mentioned, discharged domestic effluents represent the mostsignificant 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 sewagein the nanogram per liter range and occasionally in the microgram per literrange (see Table 2). The nature of the compounds present in the sewage systemdoes, however, change as a function of their transport route.

D’Ascenzo et al. [39] have recently conducted a very comprehensive studywhere the presence of the three most common natural estrogens and their conjugated forms was investigated in female urine, in a septic tank collectingdomestic 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 concentrationsof 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 thetotal conjugated estrogens content. This situation, however, changed markedlyin the condominium collecting tank. Here, significant amounts of free estro-gens were observed and the estrogen sulfate to estrogen glucuronate ratio roseto 55/45, which was attributed to the ready deconjugation of the glucuronatedestrogens by the b-glucuronidase enzyme produced presumably in largeamounts by fecal bacteria (Escherichia coli). At the WWTP entrance, free es-trogens and sulfated estrogens were the dominant species. Finally, the sewagetreatment 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 de-tected 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 beenobserved with respect to the identity of the compounds most frequently de-tected in WWTP effluents. Thus, of the various compounds most commonlymonitored – namely, estradiol, its metabolites estriol and estrone, and the synthetic estrogen ethynylestradiol – estrone is the most ubiquitous both inWWTP effluents and in environmental waters in general, while the most potentestrogens estradiol and ethynylestradiol have only occasionally been detected[26, 40–42].As for the conjugates, the very few studies that have attempted theirdetermination pointed out estrone sulfates as the most abundant, while glu-curonides 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 thecorresponding effluent and decreases with distance from the WWTP [35, 36, 39,42]. In this kind of compartment, the presence of estrogens and progestogenshas been reported to occur in the low nanogram per liter range [19, 43, 44].

In activated and digested sewage sludge, the concentration of ethynylestra-diol (17 ng g–1), estrone (37 ng g–1), and estradiol (49 ng g–1), found in one of the


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 theoccurrence 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 de-tected (>0.04–0.388 ng g–1 wet weight) [42]. Estriol and norethindrone were thecompounds most frequently detected in sediments collected in two rivers fromthe northeast of Spain, where maximum concentrations were obtained forethynylestradiol (22.8 ng g–1 dry weight) and estrone (11.9 ng g–1) [46]. In bothstudies, large concentration variations between sites and between the same sitesampled on different occasions were observed, which might be explained by thedifficulty in obtaining representative samples and the variability of factors suchas the gaseous/redox conditions influencing degradation rates. In anotherstudy, conducted in Germany, estradiol, estrone, and ethynylestradiol werefound at concentrations up to 2 ng g–1 (estrone), whereas mestranol, a prodrugfor ethynylestradiol, was not detected [45]. Finally, neither estrogens norprogestogens 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 higherthan those in water were found in the bile of a rainbow trout caged downstreamof WWTPs.

5Toxicity Identification Evaluation (TIE) Approaches

Most of the studies conducted to assess the environmental occurrence and fateof estrogens and progestogens have focused on the determination of specifictarget compounds. However, by simply following the disappearance of a sub-stance, 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 thegreat number of compounds that can possibly be generated, the high costs, andthe lack of analytical standards.

At the beginning of the 1990s the United States Environmental ProtectionAgency (USEPA) developed the so-called TIE procedures. These approacheswere originally designed to identify the presence of health and environmentallyrelevant compounds in WWs [49, 50]. However, since their introduction, theyhave become established, powerful tools for determining the causative agentsof 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

14 M. Kuster et al.

extracts into fractions where the identification of the compounds responsiblefor the effects observed (by means of suitable bioassays) is feasible using, e.g.,mass spectrometric methods. A survey of the TIE procedures used for theidentification of endocrine disrupting compounds (EDCs) has been recentlypublished by Petrovic et al. [51].

An example of a TIE approach is that described by Desbrow et al. [7]. In thiswork, the endocrine disrupting activity detected in effluents of seven UKWWTPs by means of a yeast-based screening assay [52] was mainly attributedto the presence of estradiol, estrone, and ethynylestradiol. However, to assessthe estrogenic activity different bioassays may be used, e.g., the yeast-based recombinant estrogen receptor–reporter assay (YES), the MCF-7 cell prolifer-ation (E-screen), and the estrogen receptor-mediated chemically activated


Fig. 3 General scheme of TIE procedure

16 M. Kuster et al.

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 1 1 1Estriol 3.7E-1Ethinyl estradiol 1.9E-1–1.2 1.25–1.9 1.2Diethylstilbestrol 4.5E-2–1.1 2.5Estrone 1.9E-2–1E-1 1E-2 5.6E-2

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

6Analytical 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 toboth the complex environmental matrices and the requirement of low detectionlimits. Therefore, the use of complicated, time- and labor-consuming analyticalprocedures 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 sam-ples and discuss the very few methods used up to now to analyze solid sam-ples (sediments, sludge, and soil). For more detailed data we refer to the reviews:

– Environmental behavior and analysis of veterinary and human drugs insoils, sediments, and sludge [53]

– Determination of endocrine disrupters in sewage treatment and receivingwaters [54]

– Review of analytical methods for the determination of estrogens and pro-gestogens in WWs [55]

– Liquid chromatography–(tandem) mass spectrometry of selected emergingpollutants (steroid sex hormones, drugs and alkyl phenolic surfactants) inthe aquatic environment [56]

Some of these methods are summarized in Table 4 and details concerning eachstep of the analytical process are described in Sects. 6.1–6.3.

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

ble

4A

naly

tica

l met

hods

des

crib

ed fo

r th

e de

term

inat

ion

ofes

trog

ens

and

prog

esto

gens

in e

nvir

onm

enta

l sam

ples

Sam

ple

Com

poun

dSa

mpl

e pr

epar

atio

nA

naly

tica

l met

hod

LOD

Ref

.(n

gL–1

or n

gg–1

)

WW

TP

E3

,E2,

EE,E

1Sp

eedi

sk-C

18 a

nd

GC

–MS

LOQ

0.0

4–0.

32[6

6]in

fluen

t,de

riva

tiza

tion

PFP

Aef

fluen

t,E2

,EE,

E1SP

E (C

18),

HPL

C fr

acti

on,L

LEG

C–M

S0.

2[7

]an

d ot

hers

E2,E

E,E1

,aE2

SPE

(SD

B-X

C d

isk,

C18

or

GC

–MS/

MS

0.1–

2.4

[26]

NH

2 co

lum

n),H

PLC

frac

tion

)

E2,E

1,M

ES,a

E2,E

2-17

V,SP

E (C

18),

silic

a ge

l,G

C–M

S/M

S1

[43]

16a-

OH

-E1,

E2-1

7Ade

riva

tiza

tion

E3,E

2,EE

,E1

SPE

(Car

bogr

aph-

4)H

PLC

–MS/

MS

(NI-

ESI)

LOQ

0.0

8–0.

6[3

5]

E3,E

2,EE

,E1,

DES

,NO

R,

SPE

(C18

col

umn)

HPL

C-D

AD

–MS

(NI-

ESI)

2–50

0[5

7]

LEV,

PRO

G

E2,E

E,E2

-17G

,E2-

3SSP

E (C

18 d

isk)

,hyd

roly

sis,

ELIS

A0.

1[3

6]H

PLC

frac

tion

GC

–MS/

MS

(E2)

0.2–

0.4

E3,E

2,EE

,E1,

DES

,NO

R,

On-

line

SPE

(PLR

P-s

colu

mn)

HPL

C-D

AD

–MS

(NI-

ESI)

10–2

00[6

0]

LEV,

PRO

G

E2,E

3,E1

,EE,

DES

SPE

(C18

col

umn)

HPL

C–M

S (N

I-ES

I) (S

IM)

200

[68]

HPL

C–M

S/M

S (N

I-ES

I)5

E2,E

1SP

E (L

iChr

olut

EN

+C

18 c

ol.)

+H

PLC

–MS

(NI-

ESI)

0.07

–0.1

8[6

9]im

mun

oaff

init

y ex

trac

tion

E2,E

3,E1

,EE,

E3-3

G,E

2-3G

,SP

E (C

arbo

grap

h-4)

HPL

C–M

S/M

S (N

I-ES

I)0.

003–

15[6

1]E1

-3G

,E3-

16G

,E2-

17G

,E3

-3S,

E2-3

S,E1

-3S

E1,e

stro

ne;E

1-3G

,est

rone

3-(

b-D

-glu

curo

nide

);E1

-3S,

estr

one

3-su

lfate

;16a

-OH

-E1,

16a-

hydr

oxye

stro

ne;E

2,17

b-es

trad

iol;

aE2,

17a-

estr

adio

l;E2

-3G

,17b

-est

radi

ol 3

-(b-

D-g

lucu

roni

de);

E2-3

S,17

b-es

trad

iol 3

-sul

fate

;E2-

17A

,17b

-est

radi

ol 1

7-ac

etat

e;E2

-17G

,17b

-est

radi

ol 1

7-(b

-D-g

lucu

roni

de);

E2-1

7V,

17b-

estr

adio

l 17-

vale

rate

;E3,

estr

iol;

E3-3

G,e

stri

ol 3

-(b-

D-g

lucu

roni

de);

E3-3

S,es

trio

l 3-s

ulfa

te;E

3-16

G,e

stri

ol 1

6a-(

b-D

-glu

curo

nide

);EE

,eth

ynyl

estr

adio

l;D

ES,d

ieth

ysti

lbes

trol

;MES

,mes

tran

ol;L

EV,l

evon

orge

stre

l;N

OR

,nor

ethi

ndro

ne;P

RO

G,p

roge

ster

one;

MST

FA,N

-met

hyl-

N-

(tri

met

hyls

ilyl)

trifl

uoro

acea

tam

ide;

PFPA

,pen

taflu

orop

ropi

onic

anh

ydri

de.

18 M. Kuster et al.

Tabl

e4

(con

tinu

ed)

Sam

ple

Com

poun

dSa

mpl

e pr

epar

atio

nA

naly

tica

l met

hod

LOD

Ref

.(n

gL–1

or n

gg–1

)

WW

TP

E2,E

3,E1

,EE

SPE

(Env

i-C

arb

col.)

HPL

C–M

S/M

S (P

I-A

PCI)

LOQ

0.5

–1[5

8]in

fluen

t,E2

,EE,

E1SP

E (S

DB-

XC

dis

k,C

18 o

r G

C–M

S/M

S0.

1–1.

8[5

9]ef

fluen

t,N

H2

colu

mn)

,HPL

C fr

acti

on)

and

othe

rsE3

,E2,

EE,E

1SP

E (C

arbo

grap

h-4)

HPL

C–M

S/M

S (N

I-ES

I)0.

2–0.

5[5

9]

E2,E

E,E1

Solv

ent e

xtra

ctio

n w

ith

GC

–MS/

MS

0.4–

1[4

2]so

nific

atio

n,H

PLC

frac

tion

ing,

deri

vati

zati

on

Nat

ural

E2

,EE,

E1U

ltras

onic

ext

ract

ion

wit

h G

C–M

S/M

S0.

04–5

[42]

rive

r D

CM

,HPL

C fr

acti

onin

g,se

dim

ent

deri

vati

zati

on

E3,E

2,EE

,E1,

DES

Ultr

ason

ic e

xtra

ctio

n w

ith

HPL

C–M

S (N

I-ES

I) (S

IM)

0.05

–1[4

6]M

eOH

-ace

tone

,SPE

(C18

)

E3,E

2,EE

,E1,

DES

SPE

(RA

M c

artr

idge

s (A

DS

C4)

)H

PLC

–MS

(NI-

ESI)

(SIM

)1–

5[6

3]

E2,E

E,E1

,MES

Solv

ent e

xtra

ctio

n,si

lica

gel

GC

–MS/

MS

(EI)

LOQ

0.2

–0.4

[45]

co

lum

n,SP

E (R

P-C

18),

HPL

C (R

P-C

18) c

lean

up,

deri

vati

zati

on M

STFA

NO

R,L

EV,P

RO

GU

ltras

onic

ext

ract

ion

wit

hH

PLC

–MS

(PI-

ESI)

(SIM

)0.

04[4

6]

MeO

H-a

ceto

ne,S

PE (C

18)

NO

R,L

EV,P

RO

GSP

E (R

AM

car

trid

ges

(AD

S C

4))

HPL

C–M

S (P

I-ES

I) (S

IM)

0.5

[63]

Slud

geE2

,EE,

E1,M

ESSo

lven

t ext

ract

ion,

GPC

,G

C–i

on-t

rap

MS/

MS

LOQ

2–4

[45]

silic

a ge

l col

umn,

deri

vati

zati

on

wit

h M

STFA

6.1Sampling

The choice of the sampling procedure is essential to obtain significant, repre-sentative results of the environmental occurrence of these compounds. Theplace and time of sample collection has to be planned carefully. The samplingprocedure and transport should not influence the matrix to be analyzed. Ingeneral, exposure to light, oxygen, and high temperatures should be avoideddue to the risk of transformation of the analytes or other organic componentspresent in the environmental samples.

Wastewater samples have usually been collected in precleaned amber glasscontainers. Both discrete and composite samples have been used for the analy-sis of effluents and influents of WWTPs. Unpreserved samples are normallystored at 4 °C for 48 h, or frozen [48]. Other authors add chemical agents suchas methanol, sulfuric acid, or mercuric chloride to prevent bacterial activity dur-ing 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 employedif more detailed information on the spatial distribution of the analytes isneeded. 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 orlyophilized prior to storage.

6.2Sample Pretreatment

An essential step in the analysis of trace pollutants in environmental matricesis the pretreatment procedure. Methods that are more efficient have been de-veloped in the last few years, facilitating subsequent chromatographic analysis.Because of the complexity of the matrices, the sample pretreatment procedureincludes both extraction and purification of the target analytes.

Filtration is the first step of WW sample preparation because of the highloading of organic material and suspended particles. This step is essential toprevent clogging of the adsorbent bed by suspended solids, if solid-phase extraction (SPE) is performed. In the case of immunochemical assays, previousfiltration 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 filtrationprocedure [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-lineSPE (on-line SPE has been reported by Lopez de Alda et al. [60]), using eitherdisks or, more frequently, cartridges. Octadecyl (C18)-bonded silica in both car-tridge [7, 57] and disk [36] format, graphitized carbon black [35, 58, 59], Isolut


ENV+ cartridges [48], and SDB-XC disks [26, 59] are the sorbents more widelyused. The combination of C18 and SDB has also been reported, showing good recoveries for the investigated analytes [43]. Previous adjustment of the samplepH is performed sometimes [43].

The occurrence of conjugated estrogens has only been investigated in a fewworks [26, 36, 48, 61]. The analysis of these compounds by means of immuno-assays requires their previous conversion to the corresponding free estrogensby enzymatic hydrolysis [26, 36, 48]. In this case, the concentrations of the con-jugated 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 intoestradiol, but convert only ca. 30% of estradiol sulfate into estradiol [36]. Levelsof the sulfate-conjugated hormones might therefore be underestimated. Bycontrast, LC–MS enables the simultaneous determination of the free and con-jugated 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–MSanalysis. Silyl derivatives formed for example with MSTFA [43], halogenatedalkene derivatives produced with heptafluorobutyric anhydride (HFBA) [36] orpentafluoropropionic acid [58] or anhydride (PFPA), as well as acetate deriva-tives 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, bypressurized-liquid extraction (PLE), or by simple blending or stirring of thesample with polar organic solvent solutions, e.g., methanol/acetone solutions[42, 45–47, 62, 63]. Cleanup of the extracts was performed using SPE (C18 car-tridges [45, 46], RAM cartridges (ADS C4) [63], silica gel [45]), gel-permeationchromatography (GPC) [45], and semipreparative HPLC [45]. RAMs are bifunctional sorbents that combine size exclusion and reversed-phase reten-tion mechanisms.

6.3Analyses

The analysis of WW samples has been dominated by the use of immunoassaysand GC–MS techniques. However, in recent years, LC–MS and LC–MS/MS havegained in popularity, because the above-mentioned preceding hydrolysis step(needed for immunoassay analysis) and derivatization step (needed for GC–MSanalysis) are not necessary.

Biological techniques, e.g., immunoassays, are among the most sensitive analytical methods, but are limited by the availability of the specific antiseraand are subject to cross-reactivity. Huang et al. [36] employed an enzyme-linked immunosorbent assay (ELISA) for determination of estradiol, its con-jugates, and ethynylestradiol in wastewaster treatment plant effluents (seeTable 4). The reported limit of detection (LOD) of 0.1 ng L–1 reflects the sen-

20 M. Kuster et al.

sitivity of this method. Low LODs in the range of pg L–1 to 2 ng L–1 have alsobeen achieved by using other immunoassay [64, 65] and radioimmunoassay[19, 22] protocols.

The analysis of estrogens and progestogens by GC–MS has been carried outwith 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 ofsensitivity, GC– and HPLC–tandem mass spectrometry are comparable tech-niques. However, the derivatization carried out prior to GC separation is timeconsuming 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 a100 ng mL–1 standard mixture of estrogens (in the NI mode) and progestogens (in the PImode). Column: Purospher STAR RP-18e (125¥2 mm, 5 mm, Merck). Mobile phase: gradientacetonitrile/water. Flow rate: 0.2 mL min–1

and in the positive-ion (PI) mode for progestogens and, to a lesser extent,atmospheric-pressure chemical ionization (APCI) in the PI mode have beenused. Chromatographic separation has been performed on octadecyl silica stationary phases. According to a recent study carried out to compare the per-formance 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, althoughthe 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 arein general comparable, the higher selectivity of the latter is essential to avoidfalse positive determinations in the analysis of real environmental samples.Figure 4 shows representative chromatograms obtained from the analysis ofestrogens and progestogens by LC–ESI-MS/MS.

For analysis of solid samples, GC–MS/MS [45] and more frequently HPLC–MShave been used [46, 63]. Limits of detection vary from 0.04 to 4 ng g–1.

7Conclusions

What is the real danger imposed by the presence of estrogens and progestogensin aquatic systems? The studies carried out up to now indicate that the occur-rence of these compounds in surface waters is an issue of concern, and thatwastewater treatment plant effluents play a major role in their introduction intothe environment. However, more information on their environmental presence,transport, and fate is needed to assess their ultimate ecosystem impacts. Maindata gaps are localized in the area of environmental solid samples. Possible future dangers from accumulation in sediments and soil are at present unpre-dictable and should be investigated thoroughly. Besides environmental levels,toxicological data and bioavailability and degradation studies should be avail-able.

In the field of analysis, important progress has been made in terms of sen-sitivity and selectivity. LC–ESI-MS/MS appears to be the technique of choice fortheir determination as it provides reliable results at subnanogram per liter orper gram levels. However, sample preparation is identified as the main bottle-neck in the analysis of these compounds. Quite tedious and time-consumingprocedures are still required, especially in the case of complex matrices such assewage 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 progesto-gens to the described medicinal and farming applications. However, sinceWWTP effluents constitute the most important source of these compounds, re-mediation actions can be performed at this level by introducing more advancedand efficient treatment processes, especially in plants receiving high inputs ofurban discharges from highly populated regions.

22 M. Kuster et al.

Acknowledgements This work was supported by the Energy, Environmental and SustainableDevelopment Program (Project ARTDEMO EVK1-CT2002-00114), and by the Spanish Min-istry 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 Min-istry of Science and Technology.

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Technol 34:505936. Huang CH, Sedlak DL (2001) Environ Toxicol Chem 20:13337. Lai KM, Scrimshaw MD, Lester JN (2002) Sci Total Environ 289:15938. Isobe T, Shiraishi H, Yasuda M, Shinoda A, Suzuki H, Morita M (2003) J Chromatogr A

984:19539. D’Ascenzo G, Di Corcia A, Gentili A, Mancini R, Mastropasqua R, Nazzari M, Samperi R

(2003) Sci Total Environ 302:19940. Petrovic M, Solé M, López de Alda MJ, Barceló D (2002) Environ Toxicol Chem 21:214641. Xiao XY, McCalley DV, McEvoy J (2001) J Chromatogr A 923:19542. Williams RJ, Johnson AC, Smith JJL, Kanda R (2003) Environ Sci Technol 37:174443. Ternes TA, Stumpf M, Mueller J, Haberer K, Wilken RD, Servos M (1999) Sci Total Envi-

ron 225:8144. Kolpin DW, Furlong ET, Meyer MT, Thurman EM, Zaugg SD, Barber LB, Bastón HT

(2002) Environ Sci Technol 36:120245. Ternes TA, Andersen H, Gilberg D, Bonerz M (2002) Anal Chem 74:349846. López de Alda MJ, Gil A, Paz E, Barceló D (2002) Analyst 127:129947. Céspedes R, Petrovic M, Raldúa D, Saura U, Piña B, Lacorte S, Viana P, Barceló D (2004)

Anal Bioanal Chem 378:69748. Larsson DGJ, Adolfsson-Erici M, Parkkonen J, Petterson M, Berg AH, Olsson PE, Förlin

L (1999) Aquatic Toxicol 45:9149. Ankley GT, Burkhard LP (1992) Environ Toxicol Chem 11:123550. Norberg-King TJ, Durhan EJ, Robert E, Ankley GT (1991) Environ Toxicol Chem 10:89151. Petrovic M, Eljarrat E, López de Alda MJ, Barceló D (2004) Anal Bioanal Chem 378:54952. Routledge EJ, Sumpter JP (1996) Environ Toxicol Chem 15:24153. Díaz-Cruz MS, López de Alda MJ, Barceló D (2003) Trend Anal Chem 22:34054. Gomes RL, Scrimshaw MD, Lester JN (2003) Trend Anal Chem 22:69755. López de Alda MJ, Barceló D (2001) Fresenius J Anal Chem 371:43756. López de Alda MJ, Díaz-Cruz S, Barceló D (2003) J Chromatogr A 1000:50357. López de Alda MJ, Barceló D (2000) J Chromatogr A 892:39158. Laganà A, Bacaloni A, Fago G, Marino A (2000) Rapid Commun Mass Spectrom 14:40159. Johnson AC, Belfroid A, Di Corcia A (2000) Sci Total Environ 256:16360. López de Alda MJ, Barceló D (2002) J Chromatogr A 911:20361. Gentili A, Perret D, Marchese S, Mastropasqua R, Curini R, Di Corcia A (2002) Chro-

matographia 56:2562. Williams RJ, Johnson AC, Smith JJL, Kanda R (2003) Environ Sci Technol 37:174463. Petrovic M, Tavazzi S, Barcelo D (2002) J Chromatogr A 971:3764. Shishida K, Echigo S, Kosaka K, Tabasaki M, Matsuda T, Takigami H,Yamada H, Shimizu

Y, Matsui S (2000) Environ Technol 21:55365. Aherne GW, Briggs R (1998) J Pharm Pharmacol 41:73566. Mouatassim-Souali A, Tamisier-Karolak SL, Perdiz D, Cargouet M, Levi Y (2003) J Sep Sci

26:10567. Díaz-Cruz MS, López de Alda MJ, López R, Barceló D (2003) J Mass Spectrom 38:91768. Croley TR, Hughes RJ, Koenig BG, Metcalfe CD, March RE (2000) Rapid Commun Mass

Spectrom 14:108769. Ferguson PL, Iden CR, McElroy AE, Brownawell BJ (2001) Anal Chem 73:3890

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

Organic Compounds in Paper Mill Wastewaters

A. Latorre1 · A. Rigol2 · S. Lacorte1 (✉) · D. Barceló1

1 Department of Environmental Chemistry, IIQAB-CSIC, Jordi Girona 18–26,08034 Barcelona, Catalonia, Spain [email protected]

2 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 Chemical Characterization of Pulp and Paper Mill Waters . . . . . . . . . . . . 323.1 Biocides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 343.2 Resin and Fatty Acids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 383.3 Surfactants and Plasticizers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 403.4 Lignin and Hemicelluloses . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 413.5 Chlorinated Compounds . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 42

4 Toxicity of the Effluents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

5 Ain Emissions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 45

6 Removal Strategies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 46

7 Conclusions and Future Recommendations . . . . . . . . . . . . . . . . . . . . 48

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49

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 pulpwastewater management: (1) the toxicity and high BOD5 of whitewaters and effluents; (2) thelack of knowledge on specific compounds responsible for the toxicity of the liquid and solidresidue (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 hightoxicity. This chapter attempts to give an overview of organic compounds that contribute to the toxicity of paper mill waters and effluents, their levels, toxicological characterizationand the methodologies used for their analysis. Families of compounds that are included arenatural compounds such as resin and fatty acids, lignins, lignans and carbohydrates, andadditives 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 todecrease the toxicity and BOD5 of the effluents, which are used to indirectly phase out toxicorganic pollutants from paper and pulp whitewaters (Table 1).

Keywords Paper mill · Whitewaters · Effluents · Organic compounds · Analytical methods ·Toxicity · Treatment


26 A. Latorre et al.

Table 1 List of acronyms ordered by compounds and techniques

Compounds

AOX Absorbable organic halidesAPEO Alkylphenol ethoxylateBOD Biochemical oxygen demandBPA Bisphenol ABSTFA Bis(trimethylsilyl)trifluoroacetamideCOD Chemical oxygen demandDBNPA 2,2-Dibromo-3-nitrilopropionamideDCM DichloromethaneDHA Dehydroabietic acidDS Dry solidECF Elementary chlorine- freeFA Fatty acidsKj-N Kjeldahl nitrogenLAS Linear alkylbenzene sulfonatesMBT Methylene-bis-(thyiocyanate)MFO Mixed-function oxidaseMIB MethylisoborneolMTBE Methyl tert-butyl etherNOx The sum of nitrogen oxide (NO) and nitrogen dioxide (NO2) expressed

as NO2

NP NonylphenolNPEC Nonylphenol ethoxycarboxylateOC OrganochlorineOP OctylphenolOPEC Octylphenol ethoxycarboxylatePCB Polychloroinated biphenylsPCDBT Polychlorinated dibenzothiophenePCDD Polychlorinated dibenzo-p-dioxinsPCDF Polychlorinated dibenzo-p-furansPCP PentachlorophenolPFB PentafluorobenzylRA Resin acidsTCA 2,4,6-TrichloroanisoleTCDD Tetrachloro dibenzo dioxinTCF Totally chlorine- freeTCMTB 2-(Thiocyanomethylthio)-benzothiazoleTOC Total organic carbonVSC Volatile sulphur compoundsVOC Volatile organic compounds

1Pulp 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 areemissions to water and air, sludge build-up and energy consumption. It is thepollution of water bodies, however, which is of major concern because largevolumes of wastewater are generated for each metric ton of paper produced, de-pending 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 ad-sorbable 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 extrac-tives (tannins, resin acids, lignin), others are xenobiotic compounds that areformed during the process of pulping and paper making (chlorinated lignins,resin acids and phenols, dioxins and furans) [2, 3]. Some of the pollutants listedabove, notably polychlorinated dibenzodioxins (PCDD) and dibenzofurans(PCDF), are recalcitrant to degradation and tend to persist in nature.

Organic Compounds in Paper Mill Wastewaters 27

Table 1 (continued)

Techniques

APCI Atmospheric pressure chemical ionizationCZE Capillary zone electrophoresisECD Electronic-capture detectorEROD Ethoxyresorufin-O-deethylaseESIP Electrospray ionizationFID Flame ionization detectorGC Gas chromatographyGPC Gel-permeation chromatographyHRGC High-resolution gas chromatographyHRMS High-resolution mass spectrometryIS Internal standardLC Liquid chromatographyLLE Liquid–liquid extractionLOD Limit of detectionMEKC Micellar electrokinetic chromatographyMS Mass spectrometrySPE Solid-phase extractionSPME Solid-phase micro extractionTU Toxicity units

The accumulation of organic matter in whitewaters is due to the paper-mak-ing process itself. Figure 1 depicts a mass stream overview of a pulp and papermill. In paper mills, paper is made from wood, pulp or recycled paper by mix-ing the raw material with water to a fibre suspension which is ground, dilutedand evenly distributed on the wire of a paper machine. On the wire, the pulpis 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 white-waters that must be recovered and recycled to the paper machine to achievemaximum productivity. Due to this process, whitewater allows build-up oforganic matter in the system, which causes a number of troubles in paper pro-duction such as neutralization of cationic retention chemicals due to the anionicnature of organic matter, growth of microorganisms due to high concentrationsof organic matter (causing depletion of oxygen and production of hydrogen sulphide), formation of biofilms and generation of “stickies” due to lipophilicextractives which are accumulated in the paper [4].

In addition, whitewaters produce corrosion in the paper machine. The typeand amount of organic compounds in whitewater and effluents depend on theraw 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

of paper [5]. Wastewater is either released untreated, treated or recycled. Pulpand paper mill effluents discharged into freshwater, estuarine and marineecosystems alter aquatic habitats, affect aquatic life and adversely impact human health. Chronic sublethal toxic effects measured as increased livermixed-function oxidase activity (MFO) and symptoms of altered reproductivecapacity in fish and aquatic invertebrates have been detected in the dischargeof treated pulp mill effluents [6]. In addition to that, large amounts of solidwastes are generated as by-products of the wastewater treatment plants. Meanemission levels are 0.3–1 kg/ton paper [5]. Primary sludge (solids removed dur-ing physical wastewater treatment prior to biological treatment) is rich in woodfibre and volatile solids. Secondary sludge (product of biological treatment)may contain organochlorine compounds (chlorophenols, chlorocatechols,chloroguaiacols, chlorovanillins and chlorosyringaldheydes) as well as tracelevels of some dioxin and furan congeners, generated by chlorine bleaching [7].There is ample evidence of the adverse health and environmental effects linkedto organochlorinated compounds, often related to endocrine disruption(growth retardation, thyroid dysfunction, decreased fertility, feminization ormasculinization of biota, etc.) [8].

Table 2 reports some common parameters used for the characterization ofpaper mill effluents. Biodegradable organic carbon, associated with families ofnon-chlorinated organic materials, is measured by biochemical and chemicaloxygen demand (BOD, COD) methods. Some of the compounds found in paper mill effluents and sludge, including chlorinated compounds and severalwood extractive constituents found in pulp liquors, are refractory (resistant torapid biological degradation) and thus not measurable by the BOD5 analyticalmethod.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 whitewatersand sludge

Type of pollutant Typical example Levels

Air emissions Malodorous gases of reduced 0.3–3 kg/t sulphur compounds, measured of air-dried pulpas total reduced sulphurParticulate matter 75–150 kg/tSulphur oxides/nitrogen oxides 0.5–30/1–3 kg/tVolatile organic compounds (VOCs) 15 kg/t

Liquid effluents Biochemical oxygen demand (BOD) 10–40 kg/tChemical oxygen demand (COD) 10–60 kg/tAbsorbable organic halides (AOX) 0–4 kg/tKj-N 3–13 mg/LP total 0.5–1.8 mg/L

Solid wastes Sludge from primary and 50–550 kg/tsecondary treatment

toxic organic fraction of pulp and paper mill effluents to evaluate their poten-tial effects towards the environment, and the specific compounds, groups ofcompounds or organic fractions that cause harmful effects should be deter-mined. The objectives of the proposed chapter are to describe the methodsused to characterize pulp and paper mill effluents chemically and toxicologi-cally, with the ultimate goal of obtaining a deep knowledge on the compoundsresponsible for toxicity, their concentration in the different types of industriesand treatment methods used for their removal.

2Legislation 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 pollu-tion prevention and control (IPPC Directive) [5] is implementing the bestavailable technologies to ensure a high level of protection of the environ-ment. The IPPC Directive has started an exchange of information between EUmember states and the pulp and paper industries concerning the best availabletechniques aimed at: (1) improving the quality of water by installation oftreatment plants; (2) minimizing water consumption, since the pulp and papermill industry is the second highest consumer of fresh water in Europe, gener-ating six billion m3 of wastewater annually; (3) resolving the chronic toxicityand ecotoxicity associated with paper and pulp effluents; and (4) reducing the amounts of additives used (or substituting less toxic compounds). As aresult, this sector still requires data and investment to improve water quality,which will be performed by revising and establishing emission limits. To assessthese necessities, and to support the European IPPC Directives [5], newanalytical techniques are emerging to determine those pollutants which mayinduce toxicity, may negatively influence water treatment or may affect paperproduction.

Historically, the pulp and paper industry throughout the world has been re-garded as particularly polluting to aquatic environments. Until the 1950s, it wascommon for pulp mills and many other industries to discharge untreated, toxiceffluents directly into rivers and seas [9]. Nowadays, the development of newtechnologies directed to the treatment of industrial wastewaters is a EuropeanCommunity priority, which aims to reduce the organic pollution generated byindustrial 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 asthe 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 measuresto control the sources of pollution and improve the quality of water. Some European countries (e.g. Portugal, France, Germany) have started to analyse the



Tabl

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nat

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for

prod

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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 precisedata on the pollution generated by the paper and pulp industries, and identi-fication of the specific compounds responsible for the toxicity of the effluent isa subject still to be covered.

Environmental limits or guidelines for the pulp and paper industry varysignificantly between European countries, despite efforts to create a moreuniform system [5].As an example the discharge limits for bleached Kraft andbleached sulphite pulp are given in Table 3. In some countries, such as Austriaor France, there is different legislation depending on the type of paper mill.Normally, the most restrictive corresponds to the bleached Kraft pulp. In othercountries, such as Ireland, there are no requirements or limits for some para-meters.

In 1997, the US EPA finalized a new set of federal guidelines, the so-calledcluster rules [12]. The guidelines contain limits for 12 different types of mills,each of them distinguished by four different technical levels depending on thetype of technology applied. Limit values are given as pollutant load expressedas kilograms of pollutant per tonne of product, distinguishing maximum valuesper day and average monthly values. The regulations apply to any pulp, paperor paperboard mill that discharges process waters. Compared to recent permitrequirements in Europe (see Table 3) the limitations for existing mills set in theUS 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 severalgeneral guidelines, and the actual limits and guidelines can be different in eachstate [5]. The wastewater limit systems in Canada are mostly based on loadkilograms per tonne production. Different limits are applied for different typesof mills. Some parameters are measured as concentrations. Toxicity limits arealso used.

3Chemical Characterization of Pulp and Paper Mill Waters

The analysis of organic pollutants in whitewaters and effluents requiresprocedures which can eliminate suspended matter and fibres and still permitthe extraction and efficient recovery of target analytes.Whitewaters and efflu-ents, 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 particulatematter and by the formation of microfibres which, if not eliminated, may affectthe extraction efficiency. To remove suspended matter and particles, samplefiltration through 1, 0.7 and 0.45 mm filters is necessary, or otherwise centrifu-gation 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 beextracted [13]. Due to the large amount of organic matter, the extraction pro-


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

cedure must be adapted to this matrix and the chemistry of the target com-pounds should be considered. The chemical structure of the most representa-tive compounds of each family are shown in Fig. 2. Table 4 summarizes theextraction methods for different families of chemicals generally found in paper mill effluents, pointing out the detection limit and the percentage ofrecovery.

3.1Biocides

Depending on the type of paper produced, it is common to dose biocides forwood preservation and during paper making to decrease the problems relatedto microbial, fungal and algal growth. This means an undesired handling oftoxic chemicals and the risk of negative effects in the receiving water, as bio-cides are discharged with the wastewater. Furthermore, high concentrations ofbiocides in whitewater may diminish the efficiency of secondary treatment,available in some paper mills. There are two main classes of biocides [14]. Onthe one hand, there are oxidizing agents such as chlorine dioxide and hydrogenperoxide. These happen to be the same chemicals that are widely used for pulpbleaching. The oxidizing action either kills the bacteria and fungi outright orit weakens the cell walls so that they are more susceptible to the other mainclasses of biocides. The other class involves highly toxic organic chemicals, suchas thiocyanates, isothiazolins, cyanobutane, dithiocarbamate and bromo com-pounds, which are used for wood preservation in place of traditionalchlorophenols [15].A possible third category consists of materials that have anability to inhibit biological film formation, e.g. surfactants such as alkyl sulpho-succinates.

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 theirdoses. The fate of biocides is as follows: a fraction will degrade (chemically or biologically), a fraction will remain in the circulating waters and finally, afraction will be present in the effluent or remain in the solid matter. An addi-tional problem is that, due to their physicochemical properties, some biocidesmay be fibre retentive and can accumulate in the final paper product [16].Little information is available on biocides in pulp and paper mill whitewatersand effluents due to the complexity of the water matrix, but generally LC tech-niques 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 effluentwaters [17]. Figure 3 indicates the average level of biocides found in paper millwaters. DBNPA and TCMTB were detected in process waters of a recyclingpaper 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 andprimary effluent, respectively, after SPE and LC-UV detection [17]. The samemethod was found to be suitable for determining this product in paper, after


Organic Compounds in Paper Mill Wastewaters 35Ta

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70–1

0123

Dir

ect i

njec

tion

0.5–

81.3

mg/L

75–9

5

RA

and

FA

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d re

sin

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acti

on w

ith

DC

M,

LCM

S6–

12mg

n.

r.29

follo

wed

by

inje

cted

deri

vati

zati

on s

tep

RA

and

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r-re

cycl

ing

proc

ess

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h M

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S0.

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81–1

0617

,32

wat

ers

follo

wed

by

mg/L

deri

vati

zati

on s

tep


Tabl

e4

(con

tinu

ed)

Com

poun

dsM

atri

xM

etho

dLO

DR

ecov

erie

s R

ef.

(%)

Extr

acti

onSe

para

tion

Det

ecti

on

NP 1

EC,O

P 1EC

Pa

per-

recy

clin

g SP

ELC

ESI-

MS

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0mg

/L<

9017

NP,

OP,

LAS

proc

ess

wat

ers

BPA

Was

tew

ater

from

pap

er

SPE

GC

MS

n.r.

n.r.

35pr

oduc

tion

NPE

CPa

per

mill

eff

luen

tsSP

EG

CM

S0.

2–2

mg/L

90–1

1036

Hem

icel

lulo

ses

Pulp

fibr

es 1

%A

cid

met

hano

lysi

s

GC

FID

n.r.

n.r.

38w

ith

HC

l,fo

llow

ed

by a

sily

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on s

tep

Hyd

roly

sis

and

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r.n.

r.39

met

hyla

tion

rea

ctio

n

Lign

inPa

per

mill

eff

luen

tsPe

rman

gana

te

CZE

UV

0.55

–2.9

mg/

Ln.

r.46

de

grad

atio

n,fo

llow

ed

MS

0.05

–0.3

0n.

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per

mill

eff

luen

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MS

n.r.

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llow

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y a

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tep

OC

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ents

from

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p G

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r.n.

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d pa

per

mill

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F

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ed p

ulp

mill

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ion

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CH

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10ng

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r.54

efflu

ents

filte

rs w

ith

tolu

ene


Tabl

e4

(con

tinu

ed)

Com

poun

dsM

atri

xM

etho

dLO

DR

ecov

erie

s R

ef.

(%)

Extr

acti

onSe

para

tion

Det

ecti

on

OC

,PC

B,N

esti

ng a

long

G

PCG

CED

Cn.

r.n.

r.55

PCD

D,P

CD

Fco

ntam

inat

ed r

iver

sH

RG

CH

RM

Sn.

r.n.

r.

Chl

orop

heno

ls,

Efflu

ent f

rom

ble

achi

ng

LLE

wit

h di

ethy

l G

CFI

Dn.

r.n.

r.27

ch

loro

guai

acol

s,pr

oces

ses

ethe

r/ac

eton

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llow

ed

chlo

rova

nilli

n,by

der

ivat

izat

ion

step

chlo

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tech

ol

Chl

orop

heno

ls,

Pulp

mill

eff

luen

tsSP

ELC

Am

pero

- 0.

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84–1

0037

ch

loro

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c ch

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Chl

orop

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ls,

Con

tam

inat

ed s

edim

ent

Dea

n–St

ark

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let

GC

MS

n.r.

n.r.

52ch

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guai

acol

s,ex

trac

tion

chlo

rova

nilli

n,ch

loro

cate

chol

Chl

orop

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men

tEx

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tion

wit

hG

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Sn.

r.n.

r.57

chlo

rogu

aiac

ols,

N-h

exan

e,fo

llow

ed b

y ch

loro

cate

chol

,de

riva

tiza

tion

ste

p

VSC

Air

,wat

er a

nd s

edim

ents

Cry

ogen

ic tr

apG

CFI

D10

pg/L

n.r.

68

VSC

Pulp

mill

eff

luen

tsH

S-SP

ME

GC

MS

0.7–

5ng

/Ln.

r.69

LLE

wit

h D

CM

GC

MS

n.r.

n.r.

70

MIB

,geo

smin

Riv

erLL

E w

ith

Hex

ane

GC

MS

0.5

ng/L

97–1

0371

Dri

nkin

g w

ater

HS-

SPM

EG

CM

S1.

2–3.

3ng

/Ln.

r.73


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 deter-mine TCMTB and chlorophenols [19] from lumber surfaces after extraction ofsticks (1¥2 mm) with acetonitrile. Micellar electrokinetic chromatography(MEKC) was optimized for the determination of ten biocides in paper foodpackaging, with the inherent advantages of rapidity, simplicity and no use oftoxic reagents [16].

3.2Resin and Fatty Acids

Wood extractives include lipophilic (fatty and resin acids, sterols, steryl estersand triglycerides) and hydrophilic (lignans, low-molecular-mass lignins, lignin-like substances and hemicelluloses) compounds that dissolve in whitewatersduring paper production. Among them, resin and fatty acids have a high ten-dency 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 showsthe concentrations of resin and fatty acids in whitewaters.The levels detected de-


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 effective-ness of this treatment in reducing the concentration of these type of com-pound [23].

The analysis of resin and fatty acids is fully reviewed elsewhere [24]. Thestructures of the most representative resin and fatty acids, dehydroabietic andpalmitic acids, are shown in Fig. 2. Liquid–liquid extraction (LLE) has beenused to extract resin and fatty acids [25]; methyl tert-butyl ether provides anexcellent solvent for the extraction of these compounds [26–29]. WhitewaterpH is generally between 6 and 8, and extraction can be performed under neu-tral or alkaline conditions [30], although acidic conditions avoid microbialgrowth during sample storage [28]. The determination of resin acids in woodextractives has been traditionally performed by GC with a flame ionization de-tector (GC-FID) on either a DB1 or HP5 analytical column [31]. This procedurepermits determination of the different compounds (resin acids and hemicel-luloses) at microgram per liter levels.With non-selective detectors, compoundidentification is performed by retention time comparison against a standard.However, the analysis of complex water matrices leads to interferences andcoelutions, 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 frag-ment or cluster ions are generated which provide structural information on theionized compound. With GC, derivatization of the extract is needed either using methylation agents such as diazomethane, which presents severe healthhazards, or bis(trimethylsilyl)trifluoroacetamide (BSTFA) combined with tri-methylchlorosilane. Another derivative of resin acids (RAs) is the pentafluo-robenzyl (PFB) esters. Figure 4a shows a GC–MS chromatogram which permitsthe complete separation of 15 resin and fatty acids, after derivatization withBSTFA. The main problem is that the derivatized extract has a very short half-life, 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-termperformance of the GC–MS system which will necessitate extra cleaning.GC–MS permits identification of resin and fatty acids from whitewaters of dif-ferent paper mills [32].

Given the poor volatility of resin and fatty acids, liquid chromatography cou-pled to mass spectrometry in the negative-ion mode can be used without theneed for derivatization [31, 34]. APCI and ESI are suitable interfaces, althoughno fragmentation is observed even at high fragmentor voltages. Figure 4bshows a LC–MS chromatogram. There is coelution of non-aromatic RAs whenusing either C18 or C8 columns, but this is not a problem since the concentra-tion 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


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

a

b

effluent discharge, the concentration of resin acids found in river sediment is139 mg/g. [22]. This level is ten times lower compared with the level found in thesludge of this Kraft mill. A similar behaviour was observed for the analysis ofresin and fatty acids in effluent and river water closed to the paper mill. Theconcentration decreases in most cases, due to degradation and dilution. Onlyfor some resin acids, such as dehydroabietic acid, was the concentrationobserved 11 km downstream similar to the level found in the source, due totheir high stability [26].

3.3Surfactants and Plasticizers

Surfactants, such as linear alkylbenzene sulfonates (LAS) and alkylphenolethoxylates, are present in whitewaters because of their use as cleaning agentsor as additives in antifoamers, deinkers, dispersants, etc. The non-ionic sur-factants alkylphenol ethoxylates (APEO) degrade to nonylphenol (NP) or to a

lesser extent, to octylphenol (OP), which are considered as persistent environ-mental pollutants. LAS and APEO have been detected in whitewaters of papermills at concentrations up to 5,000 mg/L for the former [16] and from 0.3 to10 mg/L for NP and OP [35]. The total concentration of nonylphenol ethoxy-carboxylates (NPEC) in paper mill effluents ranged from below detection to1,300 mg/L [36]. These compounds can also be analysed by LC–MS, with the additional advantage that long-chain alkylphenol ethoxylates and carboxylatescan be simultaneously determined [34]. Due to the complexity of the matrix,the identification and quantification of compounds should be controlled by theaddition of an adequate surrogate or internal standard. Heptylphenol can beused for the analysis of alkylphenols.

Paper mill whitewaters and effluents are rich in bisphenol A (BPA), which isused in great quantities for the production of epoxy resins and polycarbonateplastics. Its presence in effluents has been reported as a result of its use in themanufacture of thermal paper or due to migration from plastic containers atthe high water temperatures of whitewaters [35]. This compound is preferablyanalysed by GC–MS. The levels encountered in paper mill effluents are between28 and 72 mg/L [36, 37].Another study revealed levels up to 226 mg/L [33]. Spe-cial in vitro test systems and animal experiments have demonstrated a weak oestrogenicity for BPA. Since aquatic wildlife could be endangered by papermill waste discharges at the concentration that BPA is found, its survey in papermill effluents should be taken into consideration.

3.4Lignin and Hemicelluloses

Wood consists mainly of cellulose, hemicelluloses and lignin in various pro-portions. The amounts and compositions of these component groups dependprimarily on the wood species [38].

In chemical pulping, a significant part of the hemicelluloses is dissolved fromthe fibres into the pulping liquor.The rest remains in the fibre or is adsorbed intoit, significantly affecting the properties of the cellulose fibres or paper produced[39]. It has been demonstrated that the presence of soft or hardwood hemicel-luloses in the cellulosic pulp can improve some features of paper making. Theplasticity and the high superficial area conferred by hemicelluloses result in anincreased binding among the fibres and a higher tensile strength in the papersheet. However, high amounts of hemicelluloses seem to be deleterious to themechanical properties of the paper due to a decrease in the individual fibre re-sistance, and to the optical properties due to the low opacity in the paper sheet[40].Some studies have demonstrated a relationship between the degradation ofhemicellulose components, such arabinose and mannose, and wood strengthlosses. The significant reduction in strength observed during incipient decay of wood by brown rot fungi is therefore likely to be due to hemicellulose de-composition [41]. A basic method for the analysis of hemicelluloses is the de-termination of their constituent sugar residues obtained by acid hydrolysis


or methanolysis (hydrochloric acid in anhydrous methanol). The liberated mo-nosaccharides 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 methylglycosides, and allows simultaneous analysis of acid and neutral sugars by cap-illary GC–MS or LC after suitable derivatization [42].

On the other hand, lignins are natural polymers in plant cell walls and rep-resent, after cellulose, the most abundant polymer in nature, with a very com-plex structure. The lignin composition will be different not only among plantsof 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 duringpulping requires high amounts of energy and chemicals [45]. Extracted ligninsfrom non-wood fibres are potential raw materials for new industrial applica-tions [43]. As a result, lignin should be monitored.

In order to obtain information about the structure of lignin there were somestudies based on the oxidative treatment of the molecule by potassium perman-ganate [46]. This method involves the selective degradation of all aliphatic sidechains attached to aromatic groups in lignin, resulting in the formation of a mix-ture. The identification of these as well as the amount of each individual acid pro-vides information about the substitution pattern in a particular lignin. From thedegradation product obtained it is possible to deduce the structure of lignin. Upto now, the mixture of aromatic acids obtained from permanganate oxidation oflignin has been analysed by GC after esterification [47], but Javor et al. developeda new methodology using capillary zone electrophoresis (CZE), which providesrapid results with the avoidance of time-consuming preparation of esters of theresulting aromatic acids. Another methodology, based on the CuO degradationof lignin, is also considered a suitable technique for their analysis [48].After CuOdegradation of lignin, nine products corresponding to three lignin units (p-hy-droxyphenyl, guaiacyl and syringyl) could be identified. The degradation prod-ucts can be easily derivatized, separated by GC and identified by MS.

3.5Chlorinated Compounds

Since the late 1970s, much emphasis was put on the role of chlorinated sub-stances formed in the bleach plants. Bleaching effluents from bleached chem-ical pulp plants are one of the remaining pollution problems of pulp mills dueto the large amounts of chlorinated organic matter discharged into the envi-ronment [49]. The bleaching of chemical pulp is accomplished in several stages,to some of which chlorine is added in different forms. The chlorine reacts withlignin and other organic matter present in the pulp giving chlorinated com-pounds. During the last decade, there has been a drastic decrease in the use of


molecular chlorine as bleaching agent, which has been replaced by chlorinedioxide, molecular oxygen, peroxide and ozone. This has led to a decrease in ad-sorbable organic halides (AOX), which is the main parameter used by regulatoryagencies to determine the discharge of chlorinated organics [50]. Reduction ofAOX has also been achieved by the installation of treatment plants. Currenttrends are directed towards closed-cycle systems using either elementary chlo-rine-free (ECF) or totally chlorine-free (TCF) bleaching pulp. However, this isonly possible if no chlorinated agents have been used within the process. Forsome time chlorophenols, and especially pentachlorophenol (PCP), were usedas 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 re-stricted and the wood chain is organized so that wood is rapidly consumed andthe use of fungicides is minimized.

The presence of chlorine and chlorinated compounds is also the source ofdioxins 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 studyfound high concentrations of PCDD and PCDF along with PCP in nestling tis-sue (Tachycineta bicolor) collected downstream of paper pulp mills, suggestingthat the primary source of contaminants was the use of PCP for timber preser-vation [55]. In addition, it has been shown that dioxins bioaccumulate in fishdownstream of pulp and paper mills [56]. The levels of chlorinated compoundsof different families are shown in Fig. 3.

The survey of PCP and other chlorinated compounds has been traditionallyperformed by the measurement of AOX, which gives a measure of the total chlo-rinated organic compounds [57]. Typical AOX levels are between 0.01–0.1 kg/t.However, to specifically determine the different families of organochlorinatedcompounds in paper mill whitewaters and effluents, several analytical methodshave 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 electroncapture detection (ECD) or MS. However, there is a general trend to use SPE andLC to avoid the use of toxic organic solvents and derivatization procedures. Acomplete review of LC methods for the analysis of chlorophenols is given else-where [58]. Levels of chlorinated organic compounds in paper mill waters arebetween 1 and 100 mg/L, as shown Fig. 3. The analytical protocol for the analy-sis of dioxins and furans is well established and follows the EPA method 8280A.

4Toxicity 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 ofjuveniles, changes in communities and population structure, changes in growthrates, and delayed sexual maturation and reproduction, among others [2, 58, 59].


In addition,oxygen depletion is common in such effluents,causing anoxia to fishand other aquatic specimens. The toxicity of paper mill whitewaters and efflu-ents can be measured using acute toxicity tests such as Microtox or ToxAlert.These tests measure the bioluminescence inhibition of Vibrio fischeri caused bythe presence of different toxicants in the water sample. Toxic substances willcause changes in cell structures and/or metabolic pathways of marine Vibrio fischeri, which are rapidly reflected in a bioluminescence decrease. The LC50value of several surfactants, resin acids, fatty acids and biocides has been determined by ToxAlert using individual compounds and mixtures, and thecombination of chemical analysis and effect studies permitted the toxicity ofwhitewaters and effluents of several paper mills to be assessed [33]. Figure 5represents the percentage of bioluminescence inhibition using ToxAlert and thetotal organic load of sample (sum of resin and fatty acids, surfactants and bio-cides) of an untreated effluent and whitewaters corresponding to different pa-per mills which had undergone several treatments. In cases where the concen-tration of organic compounds was high, a high percentage of bioluminescenceinhibition was observed. On the other hand, in four samples the organic loadwas low, as well as the percentage of inhibition using ToxAlert. However, un-treated Kraft and print paper showed a low organic compound load and a hightoxicity. This is attributed to the presence of other compounds not considered


Fig. 5 Total organic composition (resin and fatty acids, biocides and surfactants) and per-centage of bioluminescence inhibition of several types of waters (recycle, Kraft, print boardin 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

or detected with chemical analysis. Nevertheless it can be concluded that thecombination of chemical analysis and bioluminescence inhibition assay permitsevaluation 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 ameasure of the activity of the cytochrome P-450 enzyme system, which playsa central role in the transformation and elimination of xenobiotics. IncreasedEROD activity has been shown as far as 40 km from pulp mills, and EROD in-duction in fish caused by pulp mill effluents remains after biological treatment[60]. It is specified that EROD activity and erythrocytic nuclear abnormalitiesare induced by abietic and dehydroabietic acid [60].

However, it is difficult to identify the chemical compounds that are respon-sible for these effects. LC50 values have been tested in several fish species andlevels below 2 mg/L have been reported for resin acids [61] and below 0.1 mg/Lfor some biocides used in the paper industry, such as MBT and TCMTB [34].Wood extractives (resin and fatty acids, sterols, etc.), diterpene alcohols andjuvabiones account for 70–100% of the toxicity in various paper mill effluentstreams [62]. However, toxicity depends on the treatment [63] and recent papers relate toxic effects towards aquatic biota due to the presence of resinacids in a secondary-treated bleached Kraft pulp mill effluent [64], and due tononylphenol polyethoxy carboxylate metabolites of non-ionic surfactants in aUS paper mill effluent [65]. Moreover, resin acids and, to a smaller extent, un-saturated fatty acids have been reported as major contributors to the toxicityof paper industry effluents to aquatic organisms, causing chronic sublethal tox-icity, genotoxicity and potential bioaccumulation in fish tissues [65]. Endocrinedisruption is being highlighted in modern toxicology. Relatively little is knownabout the potential endocrine effects of paper mill effluents on aquatic organ-isms. Field surveys on fish in proximity to sewage plants show hermaphrodismand laboratory studies also confirm this phenomenon [66].

5Air Emissions

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


as the pulp and paper industry is continuously faced with more stringent air emission limits concerning gaseous pollutants. Most of the techniques forthe analysis of VSC in aqueous matrices use the purge-and-trap method withcryogenic trapping of these analytes in glass tubes [68]. The use of solid-phasemicroextraction (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 and24 mg/L, with dimethyl sulphide being the main compound detected, which mayoriginate from the sodium hydrosulphite and sodium metabisulphite used asbleaching reagents during the process. Lower levels were found in a bleachedKraft pulp mill effluent, with values around 0.5–2 mg/L [71].

The release of mill wastewater effluents may be a significant contributor tomill 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 springrunoff and/or eutrophic systems [72], and are responsible for the majority of thereported taste and odour events in surface waters close to paper mills. Currentmethods 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 ma-nipulation and time consumption [73]. Watson et al. analysed different papermill 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-3-methoxypyrazine and 2-isobutyl-3-methoxypyrazine, were found downstreamof a pulp mill effluent, and were considered as off-flavours. These compoundsare by-products of chlorination, or can be produced by actinomycetes or otherbiota [74].

6Removal Strategies

For both environmental and economic reasons, many paper mills aim at lowerwater consumption and a decrease of water discharge. These can be achieved byrecycling water but unfortunately, closing the water system is far from easy because an increased recycle of whitewaters leads to an accumulation of solu-ble organic matter and salts in the paper mill. The advantages and disadvantagesof closed-cycle systems in paper mills are shown in Table 4 [61]. However, theproblems derived from a build-up of organic matter in the whitewater systemshas forced many mills that have been trying a closed-cycle approach to open uptheir systems again and continue to discharge great amounts of wastewater.

The first effect of paper mill wastewater discharge is the depletion of oxygenin the receiving waters, caused by oxygen-consuming microbial degradation of


readily biodegradable organic matter being discharged. This has led to a rapidinterest in developing new methods for in-mill treatment of whitewater to re-move organic matter.A review of the treatment of pulp and paper mill effluentsindicates processes that minimize the discharge of wastewater into the environ-ment [75]. Evaporation techniques and chemical treatment are very costlyoperations, and membrane filtration often suffers from fouling problems, whichdecrease the efficiency and increase the operating costs [76]. Biological treat-ment is undisputedly the most effective and economical way of removing greatamounts of organic matter from wastewaters.

The possibility of using in-mill biotreatment was proposed in the 1980s, andduring the 1990s biological treatment and reuse of recycle fibre mill processwater was applied in some mills [77, 78]. The objective of these treatments isto reduce BOD, which is the direct cause of oxygen consumption. However,being rather conventional, biological treatment plants operating under normalbiological conditions (<40 °C, pH around neutral) require extensive modifica-tions of the environmental conditions, such as cooling. This is costly and highlyundesirable as it causes heat losses and less efficient production in the papermill, which is optimally operated at higher temperatures. The insight into thishas led to a number of studies on the possibility of operating in-mill treatmentat higher temperatures [79–81]. It has even been possible to treat acidic white-water with high efficiency at a pH as low as 3.5 [81]. Effluents from the mill aretreated in bioprocesses such as aerated lagoons or activated sludge, whereaswhitewaters undergo an anaerobic treatment followed by activated sludge. Theefficiency of the treatments is controlled through measurements of generic parameters such as COD and BOD. It is assumed that removing as much of theorganic 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 ofbiologically treated whitewaters have been encountered:

– Biological treatment removes the bulk of the organic matter, but the fractionremaining, often dominated by lignin, makes biotreatment difficult. Thisgives a significant increase in the colour of the treated water, and unaccept-able colouring of the product for such paper qualities for which the colouris important.

– Aerobic biotreatment effectively eliminates odours from organic acids andsulphide. However, in cases where biotreated water has been reused in paper production, the product has suffered from a weak “soily” smell that isunacceptable 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 whitewa-ter 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 in-creased slime problems, rather than the decrease that is the aim of biotreat-ment [81].


– In cases where water is discharged, considerable effects on life in the receiv-ing waters, especially from chemical pulping effluents, have been encountered.Pathological changes in fish have been observed, as well as effects on physi-ological and biochemical parameters. It is obvious that these serious effectson the ecosystem are not due to readily biodegradable organic matter, butrather to compounds more resistant to biological treatment.

These factors have stopped the installation of biological treatment plants at anumber of large Kraft mills, which now continue to discharge great volumes ofuntreated effluent. However, there is a growing tendency to install advancedpost-treatment stages to deal with the remaining problems. The combinationof biological treatment and membrane filtration has found a special interest.Pauly and Kappen [82] studied the combination of thermophilic anaerobictreatment and ultrafiltration, and found the biotreatment improved the per-formance of filtration. However, the problems with odour remained. The im-proved performance of membrane filtration after biological treatment was alsoobserved by Nuortila-Jokinen [76], in which case biotreatment was to be con-sidered more of a pre-treatment and filtration the main treatment. However,combined biotreatment and advanced filtration, such as ultrafiltration andnano-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 organiccompounds responsible for the toxicity of paper and pulp effluents and relatedemissions and (2) reduction of the amount of solid waste going to landfills.

7Conclusions and Future Recommendations

The pulp and paper industry is the greatest industrial polluter in terms of waste-water volumes and organic discharge. Compounds encountered in whitewatersare natural wood components such as resin and fatty acids, and additives addedin the process such as wood preservatives, biocides and surfactants and plasti-cizers. Since the introduction of the best available technologies and accordingto the IPPC Directive, there has been an improvement in the pulp and papersector such as minimization of the use of chlorine, additives, energy and freshwater which has lead to a reduction of emissions of toxic compounds to water,air and sludge. Generic parameters such as COD, BOD, AOX, total suspendedsoils, SO2 and NOx are systematically controlled and maximum discharge lim-its are well satisfied. However, a recent IPPC Reference Document on the pulpand paper industry indicates that there is insufficient information on the organic composition of whitewaters, effluents and sludge from pulp and papermills, and on the sampling and analytical methods that should be used for their


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 com-position, and by the toxicological characterization of whitewaters, effluents andsolid waste. This chapter has attempted to give an overview of the organic com-pounds present in pulp paper mill whitewater, the levels encountered and theirtoxicological implication. It has also highlighted the treatments performed andthe tools which are nowadays used to remove COD, toxicity and organic loadof pulp and paper mill whitewater for an environmentally friendly paper pro-duction process. Recently, much effort has been devoted to correlating toxicitystudies and the chemical characterization of pulp and paper mill effluents. Thispermits a much stricter control of the treatment that should be performed andof the quality of the water, which still in many cases is discharged to the envi-ronment.

Acknowledgements This study has been supported by the EU Energy, Environmental andSustainable Development Program (CLOSEDCYCLE, Contract No. EVK1-2000-00749) andMinisterio de Ciencia y Tecnología (PPQ2000-3007-CE). T. Welander and A. Malmqvist areacknowledged for providing information on pulp and paper mill treatments.

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technology, Munich, October 25–27 1999, pp 6:1–6:19


Evaluation of Pesticides in Wastewaters. A Combined(Chemical and Biological) Analytical Approach

M. D. Hernando1 · I. Ferrer2 · A. Agüera2 · A. R. Fernandez-Alba2 (✉)1 Department of Environmental Chemistry, IIQAB-CSIC, Jordi Girona 18–26,

08034 Barcelona, Spain 2 Department of Analytical Chemistry, University of Almería, 04120 Almería, Spain

[email protected]

1 Introduction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

2 Chemical Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 552.1 Sample Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 552.1.1 Liquid–Liquid Extraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 562.1.2 Solid-Phase Extraction . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 562.1.3 Semipermeable Membrane Devices and Other Membrane Processes . . . . . . 572.2 Cleanup Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 582.3 Methods of Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 602.3.1 Gas Chromatography . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 602.3.2 Liquid Chromatography–Mass Spectrometry (LC–MS) . . . . . . . . . . . . . 64

3 Toxicity Biological Assays . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 663.1 Bioassays Applied to Evaluate the Toxicity of Pesticides . . . . . . . . . . . . . 663.1.1 Acute Toxicity Bioassays . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 673.1.2 Chronic Toxicity Bioassays . . . . . . . . . . . . . . . . . . . . . . . . . . . . 693.2 Toxicity Studies of Wastewater Containing Pesticides . . . . . . . . . . . . . . 70

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 74

Abstract The current status of the analysis of pesticides in wastewater by chromatographictechniques and toxicity bioassays is reviewed and evaluated. When using chromatographictechniques, the low concentrations of pesticides present and the complexity of the waste-water matrices require a sample concentration step prior to measurement. Also, cleanuptechniques need to be applied for better detection of the analytes and to avoid ion suppres-sion. The most commonly used methods of analysis for the detection of pesticides in waste-water samples involve GC–MS and LC–MS. However, an evaluation only based on chemicalanalysis 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 ofpesticides in complex samples such as wastewaters. They provide a response that relates tothe overall effects (synergism, antagonism) of the chemicals present in wastewaters and theyassess the short- (acute) and long-term (chronic) effects. Therefore, both chemical and biological analytical strategies are relevant to the correct evaluation of pesticides in waste-waters, their behavior during wastewater treatment, and the reuse of water resources.

Keywords Pesticides · Wastewater · GC–MS · LC–MS · Toxicity bioassays


1Introduction

Pesticides in wastewaters come typically from point sources of contaminationsuch as disposal sites and landfills where industrial or agricultural wastes areburied without any consideration, as well as discharges from industrial effluentsfrom pesticide production plants. Furthermore, nonpoint sources derived fromregular agricultural activities, especially in intensive agricultural areas, and accidental spills can also be significant. Urban use of pesticides is also possiblein large cities where the use of herbicides and insecticides may result in runoffinto the sewers. These sewers in turn may expel pesticides into wastewatertreatment plants (WWTPs).

Due to the partial to complete resistance of many pesticides to biodegrada-tion during the wastewater treatment processes, these compounds can escapeelimination in WWTPs and enter into the aquatic environment. As a conse-quence, their evaluation represents an important objective in the efficiency ofWWTPs 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 de-veloped countries,and subsequently a drastic reduction of emission was achievedafter adoption of appropriate measures [1]. Today, in industrialized countries wecan consider the presence of these compounds as having less importance. Butthey are used as effective pesticides (e.g., lindane, DDT) and still represent a bigissue in developing countries in terms of the environment and human pollution[2–4]. Awareness of the presence of nonpolar pesticides in wastewaters isachieved mainly through the use of gas chromatography.Conversely,a “new”gen-eration of pesticides considered as “emerging contaminants” with a wide rangeof structures and typically with high polarity has only been recognized for thelast few years. As a consequence of this, high polarity and sometimes thermallylabile 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 33priority substances issued in the last EU Water Framework Directive [5].

In general, we can consider the analytical methods for pesticides well doc-umented and evaluated as a consequence of the important routine monitoringprograms for food and drinking water [6–13]. Nevertheless, the complex natureof wastewaters is a great limitation to chemical analyses in their ability to totally evaluate pesticide content in the low microgram per liter range (or evenbelow that).A second point of interest is the large number of pesticides, around800, on the market with a very wide range of structures and physicochemicalproperties, which makes it very difficult to develop adequate target multiresiduemethods that cover enough of them, even without taking into consideration theformation of possible transformation products.

Consequently, there is a lack of knowledge concerning this kind of pollutionand a need to apply sophisticated and powerful analytical techniques to per-

54 M. D. Hernando et al.

form adequate identification and quantification of pesticides in wastewaters. Insuch situations, chemical analysis based on the concentrations of a limitednumber of compounds has serious limitations, even if the treated effluentsmeet 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 treatmentprocess to allow the reuse of wastewater. Furthermore, wastewaters are not pol-luted by a single chemical, but rather by a mixture of numerous chemicals, andthis 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 eachindividual chemical is below its threshold concentration because of interactiveeffects among them. It means that the combined effect of various chemicals can be the result of additive effects of individual chemicals, or they can evenproduce a greater toxic effect showing synergism [14, 15].

In the light of these limitations, effective additional tools able to assess thebiological responses of the pesticides present, as well as their interaction withthe other chemicals, have to be introduced to complete the evaluation of waste-waters. Bioassays on water samples provide a direct functional response thatcan relate to the negative effects of a single pesticide and overall toxic proper-ties of the complex mixture of compounds present in a sample [16].

This study is an overview focused on the application of the main analyticalstrategies based on chemical analysis and biological toxicity assays for pesti-cides, to be used as a combined approach for the evaluation of pesticides inwastewaters.

2Chemical Analysis

2.1Sample Treatment

Due to the predicted and previously detected low concentrations of pesticidesin environmental samples (usually around the nanogram per liter level), a pre-concentration step of the water samples is necessary prior to measurement. Inthis way, a preconcentration factor of several orders of magnitude (200–1,000-fold) is mandatory to reach the low detection limits necessary for the iden-tification of pesticides, especially in complex wastewater samples.Also, the useof surrogate standards (e.g., triphenyl phosphate) added before the extractionstep is a common practice in order to account for possible errors during the extraction process and for quantitative purposes. The commonly used extrac-tion methods for polar compounds from water matrices involve isolation usingliquid–liquid extraction (LLE) and solid-phase extraction (SPE), which are com-mented on below. Other methods such as semipermeable membrane devices(SPMD) are also mentioned.

Evaluation of Pesticides in Wastewaters 55

2.1.1Liquid–Liquid Extraction

LLE has been used in the past for the extraction of pesticides from environ-mental water samples [17]. However, its application in the extraction of waste-water samples is scarce due to the low efficiency of extraction, especially for polar analytes. Because of the vast amount of surfactants and natural productspresent in wastewater samples, emulsions are formed which complicate theprocess of extraction and lead to low extraction recoveries. However, there havebeen some useful applications of LLE to wastewater analyses. For example, LLEwas found to be effective for the isolation of herbicide and pesticide organiccompounds from industrial wastewater samples and also from complex ma-trices [18].

2.1.2Solid-Phase Extraction

SPE procedures are used not only to extract traces of organic compounds fromenvironmental samples, but also to remove the interfering components of thecomplex matrices in order to obtain a cleaner extract containing the analytesof interest. In this sense, it is a good sample treatment method for the analysisof wastewater. In the last few years, there has been a considerable interest in developing new selective and sensitive methods for extracting and isolatingcomponents from complex environmental matrices. The selectivity is the degree to which an extraction technique can separate the analyte from inter-ferences in the original sample.Accordingly, the selectivity of stationary phasesis an important parameter to be taken into account when compounds are to beextracted from wastewater samples, since the main objective is to remove interferences and facilitate further analysis by conventional analytical method-ologies such as gas chromatography (GC) or liquid chromatography (LC).

SPE using C18 or polymeric phases has been used widely for the determi-nation of pesticides in water samples [19, 20]. These stationary phases are gen-erally nonselective and can lead to difficulties with interferences coextractedfrom the wastewater matrices. Most of the polar pesticides cannot be deter-mined owing to their coelution with the matrix peak, which is obtained at thebeginning of the chromatogram when wastewater samples are analyzed bychromatographic techniques [21]. This matrix peak is a coeluting interferentcaused by humic substances present in natural waters. The chromatographicmethodologies used are commonly not selective toward the coextracted com-pounds present in environmental samples, and consequently it is of primaryimportance to use a selective sorbent for the preceding step (SPE) in the entireanalysis. 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 determinationof traces of organic contaminants in complex environmental samples. In other


words, the more selective the SPE step, the better the sensitivity achieved. Forthese reasons, efforts have been made to develop new selective sorbent mate-rials 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 toother bonded silicas, such as the C8 or CN types [22]. Applications of C18 sor-bents include the isolation of hydrophobic species from aqueous solutions. Themechanism of interaction with such sorbents depends on van der Waals forces,and secondary interactions such as hydrogen bonding and dipole–dipole in-teractions. Nevertheless, the main drawbacks of such sorbents are their limitedbreakthrough volumes for polar analytes, and their narrow pH stability range.For these reasons, reversed-phase polymeric sorbents are also used frequentlyin environmental applications for the trace enrichment of soluble moleculesthat are not isolated by reversed-phase sorbents such as C18.

The most widely used polymeric sorbents are the styrene–divinylbenzenecopolymers (SDB), which are among the classical reversed-phase sorbents introduced in the 1960s [20]. They are currently produced in purified form andare useful for the isolation of more polar solutes that have low capacities on the C18 reversed-phase sorbents. Their broader pH-stability range increases theflexibility of the method since the pH of the wastewater samples is usually inthe high range. Moreover, these kinds of sorbents have a greater surface area pergram, so they can retain the most water-soluble analytes.Another advantage ofthe aromatic sorbents derives from their selective interaction with aromaticrings in the analytes. Because the styrene–divinylbenzene structures containaromatic rings, they have the ability to sorb analytes by specific p–p interac-tions. More recently, many immunosorbents based on antigen–antibody inter-actions have been developed for the selective isolation of many pesticides inwater samples [23]. They have proven to be very suitable for the highly selec-tive preconcentration of organic contaminants from complex environmentalsamples, such as sediments and sludges. Since such sorbents are tailor-made forspecific applications, their cost is high compared to conventional sorbents [24].However, they are very limited for multiresidue applications and therefore onlyuseful 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 con-taminants in complex mixtures in forthcoming years [25].

2.1.3Semipermeable Membrane Devices and Other Membrane Processes

SPMD have gained widespread use for sampling hydrophobic chemicals fromwater. In these membranes the more hydrophobic compounds are retained andare further recovered with organic solvents. As an example, SPMD have beenapplied to the analysis of pesticides in wastewater samples [26].


Other membrane processes such as microfiltration, ultrafiltration, reverse osmosis,and colloid-enhanced ultrafiltration have been applied to the separationof beta-cypermethrin from wastewater samples [27]. In this study, a separationof above 92% was performed by reverse osmosis by the use of composite mem-branes and above 80% by colloid-enhanced ultrafiltration by the use of nonionicsurfactants.

2.2Cleanup Techniques

In environmental analysis, organic compounds are usually present at low con-centrations and are often masked by complex patterns of interfering compo-nents. Therefore, cleanup steps are necessary for the analysis of wastewater sam-ples, 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 gaschromatographic techniques. SPE can be easily applied as a cleanup method forthis kind of matrix as well. Accordingly, sequential SPE has been applied as apreconcentration and cleanup method in the analysis of some pesticides inwastewater samples before analyses by liquid chromatographic techniques [28].Other cleanup methods involve the rapid and effective anion-exchange ca-pacity of the anion-exchange phases to remove humic substances, which arepresent in complex water and soil samples [29]. The high selectivity of the SAXdisk for humic substances allows these interferents to be effectively removedfrom water samples during trace enrichment of herbicides from complex waterextracts. The concept of layering disks can be used to first remove the humicimpurities on the SAX disk with simultaneous isolation of herbicides on thelower C18 disk. The concept of stacking adsorbents for trace enrichment wasfirst introduced in the early 1980s with XAD adsorbents. Both anion-exchangeand reversed-phase methods can then be used to isolate both natural and con-taminant organic compounds from water. More recently, SPE cartridges havebeen introduced with layered adsorbents, which facilitate treatment of theaqueous 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 theintensity of the analyte signal.As can be observed in Fig. 1, analyte suppressionoccurs as a consequence of the different matrix interferences present in waste-water samples, making the identification and/or quantification process difficultor unfeasible. Even when working under selection ion monitoring (SIM) con-ditions, these matrix effects can cause ion suppression in the detection of someanalytes that are present at low levels of concentration, as seen in this figure.Several papers have reported this effect [30–32] and different alternatives toovercome these problems, such as the inclusion of a size-exclusion step [33] orsequential SPE [28], have been applied for the determination of pesticides in



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environmental matrices. The advantage of the fractionation of the extract is theseparation of the pollutant mix into different group classes, depending on theirchemical properties, which reduces matrix interferences and makes detectioneasier and more reliable.

Solid-phase microextraction (SPME) is also a useful alternative to conven-tional sample cleanup with LLE or SPE. SPME is based on the enrichment ofanalytes by a partitioning process between a polymeric phase coated on afused-silica fiber and its surrounding aqueous solution. SPME combines sam-ple preparation in terms of extraction from a matrix of interfering compoundswith an enrichment process in a single step. A method for the determinationof metazachlor in wastewater samples is described in the literature [34]. In thisstudy, SPME was shown to be a suitable and simple sample preparation methodfor the determination of metazachlor in wastewater by GC–AED.

2.3Methods of Analysis

A wide range of analytical techniques have been developed in order to identifythe organic contaminants often present at trace levels in complex environ-mental samples such as wastewaters. These techniques mainly use gas chro-matography (GC) and liquid chromatography (LC).

Most of the continuously monitored water contaminants are determined via gas chromatography–mass spectrometry (GC–MS). However, an adequateseparation of polar compounds via GC typically requires derivatization of thepolar moieties (e.g., BSTFA derivatives). In addition to this, as the analytegroups show different properties concerning the number and kind of func-tional groups, it is quite difficult to develop a universal derivatization proceduresuitable for all the target analytes. Furthermore, the presence in wastewater ofmany other organic compounds requires the use of labeled standards, whichcan make application of this method unfeasible [35].

2.3.1Gas 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 inmore detail later in this section. The flame ionization detector (FID) is anothernonselective detector that identifies compounds containing carbon but doesnot give specific information on chemical structure (but is often used for quan-tification because of the linear response and sensitivity). Other detectors arespecific and only detect certain species or groups of pesticides. They includeelectron capture, nitrogen–phosphorus, thermionic specific, and flame photo-metric detectors. The electron capture detector (ECD) is very sensitive to chlo-rinated organic pesticides, such as the organochlorine compounds (OCs, DDT,dieldrin, etc.). It has a long history of use in many environmental methods,

especially those advocated by the U.S. Environmental Protection Agency. Thenext most common is the nitrogen–phosphorus detector (called the NP de-tector), which is effective for many herbicides and insecticides because theycontain either nitrogen (triazines and acetanilides) or phosphorus (OP in-secticides). This is an inexpensive detector that has been widely used in theanalysis of pesticides in water. The flame photometric detector (FPD) works onchemiluminescence and detects pesticides that contain sulfur, phosphorus, andsome 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 com-pound identification in complex matrices, where GC–MS may have serious in-terferences. Wastewater is such an example. A specific detector for measuringelemental compositions as a percentage is the atomic emission detector (AED),which can detect all elements, except helium, separately due to its multichannelability and selectivity, making it more sensitive than the more commonly useddetectors cited above. This type of detector has been used for the detection ofpesticides in wastewater samples [34].

Finally, gas chromatography coupled to mass spectrometry (GC–MS) is themost universal technique for the analysis of pesticides in water samples [38].The high sensitivity and selectivity of modern GC–MS instruments enables lowlimits of detection depending on the matrix and in particular on the chemicalstructure of the pesticide. With most instruments, full-scan spectra can beevaluated at the low nanogram level, which means 1 or 10 pg analyte injectedinto the GC–MS system with the sample. Spectral averaging and backgroundsubtraction facilities provided by the data system are generally used to removecontributions from the matrix background or partially resolved contaminants.However, with very weak spectra, these data processing procedures may lead tocorrected mass spectra of dubious validity.Changing from full spectral scanningto selected ion monitoring using the reduced number of mass channels leadsto considerably improved detection limits for the specified target compoundions. The different types of ionization include electron impact (EI) and chem-ical ionization (CI). One advantage of negative chemical ionization (NCI) is inthe analysis of organochlorine insecticides in complex matrices, because thebackground 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 con-ditions.As an example, Fig. 3 shows the analysis of a real wastewater sample inthe NCI mode where three compounds were identified by the correspondingmass spectra.

Another approach in GC is that of using more power in the separation bydoing GC¥GC. In this approach, a second column is used with a different typeof stationary phase than the primary stationary phase, and fast chromatogra-phy using TOF-MS as the detector is carried out [39]. This technique uses onlyTOF-MS as the detector since it has the most sensitivity for fast-eluting peaks.The method has been applied to complicated matrix analysis.



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One of the known disadvantages of the use of GC is the need for previousderivatization of some of the most polar pesticides before analysis can be car-ried out [40]. These derivatization steps might produce low-efficiency resultsin complex wastewater matrices, which make the analysis rather difficult andcumbersome. However, the reproducibility in retention times when using GCtechniques is so precise, that specific identifications of pesticides can be madeeven in complex environmental samples.

Quantification is usually achieved by a standard addition method, use oflabeled internal standards, and/or external calibration curves. In order to allowfor matrix interferences the most reliable method for a correct quantitation ofthe analytes is the isotope dilution method, which takes into account intrinsicmatrix 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 usuallyachieved by participation in interlaboratory exercises and/or the analysis ofcertified reference materials [21].

2.3.2Liquid Chromatography–Mass Spectrometry (LC–MS)

Due to the high amount of interferences present in wastewater samples, UV de-tection is not possible for the identification of pesticides at low levels of con-centration (see Fig. 4).As can be noted in this figure, the humic and fulvic acidpeak at the beginning of the chromatogram masks the identification of themost polar pesticides and complicates the identification and quantitation ofthe analytes. Furthermore, a higher level of confidence (molecular or fragmentstructural information) is necessary for the correct identification of analytesin such complex matrices. In this sense, liquid chromatography coupled tomass spectrometric detection is the best choice for the analysis of pesticidesin wastewater samples. Since polar, nonvolatile, thermally unstable or high-molecular-weight compounds are unsuitable for gas chromatography–massspectrometry (GC–MS) analysis, the use of LC–MS has become a robust androutinely applicable tool in environmental laboratories [41, 42]. Non-GC-amenable compounds include 15–20% of the present-day pesticides, e.g.,phenylureas and carbamates, the phenoxyalkanoic acids, and a large majorityof all pesticide transformation products [43]. The performance of LC–MS inthe analysis of polar and thermally labile pesticides that are not amenable toGC–MS has been well demonstrated in several studies [44, 45]. In the last fewyears, interfaces based on atmospheric pressure ionization (API) have resultedin an increase in the number of applications in environmental determinations.In this respect, atmospheric pressure chemical ionization (APCI) and electro-spray ionization (ESI) have recently become the most universal techniques forenvironmental analysis due to their high sensitivity, the possibility of detectinga broad range of analytes, and the useful structural information obtained viafragmentation similar to collision-induced dissociation (CID) [46]. Compared



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

with older mass spectrometric detection techniques such as TSP and PB, APItechniques offer both structural confirmation and high sensitivity for targetcompounds in environmental samples. One of the great advantages of the ESIinterface is its high sensitivity for ionic pesticides such as many herbicidemetabolites containing a sulfonic or a carboxylic group in the chemical struc-ture [47].

The advent of high-performance liquid chromatography–mass spectrome-try (HPLC–MS) using quadrupole instruments has made analysis of polar pesticides in water a common procedure [45, 48]. Many classes of pesticides areeasily analyzed by LC–MS and a more challenging task is to identify the degra-dation products of pesticides. During the past 5 years many papers have beenpublished on the analysis of pesticides and their degradation products byHPLC–quadrupole MS [49]; however, there are several shortcomings yet to beovercome. For example, often polar pesticides give only a protonated or de-protonated molecule or a weak fragment ion, especially when the interface isESI. The fragmentor or cone voltage is used to enhance CID in the source andtransport regions of the electrospray source,and this fragmentation voltage may

vary substantially among different analytes and sources, which makes frag-mentation difficult to predict in an analysis of unknown compounds. Second,there are no universal libraries available for pesticide analysis by HPLC–MS, asin electron impact GC–MS; this problem makes identification of unknown pesticides or their degradates nearly impossible by simple HPLC–quadrupoleMS analysis.

These shortcomings may be overcome partially by the application of time-of-flight mass spectrometry (TOF-MS) and liquid chromatography–quadrupoleion-trap tandem mass spectrometry (LC–QIT-MS/MS) [50–52]. The LC–QIT-MS/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 theresult 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, theresult is MS/MS. This process may be repeated multiple times, which results in MSn. In contrast, triple quadrupole MS/MS does the isolation and frag-mentation in space, which means that the fragmentation is continuous in time,but the selected ion travels through the flight tube of the mass spectrometer tothe collision chamber where fragmentation occurs, and then on to the thirdquadrupole for the mass spectrum.

Two advantages of the ion trap are that it gives excellent sensitivity whiletrapping ions in full-scan mode, which then may be selected and fragmentedto 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 thesensitivity is too low to record ions in unknown samples. The ability to do mul-tiple isolation and fragmentation allows one to build a library of spectra usingstandard compounds, which give both characteristic fragmentations and di-agnostic ions that can then be used to identify unknown pesticides or theirdegradates. TOF-MS is also useful for identification of synthesized standards toverify the analysis of QIT-MS/MS when no commercial standards are availableand new standards are synthesized, as well as the identification of degradates inactual groundwater samples [52].

3Toxicity Biological Assays

3.1Bioassays Applied to Evaluate the Toxicity of Pesticides

The toxic effects of pesticides can be diverse and depend on the sensitivity oforganisms to these toxicants, and the pesticide concentration or bioavailability.Typically, the short- and long-term effects of pesticides have been evaluatedthrough acute or chronic toxicity bioassays, respectively, using lethality end-points and sublethal endpoints (e.g., growth and reproduction), particularlythese last in chronic bioassays.


3.1.1Acute Toxicity Bioassays

Most commonly, bioassays for the evaluation of the acute toxic effects of pesti-cides are based on single aquatic species selected to be representative of a rangeof taxonomic and functional groups, i.e., bacteria, algae, invertebrates or fish [53, 54]. Generally, toxicity evaluation using a single species is the alternative ofchoice rather than the use of multiple species, because extrapolation of effectsto 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 andmay be dependent on various factors. Some of these include simplicity, low cost,or modest material and equipment demand. However, a higher sensitivity thanother 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 concen-tration (EC50), which is the toxicity endpoint for the organisms (bacteria,crustaceans, algae, and fish) selected for the toxicity bioassays. These toxicitybioassays are usually classified according to the test species involved.

Fish assays have been extensively used for laboratory studies. Among com-monly used species are Pimphales promelas or Oncorhynchus mykiss. Thesespecies are relatively sensitive and respond to a variety of water constituentsand contaminants including pesticides. P. promelas is a widely distributedspecies in the aquatic environment, and its use for whole effluent toxicity(WET) procedures is also well established [55, 56]. Reported lethal concentra-tions for pesticides such as chlorotalonil or chlorpyriphos (EC50=22.6 and381 mg/l, respectively) showed these compounds as “harmful to aquatic organ-isms” and “not harmful”, respectively, according to toxicity categories [56, 57].Generally, in addition to the relative sensitivity (Table 1), the use of these bio-assays presents some disadvantages such as standardization problems, timeconsumption 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 cen-tral position. Therefore, their responses to natural and anthropogenic stressesare 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 usedfor many years as a standard aquatic test species and formally endorsed by themajor 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 aquaticecosystem because they are primary producers of the food chain. Several speciesof green algae are used in toxicity studies of pesticides, especially herbicidessuch as Chlorella vulgaris, Chlorella pyrenoidosa, or the standard test microalga



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

Pesticides EC50 (mg l–1) References

Bacteria Algae Crustaceans Fish

Fenamiphos 35.1a 0.005 [15]Benomil 0.05b [71]Pentachlorlophenol 0.55 [77]Paraquat 14,8001 <1.0 2, b [77]1, [68]2

Deltametrin 0.005g [58]Dichlorvos 2.10–4,e [63]Chlorpyriphos 3.21,e 3812, h [68]1, [57]2

Metalaxil 21.1b [71]Carbendazim 34.6b [71]Procymidone 0.74b [71]Zineb 0.52b [71]Chlorothalonil 0.0071, c 0.032, f 22.6 3, h [14]1, 2, [56]3

b-Cypermethrinm 0.02g [59]Dichlofluanid 0.081, a 0.132,c 1.03,f [14]1, 2, 3

Permethrin 0.2g [60]Carbofuran 31.21, a 0.022,f [15]1, 2

Diuron 0.041, c 8.62,f [14]1, 2

Isoproturon <1.0d [68]Atrazine <1.0d [68]Formetanate 7.41, a 0.07 2, f [15]1, 2

Pirimiphos-methyl 4.10–4, f [14]Malathion 1.8·10–3, f [63]Cyromazine 10.7 f [68]

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

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

Bioassays using bacteria as indicators constitute at present one of the mostwidely applied screening techniques for identifying the toxicity of substancesand commercial products [72]. The main reason is that, unlike toxicity bioas-says using more complex organisms such as fish, bacterial bioassays are muchquicker and cheaper.A great number of bioassays based on very different meth-ods and microorganisms have been described, and include studies of the effectsof toxicants on different parameters such as growth inhibition and enzymaticactivity [73, 74]. Among the bacteria employed, such as Pseudomonas putidaor Escherichia coli as indicators, V. fischeri is the most common standardizedbacteria specie used in toxicity bioassays [14, 72, 75]. The advantages are its sen-sitivity, reproducibility, and it is a rapid and simple test (Table 1). For toxicityevaluation of pesticides, there are reported data showing the sensitivity and theutility of this test [14, 75–77].

The different sensitivity of the species indicates that a single bioassay doesnot satisfy the correct evaluation of the wastewater. Thus, normally, variousspecies are used because toxic substances may produce a specific response inone species but not in another. Therefore, there is practically generalized con-sensus 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, andfish to assess the toxicity of chemicals such as pesticides in wastewater, and itwould 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, forpractical and ethical reasons, the use of fish is not frequently included in thesestudies.

3.1.2Chronic Toxicity Bioassays

Episodic pollution events can adequately be addressed by acute toxicity bioas-says, however these are not sufficient to investigate the water quality for delayedtoxicity effects of chemicals present. Chronic effects of pesticides can includecarcinogenicity, teratogenicity, mutagenicity, neurotoxicity, and reproductiveeffects (endocrine disruption).

Most insecticides, especially the organophosphate group, cause neurotoxi-city as their major mode of action. Assessment of the neurotoxicity includesneurochemical 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 endpointssuch as swimming speed [79]. Studies done in rats show the neurotoxic actionof insecticides such as dimethoate, methyl parathion, dichlorvos, ethyl parathionor propoxur after a prolonged exposure [80, 81].


Chemical carcinogenicity has been the target of a large list of scientific pub-lications, because it is one of the toxicological endpoints that poses the high-est concern. The standard bioassays in rodents used to assess the carcinogenicpotential of chemicals are extremely long and costly and require the sacrificeof 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 ortranslocation, mitotic recombination or base substitution [82]. Therefore,regular practices to evaluate the possible genetic lesions recommend the useof a battery of bioassays including a bacterial test for gene mutation, either anin vitro test for chromosomal aberrations or a mammalian cell mutagenesistest, and a general test for DNA damage [84, 85].A great number of studies onthe mutagenic activity of pesticides have been published. Examples of theseshow that the chloroacetanilides, classified as herbicides, have a consistent positive induction for gene mutations [86].

More recently, toxicity studies have shown the importance of noncancerendpoints in chronic toxicity assessment, with increasing emphasis on end-points such as endocrine disruption. The endocrine system as a target of pes-ticide toxicity can manifest reproductive consequences, particularly in termsof steroid hormone function, resulting in the manifestation of demasculiniza-tion in fish. The gonad histology and serum vitellogenin (VTG) protein levelshave been widely used as endpoints for screening and testing of potential endocrine-active compounds and are currently subject to validation by theOECD and associated scientific groups [87–89]. Some reports have demon-strated that the presence of organochlorines, such as dieldrin, heptachlor oraldrin, appears to be closely linked to the induction of VTG synthesis [90, 91].However, bioassays based on yeast strains are very promising among the testsystems available because of their physiological simplicity, easy handling,and low costs [92, 93]. In general, they rely on yeast constructs expressing anestrogen receptor which, upon binding of suitable substrates, acts as a tran-scriptional enhancer for an estrogen-responsive DNA-element-controlled re-porter gene, in most cases bacterial b-galactosidase. The activity of this enzymecan be determined photometrically by using a chromogenic substrate and thusmay serve as a measure of the estrogenic potency of the samples under in-vestigation. Several active components such herbicides and insecticides (e.g.,endosulfan, dieldrin or toxaphene) have been reported to possess estrogenicactivity [94, 95].

3.2Toxicity Studies of Wastewater Containing Pesticides

While reported data on the acute and chronic toxicity of many pesticides isplentiful, few studies have been published on toxicity bioassays applied towastewaters containing pesticides. The application of toxicity bioassays to thequality control of wastewaters offers several advantages in addition to being a


biological measure able to detect toxic effects. Among these, sensitivity, easyhandling, speed, simplicity, and low costs are the features of choice for routinepurposes. In the last few years, interest in toxicity bioassays for assessing waste-water has been increasing and recent publications are focused on this approach[96–98]. Some of these studies proved that there is no correlation betweenchemical and ecotoxicological parameters. Control based on global chemicalparameters such as biochemical oxygen demand (BOD), chemical oxygen de-mand (COD) or total organic carbon (TOC) may be insufficient, even if thetreated effluents meet the threshold concentration levels for discharge. This caseis illustrated in Fig. 5, which shows a monitoring study performed on influentand effluent wastewaters. Samples corresponding to toxic effluents showed permissible TOC levels [96].

Wastewater from agricultural areas that arrives at wastewater treatmentplants (WWTPs) is highly variable in nature. Intermittent or accidental episodesof toxic substances can have a damaging effect on the receiving waters, when the


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

b

a

influent has not been properly treated. Consequently, rapid methods of waste-water toxicity assessment represent a very useful tool, acting as an early warn-ing system. The capability of detecting toxic responses in a short time allowsquick 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 allowsdetecion of unwanted toxicity, potential problems in the treatment station,and contamination peaks in effluent toxicity before discharging it to receivingwaters [76].

The sensitivity of test species is a decisive feature in the choice of bioassaysto evaluate the toxicity. Despite the diversity of test species available, in manyregulatory schemes the invertebrate species recommended for acute and chronictesting 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 theEnvironmental Protection Agency recommend that acute data should be col-lected with Daphnia species (D. pulex and D. magna) [101]. Presumably, the focus on D. magna results from its high sensitivity to environmental contam-inants relative to other species, mainly invertebrate species. The sensitivity ofD. magna to pesticides has been demonstrated in recent publications showingits capability of detecting toxic responses at concentration levels as low asnanograms per liter [14]. This means that toxicants at environmentally realis-tic 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 pre-concentration of the wastewater have been proved to be a useful strategy forscreening and monitoring in the initial assessment of water pollution by pesti-cides [102, 103]. Even if bioassays are able to detect toxicity in nonconcentratedsamples, this strategy is a useful approach in order to obtain the toxicity end-point (e.g., effective concentration EC50) from a full concentration–response relationship. This combined methodology was applied in a screening studyfrom an agricultural area where methyl parathion, lambda-cyhalothrin, and endosulfan are the most commonly used pesticide chemicals. Acute toxicitywas detected in surface water from agricultural areas using standardized bio-assays with the algae S. capricornutum and crustacean D. magna [104].

Whole effluent toxicity (WET) monitoring offers several advantages becausethis toxicity evaluation has to account for the presence of unknown toxicants,the interactions among multiple toxicants, and the alterations in toxicantbioavailability caused by the effluent matrix. When evaluating the toxicity of the complex samples, the detection and identification of regulated or specificchemicals is a key need for controlling effluent quality. Thus, the identifica-tion of toxic compounds in complex samples has been the objective of re-ported studies using toxicity-based procedures. Combined protocols involvingchemical analysis and toxicity evaluation became known collectively as toxic-ity identification evaluation (TIE), and nowadays they are techniques well


established and developed by the USEPA [104]. TIE methods have been foundto be effective tools for characterizing and identifying toxicants in samples ofeffluents, sediments, ambient waters, and other complex mixtures [105, 106].Regarding the identification of pesticides in wastewaters or ambient waters, fewstudies have been published. The use of TIE methods allows the detection oftoxic 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 influentand effluent wastewaters from wastewater treatment plants that receivedwastewaters from agricultural areas. This approach was used to detect a pos-sible cause–effect relationship between the plant discharge and the receivingwater quality [107]. The results of this study showed the detection of lindaneand pp¢-DDE in fish, and chemical investigations revealed ammonia and mi-cropollutants as factors of WWTP effluent impact on receiving waters [108].

As was mentioned above, the interaction among multiple chemicals is oneof the main reasons for the wastewater toxicity. The application of TIE meth-ods using acute toxicity (D. magna) guided chemical analysis was applied forwater quality evaluation of agricultural land runoff and 11 pesticides widelyapplied 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 be-cause their effective toxic concentrations were from 5 to 10 mg/l. Therefore, itwas impossible for the authors to conclude in this study that the observeddaphnia toxicity resulted from a single highly toxic substance. The toxicity wasattributed 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 differ-ent studies [14, 15]. It is especially relevant to consider the combined effect of pollutants because several pesticides and other contaminants can occur inambient waters from agricultural areas [110, 111]. The global effect can have agreater negative impact than the single pollutants.

For a predictive assessment of the aquatic toxicity of pesticide mixtures, twoconcepts, concentration addition and independent action, are used. Concen-tration addition is generally regarded as a reasonable expectation for the jointtoxicity of acting substances [112]. Following this model, the concentration ofeach toxicant is expressed as a fraction of its EC50 (toxic unit, TU). In thismodel, 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 thanadditive and when it is less than additive, the substances act antagonistically.Synergism is a common interactive effect among pesticide mixtures. Experi-ments on pesticide mixtures showed a synergistic effect for 60% of the studiedcases [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.


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

Binary mixtures Bioassays

V. fischeri D. magna S. capricornutum(15 min) (48 h) (72 h)

T.U.M1 S T.U.i2 T.U.M1 S T.U.i2 T.U.M1 S T.U.i2

Irgarol 1501– 26 5.8 550 183 0.36 0.01Diruon Synergistic + Synergistic + Synergistic ++

Irgarol 1501– 7.4 333 550 581 5 5Sea nine 211 Antagonistic ++ Additive Additive

Irgarol 1501– 5.4 0.87 1100 414 0.271 2Chlorothalonil Synergistic + Synergistic Antagonistic +

Irgarol 1501– 9.4 15.6 132 160 0.152 0.046Dichlofluanid Additive Additive Synergistic +

Irgarol 1501– 333 25.6 330 151 10 1.08TCMTB Synergistic ++ Synergistic + Synergistic ++

T.U.M1 , experimental toxicity.S T.U.i2, theoretical toxicity.+=factor≥3.++=factor≥10.

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

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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 Introduction to Fragrance Materials . . . . . . . . . . . . . . . . . . . . . . . . 811.1 Use and Disposal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 811.2 Chemical Structures . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 811.3 Physical-Chemical Properties and Biodegradability . . . . . . . . . . . . . . . . 84

2 Analytical Chemistry of Fragrance Materials . . . . . . . . . . . . . . . . . . . 862.1 Laboratory Quality Control . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 862.2 Standards . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 862.3 Aqueous Matrices . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 872.4 Solid Matrices . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 912.5 Analysis . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 91

3 Sampling Wastewater Treatment Plants for Fragrance Materials . . . . . . . . . 923.1 Selection of Wastewater Treatment Plants . . . . . . . . . . . . . . . . . . . . . 923.2 Wastewater Treatment Plant Sampling . . . . . . . . . . . . . . . . . . . . . . . 94

4 Mechanisms of Fragrance Material Removal During Wastewater Treatment . . 954.1 Biodegradation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 954.2 Sorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 974.3 Volatilization . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 97

5 Measurement of Fragrance Materials in Wastewater Treatment . . . . . . . . . 985.1 Concentrations in Treatment Plants . . . . . . . . . . . . . . . . . . . . . . . . 985.2 Removal During Treatment . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 109

6 Predicting Fragrance Material Removal During Wastewater Treatment . . . . . 1136.1 Framework for Aquatic Risk Assessment . . . . . . . . . . . . . . . . . . . . . . 1136.2 Simple Treat Model . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 115

7 Conclusions . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 115

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 117


Abstract In recent years, there has been significant interest in understanding the input offragrance materials (FMs) to aquatic ecosystems, and this has driven a substantial amountof research on the removal of FMs during wastewater treatment. Because FMs are semi-volatile and have a wide range of physical-chemical properties and biodegradabilities,understanding their removal during the treatment process is complex. The mechanisms ofFM removal from wastewater include biodegradation, sorption, and/or volatilization.A widearray of analytical methods have been developed to measure FMs in wastewater influent,primary effluent, final effluent, and solids. Wastewater studies have been conducted in theU.S. and Europe. Finally, the efficient removal of FMs during wastewater treatment is not onlydependent on the biodegradability and physical-chemical properties of the FM, but is alsohighly dependent on plant operation and design.

Keywords Fragrance materials · Wastewater treatment · Polycyclic musks · Nitromusks

AbbreviationsADBI 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 extractionATII Traseolide (5-acetyl-1,1,2,6-tetramethyl-3-isopropylindene)BOD Biochemical oxygen demandCAS Chemical Abstracts ServiceDPMI Cashmeran (6,7,-dihydro-1,1,2,3,3-pentamethyl-4(5H)-indanone)ECD Electron-capture detectorFM Fragrance materialGC Gas chromatographyGC–MS Gas chromatography–mass spectrometryGC–MS/MS Gas chromatography–tandem mass spectrometryGPC Gel-permeation chromatographyHHCB Galaxolide (1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-gamma-2-

benzopyran)HRT Hydraulic retention timeKd Sorption coefficient to activated sludgeKow Octanol–water partition coefficientMA Musk ambrette (1-tert-butyl-2,4-dimethyl-6-methoxy-3,5-dinitrobenzene)MDL Method detection limitMK 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 NitromuskNPD Nitrogen–phosphorus detectorOTNE 1-(1,2,3,4,5,6,7,8-Octahydro-2,3,8,8-tetramethyl-2-naphthalenyl)ethanonePCM Polycyclic muskSOC Semivolatile organic compoundSPE Solid-phase extractionSPME Solid-phase microextractionSRT Solids retention timeTSS Total suspended solids

80 S. L. Simonich

1Introduction to Fragrance Materials

1.1Use and Disposal

Over 2,000 distinct chemicals are currently available globally for formulationinto fragrances [1]. Although the consumer is most aware of the use of thesechemicals in fine fragrances, by far the largest volume of these chemicals is usedin laundry detergents, fabric softeners, household cleaning products, and airfresheners [1]. Of these consumer products, fabric softeners and laundry de-tergents represent the largest volume uses of fragrances and the largest releaseto the environment through down-the-drain disposal by consumers followingproduct use [1].

Fragrance materials (FMs) are added to consumer products to mask mal-odors and to deliver consumer-preferred odors [2]. Although these chemicalsare used in low concentrations in consumer products, the volume of laundrydetergents and fabric softeners sold throughout the globe can result in sig-nificant volumes of FMs being released into the environment. Based on a1995–1996 survey, approximately 90% of these compounds are used globally atless than 10 metric tons per year [1], with less than 1% being used in volumesapproaching 4,000 metric tons per year [2].

Because the majority of the FM volume enters the environment throughdown-the-drain disposal of consumer products, it is important to understandthe removal and fate of these chemicals during municipal wastewater treatment.These semivolatile organic compounds (SOCs) may undergo a complex combi-nation of biodegradation, sorption, and/or volatilization during wastewatertreatment. In addition, few SOCs have been studied in wastewater treatment be-cause few of the conventional SOCs (such as pesticides and products of incom-plete combustion) enter the environment through down-the-drain disposal andwastewater treatment. The objective of this chapter is to review the state of thescience in understanding the removal of FMs during municipal wastewatertreatment. Others have reviewed the general environmental fate of FMs, in par-ticular the polycyclic musks (PCMs) and the nitromusks (NMs) [3–6].

1.2Chemical Structures

The chemical structures of the majority of FMs that have been studied inwastewater treatment are given in Figs. 1–3. Figure 1 shows a variety of FMstructures that include alcohols, aldehydes, and ketones, including: benzyl acetate (phenylmethyl ester acetic acid), methyl salicylate (2-hydroxy-methylester 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

Fragrance Materials in Wastewater Treatment 81

82 S. L. Simonich

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-methyl-4-(2,6,6-trimethyl-2-cyclohexen-1-yl)-3-buten-2-one], p-t-bucinal [4-(1,1-dimethylethyl)-a-methyl-benzenepropanal], hexylcinnamaldehyde [2-(phenyl-methylene)-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)ethan-one], and acetyl cedrene [3R-(3a,3ab,7b,8aa))-1-(2,3,4,7,8,8a-hexahydro-3,6,8,8-tetramethyl-1H-3a,7-methanoazulen-5-yl)ethan-1-one]. Figure 2 showsthe structures of the FMs categorized as nitromusks (nitroaromatic compounds),including: musk ketone or MK (3,5-dinitro-2,6-dimethyl-4-tert-butylacetophe-none), musk xylene or MX [1-(1,1-dimethylethyl)-3,5-dimethyl-2,4,6-trinitro-benzene], musk ambrette or MA (1-tert-butyl-2,4-dimethyl-6-methoxy-3,5-dinitrobenzene), musk tibetene or MT (1-tert-butyl-3,4,5-trimethyl-2,6-dinitro-benzene), and musk moskene or MM (1,1,3,3,5-pentamethyl-4,6-dinitroin-dane). Finally, Fig. 3 shows the structures of FMs categorized as polycyclic


Fig. 3 Polycyclic musks studied in wastewater treatment

Fig. 2 Nitromusks studied in wastewater treatment

musks (PCMs), including AHTN or Tonalid (7-acetyl-1,1,3,4,4,6,-hexamethyl-1,2,3,4-tetrahydronaphthalene), HHCB or Galaxolide (1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-gamma-2-benzopyran),ADBI or Celestolide(4-acetyl-1,1-dimethyl-6-tert-butylindene), AHDI or Phantolide (6-acetyl-1,1,2,3,3,5-hexamethyldihydroindene), ATII or Traseolide (5-acetyl-1,1,2,6-tetramethyl-3-isopropylindene), and DPMI or Cashmeran (6,7,-dihydro-1,1,2,3,3-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 re-search groups because they have been detected in the aquatic environment (seefor example [7–10]), indicating that they escape wastewater treatment to somedegree. 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 rangeof physical-chemical properties and biodegradability (see below).

It is clear from Figs. 1–3 that FMs have an interesting and wide range ofchemical structures. This results in a wide array of perceived odors, includingmusk, wood, fruit, and flower-like odors. The FMs in Figs. 1–3 will be the focusof this chapter.

1.3Physical-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 coeffi-cient for activated sludge, water solubility, vapor pressure, and Henry’s law con-stant), and biodegradability of selected FMs [2, 11]. It is clear that the widerange of FM chemical structures results in a wide range of physical-chemicalproperties (some properties ranging over six orders of magnitude). From thevapor pressures given in Table 1, it is also clear that most FMs can be classifiedas SOCs (having vapor pressures less than 1 Pa) [2], and have the potential topartition into a variety of environmental compartments once released into theenvironment.

The properties listed in Table 1 are of interest because they govern FM re-moval during wastewater treatment and their fate in the environment. The logoctanol–water partition coefficients (log Kow) for the selected FMs range from2.1 to 5.9. The more hydrophobic FMs (with high octanol–water partition co-efficients) are desirable in consumer products because they tend to be moresubstantive on fabrics and provide a residual fabric odor. Unfortunately, this hy-drophobicity may also result in bioaccumulation in organisms in the environ-ment. The sorption coefficient for activated sludge (Kd) is an important prop-erty to consider for removal of FMs due to sorption onto solids duringwastewater treatment. The Kd value ranges from 132 to 15,400 L kg–1 for the se-lected FMs. The water solubility and vapor pressure of the FMs listed in Table 1also vary greatly: from 0.49 to 1,687 mg L–1 for water solubility and from0.00003 to 21.9 Pa for vapor pressure. Finally, the Henry’s law constant (a mea-

84 S. L. Simonich


Tabl

e1

Sum

mar

y of

sele

cted

FM

phy

sica

l-ch

emic

al p

rope

rtie

s an

d bi

odeg

rada

bilit

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leva

nt to

was

tew

ater

trea

tmen

t [2,

11]

Frag

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ater

ial

CA

S nu

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Wlo

gK

owK

dW

ater

Va

por

Hen

ry’s

law

Bi

odeg

rada

bilit

y(g

mol

–1)

(Lkg

–1)

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nsta

nt(m

gL–1

)(P

a)(P

am

3m

ol–1

)

Ben

zyl a

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te14

0-11

-415

0.2

2.1

132

1,26

5.0

21.9

2.04

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dyM

ethy

l sal

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ate

119-

36-8

152.

22.

624

71,

687.

00.

750

0.06

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eady

Met

hyl d

ihyd

roja

smon

ate

2485

1-98

-722

6.3

3.0

408

91.7

20.

0054

90.

135

Rea

dyTe

rpin

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98-5

5-5

154.

33.

359

533

5.7

4.09

0.93

9R

eady

Ben

zyl s

alic

ylat

e11

8-58

-122

8.3

4.3

2,08

124

.59

0.00

0449

0.00

416

Inhe

rent

Isob

orny

l ace

tate

125-

12-2

196.

34.

32,

081

23.2

310

.084

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eady

g-M

ethy

l io

none

127-

51-5

192.

34.

63,

030

9.0

1.30

89.4

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rent

p-t-

Buci

nal

80-5

4-6

204.

34.

21,

836

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712

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k ke

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81-1

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able

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lene

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412

0.49

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able

Hex

ylci

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2.75

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00In

here

ntH

exyl

sal

icyl

ate

6259

-76-

322

2.3

5.5

9,35

56.

080.

0032

50.

118

Rea

dyO

TN

E54

464-

57-2

234.

45.

712

,020

2.68

0.20

331

.8N

ot b

iode

grad

able

Ace

tyl c

edre

ne32

388-

55-9

246.

45.

6-5.

912

,020

1.28

0.05

814

.7In

here

ntA

HT

N15

06-0

2-1

258.

45.

712

,020

(10,

040)

1.25

0.06

0812

.5N

ot b

iode

grad

able

HH

CB

1222

-05-

525

8.4

5.9

15,4

00 (1

2,78

0)1.

750.

0727

11.3

Not

bio

degr

adab

le

sure of air–water partitioning) for the selected FMs in Table 1 ranges from0.00416 to 89.4 Pa m3 mol–1, indicating that some FMs may undergo volatiliza-tion during wastewater treatment.

FMs also have a wide range of biodegradabilities (see Table 1). Some FMspass the OECD ready biodegradability test criteria, including the 10-day window[11]. Other FMs pass the OECD inherent biodegradability test or produce CO2in the OECD ready biodegradability test, but do not meet the 10-day window[11]. Still other FMs do not biodegrade in standard OECD biodegradation testsbut undergo biotransformation in more realistic tests [11–13].

2Analytical Chemistry of Fragrance Materials

In order to understand the removal of FMs during wastewater treatment, it isnecessary to measure these compounds throughout the wastewater treatmentprocess. 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 ofresearchers. Fortunately, the analytical techniques developed to measure tra-ditional SOCs, such as solvent extraction, extract concentration, and analysisby gas chromatography–mass spectrometry, in general also apply to FMs.

2.1Laboratory Quality Control

Because of the ubiquitous nature of FMs in consumer products, it is critical thatany analytical chemistry laboratory measuring these compounds takes extraprecautions to avoid laboratory contamination of samples. Several researchers[2, 11, 14–17] have pointed out that likely sources of FM contamination in themodern-day laboratory include the use of consumer products and fine fra-grances by laboratory workers, fragrances in soaps used to clean glassware andthe laboratory, and laboratory supplies such as gloves.

Before beginning the analysis of FMs at low concentrations, the laboratoryshould analyze several laboratory blank samples to assess the degree to whichthe laboratory is contaminated. With every set of samples analyzed, the labo-ratory should also analyze a laboratory and field blank sample. Laboratoryworkers should be advised to be aware of their personal use of fragrance-en-hanced consumer products and the potential for laboratory contamination.

2.2Standards

As interest in measuring FMs in the environment has increased, researchers haveused a variety of means to obtain FM standards for use in analytical chemistry.

86 S. L. Simonich

The sources of these standards include FM manufacturers, synthesis by re-searchers (especially in the case of NM metabolites), and specialty chemical cat-alogs.When possible, it is preferable to obtain these standards directly from theFM manufacturers in order to use the authentic material being discharged towastewater treatment. In all cases, the purchased, synthesized, or obtained stan-dards must be extensively analyzed to confirm the chemical structure and purity.

Also of importance is the appropriate use of surrogates and internal stan-dards for quantification of FMs in wastewater and environmental matrices.Ideally, stable isotope-labeled analogs (such as stable perdeuterated analogs) ofthe FMs are used for this purpose if gas chromatography–mass spectrometry(GC–MS) is the analysis technique. Simonich et al. [2, 11] synthesized eightperdeuterated FMs, including d3-benzyl acetate, d3-g-methyl ionone, d3-methyldihydrojasmonate, d3-OTNE, d4-acetyl cedrene, d6-musk xylene, d3-AHTN, andd7-musk ketone, through base-catalyzed exchange of protons with deuteriumand used these as surrogates to quantify 16 FMs in wastewater treatment ma-trices. 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 perdeuteratedpolycyclic aromatic hydrocarbons [14, 18, 19, 21, 22] as surrogates or internalstandards to measure FMs in wastewater treatment.

Finally, the amino metabolites of the NMs have been synthesized by re-searchers and used as standards. These synthesis methods include reductionof NMs with hydrogen in the presence of Pd/charcoal to form the aminometabolites [15, 16, 23] or reaction of NMs with hydrazine hydrate and Raneynickel [14, 23]. A metabolite of HHCB, HHCB-lactone, has also been synthe-sized and used as a standard [17].

2.3Aqueous Matrices

Table 2 lists a variety of analytical methods used to measure FMs in wastewatertreatment. Only three studies [2, 11, 24] have attempted to examine fragrancematerials throughout the wastewater treatment process, including the analysisof 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 insewage sludge and digested sewage sludge [15–18, 21]. Because wastewatertreatment processes consist of solid and aqueous phases, analytical methodshave been developed to measure FMs in both of these matrices.

Although some researchers have chosen to extract aqueous wastewater ma-trices with traditional methods, such as liquid–liquid extraction [5, 22], otherresearchers have used solid-phase extraction (SPE) to exhaustively extract FMsfrom these matrices [2, 8, 11, 14, 19, 20]. Simonich et al. [2, 11] used C18 Baker-bond Speedisks to extract 16 FMs, with a wide range of polarities, from 0.5-Linfluent and primary effluent and 1.0-L final effluent samples.Verbruggen et al.


88 S. L. Simonich

Tabl

e2

Sum

mar

y of

anal

ytic

al m

etho

ds u

sed

to m

easu

re F

Ms

in w

aste

wat

er tr

eatm

ent

Res

earc

hers

Frag

ranc

e m

ater

ials

WW

TP

mat

rix

Loca

tion

Ana

lyti

cal m

etho

d

Sim

onic

h 16

FM

s,in

clud

ing

In

fluen

t,pr

imar

y U

.S.,

Uni

ted

Kin

gdom

,–

Extr

acti

on o

f0.5

–1L

influ

ent,

et a

l.[2

,11]

thos

e in

Tab

le1,

efflu

ent,

acti

vate

d

and

The

Net

herl

ands

prim

ary

efflu

ent,

and

final

eff

luen

t wit

h C

18SP

EPC

Ms

(AH

TN

and

sl

udge

sol

ids,

– Ex

trac

tion

ofs

ludg

e by

acc

eler

ated

sol

vent

H

HC

B),a

nd N

Ms

and

final

eff

luen

tex

trac

tion

wit

h di

chlo

rom

etha

ne

(MX

and

MK

)–

Silic

a ge

l chr

omat

ogra

phy

– A

naly

sis

by s

tabl

e is

otop

e di

luti

on G

C–M

S us

ing

nine

per

deut

erat

ed F

Ms

– R

ecov

ery=

97–1

15%

;lim

it o

fqua

ntifi

cati

on=

0.5–

35ng

/L

Art

ola-

PC

Ms

(AH

TN

and

In

fluen

t,pr

imar

y T

he N

ethe

rlan

dsFr

eely

dis

solv

ed

Gar

ican

oH

HC

B) –

free

ly

sett

ler,

aera

tion

tank

,–

SPM

E w

ith

GC

–MS

et a

l.[2

4]di

ssol

ved

and

tota

l se

cond

ary

efflu

ent,

– Li

mit

ofq

uant

ifica

tion

=0.

1mg

/Lco

ncen

trat

ions

prim

ary

slud

ge,

Tota

l con

cent

rati

onan

d w

aste

slu

dge

– Li

quid

–liq

uid

extr

acti

on w

ith

cycl

ohex

ane

– Si

lica

gel c

hrom

atog

raph

y–

GC

–MS

– R

ecov

ery=

85–1

06%

;lim

it o

fqua

ntifi

cati

on=

0.1

mg/L

Kan

da

PCM

s an

d N

Ms

Influ

ent,

efflu

ent

Uni

ted

Kin

gdom

– Li

quid

–liq

uid

extr

acti

on w

ith

dich

loro

met

hane

et a

l.[2

2]–

GC

–MS

– R

ecov

ery=

69–9

5%;l

imit

ofd

etec

tion

=3.

7–8.

5ng

/L

Rim

kus

NM

s an

d m

ono-

In

fluen

t,ef

fluen

tG

erm

any

– Li

quid

–liq

uid

extr

acti

on w

ith

hexa

neet

al.

[23]

amin

o m

etab

olite

s–

Silic

a ge

l and

alu

min

a ch

rom

atog

raph

y–

GC

–MS/

MS,

GC

–MS,

GC

-EC

D,a

nd G

C-P

ND

– Li

mit

s of

quan

tific

atio

n=1

ng/L


Tabl

e2

(con

tinu

ed)

Res

earc

hers

Frag

ranc

e m

ater

ials

WW

TP

mat

rix

Loca

tion

Ana

lyti

cal m

etho

d

Ric

king

PC

Ms

and

NM

sW

aste

wat

er e

fflu

ent

Can

ada

and

Swed

en–

SPE

and

filtr

atio

n an

d ex

trac

tion

wit

het

al.

[19]

n-pe

ntan

e an

d di

chlo

rom

etha

ne

– Si

lica

gel c

hrom

atog

raph

y–

GC

–MS

– D

etec

tion

lim

it=

0.5

ng/L

Verb

rugg

en

PCM

s W

aste

wat

er e

fflu

ent

The

Net

herl

ands

– Bi

omim

etic

and

exh

aust

ive

extr

acti

onet

al.

[20]

(AH

TN

and

HH

CB)

– C

18Em

pore

dis

ks–

GC

–MS

– R

ecov

ery

>95

%;d

etec

tion

lim

it=

0.1

ng/L

Ose

mw

engi

e N

Ms,

PCM

s,an

d

Was

tew

ater

eff

luen

tU

.S.

– O

n-si

te 6

0-L

extr

acti

on w

ith

NEX

US

sorb

ent

et a

l.[1

4]ni

trom

usk

– Si

lica

gel a

nd g

el-p

erm

eati

on c

hrom

atog

raph

ym

etab

olite

s–

GC

–MS

– R

ecov

ery=

80–9

7%;m

etho

d de

tect

ion

limit

=0.

02–0

.3ng

/L

Buer

ge

PCM

s W

aste

wat

er e

fflu

ent

Swit

zerl

and

– M

acro

poro

us p

olys

tyre

ne–d

ivin

ylbe

nzen

e et

al.

[8]

(AH

TN

and

HH

CB)

adso

rben

t–

Silic

a ge

l chr

omat

ogra

phy

– G

C–M

S–

Rec

over

y=81

–141

%;L

OD

=10

ng/L

Ber

set

NM

s an

d am

ino

Se

wag

e sl

udge

Swit

zerl

and

– Ex

trac

tion

wit

h he

xane

by

agit

atio

net

al.

[16]

met

abol

ites

– G

el-p

erm

eati

on c

hrom

atog

raph

y an

d si

lica

gel

chro

mat

ogra

phy

– G

C–M

S/M

S,G

C–M

S (E

I,N

CI,

and

PCI)

,1 H

and

13C

NM

R–

Rec

over

y=51

–116

%;d

etec

tion

lim

it=

50ng

/L

90 S. L. Simonich

Tabl

e2

(con

tinu

ed)

Res

earc

hers

Frag

ranc

e m

ater

ials

WW

TP

mat

rix

Loca

tion

Ana

lyti

cal m

etho

d

Her

ren

NM

s,PC

Ms,

and

Se

wag

e sl

udge

Swit

zerl

and

– Ex

trac

tion

wit

h he

xane

by

agit

atio

net

al.

[15]

amin

o m

etab

olite

s–

Gel

-per

mea

tion

chr

omat

ogra

phy

and

silic

a ge

l ch

rom

atog

raph

y–

GC

–MS/

MS,

GC

–MS

(EI,

NC

I,an

d PC

I)–

Rec

over

y=50

–118

%;d

etec

tion

lim

it=

100

ng/L

Kup

per

PCM

s an

d Se

wag

e sl

udge

Swit

zerl

and

– Ex

trac

tion

wit

h he

xane

and

sti

rrin

get

al.

[17]

HH

CB-

lact

one

– G

C–M

S–

Rec

over

y=79

–108

%;L

OD

=15

–30

umg/

kg d

.m.;

LOQ

=45

–90

umg/

kgd.

m.

Stev

ens

NM

s an

d PC

Ms

Dig

este

d se

wag

e U

nite

d K

ingd

om–

Soxh

let e

xtra

ctio

n w

ith

dich

loro

met

hane

et a

l.[2

1]sl

udge

– Si

lica

gel a

nd g

el-p

erm

eati

on c

hrom

atog

raph

y–

GC

–MS

Dif

ranc

esco

22

FMs;

incl

udin

g D

iges

ted

sew

age

U.S

.–

Acc

eler

ated

sol

vent

ext

ract

ion

wit

h et

al.

[18]

PCM

s an

d N

Ms

slud

gedi

chlo

rom

etha

ne–

Silic

a ge

l chr

omat

ogra

phy

– G

C–M

S

[20] used C18 3 M Empore Disks to exhaustively and biomimetically extractPCMs (AHTN and HHCB) from 10-L effluent samples. Osemwengie et al. [14]used 6 g of Nexus sorbent [polystyrene cross-linked with 50% divinylbenzeneand 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-BeadsSM-2 (a macroporous polystyrene–divinylbenzene adsorbent) to extract HHCBand AHTN from 200-mL effluent samples. Finally, Artola-Garicano et al. [24]used 1-cm lengths of a 100-mm polydimethylsiloxane solid-phase microextrac-tion (SPME) fiber to measure free concentrations of AHTN and HHCB through-out wastewater treatment.

Many of the aqueous methods listed in Table 2 utilize adsorption chroma-tography, such as silica or alumina chromatography, to purify extracts prior toanalysis.

2.4Solid Matrices

Researchers have chosen to extract sewage sludge for FMs in its wet form usingliquid–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 sol-vent [21] or at elevated temperature and pressure using accelerated solventextraction [2, 18]. Herren and Berset [15] and Berset et al. [16] extracted 1 L ofwet sewage sludge for NMs and their metabolites with 600 mL hexane for 2 hwith vigorous agitation. Artola-Garicano et al. [24] measured the free concen-tration of AHTN and HHCB in 10 mL wet sludge by negligible depletion SPMEand the total concentration by extracting with 6 mL cyclohexane during 2 h ofshaking. 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 solventextraction (at 60 °C and 2,000 PSI) with dichloromethane to extract a widerange of FMs from centrifuged activated sludge solids and digested and de-watered sludges. In the methods used to extract centrifuged sludges, Na2SO4was used to remove water from the sample prior to solvent extraction.

Many of the sludge methods listed in Table 2 utilize gel-permeation chroma-tography (GPC) to remove high molecular weight interferences and/or ad-sorption chromatography, such as silica or alumina chromatography, to purify extracts prior to analysis.

2.5Analysis

Because FMs are semivolatile, they are amenable to analysis by gas chroma-tography (GC) and gas chromatography–mass spectrometry (GC–MS) withoutderivitization. Table 2 shows that all of the analytical methods developed tomeasure FMs in wastewater treatment to date utilize GC or GC–MS.


In general, FMs can be chromatographically resolved using a 30-m nonpolarGC 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–phosphorusdetector (NPD) to analyze for NMs and their metabolites. However, the major-ity of studies use GC–MS with electron impact ionization to detect and quan-tify the wide array of FM structures [2, 5, 8, 11, 14, 18–21, 24].An ion-trap massspectrometer has been used to analyze for FMs by GC–MS/MS and negativechemical 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 theextraction solvent to low volumes prior to analysis, and volatile solvents shouldbe used in order to minimize the loss of FMs during this step [2].

3Sampling Wastewater Treatment Plants for Fragrance Materials

3.1Selection of Wastewater Treatment Plants

Because FMs are used in consumer products, it is important that investigationsof their removal during wastewater treatment be conducted at municipal waste-water treatment plants primarily receiving domestic wastewater (>80% do-mestic wastewater) [2]. In addition, the operation of these plants should be welldocumented 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 treatmentplants should be selected to represent a range of dry and wet climates, geo-graphic locations (such as the U.S. and Europe), and plant designs (includingprimary treatment only, activated sludge, carousel, oxidation ditch, trickling fil-ter, rotating biological contactor, lagoon, etc.) in order to obtain a comprehen-sive understanding of FM removal during wastewater treatment. Plant designsshould be selected so that the most prevalent types of wastewater treatmentplant design for that geography are sampled, based on both total wastewaterflow 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, activatedsludge, carousel, trickling filter, and oxidation ditch are among the most com-mon 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 of1.4¥106–1.0¥108 L day–1. Twelve of the plants were located in different statesand regions of the U.S. and five plants were located in Europe. In addition, fiveof the 17 plants were activated sludge plants, two were carousel plants, two wereoxidation ditch plants, five were trickling filter plants, one was a rotating bio-

92 S. L. Simonich


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. Regressionsinclude 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

logical contactor plant, and two were lagoons.Artola-Garicano et al. [24] stud-ied the removal of freely dissolved and total concentrations of AHTN andHHCB in four wastewater treatment plants located in The Netherlands. Theseplants had flow rates ranging from 800 to 4,200 m3/h, HRTs ranging from 4.8 to8.9 h, and SRTs ranging from 8 to 22 days [24]. Kanda et al. [22] studied PCMsand NMs in influent and effluent from six wastewater treatment plants in theU.K. The flow rates of these plants ranged from 103 to 3,198 m3/day and in-cluded rotating biological contactor with reed beds, submerged aerated filter,oxidation ditch, biological filter bed, activated sludge, and trickling filter plantdesigns. Finally, Buerge et al. [8] measured HHCB and AHTN in effluents fromfive wastewater treatment plants in Switzerland with flow rates ranging from3,177 to 14,250 m3/day–1.

Daily BOD and TSS removal at the plant should be measured and evaluated inorder 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 FMmonitoring takes place at the plant.This is important,because Simonich et al. [11]showed that the overall removal of biodegradable, nonsorptive FMs from mostplant designs is positively correlated with plant BOD removal and that the over-all removal of nonbiodegradable, sorptive FMs is positively correlated with plantTSS 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 TSSremoval is governed by both plant design and daily operation.

3.2Wastewater Treatment Plant Sampling

When planning which wastewater compartments to measure and at what fre-quency, it is important to consider the potential for FM concentrations to varythroughout the day, within the plant. Simonich et al. [2] measured 16 FMs in theinfluent of a U.S. wastewater treatment plant every 2 h over a 24-h period. Theirdata indicate that the total FM concentration in influent varies greatly through-out a 24-h period, with a relative standard deviation in total FM concentrationof 38.9% (see Fig. 5). This variation in influent concentrations has been observedfor other consumer product chemicals, such as surfactants [2], and is a func-tion 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. andhigh influent concentrations from 9:30 a.m. to 9:30 p.m. These time periods areconsistent with consumer use of these chemicals, considering that sewer trans-port times to wastewater treatment plants can be on the order of 0.5–2 h insome locations. Simonich et al. [2] also observed that the total FM concentra-tion 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 dueto the capacity of the treatment process to treat fluctuations in influent con-centration due to the long residence times within the plant.

94 S. L. Simonich

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

Artola-Garicano et al. [24] measured the free and total concentrations ofAHTN and HHCB in the influent of a wastewater treatment plant in TheNetherlands every 2 h over a 24-h period. Their data indicate that the variationin 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. Theseauthors suggested that fluctuations in water volume cause fluctuations in totalconcentrations; however, for hydrophobic FMs such as AHTN and HHCB, thesolids act as a reservoir and stabilize the free concentrations.

If the objective of measuring FMs, or other consumer product chemicals forthat matter, in wastewater treatment is to understand FM removal and mech-anisms of removal across wastewater treatment processes, then it is importantto 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, rep-resentative grab sampling may be sufficient.

4Mechanisms of Fragrance Material Removal During Wastewater Treatment

Because of their wide range of physical-chemical properties and biodegrad-abilities (see Table 1), FMs have the potential to biodegrade, sorb to solids,and/or volatilize during wastewater treatment. The relative importance of theseremoval mechanisms will depend on the specific FM, the plant design, and thekinetics of each of these processes within the plant. It is also important to acknowledge that FMs bound to solids are not available for biodegradation orvolatilization, and it is thought that only FMs freely dissolved in the aqueousphase are available for these processes [24].

4.1Biodegradation

As shown in Table 1, many FMs meet the biodegradation criteria of a ready orinherent test. If a FM meets the criteria of a ready test, with or without acclima-tion,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 notpass ready or inherent biodegradation tests. However, this does not mean thatthese FMs do not undergo biotransformation to polar metabolites under real-istic conditions. These realistic biodegradation tests may be conducted in vitro,in bench-top die-away studies, or as continuous activated sludge and porouspot tests. Ideally, the conditions should include: (1) realistic FM concentrations


96 S. L. Simonich

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

Federle et al. [12] Artola-Garicano et al. [13]

PCM Total kbiodeg Total kbiodeg

concentration (h–1) concentration (h–1)(mg/L) (mg/L)

AHTN 5 0.015±0.001 5.25 0.02350 0.008±0.001

HHCB 5 0.010±0.002 10.33 0.07125 0.021±0.003

(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 AHTNunder realistic activated sludge conditions. Federle et al. [12] studied the bio-transformation of 14C-HHCB in activated sludge die-away tests using realisticHHCB 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 cor-responding octanol–water partition coefficients estimated by HPLC. The first-order rate constant for parent HHCB biotransformation in activated sludge wasdetermined to be 0.010–0.021 h–1, depending on HHCB concentration, in thesetests (see Table 3). Federle et al. [12] also studied the biotransformation of14C-AHTN in similar activated sludge die-away tests and in a continuous acti-vated sludge test. Polar biotransformation products of AHTN were identifiedand their octanol–water partition coefficients were estimated from HPLC data.The first-order rate constant for parent AHTN biotransformation in activatedsludge was determined to be 0.008–0.015 h–1, depending on AHTN concentra-tion in the activated sludge die-away test (see Table 3) [12]. The overall removalof AHTN (due to biotransformation, sorption, volatilization) in the continuousactivated sludge test was 86.4% and the removal of parent AHTN due to bio-transformation was estimated to be 37.4%.

Artola-Garicano et al. [13] studied the biodegradation of AHTN and HHCBin activated sludge by measuring the free concentration, using negligible de-pletion SPME, and total concentration over time. The first-order rate constantfor parent AHTN biotransformation was determined to be 0.023 h–1, while thefirst-order rate constant for parent HHCB biotransformation was 0.071 h–1 (seeTable 3) [13]. These authors also determined that microbial biodegradation ac-tivity was the rate-limiting step in biotransformation of these compounds andnot desorption from the activated sludge.

Table 3 shows that the first-order rate constants for parent AHTN and HHCBbiotransformation, determined by Federle et al. [12] and Artola-Garicano et al.

[13], are similar even though the techniques used to determine these rate con-stants were quite different and the sources of activated sludge were different (U.S.and Europe). Finally, there is empirical evidence of the biotransformation of NMsand PCMs during wastewater treatment from measurements of the aminometabolites of the NMs and the lactone of HHCB in sewage sludge [15–17].

4.2Sorption

For the hydrophobic, nonbiodegradable FMs listed in Table 1, such as the NMsand PCMs, removal due to sorption on sewage solids is a significant removalmechanism. Evidence of the significance of this removal mechanism is the mea-surement 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 salicy-late, hexylcinnamic aldehyde, AHTN, HHCB, g-methyl ionone, musk ketone,musk xylene, and OTNE in digested and dewatered sludge samples. The resultsindicate that FMs with activated sludge sorption coefficients (Kd) as low as2,000 L kg–1 have the potential to be removed to a significant degree due to sorp-tion to sewage sludge.

FM sorption coefficients for sewage sludge have most often been estimatedusing the octanol–water partition coefficient of the FM rather than measureddirectly. There is general agreement between the Kd values measured for AHTNand HHCB and the estimates based on their log Kow values (see Table 1) [11].Artola-Garicano et al. [13] determined the organic carbon normalized sorptioncoefficient for activated sludge to be 6,681 L kg–1 for HHCB and 7,018 L kg–1 forAHTN. Finally, Federle et al. [12] estimated the removal of 14C-AHTN in a con-tinuous activated sludge test to be 44.7% based on sorption to activated sludgealone.

4.3Volatilization

FMs have the potential to volatilize and enter the atmosphere during manufac-turing and consumer use and disposal. The PCMs and NMs have been detectedin ambient air [9,25].However,most FMs have atmospheric lifetimes sufficientlyshort that they are unlikely to undergo atmospheric long-range transport [26].

Because some FMs have large Henry’s law constants (Table 1) and somewastewater treatment plant designs have active aeration and large surface areasthat are exposed to the atmosphere, it is likely that some FMs volatilize duringwastewater treatment. However, the FMs with large Henry’s law constants alsohave large Kd values (see Table 1), so that in portions of the treatment processwith active aeration and high solids concentrations (such as activated sludge)it is not entirely clear whether volatilization or sorption to solids will be thedominant loss mechanism. Although there are limited experimental data onFM volatilization during wastewater treatment, volatilization appears to ac-


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 sludgetest to be 3.4% based on volatilization alone.

5Measurement of Fragrance Materials in Wastewater Treatment

5.1Concentrations in Treatment Plants

Given the variety of analytical methods used to measure FMs in wastewatertreatment and the number of researchers measuring these compounds inwastewater collected from the U.S. and Europe (see Table 2), it is important tocompare and contrast the concentrations of FMs measured throughout waste-water treatment by different researchers, in different geographies. In addition,one might expect that there would be plant-to-plant variation in the FM con-centration in wastewater treatment because of differences in the volume of FMsused (per capita FM use) and differences in the per capita water use for a givengeography. These measurements and differences are important to understandbecause, ultimately, the concentration of FMs in the final effluent and sewagesludge are used to develop aquatic and terrestrial environmental risk assess-ments for these compounds [1].

Simonich et al. [11] compared the concentration of 16 FMs in wastewater in-fluent collected from 12 U.S. treatment plants and five treatment plants fromthe U.K. and The Netherlands (see Table 4). It is important to characterize FMsin influent because it is a good representation of the relative amounts of FMsbeing used by consumers and there is minimal opportunity for biodegradation,sorption, and/or volatilization in transit to the wastewater treatment plant. Inaddition, because the same analytical methods were used by the same labora-tory to measure U.S. and European influent in the Simonich et al. [11] study, wecan directly compare concentrations between the U.S. and Europe withoutquestioning differences in laboratory procedures or analytical methodology.

Simonich et al. found that the influent concentrations of 14 of the 16 FMsmeasured were statistically similar in both U.S. and European treatment plants[11]. Only isobornyl acetate and OTNE influent concentrations were statisti-cally different, with higher concentrations of these FMs in European influent.The influent concentrations of the 16 FMs are given in Table 4. Figure 6A showsthe normalized relative profile of FMs measured in U.S. and European influentduring the [11] study. In both sets of influent, terpineol dominated the FM pro-file and had the highest concentrations, while the nitromusks (MX and MK)had the lowest concentrations. The large standard deviations in the influentconcentrations (Table 4 and Fig. 6A) indicate that there is significant varia-bility in influent concentrations within both U.S. and European treatmentplants.

98 S. L. Simonich


Tabl

e4

Sum

mar

y of

FM in

fluen

t con

cent

rati

ons

to w

aste

wat

er tr

eatm

ent

Res

earc

hers

Loca

tion

and

Ye

ars

PCM

con

cent

rati

on

NM

con

cent

rati

on

Oth

er F

M c

once

ntra

tion

nu

mbe

rsa

mpl

ed(m

g/L)

– m

ean

(mg/

L) –

mea

n

(mg/

L)–

mea

n±st

anda

rd

±st

anda

rd d

evia

tion

±

stan

dard

dev

iati

on

devi

atio

n an

d/or

ran

gean

d/or

ran

gean

d/or

ran

ge

Sim

onic

h U

.S.;

1997

–199

9A

HT

N=

12.5

±7.

35;

MK

=0.

640±

0.39

5;B

enzy

l ace

tate

=3.

74±

3.46

;et

al.

[2,1

1]n=

12 p

lant

sH

HC

B=16

.6±

10.4

MX

=0.

386±

0.29

9m

ethy

l sal

icyl

ate=

10.2

±9.

69;

met

hyl d

ihyd

roja

smon

ate

=7.

21±

4.19

;ter

pine

ol=

63.7

±36

.4;

benz

yl s

alic

ylat

e=19

.5±

10.8

;is

obor

nyl a

ceta

te=

6.47

±8.

53;

g-m

ethy

l ion

one=

3.37

±2.

56;

p-t-

buci

nal=

1.61

±0.

731;

hexy

lcin

nam

alde

hyde

=15

.3±

12.1

;he

xyl s

alic

ylat

e=5.

48±

3.56

;O

TN

E=3.

55±

1.93

;ac

etyl

ced

rene

=4.

97±

2.27

Sim

onic

h

U.K

.and

19

99–2

000

AH

TN

=5.

97±

3.88

;M

K=

0.99

6±0.

741;

Ben

zyl a

ceta

te=

9.85

±10

.2;

et a

l.[1

1]T

he N

ethe

rlan

ds;

HH

CB=

9.71

±5.

09M

X=

0.24

8±0.

136

met

hyl s

alic

ylat

e=11

.3±

13.0

;n=

5 pl

ants

met

hyl d

ihyd

roja

smon

ate=

11.9

±5.

31;

terp

ineo

l=56

.3±

33.9

;be

nzyl

sal

icyl

ate=

10.2

±4.

51;

isob

orny

l ace

tate

=37

.1±

28.4

;g-

met

hyl i

onon

e=3.

63±

1.90

;p-

t-bu

cina

l=2.

56±

1.96

;he

xylc

inna

mal

dehy

de=

12.8

±7.

27;

hexy

l sal

icyl

ate=

6.89

±3.

63;

OT

NE=

9.00

±3.

77;

acet

yl c

edre

ne=

7.15

±4.

32

100 S. L. Simonich

Tabl

e4

(con

tinu

ed)

Res

earc

hers

Loca

tion

and

Ye

ars

PCM

con

cent

rati

on

NM

con

cent

rati

on

Oth

er F

M c

once

ntra

tion

nu

mbe

rsa

mpl

ed(m

g/L)

– m

ean

(mg/

L) –

mea

n

(mg/

L)–

mea

n±st

anda

rd

±st

anda

rd d

evia

tion

±

stan

dard

dev

iati

on

devi

atio

n an

d/or

ran

gean

d/or

ran

gean

d/or

ran

ge

Art

ola-

Gar

ican

o T

he N

ethe

rlan

ds;

2001

AH

TN

=0.

54±

0.05

N

ot m

easu

red

Not

mea

sure

det

al.

[24]

– to

tal

n=4

plan

ts–1

.76±

0.09

;co

ncen

trat

ions

HH

CB=

1.42

±0.

12–4

.30±

0.23

Kan

da e

t al.

[22]

U

K;n

=6

plan

ts20

01A

HT

N=

2.2–

8.1;

MA

=<

0.01

;N

ot m

easu

red

HH

CB=

8.4–

19.2

;M

X=

<0.

01–4

.7;

DPM

I=<

0.01

–0.4

;M

M=

<0.

01;

AD

BI=

<0.

01–0

.44;

MT

=<

0.01

;A

HD

I=<

0.01

–0.1

;M

K=

<0.

01–2

.9AT

II=

<0.

01–2

.9

Rim

kus

et a

l.[2

3]

Ger

man

y;19

96N

ot m

easu

red

MX

=0.

150;

Not

mea

sure

dn=

1 pl

ant

4-N

H2-

MX

=<

0.01

;2-

NH

2-M

X=

<0.

01;

MK

=0.

55;

2-N

H2-

MX

=<

0.01


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

B

A

When we compare the PCM and NM influent concentrations measured bySimonich et al. to those of other researchers (Table 4), we can ascertain somegeneral trends. AHTN and HHCB are the highest concentration PCMs in bothU.S. and European influent, with concentrations ranging from 1 to 20 mg/L, andHHCB concentrations being greater than AHTN concentrations. The ratio ofHHCB to AHTN concentration in U.S. and European influent is in the range of1.3–3.8. In general, the concentration of other PCMs in influent are either be-low the limit of quantitation or less than 1 mg/L [22].

The NM concentrations in U.S. and European influent are significantly lessthan 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 meta-bolites of the nitromusks are not detected in influent because there is limitedopportunity for biotransformation in transit to the wastewater treatment plants.

A larger number of researchers have measured FMs in final effluent fromwastewater treatment (Table 5) because this is the most important wastewaterparameter for accessing FM discharge to aquatic ecosystems. The concentra-tion of FMs in final effluent is a function of the FM concentration in influentand the efficiency of FM removal across the plant (including plant design andoperation).

In the Simonich et al. [11] study, the concentration of 16 FMs in final efflu-ent ranged from 0.001–7.6 mg/L in the U.S. to 0.01–4.6 mg/L in Europe (seeTable 5). In addition, terpineol no longer dominated the relative FM profile infinal 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 ef-fluent (see Table 5) [14, 19, 22]. In the U.S., the concentration of AHTN in finaleffluent ranged from 0.024 to 1.7 mg/L, while the concentration of HHCB rangedfrom 0.032 to 2.2 mg/L (see Table 5) [11, 14]. In Europe, the final effluent con-centrations of AHTN ranged from 0.11 to 2.7 mg/L and the concentrations ofHHCB 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 con-centration range of <1–710 ng/L in European effluents and <MDL–112 ng/L inU.S. effluents (see Table 5). The amino metabolites of the NMs have been detectedin U.S. and European effluent in the concentration range of <MDL–250 ng/L (seeTable 5).

It is also important to consider the concentration of FMs in sewage sludgebecause this is another route in which FMs may be removed during wastewatertreatment and be released to the environment through land application ofsludge solids. Only one study has attempted to measure a large number of FMsin digested sludge. Difrancesco et al. [18] detected AHTN, HHCB, MK, g-methylionone, hexylcinnamaldehyde, hexyl salicylate, and acetyl cedrene in digestedsludge from U.S. wastewater treatment plants (see Table 6). The other FMs theystudied were not detected.As Table 6 indicates, the concentrations of these FMs

102 S. L. Simonich


Tabl

e5

Sum

mar

y of

FM fi

nal e

fflu

ent c

once

ntra

tion

s fr

om w

aste

wat

er tr

eatm

ent

Res

earc

hers

Loca

tion

and

Ye

ars

PCM

con

cent

rati

on

NM

con

cent

rati

on

Oth

er F

M c

once

ntra

tion

nu

mbe

rsa

mpl

ed(m

g/L)

– m

ean

(mg/

L) –

mea

n (m

g/L)

– m

ean±

stan

dard

±

stan

dard

dev

iati

on

±st

anda

rd d

evia

tion

de

viat

ion

and/

or r

ange

and/

or r

ange

and/

or r

ange

Sim

onic

h U

.S.;

1997

–199

9A

HT

N=

24–1

,710

;M

K=

10–6

7;B

enzy

l ace

tate

=2–

252;

et a

l.[2

,11]

n=12

pla

nts

HH

CB=

32–2

,210

MX

=1–

112

met

hyl s

alic

ylat

e=13

–693

;m

ethy

l dih

ydro

jasm

onat

e=3–

456;

terp

ineo

l=11

–1,0

79;

benz

yl s

alic

ylat

e=5–

1,02

5;is

obor

nyl a

ceta

te=

7–11

2;g-

met

hyl i

onon

e=7–

214;

p-t-

buci

nal=

13–2

58;

hexy

lcin

nam

alde

hyde

=10

–77;

hexy

l sal

icyl

ate=

1–24

3;O

TN

E=25

–615

;ac

etyl

ced

rene

=12

–1,3

59

Sim

onic

hU

.K.a

nd

1999

–200

0A

HT

N=

620–

2,67

0;M

K=

40–7

70;

Ben

zyl a

ceta

te=

60–2

60;

et a

l.[1

1]T

he N

ethe

rlan

ds;

HH

CB=

980–

4,62

0M

X=

10–1

70m

ethy

l sal

icyl

ate=

40–2

20;

n=5

plan

tsm

ethy

l dih

ydro

jasm

onat

e=26

–1,9

20;

terp

ineo

l=80

–15,

100;

benz

yl s

alic

ylat

e=20

–1,9

60;

isob

orny

l ace

tate

=10

–290

;g-

met

hyl i

onon

e=30

–730

;p-

t-bu

cina

l=40

–180

;he

xylc

inna

mal

dehy

de=

20–9

10;

hexy

l sal

icyl

ate=

10–9

10;

OT

NE=

490–

3,19

0;ac

etyl

ced

rene

=70

–1,4

30

104 S. L. Simonich

Tabl

e5

(con

tinu

ed)

Res

earc

hers

Loca

tion

and

Ye

ars

PCM

con

cent

rati

on

NM

con

cent

rati

on

Oth

er F

M c

once

ntra

tion

nu

mbe

rsa

mpl

ed(m

g/L)

– m

ean

(mg/

L) –

mea

n (m

g/L)

– m

ean±

stan

dard

±

stan

dard

dev

iati

on

±st

anda

rd d

evia

tion

de

viat

ion

and/

or r

ange

and/

or r

ange

and/

or r

ange

Art

ola-

Gar

ican

o T

he N

ethe

rlan

ds;

2001

AH

TN

=42

0±60

Not

mea

sure

dN

ot m

easu

red

et a

l.[2

4] –

tota

l n=

4 pl

ants

–1,2

00±

180;

conc

entr

atio

nsH

HC

B=1,

250±

20–2

,220

±90

Kan

da e

t al.

[22]

UK

;20

01A

HT

N=

310–

2,70

0;M

A=

<10

;N

ot m

easu

red

n=

6 pl

ants

HH

CB=

1,10

0–6,

400;

MX

=<

10–6

50;

DPM

I=<

10–1

60;

MM

=<

10;

AD

BI=

<10

–91;

MT

=<

10;

AH

DI=

<10

–48;

MK

=<

10–7

10AT

II=

<10

–790

Rim

kus

Ger

man

y;19

96N

ot m

easu

red

MX

=<

3–10

;N

ot m

easu

red

et a

l.[2

3]n=

3 pl

ants

4-N

H2-

MX

=7–

34;

2-N

H2-

MX

=<

1–10

;M

K=

6–94

;2-

NH

2-M

K=

15–2

50

Ric

king

C

anad

a;20

02A

HT

N=

42–1

04;

MX

=<

1;M

K=

<1

Not

mea

sure

det

al.

[19]

n=3

plan

tsH

HC

B=15

7–42

3;D

PMI=

<1;

AD

BI=

2–8;

AH

DI=

2–5;

ATII

=<

1


Tabl

e5

(con

tinu

ed)

Res

earc

hers

Loca

tion

and

Ye

ars

PCM

con

cent

rati

on

NM

con

cent

rati

on

Oth

er F

M c

once

ntra

tion

nu

mbe

rsa

mpl

ed(m

g/L)

– m

ean

(mg/

L) –

mea

n (m

g/L)

– m

ean±

stan

dard

±

stan

dard

dev

iati

on

±st

anda

rd d

evia

tion

de

viat

ion

and/

or r

ange

and/

or r

ange

and/

or r

ange

Ric

king

et a

l.[1

9]Sw

eden

;19

99A

HT

N=

110–

520;

MX

=<

1;M

K=

<1

Not

mea

sure

dn=

5 pl

ants

HH

CB=

205–

1,30

0;D

PMI=

<1;

AD

BI=

4–19

;A

HD

I=2–

6;AT

II=

<1

Ose

mw

engi

eU

.S.;

Unk

now

nA

HT

N=

26.6

–92.

2;M

X=

<M

DL–

1.3;

Not

mea

sure

d

et a

l.[1

4]n=

2 pl

ants

HH

CB=

35.0

–152

;M

K=

<M

DL–

27.5

;D

PMI=

<M

DL;

MA

=<

MD

L;A

DBI

=0.

3–2.

1;M

M=

<M

DL;

AH

DI=

2.4–

5;M

T=

<M

DL;

ATII

=<

MD

L–12

64-

NH

2-M

X=

<M

DL–

31.5

;2-

NH

2-M

X=

<M

DL–

0.9;

NH

2-M

K=

<M

DL

Buer

ge e

t al.

[8]

Swit

zerl

and;

2001

AH

TN

=31

0–76

0;N

ot m

easu

red

Not

mea

sure

dn=

5H

HC

B=72

0–1,

950

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 con-centration FM (in mg/L) is in parentheses. The error bars represent the normalized standarddeviation of the mean


Tabl

e6

Sum

mar

y of

FM c

once

ntra

tion

s in

was

tew

ater

slu

dge

Res

earc

hers

,Lo

cati

on a

ndYe

ars

PCM

con

cent

rati

on –

N

M c

once

ntra

tion

–O

ther

FM

con

cent

rati

on –

sam

ple

type

,nu

mbe

rsa

mpl

edm

ean±

stan

dard

m

ean±

stan

dard

m

ean±

stan

dard

dev

iati

onan

d co

ncen

trat

ion

devi

atio

n an

d/or

de

viat

ion

and/

or

and/

or r

ange

unit

sra

nge

rang

e

Art

ola-

Gar

ican

o T

he N

ethe

rlan

ds;

2001

AH

TN

=12

.38±

0.19

Not

mea

sure

dN

ot m

easu

red

et a

l.[2

4] –

tota

l n=

4 pl

ants

–82.

67±

43.0

9;co

ncen

trat

ion

HH

CB=

29.4

7±10

.84

was

te s

ludg

e in

mg/

L–2

34.6

0±11

1.86

Ber

set e

t al.

[16]

–

Swit

zerl

and;

Unk

now

nN

ot m

easu

red

MA

=nd

;N

ot m

easu

red

conc

entr

atio

nn=

10 p

lant

sM

X=

nd–3

2.5;

inmg

/kg

MM

=nd

;MT

=nd

;dr

y w

eigh

tM

K=

nd–7

.1,

amin

o m

etab

olite

s=nd

–49.

1

Her

ren

et a

l.[1

5] –

Swit

zerl

and;

Unk

now

nA

HT

N=

741–

4,16

1;M

A=

nd;

Not

mea

sure

dco

ncen

trat

ion

n=12

pla

nts

HH

CB=

2,29

3–12

,157

;M

X=

nd–3

2.5;

inmg

/kg

D

PMI=

38.4

–332

;M

M=

nd;M

T=

nd;

dry

wei

ght

AD

BI=

41–3

30;

MK

=nd

–7.0

,A

HD

I=64

.9–8

43;

amin

o m

etab

olite

s=

ATII

=n.

q.nd

–36.

2

108 S. L. Simonich

Tabl

e6

(con

tinu

ed)

Res

earc

hers

,Lo

cati

on a

ndYe

ars

PCM

con

cent

rati

on –

N

M c

once

ntra

tion

–O

ther

FM

con

cent

rati

on –

sam

ple

type

,nu

mbe

rsa

mpl

edm

ean±

stan

dard

m

ean±

stan

dard

m

ean±

stan

dard

dev

iati

onan

d co

ncen

trat

ion

devi

atio

n an

d/or

de

viat

ion

and/

or

and/

or r

ange

unit

sra

nge

rang

e

Kup

per

et a

l.[1

7] –

Swit

zerl

and;

2001

AH

TN

=2,

500–

11,2

00;

Not

mea

sure

dN

ot m

easu

red

conc

entr

atio

n in

n=16

pla

nts

HH

CB=

7,40

0–3,

600;

mg/k

g dr

y w

eigh

tA

DBI

=10

0–1,

100;

AH

DI=

200–

1,80

0;AT

II=

200–

1,00

0;H

HC

B la

cton

e=

600–

3,50

0

Stev

ens

et a

l.[2

1] –

U.K

.;U

nkno

wn

AH

TN

=12

0–16

,000

;M

A,M

X,M

M,

Not

mea

sure

d

dige

sted

slu

dge

n=14

pla

nts

HH

CB=

1,90

0–81

,000

;an

d M

T

conc

entr

atio

n in

AD

BI=

10–2

60;

not d

etec

ted

mg/k

g dr

y w

eigh

tA

HD

I=32

–1,1

00;

ATII

=44

–1,1

00;

DPM

I – n

ot d

etec

ted

Dif

ranc

esco

et a

l.U

.S.;

2000

A

HT

N=

8,10

0–51

,000

;M

K=

1,30

0;g-

Met

hyl i

onon

e=1,

100–

3,80

0;[1

8] –

dig

este

d n=

2 pl

ants

and

2002

HH

CB=

21,8

00–8

6,00

0M

X n

ot d

etec

ted

hexy

lcin

nam

alde

hyde

=4,

100;

slud

ge c

once

ntra

tion

he

xyl s

alic

ylat

e=1,

500;

inmg

/kg

dry

wei

ght

OT

NE=

7,30

0–30

,700

;ac

etyl

ced

rene

=90

0–31

,300

;re

mai

ning

FM

s no

t det

ecte

d


ranged from 900 to 86,000 mg/kg dry weight, with AHTN and HHCB having thehighest concentrations (8,100–86,000 mg/kg dry weight) [18].

The remaining studies on sewage sludge were conducted at European waste-water treatment plants (see Table 6). Of the PCMs, AHTN and HHCB had thehighest concentrations on European sludge (120–81,000 mg/kg dry weight),however the other PCMs were also detected. In one study, the lactone of HHCBwas also detected on sewage sludge [17]. The NMs and their amino metabo-lites have been detected on sewage sludge, however their concentrations (nd–1,300 mg/kg dry weight) are much lower than the PCMs.

5.2Removal During Treatment

Removal of FMs during wastewater treatment is a function of the tendency forFMs to biodegrade, sorb, and/or volatilize during treatment, plant design, andplant operation. FMs may be removed from wastewater during primary andsecondary treatment.

Simonich et al. [11] showed that FMs undergo significant removal duringprimary treatment, ranging from 14 to 50% removal (Table 7). Because bothsorptive and nonsorptive FMs are removed, these authors suggested that bothsorption and biodegradation play a role in the removal of FMs during primarytreatment. Further evidence of this is in the comparison of Figs. 6A and 6B, inwhich the relative profile of FMs does not change significantly from influent(Fig. 6A) to primary effluent (Fig. 6B), although the FM concentrations de-crease. 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 de-viations in the primary treatment removals measured by Simonich et al. [11]suggest that there is significant plant-to-plant variability in primary removalof FMs.

Simonich et al. [2, 11], Artola-Garicano et al. [24], Kanda et al. [22], andRimkus et al. [5] studied the removal of FMs following primary and secondarytreatment (see Table 7). Simonich et al. [11] showed that FM removal is de-pendent on plant design and operation and ranges from 50 to 99.9%. The over-all removal (primary + secondary treatment) of 16 FMs ranged from 87.8 to99.9% for activated sludge plants, 58.6 to 99.8% for carousel plants, 88.9 to99.9% for oxidation ditch plants, 71.3 to 98.6% for trickling filter plants, 80.8 to99.9% 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 low-est final effluent concentrations due to long retention times. The relative FMprofiles in final effluent (Fig. 7), broken down by treatment type, gives someperspective on which plant designs are best at removing biodegradable chem-icals (significant enhancement of nonbiodegradable, sorptive FMs as in thecase of activated sludge, rotating biological contactor, and oxidation ditchplants) and those which are not (enhancement of both nonbiodegradable, sorp-

110 S. L. Simonich

Tabl

e7

Sum

mar

y of

perc

ent r

emov

al o

fFM

s du

ring

was

tew

ater

trea

tmen

t

Res

earc

hers

Loca

tion

and

Ye

ars

Perc

ent P

CM

Pe

rcen

t NM

Pe

rcen

t oth

er F

M r

emov

alnu

mbe

rsa

mpl

edre

mov

alre

mov

al

Sim

onic

h U

.S.,

UK

,19

97 –

200

0Pr

imar

y re

mov

al

Prim

ary

Prim

ary

rem

oval

onl

y:et

al.

[2,1

1]T

he N

ethe

rlan

ds;

only

:re

mov

al o

nly:

benz

yl a

ceta

te=

28.2

±27

.5;

n=17

pla

nts

AH

TN

=28

.9±

20.1

;M

X=

41.2

±21

.9;

met

hyl s

alic

ylat

e=40

.4±

32.2

;H

HC

B=29

.9±

23.4

MK

=26

.6±

21.5

met

hyl d

ihyd

roja

smon

ate=

14.6

±19

.4;

terp

ineo

l=15

.5±

12.0

;Pr

imar

y+Pr

imar

y+

benz

yl s

alic

ylat

e=41

.8±

25.0

;se

cond

ary

rem

oval

:se

cond

ary

rem

oval

:is

obor

nyl a

ceta

te=

29.0

±29

.5;

AH

TN

=50

.6–9

9.9;

MX

=87

.6–9

9.9;

g-m

ethy

l ion

one=

20.8

±19

.6;

HH

CB=

63.5

–99.

7M

K=

85.2

–96.

7p-

t-bu

cina

l=50

.6±

23.4

;he

xylc

inna

mal

dehy

de=

47.1

±19

.4;

hexy

l sal

icyl

ate=

37.3

±21

.0;

OT

NE=

28.8

±22

.7;

acet

yl c

edre

ne=

31.6

±20

.3

Prim

ary+

seco

ndar

y re

mov

al:

benz

yl a

ceta

te=

86.4

–99.

9;m

ethy

l sal

icyl

ate=

92.0

–99.

9;m

ethy

l dih

ydro

jasm

onat

e=81

.9–9

9.9;

terp

ineo

l=95

.4–9

9.9;

benz

yl s

alic

ylat

e=90

.3–9

9.9;

isob

orny

l ace

tate

=84

.5–9

9.9;

g-m

ethy

l ion

one=

83.1

–99.

8;p-

t-bu

cina

l=84

.8–9

9.3;

hexy

lcin

nam

alde

hyde

=95

.3–9

9.9;

hexy

l sal

icyl

ate=

96.4

–99.

9;O

TN

E=51

.4–9

9.4;

acet

yl c

edre

ne=

71.3

–99.

9


Tabl

e7

(con

tinu

ed)

Res

earc

hers

Loca

tion

and

Ye

ars

Perc

ent P

CM

Pe

rcen

t NM

Pe

rcen

t oth

er F

M r

emov

alnu

mbe

rsa

mpl

edre

mov

alre

mov

al

Art

ola-

Gar

ican

oT

he N

ethe

rlan

ds;

2001

Prim

ary+

Not

mea

sure

dN

ot m

easu

red

et

al.

[24]

– to

tal

n=4

plan

tsse

cond

ary

rem

oval

:co

ncen

trat

ions

AH

TN

=14

.3–5

6.3;

HH

CB=

12.0

–59.

8

Kan

da e

t al.

[22]

UK

;n=

6 pl

ants

2001

Prim

ary+

Prim

ary+

N

ot m

easu

red

se

cond

ary

rem

oval

:se

cond

ary

rem

oval

:A

HT

N=

40.0

–96.

17;

MX

=80

.3–8

6.2;

HH

CB=

39.0

5–93

.49;

MK

=53

.6–6

4.5

othe

rs n

ot g

iven

Rim

kus

et a

l.[2

3]G

erm

any;

n=1

1996

Not

mea

sure

dPr

imar

y+

Not

mea

sure

dse

cond

ary

rem

oval

:M

X=

93.3

;MK

=98

.9

tive FMs and some biodegradable, nonsorptive FMs as in the case of tricklingfilter). As mentioned previously, Simonich et al. [11] showed that plant opera-tion, including the efficiency of removal of TSS and BOD, affects the removalof FMs to a significant degree (see Fig. 4 and earlier discussion) and is likely amore important variable in insuring efficient removal of FMs during waste-water treatment than is plant design alone.

The overall removal (primary + secondary treatment) of AHTN and HHCBhas been measured by Simonich et al. [11], Artola-Garicano et al. [24], andKanda et al. [22] (see Table 7). Simonich et al. measured the overall removal ofAHTN 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 TSSremoval. Artola-Garicano et al. measured the overall removal of AHTN to be14.3–56.3% and the removal of HHCB to be 12.0–59.8% at four treatment plantsin The Netherlands, based on total concentrations. Finally, Kanda et al. mea-sured the overall removal of AHTN to be 40.0–96.2% and the removal of HHCBto 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 etal. [24] appears to be significantly lower than the overall removals measured bySimonich et al. [11] and Kanda et al. [22]. This may be due, in part, to the col-lection of grab samples by Artola-Garicano et al. and the collection of flow-based composite samples by Simonich et al. and Kanda et al., to the relativelyshort hydraulic retention times in several of the treatment plants monitored byArtola-Garicano et al., and/or excessive levels of TSS in effluent from the plantsmonitored by Artola-Garicano et al.As previously mentioned, the FM influentconcentrations vary significantly throughout the day (Fig. 5).Artola-Garicanoet al. [24] confirmed that the total AHTN and HHCB concentration in influentvaried throughout the day, while the free FM concentration showed less vari-ability. Simonich et al. [11] measured relatively low overall removals of AHTNand 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 reportedby Artola-Garicano et al. [24] for other treatment plants in The Netherlands.Finally, the range of AHTN and HHCB overall removal measured by Simonichet al. [11] and Kanda et al. [22] is comparable. Both studies collected flow-com-posite samples and sampled a variety of different plant designs.

Simonich et al. [11], Kanda et al. [22], and Rimkus et al. [23] measured theremoval of MX and MK during wastewater treatment (see Table 7). Simonichet al. and Rimkus et al. measured the overall removals of MX and MK in therange of 85–99.9%. Kanda et al. measured the removal of MX in the range of80–86% and the removal of MK in the range of 53–65%, however the removalof these NMs was not reported for all six treatment plants.

112 S. L. Simonich

6Predicting Fragrance Material Removal During Wastewater Treatment

6.1Framework for Aquatic Risk Assessment

Because of the large number of FMs in commerce, a framework documentwas developed to prioritize FMs for aquatic risk assessment [1]. An integralpart of this risk assessment is an accurate prediction of FM concentration inthe aquatic environment. In order to do this, the framework document out-lines some simple calculations to estimate the concentration and removal ofFMs during wastewater treatment. This is done by first using the annual FMvolume of use and the per capita water use in the geographic area to predictan average influent concentration for the U.S. and Europe. Next, the FM re-moval during primary treatment (sorption and settling only) is predicted us-ing the octanol–water partition coefficient. Finally, the FM removal duringsecondary treatment and final effluent concentrations are predicted in thefirst tier of the framework using the octanol–water partition coefficient to es-timate sorption. Biodegradation rates are added to sorption in the second tierof the framework if a more refined exposure assessment is needed. Readilybiodegradable, inherently biodegradable, and nonbiodegradable FMs are as-sumed to have biodegradation rates of 3, 0.3, and 0 h–1, respectively [1]. Thissimple model does not account for volatilization and the assumptions andequations in the framework document are directly applicable to all primarytreatment, but only to activated sludge secondary treatment. The largest sin-gle source of error in the model is likely the estimation of annual FM volumeuse because of uses outside the fragrance industry and, in some cases, naturalsources [1].

Because of the large number of FMs being evaluated and the wide range of physical-chemical properties and biodegradabilities they represent, it is im-portant to determine if calculations outlined in the framework document areconservative for predicting FM concentration and removal during wastewatertreatment. The Simonich et al. [11] dataset for 16 FMs measured in 17 waste-water treatment plants was used to evaluate the assumptions made in theframework document. Figure 8A shows the regression of the measured percentprimary removal (Table 7) with the percent primary removal predicted by theframework. All of the plants with primary treatment studied by Simonich etal. are included in this regression. Because of the large variation in measuredprimary removal (Table 7), the correlation is not statistically significant.Also,the framework prediction does not account for potential biodegradation dur-ing primary treatment, and Simonich et al. measured significant removal ofbiodegradable, nonsorptive FMs during primary treatment (Table 7 andFig. 6B) [11].

Figure 8B shows the regression of the measured percent overall removalfrom activated sludge treatment plants only measured by Simonich et al. with


114 S. L. Simonich

Fig. 8A, B Correlation of A measured primary removal with predicted primary removal andB measured overall removal with predicted overall removal for activated sludge plants, usingthe 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 statis-tically significant; n=16. Dashed lines are the 95% confidence intervals for the regressions

A

B

the percent overall removal predicted by the second tier of the framework (accounting for sorption and biodegradation) for activated sludge plants. Thiscorrelation is significant at the 98% level, however the framework significantlyunderpredicts overall removal (slope=0.118). This is particularly true for MK,MX, OTNE,AHTN, and HHCB, which are assumed in the framework to be non-biodegradable (Table 1) and to have biodegradation rates of 0 h–1. This suggeststhat sorption alone does not account for the removals Simonich et al. measuredfor MK, MX, OTNE,AHTN, and HHCB and that biotransformation and/or vola-tilization 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 docu-ment for estimating FM concentrations in and removal from activated sludgewastewater treatment are predictive. Finally, Fig. 9 shows that the assumptionsmade in the framework document [1] are conservative for predicting final efflu-ent concentrations,regardless of treatment type and geography (US and Europe).

6.2Simple Treat Model

Artola-Garicano et al. [27] compared their measured removals of AHTN andHHCB [24] to the predicted removal of these compounds by the wastewatertreatment plant model Simple Treat 3.0. Simple Treat is a fugacity-based, nine-box model that breaks the treatment plant process into influent, primary set-tler, primary sludge, aeration tank, solid/liquid separator, effluent, and wastesludge and is a steady-state, nonequilibrium model [27]. The model inputs in-clude information on the emission scenario of the FM, FM physical-chemicalproperties, and FM biodegradation rate in activated sludge.

In general, the Simple Treat model predicted the overall removal of totalAHTN 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 Treatwere inversely related to the measured free concentrations of these compounds[27]. As previously mentioned, the overall removal of AHTN and HHCB mea-sured by Artola-Garicano et al. (14–60%) was significantly less than the overallremoval of these compounds measured by Simonich et al. [11] and Kanda et al.[22] (39–99.9%).

7Conclusions

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


116 S. L. Simonich

Fig. 9A–D Comparison of measured FM concentrations in all plants to predicted FM con-centrations 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 secondtier (sorption and biodegradation) of the framework [11]. Points above the line have mea-sured values greater than predicted by the framework model [1], while those below the linehave measured values less than predicted. The filled circles represent concentrations in theU.S. and the open circles represent concentrations in Europe

complex. The mechanisms of FM removal from wastewater include biodegrada-tion, sorption, and/or volatilization. A wide array of analytical methods havebeen developed to measure FMs in wastewater influent, primary effluent, finaleffluent, and solids, and wastewater studies have been conducted in the U.S. andEurope. Finally, the efficient removal of FMs during wastewater treatment is notonly dependent on the biodegradability and physical-chemical properties of theFM, but is also highly dependent on plant operation and design.

B

A


Fig. 9C, D (continued)

References

1. Salvito DT, Senna RJ, Federle TW (2002) Environ Toxicol Chem 21:13012. Simonich SL, Begley WM, Debaere G, Eckhoff WS (2000) Environ Sci Technol 34:9593. Balk F, Ford RA (1999) Toxicol Lett 111:574. Tas JW, Balk F, Ford RA, van de Plassche EJ (1997) Chemosphere 35:29735. Rimkus GG (1999) Toxicol Lett 111:376. Heberer T (2002) Acta Hydrochim Hydrobiol 30:2277. Oros DR, Jarman WM, Lowe T, David N, Lowe S, Davis JA (2003) Mar Pollut Bull 46:11028. Buerge IJ, Buser HR, Muller MD, Poiger T (2003) Environ Sci Technol 37:56369. Peck AM, Hornbuckle KC (2004) Environ Sci Technol 38:367

10. Standley LJ, Kaplan LA, Smith D (2000) Environ Sci Technol 34:312411. Simonich SL, Federle TW, Eckhoff WS, Rottiers A, Webb S, Sabaliunas D, De Wolf W

(2002) Environ Sci Technol 36:2839

C

D

12. Federle TW (2004) 13. Artola-Garicano E, Borkent I, Damen K, Jager T, Vaes WHJ (2003) Environ Sci Technol

37:11614. Osemwengie LI, Steinberg S (2001) J Chromatogr A 932:10715. Herren D, Berset JD (2000) Chemosphere 40:56516. Berset JD, Bigler P, Herren D (2000) Anal Chem 72:212417. Kupper T, Berset JD, Etter-Holzer R, Furrer R, Tarradellas J (2004) Chemosphere 54:111118. DiFrancesco AM, Chiu PC, Standley LJ, Allen HE, Salvito D (2004) Environ Sci Technol

38:19419. Ricking M, Schwarzbauer J, Hellou J, Svenson A, Zitko V (2003) Mar Pollut Bull 46:41020. Verbruggen EMJ, Van Loon W, Tonkes M, Van Duijn P, Seinen W, Hermens JLM (1999)

Environ Sci Technol 33:80121. Stevens JL, Northcott GL, Stern GA, Tomy GT, Jones KC (2003) Environ Sci Technol 37:46222. Kanda R, Griffin P, James HA, Fothergill J (2003) J Environ Monit 5:82323. Rimkus GG, Gatermann R, Huhnerfuss H (1999) Toxicol Lett 111:524. Artola-Garicano E, Borkent I, Hermens JLM, Vaes WHJ (2003) Environ Sci Technol

37:311125. Kallenborn R, Gatermann R, Planting S, Rimkus GG, Lund M, Schlabach M, Burkow IC

(1999) J Chromatogr A 846:29526. Aschmann SM, Arey J, Atkinson R, Simonich SL (2001) Environ Sci Technol 35:359527. Artola-Garicano E, Hermens JLM, Vaes WHJ (2003) Water Res 37:4377

118 Fragrance Materials in Wastewater Treatment

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 Immunochemical Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . 1242.1 Antibody-Based Analytical Methods . . . . . . . . . . . . . . . . . . . . . . 1372.1.1 Immunoassays . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1382.1.1.1 Enzyme-Linked Immunosorbent Assay (ELISA) . . . . . . . . . . . . . . . . 1392.1.1.2 Enzyme-Multiplied Immunoassay Technique (EMIT) . . . . . . . . . . . . . 1402.1.1.3 Polarization Fluoroimmunoassay (PFIA) . . . . . . . . . . . . . . . . . . . . 1412.1.2 Flow-Injection Immunoassay (FIIA) . . . . . . . . . . . . . . . . . . . . . . 1412.1.3 Immunosensors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1422.1.4 Immunoaffinity Chromatography (IAC) . . . . . . . . . . . . . . . . . . . . 143

3 Immunochemical Methods for Surfactants . . . . . . . . . . . . . . . . . . 1443.1 Anionic Surfactants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1483.2 Nonionic Surfactants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1503.3 Cationic Surfactants . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 152

4 Immunochemical Methods for Polychlorinated and Polybrominated Compounds . . . . . . . . . . . . . . . . . . . . . . . . 153

4.1 PCBs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1564.2 PCDDs and PCDFs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1614.3 Chlorophenols . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 163

5 Other Industrial Residues . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1675.1 Bisphenol A . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1675.2 Phthalate Esters . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 170

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

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 173

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


Immunochemical methods for anionic (LAS), nonionic (APEs and APs), and cationic sur-factants (BDD12AC and DDAC) are extensively reviewed and the features of these assays discussed, particularly if examples of their application to environmental samples have beendescribed. Similarly, a great amount of information has been collected regarding immuno-chemical determination of organochlorinated substances such as PCBs, PCDDs, PCDFs, andchlorophenols. On the contrary, immunochemical analysis of organobrominated substances,such as the BFR agents, seems to be still a goal. Immunochemical methods have also beenreported for bisphenol A and phthalates showing excellent features. The commercial avail-ability of some of these methods is also presented.

Keywords Immunochemical techniques · Surfactants · Polyhalogenated compounds ·Bisphenol A · Phthalate esters

AbbreviationsAb AntibodyADMBAC Alkyldimethylbenzylammonium compoundsAES Alkyl ether sulfatesAg AntigenAP AlkylphenolAPEC Alkylphenol ethoxy carboxylateAPE Alkylphenol polyethoxylateAS Alcohol sulfatesATMAC Alkyltrimethylammonium compoundsBBP Butylbenzyl phthalateBDD12AC Benzyldimethyldodecylammonium chlorideBFR Brominated flame retardantsBMP-IA Bacterial magnetic particle-based immmunoassayBP BromophenolBPA Bisphenol ABSA Bovine serum albuminCIA Capillary immunoassayCLIA Chemiluminescence immunoassayCP ChlorophenolCR Cross-reactivityDADMAC Dialkyldimethylammonium compoundsDBP Dibutyl phthalateDCP DichlorophenolDDAC Didecyldimethylammonium chlorideDDT DichlorodiphenyltrichloroethaneDEHP Diethylhexyl phthalateDEQ DiesterquatDMSO DimethylsulfoxideEDC Endocrine disrupter chemicalEIA Enzyme immunoassayELISA Enzyme-linked immunosorbent assayEMIT Enzyme-multiplied immunoassay techniqueEO Ethoxylene unitEPA Environmental Protection AgencyEQ EsterquatFA Fatty acidsFIA Fluorescence immunoassay

120 M.-C. Estévez et al.

FIIA Flow-injection immunoassayFOH Fatty alcoholsGC Gas chromatographyGC–MS Gas chromatography–mass spectrometryHPLC High-performance liquid chromatographyHRP Horseradish peroxidaseHTS High-throughput screeningIA ImmunoassayIAC Immunoaffinity chromatographyIC50 Concentration at 50% of signal inhibitionIgG Immunoglobulin GLAS Linear alkylbenzenesulfonatesLC Liquid chromatographyLC–MS Liquid chromatography–mass spectrometryLDS Linear 4-dodecylbenzenesulfonic acid sodium saltLIA Liposome immunoaggregationLIC Liposome immunocompetitionLIF Laser-induced fluorescenceLOD Limit of detectionMAb Monoclonal antibodyNP NonylphenolNPE Nonylphenol polyethoxylateOP OctylphenolOPE Octylphenol polyethoxylatePAb Polyclonal antibodyPBB Polybrominated biphenylPBDD Polybrominated dibenzo-p-dioxinPBDE Polybrominated diphenyl ethersPBDF Polybrominated dibenzofuransPCB Polychlorinated biphenylPCDD Polychlorinated dibenzo-p-dioxinPCDE Polychlorinated diphenyl ethersPCDF Polychlorinated dibenzofuransPCP PentachlorophenolPFIA Polarization fluoroimmunoassayQAC Quaternary ammonium compoundsQFIA Quenching fluorescence immunoassayRAb Recombinant antibodyRIA RadioimmunoassaySAS Secondary alkyl sulfonatesSDS Sodium dodecyl sulfateSPC Sulfophenyl carboxylateSTP Sewage treatment plantTBBPA Tetrabromobisphenol ATBP TribromophenolTCDD Tetrachlorodibenzo-p-dioxinTCP TrichlorophenolI-TEF Toxic equivalent factorI-TEQ Toxic equivalent quotientTtCP TetrachlorophenolWWTP Wastewater treatment plant

Immunochemical Determination of Industrial Emerging Pollutants 121

1Introduction

Chemicals and secondary by-products from industry, household chemicals,personal care products, and pharmaceuticals such as drugs, antibiotics, andhormones, are some of the substances that have been grouped under the ex-pression emerging pollutants, making reference to their recent increased useand release into the environment and to the fact that most of them were notpreviously considered as contaminants (see Table 1 for the most importantgroups of emerging pollutants). The adverse effects derived from a continu-ing exposure to these substances are often unknown, and regulations are stillnot well established [1, 2]. Much interest has been focused on these com-pounds not only because of their possible adverse effects, but also for the greatamounts that are produced worldwide. Most of the substances considered asemerging contaminants are widespread in everyday life and applied in differ-


Table 1 Some of the most important emerging pollutants divided into chemicals with an industrial origin and pharmaceuticals

Industrial chemicals Pharmaceuticals

Surfactants and their metabolites AntibioticsNonionic surfactants FluoroquinolonesAnionic surfactants SulfamidesCationic surfactants Penicillins

TetracyclinesOrganochlorinated substances Macrolides

Polychlorinated biphenylsChlorophenols Steroid hormonesDioxins Estrogens

AndrogensOrganobrominated substances Gestagens

Polybrominated biphenyl ethers CorticosteroidsBromophenolsDioxin-like compounds AnalgesicsTetrabromobisphenol A Paracetamol

AspirinIndustrial additives and others Ibuprofen

Phthalate esters DiclofenacBisphenol A

Tranquilizers (psychiatric drugs)Diazepam

Cytostatic agentsMethotrexateCyclophosphamideIfosfamide

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. Someof them, like the pesticides, arrive intentionally in the environment and theiruse and release should be theoretically controlled. However, many of them havenot been purposely produced as bioactive substances but more as componentsor additives of certain materials. Their significant growth in the chemical in-dustry has not only been produced as a consequence of the discovery of newactive principles in the pharmaceutical or pesticide area, but also because of theexpansion of new technologies (electronics, containers, textiles, plastics, resins,foams, etc.), that require the development of new materials and substances withparticular features. Most of these substances enter or are discharged to water andair sources without regulated controls. Wastewater treatment plants (WWTPs)are often not yet adapted to completely remove them, and therefore these newcompounds can be found to some extent in wastewater effluents as well as in soiland sludge.

The release into the environment of this large amount of chemicals has be-come an increasing concern for the authorities and for the scientific commu-nity. Although there is positive pressure by governmental bodies and agenciesto push chemical industries toward the development of more environmentallyfriendly compounds that are not persistent and are easily biodegradable, the final metabolites readily formed can be more ubiquitously distributed and/orpresent more toxicological effects than the parent compounds. That is the casefor alkylphenol polyethoxylates, a major group of nonionic surfactants usedworldwide, whose major breakdown products are the alkylphenols, which areconsidered relevant endocrine disrupters, especially nonylphenol. The same effects have been reported in the case of certain plastic and polymer additivessuch as bisphenol A or some phthalate esters (see BKH report [3]). Other kindsof substances are not produced intentionally but are generated as by-productsof industrial processes such as combustion or waste incineration. This is thecase for the dioxins or the PCBs. Other polyhalogenated compounds are also of concern because of their persistence in the environment. This is the case forthe chlorophenols, used for many years as insecticides, and wood and textilepreservatives. Their use is today restricted in most of the developed countriesbut residues can still be detected in many environmental compartments.Organo-brominated substances, used as flame retardant additives, have emerged as anew generation of polyhalogenated substances whose environmental and toxi-cological impacts are still not completely determined, although some evidencesuggests an endocrine disrupter action. Some of the substances considered inthis chapter do not have toxicity by themselves, but may affect the permeabilityor solubility of other pollutants present in the environment. This is the case forthe anionic surfactants such as LAS, whose presence in the environment is alsoa worry due to their high production and use.

The continuous development of more specific and sensitive analytical tech-niques has allowed the detection of traces of these substances in many com-


partments of the environment. Recently, the US Geological Survey reported thedevelopment of five new analytical methods for the detection of 95 organicwastewater 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 forthese substances,more and more rapid, sensitive,and specific methodologies ca-pable of detecting this huge amount of compounds are continuously demanded.Immunochemical techniques can fulfill all these requirements because of theirspecificity, selectivity, and demonstrated high sample throughput capabilities.

The objective of this chapter has been to collect and to provide informationon the immunochemical techniques available today for the determination of animportant number of the emerging pollutants of industrial origin. The emerg-ing pollutants considered here have been selected by attending to their toxico-logical risk, environmental relevance, or their regular use or production (theirgeneric structure and some reported data regarding their environmental levelsas well as their more probable biodegradation pathways are briefly detailed inTable 2). For many of these substances there are immunochemical methodsavailable that have been applied with significant success to the analysis ofenvironmental 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 immuno-chemical assay kits. The availability of antibodies, as the key reagents of any immunochemical technique, opens the possibility of developing a wide varietyof methods, such as those described before, depending on the necessities. How-ever, 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 im-munochemical techniques).

2Immunochemical Techniques

The key component of immunochemical techniques is the antibody (Ab). An-tibodies are globular proteins generated by the immune system as a defenseagainst foreign agents (antigens,Ag). Their structure varies depending on theirisotypes. There are five different families of immunoglobulins (IgG, IgM, IgA,IgD, and IgE) differing in their charge, size, amino acid sequence, and carbo-hydrates attached. The most abundant class in mammal serum and the mostused in immunochemical applications is the IgG subclass. The IgGs (MW150 kDa) are composed of four polypeptide chains: two identical “heavy” (H)chains that carry covalently attached oligosaccharide groups, and two identi-cal, nonglycosylated, “light” (L) chains. The heavy chains are interlinked bydisulfide bonds, and each light chain is joined by a disulfide bond to a heavy


Immunochemical Determination of Industrial Emerging Pollutants 125Ta

ble

2Su

mm

ariz

ed t

able

wit

h th

e m

ore

impo

rtan

t co

mpo

unds

incl

uded

in t

he s

urfa

ctan

ts,p

olyh

alog

enat

ed c

ompo

unds

,and

oth

er in

dust

rial

resi

dues

.The

ir g

ener

ic c

hem

ical

str

uctu

re a

nd th

e us

e or

ori

gin

are

show

n.So

me

repo

rted

dat

a re

gard

ing

thei

r en

viro

nmen

tal o

ccur

renc

e an

d th

em

ore

prob

ably

env

iron

men

tal f

ate

are

also

giv

en

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

No

nio

nic

Su

rfa

cta

nt

Alk

ylph

enol

Eth

oxyl

ates

(APE

s)

NPE

(non

ylph

enol

Hou

seho

ld a

nd–

STPs

eff

luen

ts in

Sw

itze

rlan

d:R

eadi

ly b

iode

grad

able

inet

hoxy

late

s)co

mm

erci

al d

eter

gent

sN

P1EO

(30

–65

mg L

–1),

NP2

EO

WW

TPs

und

er a

erob

ic a

ndO

PE (o

ctyl

phen

olEm

ulsi

fiers

(47–

77 m

g L–1

) and

in Ja

pan

anae

robi

c co

ndit

ions

[6] t

o et

hoxy

ates

)Te

xtile

and

leat

her

NP1

EO (

0.21

–2.9

6 mg

L–1

) [5]

shor

t cha

in a

lkyl

phen

ol

indu

stry

– SW

:usu

ally

bel

ow 1

mg L

–1et

hoxy

late

s an

d al

kylp

heno

lsPh

arm

aceu

tica

l and

and

max

imum

pea

k va

lues

H

alf-

life=

1 to

4 w

eeks

[7,8

]R

=C

8,C

9pe

rson

al c

are

prod

ucts

arou

nd 2

0mg

L–1

[5]

n=1–

40(P

PCPs

)

Alk

ylph

enol

s (A

Ps)

NP

(Non

ylph

enol

)M

ajor

com

pone

nt in

– Se

wag

e ef

fluen

ts:

Mor

e pe

rsis

tent

than

APE

.O

P (O

ctyl

phen

ol)

the

prod

ucti

on o

f0.

025–

330

mg L

–1(N

P)

Hal

f-lif

e (r

iver

wat

ers)

:30–

58A

PEs

and

brea

kdow

nan

d 0.

0022

–73

mg L

–1(O

P).

days

(NP)

and

7–5

0 da

ys

prod

uct i

n th

eir

Usu

ally

bel

ow 1

0mg

L–1

[9]

(OP)

[5]

degr

adat

ion.

– Fo

und

in a

ir [1

0] a

nd s

edi-

R=

C8,

C9

Plas

tici

zers

and

men

ts (

up to

140

00mg

Kg–1

) st

abili

zers

in p

last

ics

[5,1

1]–

SW:U

sual

ly N

P le

vels

bel

ow

1mg

L–1

.Hig

h le

vels

in S

pain

(0

.15–

644

mg L

–1) [

12] a

nd

Engl

and

rive

rs

(0.2

–180

mg L

–1) [

13].

– O

P le

vels

bel

ow 0

.1mg

L–1

[5]

DW

:dri

nkin

g w

ater

;GW

:gro

und

wat

er;S

W:s

urfa

ce w

ater

;WW

:was

te w

ater

;ST

P:se

wag

e tr

eatm

ent p

lant

;WW

TP:

was

tew

ater

trea

tmen

t pla

nt.


Tabl

e2

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Fatt

y A

lcoh

ol E

thox

ylat

es (A

E)

Hou

seho

ld a

nd–

In G

W le

vels

:(61

–189

ng

L–1R

eadi

ly b

iode

grad

able

(>80

%

laun

dry.

for

the

diff

eren

t AEO

s in

28

days

for

linea

r AE

and

R

=C

9–C

15Pu

lp a

nd p

aper

C12

EO3–

940

% fo

r br

anch

ed A

E) [1

8].

n=4–

14m

anuf

actu

ring

– To

tal c

onc.

of71

0 ng

L–1

[14]

Slow

er d

egra

dati

on (A

E>20

Te

xtile

dye

ing

– So

il in

ters

titi

al w

ater

:et

hoxy

lene

uni

ts) [

19]

Emul

sifie

rs,s

pray

(48–

73 n

g L–1

for

C12

EO3–

5 ad

juva

nts

(tot

al c

onc.

in th

e de

eper

la

yers

at 1

94 n

g L–1

) [14

]–

Trea

ted

sew

age

6.5

mg L

–1[1

5],

12.5

–300

mg

L–1[1

6]–

Sew

age

slud

ge

10–1

90 m

g K

g–1[1

7]

Fatt

y Es

ters

Pol

yeth

oxyl

ates

Emul

sifie

rsN

o da

ta fo

und

Easi

ly b

iode

grad

able

(sl

ower

Text

ile a

nd le

athe

rw

hen

it h

as m

ore

than

50

unit

sin

dust

ries

ofet

hyle

ne o

xide

uni

ts [2

0]R

=C

12–C

18C

ompo

nent

in P

PCPs

Fatt

y A

mid

es P

olye

thox

ylat

es

Foam

sta

biliz

atio

nN

o da

ta fo

und

Con

trad

icto

ry d

ata

rela

ted

toEm

ulsi

fier

they

bio

degr

adab

ility

[19]

.So

lubi

lizer

,ant

ista

tic,

Rea

dily

bio

degr

adab

le [2

0,21

]R

=C

12–C

18w

etti

ng a

gent

inPP

CPs

Hai

r sh

ampo

o,liq

uid

soap

s,sh

avin

g cr

eam

san

d ot

her

PPC

Ps


Tabl

e2

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

An

ion

ic S

urf

act

an

tsLi

near

Alk

ylbe

nzen

e Su

lfona

tes (

LAS)

Hou

seho

ld a

nd–

STPs

eff

luen

ts:1

9–71

mg L

–1 [1

5]R

eadi

ly d

egra

dabl

e w

ith

a co

mm

erci

al d

eter

gent

s–

WW

eff

luen

ts le

vels

:ha

lf-lif

e of

1–87

day

s.10

–35%

0.

09–0

.9 m

g L–1

[18]

adso

rbed

in th

e pa

rtic

ulat

e–

WW

slu

dge:

<3

mg g

–1[1

8]ar

e m

atte

r [1

8]–

SW in

Nor

th S

ea

(<0.

05–9

.4mg

L–1

) [22

] m

+n=

C10

–C14

– SW

in B

razi

l (14

–155

mg L

–1[2

3]–

SW in

Phi

lippi

nes

(2.2

–102

mg L

–1) [

24]

Sulfo

phen

yl C

arbo

xyla

tes (

SPC

s)

Maj

or m

etab

olite

in–

SW a

nd s

ewag

e ef

fluen

ts:

Foun

d m

ainl

y in

aqu

atic

LAS

biod

egra

dati

on0.

5–3.

2mg

L–1

[25]

com

part

men

ts.

– Se

awat

er:[

26]

Shor

ter

alky

l cha

in S

PCs

– D

rink

ing

wat

er:

(≤C

5) a

re e

xpec

ted

to b

e 1.

6–3.

7mg

L–1

[23]

foun

d as

the

dist

ance

from

–

Raw

riv

er w

ater

s:th

e di

scha

rge

poin

t ofL

AS

m+

n=C

3–C

111.

8–5

mg L

–1[2

7]in

crea

ses

– U

p an

d do

wns

trea

m o

fW

WT

Ps:<

1–10

1 mg

L–1

[28]


Tabl

e2

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Alk

yl S

ulfo

nate

s

Prim

ary

Alk

yl

Liqu

id d

eter

gent

sN

o da

ta fo

und

Can

be

adso

rbed

ont

o sl

udge

su

lfona

tes

R=

C11

–C17

(dis

h w

ashi

ng a

gent

s,bu

t und

er a

erob

ic c

ondi

tion

s cl

eani

ng a

gent

s,an

dar

e re

adily

bio

degr

adab

le

hair

sha

mpo

os).

(pri

mar

y de

grad

atio

n in

Se

cond

ary

Alk

ylC

omm

erci

al p

rodu

cts

WW

TP

<90

% in

3 d

ays

[18]

).Su

lfona

tes

(SA

S)ar

e al

mos

t exc

lusi

vely

Not

deg

rade

d in

ano

xic

com

pose

d of

SAS

cond

itio

nsR

+ R

1=C

12–C

18

Alk

yl S

ulfa

tes (

AS)

Laun

dry

dete

rgen

ts–

STPs

eff

luen

ts:C

12–1

5 A

S Fa

st b

iode

grad

atio

n un

der

Woo

l-w

ashi

ng a

gent

s,be

twee

n 1.

2 an

d 12

mg L

–1[1

5]ae

robi

c an

d an

aero

bic

R=

C12

–C18

soap

bar

s an

d liq

uid

cond

itio

ns.

bath

soa

ps,h

air

Effe

ctiv

e re

mov

al in

WW

TPs

sham

poos

,and

[18]

toot

hpas

tes

Alk

yl E

ther

Sul

fate

s (A

ES)

Liqu

id b

ath

soap

s,–

Efflu

ents

ofs

even

rep

rese

n-R

eadi

ly b

iode

grad

able

inha

ir s

ham

poos

,and

tati

ve S

TPs

:C12

–15

AES

:W

WT

Ps u

nder

bot

h ae

robi

c R

=C

10–C

14m

echa

nica

l3

and

12mg

L–1

[15]

an

d an

aero

bic

cond

itio

ns

m=

1–4

dish

was

hing

age

nts.

[18,

29]

Ingr

edie

nt in

indu

stri

al c

lean

ing

agen

ts


Tabl

e2

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Ca

tio

nic

Su

rfa

cta

nts

Qua

tern

ary

Am

mon

ium

Com

poun

ds (Q

AC

s)

ATM

AC

:Alk

yl-

Fabr

ic s

ofte

ner

Valu

es fo

und

for

Hig

hly

adso

rbed

ont

otr

imet

hyl-

amm

oniu

mEm

ulsi

fyin

g ag

ents

dita

llow

dim

ethy

lam

mon

ium

pa

rtic

ulat

e m

atte

r [1

8].

com

poun

dsBi

ocid

es,d

isin

fect

ants

chlo

ride

(DT

DM

AC

):Sh

ort h

alf-

life

unde

r ae

robi

cD

AD

MA

C:D

ialk

yl-

– ST

Ps-i

nflu

ents

:375

–430

0mg

L–1

cond

itio

ns.P

oorl

y an

aero

-di

met

hyl-

amm

oniu

mR

¢=M

ethy

l or

– In

SW

ofr

iver

s:2–

34mg

L–1

bica

lly b

iode

grad

ed [1

9]co

mpo

unds

benz

yl–

Efflu

ent o

fST

Ps:1

1–55

mg L

–1

AD

MBA

C:A

lkyl

-X

=C

l or

Br(r

evie

wed

in [

30])

dim

ethy

lben

zyl-

(or

Met

hyl s

ulfa

te)

– W

WT

Ps in

fluen

ts:

amm

oniu

m

340–

480

mg L

–1(U

S);

com

poun

ds≈

1000

mg L

–1[3

1]

Qua

tern

ary

Car

boxy

alky

l Am

mon

ium

Com

poun

ds (E

ster

quat

s)

Fabr

ic s

ofte

ner

No

data

foun

dEa

sily

bio

degr

aded

und

erae

robi

c co

ndit

ions

.It’s

als

oas

sum

ed th

eir

degr

adat

ion

inan

oxic

con

diti

ons

[19,

32]

R=

C16

–C18

Org

an

och

lori

na

ted

Su

bst

an

ces

Poly

chlo

rina

ted

biph

enyl

s (PC

Bs)

Insu

lati

ng fl

uid

in–

SW (u

p to

100

–500

ng

L–1)

Very

per

sist

ent i

n th

eel

ectr

ical

equ

ipm

ent

[33,

34]

envi

ronm

ent.

and

as h

ydra

ulic

flui

dsLo

w le

vels

in w

ater

and

air

.m

+n

=1–

10H

igh

leve

ls in

soi

ls,

sedi

men

ts a

nd a

nim

als


Tabl

e2

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Dio

xin

Like

Sub

stan

ces

Poly

chlo

rina

ted

– M

arin

e se

dim

ents

(Fin

land

):Ve

ry p

ersi

sten

t in

the

diph

enyl

ethe

rs (P

CD

E)<

1 pg

I-T

EQ g

–1dw

[35

,36]

envi

ronm

ent.

Very

low

leve

ls a

re fo

und

inw

ater

and

air

.Hig

h le

vels

inm

+n

=1–

10so

ils,s

edim

ents

and

ani

mal

s.Po

lych

lori

nate

d A

re n

ot in

tent

iona

lly–

Sedi

men

ts fr

om p

ollu

ted

area

s:O

ccur

renc

e in

the

envi

ron-

dibe

nzo-

prod

uced

.The

y ar

e>

20 p

g I-

TEQ

g–1

dw [3

7]m

ent r

elat

ed to

chl

orin

atio

n p-

diox

in (P

CD

D)

form

ed a

s by

prod

ucts

– Ef

fluen

ts fr

om w

aste

degr

ee o

btai

ned

by p

hoto

lysi

s:

m+

n=1–

8in

sev

eral

pro

cess

esin

cine

rato

r pl

ants

(Ja

pan)

:>

chlo

rina

tion

deg

ree>

3.3–

120,

000

pg L

–1.(

73.5

TEQ

) pe

rsis

tenc

ePo

lych

lori

nate

d(r

evie

wed

in [3

8])

dibe

nzof

uran

(PC

DF)

– So

il:0.

1–10

80 p

g I-

TEQ

g–1

and

sedi

men

ts:0

.42–

8 pg

I-T

EQ g

–1

m+

n=1–

8(S

pain

) [39

,40]

Chl

orop

heno

ls

Pres

erva

tive

s ag

ents

,–

GW

nea

r sa

wm

ills

or w

aste

M

ost o

fthe

m g

o in

to w

ater

.pe

stic

ides

site

s fr

om (0

.03

to 9

1.3

mg L

–1)

The

y ca

n be

tran

sfor

med

[4

1]by

pho

toly

sis

into

dio

xins

–

Dri

nkin

g w

ater

:[4

8,49

]m

=1–

50.

03–0

.7mg

L–1

[42–

44]

– SW

(sev

eral

cou

ntri

es):

1.6–

26.6

mg L

–1[4

5,46

]–

Sedi

men

ts (C

anad

a)

25–1

0,00

0mg

Kg–1

[47]


Tabl

e2

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Org

an

ob

rom

ina

ted

Su

bst

an

ces

Poly

brom

inat

ed b

iphe

nyls

(PBB

)

Flam

e re

tard

ant

– W

ater

Pin

e ri

ver

(USA

):C

ompo

unds

hig

hly

0.01

–3.2

mg L

–1 [5

0,51

]br

omin

ated

att

ach

stro

ngly

–

Sedi

men

ts:0

.33–

0.84

(dw

) to

sed

imen

ts.T

hey

are

slow

ly

m+

n=1–

10mg

Kg–1

[52]

degr

aded

in th

e en

viro

nmen

t –

Wat

er:<

0.05

mg L

–1an

d [5

3,54

]se

dim

ents

:<8

ng g

–1(J

apan

) in

198

9 [3

8]

Dio

xin

Like

Sub

stan

ces

Poly

brom

inat

edFl

ame

reta

rdan

t–

Sew

age

slud

ge:

Com

poun

ds h

ighl

y br

omi-

diph

enyl

ethe

rs

11–2

8 ng

g–1

(Br 1

0) [5

5]na

ted

atta

ch s

tron

gly

to

(PBD

E)–

Wat

er (

Japa

n) in

198

8:se

dim

ents

.The

y ar

e sl

owly

<0.

1mg

L–1

[38]

and

sed

imen

ts

degr

aded

in th

e en

viro

nmen

tm

+n

=1–

10(J

apan

) in

1996

<25

–580

ng

g–1

(Br 1

0) [3

8]–

SW:0

.158

ng

L–1;s

ewag

e sl

udge

:229

0–48

90 n

g g–1

;sed

i-

men

ts:1

32 n

g g–1

(rev

iew

ed

in [5

6])

– So

il (U

S) <

0.1–

31.6

ng

g–1[5

7]


Tabl

e2

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Dio

xin

Like

Sub

stan

ces

Poly

brom

inat

ed

Are

not

inte

ntio

nally

– Se

wag

e sl

udge

from

mun

icip

al

Very

per

sist

ent i

n th

edi

benz

o-p-

diox

inpr

oduc

ed.T

hey

are

WW

TPs

(G

erm

any)

:en

viro

nmen

t(P

BDD

)fo

rmed

as

bypr

oduc

ts0.

29–3

.05

mg K

g–1 (B

r 1–B

r 5) [

58]

Fast

er d

egra

dati

on th

an

m+

n=1–

8in

sev

eral

pro

cess

es–

Sedi

men

ts (

Japa

n):

PCD

Ds

[59]

.0.

03–0

.37

ng g

–1(B

r 4–B

r 6) [

38]

Very

low

leve

ls a

re fo

und

inPo

lybr

omin

ated

– R

iver

and

mar

ine

sedi

men

ts:

wat

er a

nd a

ir.H

igh

leve

ls in

dibe

nzof

uran

(PBD

F)0.

03–0

.37

mg K

g–1so

ils,s

edim

ents

and

ani

mal

s.(B

r 4–B

r 6) [

60]

Occ

urre

nce

in th

e en

viro

n-m

+n=

1+8

– Se

dim

ents

1.9

8–17

.4 n

g g–1

[38]

men

t rel

ated

to b

rom

inat

ion

degr

ee o

btai

ned

by p

hoto

-ly

sis:

>br

omin

atio

n de

gree

>pe

rsis

tenc

e

Brom

ophe

nols

Flam

e re

tard

ant

– R

aw a

nd tr

eate

d w

ater

T

hey

can

be d

ehal

ogen

ated

inW

ood

pres

erva

tive

from

pot

able

wat

er tr

eatm

ent

mar

ine

and

rive

r se

dim

ents

pl

ants

(0.6

and

1.3

ng

L–1) [

61]

[62,

63]

– Se

dim

ents

(Ja

pan)

1986

:m

=1–

5<

0.5–

4 ng

g–1

[38]

Tetr

abro

mob

isph

enol

A (

TBBP

A)

Flam

e R

etar

dant

– Se

wag

e sl

udge

T

BBPA

can

be

met

abol

ized

by

(2.9

–76

ng g

–1dw

) [55

]m

icro

bes

and

brok

en d

own

by–

Sedi

men

ts (

Japa

n) 1

988:

phot

olys

is [6

4]2–

108

ng g

–1[3

8]


Tabl

e2

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Oth

ers

Phth

alat

e Es

ters

Di-

n-bu

tyl

Plas

tici

zers

in P

VC

– SW

leve

ls a

re n

ear

to 1

0mg

L–1

;Fa

st b

iode

grad

atio

n un

der

phth

alat

e (D

BP)

prod

ucti

on.

in r

iver

s,be

twee

n 0.

5–1

mg L

–1ae

robi

c co

ndit

ions

.D

ieth

yl p

htha

late

C

ompo

nent

in th

ean

d in

sea

wat

er b

etw

een

Hal

f-lif

e in

wat

er:1

–15

days

(DEP

)m

anuf

actu

re o

f0.

005–

0.7

mg L

–1H

alf-

life

in s

oils

:7 d

ays

–Bu

tyl b

enzy

l co

smet

ics,

inks

,and

– U

S st

ream

s:2.

5mg

L–1

(DEH

P)

seve

ral m

onth

s [6

5]ph

thal

ate

(BBP

) ad

hesi

ves

and

0.25

mg

L–1(D

EP) [

4]D

i(2-

ethy

lhex

yl)

phth

alat

e (D

EHP)

Bisp

heno

l APr

oduc

tion

ofr

esin

s–

Riv

er w

ater

mea

n va

lues

:N

ot p

ersi

sten

t in

surf

ace

wat

er.

(pol

ycar

bona

te a

nd0.

016

mg L

–1(E

urop

e) a

nd

Rap

idly

bio

degr

aded

in a

quat

icep

oxy

resi

ns).

0.5

mg L

–1(U

S) [

66].

envi

ronm

ents

[68

] an

d re

mov

edC

ompo

nent

in fl

ame

– SW

:<0.

001–

1mg

L–1

[9]

in W

WT

P.re

tard

ant p

rodu

ctio

n–

WW

eff

luen

ts m

ean

valu

es:

Hal

f-lif

e:1–

4 da

ys [6

9] in

wat

er.

Ant

ioxi

dant

,1.

5mg

L–1

[67]

Acc

umul

ated

in a

noxi

cpr

eser

vati

vese

dim

ents

[9]


Tabl

e3

Som

e re

pres

enta

tive

com

mer

cial

imm

unoc

hem

ical

ass

ay k

its

for

the

mos

t im

port

ant e

mer

ging

pol

luta

nts

wit

h an

indu

stri

al o

rigi

n.T

he

supp

lier

and

the

cont

act w

eb p

age

are

also

list

ed

Ana

lyte

Imm

unoc

hem

ical

kit

Supp

lier/

man

ufac

ture

rC

onta

ct

No

nio

nic

Su

rfa

cta

nts

APE

sEL

ISA

kit

Japa

n En

viro

Che

mic

als,

Ltd.

http

://w

ww

.jech

em.c

o.jp

/eco

/A

EEL

ISA

kit

Take

da C

hem

ical

Indu

stri

es L

-EC

http

://w

ww

.take

da.c

o.jp

/ind

ex-e

.htm

lA

PEL

ISA

kit

Japa

n En

viro

Che

mic

als,

Ltd.

http

://w

ww

.jech

em.c

o.jp

/eco

/

Ani

onic

Sur

fact

ants

LAS

ELIS

A k

itJa

pan

Envi

roC

hem

ical

s,Lt

d.ht

tp://

ww

w.je

chem

.co.

jp/e

co/

Org

anoc

hlor

inat

ed su

bsta

nces

PCBs

PCB

RIS

c so

il te

st k

itEn

Sys,

Inc.

(Str

ateg

ic D

iagn

osti

c In

c.)

http

://w

ww

.sdi

x.co

m/

PCB

RIS

C li

quid

was

te te

st s

yste

mEn

Sys,

Inc.

(Str

ateg

ic D

iagn

osti

c In

c.)

http

://w

ww

.sdi

x.co

m/

PCB

imm

unoa

ssay

kit

Hac

hht

tp://

ww

w.h

ach.

com

/En

viro

Gar

d PC

B in

soi

lM

illip

ore

Inc.

(Str

ateg

ic D

iagn

osti

c In

c.)

http

://w

ww

.sdi

x.co

m/

PCB

in s

oil (

tube

ass

ay)

Envi

roLo

gix

http

://w

ww

.env

irol

ogix

.com

/D

TEC

H®

PCB

test

kit

Stra

tegi

c D

iagn

osti

c In

c.ht

tp://

ww

w.s

dix.

com

/D

ELFI

A P

CB

(soi

l and

food

)H

ybri

zym

eht

tp://

ww

w.h

ybri

zym

e.co

m/

PCB

RaP

ID A

ssay

®O

hmic

ron

Cor

p.(S

trat

egic

Dia

gnos

tics

Inc.

)ht

tp://

ww

w.s

dix.

com

/PC

Bs E

LISA

kit

,100

T

Take

da C

hem

ical

Indu

stri

es L

-EC

http

://w

ww

.take

da.c

o.jp

/ind

ex-e

.htm

l(m

agne

tic

part

icle

)


Tabl

e3

(con

tinu

ed)

Ana

lyte

Imm

unoc

hem

ical

kit

Supp

lier/

man

ufac

ture

rC

onta

ct

Org

anoc

hlor

inat

ed su

bsta

nces

PCP

PEN

TA R

ISc®

EnSy

s In

c.(S

trat

egic

Dia

gnos

tics

Inc.

)ht

tp://

ww

w.s

dix.

com

/En

viro

Gar

dTM

in s

oil

Mill

ipor

e (S

trat

egic

Dia

gnos

tics

Inc)

http

://w

ww

.sdi

x.co

m/

D T

ECH

® PC

P Te

stSt

rate

gic

Dia

gnos

tics

Inc.

http

://w

ww

.sdi

x.co

m/

PCP

RaP

ID A

ssay

®O

hmic

ron

Cor

p.(S

trat

egic

Dia

gnos

tics

Inc.

)ht

tp://

ww

w.s

dix.

com

/

Dio

xins

EnSy

s D

ioxi

n En

Sys

Inc.

(Str

ateg

ic D

iagn

osti

cs In

c.)

http

://w

ww

.sdi

x.co

m/

Hig

h-pe

rfor

man

ce d

ioxi

n/fu

ran

CA

PE T

echn

olog

ies

http

://w

ww

.cap

e-te

ch.c

om/

imm

unoa

ssay

kit

inse

rt (I

N-D

F1)

DEL

FIA

TC

DD

test

kit

Hyb

rizy

me

http

://w

ww

.hyb

rizy

me.

com

/

Oth

ers

Bisp

heno

l AEL

ISA

kit

Japa

n En

viro

Che

mic

als,

Ltd.

http

://w

ww

.jech

em.c

o.jp

/eco

/EL

ISA

kit

Japa

n En

viro

Che

mic

als,

Ltd.

http

://w

ww

.jech

em.c

o.jp

/eco

/

chain. All four chains contain defined constant and variable, locally hyper-variable, regions within their amino acid sequence (see Fig. 1). The Fc fragmentis the constant (crystallized) region and it is involved in the immune regulation,whereas the Fab (antibody binding fraction) fragment is the region that con-tains the variable fraction (Fv) with the specific binding sites that allow inter-action with the Ag.

Since the publication of the first assay based on the use of antibodies to de-termine insulin [70], development in this area has undergone rapid growth. Theapplication of immunochemical methods to detect substances not only in theclinical field, but also in food and environmental areas has led to the necessityto obtain more and more specific antibodies in considerable amounts and atlow cost. The important progress made during recent years in the molecular biology and genetic engineering fields has favored this fact. Currently, we canspeak of three different techniques to obtain antibodies, yielding what we knowas polyclonal (PAb), monoclonal (MAb), and recombinant (RAb) antibodies. Inprinciple it is possible to obtain antibodies for any kind of substance. In thecase of small molecules (i.e., molecular weight less than 1,000 Da), the designand 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 theeffort necessary in this aspect.

Polyclonal antibodies are obtained directly from the serum of the immu-nized animals (sometimes a purification step is carried out before their use).A family of clones is obtained that recognize the global structure of the haptenimmunized, 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). Itis formed by two pairs of polypeptide chains interlinked by disulfide bonds. The Fc fragmentis the constant (crystallized) region and it is involved in the immune regulation, whereas theFab (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 obtainedafter papain or pepsin digestion are also shown

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 sheephas been described. The process to finally obtain PAbs is simple once you havethe immunizing hapten but because of the animal variability, a lack of repro-ducibility can be found from animal to animal. This fact can be a problem whena constant supply of identical antisera is required.

Monoclonal antibodies are produced by the fusion of antibody-producingspleen cells from an immunized animal (mouse) with mutant tumor cells derived from myelomas [72]. In contrast with PAbs, a unique IgG molecule isobtained from a single cell clone and theoretically this technology provides anunlimited source of the antibody with identical affinity for the antigen, as longas the hybridoma line is stable. However, the screening process to isolate the desired clone is usually tedious and time-consuming, the cost of production ofMAbs is higher than for PAbs, and sometimes they have lower affinities to smallmolecules 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 technologythese problems can be in part solved by the generation of a variety of Ab frag-ments mimicking the immune response in vitro. The whole process involves the following steps: (a) the preparation of Ab encoding libraries derived fromdifferent methods (usually by isolation of mRNA from hybridoma, spleen cells,or lymphocytes of immunized mice); (b) cloning of the genes in a bacterialplasmid vector; (c) expression in bacteria (E. coli) and coinfection with helperbacteriophage virus, displaying Ab fragments on its surface as a fusion withnormally occurring coat protein; and (d) screening for antigen specificity andantigen-driven selection [73, 74]. The ability of this methodology to design theantibody polypeptidic structure, as well as to modify the existing fragments,can allow one to improve the affinity of the antibodies or even to change theirselectivity. Although this technology was developed initially for therapeuticpurposes, RAb fragments have also been used in environmental analysis. Re-combinant antibodies have been produced for insecticides like parathion [75],for dioxins [76], and for pesticides like triazines [77, 78], although at presentthey have not achieved the affinity levels of the corresponding MAbs or PAbs.

2.1Antibody-Based Analytical Methods

Immunochemical techniques are based on the immunological reaction derivedfrom the binding of the antibody to the corresponding antigen. This reactionis reversible and is stabilized by electrostatic forces, hydrogen bonds, and Vander Waals interactions. The formed complex has an affinity constant (ka) thatcan achieve values around the order of 1010 M–1. This great affinity and speci-ficity between the specific antibody and the antigen (or the analyte) haveturned these techniques into powerful analytical tools to detect and quantify


substances at low concentrations and trace levels. During recent decades manyefforts have been made and important advances have been achieved in thisfield. The advances made in microelectronics combined with the availability ofantibodies for a great variety of substances such as proteins, macromolecules,or low molecular weight molecules such as drugs, metabolites, or environ-mental, agricultural and food pollutants, have been crucial for this progress.

Despite the robustness and good detectability achieved nowadays by thechromatographic analytical methods when coupled to sensitive detectors, theyoften require expert personnel and expensive equipment. Moreover, precon-centration and cleanup procedures to remove potential interferences prior tothe analysis are usually mandatory. All these factors lead to an increase of thefinal cost of these methodologies and the analysis time. Alternatively, im-munochemical techniques are simple, fast, and very specific and sensitive.Nowadays automation and the possibility of development of high-throughputscreening (HTS) have been demonstrated. Overall we can say that they consti-tute excellent tools to be exploited in monitoring programs where a great num-ber of samples need to be analyzed. One of their drawbacks is often the fact thatmatrix effects should be carefully evaluated beforehand since, in contrast toother analytical techniques, specific and nonspecific signals are not so easy todistinguish leading to overestimation or to false positives. Contrariwise, falsenegatives are very seldom seen in these techniques. Thus, as effective screen-ing techniques, immunochemical methods are complementary to the standardanalytical techniques.

Several immunochemical techniques have been developed as analytical toolsor in sample treatment methods to separate an analyte from complex matrices.Some of the most important or more frequently used are described below.

2.1.1Immunoassays

Nowadays, immunoassay (IA) is the most extensive immunochemical method-ology.A great number of IAs have been developed for the detection of pollutantsat trace levels [71, 79–81], such as pesticides and other kinds of industrialresidues, not only in different environmental compartments (water, soils, sedi-ments, etc.) but also in food and biological matrices. Many of those developed forpesticides are commercially available and the US EPA (US Environmental Pro-tection Agency) has validated and included some of them in the SW-846 methodlist [82].Wide application has also been found in pharmaceutical, veterinary, andforensic analysis as well as, more recently, in human exposure assessment to a variety of industrial chemicals or contaminants such as polyaromatic hydrocar-bons (PAHs), polychlorinated biphenyls, (PCBs), or pesticides [83–85]. For hu-man biomonitoring, where large a number of samples are often analyzed, thepossibility of adapting immunoassays to HTS makes them particularly suited tofield studies or large-scale monitoring. In addition, the ability to recognize notonly the target analyte but also other structurally related compounds (Ab cross-


reactivity) may allow both parent molecules and their metabolites to be detectedsimultaneously. In this context, the EPA encourages the development of im-munochemical techniques for human exposure monitoring [84, 85].

In immunoassays the reaction Ab–Ag is quantified by means of labels, un-der competitive conditions. The general procedure involves a competition stepbetween a fixed concentration of a labeled derivative and the free analyte fora limited amount (low concentration) of Ab. The amount of labeled Ag can thenbe measured and therefore the amount of free Ag. Several kinds of markers canbe used as labels. The first immunoassay was based on the use of radioisotopes(radioimmunoassays, RIA) [86], but they have been replaced by more environ-mentally friendly and less hazardous substances. By using fluorescent (fluores-cein, rhodamine, etc.), chemiluminescent (i.e., luminol), or bioluminescentmarkers, techniques such as fluoroimmunoassay (FIA) and chemiluminescenceimmunoassay (CLIA) have been developed, although in FIAs, for instance, thesensitivity achieved is in many cases not as good as expected, sometimes be-cause fluorophores are exposed to many interferences that can lead to quench-ing of the signal. The use of enzyme labels (EIA, enzyme immunoassay) offersthe possibility of increasing detectability, due to the amplification produced depending on the enzyme turnover, and the option of using a variety of sub-strates producing colored, fluorescent,or chemiluminescent products.Nowadaysenzymes such as horseradish peroxidase (HRP), alkaline phosphatase (AP), andglucose oxidase (GO) are the most frequently used labels in immunoassay.

Immunoassays can be performed in solution (homogeneous format) or byimmobilization of one of the immunoreagents on a solid support (heteroge-neous format). The solid support can be tubes, nitrocellulose paper, magneticparticles, microspheres, polystyrene plates, etc. The most used supports are mi-crotiter plates where up to 96 (or in some cases up to 384) samples can beprocessed simultaneously, making use of very small sample volumes. In hetero-geneous 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. Homoge-neous assays are faster, simpler, and can be easily adapted to the available auto-mated analyzers often used in clinical chemistry. However, they are often lesssensitive and are more exposed to matrix interferences, so washing steps to helpremove these interferences must be performed. The common aspect of all theseassays when applied to the analysis of small organic molecules is the fact that theassay takes place under competitive configurations, as we will see below. In con-trast, for the determination of large substances, this is not always necessary.

2.1.1.1Enzyme-Linked Immunosorbent Assay (ELISA)

Among the heterogeneous assays, ELISA is the most common and frequentlyused for environmental monitoring. Examples of its wide applicability can befound in recent reviews [71, 79]. The most usual configurations for the analy-



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 anda 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 Indirectcompetitive ELISA.A coating antigen is immobilized on the solid support and the specific IgGand the analyte are in solution. After removal of unbound reagents, a secondary IgG labeledwith the enzyme (IgG-enzyme), which specifically recognizes the Ab, is added.After anotherwashing step the amount bound is also quantified by the addition of the substrate solution

a

b

sis of small molecules are shown in Fig. 2. In the direct format (see Fig. 2a), theAb is coated onto the solid support (usually a microtiter plate) and an equilib-rium 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 reagentsare 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 directassay can also be performed by immobilizing an analog of the analyte (coatingantigen) on the plate and performing the competition step with the free ana-lyte 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 mea-sured 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 provedto be more robust.

2.1.1.2Enzyme-Multiplied Immunoassay Technique (EMIT)

EMIT is one of the most common EIAs working under homogeneous condi-tions [87]. The principle is the competition for the specific antibody betweenthe analyte and an analog labeled with a particular enzyme (usually glucose-6-phosphate dehydrogenase, G6P-DH) such that the enzyme activity decreasesupon binding of the labeled antigen to the antibody. In this format the analyte

Direct Competitive ELISA

Indirect Competitive ELISA

concentration is directly proportional to the enzyme activity measured. Thiskind 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 pesti-cides have also been reported [93].

2.1.1.3Polarization Fluoroimmunoassay (PFIA)

PFIA also works under homogeneous conditions and makes use of fluorescentlabels (usually fluorescein). The principle is the excitation of the sample withplane 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 theabsence of the analyte, the light measured is thus very small since a great partof the labeled antigen will be bound to the antibody. The presence of the ana-lyte 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 screen-ing methodology [94–96], but also for environmental pollutants such as pesti-cides [97–99].

2.1.2Flow-Injection Immunoassay (FIIA)

In recent years the combination of immunochemical methodologies with au-tomated devices has grown in importance. In this context, flow-injection sys-tems coupled to immunoassays (FIIA) [100–102] offer rapid analysis of a greatnumber of samples, providing rapid results and good levels of sensitivity, andallowing continuous monitoring. In these devices, a small immunoreactor con-tinuously receives buffer or reagents through different valves. The techniquecan work in homogeneous (usually with fluorophores as label) or heteroge-neous conditions (usually with the Ab immobilized). The more generic FIIAworks as is shown in Fig. 3. By means of a buffer valve all the solutions are con-tinuously flowing through the system and the immunoreagents (Ag and labeledAg, both together or in sequential steps) are injected in the system. The amountof labeled Ag is detected downstream and quantified. Subsequently, the systemcan be regenerated by passing buffer solutions. Several FIIAs have been devel-oped for environmental pollutants such as pesticides and also industrial chem-icals, as will be discussed in following sections [100, 103–106].



Fig. 3 Generic FIIA system.A heterogeneous format is shown. The antibodies are immobi-lized 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 substratesolution through the system allows the determination of the amount of bound labeled Ag,which is then detected and measured

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

2.1.3Immunosensors

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


immunosensors the Ab or the Ag is immobilized on the surface of the transducerand the formation of the immunocomplex is detected by the transducer.

Although there are many approaches for direct immunosensors (the changesgenerated when the antibody binds the analyte), in the case of small organicmolecules the detection usually takes place under competitive configurationswith or without the use of labels. Several transducer principles have been de-scribed. Electrochemical immunosensors are based on the use of amperometric,potentiometric, conductimetric, or impedimetric transducers. Optical immuno-sensors make use of optical fibers, the evanescent wave, or the surface plasmonresonance principle, among others. Piezoelectric immunosensors are based onthe shift of the resonance frequency of piezoelectric crystals produced after formation of the immunocomplex. Finally some examples can also be found ofthermometric immunosensors, where the heat of the reaction produced as a con-sequence of the immunoreaction, usually coupled to an enzyme amplificationsystem, is detected. When the use of labels is required, the nature of these de-pends on the transducer principle of the immunosensor. Thus, the use of elec-troactive, fluorescent, or mass labels among others has been described. Extensiveinformation on the transducer principles, methods, and examples of all thesetypes of immunosensors can be found in recent reviews [107–110].Although theuse of biosensors has been mainly reported for clinical purposes [111], nowadaystheir application has been extended to different areas, for example the detectionof microorganisms such as viruses and bacteria [112], the detection of drugs,and also in the control of environmental pollutants [102, 113, 114].

2.1.4Immunoaffinity Chromatography (IAC)

The application of immunochemical techniques is not only in the developmentof analytical detection tools, but also the inherent specificity of the immunore-action can be exploited to develop selective extraction procedures to be usedprior to the analysis by any other analytical method (either immunochemical ora conventional one, like chromatography). Immunoaffinity chromatography(IAC) [115–117] for trace analysis of low molecular weight analytes in complexmatrices has several advantages over other solid phases.Apart from selectivity,the antibody cross-reactivity allows the extraction of both the analyte and itsmetabolites or other structurally related compounds. Preconcentration of theanalyte may assist in increasing the detectability of certain analytical methods.The essential hydrophilic media necessary when handling biomolecules allowthe purification of polar substances, which can be hindered when other con-ventional 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 thatallows removal of the nonspecifically retained compounds, the target analyteis eluted under appropriate conditions (usually by changing the buffer com-position). Finally, the column can be regenerated and reused several times.

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

3Immunochemical Methods for Surfactants

The use of surfactants in detergent formulations was extended worldwide sev-eral decades ago. More environmentally friendly compounds with surfactantproperties have gradually replaced the natural soaps. These new substances arecharacterized by the presence in the chemical structure of both hydrophilic(usually charged) and hydrophobic groups (particularly long linear alkylchains). This fact imparts unique properties to these compounds as surface-active agents. Depending on the charge of the hydrophilic part of the moleculefour clear groups can be distinguished: anionic, cationic, nonionic, and am-photeric surfactants. The production volume of these compounds has increasedconsiderably since they have been used as substitutes for soap-based detergents.Thus, in 30 years (between 1940 and 1970) the production of synthetic surfac-tants 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 theEuropean Committee of Surfactants and their Organic Intermediates (CESIO),the total production in western Europe in 2000 was 2.5¥106 metric tons, withanionic and nonionic surfactants being the most abundant ones at productionvolumes 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.


Fig. 5 Sequential steps involved in an immunoaffinity extraction procedure

Contrary to anionic and nonionic agents, they have poor detergency and areused more in the preparation of germicides, fabric softeners, and emulsifiers.Amphoteric surfactants are produced in much smaller amounts (5¥104 metrictons, near to 2% of the total production) [125]; they are biodegradable and theirecotoxicological importance can be considered low. Their environmental oc-currence up to know has been just occasional.

Liquid chromatography coupled to mass spectrometry (LC–MS) is the mostusual technique applied for the detection of anionic [126–128], nonionic


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


Tabl

e4

Imm

unoc

hem

ical

tech

niqu

es d

evel

oped

for t

he d

etec

tion

ofs

urfa

ctan

ts.T

he se

nsit

ivit

y of

the

met

hod

and

the

mat

rix

cons

ider

ed a

re sh

own

Ana

lyte

Imm

unoc

hem

ical

Se

nsit

ivit

yM

atri

xR

efer

ence

ste

chni

quea

LOD

IC50

Non

ioni

c su

rfac

tant

sA

lkyl

phen

ol e

thox

ylat

es (A

PEs)

NP1

0EO

ELIS

A (d

irec

t)8.

9mg

L–1Bu

ffer

/wat

er[1

44]

FIIA

2.4

mgL–1

24.5

mgL–1

Buff

er[1

44,1

45]

OP1

0EO

FIIA

0.5

mgL–1

17.7

mgL–1

Buff

er[1

45]

NPE

CIA

-GD

H b

iose

nsor

378

mgL–1

Buff

er[1

46]

ELIS

A10

4mg

L–1Bu

ffer

[146

]A

utom

ated

BM

P-IA

6.6

mgL–1

Buff

er[1

47]

OPE

CIA

-GD

H b

iose

nsor

605

mgL–1

Buff

er[1

46]

ELIS

A42

mgL–1

Buff

er[1

46]

aC

IA-G

DH

bio

sens

or:c

apill

ary

imm

unoa

ssay

cou

pled

to

a gl

ucos

e de

hydr

ogen

ase

bios

enso

r;EL

ISA

:enz

yme-

linke

d im

mun

osor

bent

ass

ay;

BMP-

IA:b

acte

rial

mag

neti

c pa

rtic

le-b

ased

imm

unoa

ssay

;FII

A:f

low

-inj

ecti

on im

mun

oass

ay;P

FIA

:pol

ariz

atio

n flu

oroi

mm

unoa

ssay

.b

BDD

12A

C:b

enzy

ldim

ethy

ldod

ecyl

amm

oniu

m c

hlor

ide;

DD

AC

:did

ecyl

dim

ethy

lam

mon

ium

chl

orid

e.


Tabl

e4

(con

tinu

ed)

Ana

lyte

Imm

unoc

hem

ical

Se

nsit

ivit

yM

atri

xR

efer

ence

ste

chni

quea

LOD

IC50

Non

ioni

c su

rfac

tant

sA

lkyl

phen

ols (

AP)

NP

ELIS

A10

mgL–1

Buff

er[1

48,1

49]

ELIS

A76

mgL–1

Buff

er[1

44]

ELIS

A59

0mg

L–1Bu

ffer

[144

]FI

IA52

mgL–1

1033

mgL–1

Buff

er[1

44,1

45]

PFIA

7.9

mg

L–142

mg

L–1Bu

ffer

[144

]PF

IA9

mg

L–142

mg

L–1Bu

ffer

[150

]EL

ISA

769

mgL–1

Buff

er[1

46]

CIA

-GD

H b

iose

nsor

4481

mgL–1

Buff

er[1

46]

PFIA

9m

gL–1

42m

gL–1

Buff

er[1

50]

OP

ELIS

A34

6mg

L–1Bu

ffer

[146

]C

IA-G

DH

bio

sens

or15

60mg

L–1Bu

ffer

[146

]

Ani

onic

sur

fact

ants

Line

ar a

lkyl

benz

ene

sulfo

nate

s (LA

S)

ELIS

A20

mgL–1

40mg

L–1Bu

ffer

[151

]EL

ISA

19.8

mgL–1

Buff

er[1

44]

FIIA

19.5

mgL–1

387

mgL–1

Spik

ed r

ain/

wat

er

[144

]PF

IABu

ffer

[144

]PF

IA0.

5m

gL–1

9m

gL–1

Buff

er[1

52]

Aut

omat

ed

35ng

L–1Bu

ffer

[147

]BM

P-ba

sed

IA

Cat

ioni

c su

rfac

tant

sQ

uate

rnar

y am

mon

ium

com

poun

ds (Q

AC

s)b

BDD

12A

CEL

ISA

0.04

3m

gL–1

0.66

mg

L–1Bu

ffer

,(m

ilk)

[153

]

DD

AC

ELIS

A8

mg

L–129

mg

L–1Bu

ffer

[154

]


[129–138], and cationic surfactants [139], achieving good levels of sensitivity.Gas chromatography also coupled to mass spectrometry (GC–MS) has beenused more sporadically [140–142]. As is well known, these methodologies canachieve very low limits of detection (LOD), sometimes in the range of micro-grams per liter and even nanograms per liter. However, the analytical protocolsemployed 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 proceduresare often tedious to ensure the efficient recovery of analytes from water samples[16, 143]. Moreover, usually they are not single substances but technical mixturesof several compounds and isomers (see Fig. 6 for chemical structures). Im-munochemical techniques offer in this case the advantage that the analyses aregenerally 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 4summarizes most of the immunochemical techniques that have been devel-oped for these compounds and will be briefly commented on in the next fewsections.

3.1Anionic Surfactants

Linear alkylbenzenesulfonates (LAS) represent more than 40% of all surfactantsused [18] with a production of nearly 8.5¥105 metric tons. Other anionic agentssuch as alkyl sulfonates (AS), alkyl ether sulfates (AES), or fatty alcohol sulfates(FAS) are also produced, with the same final applications but lower consump-tion (see Fig. 6 for chemical structures).All of them are easily biodegradable andefficiently removed by the WWTPs (efficiency near 99%) [28], but because oftheir large consumption they end up reaching rivers and marine environments.Both LAS and their metabolites, the sulfophenyl carboxylates (SPCs), are usuallyfound in surface waters at levels around a few micrograms per liter, although thislevel increases in areas where sewage effluents are not connected to municipalWWTPs [23–25, 27]. They have also been detected in seawater [26] and drink-ing 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 for-mation of micelles helps to dissolve and transport many nonpolar organic pol-lutants, preventing their degradation and enhancing the toxic action of othersubstances. Thus, synergistic effects are also observed with pesticides (i.e., DDTand dieldrin) and heavy metals like Cd or Hg [155].

Although several immunochemical methods have been reported for LAS, fewexamples of their application to real environmental matrices have appeared. Thefirst immunochemical method for LAS was reported by Fujita et al. [151]. It isa direct ELISA and uses MAbs generated against 5-sulfophenyl valeric acid con-jugated to BSA through the carboxylic acid, thus preserving the sulfonic group

intact. The LOD achieved is around 20 mg L–1 and the working range between20 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, andstereate), and other surfactants such as nonylphenol polyethoxylates or short-chain 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-reactin the assay; however, these compounds were not evaluated. The assay was evaluated in terms of precision and accuracy by measuring spiked river watersamples. Recoveries between 81 and 100% were obtained. Similarly the assaywas also validated with HPLC.

Franek et al. [144] also used several sulfophenyl carboxylic acids with dif-ferent chain lengths as immunizing haptens to produce a large number of poly-clonal 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 hap-ten conjugated to b-galactosidase. The sensitivity reported is quite good and inthe 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 acidsodium salt (LDS) instead of to the commercial mixture, which does not givea real picture of the detectability of these methods. The flow format allowed theanalysis of ten samples per hour. Matrix effects were also studied using spikedsurface and rainwater samples. The latter produced nonspecific interferencesin the assay whereas surface water could be directly measured without prob-lems. For PFIA [152], several antibodies and fluorescein thiocarbamyl ethyl-enediamine (EDF) tracers were screened but the sensitivity was worse. The bestcombination afforded a LOD of 0.5 mg L–1 and a dynamic range between 3 and85 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, thistechnique allows higher sample loads since about seven samples can be ana-lyzed in 10 min. SDS only cross-reacts about 5%.

Recently Matsunaga et al. [147] developed an automated immunoassay sys-tem based on the immobilization of monoclonal antibodies [151] to magneticparticles, synthesized by magnetic bacteria (Ab-BMP). The assay is performedin microtiter plates mounted in the reaction station. This kind of immunoassayallows the suspension and subsequent easy separation of antibody from the restof the immunoreagents via a magnet coupled to the tips used in the automatedpipette. The detection limit obtained in the competitive assays for LAS is verygood (35 ng L–1), similar to those obtained by GC–MS or LC–MS, and shows awide working range (between 35 ng L–1 and 35 mg L–1). The assay showed lowcross-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. Theantibodies are highly specific for LAS with alkyl chains between C8 and C12,whereas other anionic and nonionic surfactants tested showed no cross-reac-tivity.


3.2Nonionic Surfactants

During recent decades the nonionic surfactants used most worldwide havebeen the alkylphenol polyethoxylates (APEs). They are widely used in detergentformulations, both for industrial and domestic applications. They are also usedas plasticizers and stabilizers in plastics. About 5¥105 tons are produced an-nually [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 environ-ment. In the WWTPs they are readily biodegraded under both aerobic andanaerobic 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 toform alkylphenol ethoxy carboxylates (APECs). The final breakdown productsare APEs with one or two ethoxylate units (AP1EO, AP2EO), APECs, and AP(see Fig. 6 for chemical structures).All these metabolites have lost their surfac-tant properties and are much more persistent in the aquatic environment [5, 18].Several studies have indicated that whereas the parent compounds seem to pre-sent low toxicity in organisms [19, 158], APE metabolites, particularly APs(both NP and OP) and short-chain APEs, are highly toxic and their estrogenicactivity is remarkable [3, 159–163].

The most polar APECs are mainly detected in wastewater, effluents, andrivers, whereas APs, which are more lipophilic, tend to be adsorbed onto soil,sediments, and sludge. The distribution and fate of this family of compoundshave been widely reported during recent years [4, 5, 9, 164–168]. The levels ofthese pollutants in the environment may exceed the predicted no effect con-centration (PNEC of 0.33 mg L–1) [165] proposed in a risk assessment reportof the EU. As a consequence, their use for household and industrial cleaninghas been banned in several countries of the EU, being substituted by less toxicand more environmentally friendly nonionic surfactants such as alcoholethoxylates (AEs) [156]. In the USA these evidences have also led to regulatoryactions related to the domestic use of APEs. It seems that AEs are rapidlybiodegraded, although it depends on the nature of the alkyl chain, mainly theirbranching degree (the degradation decreases when the branching of the chainincreases) [19].

Immunochemical methods have been reported for both APEs and theirmetabolites,especially APs.A discussion of the immunochemical methodologiesreported to date, the effect of the immunizing haptens employed, and the fea-tures of these techniques were recently reviewed [169]. Unfortunately, the de-tectability achieved is usually far from what is necessary for direct applicationto environmental samples. Moreover, the selectivity for APs versus APEs is notalways satisfactory. Thus, Goda et al. [148] developed a direct ELISA for NP witha LOD of 10 mg L–1 and a working range between 70 and 1,000 mg L–1, but APEswith one to ten ethoxylate units are also well recognized.


MAbs for AP and APEs have been produced and commercialized by TakedaChemical Industries [170] (see Table 3) and have opened up the opportunity todevelop a great variety of assays. FIIAs and ELISAs for NP and NP10EO havebeen developed with these antibodies [144]. The detectability achieved is sim-ilar 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 NPit 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 byBadea et al., who studied the influence of different factors such as the presenceof organic solvents or heavy metals in the assay media, and its performance inseveral water matrices (tap water, surface water, and wastewater) [145]. TheLOD reported for NP was 51 mg L–1 and about 2.5 mg L–1 for NP10EO, althoughshort NPEO were also highly recognized. The method was applied to the de-termination of these surfactants in the influent and effluent of WWTPs.

The same MAbs have been immobilized on magnetic particles and used todevelop an automated immunoassay method for APEs (NPEs with 10–20 EOunits) [147]. The wide dynamic range reported (between 6.6 mg L–1 and66 mg L–1) is related to a low slope assay, which has a direct negative influenceon the immunoassay precision. NPEs with two ethoxylate units and NP are alsorecognized 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 andOP) and APEs (NPE and OPE) [146]. The competitive step takes place off-linein an antibody-coated plastic capillary and once it has been washed, it is inte-grated 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 theother formats, APEs are better recognized than APs.

Franek et al. have also produced polyclonal antibodies for NP [144]. Dif-ferent antibodies were raised using various para-alkyl hydroxyphenyl com-pounds with different alkyl chain lengths as immunizing haptens. An indirectELISA and PFIA have been developed with these antibodies. The detectabilityaccomplished is between 1 and 2 orders of magnitude higher for the ELISA for-mat. While the IC50 of the ELISA was 590 mg L–1, the limit of detection of thePFIA was around 8 mg L–1. As happened with the LAS PFIA method, the sam-ple throughput is very high but it is necessary to develop compatible sampleconcentration methods in order to apply this technique for environmentalmonitoring purposes. Subsequently, Yakovleva et al. [150] also reported the development of PFIA for NP, obtaining sensitivities in the same order of mag-nitude (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 phenoliccompounds, no APE or other related surfactants were tested, which would havebeen interesting in order to determine the capability of the method to dis-criminate between NP and its parent compounds.


3.3Cationic Surfactants

The most important cationic surfactants are those derived from quaternaryammonium salts where one or two hydrophobic groups (usually a long alkylchain, 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 ben-zyl group. The most used salts in the formulation of commercial products arequaternary ammonium compounds (QACs). Figure 6 shows a generic structureof different QACs and their abbreviated names. Alkylquats can be consideredthose structures where the hydrophobic alkyl chain is directly linked to the nitrogen atom (e.g., alkyltrimethylammonium compounds (ATMAC), dialkyl-dimethylammonium compounds (DADMAC) or alkyldimethylbenzylammo-nium compounds (ADMBAC)), whereas in the esterquats (EQ) (or diesterquats,DEQ) the hydrophobic group is linked through an ester bond.As happens withLAS or APEs, for their preparation several raw substances are obtained fromnatural oils and therefore the commercial product is often constituted of a mix-ture of different alkyl chain lengths. The properties of these surfactants includenot only their surfactant activity but also their biocide effect. They are activeprinciples in several household products such as fabric softeners or hair condi-tioners, and are also in disinfectants, biocides, emulsifiers, wetting agents, andprocessing additives.After use they are discharged to STPs or directly to surfacewaters.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]. Itseems that alkylquats are poorly degraded under anoxic conditions since thepresence 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 bio-degradation 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 comparedto other surfactants, but their poor solubility and their tendency to adsorb to solids or to complex with anionic substances considerably reduce the realrisk and adverse effects for the aquatic environment. [30, 31, 176]. The use ofalkylquats has been substituted by the more easily biodegradable and lesstoxic esterquats that are nowadays the cationic surfactants produced in highervolumes.

A competitive indirect ELISA was developed [154], using PAbs raised againstan immunizing hapten that preserved both long alkyl chains and with onemethyl group substituted by the spacer arm. A LOD of 8 mg L–1 was achievedwhen didecyldimethylammonium chloride (DDAC) was used as analyte (seeFig. 6 for structure). The specificity of the assay was evaluated by testing short-chain QACs (with methyl or ethyl groups) that were poorly or not recognized,demonstrating that the methyl group or the charged nitrogen atoms are not themain epitopes. Fatty acids (FA) or fatty alcohols (FOH) with long alkyl chains


were also tested, and high levels of cross-reactivity were observed for thosewith alkyl chain lengths between 10 and 12 carbon units. Shorter and longerchains were less recognized, which indicates that the main epitope is the decylchain. A better detectability has been reported on another ELISA for benzyl-dimethyldodecylammonium chloride (BDD12AC) [153], a component of thebenzalkonium chloride (BAK) which is a mixture of three alkyldimethylben-zylammonium 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 PAbswere raised using as immunizing hapten an analog of the analyte where amethyl group was substituted by the spacer arm. The LOD accomplished was43 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 benzyldimethy-lalkylammonium compounds (BDACs, with C6–C16 alkyl chains) showed cross-reactivity values between 42 and 106%. This ELISA has been validated by HPLCusing spiked samples and commercial products.

4Immunochemical Methods for Polychlorinated and Polybrominated Compounds

During recent decades much concern has been focused on the adverse healtheffects of organochlorinated substances. This group involves several families of compounds (chlorophenols, CP; polychlorinated biphenyls, PCB; poly-chlorinated dibenzodioxins, PCDD; polychlorinated diphenyl ethers, PCDE;polychlorinated dibenzofurans, PCDF etc.) known to be highly persistent andin some cases with a clearly proven toxicity in organisms. Recently, much moreattention is also being paid to organobrominated compounds because of theirgrowing use and production. Analog families of the organochlorinated sub-stances 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 differentdegree of halogen substitution (either Cl or Br), resulting in chemical com-pounds or commercial products containing mixtures [53]. The structures of themost common polybrominated and polychlorinated substances are shown inFig. 7.

Although the use of organochlorinated substances has been abolished inmost of the developed countries, their extensive use during the past decadesand their persistence in the environment determine their actual widespreaddistribution. Moreover, some substances are not used or commercialized butare still important synthetic intermediates for the preparation of other sub-stances. Thus, chlorophenols are important intermediates in the production ofpesticides or other chemicals. Other substances such as PCDEs or PCDDs are



Fig. 7 Generic chemical structures of polyhalogenated compounds. X=Cl, Br. (I) Poly-chlorinated biphenyls (PCBs), polybrominated biphenyls (PBBs); (II) chlorophenols (CPs),bromophenols (BPs); (III) polychlorinated diphenyl ethers (PCDE), polybrominated diphenylethers (PBDE); (IV) polychlorinated dibenzo-p-dioxin (PCDD), polybrominated dibenzo-p-dioxin (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 beformed by chemical, photochemical, or thermal reactions from precursors[177]. Chlorophenols are also unintentionally formed when water with a highcontent of organic material is disinfected with chlorine or during wood pulpbleaching processes. PCBs are used as lubricants, fire retardants, immersionoils, 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 substanceshave been extensively reviewed, although there are still questions and contro-versies on the effects after long time exposures. Moreover, the toxicity variesgreatly between families and congeners (for reviews see [180–184]). Thus, thegreatest concern exists around PCDDs, followed by PCDFs, PCDEs, and PCBs.From the PCDD family, tetrachlorodibenzo-p-dioxin (TCCD) has been identifiedas the most toxic congener [182]. Data regarding human dioxin exposure havebeen associated with an increased risk of severe skin lesions such as chloracneand hyperpigmentation, altered liver function and lipid metabolism, generalweakness associated with drastic weight loss, depression of the immune system,

and endocrine- and nervous-system abnormalities. TCDD is considered a potentcarcinogenic, teratogenic,and fetotoxic chemical in certain animals.Human pop-ulations occupationally or accidentally exposed to chemicals contaminated withdioxin have increased incidences of soft-tissue sarcoma and non-Hodgkin’s lym-phoma. On the other hand, some DDT, PCDD, PCDF, and PCB congeners havebeen classified as high-concern potential endocrine disrupting substances in thelist of the 66 priority substances according to the BKH report [3].

The use of organobrominated substances has shown extraordinary growthduring 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 aretetrabromobisphenol A, PBB, PBDE (penta-, octa-, deca-brominated diphenylether (oxide) formulations), and hexabromocyclododecane. The production,application, and potential environmental occurrence of the most importantBFRs 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 overthe same period) [186, 187]. Their impact on wildlife and the environment hasbeen reviewed by authors in different countries [38, 56, 57, 188–191].Although animportant amount of information exists on the potential risk of the organochlo-rinated substances, more studies have to be done regarding the toxicity of thepolybrominated 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 brominatedanalogs could be very similar, but the results are still very much dependent onthe assay used. Thus, toxic effects, including teratogenicity, carcinogenicity, andneurotoxicity, have been observed for some BFR congeners, in particular thePBDEs (for recent reviews see [194, 195]). The endocrine-disrupter activity ofthe BFRs is being investigated. Hence, PBBs have also been classified as of highconcern in the BKH report [3]; on the other hand, some evidence has been pro-vided on the disruption of the thyroid hormone system by BFRs, with partic-ular emphasis on the PBDEs [196].

The application of immunoassays to the analysis of organohalogenatedcompounds has not been as frequent as for more water-soluble species [197].Since immunoassays are typically aqueous-based systems, the low water solu-bility of these compounds makes the use of immunochemical techniques morechallenging since several facts have to be considered, starting from a careful se-lection of the immunizing hapten to aspects such as the preparation of sampleextracts or handling of the standards to the last stages in the assay optimiza-tion. 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 analysisof organochlorinated emerging pollutants. Conversely, to our knowledge noimmunochemical techniques have been reported for the determination of


organobrominated substances, although it has been reported that antibodiesraised against chlorophenols have the ability to recognize the correspondingbrominated analogs even more [198, 199]. Moreover, preliminary experimentsperformed in our group have shown that the application of the 2,4,6-tri-chlorophenol immunoassay to the analysis of 2,4,6-tribromophenol from woodextracts 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).

4.1PCBs

The first assays for PCBs date from the 1980s [228, 229] and since then severalother attempts have also been carried out (reviewed in [169, 230]). The de-scribed assays are usually based on the analysis of PCBs such as Aroclors, notas congeners, and use PAbs produced using an Aroclor or a single PCB as a hap-ten. RIAs have been developed and used to detect them in biological matricessuch as milk or blood [83, 207, 231], whereas other kinds of immunoassays havebeen reported to detect them in environmental samples such as water, soils, andsediments [201–203, 232, 233]. The detectability reported depends on the hap-ten used to raise antibodies and also on the congener used to calibrate the as-say. Simple extraction methods rendering extracts compatible with the aque-ous media of the immunochemical methods are sometimes reported in orderto 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].


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]

a b


Tabl

e5

Imm

unoc

hem

ical

tech

niqu

es d

evel

oped

for

the

dete

ctio

n of

orga

noch

lori

nate

d su

bsta

nces

Ana

lyte

aIm

mun

oche

mic

al

Sens

itiv

ity

Mat

rix

Ref

eren

ces

tech

niqu

eb

LOD

IC50

PCBs

Aro

clor

124

8EL

ISA

1.34

mgL–1

22mg

L–1So

il,se

dim

ents

[201

](8

.95

ngg–1

)EL

ISA

1,00

0mg

L–1So

il[2

02]

Aro

clor

124

2EL

ISA

1.57

mgL–1

25mg

L–1So

il,se

dim

ents

[201

](1

0.5

ngg–1

)EL

ISA

5–12

.9mg

L–1So

il[2

03]

FOI

10mg

L–1R

iver

wat

er,s

oil

[204

]

Aro

clor

125

4EI

A m

agne

tic

part

icle

0.2

mgL–1

Wat

er[2

05]

500

mgK

g–1So

il[2

05]

RIA

2mg

L–1Bl

ood

[83]

20mg

L–1M

ilk[8

3]EL

ISA

25mg

L–121

7mg

L–1Bu

ffer

[206

]

Aro

clor

126

0EL

ISA

38mg

L–121

2mg

L–1Bu

ffer

[206

]R

IA0.

013

mg

kg–1

Milk

[207

]

aD

CP:

dich

loro

phen

ol;P

CP:

pent

achl

orop

heno

l;2,

3,7,

8-TC

DD

:2,3

,7,8

-tet

rach

loro

dibe

nzo-

p-di

oxin

;TC

P:tr

ichl

orop

heno

l;T

MD

D:2

,3,7

-tri

chlo

ro-

8-m

ethy

ldib

enzo

-p-d

ioxi

n.b

EIA

mag

neti

c pa

rtic

le:e

nzym

e im

mun

oass

ay b

ased

on

mag

neti

c pa

rtic

les;

ELIS

A:e

nzym

e-lin

ked

imm

unos

orbe

nt a

ssay

;FO

I:fib

er o

ptic

imm

uno-

sens

or;L

IA:l

ipos

ome

imm

unoa

ggre

gati

on a

ssay

;LIC

:lip

osom

e im

mun

ocom

peti

tion

;LIF

-mic

rodr

ople

ts-Q

FIA

:las

er-i

nduc

ed fl

uore

scen

ce d

e-te

ctio

nin

mic

rodr

ople

ts w

ith

quen

chin

g- fl

uore

scen

ce im

mun

oass

ay;R

IA:r

adio

imm

unoa

ssay

.


Tabl

e5

(con

tinu

ed)

Ana

lyte

aIm

mun

oche

mic

al

Sens

itiv

ity

Mat

rix

Ref

eren

ces

tech

niqu

eb

LOD

IC50

PCBs

Aro

clor

s 12

42,1

248,

FIIA

1mg

L–1Bu

ffer

[208

]12

54,1

260

2-ch

loro

biph

enyl

LIA

0.26

pmol

Buff

er[2

09]

LIC

0.4

nmol

Buff

er[2

09]

3,4,

3¢,4

¢-tet

rach

loro

-EL

ISA

0.2

mgL–1

0.9

mgL–1

Buff

er[2

10]

bisp

heny

l

Dio

xins

PC

DD

s/PC

DFs

2,3,

7,8-

TCD

DEL

ISA

4mg

L–1Bu

ffer

[211

]EL

ISA

200

pg p

er w

ell

Buff

er[2

12]

ELIS

A1

ng p

er w

ell

Buff

er[2

13]

ELIS

A0.

1mg

L–1So

il[2

14]

0.02

5mg

L–1W

ater

[214

]EL

ISA

19.5

pg p

er tu

beFl

y as

h[2

15]

ELIS

A10

.4ng

mL–1

Buff

er[7

6]

TM

DD

ELIS

A0.

01mg

L–10.

24mg

L–1Bu

ffer

[216

]EL

ISA

0.00

4mg

L–10.

036

mgL–1

Buff

er[2

17]

ELIS

A16

ngm

L–1Bu

ffer

[218

]


Tabl

e5

(con

tinu

ed)

Ana

lyte

aIm

mun

oche

mic

al

Sens

itiv

ity

Mat

rix

Ref

eren

ces

tech

niqu

eb

LOD

IC50

Chl

orop

heno

ls

PCP

ELIS

A0.

1mg

L–12.

9mg

L–1W

ater

[219

]EI

A m

agne

tic

par

ticl

es0.

06mg

L–12.

2mg

L–1Bu

ffer

[220

]

ELIS

A50

0mg

L–1So

il[2

21]

ELIS

A30

–40

mgL–1

Buff

er

[222

]

2,4,

6-TC

PEL

ISA

0.2

mgL–1

1.53

mgL–1

Buff

er[1

98]

ELIS

A0.

2mg

L–11.

44mg

L–1Bu

ffer

[199

]EL

ISA

0.2

mgL–1

2.76

mgL–1

Buff

er

[223

]LI

F-m

icro

-0.

04mg

L–10.

45mg

L–1Bu

ffer

[224

]dr

ople

ts-Q

FIA

1.6

mgL–1

Uri

neEL

ISA

0.2

mgL–1

7.8

mgL–1

Wat

er[2

25]

2,4,

5-TC

PEL

ISA

0.05

3mg

L–10.

23mg

L–1Bu

ffer

[226

]EL

ISA

0.07

mgL–1

0.32

mgL–1

Wat

er[2

27]

0.26

mgL–1

1.28

mgL–1

Uri

ne[2

27]

0.8

mgL–1

6.46

mgL–1

Seru

m[2

27]

2,4-

DC

PEL

ISA

2mg

L–1Bu

ffer

[148

]

Two liposome-based immunomigration techniques have been developed for PCBs based on a test-strip format for on-site analysis (see Fig. 8). The li-posome immunocompetition (LIC) assay format measures the competitivereaction between analyte-tagged liposomes and the sample analyte for im-mobilized antibodies and can detect 0.4 nmol of PCB in less than 8 min. Theliposome immunoaggregation (LIA) assay is based on the principle of im-munoaggregation between anti-PCB antibodies and analyte-tagged lipo-somes, and detects the inhibition of immunospecific liposome aggregation insolution produced by the presence of the analyte. This last assay can detect2.6 pmol of PCB in less than 23 min. Both formats utilize capillary action totransport liposome-containing solutions along strips of nitrocellulose. Mea-surement of color intensity is then carried out visually or with a desktop scan-ner [209, 234].

A continuous semiautomated FIIA system has been reported [208, 235]. Inthis device the analyte-containing medium is allowed to flow through a columncontaining the antibodies immobilized on a support. First, the antibodies aresaturated with a fluorescent dye-labeled analog of the analyte. As the analytepasses through the immunosorbent, some dye-labeled antigen is displaced andis then detected in a fluorometer located downstream from the column. TheLOD 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 toxiccongeners. The ELISA developed is highly selective for PCB 77 and 126, show-ing IC50 values of 0.9 and 1.2 mg L–1, respectively. Noncoplanar PCBs, PCDDs,PCDFs, or single-ring halogenated compounds, including chlorinated benzenesand phenols, do not interfere with this assay. Johnson et al. [201] also producedPAbs for PCBs and an indirect ELISA has been optimized for the detection ofseveral Aroclors, giving LODs of about 1.5 mg L–1 that correspond to 9 ng g–1 insoils and a dynamic range between 50 and 1,333 ng g–1. Potential contaminantsusually found in soil samples such as chlorophenols and chlorobenzenes werealso tested, exhibiting cross-reactivities lower than 3%. A competitive enzymeimmunoassay for the quantification of PCBs in water has been developed us-ing PAbs covalently attached to amine-terminated superparamagnetic particlesas 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 and500 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 and1,000 mg L–1 has been achieved for the assay in methanol and allows analysis ofdiluted oils containing >35 mg mL–1 PCBs (neat). The procedure requires a pre-vious pretreatment of oil samples (either a solid-phase extraction or washingwith KOH-EtOH/sulfuric acid to remove interference).



There are some other firms that also commercialize immunoassay kits forPCBs based on different formats (see Table 3). They have been mainly validatedfor soil and wipe matrices, although the manufacturer can provide optionalprotocols for testing sediment or water samples.

A fiber optic immunosensor (FOI) has also been reported for detection ofPCBs in Aroclors [204]. The quartz fiber surface is coated with PAbs againstPCBs and the competitive assay takes place using as fluorescent tracer, an ana-log 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) methodhas been developed with the aim of improving the analytical methodology ofPCBs [236, 237]. The IAC column prepared using PAbs generated against thecoplanar toxic PCB congeners reduces the cleanup steps, time of analysis, andthe costs because of the ability to selectively retain PCBs. Moreover, this sam-ple treatment method reduces the use of hazardous organic solvents.

4.2PCDDs and PCDFs

Several attempts have been made to set up immunochemical techniques fordioxin analysis (reviewed in [230, 238, 239]). Frequently the detectability andselectivity accomplished have not been considered appropriate for the directanalysis of environmental samples. We should notice that due to the poor solubility of PCDDs and PCDFs in water, the levels of these contaminants inaqueous samples is very low. For this reason analysts usually prefer the use ofchromatographic and spectrometric methods that perform using organic sol-vents. However, the speed and high sample throughput that can be accom-plished with the immunochemical methods have prompted several researchgroups and companies to establish immunochemical methods.

The first IA for dioxins was a RIA developed by Albro et al. [240]. It was quitetime-consuming and utilized PAbs showing low specificity. MAbs were devel-oped 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 congeneris 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 optimizedassay had an IC50 around 200 pg per well [212]. Langley et al. [213] also reporteda PAb-based ELISA with an IC50 value of 1 ng of TCDD per well. Several yearslater 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 pgper tube and 25 pg per well. Sanborn et al. [218] reported the production ofPAbs using haptens containing an unsaturation in the spacer arm (between thehalogenated dibenzo-p-dioxin ring system and the protein to which it is con-jugated), which provides a rigid handle structure. The chemical structure of thehapten was similar to that of TCDD (i.e., 2,3,7,8- or 1,2,3,7,8-) but the polar

groups for hydrogen bonding were lacking. In the haptens synthesized, onesubstituent 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 ofthe 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) wasused as analytical surrogate standard in order to avoid using the toxic con-gener. 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 severalmodifications, 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-methyl-propenoic acid coupled to BSA) or the content of DMSO in the assay (up to37%). The ELISA exhibited an IC50 value of 12 pg per well (240 pg mL–1), witha working range from 2 to 240 pg per well (40 to 4,800 pg mL–1) [216]. Subse-quently, the use of a new coating antigen for the indirect assay was optimizedto finally reach a LOD of 4 ng L–1 [217]. Correlation with GC–MS was carriedout, achieving good agreements for soil samples without the necessity for anyadditional cleanup step prior to the analysis.

Intensive work in this field has continued in order to improve detectabilityand to establish reliable immunochemical protocols, including appropriatesample treatment methods, for the analysis of PCDDs and PCDFs in real sam-ples [214, 238, 243–246]. The development of selective extraction procedures aswell as solvent exchange methods to finally achieve immunoassay-compatibleextracts with solvents has been critical. Thus, an immunoaffinity extraction/cleanup protocol was developed by Harrison et al. [243] and then used on anassay employing an improved MAb (DF1). Handling of standards and sampleshas also been simplified by the use of detergent keeper in the solvent exchangeprocedure. Based on an assumption of quantitative recovery, a tenfold concen-tration of the original sample was accomplished before the immunochemicalmeasurement. The final analytical method (sample treatment plus immuno-chemical 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 anti-bodies against dioxins [76] using two hybridoma cell lines (DD1 And DD3)[211]. The option of cloning Fab fragments rather than scFv was chosen becausethey are usually more stable, which is important when environmental analysisof complex matrices is going to be carried out. Using 2,3,7,8-TCDD as standardan indirect ELISA has been developed. As in the aforementioned work, con-siderations regarding the use of organic solvent (MeOH) have been taken intoaccount and an assay with IC50 values around 10.4–14 ng mL–1 was achieved de-pending on the Fab used. These sensitivities and the results of cross-reactivitystudies are very similar to those obtained with the respective MAbs, showingtheir 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, hasbeen made available for the analysis of various types of sample extracts (fly ash,



soil, stack gas, tissue, sediment, water, etc.) after the application of conventionalextraction methods, followed by a solvent exchange step to provide hydro-al-coholic (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 ana-lyze 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 and2,3,7,8-TCDD) was also examined [254].

It is also worth mentioning that some authors have tried to demonstrate acorrelation 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 quo-tient, 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 valueof a sample by multiplying each mass concentration obtained by GC–MS by thecorresponding immunoassay cross-reactivity value [245]. Harrison and Carl-son [247] used their immunoassay to correlate PCDD/F congener recognitionprofiles with the congener toxicity in order to estimate the TEQ of real samples[243, 245].

4.3Chlorophenols

Among all chlorophenols, 2,4,6-trichlorophenol (TCP) and pentachlorophenol(PCP) are listed as priority pollutants by the US Environmental ProtectionAgency (EPA) (IRIS electronic database) and the EU [256]. In particular, PCPhas been classified as a B2 probable carcinogen for humans from animal toxi-city studies and human clinical data.

There have been various attempts to develop immunoassays for chlorophe-nols in environmental samples (soil, water). Noguera et al. [219] produced PAbsfor PCP. By using pentachlorophenoxypropionic acid as immunizing hapten (so that the five chlorine atoms are kept intact in the molecule) a direct ELISAwas developed with a LOD in water of 0.1 mg L–1 and a working range between0.3 and 30.5 mg L–1. The immunoassay is very specific for PCP since only oneof the compounds tested (2,3,5,6-tetrachlorophenol) shows a significant degreeof 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 minimalsample pretreatment for river and lake samples.

A competitive heterogeneous immunoassay based on the use of magneticparticles as solid support has also been developed using commercial PAbs forPCP [220]. Good sensitivities have been achieved with this format (the LOD isabout 60 mg L–1) and it shows low cross-reactivity with related compounds(only for tetrachlorophenols, TtCPs, is the degree of recognition significantwith CR of 54% for 2,3,5,6-TtCP and 15% for 2,3,4,6-TtCP). The assay allows the

analysis of water samples (e.g., river water, pond water, or groundwater) with-out sample pretreatment as well as quantification of soil samples including asimple extraction procedure. It provides results in 60 min and can be easilyadapted to on-site monitoring. The IA has also been compared with GC–MSand HPLC, obtaining in both cases good correlation levels for spiked water andsoil samples.

Many studies have been carried out in recent years in order to detect PCP inenvironmental samples, and most of them are based on the use of commerciallyavailable 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 byWestinghouse Bio-Analytic Systems (WBAS) have been evaluated for on-sitescreening tests of PCP in soils or aqueous matrices such as surface, drinking, orground water [221, 222, 257]. PCP RaPID-Assay was evaluated using certifiedwastewater samples, soil samples, and certified reference materials [258, 259].Acritical comparison of the ELISA method with an online liquid–solid extraction(LSE) method followed by liquid chromatography (LC-UV or LC-MS) analysiswas performed.A good correlation was found between both methods, althoughfor some samples undesirable matrix effects were also observed.

PAbs have been developed against 2,4,6- and 2,4,5-TCP after careful study ofthe chemical structure of these substance using computer-assisted molecularmodeling 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 im-munizing hapten) coupled to horseradish peroxidase as tracer. The microtiterplate 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 to56 mS cm–1) and pH values (between 5.5 and 9.5). The indirect ELISA formathas a similar detectability but it has proven to be more robust to matrix inter-ferences. This ELISA uses a heterologous hapten coupled to BSA as coatingantigen and it can be performed in 1.5 h. As with other microtiter plate ELISAmethods, the advantage is that many samples can be processed simultaneously.Both direct and indirect assays show a similar pattern of selectivity. Thus, theindirect ELISA [198] recognizes to a much lesser extent other chlorinated phe-nols, 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 forthe 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 for2,4,6-TCP [224]. The principle is the detection of laser-induced fluorescence(LIF) in single microdroplets by a homogeneous quenching fluorescence im-munoassay (QFIA). The competitive immunoassay occurs in microdroplets(d=58.4 mm) produced by a piezoelectric generator system. A continuous Arion laser (488 nm) excites the fluorescent tracer and its fluorescence is detectedby a spectrometer attached to a cooled, charge-coupled device (CCD) camera



Fig. 9 Scheme of the instrumental setup. Microdroplets are generated by a vibrating orificeaerosol generator (orifice diameter=10 mm) consisting of a syringe pump and a piezoceramicoscillator. Droplets are illuminated by a laser beam (continuous Ar ion laser, l=488 nm) thatis focused to a laser beam waist of ~200 mm at the trajectory of the droplet stream. The beamdiameter is chosen to be larger than the droplet diameter to ensure that all fluorophores inthe sample are illuminated. A holographic laser band-pass filter eliminates undesirableplasma lines from the laser source and transmits only the laser line at 488 nm. The fluo-rescence 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 thermoelectricallycooled camera with a 512¥512 pixel charge coupled device (CCD) detector.A 488-nm holo-graphic Raman notch filter placed in front of the slit of the spectrometer blocks elasticallyscattered laser radiation. Published with permission of ACS Copyright Office [224]

(see Fig. 9). The fluorescence is quenched by specific binding of the TCP PAbsto the fluorescent tracer (homologous hapten covalently coupled to fluorescein).The quenching effect is diminished by the presence of the analyte. Therefore anincrease in the signal is produced in a dose-dependent manner when TCP ispresent in the sample. The LIF-microdroplet-QFIA method shows a LOD of0.04 mg L–1 in buffer. The QFIA performed in microtiter plate format using thesame immunoreagents achieved a LOD of 0.36 mg L–1 in buffer. With the LIF-microdroplet-QFIA system, urine samples can be directly analyzed just afterbuffer dilution reaching a LOD of 1.6 mg L–1, which is sufficient detectability foroccupational exposure risk assessment.

PAbs against 2,4,5-TCP have been prepared after theoretical and molecularmodeling chemical studies [226]. Competitive direct and indirect ELISAs havebeen 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 selec-tivity of the assay is high in relation to other chlorophenols frequently presentin real samples, but as with the 2,4,6-TCP assay the brominated analogs mayalso be recognized. The 2,4,5-TCP immunoassay is stable in media with pH

values ranging from 6.6 to 10.5 and ionic strength values varying within 20 and80 mS cm–1. It shows a good accuracy and the coefficients of variation withinand between assays are around 12% or lower. The assay has been evaluated forthe analysis of water samples and complex biological matrices, such as serumand urine.While the water samples could be analyzed without any sample pre-treatment, the serum and urine samples produced important interferences.The investigation of simple sample treatment procedures compatible with theimmunochemical method allowed the establishment of reliable analytical pro-tocols for straightforward immunochemical determination of 2,4,5-TCP innatural 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 specificityfor these analytes, Noguera et al. [225] have recently developed an ELISA forscreening the total concentration of chlorophenols in environmental samples.Using 3,5-dichloro-4-hydroxyphenyl propionic acid as immunizing hapten (thestructure 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 immuno-assays developed (in both direct and indirect formats). The most sensitive onewas the direct format and after the optimization of several parameters such as surfactant concentration, ionic strength, or time of incubation a LOD of0.2 mg L–1, an IC50 value of 7.8 mg L–1, and a dynamic range between 0.7 and86.4 mg L–1 were obtained when 2,4,6-TCP was used as standard. The cross-re-activity 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%) sincethe 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-DCPwere recognized to a much lesser extent (<7%) and chlorophenols with just oneor no chlorine in the molecule or other related compounds without the phenolgroup didn’t cross-react (CR<0.09%). Owing to the broad degree of recognitionof this assay it has been tested as a potential screening tool in order to estimatecontamination by chlorophenols. Several preliminary studies have been carriedout with spiked water samples by the addition of different amounts of CPs withCR levels higher than 1%, and the results of the assay, expressed as equivalentsof 2,4,6-TCP, showed good correlation levels when a regression analysis wasperformed, with recoveries around 95%.

Goda et al. reported the only attempt carried out regarding the productionof MAbs against CPs [148]. By using a similar hapten to the aforementionedone (in this case the spacer arm has six carbon atoms instead of three, but thepattern of 2,6-dichlorophenol is also retained), a direct ELISA using 2,4-DCP asstandard was developed.A LOD of 2 mg L–1 and a working range between 2 and60 mg L–1 were achieved. This assay also shows a high degree of recognition ofseveral chlorophenols tested but, unlike the direct assay of Noguera et al [225],the pattern of recognition is different. Thus, high or moderate cross-reactivityvalues were observed for some mono- or dichlorophenols (e.g., 4-chlorophenolshowed a CR=100%), and even for phenol (24%), whereas for more substituted



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 theoverall concentration of CPs instead of the single compound 2,4-DCP, since anoverestimation can be observed in the analysis.

5Other Industrial Residues

From the wide variety of emerging pollutants of industrial origin that could beconsidered here, bisphenol A (BPA) and phthalate esters (PE) are of especial relevance not only because of the high volumes produced and their widespreaduse, but also because of their demonstrated toxicity, particularly as endocrinedisrupters. Both of them have been included in the final report of the EuropeanCommission toward the establishment of a priority list of endocrine disrupterchemicals, EDCs [3], and have been rated as of high risk of exposure for humanand wildlife populations. Because of their structural characteristics these com-pounds cannot be included in any of the groups described above, so they willbe described in this section (see Fig. 10).

5.1Bisphenol A

Bisphenol A (2,2-bis(4-hydroxydiphenyl)propane, BPA) is a man-made chem-ical mainly used in the manufacture of polycarbonate and epoxy resins. Theseplastics are used in the preparation of containers such as food and drink pack-aging as well as a great variety of products including compact discs, opticallenses, thermal paper, adhesives, powder paints, or even in dental compositefillings and sealants. To a minor extent (about 10%), BPA is used in PVC pro-duction, as an antioxidant and preservative, or as a flame retardant (i.e., tetra-bromobisphenol A, see above). This wide range of applications has led to highlevels of worldwide production (around 2.6¥106 tonnes in 1999). Besides thedischarge 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: diethyl-hexyl phthalate; BBP: butylbenzyl phthalate

release and human and wildlife exposure comes from leaching from theseproducts as a result of incomplete polymerization or hydrolysis of the poly-mers. Therefore BPA has become not only an environmental pollutant but alsoa food contaminant. Due to both its industrial and domestic applications it canbe expected to be found in sewage, influent and effluent wastewater, and sewagesludge. Their discharge into the environment from industrial emitters andcommunal 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 wasfound that almost 90% of the total load was removed, with an effluent meanconcentration 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 organismssuch 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 femalerats, fish, and mice can be produced [263–267]. BPA is mainly determined us-ing chromatographic techniques such as GC–MS [141, 142, 168, 268], HPLC-UV[269], HPLC–MS [142], or HPLC with fluorescence detection [270]. Immuno-chemical methods can improve monitoring efficiency because of their knowncapability 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 ap-plication of these immunoassays has been carried out to analyze biological matrices such as serum samples. An indirect ELISA with PAbs was developedby Zhao et al. [271] with a good LOD of 0.1 mg L–1 and a dynamic range between1 and 10,000 mg L–1 in water. The PAbs were raised using as immunizing hap-ten a bisphenolic structure preserving both phenolic groups and replacing onemethyl group by the spacer arm. No important matrix effects were found whenspiked real water samples were analyzed. Other phenolic compounds did notinterfere in this assay. When spiked serum samples were analyzed a dilutionfactor of 1:10 was required in order to obtain quantitative recoveries, placingthe LOD still at a good level (around 2 mg L–1). With these antibodies an im-munoaffinity procedure for the selective extraction of BPA from serum sam-ples has been developed providing good recovery levels (about 90%) [272].

Ohkuma et al. [273] also prepared PAbs for BPA but in this case one car-boxyalkyl ether as spacer arm was substituted for one of the phenolic groups.With these antibodies a direct competitive ELISA was developed, achieving aLOD of 0.3 mg L–1 and a working range between 0.3 and 100 mg L–1. Accuracyevaluation was carried out by analyzing spiked human serum samples; recov-ery values between 82 and 97% were obtained. The interferences caused by thematrix were negligible. Correlation studies were also done with a conventionalchromatographic technique (GC–MS), obtaining a regression coefficient of0.990, thus demonstrating the possibility of using this immunochemicalmethod to directly determine BPA in serum. Phenols and other endocrine disrupters such as alkylphenols and phthalates were not recognized in this assay.



Tabl

e6

Imm

unoc

hem

ical

met

hods

dev

elop

ed fo

r th

e de

tect

ion

and

quan

tific

atio

n of

BPA

and

pht

hala

te e

ster

s

Ana

lyte

Imm

unoc

hem

ical

Se

nsit

ivit

yM

atri

xR

efer

ence

ste

chni

quea

LOD

IC50

Oth

ers

Phth

alat

e es

ters

b

DM

P (D

EP,D

BP,B

BP

TR-

FIA

97ng

L–1Bu

ffer

[276

]an

d D

OP)

DBP

ELIS

A0.

2m

gL–1

Buff

er[1

48]

Bisp

heno

l A

ELIS

A5

mgL–1

Buff

er[1

48]

ELIS

A0.

1mg

L–1W

ater

[271

]2

mgL–1

Seru

m[2

71]

ELIS

A0.

3mg

L–1Se

rum

[273

]

ELIS

A0.

57m

gL–1

Wat

er[2

75]

Aut

omat

ed B

MP-

IA2.

3ng

L–1Bu

ffer

[147

]

aEL

ISA

:enz

yme-

linke

d im

mun

osor

bent

ass

ay;B

MP-

IA:b

acte

rial

mag

neti

c pa

rtic

le-b

ased

imm

unoa

ssay

;TR-

FIA

:tim

e-re

solv

ed fl

uoro

imm

unoa

ssay

.b

DEP

:die

thyl

pht

hala

te;D

BP:d

i-n-

buty

l pht

hala

te;B

BP:b

utyl

benz

yl p

htha

late

;DM

P:di

met

hyl p

htha

late

;DO

P:di

octy

l pht

hala

te.

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 thatshowed a working range between 0.5 and 5 mg L–1. The assay was validated byHPLC. DeMeulenaer et al. [275] developed an indirect competitive ELISA us-ing PAbs obtained from chicken egg yolk, but the assay achieved an IC50 valueof only 570 mg L–1.

The production and use of monoclonal antibodies against BPA have alsobeen reported. Nishii et al. [277] developed an ELISA with MAbs selected fortheir 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 directELISA 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 andBPA metabolites showed no cross-reactivity. Takeda Chemical Industries, Ltd.have commercialized ELISA kits using these antibodies. The immunoassayshows 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 basedon the immobilization of monoclonal antibodies on the surface of magneticparticles has been developed [147]. The sensitivities achieved were also verygood but the assay shows a wide dynamic range (between 2.3 ng L–1 and2.3 mg L–1), indicating a low slope assay as occurred with the assay developedfor APE based on the same principle (see above).

5.2Phthalate Esters

Phthalate esters are used to manufacture cosmetic products, inks, adhesives,and solvents, but they are mainly used as additives in PVC production. Thiskind 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, diethylhexylphthalate (DEHP) represents over 90% of the total phthalate production, nearto 54,000 tonnes per year [278, 279], followed by butylbenzyl phthalate (BBP),dibutyl phthalate (DBP), and dioctyl phthalate (DOP). Due to this wide rangeof applications as well as their high consumption, these compounds can enterinto the environment through different routes, increasing the risk of exposureof 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 he-modialysis [280–282]. The origin of these levels can be the leaching of phtha-lates from the tubes or the bags containing blood. These plasticizers are poorlysoluble in water, and therefore it is expected to find them not only in wastewaterbut also adsorbed onto sewage sludge and soils. The environmental fate ofphthalates has been extensively reviewed [65] and it can be concluded that theiraerobic degradation occurs rapidly, preventing their accumulation in water and


even in soil. The biodegradation is slower in anaerobic or cold environments,but several reported experiments indicate that the bioaccumulation of thesecompounds 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 con-cern has also been focused on their endocrine disrupting effects, which havebeen reviewed [289]. Several in vitro studies confirm their estrogenic activity[161, 290], mainly for BBP and DBP, although it is quite weak compared withthe natural estrogen 17b-estradiol [291]. Other in vivo experiments also showthat DBP and DEHP can produce alterations in male sexual differentiation andin 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 usualanalysis involves the use of both GC and LC coupled to MS detection, althoughsometimes 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 ofdetecting DEHP, the most used phthalate and also the most persistent one in the environment. The first immunochemical technique is based on the devel-opment of a time-resolved fluoroimmunoassay (TR-FIA) [276], a heterogeneousimmunoassay that uses europium chelates as labels. Their special fluorescentproperties, such as their longer decay time (over hundreds of milliseconds) com-pared with other more conventional organic molecules (around nanoseconds),determine the good detectability of this method. PAbs were produced against animmunogen derived from dimethyl phthalate (DMP). The assay developed usingthese antibodies showed a LOD of 97 ng L–1 and a working range between97 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 and110%).

MAbs were also produced by Goda et al. [149], and a direct ELISA has beendeveloped. The LOD accomplished for DBP was 0.2 mg L–1 with a dynamic rangeof 0.2–4 mg L–1. DEHP was not recognized (CR<1%). BBP showed a cross-reac-tivity value of 155% and dipropyl phthalate (DPrP) and dipentyl phthalate(DPnP) cross-reacted 60 and 51%, respectively.

6General Summary

It has been widely demonstrated that immunochemical techniques offer a goodalternative to conventional methodologies in many areas due to the high sen-sitivity and selectivity achieved for the antibodies toward the target analytes.


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 immuno-chemical methods and formats have been developed (RIA, ELISA, PFIA, FIIA,immunosensors, immunoaffinity extraction procedures, etc.). Several surfac-tants can readily be detected by these methods.However,although attempts havebeen made toward the immunochemical determination of NP, under suspicionbecause of its potential estrogenic activity, the methodologies reported also rec-ognize its parent compounds. Immunochemical analytical methods for PCBsand PCDDs/PCDFs have been reported, showing good detectability levels. Mostof the problems that arise while analyzing these substances are related to theirlack of solubility in the aqueous-based systems of the immunochemical meth-ods. However, protocols have been developed using organic solvents and watermixtures. BPA can be efficiently detected with the immunochemical methodsavailable today, not only in water samples, but also in more complex biologicalmatrices such as serum, with detection limits in the order of micrograms perliter or even lower. Regarding phthalates, more efforts must be directed towardthe production of specific antibodies for the main phthalate DEHP, since the im-munochemical 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 con-trol laboratories.A reason for this may lie in the lack of knowledge on the per-formance of these types of techniques by certain analytical sectors, and also inthe lack of validated protocols for a wide range of sample matrices. Im-munoassay methods may suffer from undesirable matrix effects that may leadto false positive or negative results. It is wrong to assume that the selectivity of the immunochemical method is sufficiently high to overcome nonspecificinteractions of the antibodies with the matrix components. Rigorous evaluationof 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 effi-ciency of these techniques. A close collaboration and interchange of the exper-tise of analytical chemists and immunochemists are needed to accomplish thisgoal and benefit from the advantages of these methods, to assess risk and pro-tect 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 NMP-505485-1).


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180 Immunochemical Determination of Industrial Emerging Pollutants

Immunochemical Determination of Pharmaceuticalsand 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 Antibiotics . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 1982.1 Penicillins . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2062.2 Chloramphenicol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2142.3 Tetracyclines . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2152.4 Sulfonamides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2162.5 Fluoroquinolones . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2172.6 Macrolides . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 218

3 Steroid Hormones . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2193.1 Estrogens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2233.2 Androgens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2283.3 Gestagens . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2293.4 Corticosteroids . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 230

4 Other Drugs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2334.1 Analgesics and NSAIDs . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2344.2 Cytostatic Agents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 237

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

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 239

Abstract A review on immunochemical methods for the analysis of pharmaceuticals is pre-sented. A broad range of pharmaceutical categories and personal care products may reachthe aquatic environment after excretion through industrial, domestic, and hospital waste-water. With few exceptions pharmaceuticals for human medicine are not high-productionchemicals and the expected environmental concentrations should be low. However, the useof some of these chemicals in veterinary medicine increases the probability that the con-centration values in the aquatic environment might reach higher levels. On the other handcertain 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 po-tential human risks, this review focuses on antibiotics, hormones, analgesics, nonsteroidalanti-inflammatory drugs, and cytostatic agents. Although these procedures have only beenapplied to the analysis of environmental samples on a few occasions, immunochemicalmethods for several of these substances exist and some of them are commercially availabledue to their use in clinical laboratories and forensic medicine.


Keywords Immunochemical techniques · Antibiotics · Steroid hormones · Analgesics · Cytostatic agents

AbbreviationsBSA Bovine serum albuminCAP ChloramphenicolCE Capillary electrophoresisCOD CodeineCR Cross-reactivityE1 EstroneE2 b-EstradiolE3 EstriolEE2 EthynylestradiolEIA Enzyme immunoassayELIFA Enzyme-linked immunofiltration assayELISA Enzyme-linked immunosorbent assayEMIT Enzyme-multiplied immunoassay techniqueETIA Energy transfer immunoassayEW Evanescent waveFPIA Fluorescence polarization immunoassayGC Gas chromatographyHPIAC High-performance immunoaffinity chromatographyHPLC High-performance liquid chromatographyIAC Immunoaffinity chromatographyLIF Laser-induced fluorescenceLOD Limit of detectionMAb Monoclonal antibodyMECC Micellar electrokinetic capillary chromatographyMIAC Multi-immunoaffinity chromatographyMOR MorphineMRL Maximum residue levelMS Mass spectrometryMTX MethotrexateNSAID Nonsteroidal anti-inflammatory drugPAb Polyclonal antibodyPEC Predicted environmental concentrationPFIA Polarization fluoroimmunoassayPNEC Predicted no effect concentrationPPCPs Pharmaceutical and personal care productsRIA RadioimmunoassaySPIA Sol particle immunoassaySPR Surface plasmon resonanceSTPs Sewage treatment plantsTIRF Total internal reflection fluorescence immunoassayWWTPs Wastewater treatment plants


1Introduction

Pharmaceuticals and personal care products (PPCPs) are a diverse group ofchemicals used internally or externally in the body of humans, domestic ani-mals, and plants. They include substances such as human and veterinary drugs,diagnostic agents (i.e., X-ray contrast media), nutraceuticals (bioactive foodsupplements), certain feed and food additives, sunscreen agents, fragrances,cosmetic additives, etc. Their presence in the environment is determined bytheir worldwide frequent use by multitudes of individuals or by their use as veterinary drugs. Thousands of tons of pharmacologically active substances areused yearly to treat illnesses, to prevent unwanted pregnancy, or to face thestress of modern life.For example,about 50,000 drugs are registered in Germanyfor human use, 2,700 of which are responsible for 90% of the total consumptionand which, in turn, contain about 900 different active substances [1]. In the UKapproximately 3,000 active substances are licensed [2]. In animal farming theuse of antibiotics, feed additives, and hormones has become a usual practice forcertain farmers. Substances such as the UV filters used for sunscreens and cos-metics, designed to remain on the skin, are generally washed off either duringbathing 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 envi-ronment of PPCPs. Several reviews discuss and present real data on the envi-ronmental occurrence of the most relevant PPCPs [1, 3–7]. We must moreoverremark 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 potentagents. Often they are resistant to biodegradation, as a certain metabolic sta-bility is necessary for their pharmacological action. PPCPs are released into theenvironment unaltered or as still active metabolites. Thus, human and veteri-nary drugs are frequently excreted as glucuronide or sulfate conjugates that caneasily be hydrolyzed to release again the active parent compound in the envi-ronment. Certain PPCPs or their metabolites are highly soluble in water. Thisparameter combined with a lack of biodegradability limits the removal ofPPCPs in wastewater treatment plants (WWTPs) [4, 8].Additionally, antibioticsand disinfectants are supposed to disturb the wastewater treatment process andthe microbial ecology in surface waters. Furthermore, resistant bacteria may beselected in the aeration tanks of STPs (sewage treatment plants) by the antibi-otic substances present [1]. In these cases the drugs enter the aquatic environ-ment, contaminating ground and surface waters and eventually also reachingdrinking water. Some other PPCPs are more lipophilic, thus showing a tendencyto bioaccumulate in organisms that fortuitously will be used for human con-sumption. The consequence is that PPCPs enter the environment resulting inreal exposure of humans and wildlife to these active substances [5, 9–12].

The concern produced within the scientific community, the authorities, andgovernmental bodies has prompted the establishment of certain regulations

Immunochemical Determination of Pharmaceuticals and Personal Care Products 183

[13, 14]. Thus, before any new veterinary pharmaceutical product can obtain amarketing authorization, a review must be carried out by national or EuropeanUnion (EU) authorities to ensure its efficacy, quality, and safety to public healthand the environment. The requirements for ecotoxicity testing are regulated byDirective 2001/82/EC. The European Agency for the Evaluation of MedicinalProducts (EMEA), which coordinates the evaluation and supervision of med-icinal products for both human and veterinary use, has published guidancedocuments on how to perform environmental risk assessment of these products[15]. The International Cooperation on Harmonization of Technical Require-ments for Authorization of Veterinary Medicinal Products (VICH) formed bythe EU, USA, and Japan (Australia and New Zealand participate as observers)are trying to establish uniform risk assessment criteria [16]. The general ideais to predict the environmental concentration (PEC) of these substances basedon calculations such as number of prescriptions, doses, degradation, environ-mental models, etc. [12]. The PEC value is then compared to the lowest effec-tive concentration found, for the parent compound or its metabolites, in certainecotoxicity tests performed in soil and/or water to establish what is known asthe predicted no effect concentration (PNEC). The ratio PEC/PNEC should belower than 1. Otherwise a risk to the environment is assumed and risk mitiga-tion measures should be linked to the authorization of the product [17]. Thecomplete guidance is still being developed, but already some data dealing with these parameters have been published on the most frequently used phar-maceuticals (i.e., [18, 19] and personal care products (i.e., [17]). In a study per-formed in Denmark, taking only pharmaceuticals prescribed for human medicine into consideration, ibuprofen, acetylsalicylic acid, and paracetamolexceeded the PEC/PNEC reference value [18]. Similar conclusions were drawnfor paracetamol, amoxicillin, oxytetracycline, and mefenamic acid in a study onthe aquatic environmental assessment of the top 25 English most prescribedpharmaceuticals [19].

Because of their inherent bioactivity, trace levels of these substances canhave a negative impact on the environment and public health. This fact, to-gether with the above-mentioned considerations, gives rise to substantial an-alytical problems. The different formats and benefits that the immunochemi-cal methods can confer on the analysis of trace contaminants in aqueous-basedsamples has been widely demonstrated [47–56] (see also Chap. 5 in this vol-ume). For this particular case immunochemical techniques offer an additionalbenefit derived from the possibility of developing single or class-specific meth-ods according to necessity. Environmental monitoring of PPCPs requires effi-cient methodologies able to detect trace levels of contamination caused by bothparent compounds and their metabolites. The present chapter describes someof the immunochemical methods available today for the analysis of the mostimportant PPCPs regarding their potential impact on the environment (seeTable 1). The PPCPs treated in this chapter have been selected according to theirproduction and use. Because of the recent concern about their environmentalimpact, the aim of most of the immunochemical techniques reported for ana-



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ewag

eVe

teri

nary

med

icin

efis

h in

Nor

way

189

–285

mg

L–1in

slud

ges

or s

edim

ents

.H

uman

med

icin

ese

dim

ents

in fi

sh fa

rmin

g [2

6]It

form

s st

able

com

plex

es–

Riv

er w

ater

:1 m

g L–1

[3]

wit

h C

a2+.

– U

S st

ream

s:0.

34 m

g L–1

[20]

Pers

iste

nt in

ano

xic

cond

itio

ns

Tetr

acyc

line

Aqu

acul

ture

– R

iver

wat

er:1

mg

L–1[3

]A

ccum

ulat

es in

sew

age

Vete

rina

ry m

edic

ine

– U

S st

ream

s:0.

11 m

g L–1

[20]

slud

ges

or s

edim

ent

Hum

an m

edic

ine

Deg

rada

tion

rat

e of

tetr

acyc

line

in li

quid

man

ure

is a

ppro

xim

atel

y50

% in

5 m

onth

sPh

otod

esco

mpo

siti

on

Chl

orte

trac

yclin

eA

quac

ultu

re–

SW in

US:

0.42

mg

L–1[2

0]A

ccum

ulat

es in

sew

age

Hum

an m

edic

ine

– R

iver

wat

er:1

mg

L–1[3

]sl

udge

s or

sed

imen

tG

row

th p

rom

oter

inA

fter

30

days

at 3

3 °C

,ve

teri

nary

med

icin

e44

% r

emai

ned


Tabl

e1

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Mac

rolid

es

Eryt

hrom

ycin

Vete

rina

ry m

edic

ine

– G

W in

Ger

man

y:up

to

Not

deg

rada

ble

in th

eH

uman

med

icin

e20

0 ng

L–1

[25]

envi

ronm

ent [

3].

Aqu

acul

ture

– SW

in G

erm

any:

0.15

mg

L–1 [2

3]La

bile

in a

cid

cond

itio

ns–

STP

efflu

ents

in G

erm

any.

2.5

mg L

–1[2

3]–

Riv

er w

ater

:1 m

g L–1

[3]

– R

iver

Po

in It

aly:

3.2

ng L

–1[2

7]

Cla

rith

rom

ycin

Hum

an m

edic

ine

– ST

P ef

fluen

ts in

Ger

man

y:M

ore

stab

ility

than

0.24

mg

L–1[2

3]er

ythr

omyc

in in

aci

d–

GW

in G

erm

any

cond

itio

ns0.

24–0

.87

mg L

–1[2

3]–

Riv

er P

o in

Ital

y:1.

7 ng

L–1

[27]


Tabl

e1

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Mac

rolid

es

Rox

ithr

omyc

inH

uman

med

icin

e–

US

stre

ams:

0.05

mg

L–1[2

0]N

o da

ta fo

und

Trim

etho

prim

Vete

rina

ry m

edic

ine

– ST

P ef

fluen

ts (

Ger

man

y):

Hal

f-lif

e >

1 ye

arH

uman

med

icin

e0.

32 m

g L–1

[23]

Mix

ture

wit

h–

US

stre

ams:

0.15

mg

L–1[2

0]su

lfona

mid

es

Chl

oram

phen

icol

Aqu

acul

ture

– Se

wag

e an

d su

rfac

e le

vel:

low

Hyd

roly

sis

ofEx

cept

iona

l cas

es in

at m

g L–1

[4]

chlo

ram

phen

icol

hum

an m

edic

ine

– ST

P in

ger

man

y:0.

56 m

g L–1

[23]

gluc

uron

ide

inVe

teri

nary

use

isch

lora

mph

enic

olfo

rbid

den


Tabl

e1

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Ho

rmo

nes

Estr

ogen

s

Estr

adio

l (E2

)En

doge

nous

hor

mon

e–

Sew

age

efflu

ent:

<0.

1–88

ng

L–1M

ean

Hal

f-lif

e:2.

8 da

ys.

[28,

29]

Sorp

tion

in th

e se

dim

ents

– SW

<0.

05–1

5 ng

L–1

Estr

one

(E1)

Met

abol

ite o

f–

Sew

age

efflu

ent:

Mea

n H

alf-

life:

3.0

days

.es

trad

iol

<0.

1–22

0 ng

L–1

[28]

Mai

n pr

oduc

t of

– SW

<0.

1–17

ng

L–1de

grad

atio

n of

estr

adio

l

Estr

iol (

E3)

Met

abol

ite o

f–

Sew

age

efflu

ent:

<0.

1–42

ng

L–1Av

erag

e re

mov

ales

trad

iol

(diff

eren

t cou

ntri

es) [

28]

effic

ienc

y 96

%.

– SW

<0.

1–3.

4 ng

L–1

Deg

rada

tion

not

rep

orte

d

Ethy

nyle

stra

diol

O

ral c

ontr

acep

tive

– Se

wag

e ef

fluen

t:M

ean

Hal

f-lif

e:17

day

s.(E

E2)

<0.

053–

62 n

g L–1

Hig

h so

rpti

on o

nto

(diff

eren

t cou

ntri

es)

[28,

29]

sedi

men

ts–

SW:<

0.05

3–30

.8 n

g L–1


Tabl

e1

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Estr

ogen

s

Mes

tran

ol

Ora

l con

trac

epti

ve–

Efflu

ents

:<1–

8 ng

L–1

[30]

Deg

rade

d in

aer

obic

(MeE

E2)

cond

itio

ns in

slu

dge

(80%

)an

d 7%

hyd

roly

zed

to E

E2

And

roge

ns

Test

oste

rone

(T

)En

doge

nous

hor

mon

e–

Raw

sew

age

in W

WT

P:16

toT

run

s of

fby

leac

hing

of

Gro

wth

pro

mot

er70

0 ng

mL–1

aque

ous

solu

tion

from

the

Ana

bolic

ste

roid

– G

W:1

.0 n

g L–1

soil

[31]

– R

unof

fwat

er fr

om m

anur

edfie

lds:

215

ng L

–1[1

3]

Met

hylte

sto-

Gro

wth

pro

mot

er–

Pond

Wat

er a

fter

trea

tmen

t wit

hPh

elps

et a

l.[3

2] fo

und

ster

one

(MT

)A

nabo

lic s

tero

idM

T fo

od:<

5 m

g L–1

[32]

that

MT

in th

e w

ater

retu

rned

to th

e ba

ckgr

ound

leve

ls w

ithi

n on

e w

eek

afte

r ho

rmon

ead

min

istr

atio

n

Tren

bolo

ne (

Tr)

Gro

wth

pro

mot

er–

Aft

er 5

.5 m

onth

s in

soi

lTr

acea

ble

afte

r 8

days

of

Ana

bolic

ste

roid

fert

ilize

d w

ith

liqui

d m

anur

e:fe

rtili

zati

on,n

ot d

etec

tabl

e0.

16–0

.10

mg k

g–1af

ter

40 d

ays

[31]


Tabl

e1

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Ges

tage

ns

Prog

este

rone

Endo

geno

us h

orm

one

– U

S st

ream

s:0.

11 m

g L–1

[20]

No

data

foun

d

Nor

ethi

ndro

neO

ral C

ontr

acep

tive

– R

iver

sam

ples

(UK

) 28

% B

iode

grad

atio

n in

17

ng

L–1 [3

3]6

h af

ter

the

plan

t–

Efflu

ents

:8–2

0 ng

L–1

[30]

trea

tmen

t and

com

plet

ely

in 2

4 h

[33]

Cor

tico

ster

oids

Cor

tiso

neA

nti-

infla

mm

ator

yN

o da

ta fo

und

Prac

tica

lly in

solu

ble

inag

ent

wat

er [

34]

Cor

tiso

lA

nti-

infla

mm

ator

yN

o da

ta fo

und

Very

sta

ble

[34]

agen

t


Tabl

e1

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Cor

tico

ster

oids

Bet

amet

haso

neA

nti-

infla

mm

ator

yN

o da

ta fo

und

Prac

tica

lly in

solu

ble

inag

ent

wat

er.

Gro

wth

pro

mot

erVe

ry s

tabl

e [3

4]

Dex

amet

haso

neA

nti-

infla

mm

ator

yN

o da

ta fo

und

Prac

tica

lly in

solu

ble

inag

ent

wat

er [

34]

Gro

wth

pro

mot

er


Tabl

e1

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Oth

ers

Ana

lges

ics

Para

ceta

mol

Mild

ana

lges

ic–

US

stre

ams,

max

:10

mg L

–1[2

0]R

eadi

ly d

egra

dabl

e af

ter

(ace

tom

ino-

Ant

iphl

ogis

tic

– ST

P ef

fluen

t,m

ax:6

mg

L–1ac

clim

atiz

atio

nph

en)

(Ger

man

y) [

35]

Asp

irin

(ace

tyl-

Pain

kill

er–

Sew

age

efflu

ent (

Engl

and)

:R

eadi

ly b

iode

grad

able

salic

ylic

aci

d)A

ntit

hrom

boti

c ag

ent

1 ng

ml–1

[24]

– ST

P ef

fluen

t,m

ax:1

.5 m

g L–1

and

rive

r an

d st

ream

s w

ater

(Ger

man

y):0

.34

mg L

–1[3

5]

Salic

ylic

aci

dM

etab

olite

of

– Se

wag

e ef

fluen

ts:1

3 mg

L–1

[36]

No

data

foun

dac

etyl

salic

ylic

aci

d–

STP

efflu

ent,

max

:0.1

4 mg

L–1

and

rive

r an

d st

ream

s (G

erm

any)

:4.

1 mg

L–1

[35]

– ST

P ef

fluen

t (B

erlin

):0.

04 m

g L–1

[37]

– G

W (

Ber

lin)

max

:122

5 ng

L–1

[37]

Gen

tisi

c ac

idM

etab

olite

of

– ST

P ef

fluen

ts:m

ax:0

.59

mg L

–1N

o da

ta fo

und

acet

ylsa

licyl

ic a

cid

and

rive

rs a

nd s

trea

ms

(Ger

man

y):1

.2 m

g L–1

[35]

– G

W B

erlin

max

:540

ng

L–1[3

7]

196 M.-C. Estévez et al.Ta

ble

1(c

onti

nued

)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Ana

lges

ics

Ibup

rofe

nA

nti-

infla

mm

ator

y–

Diff

eren

t Ger

man

riv

ers:

– In

here

ntly

bio

degr

ad-

agen

t17

–139

ng

L–1[3

8]ab

le (<

95%

rem

oved

in

Pain

kill

er–

STP

efflu

ent m

ax:3

.4 m

g L–1

and

WW

TPs

) [39

] A

ntip

hlog

isti

cri

ver

and

stre

ams

wat

erA

nti-

rheu

mat

ic(G

erm

any)

:0.5

3 mg

L–1

[35]

– ST

P ef

fluen

t (B

erlin

):0.

1 mg

L–1

[37]

– G

W (

Ber

lin):

max

:20

0 ng

L–1

[37]

– ST

P in

fluen

ts u

p to

3 m

g L–1

and

STP

efflu

ent:

2 ng

L–1

,riv

ers

an

d la

kes

(Ger

man

y):u

p to

8

ng L

–1 [3

9]–

Ital

ian

rive

rs:

90.6

–92.

4 ng

L–1

[40]

– Se

wag

e ef

fluen

t:1.

5,0.

87 a

nd85

mg

L–1;S

W:2

.7 m

g L–1

[36]

Dic

lofe

nac

Ant

iphl

ogis

tic

– R

iver

s:15

–500

ng

mL–1

[38]

Rea

dily

or

inhe

rent

ly–

Efflu

ent:

up to

2 n

g m

L–1[3

5,38

]bi

odeg

rada

ble;

phot

olyt

ic–

STP

influ

ents

:12–

560

ng L

–1;

degr

adat

ion

[4,4

2]ST

P ef

fluen

ts (

Gre

ece)

:10

–365

ng

L–1[4

1]–

Riv

er A

abac

h (S

wit

zerl

and)

:11

–310

ng

L–1[4

2]–

Lake

s (S

wit

zerl

and)

:<

1–12

ng

L–1[4

2]–

WW

TP

efflu

ent (

Swit

zerl

and)

:0.

99 m

g L–1

[43]


Tabl

e1

(con

tinu

ed)

Subs

tanc

eC

hem

ical

Str

uctu

reU

ses/

Ori

gin

Envi

ronm

enta

l Occ

urre

nce

Envi

ronm

enta

l Fat

e

Ana

lges

ics

Cod

eine

Ana

lges

ic–

US

stre

ams

max

:0.0

19 m

g L–1

Not

dat

a fo

und

med

ian

valu

e:0.

012

mg L

–1[2

0]

Cyto

stat

ic (a

ntin

eopl

asti

c) a

gent

s

Met

hotr

exat

eC

ance

r th

erap

y–

Riv

er a

nd p

otab

le w

ater

:Pe

rsis

tent

trea

tmen

t<

6.25

ng

L–1[3

3](c

hem

othe

rapy

)

Cyc

loph

os-

Can

cer

ther

apy

– Tr

eate

d ho

spit

al e

fflu

ent f

rom

Not

deg

rada

ble

pham

ide

trea

tmen

tST

P:14

6 ng

L–1

[44,

45]

(che

mot

hera

py)

– ST

P ef

fluen

t (G

erm

any)

max

:20

ng

L–1[3

5]

Ifos

fam

ide

Can

cer

ther

apy

– Tr

eate

d ho

spit

al e

fflu

ent:

No

data

foun

dtr

eatm

ent

24 n

g L–1

[44]

(che

mot

hera

py)

– O

ncol

ogic

hos

pita

l eff

luen

t:m

ean:

109

ng L

–1[4

6]–

STP

efflu

ent (

Ger

man

y) m

ax:

2.9

mg L

–1[3

5]

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 thecomplexity of the biological samples, their application to environmental watersamples should be straightforward. The use of those formats using radioactivelabels, initially developed for biochemical studies, should be avoided wheneverpossible due to the problems derived from handling and producing radioactivewaste.

2Antibiotics

Antibiotics are chemical substances that are able to suppress or kill the growthof bacteria. They are extensively used in human and veterinary medicine as wellas in aquaculture. They are administered in veterinary medicine for the treat-ment and control of infectious diseases such as mastitis, enteritis, peritonitis,and pneumonia. Moreover, certain antibiotic substances have also been used asgrowth promoters in food producing animals [57, 58].Antibiotic therapy beganwith the clinical use of sulfonamides in 1936 and was followed by the develop-ment of penicillins (1944), chloramphenicol (1947), tetracyclines (1948), and fluoroquinolones (1980) (see Fig. 1 for the chemical structures). Since 1950, andin parallel with the use of antibiotics in human medicine, veterinary use hasprovided control of diseases in animal farms. This was followed by their use asgrowth promoters in many countries, although at present this practice is be-coming 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 tothese data, 5,000 tons are due to veterinary purposes (3,500 tons prophylaxisand therapy, and growth promotion about 1,500 tons). The other half of pro-duction is used in medicine.Among the antibiotics used in veterinary practice,


Table 2 Veterinary consumption of therapeutic antibiotics in Europe

Product group % Share

Penicillins 9Tetracyclines 66Macrolides 12Aminoglycosides 4Fluoroquinolones 1Trimethoprim/sulfonamides 2Others 6Totala 100

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


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

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 ofantimicrobial agents used in food animals (cattle, chickens, pigs, and turkeys) inthe United States is unknown.At least 17 classes of antimicrobial agents are ap-proved for growth promotion and feed efficiency in the United States, includingtetracyclines, penicillins, macrolides, lincomycin (analog of clindamycin), andvirginiamycin (analog of quinupristin/dalfopristin). To understand the humanhealth consequences of the use of antimicrobial agents in food animals, it is important to evaluate the quantity of antimicrobial agents used in food animalsin the USA. Unfortunately, although reporting systems have recently been im-plemented in several European countries, no reporting system exists for thequantity of antimicrobial agents used in food animals in the United States. TheAnimal Health Institute, which reportedly represents 80% of the companiesthat produce antimicrobial agents for animals in the USA, has estimated thattheir member companies produced 18 million pounds of antimicrobial agentsfor therapeutic and nontherapeutic (growth promotion and disease prevention)use in food animals in the USA in 1999 [60]. An alternative report, provided bythe Union of Concerned Scientists in 2001, estimated that 29 million pounds ofantimicrobial agents are used in food animals annually in the USA, of which25 million pounds are used for nontherapeutic purposes [61]. Though moreprecise data on the quantity of antimicrobial agents used in food animals areneeded, these initial estimates provide some perspective on the quantity of an-timicrobial 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 beviewed as the ability of microorganisms of a certain species to survive or evento grow in the presence of a concentration of an antimicrobial that is usuallysufficient to inhibit or kill bacteria of the same species. In the presence of anantimicrobial, organisms with inherent or acquired resistance to the agent willbe selected. The bacterial population then comes to consist largely or entirelyof resistant bacteria, causing failure of the traditional treatments. It has beenreported that more than 70% of bacteria are insensitive against at least one antibiotic. This situation is causing a serious threat for public health, as moreand more infections can no longer be treated with the presently known anti-dotes [62–68].

The World Health Organization has recommended that, unless a risk-basedevaluation demonstrates their safety, the growth promotion use in food animalsof antimicrobial agents that belong to the same classes of antimicrobial agentsused in humans should be terminated [69]. Similar recommendations to dis-continue the use of human antimicrobial agents as growth promoters in foodanimals have been made by several independent organizations in the UnitedStates, including the Alliance of Prudent Use of Antibiotics in 2002 [70] and thedistinguished Institute of Medicine of the National Academies in 2003 [71]. Forthis reason, the EU has established the principle of using different antibiotics for


humans and animals. Since 1999 the EU has also banned some antibiotics suchas 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 localconcentrations (e.g., aquaculture, hospital effluents). Several studies have beencarried out in the USA, Germany, Switzerland, and Denmark to investigate theoccurrence and fate of the antibacterial drugs in STPs or surface waters (seeTable 1). Antibiotic resistance causes an important impact on the ecosystem,water, and soil-dwelling organisms. Moreover, some antibiotics can also pro-duce adverse effects in animals and plants. For example, sulfadimethoxine andbacitracin produce loss of weight in roots and leaves in some plants, and oxyte-tracycline and tetracycline can kill pinto bean plants at a concentration level of160 mg L–1 [3]. Chloramphenicol can produce pneumonia and sulfamethazinehas been evaluated by the WHO/FAO Expert Committee as a suspected car-cinogen.

Macrolide antibiotics (clarithromycin, dehydroerythromycin, etc.) and sul-fonamides (sulfamethoxazole, sulfadimethoxine, sulfamethazine, and sulfathi-azole) are the most prevalent antibiotics found in the environment with levelsaround a few micrograms per liter, whereas fluoroquinolones, tetracyclines, andpenicillins have been detected in fewer cases and usually at low concentrations(nanograms per liter) [3, 20, 23, 72]. This result is not surprising, since penicillinsare easily hydrolyzed and tetracyclines readily precipitate with cations such ascalcium and are accumulated in sewage sludge or sediments. Several reviewshave reported the environmental occurrence of different antibiotics in aquaticand soil compartments. Some of these data are detailed in Table 1.

Various techniques based on completely different principles have been usedto detect antibiotic residues. Traditionally, most of the tests used take advantageof the antibacterial activity of the antibiotics. These growth inhibition tests havebeen used in different animal matrices; however, they detect all residue levels of any antibiotic above the MRL and no conclusion may be drawn about theidentity of the antibiotic or its concentration [73]. On the other hand, HPLC andGC are highly specific but require extensive sample preparation, sophisticatedequipment, and skilled laboratory personnel. Therefore they cannot be used for routine screening of a large number of samples [21, 23, 72, 74]. Immuno-chemical techniques can be excellent tools to assess contamination of the en-vironment by antibiotics in different matrices due to their high detectabilityand specificity. Furthermore, immunoassays are excellent tools for screeninglarge numbers of samples in short time periods. Table 3 summarizes some ofthe immunochemical techniques reported for the detection of several familiesof antibiotics.As mentioned in the introduction, usually these techniques havebeen developed to determine antibiotic levels in biological matrices, howevertheir availability opens the door to further applications in analyzing environ-mental samples.



Tabl

e3

Som

e im

mun

oche

mic

al te

chni

ques

dev

elop

ed fo

r th

e de

tect

ion

ofan

tibi

otic

s

Ana

lyte

Imm

unoc

hem

ical

Se

nsit

ivit

yM

atri

xR

efer

ence

ste

chni

quea

LOD

IC50

Fluo

roqu

inol

ones

Cip

roflo

xaci

nEL

ISA

10pg

mL–1

cSe

rum

[75]

Sara

floxa

cin

ELIS

A7.

3mg

L–1

Buff

er[7

6]

Nor

floxa

cin

ELIS

A4.

0g

kg–1

Bovi

ne m

ilk[7

7]O

vine

kid

ney

Sulfo

nam

ides

Sulfa

thia

zole

ELIS

A1

ngm

L–16

ngm

L–1Sw

ine

liver

[78]

ELIS

A35

.5mg

L–1M

ilk[7

9]88

.0mg

L–1H

oney

[79]

ELIS

A2.

5mg

L–1Bu

ffer

[80]

Sulfa

met

hazi

neBi

acor

e Q

10

ngm

L–1Se

rum

[81]

(opt

ical

bio

sens

or)

Biac

ore

Q

7.4

ngg–1

Mus

cle

[82]

(opt

ical

bio

sens

or)

Biac

ore

1000

0.02

3mg

mL–1

0.04

1mg

mL–1

Bile

s[8

3]R

IA5

mgL–1

Wat

er[8

4]IA

C–E

LISA

0.17

ngm

L–11.

39ng

mL–1

Uri

ne[8

5]EL

ISA

10ng

g–1Ti

ssue

s[8

6]Bi

acor

e 10

00 s

yste

m0.

015

mgm

L–1Po

rcin

e bi

le[8

7]

aEI

A:e

nzym

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mun

oass

ay;E

LISA

:enz

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linke

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bent

ass

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LIFA

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linke

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mun

ofilt

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ssay

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C:i

mm

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ffin

ity

chro

mat

ogra

phy;

SPIA

:sol

par

ticl

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mun

oass

ay;S

PFIA

:sol

id-p

hase

fluo

resc

ence

imm

unoa

ssay

;SPR

:sur

face

pla

smon

res

onan

ce;R

IA:r

adio

-im

mun

oass

ay.


Tabl

e3

(con

tinu

ed)

Ana

lyte

Imm

unoc

hem

ical

Se

nsit

ivit

yM

atri

xR

efer

ence

ste

chni

quea

LOD

IC50

Sulfo

nam

ides

Sulfa

dim

etho

xine

ELIS

A1.

50mg

L–1Li

ver

tiss

ue[8

8]

Sulfa

diaz

ine

Biac

ore

Q

5.6

ngg–1

Mus

cle

[82]

(opt

ical

bio

sens

or)

Biac

ore

1000

sys

tem

0.02

8mg

mL–1

Porc

ine

bile

[87]

Sulfa

dim

idin

eSP

IA10

ngm

L–1U

rine

[89]

20 n

g m

L–1M

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Penc

illin

s

Cep

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xin

Aut

omat

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ow-t

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gh

1mg

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Raw

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[90]

ampe

rom

etri

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Am

pici

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SPR

5.9

mgL–1

48.8

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er[9

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.6mg

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ilk[9

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allu

x ki

t)50

mgL–1

Bovi

ne a

nd

[92]

porc

ine

kidn

eyEL

ISA

(Lac

Tek

kit)

Pl

asm

a[9

3]qu

alit

ativ

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Peni

cilli

n G

Biac

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5–7

mgL–1

Milk

[91]

Biac

ore

SPR

2.6

mgkg

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ilk[9

4]R

IA2

mgL–1

Wat

er[8

4]SP

FIA

(Par

allu

x ki

t)50

mgL–1

Bovi

ne a

nd

[92]

porc

ine

kidn

ey

Peni

cilli

n M

Biac

ore

SPR

30mg

L–1M

ilk[9

1]


Tabl

e3

(con

tinu

ed)

Ana

lyte

Imm

unoc

hem

ical

Se

nsit

ivit

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atri

xR

efer

ence

ste

chni

quea

LOD

IC50

Tetr

acyc

lines

Chl

orte

trac

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IA1

mgL–1

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Oxy

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SPFI

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Bovi

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nd

[92]

porc

ine

kidn

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ISA

(TC

Mic

row

ell

100

mgkg

–1M

uscl

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[95]

test

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Tetr

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SPFI

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lux

kit)

300

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Bovi

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nd

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porc

ine

kidn

eyEI

A20

mgkg

–1H

oney

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ELIS

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ualit

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eat

[95]

RIA

(Cha

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test

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ELIS

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1ng

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[98]

Mac

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acro

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ELIS

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8ng

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Buff

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Eryt

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RIA

10mg

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ater

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ELIS

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4ng

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ne m

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ELIS

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8ng

mL–1

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er[9

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Tylo

sin

ELIS

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cle

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Tabl

e3

(con

tinu

ed)

Ana

lyte

Imm

unoc

hem

ical

Se

nsit

ivit

yM

atri

xR

efer

ence

ste

chni

quea

LOD

IC50

Chl

oram

phen

icol

Chl

oram

phen

icol

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(Cha

rmII

ass

ay)

5ng

g–1Ti

ssue

[101

]20

ngm

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rine

[101

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A (R

IDA

SCR

EEN

)0.

5ng

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ssue

[101

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3ng

mL–1

Uri

ne[1

01]

ELIS

A3

ngm

L–1Sw

ine

mus

cle

tiss

ue[1

02]

ELIS

A3

ngm

L–1M

uscl

e ti

ssue

[102

]EL

IFA

0.7

ngm

L–1M

ilk[1

03]

Dip

stic

k EI

A17

ngm

L–1M

ilk[1

03]

ELIS

A (L

e C

arte

Tes

t)2

mgkg

–1M

eat

[104

]EI

A k

it (5

091C

AP1

p)0.

1mg

kg–1

Shri

mp

tiss

ue[1

05]

2.1Penicillins

Penicillins are one of the most important families of antibiotics used in vet-erinary and human medicine. But due to their rapid transformation in envi-ronmental media (easy hydrolysis of the b-lactam ring), their persistence in en-vironmental samples should be low. Thus, some works aimed at detectingantibiotic residues in water samples point out the absence of penicillin residuesin spite of this drug being widely used [24, 25].

Several immunochemical methodologies have been developed for the detec-tion 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 generatedat the surface of a metal conductor (usually Ag or Au) when excited by the im-pact of light of an appropriate wavelength at a particular angle (qp). Surfaceplasmons are generated by electrons at the metal surfaces that behave differentlyfrom those in the bulk of the metal. These electrons are excited by the incidentlight, producing an oscillation (resonance) of different frequency from that in thebulk of the metal film. The absorption of light energy by the surface plasmonsduring resonance is observed as a sharp minimum in light reflectance when thevarying angle of incidence reaches the critical value. The critical angle dependsnot only on the wavelength and polarization state of the incident light, but alsoon the dielectric properties of the medium adjacent to the metal surface andtherefore is affected by analytes binding to that surface (see Fig. 2). Thus, whenthe 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 penicillinsin milk using commercial monoclonal antibodies (MAb) against ampicillin[91]. These MAbs had a higher affinity for the open b-lactam ring compoundsthan for those with the closed ring. Therefore, the analyses had to be performedby 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 lightenergy 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 dependson the dielectric medium close to the metal surface and therefore is strongly affected by mol-ecules directly adsorbed on the metal surface. This principle allows the direct detection of theinteraction of the analyte and the antibody

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 cross-reactivity (CR) was observed with penicillin G, penicillin V, amoxicillin, andcloxacillin. Gustavsson et al. [94] tried to improve the procedure of the Biacoreusing carboxypeptidase and antibodies against a hydrolyzed peptide generatedby an enzymatic reaction. This method had the advantage of detecting only theintact b-lactam structure. Penicillins inhibit this enzyme and therefore theamount of penicillin present in the sample can be measured by a decrease inthe concentration of the hydrolyzed peptide. This method could be applied tothe analysis of penicillins in milk with limits of detection around few micro-grams per liter.

Moreover, several ELISA (enzyme-linked immunosorbent assay)-type im-munoassays have also been described for analyzing penicillins in different ma-trices (see Table 3). Thus, the determination of penicillins in plasma samplesfrom cattle by ELISA has been reported [93]. Monoclonal antibodies againstampicillin have been used to develop a direct ELISA and a multi-immuno-affinity chromatography method (MIAC) for penicillins [106]. Moreover, manyimmunoassay kits have become commercially available (see Table 4). Thus, theLacTek ELISA was established as a qualitative rapid prediction test of amoxi-cillin 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 oneanalysis.Analyte and an enzyme tracer compete for an antibody coated on thetube wall. After washing, a color developer is added to visualize the surface-bound complex. The color intensity is measured in a spectrophotometer andcompared with a penicillin standard indicating positive or negative results. Atthe moment there are also LacTek tests available to detect tetracyclines, sul-fonamides, and chloramphenicol.

The Charm II 6600/7600 is a semiquantitative radioimmunoassay (RIA) de-veloped 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 forthe antibody, the bound tracer–antibody complex is separated by centrifuga-tion from the unbound tracer. The complex is analyzed in a scintillationcounter for 1 min to obtain a resultant count. Samples with high-count resultsare considered negative while samples with low count are considered positive.The assay is very fast (about 10 min per sample) and the detection limit variesdepending on the antimicrobial drug and the matrix. For example, penicillin isdetected 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 theanalysis 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 im-munoreceptor assays developed for the screening of cow’s milk before milk intake at the laboratories of the receiving stations. The Beta Screen Test employsa fluorescence endpoint and takes about 10 min per assay, showing a high speci-



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ficity for penicillin. Finally, the Parallux, a solid-phase fluorescence immunoas-say (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. Thissystem has been developed as a multianalyte method to detect simultaneouslythe presence of a variety of antibiotics from different families (see Table 4). Thekit is based on the use of antibodies immobilized on four glass capillary tubespresented in the form of a disposable cartridge. Two different kinds of cartridgesare included: one consists of four tubes each containing the same range ofantibodies (the “cillins” multicartridge), so that four different samples can bescreened simultaneously. The other one also includes four tubes but each con-taining different antibodies (the “individual” cartridge), so that a positive sam-ple can be further identified. The sample is mixed with dried reagents, and antibiotic competitively binds to the coated tube; after the tube is washed andthe cartridge centrifuged, the fluorescence signal is measured with a limit ofdetection (LOD) for penicillin of 50 mg L–1. This test has also been applied to thedetection of tetracyclines, sulfonamides, cephapirin, and ceftiofur.

To our knowledge only one work has been reported on the use of a com-mercial 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 testin water samples proximal to a US farm. The LOD of the technique was 2 mg L–1.

2.2Chloramphenicol

Chloramphenicol (CAP) is a broad-spectrum antibiotic that was widely used inveterinary 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 promptedthe establishment of a zero tolerance for the presence of these residues in meatand animal products.As a consequence, many efforts have been made to developsensitive methodologies capable of detecting CAP residues or its metabolites.

Several qualitative and quantitative immunochemical methods for CAPanalysis 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 detectedCAP in swine muscle tissue with an IC50 value of 3 ng mL–1. This immunoassaywas improved and subsequently optimized incorporating the streptavidin–biotin amplification system. There are also several commercially available testkits (see Table 4). RIDASCREEN is a competitive enzyme immunoassay for thequantitative analysis of CAP residues in milk, eggs, and meat in a microtiterplate. 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 theanalysis of tetracyclines.

On the other hand, the 5091CAP1p test is a direct competitive enzyme im-munoassay for the quantitative analysis of these residues in milk, eggs, meat,


urine, tissue, and honey. In this test in a 96-well microtiter plate is coated withantibodies. The LOD is 0.2 ng mL–1 in urine and milk, with a cross-reactivityof 65% with chloramphenicol glucuronide. Veratox is a semiquantitative CAPELISA test designed to detect the presence of CAP in farm-raised shrimp. Thetest format provides a 48-well plate operating on the basis of competition between the enzyme conjugate and the antibiotic in the sample for the anti-body-coated wells.After the addition of a colored substrate, a semiquantitativedetermination can be made of the level of drug present. Results are obtainedin approximately 90 min showing low cross-reactivity to other antibiotics anda 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 tissuesand 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 valuesto 0.5 ng g–1 for tissues and 0.3 ng mL–1 in urine. One of the advantages of thesemethods 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 of20 mg kg–1. The assay was also applied to urine samples with a detectability of5 mg L–1. Moreover, Impens et al. [105] used the 5091CAP1p test for the screen-ing of CAP in shrimp tissues and Van de Water et al. [108] developed a MAb-based cleanup procedure, prior to HPLC, for the analysis of CAP in eggs andmilk. At present we have not found examples of the application of any of theseimmunochemical methods to environmental samples, but as mentioned in theintroduction, it can be assumed that application to wastewater samples shouldproduce fewer matrix effects than those produced by the biological samples.

2.3Tetracyclines

Tetracyclines are broad-spectrum antibiotics with activity against Gram-pos-itive and Gram-negative bacteria that have been widely used for the treatmentof infectious diseases in veterinary and human medicine, as well as additivesin animal foodstuffs. Normally tetracyclines are not found at high levels in theenvironment because they readily precipitate with cations such as calcium andare accumulated in sewage sludge or sediments [3, 20, 26].

Immunochemical methods have been developed and placed on the marketto analyze tetracycline residues (see Table 4). Thus, a qualitative EIA has beendeveloped and used to analyze tetracyclines in honey samples with a detectionlevel of 20 µg/kg–1 [96].A microplate-based indirect ELISA has been developedto analyze tetracyclines using polyclonal antibodies. The assay could measuretetracycline in the range between 0.1 and 6 ng mL–1. Other tetracycline antibi-otics such as chlortetracycline, rolitetracycline, or minocycline are also highlyrecognized in this assay [98]. Several immunoassay kits are commercially avail-able for the analysis of tetracyclines although, to our knowledge, none of them


has been applied to the analysis of environmental samples (see Table 4). The RIDASCREEN EIA (see above for CAP) recognizes four tetracyclines (tetra-cycline, chlortetracycline, minocycline, and rolitetracycline). It has been used to analyze milk samples with a LOD around 1.5 mg L–1 for tetracycline and15.5 mg L–1 for oxytetracycline. This kit can also be applied to the analysis ofmore complex matrices such as meat and honey. The Parallux system (see abovefor penicillins) has also become available to detect tetracyclines and sulfonamides (see below). This responds to the presence of tetracycline, chlorte-tracycline, 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 bovineand 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 inpork 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, whichwas initially developed to analyze tetracycline in serum, urine, and milk, wassubsequently adapted to analyze water samples at concentration levels around1 mg L–1. Thus, samples from hog lagoons, surface water samples, and ground-water samples were tested using the RIA method and the results confirmed byLC–MS.

2.4Sulfonamides

Sulfonamides are antibiotics widely used in animal husbandry and as feed ad-ditives. A large number of immunoassay screening methods for sulfonamidesin foods and other related complex matrices have been reported in the litera-ture [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 thatcan detect sulfathiazole in animal tissues with IC50 values ranging from 6 to21 ng mL–1 in swine liver samples. Verheijen et al. [89] described the develop-ment of a sol particle immunoassay (SPIA) for the detection of sulfadimidineresidues at the qualitative level. This kind of immunoassay is based on the useof dyed colloidal particles as labels (i.e., gold, carbon, or latex). The device herereported is a one-step strip test where the antigen is immobilized on the mem-brane, and is based on the use of affinity-purified PAbs anti-sulfadimidine-labeled with colloidal gold particles in the mobile phase. The use of coloredparticles as labels allows their use as a direct detector reagent. This test hasbeen applied to the analysis of urine and milk samples, giving positive resultsat concentrations above 10 and 20 ng mL–1, respectively. Situ et al. [87] evalu-ated the performance of a high-throughput SPR to simultaneously analyzeeight samples and also to detect sulfamethazine and sulfadiazine in porcine bilein an online system.

Most of the antibodies developed (both monoclonal and polyclonal) onlydetect individual sulfonamides. However, due to the widespread use of sulfon-


amides and the variety of congeners that can potentially be used, several at-tempts have been made to produce antibodies showing broader specificity [80, 110–112]. Thus, Spinks et al. [110] carried out molecular modeling studiesof the hapten structure in order to accomplish this aim, although no conclusiveresults were obtained. Korpimäki et al. [80] studied the use of protein engi-neering to modify MAbs so that they would recognize a wider range of sulfon-amides with similar affinities, and have achieved a significant improvement inthis 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, theLacTek, or the Charm II, have also been placed on the market for the analysisof sulfonamides (see Table 4). Thus, the Parallux detects sulfamethazine andsulfadimethoxine 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 to10 mg L–1, whereas LacTek is able to detect sulfamethazine. Moreover, the5101SUL1p and 5101SUDA1p tests reach LOD values for sulfamethazine andsulfadiazine 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 ofcare for “on-site” applications on farms. Moreover, commercially available antibodies can be found from several sources such as Silver Lake Research, USBiological, Cortex Biochem. Inc., Accurate Chemical Scientific, Fitzgerald In-dustries International Inc., and Biotrend Chemikalien GmbH.

We have found only one attempt to use immunoassays to detect sulfon-amides in environmental samples.As in the case of penicillins and tetracyclinesand also for fluoroquinolones (see below), Campagnolo et al. [84] measuredsulfonamides in water samples proximal to a farm in Iowa using a commercialCharm II RIA test, accomplishing a LOD of 5 mg L–1 for sulfamethazine.

2.5Fluoroquinolones

Fluoroquinolones have mainly been used in human medicine but more recentlysome fluoroquinolones (enrofloxacin, sarafloxacin, ciprofloxacin) are also beingemployed in veterinary medicine. Conventional methods like spectrofluoro-metric assays and HPLC have normally been used to detect fluoroquinolones inhuman samples,although some immunoassays have also been described [75–77,113] (see Table 3). Thus, the production of MAb and the development of an in-direct ELISA against sarafloxacin have been reported [76, 113]. The IC50 rangedfrom 7.3 to 48 mg L–1 depending on the MAb used. The antibodies obtained forsarafloxacin were also able to recognize enrofloxacin, difloxacin, norfloxacin,and trovafloxacin. With the same MAbs a high-performance immunoaffinitychromatography (HPIAC) was developed that consisted of the extraction ofthe analyte by using the immunoaffinity capture columns coupled online to areversed-phase liquid chromatography system. The optimized procedure wasapplied to the detection of fluoroquinolones in serum [114], chicken liver [115],


and milk [116]. The relative affinity of these immobilized MAbs toward the different fluoroquinolones allowed the gradual elution, separation, and indi-vidual quantification of two fluoroquinolones [117].

Snitkoff et al. [75] reported the development of an EIA for the detection ofciprofloxacin in serum, which was sensitive at picogram per milliliter levels ofthe antibiotic and no cross-reaction with its metabolites was observed. Gobboet al. [118] recently described the production of PAb for ciprofloxacin with theaim of detecting fluoroquinolones in Brazilian livestock. On the other hand,Bucknall et al. [77] produced antibodies for quinolones and fluoroquinoloneswith the aim of developing both generic and specific immunoassays. ELISAs forciprofloxacin, enrofloxacin, flumequine, and nalidixic acid were developed withsensitivity values around 4 mg kg–1 (on both the generic and specific assays) inbovine milk and ovine kidney.

Regarding commercially available immunochemical kits, we could mentionthe Charm ROSA Enrofloxacin Test that detects ciprofloxacin and enrofloxacinequally (see Table 4) and the 5101ERFX1p test. This last one is a direct compet-itive ELISA, which uses MAbs and has a LOD of 3 ng g–1 in tissues. Some othercompanies do have antibodies available as reagents for different applicationssuch 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 as-say is able to detect enrofloxcin as standard analyte with sensitivity levels of5 mg L–1.

2.6Macrolides

Macrolide antibiotics are macrocyclic lactones widely used in veterinary med-icine to treat diseases and infections and also as feed additives to promote an-imal growth. Some immunochemical methods have been developed to analyzemacrolides, 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 han-dling 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 sugarmoieties are well recognized by these antibodies, regardless of the presence ofneutral sugar residues. In contrast, little or no cross-reactivity was observed withthe acrocyclic lactone ring structure (tylactone) or macrolides containing onlyneutral sugar. Also a Japanese patent [120] reported the production of anti-bodies 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 elec-trochemical competitive ELISA for the detection of erythromycin and tylosin inbovine muscle. They used MAbs against these two macrolides and the activity


of an enzyme label was electrochemically measured with an electroactive sub-strate. The detection limits were 0.4 and 4 ng mL–1 for erythromycin and ty-losin, respectively. The specificity of the assay showed only cross-reactivity withroxithromycin, a macrolide derived from erythromycin.

The Charm II 6600/7600 system is the only commercial immunoassay testavailable for the detection of several macrolides in different matrices (seeTable 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 environ-mental media have been proposed such as microbiological assay and conven-tional chromatographic methods. However, only one example of the applicationof immunochemical methods to the analysis of environmental samples has beenreported [84]. In this case, erythromycin could be measured at concentrationshigher than 10 mg L–1 using the Charm II 6600/7600 assay in water samples prox-imal to livestock farms.

3Steroid Hormones

Steroid hormones are a group of biologically active compounds controlling human body functions related to the endocrine system and the immune sys-tem. Steroids are synthesized from cholesterol and have in common a cyclo-pentan-o-perhydrophenanthrene ring. Natural steroids are secreted by theadrenal cortex, testis, ovary, and placenta in humans and animals, and includeprogestagens, corticoids, androgens, and estrogens [121]. Estrogens (estradiol,estrone, and estriol) are predominantly female hormones which are importantfor maintaining the health of the reproductive tissues, breast, skin, and brain,are secreted from the ovary, and work to make an ovulatory phase of menstrualcycles. Gestagens (progesterone) act as hormone balancers, particularly ofestrogens; they are excreted from the ovary and act during the luteinizing phaseof the menstrual cycle and are involved in maturation of the endometrium ofthe uterus. Androgens (testosterone, dehydroepiandrosterone, and androstene-dione) play an important role in tissue regeneration, especially of the skin,bones, and muscles. Glucocorticoids (cortisol) are produced by the adrenalglands in response to stressors such as emotional upheaval, exercise, surgery,illness, and starvation [122]. All the steroid hormones cause their action bypassing through the plasma membrane and binding to intracellular receptors.Besides these endogenous hormones, many synthetic steroids have been pro-duced in order to take advantage of their high bioactivity. Table 5 summarizesboth natural and the most important synthetic steroids for each group as wellas their more usual applications either in veterinary or human medicine; theirchemical structures can be seen in Fig. 3.

As well as the endogenous steroids, the xenosteroids and their metabolitesare 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-



Table 5 Most important natural and synthetic steroid hormones

Endogenous Synthetic Usehormones hormones

Androgens Testosterone Boldenone Growth promoter Nandrolone in livestockMethyltestosteroneTrenboloneStanozolol

Estrogens Estradiol Ethynylestradiol ContraceptiveEstriol MestranolEstrone

Progestagens Progesterone Norgestrel Contraceptive17a-Hydroxy- Norethindroneprogesterone Melengestrol

Medroxyprogesterone

Corticosteroids Cortisone Dexamethasone Growth promoterCortisol Betamethasone Anti-inflammatory

DeoxymethasonePrednisonePrednisoloneFlumethasone

mal waste disposal. All of these compounds have been detected in effluents ofSTPs and surface waters. The water solubility of steroid conjugates is higherthan that of the free forms and they are also more easily metabolized by de-grading bacteria. The octanol/water partition coefficient of steroids is high andtherefore adsorption on sediments and suspended soils and accumulation takeplace. These compounds have a half-life of between 2 and 6 days in the envi-ronment depending on the environmental conditions and the climate [28]. Thefate and occurrence in the environment of the different steroid hormones areshown in Table 1.

The unconjugated compounds are the active forms responsible for a po-tential environmental impact, especially in the aquatic compartment, at verylow concentration levels [123]. Compounds like the estrogens have an en-docrine disrupting activity, causing adverse effects such as the induction ofvitellogenesis (plasma vitellogenin induction) and feminization of male fish[28]. For instance, concentrations between 4.7 and 7.9 ng L–1 of estradiol ledto induction of vitellogenin in juvenile rainbow trout [124]. Therefore, vitel-logenin 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 sexualorgans. The administration of androgens could be made in the free form or asesters (mainly acetate) [126]. After administration, the ester compound is hy-


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 com-pound.

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 practicesrelated to the use of the natural and synthetic sex hormones as anabolic sub-stances 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 ofgestagens are excreted per year by livestock in the EU [31]. These data don’t include the synthetic steroids, such as ethynylestradiol or norethindrone, whichare commonly used as oral contraceptives. The use of hormones, both naturaland synthetic, to enhance growth and as reproductive aids for synchronizationof the ovarian cycle, has been regulated for animal drugs because they alter thestructure or function of the animal. For these reasons, the EU has banned theuse of these compounds as growth promoters in food-production animals (Di-rective 85/649/EEC replaced by Directive 88/146/EEC). In order to detect the

use of legal or illegal natural hormones or xenobiotic drugs, and to prevent theinappropriate use of therapeutic drugs, veterinary and public health controllaboratories require efficient screening methods. Directive 96/23/EC and theDirective 2377/90/EC regulate the MRLs and the analytical methods to detectthem. Hormone implants are widely used in the USA, Australia, and Canadawhere their use is allowed. The use of progesterone, testosterone, estradiol,zeranol, and trenbolone acetate for animal food production has been regulatedby the US Food and Drug Administration (FDA) and by the Food and Agri-culture Organization of the World Health Organization (FAO/WHO).

The number of analytical methodologies currently available for determi-nation of steroids in water samples is limited. The methodologies are based on either biological or chromatographic techniques [30, 127, 128], which areusually accurate but time-consuming methodologies. Immunoassays can be extremely sensitive and often can be directly applied to the analysis of watersamples. Many examples can be found in the literature regarding immuno-chemical determination of steroid residues in biological matrices. In contrast,their application to environmental samples, and particularly wastewater, hasrarely been reported. However, as mentioned before their application to lesscomplex matrices such as aqueous samples can open the possibility to performmore 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 bedetected in the environment (see Table 1).

Although usually the antibodies and immunochemical methods developedcan 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 ofextraction. Thus, some of these columns have been prepared not only to selec-tively extract a single family of steroids, but also to extract a larger number ofanalytes using a multi-immunoaffinity chromatographic column (MIAC). Dif-ferent antibodies are linked to the solid support, allowing extraction of severalsteroids at once and therefore multianalyte-screening procedures using im-munochemical or conventional analytical methods. The convenience of theseprocedures for biological and environmental monitoring programs of vet-erinary drug residues is that often, animal treatments are performed usingcocktails of different substances (i.e., anabolic steroids with corticosteroids, orestrogens with gestagens). Dubois et al. [129] used MIAC combined withGC–MS detection for the screening of different anabolic steroids in urine andfeces from bovine specimens. Information about 12 different compounds couldbe obtained in just one run. The MIAC gel columns were prepared by mixingindividual gels prepared with different specific antibodies (against methyl-testosterone, nortestosterone, fluoxymesterone, zeranol, clostebol, ethynylestra-diol, diethylstilbestrol, and trenbolone). The fractions selectively eluted wereevaporated to dryness, dissolved in ethanol, derivatized, and injected into theGC–MS system. The MIAC gel could be regenerated with mixtures of methanol/


water, water, and finally PBS and stored at 4 °C. More examples can be found inthe literature regarding the use of multi-immunoaffinity columns, for exampleagainst different anabolic steroids [129–131].

3.1Estrogens

Due to their endocrine disrupting activity estrogens have been the most stud-ied regarding residues in the environment. Table 6 shows some examples of thelarge number of different immunological screening methods reported for es-trogens.A first RIA was described in 1985 for the analysis of ethynylestradiol inwater [33], and also a RIA was employed for the detection of 17b-estradiol inwastewater [132]. This method allows rapid, sensitive, and inexpensive screen-ing 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 waste-water effluents and surface water. The LODs accomplished levels around0.1 ng L–1 in wastewater effluents and 0.05 ng L–1 in surface waters. Results in-dicated that the concentrations of the estrogenic hormones 17b-estradiol and17a-ethinylestradiol discharged by WWTPs were comparable to those that induce vitellogenesis in fish. Some hormones appear to be removed by effluentfiltration, 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 certaindiscrepancy in the results obtained by both methods that was attributed to thefact that the concentrations measured were near to the LOD of the GC–MS/MSmethod. Another technique that has been applied to detect 17b-estradiol inwastewater is an electrochemical ELISA [133]. The activity of the label enzyme(horseradish peroxidase) was measured electrochemically using 3,3¢,5,5¢-tetra-methylbenzidine (TMB) as electrochemical substrate, accomplishing a LOD of5 pg mL–1. The interday and intraday precision (RSD) ranged from 1 to 3% andfrom 3 to 6%, respectively. Analysis of wastewater from three different treat-ment plants demonstrated the absence of matrix effects if an extraction withdiethyl ether–water was performed or the samples were just diluted 1:1 withbuffer. Validation of the method was performed by analyzing 36 samples andcomparing the results with those obtained by LC–ESI-MS/MS (liquid chro-matography–electrospray ionization-tandem mass spectrometry). The resultscorrelated very well with an R2 of 0.960.

Commonly the development of these techniques has been directed towardthe 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, isaddressed to the detection of total natural estrogens (ES): estrone (E1), estra-diol (E2), and estriol (E3). Depending on the MAb used and considering E2 asstandard analyte, several assays were developed with working ranges within



Tabl

e6

Imm

unoc

hem

ical

tech

niqu

es d

evel

oped

for

the

dete

ctio

n of

ster

oid

horm

ones

Ana

lyte

Imm

unoc

hem

ical

Se

nsit

ivit

yM

atri

xR

efer

ence

ste

chni

quea

LOD

IC50

Estr

ogen

s

Estr

adio

lEL

ISA

5pg

mL–1

Was

tew

ater

[133

]T

IRF-

IA0.

16mg

L–11.

84mg

L–1Bu

ffer

[134

]ET

IA0.

85mg

L–11.

2mg

L–1Bu

ffer

[134

]EL

ISA

0.1

ngL–1

Was

tew

ater

eff

luen

t[1

35]

0.05

ngL–1

Surf

ace

wat

er[1

35]

ELIS

A0.

1mg

L–1Bu

ffer

[136

]

Estr

iol

ELIS

A12

pg p

er w

ell

Saliv

a[1

37]

Che

milu

min

esce

nce

10pg

per

wel

lBu

ffer

[138

]EL

ISA

Estr

one

TIR

F-IA

0.07

mgL–1

0.51

mgL–1

Buff

er[1

34]

ETIA

0.5

mgL–1

0.81

mgL–1

Buff

er[1

34]

Estr

ogen

sEL

ISA

0.1

mgL–1

Buff

er[1

36]

Ethy

nyle

stra

diol

ELIS

A0.

1ng

L–1W

aste

wat

er e

fflu

ent

[135

]0.

05ng

L–1Su

rfac

e w

ater

[135

]T

IRF-

IA0.

07mg

L–11.

07kg

L–1Bu

ffer

[134

]ET

IA0.

01mg

L–12.

7mg

L–1Bu

ffer

[134

]R

IA5

ngL–1

Wat

er s

ampl

es[3

3]

aEI

A:e

nzym

e im

mun

oass

ay;E

LIFA

:enz

yme-

linke

d im

mun

ofilt

rati

on a

ssay

;ELI

SA:e

nzym

e-lin

ked

imm

unos

orbe

nt a

ssay

;ET

IA e

nerg

y tr

ansf

erim

mun

oass

ay;R

IA:r

adio

imm

unoa

ssay

;TIR

F-IA

:tot

al in

tern

al r

efle

ctio

n flu

ores

cenc

e im

mun

oass

ay.


Tabl

e6

(con

tinu

ed)

Ana

lyte

Imm

unoc

hem

ical

Se

nsit

ivit

yM

atri

xR

efer

ence

ste

chni

quea

LOD

IC50

And

roge

ns

Test

oste

rone

ELIS

A3.

9pg

mL–1

Hum

an s

erum

[139

]EL

ISA

2.5

pg p

er w

ell

Hum

an s

erum

[140

]EL

ISA

10pg

per

wel

lH

uman

ser

um[1

41]

Bold

enon

eEL

ISA

26pg

per

wel

lU

rine

[142

]0.

1ng

g–1

Faec

es[1

42]

Tren

bolo

neEL

ISA

0.1

mgL–1

Mea

t sam

ples

[143

]EL

ISA

0.1

ngm

L–1U

rine

[144

]0.

02ng

g–1M

uscl

e ti

ssue

[144

]

Nan

drol

one

ELIS

A1

ngm

L–1Eq

uine

uri

ne[1

45]

ELIS

A>

2ng

mL–1

Bovi

ne b

ile[1

46]

Ges

tage

ns

Prog

este

rone

ELIS

A0.

5nm

olL–1

Plas

ma

[147

]EL

ISA

3.8

pg p

er tu

beH

uman

ser

um[1

48]

RIA

5ng

L–1W

ater

[33]

Nor

ethi

ndro

neEI

A10

ngL–1

Wat

er[3

3]


Tabl

e6

(con

tinu

ed)

Ana

lyte

Imm

unoc

hem

ical

Se

nsit

ivit

yM

atri

xR

efer

ence

ste

chni

quea

LOD

IC50

Cor

tico

ster

oids

Cor

tiso

neD

irec

t lum

ines

cenc

e 2

nmol

L–1H

uman

blo

od p

lasm

a[1

49]

EIA

Cor

tiso

lEL

ISA

1.4

nmol

L–1Sa

liva

[150

]EL

ISA

2.8

ngm

L–1H

uman

ser

um[1

51]

Pred

niso

lone

ELIS

A0.

1ng

g–1Fa

eces

[152

]

Bet

amet

haso

neEL

ISA

12.5

ngm

L–1Eq

uine

uri

ne[1

53]

Dex

amet

haso

neEL

ISA

3.1

ngm

L–1Eq

uine

uri

ne[1

53]

EIA

0.51

ngm

L–1Pl

asm

a[1

54]

ELIF

A39

0ng

mL–1

Equi

ne u

rine

[155

]EL

ISA

4ng

mL–1

Uri

ne[1

56]

ELIS

A0.

01ng

mL–1

Uri

ne a

nd b

lood

[157

]EL

ISA

2ng

mL–1

Equi

ne u

rine

[158

]

Flum

etha

sone

ELIS

A2.

5ng

mL–1

Equi

ne u

rine

[153

]

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 forestrogens in a particular environmental sample. Coille et al. [134] described theuse of two fluorescence immunochemical methods to detect different estro-genic compounds in wastewater: TIRF (total internal reflection fluorescenceimmunoassay) and ETIA (energy transfer immunoassay). The first is an im-munosensor technique based on the evanescent wave (EW) phenomenon usingfluorescence detection. The antigen is bound to the transducer surface, whichusually is an optical fiber inside a flow cell, and interacts with a fluorescentcompound-labeled antibody. Light travels by the fiber optic by total internal re-flection and the associated EW interacts with the immobilized antigen. As aconsequence of the biorecognition reaction with the labeled antibody, a changeis produced in the features of the light that is traveling (see Fig. 4). The ETIAworks under homogeneous conditions. In this format the antibody is labeledwith a donor fluorescent dye, whereas the Ag is labeled with an acceptor dye viaa BSA molecule.When they form the complex a quenching of the fluorescenceof the labeled antibody by energy transfer is observed. In the presence of theanalyte an increase in the fluorescence is produced. The LODs obtained byTIRF for estrone, estradiol, and ethynylestradiol were 0.07, 0.16, and 0.07 mg L–1,respectively, whereas using ETIA the corresponding detectability accomplishedwas 0.5, 0.85, and 0.01 mg L–1.

Immunoaffinity purification procedures have also been described for es-trogens and applied to environmental and biological samples [160–163]. Fer-guson et al. [160] described a method, based on immunoaffinity extractioncoupled to LC–ESI-MS, for the determination of the steroid estrogens b-estra-diol (E2), estrone (E1), and a-ethynylestradiol (EE2) in wastewater. The use ofhighly selective immunosorbents for sample preparation prior to the analysisallowed the removal of interfering sample matrix components present in thewastewater extracts that would otherwise cause severe ionization suppressionof the estrogens during the electrospray process. The authors claim that the useof immunoextraction removed much of the isobaric noise from the selected ionmonitoring chromatograms, increasing the signal-to-noise ratios and improv-


Fig. 4 Waveguide evanescent wave (EW) principle. Light is propagated through the wave-guide (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 attenua-tion in the reflected light of the waveguide

ing the detectability of the analytical method (0.18 and 0.07 ng L–1 for E2 andE1, respectively). The optimized method was applied to the analysis of estro-gens in two wastewater effluents. Recoveries of E2 and E1 were excellent(>90%), while EE2 was not retained (recovery <2%) from effluent extracts dueto its structural differences with the immunizing hapten. The precision of themethod was high, with relative standard deviations below 5%. The concentra-tions of E2 found in wastewater were 0.77–6.4 ng L–1, while levels of E1 werehigher (1.6–18 ng L–1). Farjam et al. [163] developed an immunoaffinity precol-umn (immuno-precolumn) immobilizing antibodies directed against estrogensteroids on Sepharose. They evaluated different desorbing techniques, suitablefor online coupling to an HPLC-UV system. The most effective approach used95:5 methanol–water mixtures, although the use of cross-reactants to elute thetarget was also considered. The final system consisted of a column-switchingunit allowing preconcentration of the samples on an immunoaffinity sorbent.After elution the analytes were concentrated on a C18-bonded silica precol-umn, and then separated on a C18-bonded silica analytical column. The min-imum concentration of estrogens detected in urine using this system wasaround 200 ng L–1 with a repeatability of 6–8%. The total analysis time was45 min, which gave an estimation of about 30 analyses per day in this auto-mated unattended method [164].

Besides these works, there are a lot of commercially available immunoas-says. Their main application is directed toward clinical analysis and food qual-ity. Table 4 shows the most representative kits that can be found nowadays onthe market.

3.2Androgens

The main areas of application of immunochemical techniques for androgensis doping control in athletes, forensic chemistry, farm animals for human con-sumption, and food analysis [165–167]. Immunochemical methods for andro-gen detection have been applied to a great variety of matrices (see Table 6);however, to our knowledge, their application in the environment field has notyet been recorded. One of the anabolic steroids widely used is 19-nortestos-terone. Different ELISAs have been developed for the analysis of this compoundin feeds, food from animal origin, and for doping control [145, 146]. Similarly,ELISA methods have also been developed for other anabolic steroids such asboldenone and trenbolone, used as growth promoters [142, 144]. In the case oftestosterone, several immunochemical studies have been addressed to establishthe real physiological levels of this hormone in different animals.A highly sen-sitive 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 inhuman plasma [140, 141]. Thus, Rassie et al. [140] developed a PAb-based direct


immunoassay performed in microplates with a sensitivity of 2.5 pg per well. Theassay was very specific for testosterone and did not show any cross-reaction withother related C19 steroids tested. Replacement of immunoassay plates by poly-propylene tubes raised the detection limits to 25 pg per tube, but improved therange of testosterone that could be measured up to 10,000 pg. Similarly, Rao et al. [141] developed a direct immunoassay on microplates using PAbs andpenicillinase as tracer. The LOD was 10 pg of testosterone with a dynamic rangebetween 15 and 1,000 pg. No interference was produced by other common an-drogens, estradiol, or progesterone, whereas a low level of cross-reactivity with5a-dihydrotestosterone (6.2%) and 11b-hydroxytestosterone (1%) was observed.

Specific antibodies for androgen compounds have also been used to developimmunoaffinity columns in order to include a purification step prior to theanalysis of androgens such as stanozolol [168], methyltestosterone [169], testos-terone, trenbolone, and nortestosterone [170]. In the case of methyltestosterone,a comparison made between XAD solid-phase extraction and immunoaffinityprocedures showed that immunoaffinity could be more efficient in isolating andconcentrating anabolic steroids from complex matrices such as urine and serum[169].An immunoaffinity precolumn packed with Sepharose-immobilized PAbsagainst 17b-19-nortestosterone (b-19-NT) was used for the selective online pretreatment of raw extracts of urine, bile, and tissue samples followed byHPLC-UV detection (247 nm) [171]. b-19-NT and its metabolite 17a-19-nortestosterone (a-19-NT) could be detected in these biological samples withdetection limits around 0.05 mg kg–1. Using the same immunosorbent 17b- and17a-trenbolone were also detected. By percolating high sample volumes it waspossible to confirm these results by GC–MS. Farjam et al. [172] also evaluatedan immunoaffinity purification procedure coupled to GC. The immunosorbentcontained antibodies raised against the synthetic steroid hormone b-19-NT. Theonline connection between the immunoaffinity precolumn and the capillary GCwas performed with an interface that consisted of a 10¥2-mm reversed-phaseprecolumn and a diphenyltetramethyldisilazane-deactivated GC retention gap.After preconcentration on the immunoaffinity precolumn the analytes wereeluted and reconcentrated on the reversed-phase precolumn. Subsequently, thisprecolumn was desorbed with 75 mL of ethyl acetate, which was directly intro-duced into the retention gap by using partially concurrent solvent evaporation.The system allowed the automated pretreatment and GC analysis of 19-nor-steroids at the nanogram per liter level.

A variety of immunoassay test kits for androgens are available from com-mercial sources. Table 4 indicates some of the most important companies com-mercializing these assays.

3.3Gestagens

As with estrogens and androgens, several commercial immunoassay kits arenowadays available for the analysis of gestagens [173–175] (see Table 4). Sim-


ilarly, different research groups have also made efforts to develop assays and todemonstrate the performance of those assays in a variety of sample matrices(see Table 6). Thus,Aherne et al. described different immunoassays for detect-ing natural and synthetic steroids in water [33]. Norethindrone and proges-terone were detected at concentrations above 10 and 5 ng L–1 using an EIA anda RIA, respectively. Results below the level of detection were obtained in all thesamples examined (eight river samples and six potable supply samples), exceptfor two river samples that contained norethindrone (17 ng L–1) and one riversample and one potable water sample that were positive for progesterone(6 ng L–1). They concluded that the presence of norethindrone in river waterwas caused by the low biodegradation of norethindrone in the usual 6-h watertreatment processes. Thanks to these studies the authors found that a 24-htreatment of the sludge system was necessary.

3.4Corticosteroids

Corticosteroids are synthetic glucocorticoids that produce a strong anti-in-flammatory effect. Corticosteroids such as dexamethasone are commonly usedin veterinary practice for the treatment of illnesses such as respiratory and gastrointestinal disorders. However, corticosteroids are also used illegally asgrowth promoters in animal feed. The misuse of glucocorticoids in livestockproduction occurs, sometimes in combination with b-adrenergics in mixturesor cocktails to achieve growth promotion of food-producing farm animals. Theaim is to reduce meat fat, to increase the appetite of the animals, and to increasethe efficiency of the use of b-agonists. For consumer health and safety, the useof these compounds for this purpose is banned within the EU (Council Direc-tive 86/469/EEC). Their therapeutic use is also regulated by the MRLs that havebeen established for dexamethasone, betamethasone, prednisolone, andmethylprednisolone in different tissues (EC Regulation 2377/90).

Owing to their high potency they are very effective in low doses, which re-sults in low residue levels in biological matrices. Therefore, there is a require-ment for sensitive analytical methods for the quantification and confirmationof these compounds in biological samples. Previously GC–MS methods havebeen used for the analysis of corticosteroids. However, recent developments in liquid chromatography–mass spectrometry technology have led to reportson the use of LC-based approaches for analysis of these compounds. The useof LC–MS reduces the analysis time due to elimination of the lengthy deriva-tization or oxidation procedures necessary for GC–MS methods. Efficientscreening procedures based on ELISA methods have been described for themost important corticosteroids (see Table 6) [153, 176]. Rodriguez et al. [153]reported an ELISA using PAbs raised against flumethasone that also recognizedseveral synthetic corticosteroids such as dexamethasone (CR=80%) and be-tamethasone (CR=20%), while endogenous corticosteroids such as cortisolgave very low cross-reactivity (<0.5%). This assay can be directly applied to



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

1:10 diluted urine samples without hydrolysis of glucuronide or sulfate conju-gates or any other treatment of samples. The PAbs were obtained by immuniz-ing sheep with a flumethasone derivative linked to human serum albumin.Flumethasone, dexamethasone, and betamethasone could be detected at levelsaround 2.5, 3.1, and 12.5 ng mL–1, respectively. Similar results were obtained byRoberts et al. [158] with antibodies raised against dexamethasone using 21-hemisuccinate 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 stronglydetermines the specificity of the resulting antibodies. However, considering thestructural similarities, the presence of common epitopes determines the highcross-reactivity observed in these assays (see Fig. 5 for chemical structures).

A direct enzyme immunoassay for screening of synthetic glucocorticoids inbiological samples was also developed by Meyer et al. [154], by raising antibod-ies against dexamethasone using the same immunizing hapten as Roberts et al.(21-hemisuccinate-BSA) and using prednisolone-21-hemisuccinate horseradishperoxidase as tracer. The system had an analog cross-reactivity pattern withsimilar CR values: dexamethasone (I) (100%), flumethasone (103%), betametha-sone (45%), triamcinolone (18%), and prednisolone (17%). The natural gluco-corticoids cortisone and cortisol, the gestagens progesterone and pregnenolone,and the androgen testosterone were not recognized (CR<0.4%). Recently, there


Table 7 Cross-reactivities for selected corticosteroids using IAC–HPLC (reproduced from[180] with editor permission)

Corticosteroid Relative cross-reactivity (%)

IAC–HPLC ELISA

Dexamethasone 100 100Flumethasone 96 96Betamethasone 30 37Deoxymethasone 27 21Cortisol 0 1Prednisolone 0 4

Antiserum batch: AD60.

has also been an increasing interest in the use of saliva to detect drugs. In thiscontext, Anfossi et al. [150] described the use of an ELISA for cortisol thatachieves a LOD of 1.4 nmol L–1 in this matrix and recoveries from spiked sam-ples of between 80 and 120%.

An ELIFA (enzyme-linked immunofiltration assay) method has also beenreported [155] for the rapid detection and semiquantification of dexametha-sone in equine urine samples. The assay consists of an indirect competitiveELISA in which dexamethasone in standards or samples competes for the an-tibody binding with a dexamethasone–protein conjugate immobilized as a spoton the surface of a cellulose nitrate filter. The sheep anti-dexamethasone anti-bodies are complexed with an alkaline phosphatase-labeled second antibody.The filtration system allows rapid washing and incubation steps, so the signalcould be visualized in just 15 min by an insoluble colored dye as a spot on thefilter at the site of the immobilized drug–protein conjugate. The assay has aLOD of 390 ng mL–1 for a visual endpoint in which the color intensity of spotsdeveloped in the presence of samples is compared with those of standards.Twelve filters can be processed in a single batch consisting of two standards andten samples.

Immunoaffinity procedures have also been developed to selectively extractcorticosteroids from different sample matrices. Thus, Seymour et al. demon-strated the higher efficiency of the immunoaffinity methods compared with theconventional extraction procedures using organic solvents [177]. Immuno-sorbents have also been used for online procedures followed by HLPC-UV [178,179], HPLC–APCI-MS [179, 180], GC–MS [176, 181], or capillary electrophore-sis [182]. Poly(hydroxyethyl methacrylate) (HEMA) was evaluated as a supportmaterial for the anti-dexamethasone antibodies used in IAC. The online IAC–HPLC–MS allowed determination of dexamethasone and flumethasone inequine 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 arepresented in Table 7. Bagnati et al. developed an immunoaffinity extraction

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

4Other Drugs

This section is dedicated to providing information on the immunochemicalmethods available today to determine drugs that do not belong to the abovegroups, but that years of unrestricted emission to the environment require tobe considered [183]. From the broad range of pharmaceuticals that can reachthe environment, drugs such as analgesics and nonsteroidal anti-inflamatorydrugs (NSAIDs) are regularly employed, often even without prescription. Onthe other hand, cytostatic agents are of concern not because of their productionvolume but for their high pharmacological potency.

In Germany for instance the total quantities of acetylsalicylic acid sold peryear have been estimated to be greater than 500 tons, 75 tons for diclofenac, and180 tons for ibuprofen [35]. The same occurs in other EU countries where com-mon drugs such as paracetamol or aspirin are sold in quantities comparable tohigh 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 Den-mark 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 factthat they are mainly utilized in the hospital sector and by their intrinsic muta-genic 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 antineoplasticagents were identified by Christensen as the three most relevant groups ofchemicals concerning their potential human risk as a consequence of drug ex-posure via the environment [10]. Immunochemical methods for hormones andantibiotics have already been discussed above. In this section we will describemethods 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 excretedinto wastewater, enter the aquatic environment, and eventually can reach drink-ing water if they are not biodegradable or eliminated during sewage treatment.Data on the environmental occurrence of the pharmaceuticals treated in thissection are found in Table 1.


4.1Analgesics and NSAIDs

Acetylsalicylic acid (aspirin) is still the most widely used analgesic, anti-in-flammatory, and antipyretic agent followed by paracetamol. Ibuprofen [2-(4-isobutylphenyl)propionic acid] and diclofenac (diclofenac-Na) from theNSAID group are used extensively for the treatment of rheumatic disorders,arthritis, pain, and fever. Fentanyl is a very strong opioid with analgesic prop-erties 80 times stronger than those of morphine. The narcotics law thereforeregulates 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 illicitlyfor many years.

Most of these substances have been shown to be readily or inherently bio-degradable [11, 18, 39, 42]. Photodegradation was identified as the main elimi-nation process of diclofenac in lake water [4, 42].With a relatively high sorptioncoefficient 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 effluentsanalyzed by Richarson et al. [24]. The same authors reported that they foundthese 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 ofthese 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 forthe control of drug abuse and assessment of intoxication using body fluids,tissue extracts, post-mortem specimens, and seizure samples. For this reasonthere 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 chromatog-raphy-based immunoassays with laser-induced fluorescence detection havebeen used for the determination of salicylate and gentisic acid in urine [187].Similarly, Wey et al. [196] developed two rapid, competitive binding, elec-trokinetic capillary-based immunoassays recognizing urinary opioids (COD,codeine-6-glucuronide, dihydrocodeine (DHC), dihydrocodeine-6-glucuronide,MOR, morphine-3-glucuronide, and ethylmorphine (EMOR)).Aliquots of urineand immunoreagents of a commercial, broadly cross-reacting polarizationfluoroimmunoassay (PFIA) for opiates were combined and analyzed by capil-lary zone electrophoresis or micellar electrokinetic capillary chromatographywith laser-induced fluorescence detection. Assay sensitivities for COD and



Tabl

e8

Som

e re

pres

enta

tive

imm

unoc

hem

ical

tech

niqu

es d

evel

oped

for

the

dete

ctio

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anal

gesi

cs,N

SAID

s,an

d cy

tost

atic

age

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lyte

Imm

unoc

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ence

ste

chni

quea

LOD

IC50

Ana

lges

ics

Para

ceta

mol

M

ECC

-IA

<20

mgm

L–1Se

rum

[186

](a

ceta

min

ophe

n)

Salic

ylat

eM

ECC

-IA

Seru

m[1

86]

Salic

ylic

aci

d (S

A)

CE-

IA-L

IF10

mgm

L–1(S

A)

Uri

ne[1

87]

Gen

tisi

c ac

id (G

A)

CE-

IA-L

IF5

mgm

L–1(G

A)

Uri

ne[1

87]

Salic

ylic

aci

dEL

ISA

0.39

mmol

L–1Pl

ants

[188

]

Ibup

rofe

nEL

ISA

3.62

ngm

L–1Bu

ffer

[189

]EL

ISA

100

pg p

er a

ssay

Buff

er[1

90]

Dic

lofe

nac

ELIS

A6

ngL–1

60ng

L–1Pu

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nd s

urfa

ce

[191

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ELIS

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25ng

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Hum

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[193

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ISA

0.5

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94]

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Tabl

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(con

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ed)

Ana

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uman

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96]

mer

cial

rea

gent

s)

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ood

[197

]0.

200

mg

L–1Bi

le[1

97]

0.10

0m

gL–1

Tiss

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97]

ELIS

A40

0pg

mL–1

Equi

ne b

lood

,uri

ne[1

98]

ELIS

A10

0pg

mL–1

Buff

er[1

99]

ELIS

A10

0pg

mL–1

Uri

ne[2

00]

Cyto

stat

ic (a

ntin

eopl

asti

c)

Met

hotr

exat

eR

IA6.

25ng

L–1W

ater

sam

ples

[33]

EIA

50pg

mL–1

Seru

m[2

01]

ELIS

A5¥

10–1

2g

mL–1

Buff

er[2

02]

CE-

IA-L

IF5

pgBu

ffer

[203

]

MOR were comparable (10 ng mL–1), whereas those for DHC and EMOR wereabout fourfold lower. Furthermore, glucuronides were shown to react like thecorresponding free opioids.Validation with real urine samples was performedwith identification of the peaks by capillary electrophoresis–ion-trap massspectrometry (CE–MS) after solid-phase extraction.

A PFIA, commercialized by Abbot, is one of the most common immuno-chemical 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 hypersensi-tivity reactions, an enantioselective immunoaffinity extraction method hasbeen developed that specifically isolates peptide fragments that have beenmodified with optically active ibuprofen. The antibodies were obtained by immunizing rabbits with (S)-ibuprofen coupled to BSA through a b-alaninegroup. The elicited antibody strongly recognizes the asymmetric center and theisobutylphenyl moiety of (S)-ibuprofen and its conjugates.A 0.5-mL aliquot ofthe immunosorbent (11.5 mg of IgG per mL gel) prepared by immobilizationof the antibody was capable of retaining up to 1 mg of (S)-ibuprofen [190].

The immunochemical methods available today for the analysis of analgesicsor NSAIDs should be easily adaptable to the analysis of environmental samples,although few examples have been reported. In this context, an indirect ELISAhas been developed and applied to the determination of diclofenac in tap water, surface water, and wastewater samples [191]. The authors used a diclo-fenac-BSA conjugate as immunogen to produce antisera. The ELISA showed aLOD of 6 ng L–1 in buffer.A greater recognition for the glucuronide conjugateswas observed. In order to validate the assay the results obtained were comparedwith those from GC–MS.

4.2Cytostatic Agents

Cytostatic drugs are frequently used in chemotherapeutic treatments. Residuesof these substances should exclusively occur in hospital sewage at low concen-trations. Among the cytostatic agents more frequently employed we shouldconsider: methotrexate (MTX) (4-amino-10-methylfolic acid), a folic acid an-tagonist; the alkylating antineoplastic drug cyclophosphamide is one of the old-est known cytostatics and is one of the most frequently used agents in cancerchemotherapy; and ifosfamide is a widely used antitumor agent. Cytostaticagents fall far below the quantitative importance of other drugs. However, seenfrom the potential ecotoxicological impact viewpoint, they are an importantgroup of drugs with a high potential risk for humans and wildlife. Althoughtheir effects against higher aquatic organisms such as fish or algae have onlybeen seldom investigated [1], their carcinogenic, mutagenic, and embryotoxiceffects are clearly demonstrated [204]. Most of the active substances investigatedproved to have a low biodegradability. Therefore, the active substances are ex-


pected to pass unchanged through municipal STP and thus reach surface waters[1] when they are not eliminated by adsorption onto sewage sludge. Steger-Hart-mann 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 wasdetected 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]. Toour knowledge cytostatics have not been detected in surface waters but theyhave an estimated PEC of 0.8 ng L–1 [1, 7, 46].

Few examples of immunochemical methods for cytostatic agents have beenreported (Table 8).Within the context of work performed by Aherne et al. [33],on the use of immunochemical methods in the analysis of microcontaminantsin water samples, was reported the use of a RIA for the detection of methotrex-ate 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 andpotable water) were negative.

Ferrua et al. [201] developed an EIA with enzyme-labeled Ab and an analogantigen of MTX bound to polystyrene spheres through a methylated bovine albumin carrier. Serum samples of treated patients were analyzed, and goodagreements with other methodologies developed to measure MTX were ob-tained. Recently, the use of commercial antibodies for MTX for the developmentof an immunoassay by capillary electrophoresis with laser-induced fluorescencedetection has also been described [203], achieving good sensitivities. The pro-cedure includes an initial competition step between the immunoreagents andthe analyte followed by the separation of the species and detection, both stepscarried out simultaneously with CE. The rapid separation allows the reductionof the time per assay to a few minutes.

5General Summary

Immunochemical methods have been developed or are commercially availablefor the analysis of the most important groups of pharmaceuticals with an in-creased human risk when in contact with the environment. However, additionalwork is necessary in order to expand the number of families of compoundsthat can be detected using these methods, and especially to adapt them to theanalysis of environmental samples. Considering the complexity of the bio-logical matrices, there is great promise regarding their potential application tothe analysis of water samples. Within the advantages of implementing thesemethods for environmental monitoring purposes are their simplicity and thehigh sample processing capabilities. However, we must consider the low con-centrations expected in the environment even though, due to their high bioac-tivity, these levels may be sufficient to cause adverse effects on the ecosystem.


Acknowledgements This work has been supported by CICYT (BIO2000-0351-P4-05,AGL2001-5005-E) and by the EC: nanotechnology and nanosciences, knowledge-based multifunctionalmaterials, new production processes and devices (contract number NMP-505485–1).

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244 Immunochemical Determination of Pharmaceuticals and Personal Care Products

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 Measuring, Why and How? . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2462.1 The Main Practical Objectives of Measuring . . . . . . . . . . . . . . . . . . . 2472.2 Sampling Versus On-Site Measurement . . . . . . . . . . . . . . . . . . . . . 2482.3 Types of Results . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2492.4 Analytical Characteristics and Measurement Objectives . . . . . . . . . . . . 2502.5 Parameters and Substances . . . . . . . . . . . . . . . . . . . . . . . . . . . . 251

3 On-Site Measurement: From Needs to Solutions . . . . . . . . . . . . . . . . 2513.1 Minimizing the Measurement Error . . . . . . . . . . . . . . . . . . . . . . . 2513.2 Taking Account of Variability . . . . . . . . . . . . . . . . . . . . . . . . . . . 2523.3 Preventing Sample Aging . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2533.4 On-Site Measurement Constraints and Solutions . . . . . . . . . . . . . . . . 254

4 Parameters and Substances Monitored by On-Line Systems . . . . . . . . . . 2574.1 On-Line Monitoring:

From Laboratory-Transposed Methods to Software Sensors . . . . . . . . . . 2574.2 On-Line Monitoring Systems . . . . . . . . . . . . . . . . . . . . . . . . . . . 2584.2.1 Physical and Aggregate Properties . . . . . . . . . . . . . . . . . . . . . . . . 2584.2.2 Inorganic Constituents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2584.3 Organic Constituents . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2614.3.1 Aggregate Properties . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2614.3.2 Specific Organic Constituents . . . . . . . . . . . . . . . . . . . . . . . . . . 2634.4 Toxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2644.5 Non-Parametric Measurement for Detection of Accident or Disturbance . . . 266

5 Validation and Developments . . . . . . . . . . . . . . . . . . . . . . . . . . 2665.1 Systems Validation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2665.2 Developments . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2675.2.1 Optical Techniques . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2675.2.2 Biosensors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2685.2.3 Software Sensors . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 269

References . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 270


Abstract Sampling and laboratory analysis are not well adapted to wastewater quality mon-itoring in a process control or hazards prevention context, for which on-line/on-site mea-surement is preferable. Before considering the implementation and constraints of on-linesystems, 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 ofparameters and substances to be monitored, for regulation purposes for example, only a fewof them are measurable with on-line devices.

Keywords Wastewater quality · On-site measurement · On-line monitoring · Emerging pollutants

AbbreviationsBOD Biochemical oxygen demandCOD Chemical oxygen demandCSO Combined sewer overflowORP Oxido-reduction potentialPAH Polycyclic aromatic hydrocarbonsSAC Spectral absorption coefficient (UV 254)TKN Total Kjeldhal nitrogenTOC Total organic carbonTSS Total suspended solidsUV UltravioletUV/UV UV degradation + detectionVOC Volatile organic compoundsWW Wastewater(s)WWTP Wastewater treatment plant

1Introduction

We all agree that the continuous on-line/in situ detection of pollutants in waterand wastewater should be the best practice for true quality monitoring. This isparticularly relevant for the monitoring of emerging pollutants if we considerthat 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 inorder to be sure that all potential users or developers of on-line/in situ measure-ment systems do know the main limits and constraints of the exercise.

2Measuring, Why and How?

First of all, let us explain that the term measurement used in this chapter mustbe considered in its generally accepted meaning, including all efforts carried outfor giving end-users the quantitative or qualitative result that is more often use-

246 O. Thomas · M.-F. Pouet

ful for a decision-making process. This obviously includes all analytical systemsbut also all test kits, field-portable devices and on-/off-line instruments.

2.1The Main Practical Objectives of Measuring

In the domain of wastewater quality monitoring, the main reason for measur-ing is related to regulation compliance or contractual needs, in order to checkif concentrations of substances or parameter values are below the thresholdlimits, for example at the outlet of a treatment plant, before discharge. More-over, wastewater quality monitoring programmes are also planned for somespecific sewer parts, such as combined sewage overflow (CSO) or industrialconnections. For this application, the monitoring procedure is more oftenbased on the use of the automatic sampling–laboratory analysis scheme. Flowmeasurement is generally coupled to wastewater quality measurement not onlyfor sampling assistance, but also for daily load calculations, complementary toconcentrations or parameter threshold values.

The second main reason for wastewater quality monitoring is related toprocess control, particularly for treatment plants where analysers and sensorsare generally used with physico-chemical or biological reactors, including set-tling tanks. This application is mainly encountered for important wastewatertreatment plants, either urban (majority domestic) or industrial, where thestorage capacities are rather small with regard to the flow to be treated. Obvi-ously, on-line systems are preferable in this case, but the available instrumentsoften 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 ex-ample, or to prevent potential toxic effects on the receiving medium. This ap-plication is mainly found in industrial contexts where the presence of toxic pol-lutants may occur. In this case, on-line systems are obviously preferable forreal-time warning.

The last reason is for improving the scientific knowledge of wastewater qual-ity. 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 considerationsand economic reasons, but somewhat less for wastewater quality itself, due topoor interest. However, it is well known that wastewater constitutes a hugeproblem, even in developed countries, for environmental protection fromchemical (including emerging pollutants) and other sanitary (pathogenic) risksto human health. This is the reason why research on wastewater quality mustbe encouraged with the development of suitable on-site measurement proce-dures.

Wastewater Quality Monitoring: On-Line/On-Site Measurement 247


Fig. 1 Measurement procedures for wastewater quality monitoring

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

Sampling On-site

Regulation compliance ✓Process control ✓Hazards prevention ✓Scientific knowledge ✓ ✓

2.2Sampling 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 clas-sical procedure recommended and even required in official texts for regulationmonitoring, based on sampling and laboratory analysis, including several stepsbetween sampling and analysis: conditioning, storage, transportation and pre-treatment. The other procedure, carried out on site, is based on the existenceof on-line measurement systems or on the use of field-portable devices or testkits. Actually, the two approaches are often combined, taking into account either the scientific relevance of some practices (e.g. on-site measurement ofdissolved gases and temperature), or the availability of systems for on-linemonitoring. The constraints and methodology related to this last procedurewill be explained in more in detail in the following sections.

With respect to the measurement objectives, the two procedures are notequivalent, depending on the purpose (Table 1). If it is obvious that the classi-cal procedure will still be preferred (at least because of the lack of suitable on-line systems), the relevant control of the treatment process cannot be envisagedwithout on-line monitoring. In fact, the two procedures might generally be con-sidered as complementary for most applications.

2.3Types 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 otherapplications of the analytical sciences (in the health sector, the use of test kitsand biodiagnostic systems leads to quick and useful information, often far froma classical analytical result). In fact, quantitative analysis gives the concentra-tion 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 aggre-gate (BOD, COD, toxicity, ...) parameter. In this context, total indices are oftenproposed 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 quicktests or kits designed for on-site use. The results obtained are then of a semi-quantitative nature, often relying on a number of reagent drops or a colourchange. The sensitivity of the kits is very coarse, depending on the scale ofresponses. This approach is very interesting for on-site studies from grabsampling, as it can provide assistance for focusing on sites of interest (orsome 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 toa main assumption, as for example: is the concentration lower than a thresh-old value? This method is interesting for the detection of unknown pollutantsand can be easily carried out with screening procedures coupling real qual-itative analysis with semi-quantitative responses.

– A third mode of qualitative measurement is a non-parametric one [2]. Thisconcept is based on the direct exploitation of an analytical signal (absorbance,intensity, potential, ...) without parameter calculation, leading to the simplecharacterization of the studied sample (Fig. 2). For example, fingerprintingor image analysis can be considered as non-parametric measurement.

With respect to wastewater quality monitoring, quantitative and/or qualitativeresults can be chosen (Table 2).


Table 2 Adequacy between measurement needs and types of results

Quantitative Semi- Qualitative Non-quantitative parametric

Regulation compliance ✓ – – –Process control ✓ ? ✓ ✓Hazards prevention ✓ ✓ ✓ ?Scientific knowledge ✓ ✓ ✓ ✓

2.4Analytical Characteristics and Measurement Objectives

If we try to refine the adequacy between the measurement procedures and thepractical needs for wastewater quality monitoring, different metrological (ana-lytical) characteristics have to be considered, such as detection limit, reliabilityand robustness (Table 3). Even if it is very difficult to compare the analyticalmethods carried out in the laboratory with on-site measurements (with on-lineor tests kits), this presentation points out the main features of the measurementrequired for different needs. These characteristics define the quality of theavailable information [3], which constitutes one of the major problems that


Fig. 2 Non-parametric measurement concept

Table 3 Measurement needs versus analytical characteristics for wastewater monitoring

Regulation Process Hazards Scientific compliance control prevention knowledge

Low detection limit, ✓ ✓SensitivityRapidity ✓ ✓Reliability ✓Simplicity ✓ ✓ ✓Relevance ✓ ✓ ✓ ✓RobustnessAvailabilitya ✓ ✓ ✓

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

analysts have to face.A rapid response, even coarse, is sometimes preferable fordecision making.

In fact, there exists a great difference between the two types of needs interms of traceability. On the one hand, the regulation compliance need andmore often the scientific need require us to be as confident as possible aboutthe trueness of the results (or at least the closeness between the results and truevalues) and to store the data for further exploitation. On the other hand, processcontrol and hazards prevention are based on the real-time exploitation of theresults.

2.5Parameters and Substances

A lot of substances and components are present in wastewaters and can be mea-sured, especially the emerging pollutants. However, in practice, the aggregate pa-rameters (BOD, COD, TSS, ...) and the physico-chemical ones (temperature, pH,dissolved oxygen, conductivity, turbidity, ...) are more often monitored. The onlyspecific 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, thechoice of which is guided by environmental considerations (mainly risks forhealth), then surfactants, endocrine disruptors, pesticides, other industrial organics (PAH, aromatic amines, ...) or inorganics (sulphides, arsenic, ...) andmicrobiological indicators (pathogens) must also be considered.

3On-Site Measurement: From Needs to Solutions

The previous chapters have shown that the classical procedure based on sam-pling and laboratory analysis is not suitable for the majority of cases, especiallyin 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 areother reasons for considering on-site (on-line) measurement. Experimental errors related to the numerous steps between sampling and analysis, mediumvariability with space and time, and sample aging are some good reasons.

3.1Minimizing the Measurement Error

The usual way to get information on wastewater quality is first sampling usingan autosampler and then transportation of samples to the laboratory for analy-sis. Between sampling and analysis, several steps are needed: storage/condi-tioning, transportation, preparation (filtration, pre-concentration, cleanup, ...).



Table 4 Comparison between technical steps used for the three main measurement proce-dures

On-line/off-line In situ Classical(grab sample)

Sampling –/✓ ✓ ✓Storage ✓Transportation ✓Preparation ✓? ✓Analysis ✓ ✓ ✓Transcription/transmission ✓ ✓ ✓

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 canobviously be avoided (Table 4). By minimizing the number of technical stepsbetween sampling (or even on-line sensing) and the analytical result, the globalerror of measurement will be reduced.

3.2Taking Account of Variability

Wastewater is usually a very complex medium as explained further, the com-position of which (urban or industrial) is changing with time and space. Iftime variability is well known for the load variation between night and day forurban wastewater or between weekdays and weekends for industrial waste-water, then space variability has been less studied.We easily understand that foran urban wastewater network, the composition of the medium will change withindustrial discharge or with the presence of a hospital for example (these establishments being responsible for emerging pollutant discharges dependingon the efficiency of the wastewater treatment, if any). But space variability caneasily be studied inside industrial sites [2]. Table 5 shows the evolution of thequalitative variability of a refinery wastewater network, decreasing from up-stream units (desalter) to the outlet of the treatment plant (biofilter). This studywas based on the use of UV spectrophotometry for the estimation of phenolsand sulphides [9] and on the study of hidden isosbestic points in UV spectra


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

Effluents Comments Normalized UV spectra

Inlet of the desalterpH 7.6, – Mixture of outlets ofAmmonium 49.8, different strippers.Sulphide 12, – UV spectra veryCOD 2000, structured.Phenols 200, – lHIP=248 nmTSS 100, – V=70.8%Q 50 m3/h

Outlet of a storage basinph 8.4, – UV spectra fewAmmonium 46.4, structured.Sulphide 1.7, – lHIP=224 nmCOD 480, – V=44.4%Phenols 9.1,TSS 59,Q 30 m3/h

Outlet of the biofilterph 7.1, – Few structured spectra.Ammonium 37.8, – Presence of nitrate.Sulphide 0.1, – lHIP=224 nmCOD 83, – V=7.7%Phenols 0,TSS 15,Q 200 m3/h

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

sets [10]. Taking account of time and space variability, on-site measurement isthe only suitable procedure (unfortunately limited by the available systems).

3.3Preventing Sample Aging

Another factor to be considered is that wastewater is rarely stable in its compo-sition after sampling. Mainly due to physico-chemical or biological transfor-mation, 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, surfactantsplay a major role in aggregation/adsorption phenomena (Fig. 3). This explainswhy 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].


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

A biologically treated wastewater is more stable than a raw one or a physico-chemically treated effluent. Before biological treatment, raw urban wastewatermay show strong variation in its composition due to adsorption biodegradationeven at low temperature. This problem depends on the nature of the wastewater(a raw urban one being generally less stable than an industrial wastewater, exceptfrom the food industry), but seems to be sufficiently important for a lot of sub-stances and parameters possibly to be transferred from dissolved to colloidal orsolid phases between sampling and analysis, with a risk of adsorption–com-plexation–release for metallic compounds, and degradation into by-productsfor organics. This reinforces the interest in on-site measurement.

3.4On-Site Measurement Constraints and Solutions

On-site measurement thus seems to be the optimal solution for wastewaterquality monitoring. However, as seen previously, wastewater composition isvery complex and varyies with time and space. The implementation of on-sitemeasurement must take into account some constraints related to the risk ofsampling 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 ofsystem used.

The principal factor of complexity is obviously related to the origin of thewastewaters and to the presence of fouling and/or clogging materials (Table 6).Even if raw suspended matter seems to be at first responsible for preventingmeasurement, some other components such as soluble substances (grease, pe-troleum by-products, ...) or living organisms (from bacteria to mussels) can bethe main problem.

Except for treated wastewater, almost all raw effluents contain solid com-pounds or grease and hydrocarbons responsible for fouling/clogging. As seenpreviously, on-site measurement is thus preferable, in order to place the instru-


Table 6 Wastewater types and risk of fouling/clogging

Wastewater (WW) type Suspended Fouling/clogging Factorsolids occurrence

Raw urban WWTowns + + Grease, bacteriaSmall communities Variable + id

Industrial WWRefinery, chem. plants – ++ HydrocarbonsPulp and paper +++ +++ Fibres (cellulose)Textile + + Fibres (textile)Agrofood +++ +++ Biological solidsMetal transforming + – None

Agricultural WW + + Variable

Treated WW – – None

Fig. 4 On-site measurement types

ment as close as possible to the medium to be characterized, and three mainpossibilities exist (Fig. 4). The first is on-line measurement, with the sensorplaced inside the flow to be monitored, and in this case without sampling. Thesecond method is off-line measurement, for which the sensor is placed in a sam-pling 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 sim-pler, as measurement is carried out with a field-portable system, more often af-ter grab sampling.

On-line and off-line measurements are often grouped and considered as per-manent or continuous measurements, when results are given with a more or lessregular time interval. Taking into account the complexity of some wastewaters(mainly industrial), it is thus quite impossible to carry out perfect sampling fora complete representation of the medium. But if we assume that emerging pol-lutants are mainly present in the soluble fraction of wastewater, the only re-


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/ Photometric, colorimetric,physical method 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 waste-waters 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) froma medium, continuously or at regular time intervals. The choice and imple-mentation of an on-line sensor must take into account several items or “prop-erties” 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 Good positioning of sensorFouling Anti-fouling devicesAvailability Adapted service

Double sensing preferable

Off-line Representativity Good positioning of strainerClogging Design of the strainer

Suitable flow rate and hole diameterAvailability Adapted service

In situ Representativity Repeated samples and measurement

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

4Parameters and Substances Monitored by On-Line Systems

Before considering some existing solutions for permanent measurement ofwastewater quality, let us describe briefly the different types of approaches.

4.1On-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 laboratorymethod, generally considered as a standard (or reference) method:

– Transposed laboratory methods – This first group is historically the most im-portant as the first developments were carried out in that way. All colori-metric systems using automatic sampling feeding a fast reaction/detectionline (for example, with a flow-injection procedure) have been developedfrom classical procedures, first to increase the analytical rate in laboratoriesbefore being transposed for off-line measurement.

– Rapid equivalent methods – Generally based on a principle different fromthat of the corresponding laboratory method, alternative or surrogate sys-tems are used more and more often for on-line and off-line monitoring. Forexample, the spectrophotometric methods or biosensors proposed for themeasurement of organic compounds or electrochemical techniques formetals must be considered as alternative methods.

– Software sensors and related methods – This last group is considered be-cause of the complexity of wastewater composition and of treatmentprocess control. As all relevant parameters are not directly measurable, aswill be seen hereafter, the use of more or less complex mathematical mod-els for the calculation (estimation) of some of them is sometimes proposed.Software sensing is thus based on methods that allow calculation of thevalue of a parameter from the measurement of one or more other para-meters, the measurement principle of which is completely different from anexisting standard/reference method, or has no direct relation. Statisticalcorrelative methods can also be considered in this group. Some exampleswill be presented in the following section.


4.2On-Line Monitoring Systems

In the past 5 years, some recommendations have been made for the develop-ment of new on-site sensors/analysing equipment [12, 13]. Two recent reviewspresent the state of the art of wastewater quality monitoring. The first [14] fo-cuses on existing and innovative technologies for the main parameters used forthe measurement of organic load (BOD, COD, TOC) for regulation needs, andthe second [15] is a review of on-line monitoring equipment designed forwastewater treatment processes. These studies do not refer to emerging pollu-tants, and can be completed for the other parameters by the following consid-erations.

4.2.1Physical and Aggregate Properties

Relatively far from the present topic and well known, the on-line measurementof the physical and aggregate properties of wastewater does not present anyproblem. Conductivity, temperature, turbidity and oxido-reduction potential(ORP) are easily measured by well-designed sensors, because these parametersare also used for treatment process control. In practice, turbidity is more usedfor the treatment of natural water, and ORP for the biological treatment ofwastewater. However, conductivity and temperature are often monitored at thesame time as the other parameters in this section.

The measurement of total suspended solids (TSS) in wastewater always con-stitutes a challenge because of the difficulty of transposing the laboratory pro-cedure (filtration, drying and weighing). Turbidity measurement is sometimesused for TSS estimation by correlation, but generally without good agreement(because of solids heterogeneity and the effect of colloids). However, a non-con-tact device coupling both scattered light for TSS and fluorescence for organicload (COD) estimation has been proposed [16], giving good results around100–200 mg L–1 for crude wastewater.Another study shows the interest of con-sidering the UV absorption response of the heterogeneous fraction for TSSstudy and estimation [17].

4.2.2Inorganic Constituents

There are a lot of interesting inorganic constituents to be measured in waste-water, either metallic or non-metallic, but few of them can be measured withon-line systems. Table 9 presents a selection of recent works dealing with on-line/on-site monitoring systems for inorganic analytes.

For metallic constituents, some attempts have been made using electro-chemical techniques but without real success, because of the existing inter-ferences in wastewater and the high detection limit required with respect to


regulation needs. A recent study [18] proposed the use of stripping voltam-metry with thick-film graphite and screen-printed electrodes. The analysis isperformed in three steps: sample pre-treatment, accumulation of the analyte onthe electrode surface, and measurement. Cu, Pb and Cd can be determined witha detection limit around 5 mg L–1. Unfortunately, this technique has not beensufficiently checked for wastewater. Some commercial systems based on po-larography can be envisaged, but they not really efficient for wastewater.

Another method is explored with gene expression-based biosensors [19] forCd measurement. A Cd-responsive promoter from E. coli is fused to a pro-moterless lacZ gene and monitored with an electrochemical assay of b-galac-tosidase activity. The expected detection limit is about 0.1 mg L–1 with a responseof a few minutes.As for stripping voltammetry,no real tests have been carried outfor wastewater. However, biosensors can be considered as a promising techniquefor wastewater monitoring.

Hexavalent chromium is also a toxic compound (like lead, cadmium, mer-cury) and can be easily detected with UV spectrophotometry [20]. This systemworks for the quality control of electroplating treated wastewater with a de-tection limit of 5 mg L–1.

For non-metallic constituents, several systems exist especially for nutrientsmonitoring, considering their importance in the eutrophication phenomenon.The on-line measurement of some nitrogen compounds (nitrate and ammo-


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

Analyte Principle Reference

MetalsCu, Pb, Cd Electrochemistry [18]Cd Biosensing [19]Cr(VI) UV spectrophotometry [20]

Nitrogen compoundsNitrate UV spectrophotometry [21, 22]Nitrite Biosensing [24]Ammonium Electrochemistry [23]Nitrate + ammonium UV/UV [25, 26]TKN [26]

Phosphorus compoundsColorimetry [29]UV spectrophotometry [27]Biosensing [28]

OtherSulphide UV spectrophotometry [32]

Multi-ionsPhosphate, iron, sulphate Continuous-flow analysis [30]Chloride, sulphate, phosphate, ... Capillary electrophoresis [31]

nium) has been possible for at least 20 years with the use of selective electrodesat the beginning and UV detection more recently. For nitrate measurement, UVspectrophotometry is the best method because the UV-specific signal of nitratecan be simply extracted from the raw spectrum of a wastewater [21, 22]. An in-tercomparison study was carried out some years ago for NH4

+ on-line analyserswith several instruments using selective electrodes or UV spectrophotometry[23], showing that the performances were acceptable but installation and main-tenance are crucial.

For other forms of nitrogen, nitrite and organic constituents, few techniqueshave been proposed due to the lesser importance of nitrite in wastewater man-agement (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 con-centration in activated sludge exposed to oxic/anoxic cycles. The biosensor con-tains bacteria reducing only NO2

– into N2O, which is subsequently monitored bya built-in electrochemical sensor. Up to 90% of the response is obtained in about1 min, and the detection limit is around 5 mg L–1, a concentration sufficient fortreatment process monitoring.Unfortunately, the maximum operational lifetimeof 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 theuse of electrochemical measurement, the sensor can be immersed and requiresminimum maintenance. The tests carried out to compare its performance withthose of other procedures (including UV for nitrate) showed that the resultswere 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 formsof nitrogen (ammonium and organic) into nitrate (measured by UV spec-trophotometry), allows the selective determination of nitrate, ammonium andorganic nitrogen, and thus of TKN [26], with a detection limit of 1 mg L–1. Thismethod is commercially available.

On-line phosphate measurement is more often limited to orthophosphates,the total phosphorus measurement needing a mineralization step that is diffi-cult 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 reagentaddition). The limit of detection is rather high for this analyte (0.5 mg L–1 orhigher).

Some coupled systems allow measurement of the main N and P forms (nitrate, ammonia and orthophosphates) [22, 27, 29], among which is a systembased on membrane technology in combination with semi-micro continuous-flow analysis (mCFA) with classical colorimetry. With the same principle (clas-sical colorimetry),another system [30] proposes the measurement of phosphate,iron and sulphate by flow-injection analysis (FIA). These systems are derivedfrom laboratory procedures, as in a recent work [31] where capillary electro-phoresis (CE) was used for the separation of inorganic and organic ions fromwaters in a pulp and paper process. Chloride, thiosulphate, sulphate, oxalate,


sulphite, hydrogen sulphide, formiate, carbonate, phosphate and acetate are sep-arated in 5 min after filtration.

Except for the development of on-line systems for nutrients monitoring, themeasurement of other inorganic non-metallic constituents is rather rare. Somecommercial systems based on electrochemical sensing are proposed for themeasurement of cyanide.A simple and rapid procedure for sulphide measure-ment in crude oil refinery wastewater has been developed [32]. Based on the de-convolution of the UV spectrum of a sample, this method has a detection limitof 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 tothe number of potential pollutants), instruments or sensors for parameters usedfor treatment process control are available: UV systems for residual chlorine indeodorization, electrochemical sensors for dissolved oxygen (with nowadays aluminescent dissolved-oxygen probe utilizing a sensor coated with a lumines-cent material) and a colorimetric technique for residual ozone.

In conclusion it must be noted that a lot of developments are still needed inorder to increase the possibility of on-line/on-site monitoring of mineral con-stituents, including the speciation of metallic compounds with regard to healthrisks.

4.3Organic Constituents

The monitoring of organic constituents in wastewater concerns mainly the mea-surement of aggregate properties like oxygen demand parameters (BOD andCOD) and also the detection of specific compounds, generally expressed as thetotal sum of the concentrations of their congeners. Table 10 displays a selectionof on-line/on-site methods for the monitoring of organic constituents and related parameters.

4.3.1Aggregate Properties

Among organic constituent measurements, that of aggregate properties (BODand COD) and specific parameters (TOC for example) has been well developedfor more than 20 years. Concerning BOD, a recent review on biosensors [33] hasbeen published. BOD biofilm-based sensors as well as respirometric systems,other measuring principles, and the commercial BOD instruments are dis-cussed and compared regarding their performance characteristics like linear-ity, response time, precision, agreement between BOD values obtained from thebiosensors and the conventional 5-day test, as well as toxic resistance to vari-ous compounds and operational stability.

Some new developments are also proposed such as a system based on theuse of electrochemically active bacteria in combination with a microbial fuelcell [34], giving good responses over 60 days, or a biosensor developed for fast



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

Parameter Principle Reference

BOD Sensor array [37]Respirometer-type BOD sensors [33]Biosensor [34]Short-term BOD [35]

COD, TOC, ... UV spectrophotometry [38, 39]FTIR [40]

Hydrocarbons IR evanescent wave [41]

Organohalogenated ATR-FTIR [42]compounds

Phenols Amperometric biosensor [43]UV spectrophotometry [9]

Aromatic amines UV spectrophotometry [44, 45]SPME–HPLC [50]SPME–GC [51]

Surfactants UV spectrophotometry [46]On-line titration [47]

Explosives SPME–biosensor [49]

Screening SPE (biosorbents)–HPLC–MS [50]SPME–GC [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 withdisposable strips for BOD determination has been developed [36]. The systemincludes three pairs of light-emitting diodes and photodiodes, and the dispos-able strips are made of inexpensive, transparent polycarbonate plates, wherePseudomonas fluorescens is immobilized. Using the 2,6-dichlorophenol-indo-phenol sodium salt as chromophore, a linear relationship was observed be-tween the bioluminescence of the exposed strip and the BOD value.

Another way for BOD estimation is the use of sensor arrays [37]. An elec-tronic nose incorporating a non-specific sensor array of 12 conducting poly-mers was evaluated for its ability to monitor wastewater samples. A statisticalapproach (canonical correlation analysis) showed a linear relationship betweenthe sensor responses and BOD over 5 months for some subsets of samples,leading to the prediction of BOD values from electronic nose analysis usingneural 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

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 anotherabsorbance measurement (for example at 280 or 350 nm) for compensation ofinterferences. This method can be advantageously replaced by a semi-deter-ministic deconvolution of UV spectra, taking account of the interferences fortheir own estimation (TSS for example). Some industrial applications for thepulp and paper industry or petrochemical plants [38, 39] have shown good estimations of aggregate and specific parameters, with a good availability ofresults 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 limitedto high concentrations, this method based on mid-IR analysis and calibrationgives a rapid estimation of chemical oxygen demand (COD), total organic carbon (TOC), volatile fatty acids (VFA) and partial and total alkalinity (PA andTA) in anaerobic digestion processes for the treatment of industrial waste-waters.

4.3.2Specific Organic Constituents

The first group of interesting organic constituents is hydrocarbons. Classicallymeasured by mid-IR spectrometry after solvent extraction, they can also bemeasured with near-IR devices such as a polymer-coated quartz glass opticalfibre and direct spectrophotometric measurement of the extracted species inthe polymer through the evanescent wave [41]. The proposed system can beused for the quantitative in situ analysis of organic pollutants like chlorinatedhydrocarbons, aromatic hydrocarbons, or fuels in a broad concentration rangefrom around 200 mg L–1 up to a few hundred milligrams per litre.As for the ref-erence method, this instrument provides a signal corresponding to the sum ofconcentrations of the extracted organic compounds by measuring the integralabsorption at the C–H overtone bands in the near-IR spectral range.

Always based on the use of IR spectrophotometry, a novel attenuated totalreflection–Fourier-transform infrared (ATR–FTIR) sensor [42] was proposedfor the on-line monitoring of a dechlorination process. Organohalogenatedcompounds such as trichloroethylene (TCE), tetrachloroethylene (PCE) and car-bon tetrachloride (CT) were detected with a limit of a few milligrams per litre,after extraction on the ATR internal-reflection element coated with a hydro-phobic polymer.As for all IR techniques, partial least squares (PLS) calibrationmodels are needed. As previously, this system is promising for bioprocess con-trol and optimization.

For phenolic compounds, amperometric biosensors have recently been de-signed using bacterial cells [43]. For this purpose, Pseudomonas putida immo-bilized on the surface of thick-film working electrodes made of gold, by usinga gelatin membrane cross-linked with glutaraldehyde, was used and the respi-


ration corresponding to phenol degradation was followed with a commercialoxygen electrode. Phenol detection was performed in synthetic wastewatersamples. For refinery wastewaters, a UV spectrophotometry system, based onspectrum deconvolution, has been proposed for the simultaneous determina-tion 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 concen-trations. However, taking account of the standard solution used for the calibra-tion of the colorimetric reference method (4-nitroaniline), some attempts areproposed for the on-line specific determination of the most important singlecompounds [44, 45].

Both urban and industrial wastewater often contains high concentrations ofsurfactants. Cationic (like alkylbenzene sulphonates) and non-ionic surfactants(like alcohol ethoxylates) are among the most-used surfactants and are dis-charged into sewers in widely varying concentrations.Two on-line methods havebeen designed for the monitoring of cationic surfactants with UV spectropho-tometry [46] and non-ionic surfactants by on-line titration [47]. The detectionlimits are around 10 mg L–1.

Endocrine disruptors are nowadays considered among the most importantemerging pollutants in wastewater, but they are not actually monitored on-line.A recent study [48] described the implementation of a broad-spectrum ana-lytical scheme for the screening of more than 200 compounds (endocrine dis-ruptors, pharmaceutical compounds, ...) in urban wastewater. For other specificorganic compounds, a study concerning the improvement of immunoassayswith 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 detec-tion of several specific organic constituents. The first [50], concerning polarpollutants, 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 poly-aromatic hydrocarbons, benzidine and congeners, nitroaniline and aromaticamines. Another procedure [51], using an on-site manual step of solid-phase microextraction (SPME) before gas chromatographic analysis, was designed forthe detection of organic compounds in industrial wastewater.

4.4Toxicity

On-line or on-site toxicity evaluation is a great challenge due to the complexityof measuring the different impacts (from trouble from to death of organisms)of several substances, the effect of which is often increased by synergy. Fur-thermore for wastewater, toxicity monitoring must be implemented in severallocations: raw sewer, treatment plant or discharge point. Toxicity can be eval-


uated through some effects on organisms such as respirometric modification,bioluminescence or electrochemistry biosensing and even death (rarely usedfor on-line systems).

Historically, respirometers have been used for wastewater biodegradabilityevaluation. More recently [52], a mobile on-line respirometer was proposed andtested for monitoring the activated sludge inhibition due to industrial dischargesin a sewer network.A derived portable device called a Baroxymeter [53], basedon monitoring the respiration of a bacterial culture by pressure measurementsand using respiration inhibition as a toxicity alert, was proposed for the rapiddetection of the toxicity effect of some toxic substances.

But the most-used toxicity tests are based on bioluminescence inhibition, theresponses of which are sometimes difficult to interpret particularly for waste-water quality monitoring. A comparison between a bioluminescence test kit(Microtox) and a respirometry approach for the toxicity study of seven organicand five inorganic toxic compounds was performed [54]. The bioluminescentresponse proved to have a higher sensitivity to toxicants but was less repre-sentative of the effects on activated sludge compared to respirometry, due to thenature 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 sur-factants and other compounds (aromatic hydrocarbons, endocrine disruptors)before implementation on raw and treated wastewater, followed by the identi-fication and quantification of polar organic cytotoxic substances for sampleswith more than 20% inhibition. Furthermore, the study of their contributionto 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 toxicsubstances.

A derived combined approach uses an amperometric biosensor [57] with awhole-cell (E. coli) sensing part, for industrial application (textile and tannerywastewaters) and detection of phenolic compounds, non-ionic surfactants andbenzenesulphonate compounds. As in the previous studies, chemical analysis(SSPE followed by LC–MS) revealed the pollutants responsible for the observedtoxicity.

A portable microbial sensing system [58] was developed for detecting thetoxicity of pre-treated wastewater. The signal of the modified electrode con-taining a bacterial culture renewed every 9 h, within 8 min of contact with toxicsolutions or samples, is roughly correlated with toxicity.

A novel slow-release biosensor delivery for on-line monitoring instrumenta-tion [59] allows both simple toxicity testing and more complex toxicity finger-printing of industrial effluents, with the exploitation of kinetic (dose–response)and dynamic (response with time) signals. Furthermore, the slow release of bio-sensors immobilized in a polyvinyl alcohol (PVA) matrix greatly improvedbiosensor delivery, did not affect the sensitivity of toxicity testing, and demon-strated great potential for inclusion in on-line monitoring instrumentation.


In conclusion, a review of the application of whole-cell biosensors for earlywarning systems [60] showed that electrochemical biosensors are well suitedfor on-site use in the monitoring of general toxicity as well as hydrocarbonsand heavy metals.

4.5Non-Parametric Measurement for Detection of Accident or Disturbance

A final up-and-coming application, based on non-parametric measurement, isused more and more for process control and hazards prevention, for examplefor shock load prevention or toxic events. This qualitative approach uses inte-grated information coming from multiple physical signals:

– Several existing physico-chemical sensors, with data reduction algorithmsand 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 headspacegas from a sparged liquid sample

– Only one instrument giving several responses, such as the absorption spec-trum of a UV detector [63].

5Validation and Developments

5.1Systems Validation

Generally speaking, alternative methods (including on-line, off-line or in situmethods) may be used provided it can be demonstrated that equivalent resultswith those of reference procedures can be obtained. The experiments are gen-erally carried out with standard solutions and reference materials for the de-termination of the method characteristics. The equivalence between methodsmust be statistically verified by plotting the results (Fig. 5) and checking the coordinates of the experimental regression line (comparison of the slope andintercept values, which must be not statistically different from respectively 1and 0 values of the theoretical line).

Once the equivalence between methods is confirmed, the validation proce-dure results given for on-/off-line instruments (permanent measurement) mustbe completed, taking into account that the sampling procedure is different fora lab method compared to a permanent one. For example, considering that reg-ulatory constraints require 24 hours of composite sampling before lab analy-sis, the challenge is to obtain equivalent results with this procedure and withpermanent measurement. In this case, the results to be compared are the meanof values for each measurement during the permanent acquisition, and the ref-



Fig. 5 Comparison of reference and alternative methods

erence value of the corresponding composite sample. Then, the final choice ofthe measurement system must consider the specifications and performancetests of the selected system [11] and the future availability of the measurementestimated from other end users’ experiences [39].

For on-site measurement, such as colorimetric field kits giving semi-quan-titative responses, a simple validation can be carried out [64] taking account ofthe non-continuous distribution of measurement values and that the measur-ing steps are often non-uniformly distributed over the measuring range. A recent study dealt with new elements related to metrological analysis in the fieldof (electrical safety) testing, such as measurement uncertainty and traceability[65]. It is important that the measurement result and its uncertainty are correctlyevaluated so that the right conclusion of conformity or nonconformity withspecifications is made, as is the case for wastewater quality regulation needs.

In conclusion, one must insist on the importance of the main metrologicalcharacteristic, the traceabilility (generally of a result), ensuring a clear (un-broken) relationship between the final result and the complete measurementscheme 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].

5.2Developments

At the end of this chapter, the main methods of on-site measurement systemdevelopment are briefly presented.

5.2.1Optical Techniques

Historically used but still in progress are the optical methods, the applicationsof which in wastewater quality monitoring have recently been reviewed [67, 68].

Before seeing a (future?) whole integrated system, mixing UV–visible–infraredand fluorimetric methods, the first route is the development of UV-based microsystems, including some relevant spectral exploitation techniques such asthe semi-deterministic one [69, 70].

Fluorescence data could be used to quantify oxygen demand values (chem-ical and biochemical) and total organic carbon values. Furthermore, the fluo-rescence spectral response can be apportioned to biodegradable (BOD) andnon-biodegradable (COD-BOD) dissolved organics [71]. Other studies outlinethe advantages and drawbacks of the use of fluorescence techniques for waste-water 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 mon-itoring of highly concentrated organic pollutants [74], particularly with the development of mid-infrared transparent optical fibres and waveguides, thesurface of which can be chemically modified to enhanced analyte recognitionbased on tunable properties of enrichment or (bio)chemical recognition layers.The use of attenuated total reflection (ATR) devices with Fourier-transform IRspectroscopy 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 thegenera Lactobacillus and Staphylococcus has been recently published [76]. TheNIR spectrum acquired by the optical-fibre probe immersed in the culture isexploited using a partial least squares (PLS) calibration step, a classical methodfor IR techniques.

A final optical application deals with the measurement of intracellularnicotinamide adenine dinucleotide (NADH) by fluorescence [77], giving in-formation about the physiological status of wastewater treatment plant bio-mass. This indirect method could be envisaged for toxicity estimation.

5.2.2Biosensors

Biosensors are based on the direct spatial coupling of immobilized biologicallyactive compounds with a signal transducer and an electronic amplifier. Due to the reaction between the biorecognition molecule (receptor) and a target an-alyte, a transducer signal is produced [27]. Biosensors are promising techniquesfor environmental monitoring applications such as toxicity, bioavailability andmulti-pollutant screening, even if some limitations still exist [78].With regardto the diversity of compounds and the complexity of matrices (particularlywastewaters), further developments have to be made before consideringbiosensing as an operational solution for on-line monitoring.

Currently, a large spectrum of microbial biosensors have been developedthat 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


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 ofnitrification in environmental samples has also been developed. The biosensorconsists of a Clark oxygen probe as a transducer and an immobilized mixed nitrifying culture as the microbial component. The measuring principle isbased on the direct determination of bacterial metabolic activity by measuringthe oxygen consumption rate of the microbial immobilizate [79].

In conclusion, more than 40 years after the first electrode with an immo-bilized-enzyme membrane was produced, future developments in biosensordesign will inevitably focus upon the technology of new materials, especiallythe new copolymers that promise to solve the biocompatibility problem and offer the prospect of more widespread use of biosensors in clinical (and envi-ronmental) monitoring [80].

5.2.3Software Sensors

As seen previously for some specific applications such as wastewater treatmentplants, software sensors can be envisaged to provide on-line estimation of non-measurable variables, model parameters or to overcome measurement delays[81–83]. Software sensors have been developed mainly for monitoring bio-processes because the control system design of bioreactors is not straightfor-ward 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 theprocess.

A software sensor combines the theoretical knowledge of a system througha mathematical model, and the practical knowledge of its actual functioningthrough measurements. If the inputs acting on the system are known (and pro-vided that theoretical conditions are fulfilled) and if, moreover, the model is asufficient approximation of the real system, then the software sensor estimatesthe whole state of the system [81]. There are two types of approaches in devel-oping software sensors [85], the first estimating the required parameters on thebasis of a deterministic model and the second being a black-box approach de-pending only on the observed values. In practice, for wastewater treatment ap-plications, the main techniques available are [82]:

– The representation of the biological conservation of substrate to cell massby an overall chemical reaction. The stoichiometric relationships are thenused to calculate various rates such as cell mass concentration [83].

– The previous method supposes complete knowledge of the system and de-pends on the measurement quality of instruments (errors, availability), lead-ing to severe effects on the accuracy of the on-line estimates.Therefore,a goodnoise filtration algorithm (like the Kalman filter or derivative) should be employed to improve the reliability of the estimated values before their use.



Table 11 Software sensor main types for wastewater treatment applications

Types Objectives References

Estimation through elemental Anaerobic bioreactor parameters [81] balances estimation

Observer and filtering techniques Anaerobic bioreactor control [86]

Nitrification bioreactor control [88]

Artificial neural networks TKN estimation [85] or hybrid ANN Fluorescence spectra exploitation [89]

for organic constituents estimation

Estimation of wastewater [90]parameters (COD, NH4

+, ...)

– Another procedure uses artificial neural networks (ANN) derived from ar-tificial intelligence techniques.Among several ANN algorithms, the feed-for-ward one, made up of interconnected neurone-like elements, can modelcomplex non-linear systems easily, depending on the status of the trainingdata. If there are noise and uncertainty in the training data, a problem ofoverfitting often arises but can be solved by data pre-processing, using aprincipal component analysis (PCA) for example [85].

Table 11 gives some examples of different software applied to the monitoringof bioreactors and the estimation of wastewater parameters.

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272 Wastewater Quality Monitoring: On-Line/On-Site Measurement