Humic Substances. Nature's Most Versatile Materials

HUMIC SUBSTANCES

Nature’s Most Versatile Materials

HUMIC SUBSTANCES Nature’s Most Versatile Materials Edited by

Elham A.Ghabbour Northeastern University, Boston, USA and Soil, Water and Environmental Research Institute, Alexandria, Egypt Geoffrey Davies Northeastern University, Boston, USA

Based on the proceedings of the 11th Biennial Conference of the International Humic Substances Society and the 6th Humic Substances Seminar held on 21–27 July 2002 at Northeastern University, Boston, Massachusetts, USA

Taylor and Francis, Inc. New York

Denise T.Schanck, Vice President Robert L.Rogers, Editor Liliana Segura, Editorial Assistant Thomas Hastings, Marketing Manager Maria Corpuz, Marketing Assistant Dennis P.Teston, Production Director Anthony Mancini Jr., Production Manager Brandy Mui, STM Production Editor Mark Lerner, Art Manager Daniel Sierra, Cover Designer Published in 2004 by Taylor & Francis 29 West 35th Street New York, NY 10001 This edition published in the Taylor & Francis e-Library, 2005. “To purchase your own copy of this or any of Taylor & Francis or Routledge’s collection of thousands of eBooks please go to www.eBookstore.tandf.co.uk.”

Published in Great Britain by Taylor & Francis 11 New Fetter Lane London EC4P 4EE Copyright © 2004 by Taylor & Francis Books, Inc. All rights reserved. No part of this book may be reprinted or reproduced or utilized in any form or by any electronic, mechanical, or other means, now known or hereafter invented, including photocopying and recording, or in any information storage or retrieval system, without permission in writing from the publisher. Library of Congress Cataloging-in-Publication Data Humic substances: nature’s most versatile materials/edited by Elham A.Ghabbour, Geoffrey Davies. p, cm. “Based on the proceedings of the 11th biennial conference of the International Humic Substances Society and the 6th Humic Substances Seminar held on 21–27 July 2002 at Northeastern University, Boston, Massachusetts, USA.” Includes bibliographical references and index. ISBN 0-59169-015-3 (alk. paper) 1. Humus—Congresses. 2. Soils—Humic contents—Congresses. I. Ghabbour, Elham A. II. Davies, Geoffrey. 1942–III. International Humic Substances Society. IV. Humic Substances Seminar (6th:2002: Northeastern University) S592.8.H92 2003 631.4'′17–dc21 2003053112 ISBN 0-203-48760-5 Master e-book ISBN

ISBN 0-203-59493-2 (Adobe eReader Format)

Contents

Preface

vi

Contributors

x

PART 1. FRACTIONATION AND CHARACTERIZATION: THE STATE-OF-THE-ART 1.

Use of Radioactive Tracers for the Characterization of Humic and Fulvic Acids in High Performance Size Exclusion Chromatography Karsten Franke, Doris Rössler and Hermann Kupsch

2

2.

Interpreting Capillary Electrophoresis—Electrospray/Mass Spectrometry (CZE-ESI/MS) of Suwannee River Natural Organic Matter (NOM) Philippe Schmitt-Kopplin

6

3.

Comparison of As-Delivered and AFFFF-Size-Fractionated Suwannee River Fulvic Acid by Time-ofFlight Mass Spectrometry Wilfried Szymczak, Manfred Wolf andKlaus Wittmaack

21

4.

Molecular Fingerprinting of Aquatic Fulvic Acids by Ultra-High Resolution ESI FT-ICR Mass Spectrometry William T.Cooper and Alexandra C.Stenson

27

5.

The Macromolecular or Supramolecular Nature of Humic Substances: A Dynamic Light Scattering Study Gustavo González-Gaitano and Josemaría García-Mina

35

6.

A Proposal for the Establishment of a Database of Thermodynamic Properties of Natural Organic Matter Rossane C.DeLapp and Eugene J.LeBoeuf

40

PART 2. HYDRATION, SWELLING AND SORPTION: CONTRIBUTING FACTORS 7.

Effect of Hydration/Solvation of Organic Matter on Sorption of Organic Compounds: Conception and Sorption Isotherm Model Ellen R.Graber and Mikhail Borisover

54

8.

Swelling of Organic Matter in Soil and Peat Samples: Insights from Proton Relaxation, Water Absorption and PAH Extraction Gabriele E.Schaumann, Julia Hurrass, Martin Müller and Wolfgang Rotard

65

9.

Sorption of PAHs to Natural Sorbents: Impacts of Humic and Lipid Fractions Luc Tremblay, Scott D.Kohl, James A.Rice and Jean-Pierre Gagné

76

10.

Interactions and Conversions of Polycyclic Aromatic Compounds in the Process of Humification Matthias Hübner, Kristoffer E.N.Jonassen and Torben Nielsen

88

11.

Pyrolytic Study of the Bound Residues of 13C-Atrazine in Soil Size Fractions and Soil Humin Marie-France Dignac, Yahya Zegouagh, Ludovic Loiseau, Gérard Bardoux, Enrique Barriuso, Sylvie Derenne, André Mariotti and Claude Largeau

101

12.

Phenanthrene Sorption by Clay-Humic Complexes Kaijun Wang, Elham A.Ghabbour, Geoffrey Davies and Baoshan Xing

109

13.

Kinetics of Desorption of 2,4-dichlorophenoxyacetic Acid from Humic Acid, Metal Oxides and Metal Oxide-Humic Complexes C.LiuP.M.Huang

115

v

PART 3. METAL BINDING AND MOBILITY: THEORY, DATA AND CONSEQUENCES 14.

Exploring the Molecular Character and Heterogeneity of Humic Substances via the Study of the IonBinding Process Using an Extended Polyelectrolyte Model Josemaría García-Mina

123

15.

Study of Fulvic-Aluminum(III) Ion Complexes by 27Al Solution NMR Norman C.Y.Lee and David K.Ryan

139

16.

Investigation of Colloidal Properties and Trace Metal Complexation Characteristics of Soil-Derived Fulvic Acids by Flow Field-Flow Fractionation-Inductively Coupled Plasma-Mass Spectrometry (Flow FFF-ICPMS) Jonathan Bell, Dula Amarasiriwardena, Atitaya Siripinyanond, Baoshan Xing and Ramon M.Barnes

145

17.

Comparison of Dialysis, Polarography and Fluorimetry for Quantification of Cobalt(II) Binding by Dissolved Humic Acid Fanny Monteil-Rivera, Jean-Paul Chopart and Jacques Dumonceau

154

18.

Diffusion of Metal Cations in Humic Gels Martina Klu áková and Miloslav Peka

167

19.

The Role of Humic Substances in Trace Element Mobility in Natural Environments and Applications to Radionuclides Valerie Moulin, Badia Amekraz, Nicole Barre, Gabriel Planque, Florence Mercier, Pascal Reillerand Christophe Moulin

175

20.

Influence of Humic Substances on the Migration of Actinides in Groundwater G.Buckau, M.Wolf, S.Geyer, R.Artinger and J.I.Kim

184

21.

Catalytic Effects of Ni-Humic Complexes on the Reductive Dehalogenation of C1 and C2 Chlorinated Hydrocarbons Edward J.O’Loughlin, Huizhong Ma and David R.Burris

191

PART 4. BIOGEOCHEMICAL EFFECTS: THE GOOD, THE BAD AND THE UGLY 22.

Humic Substances and Their Direct Effects on the Physiology of Aquatic Plants 209 Stephan Pflugmacher, Constanze Pietsch, Wiete Rieger, Andrea Paul, Torsten Preuer, Elke Zwirnmann and Christian E.W.Steinberg

23.

More Evidence for Humic Substances Acting as Biogeochemicals on Organisms C.Wiegand, N.Meems, M.Timoveyev, C.E.W.Steinberg and S.Pflug macher

224

Index

233

Preface

More people, less arable land and more pollution mean greater reliance on humic substances, Nature’s most versatile materials and the major regulators of plants, soils and water. Food and pollution are everyone’s problem and understanding humic substances is in everyone’s best interest. Fortunately, we are witnessing a remarkable growth of knowledge of natural organic matter (NOM) in general and its humic components (the relatively long-lived humic and fulvic acids and humin) in particular. The Humic Substances Seminars at Northeastern University promote and acknowledge this increased understanding. This book captures the spirit of the impressive progress being made. Humic Substances Seminar VI in July 2002 immediately followed the Twentieth Anniversary Biennial Conference of the International Humic Substances Society in Boston. As a result we have been able to invite contributions from Seminar and the Conference, and by doing so cover the major advances in the field. This book reports the best current research on humic substances from around the world. Its contributions follow a natural progression of humic substances separation, characterization, sorption (including all-important interactions with water), metal binding and transport, and increasingly apparent biogeochemical effects. Each contribution encourages us to combat drought and pollution with humic substances and save our planet for this and future generations. FRACTIONATION AND CHARACTERIZATION: THE STATE-OF-THE-ART Chromatography still heads the list of methods that fractionate humic substances (HSs). The workhorses are high performance size exclusion chromatography (HPSEC) and dialysis, whose calibration and detector sensitivity are troublesome. Franke and colleagues have labeled humic and fulvic acids with two radioisotopes and used a high sensitivity germanium detector to watch 131I and 111In exiting as HPSEC column. Close similarity of the radiochromatograms confirms covalent binding of iodine and tight binding of indium, a surrogate for aluminum, which is toxic but difficult to study. This contribution illustrates the analytical benefits of radiolabeling and complements descriptions of radionuclide binding and mobility later in this book. Schmitt-Kopplin is contributing strongly to improved HSs separation, detection and characterization. The main focus of this contribution is capillary electrophoresis, the workhorse of genome and proteome research and increasingly useful for humic substances characterization. Efficient fractionation of HSs by selective methods such as capillary zone electrophoresis is still not achieved. One sensible suggestion is to avoid reactive buffers, with carbonate recommended and ammonium as the counterion to get rid of the volatile buffer prior to analysis by electrospray mass spectrometry. The interplay between mass and charge in determining electrophoretic mobility is dominated by charge. Combined electrophoretic/mass spectral data in this contribution are a prelude to other chromatographies and successful applications of mass spectrometry in later chapters. A pervading theme is that humic substances from completely different sources exist as homologous families more similar than we thought just a year ago. Scymczak and colleagues have used the relatively new asymmetric flow field flow fractionation technique to separate Suwannee River fulvic acid into fractions. Comparison of the time-of-flight secondary ion mass spectra for m/z=200 to 2000 Da of the fractions and the unseparated fulvic sample indicates remarkable similarity. This supports conclusions that fulvic acid fractions contain the same or very similar low mass molecules that may associate to form aggregates. It wasn’t long ago that getting any kind of mass spectrum of a humic sample was a cause for celebration. Now Cooper and Stenson at Florida State University and the National High Magnetic Field Laboratory have identified individual molecular components of Suwannee River fulvic acid with electrospray ionization Fourier transform—ion cyclotron resonance mass spectrometry (ESI FT-ICR MS). Virtually all the 5500 ions between m/z=300–1100 Da have been identified by this ultrasensitive method and guess what: the components are closely similar members of 266 homologous molecular families! We can’t think of better use of an ESI FT-ICR MS instrument than to confirm humic substances similarity after centuries of thinking that NOM samples have little or nothing in common. Are humic substances macromolecules or supramolecular aggregates of small molecules held together by weak intermolecular forces? This history-making debate goes on. González-Gaitano and García-Mina contribute by using dynamic light scattering to investigate fulvic samples fractionated with HPSEC. They conclude that acidification and re-alkalization

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with hydrochloric, citric and oxalic acids gives the same result: no difference in mass distribution and no evidence for supramolecule formation from the lowest molar mass constituents. Thermal analysis techniques provide a means to quantify the thermodynamic properties of heat capacity, thermal expansion coefficient and glass transition temperatures, as described in detail by DeLapp and LeBoeuf. These parameters combined impose constraints on existing molecular simulations models and help to focus attention on models with closest similarities to real HSs systems. HYDRATION, SWELLING AND SORPTION: CONTRIBUTING FACTORS Humic substances are major importers, exporters and transporters of solutes in soils and natural waters, and they play a much greater role than clays and minerals in this respect. These HSs functions depend on sorption rates and equilibria. As of now, we can only speculate on the overall contributions of HSs behavior to the global ecosystem. But now their importance is receiving the close scrutiny it deserves. The seven contributions in this section represent excellent work on sorption being conducted around the world. Interactions of humic substances with water dominate all the processes investigated. Dissolved NOM can be analyzed as a system of natural ligands with the aid of well-tried models. Sorption by solid HSs makes a larger environmental contribution, but it is more difficult to conceptualize. One important effort is to test and re-test the ‘dual-mode” sorption model, which divides interactions of solid HSs with solutes into partition (as between an organic solvent and water) and adsorption, which implies either interaction with a rigid sorbent framework or with specific (though heterogeneous) adsorption sites. The test often uses the Freundlich model to explain isotherm non-linearity due to solute adsorption. New attention is being paid to effects on sorption of wetting, swelling and hydration of solid HSs, particularly as they affect solute sorption by hydrated solid HSs from water on the one hand and by dry solid HSs from “dry” solvents on the other. In essence, we are using solvent/hydration behavior to interrogate the active sites of HSs, a worthy cause that requires no definite molecular structure but still reveals interaction mechanisms, as for uncrystallized proteins and enzymes. Graber and Borisover’s review of their latest sorption work makes fascinating reading. They focus on a link-disruption (or link-solvation) model developed from the local to general to co-operative levels. This gives a general isotherm model that relates the differences between intra- and intermolecular interactions in the dry state to competition between the solvent and solute for interactions with sites released when the linkages are broken on hydration. The model simplifies to the phenomenological Freundlich description at low solute activities that correspond to the natural conditions used in the most revealing experimental studies. Schaumann and colleagues have studied dehydration effects on the extractability of polyaromatic hydrocarbons from soils. Peat and whole organic soil hydration/swelling have been investigated with measured 1H nuclear relaxation times and gravimetry. Starch and semolina are used as standards. The relaxation times depend on chemical composition and also on the particle size of the dry materials, which affects the pore size distributions. A difference between pore size distributions in the dry and hydrated states is necessary to monitor hydration. Pores in swollen semolina are larger than those in swollen peats and soils. The data depend strongly on the individual sample. Schaumann et al. have measured the low swelling rates (time constants of days instead of hours for minerals) and suggest a relation between PAHs extractability from soil and its state of hydration for PAH-laden samples of the same contaminant age. Slow PAHs sorption on sorbents from coastal sources is the focus of the contribution from Tremblay and colleagues, who demonstrate the importance of lipid content in sediments and humins on the extent, rate and character of PAHs sorption. Removing lipid components increases the non-linearity of the sorption isotherms, consistent with unblocking of lipidrestricted adsorption sites. Re-addition of lipids increases isotherm linearity, as expected. There is a strong correlation between the capacity for PAHs and the carbon content of the sorbent, even for non-linear isotherms. Huebner and colleagues have attached humic residues to chromatographic silica and used the products as solid phases to study the sorption of polycyclic aromatic compounds (PACs) by the attached humic substances. Complementary data from solid phase microextraction and liquid-liquid extraction are reported, along with an investigation of iron(III) catalyzed humification of PACs as a means of contaminated sites remediation in concert with current bioremediation strategies. Leftover pesticides become non-extractable (by water and methanol) or ‘bound’ residues in soils, and people worry about their long-term health effects. Identifying the residues is difficult, as illustrated by the contribution from Dignac and colleagues. Analytical pyrolysis results pesticide-SOM reactions that obscure bound residue identification by PY-GC-MS. Fresh soil forms the most bound residues and it seems that microbes ‘adapt’ to metabolize atrazine, lessening the percentage of ‘bound’ pesticide in previously exposed soils. Most of the pesticide detected is unchanged atrazine, which is another example of how humic substances are able to wrap around metals and other solutes to slow or stop their desorption. A sizeable fraction of solid humic substances exist in soils as clay-humic composities, and the question is how the sorptive properties of HSs are affected when they coat a mineral surface. Wang and colleagues report that a humic coating increases the sorptive capacity of two common clays for phenanthrene and that the isotherms are non-linear, indicating specific

viii

adsorptive interactions. Two possible reasons for this behavior are 1) a more rigid HS structure induced by the mineral surface template, and 2) disruption of functional group linkages on surface binding or hydration, as implied by Graber, Borisover and other researchers. This theme of HS-mineral surface effects also is the focus of Liu and Huang’s description of the kinetics of desorption of the common pesticide 2,4-dichlorophenoxyacetic acid (2,4-D) from humified catechol on Al, Fe and Mn oxide surfaces. 2,4-D was one of the first pesticides used in Canada and exists in enormous quantities in its soils. Lui and Huang report that 2,4-D desorption is assisted by citrate and obeys the rate law expected for overall parabolic diffusion. Solutes have to fight their way through humic matrices for desorption to occur. Metal cation diffusion through humic gels that model mineral coatings is explored later in this book. METAL BINDING AND MOBILITY: THEORY, DATA AND CONSEQUENCES Wanting to understand metal binding by humic substances has three main driving forces: nutrient supply to plants, toxic metal sequestration by soils and radionuclide transport from nuclear waste sites by water. As with any solute, we need to know the thermodynamics and kinetics of metal binding, and some of the best minds in the field are working on these issues. Garcia-Mina adds to the stock of metal binding tools by describing a model based on polyelectrolyte theory that incorporates the key factors of competitive proton and metal ion binding without requiring a pre-conceived humic geometry or structural pattern. A distinctive feature of the ‘EPM’ model is that electrostatic parameters related to molecular geometry are obtained by graphical analysis of the data. This removes the need to assume a specific molecular geometry. One intriguing conclusion is that ‘hidden’ carboxylic acid groups become accesssible for metal binding due to humic molecule expansion at higher pH. This is another way in which humic substances buffer natural systems: they bind metals more strongly at higher pHs where many metals precipitate. As mentioned earlier, aluminum binding by humic substances is difficult to measure and model because of aluminum’s complicated aqueous chemistry. Lee and Ryan have studied Al binding with 27Al nuclear magnetic resonance (I=5/2), which is simple in principle but difficult in practice. Al-FA complexes form at pH ~4 but Al hydrolysis takes over at higher pH. The Al-FA and Al-oxalate NMR spectral data compare well, suggesting oxalate as a useful model for Al-fulvic acid binding in the environment. The contribution from Bell et al. spotlights renewed attention to flow fleld flow fractionation, which in principle is an ideal way of studying humic aggregation and solute binding. Fractionation is coupled to ICP-MS to study the distribution of a range of metals complexed by fulvic acids from soils with different cropping systems. The results show that metals bind to a range of FA fractions with apparent molar masses of 900 Da and hydrodynamic diameters of 2 nm. Fractionation has no large effect on metal binding, as expected for similar molecular constituents. Agricultural management practices also have little effect on the molar masses or hydro-dynamic diameters of the FA samples investigated. Measuring weak metal binding by humic substances is challenging because these natural ligands have both strong and weak binding sites and high total metal concentrations are needed to detect the weakest ones. As an example, cobalt(II) binding by humic substances is difficult to measure because the binding is weak and different measurement methods work best over different total metal concentrations. The contribution by Monteil-Rivera et al. compares the requirements and limitations of three potential ways of measuring weak cobalt(II) binding: fluorescence quenching, polarography and dialysis. The preferred method is dialysis, which is the most cumbersome of the three but gives the most straightforward results. Transport of nutrient and toxic metals and of radionuclides as mobile humic substance complexes is a major environmental concern. Metal transport by HSs depends on the rates of dissociation of metals from predominantly carboxylate ligands. It is known that the dissociation rates are much lower than for dissociation of a given metal from the same donor atoms in simple mononuclear complexes and the question is why? Analysis of data with a model for diffusion of Cuaq2+, Coaq2+, and Niaq2+ through a lignite-derived humic substance gel at 25°C in the work by Klu′ áková and Peka′ includes a metal binding-dissociation component. However, the apparent diffusion coefficients are similar to those for the same cations in water. This is because humic substance gels typically are 95% water by mass because of the very high affinity of humic substances for water. Moulin et al. review methods to measure and predict the effects of humic substances on the solubility, speciation and transport of radionuclides and other environmental metals. Humic substances affect the bioavailability and toxicity of run-off from nuclear waste sites and determine the safety of such repositories in the short and long term. Time-resolved laser induced fluorescence is a very useful tool in this work. The results confirm the existence of tight and weaker metal binding sites in humic substances and the inclusion of hydroxo and carbonato ligands along with HSs in ternary complexes formed by radionuclides and lanthanides under environmental conditions. Also discussed is experimental evidence that iodine, another nuclear waste component, is covalently bonded to aromatic constituents in its reactions with fulvic acids, and the effects on radionuclide retention of humic coatings on mineral surfaces. This contribution contains very useful insight for researchers in environmental nuclear chemistry.

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The metal transport theme is continued in the contribution by Buckau et al., who focus on the origin, concentrations, stability and mobility of humic substances in radionuclide-contaminated ground waters and the effects of HSs interactions on radionuclide transport. For example, the transport of americium by humic colloids can be described with a model consisting of two consecutive binding steps, quasi-irreversible binding and no colloid retention. Precautions necessary to obtain reliable data for HSs system modeling are carefully described and the contribution emphasizes the need for kinetic data rather than reliance on equilibrium interactions that underestimate trace metal transport. The list of known humic substances functions would easily fill this page, which is why we designate them ‘Nature’s most versatile materials.’ Work in recent years has revealed one of humic substances’ least appreciated functions: redox catalysis associated with quinoid components in HSs structures. The contribution by O’Laughlin et al. describes the products and kinetics of homogeneously catalyzed trichloroethene (TCE) and other chlorinated hydrocarbon reductions with titanium(III) citrate in water. The catalysts are Ni-Aldrich humic acid complexes and the products are non-chlorinated hydrocarbons, which is the grail of chlorohydrocarbon-contaminated water remediation. The contribution discusses likely rate-determining steps and subsequent reactions leading to the desired products. As such, it will be of interest to anyone working on remediation of halocarbon-contaminated water. Of equal importance is heightened awareness of HSs catalysis of many biochemical and environmental processes. BIOGEOCHEMICAL EFFECTS: THE GOOD, THE BAD AND THE UGLY Humic substances mostly consist of plant metabolites, and the onset of humification is associated with senescence, the twilight zone between the life and death of a plant. The link between the death of one organism and the physiology of another is through the products of death, which are humic substances. Significant levels of humic substances were first discovered in a marine alga about eight years ago, and now we are seeing more studies of the direct effects of humic substances on plants and animals. The last two contributions of this book explore the effects of humic substances on the physiology of aquatic plants and water animals. As such, they are major contributions to our understanding of yet another role of humic substances in the environment and on human health. The contribution by Pflugmacher et al. demonstrates strong effects of natural HSs and a synthetic HS on the enzyme systems and inhibited photosynthetic activity of two common aquatic plants (Ceratophyllum demersum and Vesicularia dubyana). The effects are similar in nature and magnitude to those of anthroquinone, a known inhibitor of plant photosynthesis. The HS spin densities are related to their effects on photosynthetic oxygen production, which is thus shown to be a redox reaction mediated by humic substances. A plant either synthesizes humic substances during its own lifetime or is exposed to humic substances from another organism. Put another way, an organism can induce senescence in a plant. This collection of important new work in humic substances science is completed with the contribution from Wiegand et al. on humic substances as geochemicals. Specifically, the physiological effects of humic substances on the common carp (Cyprinus carpio), the water flea (Daphnia magna) and three amphipod species are enumerated. Animal stress and changes in enzymatic activity on exposure to humic substances from different sources are demonstrated. Possible mechanisms are HSs uptake by cells or adsorption on cell surfaces. The sources of HSs affect their structures and so a quantitative structure activity relationship (QSAR) is expected as more data become available. Many participants in Humic Substances Seminar VI and the IHSS Conference commented that the scientific level was the highest they have seen at any previous forum on humic substances. Hard work on HSs over many years is beginning to pay dividends, as demonstrated by the contributions in this book. We are standing on the shoulders of giants with the prospect that Nature’s most versatile materials will one day be fully understood and appreciated. In the meantime, we encourage you to make your own best contribution to this important goal. ACKNOWLEDGEMENTS We thank the authors for their enthusiasm and co-operation, the contributors to the 20th Anniversary Conference of the International Humic Substances Society and Humic Substances Seminar VI, the staff of Northeastern University who contributed so much to their success and the sponsors who make this progress possible. Daniel Mercuri kept our computers running as we edited this work. Robert Rogers and the staff at Taylor & Francis made our work as Editors feasible and enjoyable. We are grateful. Elham A.Ghabbour Geoffrey Davies Editors Boston, Massachusetts December, 2002

Contributors

Dula Amarasiriwardena Professor, School of Natural Science, Hampshire College, Amherst, MA 01002, USA Badia Amekraz Research Scientist, CEA, Nuclear Energy Division & UMR CEA-CNRS-UEVE 8587, Analysis and Environment Laboratory, 91191 Gif-sur-Yvette Cedex, France Robert Artinger Senior Scientist, Forschungszentrum Karlsruhe, Institut für Nuklerare Entsorgung, Postfach 3640, 76021 Karlsruhe, Germany Gérard Bardoux Technical Staff, Laboratoire de Biogéochime Isotopique, Université Pierre et Marie Curie, 4 place Jussieu, 75252 Paris Cedex 05, France Ramon M.Barnes Professor and Director, University Research Institute for Analytical Chemistry, 85 North Whitney St., Amherst, MA 01002–1869, USA Nicole Barre Research Scientist, CEA, Nuclear Energy Division & UMR CEA-CNRS-UEVE 8587, Analysis and Environment Laboratory, 91191 Gif-sur-Yvette Cedex, France Enrique Barriuso Research Scientist, UMR INRA, INAP-G Environnement et Grandes Cultures, 78850 ThivervalGrignon, France Jonathan Bell Undergraduate Student, School of Natural Science, Hampshire College, Amherst, MA 01002, USA Mikhail Borisover Research Scientist, Institute of Soil, Water and Environmental Sciences, The Volcani Center, Agricultural Research Organization, P.O.B. 6, Bet Dagan 50250, Israel Gunnar Buckau Senior Scientist, Forschungszentrum Karlsruhe, Institut für Nuklerare Entsorgung, Postfach 3640, 76021 Karlsruhe, Germany David R.Burris Principal Scientist, Integrated Science and Technology Inc., 433 Harrison Avenue, Panama City, FL 32401, USA Jean-Paul Chopart William T.Cooper Geoffrey Davies Rossane C.DeLapp Sylvie Derenne Marie-France Dignac Jacques Dumonceau Karsten Franke Jean-Pierre Gagné Josemaría García-Mina

Professor, University of Reims Champagne-Ardenne, Dynamique des Transferts aux Interfaces UMR 6107 SC CNRS, UFR Sciences, Moulin de la Housse, BP 1039, 51687 Reims Cedex 2, France Associate Professor, Department of Chemistry and Biochemistry, Florida State University, Tallahassee, FL 32306–4390, USA Professor, Department of Chemistry and Chemical Biology, Northeastern University, Boston, MA 02115–5000, USA Graduate Student, Department of Civil and Environmental Engineering, Vanderbilt University, Nashville, TN 37235, USA Research Scientist, Laboratoire de Chimie Bioorganique et Organique Physique UMR CNRS 7573, ENSCP, 11 rue Pierre et Marie Curie, 75231 Paris Cedex 05, France Research Scientist, Laboratoire de Biogéochime Isotopique, Université Pierre et Marie Curie, 4 place Jussieu, 75252 Paris Cedex 05, France Professor and Group Leader, Groupe de Recherche en Chimie Inorganique, Equipe de Chimie de Coordination aux Interfaces, University of Reims Champagne-Ardenne, UFR Sciences, Moulin de la Housse, BP 1039, 51687 Reims Cedex 2, France Institute of Interdisciplinary Isotope Research, Permoserstrasse 15, 04318 Leipzig, Germany Professor, Laboratoire d’analyses et d’études en géochimie organique, Institut des Sciences de la mer de Rimouski, Université du Québec à Rimouski, Rimouski, QC G5L 3A1, Canada Associate Professor, Department of Chemistry and Soil Science, University of Navarra, and Director, R&D Department, Inabonos-Roullier Group Poligono Arazuri-Orcoyen, C/C. n° 34, 31160 Orcoyen, Spain

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Stefan Geyer

Senior Scientist, Environmental Research Center Leipzig-Halle, Hydrogeology Section, 06120 Halle, Germany Elham A.Ghabbour Senior Scientist, Department of Chemistry and Chemical Biology, Northeastern University, Boston, MA 02115–5000, USA Gustavo González-Gaitano Associate Professor, Department of Chemistry and Soil Science, University of Navarra, Pamplona, Navarra, 31080, Spain Ellen R.Graber

Research Scientist, Institute of Soil, Water and Environmental Sciences, The Volcani Center, Agricultural Research Organization, P.O.B. 6, Bet Dagan 50250, Israel P.Ming Huang Professor, Department of Soil Science, University of Saskatchewan, 51 Campus Drive, Saskatoon, SK S7N 5A8, Canada Matthias Hübner Postdoctoral Fellow, Centro Ricerche Ambientali— Montecatini, Via Ciro Menotti 48, I-48023 Marina di Ravenna, Italy Julia Hurrass Doctoral Student, Technical University Berlin, Institute of Environmental Protection, Dept. Environmental Chemistry, Sekr. KF 3, Strasse des 17. Juni 135, D-10623 Berlin, Germany Kristoffer E.N.Jonassen Doctoral Student, Plant Research Department, Risoe National Laboratory, P.O.B. 49, DK-4000 Roskilde, Denmark Jae-II Kim Professor, Forschungszentrum Karlsruhe, Institut für Nuklerare Entsorgung, Postfach 3640, 76021 Karlsruhe, Germany Martina Klu áková Assistant Professor, Institute of Physical and Applied Chemistry, Faculty of Chemistry, Brno University of Technology, Purky′ ova 118, 612 00 Brno, Czech Republic Scott D.Kohl Ph.D., Box 323, Aurora, SD 57002, USA Hermann Kupsch Ph.D., Institute of Interdisciplinary Isotope Research, Permoserstrasse 15, 04318 Leipzig, Germany Claude Largeau Research Scientist, Laboratoire de Chimie Bioorganique et Organique Physique UMR CNRS 7573, ENSCP, 11 rue Pierre et Marie Curie, 75231 Paris Cedex 05, France Eugene J.LeBoeuf Assistant Professor, Department of Civil and Environmental Engineering, Vanderbilt University, Nashville, TN 37235, USA Norman C.Y.Lee Doctoral Student, Department of Chemistry, University of Massachusetts Lowell, Lowell, MA 01854, USA Chen Liu Research Scientist, Kuo Testing Labs, Inc., 337 S. 1st Ave, Othello, WA 99344, USA Ludovic Loiseau Doctoral Student, UMR INRA, INAP-G Environnement et Grandes Cultures, 78850 ThivervalGrignon, France Huizhong Ma Postdoctoral Fellow, Air Force Research Laboratory AFRL/MLQR, 139 Barnes Drive, Tyndall Air Force Base, Florida 32403–5323, USA André Mariotti

Professor, Laboratoire de Biogéochime Isotopique, Université Pierre et Marie Curie, 4 place Jussieu, 75252 Paris Cedex 05, France N.Meems Leibniz Institute of Freshwater Ecology and Inland Fisheries, AG Detoxication/Metabolism, Müggelseedamm 301, 12561 Berlin, Germany Florence Mercier Research Scientist, UMR CEA-CNRS-UEVE 8587, Analysis and Environment Laboratory, 91191 Gif-sur-Yvette Cedex, France Fanny Monteil-Rivera Research Officer, Group of Analytical and Environmental Chemistry, National Research Council of Canada—Biotechnology Research Institute, 6100 Royalmount Avenue, Montreal, Quebec H4P 2R2, Canada Valerie Moulin Research Scientist, CEA, Nuclear Energy Division & UMR CEA-CNRS-UEVE 8587, Analysis and Environment Laboratory, 91191 Gif-sur-Yvette Cedex, France Christophe Moulin Professor and Head, Department of Physico-Chemistry, CEA, Nuclear Energy Division, 91191 Gif-sur-Yvette Cedex, France Martin Müller Postdoctoral Fellow, Technical University Berlin, Institute of Applied Geosciences, Dept. of Applied Geophysics, ACK 2, Ackerstr. 76, D-13355 Berlin, Germany Torben Nielsen Senior Scientist, Plant Research Department, Risoe National Laboratory, P.O.B. 49, DK-4000 Roskilde, Denmark Edward J.O’Loughlin Staff Scientist, Environmental Research Division, Argonne National Laboratory, 9700 South, Cass Ave, Argonne, IL 60439, USA

xii

Andrea Paul Miloslav Peka Stephan Pflugmacher Constanze Pietsch

Leibniz Institute of Freshwater Ecology and Inland Fisheries, AG Detoxication/Metabolism, Müggelseedamm 301, 12561 Berlin, Germany Associate Professor, Institute of Physical and Applied Chemistry, Faculty of Chemistry, Brno University of Technology, Purky′ ova 118, 612 00 Brno, Czech Republic Assistant Director, Leibniz Institute of Freshwater Ecology and Inland Fisheries, AG Detoxication/Metabolism, Müggelseedamm 301, 12561 Berlin, Germany Leibniz Institute of Freshwater Ecology and Inland Fisheries, AG Detoxication/Metabolism, Müggelseedamm 301, 12561 Berlin, Germany

Gabriel Plancque

Research Scientist, CEA, Nuclear Energy Division, Department of Physico-Chemistry & UMR CEA-CNRS-UEVE 8587, Analysis and Environment Laboratory, 91191 Gif-sur-Yvette Cedex, France Torsten Preuer Leibniz Institute of Freshwater Ecology and Inland Fisheries, AG Detoxication/Metabolism, Müggelseedamm 301, 12561 Berlin, Germany Pascal Reiller Research Scientist, CEA, Nuclear Energy Division, Department of Physico-Chemistry, 91191 Gif-sur-Yvette Cedex, France James A.Rice Professor and Head, Department of Chemistry and Biochemistry, South Dakota State University, Brookings, SD 57007–0896, USA Wiete Rieger Diploma Student, Leibniz Institute of Freshwater Ecology and Inland Fisheries, AG Detoxication/Metabolism, Müggelseedamm 301, 12561 Berlin, Germany Doris Rössler Institute of Interdisciplinary Isotope Research, Permoserstrasse 15, 04318 Leipzig, Germany Wolfgang Rotard Professor, Technical University Berlin, Institute of Environmental Protection, Dept. Environmental Chemistry, Sekr. KF 3, Strasse des 17. Juni 135, D-10623 Berlin, Germany David K.Ryan Professor, Department of Chemistry, University of Massachusetts Lowell, Lowell, MA 01854, USA Gabriele E.Schaumann Postdoctoral Fellow, Technical University Berlin, Institute of Environmental Protection, Dept. Environmental Chemistry, Sekr. KF 3, Strasse des 17. Juni 135, D-10623 Berlin, Germany Philippe Schmitt-Kopplin Ph.D., GSF-National Center for Environment and Health, Institute of Ecological Chemistry, Ingoldstädter Landstraße 1, D-85764 Neuherberg, Germany Atitaya Siripinyanond Lecturer, Department of Chemistry, Faculty of Science, Mahidol University, Rama 6 Rd., Rajthevee, Bangkok 10400, Thailand Christian E.W.Steinberg Professor and Director, Leibniz Institute of Freshwater Ecology and Inland Fisheries, AG Detoxication-Metabolism, Müggelseedamm 301, 12561 Berlin, Germany Alexandra C.Stenson Research Scientist, Vertex Pharmaceuticals, Inc., 130 Waverly Street, Cambridge, MA 02139– 4242, USA Wilfried Szymczak Senior Scientist, Institute of Radiation Protection, GSF-National Research Center for Environment and Health, 85758 Neuherberg, Germany M.Timoveyev Luc Tremblay

Irkutsk State University, Karl Marx 1, 664003 Irkutsk, Russia Postdoctoral Fellow, Department of Biological Sciences and Marine Science Program, University of South Carolina, Columbia, SC 29208, USA Kaijun Wang Doctoral Student, Department of Plant and Soil Sciences, University of Massachusetts, Amherst, MA 01003, USA Claudia Wiegand Postdoctoral Fellow, Institute of Biology, Humboldt-University of Berlin, Chausseestr 117, 10115 Berlin, Germany Klaus Wittmaack Head of Radiation Physics Group, Institute of Radiation Protection, GSF-National Research Center for Environment and Health, 85758 Neuherberg, Germany Manfred Wolf Senior Scientist, Institute of Hydrology, GSF-National Research Center for Environment and Health, 85764 Neuherberg, Germany Baoshan Xing Associate Professor, Department of Plant and Soil Sciences, University of Massachusetts, Amherst, MA 01003, USA Yahya Zegouagh Postdoctoral Fellow, Laboratoire de Biogéochime Isotopique, Université Pierre et Marie Curie, 4 place Jussieu, 75252 Paris Cedex 05, France

xiii

Elke Zwirnmann

Staff Scientist, Leibniz Institute of Freshwater Ecology and Inland Fisheries, AG Detoxication/ Metabolism, Müggelseedamm 301, 12561 Berlin, Germany

Part 1 FRACTIONATION AND CHARACTERIZATION: THE STATEOF-THE-ART

Chapter 1 USE OF RADIOACTIVE TRACERS FOR THE CHARACTERIZATION OF HUMIC AND FULVIC ACIDS IN HIGH PERFORMANCE SIZE EXCLUSION CHROMATOGRAPHY Karsten Franke, Doris Rössler and Hermann Kupsch Institute of Interdisciplinary Isotope Research, Permoserstrasse 15, 04318 Leipzig, Germany

1.1. INTRODUCTION Humic substances (HSs) influence many geochemical and environmental processes in soil [1–5]. As ubiquitous major components of soil organic matter, their regulation of metal interactions in soils is unquestioned. Depending on geochemical parameters, HSs can act as geochemical barriers or contribute to the non-retarded migration of metals. An important aspect in understanding these processes is the conformational nature of HSs [6,7]. Depending on circumstances, HSs can form refractory colloids [8], precipitate via aggregation or remain in solution as negatively charged complexes. Therefore, the effect of molecular size and shape of HSs must be taken into account [9]. Much effort has been devoted to the use of size exclusion chromatography (SEC) to answer some fundamental questions about HSs [1,10]. Overcoming artifacts caused by secondary interaction with the column material, high performance size exclusion chromatography (HPSEC) has become a predictive tool for investigation of humic substances [11–14]. Nevertheless, common detection methods (UV-absorption, fluorescence, ICP-MS) have different limitations, such as detection limits, signal quenching and problems caused by online coupling. This prompted us to investigate the possibility of using radioactive tracers in HPSEC. We adapted radiolabeling techniques to label HSs, leading to very good agreement of the radiochromatograms obtained with the results of classical detection methods. In addition to labeling of the carbon backbone of HSs (with 131I), our interest was focused on the interaction of humic substances with aluminum. In the absence of a suitable aluminum isotope, we used 111In as a surrogate for aluminum in our investigations. Table 1.1 Elemental compositions and amounts of humic and fulvic acids in soil Depth 0 cm–10 cm 10 cm–20 cm 20 cm–30 cm 30 cm–40 cm 40 cm–50 cm

Amount [mg/g]

HA Elemental composition

FA Elemental composition

HA

FA

%N

%C

%H

%S

%N

%C

%H

%S

12.3 15.1 12.1 8.4 2.2

2.0 2.3 2.2 2.3 1.0

4.5 3.8 3.8 3.4 3.1

41.4 33.8 34.5 33.5 32.5

5.0 4.5 4.5 4.2 3.8

1.3 0.8 0.8 0.7 0.8

3.4 3.6 3.3 3.5 3.3

41.0 40.9 42.7 41.2 41.4

4.3 4.2 4.4 4.3 4.4

0.5 0.5 2.0 0.7 0.5

1.2. MATERIALS AND METHODS 1.2.1. Humic Substances Humic substances were extracted from different soil profiles at depths of 0–50 cm. The sampling site is located in Schlema/ Alberoda, Saxony, Germany. The soil samples were taken with a 10 cm spacing. A detailed description of the sampling area can be found in [15]. The humic acids (HAs) and fulvic acids (FAs) were obtained following the IHSS procedure [16]. The elemental compositions and percent yields of the humic and fulvic acids are given in Table 1.1.

RESULTS AND DISCUSSION |

3

Figure 1.1 Reaction scheme of the 131I labeling of humic and fulvic acids

Figure 1.2 Decay scheme of 131I (a) and 111In (b) (simplified plot of the ′ -rays used)

1.2.2. Radiolabeling 131I (t =8.041d) was used for labeling of the HAs and FAs. In the reaction, the aromatic moieties of HS molecules are 1/2 labeled via electrophilic H J-substitution (Figure 1.1). Differing from the known iodination method for humic substances [17,18], 1,3,4,6-tetrachloro-3a,6a-diphenylglycouril was used as the oxidizing agent. This so-called Iodogen method is widely known in radiopharmaceutical chemistry [19]. The main advantage of the Iodogen method is the easy separation of the oxidizing agent from the sample. The radiochemical yields were measured and the stability of the labeling was proved by ultracentrifugation, dialysis and precipitation experiments. Different yields were observed depending on the labeling method and the source of HS. The yields of the labeling procedure varied between 65% and 80%. The known photo-susceptibility of iodination must be considered in following experiments. In addition to iodination, the complexing functional groups of HSs have been used for 111In labeling of the HA and FA. The labeling was performed with no-carrier-added [111In]InCl3 (t1/2=2. 81d) (Nycomed-Amersham) via complexation.

1.2.3. High Performance Size Exclusion Chromatography HPSEC was used to determine the molar mass distribution of the humic and fulvic acids (10 mg/L). The instrument was equipped with a TSK gel column (TSK-G3000PWXL, 300 mm×7.8 mm) with a size fraction range of 0.5–800 kDa. A flow of 0.5 mL/min and a sample volume of 25 µL were used for chromatography. The mobile phase consisted of 0.02 M KCl in MilliQ water with 0.05 M tris-buffer (pH 8). Sodium polystyrene sulfonates (Mn: 4.3, 8.6, 17.4 and 33.8 kDa), blue dextran (Mn: 100 kDa) and acetone were used as standards. A high purity Ge-detector was used to measure the radiochromatograms. Due to their excellent energy resolution, the following gamma rays were used for detection (131 I: ′ 1=364 keV, ′ 2=637 keV, 111In: ′ 1=171 keV, ′ 2=245 keV, Figure 1.2a,b). 1.3. RESULTS AND DISCUSSION The coincidence of the radiochromatogram of the [131I]HA and the UV absorbance chromatogram was used as a criterion for testing the iodination of HS. These signals are compared in Figure 1.3. The observed radiochromatogram is in clear correspondence with the UV signal, which means that all constituents of the HS are labeled in similar proportions. The sample is completely eluted in the size fraction range 050 µm fraction; (B) Humin; (C) (Humin)after HF; (D) (Humin)HF; (E) Refractory humin. Numbering is te same as for Figure 11.4.

spiked after separation with a dilute solution of 13C-labeled atrazine and immediately analyzed with Py/GC/C-SIRMS to determine the retention times of the products formed from atrazine. In Figures 11.4 and 11.5, the enriched peaks are numbered according to their retention times. Peaks 14 and 15 are present in almost all the spiked fractions of soil. For these two peaks, the enriched signal observed on the 13C/12C trace can be related to a clearly observed peak in the total CO2 trace. This is not the case for the other enriched peaks: they can be observed on the 13C/12C trace because they are largely enriched compared to organic matter with natural 13C abundance, but their concentration in the pyrogram is so small that no peak can be observed on the total CO2 trace. The same spiked fractions as above were analyzed by Py/GC/MS to search for the chemical structures of the labeled products observed via Py/GC/C-SIRMS. Only compounds 14 and 15 that also display a “real” peak on the total CO2 trace (Py/ GC/C-SIRMS) and thus also on the Total Ion Current (TIC) trace of the GC/MS pyrograms could be identified. For the other peaks, only observed because of their high 13C enrichment, structure determination with Py/GC/MS was not possible. The chemical structures of compounds 14 and 15 were recognized with Py/GC/MS as 13C-labeled OH-simazine (m/z 214) and 13Clabeled atrazine (m/z 218), respectively (Figure 11.3). Their mass spectra both display a fragment at m/z 70, characteristic for

MATERIALS AND METHODS |

105

Figure 11.6 Py/GC/C-SIRMS chromatograms of size fractions of the MS5 soil incubated with 13C-labeled atrazine and extracted to remove the extractable forms of atrazine-derived residues. The i) traces are the ratio of m/z=45 to m/z=44 (multiplied by 100) and the ii) traces represent total CO2. (A) >50 µm fraction; (B) 20–50 µm fraction; (C) < 20 µm fraction. Same numbering as for Figure 11.4.

Figure 11.7 Py/GC/C-SIRMS chromatograms of >50 µm size fractions of the MV6, WG8 and MG8 soils incubated with 13C-labeled atrazine and extracted to remove the extractable forms of atrazine-derived residues. The i) traces are the ratio of m/z=45 to m/z=44 (multiplied by 100) and the ii) traces represent the total CO2. (A) MV6; (B) WG8; (C) MG8. Same numbering as for Figure 11.4.

the presence of the triazine ring (Figure 11.3). The presence of atrazine and OH-simazine was confirmed in the pyrograms of most of the spiked soil fractions by monitoring this characteristic mass fragment. 11.3.2. Pyrolysis of Incubated Soils The bound residues of atrazine were searched for in the size fractions of the four incubated soils (>50 µm, 20–50 µm, 50 µm and 20–50 µm) but was not found in the more humified fraction (Figure 11.6). This compound was also found in high relative abundance in the >50 µm fractions of the three other soils (Figure 11.7). This suggests that atrazine is the main form of bound residues in these soils.

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107

The history of the soil influences the formation of bound residues. In the fractions of the soils adapted to atrazine (MS5 and MG8), that have developed a microflora able to degrade this pesticide, only one or two enriched compounds are pre sent apart from atrazine. For the “non adapted” soils atrazine is present in the pyrograms of the fractions of the incubated WG8 and MV6 soils, along with many other enriched bound products. An adaptation of the microflora therefore seems to decrease the number of chemical forms of the bound residues, which are mainly bound atrazine in that case. Barriuso and Houot [3] have shown that the total amount of bound residues formed in adapted soils is lower. Accordingly, adaptation of soil microflora appears to influence both the abundance and the nature of the bound residues of atrazine. The content of OM in the soil does not appear to influence the bound residues formed. Indeed both in the soil with the highest OM content (MS5) and in the soil with a low OM content (MG8), only a few bound residues are observed apart from atrazine. 11.4.3. Evaluation of the Analytical Technique In previous studies [3,5,6], the use of the 14C isotope to follow the bound residues of atrazine has proven to be very useful in determining the distribution of these bound residues in the different soil fractions. To examine the chemical structures of these residues with a chromatographic method, labeling of the atrazine with 13C isotope is required, as in the present work. Since the bound residues are chemically bound to OM [6], pyrolysis was supposed to be a good option to break the corresponding bonds and release the residues for molecular analyses. Analysis of the fractions of the incubated soils with Py/GC/C-SIRMS confirms that the bound enriched molecules can be released upon pyrolysis and several enriched peaks can be observed. Nevertheless, it was not possible to determine the chemical structures of the enriched compounds thus detected. The detection limit of SIRMS is very low because this method specifically detects the 13C isotope of enriched compounds. The 13C-labeled atrazine used to incubate the soils was highly enriched in 13C (99% of the triazine ring). Therefore, the 13C-enriched compounds formed from the bound residues can be detected, even when their peaks on the total CO2 trace of the GC/C-SIRMS analysis are not distinguishable from the noise or from the peaks of the pyrolysis products of soil OM. In contrast, MS was not sensitive enough to allow the detection of extremely low amounts of such compounds, especially when they are highly diluted by large quantities of the pyrolysis products arising from OM present in the soil fractions. Moreover, the enriched compounds detected via Py/GC/C-SIRMS often coeluted with much more intense peaks originating from the pyrolysis of soil OM. Even by monitoring the characteristic mass fragments of atrazine, it was not possible to confirm its presence as a bound residue by MS analysis. Thus, only the GC retention times of the detected enriched compounds could be used for tentative identification. The concentration of labeled atrazine used for incubation in the soil was 20 times the agronomic dose used for field treatments. Even this relatively large value leads to concentrations of labeled bound residues that are not high enough for GC/ MS analysis. The pyrolytic technique is satisfactory for releasing the bound residues, but problems are encountered for detection of the very low concentration of these released bound residues that eluted with much larger amounts of pyrolysis products from soil OM. New analytical techniques need to be developed to study the chemical structure of these bound residues. Such techniques should combine 1) an efficient release of the residues chemically bound to soil OM; 2) an efficient detection of the released products; and 3) a very sensitive identification technique to distinguish the chemical structures of the released residues from those of products originating from soil OM. The two first requirements were achieved in the present work by pyrolysis and SIRMS, respectively, but compound identification still has to be improved. 11.5. CONCLUSIONS 13C-labeled

atrazine was incubated with different soil samples to examine the bound residues of atrazine at a molecular level. This stable isotope labeling allows sensitive detection with SIRMS. Analytical pyrolysis was used to break the chemical bonds between the atrazine residues and the organic matrix of the soil fractions. Preliminary experiments showed the formation of secondary compounds by reaction of atrazine with products released upon pyrolysis of OM of soil fractions. Comparison of fractions from MS5 and MG8 soils spiked with 13C-labeled atrazine and immediately analyzed showed marked differences. These differences depend on the nature and relative abundance of the 13C-enriched compounds formed. Differences were observed both between the successive humic fractions obtained from a given soil by chemical fractionation and between the same type of fraction from the two soils. Among the 13C-enriched products generated upon pyrolysis of the spiked fractions of soils and detected by SIRMS, atrazine and OH-simazine could be identified by parallel Py/GC/MS analysis in almost all the spiked fractions. Analysis by Py/GC/C-SIRMS of size fractions of incubated soils for the bound residues of 13C-labeled atrazine indicated the presence of several enriched compounds with different chemical structures whose distribution changed from one fraction to the next. Comparison of the retention times of the enriched products suggested the presence of atrazine along with other

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| CHAPTER 11: PYROLYTIC STUDIES OF THE BOUND RESIDUES OF 13C-ATRAZINE

labeled compounds. Atrazine was found to be the main form of bound residues in the studied soils. The content of OM in soil did not appear to influence the number of enriched peaks detected. However, the history of the soil seemed to be important as regards the number of chemical forms of bound residues formed. Adaptation of the microflora decreased the number of bound compounds, which were mainly bound atrazine, compared to “non-adapted” soils. Parallel analysis by on-line Py/GC/C-SIRMS and on-line Py/GC/MS was used for the first time to study bound residues in soil. This work showed the necessity of developing new tools in order to be able to detect and characterize these bound residues. It appears that 1) pyrolysis effectively releases bound residues; 2) the detection limits of the bound residues at a molecular level was clearly enhanced by the use of 13C labeled compounds, which allowed the detection of trace amounts of bound residues released upon pyrolysis; and 3) the analytical technique has to be improved for the GC/MS determination of the chemical structure of products that occur in extremely low amounts and moreover are highly diluted by the pyrolysis products generated from the OM of the soil. REFERENCES 1. 2. 3. 4. 5. 6. 7. 8.

Hamaker JW, Thomson HB. Adsorption. In: Goring CAJ, Hamaker JW eds. Organic chemicals in the soil environment. New York: Dekker, 1972:49–143. Dec J, Bollag JM. Determination of covalent and non-covalent binding interactions between xenobiotic chemicals and soil. Soil Sci., 1997; 162:858–874. Barriuso E, Houot S. Rapid mineralization of the s-triazine ring of atrazine in soils in relation to soil management. Soil Biol. Biochem., 1996; 28:1341–1348. Rice JA. Humin. Soil Sci., 2001; 166:848–857. Loiseau L, Barriuso E, Zegouagh Y, Largeau C, Mariotti A. Release of the atrazine non-extractable (bound) residues of two soils using degradative techniques. Agronomie, 2000; 20:513–524. Loiseau L, Barriuso E. Characterization of atrazine’s bound (nonextractable) residues using fractionation techniques for soil organic matter. Environ. Sci. Technol., 2002; 36:683–689. Gleixner G, Poirier N, Bol R, Balesdent J. Molecular dynamics of organic matter in a cultivated soil. Org. Geochem., 2002; 33: 357–366. Hatcher PG, Dria K, Kim S, Frazier SW. Modern analytical studies of humic substances. Soil Sci., 2001; 166:770–794.

Chapter 12 PHENANTHRENE SORPTION BY CLAY-HUMIC COMPLEXES Kaijun Wang,1 Elham A.Ghabbour,2 Geoffrey Davies2 and Baoshan Xing1 1Department

of Plant and Soil Science, University of Massachusetts, Amherst, MA 01003, USA

2Chemistry

Department, Northeastern University, Boston, MA 02115, USA

12.1. INTRODUCTION Sorption of hydrophobic organic compounds (HOCs) in soils and geologic materials is of growing concern because of the strong affinity of these compounds for soil colloids and the potential risks associated with their long-term persistence in the environment. It is well accepted that soil organic matter (SOM) is the predominant sorbent of HOCs. The sorption of HOCs by humic substances has been extensively studied, and it is generally believed that partitioning and adsorption-like mechanisms are involved in the sorption process [1–5]. According to the dual-mode model [2,3], sorption in rubbery domains is anticipated to exhibit a linear (partitioning) isotherm contribution, while sorption in glassy domains is expected to represent the nonlinear (Langmuir) contribution. The heterogeneous nature of humic substances (HSs) and their colloidal aggregates give rise to numerous potential binding sites for a wide range of materials of diverse chemistry, such as metal ions, organic pollutants, and biocides used for agricultural purposes. Humic substances can also interact with clay particles, and ultimately enhance the colloidal stability of the clay particles and the structural stability of soils [6,7]. Another consequence of those interactions is to stabilize organic matter in soils as clay-humic complexes [8]. Previous studies have demonstrated that organic matter is likely to be associated with the clay fractions as clay-organic complexes in soil and sedimentary environments [9,10], most commonly in the form of coatings on solid surfaces [11,12]. In this respect, soil organic matter may play very important roles in controlling the fate and transport of organic or inorganic contaminants in natural environments. In surface soil containing as little as 1% organic mat ter, for instance, the surface chemistry was found to be controlled by organic components coated on phyllosillicate clay minerals, and on Al and iron oxides [13]. Humic substances may undergo conformational changes after associating with clay minerals [14,15], and such modification will affect their sorption behavior with HOCs. Therefore, it may be inappropriate to extrapolate the sorption behavior of HS coated on mineral particles from extracted HS alone. So far, a few studies have included sorption of HOCs to constructed clay-humic complexes [16–19]. The results show that the organic carbon normalized sorption coefficient (Koc) values are lower for mineral-bound humic materials compared with humic materials alone. It is proposed that humic acids might adopt a more condensed conformation once they were bound to a mineral surface, resulting in a decreased number of sites for HOCs sorption than for bulk humic acids [19]. In addition, Murphy et al. [17] observed nonlinearity and competitive sorption of three organic compounds by humic-coated minerals implying adsorption rather than partitioning into the surface organic phase. However, Onken and Traina [20] measured the sorption of pyrene and anthracene to three humic acid-mineral, complexes (HA-CaCO3, HA-Calcite, and HA-Na-Montmorillonite) and found that partitioning is the predominant process for the higher foc (fractional organic carbon content) system (>3×10–5), while adsorption and condensation are major processes for the lower foc system (C=O stretch of -COOH appears at 1716–1719 cm−1. Inaddition, the 1610 cm−1 band of aromatic >C=C< stretch or asymmetric -COO- stretch, the 1415 cm−1 CHdeformation of CH3 and -CH bending of CH3 groups, the 1212 cm−1 band (due to>C=O stretch and OH deformation of -COOH), and the 1070 cm−1 C-C stretch of aliphatic groups were prominent in all the fulvic acid DRIFT spectra. The peaks in the low wavenumber region are mostly due to aromatic CH out of plane bending or to organically bound mineral phases [28–30]. Table 16.3 Hydrodynamic diameters and molar masses determined by flow FFF FA Sample

Hydrodynamic diameter (nm) dh at peak maxa

FAVR1 2.01 FAVR4 2.01 FAR1 2.01 FAR4 2.01 FAC1 2.01 FAC4 2.01 a Average obtained from duplicate flow FFF runs.

Molar Mass (Da), (Mp)a 920 920 920 920 870 870

16.3.2. Flow-Field Flow Fractionation Analyses A flow FFF fractogram is a plot of UV signal as a function of emergence time (tr), as shown in Figure 16.2 for sample FAC1. The molar masses and hydrodynamic diameters at peak maximum obtained for the soil derived FAs are listed in Table 16.3.

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149

Figure 16.2 Typical UV-fractogram of a fulvic acid sample (FACl)

Figure 16.3 Fractograms of FAVR4 (with cover crops and fertilizer treatments) fulvic acid. (top left) frequency distribution of hydrodynamic diameters; (top right) Frequency distribution of molar masses; (bottom left) ion fractograms: ion intensity of 63Cu, 64Zn, 208Pb vs. hydrodynamic diameter; (bottom right) ion intensity of 63Cu, 64Zn’ 208Pb vs. calculated molar mass

Transformation of raw FFF fractograms into molar mass or hydrodynamic diameter distributions is described in the literature [22,31,32]. The time axis of the fractogram can be transformed into hydrodynamic diameter with Eq. 16.2. With the linear calibration function obtained from logarithmic tr vs. logarithmic molar mass plots, the retention time of each sample was converted to molar mass. The UV signal is directly proportional to fulvic acid concentration. The analyte signals were measured at small time interval slices (dm/dt) and transformed into a frequency function of hydrodynamic diameter (dm/dd) or molar mass distributions (dm/dM) at fixed digitized intervals with Eqs. 16.3 and 16.4, where dt is the emergence time difference for successive (16.3) (16.4) digitized points and dd and dM are the corresponding differences in the hydrody namic diameter and molar masses for those points, respectively [22,31,32]. Fractograms of molar mass and hydrodynamic diameter distributions of fulvic acid extracted from vetch and rye grown soil with nitrogen fertilizer (FAVR4) and fulvic acid derived from the control soil (FAC1, no fertilizer or cover crops used) are shown in Figures 16.3a,b and 16.4a,b, respectively.

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| CHAPTER 16: INVESTIGATION OF COLLOIDAL PROPERTIES AND TRACE METAL COMPLEXATION

Figure 16.4 Fractograms of FAC1 (without treatments) fulvic acid. (top left) frequency distribution of hydrodynamic diameters; (top right) frequency distribution of molar masses; (bottom left) ion fractograms: ion intensity of 63Cu, 64Zn, 208Pb vs. hydrodynamic diameter; (bottom right) ion intensity of 63Cu, 64Zn, 208Pb vs. calculated molar mass

16.3.3. Elemental Analyses Total trace metal concentrations in the digested fulvic acid samples are shown in the Table 16.4. 16.3.4. Investigation of Trace Metals Complexed to Fulvic Acids by Flow FFF-ICP-MS Ion fractograms of soil-derived fulvic acid were obtained for 27Al, 63Cu, 54Fe, 48Ti, Table 16.4 Elemental concentrations in soil derived fulvic acids Elemental Concentrations (µg/g) FA Sample

Cu

Zn

Fe

FAVR1 407 7 525 FAVR4 143 8 788 FAR4 387 3 542 FAC1 289 1 511 FAC4 315 3 475 Note: FAR1 was not analyzed due to limited sample size available 208Pb

Mn

Pb

Al

As

40 129 48 38 49

11 2.3 9.9 4.5 6.2

1477 1094 1534 1500 1287

1.7 .3 1.9 1.6 2.0

and 64Zn using flow FFF-ICP-MS. The ion fractograms (63Cu, shown in Figures 16.3c,d (FAVR1) and 16.4c,d (FAC1).

208Pb, 64Zn)

of size and molar mass distributions are

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151

16.4. DISCUSSION 16.4.1. Spectroscopic Characterization of Fulvic Acids The E4/E6 ratios obtained from UV-visible spectra are consistent within the values reported by Stevenson (6.0–8.5) [1] and by Chen et al., (8.4–8.9) [27] for soil fulvic acids at various concentrations (100–500 mg/L). The DRIFTS peak at ~1720 cm−1 in all fulvic acid spectra is quite intense and indicates an abundance of -COOH groups in the fulvic acids investigated. Carboxylic groups are important cation exchange and ligand sites for trace metal complexation. 16.4.2. Flow Field-Flow Fractionation of Fulvic Acids All fulvic acid samples exhibited monomodal fractograms with a skewed distribution toward large hydrodynamic diameters or molar masses. The average apparent molar mass at peak (Mp) for these fulvic acids was 920 Da except for the FAs derived from soils without cover crops (870 Da). The hydrodynamic particle diameters (dh) of all FA were ~2 nm. Identical Mp were obtained for treated (cover crops and nitrogen fertilizer) FA, and the magnitude of the molar mass is consistent with literature data obtained with flow FFF and PSS standards [15,33]. As a word of caution, the comparison of apparent molar mass obtained for fulvic or humic acids often depends on the type of calibration standards, the separation method and experimental conditions such as pH and ionic strength [33,34]. In addition, molecular aggregation or association by bridging with complexed cations, hydrogen bonding and van der Waals forces may lead to apparent large molecular size and hence large molar masses [33,34]. On the other hand, the dh values for FAs were calculated directly from Eq. 16.2 without the need for any standards. Identical apparent molar masses and hydrodynamic diameters were obtained for FAs even though they were derived from soils with various cover crops and fertilizer applications. 16.4.3. Elemental Analyses Interestingly, no discernable pattern appears in trace metals concentrations in FA that are extracted from soils treated with various soil management practices (Table 16.1). As expected, elevated levels (that is, in the upper ppm range) of Al, Fe and Cu concentrations were found in all the FAs investigated (Table 16.3). Lead concentrations in the FAs were between 2.3 and 11.0 µg/g and arsenic ranged from 1.6 to 3.3 µg/g. These results demonstrate the strong metal sequestering capability of soil fulvic acids. 16.4.4. Flow FFF-ICP-MS Investigation of Trace Metals Complexed to Fulvic Acids Our results also demonstrate that metals are complexed to a wide range of molar mass fractions, consistent with the data obtained from total elemental analyses (Table 16.4). Trace metals Zn, Cu and Pb complexed by FA molecular fractions were qualitatively identified by flow FFF-ICP-MS. The corresponding ion fractograms of 63Cu, 208Pb and 64Zn for soil-derived FAVR4 (soils treated with fertilizer and hairy vetch and rye cover crops) and FAC1 (control soil without any cover crops or fertilizer) are illustrated in Figures 16.3c,d and 16.4c,d for hydrodynamic distributions and molar mass distributions, respectively. No obvious differences appear in binding with fulvic acid size fractions among the various fulvic acids studied. All of these metals are bound to 1 to 5 nm diameter size FA fractions with maximum complexation with ~2 nm fractions (~1 kDa). As expected, intense signals were obtained for 27Al and 54Fe, indicating strong binding of these elements to soil fulvic acids. 16.5. CONCLUSIONS Our results demonstrate that flow FFF-ICP-MS can be used for characterization of trace metals complexed to various soilderived FA molecular fractions. Monomodal elemental fractograms showed that trace metals Al, Cu, Fe, Ti, Pb and Zn are bound to a broad range of FA molecular sizes. DRIFTS spectra demonstrated the presence of ligand sites, particularly the intense carboxylic sites that play an important part in metal sequestration by fulvic acids. The total trace metals present in these fulvic acids revealed the presence of iron and aluminum along with all other cations. The apparent molar mass (Mp) of

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fulvic acid obtained by flow FFF ranged from 870–920 Da. All treated soil-derived FA samples had very similar apparent molar masses Mp of 920 Da, except for the FA sample extracted from untreated soil. The hydrodynamic particle diameters (dh) of all FA studied was 2 nm. The data indicate that various soil management practices have no apparent effect on the molar masses or hydrodynamic diameters of these fulvic acid samples. Particle size and molar mass information for metalbound FA is useful in assessing the environmental behavior of biologically and toxicologically important trace elements in soil and water bodies. The small size and low molar mass of these FA molecules and their strong association with nutrient (Fe, Zn) and toxic (Pb, Al) cations is consistent with the great mobility and rapid binding [35] of these elements in the soil environment. ACKNOWLEDGEMENTS Dula Amarasiriwardena gratefully acknowledges the financial and instrument support of the National Science Foundation (BIR 951270) and the Kresge Foundation. This research was supported in part by ICP Information Newsletter, Inc., Hadley, Massachusetts. Jonathan Bell is grateful for the undergraduate research grant awarded by the Howard Hughes Medical Institute. REFERENCES 1. 2.

3. 4. 5. 6. 7. 8. 9.

10.

11. 12. 13. 14. 15. 16. 17. 18.

19.

Stevenson, FJ. Humus chemistry: Genesis, composition and reactions. 2nd Edn. New York: Wiley, 1994. Davies G, Fataftah A, Cherkasskiy A, Ghabbour EA, Radwan A, Jansen SA, Kolla S, Paciolla MD, Sein LT Jr, Buermann W, Balasubramanian M, Budnick J, Xing B. Tight metal binding by humic acids and its role in biomineralization. J. Chem. Soc., Dalton Trans., 1997; 21:4047–4060. Steelink C. Investigating humic acids in soils. Anal. Chem., 2002; 74:326A– 333A. Davies G, Ghabbour EA, Steelink C. Humic acids: Marvelous products of soil chemistry. J. Chem. Ed., 2001; 78:1609–1614. Chin YP, Aiken G, O’Loughlin E. Molecular weight, polydispersity and spectroscopic properties of aquatic humic substances. Environ. Sci. Technol., 1994; 28:1853–1858. Schmitt P, Kettrup A, Freitag D, Garrison AW. Flocculation of humic substances with metal ions as followed by capillary zone electrophoresis. Fresenius J. Anal. Chem., 1996; 354:915–920. Rottmann L, Heumann KG. Development of an on-line isotope dilution technique with HPLC/ICP-MS for the accurate determination of elemental species. Fresenius J. Anal. Chem., 1994; 350:221–227. Vogl J, Heumann KG. Determination of heavy metal complexes with humic substances by HPLC/ICP-MS coupling using an on-line isotope dilution technique. J. Anal. At. Spectrom., 1994; 359:438–441. Ruiz-Haas P, Amarasiriwardena D, Xing B. Determination of trace metals bound to soil humic acids species by size exclusion chromatography and inductively coupled plasma mass spectrometry. In: Davies G, Ghabbour EA eds. Humic substances: Structures, properties and uses. Cambridge: Royal Society of Chemistry, 1998:147–163 Bhandari SA, Amarasiriwardena D, Xing B. Application of high performance size exclusion chromatography (HPSEC) with detection by inductively coupled plasma-mass spectrometry (ICP-MS) for the study of metal complexation properties of soil derived humic acid molecular fractions. In: Ghabbour EA, Davies G eds. Understanding humic substances: Advanced methods, properties and applications. Cambridge: Royal Society of Chemistry, 1999:203–221 Giddings JC. A new separation concept based on a coupling of concentration and flow non-uniformities. Sep. Sci., 1966; 1:123–125. Giddings JC. Field-flow fractionation of macromolecules. J. Chromatog., 1989; 470:327–335. Giddings JC, Yang FJ, Myers MN. Theoretical and experimental characterization of flow field-flow fractionation. Anal. Chem., 1976; 48:1126–132. Beckett R. Field-flow fractionation-ICP-MS: A powerful new analytical tool for characterizing macromolecules and particles. At. Spectrosc., 1991; 12:228–232. Beckett R. Zhang J, Giddings JC. Determination of molecular weight distributions of fulvic and humic acids using flow field-flow fractionation. Environ. Sci. Technol, 1987; 21:289–295. Dycus PJM, Healy KD, Stearman GK, Wells MJM. Diffusion coefficients and molecular weight distributions of humic and fulvic acids determined by flow-field-flow fractionation. Sep. Sci. Technol., 1995; 30:1435–1453. Siripinyanond A, Barnes RM. Flow-field-flow fractionation-inductively coupled plasma mass spectrometry and metal speciation in proteins: A feasibility study. J. Anal. At. Spectrom., 1999; 14:1527–1531. Amarasiriwardena D, Siripinyanond A, Barnes RM. Trace elemental distribution in soil and compost-derived humic acid molecular fractions and colloidal organic matter in municipal wastewater by flow field-flow fractionation-inductively coupled plasma mass spectrometry. J. Anal. At. Spectrom., 2001; 16: 978–986. Anderson T, Shifley L, Amarasiriwardena D, Xing, B, Siripinyanond A, Barnes RM. Characterization of trace metals complexed to humic acids derived from agricultural soils, annelid composts and sediment by flow field-flow fractionation-inductively coupled plasma mass spectrometry. In: Ghabbour EA, Davies G eds. Humic substances: Structures, models and functions. Cambridge: Royal Society of Chemistry, 2001:165–177.

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Lead JR, Wilkinson KJ, Balnois E, Cutak BJ, Larive CK, Assemi S, Beckett R. Diffusion coefficients and polydispersities of the Suwannee river fulvic acid: Comparison of fluorescence correlation spectroscopy, pulsed-field gradient nuclear magnetic resonance, and flow field-flow fractionation. Environ. Sci. Technol., 2000; 34:3508–3513. Taylor HE, Garbarino JR, Murphy DM, Beckett R. Inductively coupled plasma mass spectrometry as an elemental detector for fieldflow fractionation particle speciation. Anal. Chem., 1992; 64:2036–2041. Ranville JF, Chittleborough DJ, Shanks F, Morrison RJS, Harris T, Doss F, Beckett R. Development of sedimentation field-flow fractionation-inductively coupled plasma mass spectrometry for the characterization of environmental colloids. Anal. Chim. Acta, 1999; 381:315–329. Hassellöv M, Lyvén B, Haraldsson C, Sirinawin W. Determination of continuous size and trace element distribution of colloidal material in natural water by on-line coupling of flow field-flow fractionation with ICP-MS. Anal. Chem., 1999; 71:3497–3502. Siripinyanond A, Barnes RM, Amarasiriwardena D. Flow field-flow fractionation-inductively coupled plasma mass spectrometry for sediment bound trace metal characterization. J. Anal. At. Spectrom., 2002; 17:1055–1064. Ding G, Mao J, Herbert S, Amarasiriwardena D, Xing B. Spectroscopic evaluation of humin changes in response to soil managements. In: Ghabbour EA, Davies G eds. Humic substances: Structures, models and functions. Cambridge: Royal Society of Chemistry, 2001:271–279. Ding G, Amarasiriwardena D, Herbert S, Novak J, Xing B. Effects of cover crops systems on the characteristics of soil humic substances. In: Ghabbour EA, Davies G eds. Humic substances: Versatile components of plants, soil and water. Cambridge: Royal Society of Chemistry, 2000:53–61. Chen Y, Senesi N, Schnitzer M. Information provided on humic substances by E4/E6 ratios. Soil Sci. Soc. Am. J., 1977; 41:352–358. Baes AU, Bloom PR. Diffuse reflectance and transmission Fourier transform infrared (DRIFT) spectroscopy of humic and fulvic acids. Soil Sci. Soc. Am. J., 1989; 53:695–700. Niemeyer J, Chen Y, Bollag J-M. Characterization of humic acids, composts, and peat by diffuse reflectance Fourier-transform infrared spectroscopy. Soil Sci. Soc. Am. J., 1992; 56:135–140. Wander MM, Traina SJ. Organic fractions from organically and conventionally managed soils: II. Characterization of composition. Soil Sci. Soc. Am. J., 1996; 60:1087–1094. Becket R, Hotchin DM, Hart BL. Use of field-flow fractionation to study pollutant-colloid interactions. J. Chromatogr., 1990; 517: 435–448. Chen B, Shand CA, Beckett R. Determination of total and EDTA extractable metal distributions in the colloidal fraction of contaminated soils using SdFFF-ICP-MS. J. Environ. Monit., 2001; 3:7–14. Wolf M, Buckau G, Geckeis H, Thang NM, Hoque E, Szymczak W, Kim J-I. Aspects of measurement of the hydrodynamic size and molecular mass distribution of humic and fulvic acids. In: Ghabbour EA, Davies G eds. Humic substances: Structures, models and functions. Cambridge: Royal Society of Chemistry, 2001:51–61. Schimpf ME, Wahlund K-G. Asymmetrical flow filed-flow fractionation as a method to study the behavior of humic acids in solution. J. Microcol. Sep., 1977; 9:535–543. Cabaniss SE, Zhou Q, Maurice PA, Chin Y-P, Aiken GR. A log-normal distribution model for the molecular weight of aquatic fulvic acids. Environ. Sci. Technol., 2000; 34:1103–1109.

Chapter 17 COMPARISON OF DIALYSIS, POLAROGRAPHY AND FLUORIMETRY FOR QUANTIFICATION OF COBALT(II) BINDING BY DISSOLVED HUMIC ACID Fanny Monteil-Rivera,1 Jean-Paul Chopart2 and Jacques Dumonceau1 1Université 2Université

de Reims Champagne-Ardenne, GRECI—BP 1039, 51687 Reims Cedex 2, France

de Reims Champagne-Ardenne, DTI UMR CNRS 6107, BP 1039, 51687 Reims Cedex 2, France 17.1. INTRODUCTION

The bioavailability and migration of pollutant ions (for example, radionuclides or heavy metals) in soils is closely related to the presence of humic substances (HSs). Because they strongly bind cationic species, humic substances when solid or adsorbed on mineral surfaces can retard the transport of metal ions; or conversely, dissolved HSs can enhance the mobility of these ions [1–10]. To predict speciation of metal ions in soils one must first evaluate the binding properties of both the minerals and the organic matter towards the studied metal species. Binding properties of dissolved organic matter (DOM) have been studied with numerous analytical methods including potentiometric, voltammetric, spectroscopic or physical separation based methods [2]. Among the different techniques employed, potentiometry using ion selective electrodes (ISE) is, when applicable, a method of choice due to its capacity to reach, in a direct way, the concentration of free metal ions over a wide concentration range. Most attention has been given to the complexation of divalent trace metals that can be analyzed by ISE (Cu2+, Cd2+, Pb2+ and Ca2+) and far fewer studies have been devoted to the interaction of DOM with ions like Co2+, for which there is no existing ISE. The complexation of Co(II) by HS has been studied with techniques such as fluorescence quenching [11,12], equilibrium dialysis [3], ion exchange [13,14], voltam-metry coupled with ligand exchange [15] and solvent extraction [16]. Each of these techniques may present sources of errors and difficulties of interpretation, so a com parison of results obtained by several of them would help to validate speciation tools for cobalt. In dialysis, adsorption of metal on the membrane and diffusion of organic matter through the membrane might occur and result in experimental errors. In polarography, reversibility, lability, adsorption of organic macromolecules on the mercury electrode and the coexistence of molecules of different sizes all may make polarogram interpretation difficult. In fluorescence, the interpretation of quenching curves, and more specifically the extent of complexation at the maximum quenching and the correlation between quenching and complexation (which generally is assumed to be linear) are main difficulties of the method. The aim of the present work was to evaluate the strengths, weaknesses, and applicability of dialysis, polarography and fluorimetry to study the speciation of cobalt in the presence of dissolved humic acid. Dialysis and polarography experiments were performed with similar concentrations of humic acid (~60 mgc L−1) while fluorimetry involved a more diluted medium (~3 mgc L−1). 17.2. MATERIALS AND METHODS 17.2.1. Chemicals Reagents used were Co(NO3)2.6H2O (Acros Organics), HNO3 (Normadose Prolabo 1N), KOH (Normadose Prolabo 1N), and KNO3 (Fluka). Deionized, distilled and filtered (Millipore: 0.2 µm) water was used to prepare all solutions. Leonardite humic acid (LHA) was obtained from the International Humic Substances Society (IHSS) and used as received. The elemental composition determined with a C, H, N, S Analyser (LECO CHNS-932) was C, 57.74; H, 4.17; N, 1.16; S, 0. 69 and O, 33.89%, with the latter deduced by difference after subtracting the ash content (2.35%). LHA stock solutions of about 500 mg L−1 (pH′ 7) were prepared by stirring 500 mg of LHA with 0.1 M KOH (20.8 mL) and 50 mL of water for 24 h, making up the volume with water and centrifuging twice at 26300 g for 30 minutes to remove the insoluble fraction. The total organic carbon (TOC) concentration of the supernatant determined with a Shimadzu TOC-5050 Analyzer was 271 mgc/L−1.

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The total acidity and carboxylic acid groups were determined by Ba(OH)2 and Ca(Ac)2 exchange [17], respectively, and found to be 12.0 and 7.1 mmol gc−1, respectively. In the present study, the molarity of LHA is deduced from the total acidity and the TOC concentration. 17.2.2. Instrumentation pH measurements were performed under argon with a Metrohm automatic system equipped with a combination Ag/AgCl glass electrode (Metrohm) calibrated with a solution of 10−3 M HNO3 in 0.099 M KNO3. For dialysis experiments, Co(II) concentrations were determined by inductively coupled plasma atomic emission spectrometry (ICP-AES). Differential pulse (dp) and Tast direct current (dc) polarography experiments were carried out at 25° C with a Metrohm 626 Polarecord attached to a Metrohm 663 VA Stand with a hanging mercury drop electrode (HMDE). Drop times of 1 s and scan rates of 2 mV s−1 were used. The polarograms were performed from −0.8 to −1.6 V (relative to SCE) and a 10 mV pulse amplitude was applied for dp polarography. Fluorescence measurements were performed with a Perkin-Elmer Model 50 B luminescence spectrometer with both slits set at 10 nm. An excitation wavelength of 340 nm was used for all the emission measurements, while the synchronous fluorescence measurements were made with two different offsets (′ ′ 1=20 nm and ′ ′ 2=80 nm) between the excitation and emission monochromators. 17.2.3. Procedures Dialysis Experiments. The complexation of Co2+ by LHA was determined in 0.1 M KNO3. Solutions of LHA (20 mL) were added to 1000 Da molecular weight cutoff Spectra/Por 6 dialysis tubing immersed in 150 mL of solutions containing 0.113 M KNO3 and the required amount of cobalt(II) nitrate. The pH was adjusted to steady-state values by adding small aliquots of HNO3 or KOH to the outside solution. After a 4-day equilibrium period at 25°C, TOC concentration and total concentration of Co(II) were determined inside and outside the dialysis tubing. Two different concentrations of LHA inside the tubing were tested under different sets of conditions that are listed in Table 17.1. The above procedure was preceded by preliminary experiments to either minimize or evaluate experimental artifacts associated with equilibrium dialysis. First, preservation agents (sodium azide), trace metals and sulfides in the dialysis tubing were removed by extensive washing with EDTA and hot water (80°C). Second, the extent of Co(II) binding to the tubing walls was evaluated at pH 7 by checking the mass balance. It was found to be 2% of total Co2+ added and therefore was neglected in the calculations. Third, the extent of LHA transfer through the tubing walls was measured at pH 7. After a 4-day equilibrium period, the concentration of LHA in the external compartment was found to be about 10 % of the total amount of LHA. The effect of LHA transfer on the calculation of equilibrium constants will be discussed later. Polarography Experiments. The complexation of Co2+ by LHA was determined in 0.1 M KNO3. Solutions containing the amounts of LHA and Co(II) given in Table 17.1 were prepared. The pH was adjusted to 5.0, 6.0 or 7.0 by adding small aliquots of KOH or HNO3, the solutions were then stirred for 12 h, and the pH was readjusted to the desired value (deviations of 0.5 unit pH from the initial value were observed). Measurements were made at 25°C after removing O2 by bubbling N2 through the solutions for 15 min. Although a “first-order” maximum was observed on the current voltage curves of cobalt, the experiments were performed in the absence of a maximum suppressor because the latter (gelatin) was found to react with humic acid and cause precipitation. Table 17.1 Experimental conditions used with the three different techniques Dialysis pH 4.4−7.5 4.4–6.4 ~6 Polarography 5.0 6.0

[LHA]initial, inside mgC 43.4 65.0 65.0

L−1

65.0 130.0 65.0

[Co(II)]total (M)

R=[LHA]/[Co]

M 5.21×10−4 7.80×10−4 7.80×10−4

1×10−5 1×10−5 5×10−6−5×10−4

52.1 78.0 1.56−156.0

7.80×10−4 1.56×10−3 7.80×10−4

1×10−5−2×10−4 8×10−5−4×10−4 1×10−5−2×10−4

3.90−78.0 3.90−19.5 3.90−78.0

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| CHAPTER 17: COMPARISON OF DIALYSIS, POLAROGRAPHY, AND FLUORIMETRY

Dialysis pH 7.0 Fluorescence 5.0 6.0 7.0

[LHA]initial, inside

[Co(II)]total (M)

R=[LHA]/[Co]

130.0 65.0

1.56×10−3

3×10−5

7.80×10−4

1×10−5−2×10−4

52.0 3.90−78.0

3.3 3.3 3.3

3.96×10−5 3.96×10−5 3.96×10−5

1×10−6−1.6×10−3 1×10−6−1.6×10−3 1×l0−6−1.6×10−3

0.025−39.6 0.025−39.6 0.025−39.6

17.2.3.3. Fluorescence Experiments. An LHA concentration of 3.3 mgC L−1 was selected as a good compromise between significant intensity and a linear relationship between intensity and concentration. The ionic strength was adjusted to 0.1 with KNO3. Solutions containing LHA and Co(II) at the concentrations in Table 17.1 were adjusted to pH 5.0, 6.0 or 7.0 with either HNO3 or KOH, stirred for 12 h, and readjusted to the desired pH. Emission or synchronous fluorescence spectra were collected at 25°C. 17.3. THEORY 17.3.1. Definition of Complexing Parameter K To be compared, the results obtained with the three techniques have to be expressed in a common way. Because experimental conditions (pH, ionic strength) were kept similar, the conditional stability constant, K, which is valid only for a given set of experimental conditions, was used to compare results. Humic acids are complex mixtures of acidic functional groups i, each of which can react in the presence of cobalt (II) according to Eqs. 17.1 and 17.2. (17.1) (17.2) The average stability constant, K, may be calculated from the free metal concentration [M], the sum of all individual metalhumic acid complexes concentrations [MLi] and the sum of all individual free binding sites concentrations [Li], Eq. 17.3, (17.3) where is the concentration of bound metal, [M] is the concentration of free metal and is the concentration of ligand not bound to the metal. 17.3.2. Evaluation of K from Dialysis Since hydrolysis is negligible in our experiments (pH 7, [Co(OH)+]/[Co2+]=1/1000 [18]), the [Co(II)] inside and outside the dialysis tubing were assumed to be identical. [Co(II)] also corresponds to the free [Co2+] inside the tubing at equilibrium. Knowing the free cobalt concentration ([Co2+]inside= [Co2+]outside=[Co(II)]outside), the total concentration of cobalt ([Co(II)]inside) and the total concentration of LHA ([LHA]inside) inside the tubing, one can calculate K from Eq. 17.4, where all concentraions are molar. (17.4)

17.3.3. Evaluation of K from Polarograms Both [M] and [ML] have to be known to determine K with Eq. 17.3. These concentrations can be determined for a reversible process from the values of either the potential or the current obtained for reducing waves in the presence and absence of ligand. When plotting log [(id–i)/i] against E for the reduction of Co(II) (1×10−5 M in 0.1 M KNO3), a straight line was

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obtained but with a slope a of 11.20, giving rise to a value of n=0.66 (0.0591×′ ), which is not a whole number. This shows that the reduction of free Co2+ ions on the electrode is non-reversible [19]. Irreversible cobalt reduction affects polarogram interpretation. First, for differential pulse (dp) polarography it is not possible to get quantitative information on [Co(II)] from the peak currents. Dillard et al. [20] showed that the peak current measured by dp polarography for irreversible systems, which is much less sensitive than for reversible systems, is subject to residual current from charging of the electrical double layer. Second, the half wave potentials (E1/2) for direct current (dc) polarograms, which are a function of the charge transfer rate constant, k0, vary not only because of complexation but also because of changes in the double layer [2]. For these reasons, neither dp polarograms nor E1/2 values of dc polarograms were used in the present work. Instead, [M] and [ML] were exclusively determined from diffusion currents, which is applicable to both irreversible and reversible systems. Humic acids are macromolecules, so the diffusion constant of ML, DML, should be smaller than the value for M, DM. This slower diffusion favors labile complex behavior, that is complexes that dissociate faster than they diffuse. The lability of the metal/LHA systems implies that free and bound metal ions with unequal diffusion coefficients are present in the sample solutions, including at the electrode. In this case, the limiting current, il, can be expressed by Eq. 17.5, (17.5) where A is a constant that depends on the number of electrons transferred and on experimental conditions, CM is the total bulk concentration of metal an is the mean diffusion coefficient given by Eq. 17.6. (17.6) It should be noted that Eqs. 17.5 and 17.6 only apply if is constant in the diffusion layer, that is if an excess of ligand is present. Parameter A=417 was calculated after measuring the mercury flow rate under our conditions. The slope of the calibration plot in the absence of ligand allowed determining the diffusion coefficient of free cobalt (DM) in 0.1 M KNO3. A value of 9. 9×10−6 cm2 s−1 was obtained, in relatively good agreement with the literature value 7.3×10−6 cm2 s−1 for Co2+ in water [21]. The experimental value of can then be determined from Eq. 17.5 for different values of CL and CM. Provided DML is known, the values of [ML] and [M] can then be calculated from Eq. 17.6 and the mass balance of the metal. The value of [L] is deduced from the mass balance of the ligand, and the formation constant K can be calculated. 17.3.4. Evaluation of K from Fluorescence Spectra Three methods were used to extract conditional constants from fluorescence spectra: 1) a non-linear adjustment initially applied to humic substances [11,22], 2) Stern-Volmer plots [23], and 3) a discrete log K spectrum model describing metal binding by humic matter [24]. The two first methods give a unique average conditional constant for the whole range of concentrations covered by the titration curve, while the last method gives thermodynamic constants that allow calculation of K at any desired concentration. Non Linear Adjustment Procedure. Ryan and Weber have described the fluorescence quenching of a ligand by complexation with metal ions [22]. The total intensity for a given concentration of metal, I, is expressed in Eq. 17.7, (17.7) where xL is the mole fraction of the free ligand, xML is the mole fraction of bound ligand and IL and IML are limiting fluorescence intensities at the beginning and end of the titration, respectively. As the ligand goes from the unbound to the bound form in the titration, the fraction of total ligand bound is given by Eq. 17.8, (17.8) where [MLi] is the concentration of metal bound to the ith site, and CL is the maximum capacity of the ligand under the conditions examined. Combining Eqs. 17.3 and 17.8 gives Eq. 17.9. (17.9) The best values of K, CL and IML can then be obtained by a non-linear adjustment of the experimental data in the form of relative fluorescence versus CM. In the present work, this adjustment was performed with the non-linear fitting function of Microcal Origin software, with a minimum of 10 iterations that were repeated until convergence. Stern-Volmer Plots. When a fluorescent ligand, L, forms no fluorescent complexes with a metal (static quenching) in Eq. 17.2, the following Stern-Volmer Eq. 17.10 is expected to hold [23]. (17.10)

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Fluorophores in HSs with no metal ion complexation properties but that contribute a constant factor to the overall quenching profiles cause deviation from linearity of Eq. 17.10 towards the x-axis. If f is the fraction of the initial fluorescence that corresponds to the fluorescent structures accessible for complexation, the following modified Stern-Volmer Eq. 17.11 is obtained [23]. (17.11) If the plot of IL/(IL-I) vs. 1/[M] is linear, K and f can be readily estimated from the slope and intercept of the straight line. However, only the total metal ion concentration, CM, is known experimentally, and its use instead of [M] in Eqs. 17.10 and 17. 11 requires that CM′ [M]. This condition can only be satisfied with high concentrations of metal and/or low values of K. Only the experiments involving concentrations of cobalt′ 4×10−4 M(CM>10× [LHA]) were used for the calculations of K from modified Stern-Volmer plots. Discrete log K Spectrum Model [24]. After converting the quenching to the concentration of bound metal, [ML] with Eq. 17.8 (see below for the value of CL), the free metal concentration can be calculated from the mass balance equation for the total metal, Eq. 17.12. (17.12) Knowing the concentrations of the bound and free metal species at any point in the titration, one can then apply a model specifically designed to represent metal/HS binding. From the different existing models (NICA-Donnan [25], Tipping [26]), we chose the simplest to use model of Westall et al. [24] to analyze the data. In this model, the HA is represented by an assembly of monoprotic acids with assumed pKa values, the anions of which bind cations in a 1:1 complex. Each site of LHA is thus expected to react according to Eqs. 17.1, 17.2 and 17.13. (17.13) Ka(i), KK(i) and sites concentrations are calculated from acid-base titrations whereas the stability constants for cobalt can be calculated from experiments involving Co(II). In the present study, acid-base titrations were performed in 0.1 M KNO3. Parameters were fitted with the computer program FITEQL 3.2 [27], which iteratively optimizes adjustable parameters by minimizing the differences between calculated and experimental data using a nonlinear least-squares optimization routine. Consistent with the discrete log K spectrum approach, the values of four pKa(i) were set at 4,6,8 and 10 to cover the pH range of the data. The concentration of each site as well as the values of KK(i) were adjusted. The resulting parameters are given in Table 17.2. The overall variance, WSOS/DF (weighted sum of squares of residuals divided by the degree of freedom), indicates good fitting (WSOS/DF