Progress on Drinking Water Research [1 ed.] 9781616680893, 9781604567489

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Copyright © 2008. Nova Science Publishers, Incorporated. All rights reserved. Progress on Drinking Water Research, Nova Science Publishers, Incorporated, 2008. ProQuest Ebook Central,

Copyright © 2008. Nova Science Publishers, Incorporated. All rights reserved. Progress on Drinking Water Research, Nova Science Publishers, Incorporated, 2008. ProQuest Ebook Central,

Copyright © 2008. Nova Science Publishers, Incorporated. All rights reserved.

PROGRESS ON DRINKING WATER RESEARCH

No part of this digital document may be reproduced, stored in a retrieval system or transmitted in any form or by any means. The publisher has taken reasonable care in the preparation of this digital document, but makes no expressed or implied warranty of any kind and assumes no responsibility for any errors or omissions. No liability is assumed for incidental or consequential damages in connection with or arising out of information contained herein. This digital document is sold with the clear understanding that the publisher is not engaged in rendering legal, medical any other professional services. Progress on Drinking Water Research, Nova Scienceor Publishers, Incorporated, 2008. ProQuest Ebook Central,

Copyright © 2008. Nova Science Publishers, Incorporated. All rights reserved. Progress on Drinking Water Research, Nova Science Publishers, Incorporated, 2008. ProQuest Ebook Central,

PROGRESS ON DRINKING WATER RESEARCH

MATHIS H. LEFEBVRE AND

MATHEO M. ROUX Copyright © 2008. Nova Science Publishers, Incorporated. All rights reserved.

EDITORS

Nova Science Publishers, Inc. New York

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Copyright © 2008 by Nova Science Publishers, Inc. All rights reserved. No part of this book may be reproduced, stored in a retrieval system or transmitted in any form or by any means: electronic, electrostatic, magnetic, tape, mechanical photocopying, recording or otherwise without the written permission of the Publisher. For permission to use material from this book please contact us: Telephone 631-231-7269; Fax 631-231-8175 Web Site: http://www.novapublishers.com NOTICE TO THE READER The Publisher has taken reasonable care in the preparation of this book, but makes no expressed or implied warranty of any kind and assumes no responsibility for any errors or omissions. No liability is assumed for incidental or consequential damages in connection with or arising out of information contained in this book. The Publisher shall not be liable for any special, consequential, or exemplary damages resulting, in whole or in part, from the readers’ use of, or reliance upon, this material.

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Independent verification should be sought for any data, advice or recommendations contained in this book. In addition, no responsibility is assumed by the publisher for any injury and/or damage to persons or property arising from any methods, products, instructions, ideas or otherwise contained in this publication. This publication is designed to provide accurate and authoritative information with regard to the subject matter covered herein. It is sold with the clear understanding that the Publisher is not engaged in rendering legal or any other professional services. If legal or any other expert assistance is required, the services of a competent person should be sought. FROM A DECLARATION OF PARTICIPANTS JOINTLY ADOPTED BY A COMMITTEE OF THE AMERICAN BAR ASSOCIATION AND A COMMITTEE OF PUBLISHERS. LIBRARY OF CONGRESS CATALOGING-IN-PUBLICATION DATA Progress on drinking water research / Mathis H. Lefebvre and Matheo M. Roux, editors. p. cm. Includes bibliographical references and index. ISBN 978-1-61668-089-3 (e-Book) 1. Water quality--Research. 2. Drinking water--Research. 3. Water--Analysis--Research. I. Lefebvre, Mathis H. II. Roux, Matheo M. TD370.P76 2008 628.1--dc22 2008024332

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CONTENTS Preface Chapter 1

Chapter 2

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Chapter 3

Chapter 4

Chapter 5

Chapter 6

Chapter 7

vii Carbon Paste Electrodes for the Determination of Detrimental Substances in Drinking Water Jiří Zima, Ivan Švancara, Karolína Pecková and Jiří Barek The Use of Composite Electrodes in the Analysis of Drinking Water T. Navratil, B. Yosypchuk and J. Barek Voltammetric and Amperometric Determination of Organic Pollutants in Drinking Water Using Boron Doped Diamond Film Electrodes K. Peckova, J. Musilova, J. Barek and J. Zima Amalgam Electrodes as Sensors in the Analysis of Aquatic Systems B. Yosypchuk, T. Navratil, J. Barek, K. Peckova and J. Fischer Polarographic and Voltammetric Determination of Genotoxic Substances in Drinking Water Using Mercury Electrodes Vlastimil Vyskocil, Jiří Barek, Ivan Jiranek and Jiří Zima Groundwater Toxicity Due to Natural Dissolved Radionuclides Belonging to the U and Th Decay Series Daniel Marcos Bonotto Harmonising Internal Quality Tasks (Method Validation, Quality Control and Sample Uncertainty) in Analytical Laboratories. Case Studies: Water Analysis Methods Salvador Sagrado and Laura Escuder-Gilabert

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vi Chapter 8

Chapter 9

Contents Solar Disinfection of Drinking Water: Lessons from Field Studies in India Robert H. Reed, Isaac S. Bright Singh and Shibu K. Mani Impact of Low Turbidities in the Sanitary Quality of Exploited Water Thierry Jouenne and Jean-Paul Dupont

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Index

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267 271

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PREFACE This new book focuses on worldwide research on contamination, toxicity and treatment of drinking water. Water of sufficient quality to serve as drinking water is termed potable water whether it is used as such or not. Although many sources are utilized by humans, some contain disease vectors or pathogens and cause long-term health problems if they do not meet certain water quality guidelines. Water that is not harmful for human beings is sometimes called safe water, water which is not contaminated to the extent of being unhealthy. The available supply of drinking water is an important criterion of carrying capacity, the population level that can be supported by planet Earth. Typically water supply networks deliver single or multiple qualities of water, whether it is to be used for drinking, washing or landscape irrigation; one counterexample is urban China, where drinking water can be optionally delivered by a separate tap. Chapter 1 - In this contribution, electroanalysis with carbon paste electrodes (CPEs) and chemically modified carbon paste electrodes (CMCPEs) is reviewed with respect to their applicability to the determination of various pollutants of both inorganic and organic origin in water samples, including drinking water. At first, carbon paste-based electrodes are presented via a short retrospective survey, introducing and classifying the individual types of which especially chemically modified variants may offer a large palette of electrodes, sensors or even whole detection systems applicable to the water analysis. Both CPEs and CMCPEs are briefly discussed with respect to their specifics such as the “alchemy” of their laboratory preparation or a number of different physico-chemical and electrochemical processes employed in practical measurements. This is also the case of various pre-concentration schemes for electrochemical stripping analysis, which is being the technique of choice in analyses of water samples at carbon paste-based electrodes. The key section of this chapter is a nearly complete survey of methods which have ever been developed and proposed for analysis of water samples when using carbon paste-based electrodes. It covers more than hundred methods for the determination of inorganic ions, complexes, and molecules, as well as nearly the same number of the respective procedures for organic compounds such as environmental pollutants, industrial surfactants, numerous pesticides or carcinogenic substances of the polycyclic aromatic hydrocarbon type. The most important achievements in the field are commented in more detail and typical applications illustrated on numerous examples, mostly based on the research work of the authors and withdrawn from their own archives. Last but not least, some typical trends in the

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present day's electroanalysis with CPEs and CMCPEs are outlined and future prospects given focused on specific applications in water analysis. Chapter 2 - Recent results regarding voltammetric and amperometric determination of micromolar and submicromolar concentrations of various environmentally important biologically active inorganic and organic substances in drinking water using composite solid electrodes (CSE) are described in this chapter. This type of relatively frequently used electrodes, in commercial as well as in scientific laboratories, can represent successful and environmentally friendly alternative to mercury electrodes in analysis of drinking water. The discussed electrodes belong to the group of composite electrodes with random ensembles of dispersed particles with final solid form. CSE are composed, according to their definition, of at least one conductor phase and at least one insulator phase, which particles are mutually mixed. Insulator phase is usually represented by a polymeric or a monomeric material. Conductor phase can be formed by either metallic (silver, gold, etc.) or by another nonmetallic conducting material (e.g. graphite powder) or by their mixture. Some specific properties, required for determination of chosen analytes in specified matrix, can be added into bulk of the electrode materials. Surface of these electrodes can be also modified, e.g. by Nafion, catalyst or enzyme. CSE can be applied either in batch analysis or in flow liquid systems (especially HPLC or FIA with electrochemical detection). Solid amalgam composite electrodes, prepared from solid amalgam powder and insulator phase, are another example of CSE. They proved experimentally to be a suitable, reliable and environmentally friendly substitute for the hanging mercury drop electrode (HMDE) in electrochemical analyses. The review will concentrate on the authors’ own results in the context of the general development in the field. Chapter 3 - There is a never-ending search for new electrode materials for voltammetric or amperometric determination of various detrimental substances in drinking water. The basic requirements for new electrode materials are lower noise, broader potential range, mechanical robustness enabling measurements in flowing systems, compatibility with organic solvents making them compatible with high performance liquid chromatography (HPLC), flow injection analysis (FIA) or capillary electrophoresis (CE) with electrochemical detection (HPLC/ED, FIA/ED, CE/ED) and – last but not least – resistance towards passivation and electrode fouling which is probable the biggest complication in the practical applications of electroanalytical methods. One of the most promising non-traditional electrode materials is boron doped polycrystalline diamond film on a silica support, which is suitable for the determination of a wide spectrum of organic and inorganic analytes. The electrodes based on it possess excellent electrochemical properties, such as a low and stable background current over a wide potential range, corrosion resistance, high thermal conductivity and high current densities. Furthermore, this material offers superb micro structural stability at extreme cathodic and anodic potentials and resistance to fouling because of weak adsorption of polar species on the hydrogen terminated paraffin-like surface, which results in good responsiveness for many redox analytes without any pretreatment. Boron doped diamond film electrodes are applicable to voltammetric or amperometric determination of both oxidisable and reducible substances with limit of determination down to 10-8 mol L–1 without any preconcentration step. The chapter is devoted to their practical applications for both batch voltammetric analysis and continuous flow amperometric analysis (HPLC/ED, FIA/ED, CE/ED) of organic pollutants possibly present in drinking water.

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Chapter 4 - This chapter will deal with recent results regarding voltammetric and amperometric determinations of micromolar and submicromolar concentrations of various environmentally important biologically active inorganic and organic substances in drinking water using non-traditional types of electrodes based on solid and paste amalgams, which can be environmentally friendly alternatives to mercury electrodes. Attention will be paid to polished and mercury meniscus modified solid amalgam electrodes (SAE), to silver solid amalgam composite electrodes and to amalgam paste electrodes either in batch analysis or in flow liquid systems (especially HPLC or FIA with electrochemical detection). Both polished and mercury meniscus modified silver solid amalgam electrodes (AgSAE) can be used in voltammetric and amperometric analysis of drinking water as an alternative to a hanging mercury drop electrode (HMDE). Other solid amalgam electrodes (e.g., CuSAE, AuSAE) are convenient for specific purposes, where properties of the metal, of which the solid amalgam consists, are employed. Silver solid amalgam composite electrodes, prepared from silver amalgam powder and epoxy resin, proved experimentally to be a suitable, reliable and environmentally friendly substitute for the hanging mercury drop electrode in electrochemical analyses. Amalgam paste electrodes based either on powdered amalgam mixed with a suitable organic pasting liquid or on amalgam paste can be used as non-toxic sensors with easily renewable surface and as a substitute of disposable electrodes. The review will concentrate on the authors’ own results in the context of the general development in the filed. Chapter 5 - Even 85 years after their introduction in analytical chemistry by Nobel Prize winner professor Jaroslav Heyrovsky, mercury electrodes are still the best available sensors for voltammetric monitoring of trace amounts of electrochemically reducible inorganic and organic substances. Their main advantages (atomically smooth surface, easily renewable surface diminishing problems with passivation so frequently encountered with solid working electrodes and large potential window in cathodic region) usually overbalance their disadvantages (very limited anodic potential window, limited mechanical stability complicating their application for measuring in flowing media and unsubstantiated fears of their toxicity). Modern variants of mercury electrodes, namely hanging mercury drop electrode, in combination with pulse techniques or accumulation of the analyte on the surface of working electrode (electrochemical or adsorptive) enable to reach micromolar or even nanomolar limits of determination for electrochemically reducible substances. Recently, increasing attention is paid to their application for the determination of trace amount of genotoxic substances in both drinking and surface waters. In this chapter, practical applications of polarographic and voltammetric methods on mercury electrodes for the determination of trace amount of various chemical carcinogens (namely nitrated polycyclic aromatic hydrocarbons, electrochemically reducible heterocyclic compounds, etc.) in drinking water will be reviewed and compared with our most recent experimental results in this field. Advantages and disadvantages of various polarographic and voltammetric methods in this field will be critically evaluated. Attention will be paid to their combination with preliminary separation and preconcentration using liquid or solid phase extraction. Chapter 6 - Primeval radionuclides have survived in detectable amounts since the time of nucleosynthesis and contribute to the terrestrial gamma radiation, whose major contribution comes from 40K and three radioactive series having unstable members with half-lives much shorter than that of each precursor, i.e. 232Th, 238U, and 235U. Thorium and uranium are lithophile elements, being concentrated preferentially in acid igneous rocks compared with intermediate, basic, and ultrabasic varities. 232Th decays to the stable 208Pb, after 12

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disintegrations (7 alpha-type and 5 beta-type). 238U is the principal isotope of natural U (99.72% abundance) and decays to the stable 206Pb, after 14 disintegrations (8 alpha-type and 6 beta-type). Potential health hazards from some natural radionuclides belonging to these decay series in consuming water have been considered worldwide, with many countries adopting the WHO guideline activity concentration for drinking water quality. In general, the recommendations apply to routine operational conditions of water supply systems. Several national standards for limiting radiation exposure establish maximum permissible radionuclides concentration in drinking water, where, in general, for practical purposes, 0.5 Bq/L for gross alpha and 1 Bq/L for gross beta activity have been used to routine operational conditions of existing or new water supplies. However, additional information concerning to specific radionuclides may be required in some circumstances in order to decide about the drinking water quality in terms of radiological aspects. Special attention must be given when groundwaters are utilized for public water supplies, because water-rock/soil interactions may enhance the presence of natural radionuclides in solution. This chapter reports the radiological implications of the interactions between waters and different rock types occurring in aquifer systems exploited in Brazil. The data here shown demonstrate how some natural radionuclides contribute to the groundwater toxicity evaluated in terms of the WHO guideline activity concentration for drinking water quality. Chapter 7 - Method validation, Quality Assurance and sample uncertainty estimation are essential internal tasks for accredited laboratories. In the past the statistical aspects involving these tasks have been treated separately, however, recently, an approach to harmonise the internal validation-control-uncertainty process has been reported. In this commentary, an unambiguous step-by-step protocol to evaluate and harmonise internal quality aspects of a method is defined. Such protocol involves a statement of method’s scope (analytes, matrices, concentration level) and requisites (external and internal), method’s ‘fit-for-purpose’ features selection, pre-validation (to adjust the validation protocol), accuracy validation (in intermediate precision conditions) and assessment (via Monte Carlo simulation), harmonisation of the validation-verification-control-uncertainty process (u-approach), validation of other required method’s features, validity statement in terms of ‘fit-for-purpose’ decision, Internal Quality Control tasks (including Method Verification and harmonised mean control charts), harmonised sample uncertainty estimations and short-term routine work (intending to present almost ‘ready-to-use’ methods). Decision-making aspects (in view of harmonisation), innovative criteria and impact on laboratory staff are outlined. Areas which need to be explored and some missdirections are delineated. Chapter 8 - Waterborne diseases remain a significant cause of morbidity and mortality across the developing world, due to lack of safe water supplies. Solar disinfection provides a simple, minimal cost technique for the treatment of contaminated water, making use of natural sunlight to photoinactivate pathogenic microbes. At its simplest, solar disinfection involves : (i) filling a plastic container with contaminated water, leaving a small air space (ii) shaking the container, to fully oxygenate the water (iii) exposing the water to full-strength sunlight, typically for a whole day, and then (iv) consuming the treated water, typically the following day. The authors have conducted field studies into the effectiveness of solar disinfected safe water for drinking purpose in villages in several different locations in India, each with its own problems of water quality, namely: (i) the arid zone of the Thar desert, Rajasthan, where water is a scarce resource, with direct consumption of surface water especially during summer; (ii) the flood-prone region of Uttar Pradesh, on the Indo-Gangetic

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plain, where poor sanitary practices and high population density lead to the contamination of groundwater, which is then taken via shallow hand pumps for family use; and (iii) the subtropical region of Kerala, where water is plentiful, but where households located far from sources of treated water supplies use raw surface water. In each location, a field researcher worked with villagers, first explaining the problems of waterborne infections, demonstrating contamination of local water sources using standard bacteriological testing, then illustrating how solar disinfection can inactivate the contaminant bacteria. In all test villages, the introduction of solar disinfection led to a decrease in reported incidents of diarrhoea/gastroenteritis, though the extent of the decrease varied between the sites, being lowest in Uttar Pradesh. Participants were surveyed at the end of the implementation period and they expressed positive comments about their perceptions of the effectiveness of solar water treatment, with respect to a reduction in diarrhoea/gastro-enteritis, and also in terms of an improvement in their overall well-being. These findings suggest that solar disinfection has the potential for more widespread use in India and in similar countries where a significant proportion of the rural population remains unserved by conventional water treatment system. Chapter 9 - In regions where the whole drinking water resources have a karstic origin, rain events generate turbid runoffs that cause sanitary crises and recurrent interruption of water supply. In the past decade, it became evident that aquifers harbour large populations of physiologically active micro-organisms which interact with the abiotic environment by adsorption and/or biofilm formation on surfaces. Consequently, turbid runoffs induce microbial contamination associated to suspended particles. It is the reason why the turbidity value is one of the main indicator of the water quality (the standard as been fixed at 2 NTU); However, by accessing the number of attached and (planktonic bacteria in water samples, the authors recently showed that the turbidity was not always correlated with the numbers of particle-bound bacteria and consequently cannot be used as a sole indicator of the sanitary risk. Moreover, the authors showed that some opportunistic pathogens (e.g., Pseudomoas oryzihabitans) might be associated to suspended particulate matters in water samples exhibiting turbidity values lower than 2 NTU.

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Chapter 1

CARBON PASTE ELECTRODES FOR THE DETERMINATION OF DETRIMENTAL SUBSTANCES IN DRINKING WATER Jiří Zima1*, Ivan Švancara2, Karolína Pecková1, and Jiří Barek1 1

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UNESCO Laboratory of Environmental Electrochemistry, Department of Analytical Chemistry, Charles University in Prague, Albertov 6, CZ-128 43 Prague 2, Czech Republic 2 Department of Analytical Chemistry, Faculty of Chemical Technology, University of Pardubice, Náměstí. Čs. legií 565, CZ-532 10 Pardubice, Czech Republic

ABSTRACT In this contribution, electroanalysis with carbon paste electrodes (CPEs) and chemically modified carbon paste electrodes (CMCPEs) is reviewed with respect to their applicability to the determination of various pollutants of both inorganic and organic origin in water samples, including drinking water. At first, carbon paste-based electrodes are presented via a short retrospective survey, introducing and classifying the individual types of which especially chemically modified variants may offer a large palette of electrodes, sensors or even whole detection systems applicable to the water analysis. Both CPEs and CMCPEs are briefly discussed with respect to their specifics such as the “alchemy” of their laboratory preparation or a number of different physico-chemical and electrochemical processes employed in practical measurements. This is also the case of various pre-concentration schemes for electrochemical stripping analysis, which is being the technique of choice in analyses of water samples at carbon paste-based electrodes. The key section of this chapter is a nearly complete survey of methods which have ever been developed and proposed for analysis of water samples when using carbon paste-based electrodes. It covers more than hundred methods for the determination of inorganic ions, complexes, and molecules, as well as nearly the same number of the respective procedures for organic compounds such as environmental pollutants, industrial surfactants, numerous pesticides or carcinogenic substances of the polycyclic aromatic hydrocarbon type.

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Jiří Zima, Ivan Švancara, Karolína Pecková et al. The most important achievements in the field are commented in more detail and typical applications illustrated on numerous examples, mostly based on the research work of the authors and withdrawn from their own archives. Last but not least, some typical trends in the present day's electroanalysis with CPEs and CMCPEs are outlined and future prospects given focused on specific applications in water analysis.

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ABBREVIATIONS AND SYMBOLS USED -- ... unspecified, not given, not found 2nd−DV ... second-order derivative voltammetry 2.5th-DV ... two-and-halfth derivative voltammetry μE(s) ... microelectrode(s) A ... amper; mA ... miliamper, μA ... microamper nA ... nanoamper absor. ... absorption, absorbed accum. ... accumulation, accumulated AcB ... acetate buffer adsor. ... adsorption, adsorbed AdSV ... adsorptive stripping voltammetry AdCtSV ... adsorptive / catalytic stripping voltammetry AmB ... ammonia buffer AM(D) ... amperometry (amperometric detection) anal(s) ... analysis, analysed, analyte(s) appl. ... application, applicable ASV ... anodic stripping voltammetry AuF ... gold film AuFE ... gold film electrode Bi-CPE ... bismuth powder-modified carbon paste electrode Bi2O3-CPE ... bismuth trioxide-containing carbon paste electrode BiF-CPE ... bismuth film-plated carbon paste electrode BiF ... bismuth film BiFE ... bismuth film electrode biol. ... biological BRB. ... Britton-Robinson buffer CA ... chronoamperometry CCSA ... constant current stripping analysis CE(s) ... crown-ether(s) CE-EC ... capillary elecrophoresis with electrochemical detection CNT ... carbon nanotubes CNTPE ... carbon nanotube paste electrode compl(s) ... complexation(s), complexant, complexed with C/MO ... mineral oil-made carbon paste FIA ... flow injection analysis

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Carbon Paste Electrodes for the Determination of Detrimental Substances… FIA-EC ... FIA with electrochemical detection GCE ... glassy carbon electrode GCPE ... glassy carbon paste electrode hr(s). ... hour(s) incl. ... including, includes inorg. ... inorganic interf(s) ... interference(s), interfering species LOD ... limit of detection LOQ ... limit of quantification LS(AS)V ... linear sweep (anodic stripping) voltammetry M ... molar concentration, mol l−1 mM ... 0.001 mol l−1 μM ... 1×10−6 mol l−1 nM ... 1×10−9 mol l−1 pM ... 1×10−12 mol l−1 mb. ... (protective, porous) membrane Me ... metal, metallic MeOH ... methanol MEx ... medium exchange MF ... mercury film MFE ... mercury film electrode min. ... minute(s) MO ... mineral oil modif(s). ... modifier(s), modification, modified with MWD ... microwave-assisted digestion (decomposition) n., N ... nano Nf ... Nafion® (fluoro-polymer used as protective layer) o.c. ... open circuit org. ... organic oxid(s) ... oxidant(s), oxidising, oxidised oxidn. ... oxidation PAH(s) ... polyaromatic hydrocarbon(s) phen. ... 1,10-phenanthroline PhOH ... phenol ppm ... concentration unit (part per million) ppb ... concentration unit (part per billion) poll(s). ... environmental pollutant(s) CMCPE ... chemically modified carbon paste electrode CPE ... carbon paste electrode C/Nj ... carbon paste made of Nujol® oil CP ... carbon paste CP-CWE ... carbon paste coated-wire electrode C/PO ... paraffin oil-made carbon paste CRM(s) ... certified reference material(s) C/SO ... carbon paste made of silicone oil

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Jiří Zima, Ivan Švancara, Karolína Pecková et al. CSV ... cathodic stripping voltammetry CTAB ... cetyltriethylammonium bromide C/TCP ... tricresyl phosphate-containing carbon paste C/Uv ... carbon paste made of Uvasol® CV ... cyclic voltammetry ctng. ... containing, consisting of di ... diameter D ... detector DCV ... direct current voltammetry dep. ... deposition, deposited der. ... derivative(s) detc. ... detection, detected detn(s) ... determination(s) diffn. ... differentiation, differentiating dilt. ... diluted diss. ... dissolution, dissolved DMG ... dimethyl glyoxime (Chugayev reagent) DP(A,C)SV ... differential pulse (anodic, cathodic) stripping voltammetry E, E ... electrode, electrode potential ECL-EC ... electrochemiluminescence measurement with elchem. detection EDTA ... ethylendiammin tetraacetate el. ... electrode elc. ... electrolytic, electrolysis el.chem ... electrochemistry, electrochemical EP(X) ... peak potential (for an X species) en ... ethylendiamine EtOH ... ethanol extr(s) ... extraction, extracts(s), extracted, extractable POT ... potentiometry, potentiometric pract ... practical pretrm. ... pretreatment PSA ... potentiometric stripping analysis (with chemical oxidation) py, Py ... pyridine, pyridinium Q ... quinone redn. ... reduction regnt. ... regeneration, regenerated ref. ... reference reoxidn. ... reoxidation resp. ... respective (corresponding), respectively RSD ... relative standard deviation s ... second(s) (s) ... solid, in solid state SCE ... saturated calomel electrode SbF ... antimony film SbFE ... antimony film electrode s.e. ... supporting (basic) electrolyte sep. ... separation, separated

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soln(s) ... solution(s) sp. ... species (family) SPE ... screen-printed carbon paste (electrode) subst. ... substitution, substituted suppr. ... suppression, suppressed surf(s). ... surfactant(s), surface active compound SV ... stripping voltammetry SW(A,C)SV ... square-wave (anodic, cathodic) stripping voltammetry synth. ... synthetic tACC ... accumulation (preconcentration) time TBP ... tributyl phosphate TCP ... tricresyl phosphate teor ... theoretical titr(s). ... titration(s), titrated, titrant unm. ... unmodified V ... volt(s) [unit] vs. ... versus W(I) ... wax (impregnated)

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1. INTRODUCTION - A HALF CENTURY WITH CARBON PASTE ELECTRODES In 2008, it is exactly fifty years since Professor Ralph N. Adams (1924-2003) had presented a short report with laconic title "Carbon Paste Electrodes" [1]. As his former student recalls [2], no-one – including the inventor himself – could have predicted that this event would start a new chapter in electrochemistry and electroanalysis, comprising to date nearly two thousands of scientific papers published by authors and research teams across the globe [3]. The key-topic of this tremendous scientific database is a special electrode material: carbon paste, which is a mixture of carbon (graphite) powder with a suitable liquid binder and, eventually, with additional constituent(s) contained in the basic binary formula [1,3-13]. Before carbon paste electrodes (CPEs) have reached such a prominent position among other electrodes and sensors, they underwent a very interesting development reflecting faithfully the most significant achievements in the field. The individual periods or even the entire history of using CPEs and related sensors in laboratories of academic institutions or research institutes have already been of interest in various retrospective commentaries (see e.g. [37,9,13]). Herein, the entire era of 1958-2008 can be at least briefly pointed out via the individual revolutionary moments and periods, when providing the reader with some fundamental references to the original literature concerning the respective subject(s) or theme(s): 1958: The very first report on carbon paste electrodes [1]. 1959-1963: Proposals of the first carbon paste mixtures, their basic characterisation, and initial applications [4,14-16].

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1964, 1965: First modifications of basic types of carbon pastes (i.e., desired alterations of their basic physico-chemical and electrochemical properties [4,17,18]). 1965-1980: Expansion of carbon paste electrodes in electrochemical laboratories (starting research activities out of the U.S. [19-25]). 1981-1990: Postulation of basic rules for chemical modification of carbon pastes [26,27]: The decade of chemically modified carbon paste electrodes, CMCPEs [6,8]. 1988-1990: Carbon pastes with enzymes as a new type of biosensors applicable in analysis of organic compounds and biologically important compounds, as well as some inorganic ions and molecules [28-31]. 1991: Commencing competition of classical carbon pastes with solid heterogeneous carbon composites or screen-printed sensors [10,32,33]. 1995-2005: Carbon pastes in new technologies and some novel approaches to instrumental analysis (alternate power sources, environmentally friendly procedures, tasting electronic tongues, new forms of carbon such as fullerenes, nanotubes, doped diamonds), hybrid inorganic / organic films, nanomaterials, etc. [34-40]). 2008: A half-century of the carbon paste in chemical laboratories. Thus, the authors of this chapter would also like to take a chance to remind and celebrate such a significant anniversary, which proves – in itself – the vitality of carbon paste-based electrodes and sensors in modern electrochemistry, electroanalysis or even beyond these fields. Since the problematics of both CPEs and CMCPEs has been reviewed in detail by numerous authors and authors' teams [4-13], including a very extensive chapter published within the Encyclopedia of Sensors series a couple of years ago [3] and having featured, in fact, a kind of still-missing monograph, the contribution offered herein describes carbon pastes in a very brief way only whereas the main attention is focused on their applicability in analysis of water samples – i.e., a special area which has not been yet separately reviewed. And in order to show the possibilities and limitations of both CPEs and CMCPEs in full scope, it was decided to assess not only achievements in analysis of drinking water, but also valuable experimental material concerning related natural waters that are usually the original source for mass production of drinking water and represent the analytical specimens with – more or less – similar composition.

2. CARBON PASTE AS SPECIFIC ELECTRODE MATERIAL 2.1. Basic Considerations, Classification, and Preparation of Carbon Pastes as the Electrodes For Measurements Carbon pastes, as the mixtures of carbon (graphite) powder with a suitable liquid binder ("pasting liquid"), exhibit a behaviour which reflects the type and quality of carbon powder used, as well as the character of the binder chosen. Their mechanical, physico-chemical, and electrochemical properties are related to solid electrode substrates from graphite and similar materials; but, due to the binder present, carbon paste-based electrodes and sensors possess some specific features or even unique characteristics [3-13].

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Historically, the very first definition had classified carbon paste electrodes as a special type of solid carbon electrodes [1,4,5], newer definitions sort these electrodes among heterogeneous carbon electrodes [3,9,10,13]. Apart from preference, both categorisations comprise almost all variations of carbon paste-based electrodes and sensors that have appeared in electrochemistry and electroanalysis within the mentioned five decades. With respect to the two main carbon paste constituents, the hitherto existing types of carbon paste electrodes can be categorised into the following types: (i) In accordance with the physico-chemical character of the binder: Classical CPEs ... made of pulverised graphites and liquid organic substances characterised by chemical and electrochemical inactivity. These CPEs are the most frequent, representing approx. 70-80 % of carbon paste-based electrodes described to date [3-13]. CPEEs ("CPEs with electroactive binder") ... based on paste mixtures in which the binding liquid is a strong inorganic electrolyte, mostly concentrated solutions of mineral acids or alkaline hydroxides [41,42]. (Applications of such sensors are rather unique and beyond common electroanalysis and due to this, in this chapter, CPEEs are not further considered.) (ii) With respect to the carbon paste consistence: "Dry" or "wet" CPEs ... such carbon paste mixtures differ significantly from the content of the binder, i.e., from the carbon-to-pasting liquid ratio used. At present, this categorisation is not very usual and can e.g. be found in association with some particular studies [43]. Solid (or "solid-like") CPEs ... In these mixtures, the binder is either a substance which can be easily melted and then cooled down, e.g. phenanthrene [33] or highly viscous substances such as silicone rubber [44] and polypropylene [45]; both forming the dense mixtures that resemble solid materials rather than a soft paste. (iii) Based on the carbon paste composition and eventual occurrence of another constituent: Unmodified CPEs ... These electrodes consist of mixtures made solely of the two main constituents − graphite powder and a binder. Alternatively, such types of carbon paste electrodes are also called as "bare", "native" or "virgin" CPEs [4-6]. Modified CPEs ... Binary mixtures contain additional component(s), usually the modifier itself. According to the character of the modifying agent used, the respective electrodes and sensors are divided into two fundamental groups: (A) chemically modified carbon paste electrodes (CMCPEs [3,6,9]) and (B) biologically modified analogues, widely known as enzymatic carbon paste biosensors [31]. Regarding the modifications as such, they can be accomplished in numerous ways: (a) The simplest is modification in situ, when the reagent of choice is present in the solution during the proper measurement [46,47]. (In electroanalysis, it is commonly used also for other electrodes and successful function of modifier applied in situ relies on its affinity towards the electrode – either as alone or in a form of reaction product with the analyte. In case of CPEs, the hydrophobic nature of the binder may facilitate the attachment of lipophilic molecules.) (b) Modifying agents are present in additional membranes [48,49]. (Also this principle originates from experiences with other types of electrode substrates)

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Jiří Zima, Ivan Švancara, Karolína Pecková et al. (c) Some lipophilic compounds (with voluminous molecules) can directly be dissolved in the binder [50], which represents particularly effective method of modification. (d) Water-insoluble reagents can be embedded into the paste via impregnating the graphite particles when the solvent is evaporated to dryness and modified carbon subsequently mixed with the pasting liquid chosen [51]. (This is the case of wellknown dimethyl glyoxime whose methanolic solution had been used for one of the first modifications of carbon paste [27].) (e) Rather sophisticated and often time-consuming are the procedures in which the graphite particles undergo a chemical pretreatment, resulting in a qualitatively new electrode material, interacting with the analyte through firmly bound functional groups [52]. (Also this approach belongs to pioneering attempts on how to purposely modify carbon paste in order to obtain a substrate capable of selective interaction with the target species [26].) (f) A specific way of modifying carbon pastes is anodisation or cathodisation; i.e., electrolytic treatment of their surface under extreme conditions [53,54]. Such intensive oxidation / reduction is able to remove the binder from the outer layer, which leads to principal changes in the surface character (so-called "alteration of the surface states" [6,43]) and the resultant electrode may exhibit markedly different behaviour. (g) Modifiers in solid state are mechanically admixed into the native carbon paste during or after its preparation [55]. For CPEs, such procedures are typical and mostly preferred. Apart from their simplicity or other advantages (see e.g. [3,6,9]), it allows one to modify previously made carbon pastes, including those already modified with another substance [56]. (h) Last but not least, even binary carbon pastes made of chemically active binder can be sorted among CMCPEs. For example, some pastes proposed for potentiometric titrations or electrochemical stripping analysis contain liquid ion-exchangers such as Aliquat [57] or tricresyl phosphate [58] that effectively act in highly selective ionpairing processes.

It seems that the above surveyed variability in modification of carbon pastes is the reason why the classification "CPEs vs. CMCPEs" is now used as the fundamental characterisation of the respective detection and sensing systems in almost all scientific presentations. In order to prepare unmodified carbon pastes, one can select various commercially available spectral graphite powders (for specification of some popular and commercially marketed products, see e.g. [4,9]), obeying these important criteria: (i) particle size in micrometres, (ii) uniform distribution of the particles, (iii) high chemical purity, and (iv) low adsorption capabilities. Occasionally, also other carbon materials have been tested as the electroactive moiety in carbon paste mixtures; for instance, soot [4], acetylene black [59], glassy carbon powder [60,61], pulverised diamond (of both natural and synthetic origin [62]), carbon nanotubes [63,64] or graphite intentionally grinded into the nanoparticles [65]. As a binder, binary carbon pastes contain organic liquids whose main function is the mechanical connection of the individual graphite particles. Besides, however, the liquid component determines a majority of typical properties of carbon pastes and hence, its selection is also a crucial task. Typical parameters of liquid binders for the preparation of

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carbon pastes are: (i) chemical inertness and electroinactivity, (ii) high viscosity and low volatility, (iii) minimal solubility in water and aqueous solutions, and (iv) immiscibility with organic solvents. For the proper preparation, numerous substances have already been recommended, predominantly mineral (paraffin) oils, Nujol® being apparently the most popular binder from this group. Often chosen are also silicon oils and greases; occasionally organic esters (e.g., triaryl phoshates or diaakyl phthalates [58]) or halogenated hydrocarbons (α-brom naphtalene [1,4,5]) or wax-like polymers (Kel-F®-resin [66]). Newly recommended are then substances coming from the development of new technologies and materials; a typical example being ionic liquids (e.g., n-octylpyridinum hexafluorophosphate [67]) which can also be directly mixed with graphite powder instead of ordinary binders. Graphite powder and a binder are mutually merged in quantities which are often based on empirical experience [3-13]; optimal carbon−to−pasting liquid ratio being usually within an interval of approx. 1.0 g : 0.5 ml. The two main components can be hand-mixed in a porcelain mortar by using pestle with sufficient pressure. According to some practical guides (e.g. [53,68]), it is advised to perform the operation in two, three consecutive steps, when the forming mass is ripped off from the rubbing wall to the bottom of mortar and homogenised thoroughly again. The paste is ready in a few minutes and can either be stored for later use or immediately used for preparation of the electrode. In order to employ soft, non-compact carbon pastes in proper measurements, it is necessary to choose a suitable support – properly designed electrode holder ("body"). Such an assembly, representing carbon paste electrode from a tradition point of view [3,10,13], can be designed in numerous construction variants: (i) glass or plastic tubings [69,70] and drilled plugs [1,5] for provisory or even permanent use (ii) piston-driven holders proposed by Monien et al. [21] enabling quick carbon paste renewal, (iii) carbon paste mini-electrodes [71], micro-electrodes [72,73] and ensembles of ultramicroelectrodes [65,74], (iv) carbon paste flow-cell detectors [75,76], or (v) specially designed holders in planar configuration, resembling the screen-printed electrodes (SPEs) in shape as well as by small dimensions [77] and allowing also electrical heating of the carbon paste filled in [78]. If any of the constructions mentioned in the survey above is to be recommended for starting the experiments with CPEs and CMCPEs, it would probably be the design (ii) − plastic body with piston (or micrometric screw, respectively) which offers the most effective removal of the used carbon paste, as well as periodical renewal of a thin carbon paste layer [3,53]. This construction, used for long time by the authors of this text [79], has undergone numerous practical innovations [80] with possible selection of one actual piece from a set of bodies with various capacities of loading cavities or the different diameters of the end-holes (giving the corresponding active surface). A collection of piston-driven carbon paste holders is also imaged in Figure 1, together with the individual tools and accessories for comfortable carbon paste loading, filling, or cleaning. The figure depicts: left, from top ... two identical constructions for common measurements with different end-holes (∅ = 2 and 3 mm., resp.); below ... two more robust variants with larger active surface diameter (∅ = 5 and 10 mm), devised for potentiometric and couloumetric indications; left, on bottom ... a couple of special prototypes of mini-CPEs (left: disc configuration, ∅ = 1 mm, right: planar design as plastic bar-shaped body with filling cavity, both mountable into the flow- cells for HPLC and FIA detection; in centre, above ... laboratory porcelain mortar with pestle used for carbon paste homogenisation; right ... a quintet of stainless rods with plastic handles for loading, filling,

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and cleaning CPE-holders of the a-d type; in centre, on bottom ... laboratory spatula and small office ruler (to imagine the real size proportions).

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Figure 1. Piston-driven electrode holders for carbon paste in various construction variants along with typical tools for the preparation of carbon paste and the respective electrodes. Description of the individual items shown in photo is given in the text; for other details, see [77] and [79,80]).

In the previous text, attention was paid to a role of both CPEs and CMCPEs in electroanalysis, featuring carbon paste as a unique electrode material with multi-component character and the corresponding mechanical properties. As pointed out, heterogeneous nature of carbon pastes is inevitably reflected in their specific behaviour that may differ from mixture to mixture, depending not only on the type of and quality of the two main constituents chosen and the construction variant used, but also upon the way in which the respective electrode or sensor is employed in practical experiments. All these aspects are shortly commented in the following section.

2.2. Physico-Chemical and Electrochemical Characterisation of Carbon Pastes and Carbon Paste Electrodes Despite numerous specific features, binary carbon pastes as well as their modified variants still belong to the group of solid carbon electrodes and therefore, CPEs and CMCPEs exhibit some properties typical for carbonaceous materials. Namely, one can quote the following: Chemical and electrochemical inactivity. Compared to mercury or noble metals, carbon is a fairly inert electrode material with substantial resistance against unwanted transformations of the chemical or electrolytic nature [3-5] and solely extreme conditions may cause deeper changes at the carbon electrode surface ([43,54]; see also above).

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Adhesiveness of the surface layer. Similarly to other carbon and graphite materials, the CPE surface is easily maintained as the substrate to attach any substance; e.g., modifier [3]. Adsorption capabilities of the carbon paste surface. Being perhaps the reason for the previous attribute, carbon materials exhibit a marked adsorptibility with affinity towards some neutral molecules [5]. (In the case of CPEs and CMCPEs, this ability can be suppressed or, oppositely, enhanced by the effect of the liquid binder contained in the paste [8].) Polarisation characteristics. In faradic measurements (associated with the electron transfer and current flow), common types of CPEs and CMCPEs can be polarised within a potential range of approx. −1.0 V and +1.0 V vs. SCE [3-5,13]. However, thanks to eventual choice of particular carbon paste constituents or special modifiers, the resultant potential window may be extended in both anodic and cathodic direction. (For instance, this can be accomplished using pre-treated graphite [81], atypical pasting liquid [58], or by adding a reagent, inhibiting the decomposition processes in the supporting electrolyte [3]). The properties, in which CPEs and CMCPEs differ from other electrochemical sensors, have usually been of particular interest in hitherto-published reviews and their comprehensive description can thus be found therein [3-13]. In this text, it seems sufficient to survey the individual specifics together with the essential characterisation and some representative reference(s) to the original bibliographic source(s), concerning in detail the respective feature, phenomenon or process. Structure and microstructure. The texture of common carbon pastes can be described as a "solid dispersion" of graphite particles in the liquid binder used. All the time since CPEs are known, microscopic structure of their surface was of eminent interest (see e.g. [4,82,83]). It has been repetitively proved that typical microstructure of carbon paste surface is a random array of carbon particles linked tightly together with pasting liquid whose molecules cover − as a very thin film of nanometric dimensions − practically each grain of graphite in the mixture. Instructive SEM-images of such carbon paste microstructure have been obtained a decade ago in our laboratories by scanning electron microscopy (SEM) when using carbon paste specimens made of graphite with spherical particles [60]. One shot from this set is shown in Figure 2, showing the microscopic appearance of such a carbon paste (left) and, newly, a schematic cut through the front surface layer (right). Minimal ohmic resistance. Special studies performed already in the early era of CPEs [84] showed that even intimately coated graphite particles possess − as the array – very low ohmic resistance and e.g. the carbon paste mixtures from mineral or silicone oils exhibit the resistance below 10 Ω [53,68]. (To date, it is not yet fully explained why the conductivity of carbon pastes with insulating binder is so excellent; some sources admitting e.g. tunneling effect analogical to that in solid semi-conductors [3]). Instability of carbon pastes in organic solvents. Due to the presence of oily binders in carbon pastes and their certain miscibility with organic solvents (e.g., MeOH, ACN, DMSO, and DMFA), both CPEs and CMCPEs are rather vulnerable in measurements, involving the use of such media or even aqueous solutions mixed with these solvents. The unavoidable disintegration of the paste [3-5] can partially be prevented by using carbon pastes containing more resistant pasting liquids (e.g., highly viscous silicone fluids [85]), stabilising modifiers in solid state [18] or − rather unexpectedly − via the choice of special graphite (some types of glassy carbon powder [86]).

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Figure 2. Microstructure of carbon paste made of glassy carbon powder (left) and the respective crosssection (right). Legend: graphite particles coated with pasting liquid in a) carbon paste bulk, b) the outer layer, c) graphite particles, d) thin film of liquid binder. (Redrawn from SEM images published in [60] and schematised according to [3-5,12]).

Ageing of carbon pastes. This phenomenon, in some reviews unnoticed, has been observed for carbon pastes being made of more volatile binders (such as organic esters [58]). Regarding common mineral and silicone oil-based paste mixtures, the ageing effect is less pronounced and such CPEs can be used for weeks (up to several months ) without evident changes in behaviour. Specific are freshly made carbon pastes that need a certain time for stabilisation and such "self-homogenisation" [53] may take one or even two days. Hydrophobic character of carbon pastes. Perhaps the most typical feature of CPEs and CMCPEs is lipophilicity of the electrode material. The principal consequence of such hydrophobicity is moderated reaction kinetics at carbon paste-based electrodes [3-5,43], which is reflected in irreversible behaviour of numerous organic redox-systems and may complicate their electrode transformations. Wide variability in interactions at the carbon paste surface and in the carbon paste bulk. Both CPEs and CMCPEs enable to utilise a variety of interactions and processes based on completely different principles, which can be documented on the following survey, involving classical electrochemical processes of either faradic or non-faradic character, physicochemical interactions of non-electrochemical origin, plus some synergistic mechanims such as combined and multi-step pathways: (i) Electrolytic processes. Numerous procedures rely on monitoring the electrode transformations associated with the electron transfer − i.e., oxidation and reduction of the target analyte(s) [3-5] −, involving various voltammetric, amperometric or

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coulometric techniques. In inorganic analysis with CPEs and CMCPEs, electrolysis is a key-principle of classical methods with anodic stripping voltammetry for the determination of heavy metal ions; however, electrolytic processes are yet more frequent in analyses of organic compounds, pharmaceutical derivatives and drugs, or various biologically active compounds [3,9,13].

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Afore-mentioned moderated reaction kinetics at the carbon paste surface − the so-called "quasi-reversibility" [3-5] − can be turn into more reversible behaviour by means of (a) hydrophilisation of the CPE surface by electrolytic oxidation [43,54] or with the aid of (b) special modifications [3,6,9]. (In the current-based measurements, both these approaches lead to a marked improvement of the signal-to-noise characteristics. Concerning the latter, the effect of some modifiers is outstanding and the resultant shift may be up to +/−500 mV [87]. This enables to "relocate" the response of interest, originally occurring at high potentials and obscured by high background, into optimal position within the operating potential window) (ii) Non-electrolytic processes. Carbon paste-based electrodes and sensors are also widely applicable in electrochemical experiments that are not connected with the current flow. This is the case of classical potentiometry [88] and computer-controlled stripping (chrono) potentiometry [89] with indication of the equilibrium potential, conductivity measurements [4], or chemical reactions releasing electrochemiluminescence signals [82]. (iii) Adsorption at the carbon paste surface. As already stated, liquid binder in carbon paste mixtures may reportedly enhance the already profound adsorption capabilities of the graphite particles themselves [8]. Mostly, however, these interactions seem to be more complex and their driving force is a combination of adsorption with other phenomena such as electrostatic effects, extraction, and ion-pairing [3,90-92]. (iv) Extraction (penetration) onto the carbon paste bulk. Based on controlled separation of a substance between two liquid phases, this process is one of the most valuable features of carbon paste-based electrodes and sensors [3,53,92]. The penetration of an analyte onto the carbon paste bulk is enabled by the presence of liquid binder and the more lipophilic pasting liquid the more effective extraction can be achieved [12,93]. Similar principles may also apply to other electrodes, but the necessity of an additional membrane is less attractive and technically more difficult than easy-tomade and extraction-enabling carbon paste substrates. A very important condition for successful pathway of extractive accumulations is an electrical neutrality of the analyte (uncharged form). In case of organic and biologically active compounds, this is being aimed by adjusting the proper pH at which such substances exist merely as neutral molecules [12,88,91]. Inorganic ions and complexes can be coupled with a suitable counter-ion and the corresponding ion-associates then extracted as well [3,58]. (The individual methods utilising the extractive pre-concentration often require a regular carbon paste renewal because the re-extraction process during the electrochemical detection is not fully completed and a certain amount of the analyte remains entrapped in the bulk [92]. With respect to electroanalytical applications, this phenomenon is quite reproducible (within ±5-10 % [53]) and does not hinder to

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use of traditional methods of quantitative analysis; nevertheless, it has to be considered when optimising the respective procedure.) (v) Ion-exchange and ion-pair formation. Both terms often symbolise the same process – electrostatic interactions between the charged counter-ions. In contrast to adsorption or extraction, the resultant effectiveness of such measurements depends primarily on chemical character of the species involved and the quality of electrode is of lesser importance. Also for these processes, carbon pastes represent convenient substrates thanks to a specific preparation and the presence of liquid phase. An ion-exchanger can either be added into the paste [6,57] or represented by the binder itself (used instead of common oils [58]). Ion-pairing principles can be advantageously employed in electrochemical stripping analysis to accumulate numerous inorganic anions such as halides and pseudo-halides, X− [12], or even complex structures of the [MenXm]p− type [7,9,13]. The proper pre-concentration mechanism is often a synergistic route where ion-pair formation may be accompanied by extraction or adsorption, intercalation or precipitation and such combinations then lead to an extremely high selectivity of the respective methods [3,9,12,58,89]. (vi) Electrocatalysis. Numerous electrode processes at carbon paste-based electrodes result from interactions requiring a suitable catalyst. Typically, it is a modifier which can be added directly in the carbon paste mixture, adsorbed or chemically immobilised at the electrode surface [3,6,94]. Whereas the previous processes described in paragraphs (iv) and (v) could be presented as highly selective, their sensitivity is rather worse and seldom allowing to determine species below the micromolar level [53]. In contrast to this, some procedures with electrocatalysisassisted processes belong to the most sensitive measurements achievable at CMCPEs, enabling to detect the analyte of interest down to the nanogram quantities [95]. Final remarks. Physico-chemical properties and electrochemical characteristics surveyed in sections 5.2.1 and 5.2.2 can be defined − or at least estimated − using empirically proposed experiments, representing a testing of carbon pastes [3,53,68]. The corresponding sets of measurements help to reveal typical properties of the carbon paste(s) examined and, in some cases, such a testing is the most straightforward way of how to characterise newly made carbon pastes, including mixtures made from constituents with hitherto unknown properties. The strategy of such tests involves sequentially arranged experiments, comprising various procedures and comparative studies. The whole scheme can be flexibly adapted, but the individual tests should include measurements with suitable model systems (e.g., Ag+ / Ag0; Fe(CN)63− / Fe(CN)64−; I2 / I− ; quinone / hydroquinone, Q / HQ; ascorbic acid / dehydroascorbic acid, AA / DHA, etc. [5,22,23,43]) for which the electrode behaviour at common carbon pastes as well as other carbonaceous substrates is well known and experimentally defined. Besides the specification of typical properties and parameters of a CPE under test, properly selected experiments may serve to assess the quality of both main carbon paste components and their optimal ratio. Herein, it can be concluded that testing procedures are nearly ideal platform to obtain the fundamental information about the analytical performance of a CPE or a CMCPE − i.e.,

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characterisation of the operating potential window, the background level, reproducibility, detection capabilities for analyte(s) of interest, interference effects from foreign species, etc. These data can then be exploited for successful use of carbon paste-based electrodes, sensors or detectors of all types and variants in practical electroanalytical measurements.

3. CARBON PASTE-BASED ELECTRODES AND SENSORS IN WATER ANALYSIS

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In order to demonstrate the possibilities and limitations of an electrode or a group of electrodes, it is obvious to survey the respective applications. In this article, the authors' team has elaborated a compilation in which almost all the methods and procedures concerning CPEs and CMCPEs in analysis of water samples are gathered. By summarising this extensive collection of scientific papers, the most typical achievements, somehow interesting or even unique results are highlighted with particular emphasis on practical aspects in analysis of drinking water and related samples. As already mentioned in the introductory part, such a choice has been made intentionally as it allows us to illustrate the versatility, vitality, and practical usefulness of carbon paste-based electrodes, sensors or detectors in their entirety. The pivotal part of this practical guide is a set of tables that systematically list inorganic ions, complexes, and molecules, as well as a variety of organic substances – from native compounds, via related derivatives, up to environmental pollutants or other toxic compounds of biological importance. To survey such abundant data in a comprehensive way, the types of the individual electrodes, methodical principles, typical experimental parameters, and other characteristics had to be surveyed by means of simplified schemes, abbreviations or symbols; all of them being denoted in section "Abbreviations and Symbols Used" (see above).

3.1. Determination of Inorganic Ions, Complexes, and Molecules Up until now, the database of practical procedures within the inorganic analysis of drinking water and related samples with both CPEs and CMCPEs comprises approx. one hundred contributions (see Table I, refs. [96-200]), having reported on the determination of about thirty elements across the whole periodical system, plus other two tens of inorganic compounds such as various anions, complex species or some neutral molecules. Namely, the commentary in the following text deals with noble metals (i.e., Ag [96-98] and Hg [99-109]), heavy metals (Cu [110-120], Zn [121-123], Cd [120,123-135], Pb [120,123-132,136-148], Tl [149], In and Sn [150,151], Sb and Bi [152-155]), metalloids (As [156,176,177]), iron-group metals (Fe [157-159], Co and Ni [160-164]), metals from the VI. and VII. groups (Cr [165,174], Mo [175], and Mn [166]), or other metallic elements, represented by alkaline and alkaline-earth metals (Li [167-169], Mg [170], Ca [171]), platinum metals (Pd [172]), or naturally occurring uranium (U [173]).

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Table I. Determination of Metal Ions and Complexes in Water Samples at Carbon Paste Electrodes and Sensors. Survey of Methods Form (species)

Type of CPE (specification) [modifier]

Technique (mode,ramp set-up)

Measuring principles; (method sequences)

Linearity range (det. limit; tACC; other specification)

Water sample (specification) [other samples]

AgI

C/TCP [C7H15SO3Na]

DPASV

- ion-pairing via modif. in situ; redn. + anodic detc.

3×10−12- 2×10−5 M (< 1 pM; 120 min.)

tap water

AgI ; AgnLm

C/MO (unm.)

SWSV

(5:1) - diffn. of AsIII and AsV possible - s.e.: 0.01 M KCl + 10−4 M EDTA - no interfs from common ions

1×10−8 - 5×10−7 M (5 min.)

natural water, water-CRMs

- redn.: FeIII → FeII with NH3OH+; - interfs supprd. by using Nf-mb.

[158]

1×10−8 - 2×10−6 M (3 min.) 2×10−8 - 1×10−5 M Ni 1×10−8 - 1×10−6 M Co (5×10−9 M Me, 3 min.) 8×10−9 - 7×10−7 M (1×10−9 M, 7 min.)

natural water, water-CRMs

- CMCPE chem. regnt.

[159]

sea water (synth. and real)

- NN: 1-nitroso-2-naphtol; s.e.: 0.5 M monia buffer (pH 9); anals. under expedition conds. (at ship board)

[160]

0.05 - 5 ppb (10 min.)

natural water, [fly ash]

1-6000 μg l−1 (5 ng l−1, 12 min.)

tap water, mineral water

5×10−8 - 2×10−5 M 4×10−9 - 1×10−6 M (1×10−10 M, 5 min.) 1×10−9- 5×10−7 M (Bi: 5×10−10 M; Sb: 1×10−9 M; 3 min.) 5-250 ppb (3 ppb; 15 s.) 2×10−6 - 7×10−6 M

C/PO [1,10-phen + Nf]

CV, DPV

C/PO [1,10-bipy + Nf] C/SO (+ MF) [Ni: DMG, in situ Co: α-NN, in situ] C/MO (+ phen.) [HTTA, in situ)

CV, DPV DPCSV

NiII

C/Nj [DMG]

DPCSV, FIA-EC

NiII

C/Nj [Dowex 50W]

AdSV

CrIII, (CrVI)

C/Uv [DPC]

AdSV

- accum. via modifier; - oxidn. of chelate formed

(< 1 mgl−1)

sea water

C/MO [PAN]

DPCSV

- accum. via compl.; - oxidn. + cathodic redn.

1×10−8 - 1×10−7 M (7×10−9 M, 200 s)

sea water (CRM)

FeIII FeIII CoII, NiII CoII

MnII, (MnVII)

AdSV, 2.5th-DV

- accum. via form. of ternary compl.; redn. - accum. By chelating + adsorp.; - cathodic redn. - o.c.accum by ion-pairing - redn. + reoxidn.

natural water

- HTTA: thiophen-carboxylic trifluoroacetate; s.e.: 0.1 M AcB - DMG: dimethylglyoxime; CMCPE regnt. in 1 M HNO3; s.e.: ammonia - buffer (pH 8) ; no interfs. from O2 - s.e.: 0.005 M HCl (pH 3); interfs. from HgII and AgI - DPC: 1,5-diphenylcarbazide; high selectivity; removal of oxygen not necessary - MnVII chem. red. with H2O2; some interfs can be supprd. by masking

[156] [157]

[161] [162] [163] [164] [165]

[166]

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LiI

C/Nj [λ-MnO2 , spinel]

DPASV

- intercalation via modif; - anodic reoxidn.

8×10−5 - 0.01 M (6×10−7 M, 30 s)

natural waters, [pharm. forms]

MgII

C/Nj (+ MF) [Na-TP, in situ]

SWV

- adsorp. accum. by compl. - cathodic redn.

6×10−9 - 9×10−8 M (5×10−10 M, 60 s)

tap water, [human urine]

CaII

C/MO [AR-S, in situ]

AdSV, (2nd-DV)

- formation + adsorption of complex formed - cathodic redn. - adsort. accum. by compl. - anodic reoxidn. - accum. By chelating - anodic reoxidn.

3×10−8- 2×10−6 M (9×10−9 M; 90 s)

water samples [soils]

1-450 μg l−1 (1 μg l−1, 120 s)

fresh water, [catalysts] natural water (ground origin)

PdII UVI

C/PO [TR, in situ] C/SO [propyl-gallate]

DPASV DPV

(< 1×10−7 M)

- t.: type; s.e.: borate buffer or TRIS (pH 8-9); no interfs. from Me(a) *) - TP: thiopentone; s.e.: phosphate buffer (pH 10.8); no interfs. from Me(ae), Me(a), AlIII, or biol. matrs. - AR-S: Alizarin Red "S"; 0.015 M KOH + 2×10−5 modif.; EP = −0.89 V vs. ref.; rgn.: 0.2 M HCl (120 s). - TR: thioridazine; incl. studies on interfs. from related metals - CPE designed for remote sensing (connected to long shielded cable)

[167] [168] [169] [170]

[171] [172] [173]

Table II. Determination of Inorganic Anions and Molecules in Water Samples at Carbon Paste Electrodes and Sensors. Survey of Methods Form (species)

Type of CPE (specification) [modifier]

Technique (mode,ramp set-up)

Measuring principles; (method sequences)

Linearity range (det. limit; tACC; other specification)

Water sample (specification) [other samples]

CrO42−

C/MO [DPC]

ASV

- accumulation via modifier - oxidn. of complex formed

(< 1 mgl−1)

sea water

- ion-pairing with modif. - cathodic redn. - accum. via redox-compl. - anodic reoxidn. - accum. As chelate; oxidn. - cathodic redn. - electrolytic. accum. - oxidn. by const. current - formation of ion-pairs; - indication of changes in equilibrium potential

5×10−7 - 5×10−5 M (5×10−8 M, 300 s) 2×10−7 - 5×10−6 M (1×10−7 M, 20 min.) 1×10−8 - 1×10−7 M (7×10−9 M, 200 s) 3-250 ppb ( 1, where θ is the surface coverage and fMe activity coefficient. After insertion of these expressions into the Eq. (1) one obtains

ML

Er (θ ) = E0 +

RT c Me z+ f Me z+ ln > Er θ f Me z+ zF

(2)

I [μA]

where fMez+ means activity coefficient of metal ion in the solution; fMe activity coefficient of metal ion in the bulk and cMez+ concentration of metal ion. -6

B -4

A

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-2

δ 1/2

0

A

B

2

ΔEp 4 -300

-500

E [mV]

-700

Figure 4. Cyclic voltammogram of lead(II) in 0.5M potassium chloride. Silver solid composite electrode CAgE20, in absence of oxygen, scan rate 25 mV.s-1, 25 mg.L-1 lead(II). Peaks A correspond to monolayer formation and dissolution, respectively; Peaks B correspond to bulk formation and dissolution, respectively.

In the presence of chloride ions (in comparison with the situation in absence of chloride ions) the underpotential shift ΔEp of the anodic peaks decreases (in acidic as well as neutral solutions), the area under the anodic bulk peak is enlarged more than two-times and the symmetry of both anodic peaks is increased considerably [58]. There are some other possibilities of mathematic description of this effect, but practically all of them are purely formalistic and do not involve all the parameters: influencing metal

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ions concentration, nature of the electrolyte, specific interaction occurring between the electrolyte and the solid phase. The theory of the UPD effect is more complicated and extended, however for purposes of analytical applications of composite electrodes in waters this simple qualitative and quantitative description seems to be sufficient.

2.2. Voltammetric Determination of Inorganic Compounds in Drinking water Using Composite Solid Electrodes

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2.2.1. Determinations Using Silver Solid Composite Electrode The applicability of this electrode in inorganic voltammetric analysis is very wide. 2.2.1.1. Determination of Lead, Cadmium, Copper, Thallium, and Bismuth, their mutual interferences, effect of presence of other metal ions Determinations of lead in waters using silver solid composite electrodes have been described in many papers, e.g., [10, 42, 58]. The first two of mentioned papers dealt with the use of this type of composite electrodes for explanation of some interesting effects connected with the UPD effect and lead was utilized as a very suitable analyte for their explanation. The composite silver electrodes, like other solid electrodes, suffer from lower reproducibility and worse renewable surfaces in comparison with HMDE. It is possible to improve their quality by polishing with alumina and/or by application of electrochemical pretreatment (cleaning cycles or cleaning potential) before voltammetric measurements. To improve the reproducibility of the determination of Pb(II), Cd(II), Cu(I), Cu(II), Tl(I), two procedures have been found advantageous on CAgEs [10, 42, 58, 63-65]: The first procedure consisted of repeated potential cycling of the CAgE before the ASV measurement. The program of the measurement thus included an adjusted number of the cleaning voltammetric cycles (from -900 to -150 mV and back), and 2 s after the end of the last cycle the anodic stripping voltammetric measurement started. During the second process the potential of the composite electrode was kept at a positive potential value (+250 mV) for a definite time (5 - 25 s) and after a short interval (2 s) the anodic stripping voltammetric measurement was started. The repeatability was tested by twelve times repeated voltammetric scan of lead(II) solution. According to the statistical tests [70] (Dean-Dix or Grubs tests on the confidence level 0.95), the first two records had to be omitted to enable the use of arithmetic mean, i.e., to achieve the normal distribution of results. It is advisable to apply the median instead of arithmetic mean (calculated from the third and subsequent measured records) for the evaluation. The achieved skewness amounted to –0.5 - +0.5 and excess from +2.1 to +3.5. The best results yielded insertion of 20 cleaning cycles or insertion of the cleaning potential +250 mV for 13 s before the start of measurement (the relative standard deviation was improved from 1.01 % (without cleaning) to 0.81 % and 0.84 %, respectively, but after leaving out the first two results the confidence interval was not improved practically at all) [58]. The reached reproducibility of the voltammetric measurement is sufficient and confirms that the activity of the electrode surface remains constant even during a long time measurement without mechanical pretreatment of the electrode surface.

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According to the literature sources ([10, 42, 58]), the results of anodic stripping voltammetry of lead(II) in acidic or neutral media have shown that the presence of graphite powder in the electrode material did not influence the response of the solid composite electrode directly, but its higher conductivity improved the quality of the recorded curves in contrast to the electrodes without graphite [58]. The UPD effects observed on the electrodes with and without graphite were very similar and so we can conclude that the electrochemical reaction occurs on the metallic particles in the electrode mixture only. This fact is in a good agreement with literature [69], p. 145, i.e., the UPD effect on graphite is weaker (see below) in comparison with composite or metal electrodes and less reproducible [71-73]. On the other hand, the results obtained further showed an UPD effect on composite electrodes consisting only of graphite and methacrylate resin. Two different electrodes with inner diameters of 1 and 2.5 mm were tested in [58]. The ratio of their surfaces was approximately 1:6, the voltammetric signal corresponding to the monolayer dissolution increased in the case of the electrode of 2.5 mm diameter, but its signal reproducibility was worse [58]. Most thoroughly studied was the determination of a few heavy metals described in [58]. The results of ASV of lead using the CAgE have shown that the sensitivity of the measurement increased with increasing amount of silver in the composite electrode. The dependence of the anodic peak current (monolayer peak) on the concentration of lead(II) ions was measured in the range from 10 to 150 μg.L-1 using the silver solid composite electrodes with the content of silver 15, 20 and 40 % (m/m), respectively. In all cases linear dependences of anodic peak currents vs. lead(II) ion concentrations were obtained; the slopes of these dependences had the values 6.92 nA.L.μg –1 for the 15 % electrode, 36.10 nA.L.μg –1 for the 20 % (m/m) electrode and 46.30 nA.L.μg –1 for the 40 % (m/m) electrode. The silver composite electrode containing 20 % (m/m) of silver and of 1 mm inner diameter was selected for all further measurements because of higher sensitivity of measurement and good reproducibility compared with the electrode containing 15 % (m/m) of silver. The measurements carried out with electrode containing 40 % (m/m) of silver were more sensitive, but the value of the background current was too high [58]. As it was reported in [10, 42], the optimum conditions applicable for the DP anodic stripping voltammetric determination of lead using CAgE are: pulse amplitude 25 - 90 mV, pulse width 100 ms, interval between pulses 200 ms, scan rate 20 mV.s–1, the deposition step realized by stirring of the sample at the deposition potential –850 mV and 5 s rest time period without stirring at the same potential. The resulting anodic polarization carried out in acidic media yielded well-developed anodic peaks. Analogically to the differential pulse stripping voltammetry (DPASV) measurement with metallic silver electrode, the dissolved oxygen, present in the measured solution, influences the shape of the resulting polarization curves recorded with the composite silver electrode. Furthermore, it has been found that the DPASV of lead is strongly affected also by the hydrogen and chloride ions concentration. The chloride ions play the most important role in the lowest concentration range (“ppb” and “sub ppb” ranges). The influence of chloride ions on deposition and stripping of lead could be used effectively, from the analytical point of view, in stripping voltammetry [10, 42, 58]. The height of the more positive anodic peak of lead in perchloric acid solution increases linearly with increasing Cl- ions concentration up to 0.6 g.L-1 (this limit depends on the accumulation time (tacc)) and nonlinearly at higher concentrations and simultaneously the peak potential is

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shifted to more positive potential values [10, 58]. The DPASV measurement, carried out under identical conditions as described above in 0.1 M HClO4 and 0.01 M KCl medium, showed that the dependence of the value of the anodic current of the peak at the potential -390 mV on lead ions concentration (from 10 to 150 μg.L-1, 200 s deposition time) was linear with the slope 14.9 nA.L.μg-1 [10, 58], i.e., more than 4 times higher in comparison with the slope obtained in absence of chloride ions. The more negative peak (i.e. bulk stripping) was not evaluable and thus was not changed with increasing lead(II) ions concentration as in absence of chloride ions. In the concentration range from 0.1 to 2.4 mg.L-1 of lead(II) (60 s deposition time), dependence of the anodic peak current at the potential about –400 mV on lead(II) ions concentration was found linear with the slope 2.12 μA.L.mg-1, i.e., approximately comparable with that one in absence of chloride ions. Under these conditions, also the second stripping peak appeared at the potential -540 mV, the corresponding anodic current increased nonlinearly with increasing lead(II) ions concentration and its maximum height reached 10 % of the height of the monolayer stripping peak [10, 58]. As follows from the presented data, the DPASV curves obtained by composite silver electrode differ substantially from similar curves obtained by metallic silver electrode. The difference in the course of DPASV curves obtained on metallic and composite silver electrodes results probably from the different surface structure of both types of electrodes. From the analytical point of view, the application of composite silver electrode in acidic solutions containing chloride ions is preferable [10, 42, 58]. On silver metallic electrodeonly the monolayer stripping peak can be used for ASV determination of lead only at low concentrations and in very narrow concentration interval (0.1-1.0 μg.L-1). In the case of the composite silver electrode only one anodic peak is applicable for the determination of lead as well, but in a wider concentration range (0.1 μg.L-1 - 0.1 mg.L-1). In article [58] a new interesting and relatively not so frequently applied method for evaluation of peaks recorded using CAgE was described: differential pulse anodic stripping subtractive voltammetry (DPASSV) [74]. It is one of the methods suggested for easier evaluation of results at low concentration of the analyte. The improvement achieved by such background correction can be expected especially in relatively low concentrations. First, the voltammogram is recorded with chosen accumulation time, and after a defined time interval the same measurement is repeated with zero accumulation time. Finally, the obtained i-E curve is mathematically subtracted from the first one. An application of this method was tested on the solutions containing from 2.5 to 37.0 μg.L-1 of lead(II) ions (accumulation time 90 s). Thus obtained calibration curve was linear in the case of DPASV as well as of DPASSV (correlation coefficient 0.981 and 0.984, respectively), but the slope of the DPASV calibration curve was 2 times higher than that of the DPASSV one (7.65 nA.L.μg-1 and 3 nA.L.μg-1, respectively). The symmetry and shape of the evaluated DPASSV peaks were much better and it was not affected by supporting electrolyte decomposition. The application of DPASSV for the determination of lead is advantageous when its concentration in the analyzed solution is below 10 μg.L-1 [58]. In [58] the determination of lead using CAgE20 under presence of other metals was described. This determination is not as easy as in the case of HMDE or non-composite silver metallic electrode. The monolayer as well as bulk peaks are affected by the presence of other metals. Very important role plays the presence of copper in the analyzed sample. Using the DPASV method the copper yields an anodic peak at the potential about –180 mV, which did

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not overlap either the monolayer peak of lead at –430 mV or its bulk peak at –590 mV. Most probably, copper is not deposited underpotentially on silver [69, 75], and thus the copper deposits in three dimensional nuclei of pure copper, occupying a small fraction of the surface and leaving enough space for the submonolayer deposition of lead and cadmium [75], i.e., the copper monolayer is not completely formed prior to formation of second and other layers. According to another theory [69], lead and cadmium are deposited on copper, after the electrode surface is covered by the copper monolayer and Cd and Pb underpotential shifts have practically the same values on copper as on silver surface ([69], p. 158). However, this explanation seems to be a bit complicated and not probable. Under identical conditions (deposition potential and time) and identical concentration of lead(II) and copper(II) ions, the anodic peak of copper, obtained by measurements with the silver composite electrode, is lower than the anodic peak of lead, i.e. the sensitivity to copper is lower then the sensitivity to lead (e.g. at accumulation time 60 s at the potential –850 mV the peak corresponding to the concentration of 1 mg.L-1 of copper(II) ions was only two times higher than the peak corresponding to the concentration of 0.15 mg.L-1 of lead(II)ions). This result is consistent with the above mentioned theory that copper occupies only a small fraction of the surface because copper ions are adsorbed in three-dimensional layers [58, 69, 75]. The copper accumulated on the electrode surface affects positively anodic current of lead - the monolayer peak. Using DPASV for the determination of lead (100 μg.L-1) in presence of copper(II) ions (from 100 to 1000 μg.L-1), the monolayer peak of lead increased linearly with increasing copper concentration (slope 0.39 nA.L.μg-1 of copper(II)). With increasing copper(II) ion concentration the bulk peak of lead (-560 mV) decreased and reached zero value at copper(II) concentration of approx. 1000 μg.L-1. In the presence of high excess of copper, the bulk-peak of lead increased nonlinearly and reached maximally one half of the lead monolayer peak height [58]. The influence of bismuth(III) ions on the DPASV response of lead is somewhat different. The anodic peak of bismuth, registered by DPASV under the above described conditions, appeared at the potential -70 mV and is well separated from the anodic monolayer peak of lead. It was, however, found that the deposited bismuth caused an increase of the bulk layer peak current of lead, and a decrease of the monolayer peak of lead at higher bismuth concentration. Further results have shown that the described effect of bismuth on the DPASV response of lead(II) ions was eliminated by the simultaneous deposition of copper(II) and bismuth(III) ions [58]. Under the following conditions: 0.1 M KCl and 0.005 M HCl, DPASV, deposition potential -900 mV, cadmium(II) produces an anodic peak at the potential -530 mV, very close to the value of the bulk peak of lead at -550 mV. The monolayer peak of lead (-420 mV) is, however, not affected by the described anodic peak of cadmium. In the concentration range of lead(II) from 25 to 200 μg.L-1 and in the presence of 300 μg.L-1 of cadmium, the dependences of both anodic dissolution current values concentration were linear. The anodic bulk layer peak current can be measured with an acceptable accuracy only if cadmium is present in maximally six-fold excess over lead. Lead can be, however, determined by the measurement of the monolayer peak current. Some achieved results are presented in Table III. The effect of copper(II) ions on the ASV response of lead and cadmium corresponds to the effect of copper ions on the response of lead itself. Simultaneous deposition of lead, copper and cadmium ions leads to an increase of the monolayer peak of lead and to a decrease of the bulk peak of lead and to the decrease of the cadmium anodic peak. The results have shown that the monolayer

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peak of lead increased linearly with increasing lead(II) ion concentration, and the bulk peak of lead decreased. The bulk peak of lead at the potential –580 mV actually corresponds to the sum of the lead bulk layer peak and the cadmium peak; both these peaks cannot be measured separately. The high concentration of cadmium (100times higher than lead concentration) excluded the determination of lead, because the bulk peak of lead is overlapped or deformed by the monolayer and bulk peaks of cadmium (at the potential -530 mV and -810 mV, respectively) and an accurate measurement of the monolayer peak at -410 mV would be complicated [58]. NaF, as a supporting electrolyte seems much more suitable than KCl for the DPASV determination of cadmium using the CAgE. The calibration curve of cadmium (0.05M NaF supporting electrolyte, deposition potential -1150 mV, deposition time 50 s) is linear in a wide concentration range (slope 9.00 nA.L.μg–1, correlation coefficient 0.996) and its peak is not affected by the presence of lead(II) ions (minimally up to 750 μg.L-1 of cadmium(II) and 125 μg.L-1 of lead(II)) [58]. Under the described conditions, the DPASV of thallium yields an anodic peak at the potential -635 mV and practically does not affect the monolayer peak of lead(II) ions and only partly the bulk peak, but on the contrary, its peak is disturbed by the presence of higher concentration of lead. In the presence of 250 μg.L-1 of thallium(I) ions, the monolayer lead peak increases linearly with increasing lead(II) ion concentration in the range from 20 to 250 μg.L-1. The bulk peak increases also linearly, but only at higher concentrations (approximately from 75 μg.L-1) of lead(II), while the peak corresponding to thallium stripping decreases linearly and if the concentration of thallium(I) and lead(II) were equal, the peak of thallium practically disappeared [58]. At higher thallium concentration (1000 μg.L-1) and in the concentration range of lead(II) ions from 12.5 to 137.5 μg.L-1, a similar ASV response was obtained. The error of ASV determination of lead(II) in the presence of an excess (100times) of thallium is (under the given conditions) acceptable, see Table III. In the presence of higher concentration of copper (approx. 1 mg.L-1) and thallium (approx. 1 mg.L-1), and in the concentration range of lead(II) from 20 to 150 μg.L-1, the corresponding differential pulse anodic stripping voltammogram (accumulation time 100 s) contained three peaks: monolayer peak of lead at -430 mV (its anodic current increased linearly with increasing concentration of lead(II) – slope value 8.70 nA.L.μg-1), peak of thallium at -635 mV (its anodic current decreased linearly with increasing concentration of lead(II) – slope -3.50 nA.L.μg-1) and finally the peak of copper, which practically did not depend on the lead(II) ion concentration [58]. No significant effect of other metal ions (such as Zn2+, Mn2+, Ni2+, Co2+, Se4+, As3+, Fe3+) on described DPASV determination of lead(II) ions was observed. Lead in the presence of many other metal ions can be thus determined by DPASV using silver composite electrode with an acceptable accuracy as shown in Table III [58]. By the determination of lead in higher concentration ranges after removing oxygen, both peaks, the monolayer as well as the bulk, were observed on the voltammogram. If the same voltammogram was recorded in the presence of oxygen, the bulk peak considerably decreased, while the half height width (δ1/2) of the monolayer peak increased.

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Table III. Effect of presence of various metals on the determination of lead by DPASV using silver composite electrode CAgE Lead present [μg.L-1] 10.0 20.0 112.0 20.0 20.0 12.5 25.0 112.0 12.5 12.5 25.0 25.0 25.0 12.5 12.5 200.0 25.0

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25.0

50.0 50.0

200.0

Other metals [μg.L-1] Me 10.0 Me 25.0 Cu2+ 1000.0 Cu2+ 1500.0 Cu2+ 1500.0 Cd2+ 12.5 Cd2+ 12.5 Cd2+ 25.0 Cd2+ 25.0 Cd2+ 300.0 Cd2+ 25.0 Cu2+ 1000.0 Tl+ 1000.0 Cd2+ 25.0 Cu2+ 3500 Me 100.0 Cd2+ 50.0 Cu2+ 1000 Me 100.0 Cd2+ 50.0 Cu2+ 2000 Cd2+ 25.0 Me 100.0 Cd2+ 25.0 Cu2+ 4000 Cd2+ 25.0 Cu2+ 4000

Lead determined [μg.L-1] 10.1±2.2 20.1±2.6 109.0±8.2 20.2±2.8 20.3±2.6 13.3±1.5 24.9±0.9 110.0±6.5 11.7±1.3 13.1±2.4 26.5±4.5 25.5±3.8 24.8±3.2 11.3±2.3

Difference [%] 1.00 % 0.50 % -2.68 % 1.00 % 1.50 % 6.40 % -0.40 % -1.79 % -6.40 % 4.80 % 6.00 % 2.00 % -0.80 % -9.60 %

13.1±2.0 201.0±1.6

4.80 % 0.50 %

24.4±4.0

-2.40 %

25.1±1.2

0.40 %

50.1±7.1 48.7±4.5

0.20 % -2.60 %

201.0±1.8

0.50 %

Me – the mixture of Zn2+, Mn2+, Ni2+, Co2+, Bi3+, Se4+, As3+, Fe3+, Cd2+, Cu2+, Tl+.

The effect of the presence of surface-active substances on the described DPASV response of lead was examined. The DPASV measurement of lead at the concentration level of tens of μg.L-1 showed that Triton X-100 had no significant effect on the determination results up to 2 - 5.10-4 % (m/m), where a small increase (approx. 6 %) of the monolayer peak current of lead was observed. At higher Triton X-100 concentration (approx. 10-3 % (m/m)), the increase of the monolayer peak current is close to 10 %. These results confirmed the possibility of the determination of low amounts of lead in natural water samples by DPASV using the CAgE without sample pretreatment [58].

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Example of Pb(II) determination in drinking (tap) water was described in [58]: To 10 mL of freshly sampled tap water 1 mL of 1 M HClO4 and 0.01 mL of 1 M KCl were added, and thus prepared solution was analyzed using silver composite electrode by DPASV under the following conditions: deposition potential –850 mV, deposition time 500 s, rest time 5 s, scan rate 20 mV.s–1, cleaning time 3 s at cleaning potential 250 mV. The obtained lead(II) ions concentration was 5.65±0.50 μg.L-1. The sample was also analyzed with the addition of copper(II) ions (1 mg.L-1), and the result of 6.15±0.45 μg.L-1 was achieved. The result 6.02±0.40 μg.L-1 was obtained by the measurement of the same sample using HMDE after removing the surface active substances (SAS) by microwave digestion. A few statistical parameters were calculated by evaluation of the obtained results in the same paper: In the case of sample in absence of copper the following parameters were calculated: limit of decision 1.5 μg.L-1, limit of detection 2.2 μg.L-1; limit of determination 3.5 μg.L-1 and in the case of presence of copper: limit of decision 1.1 μg.L-1, limit of detection 3.1 μg.L-1; limit of determination 4.0 μg.L-1 (both determined concentrations of lead(II) ions in tap water were above the limits of determination). The attainable limits of detection on composite electrodes are higher in comparison with mercury electrodes; nevertheless, they are sufficient from the point of view of analysis of environmental samples.

2.2.1.2. Determination of Nitrates and Nitrites In papers [10, 62] the way of nitrates determination in aqueous solutions using CAgE20 was described. After the mechanical treatment, the composite electrode must be activated by cyclic voltammetry (twenty five cycles from –100 to –1500 mV, scan rate 250 mV.s-1, supporting electrolyte 0.1 M NaOH). The activation of the electrode surface by time and potential defined polarization before each measurement assures good reproducibility of the obtained results. Using differential pulse voltammetry (DPV) the reduction peak of nitrate ions is registered at the potential –800 mV and the reduction peak of nitrite ions at the potential –1250 mV. To improve the repeatability of the measurement, the potential of the electrode before each measurement is to be kept for 0.2 s at the potential of +100 mV and then for the same time interval at the potential of -1500 mV. This potential jump must be repeated ten times and immediately after that the potential scan from –100 to –1400 mV can be started [10, 62]. The height of this peak at the potential –800 mV depended linearly on the nitrate ion concentration in the range from 5 to 400 mg.L-1. According to the results published in [62], the limit of detection amounted to 7 mg.L-1. While the silver disc (non-composite) electrodes exhibited peak-shaped responses when linear scan voltammetry (DC) was applied, the response of DC voltammetric measurement of nitrate ions in sodium hydroxide media with solid silver composite electrode had a sigmoidal form. In this case the limiting current increased nonlinearly (parabolically) with increasing scan rate. The same dependence measured with silver metal electrode is, on the other hand, linear. The obtained results led to the confirmation of presumption stated above that the solid composite electrode with random distribution of silver particles on the electrode surface exhibits the electrochemical behavior similar to the ensemble of microelectrodes [6, 62, 67]. Using this electrode, the content of nitrate ions was determined in the samples of river and tap waters. The obtained results were in good agreement with the results obtained by a spectrophotometric method (Table IV) [62, 67].

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Table IV. Voltammetric determination of NO3- ions in waters using CAgE20 and metallic (non-composite) AgE compared with spectrophotometric determination of NO3NO3- in [mg.L-1] Tap water, Prague

CAgE20 22.1

AgE 22.9

Spectrophotometry 24.1

Spring water, North Bohemia Waste water

138.0 29.8

139.5 30.5

132.0 26.5

2.2.1.3. Determination of Halides Cathodic direct current as well as differential pulse stripping voltammetry with the use of the silver composite electrode prepared from silver, graphite powder and methacrylate resin enables direct determination of chloride, bromide as well as iodide ions in waters and aqueous solutions. The achieved limits of detection for all determined ions are sufficient for practical analytical purposes (tap water, natural water etc.) (See Table V) [35, 63, 65]. The achieved results of their analysis are fully comparable with results achieved using other electrodes [35]. The results of experiments with silver composite electrodes have shown that a welldeveloped voltammetric response of halide ions was obtained by CSV measurements in acid electrolytes, as optimum medium 0.1 M nitric acid was found [65]. Both types of the tested silver composite electrodes (CAgE15 and CAgE20) were used for the voltammetric measurement, better developed current/potential curves were achieved by CAgE20 electrode. Similarly to all other above and below described determinations, Ag/AgCl/3 M KCl electrode was used, but in case of halides it was separated by a salt bridge filled with saturated NaNO3 solution.

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Table V. Statistical Evaluation of Results Obtained by DC or DP Voltammetric Determination of Halides Using Silver Composite Electrode CAgE20 [65]

Rel. standard deviation [%] Skewness Excess Limit of decision [mg.L-1] Limit of detection [mg.L-1] Limit of determination [mg.L-1]

Chlorides DPV 13.0 -0.58 2.99 0.76 36.9 2.10

Chlorides DCV 5.0 -0.02 1.46 0.32 60 0.78

Bromides DCV 0.54 0.27 2.32 0.04 0.08 0.32

Iodides DCV 5.20 0.52 2.4 0.27 0.78 2.6

Iodides DPV 9.30 -0.56 2.19 0.12 0.26 0.84

The potentials of the cathodic stripping peaks of halide ions measured in 0.1 M nitric acid exhibit the values: +225 mV for chloride ions (DPV) (Eacc = +350 mV), -200 mV for bromide ions (DCV) (Eacc = +170 mV) and -300 mV (DPV) and -500 mV (DCV) for iodide ions (Eacc = +400 mV), where Eacc denotes the applied accumulation potential [65]. The results of measurements are independent of the method used (DCV or DPV), but the application of DPV is preferable for the determination of chlorides, due to a very good elimination of the background current. For the determination of iodides and bromides, on the other hand, both DPV and DCV techniques can be applied with good results. In order to obtain reproducible results it is advisable to apply some mode of electrochemical pretreatment of the electrode surface. The following pretreatment was found to be optimal: insertion of 50

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potential cycles (cyclic voltammetry, initial potential Ein = +350 mV, final potential Efin = -900 mV, scan rate 500 mV.s-1 with 0.1 s time delays at initial and final potential values) before the beginning of measurement of each new sample. It is also advisable to insert one cleaning cycle (under identical experimental conditions as presented above) before each recorded scan and 5 s long delay at the accumulation potential [65]. At the concentration of chlorides and bromides 5 mg.L-1 and at the concentration of iodides 10 mg.L-1 the dependence of the peak heights of the halide ions on the time of accumulation was tested, using the given values of accumulation potential. The dependence of the cathodic peak-height (DPCSV) of chlorides on the time of accumulation proved to be linear in the range from 0 to 60 s with the slope 9.30 nA.s-1 and intercept –170 μA, correlation coefficient 0.9957 [65]. The dependences of the DC and DP peak-heights of iodides on the time of accumulation were not linear in whole time interval and this dependence had the sigmoidal shape of a Frumkin isotherm (the surface of the electrode was fully occupied by the deposited substance in the monolayer and formation of second and further layers started using longer tacc). From the analytical point of view, the suitable time of accumulation should be situated in the linear part of the dependence, e.g., the accumulation time for iodides 10 mg.L-1 was 50 s for both DPV and DCV. The dependence of cathodic peak-heights of bromides (5 mg.L-1) on the time of accumulation started to be linear at longer accumulation times (more than 40 s) due to the particle-particle interactions on the electrode surface. The suitable accumulation time for bromides in this concentration was therefore 60 s for DCV. Under the above-mentioned conditions the dependences of the peak-height on the concentration of chlorides, bromides and iodides in aqueous samples on the silver composite electrode CAgE20 were studied in [65]. Method of standard addition was used for evaluation of concentrations of these halide ions to eliminate the matrix effect of the sample. The chlorides were measured in the concentration interval of the order of mg.L-1 (DPV, Eacc = +350 mV vs. 1 M Ag/AgCl, tacc = 60 s) in the supporting electrolyte (0.1 M HNO3). The peak of chlorides has shifted to more negative potentials with increasing chloride ions concentration. The dependence of the peak-height on concentration of chlorides was linear with the slope -171.2 nA.L.mg-1 and with the correlation coefficient 0.9989. The determination of bromides was described in [65] in concentration interval 1 – 50 mg.L-1 in the supporting electrolyte (0.1 M HNO3) (DCV, Eacc = 170 mV vs. 1 M Ag/AgCl, tacc = 60 s). The peak of bromides was located at the potential –200 mV and with increasing bromide ions concentration it has shifted to more negative potentials. The signal of bromides has the shape of two overlapping peaks (probably due to formation of various complexes with silver). The dependence of the peak-height (for the purpose of determination of bromides the more negative peak was used because of its better evaluability) on the concentration of analyte in lower concentration range is not linear – it shows nonlinear shape in the region of small concentrations (up to 10 mg.L-1, where the peaks obtained were too small and their evaluation was uncertain). Such nonlinear shape of the concentration dependence could be explained by particle-particle interactions; it corresponds to the sigmoidal shape of the Frumkin isotherm. In higher concentration range (10 – 50 mg.L-1) the dependence of the peak-height on the concentration of bromides is better approximable by a linear equation with the correlation coefficient 0.9952 (slope -357 nA.L.mg-1). The iodides were in the same paper determined in the concentration interval from 1.2 to 14.3 mg.L-1 in the supporting electrolyte 0.1 M HNO3. First, DPV was used (Eacc = 400 mV vs. 1 M Ag/AgCl, tacc = 50 s). The reduction peak of iodides was located at the potential

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–300 mV and with increasing concentration (from 5 mg.L-1) it has shifted to more negative potential. The dependence of the peak-height on concentration of iodides in concentration range 1.26 – 3.11 mg.L-1 was linear (slope -99.8 nA.L.mg-1) with the correlation coefficient 0.9953. In the concentration interval from 5 to 15 mg.L-1 of iodide ions the dependence of the peak-height on the concentration of iodides had a sigmoidal shape of Frumkin isotherm (probably more than one layer was deposited at higher iodide concentrations) [65]. The same results were obtained using DCV, but the peaks were smoother in lower concentration intervals in comparison with DPV. In the iodide concentration range from 1 to 5 mg.L-1, the dependence of the peak-height on concentration was linear (slope -500.1 nA.L.mg-1) with the correlation coefficient 0.9951. In the concentration interval from 5 to 15 mg.L-1 of iodides, the dependence of the peak-height on concentration had the sigmoidal shape of Frumkin isotherm similarly as in the DPV measurement. The process of interaction of halide ions with the electrode surface was studied in the case of mercury electrodes in details. The halide ions react with mercury ions on the surface of the mercury electrode, producing various complexes. Therefore we can call the observed process chemisorption. This process is similar to the process occurring on the surface of silver (solid composite and metallic) electrodes [69, 74, 75], where the formations of the complexes of Cl- with Ag+ take place. The surface concentration of chlorides on Ag is about 1.2 – 1.5 nmol.cm-2 ([69], p. 180). Formed bond is too strong, and therefore this process cannot be explained as a physical sorption only, but as formation of a compound [65]. The halide ions play an interesting role in the case of deposition of metal ions (e.g. lead) on the electrode surface. If the metal ions are deposited on the surface of the metal or of composite electrode (without the presence of halides), their monolayer coverage (defined as the ratio of really occupied to all available places on the surface of the electrode at the moment, when the formation of the second layer starts) reaches the value θ about 0.2 (θ = relative surface coverage). However, it was proved that the presence of halide ions causes a compression of the monolayer and an overlapping of adsorbed islands [10], so that the surface coverage of metal reaches about θ ≈ 0.5. If this process is more complicated, it is possible to observe either more than one monolayer peak on the voltammetric record or a non-monotonous shape of the calibration curve (8). To the advantages of the CAgE belongs the fact that the effect of the presence of oxygen in the analyzed solution on the results of halide measurement is minimized. The mentioned oxygen effect is more pronounced when the solid composite silver electrode is used for the measurement. When chloride ions were determined by DPCSV, the first two records proved to be suitable for the determination of chloride ions (potential of accumulation +350 mV, time of accumulation 60 s, without elimination of air oxygen from the analyzed solution). During the third and following records, the height of the reduction peak of oxygen gradually increased and this peak was situated a bit more positively to the peak-potential of chlorides (+250 mV), i.e., the reduction peak of oxygen deformed and overlapped the reduction peak of halides and complicated its evaluation. The dependence of the peak-heights (first two records) of chloride ions on the concentration was fairly linear even without elimination of oxygen. The slope of this dependence (concentration of chlorides of the order of mg.L-1) was about 20 % higher in the presence of oxygen in comparison with the results obtained in the absence of oxygen. When the DPCSV determination of chloride ions was carried out in the presence

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and in the absence of oxygen, both results obtained were acceptable (on the level of significance 0.95) [65]. The possibilities of halide determinations with solid silver composite electrodes with different contents of silver were compared [65], namely CAgE20 and CAgE15, i.e. 15 % and 20 % of silver. In the case of CAgE15 the electrochemical pretreatment played a more important role than in case of CAgE20, because the reduction product remained adsorbed on the electrode surface, and thus could negatively affect the measurements. The testing measurements of this electrode were performed on the reduction of iodide ions, because iodides yielded the highest and best evaluable peak in comparison with other halides at the same concentration level. The peak-current values of iodides on CAgE15 were in comparison with CAgE20 lower and peaks had the shape of a double peak (two overlapping peaks). DPV and DCV were applied for the determination of iodides. The more positive peak was used for the measurement due to its better evaluability. The optimum potential of accumulation for iodides on CAgE15 was found to be 0 mV for both DPV and DCV. The dependences of DC and DP peak-heights of iodides on the time of accumulation (concentration of iodides 5 mg.L-1) were linear in the time interval from 0 to 60 s. At the time of accumulation 50 s there was a break on this dependence. For DPV or DCV determination of iodides at the iodide concentration level 5 mg.L-1 the accumulation times below 60 s should be therefore applied. The reduction of iodides was measured in concentration intervals from 5 to 11.9 mg.L-1 in the supporting electrolyte 0.1 M HNO3 using CAgE15, tacc=50 s, Eacc=0 s. First, the DPV was used. Peaks of iodides (two overlapped peaks were registered) were located at the potentials –150 mV and –250 mV. Both peaks were evaluated and dependences of the peak heights on the concentration of iodides were nonlinear in both cases. These dependences have the shapes of the Frumkin isotherm up to the concentration of iodides 9.7 mg.L-1. If DC voltammetry was applied (the measured concentrations of iodides were in the same concentration interval as in the case of DPV, i.e. 5.0 – 11.9 mg.L-1), linear dependences of the peak-height on the concentration of iodides were obtained both for the more positively situated and for the more negatively situated peak (correlation coefficients 0.9960 and 0.9964, respectively). The peaks of iodides were located at the potentials –300 mV and –400 mV. If slower DC scan rate was applied (5 mV.s-1) using CAgE15, only one reduction peak of iodides was obtained at the potential –280 mV. The iodides were measured in concentration intervals from 5.0 to 16.1 mg.L-1 in the supporting electrolyte 0.1 M HNO3. Using these parameters the height of the background current was decreased approximately ten times. In comparison with CAgE20, the CAgE15 electrode offers lower background current (it follows from the construction of the electrode), the separation of the overlapping reduction peaks of iodides can be achieved only with slower DCV, but the peak-heights are lower than in the case of CAgE20. The optimal potential of the accumulation on CAgE15 (0 mV) is shifted more negatively than in the case of CAgE20 (+400 mV) using DCV or DPV. The relative influence of the time of accumulation on the peak-height was the same for both electrodes, but the absolute peakheights of the reduction peaks of iodides (i.e., the sensitivity of the determination) on the CAgE15 was smaller. It is well known that the presence of surface-active substances (SAS) can negatively affect the voltammetric measurements, but, as it was reported earlier (e.g., [10, 42, 63, 65]), the silver composite electrodes enable voltammetric measurement even in the presence of SAS. To the supporting electrolyte 0.1 M HNO3 containing 5 mg.L-1 of chloride ions, Triton X-100 was added up to the concentration 5.10-2 %. When the concentration of Triton X-100

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was less than 10-2 %, the peak-height of chloride ions increased by about 10 %, and at higher SAS concentration the peak-height started to decrease (the decrease was almost 50 % in the case of concentration of Triton X-100 corresponding to 5.10-2 %). The peak of chlorides was smoother, narrower and shifted to more negative potentials after the addition of Triton X-100 [65]. The break in the dependence of the chloride ions peak height on the Triton X-100 concentration can be explained by reaching the critical micellar concentration (CMC) [76]. Analogous results were supposed to play a role also for other halide ions. The analysis of real samples was described in [65] as well: the determination of chlorides in a) Sample of drinking water from the well (Polabí, middle part of the Czech Republic), b) mineral water with declared content of chlorides maximally 4.5 mg.L-1. Both samples were analyzed directly, without any pretreatment or mineralization, the nitrogen gas was introduced into each sample for 5 minutes to remove the dissolved oxygen. The analyses were performed using silver composite electrode CAgE20, and the results were compared with the results obtained using HMDE. Both samples were evaluated using standard addition method. To 9 mL of the sample 1 mL of 1 M HNO3 was added and pH was adjusted to 1 (if it was necessary). By the silver composite electrode it was determined that the sample of drinking water contained 4.78 ± 0.58 mg.L-1 of chloride ions (on the level of significance 0.95). The results were confirmed by the analysis realized on the HMDE, where the concentration of chloride ions 4.10 ± 0.51 mg.L-1 (on the level of significance 0.95) was obtained. It is possible to conclude, that the results were fully comparable, the confidence intervals partly overlapped and the results corresponded to the requirements for drinking water, i.e. the obtained results are approximately 25 times lower than the value given by the Czech standard for drinking water (100 mg.L-1). The sample of mineral water was analyzed under the same conditions. Using the silver composite electrode the concentration of chlorides 2.82 ± 0.16 mg.L-1 (on the level of significance 0.95) was found, and using HMDE the concentration of chlorides 3.10 ± 0.42 mg.L-1 (on the level of significance 0.95) was found. Therefore we can conclude that the obtained results based on the measurement with silver composite electrode were in fair accordance with the results obtained with HMDE (the confidence intervals overlapped) and the results corresponded to the requirements for mineral waters (500 mg.L-1) [65]. It is well known that the selectivity of electrochemical determination of halides is very low. The only possibility to improve it, is the variation of the accumulation potential [65]. In the case of the mixture of bromides and chlorides measured in the supporting electrolyte 0.1M HNO3, the applied potential of accumulation (Eacc = Ein) was equal to the above mentioned potential of accumulation, which was used for the DCV determination of bromides, i.e. +170 mV. The concentration of bromides was 10 mg.L-1, the concentration of added chloride ions was from 1.5 to 5.8 mg.L-1. It was not possible to observe the peak of chlorides, because this potential of accumulation did not correspond to the potential of accumulation of chlorides, but the peak of bromides increased after addition of chlorides as well as in the case of addition of bromide ions and in consequence of it, it was not possible to separate the signals of bromide and chloride ions in their mixture. If the potential of accumulation (Eacc = Ein), which was used for DCV determination of chlorides (+300 mV), was applied and if only bromides were present in the solution, their peak could not be observed at all. The peak of bromides appeared after addition of chloride ions as well as the peak corresponding to the chemisorption of chlorides, but the peak-height of bromides did not

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practically change (as in the case of addition of bromide ions), therefore it was possible relatively easy to separate the two peaks in this case [65]. When the mixture of iodides and chlorides was measured in the supporting electrolyte 0.1 M HNO3, the potential of accumulation (Eac = Ein) (which was used in the case of DCV determination of iodides (+400 mV)) was applied. An excess of iodides (63.5 mg.L-1) was added to 2.65 mg.L-1 of chlorides and then the increase of the iodide peak-height was observed in dependence on addition of chloride ions from 4.38 to 7.74 mg.L-1. In this case it was again not possible to observe the peak corresponding to the reduction of chlorides only. With the first addition of chloride ions the peak of iodides decreased about 1000 nA (15 %) and with the next additions of chlorides the peak of iodides started to increase again. When the potential of accumulation (Eacc = Ein) used for DCV determination of chlorides (+300 mV) was applied, and when the 10-times higher excess of iodides was present, the additions of chlorides obscured the peak of iodides, which disappeared after additions of chlorides. The peak of chloride ions increased linearly [65].

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2.2.2. Determinations Using Gold Solid Composite Electrode The application of gold solid composite electrode, constructed as described above, was tested on the determination of arsenic only, but its applicability is surely much wider. 2.2.2.1. Determination of Arsenic A composite gold electrode CAuE20, disk diameter 2-mm, can be applied for differential pulse anodic stripping voltammetry of arsenic(III) [13, 77]. Because only arsenic(III) is electrochemically active, in the analysis of natural water samples it is necessary to prevent the oxidation of arsenic(III) (to arsenic(V)) immediately after sampling [78]. This can be done by the acidification of the sample and by the addition of hydrazine chloride as mentioned in [79]. Solution of 0.7 mol.L-1 KCl and of 0.1 mol.L-1 H2SO4 seems to be suitable as supporting electrolyte for arsenic(III) determination. It has been found that the reproducibility of the measurement was improved when the working electrode was kept for several seconds at the potential of +2500 mV (cleaning potential) before each measurement. It was possible to construct the optimal sequence of potential changes of the working electrode and the corresponding time intervals during the measurement [13]: cleaning potential Ecl +2500 mV with cleaning time tcl 5 s, accumulation potential Eacc -500 mV with accumulation time tacc 1 – 500 s, rest period tres 30 s at the accumulation potential Eacc and finally anodic potential scan from -400 mV to +500 mV. During the cleaning time and accumulation time the stirrer is preferably on. It has been found that the anodic current of arsenic(III) had practically constant value in the interval of the accumulation potential between –450 mV and –650 mV and slowly decreased at more positive and at more negative accumulation potential potential values. Linear dynamic range of this method was found in the range from 0.4 to 250 μg.L-1. The limit of detection of the procedure using the electrolysis time 200 s was 0.32 μg.L-1. The relative standard deviation of arsenic(III) determination was 12 % in the arsenic(III) concentration range from 5 to 100 μg.L-1 and below 8 % in the range from 100 to 500 μg.L-1 [13]. The results of DPASV determination are not influenced by the presence of surface active compounds (under above given conditions up to the Triton X-100 concentration 8 mg.L-1 no change of the peak current or peak potential values or peak shape was observed compared

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with the results obtained in the absence of Triton X-100) and by an excess of chloride ions. The only interference is caused by the presence of an excess of copper(II) ions. This interference can be eliminated either by the complexation with EDTA (the copper - EDTA chelate yielded only one dissolution peak at the potential 350 mV) or by the reduction with hydrazine and potassium bromide combined with the extraction with toluene and reextraction into the hydrochloric acid solution [80]. This method can be applied for the analysis of drinking water as well as of natural waters of various types. This determination can be realized in seawater as well. In [13] the results of the determination of As in distilled water, tap and river waters were published (recovery 98 to 105 %).

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2.2.3. Determinations Using Graphite Solid Composite Electrodes 2.2.3.1. Determination of Lead, Cadmium, Copper, Thallium and Bismuth, their Mutual Interferences, Effect of Presence of Other Metal Ions The UPD effect, which appears at metallic electrodes, was clearly observed on this type of electrode as well. In the case of a simultaneous deposition of two metals on the surface of the composite electrode, the anodic dissolution of the metal, which is anodically dissolved at more negative potential, is substantially affected by the presence of the other deposited metal. This effect was exploited for the determination of lead in the presence of other metals by differential pulse anodic stripping voltammetry [45]. If differential pulse anodic stripping voltammetry of lead was realized in the supporting electrolyte 0.1 M KCl + 0.01 M HCl using CCE30, the response was represented by one monolayer and one bulk peak of lead at the potential values -430 mV and -540 mV. The linear dependence of the peak current of monolayer as well as bulk peak on the lead ions concentration was obtained applying the accumulation time tacc = 180 s and the accumulation potential Eacc = -0.85 V. The linearity was verified in the range from 25 to 150 μg.L-1 with the slope 1.05 nA.L.μg-1 in the case of the monolayer peak and from 50 to 350 μg.L-1 with the slope 1.91 nA.L.μg-1 in the case of the bulk peak. Due to the low difference between the potential values of the anodic bulk and monolayer peaks of lead, the measurement of the mentioned peak heights is somewhat complicated. Therefore, the influence of Cu(II) ions on the nature of anodic voltammetric curves of lead(II), which is well known at silver solid composite electrode [58], must be taken into account. The results of cyclic voltammetric measurement of Pb(II) ions in the presence of identical concentration of Cu(II) ions have shown a decrease of the anodic bulk peak and an increase of anodic monolayer peak of lead. This result is more pronounced in DPASV measurement, where the deposition of copper on the electrode surface is more pronounced and thus it can effectively affect the deposition and dissolution processes of Pb(II) ions. More pronounced is the influence of Cu(II) ions on the DPASV response of Pb(II) ions present at a very low concentration level. It is evident that monolayer peak of lead increases with increasing concentration of Pb(II) ions in the range from 10 to 150 μg.L-1 linearly (slope 9.51 nA.L.μg-1) and the bulk peak under these conditions disappears [45]. During the increase of the Pb(II) ions concentration, the peak height of the anodic monolayer peak of copper is not constant. It is evident that the described effect of Pb(II) ions is limited in the concentration range up to 50 μg.L-1; further increase of the Pb(II) ions concentration did not influence the monolayer peak of copper and its height stayed practically constant [45].

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In 0.1 M KCl medium, the influence of Cu(II) ions on the DPASV response of Pb(II) ions is identical as that obtained in more acidic medium, but the measured signal is lower. Under identical conditions, the slope of the linear dependence of the height of the lead monolayer dissolution peak on the Pb(II) ions concentration (from 5 to 100 μg.L-1) has the value of 4.26 nA.L.μg-1 (9.55 nA.L.μg-1 for the KCl/HCl medium). With linear increase of the monolayer peak of lead in dependence on the Pb(II) ions concentration, the anodic monolayer peak simultaneously decreases as well as the bulk peak of copper. This decrease depends linearly on the Pb(II) concentration using longer electrolysis time (slope of the dependence of the monolayer peak height has the value -13.25 nA.L.μg-1 and of the bulk the value -3.95 nA.L.μg-1, electrolysis time 180 s) [45]. If the ASV determination of lead is realized in the presence of a metal deposited electrochemically together with lead at a negative potential value, but if anodic dissolution occurs at a more negative potential than anodic dissolution of lead (e.g., Cd, Tl), no effect of such metal on the ASV response of lead is observed [45]. By ASV measurement of very high concentrations of Cd(II) ions (3 mg.L-1) and Pb(II) ions (50 - 175 μg.L-1) in 0.1 M KCl solution, only the anodic peak of cadmium at -750 mV appears together with the bulk peak of lead (-530 mV) and monolayer peak of lead (-400 mV), i.e. the anodic dissolution of lead starts at the potential, at which the deposited cadmium is already anodically dissolved. The measurement of the ASV response of lead in the presence of cadmium is more advantageous at higher concentration of hydrogen ions, e.g., 0.1 M KCl + 0.1 M HCl, where the bulk peak of lead is better separated from the peak of cadmium. The effect of the presence of high concentration of Cu(II) ions on the DPASV response of lead(II) in the presence of cadmium ions is identical as described above (increase of the monolayer peak of lead). In the electrolyte used (0.1 M KCl + 0.01 M HCl), the DPASV response of Tl(I) ions under the described conditions presents an anodic monolayer peak at the potential value -610 mV and at a higher concentration also a bulk peak at the potential -860 mV. In the concentration range of Tl(I) ions from 50 to 1000 µg.L-1, the monolayer peak height increases linearly with increasing Tl(I) ions concentration with the slope 1.13 nA.L.µg-1 and the bulk peak increases in the range from 500 to 1000 µg.L-1 with the slope 0.35 nA.L.µg-1 (60 s time of electrolysis). At the DPASV determination of lead in the presence of thallium, the anodic dissolution process of deposited lead is not influenced by the dissolution process of thallium, which occurs at more negative potential. However, due to a small difference between the bulk peak potential of lead and monolayer peak potential of thallium, the measurement of the peak heights is not exact enough. The influence of Cu(II) ions on the DPASV response of thallium is not very significant - a twenty times increase of Cu(II) ions concentration leads to an increase of the thallium monolayer peak height by approx. 7 % of the original height. Because the bulk peak potential of copper is very close to the value -90 mV, also the bulk peak height of thallium increases with increasing Cu(II) ions concentration. It can be concluded that it is possible to record the ASV monolayer peak of lead (at the potential value -410 mV) even in the presence of thallium and copper using this electrode type. The obtained results confirm the linear dependence of the monolayer peak height on Pb(II) ions concentration in the range from 50 to 500 µg.L-1 in the presence of Cu(II) ions (4 mg.L-1) and Tl(I) ions (0.4 mg.L-1) with the slope value 1.94 nA.L.µg-1 (accumulation time 60 s). It should be mentioned that the slope value depends on Cu(II) ions concentration: with decreasing Cu(II) concentration decreases also the slope value - at the Cu(II) ions concentration 2.5 mg.L-1 the slope value decreased to 1.13 nA.L.µg-1 under identical experimental conditions [45].

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DPASV of Bi(III) ions yields a not very well developed anodic monolayer peak at the potential -111 mV and a bulk peak at the potential -850 mV in the KCl/HCl solution. For the voltammetric determination of bismuth the application of an acid supporting electrolyte is evidently more convenient. On the other hand, similarly as described above, the presence of Cu(II) ions also causes well acceptable conditions for the DPASV determination of lead in the medium containing Bi(III) ions. For the DPASV determination of lead in the presence of Bi(III) ions, the described method, which is based on the increase of the anodic monolayer peak of lead by simultaneous deposition of copper(II) ions, can be applied [45]. The affection of DPASV determination of Pb(II) ions by the presence of the surface active substances was studied using Triton X-100 as a model substance [45]. Low concentration of Triton X-100 causes only a decrease of the lead monolayer peak height without any deformation. The presence of 1 mg.L-1 of Triton X-100 causes the decrease of the peak height about 40 % (the dependence: peak height vs. Triton X-100 concentration was practically up to this concentration linear), but the further increase of Triton X-100 concentration does not influence this parameter substantially (10 % decrease in the range from 1 to 4 mg.L-1. On the other hand, the monolayer peak current of lead is otherwise decreased by the presence of Triton X-100 (generally any SAS), which concentration is in analyzed sample constant, but it increases linearly with increasing Pb(II) ions concentration [45]. The determinations of lead(II) ions in model as well as real samples was described in [45]: To 10 mL of analyzed sample, a supporting electrolyte (0.1 M KCl + 0.008 M HCl) was added. At first, two model samples, containing 50 and 25 μg.L-1 of Pb(II) ions, were analyzed under the above described conditions (accumulation time 180 s). Solution of cupric ions (final concentration 2 mg.L-1) was added to the sample. The standard addition method was used to quantify the lead concentration. The following concentrations of Pb(II) ions were found: 49.70±0.41 μg.L-1 and 25.51± 0.72 μg.L-1 (confidence intervals were calculated with the probability 95 %). Similarly, to 10 mL of freshly sampled tap water, the same supporting electrolyte (0.1 M KCl + 0.008 M HCl) was added and thus prepared solution was analyzed by DPASV under the above described conditions (accumulation time 180 s). To eliminate the possible matrix effect, standard addition method was used to quantify the Pb(II) ions concentration. The obtained result was 6.45±1.01 μg of Pb(II) ions in 1 L. The sample was also analyzed with the addition of cupric ions (2 mg.L-1) and the result of 7.14±0.84 μg.L-1 was achieved. The result 7.25±0.42 μg.L-1 was obtained by the measurement of the same sample using HMDE after removing the surface active substances by microwave digestion. Therefore we can conclude that the confidence intervals of all results overlapped with the probability of 95 % and that the surface active substances did not affect the determination process significantly, which is in agreement with conclusions of the previous paragraphs. In the case of a sample in absence of copper, the following parameters were calculated: limit of decision 2.21 μg.L-1, limit of detection 3.09 μg.L-1; limit of determination 5.64 μg.L-1 and in the presence of copper: limit of decision 1.90 μg.L-1, limit of detection 3.46 μg.L-1; limit of determination 5.12 μg.L-1 (i.e., both determined concentrations of Pb(II) ions in tap water were above the limits of determination). It is possible to conclude that the obtainable limits of detection on graphite composite electrodes are higher in comparison with mercury ones; nevertheless, they are sufficient according to the requirements for analysis of environmental samples (e.g., according to the valid standards for drinking water).

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2.2.3.2. Determination of Manganese Determination of manganese ions can illustrate the applicability of the graphite composite electrode in positive potential area. It follows from the Table I that the most interesting part of voltammetric applications of CCE lays in area of positive potentials, in which occures Mn2+ oxidation. The reaction mechanism of the determination of manganese was described e.g. in paper [81]. The described determination of manganese is based on oxidation of Mn2+ ions to an insoluble hydrated dioxide, which is accumulated on the electrode and offers a well-defined peak during its electrochemical reduction [14]. The determination can be realized using cathodic stripping voltammetry with graphite composite electrode containing 30 % of graphite (denoted as CCE30), applying potential of accumulation +700 mV. At this potential of accumulation is Mn2+ oxidized to insoluble Mn4+, which is accumulated on the electrode surface. If the potential of accumulation is higher than +700 mV, Mn2+ is oxidized up to Mn6+, which is soluble and cannot be accumulated on the electrode surface [81]. If the potential of accumulation is lower than +700 mV, Mn2+ is not oxidized to Mn4+ and no reaction occurs [14]. Concentration dependence of Mn2+ in waters from 0.3 to 3.3 mg.L-1 was measured in [14]. Peak at the potential +280 mV corresponds to the reduction of Mn4+ to Mn2+. The second peak, corresponding to the reduction of Mn2+ to Mn0 can arise in the area of negative potentials (approximately at the potential –1500 mV on different electrodes). In case of CCE30 this peak was totally overlapped by supporting electrolyte decomposition. Measuring procedure is composed from the following steps: Mn2+ → Mn4+ → accumulation of Mn4+ → Mn2+ → Mn0. Except the last step, it is possible to observe these processes also using CCE30. Statistical results were calculated in [14] from 15 repeated measurements and from the calibration curve [70, 82], which was linear in whole tested concentration range (0.3 to 3.3 mg.L-1): relative standard deviation 16.1 %, skewness 0.31, excess 2.01; limit of decision 0.12 mg.L-1; limit of detection 0.15 mg.L-1; limit of determination 0.52 mg.L-1. 2.2.4. Determinations Using Solid Amalgam Composite Electrodes The field of applications of solid amalgam composite electrodes [41] is similar to that of solid amalgam electrodes [83-88]. The first experiments in determination of inorganic compounds were realized with electrode material based on silver amalgam (AgSA-CE) [41]. As it was mentioned above, the relatively typical phenomenon at solid surfaces is UPD effect [10, 42, 58, 69]. It is well known that this effect was not observed at mercury (“metallic” as well as “amalgam”) electrodes. Therefore, it seemed interesting to verify the existence of this phenomenon on SA-CE, namely it was proved on AgSA-CE [41]. Lead, as very often voltammetrically studied inorganic analyte, was used for these purposes. All below given potentials were referred to Ag/AgCl/3M KCl electrode. When cyclic voltammograms of Pb(II) ions solution were recorded with the p-AgSA-CE in 0.1 M KCl, only one anodic peak appeared (at -407 mV). After short dipping of this electrode in liquid mercury (1 minute long amalgamation), only one peak at -361 mV, corresponding to the lead layer oxidation (dissolution) was recorded in low concentrations of Pb(II) in solution. With increasing concentration of Pb(II), the symmetry of this peak was disturbed and two badly separated overlapping peaks were observed (-428 mV and -362 mV). With subsequent standard additions of Pb(II), the more positive peak disappeared and it was overlapped by the more negative one. If the surface was covered either by mercury film (formed by 30 minute long electrolysis) or by mercury “meniscus” (formed by 10 minute long amalgamation), such

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splitting was not observed (peak was observed at about -360 mV). Therefore, it is possible to conclude that this phenomenon cannot be ascribed to the UPD effect but to simple deposition of lead layers at various surfaces (covered or uncovered amalgam particles by metallic mercury). Verification, whether the surface of the AgSE-CA [41, 89] corresponds to the array of microelectrodes was realized using determination of Tl(I) ions. It is well known from the theory of microelectrodes that the voltammetric signal is practically independent of solution stirring [90]. Therefore, the anodic stripping voltammograms of 0.039 mg.L-1 of Tl(I) after accumulation time tacc = 180 s, during which the solution was stirred, were recorded using m-AgSAE and AgSA-CE. In the case of m-AgSAE, the current was 5.3 times higher if the solution was stirred during the accumulation in comparison with unstirred solution. In the case of AgSA-CE this ratio was 2.2 only. This fact confirms that some parts of conducting segments of composite electrode exhibit the character of microelectrodes, whereas the other part exhibit the “classical” character of macro electrodes, which responses depend on stirring (heterogeneity of the surface was confirmed by microscopic electrode investigations as seen on Figure 3).

2.2.4.1. Determination of Cadmium, Lead, Copper, Thallium and Iodates Voltammetric determinations of the (probably) most common and most frequently determined inorganic elements using AgSA-CE were described e.g. in [41, 89]. The parameters of their determinations and the achieved statistical results at p-AgSA-CE (without accumulation) are listed in Table VI. It is possible to utilize high hydrogen overpotential of AgSA-CE for the reduction of iodates in alkaline solution (0.1 M NaOH). Their peak potential is situated at about -1250 V. The concentration dependence is linear in wide range (e.g. in narrower interval from 1 mg.L-1 to 10 mg.L-1, the correlation coefficient amounted to 0.9965). This composite electrode proved to be suitable for the determination of heavy metal cations in aqueous solutions by their direct reduction or by anodic stripping of the reduced metal. Combination of the accumulation of an analyte on the surface of the working electrode with consecutive voltammetric scan usually leads to the increase of sensitivity and the decrease of limits of detection, sometimes by several concentration orders. Thus, anodic stripping voltammetry (ASV) or cathodic stripping voltammetry (CSV) are suitable techniques for the determinations of very low concentrations of analyzed compounds. ASV is mostly applied for determination of heavy metals cations and CSV for anions and organic compounds. MeSAEs in traditional arrangement proved to be suitable alternative to mercury electrodes in ASV as well as in CSV [68]. Therefore, the described AgSA-CEs were tested for methods comprising the accumulation step [41]. Set of curves, corresponding to the calibration dependence, was recorded under optimized conditions using ASV at m-AgSA-CE for the of Tl(I). Electrochemical regeneration of the electrode surface was applied before each measurement (EReg1 = -1500 mV, tReg1 = 0.3 s, EReg2 = -200 mV, tReg2 = 0.2 s, N=50). The shape of DPAS voltammograms and wide linear dynamic range of calibration graph (its part from 0.005 to 0.08 mg.L-1 was statistically evaluated in [41]: slope (2244±35) nA.L.mg-1, r = 0.9998)) indicate that the composite electrode constructed on the basis of solid amalgam is suitable for ASV determination of heavy metals ions. Similar results were obtained for Cd, Pb, and some other metallic cations in aqueous solutions [89].

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Table VI. Voltammetric Determination of Cd(II), Pb(II), Cu(II) and KIO3 at AgSA-CE (all DPCV); N = 11, v = 20 mV.s-1, tacc = 0 s Parameter Cd Pb Cu KIO3 Analyte concentration [mg.L-1] 0.1 0.1 0.99 3.85 Supporting electrolyte acetate buffer 0.1 M, pH 4.8 NaOH 0.1 M Average peak height [nA] 14.25 2.64 51.66 203.01 Confidence interval (α=0.05) 0.20 0.08 0.99 0.78 [nA] Standard deviation [nA] 0.30 0.13 1.50 1.18 Relative standard deviation [%] 2.12 4.73 2.90 0.58 0.006 0.014 0.10 Limit of detection (3xstandard deviation) [mg.L-1] 0.017 0.12 0.15 Limit of detection (Direct method 0.015 -1 of signal, IUPAC) [mg.L ] 0.024 0.026 0.24 0.23 Limit of determination (Direct method of signal, IUPAC) [mg.L-1]

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2.2.5. Determinations Using Silica Gel Modified Carbon Composite Electrodes Silica gel modified carbon composite electrode proved to be suitable sensor for determination of silver, copper, cadmium, lead ions as well as ferrocyanides in waters [67], i.e. in direct as well as in anodic stripping voltammetry. Moreover, this electrode can be successfully applied in organic analysis as well (see corresponding paragraph of this chapter). 2.2.5.1. Determination of Silver Silica gel modified carbon composite electrode proved to be a suitable sensors for the determination of Ag(I) ions in water. The relative standard deviation amounted to 1.1 % for a concentration of Ag(I) ions of 8.10-5 mol.L-1, using an electrode modified with 10 % silica gel with a specific surface area of 70 m2.g-1. The dependence of the limiting current on the Ag(I) ion concentration from 10 to 500 μmol.L-1 had a slope of 5.50 μA.L.mmol-1 and a correlation coefficient of 0.9998 [67]. These modified electrodes have also been tested in anodic stripping voltammetry (ASV) [67]. It follows from the results obtained in ASV of silver (an electrolyte of 0.05 M NH3, 0.05 M NH4NO3, 0.001 M NH4CNS, an electrolyte with 10 % of the 70 m2.g-1 silica gel) that stirring has little effect on the amount of the silver deposited and thus on the height of the appropriate anodic peak in case of shorter accumulation times. At Ag(I) concentration of 2.5·10-8 mol.L-1, the peak obtained with stirring during deposition was twice as high as that obtained without stirring. When the electrolysis time was longer than 120 seconds, the effect of stirring was more pronounced and thus the attainable levels of determination could be lower [67]. 2.2.5.2. Determination of Cadmium, Lead and Copper Carbon composite electrodes modified with 10 % silica gel were also used as supports for the preparation of mercury film electrodes (MFE) applicable for determination of amalgam forming metals in waters [67]. This modified MFE was prepared by depositing mercury in situ and was used for an ASV determination of Cu, Pb and Cd in the usual manner in 0.1 M HNO3. Similar to voltammetric measurements with the modified carbon composite electrode,

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measurement with the modified MFE exhibited during electrolysis a smaller stirring effect on the stripping peak currents. The background current also decreased in proportion to the specific surface area of the silica gel used (from 86 nA in the case of electrode without silica gel to 0.25 nA in the case of 640 m2.g-1 silicagel (ASV of lead, electrolyte 0.1 M nitric acid, the modified MFE was prepared in situ, time of accumulation 360 seconds, potential of accumulation -900 mV; background current measured at -500 mV)). The use of silica gel modified MFE did not dramatically lowered the detection limits of ASV determination of amalgam forming metals. Application of these electrodes was, however, advantageous for ASV determination of minor concentrations of copper, lead, and zinc in acetate buffer medium without the removal of dissolved oxygen from the analytical solutions [91].

2.3. Voltammetric Determination of Organic Compounds Using Composite Solid Electrodes

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2.3.1. Determinations Using Silver Solid Composite Electrode 2.3.1.1. Determination of 2-Nitronaphtalene Electrodes with varied content of silver CAgE10, CAgE15 and CAgE20 (i.e., 10, 15, 20 m/m %) were used for the determination of 2-nitronaphtalene by differential pulse (DPV) and direct current voltammetry (DCV) [40]. The application of this electrode in differential pulse cathodic stripping voltammetry enables the direct determination of nitro compounds without elimination of the presence of dissolved oxygen. The easy reducibility of nitro-group enables the application of electrochemical methods for the determination of trace amounts of nitrated polycyclic aromatic hydrocarbons (NPAH), to which belongs 2-nitronapthalene as well. It is highly dangerous and mutagenic compound associated with the increased occurrence of cancer. Mechanism of their formation and mechanism of polarographic reduction on various electrodes was outlined in [40]. As an optimal mode of electrochemical cleaning and pretreatment proved to be the insertion of 50 and more cleaning cycles (from +350 mV to –900 mV with time delays 0.1 s at the initial and final potential, scan rate 500 mV.s-1) before the start of measurement of each new sample (which could consist of several scans). One cleaning cycle between initial and final potential using scan rate 500 mV.s-1 should be inserted before each registered scan. The pH dependences showed that the optimal pH value is about pH 8 in the case of CAgE10 and CAgE15 electrode and pH 9 in the case of CAgE20, because at these pH values the peaks exhibited the maximal heights and good symmetries. The results of experiments realized using Britton-Robinson buffer were similar to those realized using borax buffer (in the used pH range 8-9). The peak-heights as well as peak-potentials are independent on the time of accumulation for all types of electrodes, thus all measurements can be carried out with zero time of accumulation. Determination of 2-nitronaphtalene in two concentration ranges: 10 – 100 μmol.L-1 and 1 – 10 μmol.L-1 was described in [40]. The most symmetric and the best-evaluable peaks were obtained in higher concentration intervals on all electrodes. The peak-heights exhibited linear dependence on analyte concentration in the range from 1 to 10 μmol.L-1. Nevertheless, the records were not smooth and in the lowest concentrations it was difficult to evaluate them.

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The values of background currents were different for different types of electrodes; typical average values were - 4 μA for CAgE10, - 8 μA for CAgE15 in case of DPV and - 13 μA in case of DCV and - 5 μA for CAgE20, i.e., the background current increased almost linearly with silver content in the electrode. DP peak of 2-nitronaphtalene was located at the potential –650 mV at CAgE 10. For the peak-height dependence on concentration in range 1 – 100 μmol.L-1 the linear dependence was found (slope -15.6 nA.L.μmol-1, r = 0.9963). DPV peaks on the CAgE15 increased in concentration range from 1-100 μmol.L-1 linearly (slope -35.6 nA.L.μmol-1, r = 0.9966) and DCV waves on the CAgE15 in concentration range 10 - 100 μmol.L-1 (slope -63.8 nA.L.μmol-1, r = 0.9930). DCV wave of 2-nitronaphtalene was located about 150 mV more negatively (- 800 mV) than in the case of DPV peak. DPV peak of 2-nitronaphtalene obtained using CAgE20 was located at the potential –650 mV. For the peak-height dependence in the concentration range 1 – 100 μmol.L-1 the linear dependence was found (slope -11.2 nA.L.μmol-1, r = 0.9920).

The repeatability of determinations realized using all three tested electrodes was tested by fifteen times repeated measurements at the constant concentration 6 μmol.L-1 of 2-nitronaphtalene (Table VII) [40]. Generally, the first peak of each measurement (which consisted of 3 repeated scans and in the case of repeatability test of 15 scans) was higher than the other ones and therefore it should be omitted. This fact was confirmed using the statistical tests [82] (Dean-Dix or Grubs tests at the confidence level 0.95): the first record on CAgE10 had to be omitted in order to enable the use of arithmetic mean. It is advisable to apply the median instead of arithmetic means (calculated from the second and subsequent measured records in case of CAgE10) for the evaluation.

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Table VII. Statistical evaluation of results obtained by fifteen times repeated measurement of 6.10-6 mol.L-1 of 2-nitronaphtalene with CAgE10, CAgE15, CAgE20 CAgE10

CAgE15

CAgE20

Arithmetic mean [nA]

-80.9

-262.5

-73.6

Median [nA]

-77.6

-262.5

-75.0

Confidence interval [nA]

3.7

7.6

3.4

Standard deviation [nA]

6.0

13.2

6.1

Relative standard deviation [%]

7.8

5.0

8.3

Skewness

-0.5

-0.1

-0.3

2.6

2.9

3.4

Limit of decision [μmol.L ]

0.51

0.42

0.64

Limit of detection [nA]

18.1

39.5

18.4

1.34

0.90

1.50

2.61

1.89

3.10

Excess -1

-1

Limit of detection [μmol.L ] -1

Limit of determination [μmol ]

In the case of this electrode, the statistical distribution was very close to the normal distribution (according to skewness and excess values). The limit of detection of

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2-nitronaphtalene, achieved in [40] using DPV on CAgE15, is around 0.90 μmol.L-1, which is higher than for differential pulse voltammetry on a HMDE (LODHMDE 0.01 μmol.L-1[92, 93]). The developed analytical method was tested by analysis of the model sample (2 μmol.L-1 of 2-nitronaphtalene in drinking water) using CAgE15 and HMDE [40]. DPCV method was used in supporting electrolyte 0.05 M borax buffer (pH 8), scan rate 20 mV.s–1. Under these conditions the peak of 2-nitronaphtalene was situated at the potential -600 mV on HMDE and at the potential -650 mV on silver composite electrode. The results were evaluated using standard addition method (4 standard additions) in both cases. Using this electrode, the sample concentration 1.95 ± 0.26 μmol.L-1 was determined and using HMDE the sample concentration 2.01 ± 0.14 μmol.L-1 was determined. Therefore we can conclude that the obtained results on CAgE15 were in fair agreement with those obtained by DPCV determination on HMDE (the confidence intervals overlapped on 95 % confidence level) [40]. It was concluded in [40] that the optimal composition of the electrode for voltammetric nitro compound determination is CAgE15: 15 % of silver, 25 % of graphite and 60 % of methacrylate resin, inner diameter 1 mm, because of very good reproducibility of records, heights of peaks, symmetry, slope and limit of detection, etc. Comparison with chemometric parameters achieved using other types of electrodes (CCE30 and silver metallic electrode AgE (i.e., non-composite)) is listed in Table VIII. The selectivity of determination of 2-nitronaphtalene in mixture with other NPAH is not too good, because all NPAH exhibited their reduction peaks on composite electrode in the similar potential region (about -700 mV vs. Ag/AgCl.), but this disadvantage could be compensated by chromatographic preliminary separation of the analyte.

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Table VIII. Statistical evaluation of voltammetric determination of 2-nitronaphtalene on AgE, CAgE15 and CCE30. The limits were calculated from the calibration curves in concentration interval from 1 to 10 μmol.L-1 using all types of working electrodes Electrode type Limit of decision [μmol.L-1] Limit of detection [nA] Limit of detection [μmol.L-1] Limit of determination [μmol.L-1] Relative standard deviation [%]

AgE 0.21 5.7 0.36 0.85 3.48

CAgE15 0.42 39.5 0.90 1.89 4.99

CCE30 0.89 0.3 0.90 1.60 7.64

2.3.1.2. Determination of 1-Nitronaphtalene Similar experiments as described in the above mentioned paragraph on determination of 2-nitronaphtalene were performed with of 1-nitronaphtalene on CAgE20 [15]. As an optimum electrochemical pretreatment of the electrode polarization by cyclic voltammetry was suggested: 10-20 cycles from -100 to -1500 mV, scan rate 100 mV.s-1 in the case of silver composite electrode in the chosen supporting electrolyte. This type of pretreatment is to be realized each day before the measurement was started. It is advisable to insert a few (e.g. 10 or more) cleaning cycles (between +200 and -1500 mV) before the start of the measurement of each new sample and the electrode was polarized for 4 s by the more positive cleaning potential at the start of each scan. Alkaline pH 12 was used for the determinations in [15]. The determinations were realized in concentration intervals 20-100 μmol.L-1 and 2-10

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μmol.L-1 (Ein = -400 mV, Efin=-900 mV, v=20 mV.s-1). The achieved limits of detection amounted to 3.10-5 μmol.L-1 for DC (slope 0.20 nA.L.μmol-1) and 3.10-6 μmol.L-1 for DP (slope 0.63 nA.L.μmol-1) voltammetry.

2.3.1.3. Determination of 6-Nitroquinoline As an optimum electrochemical pretreatment of the electrode for determination of 6-nitroquinoline polarization by cyclic voltammetry was suggested: 100-200 cycles from +350 to -900 mV, scan rate 500 mV.s-1 in the case of silver solid composite electrode in the chosen supporting electrolyte [94]. This type of pretreatment is to be realized each day before the measurement was started. It is advisable to insert a few (e.g., 50 or more) cleaning cycles before the start of the measurement of each new sample. DP voltammetry was used in the case of determination of 6-nitroquinoline using CAgE20 (Ein = +100 mV and Efin = -1000 mV) [94]. At pH > 7 two peaks were observed, whereas below this value only one peak was observed. Britton-Robinson buffer at pH 10 was chosen as the optimal supporting electrolyte (due to sensitivity, linearity of the calibration line and symmetry of both peaks). Concentration interval of 6-nitroquinoline from 2 to 100 μmol.L-1 was investigated and two peaks situated at potentials –300 and –680 mV were obtained. The dependences of the peak-heights on the concentration were linear in this range (for the more positive peak: slope -4.87 nA.L.μmol-1, correlation coefficient 0.9977 and for the negative peak slope -6.18 nA.L.μmol-1, correlation coefficient 0.9985). The achieved chemometric parameters [82] are listed in Table IX. On the basis of the Grubbs test [70] (for identification of outlying points) the first point from 15 recorded was omitted (on the confidence level 0.95) in the case of silver composite electrode CAgE20.

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Table IX. Statistical evaluation of voltammetric determination of 6-nitroquinoline on AgE, CAgE20 and CCE30. The limits were calculated from the calibration curves in concentration interval from 1 to 10 μmol.L-1 using all types of working electrodes Electrode type Limit of decision [μmol.L-1] Limit of detection [nA] Limit of detection [μmol.L-1] Limit of determination [μmol.L-1] Relative standard deviation [%]

AgE 0.11 19.9 0.22 1.04 10.48

CAgE20 0.33 8.5 0.49 0.81 4.10

CCE30 0.17 0.6 0.40 0.69 7.29

2.3.1.4. Determination of 5-Nitrobenzimidazole As an optimum electrochemical pretreatment of the electrode for 5-nitrobenzimidazole polarization by cyclic voltammetry was suggested: 100-200 cycles from +350 to -900 mV, scan rate 500 mV.s-1 in the case of silver composite electrode in the chosen supporting electrolyte [94]. This type of pretreatment is to be realized each day before the measurement was started. It is advisable to insert a few (e.g., 50 or more) cleaning cycles before the start of the measurement of each new sample. Only DP voltammetry can be used for the determination of 5-nitrobenzimidazole on CAgE20 (Ein = 0 mV and Efin = -1000 mV) (DC waves are worse reproducible and more difficult to evaluate). pH 9 proved to be optimal for its determination using this electrode [94]. Below pH 7 only one peak was registered at the potential about -750 mV. If pH value

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was higher than 7, at -300 mV one more positively situated peak was recorded (Its increase was non-linear in the investigated range) [94]. Concentration dependence of 5-nitrobenzimidazole was investigated in concentration interval 1 – 100 μmol.L-1 in [94]. This dependence was linear: slope -5.70 nA.L.μmol-1, correlation coefficient 0.9993. The achieved parameters (calculated according to [82]) are listed in Table X. Table X. Statistical evaluation of voltammetric determination of 5-nitrobenzimidazole on AgE, CAgE15 and CCE30. The limits were calculated from the calibration curves in concentration interval from 1 to 10 μmol.L-1 using all types of working electrodes

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Electrode type Limit of decision [μmol.L-1] Limit of detection [nA] Limit of detection [μmol.L-1] Limit of determination [μmol.L-1] Relative standard deviation [%]

AgE 0.63 13.8 1.07 5.19 10.41

CAgE15 0.31 2.3 0.69 2.02 5.21

CCE30 0.41 0.5 0.98 1.43 8.29

2.3.1.5. Determination of Azo Compounds Polarographic and voltammetric studies of aromatic azo compounds are frequent with respect to the importance of this group in the industry of dyes and because they can pollute surface waters in industrial areas. The mechanism of polarographic reduction was described in e.g. paper [95]. It is affected by different substituents on aromatic rings of azo compounds and it also depends on reaction conditions [96]. Reduction of azo compounds includes 4-electron reduction of azo group to hydrazo group and subsequent 2-electron reduction of hydrazo group to amino group [46]. Alizarine black PT can be mentioned as an example of azo compound determination. For the analysis of this compound CAgE20 and CCE30 (see below) in comparison with HMDE can be used [46]. After mechanical pretreatment of the CAgE20 by alumina, the electrochemical activation in chosen supporting electrolyte was to be performed by application of 100 - 200 DC cycles from +350 mV to -950 mV at scan rate 500 mV.s-1. Electrochemical renovation of the electrode before each measurement consisted of insertion of 50 cleaning cycles minimally (+350 mV for 0.1 s and –950 mV for 0.1 s). This compound was determined using cathodic voltammetry from -200 mV to -900 mV, DPV scan rate 20 mV.s-1. Britton-Robinson buffer, pH 7 proved to be the optimum supporting electrolyte in the case of CAgE20 [46]. Linear dynamic range was found from 10 to 100 μmol.L-1, in higher concentration ranges the peak was splitted. In lower range (2 – 16 μmol.L-1) the concentration dependence is non-linear and more complicated [35, 46]. The same compound was determined using HMDE [46]. Britton-Robinson buffer pH 10 was used as supporting electrolyte. Linear dynamic range was found from 0.1 to 100 μmol.L-1. It was confirmed that diffusion is the controlling process in the range from 0.1 to 70 μmol.L-1 [46]. Statistical evaluation, calculated from 15 repeated determinations of alizarine chrome black PT in the case of CAgE20 and CCE30 and from 11 repeated measurements in the case of HMDE are summarized in Table XI.

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T. Navratil, B. Yosypchuk and J. Barek Table XI. Statistical Evaluation of Voltammetric Determination of alizarin chrome black PT on CAgE20, CCE30 and HMDE Electrode

CAgE20

CCE30

HMDE

Relative standard deviation [%]

17.99

3.95

6.90

Skewness Excess

-0.01 1.69

-1.30 3.57

-0.32 2.96

0.55

0.23

0.21

1.77

0.41

1.05

8.76

8.92

4.26

Limit of decision [μmol.L-1] -1

Limit of detection [μmol.L ] -1

Limit of determination [μmol.L ]

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2.3.2. Determinations Using Graphite Solid Composite Electrode Graphite solid composite electrode found its field of application not only in determination of inorganic elements and compounds, but many organic compounds in drinking waters as well as in other aqueous solutions can be determined using this type of solid composite electrode. 2.3.2.1. Determination of 2-Nitronaphtalene As an optimum electrochemical pretreatment of the electrode polarization by cyclic voltammetry was suggested: 100-200 cycles or from +1200 to –950 mV, scan rate 500 mV.s-1 in the case of graphite composite electrode in chosen supporting electrolyte [94]. This type of pretreatment is to be realized each day, before the measurement was started. It is advisable to insert a few (e.g., 50 or more) cleaning cycles before the start of the measurement of each new sample. The optimal composition of CCE for 2-nitronaphtalene determination contained 30 % of graphite powder dispersed in the epoxy resin (i.e. CCE30). DP voltammetry proved to be most suitable for the determination of 2-nitronaphtalene on this electrode [94]. DCV can be applied in the region of higher concentrations only (i.e. 10 – 100 μmol.L-1). The measurements were realized in the potential range from Ein = 0 mV to Efin = -1200 mV. Only one peak was observed below pH 7, the other peak appeared at pH > 7. pH 10 was chosen as an optimal pH value (symmetry, sensitivity of the determination, etc.). The dependence of the peak-height on the concentration of 2-nitronaphtalene was tested in concentration intervals 2 – 100 μmol.L-1 (DPV) and 10 – 100 μmol.L-1 (DCV) [94]. The more negative peak at the potential -720 mV was used for the evaluation, because it was better evaluable than peak situated at the potential about -260 mV. DC waves of 2-nitronaphtalene were situated at the potential -370 mV and -840 mV, respectively. Dependence of the height of the DPV peak at -720 mV on concentration had the shape of Frumkin isotherm. Its segment from 2 to 30 μmol.L-1 could be approximated by a straight line (slope -0.48 nA.L.μmol-1, correlation coefficient 0.9990). Similar dependence of the more positively situated peak was non linear in the whole investigated range [94]. DC voltammetry was used for investigation of the dependence of the peak-height of 2-nitronaphtalene on scan rate (with solution containing 40 μmol.L-1 of 2-nitronaphtalene). This was linear in the range from 100 to 1000 mV.s-1 (ip [nA] = -0.02 v [mV.s-1] – 6.45; correlation coefficient 0.9984) and it could bring us to the conclusion that controlling process

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of this process is adsorption. However, this conclusion could be confusing, because this peak represents the total reduction (4 electron process, which is realized in a few subsequent steps). Therefore, the Elimination Voltammetry with Linear Scan (EVLS) [47, 97-99] was applied to the revealing of the processes on the electrode surface and to the calculation of coefficient αn of the controlling step [97] (the set of scan rates 100, 200, 400 and 800 mV.s-1 was used). This method belongs to the group of methods based on elimination of some particular currents registered using linear scan voltammetry (d. c. voltammetry). The EVLS is based on the assumption that the total registered current is composed from various contributions and the EVLS eliminates or conserves its particular contributions or their combinations. Two peaks were obtained on the elimination voltammograms, which corresponded to one-electron reduction (the more positive) and subsequent three-electron reduction of 2-nitronaphtalene to 2-hydroxylaminonaphtalene (the more negative) peak. According to the obtained results it is possible to conclude that controlling process of the more positive peak is in this case diffusion influenced by a capacity current (adsorption). Coefficient αn of the first step of reduction was calculated as 0.64, which confirmed the assumption of the one-electron reduction process with coefficient of the charge transfer (α=) 0.64. Achieved statistical evaluations [82] are summarized in the Table VIII.

2.3.2.2. Determination of 6-Nitroquinoline As an optimum electrochemical pretreatment of the electrode for determination of 6-nitroquinoline polarization by cyclic voltammetry was suggested: 100-200 cycles or from +1200 to –950 mV, scan rate 500 mV.s-1 in the case of graphite composite electrode in chosen supporting electrolyte [94]. This type of pretreatment is to be realized each day before the measurement was started. It is advisable to insert a few (e.g., 50 or more) cleaning cycles before the start of the measurement of each new sample. Determinations of 6-nitroquinoline were realized using DP as well as DC voltammetry (Ein = +100 mV (DPV), -200 mV (DCV) respectively; Efin = -1000 mV) with CCE30 [94]. pH 11 was chosen as the most suitable (in studied region from 3 to 12). At pH below 7 only one peak was registered, at pH > 7 the second peak appeared at the potential, which was situated about 300 mV more positively than the main peak. The dependence of the peak-height on the concentration of 6-nitroquinoline was tested in concentration intervals 0.5 – 100 μmol.L-1 (DPV) and 2 – 100 μmol.L-1 (DCV). For the evaluation of DP measurements the more negative peak was used (at about -650 mV), due to its better evaluability in comparison with the peak situated at about -300 mV. The dependence of the peak-height on the concentration of analyzed substance was linear in the whole interval from 0.5 to 100 μmol.L-1 (slope -6.54 nA.L.μmol-1, correlation coefficient 0.9994). Similarly, for the evaluation of DC-voltammetric measurements the more negative peak (from those recorded at -380 mV and -750 mV) can be utilized. DCV current - concentration dependence in concentration interval from 2 to 100 μmol.L-1 had the shape of a Frumkin isotherm. Only its segment from 2 to 10 μmol.L-1 was possible to approximate by a straight line (slope -2.27 nA.L.μmol-1, correlation coefficient 0.9997).The achieved statistical results, calculated from the calibration curve in concentration interval from 1 to 10 μmol.L-1 are listed in Table IX. The dependence of the DC-wave-height (at -650 mV) on the scan rate (concentration 10 μmol.L-1 of 6-nitroquinoline) was linear (slope -0.04 s.mV-1, correlation coefficient 0.9983

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in the range from 100 to 1000 mV.s-1). On the basis of this fact it would be possible to make the conclusion that controlling process of the electrode process on CCE30 is adsorption, but this conclusion could be again confusing, because this peak represents the total reduction occurring at chosen pH, which is realized in a few subsequent steps. Therefore, the EVLS (set of scan rates: 100, 200, 400, 800 mV.s-1) was applied for the calculation of αn coefficients of the controlling step [47, 97-99]. There were two reduction steps observed on the elimination voltammograms. The first of them corresponds probably to the one-electron reduction in adsorbed state [98] with charge transfer coefficient 0.80 and the second one probably to the three-electron reduction of 6-nitroquinoline to 6-hydroxylaminquinoline, which process is diffusion controlled.

2.3.2.3. Determination of 5-Nitrobenzimidazole As an optimum electrochemical pretreatment of the electrode for determination of 5-nitrobenzimidazole polarization by cyclic voltammetry was suggested: 100-200 cycles or from +1200 to –950 mV, scan rate 500 mV.s-1 in the case of graphite composite electrode in chosen supporting electrolyte [94]. This type of pretreatment is to be realized each day before the measurement was started. It is advisable to insert a few (e.g. 50 or more) cleaning cycles before the start of the measurement of each new sample. DC voltammetry was not sensitive enough for the determination of 5-nitrobenzimidazole and therefore DPV (Ein = -200 mV and Efin = -1200 mV) have to be used for the determination of this substance using CCE30 [94]. The pH value 9.5 proved to be most suitable for its determination with this electrode. This substance yielded only one peak in the whole range of pH. Current - concentration dependence of the peak-height of 5-nitrobenzimidazole was measured in concentration interval from 2 to 100 μmol.L-1. This dependence had the shape of stairs with two linear segments (from 2 to 8 μmol.L-1 and from 14 to 100 μmol.L-1). The slope of the second segment of this dependence amounted to -0.20 nA.L.μmol-1 and correlation coefficient 0.9994. The achieved statistical results, calculated from the calibration curve in concentration interval from 1 to 10 μmol.L-1 are listed in Table X. 2.3.2.4. Determination of Amino Compounds The use of solid graphite composite electrode in positive potential region can be documented by the oxidative determination of 2-aminonaphtalene [14]. Amino derivates of polycyclic aromatic hydrocarbons (APAH) are important chemical carcinogens. The carbon paste electrode has been mostly used for their determination, but this electrode is less stable in organic solvents in contrast to CCE30. The process of regeneration of the electrode surface is rather complicated in this case. Electrochemical regeneration seems not to be sufficient enough in this case. More repeatable results can be achieved after the mechanic regeneration of the surface by wiping with wet filter paper, what is in good agreement with the literature [100, 101]. This renovation did not influence the accuracy of the measurement as confirmed by statistical evaluation of 5 repeated measurements (between each record it was necessary to wipe the electrode surface with wet filter paper). According to the realized pH study, it is suitable to perform all measurements in BrittonRobinson buffer at pH 8. Measurements are realized without accumulation, because it did not

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increase the sensitivity of determination. Concentration dependences were measured in concentration intervals 5 - 35 μmol.L-1 and 10 - 90 μmol.L-1 [14]. The dependence of DPV peak-height on concentration can be approximated by a straight line in the case of first concentration interval (slope 0.31 nA.L.μmol-1 with correlation coefficient 0.9984). Peak corresponding to anodic oxidation of 2-aminonaphtalene was situated at the potential +600 mV. The dependence of DPV peak-height on the concentration of 2-aminonaphtalene exhibited sigmoidal shape of Frumkin isotherm in the region of higher concentrations (10 – 90 μmol.L-1) and this fact led to the conclusion that surface of the electrode was fully occupied by the analyzed substance [14]. Statistical results calculated from 15 repeated measurements were as follows: Relative standard deviation 7.1 %, skewness -0.45, excess 1.44; limit of decision 0.86 μmol.L-1; limit of detection 0.92 μmol.L-1; limit of determination 1.95 μmol.L-1 [14].

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2.3.2.5. Determination of Azo Compounds Alizarine black PT was determined using CCE, namely CCE30 [46]. The results were compared with those achieved with HMDE [46]. After mechanical pretreatment of the electrode by alumina, the electrochemical activation in chosen supporting electrolyte was to be performed by application of 100 - 200 DC cycles from +1200 mV to -950 mV, scan rate 500 mV.s-1. Electrochemical renovation of the electrode before each measurement consisted of insertion of 50 cleaning cycles minimally (+1200 mV for 0.1 s and –950 mV for 0.1 s). Britton-Robinson buffer, pH 8 proved to be the optimum supporting electrolyte for alizarine black PT determination using CCE30. Linear dynamic range was found from 2.10-6 to 1.10-4 mol.L-1. Peak potential appeared at -750 mV vs. 1 M Ag/AgCl, but it was slightly shifted to more negative potentials with increasing concentration. At higher concentrations (from 7.10-5 mol.L-1) another voltammetric peak appeared at -380 mV [35, 46]. The achieved results for all three electrodes used for determination of this analyte in waters are summarized in Table XI. 2.3.3. Determinations Using Solid Amalgam Composite Electrode In the literature [41], there are described the applications of SA-CE for determinations of some organic analytes (e.g. p-nitrophenol and ascorbic acid) in aqueous solutions using cathodic as well as anodic voltammetry. The conditions and the statistical parameters of their determinations are listed in Table XII. Due to the fact that the SA-CE, similarly as solid amalgam electrodes, can be miniaturized, they can be built-in in a multisensor, which enables the analysis of small volumes as described e.g. in [102, 103]. Such multisensor enables simultaneous analysis of one sample using set of different electrodes. The analyzed volume can amount to from 20 to 500 μL [41]. 2.3.3.1. Determination of Ascorbic Acid The suitability of m-AgSA-CE for oxidative determination in drinking waters and beverages was illustrated on ascorbic acid. The linear concentration dependence was obtained in the range from 10 to 767 μmol.L-1 with correlation coefficient r = 0.9999. The voltammograms are depicted in Figure 5A. The solution of ascorbic acid (concentration 300 μmol.L-1) was analyzed using m-AgSA-CE and for comparison using HMDE, and two

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silver solid amalgam electrodes: m-AgSAE and p-AgSAE (Figure 5B). The peak of ascorbic acid exhibits not so high symmetry as in case of other electrodes, but its height is comparable with peaks registered with other electrodes. Its shape is similar to that of p-AgSAE, which fact can be explained by similar constructions of both electrodes. From this picture it is evident that the sensitivity of all tested electrodes is almost equivalent. The conditions and the chemometric parameters of their determinations are listed in Table XII. Table XII. Comparison of achieved results at AgSA-CE for the determination of p-nitrophenol (DPCV) and ascorbic acid (DPCV), N = 11, v = 20 mV.s-1, tacc = 0 s

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Parameter Analyte concentration [μmol.L-1] Supporting electrolyte Average peak height [nA] Confidence interval (α=0.05) [nA] Standard deviation [nA] Relative standard deviation [%] Limit of detection (3xstandard deviation) [μmol.L-1] Limit of detection (Direct method of signal, IUPAC) [μmol.L-1] Limit of determination (Direct method of signal, IUPAC) [μmol.L-1]

p-nitrophenol 29 acetate buffer 0.1 M, pH 4.8 118.31 0.56 0.85 0.71 0.62 0.71

Ascorbic acid 30 NaOH 0.1 M 106.10 1.31 1.97 1.86 1.7 0.13

1.12

0.19

Figure 5. Differential pulse voltammograms of ascorbic acid on m-AgSA-CE; concentration dependence from 10 µmol.L-1 to 767 µmol.L-1; B – comparison of differential pulse voltammograms of ascorbic acid (300 µmol.L-1) on various electrodes: 1 – m-AgSA-CE, 2 - HMDE, 3 - m-AgSAE, 4 - pAgSAE; DPV; supporting electrolyte – 0.1 M acetate buffer, pH 4.8; v = 20 mV.s-1.

2.3.3.2. Determination of p-Nitrophenol Among other organic compounds, determination of nitrophenol, representing a compound containing one reducible nitro group, was described in [41]. There was depicted the concentration dependence of p-nitrophenol, measured with m-AgSA-CE in 0.1 M acetate buffer, pH 4.8. The shapes of peaks, high correlation coefficient (r = 0.9990) and chemometric parameters (listed in Table XII) indicate that the investigated electrode is suitable for the determination of nitro compounds as well. Comparison of peaks of p-nitrophenol registered with different working electrodes shows that m-AgSA-CE exhibits the characteristics similar to solid amalgam electrodes [41].

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2.3.4. Determinations Using Silica Gel Modified Carbon Composite Electrodes Electrodes of this type were successfully used for determination of pentachlorophenol [104] and of antihypertensive drug todralazine, which can be unmetabolized in human body and transported to municipal waste waters [105, 106]. Combination of organic and inorganic analysis is their utilization in investigation of metallothioneins and phytochelatins [30, 31]. In this field of analysis is determined cadmium, zinc or eventually other metals in various oxidation states, they are free or complexed in various way. The results help to elucidate the reactions and complexation processes of metallothioneins and phytochelatins with these metals [30, 31]. In [107], the application of this type of electrode for determination of phenolic antioxidants – flavonoids, concretely rutin, quercetin-3-O-rutinosa, kaempherol, morin, hesperidin, rhamnetin, flavone and 7-hydroxyflavone in aqueous solutions, teas and in wines using solid silica gel modified carbon composite electrodes is described. The electrode was activated at +800 mV, accumulation potential +200 mV, Ein +100 mV, Efin +800 mV, scan rate 10 mV.s-1. Rutin and quercetin exhibited their oxidation peak at +380 mV vs. Ag/AgCl/3M KCl in waters. This type of electrode is suitable for the determination of oxalic acid which belongs to the group of low molecular weight organic acids (LMWOAs), and its complexes with various metals (e.g. Cd, Pb) in aqueous and soil solutions as well [108].

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3. CONCLUSION On the basis of all above mentioned examples and other literature sources, it is possible to conclude that the solid composite electrodes have a wide application in analysis of not only drinking water, but in analysis of other types of water and of other environmental samples as well. This type of solid electrodes can successfully replace the traditionally used electrodes containing liquid mercury as well as other frequently applied solid and paste electrodes. They offer certain advantages over metallic or glassy carbon electrodes (i.e. electrodes consisting of one conducting phase only), e.g. lower cost, lower weight, higher signal-to-noise ratio, and the possibility of chemical modification of the conductor or of the insulator phase. A wellprepared electrode has long-term stability; it can be used for daily measurements without mechanical treatment of the surface for at least one month. In comparison with metallic silver electrode the preparation of the composite electrode is simple and economically acceptable, the electrode has very good stability and the activation of the surface is easy. This type of electrodes can be applied not only in voltammetry or chronopotentiometry, but in separation techniques with electrochemical detection as well (e.g. chromatographic and electrophoretic techniques, HPLC, FIA). Construction and preparation of these electrodes for the measurement is simple and wellprepared electrode provides simple handling and long-term mechanical stability. The activation of the electrode surface is achieved by polishing and further by repeated polarization. The following electrochemical activation of the electrode surface is realized simply by repeated cyclic polarization every day before the measurement. The reproducibility (repeatability) of the results is improved by keeping the electrode potential at a definite value for a definite time interval before each measurement.

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The selectivity of determination of organic compounds in their mixture is relatively low (e.g., all recordable reduction peaks nitro compounds on a composite electrode are located in the same potential area under equal experimental conditions), similarly as in case of other solid electrodes, but this disadvantage could be compensated by chromatographic preliminary separation of the sample. The CSEs can be utilized in batch as well as in flow through analysis. In such a case, their surface can be easily renewed or they can be applied in the form of disposable electrodes, which do not represent any environmental litter. These electrodes can be easily miniaturized and used in the form of mini- or microsensors and grouped together with other electrodes to a multisensor.

ACKNOWLEDGMENTS Financial support of this work was provided by the Grant Agency of the Czech Republic (project No. 203/07/1195), by the Grant Agency of the Academy of Sciences of the CR (project No. IAA400400806), by the Ministry of Education, Youth and Sports of the Czech Republic (projects LC 06035 and MSM 0021620857).

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[82] Meloun, M.; Militky, J.; Forina, M. Chemometrics for analytical chemistry, Volume 1: PC-Aided Statistical Data Analysis, Volume 2: PC-Aided Regression and Related Methods; Ellis Horwood: Chichester, 1992. [83] Yosypchuk, B.; Heyrovsky, M.; Palecek, E.; Novotny, L. Electroanal. 2002, 14, 14881493. [84] Yosypchuk, B. Ph.D. Thesis, University Pardubice, Pardubice, 2003. [85] Yosypchuk, B.; Novotny, L. Chem. Listy 2002, 96, 886-888. [86] Yosypchuk, B.; Novotny, L. Electroanal. 2002, 14, 1733-1738. [87] Yosypchuk, B.; Novotny, L. Electroanal. 2002, 14, 1739-1741. [88] Yosypchuk, B.; Novotny, L. Chem. Listy 2002, 96, 756-760. [89] Navratil, T.; Yosypchuk, B.; Lukina, A. N.; Peckova, K.; Barek, J. ChemZi 2007, 3, 5657. [90] Wightman, R. M. Science 1988, 240, 415-420. [91] Kopanica, M.; Stara, V. Electroanal. 1991, 3, 925-928. [92] Peckova, K.; Barek, J.; Moreira, J. C.; Zima, J. Anal. Bioanal. Chem. 2005, 381, 520525. [93] Peckova, K.; Barek J.; Zima, J. Proc. 45. Zjazd PTCh i SITPCh, M. Drach, K. Nieszporek, Eds.,p.1236. Primooffset, Lublin 2003. [94] Navratil, T.; Barek, J; Fasinova-Sebkova, S. Electroanalysis 2008, Submitted. [95] Stradyns, J.; Glezer, V., Eds. Encyclopedia of the electrochemistry of Elements; M. Dekker: New York, 1979. [96] Zanoni, M. V. B.; Fogg, A. G.; Barek, J.; Zima, J. Anal. Chim. Acta 1997, 349, 101109. [97] Dracka, O. J. Electroanal. Chem. 1996, 402, 19-28. [98] Sander, S.; Navratil, T.; Novotny, L. Electroanal. 2003, 15, 1513-1521. [99] Fadrna, R.; Yosypchuk, B.; Fojta, M.; Navratil, T.; Novotny, L. Anal. Lett. 2004, 37, 399-413. [100] Barek, J.; Cvacka, J.; Muck, A.; Quaiserova, V.; Zima, J. Electroanal. 2001, 13, 799803. [101] Armalis, S.; Novikova, N.; Kubiliene, E.; Zima, J.; Barek, J. Anal. Lett. 2002, 35, 15511559. [102] Navratil, T.; Yosypchuk, B.; Fojta, M., The Way of Automated Measurement of Electrochemical Signal on Number Sets Electrodes and Apparatus for the Realization of this Way, Application No. PV 2007-40, 2007. [103] Navratil, T.; Yosypchuk, B.; Barek, J. Chemia Analityczna 2008, Submitted. [104] Barrio, R. J.; Debalugera, Z. G.; Goicolea, M. A. Anal. Chim. Acta 1993, 273, 93-99. [105] Walcarius, A.; Lamberts, L.; Derouane, E. G. Electrochim. Acta 1993, 38, 2257-2266. [106] Walcarius, A.; Lamberts, L.; Derouane, E. G. Electrochim. Acta 1993, 38, 2267-2276. [107] Sulc, M.; Sestakova, I., Proc.Conf. Modern Electrochemical Methods XXVII, J.Barek, T. Navrátil, Eds, p. 159-162. Czech Chemical Society, Prague 2007. [108] Jaklova Dytrtova, J.; Sestakova, I.; Jakl, M.; Navratil, T. Electroanalysis 2008, Submitted.

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Chapter 3

VOLTAMMETRIC AND AMPEROMETRIC DETERMINATION OF ORGANIC POLLUTANTS IN DRINKING WATER USING BORON DOPED DIAMOND FILM ELECTRODES K. Peckova∗, J. Musilova, J. Barek and J. Zima Charles University in Prague, Faculty of Science, Department of Analytical Chemistry, UNESCO Laboratory of Environmental Electrochemistry, Albertov 6, CZ-128 43 Prague 2, Czech Republic

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ABSTRACT There is a never-ending search for new electrode materials for voltammetric or amperometric determination of various detrimental substances in drinking water. The basic requirements for new electrode materials are lower noise, broader potential range, mechanical robustness enabling measurements in flowing systems, compatibility with organic solvents making them compatible with high performance liquid chromatography (HPLC), flow injection analysis (FIA) or capillary electrophoresis (CE) with electrochemical detection (HPLC/ED, FIA/ED, CE/ED) and – last but not least – resistance towards passivation and electrode fouling which is probable the biggest complication in the practical applications of electroanalytical methods. One of the most promising non-traditional electrode materials is boron doped polycrystalline diamond film on a silica support, which is suitable for the determination of a wide spectrum of organic and inorganic analytes. The electrodes based on it possess excellent electrochemical properties, such as a low and stable background current over a wide potential range, corrosion resistance, high thermal conductivity and high current densities. Furthermore, this material offers superb micro structural stability at extreme cathodic and anodic potentials and resistance to fouling because of weak adsorption of polar species on the hydrogen terminated paraffin-like surface, which results in good responsiveness for many redox analytes without any pretreatment. Boron doped diamond ∗

E-mail:[email protected]

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K. Peckova, J. Musilova, J. Barek and J. Zima film electrodes are applicable to voltammetric or amperometric determination of both oxidisable and reducible substances with limit of determination down to 10-8 mol L–1 without any preconcentration step. The chapter is devoted to their practical applications for both batch voltammetric analysis and continuous flow amperometric analysis (HPLC/ED, FIA/ED, CE/ED) of organic pollutants possibly present in drinking water.

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LIST OF SYMBOLS AND ABBREVIATIONS 1,2-PDA 2,4,5-T 2,4,6-TCP 2,4-DB 2,4-DCP 2,4-DP 2-AN 2-CA 2-CP 3-CP 4-AP 4-CP 4-NP BDD c CE CP CV CVD DCP DPV DW ECD ED EP EU FIA FID FPD GC GCE HMDE HPLC HQ i.d. IARC LDR

1,2-phenylenediamine 2,4,5-trichlorophenoxyacetic acid 2,4,6-trichlorophenol 4-(2,4-dichlorophenoxy)butyric acid 2,4-dichlorophenol 2,4-dichlorophenoxypropionic acid 2-aminonaphthalene 2-chloroaniline 2-chlorophenol 3-chlorophenol 4-aminophenol 4-chlorophenol 4-nitrophenol boron-doped diamond molar concentration capillary electrophoresis chlorinated phenol cyclic voltammetry chemical vapour deposition dichlorophenol differential pulse voltammetry drinking water electron capture detector electrochemical detection environmental pollutant European Union flow injection analysis flame ionization detector flame photometric detector gas chromatography glassy carbon electrode hanging mercury drop electrode high performance liquid chromatography hydroquinone inner diameter International Agency for Research on Cancer linear dynamic range

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Voltammetric and Amperometric Determination of Organic Pollutants… LOD LOQ MCL MCPA MCPB MCPP MS Nd:YAG o-ABA OP PAH PCP Ph Pt/BDD ROS RSD S/N SCE SEM SHE Si/BDD SPE TCP US EPA UV VIS vs. WHO

105

limit of detection limit of quantitation maximum allowable contaminant level 4-chloro-2-methylphenoxyacetic acid 4(2-methyl-4-chlorophenoxy)butyric acid 2(2-methyl-4-chlorophenoxy)propionic acid mass spectrometry neodymium-doped yttrium aluminium garnet o-aminobenzoic acid organophosphorous pesticide polycyclic aromatic hydrocarbon pentachlorophenol phenol boron-doped diamond deposited on platinum reactive oxygen species relative standard deviation signal-to-noise ratio saturated calomel electrode scanning electron microscopy standard hydrogen electrode boron-doped diamond deposited on silica solid phase extraction trichlorophenol United States Environmental Protection Agency ultraviolet part of the spectrum visible part of the spectrum versus World Health Organization

1. INTRODUCTION Access to safe drinking-water (DW) is important as a health and development issue at a national, regional and local level. This worldwide concern is respected by World Health Organization (WHO), which has published since 1958 the WHO International Standards, in 1983 extended to Guidelines for Drinking-water Quality. The Guidelines are recognized as representing the position of the United Nations (UN) system on issues of DW quality and health by “UN-Water”, the body that coordinates programs concerned with water issues amongst the 24 UN agencies. The last, third edition of the Guidelines for Drinking-water Quality (further denoted as Guidelines), released in 2004 and revised in 2005 and 2006 [1] explains requirements to ensure DW safety, how these requirements are intended to be used and includes specific guideline values and fact sheets for significant chemicals and microbial hazards. The volume also describes the approaches used in deriving the guideline values. The Guidelines are also accompanied by other publications explaining the scientific basis of their development and

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providing guidance on good practice in implementation. Safe DW, as defined by the Guidelines, does not represent any significant risk to health over a lifetime of consumption, including different sensitivities that may occur between life stages. The WHO Guidelines is a basic document, which can serve to governmental authorities and institutions, when releasing their own documents and mandatory standards on DW safety respecting the local or national environmental, social, economic and cultural conditions. Exact description of analytical methods suitable for the determination of chemical hazards in DW is not a part of the Guidelines, only recommended methods are briefly listed. However, many national agencies setting the enforceable or recommended standards on DW quality including United States Environmental Protection Agency (US EPA) [2] and Commissions of the European Union (EU) [3] offer a number of validated methods for the determination of selected chemicals or group of chemicals. The electroanalytical methods are rarely among them, as they are often assessed by institutional authorities and laboratory personal as too demanding on trained personal to give reliable results. Nevertheless, the fundamental and application research continues to search for new electrode materials, methods and their applications for analysis of a wide variety of matrices and simultaneously it is struggling on the popularization of electroanalytical methods and their implementation into the praxis. The introduction of diamond based electrodes by Iwaki in 1983 [4] and Pleskov in 1987 [5], boron-doped diamond (BDD) thin films in 1992 by Fujishima [6] and their applications a year later [7-9] started a new era in electrochemical research. Over the past fifteen years, it has become apparent that the diamond-based electrodes are in many ways ideal as electrode materials for electrochemistry and thus can be used for high-sensitivity analytical measurements of a wide variety of organic and inorganic species. The biggest popularity have gained polycrystalline BDD thin films. Due to the relatively short history, a limited spectrum of environmental pollutants (EPs) has been so far investigated by means of BDD electrodes. The overview of organic pollutants of DW referenced in Guidelines and investigated so far by a voltammetric or amperometric method using a BDD electrode is summarized in Table 1. For better orientation, there is listed the source of the chemical as DW pollutant, its guideline value and the treatment method used to decrease its concentration preferably below it, further analytical methods recommended for the determination of the chemical in DW and finally the electroanalytical methods applied at BDD electrodes with corresponding reference. The following part of introduction first briefly describes the preparation and characterization of BDD thin films and their use in electrochemistry in general. The main part of this chapter is devoted to the results achieved with BDD thin film electrodes in the field of batch voltammetric analysis and electrochemical, preferably amperometric detection (ED) of organic DW pollutants in connection with liquid flow methods (high performance liquid chromatography (HPLC), flow injection analysis (FIA), and capillary electrophoresis (CE)). Potential of these methods for other environmental applications is also briefly shown.

1.1. Preparation and Characterization of Boron Doped Diamond Thin Films Diamond thin films can be grown from dilute mixtures of a hydrocarbon gas (e.g. methane) in hydrogen using one of several energy-assisted chemical vapor deposition (CVD) methods, the most popular being hot-filament and microwave plasma assisted CVD.

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Table 1. Summary of organic drinking-water (DW) pollutants and their recommended guideline values for DW according to WHO Guidelines [1] investigated so far by means of boron-doped diamond (BDD) electrodes Treatment method to achieve guideline value in DW Air stripping; Activated carbon; Ozonation

Recommended analytical method GC/ECD; GC/FID; PT-GC/MS

BDD application

Reference for BDD application

BDD microelectronic gas sensor; CV + destruction

[10]

GC/FID; PT-GC/MS

BDD microelectronic gas sensor

0.7 (NT)

Air stripping; Activated carbon; Ozonation; Advanced oxidation ----a

CV + HPLC/ED

[12]

200 (C)

----a

GC/MS; HPLC/FD GC/ECD; GC/MS

SWV CE/ED; FIA and HPLC/ED SWV; CV + FIA and HPLC/ED; CE/ED CV + FIA and HPLC/ED

[13] [14-16] [17] [18, 19] [20]

CV + ChA

[22]

CV + ChA

[22]

CV + HPLC/ED

[12]

WHO referenced DW pollutant

Source of DW pollution

Guideline value for DW [μg L–1]

Benzene

Industrial sources and human dwellings

10 (NT)

Toluene

Industrial sources and human dwellings

700 (C)

Benzo[a]pyrene

Contaminants from pipes and fittings Disinfection by-product

2,4,6Trichlorophenol PCP

Industrial sources and human dwellings

9 (P, NT)

Activated carbon

GC/ECD; PT-GC/MS; HPLC

Carbofuran

Agricultural activities

7

Activated carbon; Membranes

GC/TID; HPLC/FD

MCPA

Agricultural activities

2

Activated carbon; Ozonation

MCPP

Agricultural activities

10

Activated carbon; Ozonation

Fluoranthene

Contaminants from pipes and fittings

LC

----a,b

GC/ECD; GC/MS GC/ECD; GC/MS; HPLC/UVPAD ----c

[11] [10]

[14, 15] [21]

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Table 1. Continued WHO referenced DW pollutant 2-Chlorophenol

2,4Dichlorophenol

Parathion

Source of DW pollution

Guideline value for DW [μg L–1]

Treatment method to achieve guideline value in DW

Disinfection byproduct

ID US EPA: 40 (lifetime exposure)

----a

ID US EPA: 20 (lifetime exposure)

----a

Disinfection byproduct

Agricultural activities

LC

----b

Recommended analytical method ----c

----c

----c

BDD application

Reference for BDD application

CV + FIA and HPLC/ED; FIA/ED; CE/ED; microchip CE/ED SWV CV + FIA and HPLC/ED; FIA/ED; CE/ED; microchip CE/ED SWV

[20] [23] [14-16] [24] [13] [17] [23] [14-16] [24] [25]

PCP – pentachlorophenol; MCPA – 4-chloro-2-methylphenoxyacetic acid; MCPP – 2(2-methyl-4-chlorophenoxy)propionic acid. US EPA – Environmental Protection Agency of United States; P – provisional guideline value, as there is evidence of a hazard, but the available information on health effects is limited; C – concentrations of the substance at or below the health-based guideline value may affect the appearance, taste or odor of the water, leading to consumer complaints; NT - non-threshold substance, the guideline value is the concentration in DW associated with an upper-bound excess lifetime cancer risk of 10-5 (one additional cancer per 100 000 of the population ingesting DW containing the substance at the guideline value for 70 years). ID – available data inadequate to permit derivation of health-based guideline value; LC – occurs in DW at concentrations well below those at which toxic effects may occur. Treatment method: a no data available, chemicals entering DW from materials in contact with it or directly during water treatment; b no data available, expected natural concentration of the chemical in DW is low. GC – gas chromatography/ECD - electron capture detector, MS – mass spectrometry, FID – flame ionization detector, TID – thermal ionization detector; PT-GC/MS – purge-and-trap gas chromatography/mass spectrometry; HPLC – high performance liquid chromatography/FD fluorescence detector, UVPAD – ultraviolet photodiode array detector, c no method recommended in Guidelines. CV – cyclic voltammetry; ChA – chronoamperometry, HPLC/ED; FIA/ED; CE/ED – high performance liquid chromatography, flow injection analysis, capillary electrophoresis with electrochemical detection (in all cases using BDD electrode); SWV – square wave voltammetry.

Voltammetric and Amperometric Determination of Organic Pollutants…

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In these processes, a carbon containing gas is energetically activated to decompose the molecules into methyl-radicals and atomic hydrogen if a methane-hydrogen source gas mixture is used. The substrate, typically silica wafers, is pretreated by cleaning with a series of solvents and “seeding” with small particles by polishing the substrate with diamond powder. The embedded particles serve as nucleation centers for film growth. The growth methods mainly differ in the manner in which the gas activation is accomplished. Typical growth conditions are C/H ratios of 0.5-2 %, pressures of 10-150 torr, substrate temperatures of 700-1000 °C, and microwave powers of 1000-1300 W, or filament temperatures up to ~ 2800 °C, depending on the methods used. The film grows by nucleation at rates in the 0.1-2 μm h–1 range. For the substrates to be continuously coated with diamond, the nominal film thickness must be ~ 1 μm. Boron doping is accomplished from the gas phase by mixing boron-containing compounds such as B2H6, trimethylborane or B2O3 with the source gases, or from the solid state by gasifying a piece of hexagonal-boron nitride (h-BN). In the solid state approach, diborane is produced by the interaction of atomic hydrogen with h-BN and is then incorporated into the gas flux to the substrate. The doping level can be as high as 10 000 ppm of boron, resulting in film resistivities < 0.1 Ω cm [26, 27]. Several analytical techniques are routinely used to characterize the morphological, optical, chemical and electronic properties of diamond thin films. Figure 1 represents a scanning electron micrograph (SEM) of a BDD thin film grown on p-type Si by microwave assisted CVD. The film is composed of sharp, well faceted microcrystallites ranging in size from 0.5 to 3 μm with no obvious preferential orientation.

Figure 1. Scanning electron micrograph of microcrystalline boron-doped diamond thin film on p-Si deposited by microwave assisted CVD. Courtesy of prof. G. M. Swain, Michigan State University, USA.

The Raman spectroscopy is quite sensitive to the presence of non-diamond carbon impurities. These consists likely of mixture of sp3- and sp2-hybridized bonding, similar to diamond-like and amorphous carbon and causes weak scattering intensity in the region of about 1520 cm–1. Typical Raman spectra for high-quality diamond and BDD films feature one intense band at 1332 cm–1.

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Scanning tunneling and atomic force microscopy are used to evaluate the film morphology and probe the local electronic properties. Powder X-ray diffraction analysis is used to investigate the preferential crystallite orientation of the films. Surface elemental composition is determined by Auger electron and X-ray photoelectron spectroscopies. Secondary ion mass spectrometry (MS) is used to quantify the boron dopant concentration and probe the spatial distribution of the dopant species over the surface and within the bulk [26].

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1.2. BDD Thin Films as Electrode Material The use of conductive BDD thin films as an electrode material is of increasing interest in scientific and commercial sphere as the electrodes possess several excellent electrochemical properties: Low and stable background current over a wide potential range, corrosion resistance, high thermal conductivity and high current densities. Further, this material offers superb micro structural stability at extreme cathodic and anodic potentials and resistance to fouling because of weak adsorption of polar species on the hydrogen terminated paraffin-like surface, which results in good responsiveness for many redox analytes without pretreatment [26, 28-30]. There exist four main directions in the use of BDD electrodes in electrochemistry: (i) Electrochemical oxidation of EPs at BDD anodes proposed for their quantitative conversion or destruction in wastewaters, (ii) electrochemical disinfection of drinking and bathing water, (iii) use of BDD electrodes in electroanalysis as electrochemical sensors employed in voltammetric methods or liquid flow methods (HPLC, FIA, CE) for detection of organic and inorganic species in environmental, biological and pharmaceutical matrices, and (iv) electrochemical synthesis, in particular in the production of strong oxidizing agents (e.g., peroxodisulfuric acid [31, 32], hydrogen peroxide, ozone, chlorine [33] or ferrates [34]), in electroorganic synthesis [35-38] and technical galvanic applications such as lead free chroming or recycling processes [39]. The first two points (i, ii) interfering with the thematic scope of this book – DW quality – are briefly reviewed in following paragraphs. (i) The principal aim of the wastewater treatment is the complete oxidation of EPs to CO2 or their conversion to biocompatible compounds. BDD electrodes are extremely suitable for this purpose, because the wide overpotential of BDD for the water oxidation favours the generation of large quantities of hydroxyl radicals, which are powerful oxidants ensuring the direct oxidation of the organics on the BDD anode surface so that its passivation is prevented. These studies are often accompanied by voltammetric studies in order to elucidate the mechanism involved in the electrochemical oxidation and to elucidate the influence of waste and operative characteristics (initial concentration, pH, temperature, current density, supporting electrolyte and its stirring rate, cell design) in the process. Several reviews were devoted to this topic recently [40-43] showing that the methods lead to almost complete mineralization of EPs. Precisely, destruction methods using BDD anodes were proposed for following organic DW pollutants included in WHO Guidelines: Selected chlorophenols [44-46], pesticide parathion [25], selected chlorophenoxy herbicides [22, 47-49], benzene [11], and tributyltin [50]. Other important EPs treated by anodic oxidation using BDD anodes include: Pesticides and their degradation products, e.g. methylparathion [51], amitrole [52], clofibric acid [53, 54], 4-chlorophenoxyacetic acid [55], 2,4-dinitrophenol [56] and

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Voltammetric and Amperometric Determination of Organic Pollutants…

111

4-nitrophenol [57], further aniline [58], various dyes [59-65], surfactants [60, 66], polyaromatics and their derivatives [67, 68], phenolic and benzoic compounds [11, 36, 69-74], and selected drugs [75]. Most of these studies so far were performed at silica supported devices, in spite of the difficulties related to their industrial transposition due to the fragility and the relatively low conductivity of the Si substrate. Nowadays the stress is put on the search for new substrates for diamond depositions. Metals like Nb, Ta, W and substantially cheaper Ti are promising [76, 77] and were tested for the destruction of several EPs, e.g. dyes [78-81], carboxylic acids [80], phenol [79, 80], and organotin compounds [50]. (ii) The application of BDD for water disinfection purposes represents a less explored field. It is accepted that •OH radicals are produced at sufficient high anodic polarization [37]; subsequently, the complex chemical and electrochemical reactions lead to production of other strong oxidants, so called reactive oxygen species (ROS, e.g. •OH, H2O2, O3, •O2–). These contribute significantly to destruction of organic micropollutants and microorganism, such as bacteria, viruses, protozoa and helminths. It was shown that pathogens typically selected as indicators of water contamination (Escherichia coli, Enterococcus faecalis, Legionella pneumophila, coliform bacteria) are inactivated in chloride-free media after electrochemical disinfection using BDD anode [82-84]. However, the presence of chlorides for practical applications should be considered as they are ubiquitous and no health-based guideline value is proposed by WHO – higher concentrations just affect the acceptability of DW. The electrolysis of chlorides results preferably in production of free or active chlorine components (free available chlorine, dissolved Cl2, HOCl and ClO–) having high disinfecting ability. However, under uncontrolled conditions chlorite, chlorate and perchlorate may be formed by bulk reactions and further electrochemical oxidation [85]. Perchlorate, suspicious from carcinogenity [86], was identified as product by long-term electrolysis of DW at BDD anodes [85, 87], chlorite and chlorate appear as intermediates and by-products. Due to their adverse health effects their concentrations in DW are usually regulated by national legislations. Thus, the control of these toxic by-products is of great concern and optimal operative conditions of electrolysis guaranteeing their bellow limit concentrations must be achieved with simultaneous high production of disinfection chlorine components (i.e. hypochlorite ClO–). Further, the fate of common water components, e.g. nitrates, sulfates, bicarbonates, and phosphates must be investigated under various electrolysis conditions, as these may react with ROS and active chlorine species under formation of various, possibly undesirable by-products (e.g., nitrite, ammonium ions, chloramines). To sum up, a more complex and careful approach to BDD applications for water disinfection purposes is recommended. The most important application of diamond electrodes with regard to the content of this book are electroanalytical applications (iii), described so far mostly on devices using BDD thin films deposited on silica (Si/BDD). Nowadays, there are several commercial suppliers of Si/BDD materials, including Element Six (formerly De Beers Industrial Diamond, UK), Windsor Scientific (UK), Adamant Technologies (Switzerland), Condias (Germany), Sumitomo (Japan), and sp3 Diamond Technologies (USA). Many other research laboratories produce the BDD materials for their own purposes. In general, attention is paid to both inorganic and organic species. In the next section, details will be given on the use of Si/BDD electrodes as voltammetric and amperometric sensors for determination of potential organic DW pollutants. The pollutants widely examined involve nonchlorinated and chlorinated phenols, pesticides and their degradation products and monocyclic and polycyclic aromatic

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hydrocarbons. Analytical applications of BDD electrodes were also subject to several reviews in last five years [28, 88-93].

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2. VOLTAMMETRY AT BDD THIN FILM ELECTRODES IN BULK MEASUREMENTS Voltammetric methods are used in electrochemistry to investigate electrochemical processes at the electrode surface and as analytical tool for quantitation of analytes. However, when analyzing complex matrices the selectivity of batch voltammetric methods is often insufficient due to relatively narrow potential window and the near oxidation / reduction potentials for structurally relative group of organic compounds, which are often found together in an environmental or biological matrix. Thus, preliminary on- or off-line separation of analytes may be required complicating and prolonging the analysis. Therefore, in the relatively short history of BDD electrodes, voltammetric methods developed for the qualitative or quantitative determination of organic analytes in various matrices and characterized by exact analytical figures of merit (i.e., linear dynamic range (LDR) and slope and intercept for linear calibration dependences, limit of detection and quantitation (LOD and LOQ), repeatability / reproducibility of the electrode signal and others) are so far in minority against voltammetric investigations concerning basic electrochemical properties of selected groups of analytes. In this case, cyclic voltammetry (CV) is usually the first choice for electrochemists investigating the reaction mechanism and its kinetics, electrode pretreatment, passivation of the electrode surface and its remediation. These studies usually precede further applications of BDD electrodes either for anodic decomposition of organic compounds or their qualitative and quantitative determination using voltammetry or more often amperometry in connection with liquid flow methods. In Table 2, there is an overview of organic DW contaminants (according WHO Guidelines) investigated so far for any purpose by batch voltammetric methods using BDD electrodes. These contaminants include chlorinated phenols (CPs), several pesticides, and monocyclic and polycyclic aromatic hydrocarbons and their derivatives. The Table 2 contains for each analyte the purpose of voltammetric study and the methods and conditions used, further characterization of used BDD electrode and achieved LOQ and LDR (when given). Each referred publication is further shortly introduced in the following text. The electrochemistry of other EPs and biological active compounds not likely occurring in DW but investigated at BDD electrodes is not included in detail and more on these studies can be found in several reviews and monographs published recently [28, 88-93].

2.1. Chlorinated Phenols Phenols and chlorinated phenols (CPs) enter the aquatic environment as by-products of industrial processes, such as the production of antioxidants, dyes, and drugs. In DW they occur as a result of the chlorination of phenols, as by-products of reactions with phenolic acids, as biocides or as degradation products of phenoxy herbicides and chlorinated bleaching of paper [94-96].

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Table 2. Overview of batch voltammetric methods using BDD electrodes for organic drinking-water (DW) pollutants according to WHO Guidelines [1] WHO referenced DW pollutant

BDD electrode and used pretreatment

Purpose of the study

Voltammetric method and conditions

2,4-DCP

Study of oxidation mechanism and electrode passivation

CV, BR buffer, pH 2.0.

2,4-DCP 2,4,6-TCP

"As-deposited" and anodically oxidized (+2.64 V for 4 min in BR buffer, pH 2) MPA CVD deposited BDD, A = 0.09 cm2. HF CVD deposited polycrystalline BDD, anodic treatment at +3.0 V for 30 min followed by cathodic treatment at –3.0 V for 30 min; A = 0.62 cm2.

Optimization of conditions for direct determination in pure and river water for mixture of analytes (4-CP, 2,4-DCP, and 2,4,6-TCP).

2-CP PCP

nanocrystalline and microcrystalline MPA CVD deposited BDD, A = 0.09 cm2.

Study of oxidation mechanism and electrode passivation.

PCP

HF CVD deposited polycrystalline BDD, anodic treatment at +3.0 V for 30 min followed by cathodic treatment at –3.0 V for 30 min; A = 0.62 cm2.

Optimization of conditions for direct determination in pure, river water and soil.

SWV combined with mathematical deconvolution procedure, BR buffer, pH 6, pulse amplitude 50 mV, frequency 100 s–1, scan increment 2 mV. CV, 0.05 mol L–1 phosphate buffer, pH 3.5. SWV, BR buffer, pH 5.5, pulse amplitude 50 mV, frequency 100 s–1, scan increment 2 mV.

Parathion

MCPA, MCPP

HF CVD deposited polycrystalline BDD, anodic treatment at +3.0 V for 30 min followed by cathodic treatment at –3.0 V for 30 min; A = 0.62 cm2. HF CVD deposited polycrystalline BDD, anodic treatment in 1 mol L–1 HClO4 for 5 min at i = 30 mA cm–1, A = 1 cm2.

Optimization of conditions for direct determination in pure and river water

SWV, BR buffer, pH 7.0, pulse amplitude 40 mV, frequency 150 s–1, scan increment 2 mV.

Study of oxidation mechanism and electrochemical decomposition.

Chronoamperometry and CV, 1 mol L–1 HClO4.

Ref.

LOQ [μmol L–1], (μg L–1) for given matrix not given

[17]

not given

[13]

not given

[20]

[0.067], (18) in pure water; [0.056], (15.5) in river water (all LOD); for soil not given. [0.030], (8.72) in pure water; [0.112], (32.65) in river water.

[18]

not given

[19] [25]

[22]

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Table 2. Continued WHO referenced DW pollutant

BDD electrode and used pretreatment

Purpose of the study

Voltammetric method and conditions

Benzene

HF CVD deposited polycrystalline BDD, anodic treatment at +3.0 V for 30 min followed by cathodic treatment at –3.0V for 30 min, A = 0.2 cm2. commercial BDD (Windsor Scientific), 3 mm diameter, anodic treatment in phosphoric acid / acetonitrile at +2.5 V for 10 min.

Study of oxidation mechanism and electrochemical decomposition.

CV, 0.5 mol L–1 H2SO4.

Study of oxidation mechanism.

CV in 0.04 mol L–1 phosphoric acid / acetonitrile mixture

BaP, Fluoranthene

LOQ [μmol L–1], (μg L–1) for given matrix only LDR given: (0.361.05)·10-3 mol L–1 not given

Ref.

[11]

[12]

2-CP – 2-chlorophenol; 4-CP – 4-chlorophenol; 2,4-DCP – 2,4-dichlorophenol; PCP – pentachlorophenol; MCPA – 4-chloro-2-methylphenoxyacetic acid; MCPP – 2(2-methyl-4-chlorophenoxy)propionic acid; BaP – benzo[a]pyrene. MPA CVD – microwave plasma-assisted chemical vapour deposition; HF CVD – hot filament chemical vapour deposition; A – geometric surface of the electrode; i – current density; BR buffer – Britton-Robinson buffer. CV – cyclic voltammetry; SWV – square wave voltammetry. LOQ – limit of quantitation; LOD – limit of determination; LDR – linear dynamic range.

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CPs concentration in DW usually does not exceed 1 μg L-1. CPs mentioned in WHO Guidelines include: 2-chlorophenol (2-CP), 2,4-dichlorophenol (2,4-DCP) and 2,4,6-trichlorophenol (2,4,6-TCP) – by-products of DW chlorination – and pentachlorophenol (PCP) which is used for protecting wood from fungal growth. The data on toxicity on humans of most CPs are limited, the International Agency for Research on Cancer (IARC) [97] – part of WHO – has classified combined exposures to polychlorophenols or to their sodium salts as possibly carcinogenic to humans (Group 2B) [98]. The WHO set the limit for DW only for 2,4,6-TCP, which has shown week mutagenic activity in several in vitro and in vivo studies, and PCP, which is carcinogenic to animals. Their WHO guideline values are associated with a 10–5 upper-bound excess lifetime cancer risk (or one additional cancer per 100 000 of the population ingesting DW containing the substance at the guideline value for 70 years), similar as guideline values for other genotoxic carcinogens, so called non-threshold chemicals. For 2,4,6-TCP the guideline value is 200 μg L–1. Nevertheless, the lowest reported taste threshold limit for the chemical is 2 μg L–1 [96]. For PCP the WHO guideline value of 9 μg L–1 is recommended. Some national agencies prescribe even lower values; EPAs' maximum allowable contaminant level (MCL) for DW is 1 μg L–1 [99]. Beside analytical methods recommended by WHO for each of the analyte (Table 1), this agency recommends isotope dilution gas chromatography – mass spectrommetry (GC/MS, EPA Method 1625) with detection limits of tens of μg L–1 for determination of CPs and other semivolatile organic compounds and more sensitive GC with flame ionization detector (GC/FID) or electron capture detector (ECD/GC) (both Method 604, detection limits hundreds of ng L–1), or GC/MS (method 625, detection limits units of μg L–1) after liquid-liquid extraction with methylene chloride for analysis of municipal or industrial wastewaters [100]. Electroanalysis of phenol derivatives including CPs is commonly problematic at most solid electrodes mainly due to the electrode passivation resulting from oxidation products. The mechanistic work by Gatrell and Kirk revealed that the phenol oxidation reaction at platinum electrodes proceeds initially through the one-electron, one-proton formation of a phenoxy radical species, which can subsequently undergo radical-radical coupling to form oligomeric and polymeric species [101, 102]. Follow-up direct and indirect oxidation reactions producing soluble products, such as hydroquinone and catechol, are also possible. Similarly, the studies on reaction mechanism of electrooxidation of phenol [36, 69, 74] and CPs [17, 20] in acidic media at BDD electrodes revealed that in the potential region of water stability direct electron transfer can occur on BDD surface at the potential of +800 to +1400 mV forming phenoxy radical species with possible electrode fouling due to the formation of a polymeric film on its surface. The formation of o- and p-benzoquinone as the main products of phenol oxidation at BDD electrodes was confirmed by Terashima et al. [17]. They published an extended comparative study on electrooxidation of 2,4-DCP, phenol, hydroquinone, and catechol at ordinary glassy carbon electrode (GCE), anodically pretreated and at "as deposited" BDD electrodes. Reproducible, well-defined CVs were obtained even at high 2,4-DCP concentration (5 mmol L–1) at anodized BDD and GCE, however, upon repetitive cycling the signals disappeared due to formation of a passivation film. In contrary to GCE, the partially deactivated diamond surface could be reactivated on-line by applying a highly anodic potential (+2.64 V vs. saturated calomel electrode (SCE)) for 4 min, which enabled the destruction of the passivating layer on the surface. Hydroxyl radicals produced by the high applied potential are believed to be responsible for the oxidation of the passivating

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layer. The analytical results on coupling of the electrodes with FIA and HPLC are discussed in the next chapter. Several other voltammetric studies were published concerning oxidative detection and remediation of CPs at BDD electrodes [20, 103, 104] with different approaches to overcome the fouling problems. Saterlay et al. [103] observed that the use of power ultrasound during the electrooxidation of 4-chlorophenol (4-CP) in aqueous acidic media facilitates its voltammetric determination due to surface regeneration during measurement. Moreover, the ultrasonic field also yields an enhanced oxidation current signal due to more efficient mass transport of 4-CP. Linear dynamic range of 1-300 μmol L–1 and a detection limit of 1 μmol L–1 (128.6 μg L–1) was achieved in this study. In another study by Pedrosa et al. [104], direct determination of this pollutant using square wave voltammetry (SWV) at BDD electrode in Britton-Robinson (BR) buffer, pH 6 without any electrode pretreatment was reported with LOQ of 6.4 μg L–1 for pure and 21.5 μg L–1 for polluted river water. These achieved detection limits are satisfactory for the EPAs' concentration limit of 100 μg L–1 for surface waters [105] and show that the technique is analytically useful over a concentration range where aquatic 4-CP pollution is known to occur. In the next study the same authors focused on the use of SWV for simultaneous determination of 4-CP, 2,4-DCP and 2,4,6-TCP [13]. The contribution of each individual analyte to the unique voltammetric peak was calculated by mathematical deconvolution procedure. The LOQs for 4-CP in the presence of the other analytes of 0.072 μmol L–1 (9.2 μg L–1) for pure water and 0.329 μmol L–1 (42.2 μg L–1) for river water were achieved, the results were in good agreement with values obtained by HPLC/UV. Also CV experiments of Muna et al. [20] revealed no significant electrode fouling at microcrystalline and nanocrystalline BDD electrodes without any reactivation, when working with phenol and CPs (2-CP, 3-CP, 4-CP, PCP) in phosphate buffer, pH 3.5. However, these authors were working with monochlorinated phenols in contrary to Terashima [17], who reported fouling of BDD electrodes by electrooxidation of di- and trichlorophenols as already mentioned. According to Muna studies, the electrooxidation of phenol, 2-CP and 4-CP yields similar responses corresponding to the formation of surface-confined hydroquinone/p-benzoquinone and catechol/o-benzoquinone redox couples and/or polar oligomeric / polymeric species, which can be easily removed from the electrode surface by copiously rinsing with ultrapure water and exposure to fresh electrolyte. The electrooxidation of 3-CP and PCP differs, PCP oxidation involves a one-electron, one proton transfer to form pentachlorophenoxyradical, which undergoes radical-radical coupling to form various dimers. For 3-CP, first the 3-chlorophenoxyradical is formed, however, its fate is unclear, the surface-confined electroactive quinones are not by-products. Both analytes, PCP and 3-CP or their products adsorb on the BDD surface more strongly than do the other examined CPs and limited potential cycling into the oxygen evolution region was necessary to remove all traces of reaction products. Specialized methods were developed for PCP, as this persistent pesticide is widely used for wood preservation and can be used as model compound for the development of new analytical techniques. Codognoto et. al [18] developed a SWV method for direct determination of PCP in natural waters. When using BR buffer, pH 5.5 as the supporting electrolyte, PCP oxidation occurs at +0.80 V vs. Ag/AgCl in a two-electron process controlled by adsorption of the species. The SWV signal reproducibility (relative standard

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deviation (RSD) 1.3 % for c(PCP) = 5·10–5 mol L–1, n = 5) was achieved by rinsing of the electrode surface with water and polarization at –3.0 V for 30 s after each measurement. The detection limits obtained were 2.0·10–8 mol L–1 (5.5 μg L–1) in pure water and 5.6·10–8 mol L–1 (15.5 μg L–1) for contaminated river water which is just about the WHO guideline value for DW (9 μg L–1). However, extraction and preconcentration of PCP from this matrix followed by analogous determination at BDD electrode should be easily realizable. In the next study, the same authors focused on the determination of PCP in contaminated soil from a chemical plant [19]. The soil samples were extracted with hexane for three hours in Soxhlet apparatus. The organic extract was evaporated and reconstituted in 1.5 mL of acetonitrile, which was analyzed by SWV at BDD or Au microelectrode as well as by HPLC/UV or GC/MS for validation of the results and identification of other soil components. For SWV at BDD, 0.25 mL of the acetonitrile fraction was simply added to 20 mL of BR buffer, pH 5 and analyzed using standard addition method. In addition to the PCP, the peak of its degradation product o-tetrachlorobenzoquinone (o-chloranil) was identified. Recovering experiments for PCP quantitation in soil showed good agreement of results obtained by electroanalytical determinations (27.5 mg kg–1) and HPLC/UV analysis (26.8 mg kg–1). This application reveals that a simple voltammetric method can represent useful alternative to officially approved, but more expensive and time demanding method.

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2.2. Chemicals from Agricultural Activities According to WHO, chemicals from agricultural activities include chemicals used in agriculture and in animal husbandry. Most chemicals that may arise from agriculture are pesticides. Contamination of DW can result from their application and subsequent movement following rainfall or from inappropriate disposal method [106]. Beside DW, the content of pesticides in surface, ground and wastewater, soil, sludge, sediments, feed or foodstuffs may be under control of national health agencies. The few WHO referenced pesticides investigated so far on BDD electrodes include carbofuran, parathion and some of the chlorophenoxy pesticides (MCPA, MCPP, see Table 1). Carbofuran belongs to carbamate pesticides, the N-substituted esters of carbamic acid. Their general formula is R1NH-CO-OR2, where R2 is an aromatic or aliphatic moiety. Three main classes of carbamate pesticides are known: (i) Carbamate insecticides; R1 is a methyl group (carbaryl, bendiocarb, propoxur); (ii) Carbamate herbicides; R1 is an aromatic moiety (asulam, carbetamide, swep); (iii) Carbamate fungicides; R1 is a benzimidazole moiety (methyl benzimidazole carbamate, benomyl, carbendazim). The carbamates' toxicological properties, usage, fate in environment and analytical methods for their determination are summarized by WHO [107]. In general, the toxicity of carbamates for wildlife is low, but exceptions exist, e.g. for bees. Carbofuran is a N-methylcarbamate insecticide used to control insects in a wide variety of field crops, including potatoes, corn and soybeans. It has also contact activity against pests. It has one of the highest acute toxicities to humans of any insecticide widely used on field crops. Toxic

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effects are due to its activity as a cholinesterase inhibitor. Carbofuran is also known to be highly toxic to birds, therefore it was banned by US EPA in 1994 in its granular form, liquid solutions can still be used [108]. The analytical methods for carbamate pesticides determination rely more on HPLC, as especially N-methylcarbamates' polarity and thermal instability makes the use of GC problematic. HPLC methods with MS [109], UV, electrochemical or fluorescence detection [110-112] were described, the last one being recommended by WHO and EPA (Method 531.1) for determination in ground and drinking water after post-column N-methylcarbamates' derivatization (detection limit units of μg L–1, carbofuran 1.5 μg L–1) [113]. The electrochemical methods for determination of carbamate pesticides rely on the oxidation of amide nitrogen in the carbamate molecule or on determination of phenolic derivatives produced by alkaline hydrolysis of some carbamate pesticides [114-118]. Both approaches were used by Rao et al. [21] in their pioneering work dealing with voltammetric characterization and HPLC/ED determination of carbamate pesticides at BDD electrodes. CV experiments in 0.1 mol L–1 phosphate buffer, pH 7.2 were performed with carbaryl, an N-methylcarbamate similar to carbofuran. The relatively well defined CVs were obtained at BDD electrode due to direct oxidation of carbaryl observable at potentials higher than +1.2 V (vs. SCE). No electrode fouling was observed, when stirring the solution between each CV cycle. The signal-to background ratio was found to be 5-10 times higher than that of GCE, where in this potential region the oxygen evolution reactions coincide with carbaryl oxidation. After alkaline hydrolysis of carbaryl to 1-naphthol, its oxidation peak appeared at +0.7 V and fouling of the electrode was observed under repetitive cycling due to the formation of polynaphthol film. It seems that this film adsorbs more strongly at the BDD surface than polyphenol films, observed by the same authors when oxidizing CPs [17]. However, the electrode could be perfectly cleaned by treating it at the anodic potential of +2.5 V for 10 min. The amperometric determination of carbamate pesticides further reported in this study is described in the next chapter. Direct voltammetric determination of carbaryl in polluted urban creek with LOD of 9.3 μg L–1 (LOQ = 31.2 μg L–1) succeeded in a study by Codognoto [119]. Spiked samples were analyzed using SWV at Si/BDD without any previous extraction, clean-up or preconcentration steps that could impede in situ and real time determinations. Other pesticide investigated by means of BDD electrodes was parathion. It is an organophosphorous pesticide (OP), which inhibits cholinesterase like carbamates. Like many other OPs, parathion is a very potent insecticide and acaricide, generally applied by spraying, and often used on cotton, rice and fruit trees. As it is relatively persistent and one of the most dangerous pesticide with high acute toxicity to wildlife and humans and suspected carcinogenity, the search for safer and less toxic alternatives continues and numerous environmental organizations propose its general and global ban [120]. Since 2007, in the USA there are no longer any pesticide products registered containing parathion [121]. However, the chemical is still produced and used in other countries. As parathion disappears from surface water in about one week, the main source of exposure is generally food. Therefore its health-based guideline value was not established by WHO for DW, as parathion concentration are likely to be far under the hazardous level of 10 μg L–1 [122].

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The standard analytical methodology for the determination of parathion and its metabolites (including 4-nitrophenol (4-NP)) in environmental samples and foodstuffs is GC. The detection methods most frequently used are coupled with electron capture, flame photometric (FPD) or flame ionization detector. HPLC coupled with UV, MS or electrochemical detection has also been widely applied [21, 110]. EPA recommends GC/FPD for OPs including parathions' determination after liquid-liquid extraction from municipal and industrial wastewater (detection limit parathion 0.012 μg L–1, Method 614 or 1657) [113]. Electroanalytical determinations of parathion are based on the reduction of the nitro group in its structure. SWV in BR buffer was selected for the determination of parathion in pure and polluted river waters at Si/BDD by Pedrosa et al. [25]. The reduction signals were analyzed and compared with those previously obtained using a hanging mercury drop electrode (HMDE). LOD and LOQ were calculated from the calibration curves for both BDD electrode and HMDE in deionized water (BDD 2.4 and 7.9 μg L–1, HMDE 3.9 and 12.8 μg L–1, respectively) showing only a slight improvement when using BDD electrode. However, if the application involves polluted river waters the improvement is accentuated due to the very low adsorption on BDD, which prevents the fouling of electrode surface by organic pollutants (LOD for BDD ~ 10 μg L–1). The BDD was also used as anode for electrochemical remediation of parathion contamination. In this case, electrolysis was carried out at highly positive potential (+3.0 V) and led to the electrochemical combustion of parathion to CO2 and H2O, as measured by the decrease of total organic carbon in the electrolyte. Further BDD voltammetric studies were devoted to nitrophenols, chemicals not mentioned among WHO DW pollutants. However, these compounds originating from oxidation of many pesticides or as reactants or intermediates in pesticide production are important soil, agricultural, industrial, and municipal water pollutants and many national health agencies set their own limits for them in these matrices. 4-NP (EPA general water limit 30 μg L–1 [123]), the main degradation product of parathion and other OPs [124-126] was determined at BDD electrodes using both the hydroxyl group oxidation and nitro group reduction. Using SWV in BR buffer, pH 6 as the electrolyte, the LOQs varied between 9.4 and 53.1 μg L–1 for the oxidation process and between 14.1 and 61.3 μg L–1 for the reduction process, depending on water sample (either deionized or river water), which was comparable to results achieved by SWV at HMDE [123, 127, 128]. Codognoto at al. [119] demonstrated that using SWV, the oxidation peak of 4-NP does not coincide with signals of some organophosphorus or carbamate pesticides. Lei and Zhao [129, 130] developed a CV and differential pulse voltammetric (DPV) method for the simultaneous direct determination of hydroquinone (HQ), phenol (Ph), and 4-NP in wastewater. These EPs gave consecutive oxidation peaks at the potentials of +0.76 V, +1.24 V and +1.52 V and detection limits of Ph, HQ and 4-NP were estimated to be 1.82 μmol L–1 (184 μg L–1), 1.67 μmol L–1 (171 μg L–1), and 1.44 μmol L–1 (200 μg L–1), respectively. The possibility of voltammetric oxidation at BDD was also reported for 2,4-dinitrophenol [56], the common starting material in the production of pesticides and dyes [124]. All mentioned methods are promising regarding the water analysis, as they are simple, convenient and inexpensive. The last group of pesticides so far investigated at BDD electrodes are chlorophenoxy herbicides. They are used extensively throughout the world for the control of broad-leaved annual and perennial weeds in a variety of agricultural crops. They are also used in brush

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control in non-agricultural areas, to control broad-leaved aquatic weeds, and as a pre-harvest treatment to reduce early drop in apple orchards. They are derived from CPs and these compounds are often their degradation products in environment. As a group, chlorophenoxy herbicides have been classified in Group 2B (possibly carcinogenic to humans) by IARC [131]. However, the available data from studies on exposed populations and animals do not permit assessment of the carcinogenic potential to humans of any specific chlorophenoxy herbicide. Therefore, DW guidelines for these compounds are based on a threshold approach for other toxic effects. WHO recommends guideline values for 4-chloro-2-methylphenoxyacetic acid (MCPA), 2-(2-methyl-4-chlorophenoxy)propionic acid (MCPP, mecoprop), 2,4-dichlorophenoxypropionic acid (dichlorprop, 2,4-DP), 4-(2,4-dichlorophenoxy)butyric acid (2,4-DB), 2,4,5-trichlorophenoxyacetic acid (2,4,5-T), 2,4,5-trichlorophenoxypropionic acid (2,4,5-TP, fenoprop, silvex), and 4-(2-methyl-4-chlorophenoxy)butyric acid (MCPB) in the range of 2 μg L–1 (MCPA) to 100 μg L–1 (2,4-DP) [132]. MCPP (guideline value 10 μg L–1) was reported to be the most often found herbicide in DW and can reach concentrations up to 0.4 mg L–1 in natural waters [22, 133, 134]. Common methods for the determination of chlorophenoxy herbicides in water include solvent extraction, separation by GC, gas-liquid chromatography, thin-layer chromatography, or HPLC, with ED or UV detection. Detection limits range from 1 μg L–1 to 1 mg L–1 [135]. Chemical derivatization of chlorophenoxy acids and salts is often necessary, as they are practically nonvolatile and too polar to chromatograph them directly. No study was published so far on quantitation of chlorophenoxy herbicides based on voltammetric or amperometric detection using BDD electrodes. Nevertheless, few studies have appeared concerning the electrochemical destruction at BDD electrodes (MCPA, MCPP [22]; 2,4-DP [47]; MCPA, 2,4-D, 2,4,5-T [48]; 2,4,5-T [49]) and basic voltammetric and chronoamperometric studies were performed by Boye et al. [22] for MCPA, MCPP and structurally related clofibric acid in 1 mol L–1 HClO4. They showed that these compounds display one diffusion controlled oxidation peak at about +1.6 V (vs. standard hydrogen electrode (SHE)). A fast electrode fouling was observed under repetitive cycling. However, the formed film can be easily removed by applying a positive potential (+2.6 V vs. SHE) in the region of water oxidation (generation of OH•). The film seems to be formed by polymerization of a phenoxy type radical and/or its products. This radical is formed in the first one-electron reaction step, leading to the formation of herbicides´ radical cation, and its consecutive hydrolysis to corresponding phenoxy radical and 2-hydroxycarboxylic acid. As mentioned above, the phenoxy radicals may be responsible for BDD passivation also when oxidizing chlorophenols, and the number of methods for their determination using BDD electrodes reveals that this should be no serious obstacle when developing similar applications for chlorophenoxy herbicides. Thus, other studies on the use of BDD electrodes for quantitation of chlorophenoxy herbicides can be expected in the near future.

2.3. Monocyclic and Polycyclic Aromatic Hydrocarbons Monocyclic and polycyclic aromatic hydrocarbons (PAHs) represent a serious potential contaminant of DW. The simple aromatics are usually used as solvents, chemical reactants or

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intermediates and for blending of petrol, so they enter the environment usually from industrial effluents or atmospheric pollution. The humans' exposure to them may be increased due to smoking or heavy traffic. For DW, WHO recommends guideline values for benzene, toluen, xylenes and styrene. The lowest limit (10 μg L–1) is set for benzene, a proved human carcinogen (IARC group 1, ref. [131]). The common concentrations of these compounds in drinking and surface water are a few μg L–1, higher concentrations were found due to point emissions [135]. PAHs enter the environment via the atmosphere from a variety of combustion processes and pyrolysis sources. In contrary to simple aromatics, they are owing to their low water solubility and high affinity for particulate matter usually not found in natural water in notable concentrations, their levels are usually in the range of 0-5 ng L–1. The main source of PAHs contamination in DW is usually the coal-tar coating of DW distribution pipes used to protect these from corrosion. In contaminated groundwater their concentration may exceed 10 μg L–1 [136]. This value also represents the limit for the sum of selected PAHs of some national agencies [137]. Beside the sum of PAHs, the limits for individual PAHs may be enforceable. WHO recommends the guideline values for benzo[a]pyrene (IARC group 1, ref. [131]), fluoranthene is evidenced, but the limit is not given as usually it occurs in DW at concentrations well below those at which toxic effects may occur. So far, there exist only a few reports on the use of BDD electrodes for oxidation of simple aromatics and PAHs. To oxidize these compounds, it seems that special treatment of the electrode surface is necessary to overcome its hydrophobic character. The anodic oxidation of BDD surface at high positive potentials is believed to chemically modify it by carbonyl and carboxyl groups [138]. Such BDD electrodes were used in reports on benzene [11] and selected PAHs [12] oxidation. The first study on benzene electrochemical oxidation was published by Oliveira et al. [11]. They used anodically oxidized electrodes (removal of the hydrophobic film) consecutively polarized at high negative potential (–3.0 V) for surface conditioning. This cathodic polarization enhances and stabilizes the electrochemical response, as shown previously on the example of PCP and 4-CP [13, 18, 139]. The CV studies indicate that benzene is irreversibly oxidized in acidic medium (0.5 mol L–1 H2SO4) on BDD at +2.0 V vs. Ag/AgCl in a diffusion controlled process, the benzene peak-height depends linearly on its concentration. During the cycling, other products are generated, and a pair of peaks was observed that can be associated with the redox processes of any of the following species: Hydroquinone, benzoquinone, resorcinol or catechol. Together with phenol, these compounds are the main products of benzene electrolysis at +2.5 V. At higher potentials, the total oxidation to CO2 was confirmed when stirring the electrolyzed solution. PAHs were firstly investigated recently by Bouvrette et al. [12], who employed electrochemically anodized BDD electrode for voltammetric characterization of sixteen priority PAHs according to US EPA. While about one half of the investigated PAHs (e.g. naphthalene, anthracene) showed one or two clear oxidation peaks in phosphoric acid / acetonitrile mixtures at potentials from +1.1 V to +1.8 V, the remaining half (including benzo[a]pyrene and fluoranthene) did not display any peak. Nevertheless, these PAHs exhibited an increasing signal / background ratio at potentials higher than ~ +1.1 V. Other experiments together with UV-VIS and fluorescence spectrometry proved that the oxidation of PAHs at BDD in mixed phosphoric acid / acetonitrile media, presumably initiated by

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formation of a radical cation, does not lead to products showing quinone structure. Pyrene as the only PAH was quantified in this study by hydrodynamic amperometry (detection potential +1.8 V vs. Ag/AgCl) at rotating BDD electrode with detection limit of 0.6 nmol L–1 (0.12 μg L–1). The results on HPLC/ED from this study are described in the next chapter. The last interesting example of the use of BDD in aromatic hydrocarbons determination is a BDD microelectronic gas sensor constructed by Gurbuz et al. [10], who proposed it to detect volatile organics, i.e. benzene and toluene in soil/subsoil and so locate potential oil reservoirs or environmental pollution. This seems usable with regard to DW protection, as these compounds may also penetrate from contaminated soil through plastic pipes into it. The sensor consists of three important layers – catalytic metal (Pd), intrinsic-diamond, and BDD layer and registries changes of transient characteristics after analytes adsorption at the Pd/i-diamond interface.

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3. BDD THIN FILM ELECTRODES AS SENSORS IN LIQUID FLOW METHODS There are two main reasons, why the electrochemical, preferably amperometric detection of organic pollutants in connection with liquid flow methods is preferred to batch voltammetric analysis: (i) The possible electrode passivation can be prevented as the flowing liquid possibly removes immediately from the electrode vicinity the electrode reaction products and intermediates creating the passivation films. (ii) Most of the methods (HPLC, CZE) enable the separation of analytes, which is of great concern in analysis of complex mixtures, such as PAHs, pesticides, and others. Moreover, BDD electrodes offer many advantages compared to other solid electrodes usually used in flow-through amperometric detection cells: (iii) Usually no mechanical or electrochemical pretreatment of BDD electrode is needed; (iv) The embarrassed creation of passivation films in flowing liquids mentioned under (i) seems to be even more difficult at BDD electrodes due to decreased proclivity of many reaction intermediates and products to the adsorption at their relatively hydrophobic (possibly H-terminated) surface. (v) The background current stabilizes within seconds to few minutes after detector turnon in contrast to solid, especially other carbon based electrodes, where it takes frequently about one hour to reach a constant current value. (vi) As result of (iii) and (v) shorter overall analysis time saving material and energetic costs can be achieved. Therefore, for many of the analytes discussed in previous chapter, a method for their amperometric detection after FIA or HPLC separation was proposed. For this purpose, either commercial flow-through cells are used, or the authors rely on laboratory-made devices. Less common is the CE/ED coupling, as this requires the technically exacting miniaturization of BDD electrodes and adaptation of the appropriate electrophoretic system. The overview of amperometric methods with BDD electrodes for WHO registered DW organic pollutants is given in Table 3 with characterization of used BDD electrodes, cells and methods, achieved LOQs and LDR. Referred publications are further discussed in the following chapter together

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with other applications of amperometry at BDD electrodes suitable for determination of other potential DW pollutants.

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3.1. Chlorinated Phenols The phenolic compounds including CPs represent an unceasing challenge for electroanalysis due to possible electrode passivation during electrooxidation, as highlighted in previous chapter. However, it seems not to be the case in liquid flow methods with amperometric detection at BDD based detectors – a number of methods for their determination without any electrode activation was reported (Table 3). Muna et al. focused after CV characterization of phenol and CPs (2-CP, 3-CP, 4-CP, PCP) [20] on their determination using FIA and HPLC coupled with a home-made thin layer cell with microcrystalline or nanocrystalline H-terminated BDD electrodes (however, the presence of oxygen containing groups is always questionable at this type of BDD electrodes when oxidizing them to positive potentials during experiments). Their performance for the amperometric detection of these contaminants in FIA/ED and HPLC/ED was evaluated in terms of achieved LDR, LOQs (Table 3), sensitivity, response precision, and response stability. Both diamond types yielded concentration LOQs from 0.05 to 0.6 μmol L–1 for all the phenolic compounds in FIA/ED and 0.1 μmol L–1 for all the compounds in HPLC/ED. The highest and the lowest LOQs and sensitivity were obtained for 2-CP and PCP, the WHO-referenced analytes. These achieved LOQs are comparable with those achieved by EPA Method 1625 [100]. Both electrode types also exhibited excellent response reproducibility (1.4-5.8 % for FIA/ED and 2.9-5.6 % for HPLC/ED for 20 consecutive measurements) and could be used from days to weeks in the measurement with only a periodic soak in distilled 2-propanol required to maintain optimum performance. Both types of diamond outperformed GCE, which exhibited short-lived responsiveness as a consequence of fouling by reaction products and potential-dependent changes in the electrodes' physicochemical properties. The use of the HPLC/ED assay for the determination of 2-CP in a contaminated soil sample was also demonstrated – 2-CP was detected among sixty other organic pollutants in a commercially supplied contaminated soil sample after their HPLC analysis (details in Table 3). The other study dealing with CPs was performed by Terashima et al. [17] at BDD electrodes intentionally oxidized by polarization at +2.64 V (vs. SCE) in pH 2 BR buffer. In voltammetric studies of selected phenols, these electrodes performed more reliable than "as-deposited" BDD (see previous chapter). When coupled with FIA anodized diamond exhibited excellent stability, with a response variability of 2.3 % (n = 100) for the oxidation of a high concentration (5 mmol L–1) of 2,4-DCP. In contrast, GCE exhibited a response variability of 39.1 %. After 100 injections, relative peak intensity for GCE decreased by drastic 70 %, while a slight decrease of 10 % was observed for BDD. Moreover, anodized BDD electrode could be reactivated in situ by applying a high positive potential +2.64 V. The LOQs obtained in the FIA mode for 2,4-DCP was found to be 20 nmol L–1 (S/N = 3), with a LDR up to 100 μmol L–1. By coupling with HPLC and column-switching technique, which enabled on-line preconcentration (50 times), the detection limit was lowered to 0.4 nmol L–1. Several other DCPs and TCPs including 2,4-DCP and 2,4,6-TCP were determined using this

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technique with LOQs ranging from 0.038 μg L–1 (0.23 nmol L–1) for 2,6-DCP to 0.361 μg L–1 (2.23 nmol L–1) for 3,5-DCP. Further, drain water that was condensed from the flue gas of waste incinerators was directly injected into the HPLC/ED assay and eleven DCPs and TCPs were quantified at concentrations from 9.1 μg L–1 (3,5-DCP) to 85.1 μg L–1 (2,4,5-TCP). The HPLC assay presented in this study fulfills requirements not only for DW analysis, but even more complicated matrices such as industrial or municipal discharges, as it offers low LOQs and enables simple direct injection of the analyte. Prado et al. [23] used a hydrodynamic channel flow cell with a BDD electrode for the detection of CPs (2-CP, 4-CP, 2,4-DCP and 4-chloro-3-methylphenol) in aqueous solution of 0.1 mol L–1 HNO3. Laser ablation voltammetry using a 532 nm Nd:YAG (neodymium-doped yttrium aluminium garnet) laser produced no significant change in the signal of the analyte currents observed at irradiated and bare electrode and confirmed the lack of any significant electrode passivation for the concentrations studied (up to 50 μmol L–1 for 4-CP). The results showed similar sensitivities and linear ranges up to 20 μmol L–1 for all CPs studied separately. This enables to measure the total CPs concentration as in the measurements of their mixtures only a single peak was observed. No interferences were observed for nitrite or sulfite. Undoubtedly, all FIA/ED or HPLC/ED assays presented in these three studies represent a valuable contribution to the CPs detection. Achieved LOQs for 2-CP, 2,4-DCP and 2,4,6-TCP under optimized conditions are far lower than guideline values recommended by WHO or limits required by other agencies. Moreover, only in the HPLC/ED determination of Terashima et al. [17] a preconcentration step was employed; so this possibility remains open for the other methods, because DW is a relatively simple matrix. CPs were also among the few model analytes tested in electrophoretic methods. The first CE/ED setups with BDD microelectrodes were described in 2003 [16, 24, 28, 141]. Shin et al. [16] prepared a diamond microline electrode by sandwiching freestanding diamond CVD films between two glass slides with UV adhesive. In order to obtain a structure with dimensions appropriate for the inner diameter (i.d.) of the fused silica capillary, the cross section (50 x 300 to 500 μm) of the diamond thin film was exposed as an electrode surface area by polishing the glass-diamond-glass sandwich structure. The separation efficiency and analytical performance of such BDD microline electrode in end-column CE/ED was evaluated for the determination of a catecholamine mixture (dopamine, norepinephrine, epinephrine). The high number of theoretical plates (> 138 000), quite low LOQs (~ 20 nmol L–1 at S/N = 3) and satisfactory repeatability (RSD < 4.4 %) for all analytes encouraged the authors to use the setup for determination of phenol and CPs mixture (2-CP, 2,4-DCP, 2,3-DCP, 2,4,6-TCP). Compared to carbon fiber electrodes, suffering an increase of noise in the baseline current over the span of several repeated injections, "as deposited" BDD electrodes showed low, stable noise levels (~1-1.5 pA) and only slight decrease in the peak currents throughout repeated injections without any electrode pretreatment.

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Table 3. Overview of liquid flow methods coupled with amperometric detection at BDD electrodes for organic drinking-water (DW) pollutants according to WHO Guidelines [1] WHO referenced DW pollutant (+ other analytes)

2-CP 2,4-DCP 2,4,6-TCP PCP (3-CP, 4-CP, Ph) 2-CP 2,4-DCP (3-CP) 2-CP 2,4-DCP 2,4,6-TCP PCP (3-CP, 4-CP) 2-CP 2,4-DCP 2,4,6-TCP (Ph, 2,3-DCP)

LOQ (for S/N = 3) [μmol L–1], (μg L–1) for given matrix a [0.1], (13) [0.5], (82) [0.5], (99) [0.5], (133)

LDR [μmol L–1]

Analytical method and BDD electrode characterization

Injection, separation and detection conditions

Ref .

0.1-100 0.5-100 0.5-100 0.5-100

CZE/ED, end column direct amperometric detection at MPA CVD deposited BDD/Pt conical microelectrode, A = 2.2 (±0.7)·10-4 cm2.

Fused-silica capillary 27 μm i.d., length 76 cm, 0.01 mol L–1/0.02 mol L–1 mixed borate/phosphate, pH 8.4 run buffer, separation voltage 20 kV, Edet = +1.05 V, injection 8 kV for 3 s.

[14]

[30], (3860) [50], (8150) (S/N=6)

30-600 50-600

CZE/ED, end column indirect amperometric detection, the same BDD/Pt microelectrode as in [14].

[14]

[0.00016], (0.02) [0.00025], (0.04) [0.0010], (0.20) [0.00019], (0.05) for river water not given

0.00016-0.78 0.00025-0.80 0.0010-0.76 0.00019-0.76 for river water

CZE/ED end column direct amperometric detection after off-line SPE, BDD/Pt microelectrode as in [14].

Fused-silica capillary 30 μm i.d., length 76 cm, 0.8 mmol L–1 ferrocene carboxylic acid in 0.01 mol L–1 phosphate run buffer, pH 8.1, separation voltage 20 kV, Edet = +0.45 V, injection 8 kV for 3 s. Capillary + run buffer as in [14], injection 10 kV for 3 s. Off line SPE: ENVI-Chrom P (250 mg, Supelco) cartridges, spiked river water samples acidified to pH 2-3 for loading, elution 2 mL of ethyl acetate. Preconcentration factor 250:1.

not given

CZE/ED, end column amperometric detection at MPA CVD deposited BDD microline electrode, A = 50 x ~300-500 μm.

Fused-silica capillary 25 μm i.d., length 75 cm, 10 mmol L–1/10 mmol L–1 mixed borate/phosphate, pH 7.8 run buffer, separation voltage 25 kV, Edet = +1.0 V, injection 7 kV for 10 s.

[16]

[15]

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Table 3. Continued 2-CP 2,4-DCP (Ph)

not given

not given

Electrophoretic microchip, amperometric detection at BDD film band electrode, A = 0.3 x 6 mm.

2,4-DCP

[0.02], (3.26)

0.02-100

2,4-DCP 2,4,6-TCP (and other 5 DCPs and 4 TCPs)

[0.0004], (0.065) [0.0005], (0.102) - for drain water no LOQ given [0.05], (6.43) c,d [0.6], (159) c,d

not given

FIA/ED, amperometry, BDD anodically oxidized (+2.64 V for 4 min in BR buffer, pH 2) in thin layer cell (GL Sciences) or wall jet arrangement, A = 0.4 cm2. HPLC/ED, amperometry, BDD anodically oxidized (+2.64 V for 4 min in BR buffer, pH 2) in thin layer cell (ED623, GL Sciences).

0.05-200 c,d 0.6-1200 c,d

FIA/ED

2-CP PCP (Ph, 3-CP)

[0.1], (12.9) c,e [0.1], (26.6) c,e

0.1-60 c,e 0.1-80 c,e

HPLC/ED both methods amperometry at MPA CVD deposited microcrystalline and nanocrystalline BDD, home-made thin layer cell [140].

2-CP 2,4-DCP (Ph, 4-CP, 4-C-3-MP)

[0.01], (1.3) [0.01], (1.6)

0.01-10 0.01-20

LSV and ChA, home-made channel flow cell (4 x 0,7 x 0.04 cm) with commercial BDD film (DeBeers Industrial Diamond Division, 5 x 5 x 0.535 mm), either bare or irradiated for 30 s with 532 nm Nd:YAG laser at 1.6 W cm–2.

2-CP PCP

Separation medium 0.01 mol L–1/0.02 mol L–1 mixed borate/phosphate buffer, pH 8.0, separation voltage +1.5 V, Edet = +0.95 V, injection 1.5 V for 2 s. MP: 60% methanol / 0.5% phosphoric acid, Fm = 0.2 ml min–1, Edet = +1.2 V, injection volume 10 μl.

[24]

Separation column: Inertsil ODS3 (150 mm x 2.1 mm i.d.). Clean-up column: (50 mm x 2.1 mm i.d.), 35°C, column switching technique for preconcentration (50 x), MP, Fm and Edet as for FIA/ED, injection volume 500 μl.

[17]

FIA/ED: MP: 0.05 mol L–1 phosphate buffer, pH 3.5, injection volume 20 μl. HPLC/ED: C18 RP phase ((X-Terra, Waters, 5 μm particle size), MP: two-step gradient elution – 0.05 mol L–1 phosphate buffer, pH 3.5/acetonitrile 65:35 (v/v) for 10 min, after change to 20:80 (v/v). FIA+HPLC/ED: injection volume 20 μl, Fm = 0.6 ml min–1, Edet = +1.1 V (+1.2 V) microcrystalline (nanocrystalline) BDD. 2-CP determination in soil: Ultrasonical extraction from soil to acetonitrile followed by SPE on C18 phase. LSV and ChA in hydrodynamic flow, MP: 0.1 mol L–1 HNO3, gravity controlled Fm = 4.0 ml min–1. ChA: Edet = +1.5 V. LSV: Scan rate 2 mV s–1.

[20]

[17]

[23]

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WHO referenced DW pollutant (+ other analytes)

Carbofuran (carbaryl, bendiocarb,dichloron, MBIC)

LOQ (for S/N = 3) [μmol L–1], (μg L–1)for given matrix a HPLC direct determination [0.06], (13.3) HPLC indirect determination [0.005], (1.1) (S/N = 2) For FIA not given

BaP, Fla (and other 14 PAHs)

[0.0123], (3.1) f [0.0141], (2.8) f

LDR [μmol L–1]

Analytical method and BDD electrode characterization

Injection, separation and detection conditions

Ref .

FIA direct determination 0.1-100 For other methods not given

FIA/ED, HPLC/ED, direct determination and determination after alkali hydrolysis to phenols. MPA CVD deposited BDD, thin layer flow cell (BAS), wall jet cell, A = 0.64 cm2, cell usage not differenced.

[21]

0.085-50 0.075-10

HPLC/ED, home-made wall jet cell with commercial BDD (Windsor Scientific), 3 mm diameter after anodic treatment in phosphoric acid / acetonitrile at +2.5 V for 10 min.

Direct determination: FIA MP 0.1 mol L–1 phosphate buffer, pH 2.25; HPLC/ED: Inertsil CN-3 (150 mm x 4.6 mm i.d., 5 μm) column, MP 0.1 mol L–1 phosphate buffer, pH 2.25/acetonitrile (80%, 20%), Edet = +1.45 V. Indirect determination: HPLC: ODS column (not specified); HPLC+FIA: MP 0.01 mol L–1 NaClO4 in acetic acid / acetonitrile / water / (0.5 %, 40 %, 59.5 %), Edet = +0.9 V. For all methods: Injection volume 20 μl, Fm = 1 ml min–1. Separation column: C18 RP phase Supelcosil LC-PAH (50 mm x 4.6 mm i.d., 3 μm). MP: gradient elution 0.04 mol L–1 phosphoric acid / acetonitrile from 50:50 to 10:90 (v/v) in 10 min, after kept at 10:90. Edet = +1.9 V, injection volume 10 μl, Fm = 2 ml min–1.

[12]

CP – chlorphenol; DCP – dichlorphenol; TCP – trichlorophenol; 4-C-3-MP – 4-chloro-3methyfenol; Ph – phenol; MBIC – methyl benzimidazolecarbamate; BaP – benzo[a]pyrene; Fla – fluoranthene. a if no matrix given, listed limits of quantitation (LOQ) are for model experiments in deionized water; S/N – signal-to-noise ratio; LDR – linear dynamic range; c for both microcrystalline and nanocrystalline diamond; d FIA/ED; e HPLC/ED; f detection limit equal to kSBm-1, where k = 3, SB is the standard deviation of the blank and m is the slope of the calibration curve. HPLC/ED; FIA/ED; CZE/ED; CE/ED – high performance liquid chromatography, flow injection analysis, capillary zone electrophoresis, capillary electrophoresis with electrochemical detection; MPA CVD – microwave plasma-assisted chemical vapor deposition, Nd:YAG – neodymium-doped yttrium aluminium garnet; SPE solid phase extraction; BR buffer – Britton-Robinson buffer; A – geometric surface of the BDD electrode. MP – mobile phase; RP – reversed phase; i.d. – inner diameter; LSV – linear sweep voltammetry; ChA – chronoamperometry; Fm – flow rate; Edet – detection potential.

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Table 4. An overview of the use of BDD electrodes in electroanalysis of PAH derivatives and other environmental pollutants at MPA CVD deposited BDD electrodes (summary of results of UNESCO Laboratory of Environmental Electrochemistry) Electrode / Technique

Analyte

Supporting electrolyte or separation and detection conditions

LOQ LDR Ref. [mol L-1] [mol L-1]

Si/BDDnanocrystalline / DPV

3-nitrofluoranthene BR buffer pH 12.0 / MeOH (1:1)

2·10-81·106

3·10-8

[145 ]

Si/BDDnanocrystalline / DPV

3BR buffer pH 4.0 / MeOH (1:1) aminofluoranthene

2·10-71·105

2·10-7

[145 ]

Si/BDD -microcrystalline / HPLC/ED

3MP: MeOH / phosphate buffer pH 4 aminofluoranthene (9:1, v/v); Fm = 0.5 ml min-1, Edet = +1.0 V

2·10-81·104

5·10-8

[145 ]

1.2·10-

[146 ]

Si/BDD-microcrystalline / DPV 2-aminobiphenyl 3-aminobiphenyl 4-aminobipheny

BR buffer pH 7.0 BR buffer pH 8.0 BR buffer pH 9.0

1·10-71·105

7

1.3·10-

7

2.5·10-

7

Si/BDD -nanocrystalline / HPLC/ED

Si/BDD -microcrystalline / HPLC/ED

2-aminobiphenyl 3-aminobiphenyl 4-aminobiphenyl

MP: 0.01 mol L–1 acetate buffer, pH 5 / acetonitrile / MeOH (40/30/30; v/v/v); Fm = 1 mL min–1; Edet = +1.2 V

1-aminonaphthalene MP: MeOH / 0.01 M phosphate buffer, 2-aminobiphenyl pH 6 (3:7, v/v); Fm = 1.5 ml min-1, Edet = 1.0 V

(4100)·107 (2100)·107 (2100)·107 1·10-71·104 1·10-71·104

2.0·10-

7

3.2·10-

[146 ]

7

5.1·10-

7

1.3·10-

7

1.2·10-

7

[147 ]

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Table 4. Continued Electrode / Technique Pt/BDD microcrystalline / CZE/AD

Analyte 1-naphtol 2-naphtol

LOQ Ref. [mol L-1]

Supporting electrolyte or separation and detection conditions

LDR [mol L-1]

0.04 mol L–1 borate run buffer, pH 9.2; Edet = +0.9 V

9·10-8-4·10-5 9.0·109·10-8-4·10-5 8 9.5·10-

[148 ] [149 ]

2·10-7-2·10-4 3·10-7 2·10-7-3·10-4 3·10-7

[150 ]

2·10-7-1·10-5 7·10-6 a 3·10-7 b -7 -5 3·10 -1·10

[151 ]

2·10-6-4·10-5 6·10-6 a 4·10-7 b -7 -4 2·10 -1·10

[152 ]

8

Si/BDD - microcrystalline / HPLC/ED

1-aminopyrene 1-hydroxypyrene

Si/BDD -microcrystalline / DPV 2-methyl-4,6dinitrophenol

MP: MeOH / 0.02 mol L–1 acetate buffer, pH 4 (2:8, v/v); Fm = 0.8 ml min-1, Edet = +0.9 V BR buffer pH 8.0 a BR buffer pH 5.0 b

a

b

commercial BDD (Windsor Scientific), 3 mm diameter

2-nitrophenol

BR buffer pH 11.0 a BR buffer pH 4.0 b

a

b

a

For oxidative determination; b for reductive determination. Si/BDD – boron-doped diamond deposited on silica substrate; Pt/BDD boron-doped diamond deposited on platinum substrate; DPV – differential pulse voltammetry; CZE – capillary zone electrophoresis; HPLC/ED – high performance liquid chromatography with amperometric detection in home-made thin layer flow cell [140]; CZE/ED – capillary zone electrophoresis with amperometric detection in end column arrangement [141]; LDR – linear dynamic range; LOQ – limit of quantitation; MP – mobile phase; MeOH – methanole; Fm – flow rate of the mobile phase; Edet – detection potential.

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The Swains' group solved the problem of miniaturization by deposition of microcrystalline BDD on electrochemically sharpened platinum wires (76 μm diameter, Pt/BDD) [141]. The BDD-coated wires were then attached to copper wires and sealed in a polypropylene pipette tip (Fig. 2). Resulting electrode had conically-shaped microcylindrical geometry with a maximum diameter of 85 μm (cylindrical portion) and an overall length of 150-200 μm. The cone dimensions were determined from the SEM images. The surface area of such electrodes can be calculated from the limiting current of the sigmoidal curve of [Fe(CN)6]3–/4– in 1 mol L–1 KCl (scan rate 10 mV s–1) and are ~ (1-3)·10-4 cm2. The electrodes were placed in end-column arrangement in a detection cell fabricated from a glass vial. The preliminary tests of this CE/ED setup were performed using dopamine, catechol and ascorbic acid in 10 mmol L–1 phosphate buffer, pH 6.0 run buffer and 30 cm long fused-silica capillary with 75 μm i.d. The background current (~100 pA) and noise (~3 pA) were measured at various detection potentials and are quite stable. Limits of detection for dopamine and catechol are 78 nmol L–1 and 120 nmol L–1 [141]. The separation efficiency for the system is influenced by the dimensions of the electrode and the precision of the microelectrode fixation opposite the column end. The system was further used by Muna et al. [14] for the analysis of aqueous solutions of phenol and six CPs (2-CP, 3-CP, 4-CP, 2,4-DCP, 2,4,6-TCP, PCP) using direct and indirect amperometric detection. The separation of all analytes was accomplished in 10 mmol L–1/20 mmol L–1 mixed borate/phosphate, pH 8.4 run buffer (separation voltage 20 kV, fused-silica capillary 27 μm i.d., length 76 cm) in 15 min with separation efficiency greater than 100 000 plates/m for all analytes. No electrode activation was needed and no fouling was observed using direct amperometric detection at +1.05 V. The LDR reaches from LOQ (2-CP, 3-CP, and 4-CP 100 nmol L–1; phenol, 2,4-DCP, 2,4,6-TCP, and PCP 500 nmol L–1) to 100 μmol L–1. For indirect amperometric detection, ferrocene carboxylic acid was added to the run buffer (10 mmol L–1 phosphate, pH 8.1) as electrophore. Good detection figures of merit using separation voltage 20 kV, fused-silica capillary 30 μm i.d., length 76 cm and detection potential of +0.45 V were achieved for 2-CP, 3-CP, and 2,4-DCP, although the LDR was not as wide as in the direct mode and LOQs were higher, in the mid micromolar range. However, this approach promises the detection of electroinactive analytes. Further, Muna et al. focused on the detection of the same CPs after their off-line solid phase extraction (SPE) from model river water samples using the same system in direct amperometric detection mode [15]. With ENVI-Chrom P, a highly cross-linked styrene-divinylbenzene copolymere SPE cartridges, and elution by ethyl acetate, CPs extraction recoveries in the 95-100 % range with the RSDs of 1-4 % were achieved. Using a preconcentration factor of 250:1, the LOQs for all six analytes ranged from 0.02 to 0.2 μg L–1. In general, the response precisions of the Pt/BDD microelectrodes in Swains' group studies were ~ 5 % (RSD) for tested CPs. Other utilization of BDD microelectrode was developed by Wang et al. [24, 28]. They coupled in end column mode a BDD film band electrode (0.3 x 6.0 mm) and a thick screen-printed carbon electrode with the electrophoretic microchip and compared the results for several groups of analytes. The LOQs for CPs using the latter electrode were in the 1 to 2 μmol L–1 range, unfortunately, no detection figures of merit were reported for the BDD microelectrode. However, it provided higher sensitivity, lower noise, better resistance to fouling, sharper peaks and enhanced resolution than did a screen-printed carbon electrode not only for CPs, but also for nitroaromatic explosives and organophosphate nerve agents. The

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attractive features of this system were further applied for the separation and detection of dye-related amino-substituted aromatic compounds 4-aminophenol (4-AP), 1,2-phenylenediamine (1,2-PDA), 2-aminonaphthalene (2-AN), 2-chloroaniline (2-CA), and o-aminobenzoic acid (o-ABA) [142]. These highly toxic and suspected or proven carcinogens do not represent direct threat for DW, however, they may occur in waste waters and effluents with respect to their extended use in manufacturing of dyes, as additives to polymer and rubber compounds, or as intermediates in the manufacturing of industrial chemicals. The diamond detector displayed detection limits of 2.0 μmol L–1 and 1.3 μmol L–1 for 4-AP and 2-AN, respectively, and a wide linear response for these compounds over the 2-50 μmol L–1 range. The enhanced stability was demonstrated by RSD values of 1.4 % and 4.7 % for 100 μmol L–1 1,2-PDA and 200 μmol L–1 2-CA, respectively. Besides, the simultaneously observed current decrease was 2.4 % and 9.1 % for 1,2-PDA and 2-CA, respectively (compared to 21.8 % and 41.0 % at the screen-printed carbon electrode and 28.3 % and 34.1 % at GCE). The above mentioned favorable properties of the diamond electrodes indicate great promise for environmental applications in CE and other microchip devices. Ongoing research in this area expects more diverse analytical applications, including field-deployable diamond-based microchip detection systems for on-site environmental monitoring.

Figure 2. Diagram of the electrochemical detection cell for end-column detection in capillary zone electrophoresis (A) and detail of the working BDD/Pt microelectrode (B). 1 – working BDD/Pt microelectrode insulated in pipette tip, 2 – reference electrode, 3 – auxiliary Pt electrode, 4 – end portion of the separation capillary, 5 – bolts for adjusting the capillary position just opposite the working microelectrode. Reprinted with permission from [141], Cvacka, J.; Quaiserova, V.; Park, J.; Show, Y.; Muck, A.; Swain, G. M. Anal. Chem. 2003, 75, 2678-2687. © 2003 American Chemical Society.

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3.2. Chemicals from Agricultural Activities In contrast to batch voltammetric studies, the chemicals determined by amperometric methods in FIA or HPLC include so far only carbamate pesticides with carbofuran referenced by WHO. In the study of Rao [21], FIA and HPLC/ED at BDD electrodes were used for determination of selected carbamate pesticides (carbofuran, carbaryl, methyl-2-benzimidazolecarbamate, bendiocarb). As in CV studies, two kinds of detection methods were adopted. In the first method, a direct detection of underivatized pesticides was carried out at an operating potential of +1.45 V versus Ag/AgCl, which resulted in linear calibration plots for FIA/ED in the concentration range of 100 nmol L–1-100 μmol L–1 and the LOQs of 5-20 ng mL–1 for HPLC/ED. In the second method, the LOQs were improved by subjecting the pesticide samples to alkaline hydrolysis in a separate step prior to injection. The obtained phenolic derivatives oxidize at a relatively lower potential (+0.9 V), which increases the sensitivity drastically. This method gives the LOQs of 5 nmol L–1 (1 μg L–1) for carbofuran, 3 nmol L–1 (0.6 μg L–1) for carbaryl, and 10 nmol L–1 (2 μg L–1) for bendiocarb. The achieved LOQ of carbofuran is below its WHO guideline value (7 μg L–1) and is in the same range as reported by using validated HPLC/FD method. On-line reactivation of the diamond electrode surface was shown to be possible by an anodic treatment of the electrode at +3.0 V for 30 min in case of electrode fouling, which may occur after a prolonged use. Such a treatment damages GCE and metal electrodes, while the diamond electrode remains stable. These results suggest that the diamond electrode is superior to the other previously used electrodes such as GCE and Kelgraf type for highly sensitive and stable detection of carbamate pesticides.

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3.3. Monocyclic and Polycyclic Aromatic Hydrocarbons So far, the only study on amperometric determination of aromatic hydrocarbons was published by Bouvrette [12]. Anodized BDD in wall jet arrangement operating in amperometric mode at detection potential of +1.9 V proved satisfactory for detection of sixteen priority PAHs according to US EPA after their HPLC separation using gradient elution with 40 mmol L-1 phosphoric acid/acetonitrile mobile phase. The detection limits (tens of nmol L-1 ~ units of μg L-1) meet or exceed those specified in HPLC methods using UV or laser-induced fluorescence or the official US EPA Method 550 (liquid-liquid extraction and HPLC with coupled UV and fluorescence detection [143]), even for PAHs displaying no clear oxidation peak in CV experiments. This study undoubtedly represents a promise for the determination of PAHs not only in DW, but in more complicated matrices and is an excellent example of the extraordinary versatile use of BDD electrodes in electroanalysis, as PAHs are not determinable at other electrode materials.

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3.4. Other Organic Pollutants

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Not only the compounds listed in WHO Guidelines, but also other EPs can represent a serious danger for water sources. This is valid especially for genotoxic compounds, where the exposure limits are questionable, as sometimes a minimal amount of the compound can initiate processes leading to unavoidable health damage. In this chapter, a brief overview on the use of BDD for both voltammetric and amperometric determinations of other important EPs is given. Determination of oxidizable and reducible PAH derivates, particularly amino-, hydroxyand nitro- derivatives of PAHs at BDD electrodes is studied by our group (UNESCO Laboratory of Environmental Electrochemistry in Prague, Czech Republic) in cooperation with the group of G. M. Swain from Michigan State University, USA. Pollutants determined so far by voltammetric methods using different designs of BDD electrodes (Figure 3) or amperometric detection after HPLC separation include derivatives of naphthalene, fluoranthene, biphenyl and pyrene. Obtained results are summarized in Table 4.

Figure 3. The detailed scheme of laboratory made BDD electrodes used in UNESCO Laboratory of Environmental Electrochemistry in Prague: (A) Teflon body electrode: Electrode body made of Teflon (1), stainless steel (2), screw attachment (3), small metal spring (4), brassy sheet (5), Si/BDD thin-film electrode (6), Viton® gasket (7), and access for solution (8); (B) Glass cell with clamped BDD electrode: Glass cell (1), Viton® gasket (2), Si/BDD thin-film electrode (3), Cu current collecting plate (4), insulating pad (5), clamp (6).

In Figure 4, there is depicted a chromatogram of DW analysis on 2-, 3-, and 4-aminobiphenyl content recorded at a BDD electrode in thin layer flow-through cell (for cell design see ref. [140]). The LOQs as low as ~ 0.2 μmol L-1 were achieved not only for drinking, but also for river water. Other studies on dye-related aromatic amines were performed by Shin [142], who used thin-film BDD detector after microchip CE separation with detection limit of 1.3 μmol L-1 for 2-aminonaphthalene. Other studies concern the electrochemical behavior of derivatives of simple aromatic compounds. CV at Si/BDD was used to monitor the destruction of phenolic compounds after the oxidative electrolysis of olive mill wastewaters [70]. Different voltammetric behavior of

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aniline [58], benzoic and salicylic acids and phthalic anhydride [71], and of hydroquinone, resorcinol and catechol [72] was reported at Si/BDD. Nevertheless, bulk electrolysis of these pollutants at BDD electrode leads to their complete mineralization with exception of aniline with about 80 % conversion. Other study was devoted to the determination of sulfonamides residues in egg samples by HPLC with amperometric detection at Si/BDD [144].

I, nA

220 2-AB

3-AB 5

210

4

4-AB

3 2

200

1

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3

4

5

6

t, min

Figure 4. HPLC separation of the mixture of 2-aminobiphenyl (2-AB), 3-aminobiphenyl (3-AB), and 4-aminobiphenyl (4-AB) after direct injection of model DW samples with analytes concentration 2·10–7 mol L–1 (1), 4·10–7 mol L–1 (2), 6·10–7 mol L–1 (3), 8·10–7 mol L–1 (4), 10·10–7 mol L–1 (5). Mobile phase 0.01 mol L–1 acetate buffer, pH 5 / acetonitrile / methanol (40/30/30; v/v/v), flow rate = 1 mL min–1; detection potential +1.2 V vs. Ag/AgCl. ChiraDex® column with chemically bonded β-cyclodextrin (250 × 4 mm, 5 μm, Merck, Germany). Amperometric detection at microcrystalline BDD electrode in thin layer amperometric cell.

CONCLUSION Conductive diamond, in particular BDD thin films as an electrode and electrochemical sensor material has gained a lot of attention since its introduction in early nineties. Based on fundamental research, aimed at gaining and understanding of the outstanding electrochemical properties of this material, the application-oriented research progressed rapidly. Many analytical methods for the determination of organic and inorganic species in biological, environmental and pharmaceutical matrices have been published. BDD electrodes were soon commercialized, which induced further development in this area. In this review, the range of possible analytes was restricted to organic DW contaminants. Basic voltammetric studies were performed for a number of them, including chlorophenols, selected pesticides, chlorophenoxy herbicides, monocyclic and polycyclic aromatic hydrocarbons and their derivatives demonstrating the possibility of their oxidation / reduction at BDD thin films. Although only few voltammetric or liquid flow methods with amperometric detection at BDD electrodes so far were developed exactly for the determination of these analytes or group of

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analytes in drinking-water, under optimized conditions in pure solvents in many cases notable reproducibility, high sensitivity and linear dynamic range over three orders of magnitude were reported without any previous preconcentration of the analytes. The achieved LOQ usually meet or are bellow the guideline values set by WHO and enforceable or recommended concentration limits of other national health agencies. Moreover, in many cases they are comparable or overcome the detection limits of GC and HPLC methods usually recommended for the determination of drinking-water pollutants. The potential of the usage of BDD electrodes in analysis of drinking water is great even for pollutants passivating other solid electrode materials (phenolic compounds, aromatic amines, and others). Thus, it can be concluded that diamond electrodes have proven useful in overcoming the limitations of conventional carbon and other solid electrodes and more applications for analysis of different matrices, including drinking-water, may be expected in near future.

ACKNOWLEDGMENTS K. P. thanks to the Grant Agency of the Czech Republic (grant 203/07/P261), J.M. thanks to the Grant Agency of Charles University (grant 6104/2007/B-Ch/PrF), the project was further financially supported by the Czech Ministry of Education, Youth and Sports (projects LC 06035 a MSM 0021620857).

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[116] Krause, R. T. J. Chromatogr. 1988, 442, 333-343. [117] Diaz, T. G.; Guiberteau, A.; Salinas, F.; Ortiz, J. M. J. Liq. Chromatogr. Relat. Technol. 1996, 19, 2681-2690. [118] Guiberteau, A.; Diaz, T. G.; Salinas, F.; Ortiz, J. M. Anal. Chim. Acta 1995, 305, 219226. [119] Codognoto, L.; Tanimoto, S. T.; Pedrosa, V. A.; Suffredini, H. B.; Machado, S. A. S.; Avaca, L. A. Electroanalysis 2006, 18, 253-258. [120] Kegley, S.; Hill, B.; Orme, S., PAN Pesticide Database; Pesticide Action Network: 2007, http://www.pesticideinfo.org, 7. 2. 2008. [121] US EPA21, http://www.epa.gov/EPA-PEST/2007/November/Day-16/p22374.htm, 31. 12. 2007. [122] WHO, Parathion. In Guidelines for Drinking-water Quality: First addendum to Third Edition; WHO, 2006; Vol. 1 – Recommendations, pp 421-422. [123] Pedrosa, V. A.; Codognoto, L.; Machado, S. A. S.; Avaca, L. A. J. Electroanal. Chem. 2004, 573, 11-18. [124] US EPA, Ambient Water Quality Criteria for Nitrophenols; US EPA: 1980, http://www.epa.gov/waterscience/criteria/library/ambientwqc/nitrophenols80.pdf, 7. 2. 2008. [125] Castillo, M.; Domingues, R.; Alpendurada, M. F.; Barcelo, D. Anal. Chim. Acta 1997, 353, 133-142. [126] Dzyadevych, S. V.; Chovelon, J. M. Mater. Sci. Eng., C 2002, 21, 55-60. [127] Pedrosa, V. A.; Suffredini, H. B.; Codognoto, L.; Tanimoto, S. T.; Machado, S. A. S.; Avaca, L. A. Anal. Lett. 2005, 38, 1115-1125. [128] Pedrosa, V. D.; Codognoto, L.; Avaca, L. A. J. Braz. Chem. Soc. 2003, 14, 530-535. [129] Zhao, G. H.; Tang, Y. T.; Liu, M. C.; Lei, Y. Z.; Xiao, X. E. Chin. J. Chem. 2007, 25, 1445-1450. [130] Lei, Y. Z.; Zhao, G. H.; Liu, M. C.; Xiao, X.; Tang, Y. T.; Li, D. M. Electroanalysis 2007, 19, 1933-1938. [131] IARC, IARC Monographs on the Evaluation of Carcinogenic Risks to Humans. Overall Evaluations of Carcinogenicity: An Updating of IARC Monographs Volumes 1 to 42. Supplement 7; IARC: Geneva, 1987, http://monographs.iarc.fr/ENG/Monographs /suppl7/suppl7.pdf, 7. 2. 2008. [132] WHO, Chlorophenoxy Herbicides (Excluding 2,4-D and MCPA) in Drinking-water: Background Document for Development of WHO Guidelines for Drinking-water Quality; WHO: Geneva, 2003, http://www.who.int/water_sanitation_health/dwq/ chemicals/chlorophenoxyherb.pdf. [133] Reffstrup, T. K.; Sorensen, H.; Helweg, A. Pestic. Sci. 1998, 52, 126-132. [134] Albrechtsen, H. J.; Mills, M. S.; Aamand, J.; Bjerg, P. L. Pest Manage. Sci. 2001, 57, 341-350. [135] WHO, Chemical Fact Sheets. In Guidelines for Drinking-water Quality: First addendum to Third Edition; WHO, 2006; Vol. 1 – Recommendations, pp 296-460. [136] WHO, Polynuclear Aromatic Hydrocarbons. In Guidelines for Drinking-water Quality: First addendum to Third Edition; WHO, 2006; Vol. 1 – Recommendations, pp 428-430. [137] The Council of the European Union, Off. J. Europ. Commun. 1998, 32-54.

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[138] Notsu, H.; Tatsuma, T.; Fujishima, A., Characterization of Oxygenated Diamond Electrodes. In Diamond Electrochemistry; Fujishima, A., Einaga, Y., Rao, T. N., Tryk, D. A., Eds.; Elsevier: Amsterdam, 2005, pp 11-25. [139] Suffredini, H. B.; Pedrosa, V. A.; Codognoto, L.; Machado, S. A. S.; Rocha-Filho, R. C.; Avaca, L. A. Electrochim. Acta 2004, 49, 4021-4026. [140] Jolley, S.; Koppang, M.; Jackson, T.; Swain, G. M. Anal. Chem. 1997, 69, 4099-4107. [141] Cvacka, J.; Quaiserova, V.; Park, J.; Show, Y.; Muck, A.; Swain, G. M. Anal. Chem. 2003, 75, 2678-2687. [142] Shin, D. C.; Tryk, D. A.; Fujishima, A.; Muck, A.; Chen, G.; Wang, J. Electrophoresis 2004, 25, 3017-3023. [143] US EPA, http://www.epa.gov/safewater/methods/epachem.html, 4. 2. 2008. [144] Preechaworapun, A.; Chuanuwatanakul, S.; Einaga, Y.; Grudpan, K.; Motomizu, S.; Chailapakul, O. Talanta 2006, 68, 1726-1731. [145] Čížek, K. Ph.D. Thesis; Charles University, Prague, 2006. [146] Jandová, K. M.Sc. Thesis; Charles University, Prague, 2007. [147] Cvačka, J.; Swain, G. M.; Barek, J.; Zima, J. Chem. Listy 2002, 96, 33-38. [148] Mocko, V. Ph.D. Thesis; Charles University, Prague, 2004. [149] Peckova, K.; Mocko, V.; Opekar, F.; Swain, G. M.; Zima, J.; Barek, J. Chem. Listy 2006, 100, 124-132. [150] Muck, A. Ph.D. Thesis; Charles University, Prague, 2002. [151] Urbanová, M. Bc. Thesis; Charles University, Prague, 2007. [152] Musilova, J.; Barek, J.; Drasar, P.; Peckova, K., Book of proceedings - Modern Analytical Chemistry: 4th International Student Conference; Prague, 28. - 29. 1. 2008; Charles University in Prague: 2008, pp147-152.

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In: Progress on Drinking Water Research Editors: M. H. Lefebvre and M. M. Roux

ISBN: 978-1-60456-748-9 © 2008 Nova Science Publishers, Inc.

Chapter 4

AMALGAM ELECTRODES AS SENSORS IN THE ANALYSIS OF AQUATIC SYSTEMS B. Yosypchuk∗a, T. Navratila, J. Barekb, K. Peckovab and J. Fischer b a

J. Heyrovský Institute of Physical Chemistry of AS CR, v.v.i., Dolejškova 3, CZ-182 23 Prague 8, Czech Republic b Charles University in Prague, Faculty of Science, Department of Analytical Chemistry, UNESCO Laboratory of Environmental Electrochemistry, Albertov 6, CZ-128 43 Prague 2, Czech Republic

ABSTRACT Copyright © 2008. Nova Science Publishers, Incorporated. All rights reserved.

This chapter will deal with recent results regarding voltammetric and amperometric determinations of micromolar and submicromolar concentrations of various environmentally important biologically active inorganic and organic substances in drinking water using non-traditional types of electrodes based on solid and paste amalgams, which can be environmentally friendly alternatives to mercury electrodes. Attention will be paid to polished and mercury meniscus modified solid amalgam electrodes (SAE), to silver solid amalgam composite electrodes and to amalgam paste electrodes either in batch analysis or in flow liquid systems (especially HPLC or FIA with electrochemical detection). Both polished and mercury meniscus modified silver solid amalgam electrodes (AgSAE) can be used in voltammetric and amperometric analysis of drinking water as an alternative to a hanging mercury drop electrode (HMDE). Other solid amalgam electrodes (e.g., CuSAE, AuSAE) are convenient for specific purposes, where properties of the metal, of which the solid amalgam consists, are employed. Silver solid amalgam composite electrodes, prepared from silver amalgam powder and epoxy resin, proved experimentally to be a suitable, reliable and environmentally friendly substitute for the hanging mercury drop electrode in electrochemical analyses. Amalgam paste electrodes based either on powdered amalgam mixed with a suitable organic pasting liquid or on amalgam paste can be used as nontoxic sensors with easily renewable surface and as a substitute of disposable electrodes. ∗

E-mail: [email protected]

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B. Yosypchuk, T. Navratil, J. Barek et al. The review will concentrate on our own results in the context of the general development in the field.

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1. INTRODUCTION Various modifications of liquid mercury electrodes and electrodes from solid and paste materials are important electrochemical sensors applied, e.g., in voltammetric, potentiometric, coulometric and conductometric detections. The working electrode (WE) does not allow to gain the required information alone; the reference electrode is necessary, as well as auxiliary electrode, eventually. Because the reference and auxiliary electrodes do not influence the WE signal essentially, the working electrode will be taken as sensor in this text. Polarizable electrode in polarography and in methods derived from it (voltammetry, chronopotentiometry, etc.) represents the source of electrochemical signal. From the point of view of electrode surface and double layer renovation, dropping mercury electrode (DME), which was introduced for polarographic methods by J. Heyrovský, seems to be ideal working electrode [1, 2]. Relatively high limit of detection achieved using DME (10–5 – 10–6 mol.L–1) on one hand and limited possibilities of the applications in positive potential area (due to mercury dissolution) on the other hand, accelerated the necessity of stationary electrodes development. Introduction of the hanging mercury drop electrode (HMDE) as well as modern sensitive techniques (e.g., stripping voltammetry and chronopotentiommetry) enabled to decrease the limit of determination down to 10–10 – 10–12 mol.L–1. The work in positive potential area is possible with non mercury electrodes generally, most often constructed from various types of carbon [3-5] and from noble metals [6]. The most important complications of all non-mercuric electrodes are low hydrogen overpotential and insufficient regeneration of the electrode surface, often realized by periodical polishing. The electrodes from compact metals covered by a mercury layer, mostly Ag/Hg, Pt/Hg, Au/Hg, Ir/Hg, Cu/Hg, represent a certain interlink between solid and pure mercury electrodes. This type of electrodes exhibits high hydrogen overpotential and enables as convenient work as with solid electrodes. Regeneration of their surface is mostly realized in an electrochemical way. On the other hand, an analysis can be complicated by the formation of intermetallic compounds, by instability of mercuric coverage due to gradual dissolution of a metallic substrate in mercury, accompanied by formation of a solid amalgam, by technical problems connected with metals fixation in an electrode body (with exception of platinum, other metals can be hardly sealed in glass), etc. Development of new solid electrode materials is the result of compromise between achieving the desired characteristics and keeping the measurement reproducibility on a required level. Modern trends in analytical applications (flow analysis, automatic systems, application of monitoring devices, electrochemical detection in HPLC, etc.) (e.g., [7]) require introduction of solid electrodes and practically exclude the possibility of mechanical polishing or chemical surface regeneration of the working electrode. Qualitatively close alternative to liquid mercury electrodes should be solid electrode with a wide interval of working potential, comparable with HMDE, which can be regenerated electrochemically with sufficient effectiveness. The use of modern, computer controlled devices, enable the realization of the automatic electrochemical regeneration before each measurement in a chosen way (e.g., [7-14]). A few years ago, solid amalgams [15-17], formed by amalgamation of fine powder of corresponding metal (MeSAE – where Me means Ag, Au, Ir, Cu, Bi, Cd,

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etc.) were introduced as an alternative material for electrode construction. These amalgams are analogous to the dental amalgams, which are, as it is well known, non toxic [18]. The amount of free liquid metallic mercury on their surface used for their modification is relatively small. The removal of mercury is complicated. Free mercury forms with metallic substrate nontoxic solid amalgam gradually. Some MeSAE have, as mentioned below, only somewhat narrower working potential window than HMDE, and thus they enable most of determinations realizable with mercury WE. In dependence on the ratio of mercury : metal, the amalgam can be liquid or solid. Liquid amalgam is used for filling of a dropping electrode or of the electrode with stationary drop. The solid amalgam electrodes can be prepared from solid amalgam. There is a relatively narrow interval of metal content in mercury, where this material has the paste form. This paste is pure metallic and does not contain any pasting liquid. The electrode on the basis of the paste amalgam connects advantages of metallic and paste electrodes. It is possible to sort the solid and paste amalgam electrodes (AE) according to the surface forms as follows: polished – liquid mercury free solid amalgam electrode (p-MeSAE); mercury film modified – polished MeSAE covered by mercury film (MF-MeSAE); mercury meniscus modified– polished MeSAE covered by mercury meniscus (m-MeSAE); composite – based on a fine solid amalgam powder and a solid polymer (MeSACE); paste – type I – based on a fine solid amalgam powder and proper pasting liquid (MeSA-PE) paste – type II – based on paste amalgam (MeA-PE) The above mentioned AEs and possible other types of AEs can be divided into two main groups:

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1. 2.

Amalgam form metal (or metals in the case of multi component amalgam) which is electrochemically less active than mercury (e.g., AgSAE, AgA-PE, AuSAE, IrSAE); Amalgam form metal (or metals in the case of multi component amalgam) which is electrochemically more active than mercury (e.g., CuSAE, BiSAE, BiAgSAE, CdAgSAE).

The working electrodes of the first group are more similar to HMDE according to their nature. This is confirmed by values of working potentials (see Table I) which are similar to those achieved at HMDE. The values of the background currents of various AEs and HMDE differ only slightly. The small differences are caused by unequal geometry of the active surface of WEs. The working electrodes from amalgams containing metal which is electrochemically more active than mercury can be applied for specific purposes, where the interaction of the analyte with this metal is used (e.g., interaction of adenine and cysteine with copper which is dissolved in meniscus of m-CuSAE). MeSAE, prepared as described in paragraph 2.1., can be used directly after surface polishing (p-MeSAE). The potential window of p-AgSAE is for the electrodes, which do not contain any liquid mercury extremely wide, often comparable with HMDE (see Table I).

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Table I. Range of working potentials (in V) of different electrodes in selected supporting electrolytes. Experimental results obtained by DCV using saturated calomel reference electrode; scan rate 0.02 V s–1; the potential limits correspond to 1 µA current level for the electrodes with diameter less than 1 mm and 20 µA for the larger electrodes; air oxygen was removed by purging with nitrogen

Electrode (disc diameter, mm; electrode surface, mm2) HMDE (1.01 mm2) PtE (0.40; 0.13) AuE (0.40; 0.13) AgE (0.40; 0.13) CuE (0.60; 0,28) p-AgSAE (0.70; 0.23) m-AgSAE (0.54; 0.46) m-AuSAE (0.40; 0.25) m-IrSAE (0.67; 0.71) m-CdAgSAE (0.53; 0.41) m-CuSAE (0.48; 0.36) m-BiAgSAE (0.70; 0.77) AgSA-CE (2.9 ) AgSA-PE (3.0) AgA-PE (0.50; 0.39) AgA-PE (2.0; 3.14)

0.1M HClO4

0.1M HCl

Potential range, V 0.2M acetate 0.05M Na2EDTA, 0.2M buffer, pH 4.8 acetate buffer, pH 4.8

–1.19 … +0.44 –0.32 … +1.37 –0.54 … +1.69 –0.64 … +0.39 –0.95 … +0.02 –1.12 … +0.45 –1.11 … +0.44 –1.12 … +0.45 –1.01 … +0.43 –1.07 … –0.66 –1.17 … +0.06 –0.97 … –0.02 –1.09 … +0.45 –1.04 … +0.36 –1.15 … +0.43 –1.24 … +0.48

–1.27 … +0.11 –0.30 … +1.11 –0.55 … +0.93 –0.81 … +0.08 –0.80 … –0.12 –1.12 … +0.11 –1.12 … +0.12 –1.11 … +0.12 –1.01 … +0.12 –1.06 … –0.69 –1.18 … –0.07 –0.96 … –0.09 –1.07 … +0.15 –0.90 … +0.21 –1.18 … +0.11 –1.24 … +0.15

–1.70 … +0.31 –0.51 … +1.28 –0.91 … +1.49 –0.99 … +0.36 –0.99 … –0.01 –1.51 … +0.31 –1.39 … +0.30 –1.47 … +0.31 –1.32 … +0.29 –1.29 … –0.68 –1.44 … –0.03 –1.25 … –0.18 –1.34 … +0.45 –1.15 … +0.38 –1.67 … +0.30 –1.47 … +0.36

–1.55 … +0.09 –0.51 … +1.05 –0.84 … +0.75 –0.99 … +0.35 –0.97 … –0.04 –1.45 … +0.11 –1.39 … +0.10 –1.45 … +0.11 –1.34 … +0.10 –1.29 … –0.73 –1.43 … –0.13 –1.29 … –0.17 –1.33 … +0.45 –1.09 … +0.36 –1.58 … +0.10 –1.46 … +0.15

0.05M Na2B4O7, pH 9.2

0.1M NaOH

–1.98 … +0.15 –0.77 … +1.16 –1.39 … +1.15 –1.20 … +0.38 –1.17 … +0.97 –1.88 … +0.16 –1.92 … +0.15 –1.90 … +0.16 –1.63 … +0.15 –1.93 … –0.69 –1.75 … +0.95 –1.84 … –0.29 –1.63 … +0.52 –1.50 … +0.24 –1.85 … +0.15 –1.90 … +0.21

–1.97 … –0.07 –0.96 … +0.85 –1.62 … +0.81 –1.41 … +0.19 –1.34 … –0.19 –1.96 … –0.06 –1.99 … –0.06 –1.91 … –0.05 –1.72 … –0.06 –1.93 … –0.84 –1.86 … –0.24 –1.82 … –0.48 –1.85 … +0.17 –1.45 … +0.30 –1.95 … –0.07 –1.83 … +0.01

Amalgam Electrodes As Sensors in the Analysis of Aquatic Systems

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This fact enables to apply p-AgSAE for determination of very electronegative compounds (Zn2+, Fe2+, Mn2+, IO3−, etc.), for investigation of electrochemical processes in area of very negative potentials (catalytic hydrogen evolution, adsorption-desorption of organic compounds, etc.), and in such cases when the use of liquid mercury containing electrodes is unsuitable. The solid surface of p-MeSAE can be modified in various ways, which substantially extends its applicability. Very important is the fact that it is possible to prepare relatively easy the electrodes from solid amalgams of various metals and subsequently to utilize specific interaction of these metals with investigated compound. From the analytical point of view, AEs modified by mercury proved to be the best, namely modified by mercury meniscus (m-MeSAE) or by mercury film (MF-MeSAE). Their liquid surface is ideally smooth and homogenous, which solves the principal problem of all solid electrodes – the mechanical regeneration of their surfaces and with this complication often connected insufficient reproducibility of repeated measurements. Electrochemical regeneration of the surface of meniscus or film electrode before each measurement can be carried out by means of a computer controlled analyzer (where it is possible to built-in this operation relatively easy into controlling software). It enables to reach RSD of repeated determinations below 2–3 %. This fact enables to utilize MeSAE in flow-through system, in automated system with autosampler, as amperometric detector in HPLC, etc.

2. PREPARATION AND PRETREATMENT OF AMALGAM ELECTRODES

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2.1. Preparation of Working Solid Amalgam Electrodes 2.1.1. Liquid Mercury Free Polished Solid Amalgam Electrode (p-MeSAE) A fine powder of the given metal (Ag, Cu, Au, etc.) is pressed into the bottom part of the tube 1 (see Figure 1) so that the height of the formed column 2 amounts to 0.5 – 1 cm. The platinum wire 3 (φ 0.1 mm) serving as a contact, is introduced into the upper part of the column. The bottom of the tube is dipped into a small bottle with 0.5 – 1 mL of waterless liquid mercury until the whole metal powder column will be wetted by mercury (usually just a few minutes). The mercury soaks in between the fine particles and forms within couple of hours a phase of compact solid amalgam. The bottle with mercury must be hermetically closed. The upper part of the platinum wire is connected with the electrical contact 4 and upper part of the tube is closed by a cover. The bottom of the tube is first brushed on soft emery paper followed by subsequent polishing with a wet powder of aluminium oxide (alumina). 2.1.2. Mercury Film Modified Solid Amalgam Electrode (MF-MeSAE) On the polished surface of the amalgam, it is possible to deposit electrolytically mercury, which adheres very well on the solid amalgam and forms the liquid film. One of the best electrolytes for these purposes is solution of 0.01M HgCl2 and 0.1M HCl and the optimum reduction potential Hg(II) is about –800 mV. The thickness of the film depends on the time of electrolysis (time of the film formation). E.g., in the case of accumulation time 300 s on p-AgSAE, disk diameter about 0.5 mm, the mercury film with thickness of about 1 µm is

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formed. MF-AgSAE with 1 µm mercury film assures the stable measurement for one day, and then the new film must be formed again.

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Figure 1. A. Schema of MeSAE: 1 – tube, 2 – solid amalgam, 3 – platinum wire, 4 – wire with connector. B. Multisensor with 8 working electrodes (m-AgSAE, MF-AgSAE, p-AgSAE, CuE, m-CuSAE, AuE, PtE, GCE) and with platinum auxiliary electrode (in centre). C. Working MeSAE. D. Magnified segment of a bottom part of m-MeSAE.

2.1.3. Mercury Meniscus Modified Solid Amalgam Electrode (m-MeSAE) Mercury meniscus is formed on the surface of the solid amalgam by dipping of the bottom part of the polished electrode into liquid mercury. Meniscus is stable during several months but due to the simplicity of amalgamation it is advisable to repeat this operation once a week.

2.2. Preparation of Composite Electrodes from Silver Solid Amalgam Powder The silver amalgam powder was prepared according to the procedure described in paragraph 2.3. An appropriate amount of solid amalgam powder was added to the epoxy resin and the mixture was thoroughly homogenized [19]. The composite mixture was pressed into the electrode body (in our case: 90 mm long glass tube of 2.9 mm inner diameter, the height of the column amounted to 8 – 10 mm). The electric contact of the electrode was realized by copper wire inserted into soft composite mixture. After solidification of the electrode material (1 day), the electrode surface was prepared for measurement: it was brushed by emery paper of various granulation and at the end it was gently polished by alumina paste of various granulation (up to 0.05 μm). Thus prepared and polished silver solid amalgam composite electrode (p-AgSA-CE) [19] can be eventually modified, (similarly as p-AgSAE) by various

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ways (e.g., by electrolytic deposition of the mercury film on the solid surface to prepare MF-AgSAE-CE or by formation of mercury meniscus by short dipping of the electrode in the liquid mercury to prepare m-AgSA-CE).

2.3. Preparation of Paste Electrodes from Silver Solid Amalgam Powder The silver amalgam was prepared by mixing mercury and silver powder (3:2, w/w) by 1 min shaking in dental amalgamator (Dentomat compact, Degussa, Brazil). The amalgam was comminuted to fine powder in agate dish and let to solidify for one day at room temperature. Afterwards, the formed amalgam was powdered once again to very fine form. The pastes were prepared by mixing suitable weight ratio of the amalgam powder and the pasting liquid (e.g., mineral oil, silicone oil, paraffin oil, tricresylphosphate, etc.). This ratio must be chosen to prepare paste of optimal consistency. The pastes were stuffed into the Teflon body of the electrode.

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2.4. Preparation of Paste Electrodes from Paste Silver Amalgam Mixture of mercury and 10 to 15% of fine silver powder (particle size 2 – 3.5 µm, Aldrich) was vigorously mixed for 60 s in dental amalgams preparation unit Dentomat compact (Degussa, Brazil). Teflon holder with movable piston, designed for carbon paste electrodes [20-24] was filled with paste amalgam. Before use, about 1 mm layer of paste was pressed out and discarded, and new surface was smoothed on a glass plate. In dependence on silver content, smoothed surface can resemble either mercury film electrode (10 – 12% Ag) or solid electrode (more than 12% Ag). In the first case, better reproducibility can be achieved. The surface of AgA-PE should be flat, glossy and it should be similar to the mercury film electrode. In Figure 2 the electrode from silver paste amalgam (A) and magnified segment of its working part (B) are depicted.

Figure 2. Working electrode from the paste amalgam (A) and increased segment of its active working surface (B; disk diameter of the paste amalgam is 2.0 mm). ). Reproduced with permission from Electroanalysis 2008, 20, 426–433.

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2.5. Pretreatment of Working Amalgam Electrodes Three operations are necessary to be performed for successful AEs application: 1.

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2. 3.

Amalgamation in the case of m-MeSAE, formation of the mercury film in the case of MF-MeSAE, polishing in the cases of p-MeSAE and composite electrode, removing of surface layer of the paste and smoothing of the surface for paste electrodes; Activation prior the set of measurrements; Regeneration before each measurement.

Amalgamation of m-MeSAE and MeSA-CE is to be realized once a week approximately, or even more frequently if necessary (e.g., in the case of sensitivity or reproducibility deterioration or if the liquid mercury meniscus is not distinct on the electrode surface). Into the bottle (volume 5 – 10 ml), 0.5 – 1 mL of metallic mercury and 2 – 5 ml of redistilled water are placed. Bottom part of AE is dipped into mercury and it is intensively stirred for 15 s. Then the bottle is hermetically closed and stored for future use (the mentioned amount of mercury is sufficient for repeated electrode amalgamation for many years). AE is rinsed by redistilled water and the presence of meniscus on the amalgam surface is optically verified, preferably using a magnifying glass. If the meniscus is not visible, the whole operation must be repeated. Formation of mercury film for MF-MeSAE and MeSA-CE is carried out electrochemically from solutions containing mercuric ions (e.g. 0.01 M HgCl2, 0.1 M HCl; Eacc = −800 mV, tacc = 300 s). Polishing of p-MeSAE and MeSA-CE is realized using a wet alumina powder with granulity 0.3 – 0.05 µm for 1 – 3 min. Activation of AE takes about 5 min and it is done at the beginning of the working day, after a pause in measurements longer than 1 hour, and after amalgamation, film formation or polishing. The oxides and the adsorbed compounds are removed from the surface of the AE during the activation (0.2 M KCl; Eactivation = −2 200 mV; tactivation = 300 s; oxygen is not removed from solution). This process improves the sensitivity and reproducibility of measurements. The electrode surface covered by mercury corresponds in fact to the saturated amalgam of the metal after a few minutes activation. Therefore, for instance m-CuSAE exhibits both the features of metallic copper electrode and of stationary drop electrode formed by liquid copper amalgam. Regeneration of AE takes about 30 s, it is carried out in analyzed solution before each measurement automatically (or according to the instructions of the controlling software); the sufficient reproducibility of results is thereby achieved (RSD is usually less than 2–3 %). The parameters of the regeneration of AE can be predefined in analyzer controlling software and their setup or change is realized using this software. For AE surface renovation it is usually sufficient to insert for 20 – 30 s the electrode potential 50 - 100 mV more positive than the hydrogen evaluation potential or decomposition potential of the supporting electrolyte. At this potential the oxides of amalgam forming metals are reduced (in the case of AgSAE these are Hg and Ag) and the adsorbed substances are removed. Simultaneously, the most of metals, present in analyzed solution, are accumulated on the surface. To prevent uncontrolled process of this accumulation, the jumps between negative and positive potentials, at which are the

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accumulated metals stripped, are imposed. The value of more positive regeneration potential should not be too high so that the electrode material could be dissolved or the supporting electrolyte decomposed. For instance, in the case of determination of Cu, Cd, Pb, Zn, and Tl, the process of m-AgSAE regeneration in 0.2 M acetate buffer (pH 4.6 – 5.0) consists in application of 50 polarization jumps, in which the potentials 0 mV and −1300 mV are applied alternately, each for 0.3 s.

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3. DETERMINATION OF INORGANIC COMPOUNDS USING AMALGAM ELECTRODES Cations determination. The amalgam electrodes can find very wide field of application for determination of cations, namely of all heavy metals. Determination of metallic ions is essential part of environment monitoring, tracing of harmful substances occurrence and biogenous elements in drinking water, beverages, body fluids, in control of technological processes, etc. Hanging mercury drop electrode, electrodes from noble metals, eventually modified by mercury film, by mercury meniscus or by biologically active substances, composite electrodes and electrodes from various types of carbon, including graphite paste electrodes, are most frequently used in voltammetry for these purposes. We have tested the possibility of determination of As(III), Cd(II), Co(II), Cr(III), Cu(II), Fe(III), In(III), Mn(II), Ni(II), Pb(II), Sn(II), Tl(I), Zn(II), etc. using various AEs (e.g., [16, 17, 25-28]). Anions determination. For determination realized with AEs it is possible to utilize various electrochemical processes, such as reduction (IO3−, Nb(V)), catalytic effects (NO3−), chemisorption in combination with follow up cathodic scan (Cl−, Br−, I−, S2−, CNS− etc.), or adsorption (Cr(VI)). Samples mineralization (digestion). Organic compounds, contained in natural and waste waters, in soil extracts etc., can principally affect the course of voltammetric measurement and the results of analysis. Mineralization (called digestion as well) of such samples is carried out using various ways, e.g., using various oxidizers. Drinking water should not contain such a high amount of organic compounds, which could affect the results of analysis substantially, and therefore it is not necessary to mineralize the samples at all. For digestion of water samples it is possible to recommend the following process: 1 mL of concentrated nitric acid and 1 mL of hydrogen peroxide solution (30 %) are added to 5 – – 50 mL of liquid sample in a flask or in a bottle. The mixture is then evaporated to the dryness. The residue is cooled down to the room temperature, 1 mL of concentrated nitric acid and 1 mL of hydrogen peroxide solution are added again and the solution is evaporated to the dryness again. If the residue is white or light yellow (e.g., iron compounds), the mineralization is finished. If the dry residue is brown, the addition and evaporation of oxidants must be repeated. In such dry form, it is possible to store and to transport the samples safely, if it is not possible to carry out the analysis in-situ or in the place of digestion. 10-25 ml of 0.1 M HCl is added to the dry residue, the solution is heated and stirred, until light steam is not observable; then the solution is cooled down, and, if necessary, it is filtrated. Thus prepared water sample can be used for voltammetric determinations.

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3.1. Voltammetric determination of Cu, Pb, Cd, Zn and Tl using silver solid amalgam electrode This determination is described in detail in [27]. Silver solid amalgam electrode enables, similarly as HMDE, to determine copper, lead, cadmium and zinc in one potential scan. Concentration of these ions in drinking water can vary from µg/L level for Cd and Tl to a few mg/L for Cu and Zn. Anodic stripping voltammetry is usually used for these determinations – the determined metals are electrolytically deposited (accumulated) on the working electrode surface in the first step and they are stripped out in the second step, which process is connected with registration of passing currents. The time of accumulation depends on analyte concentration. It usually varies from a few minutes (for μg.L–1 levels) to a few seconds (for mg.L–1 levels). If the concentration of determined ions in analyzed solution is about 0.5 mg.L–1 and higher, it is possible to use direct cathodic scan and to leave out the accumulation process. For analytical purposes differential pulse voltammetry is most frequently used, which exhibits the higher sensitivity and yields well developed and symmetric, easy evaluable peaks.

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Preparation of Samples for Voltammetric Measurements Simultaneous determination of Cu(II), Pb(II), Cd(II) and Zn(II). 1 – 6 mL of drinking water or mineralizate (i.e., of analyzed sample) – in dependence on supposed concentration of ions of determined metal – are placed into voltammetric cell, filled up to 6 mL with redistilled water, and 4 mL of 1 M acetate buffer, pH 4.8 – 5.0, are added. Determination of Tl(I). 1 – 5 mL of drinking water (i.e., of analyzed sample) – in dependence on supposed concentration of ions of determined metal – are placed into voltammetric cell, filled up to 5 mL with redistilled water and 4 mL of 1 M acetate buffer, pH 4.8 – 5.0, and 1.0 mL 0.1 M Na2EDTA are added. Voltammetric Measurements Parameters of simultaneous determination of Cu(II), Pb(II), Cd(II) and Zn(II): DPV; m-AgSAE; Ein = −1 300 mV; Efin = +50 mV; Eacc = −1 300 mV; tacc = 30-300 s in stirred solution; regeneration m-AgSAE: 50 polarization jumps, in which the potentials 0 mV and −1300 mV are applied alternately, each for 0.3 s, between working and reference electrode (automatically before each measurement); oxygen is removed by passage of an inert gas (pure nitrogen or argon) for 300 s through the sample; scan rate 20 mV.s−1; evaluation is realized using standard addition method. Parameters of determination of Tl(I): DPV; m-AgSAE; Ein = −800 mV; Efin = −300 mV; Eacc = −800 mV; tacc = 30 – 300 s in stirred solution; regeneration m-AgSAE: 50 polarization jumps, in which the potentials -50 mV and −1300 mV are applied alternately, each for 0.3 s, between working and reference electrode (automatically before each measurement); oxygen is removed by passage of an inert gas (pure nitrogen or argon) for 300 s through the sample; scan rate 20 mV.s−1; evaluation is realized using standard addition method. In Figure 3 voltammograms of mentioned metals in various concentrations and their calibration curves (straight lines) are depicted. The results of statistical processing of 11 repeated measurements in model sample are given in Table II. Linear dependence of the peak

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.

heights (ip) on concentration of determined ions and results of statistical evaluation of repeated measurements prove the applicability of m-AgSAE for analytical purposes. It is known that it is possible to use an analog recorder for recording of simultaneous determination of some compound on the assumption that the concentrations of these compounds do not differ by more than one order (when computer controlled device is used, the possibilities are wider). Quite often this condition is not fulfilled, and therefore it is necessary to realize the analysis separately; it requires to setup correctly the optimum value of potential of accumulation (Eacc) and of final potential (Efin).

i, nA

i p, nA

Cu

Cu

1000 1000

Cd

Cd

500 500

Zn

Zn

Pb

Pb

0

0 -300

-800

E, mV

-1300

0

40

80

c, ppb

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Figure 3. Concentration dependences and the corresponding calibration graphs of Cu(II), Pb(II), Cd(II) and Zn(II) recorded with m-AgSAE [27]. Experimental conditions: DPV; supporting electrolyte 0.4 M acetate buffer (pH 4.8); Eacc = -1300 mV; tacc = 180 s in stirred solution; regeneration of m-AgSAE during 30 s automatically before each measurement; scan rate, v = 20 mV⋅s-1; concentration of cations: 0, 10, 20, 30, 40, 50, 60, 70, 80, 90 and 100 ppb. ip Cu = 11.47 c + 49.55, R = 0.9997; ip Pb = 3.83 c + 15.78, R = 0.9990; ip Cd = 8.44 c + 25.24, R = 0.9992; ip Zn = 3.76 c + 42.16, R = 0.9992. Reproduced with permission from Chem. Listy 2008, 96, 756-760.

Table II. Statistical evaluation of repeated measurements of Cu(II), Pb(II), Cd(II), Zn(II) and Tl(I) [27]. Experimental conditions for simultaneous determination of Cu(II), Pb(II), Cd(II) and Zn(II): DPV; supporting electrolyte 0.4 M acetate buffer, 0.06 M NaCl (pH 4.6); Eacc = -1300 mV; tacc = 300 s in stirred solution; regeneration of m-AgSAE for 30 s automatically before each measurement; scan rate, v = 20 mV⋅s-1; concentration of cations 20 ppb; N = 11. Experimental conditions for determination of Tl(I): DPV; supporting electrolyte 0.4 M acetate buffer, 0.05 M NaCl, 0.01 M Na2EDTA (pH 4.6); Eacc = -800 mV; tacc = 180 s in stirred solution; regeneration of m-AgSAE for 30 s automatically before each measurement; scan rate, v = 20 mV⋅s-1; concentration of Tl(I) 20 ppb; N = 11 Parameter Average current, [nA] Confidence interval, [nA] Standard deviation, SD, [nA] Relative standard deviation, RSD, [%] Detection limit; 3⋅SD, [μg.L-1]

Cu(II) 216.5 0.8 1.1

Pb(II) 77.3 0.6 0.9

Cd(II) 157.2 1.8 2.8

Zn(II) 127.5 1.3 1.9

Tl(I) 50.8 0.2 0.3

0.5

1.2

1.8

1.5

0.7

0.3

0.7

1.1

0.9

0.4

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These potentials amounted to −400 and +50 mV for Cu(II); −800 and −200 mV for Pb(II); −1 000 and −500 mV for Cd(II) and −1 300 and −700 mV for Zn(II). If the concentration of given ion in solution is ≥ 0.5 mg.L−1, it is possible to register the voltammetric scan from initial potential Ein (which is equal to above mentioned value of Efin) without application accumulation (tacc = 0 s).

3.2. Voltammetric Determination of Ni with Silver Solid Amalgam Electrode This determination was described in detail in [25]. Adsorption stripping voltammetry belongs to one of the most sensitive methods, enabling to determine the concentrations of, e.g., chromium, uranium, vanadium, aluminium on the level of 10-10 – 10-11 mol.L-1. Selection of the proper complexing compound can essentially affect the analysis selectivity. Determination of nickel in ammonium buffer in the presence of dimethylglyoxime represents a classical example of the adsorptive stripping voltammetry [29]. Nickel forms sufficiently selectively complex with dimethylglyoxime, which is adsorbed on the electrode surface. During the cathodic scan (towards more negative potentials) nickel is reduced in adsorbed complex and it yields a well developed peak. It is valid in nickel determination (similarly as in the case of other adsorptive determinations) that at high analyte concentrations, the obtained concentration dependence is not linear any more.

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Preparation of Solutions for Voltammetric Measurements 1 - 9 mL of drinking water or mineralizate (i.e., of analyzed sample) - in dependence on supposed nickel concentration - is placed into voltammetric cell, filled up to 9 mL with redistilled water and 1 mL of 1 M ammonium buffer, pH 9.8, and 0.1 ml 0.1 % dimethylglyoxime are added. Oxygen is removed by passage of an inert gas (pure nitrogen or argon) for 300 s through the sample. Voltammetric Measurements Parameters of determination of Ni(II): see legend for Figure 4. Concentration dependence and calibration curve are depicted in Figure 4. The best linearity was achieved (under used parameters) in nickel concentration interval from 10 to 100 μg.L-1. The low limits of detection and RSD (see Table III) enable reliable determination of current nickel concentration in drinking water. Even at lower nickel concentrations (1 10 μ.L-1) the registered current is well measurable, however, the linearity of concentration dependences is somewhat worse (R = 0.9928). Moreover, the peak height decreases gradually (probably due to the adsorption of nickel dimethylglyoximate on the surface of the electrochemical cell components). At nickel concentrations 100 μg.L-1 and higher it is possible to register second, more negatively situated peak (see Figure 4), corresponding to the nickel reduction from non adsorbed state (diffusion of the complex from the bulk of the solution to the WE is the controlling process in such a case). Due to the fact that it is not possible to secure the unchanged state of the solid electrode for long time, the current response of nickel on m-AgSAE can be changed. Therefore, it is advisable to use standard addition method for the result evaluations. Determination of nickel on the level of μg.L-1 can be successful only under

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condition that this determination is carried out after necessary reaction time after dimethylglyoxime addition into the analyzed solution. Table III. Statistical evaluation of repeated measurements of Ni(II), Fe(III), NO3- [25], IO3- [28]. Experimental conditions for determination of Ni(II): concentration of Ni(II) 10 μg.L-1; N = 11; other parameters see Figure 4. Experimental conditions for determination of Fe(III): concentration of Fe(III) 3 mg.L-1; N = 11; other parameters see Figure 5. Experimental conditions for determination of NO3-: concentration of NO310 mg.L-1; N = 11; other parameters see Figure 6. Experimental conditions for determination of IO3-: concentration of IO3- 9.9 mg.L-1; N = 11; other parameters see Figure 7 Parameter

Ni(II) [μg.L1 ]

Fe(III) [mg.L1 ]

NO3[mg.L1 ]

IO3m-AgSA E [mg.L-1]

Average current, [nA] Confidence interval, [nA] Standard deviation, SD, [nA] Relative standard deviation, RSD, [%] Detection limit; 3⋅SD

-11.4 0.2

-49.6 0.3

-274.8 7.4

-644.4 2.3

IO3p-AgSA E [mg.L1 ] -283.9 2.4

0.3

1.1

11.2

3.1

3.3

3.0

2.3

4.1

0.5

1.2

0.9

0.2

1.2

0.1

0.3

-100

-90

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i, nA

i p , nA -60

-60

-30 -20

0 -800

-900

-1000

E, mV

-1100

0

20

40

60

80

c, ppb

100

Figure 4. Concentration dependence and the corresponding calibration graph of Ni(II) recorded with m-AgSAE. Experimental conditions: DPV; supporting electrolyte 0.1 M [NH4Cl + NH3], 0.001 % dimethylglyoxime (pH 9.8); Eacc = -125 mV; tacc = 60 s in stirred solution; Ein = -800 mV; Efin = -1100 mV; Ereg = -1500 mV, regeneration of m-AgSAE for 40 s, automatically before each measurement; scan rate, v = 20 mV⋅s-1; concentration of Ni(II) 0, 10, 20, 30, 40, 50, 60, 70, 80, 90 and 100 μg.L-1. R = 0.9998; ip = -8.11 c [μ.L-1] – 1.20.

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3.3. Voltammetric Determination of Fe(III) Using Silver Solid Amalgam Electrode This determination was described in [25]. The well developed reduction wave corresponding to the process Fe(III) → Fe(II), recorded on mercury electrodes from triethanolamine complex in alkaline solutions or in the presence of ammonia, is mostly used for analytical purposes. Due to the fact that the accumulation of iron is not possible on the surface of the mercury working electrode, the sensitivity of such determinations is not too high. With respect to low iron toxicity, this process is sufficient for its determination at limiting value concentrations (according to the valid standards) in drinking water. Alkaline supporting electrolyte containing triethanolamine was chosen as the most suitable for iron determination with m-AgSAE. To eliminate the interfering effect of zinc and manganese, eventually, which can occur in waters in concentrations comparable with iron concentration, complexing agent Na2EDTA was added to this solution. Peak potentials of zinc and manganese in EDTA-complex were thus shifted to very negative values and these ions cannot affect the peak of Fe(III). Statistical treatment is given in Table III. The achieved acceptable repeatability and sufficient sensitivity confirm the possibility of the use of m-AgSAE for those analytical purposes. The concentration dependence of the heights of well developed peaks of iron on its concentration, depicted in Figure 5, exhibit linear course in wide concentration range of Fe(III).

-600 ip , nA

-700 i, nA

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Preparation of Solutions for Voltammetric Measurements 1 – 5 mL of drinking water or mineralizate (i.e., of analyzed sample) - in dependence on supposed Fe(III) concentration – are measured into voltammetric cell, filled up to 5 mL with redistilled water, 3 mL of 1M NaOH, 2 mL of triethanolamine and 1 mL 0.1M Na2EDTA are added.

-500

-400

-300 -200

-100 0

-900

-1200

E, mV

-1500

0

10

20

30

c, ppm

40

Figure 5. Concentration dependence and the corresponding calibration graph for the determination of Fe(III) ions with m-AgSAE. Experimental conditions: DPV; supporting electrolyte: 0.27M NaOH, 2.7 % triethanolamine, 0.009M Na2EDTA; Ein = -800 mV; Efin = -1500 mV; regeneration of m-AgSAE for 30 s automatically before each measurement; v = 20 mV.s-1; concentration of Fe(III): 0; 0.5; 1.0; 1.5; 2.0; 2.5; 3.0; 3.5; 4.0; 4.5; 5.0; 6.0; 7.0; 8.0; 9.0; 10; 15; 20; 25; 30; 35; 40 mg.L-1. R = 0.9993; ip = -14.22 c[mg.L-1] – 10.19.

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Voltammetric Measurements Parameters of Fe(III) determination: see legend for Figure 5; regeneration m-AgSAE: 50 polarization jumps, in which the potentials -200 and −1200 mV are applied alternately, each for 0.3 s, between working and reference electrode (automatically before each measurement); oxygen is removed by passage of an inert gas (pure nitrogen or argon) for 300 s through the sample; evaluation is realized using standard addition method.

3.4. Voltammetric Determination of NO3- Using Silver Solid Amalgam Electrode This determination was described in [25]. Since the nitrate ions do not reduce on mercury electrodes directly, the catalytic reoxidation in the presence of multivalent elements (La(III), Ce(III), Yb(III), Zr(IV), U(VI), etc.) is utilized. The cerium solution, which proved suitable with HMDE, was applied for testing of possibilities of nitrates determination with m-AgSAE. In Figure 6, the concentration dependence of NO3- ions recorded using m-AgSAE in range 1 – 56.6 mg.L-1 is depicted. The linearity of the current response in dependence on concentration seems very good (the value of the calculated correlation coefficient is near to 1) and it attests possible application of m-AgSAE for nitrates analysis. The right side of the apex of peak starts to increase substantially and the ip – c-dependence becomes non-linear at higher concentrations of NO3- ions (see the highest peaks in Figure 6). The linear part of the concentration dependence includes the limit values of nitrates concentration, which are defined in standards for different types of waters. Nevertheless, if the concentration of NO3ions exceeds the mentioned linear part, the analyzed sample can be easily diluted by distilled water. -1400 ip , nA

i, nA

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-1400

-900

-900

-400

-400

-800

-1200

E, mV

-1600

0

20

40

c, ppm

60

Figure 6. Concentration dependence and the corresponding calibration graph for the determination of nitrates with m-AgSAE. Experimental conditions: DPV; supporting electrolyte: 0.02 M CeCl3, pH 5.0; Ein = -800 mV; Efin = -1600 mV; Ereg = -1600 mV, regeneration of m-AgSAE for 15 s automatically before each measurement; v = 20 mV.s-1; concentration of NO3-: 0; 1.0; 2.0; 3.0; 4.0; 5.0; 6.0; 7.0; 8.0; 9.0; 10; 19.6; 29.1; 38.5; 47.6; 56.6 mg.L-1. R = 0.9999; ip = -24.64 c [mg.L-1] + 0.24.

The current response of nitrates reactions (as well as of the most other catalytic processes) is connected with the state of the working electrode surface. It means that the

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effective and reliable electrode surface regeneration plays very important role for the reproducibility of analyses. In contrast to HMDE, where the drop is renewed before each measurement, each following measurement is affected by all previous ones due to the stationary surface in the case of solid electrodes. This can be confirmed by the value of relative standard deviation on the level of 4 % (see Table 3), which is substantially higher, than usually achieved 1 – 2 % with mercury electrodes for processes unaffected by a catalytic process. It is necessary to mention that the peak current irregularly increases for repeated measurements. Though the increase is not too dramatic, it is advisable to use the standard addition method for the evaluation of results. We were not successful in finding a simple and reliable way of electrochemical regeneration of m-AgSAE, which would assure the stable current response of nitrates for long time. Obviously, the insoluble compounds of cerium are accumulated on the surface of the electrode and thereby the catalytic activity of nitrates reduction is changed. Washing of the electrochemical cell and of the electrode system by 0.1 M HCl solution for 5 min. enables to restore the original behavior of the m-AgSAE.

Preparation of Solutions for Voltammetric Measurements 1 – 9.5 mL of drinking water or mineralizate (i.e., of analyzed sample) - in dependence on supposed nitrates concentration – are placed into voltammetric cell, filled up to 9.5 mL with redistilled water and 0.5 mL of 0.4M CeCl3 is added.

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Voltammetric Measurements Parameters of NO3- determination: see legend for Figure 6; evaluation is realized using standard addition method.

3.5. Voltammetric Determination of IO3- Using Silver Solid Amalgam Electrode This determination was described in detail in [28]. The direct reduction or the chemical reduction to iodides (e.g., by addition of hydrazine hydrochloride) with following application of voltammetric stripping methods or UV-spectroscopy is frequently used for determination of iodates. The iodates reduction is carried out at relatively highly negative potential area and therefore it was mostly carried out on classic mercury electrodes. It was proved that it is possible to carry out the iodates determination using p-AgSAE containing no liquid mercury (however, sensitivity and accuracy of analysis is in the case of m-AgSAE better). The suggested fast and very simple method offers a very suitable possibility of iodates determination in real samples. It was found by testing of various amalgam electrodes in various supporting electrolytes and at various pH, that AgSAE is the best electrode and 0.1 M NaOH is the best supporting electrolyte for the determination of iodates. The potential window of m-AgSAE, p-AgSAE and HMDE in such solution is almost identical (see Table I). The dependence of the iodates reduction peak height on their concentration, registered with m-AgSAE is depicted in Figure 7A. Very similar shape of voltammetric curves, with slightly different peak potential only,

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was recorded both with m-AgSAE and with p-AgSAE. These concentration dependences are depicted in Figure 7B. It is obvious that the response of m-AgSAE to the same iodates addition is substantially higher, than that of p-AgSAE. It is possible to note for both electrodes that the linearity of achieved calibration curves is excellent and herewith one of the most important conditions for application of this method in real analysis of iodates is fulfilled. The results of statistical evaluation of repeated determinations of the same sample solution if iodates on both electrodes are given in Table III. The best results of electrode renovation are achieved with m-AgSAE. -800

A

i, nA

-800

B

i p , nA

-600

-600 m-AgSAE

-400

-400

-200

-200 p-AgSAE

0

0

-1100

-1400

E, mV

-1700

0

4

8

c, ppm

12

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Figure 7. Concentration dependence for the determination of iodates with m-AgSAE(A) and calibration curves of iodates achieved with m-AgSAE a p-AgSAE (B). Experimental conditions: DPV; supporting electrolyte 0.1M NaOH; starting potential -900 mV; final potential -1600 mV; scan rate 20 mV.s-1; AgSAE regeneration for 30 s automatically before each measurement. Iodates concentration: 0 – 9.01 mg.L-1. ip m-AgSAE = -68.43 c – 11.37, R = 0.9997; ip p-AgSAE = -26.83 c + 1.67, R = 0.9995. Reproduced with permission from Electroanalysis 2002, 14, 1138-1142.

This fact is explainable by presence of ideal surface of liquid silver amalgam which forms meniscus. Meniscus of AgSAE does not require any chemical or mechanical step for its regeneration, and therefore it can be used for determinations realized in flow-through systems or in an automatic analyzer with autosampler as well. It is evident that the quality of the p-AgSAE surface, even in the case of ideal polishing, cannot be compared with that of liquid meniscus surface. Worse defined surface and chemical composition of solid amalgam, in comparison with saturated liquid amalgam, can explain worse sensitivity and repeatability of p-AgSAE. In contrast to m-AgSAE, p-AgSAE is to be polished periodically to remove the oxidation products of solid amalgams and undesired substances accumulated from analyzed solutions during preceding measurements. Despite of the fact that p-AgSAE yields somewhat worse analytical results in comparison with m-AgSAE, this electrode can be successfully applied for iodates determination, especially under conditions, when the presence of any amount of mercury is undesired.

Preparation of Solutions for Voltammetric Measurements 1 – 9 mL of drinking water or mineralizate (i.e., of analyzed sample) - in dependence on supposed IO3- concentration – are placed into voltammetric cell, filled up to 9.0 mL with redistilled water and 1 mL of 1.0M NaOH is added.

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Voltammetric Measurements Parameters of IO3- determination: see Figure 7 legend; AgSAE regeneration: 50 polarization jumps, in which the potentials -500 and −1500 mV are applied alternately, each for 0.3 s, between working and reference electrode (automatically before ach measurement); evaluation is realized using standard addition method.

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4. DETERMINATION OF ORGANIC COMPOUNDS USING SOLID AMALAGAM ELECTRODES Various modes of voltammetry (see Table IV) can be used for the determination of organic compounds in dependence on their electrochemical properties. Most of bellow mentioned compounds are reduced during their scan and exhibit linear dependences of the peak height on their concentration in ranges of 2 – 3 concentration orders. Limits of detection of these methods are about 10-6 mol.L-1. Therefore, it is possible to determine various biologically active organic compounds, pesticides, agrochemicals, drugs or derivatives of these substances with amalgam electrodes using their direct reduction (see Table IV). Usually some functional group is reduced (e.g., nitro group, azo group, single bond (e.g., C-Cl, C-S) or double, or triple bond). Therefore, the selectivity of such determinations is not too high. Selectivity and sensitivity of these measurements can be substantially improved by proper sample pretreatment, by combination of voltammetry with pre-separation using HPLC or by separation and preconcentration using solid phase extraction (SPE) [30]. SPE enables determination of nanomolar concentrations of tested substances. Amalgam electrodes, similarly as mercury electrode, do not possess great potentiality for oxidation processes due to dissolution of electrode material at relatively low positive potentials (see Table I). Suitable compound for such measurements is ascorbic acid in beverages and waters. In acetate buffer of pH 4.8, ascorbic acid DPV peak is at about +55 mV on AgA-PE, +67 mV on m-AgSAE, +76 mV on p-AgSAE, +47 mV on m-AuSAE, +85 mV on m-IrSAE and +5 mV on HMDE. On AgA-PE, the response was linear function of the concentration of ascorbic acid in region 1·10-5 – 9.1·10-4 mol.L-1 (R2 = 0.9999, ip = 576.65 c – 5.90). Appropriate repeatability was provided by electrochemical regeneration, no exchange of active surface was necessary. Cathodic stripping voltammetry (CSV) is a very sensitive method for the determination of organic compounds. CSV, applied in below mentioned analyses, is based on the principle that the material of the working electrode is dissolved during fixed period at defined potential at first, the resulting cations (Ag(I) and/or Hg(I) in case of AgSAE; Cu(I) in case of CuSAE, etc.) form the compounds with the determined analyte, which are adsorbed at the electrode. The reduction of such metal ions from adsorbed component is observable as a voltammetric peak. This procedure is applicable for indirect voltammetric analysis of such substances, which can react with metal cations, released from the working electrode (e.g., compounds containing –SH or –SS- groups). To the advantages of amalgam electrodes belongs the fact that it is possible to prepare the electrode material directly with the metal, which interacts with the analyzed compound and by this procedure to increase the selectivity of discussed electrochemical measurements. Since the analyte is accumulated at the working electrode in case of CSV, the limit of detection of thus realized determinations is commonly on the

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level of 10-8-10-9 mol.L-1 (Table IV). Some disadvantage of CSV consists in relatively narrow linear part of the dependence of the peak height on the concentration which represents usually one, rarely two concentration orders (for comparison, similar dependence in case of direct analyte reduction or in case of anodic stripping voltammetry of heavy metals is linear in range of 2–3 or even more concentration orders). Some substances are strongly adsorbed at the surface of the working electrodes and they can be reduced, oxidized or desorbed in such adsorbed state. Each of these processes is accompanied by the change of the registered signals. On favorable terms is the current change proportional to the amount of the analyte at the electrode surface. The adsorption at the surface of discussed amalgam electrodes is so much stable in case of some biological macromolecules that it is possible to talk about electrodes modified by corresponding macromolecule. The electrode with strongly adsorbed molecules can be rinsed and transferred into the cell with clean supporting electrolyte, which does not contain any pollutants, which could negatively affect the determination. Using proper electrochemical method, it is possible to record the electrochemical signal. This procedure is usually called adsorptive transfer stripping voltammetry (AdTSV) [31, 32]. One of the most pronounced advantages of the use of the solid amalgam electrodes in AdTSV consists in fact that in some special cases 1–2 μl of the sample is sufficient amount for analysis. This small volume of the substance is put on directly at the active working surface of the electrode by a micropipette (for the traditional analysis at least 1 ml of the sample is necessary) [33]. One of the ways to substantially increase the sensitivity of the measurement is the adsorption of a compound, which is catalytically active. The catalytic currents are several orders of magnitude higher than those controlled by diffusion, and it results in sensitivity increase. Table IV. The use of solid and paste amalgam electrodes for the determination of biologically active substances. LQ – limit of quantitation; DPV – differential pulse voltammetry; DCV – direct current voltammetry; SWV – square-wave voltammetry; ASV – anodic stripping voltammetry; CSV – cathodic stripping voltammetry; AdSV – adsorptive stripping voltammetry; FIA-ED – flow injection analysis with electrochemical detection; catalyt. – catalytic evolution of hydrogen; BR – BrittonRobinson; MeOH – methanol Substance Cu2+ Pb2+ Pb2+ Cd2+ Cd2+ Zn2+ Tl+ Ni2+

Electrode/technique m-AgSAE/ASV m-AgSAE/ASV m-CuSAE/ASV m-AgSAE/ASV m-CuSAE/ASV m-AgSAE/ASV m-AgSAE/ASV m-AgSAE/AdSV

Fe3+

m-AgSAE/DPV

Mn2+ IO3IO3-

m-CuSAE/CSV m-AgSAE/DPV p-AgSAE/DPV

Supporting electrolyte 0.4M acetate buffer (pH 4.8) 0.4M acetate buffer (pH 4.8) 0.2M acetate buffer (pH 4.8) 0.4M acetate buffer (pH 4.8) 0.2M acetate buffer (pH 4.8) 0.4M acetate buffer (pH 4.8) 0.4M acetate buffer (pH 4.8) 0.1M [NH4Cl + NH3] (pH 9.8), 0.001 % dimethylglyoxime 0.35M NaOH, 3 % triethanolamine, 0.01M EDTA 0.05M Na2B4O7 (pH 9.2) 0.1M NaOH 0.1M NaOH

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LQ 3·10-4 mg L-1 7·10-4 mg L-1 2·10-3 mg L-1 1·10-3 mg L-1 2·10-3 mg L-1 9·10-4 mg L-1 4·10-4 mg L-1 9·10-4 mg L-1

Ref. [27] [27] [34] [27] [34] [27] [27] [25]

0.20 mg L-1

[25]

0.03 mg L-1 0.14 mg L-1 0.34 mg L-1

[34] [35] [25]

162

B. Yosypchuk, T. Navratil, J. Barek et al.

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Table IV. Continued Substance IO3NO31,3-Dinitronaphtalene

Electrode/technique AgA-PE/DPV m-AgSAE/DPV m-AgSAE/DPV

1,5-Dinitronaphtalene

m-AgSAE/DPV

1,8-Dinitronaphtalene

m-AgSAE/DPV

3-Nitrofluoranthen 3-Nitrofluoranthen Nitrobenzene 1,3-Dinitrobenzene 2-Nitrophenol 2-Nitrophenol 4-Nitrophenol 4-Nitrophenol 4-Nitrophenol 4-Nitrophenol 2,4-Dinitrophenol 2,4-Dinitrophenol 2-Methyl-4,6dinitrophenol Pendimethalin Oxyfluorene

m-AgSAE/DPV m-AgSAE/AdSV m-AgSAE/DPV m-AgSAE/DPV m-AgSAE/DPV p-AgSAE/DPV m-AgSAE/DPV p-AgSAE/DPV m-AgSAE/FIA+ED AgA-PE/DPV m-AgSAE/DPV p-AgSAE/DPV m-AgSAE/DPV

Bifenox N,N-dimethyl-4-amino -2’carboxyazobenzene Doxorubicin Doxorubicin Atrazine Ametrine Ostazine orange Ascorbic acid

Supporting electrolyte 0.1M NaOH 0.02M CeCl3 BR buffer (pH 10) – MeOH, 1:1 BR buffer (pH 10) – MeOH, 1:1 BR buffer (pH 10) – MeOH, 1:1 0.01M NaOH – MeOH, 1:9 0.01M NaOH – MeOH, 1:1 0.1M acetate buffer (pH 4.8) 0.1M acetate buffer (pH 4.8) BR buffer (pH 8) BR buffer (pH 5) BR buffer (pH 6) BR buffer (pH 6) BR buffer (pH 6) 0.1M acetate buffer (pH 4.8) BR buffer (pH 4) BR buffer (pH 5) BR buffer (pH 4)

LQ 0.03 mg L-1 1.20 mg L-1 2·10-6 mol L-1

Ref. [36] [25] [37]

1·10-6 mol L-1

[37]

5·10-7 mol L-1

[37]

4·10-7 mol L-1 3·10-8 mol L-1 6·10-7 mol L-1 3·10-7 mol L-1 1·10-6 mol L-1 1·10-6 mol L-1 1·10-6 mol L-1 3·10-6 mol L-1 3·10-6 mol L-1 0.03 mg L-1 2·10-6 mol L-1 3·10-6 mol L-1 2·10-7 mol L-1

[38] [38] [25] [25] [30] [30] [30] [30] [39] [36] [30] [30] [40]

3·10-7 mol L-1 7·10-7 mol L-1

[38] [41]

m-AgSAE/DPV m-AgSAE/DPV

BR buffer (pH 7) – MeOH, 1:1 BR buffer (pH 10) – MeOH, 1:1 BR buffer (pH 7) – MeOH, 1:9 BR buffer (pH 5)

4·10-7 mol L-1 4·10-7 mol L-1

[42] [43]

m-AgSAE/DPV m-AgSAE/AdSV p-CuSAE/SWV p-CuSAE/SWV m-AgSAE/AdSV AgA-PE/DPV

BR buffer (pH 6) BR buffer (pH 6) 0.1M Na2SO4 (pH 2.4) 0.1M Na2SO4 (pH 2.4) 0.01M NaOH 0.1M acetate buffer (pH 4.8)

3·10-7 mol L-1 9·10-9 mol L-1 5·10-8 mol L-1 6·10-8 mol L-1 2 · 10-7 mol L-1 0.60 mg L-1

[44] [44] [45] [46] [47] [36]

m-AgSAE/DPV m-AgSAE/DPV

4.1. Determination of Dinitronaphthalenes Using DPV in Drinking Water This topic was described in more details in [37, 48]. Dinitronaphthalenes (DNN) belong to the group of genotoxic nitrated polycyclic aromatic hydrocarbons. They find their field of application as intermediate product in dye industry, in organic synthesis and 1,5-DNN in the explosives production. Environment is contaminated with these compounds first of all due to combustion processes of fossil fuels and by photochemical reactions of polycyclic aromatic hydrocarbons with nitrogen oxides in the atmosphere. Low levels of these substances are usually determined by expensive devices as GC-MS and HPLC with fluorescence or chemiluminiscence detection. In consequence of the presence of electrochemically reducible

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nitro groups it is possible to apply modern electroanalytical methods for DNN determination [49, 50]. Optimized determinations of 1,3-, 1,5- and 1,8-dinitronaphthalenes using DPV with m-AgSAE were used for their determinations in drinking water as model matrix. This procedure is not realizable in Britton-Robinson buffer, probably because of the precipitation of calcium and magnesium ions. The determination was realized in solution of 0.001M NaOH, pH 10.7. Under such conditions, DP voltammograms of 1,3-DNN and 1,8-DNN in dependence on concentration were recorded and plotted in corresponding calibration graph (see Figure 8). Probably due to the lower solubility of non-polar 1,5-DNN, this compound was in 0.001M NaOH precipitated and therefore it cannot be determined. Parameters of calibration dependences and limits of detections of 1,3-DNN and of 1,8-DNN are listed in Table V. -45

-50 I [nA] -40

7

A

Ip [nA]

6

-30

5

-30

B

4 -20 -15

3 -10

2 1

0

0

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-400

-600

E [mV]

-800

0

5

c [umol.L-1]

10

Figure 8. (A) DP voltammograms of 1,8-DNN with m-AgSAE in model sample of drinking water in mixture with 0.002M NaOH (1:1); pH 10.7; c(1,8-DNN) in analyzed solution [mol.L-1]: (1) 0supporting electrolyte 0.001M NaOH; (2) 1·10-6; (3) 2·10-6; (4) 4·10-6; (5) 6·10-6; (6) 8·10-6; (7) 10·10-6. (B) Calibration line for 1,8-DNN determination using DPV with m-AgSAE in model sample of drinking water in concentration range 1-10 µmol.L-1, supporting electrolyte 0.001M NaOH, error bars were calculated from average value and standard deviation from 5 measurements for each particular concentrations.

Table V. Parameters of calibration lines for DNN determination in drinking water as a model matrix

Compound

Peak

Slope [nA.L.μmol-1]

Intercept [nA]

r

LQ[μmol.L-1]

Concentration Range [μmol.L-1]

1,3-DNN

1.

-1.05

0.23

0.9969

1

1-10

1.

-4.26

-0.32

0.9970

-

1-10

1.

-3.37

0.9814

0.5

0.25-1.00

1,8-DNN

0.32

Preparation of Samples for Voltammetric Measurements 1 - 10 mL of analyzed water, (i.e., of analyzed sample) - in dependence on supposed concentration of determined nitronaphthalenes – are measured into voltammetric vessel, filled up to 10 mL with redistilled water and 0.1 mL of 0.1M NaOH is added. Oxygen is removed by passage of an inert gas (pure nitrogen or argon) for 300 s through the sample.

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Voltammetric Measurements Parameters of DDN determination: Ein = -100 mV; Efin = -1800 mV; regeneration of AgSAEs: 50 polarization jumps, in which the potentials -100 mV and −1800 mV are applied alternately, each for 0.3 s (automatically before each measurement), between working and reference electrode. The evaluation is realized using standard addition method. It can be concluded that it is possible to determine 1,3-dinitronaphtalene in concentration range 1-10 μmol.L-1, with limit of detection 1 μmol.L-1 and 1,8-dinitronaphtalene in concentration range 0.5-10 μmol.L-1, with limit of detection 0.5 μmol.L-1 in drinking water as model matrix and in 0.001M NaOH as supporting electrolyte.

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4.2. Voltammetric Determination of 2-nitrophenol at m-AgSAE Voltammetric determination of 2-nitrophenol at m-AgSAE was described in details in [30]. Nitrophenols coming from pesticide degradation products, car exhausts, and industrial wastes are listed as priority pollutants by the US Environmental Protection Agency [51, 52]. Pesticides based on simple nitrophenols are generally not allowed today but some of tham are still used as growth stimulators in agriculture [53]. They are potential carcinogens, teratogens, and mutagens [54]. Because of their toxicity and vast scale distribution in the environment, their determinations have become one of the important goals of environmental analysis. The practical applicability of these methods after model experiments with deionized water was confirmed by determination of 2-NP in drinking water as a simple environmental matrix. To improve limit of determination, preconcentration by solid phase extraction (SPE) from 100 mL and 1000 mL samples was used. Lichrolut EN cartridges containing polymeric sorbent with large specific surface and the adsorption capacity for polar organic substances (like e. g. nitrophenols [55]) were used. Recovery of 2-nitrophenol using SPE was calculated from the ratio of the peak height of the substance after SPE and peak height of 2-nitrophenol standard solution at concentration corresponding to expected concentration after extraction. Parameters of calibration dependences are summarized in Table VI. Table VI. Parameters of the calibration straight lines for the determination of 2-nitrophenol in drinking water after solid phase extraction using DPV at m-AgSAE. Concentrations range 20-1000 nmol.L-1. After SPE, the sample was eluted with 6 mL of methanol and filled up to 10 mL by Britton-Robinson buffer pH 8. Potentials for regeneration were Ereg1 = -100 mV and Ereg2 = -1300 mV Matrix Deonized water Drinking water

Volume [mL] 100 1000 100 1000

Recovery [%] 98.0 95.1 96.2 95.1

Concentration [mol.L-1] (2-10)·10-7 (2-10)·10-8 (2-10)·10-7 (2-10)·10-8

Slope [nA.mol-1.L] -1.04·107 -1.05·108 -0.95·108 -1.05·108

Intercept [nA] -0.6 -0.2 -0.4 -1.2

R 0.9987 0.9989 0.9983 0.9987

LQ [mol.L-1] 2.1·10-7 2.2·10-8 2.7·10-7 2.1·10-8

Mercury meniscus modified AgSAE is a suitable sensor for the determination of micromolar concentrations of nitrophenols. The limits of determination of nitrophenols using DPV at m-AgSAE and p-AgSAE are about 2 μmol.L-1, but for m-AgSAE larger signal

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stability given by lower passivation of electrode was observed. Furthermore, m-AgSAE has a better reproducibility and thus in many cases represents an effective and simpler alternative to the HMDE. Therefore, for coupling with solid phase extraction (SPE) more reliable m-AgSAE was used. It was shown that in combination with a preliminary separation and preconcentration by SPE it is possible to determine concentration of 2-nitrophenol down to 20 nmol.L-1. For methods without pre-concentration, limits of determination are similar as for differential pulse polarography on a dropping mercury electrode (2 μmol.L-1, [56]) and higher than for differential pulse voltammetry on a hanging mercury drop electrode (0.1 μmol.L-1, [57]). As with polarographic and voltammetric methods on mercury electrodes, substances reducible at the same potential will interfere.

Figure 9. “Wall-jet” cell based on meniscus modified silver solid amalgam electrode 1 – 4 m-AgSAE; 1 Electric contact; 2 Glass tube; 3 Ag amalgam; 4 Hg meniscus; 5 Reference electrode; 6 Auxiliary Pt electrode; 7 Teflon tubing; 8 Glass overflow vessel 9 Overflow. Reproduced with permission from Electroanalysis 2007, 19, 2003-2014.

5. THE USE OF SOLID AMALGAM ELECTRRODES IN FLOW-THROUGH SYSTEMS A combination of electrochemical detection with flow-injection system vastly improves the productivity by substantially decreasing the time per one analysis. An important issue concerning electrochemical detection at solid electrodes is electrode fouling due to the irreversible adsorption of reaction products and intermediates, which results in a need for electrode pretreatment. Moreover, the electrode itself should be mechanically stable which complicates the use of mercury electrodes in flow techniques. Novel electrode materials based on non-toxic solid amalgam were successfully tested for FIA-ED and HPLC-ED determination of trace amounts of electrochemically reducible environmental active organic

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B. Yosypchuk, T. Navratil, J. Barek et al.

compounds in water samples. It was shown that mechanically stable silver solid amalgam electrode can be used for determination of organic water pollutants in flowing systems. “Wall-jet” arrangement (Figure 9) with silver solid amalgam electrode (AgSAE) can be used for determination of model samples of electrochemically active pesticides (2-methyl-4,6dinitrophenol), [58] grow stimulators (4-nitrophenol) [39] and carcinogens (5-nitroquinoline). [59] For analysis in flowing systems it is generally possible directly use optimal condition found for batch analysis and optimize only detection potential, flow rate and injection volume (see Figure 10). The main advantage of the use of AgSAE in flowing system is its mechanical stability. Moreover, the effect of passivation is very low, because the removal of passivating products of electrode reaction from working electrode surface. Even if there are problems with passivation, they can be minimized by an electrochemical pretreatment of the working electrode. The presence of oxygen increases the background current and therefore it is necessary to remove it from measured solutions. -60 I [µA] -40

1

-20 2

0 0,0

-0,5

-1,0

-1,5

50

30 20

I [µA]

40

10 06 5

1

40 20

0

]

[m

V [ inj µl]

in

2

60

-1

4 3

80

l.m

100

Q

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E [mV]

Figure 10. Optimization of detection potential (A), flow rate and injection volume (B) for flow injection analysis (“wall-jet” m-AgSAE amperometric detector) of 2-methyl-4,6-dinitrophenol in BrittonRobinson buffer pH 4. 1 – substance signal, 2 – background current.

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Voltammetry at gold amalgam working electrodes was applied to measure simultaneously oxygen and dissolved sulfur species (S8,Sx2– , HS−/H2S) in an on-deck flow cell attached [60] when real-time redox profile of water in the Black Sea was monitored. The flow system has been used for adsorptive stripping voltammetric determinations of -9 10 mol.L-1 concentrations of nickel at gold amalgam microelectrodes [61]. Copper amalgam electrode allowed amperometric determination of metal-ions by flow injection system in 0.2 M ammoniacal buffer. [62] Flow cell constructed from Perspex was used for detection of Cu2+, Cd2+ and Zn2+ with limit of detection at 10-11mol.L-1 level. In this case the samples were not deaerated before injection, but contained sodium sulphite (8.10-3 mol.L-1) to reduce oxygen.

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CONCLUSIONS Solid and paste amalgam electrodes as environmentally friendly alternatives to mercury electrodes are suitable both for batch analysis and for HPLC with electrochemical detection (HPLC-ED) or flow injection analysis with electrochemical detection (FIA-ED) of electrochemically reducible substances with limit of quantitation (LQ) down to 10-7 mol L-1 [8]. They can be easily prepared in any laboratory and their simple electrochemical pretreatment in many cases eliminates problems with their passivation. Application of AEs presents a very suitable alternative to traditionally used hanging mercury drop electrode in connection with the differential pulse anodic stripping voltammetry method for determination of mentioned analytes in practically all types of water, drinking water including, as well as in relatively complicated matrices [63]. The achieved results are comparable with those achieved with HMDE or using AAS method in the case of trace metals determination. Results of DPASV with m-AgSAE are precise and accurate [63]. Because the amalgam is solid crystalline material, insoluble in most solvents, we can consider the electrodes based on this material as absolutely non-toxic (the toxicity under normal temperature is comparable with dental amalgams). These electrodes can be applied where the work with liquid mercury is forbidden or undesirable.

ACKNOWLEDGMENT Financial support of this work was provided by the Grant Agency of the Czech Republic (project No. 203/07/1195), by the Grant Agency of the Academy of Sciences of the CR (Project No. IAA400400806) and by the Ministry of Education, Youth and Sports of the Czech Republic (projects LC 06035 and MSM 0021620857).

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[39] Fischer, J.; Vanourkova, L.; Barek, J.; Navratil, T.; Novotny, L.; Yosypchuk, B.; Zima, J. Chem. Listy 2004, 98, 612-613. [40] Fischer, J.; Barek, J.; Yosypchuk, B.; Navratil, T. Electroanalysis 2006, 18, 127-130. [41] Novotny, V. B.Sc. Thesis, Charles University in Prague, Prague, 2006. [42] Cabalkova, D. B.Sc. Thesis, Charles Univesity in Prague, Prague, 2006. [43] Barek, J.; Dodova, E.; Navratil, T.; Yosypchuk, B.; Novotny, L.; Zima, J. Electroanalysis 2003, 15, 1778-1781. [44] Konecna, B. M.Sc. Thesis, Charles Univesity in Prague, Prague, 2005. [45] De Souza, D.; de Toledo, R. A.; Mazo, L. H.; Machado, S. A. S. Electroanalysis 2005, 17, 2090-2094. [46] De Souza, D.; de Toledo, R. A.; Suffredini, H. B.; Mazo, L. H.; Machado, S. A. S. Electroanalysis 2006, 18, 605-612. [47] Barek, J.; Fischer, J.; Navratil, T.; Peckova, K.; Yosypchuk, B. Sensors 2006, 6, 445452. [48] Daňhel, A. M.Sc. Thesis, Charles Univesity in Prague, Prague, 2006. [49] Shanmugam, K. PhD. Thesis, Charels University in Prague, Prague, 2004. [50] Shanmugam, K.; Barek, J.; Zima, J. Collect. Czech. Chem. Commun. 2004, 69, 20212035. [51] U.S. Environmental Protection Agency; Federal Register, 1989. [52] Luttke, J.; Scheer, V.; Levsen, K.; Wunsch, G.; Cape, J. N.; Hargreaves, K. J.; StoretonWest, R. L.; Acker, K.; Wieprecht, W.; Jones, B. Atmos. Environ. 1997, 31, 2637-2648. [53] List of the Registered Plant Protection Products in Czech Republic; SRS - The state phytosanitary administration: Brno, 2006. [54] Toxicological profile for nitrophenols; Agency for Toxic Substances and Disease Registry (ATSDR): Atlanta, 1992. [55] ChromBook 06/07; Merck KGaA: Darmstadt, 2006. [56] Zietek, M. Mikrochim. Acta 1975, 2, 463-470. [57] Barek, J.; Ebertova, H.; Mejstrik, V.; Zima, J. Collect. Czech. Chem. Commun. 1994, 59, 1761-1771. [58] Fischer, J.; Barek, J.; Zima, J.; Peckova, K.; Yosypchuk, B.; Navratil, T. Proc. 59. Zjazd chemickov, pp. 159-159. Slovenská chemická společnosť - ChemZi 2007. [59] Peckova, K.; Barek, J.; Cizek, K.; Jiranek, I.; Zima, J. In Sensing in electroanalysis; K. Vytřas, K. Kalcher, Eds.; University of Pardubice: Pardubice, 2007; Vol. 2, pp. 209216. [60] Glazer, B. T.; Luther, G. W.; Konovalov, S. K.; Friederich, G. E.; Nuzzio, D. B.; Trouwborst, R. E.; Tebo, B. M.; Clement, B.; Murray, K.; Romanov, A. S. Deep-Sea Res. Part II-Top. Stud. Oceanogr. 2006, 53, 1740-1755. [61] Bjorefors, F.; Nyholm, L. Anal. Chim. Acta 1996, 325, 11-24. [62] Alexander, P. W.; Akapongkul, U. Anal. Chim. Acta 1983, 148, 103-109. [63] Cizkova, P.; Navratil, T.; Sestakova, I.; Yosypchuk, B. Electroanalysis 2007, 19, 161171.

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Chapter 5

POLAROGRAPHIC AND VOLTAMMETRIC DETERMINATION OF GENOTOXIC SUBSTANCES IN DRINKING WATER USING MERCURY ELECTRODES Vlastimil Vyskocil, Jiří Barek, Ivan Jiranek and Jiří Zima Charles University in Prague, Faculty of Science, Department of Analytical Chemistry, UNESCO Laboratory of Environmental Electrochemistry, Albertov 6, CZ-128 43 Prague 2, Czech Republic

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ABSTRACT Even 85 years after their introduction in analytical chemistry by Nobel Prize winner professor Jaroslav Heyrovsky, mercury electrodes are still the best available sensors for voltammetric monitoring of trace amounts of electrochemically reducible inorganic and organic substances. Their main advantages (atomically smooth surface, easily renewable surface diminishing problems with passivation so frequently encountered with solid working electrodes and large potential window in cathodic region) usually overbalance their disadvantages (very limited anodic potential window, limited mechanical stability complicating their application for measuring in flowing media and unsubstantiated fears of their toxicity). Modern variants of mercury electrodes, namely hanging mercury drop electrode, in combination with pulse techniques or accumulation of the analyte on the surface of working electrode (electrochemical or adsorptive) enable to reach micromolar or even nanomolar limits of determination for electrochemically reducible substances. Recently, increasing attention is paid to their application for the determination of trace amount of genotoxic substances in both drinking and surface waters. In this chapter, practical applications of polarographic and voltammetric methods on mercury electrodes for the determination of trace amount of various chemical carcinogens (namely nitrated polycyclic aromatic hydrocarbons, electrochemically reducible heterocyclic compounds, etc.) in drinking water will be reviewed and compared with our most recent experimental results in this field. Advantages and disadvantages of various polarographic and voltammetric methods in this field will be critically evaluated. Attention will be paid to their combination with preliminary separation and preconcentration using liquid or solid phase extraction.

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Keywords: Polarography; Voltammetry, Dropping mercury electrode; Hanging mercury drop electrode; Chemical carcinogens.

ABBREVIATIONS AdSV BR buffer DC DCTP DME DPP DPASV DPV Eacc EPA FIA-AD GC-MS HMDE HPLC HPLC-ED

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HPLC-MS LD LLE LQ MeOH MFE MNQ n NPAH PAH SCE SPE SWV tacc τ TP

adsorptive stripping voltammetry Britton-Robinson buffer direct current DC tast polarography dropping mercury electrode differential pulse polarography differential pulse anodic stripping voltammetry differential pulse voltammetry potential of accumulation Environmental Protection Agency flow injection analysis with amperometric detection gas chromatography with mass spectrometric detection hanging mercury drop electrode high performance liquid chromatography high performance liquid chromatography with electrochemical detection high performance liquid chromatography with mass spectrometric detection limit of detection liquid–liquid extraction limit of quantification methanol mercury film electrode 6-methyl-5-nitroquinoline number of measurements nitrated polycyclic aromatic hydrocarbon polycyclic aromatic hydrocarbon saturated calomel electrode solid phase extraction square wave voltammetry time of accumulation time of the drop tetraethylammonium perchlorate

1. INTRODUCTION

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There is an ever increasing demand for new, more selective and more sensitive methods for the determination of trace amounts of various genotoxic substances in drinking water and surface waters. Some polycyclic aromatic hydrocarbons and their nitro derivatives, mycotoxins, aromatic amines, N-nitroso compounds, azo compounds, triazines, hydrazines, etc. can serve as an example. The objective carcinogenity evaluation can be found in monographic series of International Agency for Research on Cancer in Lyon [1]. The fact that even extremely small amounts of these substances can start various forms of cancer puts their monitoring on the top of the priority lists of most environmental agencies. Their determination in various environmental or biological matrices presents a challenge to modern analytical chemistry. The analytical methods used for their determination must be both sensitive and selective. Modern polarographic and voltammetric methods are among those capable to fulfill those stringent demands [2]. Modern polarographic and voltammetric techniques were successfully used for the determination of genotoxic substances containing electrochemically reducible functional groups (e.g. azo compounds, N-nitroso compounds, etc.) [3-7]. The use of classical polarography is limited to concentrations above 10-5 mol L-1. Much lower limits of detection can be achieved using modern polarographic methods, namely differential pulse polarography [8, 9], differential pulse voltammetry [8, 10] and adsorptive stripping voltammetry [8, 11]. The principles of these techniques and their practical applications were reviewed [12-14] and the general overview will be described in paragraph 3 of this chapter. The limit of quantification of these methods is 10-7 to 10-10 mol L-1 and their selectivity can be increased using preliminary separation with liquid-liquid or solid phase extraction or using column, paper or thin layer chromatography. For determination of some electrochemically nonreducible carcinogens (e.g. aromatic amines), their anodic oxidation at a suitable solid electrode can be used. High performance liquid chromatography with electrochemical detection combines powerful separation tool with sensitivity of electrochemical detection [15]. The application of polarographic detector enables analysis of polarographically reducible carcinogens (e.g. N-nitroso compounds, azo compounds, etc.), whereas the voltammetric detector can be used for analysis of oxidisable substances (e.g. aromatic amines). The combination of both detectors and/or their connection with photometric or refractometric detector can often increase the selectivity which is necessary especially for analysis of biological materials.

2. MERCURY WORKING ELECTRODES The performance of voltammetry is strongly influenced by the working electrode material. Ideally the electrode should provide a high signal-to-noise ratio as well as a reproducible response. In polarography and voltammetry, different working mercury electrodes can be used. Two types of mercury electrodes are commonly used in determination of low concentrations of genotoxic pollutants – the dropping mercury electrode and hanging mercury drop electrode. The greatest advantage of mercury electrodes is the fact that new drops or new thin mercury films can be readily formed and this renewal process diminishes problems with electrode passivation. This is not generally the case of electrodes made from other materials, with the possible exception of carbon paste electrodes, where the electrode

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surface renewal can be made by cutting off a thin layer of the previous electrode surface. Important features of mercury electrodes in polarography and voltammetry are summarized in Table I. Table I. Summary of commonly used working mercury electrodes Working Electrode

DME

HMDE

Characteristics - mercury freely dropping from a capillary, τ=1-5s

Advantages - simplicity - reliability - renewable surface

- valve mechanics and a hammer - electrode surface not renewed during one analysis - whole analysis on one drop

- LD ~ 10-7 - 10-10 mol L-1 - high reproducibility - low consumption of mercury - adsorptive or electrolytic accumulation - possibility of chemical modifications

Disadvantages - LD ~ 10-5 mol L-1 - high consumption of mercury - higher charging current - demands on stand mechanics - increased danger of passivation - more complex mechanics and electronics

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2.1. Dropping Mercury Electrode In classical polarography, the dropping mercury electrode is used as a working electrode. The mercury drop is formed at the end of a glass capillary (length 10 – 20 cm, inside diameter 0.05 mm) as the capillary is connected by flexible tube to a reservoir about 50 cm in height. The mercury droplets are of highly reproducible diameter and of a lifetime ranging from 2 to 6 s. The drop-time is controlled by the height of the mercury reservoir (i.e. by the hydrostatic pressure of mercury). The device is usually connected to a mechanical dislodge knocker to control the lifetime of mercury drops. The advantages of the dropping mercury electrode are as follows: the constant renewal of the electrode surface eliminates the contamination of the electrode surface, which results in reproducible current-potential data. The charge transfer overvoltage of hydrogen ions present in the aqueous medium is high on mercury, thus its reduction does not disturb the study of the reduction processes of the electroactive species having more negative potential. In acidic solution at a potential more negative than -1.2 V, hydrogen gas evaluation is observed [16]. The polarization range of the mercury electrode vs. saturated calomel electrode in aqueous solution in the absence of oxygen is between +0.3 and about -2.0 V. The cathodic potential limit is determined by the reduction of cations of the so-called supporting electrolyte (a supporting electrolyte is added in order to ensure diffusion-controlled mass transport conditions). The most negative polarization limit can be achieved in the presence of 2+

tetraalkylammonium cations, while the oxidation of metal mercury to Hg 2

sets a limit of

polarization in the positive potential range. From this it can be concluded that reducible electroactive substances can be primarily analyzed with the mercury drop electrode [14].

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2.2. Hanging Mercury Drop Electrode This electrode construction features the same elements as the dropping mercury electrode, e.g. mercury reservoir and glass capillary, but the reproducibility of drop size and drop time are controlled more precisely. This is ensured by incorporating a solenoid-activated valve between the mercury reservoir and the capillary, and a mechanical knocker attached to the capillary. Thus, by appropriate setting of the opening and closing time of the valve, the time of dislodging of the drop can be set, and the device can be operated as a hanging mercury drop or in a dropping operation mode with controlled drop time [17]. The hanging mercury drop electrode has all the characteristics of the dropping mercury electrode that make it especially advantageous for routine analytical work. An additional feature is due to the constant surface area of the drop, which will be discussed in connection with highly sensitive, current-sampled pulse techniques [17].

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3. POLAROGRAPHIC AND VOLTAMMETRIC TECHNIQUES Simple principle of polarography is the study of solutions or of electrode processes by means of electrolysis with two electrodes, one polarizable and one unpolarizable, the former formed by mercury regularly dropping from a capillary tube. In the case of the common voltage-controlled (or, with three-electrode systems, potential-controlled, or "potentiostatic") electrolysis, the polarographic curve is a current-voltage (or current-potential) curve showing the dependence of the current, passing through the system, on the voltage applied to the electrodes, or on the potential of the dropping electrode. The dropping mercury electrode keeps the interface always fresh between its constantly renewed surface and the solution, independent of the processes that were taking place at previous drops, and thereby the measurements taken with it are perfectly reproducible. Polarography was developed in 1922 when Jaroslav Heyrovsky used dropping mercury electrode for measuring polarization curves. Since this time, the polarography has come through many innovations. Once the recording of polarization curves was made automatic, the relative simplicity of the apparatus and the rapidity of gaining results attracted attention of analytical chemists, and, especially after an exact theoretical background was worked out, polarography has become and remains an appreciated and trusted method in laboratories all over the world, thanks to the high reproducibility of the curves. Technical progress has continued modifying the classical, original method to meet increasing requirements of science and technology; the latest versions derived from the simple electrolysis with dropping mercury electrode (e.g. differential pulse voltammetry or adsorptive stripping voltammetry at hanging mercury drop electrode) belong to the most sensitive and versatile methods of analysis, and so are most appropriate for trace analysis of micromolar or even nanomolar concentrations of genotoxic organic pollutants of drinking water or its sources [8]. The detailed fundamentals and theory of polarographic and voltammetric techniques are well described in monographs [13, 14, 18]. Voltammetric methods used today in analytical laboratories comprise a suite of techniques, the creation of which was made possible by rapid advances in instrumentation, by the computerized processing of analytical data, and particularly by innovative electrochemists. Advances in microelectronics and in particular the

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early introduction of operational amplifiers and feedback loops have led to major changes in electroanalytical instrumentation. Indeed, many functions can be performed now by small and reliable integrated circuits. Voltammetric analyzers consist of two such circuits: a polarizing circuit that applies the potential to the cell and a measuring circuit that monitors the cell current. The working electrode is potentiostatically controlled, and this minimizes errors from cell resistance. Electroanalytical procedures can be fully programmed and can be driven automatically by means of a personal computer with a user-friendly software [13]. All this results in the possibility of fast ”time-resolved” sampling of the current from dropping mercury electrode. The mercury drop emerging from the capillary monitors current which consists of that due to charging of the double layer and the faradaic current produced by reduction or (less frequently) oxidation of the analyte in solution. The contribution of the capacitance current becomes less as the drop increases in size and the rate of increase in area becomes much smaller. Thus, if the current is sampled at a long enough time after the drop has started to emerge from the capillary, the capacitance current is discriminated against to the faradaic current: this is utilized in its simplest form in tast polarography, but it is utilized also when more advanced pulse waveforms are used. All commonly used pulse techniques (DPP, DPV and SWV) are chronoamperometric and are based on a sampled current potentialstep experiment. After the potential is stepped, the charging current decreases rapidly (exponentially), while the faradaic current decays more slowly [14]. Stripping analysis is one of the most sensitive voltammetric methods. A detailed description of stripping voltammetry has been given in a monograph by Wang [19]. Its high sensitivity is due to the combination of an effective preconcentration step (electrolytic or adsorptive) with advanced measurement procedure. Because analytes are preconcentrated onto the electrode by factors of 100 - 1000, detection limits are lowered by 2 - 3 orders of magnitude to those voltammetric measurements which do not utilize preliminary accumulation. A survey of the theory and practical applications of the preconcentration methods can be found in monographs [20, 21] and reviews [22, 23]. For the trace analysis of genotoxic substances in water solutions is the most used technique adsorptive stripping voltammetry [19]. AdSV uses nonelectrolytic adsorptive preconcentration where the analyte accumulation is a result of its adsorption on the electrode surface or that of a surface active complex of the analyte. It exploits the reduction of a metal or of a ligand in the adsorbed complex. The adsorption can be coupled in some cases with catalytic reactions. The theoretical aspects of electrocatalysis on HMDE are described in ref. [24]. AdSV proved to be suitable for measuring trace amounts of metals in complexes with chelating agents and of many surface active organic compounds (drugs, vitamins, pollutants and many others). It has been applied in many environmental and clinical studies [25] as well as in drug analysis [26]. Possibilities of stripping voltammetry with an emphasis on adsorptive stripping voltammetry and on the use of modified or ultramicro electrodes [27] and chemically modified electrodes, including mercury ones [28] have been reviewed. The pros and cons of the reactant adsorption in pulse techniques together with the survey of phenomena due to reactant adsorption and with practical guidelines of treating it have also been discussed [29]. Recent applications of AdSV can be found in [11]. It follows from the results shown in paragraph 4 of this chapter that the most frequently used techniques for the genotoxic water pollutants determination are DPP at DME, DPV at HMDE, SWV at HMDE and AdSV at HMDE. The comparison of LD with other voltammetric techniques at mercury electrodes [8] is shown in Table II.

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Table II. Basic parameters of modern polarographic and voltammetric techniques Technique

Working Electrode

LD [mol L-1]

DC tast polarography

DME

~ 10-6

Normal pulse polarography

DME

~ 10-7

Normal pulse voltammetry

HMDE

~ 10-7

Staircase voltammetry

HMDE

~ 10-7

Differential pulse polarography

DME

~ 10-7

Differential pulse voltammetry

HMDE

~ 10-8

Square wave polarography

DME

~ 10-8

Square wave voltammetry

HMDE

~ 10-8

Alternating current polarography

DME

~ 10-7

Alternating current voltammetry

HMDE

~ 10-8

Anodic stripping voltammetry

HMDE

~ 10-10

Cathodic stripping voltammetry

MFE

~ 10-9

HMDE

~ 10-11

MFE

~ 10-12

MFE

~ 10-12

Adsorptive stripping voltammetry Potentiometric stripping analysis

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3.1. Methods of Preliminary Separation and Preconcentration It can be seen in Table II that sensitivity of polarographic and voltammetric techniques at mercury electrodes is above-average for the determination of genotoxic pollutants in drinking water. However, the selectivity of these techniques is limited by the width of potential window. For the determination in simple matrices (e.g. drinking water or rain water), the direct determination can be used. In the case the matrix is more complicated (e.g. river or sea water), the preliminary separation step is required. This step can be simultaneously used for analyte preconcentration, too. The LLE or SPE are most frequently used techniques. LLEs are usually accomplished with a separatory funnel. However, they also may be carried out in the sample container by adding the extracting solvent when the sample is collected. Pesticides in water, for example, may be preserved for longer periods by extracting into a small volume of hexane added to the sample in the field. Liquid-liquid microextractions, in which the extracting phase is a 1-mL drop suspended from a microsyringe have also been described [30]. A LLE is one of the most important separation techniques used in environmental, clinical, and industrial laboratories. For example, public drinking water supplies are routinely monitored for trihalomethanes (CHCl3, CHBrCl2, CHBr2Cl, and CHBr3) because of their known or suspected carcinogenity. Before their analysis by gas chromatography, trihalomethanes are separated from their aqueous matrix by a liquid–liquid extraction using pentane [31].

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In a simple liquid–liquid extraction the solute is partitioned between two immiscible phases. In most cases one of the phases is aqueous, and the other phase is an organic solvent such as diethyl ether or chloroform [32]. In SPE the sample is passed through a cartridge containing solid particulates that serve as the adsorbent material. For liquid samples the solid adsorbent is isolated in either a disk cartridge or a column. The choice of adsorbent is determined by the properties of the species being retained and the matrix in which it is found. Representative solid adsorbents are silica or alumina, eventually modified by cyanopropyl, diol, octadecyl (C-18), octyl (C-8), styrene, etc. Typically a 500-mL of sample is passed through the cartridge. The cartridge is then washed with distilled water to remove any residual traces of the matrix. Finally, the retained analytes are eluted from the cartridge by a solid–liquid extraction using suitable solvent. For many analyses, SPEs are replacing LLEs due to their ease of use, faster extraction times, decreased volumes of solvent used, and their superior ability to concentrate the analytes [32]. The multicomponent samples or samples in more complicated matrix (e.g. waste waters, industrial waters) are unsuitable to be analysed by less selective electrochemical methods. The preliminary separation step is set before proper electrochemical determination to eliminate selectivity problem. In analysis of liquid samples the most suitable and powerful technique is HPLC. In combination with electrochemical detection, we obtain very sensitive and selective method for the determination of wide group of organic compounds [33]. Nevertheless, HPLC in combination with polarographic or voltammetric detection using mercury electrodes is not often used. An article by Bersier and Bersier with many references reviews polarography, voltammetry and HPLC-ED of pharmaceuticals [15, 34]. Newly developed electrode materials get an opportunity to excel in electrochemical detection area. Especially, boron-doped diamond electrodes, solid amalgam electrodes, carbon paste electrode or graphite composite electrodes are recently used for electrochemical detection [2, 35]. Although mercury electrodes are more suitable for batch analysis, HPLC offers the possibility to separate particular mixture components and than to determine them.

4. PRACTICAL APPLICATIONS National Primary Drinking Water Regulations [36] apply to public water systems and are legally enforceable standards. These primary standards are intended to protect public health by limiting the levels of contaminants that can be found in drinking water. Although these standards are applicable to public water systems (i.e., at the tap), they are often applied by remediation regulators in the aquifer (i.e., at the monitoring wellhead). Table III summarizes the drinking water standards imposed by the U.S. Environmental Protection Agency [37]. Substances in bold are amenable to polarographic/voltammetric determination at mercury electrodes (appropriate references of polarographic activity are incorporated). Additional information regarding potential health impacts and sources of contamination can also be found at US EPA web site [36]. It is obvious from this list that the variability of dangerous drinking water pollutants is very wide in structure as well as in toxicity. Although in the case of electroanalytical methods we can speak about sensitive and inexpensive methods, in the area of drinking water monitoring and analysis they are not too often used. However, the only prerequisite to voltammetric determination is the presence of reducible or oxidisable moieties.

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Thus, the research in polarographic and voltammetric determination of genotoxic pollutants in drinking water should further continue. Two groups of organic genotoxic compounds are studied in connection with drinking water monitoring using mercury working electrodes. Those are pesticides and environmental carcinogens (i.e., NPAHs, heterocyclic compounds, etc.). During methods development, the optimum conditions are found (e.g. optimum composition of supporting electrolyte, optimum working electrode, optimum technique, utilization of accumulation etc.) than the method is tested at model samples of drinking water spiked with the genotoxic pollutant. The examples of mercury electrodes applications in drinking water contaminants determination are given in this review.

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4.1. Examples of Electrochemically Reducible Genotoxic Substances Polycyclic aromatic hydrocarbons are mostly determined using gas chromatographic or spectrofluorometric methods [59]. Polarographic techniques are not too suitable, because of these compounds are reduced at considerably negative potential (about -2.5 V vs. SCE). Demanded sensitivity and accuracy cannot be reached at this potential. It documents e.g. an attempt for the polarographic determination of 3,4-benzopyrene in air; this determination does not suit for the practical application because of its inaccuracy [60]. Other possible drinking water contaminants are the substances from the group of mycotoxins. Metabolites of some moulds belong to this group. These genotoxic compounds, including their structure, genotoxic properties and methods of their determination in different types of matrices, are reviewed [61]. The polarographic and voltammetric methods of their determination are discussed in review [62]. Mitomycin B and C are reversibly polarographically reducible [63]. Among another substances produced by the some type of moulds are active, clinically used cytostatics, as e.g. doxorubicin or daunorubicin. These compounds are polarographically reducible as well [64]. About 300 substances from the group of N-nitroso compounds were tested on carcinogenity. Around 87% of them gave the positive results during animal tests. The gas or liquid chromatography is mainly used for the N-nitroso compounds determination; thermal energy analyzer is suitable detector because of the high sensitivity and selectivity to N-nitroso compounds [65]. Polarographic behavior of N-nitroso compounds from the physical and analytical chemistry point of view was reviewed in [10, 66]. Carcinogenic N-nitrosomorpholine that is produced in Diesel motor combustion can be determined by square wave polarography [67]. The polarography without preliminary separation or preconcentration step enables N-nitroso amines determination in simple matrices. In the case of more complicated samples, the suitable extraction or separation step is necessary prior to polarographic or voltammetric determination. N-nitroso compounds do not adsorb on HMDE and thus cannot be determined using AdSV [9]. The methods used for the determination of carcinogenic aromatic amines are summarized in [68-71]. The polarography enables the amino compounds determination after their conversion to derivatives which are reducible at mercury electrodes. Derivatization complicates the determination; therefore, the anodic oxidation at solid electrodes is more frequently used. The same holds for hydrazine derivatives [72]. Hydrazines alone are not polarographically reducible. Nevertheless, they can be determined after reaction with

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carbonyl group leading to polarographically reducible compounds. The electrochemical behavior of azo compounds was investigated in detail and the polarography and voltammetry at mercury electrodes is frequently used for the azo compounds determination [73]. Nitrated polycyclic aromatic hydrocarbons are a relatively new class of environmental carcinogens [3, 74]. Nevertheless, there is an ever increasing demand for the determination of their trace concentrations. So far mostly chromatographic methods, such as GC-MS or HPLC with fluorimetric detection have been used for the purpose [35]. Because of easy polarographic reducibility of nitro group [75-77], modern electroanalytical methods, especially very sensitive techniques at mercury electrodes, could be expected to satisfy high requirements for sensitivity of the determination. Nevertheless, the use of these methods for the determination of NPAHs has not been properly investigated so far [7], even though they are much cheaper as far as investment and running costs are concerned and they present an independent alternative, which is very important from the legal point of view. From this short review of applicability of polarography and voltammetry at mercury electrodes it is obvious that these methods can be used for different types of genotoxic compounds and their application for the drinking water organic genotoxic contaminants can be evident.

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4.2. Determination of Genotoxic Substances in Drinking Water 4.2.1. Nitrated Polycyclic Aromatic Hydrocarbons NPAHs belong among substances whose occurrence in the environment can be related to an increased cancer rate. Three main sources of NPAHs are believed to contribute significantly to the NPAHs occurrence in environment: the industrial production, NPAHs originating as direct or indirect products of incomplete combustion and NPAHs formed in the atmosphere from the gas-phase reactions of PAHs (generally adsorbed on particulate matter and themselves products of incomplete combustion) with four rings or less. Because of their easy polarographic reduction, they can be determined using modern polarographic and voltammetric techniques at nanomolar and subnanomolar levels, including determination in drinking or river water samples [6, 7, 9]. Polarography and voltammetry were successfully applied for monitoring the efficiency of the destruction of chemical carcinogens, which had been carried out by chemical oxidants, reduction agents or UV irradiation [7]. For example, the polarographic behavior of 3-nitrofluoranthene was investigated by DC tast polarography and DPP, both at a DME, DPV and by AdSVat a HMDE [78]. Optimum conditions have been found for its determination by the given methods in the concentration ranges of 1×10-6 – 1×10-4 mol L-1 (DCTP), 1×10-7 – 1×10-4 mol L-1 (DPP), 1×10-8 – 1×10-6 mol L-1 (DPV) and 1×10-9 – 1×10-7 mol L-1 (AdSV), respectively. Practical applicability of these techniques was demonstrated on the determination of 3-nitrofluoranthene in drinking and river water after its preliminary separation and preconcentration using LLE and SPE with the limits of determination 4×10-10 mol L-1 (drinking water) and 2×10-9 mol L-1 (river water) [78]. DPV at HMDE was successfully used for sensitive determination of another genotoxic NPAHs in drinking water using direct determination or in combination with preconcentration as well. The determined substances were nitro derivatives of naphthalene, anthracene, biphenyl and fluorenone, respectively. Reached limits of quantification cover from

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submicromolar and nanomolar concentrations even to the subnanomolar concentrations. Methods developed in UNESCO Laboratory of Environmental Electrochemistry for the determination of NPAHs are summarized in Table IV. It is evident that voltammetric techniques at mercury electrodes are suitable for the determination of potential drinking water pollutants from the family of NPAHs.

4.2.2. Heterocyclic Compounds While homocyclic polyaromatic hydrocarbons including their oxy-, hydroxy-, chloroand nitro-derivates have been of a major concern since 1970s, N-, O-, and S-heterocyclic PAHs have been studied relatively recently. A little interest in these heterocyclic compounds was probably caused by their low concentrations in the environment where they are at levels one or two orders lower than those of their homocyclic analogues [92]. However, the first studies focused on biological effects and toxicity of heterocyclic PAHs together with determined environmental concentrations, in soil, sediments, water and air, have started discussions on their possible human health risks as well as ecological risks [93]. The most studied group of these “new environmental pollutants” are N-heterocyclic PAHs. Heteroaromatic compounds with a π-electron deficit are polarographically reducible and can be determined using modern polarographic and voltammetric techniques in aqueousmethanolic solutions [9]. The practical procedure of method development, including obtained results, can be demonstrated on the case of determination of 6-methyl-5-nitroquinoline (MNQ) [94]. The limits of quantification for the determination of this compound in drinking water are around 10-9 mol L-1 (see Table V). These concentrations are comparable with limits given by EPA in their National Primary Drinking Water Regulations (see Table III). MNQ, compound with proven mutagenic and cytotoxic properties [95-97], has been studied polarographically and voltammetrically at mercury electrodes. The method of MNQ determination using DC tast polarography at DME (LQ ~ 1×10-6 mol L-1), DPP at DME (LQ ~ 2×10-7 mol L-1), DPV at HMDE (LQ ~ 2×10-7 mol L-1) and AdSV at HMDE (LQ ~ 2×10-8 mol L-1) has been developed. It has been verified that MNQ can be directly determined (without preliminary extraction) in drinking water using DPV at HMDE. The procedure was following: to 4.0 mL of spiked drinking water 5.0 mL of methanol were added and the solution was filled up to 10.0 ml by BR buffer pH 5.0 (resulting pH 6.2). Corresponding DP voltammograms are depicted in Figure 1 for the sake of illustration. The reached LQ was 3×10-7 mol L-1. To further increase the sensitivity, AdSV at HMDE has been used for the direct determination in drinking water. The procedure was as follows: to 9.0 mL of spiked drinking water 0.1 mL of methanol was added and the solution was filled up to 10.0 mL with BR buffer pH 5.0 (resulting pH 5.8). Corresponding AdS voltammograms are depicted in Figure 2. The reached LQ was 2×10-8 mol L-1.

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Table III. EPA national primary drinking water standards for organic chemicals. Polarographically active compounds are in bold Max. Contaminant Levelb [mg L-1] Treatment technology* 0.002

Contaminant

Max. Contaminant Level Goala [mg L-1]

Acrylamide

zero

Alachlor

zero

Atrazine

0.003

0.003

Cardiovascular system or reproductive problems

Benzene

zero

0.005

Anemia; decrease in blood platelets; increased risk of cancer

Benzo(a)pyrene (PAHs)

zero

0.0002

Reproductive difficulties; increased risk of cancer

Carbofuran

0.04

0.04

Problems with blood, nervous system, or reproductive system

Carbon tetrachloride

zero

0.005

Liver problems; increased risk of cancer

Chlordane

zero

0.002

Liver or nervous system problems; increased risk of cancer

Chlorobenzene

0.1

0.1

Liver or kidney problems

2,4-D

0.07

0.07

Kidney, liver, or adrenal gland problems

Dalapon

0.2

0.2

Minor kidney changes

1,2-Dibromo-3chloropropane (DBCP)

zero

0.0002

Reproductive difficulties; increased risk of cancer

o-Dichlorobenzene

0.6

0.6

Liver, kidney, or circulatory system problems

p-Dichlorobenzene

0.075

0.075

Anemia; liver, kidney or spleen damage; changes in blood

Potential Health Effects from Ingestion of Water Nervous system or blood problems; increased risk of cancer Eye, liver, kidney or spleen problems; anemia; increased risk of cancer

Sources of Contaminant in Drinking Water Added to water during sewage/wastewater treatment Runoff from herbicide used on row crops Runoff from herbicide used on row crops Discharge from factories; leaching from gas storage tanks and landfills Leaching from linings of water storage tanks and distribution lines Leaching of soil fumigant used on rice and alfalfa Discharge from chemical plants and other industrial activities Residue of banned termiticide Discharge from chemical and agricultural chemical factories Runoff from herbicide used on row crops Runoff from herbicide used on rights of way Runoff/leaching from soil fumigant used on soybeans, cotton, pineapples, and orchards Discharge from industrial chemical factories Discharge from industrial chemical factories

Ref. [38] [39] [40] –

– [41] [42] – – [43] [44] – – –

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Table III. Continued Contaminant

Max. Contaminant Level Goala [mg L-1]

Max. Contaminant Levelb [mg L-1]

Potential Health Effects from Ingestion of Water

1,2-Dichloroethane

zero

0.005

Increased risk of cancer

1,1-Dichloroethylene

0.007

0.007

Liver problems

cis-1,2-Dichloroethylene

0.07

0.07

Liver problems

trans-1,2-Dichloroethylene

0.1

0.1

Liver problems

Dichloromethane

zero

0.005

Liver problems; increased risk of cancer

1,2-Dichloropropane

zero

0.005

Increased risk of cancer

Di(2-ethylhexyl) adipate

0.4

0.4

Di(2-ethylhexyl) phthalate

zero

0.006

Dinoseb

0.007

0.007

Reproductive difficulties

Dioxin (2,3,7,8-TCDD)

zero

0.00000003

Reproductive difficulties; increased risk of cancer

Diquat Endothall Endrin

0.02 0.1 0.002

0.02 0.1 0.002

Cataracts Stomach and intestinal problems Liver problems

Epichlorhydrin

zero

Treatment technology*

Increased cancer risk, and over a long period of time, stomach problems

Ethylbenzene

0.7

0.7

Liver or kidneys problems

Weight loss, liver problems, or possible reproductive difficulties. Reproductive difficulties; liver problems; increased risk of cancer

Sources of Contaminant in Drinking Water Discharge from industrial chemical factories Discharge from industrial chemical factories Discharge from industrial chemical factories Discharge from industrial chemical factories Discharge from drug and chemical factories Discharge from industrial chemical factories Discharge from chemical factories Discharge from rubber and chemical factories Runoff from herbicide used on soybeans and vegetables Emissions from waste incineration and other combustion; discharge from chemical factories Runoff from herbicide use Runoff from herbicide use Residue of banned insecticide Discharge from industrial chemical factories; an impurity of some water treatment chemicals Discharge from petroleum refineries

Ref. – – – – – – – – [45]

– [46] – [47] – –

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Contaminant

Max. Contaminant Level Goala [mg L-1]

Max. Contaminant Levelb [mg L-1]

Potential Health Effects from Ingestion of Water Problems with liver, stomach, reproductive system, or kidneys; increased risk of cancer Kidney problems; reproductive difficulties Liver damage; increased risk of cancer

Sources of Contaminant in Drinking Water Discharge from petroleum refineries Runoff from herbicide use Residue of banned termiticide

Ethylene dibromide

zero

0.00005

Glyphosate Heptachlor

0.7 zero

Heptachlor epoxide

zero

0.7 0.0004 0.0002

[49] [50]

Liver damage; increased risk of cancer

Breakdown of heptachlor

–

Hexachlorobenzene

zero

0.001

Liver or kidney problems; reproductive difficulties; increased risk of cancer

Hexachlorocyclopentadiene

0.05

0.05

Kidney or stomach problems

Lindane

0.0002

0.0002

Liver or kidney problems

Methoxychlor

0.04

0.04

Reproductive difficulties

Oxamyl (Vydate)

0.2

0.2

Slight nervous system effects

Polychlorinated biphenyls (PCBs)

zero

0.0005

Skin changes; thymus gland problems; immune deficiencies; reproductive or nervous system difficulties; increased risk of cancer

Pentachlorophenol

zero

0.001

Liver or kidney problems; increased cancer risk

Picloram Simazine

0.5 0.004

0.5 0.004

Liver problems Problems with blood

Styrene

0.1

0.1

Liver, kidney, or circulatory system problems

Tetrachloroethylene

zero

0.005

Liver problems; increased risk of cancer

Discharge from metal refineries and agricultural chemical factories Discharge from chemical factories Runoff/leaching from insecticide used on cattle, lumber, gardens Runoff/leaching from insecticide used on fruits, vegetables, alfalfa, livestock Runoff/leaching from insecticide used on apples, potatoes, and tomatoes Runoff from landfills; discharge of waste chemicals Discharge from wood preserving factories Herbicide runoff Herbicide runoff Discharge from rubber and plastic factories; leaching from landfills Discharge from factories and dry cleaners

Ref. [48]

– [51] [52] [53]

[54]

[55] [56] [57] [40] – –

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Table III. Continued

a

Contaminant

Max. Contaminant Level Goala [mg L-1]

Max. Contaminant Levelb [mg L-1]

Potential Health Effects from Ingestion of Water

Toluene

1

1

Nervous system, kidney, or liver problems

Toxaphene

zero

0.003

2,4,5-TP (Silvex)

0.05

0.05

Kidney, liver, or thyroid problems; increased risk of cancer Liver problems

1,2,4-Trichlorobenzene

0.07

0.07

Changes in adrenal glands

1,1,1-Trichloroethane

0.20

0.2

Liver, nervous system, or circulatory problems

1,1,2-Trichloroethane

0.003

0.005

Liver, kidney, or immune system problems

Trichloroethylene

zero

0.005

Liver problems; increased risk of cancer

Vinyl chloride

zero

0.002

Increased risk of cancer

Xylenes (total)

10

10

Nervous system damage

Sources of Contaminant in Drinking Water Discharge from petroleum factories Runoff/leaching from insecticide used on cotton and cattle Residue of banned herbicide Discharge from textile finishing factories Discharge from metal degreasing sites and other factories Discharge from industrial chemical factories Discharge from metal degreasing sites and other factories Leaching from PVC pipes; discharge from plastic factories Discharge from petroleum factories; discharge from chemical factories

Ref. – –

– – – – [58] –

– the level of a contaminant in drinking water below which there is no known or expected risk to health (maximum contaminant level goals allow for a margin of safety and are non-enforceable public health goals); b – the highest level of a contaminant that is allowed in drinking water (maximum contaminant levels are set as close to maximum contaminant level goals as feasible using the best available treatment technology and taking cost into consideration; maximum contaminant levels are enforceable standards). * – each water system must certify, in writing, to the state (using third-party or manufacturer's certification) that when acrylamide and epichlorohydrin are used in drinking water systems, the combination (or product) of dose and monomer level does not exceed the levels specified, as follows: acrylamide = 0.05% dosed at 1 mg L-1 (or equivalent); epichlorohydrin = 0.01% dosed at 20 mg L-1 (or equivalent).

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Table IV. Voltammetric determination of selected genotoxic nitrated polycyclic aromatic hydrocarbons in drinking water Substance

Technique

Preconcentration

Medium

LQ [mol L-1]

DPV at HMDE

direct determination

drinking water – 0.1 mol L-1 NaOH (9:1)

2×10-7

DPV at HMDE

SPE (from 100 mL)

MeOH – 0.01 mol L-1 NaOH (1:1)

2×10-8

DPV at HMDE

direct determination

drinking water

2×10-8

DPV at HMDE

SPE (from 500 mL)

MeOH – BR buffer pH 11 (1:1)

2×10-9

DPV at HMDE

SPE (from 500 mL)

BR buffer pH 7

4×10-7

2,2’-Dinitrobiphenyl

2,7-Dinitro-9-fluorenone 9-Nitroanthracene 3-Nitrobiphenyl 3-Nitrofluoranthene 2-Nitro-9-fluorenone

1-Nitronaphthalene

direct determination

drinking water – MeOH – 1 mol L NaOH (50:49:1)

2×10

DPV at HMDE

hexane LE

MeOH – BR buffer pH 12 (1:1)

2×10-9

DPV at HMDE

hexane LE

MeOH – BR buffer pH 3 (9:1)

7×10-9

direct determination

drinking water – MeOH – BR buffer pH 11 (4:5:1)

2×10

DPV at HMDE

SPE (from 50 mL)

MeOH – BR buffer pH 11 (1:1)

2×10-8

DPV at HMDE

SPE (from 500 mL)

MeOH – BR buffer pH 11 (1:1)

2×10-9

DPV at HMDE

direct determination

drinking water – 0.1 mol L-1 NaOH (9:1)

2×10-8

DPV at HMDE

SPE (from 100 mL)

MeOH – 0.01 mol L-1 NaOH (1:9)

2×10-9

DPV at HMDE

SPE (from 1000 mL)

MeOH – 0.01 mol L NaOH (1:9)

2×10-10

DPV at HMDE

drinking water – 0.1 mol L-1 NaOH (9:1)

2×10-8

drinking water – 0.01 mol L-1 NaOH (9:1)

5×10-9

drinking water – 0.01 mol L-1 NaOH (9:1)

4×10-10

DPV at HMDE

direct determination hexane LE (from 100 mL) hexane LE (from 1000 mL) SPE (from 100 mL)

MeOH – 0.01 mol L-1 NaOH (1:9)

3×10-9

DPV at HMDE

SPE (from 1000 mL)

MeOH – 0.01 mol L-1 NaOH (1:9)

3×10-10

DPV at HMDE

[80, 81] [82-84] [85, 86] [78, 87]

-8

DPV at HMDE

-1

[79]

-8

DPV at HMDE

DPV at HMDE 2-Nitronaphthalene

-1

Ref.

[88]

[89, 90]

[89, 91]

Polarographic and Voltammetric Determination of Genotoxic Substances…

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Figure 1. DP voltammograms of MNQ in BR buffer pH 5.0 – methanol mixture (1:1), resulting pH 6.2. Concentration of MNQ in drinking water (mol L-1): 0 (1), 0.5×10-6 (2), 1.0×10-6 (3), 1.5×10-6 (4), 2.0×10-6 (5), 2.5×10-6 (6).

Figure 2. AdS voltammograms of MNQ in BR buffer pH 5.0 – methanol mixture (99:1), resulting pH 5.8. Concentration of MNQ in drinking water (mol L-1): 0 (1), 2×10-8 (2), 4×10-8 (3), 6×10-8 (4), 8×10-8 (5), 1×10-7 (6); Eacc = -200 mV, tacc = 30 s.

LLE with hexane has been used to preconcentrate MNQ from drinking water to reach lower LQ. 1000 mL of MNQ solution in drinking water in concentrations (2 – 10)×10-9 or (2 – 10)×10-10 mol L-1 was extracted to 50 mL of hexane. Than hexane was evaporated to dryness

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and the residue after evaporation was dissolved in 0.5 or 5.0 mL of methanol (using Vortex stirrer), respectively, and the sample was filled up with BR buffer pH 5.0 to the volume of 1.0 or 10.0 mL, respectively (resulting pH 5.9). The samples were transferred to the polarographic vessel and DP voltammograms were measured (after the oxygen removal). Corresponding voltammograms are depicted in Figure 3 (residue after evaporation dissolved in 1 mL of supporting electrolyte) and Figure 4 (residue after evaporation dissolved in 10 mL of supporting electrolyte).

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Figure 3. DP voltammograms of MNQ in BR buffer pH 5.0 – methanol mixture (1:1), resulting pH 5.9. Concentration of MNQ in drinking water (mol L-1): 0 (1), 2×10-9 (2), 4×10-9 (3), 6×10-9 (4), 8×10-9 (5), 1×10-8 (6).

Figure 4. DP voltammograms of MNQ in BR buffer pH 5.0 – methanol mixture (1:1), resulting pH 5.9. Concentration of MNQ in drinking water (mol L-1): 0 (1), 2×10-10 (2), 4×10-10 (3), 6×10-10 (4), 8×10-10 (5), 1×10-9 (6).

Described sensitive determination of MNQ represents one of many developed voltammetric methods for the determination of genotoxic substances in drinking water. A

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review of methods developed in UNESCO Laboratory of Environmental Electrochemistry is summarized in Table V.

4.2.3. Insecticides, Herbicides and Agrochemicals The determination of trace amounts of various agrochemicals (pesticides, growth stimulators etc.) in foodstuffs, soils, natural waters and body fluids continues to be important task of analytical chemistry. However, despite the fact, that many agrochemicals are directly reducible at the DME, relatively few determinations appeared in recent literature [101, 102]. Polarography and voltammetry are usually used after appropriate sample preparation (preliminary separation, clean-up and preconcentration) [103]. Nitrated dipropylaniline, ethylaniline, pentylaniline, phenyloxime and glycolate pesticides were determined in artificially contaminated soils by DPV and AdSV [104] and some substituted s-triazine herbicides [105, 106] were determined at nanomolar concentrations by these methods at the HMDE. Many other examples can be found in reviews [102, 103]. DPP method was applied for the determination of dinobuton in agricultural formulations and in spiked water samples [107] and for monitoring of the photochemical degradation of metamitron and imidacloprid [108]. Oil-in-water emulsions were used as suitable working media for the direct polarographic determination of aziprotryne and desmetryne from its organic extracts in water samples [109]. Polarographic determination of insecticides is demonstrated on two different types of insecticides, i.e. cyfluthrin and buprofezin. Both developed methods were successfully applied for the determination in spiked drinking water. The polarographic behavior of cyfluthrin, an α-cynoester pyrethroid, was studied using a DME and HMDE in methanolic BR buffer of pH 2.0 – 12.0 with different ionic media. The nature of the electrode process was examined, the number of electrons was evaluated, and the reduction mechanism was proposed. Quantitative determination was achieved in the concentration range of 6.0×10-8 to 1.15×10-5 mol L-1 using a DPP method with detection limit of 2.4×10-8 mol L-1. The proposed method was successfully applied for the determination of cyfluthrin in formulations, grains, soils, and spiked water samples [110]. A sensitive method for the determination of buprofezin at nanomolar level by AdSV at a HMDE was described. The cyclic voltammograms demonstrate the adsorption of this compound at the mercury electrode. A systematic study of the various operational parameters that affect the stripping response was carried out by DPV. With an Eacc = -0.8 V and a tacc = 60 s, the limit of detection was 2.2 μg L-1 (7.2×10-9 mol L-1), and the relative standard deviation (n = 5) was 3.4% at the concentration level of 5.0×10-7 mol L-1 of buprofezin. The degree of interference from diverse ions and some other pesticides on the differential pulse stripping signal for buprofezin was evaluated. The method was applied for determination of buprofezin in spiked soil, drinking water, and treated wastewater [111]. The methods of determination in drinking water have been developed for herbicides propazine, bromfenoxim, chloridazon, fluchloralin, methyltin, diquat, fenchlorazole-ethyl, propazine, terbutryn and 2-methyl-4,6-dinitrophenol (summary of obtained results see in Table VI).

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Table V. Voltammetric determination of selected genotoxic heterocyclic compounds in drinking water Substance

6-Methyl-5-nitroquinoline

6-Methyl-5-nitrouracil

Preconcentration

DPV at HMDE

direct determination

AdSV at HMDEa

direct determination

5-Nitroindazole

5-Nitroquinoline

8-Nitroquinoline b

LQ [mol L-1]

Medium

DPV at HMDE

b

hexane LLE

drinking water – MeOH – BR buffer pH 5 (4:5:1) drinking water – MeOH – BR buffer pH 5 (90:1:9) MeOH – BR buffer pH 5 (1:1)

DPV at HMDE

c

hexane LLE

MeOH – BR buffer pH 5 (5:95)

2×10-10

direct determination

drinking water – BR buffer pH 7 (9:1)

3×10-7

DPV at HMDE

5-Nitrobenzimidazole

a

Technique

-1

DPV at HMDE

direct determination

drinking water – 0.1 mol L NaOH (9:1) drinking water – MeOH – 0.1 mol L-1 NaOH (4:5:1) MeOH – 0.01 mol L-1 NaOH (1:1)

DPV at HMDE

direct determination

DPV at HMDE

SPE (from 100 mL)

DPV at HMDE

SPE (from 500 mL)

DPV at HMDE

direct determination

DPV at HMDE

SPE (from 100 mL)

DPV at HMDE

direct determination

DPV at HMDE

SPE (from 100 mL)

MeOH – 0.01 mol L-1 NaOH (1:1) drinking water – 2 mol L-1 NaOH (9:1) MeOH – 0.2 mol L-1 NaOH (1:9) drinking water – MeOH – BR buffer pH 4 (5:1:4) MeOH – BR buffer pH 4 (1:1)

DPV at HMDE

SPE (from 200 mL)

MeOH – BR buffer pH 4 (1:1) c

3×10-7 2×10-8

Ref.

[94]

-9

4×10

4×10

-8

[94] –

1×10-7 2×10-8

[98]

2×10-9 2×10-8

[81, 99]

3×10-9 9×10-8 1×10-8

[100]

2×10-9

– Eacc = -200 mV, tacc = 30 s; – substance after evaporation dissolved in 10 mL of supporting electrolyte; – substance after evaporation dissolved in 1 mL of supporting electrolyte.

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Table VI. Polarographic and voltammetric determination of selected pesticides in drinking water Substance

Technique

Preconcentration

Medium -1

bromofenoxim buprofezin chloridazon

SPE (from 100 mL) a

AdSV at HMDE

fast-scan DPV at HMDE

direct determination dichloromethane LLE

0.1 mol L LiClO4 in acetonitrile – 10-3 mol L-1 HClO4 EtOH – BR buffer pH 7 (2:8) citric acid or HCl at pH 2.3

Ref.

1×10-10

[113]

-9 x

[111]

-8

[114]

-8 x

7×10

3×10

cyfluthrin

DPP at DME

direct determination

MeOH – BR buffer (2 – 12)

2×10

[110]

diquat

AdSV at HMDEb

direct determination

water pH 7, ionic strength 0.1 mol L-1

3×10-8 x

[46]

2-methyl-4,6-dinitrophenol fenchlorazole-ethyl methyltin (monomethyltin) methyltin (dimethyltin) methyltin (trimethyltin) propazine terbutryn a

SWV at HMDE

LD [mol L-1]

-8

DPV at HMDE

direct determination

drinking water – BR buffer pH 6 (1:9)

6×10

DPP at DME

SPE (from 100 mL)

BR buffer pH 7

2×10-8

DPP at DME

SPE (from 500 mL)

BR buffer pH 7

2×10-9

AdSV at HMDEc

SPE (from 1000 mL)

aqueous buffered medium

5×10-10 x

-1

DPASV at HMDE

direct determination

MeOH – 0.05 mol L TP (2:8)

1×10

direct determination

MeOH – 0.05 mol L-1 TP (2:8)

7×10-7 x

DPASV at HMDE

direct determination

MeOH – 0.05 mol L-1 TP (2:8)

2×10-7 x

DPP at DME

direct determination

MeOH – 0.05 mol L TP (2:8)

4×10

direct determination

MeOH – 0.05 mol L-1 TP (2:8)

1×10-6 x

DPP at DME

SPE

micellar media – ethyl acetate

2×10-7

d

direct determination

micellar media – 0.1 mol L HClO4

2×10

AdSV at HMDEe

direct determination

emulsified media – 0.1 mol L-1 HClO4

1×10-9

AdSV at HMDE

[116]

-6 x

DPASV at HMDE

-1

[117]

-7 x

DPP at DME

-1

[119-121]

[112]

-9

[118]

– Eacc = -800 mV, tacc = 60 s; b – Eacc = -800 mV, tacc = 60 s; c – Eacc = -100 mV, tacc = 600 s; d – Eacc = -600 mV, tacc = 50 s; e – Eacc = -600 mV, tacc = 70 s; x – the limit of quantification in distilled water, the method was applied for the determination of the substance in spiked drinking water.

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An electroanalytical study of the herbicide propazine's reduction process in micellar solutions and oil-in-water emulsions is reported [112]. The anionic surfactant, sodium pentanesulfonate, was chosen as the most suitable one. DP polarograms of micellar solutions had 2 reduction peaks at pH < 2.0, whereas only 1 peak was obtained at pH > 2.0. Ethylacetate was chosen as the organic solvent to form propazine emulsions. Unlike in micellar solutions, the DP polarograms of propazine emulsions showed only 1 peak even at pH < 2.0, suggesting that propazine hydrolysis was hindered in the emulsified medium. The limiting current is diffusion-controlled and the electrode process is irreversible. Propazine can be determined by DPP over the 1.0×10-6 to 1.0×10-5 mol L-1 and 1.0×10-5 to 4.0×10-5 mol L-1 concentration ranges; LD was 2.8×10-7. Of potential interferents, simazine, methoprotryne, and terbutryn (all s-traizines), thiram (a dithiocarbamate), dinoseb (nitrophenol), and heptachlor (chlorinated cyclodiene herbicide), only the first 2 were significant (10% error for equimolar concentrations). The method was used for determination of propazine in spiked drinking water. At a concentration of 2.0×10-7 mol L-1, a recovery of 94 ± 6% was obtained, after tenfold pre-concentration on Sep-Pak [112]. The application of SPE as a preconcentration and clean-up step with subsequent off-line flow injection analysis with amperometric detection (FIA-AD) or batch SWV detection of the herbicide bromofenoxim was developed. The selection of an appropriate organic eluent, some parameters influencing the efficiency of the SPE and the electrochemical detection of bromofenoxim in the organic effluent solution were thoroughly investigated. Undiluted acetonitrile for SWV and acetonitrile-water (80:20) for FIA-AD, both containing 0.1 mol L-1 LiClO4, were chosen as the most appropriate SPE eluents. The addition of LiClO4 as supporting electrolyte to the eluent and acidification of a water sample to 1x10-3 mol L-1 HClO4 (pH 3) prior to SPE procedure improved greatly the current response on mercury drop (for SWV detection) and mercury film (for FIA-AD) electrodes. Subsequent to SPE procedure, the effluents were transferred to the voltammetric cell or injected into the flow injection system without any further treatment. The calibration plots obtained for bromofenoxim in pure water samples were linear over the ranges 0.2 – 12.0 μg L-1 and 3.0 – 120 μg L-1, with calculated LD of 0.05 and 1.5 μg L-1 (100 mL samples), for the SPE-SWV and SPE-FIA-AD procedures, respectively. The actual detection capabilities of the proposed methods depend on the water sample volumes applied to the extraction cartridges. The recoveries of the over-all procedures, applying spiked drinking water samples, and the corresponding relative standard deviations were 92% and 6% (n = 6) and 121% and 9% (n = 7) for SPE-SWV and SPE-FIA-AD, respectively. The practical applicability of the proposed methods for the analysis of ground and drinking water samples was confirmed via an interlaboratory test on a drinking water sample containing five common pesticides including bromofenoxim [113]. The electrochemical behavior of the herbicide chloridazon (pyrazon), at different pH was described [114]. The electrode reaction (one wave in acidic media and another one in alkaline media), investigated using direct current and pulse voltammetry, controlled-potential coulometry, and HPLC-MS, is a combination of the reduction (two-electron in the first step) and a kinetic process as a result of which simple compounds (HCl, NH3) are released and, moreover, a five-member pyrrole cycle is formed in strongly acidic media. Products of subsequent reaction are further reducible. Fast-scan differential pulse voltammetry was used for determination of chloridazon; the detection limit was 2.7×10-8 mol L-1 (0.006 μg L-1) at pH 2.3. Chloridazon was determined in spiked drinking and river water [114].

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The electrochemical reduction of fluchloralin was studied employing DC polarography, cyclic voltammetry, and coulometry in 25% N,N-dimethylformamide in the pH range 3.0 – 8.0. A single, well defined, irreversible, and diffusion controlled, 8-electron cathodic wave/peak was observed, due to the simultaneous reduction of the two nitro groups, to give dihydroxylamine. The method is applicable to the analysis of drinking water samples [115]. The electrochemical behavior of monomethyltin, dimethyltin and trimethyltin compounds in 20% (v/v) methanol – water solution, 0.05 mol L-1 in tetraethylammonium perchlorate at pH 2.5, was studied by DPP and DPASV. In DPP, dimethyltin and trimethyltin gave one reversible peak at -0.70 V and -1.07 V, respectively. Detection limits were 6.6×10-7 mol L-1 for dimethyltin and 4.1×10-6 mol L-1 for trimethyltin. The electrochemistry of monomethyltin is more complex. Using DPASV, monomethyltin, dimethyltin and trimethyltin produced distinct stripping peaks (-0.39 V, -0.75 V and -1.14 V, respectively), which enabled determination of those compounds at trace levels. Detection limits were: 1.2×10-7 mol L-1 for monomethyltin, 1.7×10-7 mol L-1 for dimethyltin and 1.4×10-6 mol L-1 for trimethyltin. For monomethyltin, the second peak (E = -0.60 V), producing detection limit of 2.5×10-6 mol L-1, was also observed at concentrations > 4.2×10-7 mol L-1. Recoveries of methyltin compounds added separately to drinking water samples at the 0.42 – 16.9 μmol L-1 level ranged from 84.5 to 99.8% depending upon the methyltin species [116]. A sensitive analytical procedure for diquat herbicide in drinking water and soil samples, using AdSV has been developed. This involved the adsorptive accumulation of a diquat onto a HMDE (Eacc = -0.8 V vs Ag/AgCl, tacc = 60 s, pH 7.0, ionic strength 0.1 mol L-1) for AdSV measurement. The LD for diquat herbicide was as low as 0.034 μmol L-1 [46]. A very sensitive procedure was presented for determination of fenchlorazole-ethyl by AdSV using disposable Carbopack SPE columns for its isolation from drinking water. The stripping response was evaluated with respect to pH, accumulation time, potential, and mercury drop size. The detectable level for fenchlorazol-ethyl after 10 min accumulation at -0.1 V was 0.2 μg L-1 in 1 L of water. The procedure was applied to spiked drinking water [117]. The possibility of applying AdSV in dispersed media has been evaluated. Sensitive methods for the determination of the herbicide terbutryn at nanomolar levels by AdSV at the HMDE in micellar and emulsified media were proposed. The anionic surfactant sodium pentanesulfonate was chosen as the most suitable agent for micellar solutions and oil-water emulsions. Ethylacetate was used as the organic solvent to form terbutryn emulsions. A systematic study of the various experimental parameters that affect the stripping response was carried out by DPV. When working in a 0.1 mol L-1 perchloric acid medium with Eacc = -0.60 V, terbutryn could be determined over the 6.0×10-9 – 4.0×10-7 and 6.0×10-9 – 2.0×10-7 mol L-1 ranges in micellar (tacc = 50 s) and emulsified (tacc = 70 s) media, respectively. The LD were 2.2×10-9 and 1.1×10-9 mol L-1, and the relative standard deviations (n = 10) were 6.3 and 4.3% in micellar and emulsified media, respectively. The effect of other herbicides on the terbutryn stripping peak was evaluated. The developed method in emulsified medium was applied to the determination of terbutryn in spiked drinking waters [118]. Optimum conditions are described for the determination of the toxic pesticide 2-methyl4,6-dinitrophenol using DPP at a DME after preliminary separation and preconcentration using SPE on Lichrolut EN (200 mg) SPE column. The LD is around 2×10-9 mol L-1 both for drinking and river water. DPV at HMDE was used to the direct determination of 2-methyl4,6-dinitrophenol in drinking water (LD is around 6×10-8 mol L-1) [119-121].

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CONCLUSION Drinking water, as the most important liquid in the human life, needs a permanent quality monitoring. Especially in present days, when the living environment is daily attacked by various pollutants, truly said mainly from human activities. For the sake of environmental protection, reliable methods for drinking water quality monitoring are requested. It was the aim of this article to show that for some dangerous analytes and drinking water possible pollutants, polarographic and voltammetric methods at mercury electrodes may be the "best method" and can successfully compete with more widely used separation and spectrometric techniques. Moreover, in many other cases, modern polarographic and voltammetric techniques can be among “fit for the purpose” methods. Lower investment and running costs, speed, sensitivity, universality and wide applicability speaks in favor of polarographic techniques in spite of their limited selectivity. To increase the use of polarography and voltammetry in modern analytical laboratories and in environmental analysis would require to improve education in this field and to pay more attention to the validation of newly developed methods. This fact concerns not only to drinking water analysis, but to environmental or industrial analysis as well. Only then it can be expected that polarographic and voltammetric methods will play a useful role in analytical laboratories even in the third millennium.

ACKNOWLEDGMENT

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Authors thank the Ministry of Education, Youth and Sports of the Czech Republic (projects LC06035 and MSM 0021620857) and the Grant Agency of Charles University (project 6107/2007/B-Ch/PrF) for the financial support.

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In: Progress on Drinking Water Research Editors: M. H. Lefebvre and M. M. Roux

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Chapter 6

GROUNDWATER TOXICITY DUE TO NATURAL DISSOLVED RADIONUCLIDES BELONGING TO THE U AND TH DECAY SERIES Daniel Marcos Bonotto∗ Departamento de Petrologia e Metalogenia, Universidade Estadual Paulista (UNESP), Câmpus de Rio Claro, Av. 24-A No.1515, C.P. 178, CEP 13506-900, Rio Claro, São Paulo, Brasil

ABSTRACT

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Primeval radionuclides have survived in detectable amounts since the time of nucleosynthesis and contribute to the terrestrial gamma radiation, whose major contribution comes from 40K and three radioactive series having unstable members with half-lives much shorter than that of each precursor, i.e. 232Th, 238U, and 235U. Thorium and uranium are lithophile elements, being concentrated preferentially in acid igneous rocks compared with intermediate, basic, and ultrabasic varities. 232Th decays to the stable 208Pb, after 12 disintegrations (7 alpha-type and 5 beta-type). 238U is the principal isotope of natural U (99.72% abundance) and decays to the stable 206Pb, after 14 disintegrations (8 alpha-type and 6 beta-type). Potential health hazards from some natural radionuclides belonging to these decay series in consuming water have been considered worldwide, with many countries adopting the WHO guideline activity concentration for drinking water quality. In general, the recommendations apply to routine operational conditions of water supply systems. Several national standards for limiting radiation exposure establish maximum permissible radionuclides concentration in drinking water, where, in general, for practical purposes, 0.5 Bq/L for gross alpha and 1 Bq/L for gross beta activity have been used to routine operational conditions of existing or new water supplies. However, additional information concerning to specific radionuclides may be required in some circumstances in order to decide about the drinking water quality in terms of radiological aspects. Special attention must be given when groundwaters are utilized for public water supplies, because water-rock/soil interactions may enhance the presence of natural radionuclides in solution. This chapter reports the radiological ∗

E-mail: [email protected]

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Daniel Marcos Bonotto implications of the interactions between waters and different rock types occurring in aquifer systems exploited in Brazil. The data here shown demonstrate how some natural radionuclides contribute to the groundwater toxicity evaluated in terms of the WHO guideline activity concentration for drinking water quality.

INTRODUCTION Primeval radionuclides have survived in detectable amounts since the time of nucleosynthesis and contribute to the natural terrestrial radioactivity. The major contribution comes from 40K and from radionuclides generated through the sequence of decay transformations of the three alpha emitting radionuclides: 232Th, 238U, and 235U. Each radioactive series ends by a stable Pb isotope after passing through several unstable members having half-lives much shorter than that of the respective precursors. The members in each decay chain have atomic numbers between 81 and 92. Potassium is a major element widely distributed in crustal rocks, for instance, calcium rich granites may contain up to 2.5% K (Cox, 1991). Thus, potassium occurs in various minerals (like the feldspars orthoclase and microcline) and clays, from which it may be dissolved through weathering processes and transferred into the liquid phase. Potassium reacts rapidly and intensely with water, forming a colourless basic potassium hydroxide solution and hydrogen gas, according to the following reaction mechanism (Greenwood and Earnshaw, 2002): 2K(s) + 2H2O(l) → 2KOH(aq) + H2(g)

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40

(1)

K is the only radioactive isotope of potassium and is present in an amount of 0.0119% in this natural element (Davis, 1963). Further potassium includes isotopes 39K and 41K with abundancies of 93% and 6.9%, respectively (Davis, 1963). 40K decays directly to 40Ca in the ground state through β- emission (89%) and also to 40Ar in a 1.46 MeV excited state followed by a prompt 1.46 MeV gamma emission through electron capture (11%) (Adams and Gasparini, 1970). As a consequence of water/rock-soil interactions, 40K is released to water bodies, contributing to the presence of radioactive constituents of drinking water. In general, it is often impractical to use a radioactive measurement technique to determine the concentration of 40K in a water sample due to the lack of sensitivity in gamma-ray analysis and the difficulty of chemically isolating the radionuclide from solution. Therefore, WHO (2004) recommended the chemical analysis for potassium by traditional methods like atomic absorption spectrophotometry or specific ion electrode and the use of an appropriate factor to convert the total K to the 40K activity concentration, because of the fixed ratio between 40K and stable K. The natural radioelements uranium and thorium are lithophile elements widely distributed in crustal rocks, being concentrated preferentially in acid igneous rocks compared with intermediate, basic, and ultrabasic varities. There are three natural isotopes of uranium: 238U, 235 U, and 234U. 238U is the principal isotope of natural U in a mixture (about 99.3% abundance) and is the parent nuclide in the mass number (4n+2) decay series of radionuclides, the longest known series. 234U is radiogenic, being the fourth member of the

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201

U radioactive decay chain; its abundance in a mixture of natural U is 0.0054%. 235U is the primary nuclide in the mass number (4n+3) decay series of radioelements, and its proportion in a mixture of natural U can often be considered as 0.72% if exceptional cases like the 235U fission reactor site at Oklo, Gabon (Cowan, 1979) are disregarded. The 238U decay series finishes at the stable 206Pb, according to the sequence (Adams and Gasparini, 1970): 238U (4.49 Ga, α) → 234Th (24.1 d, β-) → 234Pa (1.18 min, β- ) → 234U (0.248 Ma, α) → 230Th (75.2 ka, α) → 226Ra (1622 a, α) → 222Rn (3.83 d, α) → 218Po (3.05 min, α) → 214Pb (26.8 min, β-) → 214Bi (19.7 min, β-) → 214Po (0.16 ms, α) → 210Pb (22.26 a, β-) → 210Bi (5 d, β-) → 210Po (138 d, α) → 206Pb. The 235U decay series finishes at the stable 207Pb, according to the sequence (Adams and Gasparini, 1970): 235U (0.71 Ga, α) → 231Th (25.5 h, β-) → 231Pa (34.8 ka, α) → 227Ac (21.6 a, β- ) → 227Th (18.2 d, α) → 223Ra (11.4 d, α) → 219Rn (4 s, α) → 215Po (1.8 ms, α) → 211Pb (36.1 min, β-) → 211Bi (2.16 min, α) → 207Tl (4.8 min, β-) → 207Pb. Since its discovery by Joly in 1908, radioactive disequilibrium in the 238U decay series has been extensively studied in almost all superficial environments. Because 234U is a heavy radiogenic isotope, it is essentially immune to mass-fractionation effects. Despite this, studies on secondary minerals and groundwaters have found radioactive disequilibrium between 238U and 234U (Cherdyntsev, 1971). The isotopes 238U and 234U are in secular equilibrium only in minerals and rocks which are closed systems for U over a time-scale greater than 106 years. Considering that the 234U/238U activity ratio is defined as the ratio of the 234U daughter activity to the 238U parent activity, an unit value for this parameter is achieved in the bulk of rock/mineral matrices under these conditions. However, water/rock-soil interactions frequently result in 234U/238U activity ratios for dissolved uranium which are greater than unity (Osmond and Cowart, 1976; Ivanovich and Harmon, 1992). Because the 235U natural isotopic ratio relative to 238U is often 0.72 and, consequently, the disintegration rate of 235U in natural U occurring in minerals, soils and rocks is only 4.6% of that of 238U, the presence of 235U and descendants in the liquid phase during water/rock-soil interactions has been generally disregarded in guidelines for drinking water quality due to natural radioactivity (WHO, 2004). Thorium is an element ~ 3-4 times more abundant than uranium in crustal rocks because it is less susceptible to mobilization in the supergene environment. It occurs predominantly as a tetravalent cation and a trace constituent in phosphates, simple and multiple oxides and silicates, as well in the major, rock-forming minerals such as monazite, thorianite (ThO2) and thorite (ThSiO4), among others (Gascoyne, 1992). The major natural Th isotope is 232Th that is precursor of the mass number 4n decay series. The 232Th half life is 14.1×109 years, which is longer than that of 238U (4.5×109 years) and 235U (0.25×109 years). The 232Th decay series finishes at the stable 208Pb, according to the sequence (Adams and Gasparini, 1970): 232Th (14.1 Ga, α) → 228Ra (6.7 a, β-) → 228Ac (6.1 h, β- ) → 228Th (1.9 a, α) → 224Ra (3.6 d, α) → 220Rn (55.3 s, α) → 216Po (0.14 s, α) → 212Pb (10.6 h, β-) → 212Bi (60.6 min, β-) → 212Po (0.3 µs, α) → 208Pb (or … → 212Bi (60.6 min, α) → 208Tl (3.1 min, β-) → 208Pb). WHO (2004) defined guidance levels in drinking water for both 232Th and 238U, as well for several of their daughters, for example, 228Th, 228Ra, 234U, 230Th, 226Ra, 222Rn, 210Pb, and 210 Po. Therefore, if it is warranted to determine the activity concentration of each radionuclide, then, it is necessary to utilize several sophisticated and time-consuming

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procedures. Thus, for practical purposes, the national standards for limiting radiation exposure establish maximum permissible radionuclides concentration in drinking water corresponding to 0.5 Bq/L for gross alpha and 1 Bq/L for gross beta activity (WHO, 2004). The detection of gross alpha and beta radioactivities in water is an US Environmental Protection Agency mandated program in the context of National Primary Drinking Water Regulations (EPA, 1997). The radiological criteria for drinking water quality in Brazil are often related to the Rule No. 36 (19 January 1990) of the Health Ministry, which defines that the identification of the dissolved radionuclides and the measurement of their activity concentrations in the samples should be performed only when the values found in them are greater than 0.1 Bq/L for the gross alpha and 1 Bq/L for the gross beta activity concentration. Therefore, only when the results of a previous screening are positive, it is performed the identification of specific radionuclides. As a consequence of the technical difficulties, in general, there is a lack of integrated studies focusing the gross alpha and beta measurements coupled to the major radionuclides dissolved in water bodies occurring in South America. This chapter presents data on the activity concentration of some specific radionuclides providing from different aquifer systems exploited in Brazil for supplying water for human consumption. The implications of the reported data in terms of standards established for defining the drinking water quality are also presented.

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DETECTION OF THE PARTICLES AND RADIATION EMITTED BY MAJOR NATURALLY OCCURRING RADIONUCLIDES The parent nuclides of the 238U, 235U and 232Th decay series are emitters of alpha particles. Disregarding the main complexities on the decay modes in the 238U decay series, the stable 206Pb is attained after 14 disintegrations, 8 alpha-type and 6 beta-type. The stable 207Pb in the 235U decay series is attained after 11 disintegrations (7 alpha-type and 4 beta-type), whereas the stable 208Pb in the 232Th decay series is attained after 12 disintegrations (7 alphatype and 5 beta-type). Following the alpha and beta disintegrations, it is also verified the occurrence of gamma emissions in each decay series. Gamma rays are generally characterized as electromagnetic radiation having the highest frequency and energy, and also the shortest wavelength, within the electromagnetic spectrum, i.e. high energy photons. Due to their high energy content, they can cause serious damage when absorbed by living cells. Therefore, the major contribution to the terrestrial gamma radiation field comes from 40K and from radionuclides generated through the sequence of decay transformations of 238U, 235U, and 232Th. The gamma rays spectrometry is a non-destructive method that has a lot of advantages from the technical point of view over beta spectrometry, inclusive allowing the identification and quantification of β—emitters radionuclides. The NaI(Tl) scintillation detectors have been extensively used for characterizing the natural gamma radiation, which interacts with the crystal atoms through three major processes: photoelectric effect, Compton effect, and pair production. Figure 1 illustrates a complete spectrometric system utilizing a 2”×2” NaI(Tl) scintillation detector coupled to a 2,048 multichannel analyzer.

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Figure 1. A typical spectrometric system for detecting gamma rays.

The spectrometric system for performing gamma readings must be calibrated in energy by the use of different radioactive sources. Among the typical sources often utilized are: 137Cs (γ-rays energy = 0.66 MeV), 60Co (γ-rays energy = 1.17 and 1.33 MeV), solutions containing 133 Ba (γ-rays energy = 0.36 and 0.39 MeV), and pure powdered KCl (52 wt % in K) as a source of 40K (γ-rays energy = 1.46 MeV). Some gamma spectra obtained for these radionuclides are shown in Figures 2 to 5. The channels identified in the gamma spectra and the corresponding gamma rays energy for any specific radionuclide are used to trace the energy calibration curve of the gamma spectrometer, as illustrated in Figure 6, which may be written according to the equation E = 0.001256 Ch, where E is the energy (in MeV) and Ch is the channel number in the multichannel analyzer. The major β--emitting nuclides in the 238U decay series are 234Th, 234Pa, 214Pb, 214Bi, 210 Pb, and 210Bi. Because no gamma emission has been reported by Adams and Gasparini (1970) for 234Pa and 210Bi, according to data reported in Table 1, the most appropriate gamma rays lines for identifying β--emitting nuclides in the 238U decay series are that of 0.295 MeV (214Pb), 0.352 MeV (214Pb), 0.609 MeV (214Bi), 1.12 MeV (214Bi), 1.26 MeV (214Bi), 1.40 MeV (214Bi), and 1.76 MeV (214Bi). Figure 7 shows a gamma spectrum obtained for a pitchblende sample from Poços de Caldas plateau, Minas Gerais State, Brazil. Several 214Bi photopeaks have been identified in Figure 7, as this radionuclide is a 238U-descendant with many gamma ray emissions. The Poços de Caldas plateau is an area well known in the literature because the levels of background radiation are relatively high, i.e. the exposure for the general population may be up to 10 times higher than the average background level of 2.4 mSv (WHO, 2004). Pitchblende is a uranium mineral, and the short counting time yielded high counts in the 214Bi photopeaks shown in Figure 7.

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Figure 2. Typical spectrum for radionuclide 137Cs utilized for calibrating a NaI(Tl) gamma spectrometer.

Figure 3. Typical spectrum for radionuclide 60Cs utilized for calibrating a NaI(Tl) gamma spectrometer.

Figure 4. Typical spectrum for radionuclide 133Ba utilized for calibrating a NaI(Tl) gamma spectrometer.

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Figure 5. Typical spectrum for radionuclide 40K utilized for calibrating a NaI(Tl) gamma spectrometer.

Figure 6. The typical calibration curve of a NaI(Tl) spectrometer for γ-rays energy readings.

Figure 7. Gamma spectrum obtained for a pitchblende sample from Poços de Caldas plateau, Minas Gerais State, Brazil.

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Table 1. Gamma emissions related to β--emitting radionuclides in the 238U, 235U and 232 Th decay series. Data reported by Adams and Gasparini (1970) 232

238

U decay series Nuclide Energy

Absolut

Nuclide

Energy

Absolute

U decay series Nuclide Energy

(MeV)

Intensity

(MeV)

Intensity

Th

0.026

12

214

Pb

214

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Bi

210

Pb

Absolute

(MeV)

Intensity

0.030

*

228

0.010

*

0.064

3.5

228

0.058

*

0.059

*

0.093

4.0

0.129

4.1

0.082

*

(%) 234

235

Th decay series

(%) Ra Ac

0.053

~1.0

0.209

1.0

0.242

4.0

0.270

3.0

0.295

19.0

0.328

4.0

0.352

36.0

0.338

11.0

0.609

47.0

0.908

25.0

0.665

2.3

0.960 + +

20.0

0.769

5.3

0.966

0.787

1.2

0.805

1.5

0.935

3.3

1.120

16.0

1.155

1.8

212

Pb

212

Bi

208

Tl

0.115

*

0.239

47.0

0.300

3.2

0.785

1.1

1.620

1.8

1.260

7.7

0.277

7.0

1.400

8.8

0.511

23.0

1.510

2.4

0.583

86.0

1.660

1.2

0.763

2.0

1.730

3.2

0.860

12.0

1.760

17.0

2.615

100.0

1.850

2.3

2.120

1.3

2.200

6.0

2.440

2.0

0.047

4.1

(%) 231

Th

*non-reported value.

The major β--emitting nuclides in the 235U decay series are 231Th, 227Ac, 211Pb, and 207Tl. Because no gamma emission has been reported by Adams and Gasparini (1970) for 227Ac, 211 Pb, and 207Tl, according to data reported in Table 1, the gamma rays line of 0.026 MeV for 231 Th could be used for identifying beta radioactivity in the 235U decay series, but such value is not appropriate because it is very low, i.e. it is inserted within the energy range of the Xrays. The main β--emitting nuclides in the 232Th decay series are 228Ra, 228Ac, 212Pb, 212Bi, and 208 Tl. Table 1 reports the gamma rays energy for the β--emitting nuclides in the 232Th decay

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series. The best gamma rays lines for identifying β--emitting nuclides in the 232Th decay series are that of 0.239 MeV (212Pb), 0.277 MeV (208Tl), 0.338 MeV (228Ac), 0.511 MeV (208Tl), 0.583 MeV (208Tl), 0.860 MeV (208Tl), 0.908 MeV (228Ac), 0.960 + 0.966 MeV (228Ac), and 2.615 MeV (208Tl). Information about the presence of 228Ra in waters may be obtained from 228Ac by gamma spectrometry as also demonstrated Mancini and Bonotto (2002). Figure 8 shows a gamma spectrum obtained for monazite sand from Guarapari beaches, Espírito Santo State, Brazil. Several 208Tl photopeaks have been identified in Figure 8, as this radionuclide is a 232Th-descendant with many gamma ray emissions. The Guarapari beaches are also known in the literature due to the high levels of background radiation (WHO, 2004). Monazite sand exhibits thorium accumulation, justifying the high counts in the 208Tl photopeaks shown in Figure 8.

Figure 8. Gamma spectrum obtained for a monazite sand sample from Guarapari beaches, Espírito Santo State, Brazil.

The alpha spectrometry is another useful technique often utilized to characterize the presence of natural dissolved radionuclides in groundwater. It is based on the direct measurement of the alpha particles generated in the 238U, 235U and 232Th decay series (Table 2). The Si(Au) surface barrier detectors have been widely used for characterizing the natural radionuclides, which are inserted within vacuum chambers in order to avoid absorption of alpha particles by the atmosphere. Figure 9 illustrates a complete spectrometric system utilizing a detector coupled to a 1,024 multichannel analyzer. The alpha spectrometric systems are available for a variable number of Si(Au) surface barrier detectors, depletion depths and surface area in order to yield appropriate spectra for providing information about specific radionuclides. The α-spectra have been sometimes recorded on EGandG ORTEC 919 Spectrum Master Multichannel Buffer that provides four 1,024-channel analyzers controlled by MAESTRO software.

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Daniel Marcos Bonotto Table 2. Alpha particles generated in the 238U, 235U and 232Th decay series. Data reported by Adams and Gasparini (1970)

238

U decay series Nuclide Energy (MeV) 238

U

234

U

230

Th

226

Ra

222

Rn Po 214 Po 210 Po 218

4.15 4.20 4.72 4.77 4.62 4.69 4.60 4.78 5.49 6.00 7.69 5.31

235

Absolute Intensity (%) 23.0 77.0 28.0 72.0 24.0 76.0 5.4 94.6 100.0 ~100.0 100.0 100.0

U decay series Nuclide Energy (MeV)

235

U

231

Pa

227

Th

223

Ra

219

Rn

215

Po Bi

Absolute Intensity (%) 84 10.5 11 22 47 10 14 21 24 23 35 63 8 11 81 100 16 84

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4.40 4.55 4.73 4.95 5.02 5.06 5.71 5.76 5.98 6.04 5.57 5.73 6.42 6.55 6.82 7.38 6.28 6.62

232

Figure 9. A typical spectrometric system for detecting alpha particles.

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Th decay series Nuclide Energy (MeV)

232

Th

228

Th

224

Ra

220

Rn Po 212 Bi 216

212

Po

3.95 4.01 5.34 5.43 5.45 5.68 6.29 6.78 6.05 6.09 8.78

Absolute Intensity (%) 24.0 76.0 28.0 71.0 6.0 94.0 100.0 100.0 23.5 9.0 100.0

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209

The system must also be calibrated in energy on using different radioactive sources. One appropriate source has been prepared for this purpose at the Centre de Faibles Radioactivités, CNRS-CEA, Gif-sur-Yvette, France, and contains 0.91 Bq of 238U. In addition, the same source includes 234U (α-particles energy = 4.77 MeV) and 0.17 Bq of 232U, that is an artificial uranium isotope produced in nuclear reactor by neutron activation of 232Th, has a half-life of 72 years, and originates 228Th by the emission of α-particles of 5.3 MeV energy (Lederer et al., 1967). The 228Th-daughters correspond to those occurring in the 232Th decay series. The alpha spectrum obtained for natural uranium isotopes 238U and 234U plus 232U and descendants is shown in Figure 10. The channels identified in the alpha spectrum and the corresponding α-particles energy for the identified radionuclides allows to trace the energy calibration curve of the alpha spectrometer (Figure 11), which may be written according to the equation E = 0.01075 Ch, where E is the energy (in MeV) and Ch is the channel number in the multichannel analyzer.

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Figure 10. The spectrum for α-particles emitted by 238U, 234U, and 232U+daughters in a radioactive source prepared at the Centre de Faibles Radioactivités, France.

Figure 11. The typical calibration curve of a Si(Au) spectrometer for α-particles energy readings.

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The use of the alpha spectrometry allows the direct identification of several radionuclides belonging to 238U and 235U decays series. Bonotto et al. (1993) utilized α-spectroscopy with a gridded ionization chamber over the 3.5-8 MeV range of energy to study nodular uraninite from Campo do Cercado, Poços de Caldas plateau. A typical alpha spectrum from a 300 mesh crushed sample deposited on stainless steel disk (0.21 mg/cm2) is shown in Figure 12. All high intensity alpha-emitter nuclides of the 238U decay series are displayed, and 230Th, 234U and 226Ra can be viewed only as one peak due to the proximity of the energy of their alphaparticles and the resolution of the detector used. 211Bi and 219Rn are the unique nuclides of the 235 U decay series which were identified, because the energy of their alpha-particles is very distinct relatively to those of the alpha-emitter nuclides belonging to the 238U decay series.

Figure 12. The alpha spectrum for nodular uraninite from Campo do Cercado, Poços de Caldas plateau, Brazil.

Gamma emissions related to α-emitting radionuclides also occur in the 238U, 235U and Th decay series, as reported in Table 3. In some cases, there is similarity among the gamma rays energy derived by beta and alpha emitters radionuclides, for instance, 212Pb and 224Ra, 214 Pb and 227Th, 214Pb and 231Pa, 214Pb and 211Bi, etc. Despite gamma spectra of natural 232

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samples often display the overlap of photopeaks related to 40K, 232Th+daughters, 238 U+daughters, and 235U+daughters, the presence of K, Th, and U may be unequivocally detected even in low (8-10%) resolution systems based on NaI(Tl) scintillation detectors due to the presence of some distinct gamma rays energy, for instance, 1.46 MeV [K (40K)], 1.76 MeV [U (214Bi)], and 2.615 MeV [Th (208Tl)] (Adams and Gasparini, 1970). Table 3. Gamma emissions related to α-emitting radionuclides in the 238U, 235U and 232Th decay series. Data reported by Adams and Gasparini (1970) 232

238

U decay series Nuclide Energy (MeV)

Absolute (M V( Intensity

Nuclide

Energy

Absolute

U decay series Nuclide Energy

Absolute

(MeV)

Intensity

(MeV)

Intensity

(%) 238

U

234 230

U Th

226

Ra

235

Th decay series

(%)

(%)

*

232

0.053

*

228

0.084

1.6

0.068

*

224

0.241

3.7

231

4.0

212

0.040 0.727

0.048

0.186

Th Th Ra Bi

0.059

*

235

U

0.143

11.0

0.185

54.0

0.290

6.0

2.0

227

0.237

15.0

7.1

223

0.147

12.0

0.270

10.0

211

0.351

14.0

Pa Th Ra Bi

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*non-reported value.

The dissolved radionuclides present in aquifer systems have different chemical properties and specific procedures must be used for performing their extraction from water samples before conducting gamma-rays or alpha-particles analyses. They are generally found in very low concentrations and, therefore, it is a common practice to collect great volume (20-30 L) of water samples for the analysis of each radionuclide. The samples are stored in polyethylene bottles and often filtered through a 0.45 μm Millipore membrane. The volume sampled for 228Ra analysis should be divided into two aliquots of equal weight, with the 133Ba radioactive tracer being added to one aliquot. Both aliquots are acidified to pH less than 2 using HCl, and about 500 mg of FeCl3 are added to each aliquot. Radium is co-precipitated on Fe(OH)3 by increasing the pH to 7-8 through addition of concentrated NH4OH solution. The precipitate is recovered, dissolved in 8M HCl, and Fe3+ extracted into an equal volume of isopropyl ether. The acid solution containing radium has to be evaporated to dryness, and the dry residue dissolved with pure distilled water to a volume of 15 mL. The solution is stored in an appropriate cylindrical glass bottle that is inserted into a NaI(Tl) well-type scintillation detector. The data acquisition is performed taking into account the condition of secular radioactive equilibrium between 228Ra and its direct descendant, 228Ac, which is reached in approximately 45 h. The aliquots in duplicate allow correct the overlap of the 133Ba peaks with the 228Ac low-energy peak, according to the procedure described by Mancini and Bonotto (2002). The aliquots (20-25 kg) for U and Th analysis are also acidified to pH less than 2 on using HCl. About 500 mg of FeCl3 plus 56 mBq of 232U-228Th are added, and U-Th coprecipitated on Fe(OH)3 by increasing the pH to 7-8 through addition of concentrated NH4OH solution. The precipitated is recovered, dissolved in 8M HCl and Fe3+ extracted into an equal

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volume of isopropyl ether. U is separated from Th by anion exchange. The acid U and Thbearing solution may be purified by anion exchange, first on a Cl- and then on a NO3- column of 50-100 mesh Rexyn 201 resin. U is finally eluted from the NO3- column with 0.1 M HCl and after evaporation to dryness is dissolved in 10 mL of 2M (NH4)2SO4 electrolyte and transferred to an electrodeposition cell. Th in eluent may be again submitted to ion exchange steps using different reagents. The sorbed Th may be eluted with 2 M HCl and after evaporation to dryness dissolved in 10 mL of 2M (NH4)2SO4 electrolyte and transferred to the same electrodeposition cell. The pH should be adjusted to 2.4 and electrodeposition of U on a stainless steel planchet is complete after 3 hours at a current density of 1 Acm-2. The Th deposition on the planchet is complete after 5 hours at a pH = 2. The 238U and 232Th activity concentration may be conveniently determined by alpha spectrometry. The data may be obtained by isotope dilution from the counting rates of 238U, 232 Th, 232U and 228Th peaks in the alpha spectra for U and Th extracted. The 234U activity concentration is calculated from the 234U/238U activity ratio, whereas the 228Th activity concentration may be determined from the 228Th/232Th activity ratio. The presence of natural 228 Th in the samples may be properly evaluated if the water samples are processed in duplicate. 230Th may be determined from the 232Th activity concentration. More analytical details for these measurements have been reported elsewhere (Ivanovich and Harmon, 1992).

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THE EXTRACTION AND DETECTION OF DISSOLVED RADON Radon (222Rn, half-life 3.84 days) is a colorless, odorless, tasteless, chemically inert and radioactive volatile noble gas produced continuously in rocks and soils through α-decay of 226 Ra, with some atoms escaping to the surrounding fluid phase, such as groundwater and air. Its characterization in waters is important because it may utilized as an isotopic tracer in hydrologic investigations (Hoehn and von Gunten, 1989) and also in order to assure the drinking water quality, since although radon in water is not a well-documented health risk, it does contribute to radon in indoor air, which has been established as a health threat (Nazaroff and Nero, 1988). A convenient method often used for quantifying radon dissolved in water is based on the emanation procedure (Lucas, 1957) that consists on its removal from the sample, its transfer to a scintillation flask, and its detection by alpha-scintillation counting. The α-counting is carried out following a delay of about 3 h from the Rn transfer to permit total ingrowth of the short-lived 222Rn daughters, with the flask being placed on the end of a photomultiplier tube covered with a light-tight lid. A counter/timer is then used for recording the total number of pulses, being performed a calibration in order to convert this parameter in a value of activity concentration. The generation of 222Rn data through this technique is based on the premisse of ingrowth of their short-lived daughters up to an activity value corresponding to that of the secular radioactive equilibrium, i.e. the short-lived 222Rn daughters are not directly identified and quantified during the measurement process. The groundwater samples (1 kg) are collected in glass bottles fitted with inlet and outlet stopcocks (Zereshki, 1983). Radon is outgassed from the water samples by passing a stream of 222Rn-free air through the sample. Quantitative recovery of 222Rn is achieved after passing a volume of air equal to 10 times the water volume, where the gas was passed through the

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sample at a rate of about one liter per minute. Three stages are used for drying the gas stream before reaching a charcoal trap. In the first stage, the gas is passed through a tube immersed in an ice container to condense the major part of the water vapor present. The second stage is a drying tube full of 3-8 mesh CaCl2, and the third drying stage is a Dreschel bottle containing 300 mL concentrated sulfuric acid. After drying the gas stream, it is passed through a charcoal trap at -80 °C (solid CO2/acetone) to absorb the 222Rn (Figure 13).

Figure 13. Sketch diagram of a typical system for absorbing radon from water samples.

A 50 mL conical flask, fitted with an appropriate socket, can be used to load radon (Figure 14). A few grammes of ZnS(Ag), an α-sensitive scintillator, are placed inside the flask. Before loading, each flask is evacuated to less than 50 μm of Hg and then connected to the charcoal trap at -80 °C, containing the radon. After taking care with some connections to the vacuum system, the charcoal trap is taken out of the cold trap and placed inside a tube furnace at 200 °C for 15 minutes. A needle valve allows flush the last of the radon into the scintillation flask (Figure 15). A suitable technique used for quantifying 222Rn is the alpha-scintillation counting. The αcounting has to be carried out following a delay of about 3 h from the Rn transfer to permit total ingrowth of the short-lived 222Rn daughters. The flask is placed on the end of a 3” diameter photomultiplier tube covered with a light-tight lid. A high voltage of 1,000 V is applied to the photomultiplier assembly, the pulses amplified and a counter/timer used for recording the data. Immediately before and after each sample count, a standard 232U-228Th scintillation source or equivalent should be counted to take account of any changes in the response of the instrumentation.

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Figure 14. Sketch diagram of a scintillation flask for storing radon.

Figure 15. Sketch diagram of the system for transferring radon to the scintillation flasks.

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The scintillation flasks are calibrated on using standard 226Ra solutions of activity corresponding to 37 Bq. Since 226Ra is progenitor of 222Rn in the 238U-decay series, the time te for 222Rn to reach radioactive equilibrium with 226Ra is: te = - (1/λ222) × ln (1 - A2/A6)

(2)

where λ222 is the 222Rn decay constant, and A2 and A6 correspond to the activities of 222Rn and 226 Ra, respectively. This time is about 25 days, after of which it must be performed the extraction and transfer of radon to the scintillation flask. The counting rate of the background is also evaluated, and often varies between 1 and 3 cpm. It allows to calculate the normalized activity (An) (the counting rate due to the standard minus the counting rate due to the background) and, consequently, the corrected activity (Ad) due to the decay of 222Rn during the time t elapsed between the beginning and final of the extraction procedure, i.e.:

Ad = An e λ222

t

(3)

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The activity Ad allows finding the factor Fk representing the calibration of the flask. For the groundwaters analyzed, the parameter t in equation 3 corresponds to the time elapsed between the sampling and ending of the measurement. The activity concentration of the samples is obtained dividing Ad by Fk. The statistical uncertainty of the activity determinations is often within 5% and for most measurements is 3%. 222 Rn data obtained in 5-L aliquots provided 226Ra activity concentration values. The samples were inserted in glass bottles fitted with inlet and outlet stopcocks, as well outgassed with 222Rn-free N2 to remove the 222Rn originally present in the sample. The radon removal was performed only after waiting a time of about 25 days for 222Rn to reach radioactive equilibrium with 226Ra.

THE TRANSFER OF NATURAL RADIONUCLIDES TO THE WATERS The uranium distribution in terrestrial crust is directly related to magmatic activities that caused its mobilization since the formation of primeval Earth. The transfer process is global rather than local and remains active even in actual days. According to meteoritic models for the Earth composition, uranium is enriched about 230 times in crust and 3 times in upper mantle, whereas it is depleted 3 times in lower mantle and 25 times in nucleus (Gabelman, 1977). Therefore, uranium is a lithophile element that migrates towards crust accompanying light silicates instead of directing to nucleus under gravity action. Such anomalous behavior is a consequence of its ionic radius and tendency to volatize or forming volatile combinations at lower temperatures comparatively to other metals in nucleus, as well to its strong tendency to combine with fluoride, chlorine and oxygen (Gabelman, 1977). Uranium occurs in crustal rocks at an average concentration of about 2.5 µg/g (Bowie and Plant, 1983). Other rock types exhibit the following average U concentration (Gabelman, 1977): sandstone = 1.4 µg/g; grayish schist = 4.2 µg/g; carbonaceous schist = 53 µg/g; limestone = 1.9 µg/g; riolite = 5.0 µg/g; granite = 3.6 µg/g; phonolite and syenite = 6.5 µg/g; alkali basalt = 0.99 µg/g; gabbro = 0.84 µg/g; andesite =0.79 µg/g; peridotite = 0.01 µg/g.

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Uranium may reach a maximum enrichment of 500 times in granite and 650 times in syenite relatively to rocks representing the mantle composition like amphibolite, granulite, eclogite and dunite (Gabelman, 1977). In crystalline rocks, the most of the uranium is incorporated into accessory minerals such as monazite, allanite, sphene, and zircon so that uranium is not readily accessible for solution and available to secondary mineralization processes (Gabelman, 1977; Speer et al., 1981). The typical uranium concentration in some minerals is (Gabelman, 1977): quartz = 1.7 µg/g; feldspars = 2.7 µg/g; biotite = 8.1 µg/g; muscovite = 2.8 µg/g; hornblende = 0.2 – 60 µg/g; pyroxene = 0.1 – 50 µg/g; olivine = 0.05 µg/g; allanite = 30 – 1000 µg/g; apatite = 10 – 100 µg/g; epidote = 20 – 200 µg/g; garnet = 6 – 30 µg/g; huttonite = 3 – 70000 µg/g; magnetite = 1 – 30 µg/g; monazite = 500 – 3000 µg/g; titanite = 10 – 700 µg/g; xenotime = 300 – 40000 µg/g; zircon = 100 – 6000 µg/g. Uranium is also a major constituent in almost 100 minerals, occurring mainly in uraninite, UO2, and its non crystalline variety pitchblende. The ratio oxygen/uranium may alter in dioxide (UO2), modifying the composition to U3O8 (Krauskopf and Bird, 1995). The presence of U6+ in uraninite is a consequence of the oxidation process (Goldschmidt, 1954). Additionally, uranium may also associate genetically to limestones, phosphates, vanadates, silicates, sulphides, sulfates, among others compounds (Frondel, 1956). Uranium in the major, rock-forming minerals is more susceptible to leaching, particularly as the rock disintegrates during weathering (Gavshin et al., 1997). Brown and Silver (1955) studied the distribution of uranium and thorium in igneous rocks and concluded that less than one-third of the uranium is present as interstitial oxide or cryptocrystalline aggregate and available for leaching. Adams et al. (1959) estimated that 60-85% of the Th and U in igneous rocks is contained in resistates that are incorporated intact into sedimentary derivatives. Several alternative mechanisms have been suggested to explain the generation of the enhanced 234U/238U activity ratios in solution. For instance, Rosholt et al. (1963) proposed the occurrence of enhanced chemical solution of 234U due to radiation damage of crystal lattices or to autoxidation from U4+ to U6+ on decay of the parent 238U. This model was supported by the results of Chalov and Merkulova (1966) who obtained activity ratios of up to 1.3 by leaching fresh, unaltered igneous rocks and showed that in 6 of 12 samples of uranium minerals 234U6+ is enhanced relative to 238U6+. The experiments performed by Kigoshi (1971) and Fleischer and Raabe (1978a,b) showed that alpha-particle recoil ejection of the 234U precursor, 234Th, into solution may also generate enhanced activity ratios in the liquid phase. According to this model, if the interstices of a rock or assemblage of mineral grains are permeated with water, the water will absorb the recoiling, short lived, 234Th nucleus from 238U decay, thereby enriching the solution and depleting the solid in 234U. Kigoshi (1971) also recognized the possibility that alpha-recoil tracks intersecting the surface of grains could provide paths of rapid diffusion that could allow 234U to diffuse outwards and be accessible to water that may later enter the pore spaces. Fleischer (1975) suggested a similar model involving the presence of alpha-recoil tracks, except that the damage track is removed chemically by the intergranular liquid: the recoil 234U nuclei are implanted when the interstitial space of the solid phase was dry during 238U decay, and chemical dissolution of the damage track after infiltration of water removes the 234U by etch solution. Some experimental evidence for this process was given by Fleischer (1980, 1982). Leaching experiments on igneous rocks have been performed under situations that approximate natural conditions, with the aim of determining the mobility of uranium

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(Kovalev and Malyasova, 1971; Zielinski et al., 1981; Michel, 1984; Eyal and Olander, 1990). An acid (1-6N HCl) or carbonate [0.05-5% NaHCO3 or (NH4)2CO3] solution has been generally utilized for leaching powdered samples in order to characterize the mobile uranium. For example, Kovalev and Malyasova (1971) in multiple leaching experiments on various igneous rock types (granite, andesite, diorite, gabbro, basalt, etc.) using a solution of 5% (NH4)2CO3 found the greatest values (20-80%) for uranium leached from granites. Using 2% NaHCO3, a substantial amount of U in solution was readsorbed onto the rock particles, implying that adsorbed, mobile, leachable, and acid-soluble uranium cannot be distinguished, with acid-soluble uranium being considered to range between 20 and 70% (Szalay and Sansoni, 1966). Michel (1984) considered leachable uranium to be the fraction that is soluble under the experimental conditions, no matter how it is distributed in the rock or its behavior in nature, and thus the leachable uranium from parent rock may or may not represent uranium available for groundwater dissolution and transport. Leaching experiments involving 238U and 234 U uranium isotopes were reported by Zielinski et al. (1981) who found that after 20 hours, 2-6% of the total U was removed, with an average 234U/238U activity ratio of 1.72 in leach solutions. Eyal and Olander (1990) studied eight separate leaching series of monazite specimens in a bicarbonate-carbonate solution for durations up to 6.8 years, and observed a fractionation of 234U relative to 238U by a factor of 1.1 to 2 during the initial leaching period. The effect of chemical etching and leaching on the 234U/238U activity ratio of dissolved U in the etch/leach solution in karstic limestone matrices from Mendip Hills, England, and in Carnmenellis granite gravels from Cornwall, England, has been investigated on a laboratory time-scale by Bonotto and Andrews (2000, 2001). The implications of the results obtained in the laboratorial experiments for the production of enhanced activity ratios in natural groundwaters from the studied areas were also evaluated. Uranium has several oxidation states, but the most extensively investigated are +4 and +6 (Langmuir, 1978). It is an element very sensitive to modifications on redox potentials, since if oxidizing conditions prevail, active solution of U occurs, and if the redox character changes towards reducing conditions, then, deposition of U takes place; at low temperatures, the U(IV) amorphous oxyhydroxides generally precipitate instead of the crystalline UO2 (Langmuir, 1978). Uranium that goes into solution can migrate over long distances, mainly due to its ability to form complexes with bicarbonate and carbonate ions, which are most stable in solutions where the pH is greater than 7.5 (Langmuir, 1978). In general, complexation of the uranyl ion (UO22+) with bicarbonate/carbonate ions has been recognized as a very important mechanism for the transport of U in solution (Langmuir, 1978). Thorium is an element ~ 3-4 times more abundant than uranium in crustal rocks because it is less susceptible to mobilization in the supergene environment. It occurs predominantly as a tetravalent cation and a trace constituent in phosphates, simple and multiple oxides and silicates, as well in the major, rock-forming minerals such as monazite, thorianite (ThO2) and thorite (ThSiO4), among others (Gascoyne, 1992). Thorium has been considered an element highly insoluble in water due to its presence in minerals of difficult dissolution, however, the migration is enhanced when inorganic/organic complexes are formed with Cl , NO3 , H3PO40, -

2

-

-

2-

H2PO4 , SO4 , F , OH e HPO4 , oxalates, citrates, and EDTA (Langmuir and Herman, 1980). Thorium mobilization also is affected by pH and ionic strength, and this element also occurs

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adsorbed in inorganic (clays and colloids) and organic suspended matter (Langmuir and Herman, 1980; Ferronsky and Polyakov, 1982). Radium has about 25 isotopes with mass numbers between 206 and 230. All are unstable and only four occur in the natural series of radioactive decay: 226Ra (half life=1622 years) in the 238U series, 223Ra (half life= 11.1 days) in the 235U series, 228Ra (half life= 6.7 years) and 224 Ra (half life=3.64 days) in the 232Th series. Among these radium isotopes, only 228Ra is a beta emitter, the others being alpha emitters (Ku and Broecker, 1976). Radium in the environment may be released due to the interaction of waters with rocks, soils or mineralized bodies, where the mining and processing of phosphate minerals, apatite, copper, gold, lignite, coal and bauxite also can contribute to the enrichment of Ra in the superficial and underground waters, since it is present in the U–Th decay series (Iyengar, 1984, 1990; Jaworowski, 1990; Dickson, 1990). The radioactive gas 222Rn is produced continuously in rocks and soils through α-decay of 226 Ra, with some atoms escaping to the surrounding fluid phase, such as groundwater and air. It is subjected to recoil at “birth”, with the emanated fraction relatively to that produced in the solid phase being dependent on factors such as total surface area of solids and concentration/distribution of 238U (226Ra) in the minerals (Flügge and Zimens, 1939). High 222Rn concentrations occur in groundwaters in many areas where wells are used for domestic water supply, inclusive in small rural water supplies (Petrie and Cram, 1999). Exposure to radon and its progeny is believed to be associated with increased risks of several kinds of cancer (USEPA, 1999). When radon or its progeny are inhaled, lung cancer accounts for most of the total incremental cancer risk, while ingestion of radon in water is suspected of being associated with increased risk of tumors of several internal organs, primarily the stomach (USEPA, 1999). Inhalation of radon progeny accounts for about 89% of the individual risk associated with domestic water use, with almost 11% resulting from directly ingesting radon in drinking water (USEPA, 1999). Gesell and Prichard (1975, 1980), Partridge et al. (1979), Bruno (1983), Prichard (1987), Nazaroff and Nero (1988) and Nazaroff et al. (1988) also pointed out that although radon in water is not a well-documented health risk, it does contribute to radon in indoor air, which has been established as a health threat. Furthermore, high concentrations of 222Rn in groundwaters also indicate the presence of radon’s parent nuclides, 238U and 226Ra, in the water-rock system, which are known health risks when ingested in drinking water (Aieta et al., 1987). Beyond the importance to develop effective methods for reducing or eliminating radon from homes, the knowledge of the geochemical and hydrogeological mechanisms that control the movement of 222Rn through the groundwater system may be informative of the natural processes related to its high concentration associated with low transmissivity zones (Lawrence et al., 1991), with the uranium content of the source rock, severe chemical weathering, hydrothermal solution, deposition, or extensive fracturing (Nelson et al., 1983), and with variations in stress in rocks associated with seismicity, where in some cases temporal variations have preceded large earthquakes (Shishkevich, 1971; Wollenberg et al., 1984; Smith et al., 1976). Several field and laboratory studies have been conducted and many models for radon generation have been proposed, most of which are applicable for a single field area or have been developed using microscopic properties. For instance, Nelson et al. (1983) proposed a simple crack model to understand the relation between flow rate and radon concentration, where water flows radially inward from all directions through a thin flat disk into the

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borehole, to be collected within the central hole; different assumptions for mixing and for entry point locations were performed. Rama and Moore (1984) suggested that radon diffuses to the atmosphere or water through a network of 100 - 200 Å wide nanopores in rocks and minerals, many of which intersect the grain surface, and give rise to a large internal surface area of the solids (in the range of about 10 m2 per cm3), orders of magnitude higher than the external surface calculated from grain size. Krishnaswami and Seidemann (1988) pointed out that the high 222Rn leakage from rocks and minerals responsible for regulating the supply of this gas to the surrounding fluid phases such as air and groundwater is related to 226Ra enrichment in grain boundaries or crevices instead of an extensive network of nanopores, which could increase their total surface area several fold. Semkow (1990) evaluated the mechanism of α-recoil to account for Rn emanation from solids, emphasizing the basic phenomena related to single mineral grains, since soils and rocks could be considered multigrain aggregates; the author derived formulas for the emanation from cylinders and spheres having non-uniform Ra distribution or variable thickness of Ra-containing material. The use of models emphasizing mainly microscopic properties such as the network of nanopores, inhomogeneous 226Ra distribution in the solid or surface roughness in natural systems is, in general, difficult or impossible to be carried out. Thus, more practical theories for radon emanation from rocks to groundwater may have applications, because they could allow prediction of levels of radon in groundwater if the uranium concentration of the aquifer rock and other parameters of simple evaluation are known. Such approach was successfully utilized by Bonotto and Andrews (1997) and Bonotto and Caprioglio (2002) to investigate different aquifer systems occurring in England and Brazil. As a consequence of the several mechanisms that have been described in this item, the transfer of the natural radionuclides from rocks and soils to the waters may occur at rates causing their presence in concentrations exceeding the maximum values established by authorities responsible to control the water distribution for human consumption. For instance, WHO (2004) established guidance levels for several radionuclides in drinking water, assuming that 730 liters is the annual ingested volume of water. WHO (2004) proposed the following guidance levels for the activity concentration in drinking water: 238U and 234U = 10 Bq/L; 226Ra, 230Th, 232Th and 228Th = 1 Bq/L; 228Ra = 0.1 Bq/L. In the case of 222Rn, WHO (2004) proposed that controls should be implemented if the 222Rn concentration of drinkingwater for public water supplies exceeds 100 Bq/L, i.e. treatment of the water source should be undertaken to reduce the 222Rn levels to well below 100 Bq/L. These guideline values will be compared with the activity concentration obtained for several groundwater samples collected in different aquifer systems exploited in Brazil.

THE MAJOR FEATURES OF THE AREAS INVESTIGATED In order to evaluate the water quality in terms of natural dissolved radionuclides, several groundwater samples were collected at São Paulo and Minas Gerais States in Brazil. The water samples were taken from springs, monitoring wells and taps installed in pumped tubular wells drilled at different aquifer systems. The sampling sites were chosen in accordance with the hydrochemical information available from previous studies, as well with the facilities available for sampling. The collection sites are located at Poços de Caldas city in Minas

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Gerais State and at the following municipalities in São Paulo State: Águas de Lindóia, Águas da Prata, Águas de São Pedro and Rio Claro (Figure 16).

Figure 16. Simplified sketch map showing the location of the sampling points for groundwater analysis of dissolved radionuclides.

The areas studied include rocks of the crystalline basement varying from Paleo to Neoproterozoic, volcanic rocks associated to the caldera subsidence at Poços de Caldas alkaline massif and rocks belonging to the sedimentary sequence at the Paraná basin. Most of the water sources are used for drinking purposes or health treatment in thermal and nonthermal spas, and their chemical composition is attained due to processes occurring at the liquid-solid interface when different rock matrices are extensively leached, for instance, sandstones, conglomerates, diamictites, siltstones, shales, mudstones, limestones, basalts, diabases, granites, migmatites, phonolites, nepheline syenites, etc. The evolution of Águas de Lindóia region is characterized by the occurrence of several phases and cycles, involving different aspects of metamorphism, deformation and magmatism which difficult its delineation, reconstitution of the sequences and primary characterization of the rocks (Ebert, 1955). These events acted from the Archean to the Upper Proterozoic times and affected rocks characterized by high metamorphic grade, generally of granulite and amphibolites facies (Almeida and Hasui, 1984). The first geochronological studies were

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realized by Delhal et al. (1969) and Cordani et al. (1973) that attributed Archean 2,800 Ma), Transamazônica (2,070 Ma) and Brasiliana (620 – 470 Ma) ages. Wernick and Penalva (1974) found an age of 2,010 Ma for migmatites dated by the Rb/Sr method and Artur et al. (1979) determined an older age of 2,500 Ma for biotite/hornblende gneisses and granitic gneisses. Zanardo (1987) described the major rock types occurring at Águas de Lindóia region. recent alluvium related to flood plains and consisting on immature sediments like sand, silt, clays and organic matter; various types of milonites cutting diagonally the area in a strip 5001000 m thick; quartzites, quartz schists, schists, gneisses, calcium-silicate rocks, anfibolites, ultramafic rocks; migmatites and syn tectonic granites; biotite and hornblende gneisses with homogeneous structure; gneisses and pink migmatites with homogeneous structure; grayish gneissed migmatites with folded structure tending to homogeneous types. The Monte Sião fault is an important structural element to understand the geological evolution of Águas de Lindóia area (Campanha et al., 1982). It begins at the north border of Morungaba granitic massif, has N10–20E direction, and involves a group of well deformed rocks, the majority represented by specimens enriched in quartz. The lineaments of the faults are disposed in the directions of the Monte Sião fault and control the drainage system at southeast and northwest of the area. Rio Claro and Águas de São Pedro cities are geologically located at the Paraná sedimentary basin. The basin is located in the South American continent, between parallels 10°-20° southern latitude and meridians 47°-64° western longitude. Its surface area corresponds to 1.6×106 km2, comprising southern Brazil (1×106 km2 in the states of Mato Grosso, Mato Grosso do Sul, Goiás, Minas Gerais, São Paulo, Paraná, Santa Catarina and Rio Grande do Sul), eastern Paraguay (0.1×106 km2), NW Uruguay (0.1×106 km2) and the northeastern extreme corner of Argentina (0.4×106 km2) (Araújo et al., 1995). The altitudes in Paraná basin vary from 100 to 1800 m, the annual average rainfall corresponds to 100-240 cm, the soils are mainly lateritic, and the vegetation is often represented by grass-covered fields used for pasture and cerrado fields. The Paraná basin is intercratonic, where the sedimentary sequence covers since the Silurian-Devonian up to the Cretaceous periods. The sequence is almost undisturbed, with gentle dips towards the center of the basin. Local faults may have served as channels for the extruding basalt flows of Jurassic-Cretaceous age (Serra Geral Formation), which cover about 1×106 km2 of the basin. Renne et al. (1992) utilized the 40Ar-39Ar technique to estimate volcanism ages between 127 and 137 Ma, and Turner et al. (1994) suggested that the duration of the volcanism corresponded to 10 Ma, occurring at a rate of 0.1 km3yr-1. The major stratigraphic units occurring at the Paraná basin are (IPT, 1981): the Tubarão Group comprising the Itararé Subgroup (sandstones, conglomerates, diamictites, tillites, siltstones, shales and rythmites) and Tatuí Formation (siltstones, shales, silex and sandstones with local concretions); the Passa Dois Group comprising the Irati Formation (siltstones, mudstones, black betuminous shales and limestones) and Corumbataí Formation (mudstones, shales and siltstones); the São Bento Group comprising the Pirambóia Formation (sandstones, shales and muddy sandstones), Botucatu Formation (sandstones and muddy sandstones), Serra Geral Formation (basalts and diabases) and related basic intrusives; different types of Cenozoic covers like the recent deposits, terrace sediments and the Rio Claro Formation (sandstones, conglomerate sandstones and muddy sandstones).

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Águas da Prata and Poços de Caldas cities are important spas located in São Paulo and Minas Gerais States, respectively. They are geologically situated in the Poços de Caldas alkaline complex that is circularly shaped (Figure 17), with a mean diameter of 33 km. The total surface area is about 800 km2, the altitude varies between 1300 and 1600 m, and the topography is characterized by valleys, mountains, and gentle grass-covered hills. The plateau is a ring structure of Mesozoic age comprising a suite of alkaline volcanic and plutonic rocks, mainly phonolites and nepheline syenites. The evolutionary history, according to Ellert (1959), starts with major early volcanism involving ankaratrites (biotite-bearing nephelinite), phonolite lavas, and volcano-clastics, followed by caldera subsidence and nepheline syenite intrusions forming minor ring dykes and circular structures and, finally, the intrusion of eudialite-bearing nepheline syenites. This early model has been partly confirmed by the geochronological work of Bushee (1971) and the structural interpretations of Almeida Filho and Paradella (1977).

Figure 17. Location of Águas da Prata and Poços de Caldas cities at the Poços de Caldas alkaline massif.

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The average annual temperature at the Poços de Caldas plateau is 19°C, with maximum around 36°C and minimum around 1°C, indicating the possibility of frost during colder climatic phases in the past (Amaral et al., 1985). The regional weather pattern is dominated by wet and dry seasons, with more than 80% of the precipitation falling between October and March (Lei, 1984). Amaral et al. (1985) record 1500 mm as the total maximum annual rainfall, with more than 120 days of rain each year. The regional 30-yr annual average precipitation corresponds to 170 cm (Lei, 1984). Soils tend to be uniform at the Poços de Caldas plateau, due to deep weathering, and the land surface is homogeneous in character. The shrub “campo cerrado” vegetation tends to be remarkably uniform, occurring mainly in the southern portion of the plateau (Holmes et al., 1992). The other main vegetation units are evergreen forest, grass and woodland savanna, however, the annual fire practices reduce the presence of humus in the terrain. Soils are predominantly silts and clays, and because of the intense weathering in a tropical climate, surface soils tend to be a fine-textured mixture of kaolinite, gibbsite and limonite. Fault patterns are an important control on the drainage system of the plateau, occurring different fault systems within the Poços de Caldas complex. Maciel and Cruz (1973) suggested that two large faults striking N30°W and N50°E and crossing close to the center of the caldera delimits two distinct relief areas. Miranda Filho (1983) and Holmes et al. (1992) pointed out three major fault systems within the alkaline massif: the first, which strikes N50°60°W, is the major regional trend as it extends beyond the complex, being reactivated during uplift; the second trend strikes N30°-45°E and is genetically related to the formation of the collapsed caldera, whereas, in the third, there are radial and sub-circular faults, related to various intrusives. Miranda Filho (1983) also considered other secondary fault systems striking N10°E, N20°E, and E-W, which may have controlled the uranium mineralization in the plateau (Osamu Utsumi mine). The Morro do Ferro is a high grade thorium and rare earth elements (REEs) deposit located about 20 km south of the city of Poços de Caldas. The local geology of the Morro do Ferro is characterized by hydrothermally altered country rocks termed "potassic rocks" overlain by a deep weathering cover. Dominant rock-types identified around Morro do Ferro are phonolites, tinguaites and foyaites mineralogically and geochemically comparable to the rocks exposed all over the Poços de Caldas plateau (Barreto and Fujimori, 1986; Ulbrich, 1984; Schorscher and Shea, 1991). The Th-REEs ore body is a 410 m long, 215 m wide and 10-35 m thick zone of NW-SE elongated argillaceous lenses extending from the summit of the hill along its south-eastern slope (Barreto and Fujimori, 1986). The magnetite veins are the dominant features at the surface of Morro do Ferro, being concentrated in the uppermost 40 m of the hill. Águas de Lindóia is another important thermal spa in São Paulo, whose waters are mainly provided by abundant springs occurring there, most of them associated to Monte Sião fault. The water circulation generally realizes through fractures and the flow is from the higher to lower altitudes (Del Rey, 1989). Eight springs located at the urban area were sampled for analyses of 238U and 234U. They are described in Table 4, which shows that Lindália is in fact a well drilled to exploit groundwater. In terms of lithologies, the waters are leaching migmatite (Lindália and Santa Isabel), quartzite (Levíssima I and II, Curie, Filomena and Beleza) and milonite/quartzite (São Roque). São Roque, Curie, Filomena and Beleza

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springs are located in Prof. Dr. João de Aguiar Pupo spa, whereas the others have been utilized by private companies for commercializing mineral waters. Table 4. The springs sampled at Águas de Lindóia city for analysis of the uranium isotopes 238U and 234U. The data are from Del Rey (1989) and Yoshinaga (1990)

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SPRING Lindália Levíssima I Levíssima II Santa Isabel Curie Filomena Beleza São Roque

LITHOLOGY Migmatite Quartzite Quartzite Migmatite Quartzite Quartzite Quartzite Milonite/Quartzite

SURGENCE Well Fracture N80W Fracture N30W Fracture N30W Fracture N50-60W Fracture N35E Fracture N30E Fracture N35E

Rio Claro city is located at the northeastern edge of the Paraná sedimentary basin, cropping out several stratigraphical units of the basin in the region, for example: Itararé SbGroup; Aquidauana and Tatuí from Tubarão Group; Irati and Corumbataí formations from Passa Dois Group; Pirambóia, Botucatu and Serra Geral formations from São Bento Group; intrusive rocks related to the lava flows from Serra Geral Formation and different types of Cenozoic covers such as the Rio Claro Formation, where the urban area is situated. Three aquifer systems were investigated for analysis of the uranium isotopes 238U and 234U. The first consists of a phreatic aquifer comprising a cover of unconsolidated materials from Rio Claro Formation; impermeable sediments from Corumbataí Formation underlie this cover. A dug well for individual water supply intersected the water table of this aquifer at about 10-11 m in the beginning of the wet period however variations between 1 and 3 m have been identified due to seasonal variations (Cottas, 1983). The second consists of sediments from Tubarão Group (Tatuí Formation and upper layers from Itararé Sub Group), and comprises a confined aquifer about 40 m thick, which has a transmissivity of 12.3 m2/h (Cottas, 1983). A 198.5 m deep well for supplying water to a local club was drilled in April 1970 and intersected the water table at 34 m. The third groundwater system comprises deep bodies of diabase, which store water within their fractures. A 60 m deep well for supplying water to the rail company was drilled and intersected the water table at 30 m. Águas de São Pedro is a much known spa at São Paulo State and is geologically situated in the Paraná sedimentary basin. The groundwater samples were collected for analysis of the uranium isotopes 238U and 234U and radioactive noble gas 222Rn from tubular wells drilled in 1936 by DNPM (Departamento Nacional de Produção Mineral) for petroleum exploration. These groundwaters have often been referred to as springs (Kimmelmann et al., 1987), whose names are Gioconda, Juventude, and Almeida Salles. The major characteristics of the wells are shown in Table 5. Several springs were sampled at Águas da Prata city for analyses of the uranium isotopes 238 U and 234U, dissolved gas 222Rn and thorium isotopes 232Th, 228Th and 230Th: Villela, Vitória, Platina, Prata Antiga, Prata Nova, Prata Radioativa, Boi and Paiol. Villela spring discharges through vertical and horizontal open fractures into a well silicified and lightly folded sandstone (Szikszay, 1981; Szikszay and Sampa, 1982).

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Table 5. Characteristics of the deep wells drilled at the Tubarão Group and results of chemical analyses. The data are from Kimmelmann et al. (1987) PARAMETER Depth of the well Depth of the top of the aquifer Potentiometric level Depth of the mantle of weathering Depth of Pirambóia Formation Depth of Serra Geral Formation Depth of Passa-Dois Group Depth of Tubarão Group

UNIT m m m m m m m m

GIOCONDA 625 255 >530 0 up to 3 3 up to 68 68 up to 275 275 up to 625

JUVENTUDE 469 290 0 up to 3 3 up to 87 87 up to 240 240 up to 469

This spring is very much known in Brazil due to its high radioactivity related to dissolved Rn (Longo, 1967). Vitória spring discharges through fissures in diabase by ascendant water circulation (Szikszay, 1981). Platina spring discharges through vertical fractures in outcropping phonolites by ascendant water circulation (Szikszay, 1981). Prata Antiga spring was the first identified in the city and the water discharges through fissures and fractures in diabase (Szikszay, 1981). Prata Nova spring corresponds in fact to some wells about 16 m deep which cut diabase and phonolites (Szikszay, 1981). The water in Prata Radioativa spring discharges through fractures in a silicified and recrystallized sandstone, which is the same rock type associated to Boi spring located nearby (Szikszay, 1981). The discharge of Paiol spring occurs through alkaline rocks like volcanic tuffs, phonolites and eudialite-bearing nepheline syenites (Szikszay, 1981). Despite the occurrence of superficial granular aquifers at the Poços de Caldas city area, the drilling companies have mainly exploited aquifers in crystalline fractured rocks due to the better yielding of the wells. Groundwater also occurs in diffuse and punctual thermal and non-thermal springs that discharge at the depression in Poços de Caldas city. Cruz (1987) and Cruz and Peixoto (1989) utilized the water temperature and probable circulation depth as criteria to suggest three main aquifer zones at the Poços de Caldas city area, i.e. a shallower, an intermediate, and a deeper one. The most superficial zone is represented by risings/spring waters whose discharge temperature corresponds to a maximum of 24°C (non-thermal waters). Araújo (1980) suggested these water temperatures are affected by air temperature, whose influence at the Poços de Caldas plateau occurs up to 20-25 m depth. The intermediate aquifer zone is generally exploited by tubular wells drilled up to 200 m depth, whereas the deeper zone is represented by naturally discharging thermal (temperatures higher than 25°C) springs, as the costs for drilling deep wells are high. Etchebehere (1990) proposed a conceptual model for groundwater circulation at the Poços de Caldas city area considering the topographic heights and lineament data provided by the photogeological analysis. Figure 18 illustrates the model which indicates that the thermal waters of Macacos spring and other famous thermal springs discharge at a depressed area in the plateau (altitude = 1190 m) (Etchebehere, 1990). Thermal and non-thermal springs from Poços de Caldas city were sampled for analysis of the uranium isotopes 238U and 234U. Morro do Ferro is one of the highest points on the Poços de Caldas caldera, rising some 140 m above the base (Figure 2). Several boreholes were drilled to investigate the hydrological and chemical character of the subsurface flow on the aquifer system developed in the weathered profile of the Morro do Ferro area. 222

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ALMEIDA SALLES 329 323 >530 0 up to 8 8 up to 22 22 up to 207 207 up to 329

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Figure 18. Conceptual model of groundwater circulation at Poços de Caldas alkaline massif, according to Etchebehere (1990).

Nine were drilled by IPT (1982) to sample the upper part of the saturated zone, from which three in the ore body zone (SR-5, SR-7 and SR-8), four around the ore body zone (SR2, SR-4, SR-6 and SR-9) and two distant from the ore body zone (SR-1 and SR-3). Four other boreholes were drilled along the southeastern slope of Morro do Ferro to study the lower part of the saturated zone (Holmes et al., 1992): MF 10, situated in the ore body zone; MF 11, immediately adjacent to MF 10, to sample groundwater in the vicinity of the water-table for comparison with the deeper water sampled by MF 10; MF 12, at the bottom of the southeastern slope near the south stream, and MF 13, between boreholes MF 10 and MF 12. The water table is a subdued replica of the topography, with recharge on the higher elevations from precipitation and discharge locally to streams (Eisenbud et al., 1979). At the top of the hill, the water table is at least 80 m below the surface and in the valley bottoms is at or near the surface, occurring as seepages or discrete springs. The groundwater samples were analyzed for the uranium isotopes 238U and 234U and radium isotopes 226Ra and 228Ra.

THE WATER QUALITY DUE TO DISSOLVED RADIONUCLIDES Worldwide soluble U content generally ranges from 0.1 to 10 μg/L in rivers, lakes and groundwaters (Fritz and Fontes, 1980; Ivanovich and Harmon, 1992). In terms of 234U/238U activity ratio data, values between 1 and 2 define “normal” worldwide situation reflecting rock-water-rock/soil interactions that frequently result in ARs for dissolved uranium that are greater than unity (Osmond and Cowart, 1976; Ivanovich and Harmon, 1992). The results obtained on the analysis of 238U and 234U in groundwaters of Águas de Lindóia are reported in Tables 6 and 7. The U content is between 0.01 and 0.51 μg/L and the 234U/238U ranges from 0.81 to 3.76 that are not exactly within the range of the most frequent values elsewhere. However, the highest 238U and 234U activity concentration corresponded to 6 and 5 mBq/L,

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respectively, that are below the maximum guideline value of 10 Bq/L estalished by WHO (2004). Therefore, these spring waters have no restrictions for ingestion of dissolved 238U and 234 U.

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Figure 19. Profile across Morro do Ferro showing the mineralization of Th and REEs and the position of the water table during the dry season . Adapted from IPT (1982) and Waber (1991).

The pH of the spring waters of Águas de Lindóia varies between 5.5 and 7.3, whereas the Eh potential redox between +61 and +229 mV. Bicarbonate is the dominant dissolved anion, varying between 6 and 67 mg/L (Yoshinaga, 1990). The dissolved CO2 data reported by Yoshinaga (1990) for these waters allow calculate values between 10-2.5 and 10-1.5 atm (average = 10-1.9 atm) for the CO2 partial pressure. The mean value is close to 10-2 atm that is the typical CO2 partial pressure reported for groundwaters (Langmuir, 1978). Therefore, the pH and Eh data for the spring waters of Águas de Lindóia may be inserted in the Eh-pH diagram representing the stability fields for U4+, U5+ and UO22+ as proposed by Langmuir (1978). The diagram indicates the importance of bicarbonate for the mobilization of uranium in the aquifers occurring at Águas de Lindóia, as it suggests the formation of the soluble complexes UO2CO3° e UO2(CO3)22-. However, despite this, the values obtained in the liquid phase are not high enough to exceed the maximum established by WHO (2004) for the quality of drinking water in terms of the uranium isotopes 238U and 234U. The dissolved U content and 234U/238U activity ratios shown in Table 7 may also be inserted in the diagram proposed by Cowart and Osmond (1980) for the hydrogeochemical prospection of concealed U deposits. The classification “normal reducing” may be attributed to the hydrological system studied, implying that the reducing character dominates the

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environment of the water circulation that would be occurring through aquifer strata exhibiting low U concentration, i.e. not containing any significant accumulation of this element. Table 6. Counts obtained in the major peaks occurring in the alpha spectra for uranium extracted from spring waters of Águas de Lindóia, São Paulo State, Brazil

SPRING

VOLUME (L) COUNTING TIME (s)

NUMBER OF COUNTS 234 238 U U U

232

Santa Isabel

20.6

244,770

1974

168

146

Curie

20.2

240,086

497

71

65

Lindália

20.6

190,363

1764

192

109

Filomena

20.1

343,993

1363

195

187

Levíssima I

20.3

162,768

1876

319

185

Beleza

20.3

90,741

1771

277

194

São Roque

17.1

261,188

1229

201

143

Levíssima II

19.5

181,196

1004

187

259

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Table 7. The activity concentration for 238U and 234U in spring waters of Águas de Lindóia, São Paulo State, Brazil SPRING Santa Isabel Curie Lindália Filomena Levíssima I Beleza São Roque Levíssima II

U (µg/L) 0.07 < 0.01 0.06 0.13 0.16 0.21 0.13 0.51

234

U/238U A.R. 2.12 3.76 2.08 2.28 1.62 3.04 0.81

238

U (mBq/L) 0.84 < 0.12 0.72 1.56 1.92 2.52 1.56 6.12

234

U (mBq/L) 1.78 2.71 3.24 4.38 4.08 4.74 4.96

The results obtained on the analysis of 238U and 234U in groundwaters of Rio Claro are reported in Table 8. The highest 238U and 234U activity concentration corresponded to 3 mBq/L that is also below the maximum guideline value of 10 Bq/L estalished by WHO (2004). Therefore, these spring waters as well have no restrictions for ingestion of dissolved 238 U and 234U. In Brazil, uranium exploration started in 1952 through a collaborative program established with USA (White and Pierson, 1974) and, as a consequence, enriched U concentrations were characterized in Paleozoic sediments from Paraná sedimentary basin,

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among other occurrences. It is situated at Figueira area in the charcoal region of Peixe river (Paraná State), and comprises an area of 700 km2 at southeast of Paraná basin (Ramos and Fraenkel, 1974). Table 8. The activity concentration for 238U and 234U in groundwaters occurring at Rio Claro city, São Paulo State, Brazil Aquifer Rio Claro Formation

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Fractured Diabase

Tubarão Group

Sample No. 1 2 3 4 5 6 7 8 9 1 2 3 4 5 6 7 8 9 1 2 3 4 5 6 7 8 9

Weight (kg) 17.1 16.7 17.5 18.3 17.4 17.6 17.3 17.2 17.2 17.7 17.4 17.5 16.6 19.5 20.6 19.7 19.2 20.6 22.5 20.1 20.0 20.2 18.5 20.4 20.2 20.0 20.0

U (µg/L)