Adsorption and its Applications in Industry and Environmental Protection : Applications in Environmental Protection (Studies in Surface Science and Catalysis, 120 a-B.)

Adsorption and Its Applications in Industry and Enviromntal Protection: Vol. 2: Applications in Environmental Protection, by A. Dabrowski

• ISBN: 0444501665 • Publisher: Elsevier Science & Technology Books • Pub. Date: January 1999

PREFACE Every aspect of h u m a n activity is closely connected with the natural environment. Whether or not we are aware, or care, every day each of us interacts with and affects our environment. The rapid development of technology, especially at the end of 20th century, has increased enormously man's ability to produce goods which, in turn, have enhanced his standard of living. On the other hand, this development has also generated a secondary phenomenon, the environment pollution. Such effect led to deterioration of life quality. Thus, improvement of the life quality owing to innovative technologies caused negative effects for the environment. In order to keep the balance between technology development and main components of the man's environment the appropriate technologies should be used which appear to be a powerful force for the improvement of the environment. The relevant activities for upgrading the quality of ground water, drinking water, soil and air have to be developed. The environmental changes affect also the h u m a n health. Only few chemical compounds present in the h u m a n close surrounding may be considered as beneficial for health. The majority of them act harmfully on humans, even in minimal doses. They occur in our environmental media - air, water and soil and that is why we observe the increasing efforts devoted to the h u m a n environmental protection. One of the most important factors in this field are the possibilities and results of modern chemical analyses of pollutants in biological fluids to maintain h u m a n health. Water is one of the most important components of our environment. Nowadays, the drinking water is becoming more and more scarce, but our demand for water is becoming greater and greater. A very important problem is concerned with the rising levels of nutrients such as nitrates and phosphates in the surface water. Their presence has caused a serious deterioration in the water quality of many rivers, lakes and reservoirs. Therefore the attention has to be given to the removal of nutrients originating from sewages and fertilizers by adsorption methods, ion-exchange and relevant biotechnological techniques. Phosphorous and its compounds dissolved in the ground waters are responsible for the eutrophication in the closed water system, especially in lakes and highly enclosed bays where water is stagnant. Slag media, wasted by - products from steel industries, are effective adsorbents for phosphorous and its compounds. The earth atmosphere along with water, is the main component of our environment. One essential cause of pollution of the air is the tendency to decrease the cost of manufacturing goods by the use of contaminated raw materials without purifying or enriching them before their application. A preliminary desulfurization of coal is still rare. When air is used as a source of

vi oxygen, nitrogen in the air is a diluent which, after the oxygen consumption, is discharged into the atmosphere together with other impurities. Dusts and smogs are another group of air contaminants. The modern technologies should restrict emissions of carbon dioxide to prevent from increasing the amount of heat being dispersed into the atmosphere. This increase, leading to a change of climate, is the greenhouse effect. The other fundamental problem is connected with the removal of volatile organic chloride (VOC) compounds from ground water and recovery of chlorofluorocarbons (CFCs), which are still used in refrigeration and cooling systems. Emission control of ozone depletion by CFCs is very urgent. The pressure on industry to decrease the emission of various pollutants into the environment is increasing. A broad range of methods is available and developed to control and remove both natural and anthropogenic, municipal, agricultural and other pollutants. In relation to the price/performance, adsorption technologies are the most important techniques to overcome the degradation of environmental quality. They play a significant role both in environmental and h u m a n health control and in prevention from global warming and ozone layer depletion. The neccessity to reduce the ozone depletion gases like CFCs and the demand for primary energy diversification in the air conditioning sector, are the main reasons for the increasing interest in adsorption devices considered as alternative to the traditional compressor heat pumps in the cooling systems. Adsorption processes are the ,,heart" of several safety energy technologies which can find suitable applications in the domestic sectors as reversible heat pumps, and in the industrial sectors as refrigerating systems and heat trasnformers using industrial waste heat as the primary energy source. They can also be used for technologies to be applied in the transportation sectors, for automobile air conditioning or for food preservation in trucks. The adsorption dessicant dehumidification technology is also emerging as an alternative to vapour compression systems for cooling and conditioning air for a space. Dessicant base systems can improve indoor air quality and remove air pollutants due to their coadsorption by the dessicant materials. Moreover, a number of microorganisms are removed or killed by the dessicant. Other problems are production of drinking water, removal of anthropogenic pollutants from air, soil and water as well as removal of microorganisms from the indoor air and other important tasks to solve in terms of adsorption technologies. Adsorption can also be expected to play a significant role in the environmental control and life supporting systems or planetary bases, where sorbents may be used to process the habitat air or to recover useful substances from the local environments. Another environmental dilemma deals with the removal of thermal SOx and NOx from hot combustion gases. The above mentioned problems may be solved by advanced adsorption techniques. Among them, the rapid pressure swing adsorption (PSA) methods are very efficient for solving both global and local environmental issues. By the term of global environmental problem is meant emission of ozone depletion gases like CFCs, VOC and emission of green-house gases (CO2, CH4, N20, etc.), but the term local environmental problem deals with flue gas recovery (SOx and NOx),

vii solvent vapour fractionation and solvent vapour recovery, wastewater treatment and drinking water production. Other environmental issues concern the industrial solid aerosols, which are the incomplete combustion products. They are harmful as precursors to the synthesis of strong toxins, carcinogenes and mutagenes. Automobiles contribute substantially to man-made hydrocarbon emissions. A new type of activated carbon filtres for the application in Evaporative Loss Central Devices (ELCD) were developed by NORIT. Automobiles had to pass the so-called SHED emission test, which was legislated in Europe in 1992. Adsorption of metals into living or dead cells has been termed biosorption. Biosorption dealing with the metal - microbe interactions include both terrestrial and marine environments. Biosorption by the sea bacteria plays a significant role in detoxification of heavy metals in the aqueous systems. The literature on the influence of biosorption in metal crystal formation is rather scant. The subject of microbe participation in nucleation and halite crystal growth is important with regard to the influence of cell surface layer (S-layer) components on the crystal habit. As follows from the above considerations, the subject of utility of modern adsorption technologies has enormous environmental, economic and legal importance and constitutes a serious challenge with the prospects for further intense development. Likwise to volume I which contains the most important industrial applications of adsorption, this volume includes the chapters written by authoritative specialists on the broad spectrum of environmental topics to find a way for intense anthropogenic activities to coexist with the natural environment. Some of the topics presented in this volume were mentioned above. However, I would like to highlight a wide spectrum of themes referring to the environmental analysis and environmental control, molecular modelling of both sorbents and adsorption environmentally friendly processes, new trends in applications of colloidal science for protecting soil systems, purification and production of drinking water, water and ground water treatment, new environmental adsorbents for removal of pollutants from waste waters and sewages, selective sorbents for hot combustion gases, some corrosion aspects and ecological adsorption of heating and cooling pumps. This book is divided into two volumes, consisting of chapters arranged in a consistent order, though some chapters could be connected with the industrial (volume I) or environmental (volume II) fields. In order to highlight for readers all topics and considerations each volume of the monograph comprises the complete contents and the complete list of authors, but ncludes its own subject index only. It should be emphasized that all contributions were subjected to a rigorous review process, with almost all papers receiving two reviews from a panel of approximately fifty reviewers. The presented chapters give not only brief current knowledge about the studied problems, but are also a source of topical literature on it. Thus each chapter constitutes an excellent literature guide for a given topic and encourages

viii the potential reader to get to know a problem in detail and for further specialistic studies. At the end of the volume the comprehensive bibliography on adsorptive separations, environmental applications, PSA, parametric pumping, ion-exchange and chromatography is presented which includes the period 1967-1997.All the articles give both the scientific background of the phenomena discussed and indicate practical aspects to a great extent. Consequently, this monograph is addressed to a large group of research workers both in academic institutions and industrial laboratories, whose professional activities are related to widely understood surface environmental problems, including environmental analysis, environmental catalysis and biocatalysis,modern adsorption ecologicallyfriendlly technologies, etc. This book is meant also for students of graduate and postgraduate courses. I am aware, that the panorama of the researches presented is incomplete.On the other hand, I believe that this monograph is a substantial step presenting the current trends and the state of the art. I would like to express my warmest thanks to all the contributors for their efforts to develop the topical environmental fields of great importance. Finally, I wish acknowledge the great help I had my wife, Mrs. Iwona D@rowska, during all stages of the growth of the monograph.Her patience, encouragment and support made it possible to appear this book in present form.

Lublin, September, 1998.

A.Dqbrowski (ed.)

Complete List of Authors

1.

A l e x a n d r a t o s S.D. Department of Chemistry, University of Tennessee at Knoxville, Knoxville, TN 37996-1600, USA 2. A n d r u s h k o v a O.V. Department of Total and Bioorganic Chemistry, Novosibirsk Medical Institute, Krasny Prospekt 52, Novosibirsk 630091, Russia 3. Baldini F. Instituto di Ricerca sulle Onde Elettromagnetiche ,,Nello Carrara", CNR, Via Panciatichi 64, 50127 Firenze, Italy 4. B a n d o s z T.J. Department of Chemistry, City College of New York, New York, NY 10031, USA 5. Blom J. Tauw Milieu P.O.Box 133, 7400 AC Deventer, The Netherlands 6. Bl~dek J. Institute of Chemistry, Military University of Technology, Kaliskiego 2, 01-489 Warsaw, Poland 7. Boere J.A. NORIT N.V., Research & Development, Nijverheidsweg - Noord 72, P.O.Box 105, 3800 AC Amersfoort, The Netherlands 8. Bogillo V.I. Institute of Surface Chemistry, National Academy of Sciences, Prospekt Nauki 31, 252022 Kiev, Ukraine 9. B r a c c i S . Centro di Studio sulle Cause di Deperimento e Metodi di Conservazione Opere d'Arte, CNR, Via G.Capponi 9, 50121 Firenze, Italy 10. Billow M. The BOC Group Gases Technical Center, 100 Mountain Ave., Murray Hill, NJ 07974, USA 11. B u c z e k B. Faculty of Fuels and Energy, University of Mining and Metallurgy, 30-059 Cracow, Poland

12. B u r k e M. University of Arizona, Old Chemistry Bldg., Tucson, AZ 85721, USA 13. Cacciola G. National Council of Research, Institute for Research on Chemical Methods and Processes for Energy Storage and Transformation, S.Lucia sopra Contesse, 98126 Messina, Italy 14. Carey T.R. Radian International, LLC, 8501 N.Mopac Blvd., Austin, TX 78759, USA 15. Cerofolini G.F. SGS-THOMSON Microelectronics, 20041 Agrate MI, Italy 16. C h a n g R. Electric Power Research Institute, 3412 Hillview Ave., Palo Alto, CA 94403, USA 17. Chen J. Georgia Institute of Technology, School of Civil and Environmental Engineering, Atlanta, GA 30332-0512, USA 18. Chen S. Illinois State Geological Survey, 615 E. Peabody Dr. Champaign, IL 61820, USA 19. D a b o u X. Chemical Process Engineering Laboratory, Department of Chemical Engineering, Aristotle University of Thessaloniki and Chemical Process Engineering Research Institute, PO Box 1520, Thessaloniki 54006, Greece 20. D a l l B a u m a n L.A. NASA Johnson Space Center, Houston, TX 77058, USA 21. D ~ b r o w s k i A. Faculty of Chemistry, M.Curie-Sktodowska University, 20031 Lublin, Poland 22. Deka R.C. India Catalysis Division, National Chemical Laboratory, Pune - 411008, India 23. Deng S.G. USA Department of Chemical Engineering, University of Cincinnati, Cincinnati, Ohio 45221, USA 24. D o b r o w o l s k i R. Faculty of Chemistry, M.Curie-Sklodowska University, 20031 Lublin, Poland 25. D o m i n g o - G a r c i a M. Grupo de InvestigaciSn en Carbones, Dpto. de Quimica Inorganica, Fac. de Ciencias, Universidad de Granada, 18071 Granada, Spain

xi 26. D y b k o A. Department of Chemistry, Warsaw University of Technology, Noakowskiego 3, 00-664 Warsaw, Poland 27. F a d o n i M. Department of Physical Chemistry and Electrochemistry, University of Milan, Via Golgi 19, 20133 Milan, Italy 28. F e r n a n d e z - M o r a l e s I. Grupo de InvestigaciSn en Carbones, Dpto. de Quimica Inorganica, Fac. de Ciencias, Universidad de Granada, 18071 Granada, Spain 29. F i n n J.E. NASA Ames Research Center, Moffett Field CA, USA 30. F l e m i n g H. Cochrane Inc., 800 3nd Avenue, King of Prussia, 19406 PA, USA 31. G h o s h T.K. Particulate Systems Research Center, Nuclear Engineering Program, E 2434 Engineering Building East, University of Missouri-Columbia, Columbia, MO 65211, USA 32. G h z a o u i A.E1. UM II LAMMI ESA 5079, Case 015, Place Eugene Bataillon, 34095 Montpellier Cedex 5, France 33. Golden T.C. Air Products and Chemicals, Inc., 7201 Hamilton Boulevard, Allentown, PA 18195-1501, USA 34. G r o s z e k A.J. MICROSCAL LTD, 79 Southern Row, London W 10 5 AL, UK 35. H a u k k a S. Microchemistry Ltd., P.O.Box 132, FIN-02631 Espoo, Finland 36. H e i j m a n S.G.J. KIWA Research and Consultancy, P.O.Box 1072, 3430 BB Nieuwegein, The Netherlands 37. Hines A.L. Honda of America Mfg.Inc., 24 000 Honda Parkway, Marysville, OH 43040, USA 38. H o p m a n R. KIWA Research and Consultancy, P.O.Box 1072, 3430 BB Nieuwegein, The Netherlands 39. H o r v a t h G. University of Veszprem, H-8201 Veszprem, P.O.Box 158, Egyetem u.10, Hungary

xii 40. H s i H-C.

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University of Illinois, Environmental Enegineering Program, 205 N.Mathews Ave., Urbana, IL 61801, USA H u b i c k i Z. Faculty of Chemistry, M.Curie-Sktodowska University, 20031 Lublin, Poland I s u p o v V.P. Institute of Solid State Chemistry and Raw Mineral Processing Kutateladze-18, 630128, Novosibirsk, Russia I v e r s o n I. Department of Chemistry, University of Nevada, Reno, NV 89557, USA Izmailova V.N. Moscow State University, Department Colloid Chemistry, Vorob'evy Gory, 119899 Moscow, Russia J a k o w i c z A. Faculty of Chemistry, M.Curie-Sklodowska University, 20031 Lublin, Poland J a n u s z W. Faculty of Chemistry, M.Curie-Sktodowska University, 20031 Lublin, Poland Kalvoda R. J.Heyrovsky Inst.Phys.Chem., Czech Acad. Scis, Dolejskova 3, 18223 Prague 8, Czech Republic K a n e k o K. Chiba University, Department of Chemistry, Faculty of Science, 1-33 Yayoi, Inage, Chiba 263, Japan Kanellopoulos N. Institute of Physical Chemistry NCSR ,,DEMOKRITOS", Aghia Paraskevi Attikis, GR- 153 10, Athenes, Greece K i k k i n i d e s E.S. Institute of Physical Chemistry NCSR ,,DEMOKRITOS", Aghia Paraskevi Attikis, GR- 153 10, Athenes, Greece K i r i c h i e n k o O.A. Institute of Solid State Chemistry, SB RAS, Kutateladze 18, Novosibirsk 630128, Russia Kleut D.v.d. NORIT N.V., Research & Development, Nijverheidsweg - Noord 72, P.O.Box 105, 3800 AC Amersfoort, The Netherlands Kobal I. Department of Physical and Environmental Chemistry, J.Stefan Institute, 61000 Ljubljana, Slovenia

xiii 54. Kotsupalo N.P. Ekostar - Nautech Company, B.Chmielnitsky 2, 630075 Novosibirsk, Russia 55. Krebs K.-F. Merck KGaA, LAB CHROM Synthese, D-64271 Darmstadt, Germany 56. Kubo M. Department of Molecular Chemistry and Engineering, Faculty of Engineering, Tohoku University, Sendai 980-77, Japan 57. L a k o m a a E.-L. Neste Oy, Technology Center, P.O.Box 310, FIN-06101 Porvoo, Finland 58. Lemcoff N.O. The BOC Group, 100 Mountain Avenue, Murray Hill, NJ 07974, USA 59. Lin Y.S. USA Department of Chemical Engineering, University of Cincinnati, Cincinnati, Ohio 45221, USA 60. Liu Y. Department of Chemical Engineering, Swearingen Engineering Center, University of South Carolina, Columbia, SC 29208, USA 61. Long R. Department of Chemical Engineering, The University of Michigan, Ann Arbor, Michigan 48109-2136, USA 62. Lopez-Cortes A. Center for Biological Research, P.O. Box 128, La Paz 23000, BCS, Mexico 63. Lopez-Garzon F.J. Grupo de Investigaci6n en Carbones, Dpto. de Quimica Inorganica, Fac. de Ciencias, Universidad de Granada, 18071 Granada, Spain 64. Lucarelli L. ThermoQuest Italy S.p.A., Strada Rivoltana, 20090 Rodano (Milan), Italy 65. L u o R.G.

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Department of Chemical Engineering, Chemistry and Environmental Science, New Jersey Institute of Technology, University Heights, Newark, NJ 07102-1982, USA Lutz W. Holzmarktstrasse 73, D-10179 Berlin, Germany Lodyga A. Fertilizers Research Institute, 24110 Putawy, Poland L u k a s z e w s k i Z. Poznafl University of Technology, Institute of Chemistry and Technical Electrochemistry, Piotrowo 3, 60-965 Poznafl, Poland MacDowall J.D. NORIT United Kingdom Ltd., Clydesmill Place, Cambuslang Industrial Estate, Glasgow G32 8RF, Scotland

xiv 70. Matyska M. Department of Chemistry, San Jose State University, San Jose, CA 95192 USA 71. Matijevic E. Center for Advanced Materials Processing, Clarkson University, P.O.Box 5814, Potsdam, New York 13699-5814, USA 72. Meda L. EniChem - Istituto Guido Donegani, 28100 Novara NO, Italy 73. Menzeres L.T. Ekostar - Nautech Company, B.Chmielnitsky 2, 630075 Novosibirsk, Russia 74. Meyer K. Bundesanstalt ffir Materialforschung und -prfifung (BAM), Zweiggelande Adlershof, Rudower Chaussee 5, D-12489 Berlin, Germany 75. Mitropoulos A.Ch. Institute of Physical Chemistry NCSR ,,DEMOKRITOS", Aghia Paraskevi Attikis, GR- 153 10, Athenes, Greece 76. Miyamoto A. Department of Molecular Chemistry and Engineering, Faculty of Engineering, Tohoku University, Sendai 980-77, Japan 77. M i z u k a m i K. Department of Molecular Chemistry and Engineering, Faculty of Engineering, Tohoku University, Sendai 980-77, Japan 78. Moon H. Department of Chemical Technology, Chonnam National University, Kwangju 500-757, Korea 79. Moreno-Castilla C. Grupo de Investigaci6n en Carbones, Dpto. de Quimica Inorganica, Fac. de Ciencias, Universidad de Granada, 18071 Granada, Spain 80. Neffe S. Institute of Chemistry, Military University of Technology, Kaliskiego 2, 01-489 Warsaw, Poland 81. N e m u d r y A.P. Institute of Solid State Chemistry and Raw Mineral Processing, Kutateladze-18, 630128, Novosibirsk, Russia 82. Nijdam D. Tauw Milieu, P.O.Box 133, 7400 AC Deventer, The Netherlands 83. Ochoa J.L. Center for Biological Research, P.O.Box 128, La Paz 23000, BCS, Mexico 84. P a n G. Department of Earth Sciences, University of Leeds, Leeds LS2 9JT, UK

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85. P a r t y k a S. UM II LAMMI ESA 5079, Case 015, Place Eugene Bataillon, 34095 Montpellier Cedex 5, France 86. P a t e l D.C. Department of Chemical Engineering, Chemistry and Environmental Science, New Jersey Institute of Technology, University Heights, Newark, NJ 07102-1982, USA 87. P e s e k J. Department of Chemistry, San Jose State University, San Jose, CA 95192, USA 88. P o k r o v s k i y V.A. Institute of Surface Chemistry, National Academy of Sciences, Prospekt Nauki 31, 252022 Kiev, Ukraine 89. Raisglid M. University of Arizona, Old Chemistry Bldg., Tucson, AZ 85721, USA 90. R a m a r a o B.V. Syracuse University, Faculty of Paper Science and Engineering and Engineering, SUNY, College of Environmental Science and Forestry, Syracuse, NY 13210, USA 91. R a o M.B.

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Air Products and Chemicals, Inc., 7201 Hamilton Boulevard, Allentown, PA 18195-1501, USA Ray M.S. Department of Chemical Engineering, Curtin University of Technology, GPO Box U1987, Perth 6845, Western Australia R e i m e r i n k W.M.T.M. NORIT N.V., Research & Development, Nijverheidsweg - Noord 72, P.O.Box 105, 3800 Ac Amersfoort, The Netherlands R e s t u c c i a G. National Council of Research, Institute for Research on Chemical Methods and Processes for Energy Storage and Transformation, S.Lucia sopra Contesse, 98126 Messina, Italy R i c h a r d s o n C.F. Radian International, LLC, 8501 N.Mopac Blvd., Austin, TX 78759, USA R i p p e r g e r K.P. Department of Chemistry, University of Tennessee at Knoxville, Knoxville, TN 37996-1600, USA R i t t e r J.A. University of South Carolina, Department of Chemical Engineering, Swearingen Engineering Center, Columbia, South Carolina 29208, USA

xvi 98. Robens E. Institut ffir Anorganische Chemie und Analytische Chemie der J.Gutenberg-Universitat D-55099 Mainz, Germany 99. R o d r i g u e s A.E. Laboratory of Separation and Reaction Engineering, University of Porto, 4099 Porto Codex, Portugal 100. Rood M. University of Illinois, Environmental Engineering Program, 205 N.Mathews Ave., Urbana, IL 61801, USA 101. R o s e n h o o v e r W. CONSOL, 4000 Brownsville Rd., Library, PA 15129, USA 102. Rostam-Abadi M. Illinois State Geological Survey, 615 E. Peabody Dr. Champaign, IL 61820, USA 103. R u l e J.

College of Sciences, Old Dominion University, Norfolk, VA 23529-0163, USA 104. Saba J. Faculty of Chemistry, M.Curie-Sklodowska University, 20031 Lublin, Poland 105. Sakellaropoulos G.P. Chemical Process Engineering Laboratory, Department of Chemical Engineering, Aristotle University of Thessaloniki and Chemical Process Engineering Research Institute, PO Box 1520, Thessaloniki 54006, Greece 106. S a m a r a s P. Chemical Process Engineering Laboratory, Department of Chemical Engineering, Aristotle University of Thessaloniki and Chemical Process Engineering Research Institute, P.O. Box 1520, Thessaloniki 54006, Greece 107. S h i n t a n i H.

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National Institute of Hygienic Sciences, 18-1 Kamiyoga 1-Chome, Setagaya-ku, Tokyo 158, Japan Silva da F.A. Laboratory of Separation and Reaction Engineering, University of Porto, 4099 Porto Codex, Portugal Silva J.A.C. Laboratory of Separation and Reaction Engineering, University of Porto, 4099 Porto Codex, Portugal Sircar S. Air Products and Chemicals, Inc., 7201 Hamilton Boulevard, Allentown, PA 18195-1501, USA S i v a s a n k e r S. Catalysis Division, National Chemical Laboratory, Pune - 411008, India

xvii 112. Stubos A.K. Institute of Nuclear Technology and Radiation Protection, NCSR ,,DEMOKRITOS", 15310 Aghia Paraskevi Attikis, GR-15310, Athenes, Greece 113. S u b r a m a n i a n D. University of South Carolina, Department of Chemical Engineering, Swearingen Engineering Center, Columbia, South Carolina 29208, USA 114. Suckow M. Fachhochschule Lausitz, Grossenhainer Strasse, D-01968 Senftenberg, Germany 115. S u n t o l a T. Microchemistry Ltd., P.O.Box 132, FIN-02631 Espoo, Finland 116. Suzuki M. Institute of Industrial Science, University of Tokyo, 7-221 Roppongi, Minato-ku, Tokyo 106, Japan 117. Szczypa J. Faculty of Chemistry, M.Curie-Sktodowska University, 20031 Lublin, Poland 118. S y k u t K. Faculty of Chemistry, M.Curie-Sklodowska University, 20031 Lublin, Poland 119. Swi~ttkowski A. Institute of Chemistry, Military Technical Academy, Kaliskiego 2, 01-489 Warsaw, Poland 120. T a k a b a H. Department of Molecular Chemistry and Engineering, Faculty of Engineering, Tohoku University, Sendai 980-77, Japan 121. T a m - C h a n g S.-W. Department of Chemistry, University of Nevada, Reno, NV 89557, USA 122. T a r a s e v i c h Yu.I. Institute of Colloid Chemistry and Chemistry of Water, 42 Vernadsky avenue, Kiev 252680, Ukraine 123. TSth J. Hungarian Academy of Sciences, Research Laboratory for Mining Chemistry, 3515 Miskolc-Egyetemvaros, P.O. Box 2, Hungary 124. Tzevelekos K.P. Institute of Physical Chemistry NCSR ,,DEMOKRITOS", Aghia Paraskevi Attikis, GR-153 10, Athenes, Greece 125. Unger K.K. Institut ffir Anorganische Chemic und Analytische Chemic der J.Gutenberg-Universitat, D-55099 Mainz, Germany

xviii 126. U s h a k o v V.A. Institute of Solid State Chemistry, SB RAS, Kutateladze 18, Novosibirsk 630128, Russia 127. V a n s a n t E.F. Laboratory of Inorganic Chemistry, University of Antwerpen (U.I.A.), Universiteitsplein 1, 2610 Wilrijk, Belgium 128. Vetrivel R. Catalysis Division, National Chemical Laboratory, Pune - 411008, India 129. V i g n e s w a r a n S. University of Technology, Sydney, Faculty of Engineering, Building 2, Level 5 P.O.Box 123 Broadway, NSW 2007, Australia 130. W a g h m o d e S.B. Catalysis Division, National Chemical Laboratory, Pune - 411008, India 131. W r 6 b l e w s k i W. Department of Chemistry, Warsaw University of Technology, Noakowskiego 3, 00-664 Warsaw, Poland 132. Y a m p o l s k a y a G.P. Moscow State University, Department Colloid Chemistry, Vorob'evy Gory, 119899 Moscow, Russia 133. Y a n g R.T. Department of Chemical Engineering, The University of Michigan, Ann Arbor, Michigan 48109-2136, USA 134. Y i a c o u m i S. Georgia Institute of Technology, School of Civil and Environmental Engineering, Atlanta, GA 30332-0512, USA

xix

Contents of V o l u m e I Preface Complete List of A u t h o r s

v IX

F u n d a m e n t a l s of A d s o r p t i o n 1. Adsorption - its development and applications for practical purposes (A.D@rowski) 2. Industrial carbon adsorbents (A.Swi~tkowski) 3. Standarization of sorption measurements and reference materials for dispersed and porous solids (E.Robens, K.-F.Krebs, K.Meyer, K.K.Unger) 4. Spectroscopic characterization of chemically modified oxide surfaces (J.Pesek, M.Matyska) 5. Advances in characterisation of adsorbents by flow adsorption microcalorimetry (A.J.Groszek) 6. Temperature programmed desorption, reduction, oxidation and flow chemisorption for the characterisation of heterogeneous catalysts. Theoretical aspects, instrumentation and applications (M.Fadoni, L.Lucarelli) 7. Adsorption with soft adsorbents and adsorbates. Theory and practice (G.F.Cerofolini, L.Meda, T.J.Bandosz)

3 69 95 117 143

177 227

A p p l i c a t i o n in I n d u s t r y 1. Advanced technical tools for the solution of high capacity adsorption separation (G.Horvath, M.Suzuki) 2. The mutual transformation of hydrogen sulphide and carbonyl sulphide and its role for gas desulphurization processes with zeolitic molecular sieve sorbents (M.B(ilow, W.Lutz, M.Suckow) 3. Nitrogen separation from air by pressure swing adsorption (N.O.Lemcoff) 4. Methodology of gas adsorption process design. Separation of propane/propylene and rgiso- paraffins mixtures (Jose A.C.Silva, F.Avelino da Silva, Alirio E.Rodrigues) 5. Fractionation of air by zeolites (S.Sircar, M.B.Rao, T.C.Golden) 6. Production, characterization and applications of carbon molecular sieves from a high ash Greek lignite (P.Samaras, X.Dabou, G.P.Sakellaropoulos) 7. Development of carbon-based adsorbents for removal of mercury emissions from coal combustion flue gas (M.Rostam-Abadi, H-C.Hsi, S.Chen, M.Rood, R.Chang, T.R.Carey, C.F.Richardson, W.Rosenhoover) 8. Sorption properties of gas/coal systems, degasification of coal seams (J.TSth) 9. The influence of properties within particles of active carbons on selected adsorption processes (B.Buczek)

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301 347

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459 485 507

Adsorption and its Applications in Industry and Environmental Protection Studies in Surface Science and Catalysis, Vol. 120 A. Dabrowski (Editor) 9 1998 Elsevier Science B.V. All rights reserved.

Environmental pollutants and application of the adsorption phenomena for their analyses J. Btadek and S. Neffe Institute of Chemistry, Military University of Technology, 00-908 Warsaw, Kaliskiego St., 2, Poland 1. I N T R O D U C T I O N H u m a n activity is now altering the global environment on an unprecedented scale and thus contributes to the environmental change affecting h u m a n health. Only few of chemical compounds present in direct h u m a n surrounding may be considered as beneficial for health; the majority of them act harmfully on humans, even in minimal doses. They occur in all environmental media (air, water and soil) and that is why we observe the increasing attention to the environmental protection. One of the most important factors in this field are the results of chemical analysis of pollutants. It is obvious that only reliable analytical data obtained during monitoring can be a base for environmental protection activities. The term monitoring means systematic and planned collection of analytical activities realised in any space to define the quality of air, water and soil. Volatile organic compounds, pesticides, polycyclic aromatic hydrocarbons, polycyclic aromatic heterocycles, phenols, polychlorinated biphenyls, organotins, chemical warfare agents and inorganic pollutants belong to the most important environmental pollutants. The need of monitoring leads to the development of independent branch of instrumental analysis - environmental analysis. It is a discrete, and sophisticated branch of instrumental analysis which concerns the treatment of environmental samples from their sampling to receiving the final result of analysis. The fundamental requirement of environmental analysis is for a fast, modern and reliable methodology, especially as the data produced are increasingly drawn upon as the decisive basis for regulatory measures. Consequently, specific conditions need to be fulfilled for the detection of pollutants in trace and ultra-trace quantities, within a short time and with a high degree of precision. To define the analytical process, Skoog and co-workers [1] mention the following steps: selecting method of sampling, obtaining representative samples, preparing laboratory samples, defining replicate samples, dissolving samples,

eliminating interference and measuring features of analyses. The aims of these activities are: 9 making the sample suitable physical parameters, removing interference and transferring the analytes to matrix being compatible with analytical technique; liquid, gas, solid phase and supercritical fluid extraction is usually applied for transferring analytes directly from samples into media being subjected to final instrumental analysis, as well as to liberate analytes trapped on sorbents during preconcentration steps; 9 cleaning-up the analytical samples and analytes enrichment; liquid-liquid partitioning, solid phase extraction, preparative column and thin layer chromatography are usually applied as clean-up and preconcentration techniques, 9 separation of sample components to obtain the chemical individuals; in environmental analyses the partition of analysed mixtures is most often realised by chromatographic methods, 9 detection, identification and quantitation; detectors which are parts of chromatographic apparatus or can co-operate with them in on-line mode are predominantly used. There are many various methods of sampling, sample preparation and analyses, which w a r r a n t correctness of obtained analytical results. Extraction, chemisorption, absorption, adsorption, distillation or freezing are used in them inter alia. Features and applications of these methods are presented in numerous compilations and monographs. In this elaboration we present only these techniques in which phenomena of adsorption are used. They are applied mainly to the sampling of pollutants in fluid, sample preparation and such analytical techniques, which w a r r a n t separation of components of analysed mixture (mainly chromatographic techniques of analyses). In these processes compounds of interest are selectively removed from the bulk sample matrix, preconcentrated, cleaned-up~ separated into individual substances and analysed.

2. SHORT CHARACTERISTIC OF MONITORED S U B S T A N C E S The term environmental pollution means any physical, chemical, or biological change disturbing ecological equilibrium in the environment. It may be a result of random, accidental events, emission of certain pollutants due to activity of nature itself, or h u m a n activities. As a result of the activity of nature, natural pollutants are emitted into atmosphere; h u m a n activity leads to the emission of pollutants called anthropogenic pollutants. Natural and anthropogenic pollutants emitted from a given source are called primary pollutants. A number of primary pollutants can undergo some changes due to reactions with other pollutants, as well as with some components of the environment. In this way, new compounds, often of higher toxicity, can be formed. They are called secondary pollutants. Primary as well as secondary pollutants occur in all of the environment media:

atmosphere, hydrosphere, and soil. The following groups of substances are considered as the most important environmental pollutants: 9 Volatile organic c o m p o u n d s . Volatile organic compounds (VOCs), originating from anthropogenic sources, are the monocyclic aromatic hydrocarbons and the volatile chlorinated hydrocarbons. Both groups of compounds are considered as priority pollutants; they are present in all parts of the environment. Monocyclic aromatic hydrocarbons are mainly emitted by industrial processes and combustion of fossil fuels, while chlorinated hydrocarbons are widely applied as solvents for dry cleaning, as degreasing agents in metal industries or as fumigants [2]. Due to their lipophilic properties, they can be taken up in lipophilic matrices. Uptake of xenobiotic VOCs in plants used for h u m a n nutrition (vegetables and fruits), results in an exposure of man through the food chain, next to a direct exposure to air pollutants through inhalation. VOCs are also the most frequently encountered contaminants at hazardous waste sites. 9 Pesticides comprise a group of compounds that are given great attention in environmental studies. They are introduced into environment due to wilful h u m a n activity; economic production in the cultivation of vegetables and fruits, as well as in agriculture, can not be achieved without pesticides. Pesticides belong to different chemical groups of compounds; the most important of them are: organophosphorous, organochloride, carbamate, triazine compounds and chlorophenoxy acids. With respect to the biological activity they are classified as insecticides, herbicides and fungicides. Well known compounds such as DDT, lindane or aldrin belong to the organochloride group which, in the past, was widely used all over the world. Although their manufacture and application are now largely prohibited, they can still, due to their persistence, be found in the soil, in animals, plants and food products. Pesticides are poisons; some of them or their degradation products also demonstrate carcinogenic potential and teratogenic activity. They are present in all parts of the environment. 9 Polycyclic a r o m a t i c h y d r o c a r b o n s (PAHs) are compounds whose molecules can contain 2-13 aromatic rings arranged in linear, cluster, or angular shapes. They may contain some number of alkyl substituents. PAHs arouse much interest mainly due to their carcinogenic and mutagenic properties. They are widespread environmental contaminants emitted from a variety of sources, including industrial combustion and discharge of fossil fuels, residential heating, or motor vehicle exhaust. In processes of monitoring, PAHs have been measured in a variety of environmental matrices including air, water, soil, sediments and tissue samples. 9 Polycyclic a r o m a t i c heterocycles. In the environment, carbon atoms in PAHs rings can be substituted with oxygen, sulphur, or nitrogen atoms. In this way polycyclic aromatic heterocycles are formed, and they usually occur together with PAHs. The most dangerous of these, polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans, are by-products formed during the

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manufacture of chlorophenols and related products; other sources include the pulp and paper industry and accidental fires that release polychlorinated biphenyls. Dibenzotiophene and some of its methyl-substituted compounds are persistent residues in sea environment after oil accidents. In the natural environment, polychlorinated thianthrenes and polychlorinated dibenzothiophenes also exist. As with their oxygen analogues, they are hazardous substances. Azaarenes, mainly benz(c)acridine and many of its related compounds, have been shown to exhibit carcinogenic activity. Nitrorelated compounds are mutagenic and carcinogenic. Polycyclic aromatic heterocycles are continually found in many natural and environmental samples. P h e n o l s form a group of aromatic compounds with one or more hydroxyl groups. Phenols and substituted phenols are products of manufacturing processes used in plastics, dyes, drugs, antioxidants, and pesticides industries. They pose the serious danger to the environment, especially when they enter the food chain as water pollutants. Even at very low concentration phenols affect the taste and odour of fishes and drinking water. Because of this, many phenol derivatives (mainly nitrophenols and chlorophenols, which are also poisons) are considered as priority pollutants of the environment. P o l y c h l o r i n a t e d b i p h e n y l s (PCBs) are a group of compounds derived from biphenyl by substituting one to ten hydrogen atoms with chlorine. There are 209 possible PCB configurations. They have extensive application because of high chemical and thermal stability, low or no flammability, low vapour pressure at ambient temperature and high permeability. PCBs are utilised alone or in mixtures as heat-transfer fluids, dielectrics for capacitors and transformers, hydraulic fluids, lubricants, additives in plastics and dyes, etc. PCBs are different in their physical and chemical properties as well as toxic potencies; some of them are inducers of drug-metabolising enzymes also being able to affect various physiological processes such as reproduction, carcinogenesis or embryonic development. O r g a n o t i n s . These compounds have been widely used as biocides incorporated in antifouling paints, and are accumulated by the biota, especially by filtrating organisms. The organotins are much more toxic than inorganic tin. Contamination of the marine environment by organotins has been well documented. Tributyltin is the most often used organotin compound, followed by triphenyltin. In water these substances can be step-wise decomposed to less substituted and down to inorganic tin, absorbed by lipid fraction of organisms or adsorbed onto particulate matter. C h e m i c a l w a r f a r e a g e n t s . The need of the monitoring on the presence of these substances in the environment results not only from the need of the verification of the Chemical Weapon Convention [3] but also because certain chemical warfare agents can be spread in the environment as the old or abandoned chemical warfare agents. Out of this group of compounds organophosphorous (O-ethyl S-2-diisopropylaminoethyl methyl phosphono-

thiolate, O-pinacolyl methylphosphono-fluoridate, etc.) and bis(2-chloroethyl) sulfide ( m u s t a r d gas), tris(2-chloroethyl) amine (nitrogen gas), 10-chloro-5,10dihydrophenarsazine (adamsite) have importance due to their toxicity or persistence in the environment. 9 E x p l o s i v e s . 2,4,6-trinitrotoluene (TNT) is known first of all as an explosive, but it appears t h a t this compound and its degradation products have been found as c o n t a m i n a n t s in w a t e r and soil. TNT and its degradation products have been identified in the blood and urine of the explosives m a n u f a c t u r i n g plants personnel. Because of the m u t a g e n i t y of these compounds, environmental t r e a t m e n t of TNT and its degradation products (2- and 4-monoaminodinitrotoluenes as well as 2,4- and 2,6-diaminonitrotoluenes) is an i m p o r t a n t issue. 9 I n o r g a n i c p o l l u t a n t s . Among inorganic environmental pollutants aerosols, heavy metals, radionuclides and some anions are monitored. Aerosol or particulate m a t t e r refer to any substance, except pure water, t h a t exists as a liquid or solid in the atmosphere under normal conditions and is in microscopic or submicroscopic size. Even non-toxic aerosols are harmful; they can cause eye or t h r o a t irritation, bronchitis or lung damage. Heavy metals (mainly As, Cd, Cr, Cu, Se, Ni, Mo, Hg and Pb) can pose serious t h r e a t s to the h u m a n health even at very low concentrations in air and water. For instance, lead causes damage of brain, mercury affects several areas of the brain, as well as the kidneys and bowels, arsenic causes cancer etc. After pollution of soil they can be incorporated into the food cycle via vegetables or, alternatively, be washed towards surface or underground water. Farming, industrial and u r b a n activities are most often mentioned as pollution sources of heavy metals. The radioactivity in environment originates from both n a t u r a l sources and h u m a n activities. The l a t t e r include operations concerned with the nuclear fuel cycles, from mining to reprocessing, medical uses etc. Radionuclides cause cancer. The common anions, such as cyanides (CN-), halides (Br-, CI-, F-) or the oxy-ions (SO3-, 304-, NO2-or NO3-) are monitored mainly in w a t e r and wastewaters. When listing the most i m p o r t a n t environmental pollutants it is impossible to forget industrial gases such as SO2, NOx, CO2, etc., which are emitted in huge quantities to the atmosphere. First two of t h e m cause respiratory illness and lung damage. They also cause the acid rains which are responsible for corrosion of metals, acidification of soil and surface waters, as well as degradation of forests. NO2 and CO2 are, like as CH4, tropospheric 03 and chlorofluorocarbons, greenhouse gases. These gases absorb in the spectral range where t h e r m a l energy r a d i a t e d from the e a r t h is at a m a x i m u m . All of them, analogically as above mentioned organic and non-organic pollutants m u s t be systematically monitored.

3. A D S O R P T I O N IN S A M P L I N G A N D S A M P L E P R E P A R A T I O N Basic feature which distinguishes environmental analysis is the need of sampling and sample preparation of substances existing in matrix on trace levels. Monitoring of polychlorinated dibenzo-p-dioxins and polychlorinated dibenzofurans can be a good example of such needs. Because of high toxicity the level of quantitation of these substances equals 10 -~2 g/kg; it is also important that these substances usually exist in natural environment in neighbourhood of other organic chlorine compounds whose concentration can be twice or three times higher. So to cope with the demands of environmental analyses such as techniques of sampling, sample preparation and analyses, which have proper ability to separation, high sensitivity, good selectivity, ability to generate reliable identification data should be applied. Adsorption phenomena play an important, if not decisive, role in many of these processes. 3.1. S a m p l i n g The term sampling is used for the description of the process by which a representative fraction of matrix is acquired. In environmental analyses various sampling techniques (and equipment submitting them) are used; adsorption phenomena are usually applied for the sampling of air, surface water and wastewater; in these processes sampling is realised together with the enrichment of analytes. Owing to the adsorption processes compounds of interest are selectively removed from the bulk sample matrix and preconcentrated (an enrichment factor of 103-107 can be usually obtained). There are two main groups of sampling and preconcentration methods of air samples: passive and aspirative (denudatic or dynamic) [4]. The idea of passive method is diffusion or permeation of analytes to the trapped medium surface. Analytes which are present in the nearest surrounding of the enriching device (dosimeter) are transferred due to the molecular diffusion forces towards the semipermeable membrane and are penetrating through it. Phenomena of absorption, chemisorption and adsorption are used in aspirative methods. Passive samplers are suitable for large scale measurements. As they do not require pumping of air during sampling they can be employed at virtually every location. Passive samplers can be sent by mail and stored before and after sampling for periods of several months. On the other hand, passive samplers require at least 24-hour exposure and therefore cannot be used for short-term sampling. Aspirative denudatic method of preconcentration consist in a junction of a forced gas stream flow and diffusive transfer of analytes in the direction of denuder wall which acts as an analyte trap. The advantage of denudating techniques is the possibility of differentiation of so called physical speciation of analytes, it means differentiation between gaseous and aerosol form of preconcentrated substances. Aspirative dynamic enrichment is the oldest method of air sampling. It allows to determine the time weighted average concentration or short term exposure level. Absorption in liquid solutions, freezing out in

cryogenic traps and adsorption belongs to these methods. Adsorption aspirative dynamic methods are used to separate the volatile and non-volatile organic pollutants. The applied techniques differ from each other in volume of sample, shape of sorbent container, and first of all in dissimilarity of used sorbents (usually they are carbonaceous, inorganic or polymeric sorbents). The scheme of the set for sampling and preconcentration of atmospheric air pollutants on adsorbent is presented in Figure la. In Figure lb the crossection of adsorption tube is shown.

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Figure la. The set for collecting samples. 1-probe, 2- adsorption tube, 3- filter, 4-capillary tubes 5-vacuum-gauge, 6-flow controller, 7-pressure reducing valve, 8-vacuum pump. Reprinted from [4].

1 2 .~/3 #y5. ~_.#6#'L.~ 8

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Figure lb. Adsorption tube. 1-plastic caps, 2-fused ends of tube (they are broken before using), 3-glass sorption tube, 4- spring, 5-glass wool, 6-adsorbent layer, 7-polyurethane plug, 8-adsorbent protective layer. Reprinted from [4].

Among carbon sorbents active carbons and carbon molecular sieves with specific surface area between 600 and 1200 me/g, and relatively high adsorption

10 capacity for most organic compounds are used. For specific non-polar analytes graphitized carbon blacks with a small specific surface area are used. Disadvantage of carbon adsorbents is an irreversible adsorption of many analytes and substantial variability of adsorption properties between different batches of the same product. Detailed description of application of carbon adsorbents in analyses of organic environmental pollutants is presented in work of Matiskowa and Skrabakov~ [5]. Among the inorganic sorbents, silica is the most widely used. Chromatographic silica is amorphous, porous solid which can be prepared in a wide range of surface areas and average pore diameters. Variation of solution pH during the acid gelation of sodium silicate yields silica with surface areas varying from about 200 m2/g (pH ~ 10) to 800 m2/g (pH < 4). Silica may be treated as a typical polar adsorbent. The raw material for the production of chromatographic alumina (aluminium oxides) are different aluminium hydroxides, e.g. hydrargillite. Like silica, alumina can be regarded as a typical polar adsorbent, and sample separation order on alumina and silica is generally similar. The presence of carbon-carbon double bonds in a pollutant molecule generally increases adsorption energy on alumina more than on silica. Aromatic hydrocarbons which contain different numbers of aromatic carbon atoms are much better separated on alumina t h a n on silica. Adsorption sites are used for the selective adsorption of u n s a t u r a t e d or polar molecules onto a hydroxylated silica surface. Three distinct site types can be recognised on the alumina surface: acidic or positive field sites, basic or proton acceptor sites and electron acceptor (charge transfer) sites [6]. Each of these is important in the adsorption of certain samples on alumina. Florisil is co-precipitate of silica and magnesia and this is why the retention and separation on its surface is generally intermediate between alumina and silica. Inorganic adsorbents have a high adsorption capability, even to polar and volatile organic compounds. This property is limited in the case of moisture samples (adsorption of water vapours cause the deactivation of adsorption centres and lowers the retention of analytes). Porous polymers and co-polymers are the most universal group of adsorbents used for sampling of air; they are synthesised in the processes of the bead polymerisation. A suitable selection of cross-linking polymers and other polymerisation parameters allows to control polymerisation processes. Therefore it is possible to obtain adsorbents with desirable specific surface area, porosity and polarity (for example Tenax | Porapak | Chromosorb | or XAD| Tenax is a porous polymer based on 2,6-diphenyl-p-phenylene oxide. The high thermal stability and its compatibility with alcohols, amines, amides, acids and bases together with good recovery characteristics make Tenax very suitable as sorbent medium in air and water analysis [7]. Porapak is a series of cross-linked porous polymers, for example divinylbenzene/ethylene glycol dimethacrylate (Porapak N). That sorbent is used for preconcentration of many substances [8]. Porapaks have the following polarity: N>S>P>Q, T>R. Chromosorbs or XAD are produced by copolymerising monofunctional monomers with bifunctional monomers. For

11 instance Chromosorb 102 is a styrene/divinylbenzene copolymer with specific surface area in the range of 300-400 m2/g; the surface is non-polar. Chromosorb 108 is moderately polar acrylic ester resin with the specific surface area between 100 and 200 m2/g. They are also commonly used for air sampling and preconcentration of analytes [9, 10]. Disadvantage of polymeric adsorbents is their sensibility to oxidative action of ozone or chlorine. Among the adsorption methods applied for isolating analytes from liquid matrixes (mainly from water) and for their preconcentration, practical importance has the solid phase extraction (SPE) technique. The idea of this technique consists in retention of analytes from a large sample volume on a small bed of adsorbent (placed in cartridge or shaped in the disk form), and following elution of analytes, with a small volume of solvent. The selection of appropriate parameters of adsorbents and solvents is the basic condition for successful employment of this method. Details on the SPE are presented in chapter 23, vol. 2 of this book. An alternative to the SPE, solvent-free sampling technique is a solid phase microextraction [SPME]. Typically, a fused-silica fibre, which is coated with a thin layer of polymeric stationary phase, is used to extract analytes from fluid (for analysis the retained analytes are thermally desorbed). The application of the SPME for sampling of polycyclic aromatic hydrocarbons [PAHs] from aqueous samples is presented in the work of Yu Liu et al. [11]. The porous layer coatings were prepared by the use of silica particles (5 ~m diameter) bonded with phenyl, Cs, and monomeric or polymeric Cls stationary phases. It was proved that several factors affected the selectivity for extraction of PAHs, including functional group in the bonded phase, and phase type (monomeric or polymeric). The distribution coefficients of PAHs in the porous layer increased with an increasing number of carbon atoms. A greater selectivity towards solute molecular size and shape were obtained using a polymeric Cls porous layer. The effect of solution ionic strength on recovery was also investigated. There are many papers describing the testing of usefulness of various adsorbents for fluid sampling [12, 13]. Adsorption capacity for the defined groups of the analytes, breakthrough capacity and influence of adsorbent bed length, as well as enrichment conditions on these parameters were investigated. The recovery of analytes by their thermal or liquid desorption is an essential element of such investigations.

3.2. Sample preparation Only few analytical methods provide the possibility for examination of samples in their original state, without preliminary preparation. In case of the environmental analysis such examinations are in practice almost impossible. Complexity of environmental samples is the reason why analytical processes are very difficult and usually multistages. Analytes need to be determined at extremely low concentrations over a wide polarity range, and frequently there is little or no information about the analysed sample. This is why the sample

12 preparation is the most important and often the most difficult step of analysis in environmental studies. The experiment described by Falcon et al. [14] can be a very good example of complications of sample preparation process. They developed the procedure for trace enrichment of benzo(a)pyrene in extracts of smoked food products. All steps of this analysis are presented in Figure 2. As it was mentioned above, the

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Figure 2. Flow-chart summarising treatment sample prior to the HPLC analysis of benzo(a)pyrene. Reprinted from [ 13].

13 transfer of analytes to matrix being compatible with analytical technique, usually by means of liquid, gas or supercritical fluids extraction, is one of the steps of sample preparation process. Unfortunately, in this process very undesirable substances (interferences) penetrate to the matrix. This is why a cleaning-up of analytical samples, connected usually with preconcetration of analytes, is a very essential step of environmental analyses. Among the adsorption methods, preparative column chromatography and thin layer chromatography are commonly used. Aluminium oxide, cellulose powder or microcrystalline cellulose, silica, diatomaceous earth (Kieselguhr), polyamide and Florisil | are employed in column or layer preparation. Nowadays, a large variety of chemicaly bonded stationary phases are applied. Such phases are prepared by anchoring specific organic moieties to inorganic oxides (mainly silica), under defined reaction conditions. Organic moieties can be attached to the silica by mono-, di-, or trifunctional silane reaction. After derivatisation of the silica substrate to yield a bonded phase, a network of socalled structure elements can be distinguished at the silica surface. This includes organic moieties bound to the surface, like cyano-, NH2-, phenyl-, octyl-, or octadecyl groups. The residual silanols, approximately 50% of the originally present silanols, have different properties as they consist of lone, vicinal and geminal groups. Consequently, besides the attached organic ligands, also the residual silanols play an important role in the final properties of the chemically bonded stationary phases. Carlsson and Ostman [15] presented a method for the isolation of polycyclic aromatic nitrogen heterocyclic (PANHs) compounds from complex sample matrix. They are known to be mutagenic and /or carcinogenic. PANHs with a single endocyclic nitrogen heteroatoms can be divided into two classes: acridines (containing a pyridine ring) and carbazoles (containing a pyrrole ring). They were isolated and separated as carbazole and acridine type PANHs with an absolute recovery in the range between 79-98%. The open column chromatography was used as an initial step for isolating a PANH fraction. By applying a normal-phase liquid chromatography using a dimethylaminopropyl silica stationary phase and utilising back-flush technique it was possible to separate the PANHs fraction into two fractions containing acridine type and carbazole type PANHs respectively. The method applied on a sample of solvent refined coal heavy distillate; acridines and carbazoles were identified by gas chromatography (GC). Rimmer's and co-workers work [16] is a good example of application of highresolution gel permeation chromatographic clean-up technique (prior to GC). The method for the determination of phenoxy acid herbicides in vegetable samples was presented. Macerated samples were extracted with acetone, filtered and acidified; the herbicides were then partitioned into dichloromethane, cleaned-up using high-resolution gel permeation chromatography before undergoing rapid and efficient methylation using trimethyl-silyldiazomethane. The resultant methyl esters were than selectively and sensitively analysed by GC/MS

14 technique. The procedure has been applied for grass samples spiked with four phenoxy acid herbicides: 2,4-D, dichlorprop, MCPA and mecroprop. Environmental monitoring is often realised by using the non-direct methods; in such investigations the results of contamination, e.g. presence of pollutants or products of their transformation in food are determined. For example milk; being at one of the highest levels of the tropic chain and due to its lipophilic nature, milk has been usually studied as an indicator of the bioconcentration process of environmentally persistent organic micropollutants. Di Muccio and co-workers [17] developed a rapid procedure that allows a single step selective extraction and clean-up of organophosphate pesticide residues from milk, dispersed on solid matrix diatomaceous material into disposable cartridges by means of light petroleum saturated with acetonitrile and ethanol. Recovery experiments were carried out on homogenised commercial milk spiked with solutions of 24 pesticides. Bernal and co-workers [18] presented a method for determination of vinclozolin (agrochemical fungicide) in honey and bee larvae. LL or SPE extraction techniques were used and two clean-up procedures (chromatography on Florisil or Cls column) were assayed after the solvent extraction. A clean-up method for organochloride compounds in fatty samples based on normal-phase liquid chromatography is described in work of van der Hoff et al. [19]. The use of liquid chromatography column packed with silica enables complete fat/organochloride pesticide separation in total fraction volume of 12 ml and results in a fully automated clean-up procedure. Adsorption phenomena in the soil sampling and sample preparation is rarely applied; it is used mainly to the clean-up of extracts. 4.

TIIE C H R O M A T O G R A P H I C M E T H O D S

The detection and determination of pollutants in complex environmental systems by conventional and biochemical methods is difficult and timeconsuming, and the results are often doubtful. These methods are now being systematically replaced by instrumental analytical methods, among which adsorption procedures play an imporatan role; crucial meaning have the chromatographic methods. The idea of all chromatographic methods is the partition of components of analysed mixture between two phases. One of these phases is stationary; the second is the mobile phase which moves along the stationary phase. Gas, liquid or supercritical fluid can be the mobile phase; the separation techniques which use these phases are called respectively: gas chromatography (GC), liquid chromatography (LC) and supercritical fluid chromatography (SFC). A solid or liquid can be the stationary phase; in the first case it is adsorption chromatography (GSC), in the second one - partitioning chromatography (GLC). If the stationary phase is in a column we call it column chromatography (GC or High Performance Liquid Chromatography - HPLC). In the case when adsorbent

15 is spread on a solid carrier plate in the form of thin layer and attached to it we call it thin-layer chromatography (TLC). In every case the separation is achieved by repeating distribution of analytes between two phases of given chromatographic system. In the column chromatography the compounds are eluted with the mobile phase to a detector (universal or selective), which produces a signal proportional to the amount of a particular substance in this phase. The proper choice of column, injection technique and temperature program will ensure the separation of interesting substances from the background ones. Good separation efficiency is one of the most critical parameters for reliable identification of pollutants by a detector. Pollutants can be identified by means of the absolute or relative retention times; a very useful parameter of identification is also retention index. Quantitation can be realised by internal or external standards. In cases of environmental analyses very frequently compounds cannot be separated from each other. These problems can often be solved by chromatographic technique utilising two or more columns. In multicolumn chromatography the columns may have widely varying measurements and separation characteristics. The columns may be connected either off-line or, nowadays much more often, on-line technique. Volatile or semi-volatile environmental pollutants which are the subject of monitoring are usually analysed by GC. In this technique sensitive and selective detectors such as the electron capture detector (ECD) or the mass spectrometer (MS) are used. They enable identification and quantitation of trace components in complex mixtures. HPLC has been recommended for the analyses of thermally labile, non-volatile and highly polar compounds. Application of high performance adsorbents in TLC and sophisticated equipment (apparatus for automatically spotting and developing chromatograms, scanning densitometry) caused, that present instrumental TLC can compete with the HPLC in terms of analytical efficiency, sensitivity, and precision. Other chromatographic methods such as SCFC and capillary electromigration have been currently developed but for the time being their application in environmental analysis is limited. The studies on applications of chromatographic methods for environmental investigation can be classified on the criteria of goals of experiments. According to this criterion they can be divided into three groups. These ones which refer to the monitoring are represented the most frequently. The reports which can be entitled "behaviour" are relatively numerous too. They refer to behaviour (in term of resolution possibilities) of pollutants in various chromatographic systems. The third group consists of the works in which physical and chemical properties of pollutants, i.e. their mobility, bioaccumulation, biotransformation etc. are examined.

4.1. High Performance Liquid Chromatography High Performance Liquid Chromatography (HPLC) is a form of column liquid chromatography. In this technique the mobile phase is pumped through the

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packed column at high pressure and therefore HPLC is also called High Pressure Liquid Chromatography. Columns are made of stainless steel tubes 10-, 20 cm long and internal diameters (I.D.) of a few millimetres. Depending on the type of interaction between stationary phase, mobile phase and a sample, following separation mechanisms can take place: adsorption, partition, ion exchange, ionpair and size exclusion. In adsorption liquid chromatography mainly silica and (rarely) aluminium oxide, cellulose and polyamide are used as stationary phases. The separated molecules are reversibly bonded to the solid surface by dipole-dipole interactions. Because the strength of interaction is different for different molecules, residence time at the stationary phase varies for different compounds; thus, separation can be achieved. This technique is used mainly for resolution of polar, non-ionic substances; in environmental analyses it is used occasionally. In the case of liquid- liquid partition chromatography stationary phases (liquids) can coat a support or can be chemically bonded to that support. Distribution mechanism is called partitioning because separation is based on the use of relative solubility differences of the sample in the two phases (in fact the separation is also achieved through the adsorption by non-protected silanol groups). In the normal phase (NP) liquid-liquid partition chromatography, the stationary phase is more polar than the mobile phase, in the reversed phase (RP) liquid-liquid partition chromatography, the mobile phase is more polar than stationary phase. The NP liquid-liquid partition chromatography is used for separation of very polar organic substances, while the RP chromatography (nowadays more popular technique) is used for the non-polar or weakly polar compounds. An example of using the liquid-liquid partition chromatography for the environmental analyses can be the above mentioned work of FalcSn et al. [14]. They used a HPLC-fluorescent detection method for the determination of benzo[a]pyrene in the enriched extract of the smoked food products. It should be stressed that the determination of polycyclic aromatic hydrocarbons (PAHs) by HPLC requires separation columns of high selectivity and efficiency. Reupert and co-workers [20] proposed a method for the separation of PAHs by the application of PAH 16-Plus column under optimal operating conditions. A very good separation of 16 PAHs was obtained (Figure 3). Liquid-liquid partition chromatography is often employed in the analysis of pesticides. The analysis of pesticide residues in the environment is of great current interest due to the possible risks that may arise from the exposure of humans and animals to such agents. From among the latest papers concerning that problem the special issue of Journal of Chromatography "Chromatography and Electrophoresis in Environmental Analysis: Pesticide Residues" is worthy to notice [21]. A good example of taking advantage of liquid-liquid partition HPLC can be the paper by Somsen and co-workers [22]. Precolumn packed with Cls (Polygosil) material for the enrichment of herbicides was combined on-line with the column liquid chromatography and Fourier-transform infrared spectrometry.

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Time, min Figure 3. HPLC chromatogram of 10 ~tl PAHs standard (EPA) in CH3CN; concentration of individual substances 90 pg/~tl. Emission signals. Column- Bakerbond PAH 16-Plus; mobile phase H20 - CH3CN (gradient elution). 1-naphthalene, 2-acenaphthene, 3-fluorene, 4-phenanthrene, 5-anthracene, 6-fluoranthene, 7-pyrene, 8-benzo[a] anthracene, 9-chrysene, 10-benzo[e]pyrene, 11-benzo[b]fluoranthene, 12-benzo[k]fluoranthene, 13-benzo[a]pyrene, 14-dibenzo[a,h]anthracene, 15-benzo[g,h,i] perylene, 16-indeno[ 1,2,3oc,d]pyrene. Reprinted from [20].

The isocratic separation was carried out on a 200x2.1 mm I.D. C18 column (Rosil) using acetonitrile-phosphate buffer (40:60) as eluent. The method was based on post-column on-line liquid-liquid extraction and solvent elimination, followed by Fourier-transform infrared spectroscopy. The feasibility of the complete system was demonstrated by analysing river water spiked with triazines and phenylureas at the ~g/1 level. Identifiable spectra were obtained for all analytes. The authors showed that on-line trace enrichment in combination with column liquid chromatography and Fourier-transform infrared detector offers a selective method for the characterisation of moderately polar analytes such as phenylureas and triazines in water samples. In the analysis of pesticides the degradation products also have to be determined because these products will often possess such activities as the parent pesticides. One ought to emphasise that the analysis of pesticide degradation in environmental samples is often difficult to perform due to the different polarities and lower concentrations of the degradation products relative to the parent compounds. Taking into account these difficulties Rollang, Beck-Westermeyer and Hage [23] applied the RP liquid-liquid partition chromatography and the high performance immunoaffinity chromatography for determining the degradation products of the herbicide atrazine in water. A high performance

18 immunoaffinity chromatography column containing anti-triazine antibodies was first used to extract the degradation products of interest from samples, followed by the on-line separation of the retained components on C18 analytical column. The limits of detection for hydroxyatriazine, deethylatriazine and deisopropylatriazine were 20-30 ng/1. Usefulness of this method was demonstrated in the analysis of both river water and groundwater samples. Rapid methods for the isolation and determination of alkylphenols from crude oils with the use of partitioning chromatography were described by Bennett et al. [24]. Determinations were performed by RP liquid-liquid partition HPLC. The authors have proved that the method affords rapid and accurate quantitation of phenol, cresols, dimethylphenols and is suitable for screening large number of samples. They illustrated the methods with two petroleum geochemical examples: determination of the partition coefficients of alkylphenols in oil/brine systems under high pressure and temperature conditions. Leira, Botana and Cela [25] applied an effect of differences in the retention capacity and selectivity of C18 and graphitized carbon column to resolve complex mixtures of non-flavonoid polyphenols (Table 1). Separation of mixture components was accomplished in a single switching operation by using mobile phase of the same composition but a different eluting strength in both separation steps. The elution conditions used in both columns were simplified by means of simulation software in order to obtain multiple fractions. The potential of this technique was realised by resolving a mixture of 38 very similar species (Figure 4).

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'

19 Table 1 Listing of the non-flavonoid species studied; key numbers match the spectrum labels in the figures, and heart-cut groups the labels in Figure 4 Key Compound Heart-cut number Group 3-Hydroxybenzoic acid III 4-Hydroxybenzoic acid II 2,4 Dihydroxybenzoic acid (~-resorcylic acid) II 2,5 Dihydroxibenzoic acid (gentisic acid) II 2,6 Dihydroxybenzoic acid (7-resorcylic acid) I 6 3,4 Dihydroxybenzenzoic acid (protocatechuic acid) I 7 3,3 Dihydroxybenzoic acid (a-resorcylic acid) I 3,4,5-Trihydroxybenzoic acid (gallic acid) 8 4-Hydroxy-3-methoxybenzoic acid (vanillic acid) III 9 10 3-Hydroxy-4-methoxybenzoic acid (isovanillic acid) III 11 4-Hydroxy-3,5-dimethoxybenzoic acid (syringic acid IV 2,4 Dimethoxybenzoic acid 12 2,6 Dimethoxybenzoic acid IV 13 14 3,4-Dimethoxybenzoic acid V 15 3,5- Dimethoxybenzoic acid VII 2-Hydroxycinnamic acid (o-coumaric acid) VI 16 3-Hydroxycinnamic acid (o-coumaric acid) V 17 18 4-Hydroxycinnamic acid (p-coumaric acid) 19 3,4-Dihydroxycinnamic acid (caffeic acid) III 20 4-Hydroxy-3-methoxycinnamic acid (ferulic acid) V 21 3,5-Dimethoxy-4-hydroxycinnamic acid (sinapic acid) V 22 3,4,5-Trimethoxycinnamic acid VII 23 2-Hydroxybenzaldehyde (salicyl aldehyde) V 24 3- Hydroxybenzaldehyde III 25 4-Hydroxybenzaldehyde III 26 2,5-Dihydroxybenzaldehyde III 27 3,4-Dihydroxybenzaldehyde(protocatechialdehyde) III 28 3,5-Dimethoxy-4-hydroxybenzaldehyde 29 2-Hydroxy-3-methoxybenzaldehyde (o-vanillin) V 30 4-Hydroxy-3-methoxybenzaldehyde (vanillin) IV 31 3-Hydroxy-4-methoxybenzaldehyde (isovanillin) IV 32 2,4-Dimethoxybenzaldehyde 33 3,4-Dimethoxybenzaldehyde (veratraldehyde) V 34 3,5-Dimethoxybenzaldehyde VII 35 3-Methoxybenzaldehyde (m.-anisaldehyde) VI 4- Methoxybenzaldehyde (p-anisaldehyde) VI 36 37 3,4,5-Trimethoxybenzaldehyde VI Chlorogenic acid II 38 Reprinted from [25].

20

Ion-exchange chromatography is a separation procedure in which ions of similar charges are separated by elution from a column packed with a finely divided resin. The stationary phase consists of acidic or basic functional groups bonded to the surface of the polymer matrix. Charged species present in the mobile phase are attracted to appropriate functional groups in the ion exchanger and separated. Mixtures of bases and acids can be separated by this technique. The stationary phases used in ion-pair chromatography are the same as in RP chromatography. Ionic organic compounds (e.g. C7H15803- - heptane sulfonic ion for bases or Bu2N § - tetrabutyl ammonium ion for acids), which form the ion-pair with the analysed sample component of opposite charge, are added to the mobile phase. This ion-pair is a salt, which behaves chromatographically like a non-ionic organic molecule that can be separated by RP chromatography. These methods found only limited application in environmental analysis. D. Krochmal and A. Kalina [26] proved that coupling the ion-exchange chromatography with active or passive sampling of air pollutants gives the possibility of simultaneous determination of sulphur dioxide and nitrogen dioxide. Both gases can be quantitatively absorbed in aqueous solution of triethanolamine and subsequently determined with ion chromatography as sulphates and nitrates. Absorbing solutions were analysed with a single column ion chromatograph equipped with a packed column. Size-exclusion chromatography is a powerful technique applicable for separation of high-molecular-weight pollutants. Packing material for sizeexclusion chromatography consists of a small silica or polymer particles containing network of uniform pores into which solute and solvent molecules can diffuse. In the chromatographic process molecules are effectively trapped in pores and removed by the flowing mobile phase. The compounds with higher molecular weight cannot penetrate into the pores and are retained to a less extend t h a n smaller ones. Some of size-exclusion packing materials are hydrophilic and are used with aqueous mobile phase (gel filtration); others are hydrophobic and are used with non-polar, organic solvents (gel permeation). In environmental analyses size-exclusion chromatography is used for sample clean-up and fractionation. For example, gel permeation chromatography is a standard technique for the isolation of herbicides and fungicides from samples that contain high-molecular-weight interferences, such as solid waste extract, oil or fats [16]. In the case of environmental analyses information about pollutants may be obtained not only from environmental matrix. Kabzifiski [27] proposed a new analytical method for the quantitative determination of metallothioneins protein in h u m a n body fluids and tissues, in order to determine the level of environmental and industrial exposition to heavy metals. For metallothioneins isolation covalent affinity chromatography with thiol-disulfide interchange was applied, which is a modern technique of separation of high affinity, good repeatibly and reproducibility, allowing specific isolation of the thiolproteins and metallothiolproteins. Fundamentals of indirect determination of the contents of metallothioneins protein were worked out throughout estimation of the

21 quantities of metals bound with metallothionein protein and adsorbed on covalent affinity chromatography gel as on the solid-phase extraction support during a separating process.

4.2. Gas c h r o m a t o g r a p h y The term gas chromatography (GC) is used to denote the chromatographic techniques in which the mobile phase is a gas (the carrier gas, mostly N2, H2, Ar or He). The stationary phase is placed in the column; it may be a porous solid (GSC-gas solid chromatography, adsorption chromatography) or viscous liquid (GLC-gas liquid chromatography, partition chromatography). In both cases the transport of components of analysed mixtures (adsorbates, analytes) is realised exclusively in the gas phase, separation - exclusively in the stationary phase. The time of passing of particular analytes through the stationary phase and the frequency of interactions of analytes with this phase are the decisive factors in the separation process. In case of GSC separation occurs because of differences in the adsorption equlibria between the gaseous components of the sample and the solid surface of the stationary phase. In case of GLC, in contrast to HPLC, there is no interaction between the mobile phase and the analyte. Glass, metal (copper, aluminium, stainless steel) or Teflon tubes long 2-3 m and I.D. 2-4 mm are used for making the packed columns to GC. Open tubular columns (capillary columns) are of two basic types: wall coated open tubular (WCOT) and support coated open tubular (SCOT). WCOT is the traditional capillary column made of glass or stainless steel. The liquid phase is applied as a continuous thin film on the inside walls of the tube. The newest WCOT are fused silica open tubular columns (FSOT). This is a very small outer diameter thin wall column that is inherently a straight tube but is flexible enough to be coiled to diameters c.a. 20 cm. FSOT are drawn from specially purified silica that contains minimal amounts of metal oxides. Compared to packed column these capillaries show inert surfaces and higher reproducibility with at last equal separation efficiencies. PLOT (porous layer open tubular) column is similar to a SCOT except for the fact, t h a t the support material is responsible for the separation through the adsorption process. In a PLOT columns there is no coating liquid phases. There are two basic types of packing materials employed in GC. The first type is porous materials used in GSC. The second type are the support materials which are covered with a layer of liquid phase used in GLC. The adsorption properties and selectivity of adsorbents applied in GSC depend first of all on the chemical composition and geometrical structure of their surface. There are several kinds of attractive adsorbate-adsorbent interactions occurring during the separation of mixturecomponents. The most important interactions are: dispersion or London forces, electrostatic forces, induction forces and specific interaction (mainly charge-transfer, which occur between one component with nbonding electrons and showing small ionisation energy and the second component showing high electron affinity). Among dozens of different solids which have been

22 used in adsorption chromatography only few adsorbent types have wide application today. Non-organic adsorbents such as silica, aluminium oxide or Florisil and polymeric adsorbents type of Tenax, Chromosorb or Porapak belong to the porous packings (which do not need to be coated with stationary phases). They can directly be used for adsorption chromatography. The carbonaceous adsorbents are today used in gas adsorption chromatography rather occasionally. In case of GLC the stationary phase is a liquid (often rubber-like), it is immobilised on the surface of a solid support by adsorption or by chemical bonding. Liquid stationary phases are applied both in packed and capillary columns. Packed columns are completely filled with a packing, liquid stationary phases coating an inert support such as diatomite (Kieselguhr), rarely Teflon or glass spheres. Capillary columns do not require a support because their inert walls are coated with the stationary phases. The most important feature of liquid stationary phase is its polarity. The very popular non-polar phases are Squalane (hexamethyltetracosane) and Apolane-87 (24,14-diethyl-19,29-dioctadecylhaptatetracontane). Squalane is used as reference for determination of polarity of other liquid phases in packed column. Apolane-87 is high temperature standard phase used in capillary chromatography. In environmental analyses semipolar phases are used most often. That group of phases is mainly represented by Silicones. Depending on the kind of substituent in oxosilanes chain (dimethyl-, phenyl-, trifluoropropyl-, cyano- etc.), the weak-, medium- and strong polar phases can be prepared. Polygethylenelycol is an example of strong-polar phase. Among specific liquid phases a family of polysiloxane stationary phases (Chirasil), developed for the separation of optical enantiomers, has a great practical importance. Chemicaly bonded phases used in GLC are identical as twere used in HPLC. Barrefors et al. [28] showed that furan and alkylfurans might be selectively analysed on PLOT (aluminium oxide) columns, since other oxygen-containing compounds are normally not eluted. Furan, 2-methylfuran, 3-methylfuran, 2,5-dimethylfuran and the five isomeric C6 alkylfurans, two C7 and three C6-C7 alkenylfurans were determined by adsorbent sampling and GC/MS technique. Separation on PLOT column is presented on Figure 5. Furan elutes after isoprene and cyclopentadiene in the same region as minor pentadienes and branched hexanes. Several minor C6 and C7 furans appear.in the chromatographic range before and after methylbenzene. The purpose of this study was to characterise volatile furans in birdwood smoke which may be of interest with respect to human exposure and as indoor and outdoor wood-smoke tracers in studies of air pollutants. An analytical method to determine highly volatile saturated aldehydes, degradation products of lipid peroxidation, was developed for the capillary GC [29]. The carbonyl compounds were derivatized quantitatively with 2-hydrazinobenzothiazole at room temperature to form their corresponding water-insoluble hydrazones. The derivatives were extracted and detected with high selectivity (Figure 6) by high-resolution GC with nitrogen-phosphorous

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34 ISTD,

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Figure 6. Typical gas chromatogram of the 2-hydrazinobenzothiazole-derivative aldehydes. ISTDinternal standard: 2,4 pentanedione-2-hydrazinobenzothiazole-derivative. Reprinted from [29].

24 detection due to their high nitrogen content. Analyte concentration, pH and type of extraction technique (LLE and SPE) were studied to determine optimal recovery conditions. The method was applied to the analysis of the volatile aldehydes generated during the t h e r m a l oxidation of olive oil at 220~ Begerov and co-workers [30] applied the screen method for the simultaneous d e t e r m i n a t i o n of 28 volatile organic compounds in the indoor and outdoor air at environmental concentrations. Using passive (sorption-diffusive) samplers, the volatile organic compounds were adsorbed onto charcoal during a four-week sampling period and subsequently desorbed with carbon disulphide. The eluate was split via an Y-connector and led onto two capillary columns of different polarity switched in parallel. This dual column configuration provides additional information about the volatile organic compound components and can be obtained for verification purpose. Detection was in both cases performed by connecting each column with a non-destructive electron-capture detector and a flame ionisation detector switched in series. The procedure has been successfully applied in the context of a large field study to measure outdoor air concentration in three areas with different traffic density (Figure 7). It is applicable to indoor air m e a s u r e m e n t s in a similar manner.

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Time (min) Figure 7. Typical gas chromatograms of an indoor air sample obtained by flame ionisation detection. (a) more polar column, (b) less polar column. Reprinted from [30]. Lobiafiski et al. [31] studied the potential of the microwave-induced p l a s m a atomic emission detector for capillary GC (GC/AED) as a tool for the specification of organotin compounds in environmental samples. The operational variables are optimised for chromatographic resolution and detection limits. A comprehensive

25 method for the determination of mono-, di-, tri-, and some tetraalkylated organotin compounds in water and sediments by GC/AED was developed. Ionic organotin compounds were extracted as diethyl dithiocarbamates into pentane and, after its evaporation, dissolved in a small volume of octane and derivatized by pentylmagnesium bromide to give the solution suitable for gas chromatography. The phenoxy acids were first introduced as herbicides in the late 1940s. They have found widespread usage in the post-emergence control of annual and perennial broad leafed weeds cereals and grasses. Functioning as synthetic plant growth regulators these herbicides accumulate in the roots and stems of the plants. A method for the determination of phenoxy acid herbicides in vegetation samples is described among other things in the work of Rimmer and co-workers [32]. Macerated samples were extracted with acetone. After filtration and acidification they were introduced into dichloromethane. The herbicides were than cleaned-up using high-resolution gel permeation chromatography. Analysis of PCB normally includes extensive sample clean-up and preconcentration followed by high resolution capillary GC either with electron capture or mass-selective detection. Although both techniques provide the high sensitivity required for PCB investigations, quantitative analysis is complicated by structural variations of detectors-response factors. The quantitative aspect of GC with atomic emission detection (GC/AED) used for the analysis of PCB is presented in work of Bjergaard et al. [33]. Since Cl-responses were almost independent on the PCB structure, individual PCBs were quantitated with an accuracy not better than 10% by utilising a Cl-calibration plot based on a single randomly selected congener (universal calibration). In addition, within 5-10% accuracy, GC/AED enabled estimation of total PCB residue levels and calculation of the percentage by weight of chlorine in mixtures containig PCB. Thus PCB detection limits were higher with GC/AED than GC/ECD. The GC/AED technique was very attractive for PCB and enabled significant simplification of PCB quantitation. A fraction of polycyclic aromatic nitrogen heterocycles (PANHs) from the environmental samples consists of a complex mixture of compounds, due to the large number of isomers. This cause problems with co-eluting peaks when using chromatographic separation technique. Thus, chemical analysis of PANHs requires a group separation of acridine- and carbazole-type compounds in order to facilitate identification, as well as quantitation. The method of solving this problem is presented above [15]. Separated acridine and carbazole groups were analysed by means of GC technique (Figure 8) with using nitrogen-phosphorous detector (NPD). There was no overlap between the PAH and PANH fractions. PAHs were detected with coupled LC/GC-flame ionisation detection. Isomeric selectivity (for the separation of anthracene/phenanthrene) of new monomeric and polymeric liquid crystalline stationary phases, as well as of common non-polar and polar stationary phases in capilary gas chromatography were compared in the work of Kraus and co-workers [34]. The high isomeric

26

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Figure 8. GC/NPD chromatogram of acridine fraction from solvent refined coal heavy distillate. Reprinted from [15].

selectivity of monomeric liquid crystalline phase [4,4'-biphenylene-bis-(4-n-butyloxybenxoate] was used for the separation of critical pairs of polycyclic aromatic hydrocarbons. Liquid crystal stationary phases are characterized by their chemical structure, as well as by the ordered structure of the mesophase type at given temperature. This is why using liquid crystal stationary phase the solutes retention does not only depend on their vapour pressure and their interaction with the chemical structure of the stationary phase but also on their molecular shape. The isomers having higher length to breadth ratio of the molecule and/or planar molecular shape have an increased retention in the liquid crystalline mesophase. The advantage of using liquid crystals as stationary phases in GC for separation of isomeric PAHs was also demonstrated by Apfel at al. [35] and Hahne [36].

4.3. Thin Layer Chromatography TLC is one of the best known and thoroughly tested method of the analysis of environmental pollutants [37]. The one-dimensional ascending or horizontal techniques have usually been applied for the development of chromatograms in a closed chamber. Multiple and two-dimensional development techniques have seldom been used. Recent, instrumental techniques such as forced flow planar layer chromatography (FFPC), automated multiple development (AMD) and gradient development techniques are becoming more frequently used. Multimodal TLC (TLC/GC, TLC/MS, etc.) have occasionally been applied. Details concerning the new TLC techniques were described in the review of Jork [38].

27 The separation of environmental mixture samples is usually performed on commercial chromatographic plates. The adsorbent is spread as a thin layer onto a suitable solid support (e.g. glass plate, polyester or aluminium sheet). In TLC the same adsorbents are used as in the HPLC technique. Most often separations are performed with silica 60 (pore diameter = 6 nm). Other commercial adsorbents are Kieselguhr, aluminium oxide, cellulose, polyamide and ion exchangers. At present the modified silica (amino -NH2 or cyano -CN and reversed phases such as octyl RP-Cs, octadecyl RP-Cls,) are used. Impregnated silica is mainly applied for analysis of PAHs and heavy metals. H. Engelhardt and P. Engeld [39] showed the possibility of application of TLC technique to the quantitative determination of hydrocarbons in waste water after extraction with n-heptane by means of a micro separator. Chromatographic development was carried out with n-heptane. Quantitation was done by IR spectroscopy or after dyeing with acid violet reagent. Application of acid violet reagent results in no difficulties relating to the proper selection of the hydrocarbon standards in confrontation with IR quantitation technique, and the standard TLC scanner could be used for quantitation. As it was mentioned above, aminoarenes are a class of compounds usually accompanying PAHs in environmental samples. Janoszka an co-workers [40] applied semipreparative TLC to separate these substances from other polar compounds present in sludge extract isolated by SPE technique. Chromatograms were developed on A1203 plates with ternary mixture of organic solvents to distance of 9 cm in DS chamber. The data obtained by using TLC were confirmed by GC/MS analysis. TLC has found frequent extreme application for environmental analysis of pesticides. Rathore and Begun [41] refer to ca. 300 papers in their recapitulation of TLC methods for pesticide residue analysis. Advances in the application of TLC for separation, detection, and qualitative and quantitative determination of pesticides, other agrochemicals, and related compounds are reviewed in Sherma's article [42]. The author showed the possibility of application of TLC for pesticide analysis in different matrices such as food and environmental samples, and for analyses of residue pesticides of various types, including insecticides, herbicides, and fungicides, belonging to different chemical classes. Bt~dek and co-workers [43] proved the possibility of application of thin layer chromatography and SPE to the analyses of pesticide residues in strongly contaminated samples of soil. Modern TLC equipment was used in these investigations. Chromatograms were developed in a normal-phase system by automated multiple development gradient elution. Limitations of detectability by TLC were compensated for by the application of relatively large volumes (by spray-on technique) of analysed solutions on start lines. Quantitative assessment was achieved by UV absorption measurement scanning of the chromatograms by a ,,zig-zag" technique (Figure 9) Recovery and error of the method was estimated - the recovery level was 80% and the R.S.D. was less than 9%.

28

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Petrovi6 and co-workers [44] described a study on the retention behaviour of series of thiosemicarbazide derivatives and their 1,2,4-triazoline-3-thiones by normal- and reversed- phase TLC using silica gel, alumina, and C ls- modified silica gel layers, as well as non-aqueous and aqueous eluents. Two types of compounds were studied: thiosemicarbazide derivatives and 1,2,4-triazoline-3thione derivatives. Retention data were discussed in relation to the molecular structure of the solute, and the nature of the stationary phase and eluent. The study can be an example of application of TLC for cognitive purposes. Substituted phenylureas (such as diuron, linuron, metoxuron etc.) are widely used as selective herbicides. Lautie and Stankovic [45] applied an instrumental TLC for determination of six phenylurea herbicides in food. The pesticides were extracted with acetone and purified by SPE. Analysed and standard solutions were spotted to the plates by means of spray-on technique. Chromatograms were developed in 25 steps with the use of gradient elution. Quantitative analysis were based on the measurements of UV (k = 245 nm) absorbance by using a scanner densitometer. The high selectivity, high detectability and reliability of analysis under fairly simple conditions contribute to the effective use of TLC for the detection of chemical warfare agents. It is proper to add that the problem connected with the determination of substances classified as potential warfare agents lie also in the non-military sphere of interest. This concerns, for instance, the uncontrolled spread of toxic substances as a result of industrial break-down or agrotechnical operations, and the generation of poisons, i.e. fluoroacetic acid in plants or phosgene in the troposphere. Application of TLC for military purposes, including analytical procedures for chemical warfare agents, has been recommended by many workers. The review on application of chromatographic methods for

29 chemical warfare agents analysis has been done by Witkiewicz at al. [46]. The same authors, testing a new instrument to the overpressured thin layer chromatography, proved [47] that the instrument can be used (among many other applications) to the analyses of organophosphorous chemical warfare agents (Figure 10). Currently, due to the Chemical Weapon Convention requirements, interest in the TLC technique is increasing; it can be used as "screening method"[48].

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Listing the application of adsorption TLC for military purposes, it is impossible to forget about explosives. The most widely found explosives are trinitrotoluene, hexogen, oktagen, tetryl and pentrit. Application of TLC for determination of high explosive residues in water and soil samples is presented in work of Bt~dek and co-workers [49]. In this group of compounds the most dangerous (because of

30 toxicity and wide application) is trinitrotoluene (TNT); in environment it can be reduced by bacteria to aminodinitrotoluenes, dinitrotoluenes and both di- and trinitrobenzenes. The most dangerous are aminodinitrotoluenes; they show severe toxicity and mutagenity [50]. TLC has been also used to solve a variety of analytical problems relating to the identification, separation and determination of inorganic compounds in environmental samples (mainly heavy metals). The review by Mohammad [51] summarises the application of TLC in the analysis of environmental samples for harmful and toxic metals (inorganic and organic substances containing Ag, A1, As, Be, Bi, Cd, Co, Cr, Cu, Fe, Hg, Mn, Ni, Sb, Sn, and Zn). In most cases, extraction of analytes from the matrix, clean-up of the extracts and concentration of the analytes precede TLC analysis. The author showed that spectrophotometry, titrimetry, atomic absorption spectroscopy, densitometry, fluorimetry and solvent extraction techniques have been combined with TLC for sensitive identification, quantification, and selective separation of toxic heavy metals present in rivers and sea water, waste water, sludge, aquatic plants, cosmic dust, air and airborne dust.

4.4. Supercritical Fluid Chromatography Supercritical fluid chromatography (SFC) combines some characteristics of both gas and liquid chromatography. In recent years the interest in the use of SFC as a separation technique has been increasing rapidly because of the unique properties of supercritical fluid; its higher diffusivity and lower viscosity enable analysis to be 3-10 times faster than HPLC. On the other hand, it has relatively similar density to liquid and viscosity comparable to gas, so SFC can be used to analyse a wide range of compounds, particularly those that are thermally labile, non-volatile and of high molecular mass, that cannot be satisfactorily analysed by GC. The most widely used supercritical fluid is CO2, however, for analyses of polar and high-molecular-mass solutes, polar modifiers such as methanol must be incorporated to increase the solvent polarity. Supercritical CO2 is an ideal solvent for preparing samples for GC, LC, SFC and other analyses. CO2 has easily accessible critical point and its solvating power can be controlled to match a wide range of hazardous organic solvents. Yet even as it assumes liquid-like solvent properties, supercritical CO2 retains gas-like viscosity and penetrates solid samples quickly. E. Pocurull and co-workers [52] studied the possibility of application of SFC with diode array detection system to determine phenol and nitrophenol in water. Several columns and the influence of chromatographic conditions (temperature, pressure, flow-rate and adding methanol in the mobile phase) were studied in order to separate the compounds. To decrease the detection limits of the method, SPE on-line coupled to SFC was tested. Tetrabutylammonium bromide was used as ion-pair reagent in the extraction process to increase the breakthrough volumes. The separation of five phenolic compounds, in the time period less than

31 6 min., with good resolution for all compounds was achieved. The performance of the method was checked for tap and river water samples. 5. MISCELLANEOUS METHODS In recent years, capillary electrophoresis (CE) has developed into an versatile and powerful technique. It was applied for separation of a wide variety of compounds, ranging in size from small ions to large biomolecules. The main separation modes of CE are: capillary zone electrophoresis (CZE), capillary gel electrophoresis (CGE), micellar electrokinetic capillary chromatography (MECC), capillary electrochromatography (CEC), capillary isoelectric focusing (CIEF) and capillary isotachophoresis (CITP). Various modes of separation, high resolving power and small sample requirement (CE is know as a nanoscale technique) have made possible a wide range of applications. CE is an analytical tool, possessing some typical features of chromatographic techniques and some features of traditional slab gel electrophoresis [53]. For example, micellar electrokinetic capillary chromatography couples both the electrophoresis and chromatographic partitioning element for simultaneous separation of charged and neutral compounds. In Figure 11 the versatility and

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I

12

Time (min) Figure 11. MECC separation of eleven herbicides in distilled water at the mg/l level. 1-tribenuron, 2-chlorsulfuron, 3-metsulfuron, 4-paraquat, 5-simazine, 6-atrazine, 7-1inuron, 8-terbuthylazine, 9-alachlor, 10-metolachlor, 11-trifluralin. Reprinted from [53].

32 the efficiency of CE in the separation of herbicides is shown (conditions of experiment: herbicide concentration range 1-2 mg/1, capillary 500 mm x 7,5 mm I.D.; operating voltage 25 kV at 30~ UV detection at 214 nm; separation buffer: 30 mM sodium borate and 30 mM sodium dodecyl sulphate, pH 8.0). During application of CE for monitoring analyses and environmental studies of pollutants, particular attention has to be devoted to the optimisation of the separation process in order to obtain the best selectivity in a complex matrix, where many potential compounds may interfere. In order to optimise CE separation, several parameters have to be taken into account, such as electrolyte buffer composition, capillary dimensions, capillary temperature, applied voltage and mode of injection and detection. J un Liu et al.. [54] employed in CE a palladium particle-modified carbon microdisk electrode for the simultaneous detection of hydrazine, methylhydrazine and isoniazid. Analytes were separated by CZE and MECC techniques with the Pd-modified microdisk electrode; it had high catalytic activity for hydrazine and exhibited good reproducibility and stability. Quantitative determination of total molecular concentration of bioaccumulatable organic micropollutants in water, using Cls empore disk is presented in van Loon and co-workers paper [55]. Chemical group parameters such as dissolved organic carbon, absorbable organohalogen and chemical oxygen demand determination are routinely used for water quality monitoring. These parameters gave information about the degree of anthropogenic contamination and potential aquatic toxicity of water systems. A new, highly sensitive and quantitative group parameter to determine total molar concentrations of organic micropollutants that can bioaccumulate in the lipid phase of aquatic organisms from effluents, surface water and drinking water has been developed in this work. Cls empore disk was used as a surrogate lipid phase. The partition between water and Cls empore disk was employed to simulate the bioaccumulation process. After partition extraction of the water sample, the empore disk was extracted with cyclohexane and total molar concentration of organic micropollutants was determined. Vapour pressure osmometry and GC/MS were used in these investigations. Measurements of pesticides mobility in soil by application of TLC are presented in work of Camazano et al. [56]. The effect of soil improvement by using urban compost, agricultural organic amendments and surfactants on the mobility of two sparingly-soluble pesticides (diazinon and linuron) was studied. The modifications in Rf values due to the addition of the amendments were similar for both pesticides. No significant correlation was found between the Rf values and the content of total organic carbon in the amended soils. Authors demonstrated that not only the organic carbon content of amended soils but also the amendment nature, especially their contents in a soluble fraction play a very important role in the pesticide mobility. The surfactants gave rise to important alterations in pesticide mobility. The mobility of pesticides changed from being immobile in the soil sample modified with tetradecyltrimethylammonium

33 bromide to being slightly mobile in natural soil and to being mobile in the soil sample amended with sodium dodecyl sulphate. In recent years repid development of small, simple and portable devices for detection and quantitative determination of pollutants in air, water, soil and biological materials is observed. These devices are composed of detecting system and electronic part with displaying possibilities. Adsorption of analytes at the detector surface results in changing of its physical, chemical or biological properties which can be easily converted to the changes of electric signals. In o p t i c a l s e n s o r s a change of the factor of light refraction or effect of fluorescent quenching because of adsorption of analytes are usually applied. E l e c t r o c h e m i c a l s e n s o r s are based on the changes of cell electromotance due to adsorption of analytes at the surface of ion selective electrodes. In the construction of piezoelectric sensors quartz resonators are used. They are called quartz microbalances or adsorption detectors because changes of crystal frequencies are caused due the adsorption of analysed pollutants. To increase the sensitivity and selectivity of such sensors, special liquid coating materials are used. Piezoelectric sensors for SO2, NH3, H2S, Hg, nitrocompounds, chlorinated hydrocarbons and other environmental pollutants are commonly used. The principle of s u r f a c e a c o u s t i c w a v e s e n s o r s action is similar. CONCLUSIONS Environmental analysis is a broad branch of analytical chemistry in which different analytical techniques are applied. Besides chromatographic methods described above, the spectrophotometric, spectrometric and electrochemical methods are of great importance. In this work we focused our attention only on processes and methods in which the adsorption phenomena play an important role. Quoted examples of application of adsorption phenomena in sampling, preconcentration, clean-up and analyses processes are a small part of the huge collection of works devoted to the discussed problems. These examples confirm that adsorption phenomena present one of the most important and promising tools for characterisation, identification and determination of trace pollutants in various environmental samples.

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35 28. G. Barrefors, S. Bj0rkqvist, O. Ramn~is and G. Petersson, J. Chromatogr., 753 (1996) 151. 29. E. E. Stashenko, J. W. Wong, J. R. Martinez, A. Mateus and T. Shibamoto, J. Chromatogr., 752 (1996) 209. 30. J. Begerow, E. Jermann, T. Keles, T. Koch and L. Dunemann, J. Chromatogr., 749 (1996) 181. 31. R. Lobiahski, W. M. R. Dirkx, M. Ceulemans and F. C. Adams, Anal. Chem., 64 (1992) 159. 32. A. Rimmer, P. D. Johnson and R. H. Brown, J. Chromatogr., 755 (1996) 245. 33. S. Pedersen-Bjergaard, S. I. Semb, E. M. Brevik and T. Greibrokk, J. Chromatogr., 723 (1996) 337. 34. A. Kraus, G. Kraus, R. Kubinec, I. Ostrovsk~ and L. Soj~k, Chem. Anal. (Warsaw), 42 (1997) 497. 35. A. Apfel, H. Finckelmann, G. M. Janini, R. J. Laub, B. L(ihmann, A. Price, W. L. Roberts, T. J. Shaw and C. A. Smith, Anal Chem., 57 (1985) 651. 36. F. Hahne, Disertation, Martin-Luther-Universit~it, Halle 1990. 37. J. Bt~dek, Thin Layer Chromatography in Environmental Analysis, in: Practical Thin Layer Chromatography, B. Fried and J. Sherma (eds.), CRC Press, Boca Raton, 1996, 153. 38. H. Jork, J. Planar Chromatogr., 5 (1992) 4. 39. H. Engelhardt and P. Engeld, J. Planar Chromatogr., 5 (1997) 336. 40. B. Janoszka, K. Trypieh and D. Bodzek, J. Planar Chromatogr., 6 (1996) 450. 41. H. S. Rathore and T. Begum, J. Chromatogr., 643 (1993) 271. 42. J. Sherma, J. Planar Chromatogr., 2 (1997) 80. 43. J. Bt~dek, A. Rostkowski and M. Miszczak, J. Chromatogr., 754 (1996), 273. 44. S. M. PetroviS, E. LonSar, N. U. PrisiS-Janjid and M. LazareviS, J. Planar Chromatogr., 1 (1997) 26. 45. J. P. Lautie and V. Stankovic, J. Planar Chromatogr., 2 (1996) 113. 46. Z. Witkiewicz, M. Mazurek and J. Szulc, J. Chromatogr. 503 (1990) 293. 47. Z. Witkiewicz, M. Mazurek and J. Bt~dek, J. Planar Chromatogr., 6 (1993) 407. 48. J. Bt~dek, S. Neffe and A. Rostkowski, Estimation of the Possibilities of Application of Thin-Layer Chromatography for Chemical Weapon Convention, NATO SICA Meeting in Baltimore, May 1997. 49. J. Bt~dek, A. Paplihski, S. Neffe and A. Rostkowski, Chem. Anal. (Warsaw), in press. 50. N. G. McCornick, F. E. Feeherry and H. S. Levenson, Appl. Environ. Microbiol., 31 (1976) 949. 51. A. Mohammad, J. Planar Chromatogr., 1 (1997) 48. 52. E. Pocurull, R. M. Marc~, F. Borrull, J. L. Bernal, L. Toribio and M. L. Serna, J.Chromatogr., 755 (1996) 67. 53. G. Dinelli, A. Vicari and P. Catizone, J. Chromatogr., 733 (1996) 337. 54. Jun Liu, W. Zhou, T. You, F. Li, E. Wang and S. Dong, Anal. Chem., 68 (1996) 3353.

35 55. W. M. G. M. van Loon, F. G. Wijnker, M. E. Verwoerd and J. L. M. Hermes, Anal. Chem., 68 (1996) 2916. 56. M. Sanches-Camazano, M. J. Sanches-Martin, E. Poweda and E. Iglesias-Jimdnez, J. Chromatogr., 754 (1996) 279.

Adsorption and its Applications in Industry and Environmental Protection Studies in Surface Science and Catalysis, Vol. 120 A. Dabrowski (Editor) 9 1998 Elsevier Science B.V. All rights reserved.

37

F u n d a m e n t a l s of s o l i d p h a s e e x t r a c t i o n a n d i t s a p p l i c a t i o n to environmental analyses M. E. Raisglid and M. F. Burke Department of Chemistry, University of Arizona, Tucson, AZ, USA 1. S A M P L E P R E P A R A T I O N O V E R V I E W

It has been the convention to extract an analyte as a means of sample preparation using a liquid-liquid extraction technique. In liquid-liquid extraction, the analytes are partitioned between to immiscible solvents. The partitioning is based on a difference in solubility of the analyte in each of the phases, and is a very non-selective process. Efficient separation is dependent on obtaining intimate contact between the two immiscible phases. One limitation of the liquidliquid extraction method is that for intimate contact to be achieved, the phases must be shaken vigorously, which often results in an emulsion that can be difficult to break. An additional limitation is that large sample and extraction volumes are required. After the extraction step, further concentration of the extraction solvent is often necessary in order to obtain detectable levels of the analyte. LLE can be a very time consuming process. The type of manipulations that are required to perform a liquid-liquid extractions lend this technique to being a serial and largely manual process. A different approach to sample preparation is to exploit the interactions at the liquid - solid interface. This has come to be known as solid phase extraction, or SPE. This technique has found its roots in liquid-solid extraction (LSE) which was classically carried out by adding the sorbent to a vessel containing the analytes in a liquid phase, and shaking for a controlled length of time. After the distribution of the analytes reached equilibrium, the phases were separated by either filtering or decantation. The analytes could then be desorbed using a suitable solvent [1]. 2. SOLID P H A S E E X T R A C T I O N O V E R V I E W

There are a wide variety of applications for solid phase extraction as a sample preparation technique, including many environmental applications such as the determination of pesticides and other contaminants in drinking water, waste water, soils and sludge. SPE can be defined as the separation or removal of an analyte or analytes from a mixture of compounds by selectively partitioning the compounds between a

38

stationary solid phase (sorbent) and a mobile liquid phase (solvent). This process is dependent first on our ability to extract the analytes from their matrix and retain them on a solid surface. Once the analyte is extracted from the matrix, it can be eluted from the sorbent using a selective solvent. In solid phase extraction, a "liquid" phase is immobilized onto a solid surface. The analyte is partitioned between the liquid phase (sample matrix) and the modified solid surface. There are a variety of species that can be immobilized on the solid surface, ranging from hydrophobic, to polar, to ionizable species, and a particular solid phase can be chosen to retain analytes having a specific functional group. Recoveries are often improved using solid phase extraction versus LLE. In a review article, Chladek and Marano compared a solid phase extraction procedure with the standard EPA method 625 liquid -liquid extraction and obtained an average recovery with C8 that was 20% higher t h a n that obtained with the standard method [2]. The emulsion problem which often plagues liquid-liquid extractions is eliminated, since the extraction no longer requires the mixing of two liquid phases. Intimate contact is obtained between the two phases by utilization of a solid phase that is made up of small particles packed into a bed. Typically the solid phase consists of particles that are nominally 50 microns in diameter. The particles are very porous, which provides the large surface area. A sample containing the analytes of interest is then passed through the packed bed of particles. The small packed bed provides an inherent concentration step since the sorbent has a relatively small void volume (-1.2 mL/g). The solvent volume required to elute the analytes from the solid phase is correspondingly small (typically two bed volumes), and so the concentration step which is often required for liquid-liquid extractions is eliminated or greatly reduced for solid phase extractions. Solid phase extraction, unlike liquid-liquid extraction is also very amenable to batch processing and to automation. One of the important functions that solid phase extraction can serve is the isolation and purification of an analyte. Analytes are often present in a matrix that contains a significant number of interfering species. If a sample is analyzed directly (without an extraction step), the interfering species may generate a background that makes it difficult to quantify the analyte of interest (Figure 1). Since SPE is a selective process, the interfering species can be either passed through the column during the loading step while the analyte of interest is retained, removed during a rinse step, or retained on the column during the elution of the analyte. Another important function that SPE can serve is the trace enrichment of an analyte. Often the limiting factor in quantifying a compound is that the species is at or below the detection limit. Using solid phase extraction, the analyte can be extracted from a large volume onto a small sorbent mass, and eluted using a small volume of solvent. Concentration factors as great as 500 to 1000 fold can be achieved. In solid phase extraction, both the extraction and elution steps are impacted by interactions between the analytes and the sorbent, the analytes and the

39

, > _

q

I

I

'-.._ [

Figure 1. Removal of background interferences using SPE.

matrix, and finally the matrix and the sorbent. Therefore, the three components to a solid phase extraction that need to be considered are the sample matrix, the analyte and the sorbent. In order to retain the analytes efficiently, it is necessary to optimize the interactions between the analytes and the sorbent while minimizing the interaction between the analytes and the matrix, as well as the sorbent and the matrix. During elution, it is important to optimize interactions between the matrix and sorbent, and matrix and analyte, while minimizing interactions between the sorbent and analyte. It is instructive to compare solid phase extraction to elution chromatography. In liquid chromatography, the volume of solvent required for the elution of a species (Vr), is equal to the volume of the mobile phase plus the product of a constant and the volume of the stationery phase (Vr = V m + KVs). Under the conditions of elution chromatography, it is desirable for the constant, K, to be greater than 0.2 so that separation between species can be obtained, but less than 200, so that the species are eluted within a reasonable amount of time. In extraction chromatography (SPE), ideally the species would behave digitally, stopping at the top of the column upon being loaded (K> 1000), and traveling with the solvent front during the elution (KAg+>Ca+2>Zn+2>K+>NH4+>H+>Li § The following series lists anions on the right that will displace those on the left: O H > acetate> formate> HPO4>HCO~>CI-> HS03> Citrate 2.5.2.

Column solvation After sample pre-treatment, the next step in an SPE procedure is to solvate (condition or wet) the sorbent. It has been demonstrated that the modified surface must be conditioned in order for it to be active and available to the analyte [6]. If the surface has been modified with a hydrophobic species such as octadecylchlorosilane (C18), and then is placed in an aqueous environment, the long hydrophobic chains will collapse upon themselves. The surface can be conditioned with an organic solvent, such as methanol. This will result in solvation of the hydrophobic chains, allowing the chains to become extended, and hence available to interact with the analyte (Figure 11). When sample volume

Unconditioned

Conditioned

Figure 11. Impact of conditioning modified silica surface.

48 exceed 100 mL, it is often pretreated with methanol, which is added as a "wetting agent" to keep the chains extend during the extraction process. Methanol is a good choice for the solvent since the hydrophobic end of the molecule can interact with the hydrocarbon chains, while the polar end can interact with the surface silanols. If no wetting agent is present in the sample, there will be a tendency for large volume aqueous samples to drag the conditioning solvent away from the surface of the sorbent, allowing the chains to collapse. With an adequate concentration of wetting agent present, the organic solvent can remain in a steady state concentration around the immobilized chains. The concentration of the wetting agent needs to be high enough to maintain a conditioned surface, and is typically added at a concentration of 0.5 to 1%.

2.5.3.

Column equilibration

The third step in a solid phase extraction procedure is typically column equilibration. Here, excess organic solvent is removed from the sorbent so that that analyte retention is not hindered. If conditioning solvent is present during sample loading, analytes may remain in a highly organic mobile phase, and pass through the column. In addition to removing excess conditioning solvent, the equilibration step serves to set the conditions of the column, such as pH and ionic strength. These parameters are normally adjusted to the conditions of the pre-treated sample, since the same considerations need to be made with respect to the charge on the analyte and sorbent surface. Equilibrating the surface is necessary to ensure that analytes reaching the sorbent during the early part of sample loading are extracted under the same conditions as those loaded at the end of the extraction.

2.5.4.

Sample loading

In step four of the solid phase extraction procedure, the sample is loaded onto the column. The loading rate necessary may vary significantly depending on the nature of the analytes, and on which type of interactions are being relied to extract them. For compounds having a strong affinity for the sorbent, sorption can take place in a small segment of the sorbent bed, allowing large volumes to be handled at high sampling speeds [7]. The important aspect of sample loading in solid phase extraction is the amount of contact time that an analyte is allowed with the serbent. A sufficient amount of time is required for an interaction to occur. At high flow rates, non-equilibrium conditions may exist, resulting in lower partition coefficients [8]. This will result in analyte breakthrough and consequently reduced recoveries (Figure 12). For analytes that are retained primarily through hydrophobic interactions, significantly higher loading rates are possible as compared to analytes retained through an ion exchange mechanism. Sample loading rates for the retention of hydrophobic species as high as 120 mL per minute have yielded quantitative recoveries on a 6 mL extraction cartridge [9]. It is recommended that a loading rate of five to ten mL per minute not be exceeded for a 1 mL, 100 mg sorbent bed [6]. Five mL per minute represents a linear velocity of 0.42 cm/sec, and a

49

100 % Recovery 50

Increasing Flow Rate Figure 12. Impact of loading rate. residence time of 1.9 seconds. For a 100 mg column, having a bed height of eight millimeters, there is an estimated residence time of 0.19 seconds. For ionic species, the analytes can be surrounded by a solvent sphere, hindering the interaction of analytes with the sorbent surface. Therefore, for analytes that are retained by ion exchange, the linear velocity for sample loading should be to the order of five cm per minute (one mL per minute on a one mL extraction cartridge) [6].

2.5.5. I n t e r f e r e n c e r e m o v a l An interference removal step generally follows sample loading. The extraction cartridge is treated to remove species that could interfere with the analytical determination. This usually involves rinsing the cartridge with a suitable solvent (one that will remove interferences without loss of the analyte). Appropriate retention conditions, such as pH and ionic strength, should be maintained to avoid loss of analyte. The equilibration solvent is commonly used in the interference removal step. If the analytes are well retained, a fraction of organic solvent can be added to the equilibration solvent to remove additional interferences. Drying the sorbent under vacuum or with a gas such as nitrogen or carbon dioxide may be required to remove water. Water may interfere with the elution of analytes from the sorbent if water immiscible solvents are used. Alternatively, interferences may be removed by selecting a solvent that is able to elute analytes from the sorbent, while interferences are retained on the column. Removal of water is critical if the analytical determination is gravimetric. Water that remains on the column can be eluted with the analyte and contribute to the final weight. 2.5.6. Analyte e l u t i o n The final step of the SPE procedure is elution of the analytes from the sorbent. A suitable solvent will preferably have a low viscosity, be readily available in a

50 pure form and have low flammability and toxicity. Important properties are strength and selectivity [9]. The solvent should be one that can to some degree solubilize the analyte. An appropriate solvent must be chosen to enhance the matrix~-~sorbent, and matrix~-~analyte interactions, while minimizing s o r b e n t ~ a n a l y t e interactions. The strength of the elution solvent is related to the mechanism by which the analyte is being retained. Table 1 shows the relationship between solvent strength and type of mechanism by which an analyte is being retained.

Table 1 Mechanism versus solvent strength

Mechanism

Increasing Solvent Strength

Non-polar

Water =v Methanol r Hexane

Polar

Hexane ~ Methanol =~ Water Methanol or water r Methanol and water

Multiple interaction

Since an analyte may be retained through multiple interactions on a heterogeneous surface, a mixed elution solvent is often most effective. An important consideration when choosing an elution solvent is the final analysis. If the analyte concentration is to be determined by HPLC, the mobile phase can often be used to elute the analyte. If the elution solvent must be evaporated to dryness and reconstituted, then a volatile solvent should be selected. Appropriate elution conditions, such as pH and ionic strength, should be considered. It may be necessary to adjust the pH of the elution solvent to neutralize the analyte or surface of the sorbent. Ionic species are typically eluted by adjusting the ionic strength of the elution solvent to 0.1 molar for monovalent analytes and 0.2 molar for divalent analytes. It may be useful to elute the analyte with a buffer containing counter ions that are better retained on the surface than the analyte. Refer to section 2.5.1 for ion selectivities. In addition to selecting an appropriate solvent, there should be sufficient contact time between the sorbent and solvent to ensure quantitative removal of the analyte from the surface. In section 2.1, the porous nature of the sorbent was described. During the loading step, analytes can diffuse deeply into the pores. When possible, the elution should be preformed using two aliquots of solvent, and allowing a 1-2 minute soak step between the two elutions. This allows sufficient time analytes to diffuse back out of the pores. In addition, the inclusion of a soak step allows the analyte to be eluted using minimal solvent volumes, which maximizes trace enrichment.

51

3. IMPACT OF VARIOUS FACTORS ON SOLID P H A S E EXTRACTION In the previous section, each of the steps involved in an SPE procedure was described. There are a variety of factors that can have an impact on the selectivity and efficiency of the extraction. Some of these factors include the sample loading and elution rates, choice of sorbent materials, as well as selection and volume of conditioning, equilibration, rinse and elution solvents. The influence of sample pH and the choice of solvents added to samples prior to loading on to the SPE column can have a significant impact on analyte recovery. Failure to consider these various aspects in solid phase extraction procedures can result in non-robust methods, lengthy development times and excessive costs.

3.1. I m p a c t of v a r i o u s b o n d e d p h a s e c h a i n l e n g t h s on a n a l y t e s e l e c t i v i t y It is known that C18 is the most hydrophobic phase of the bonded silicas, while C2 is the most polar of the hydrophobic phases [6]. It is instructive to look at the impact of a variety of bonded phases on analyte selectivity. By understanding the extent of these interactions, analyte recoveries can be optimized. The impact of bonded phase chain length can be examined by utilizing the EPA Method 525.1 analytes as probes [9]. Samples loaded on ISOLUTE | bonded silica modified with hydrophobic phases C2, phenyl, C8 and C18 were used to generate the data in Table 2. The behavior of these analytes varied with respect to the differences in their structures (size and degree of hydrophobicity), and the type of sorbent selected. For the purposes of analyzing the data, this discussion will characterize the analytes by grouping together those compounds that tended to behave similarly. Those compounds that were eluted early in the gas chromatographic run are grouped together. These compounds were generally small and/or fairly polar. The compounds eluted in the middle of the run were also grouped together, and then again, the late eluting compounds. The late eluting compounds tended to be fairly large and hydrophobic in character.

Table 2 Impact of sorbent type on recovery of EPA 525.1 analytes COMPOUND CI8(EC) CI8(EC) 89 lg 1 Hexachlorocyclopentadiene 62 66 2 Dimethylphthalate 92 100 3 Acenaphthalene 94 92 4 Acenaphthene-dl0 i n t e r n a 1 st 5 2-chlorobiphenyl 87 85 6 Diethylphthalate 100 100 7 Fluorene 89 88 8 2,3-dichlorobiphenyl 79 77 9 Hexachlorobenzene 66 64 10 Simazine 78 69

PH lg 41 8 27 andard 80 58 61 78 64 10

C2 lg 27 1 8 26 10 22 54 53 0

52 Table 2 (continued) I m p a c t of sorbent type on recovery of EPA 525.1 analytes COMPOUND C18(EC) CI8(EC) 89 lg 11 Atrazine 60 33 12 Pentachlorophenol 13 Lindane 99 96 internal 14 P h e n a n t h r e n e - d l 0 15 P h e n a n t h r e n e 83 82 16 Anthracene 76 75 17 2,4,5-trichlorobiphenyl 70 66 100 97 18 Alachlor 76 72 19 Heptachlor 95 93 20 di-n-butylphthalate 21 2,2', 4,4'-tetrachlorobiphenyl 74 71 22 Aldrin 70 63 23 Heptachlor epoxide 88 85 24 2,2', 3',4,6-pentachlorobiphenyl 72 70 25 G a m m a - c h l o r d a n e 78 74 76 74 26 Pyrene 79 74 27 alpha-chlordane 76 72 28 t r a n s nonachlor 29 2,2', 4,4', 5,6'-hexachlorobiphenyl 67 69 62 67 30 Endrin 31 B u t y l b e n z y l p h t h a l a t e 83 78 83 92 32 di(2 -ethylhexyl)adipate 70 71 33 benz[a] a n t h r a c e n e internal 34 Chrysene d-12 76 78 35 Chrysene 77 78 36 2,2'3,3',4,4',6-heptachlorobiphenyl 83 83 37 Methoxychlor 76 79 38 2,2',3,3',4,5',6,6'-octachlorobenzene 78 79 39 di(2-ethylhexyl)phthalate 72 71 40 benzo[b]fluoranthene 68 70 41 benzo [k]fluoranthene 58 60 42 benzo[a]pyrene 83 85 43 perylene-d12 57 58 44 indeno[1,2,3,c,d]pyrene 58 58 45 dibenz [a,h] a n t h r a c e n e 61 61 46 Benzo[g,h,i]perylene Avg: 77 76 early eluting: mid eluting: late eluting:

87 78 65

89 75 66

PH lg 1 24 st andard 80 75 66 94 72 97 73 67 87 74 78 80 79 76 78 56 90 82 75 st andard 82 78 89 76 100 78 79 72 99 71 75 79 69 46 72 79

C2 lg 0 3 38 40 61 16 66 95 70 66 52 73 78 68 78 78 74 32 100 86 74 77 70 100 68 100 80 78 75 100 81 87 82 58 16 61 83

53 For those compounds that were eluted early in the chromatographic run, there was a significant reduction in recoveries as the chain length of the bonded phase decreased. Recoveries for these earlier eluting compounds averaged 89% when extracted onto one gram of C18 material, dropping to 46% when extracted using a phenyl phase, and down to 16% when extracted onto C2. This suggests that the earlier eluting compounds require a sorbent having long hydrophobic chains in order to be extracted from a very polar matrix, since the extraction is based primarily on non-polar interactions. As the hydrocarbon chains on the sorbent get shorter, the silanols on the surface become more accessible causing the surface to become more polar in nature, and the recoveries of the earlier eluting compounds dropped off. For those compounds that were eluted in the middle group of the gas chromatographic run, there was not a significant difference in the recoveries from the phenyl, C8 and C18 phases, ranging from 72 to 78%. There was a drop in recoveries on the C2 phase, averaging 61%. The decline in the recoveries for these somewhat larger and less polar compounds was not as dramatic as that for the earlier eluting compounds, suggesting that the choice of bonded phase for analytes of intermediate polarity is less critical. It can be seen from the data that for the later eluting compounds, the best recoveries were obtained when the extraction was performed using bonded phases with modified with shorter hydrocarbon chains. In contrast to the earlier eluting compounds, as the hydroca,'bon chain length of the bonded phase increased, the recovery of these analytes dropped. For these later eluting compounds, it is energetically less favorable to remain in an aqueous matrix as compared to the earlier eluting compounds. Since this group of compounds tends to be large and quite hydrophobic, they are retained very strongly to a C18 surface during sample loading. This strong interaction between the analytes and sorbent also makes it difficult to remove these compounds from the surface during the elution step. In the case of the EPA Method 525.1 analytes, where a broad range of compounds present, it is possible to select a "compromise phase" such as C8. In this case, the chain length of the bonded phases is sufficiently long to provide a reasonable amount of hydrophobic interaction with the earlier eluting compounds, while having a surface that is sufficiently polar to allow for the elution of the later eluting compounds.

3.2. I m p a c t o f t e m p e r a t u r e on a n a l y t e r e c o v e r y The EPA Method 525.1 compounds were extracted at room temperature (20~ as well as at 4~ An improvement in recoveries can be observed for analytes extracted at a lewer temperature. For analytes extracted using an ISOLUTE | C18 phase, average overall recoveries improved from 75% when extracted at 20~ to 85% when extracted at 4~ The impact of temperature was less dramatic for analytes extracted on ISOLUTE | C8, where the extraction at 20~ gave an average recovery of 85% versus 90% when extracted at 4~ The impact of analyte solubility was considered when examining the phenomenon of improved recoveries at lower loading temperatures. It is likely

54 t h a t the analytes are less soluble at lower t e m p e r a t u r e s , and could therefore be more easily extracted from an aqueous matrix. However, when samples were extracted on cartridges stacked in series at the elevated t e m p e r a t u r e , there was no evidence of analyte breakthrough. Since lower recoveries can not be a t t r i b u t e d to analyte breakthrough, this implies t h a t lower t e m p e r a t u r e s do not improve analyte retention, and the i m p r o v e m e n t in recoveries m u s t therefore be a t t r i b u t e d to an i m p r o v e m e n t in analyte elution. An alternative explanation would be to consider the impact of t e m p e r a t u r e on the diffusion r a t e of analytes. It is known t h a t at elevated t e m p e r a t u r e s , the rate of diffusion of most molecules increases. T h a t the diffusion rate of a molecule is a function of t e m p e r a t u r e and is described by Fick's Law, where it is proportional to the square root of t e m p e r a t u r e . Calculating the difference in diffusion rate at 20~ versus 4~ gives the following: D293=k x (294) 2 D277=k x (277) 2 D293 / D277 = 1.13 It can be therefore estimated t h a t there is an increase in diffusion rate by 13%. For those samples extracted at a higher t e m p e r a t u r e , as the rate of diffusion of the molecules increases, analytes can diffuse more deeply into the pores of the silica. The more deeply the analytes are retained in the pores, the more time t h a t is required to allow the molecules to diffuse back out of the pores (Figure 13). This is consistent with experiments described in the section 3.4, addressing the inclusion of a soak step during analyte elution.

Figure 13. Diffusion into pores of silica.

55

3.3. I m p a c t o f e l u t i o n s o l v e n t The broad range of analyte characteristics of the EPA Method 525.1 analytes m a k e t h e m suitable to illustrate the impact of the elution solvent on analyte recovery. A variety of solvents were tested, both as pure and as mixed solvents as shown in Table 3. Two elutions per column were performed. It can be seen from the data t h a t the average recovery for elution solvents one through four in Table 3 is 69%. In each case, the elution was performed using a pure solvent for both elution steps, although the solvent wasn't necessarily the same for each step. Elutions 5-13 were done with mixed solvents, with the average recovery being 79% and a s t a n d a r d deviation of 5%. From these results it can be seen t h a t the highest recoveries were obtained when mixed solvents were used, as compared to pure solvents. This is consistent with our picture of the mechanisms by which analytes are r e t a i n e d on the surface of the silica. Since compounds may be retained through multiple interactions, analyte elution can be optimized by i n t e r r u p t i n g those interactions with a solvent system t h a t can solvate the analyte using multiple interactions. If a very polar elution solvent is used, analytes can be retained through hydrophobic interactions with the bonded phase. If a nonpolar solvent is used, analytes can be retained through interactions with silanol groups on the silica surface. Solvents five, seven and eight in Table 3 were mixtures of ethyl acetate and acetone in ratios of three to one, one to one, and one to three, respectively. The average recovery for each was 86, 86 and 79 percent, with a relative s t a n d a r d deviation of 5%. Therefore it can be noted t h a t relative concentrations of the mixed solvents is not critical for this solvent combination. For achieving lower detection limits, however, the fraction of w a t e r miscible solvent should be considered with respect to the limit to which the solvent can be concentrated. If it is necessary to concentrate the extract after elution, there m u s t be a sufficient volume of w a t e r miscible solvent present to prevent phase separation if traces of water are eluted along with the analytes. 3.4. I m p a c t o f e l u t i o n s o a k step on a n a l y t e r e c o v e r y Silica gels are an agglomeration of particles resulting in a very porous in structure. The structure of the silica varies with respect to pore size and surface area. The presence and dimensions of pores has a significant impact on the surface area t h a t is available with which an analyte can interact, dramatically improving the efficiency of the extraction. Commercially available silica used in solid phase extraction typically have a nominal pore size of fifty to sixty angstroms. This section addresses the impact t h a t the pores have on analyte recoveries, due to the diffusion of molecules into and out of the pores. A series of experiments were devised to study the impact t h a t including a soak step between elution volumes would have on analyte recoveries [9]. It can be seen from the d a t a in Table 4 t h a t for equal volumes of elution solvent, there are an improvements in recoverieswhen the elution is performed using two separate elution volumes with a two minute soak step between elutions, versus eluting the

56

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Table 4 % Rec. of EPA 525.1 analytes from ISOLUTE | C18 with and without 2 min soak step Two aliquots One Aliquot 3 + 2 mL 2 + 2 mL 1 . 5 + 1 . 5 m L 3 mL 4 mL 5 mL 82 77 80 74 72 74

columns with a single solvent volume. The average recovery for analytes eluted in a single step is shown to be 72%. When the elution was performed in two steps, which included a two minute soak step, the average recovery increased to 78%. The data suggests t h a t for analytes t h a t have diffused more deeply into the pores of the silica, the addition of a soak step allows more time for the analytes to diffuse back out of the pores, and recoveries are improved. This indicates a dependence on time required for the analytes to diffuse out of the pores of the silica. Simply increasing the elution volume did not have a significant impact on recoveries, as seen for the three mL versus five mL elution, where the recoveries were 73 and 72 percent respectively. An additional experiment was performed to determine the impact of elution rate on analyte recovery. Analytes were eluted using flow rates from five to forty mL per minute. Recoveries at five mL per minute average 79% while those at forty mL per minute averaged 77%, S t a n d a r d deviations are four to five percent, and so the difference in recoveries between five and forty mL per minute are statistically insignificant. The most i m p o r t a n t aspect of the elution is the contact time t h a t the elution solvent has with the surface of the silica. For samples eluted with a total volume of five mL, the difference between an elution rate of five versus forty mL per m i n u t e is only a difference in contact time of 60 seconds versus eight seconds. Both of these contact times are less t h a n half of the total contact time allowed when a two minute soak step is included between elution volumes, and so the benefit of reducing the elution rate is insignificant when a soak step is included. It has been shown t h a t the porous n a t u r e of silica used in solid phase extraction plays and i m p o r t a n t role in the extraction as well as in the elution of analytes. It is known t h a t the porosity of the silica provides the analyte with sufficient surface area to interact with the sorbent during analyte retention. During the extraction step, molecules can diffuse deeply into the pores of the silica. As a result, t h a t sufficient contact time m u s t be allowed between the solvent and sorbent during the elution step to provide analyte molecules to diffuse back out of the pores. It has been shown t h a t when samples are loaded at a reduced t e m p e r a t u r e , decreasing the diffusion rate of the analytes, the depth to which analytes p e n e t r a t e into the pores is reduced, and recoveries are improved. It has also been shown t h a t including a soak step between elution volumes can improve recoveries by allowing analytes sufficient time to diffuse back out of the pores.

59 3.5. I m p a c t o f c h a i n l e n g t h o n t h e r e t e n t i o n o f w a t e r In section 3.1, the impact of the chain length on the bonded phase on the retention of analytes was discussed. This section will address the impact that the hydrocarbon chain length has on the water retention characteristics of the sorbent. Earlier in this chapter, it was described that the silanol groups on a C2 surface are well exposed, and there is a significant amount of water associated with this surface. On a C8 surface, the hydrocarbon chains are longer. Access to the surface is somewhat restricted, but the chain lengths are still short relative to the distance between the chains. The silanol groups are still exposed, and the surface is still somewhat polar. On a C18 surface, the long chains, which can reach each other across silanol groups, serve to "water proof' the surface. Although it is still possible for water to reach the surface of the silica, the hydrocarbon chains are mostly associated with organic solvent molecules.

Nitrogen Drying of Bonded Phases 20 PSIG at AutoTrace (4.2 L/min) W 1.6 C 12 (WP)

A

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10

15

20

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Time, min Figure 14. Comparison of drying times for various ISOLUTE| phases.

Various phases were tested with respect the time required to dry the sorbent to constant weight (Figure 14). It has been demonstrated by Fung Kee Fung using differential scanning calorimetry that the silica surface can organize water in the near surface region, imparting an "ice-like" structure [11]. Since silanol groups are more easily accessed by water on sorbents modified with shorter hydrocarbon chains, it is expected that the surface would have a greater impact on structuring the water in the near surface region than those surfaces modified with longer hydrocarbons. It can be seen from the graph in Figure 14 that the bonded phase having the shortest chains required longer drying times, while those with longer hydrophobic chains (C8, C18) required increasingly shorter drying times.

60 Surface Coverage, ~tM2/m Hydrocarbon Volume, mL/t

Void Volume, mL/g: 0.67

2.3

1.6

0.08

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Void Volu~ mL/g: 0.55

Figure 15. Comparison of void volumes for C2 versus C 18.

Figure 15 illustrates the difference in pore volumes for a C2 versus a C18 bonded silica. The pore void volume is decreased for a C18 phase, corresponding to the increase in volume of hydrocarbon present. There is only a 20% difference in volume between the two phases, which is insufficient to account for the two to three fold differences in drying times. For each of the curves in Figure 14, there are two distinct regions. The region having the greatest slope for time versus grams of water removed represent water that is being removed form between sorbent particles, and from the pores of the silica. The less steep region of each of the curves represents water removed form the near surface region. The slope of each line decreases with decreasing hydrocarbon chain length, indicating that it becomes more difficult to remove more highly structured water from the near surface region. The y-intercept gives an indication of the amount of water that is associated with the surface after the bulk (water between particles and in the pores) water is removed. Using these volumes, the depth of the water in the near surface region may be calculated from the known surface areas of the material (approximately 550 m2). The endcapped material dried slightly more quickly than the uncapped material, which would be expected, since there are fewer silanol groups on the endcapped material, reducing the ability to structure water in the near surface region. 3.6. I m p a c t o f p h y s i c a l p a r a m e t e r s on d r y i n g t i m e s In the previous section, the difference in water retention characteristics on various phases was examined. The differences observed were due to differences in surface chemistry. In this section, an experiment is described that examines the difference in the retention of water due to physical differences in the particles, specifically, the impact of the distribution of particle sizes. Sorbent beds

61

versus

versus

Figure 16. Flow characteristics through columns cotaining fines versus fines-free. should ideally contain material that has a narrow particle size distribution (nominally fifty microns), with very few fines present (particle size less than 20 microns). When fines are present in the bulk silica, there is a tendency for the smaller particles to agglomerate. When this material is packed into a cartridge, agglomerations of large and small particles can result in poor flow characteristics (channeling of flow through the bed) as shown in Figure 16. When the sorbent is being dried, channeling results in areas of the bed where there is the least resistance to flow. This results in those areas having contact with the greatest volume of gas, and so drying occurs quickly. In regions of the bed where there is a greater pressure drop (where smaller particles have agglomerated), there is exposure to a smaller volume of gas, and drying times are greatly extended. The drying data for silica from different manufactures is consistent with the data obtained for particle size distribution. The presence of fines could have the same impact on liquid flow as well as on gas flow, resulting in increased volume or soak requirements for conditioning and elution steps, and channeling during loading steps. 3.7. I m p a c t o f w a t e r o n a n a l y t e r e c o v e r y

The importance of removing water from the sorbent bed varies, depending on the type of analysis being performed. If water is present in the elution solvent, additional steps may be required to remove water from the elution solvent, such as passing the extract through sodium sulfate. This can contribute to loss of analytes, as well as being a source of contamination. For samples that require evaporating to dryness, the presence of water can greatly extend the drying time. In addition, the presence of water may greatly limit the ability to use derivatization chemistry. Adequate drying time for the extraction cartridge is essential after an aqueous sample has been loaded, if the sorbent is to subsequently be eluted with a water immiscible solvent. Insufficient drying of the cartridge can result in inadequate contact between the elution solvent and the sorbent (Figure 17), as well as partitioning of the analytes between phases.

62

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The sorbent is often dried with a gas such as nitrogen or carbon dioxide. An a l t e r n a t i v e to e x h a u s t i v e l y drying the sorbent with a gas is to use a watermiscible solvent, such as acetone, as a component of an elution solvent mixture. A solvent such as acetone is able to bridge the properties b e t w e e n two immiscible solvents, such as w a t e r a n d m e t h y l e n e chloride. W h e n such a solvent s y s t e m is utilized, the r e s u l t is to chemically dry the sorbent, removing w a t e r from the pores, a n d a w a y from the surface region of the sorbent. In this m a n n e r , good contact is achieved b e t w e e n the solvent a n d a n a l y t e s (Figure 18). An e x p e r i m e n t was p e r f o r m e d to d e t e r m i n e if there was an i m p a c t on the r e s u l t s w h e n the cartridge drying time was reduced from ten m i n u t e s to 0.5 m i n u t e s prior to an elution with a w a t e r miscible solvent (acetone/ethyl a c e t a t e

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53 75:25) [9]. A half m i n u t e was shown to be sufficient to displace bulk w a t e r from the silica bed. There was no significant difference between the two sets of data, each having an average 82% recovery, with a five percent s t a n d a r d deviation. It h a s been shown t h a t the drying times for bonded phases vary significantly, depending on the chain length of the hydrocarbon phase used for bonding, and the type of silica used. For m a t e r i a l t h a t has been bonded with short hydrocarbon chains (e.g., C2), drying times are significantly longer t h a n those for silica bonded with long hydrocarbon chains (e.g., C18). These results are consistent with the idea t h a t a surface such as C2 is very polar clue to easily accessible silanol groups, and t h a t the impact from this more polar surface is to i m p a r t a higher degree of s t r u c t u r e on the w a t e r in the near surface region. The silanol groups on a C18 surface are buried under the long hydrocarbon chains, which makes this bonded phase the least polar of the modified silicas, having the least ability to organize w a t e r in the n e a r surface region.

3.8. Impact of wetting agent A series of experiments were performed to study the impact on analyte retention of increasing the concentration of wetting agent in an aqueous sample [9]. It was shown t h a t the results obtained for EPA Method 525.1 analytes vary for different types of compounds as the concentration of wetting agent is increased. For smaller and more polar compounds, such as dimethylphthalate, diethylphthalate, simazine and atrazine, the recoveries decreased significantly as the concentration of wetting agent increased, averaging 93% recovery at 0.5% isopropanol, and dropping to an average of 12% recovery at a concentration of 20% isopropanol. This behavior is consistent with w h a t would be predicted, since as the m a t r i x becomes more organic, it competes with the surface for analytes t h a t are strongly dependent on hydrophobic interactions for retention, as the m a t r i x - a n a l y t e interactions are enhanced. For compounds t h a t are eluted in the middle of the chromatographic run, which tend to be of moderate size and/or polar character, there was very little impact on recoveries as the concentration of organic solvent was increased from 0.5 to 20%. The average recovery for this group of compounds over the range of concentrations tested was 79%, with a s t a n d a r d deviation of 3%. In the case of the later eluting compounds, which are the larger and more hydrophobic analytes, recoveries improved as the concentration of organic solvent was increased in the sample matrix, the average recovery for the compounds such as di(2-ethylhexyl)phthalate to benzo[g,h,i]perylene increased from 73% at a concentration of 0.5% IPA to 82% at a concentration of 20% IPA. The fact t h a t the recoveries improved for the more hydrophobic analytes at higher concentrations of solvent could not be a t t r i b u t e d strictly to m a i n t a i n i n g an active silica surface, since 0.5% m e t h a n o l is adequate for this purpose [12]. It would be expected t h a t analytes with more hydrophobic character would be well retained on a C18 surface w h e n extracted from an aqueous, or even partially aqueous matrix. As the concentration of organic solvent is increased in the sample, the analytes themselves are better solvated, and can be more easily eluted from a C18 surface.

64 The experiment of increasing the concentration of organic solvent in the sample was repeated, using polychlorinated biphenyls as the probe set and 200 mg of polystyrene divinylbenzene as the sorbent. Polystyrene divinylbenzene differs from bonded silica, in that the surface does not need to be conditioned with an organic solvent to remain active. Therefore, any effect from the addition of organic solvent to the sample must be due to an impact on the analytes rather than on the sorbent. The average recovery of polychlorinated biphenyls from water at a concentration of 2.0 PPB was 88%. A second set of samples was spiked to a concentration of 10.0 PPB. As the samples were extracted, the effluent water was collected and subsequently re-extracted using methylene chloride in a liquidliquid extraction procedure. For the earlier eluting compounds (2-chlorobiphenyl, 2,3-dichlorobiphenyl and 2,4,5-trichlorobiphenyl), the average breakthrough was eleven percent. For the later eluting compounds (2,2',4,4',5,6-hexachlorobiphenyl, 2 ,2', 3, 3', 4, 4', 6-heptachlorobiphenyl and 2,2',3,3',4,5,6,6'-octachlorobiphenyl), the average breakthrough was fifty two percent. It was demonstrated that as the analytes become larger and more hydrophobic in character, the tendency for breaking through the extraction cartridge increased. The PCB solid phase extraction at 10 PPB was repeated using increasing concentrations of isopropyl alcohol as the wetting agent. The phenomenon that was observed was the same as that seen with the EPA Method 525.1 analytes, where the recoveries of more hydrophobic analytes improved as the concentration of IPA was increased. The improvement in recoveries for these analytes was observed for IPA concentrations as high as 30%, where the average recovery was 97%, dropping to 66% at an IPA concentration of 40%. The experiment was repeated using methanol as the wetting agent. It was shown that recoveries for higher molecular weight compounds do not decrease until the concentration of methanol is increased to from 50% to 70%, where the average recovery dropped to from 97% to 55%. As the concentration of less water-soluble compounds is increased, there is a tendency for the analytes to agglomerate. This behavior becomes more apparent for the later eluting analytes, since these are also the compounds that are less soluble in water. The earliest eluter, 2-chlorobiphenyl, has a solubility of 5.9 PPM [4]. A compound such as 2,2',4,4'-tetrachlorobiphenyl, which elutes at an intermediate time, has a solubility of 0.068 PPM, while octachlorobiphenyls have solubilities below 0.001 PPM [5]. The efficiency of solid phase extraction is dependent on the availability of a large surface area of the sorbent, which is provided by very porous material. When the size of an agglomerated analyte exceeds the size of the pores that are available, the agglomeration can no longer diffuse into the pores. Retention is greatly reduced, since the effective surface area is significantly reduced (from 1100 square meters to approximately 0.1 square meters). When an organic solvent is added, the compounds become better solvated, and the formation of agglomerations is reduced. The molecular compounds are then able to diffuse into the pores, and the effective surface area is increased.

55

4. THE A P P L I C A T I O N OF S P E TO E N V I R O N M E N T A L A N A L Y S E S The use of solid phase extraction (SPE) for environmental analyses is a rapidly growing area in analytical chemistry. The challenge in developing SPE procedures is to selectively concentrate the analytes of interest, maximize their recovery and minimize interferences. For the analysis of a broad range of analytes, conditions must be selected to optimize the recoveries of compounds that are quite varied in properties.

4.1. S e l e c t i v e e l u t i o n o f a n a l y t e s for e n v i r o n m e n t a l a p p l i c a t i o n In the previous sections, the impact of the bonded phase on the retention and elution of a broad range of analytes was examined. In the current section, experiments are described in which advantage is taken of the heterogeneous nature of bonded phases to retain analytes of differing hydrophobicity, and elute them selectively based on differences in functional groups. In this section, different sorbents types are layered for the purpose of extending the capacity to retain polar compounds during an elution with a non-polar solvent. The phenomenon being exploited is the selective elution of particular types of analytes. The analytes used are limited to one compound for which the retention mechanism is purely hydrophobic, and one capable of hydrophobic interaction as well as containing a functional group capable of polar interaction (hydrogen bonding with surface silanols). The analytes are the standards suggested by the EPA for the determination of Oil and Grease in Method 1664, which are hexadecane and stearic acid. Total Oil and Grease is operationally defined by EPA Method 1664 as those compounds that can be extracted from a sample of water using hexane as the extraction solvent (identified as hexane extractable material, or HEM). A subfraction of that material is further defined as silica gel treated hexane extractable material, or SGT-HEM. This fraction is the non-polar material, and is represented in the standard by hexadecane. The polar fraction is represented by stearic acid. The EPA method describes a liquid-liquid extraction, where one liter of sample is shaken vigorously with several portions of hexane, totaling about one hundred mL. Residual water is removed from the extract by passing it over solid sodium sulfate. The solvent is then evaporated, the residue is purged with air and then weighed to determine Total Oil & Grease. The residue is then redissolved in hexane, treated with three grams of silica gel to remove polar components, re-evaporated, purged again, and re-weighed. This residue is designated as the SGT-HEM fraction, also referred to as the Total Petroleum Hydrocarbons, or TPH fraction. If the selectivity of the modified silica surface can be exploited, the solid phase extraction can be used as an alternative to the liquid-liquid extraction procedure. Step one in the six-step SPE procedure involves pre-treating the sample. In the case of Oil and Grease, the sample is acidified to a pH between 1.9 and 2.1 to protonate the acid functions of the fatty acid. Methanol was added as a wetting agent at a concentration of 1%. Although isopropyl alcohol can be used in

66 pesticide work, it can not be used for Oil and Grease, since Oil and Grease is determined using a gravimetric finish. IPA has a higher boiling point t h a n methanol, and adds to the final weight of residue. For the same reason, the cartridge is conditioned with methanol. The equilibration step is performed with reagent water, acidified to the same pH as the sample. The sample is then loaded at rates ranging form 30 to 100 mL per minute. It was found that analytes did not break through the extraction cartridge at this loading rate, which was consistent with earlier work with pesticides [2]. After loading the sample, the cartridge is rinsed with water acidified to pH-2, to remove interferences such as inorganic salts. When extracted on a C18 phase, hexadecane is retained by hydrophobic interactions, and stearic acid is retained by both hydrophobic and polar interactions (Figure 19). In solid phase extraction, advantage can be taken of our

~ S j ~

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Figure ]9. iLnalytes held by multiple versus single retention mechanisms.

ability selectively choose the characteristics of the bonded phase and elution solvents to retain and elute analytes based on differences in functional groups. The analytes can then be selectively eluted from the sorbent by first using hexane to interrupt the hydrophobic interactions to elute the hexadecane. During the hexane elution, the stearic acid continues to be retained in the sorbent through polar interactions between silanol groups on the surface of the silica, and the -OH group on the acid. Then a mixture of hexane and THF (1:1) can be used to elute the stearic acid fraction, where the THF serves to interrupt the polar interactions, and the hexane disrupts the hydrophobic interactions. Each of these fractions can be quantitatively eluted from the sorbent and collected separately. Ninety-six and one hundred thirteen percent recoveries were obtained for the non-polar and polar fractions respectively, from a 20 PPM spike of one liter of sample. In subsequent experiments, the concentration of the spike was increased, but several problems were encountered. Stearic acid began eluting with hexadecane

67 during the hexane extraction. It appeared t h a t the capacity of the sorbent for retaining polar compounds was being exceeded at spikes greater t h a n 40 PPM. Additionally, stearic acid, which is a waxy substance, has a tendency to precipitate from solution as it is spiked into reagent water. W h e n the samples were extracted using an a u t o m a t e d system (Tekmar AutoTrace), the sample lines became plugged coated with insoluble particles of stearic acid. Some samples were extracted m a n u a l l y by attaching a Teflon cap to the sample bottle, to which an extraction cartridge could be fitted. The cartridge was held in a lure tip syringe in a rubber stopper. A -27 inch v a c u u m drew the sample from the sample bottle into a vacuum flask (Figure 20). When the samples were extracted manually, plugging of the extraction cartridge occurred. Recoveries for both fractions, polar and non-polar, were low. At a concentration of forty mg/liter, Total Oil and Grease recovery from the AutoTrace dropped to 79%. A spike of 120 P P M gave a recovery of 64%. A new cartridge was designed for the


"

-4,5 2

4

6

8

10

12

14

16

Ceq, mmol kg-I Figure 14. Adsorption isotherm of SDS and the corresponding electrophoretic mobility curve for the washed soot at 298 K. The common characteristic for all the surfactants studied is a very clear isotherm plateau which is reached around the corresponding cmc values. The amounts adsorbed at the plateau (Fmax) are reasonable and they are similar to the data obtained for hydrophobic or hydrophilic adsorbents [9,10]. A very high Fmax is observed for SDS surfactant (8.4 gmol m-2), although the significant increase in electrophoretic mobility indicates a progressive increase of the negatively charged surface patches. This tremendous amount adsorbed means that in the interfacial region, giant aggregates and may be 3D dimensioned aggregates are present. The isotherms of adsorption of TTAB and DTAB have quite similar shapes to those for hydrophilic silica surfaces [11,12]. The Fmax in both cases is identical. The analysis of the corresponding electrophoretic mobility curve suggests an electrostatic adsorption at the beginning of the isotherm. This recalls that the soot surface is initially negatively charged and becomes positively charged even at the very low adsorption coverage.

191 4

.............................................................................................................................................................................. 3,5 lID

3,5 '~

e e o'~ o

9

O

o o

3

3

2,5 _

2

o 1,5

1,5 1

::[

1

0,5

0,5

0

0 0

3

6

9

12

15

Ceq, mmol kgl Figure 15. Adsorption isotherm of TTAB and the corresponding electrophoretic mobility curve for the washed soot at 298 K.

From an equilibrium molality of about 3 mmol'kg -1, the electrophoretic mobility is then positive and constant. The adsorption mechanism is therefore composed of two simultaneous adsorption processes: firstly the cationic surfactant molecules adsorb on all available negatively charged sites situated on the different oxides, which causes the changing of the sign from negative to weakly positive, while secondly one observes adsorption of surfactants on the hydrophobic part of the surface via dispersion hydrophobic interactions between the alkyl moiety of the surfactant and the surface. Since the length of the alkyl chain of TTAB is larger t h a n that of DTAB, the cmc appears at smaller molality which is a main reason to note that the saturation plateau of adsorption of TTAB is formed earlier. We can underline that Fmax is much higher for TTAB. The cationic surfactants form aggregates at the interface between the soot and the aqueous solution of surfactant. The larger the aliphatic tail of the surfactant, the higher is the aggregation number, and by analogy the interfacial aggregate will be bigger. Consequently the amount adsorbed is higher for TTAB than DTAB (Figure 16). The shape of the adsorption isotherm of the nonionic surfactants on the soot material ressembles the isotherms of the same surfactants adsorbed on the activated carbons [13-16]. They are of the Langmuirian type. The amount adsorbed at the plateau, Fmax, depends strongly on the length of the polar chain

192 1,8

4 -

(a) 9

1,5

"

,t

@ 3 o-

(b)

9

9

0

0

0

1,2

0,9 0

0

0

2 i

0

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0

D

0,6 0,3

p

0

"0

0

lqP

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0o o

""

,g

0

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0

1

0

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I

I

0,2

0,4

0,6

Ceq, mmol kg

0,8

1

0 q

I

I

I

I

0

5

10

15

20

-1

25

Ceq, mmol kg"

Figure 16. Adsorption isotherms of TX100 and TX165 on washed soot and the corresponding electrophoretic mobility curve (panel a) and those for TTAB and DTAB at 298 K (panel b). in such a way that Fmax increases when the polar chain decreases (Figure 16). Conversely, for a polar group, when the length of the alkyl tail increases, ['max increases. This phenomenon appears at a smaller molality. The general trend of nonionic surfactant adsorption is characterised by two steps: for low coverage values the surfactants are adsorbed by their hydrophilic and hydrophobic moieties. In this first stage the molecules are certainly flat on the surface. When the coverage value increases, there is a second step; the hydrophilic parts of the molecules are repelled from the surface and the surface phase becomes thicker. We can note that the negative charge of the surface is not affected by adsorption of nonionic surfactant (Figure 17).

'7, 0

1,8 1,6 1,4 1,2 1 0,8 0,6 0,4 0,2 0

0 -0,05 -0,1 -0,15 0 O0

00

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>

-0,2 -0,25 -0,3 -0,35

0,2

0,4 0,6 Ceq, mmol kgl

-0,4 0,8

Figure 17. Adsorption isotherm of TN111 and the corresponding mobility curve for the washed soot at 298 K.

193 The adsorption of zwitterionic surfactant, NDB, on the soot surface is initially characterised by the strong adsorption at low surface coverage (Figure 18). One observes a linear increase of adsorption until the isotherm plateau is reached in the cmc region. The simultaneous measurements of the electrophoretic mobility show the changing of the sign at low coverage. This change is well correlated with the vertical part of the isotherm and suggests adsorption of surfactant via electrostatic interactions between negatively charged surface sites and the cationic group of NDB. Then, in turn, the mechanism of adsorption alreasy described above starts with the alkyl chain of surfactant on the hydrophobic patches of the surface and simultaneous formation of interfacial aggregates. The amount adsorbed at the plateau, ['max, is relatively high and comparable to the value obtained for this kind of surfactant on hydrophilic surfaces [17]. 0,6 O 9 O

0

0

0

0 9

O 9

0,5

9

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0 E

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o

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' 0

2

,

i

4

6

-0,2 8

10

12

14

Ceq, mmol kg1 Figure 18. Adsorption isotherm of NDB and the corresponding mobility curve for the washed soot at 298 K.

To summarise the results of surfactant adsorption on the soot surface, we can draw the following observations: - the soot is composed of two types of surface: hydrophilic and hydrophobic, - the surface is initially negatively charged, the negative charge of the surface can be altered by the adsorption of cationic and amphoteric surfactants while adsorption of anionic surfactant increases the negative charge of the surface, - the adsorption of nonionic surfactant does not influence the charge of the surface.

194 Furthermore the kinetic investigation of surfactants adsorption on this kind of material shows that the equilibrium of adsorption is reached after about 5 min (Figure 19). This time is relatively long compared with the time necessary for car washing which is usually from 10 to 15 min. 0,9 0,8

-

r ,

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0,70,6]0,5I

0,4 30

60

I

I

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90 120 150 180 210 240 270 300 temps, s

Figure 19. Adsorption kinetic of TN 111 on the washed soot at 298 K.

4.

A D H E S I O N OF S O O T P A R T I C L E S TO P O L Y M E R I C S U R F A C E

At the solid-liquid interface, the strength of adhesion can be expressed quantitatively as the free energy of adhesion. In the case of solid-solid interface, adhesion is much more complicated and depends on elastic and plastic properties of the solids coming into contact, because the adherents deform each other at the region of contact. Particle adhesion is the result of forces which exist between particles and the substrate. Generally, the main cause of adhesion is believed to be the London dispersion force. Excess electric charges generated by frictional electricity can sometimes enhance the rate of soiling by catching dust particles from the air. Soil particles such as soot are not only able to adhere to the substrate by adhesive bonds but they may also be occluded in holes or crevices of the substrate. The main forces involved in the adhesion of soot particles on a carbody surface which is a polymeric surface are capillary, van der Waals, elastic and ionic type.

195

4.1. Capillary force Under humid conditions, a liquid bridge between particle and surface can be formed in two different ways: by spontaneous capillary condensation of vapours and by directly dipping the particle into a wetting film which is present on the substrate, this is the capillary force [18-25]. Due to surface tension, a liquid bridge between the particle and the surface results in a mutual attraction. At thermodynamic equilibrium the meniscus radius represented in Figure 20, is related to the relative vapour pressure by the well-known Kelvin equation: ~+~=~ln ~P r1 r2 2~,M \ Ps J

(5)

where rl and r2, R, T, p, ~,, M and ~P are the meniscus radius, the gas constant, ps the temperature, the density, the surface tension of the liquid, the molecular weight and the relative vapour pressure respectively.

/ ----r2

liquid--

f

Soot particle

rl

t_~

polymeric surface Figure 20. The model of capillary interaction between the soot particle and the polymeric surface which is covered with a liquid adsorbate. The mutual attraction between the particle and the surface results from the Laplace pressure p:

p= ~-~rk where rk is the Kelvin mean radius

(6)

196 rl r2 rk = ~ r 1 + r2

(7)

If the liquid wets perfectly, t h e n the total capillary force exerted on the particle is given by:

F(d) = nx 2(d) ~' rk

(8)

where (x) is the radius of the area. From geometrical considerations one obtains x 2 = 2Rz and for t 100000), insoluble in both acid and alkali and most resistant to microbial decomposition and darkly colored[4,6]. The nonhumic portion is composed of (1) polysaccharides, polymers which have sugarlike structures and a general formula of Cn(H20)m, where n and m are variable; (2) polyuronides, not found in plants but synthesized by soil microorganisms; and (3) low molecular weight organic acids and some protein-like materials, present in relatively small quantities [4]. Moderate amounts of most trace metals are associated with the organic fraction in most mineral soils unless the texture is very sandy, or the organic content is unusually high. In these soils, a high proportion of the trace metals is bound to the organic materials. 4.6. Sulfide p h a s e Significant concentrations of trace metal sulfides are not common in soils, especially those having good drainage and aeration but could be expected in reduced or partially reduced soils that contain significant amounts of S. Thus sulfides would not be expected in normal agricultural soils but may be encountered in rice paddy or wetland soils. Common forms might be pyrite or sulfides of any of the trace metals present in notable concentrations. 4.7. R e s i d u a l p h a s e The most important primary minerals that might be present, especially in young soils, are olivine, pyroxenes, amphiboles, micas, feldspars and the silica group, including quartz. Important secondary minerals are the silicate clay minerals (phyllosilicates): illite, chlorite, vermiculite, smectites, especially montmorillonite, and kaolinite. As mentioned above, while soils age and weather, the proportion of primary minerals decreases and the proportion of secondary minerals and amorphous forms increase. 5.

SELECTIVE SEQUENTIAL EXTRACTIONS

Numerous schemes have been devised to selectively extract different geochemical phases from soils and determine metals present in each phase. Most commonly used are schemes of sequential selective chemical extractions which

332 were first developed for use with sediments in geochemical exploration and contaminated sediment characterization [26-29]. Other methods were developed directly for soil extractions [30-33] but a sediment-developed extraction sequence [29] is commonly cited as the basis for most soil extraction procedures. These methods are ideally designed to affect only the target phase or species with minimal influence on other phases, which is difficult to accomplish. A few studies have shown that although multiple phases are affected by each extractant, the influences are often minimal and not sufficient to preclude beneficial use of this technique [34,35]. The order in which the extraction reagents are used is critical to obtaining the most phase specific data since some of the chemicals will affect more than one phase if added to a previously untreated soil sample [33,36,37]. For example, the strong H202 used to oxidize organic matter will also oxidize MnO2 unless the Mn02 has been removed with a prior extractant. One of the most complete extraction schemes was outlined in [38] as: 1. soluble in water (aqueous phase) 2. exchangeable or unspecifically adsorbed 3. specifically adsorbed 4. bound to carbonates 5. bound to organic matter 6. bound to Mn oxides and hydroxides 7. bound to amorphous Fe (and A1) oxides and hydroxides 8. bound to crystalline Fe (and A1) oxides and hydroxides 9. bound to sulfides 10. bound to silicates (residual fraction). A discussion of the utility and general procedures for determination of each of these phases is helpful in evaluating the applicability of determining a given fraction for particular studies. Following this general discussion, a practical extraction scheme for evaluating contaminated soil is presented. 5.1. Water s o l u b l e p h a s e Although the water soluble fraction is the most bioavailable and subject to leaching, most researchers do not include this aqueous phase in their extraction schemes. The rationale is that concentrations of trace metals in the soil solution are extremely low and generally not of significance when compared to the exchangeable fraction. Additionally, the water soluble phase is co-extracted with the exchangeable phase metals. While this is true for most soils, contaminated soils may have notable water soluble concentrations and this fraction should be determined when investigating soils that might have been subjected to anthropogenic inputs. For example, an unusual characteristic of the soils at a contaminated site may be uncommonly high levels of Pb in the aqueous phase which should trigger a detailed investigation of the abundance and forms of Pb in the soils. Some researchers [39,40] have reported significant concentrations of Cd, Ni, Pb and Zn in the water soluble fraction of trace metal contaminated soils. This simple step involves extraction of the sample with deionized water.

333

5.2. Exchangeable phase Exchangeable cations are measured after displacement with a cation from a neutral salt solution equilibration. Several different types of salts t h a t were used in earlier procedures have largely been discarded due to various difficulties. Reagents should be carefully evaluated depending on specific objectives of the extraction scheme. One of the first extractants was NH4OAc because of its widespread use as a soil test reagent. This extractant is now rarely used to determine exchangeable cations due to the analytical interferences and the possibility that carbonates and hydroxides may be affected [29]. Calcium and Mg chlorides are commonly used to determine exchangeable cations but due to the formation of chloride complexes of Cd and Pb, studies suggest t h a t phases other that simple exchangeable forms are affected [41]. Mg is a harder Lewis acid (a species that can accept an electron pair) than Ca which indicates that Mg may displace specifically adsorbed trace metals [42]. This consideration that may discourage its use as an extractant. Both Mg(NOa)2 and NH4NOa were proposed as extractants because of their efficiency for displacing adsorbed cations as well as volatility which decreases background interferences during AAS analysis [37,43]. Extraction with NH4NOa seems to be the most overall desirable relative to extraction efficiency and low analytical interferences.

5.3. Specifically adsorbed phase Very few studies methods are reported for determination of specifically adsorbed (chemisorbed) metals [33,44-46]. Most of these methods utilize reagents (NH4OAc, NH4NOa, Na2EDTA, HOAc) that were also used for other phases and are not specific for extraction of this phase. A few researchers [33] and [45], who modified the method of [33], utilized Pb(NOa)2 in dilute CaC12 to determine specifically adsorbed metal cations. The Pb §247 is an appropriate cation for displacing most other trace metals due to its low pK (7.7) and large atomic radius which suggest that Pb would be effective in determining specifically sorbed ions. If Pb were of interest in the soil, another cation such as Cu (pK = 7.7, smaller atomic radius than Pb) would have to be utilized for the extraction.

5.4. Carbonate phase The extractants often employed to remove trace metals bound to carbonates are the acids HC1 and HOAc (pH 3-3.5), a buffer solution of HOAc/NaOAc (pH 5) and a buffered complexing agent Na2EDTA (pH 4.6) [29,47,48] The most frequently used is the buffer solution of HOAc/NaOAc (pH 5) which is thought to have minimal effect on other soil phases. Hydroxide phases of alkali, alkali-earth and trace metals are also affected by this extractant and may explain why notable amounts of trace metals are extracted in this phase in slightly acidic soils.

334

5.5. Organic/sulfide phase Three reagents most frequently used for extraction of the organic phase in soils are: acidified 30% H202, K4P207 and NaOC1 [29,49]. The H202 procedure has long been used for extraction or removal of organic matter from soils and sediments and is the most often reported method used for speciation of trace metals in soil organic materials [36,37,40]. F o l l o w i n g the oxidation step with the H202, extraction with NH4OAc is conducted to prevent re-adsorption of the liberated trace metals onto the remaining solid phases. Sulfides are solubilized during the same extraction process as organic matter in oxidative procedures. Methods to separately extract organic forms from sulfides are difficult and their use has rarely been reported with the exception of the use of K4P207. The most simple manner to estimate the proportion of metals associated with each of these two phases is to measure organic carbon and apply a correction. However, the very low solubility metal sulfides are likely to predominate over organic forms in soils where both are present. Another method is to separately determine the acid volatile sulfides and associated trace metals and assign the balance of the metal concentration to the organic form.

5.6. Mn oxide phase The Mn oxides can be either extracted simultaneously with the Fe oxides or as a separate phase. When extracted as a separate phase the reagent of choice for Mn oxides is 0.1M NH2OH-HC1 in 0.01M HNO3 [32,33,40,50].

5.7. Fe oxide phase When Fe-Mn oxides are simultaneously extracted, commonly used reagents are: (1) a heated mixture of 0.04 M NH2OH'HC1 in 25% (v/v) HOAC or (2) 0.175 M (NH4)2C204 + 0.1 M H2C204 (acid ammonium oxalate). The use of Na2S204 in a heated Na-citrate buffer solution for determination of the combined oxides is limited because of analytical interferences. Amorphous Fe oxides are generally extracted with either (1) 0.175 M (NH4)2C204 + 0.1 M H2C204 in dark; (2) 0.04 M NH2OH-HC1 in 25% (v/v) HOAC (heated); or (3) 0.25 M NH2OH-HC1 + 0.25 M HC1. To distinguish crystalline Fe oxides, the reagents of choice are usually (1) acid ammonium oxalate + ascorbic acid (heated); (2) acid ammonium oxalate under UV radiation; or (3) Na2S204 in a Na-citrate buffer solution [29,34,51]. The assortment of reagents, their concentrations and specific procedural conditions used for extraction of the Mn and Fe oxides are rather inconsistent and lead to variable results. Despite these differences in extraction parameters, the methods are all fairly selective as long as each procedure is carefully followed and the proper sequence of extraction is observed: Mn oxides prior to amorphous Fe oxides prior to crystalline Fe oxides.

5.8. Residual phase The residual phase is generally described as silicate structures and resistant minerals that remain after selective removal of the more susceptible geochemical

335 phases. In fact, this phase may contain highly r e s i s t a n t m a t e r i a l s from other phases t h a t were not completely removed. This is especially true for m a n y of the commonly used sequences t h a t do not include the extraction step for crystalline Fe oxides. The Fe content in the residual fraction is usually high in the selective extraction procedure. The types/strengths of mineral acids utilized for extraction/dissolution of the residual phase vary with two p r i m a r y schools of thought. Some researchers deem it necessary to conduct complete dissolution of the residue in order to establish a complete inventory of trace metals in the soil. Other investigators claim t h a t use of a very strong mineral acid will strip away all metals from any mineral residue t h a t may be released to the environment even over short geologic time [29]. The procedures used for these digestions also play a role in the choice of acids. For total dissolution, highly corrosive mixtures of HNO3, HF and usually HC104 are employed, either in open vessels, P a r r bombs or microwave digestion vessels. For strong acid digestion, conc. HNO3, HC1 and HC104 are used singularly or in combination and a few procedures use 30% H202. An abbreviated version of a general procedure for extraction of contaminated soils is as follows. 6.

SELECTIVE EXTRACTION FOR SOILS

1. W a t e r soluble P h a s e (WSP): Add 24 mL of ASTM Type I deionized H20 to 3.0 g sample, shake tubes end-to-end for I hour. 2. Exchangeable P h a s e (EP): Add 24 mL of 1.0 M, pH 7 NH4NO~ to 3.0 g sample, shake tubes end-to-end for 1 hour. 3. Carbonate P h a s e (CP): Add 24 mL of 1.0 M, pH 5.0 NaOAc and place on shaker for 5 hours. 4. Easily Reducible P h a s e (ERP) - predominantly Mn oxides: Add 24 mL of 0.1 M NH2OH-HC1 in 0.01 M HNO3 and place on shaker for 1 hour. 5. Moderately Reducible P h a s e (MRP) - predominantly Fe oxides: add 24 mL of 0.04 M NH2OH-HC1 in 25% HOAc and place samples in a 96 ~ w a t e r b a t h for 6 hours. 6. Organic (Sulfide) P h a s e (OSP): a) add 5 mL 0.02 M HNO3 + 5 mL 30% H202, pH 2.0 and place in 85 ~ w a t e r bath. Over the next 5 hrs, add 3 x 5 mL 30% H202. b) Extract with 7 mL of 4.0 M NaOAc in 20%(v/v) HNO3 for 30 minutes. NOTE: the reaction with peroxide can be very vigorous. 7. Acid Extractable P h a s e (AEP): Digest residue in 150 mL tall form Pyrex beaker with 15 mL conc Trace Metal Grade HNO3 and 5 mL 30% H202 at ~ 100 ~ for four hours on a hotplate. 8. General comments: a) Centrifuge the samples at 7,000 rpm in the HS-4 rotor (rcf = 9600g) in a Sorvall RC2-B refrigerated centrifuge 15 minutes, after each e x t r a c t a n t and after the w a t e r wash.

336 b) Wash the samples with 15 mL of ASTM Type I deionized water after each extractant. c) Remove s u p e r n a t a n t extract with a 5-10 mL Macro-Set or similar pipet.

B

P H A S E D I S T R I B U T I O N O F M E T A L S IN N A T U R A L AND CONTAMINATED SOILS

The origin and forms of trace metals strongly influence their behavior and availability in soils. Lithogenic metals are only slightly mobile and are potentially available to plants only under specific conditions. Pedogenic metal actions reflect particular soil geochemical conditions and anthropogenic metals are generally the most mobile of these three groups [52]. Knowledge of the distribution of trace metals in all of the geochemical phases is often extremely valuable in determining the amount and impact of anthropogenic metals in soils. In uncontaminated soils, a major reservoir for many trace metals is the residual phase while for anthropogenically influenced soils, a much greater percentage of the metals occurs in more mobile forms. Exchangeable metals are the most bioavailable and mobile of the solid phase species, but metals in other geochemical phases are in equilibrium with the exchangeable phase. Multiple and complex chemical forms may be added as contaminants to soils and these chemicals will undergo alterations over time as the soil phases alter towards a new state of equilibrium. Contaminant alteration may produce a supply of metal ions that can be sorbed to exchange sites over time. Measuring only the exchangeable ions at one point in time does not provide a sufficient evaluation of the contaminate status of the soil system. 7.1. M e t a l s in m u l t i p l e f r a c t i o n s Simultaneous examination of trace metals in soil phases is very useful in evaluating contaminant behavior in soils. Very often, patterns of soil enrichment of metals from similar sources or processes, as well as lithogenic or pedogenic metals can be determined. A recent study of the trace metal distribution in soils of four national parks in Poland did not reveal any concentrations high enough to classify the soils as "contaminated". However, enriched concentrations of Cd, Cr, Cu, Ni, Pb and Zn were discovered in soils near industrialized regions. Much of the Cd and, especially, Pb resided in the exchangeable and carbonate fractions indicating recent deposition and relatively high mobility of these two metals. The proportion of exchangeable metal fraction for Pb was greater t h a n for Cd, Ni and Zn indicating greater enrichment of Pb relative to the other three metals. Most of the soil Cu and Ni occurred in the acid extractable (residual) phase suggesting that lithogenic materials were the primary source of these two elements [53]. One new technique that offers unique insights into sorption processes is combined sequential extraction-sorption analysis (CSSA) and the method is well worth utilizing for trace metal sorption studies. Soil or sediment samples are

337 first subjected to standard sorption isotherm methods after which they are sequentially extracted determine metal partitioning into the different geochemical phases. In one study, calcareous clay sediments (pH 7.6-8.0) were individually treated with Cd, Ni or Pb at solution concentrations from 50 to 11000 mg L -1 [54]. Each sample was then sequentially extracted for the following phases: exchangeable, carbonate, Mn oxide, organic and Fe oxide; bulk metals were separately measured. For the untreated samples, Cd was detected only in the oxide phase, an unusual occurrence for this metal, while Ni and Pb were found mainly in the Fe oxide and carbonate phases. After metal sorption, the exchangeable and carbonate phases were dominant for Cd, with the carbonate phase prevailing at sediment concentrations exch>organic~carbonate>Fe-Mn oxide Fe-Mn oxide>residual>>organic>carbonate>exch residual>Fe-Mn oxide>organic>carbonate>exch residual>Fe-Mn oxide>exch~organic>carbonate residual>>carbonate=Fe-Mn oxide>organic>exch carbonate>Fe-Mn oxide>exch>organic>residual exch>Fe-Mn oxide>residual>organic>carbonate residual> Fe-Mn oxide>organic>carbonate>exch ....... Contaminated Soils ....... Cd # exch>residual>Fe-Mn oxide>carbonate>organic Pb # oxide>>carbonate~organic~residual>exch Zn # Fe-Mn oxide~residual>organic>carbonate~exch Cd* residual>carbonate>Fe-Mn oxides>organic>exch Ni* residual>>carbonate>Fe-Mn oxide>organic>exch Zn* carbonate>Fe-Mn oxide>exch>organic>residual Cd residual>>exch>carbonate>Fe-Mn oxide>organic Cu residual>organic> Fe-Mn oxide>carbonate>exch Ni residual>>exch> Fe-Mn oxide>organic>carbonate Zn residual>Fe-Mn oxide>carbonate>organic>exch Cd* carbonate>exch>Fe-Mn oxides>residual>organic Cu* Fe-Mn oxide>carbonate>residual>organic>exch Ni* residual>Fe-Mn oxide>carbonate>organic>exch Zn* Fe-Mn oxide>carbonate>residual>organic>exch Cd # exch>fe-Mn oxide>residual>carbonate>organic Cu # Fe-Mn oxide>organic>carbonate~residual>exch Ni # residual>>Fe-Mn oxide>organic>carbonate>exch Zn # Fe-Mn oxide>residual>carbonate>exch~organic #smelter contaminated soils; *sludge-treated soils Cd Pb Zn Cd Ni Zn Cd Zn

Refs [57] [57] [57] [59] [59] [59] [58] [58] [57] [57] [57] [59] [59] [59] [61] [61] [61] [61] [60] [60] [60] [60] [60] [60] [60] [60]

Very small proportions of Cu, usually < 1%, are found as exchangeable in both native and contaminated soils. Exchangeable Cu ranged from 0.2 - 1% in nine different contaminated soils [60]. About 2% of the total Cu present in contaminated soils near Sudbury, Ontario was measured as exchangeable [61], and the same proportion was noted in the surface horizon of two forested soils in a virtually unpolluted area of Switzerland [36]. Less than 5% of the total Cu was reported to be in the exchangeable form in a study of biosolids treated soils [62]. Relatively small percentages of exchangeable Ni are found in soils, but the proportions are often greater than those for Cu. Exchangeable Ni ranged from

339 8-11% in nine soils that received contaminants from five different sources [60]. Exchangeable Ni was reported to have varied only from 5-8% in both an untreated and sludge treated soils [58]. Even though the concentration of exchangeable Ni doubled in the sludge treated soil, the proportion on the exchange sites remained constant. In one study, after long-term application of biosolids, it was reported that about 10% of the total Ni in soil existed as exchangeable [62]. In uncontaminated and many contaminated soils the fraction of exchangeable Pb is very low, about 5-6% [52,58,62], but in some contaminated soils the exchangeable fraction may be substantial, as high as 25% [39,40]. The proportion of exchangeable Zn varies notably in both native and contaminated soils. These percentages are usually greater t h a n for Cu, Ni and Pb but less than for Cd. No clear pattern of controlling factors is evident from the studies reported in the literature. In an investigation of nine soils that received contaminants from five different sources, a range of 1-10% exchangeable Zn was reported [60]; A range of 3-12% exchangeable Zn was found in both native and contaminated soils in southwest Poland [58]; while another investigation noted that exchangeable Zn was < 15% of the total for a variety of biosolids treated soils [62]. However, results of one study noted that exchangeable Zn increased from 3% to 26% after the soil was treated with sewage sludge [58]. One review paper [63] reported that the order of prevalence of metals in the Easily Soluble/Exchangeable fraction of natural soils was: Cd > Zn > Ni > Pb > Cu. 7.3. M e t a l s in the c a r b o n a t e f r a c t i o n Carbonates, hydroxide and oxide forms of trace metals are extracted by the Carbonate Phase reagent. This phase has been found to contain notable proportions of Cd, Ni, Pb and Zn in both uncontaminated and contaminated soils, even at moderate soil pH values [40,58]. Many soil contaminants originate from various incinerator processes (e.g., fly ash and flue dust) and the metals occur in the forms of oxides which obviously enriches this soil phase. Depending upon the counter ions available, some of these metals may accumulate as sorbed ions as the original forms are altered. Many studies, however, have found rather low concentrations of several of the trace metals in the carbonate fraction. This is the most pH sensitive soil phase and has a notable presence at near neutral or higher pH values and a minor abundance in acidic soils. One study reported 1-3% Cd and 1-10% Zn occurred in this phase in cultivated soils, averaging near pH 6, in Norway [57]. It was noted that < 2% of Cu, Ni and Zn was found in the carbonate phase in a study of contaminated Sudbury soils but these were very acidic soils [61]. The following results for metals in the carbonate phase was reported from a study of nine different contaminated soils, with a pH range of 5.8 to 7.5: Cd: -10%, Cu: 3-11%, Ni: 0.5-2% and Zn: -10% [60]. Biosolids are often lime-stabilized and therefore soils which have received applications of biosolids may have notable proportions of trace metals in the

340 carbonate phase. The following increases of soil metals in this fraction were reported after application of sewage sludge: Cd, from 7 to 25%; Ni, from 14 to 16% and Zn, from 16 to 41% [58].

7.4. M e t a l s in t h e Fe-Mn o x i d e f r a c t i o n Although there is a difference in the selectivity of trace metals by Mn and Fe oxides (Table 2), most researchers use a reagent t h a t simultaneously extracts both oxides. The low abundance of Mn oxides in m a n y soils, especially those of sandy textures, m a y be a reason t h a t this phase is not extracted s e p a r a t e l y from the Fe oxides. The combined oxide phase is generally an i m p o r t a n t reservoir for most of the trace metal cations. Moderate a m o u n t s of Cd and Cu are associated with the oxide phase, in both u n c o n t a m i n a t e d and contaminated soils. Moderate to high proportions of Ni, Pb and Zn are common in the oxide fraction in native and c o n t a m i n a t e d soils. The Fe-Mn oxide fraction in several native soils contained from 29-48% Cd, ~ 4 8 % Pb and 19-39% Zn [56,57]. Distribution of most metals in soils c o n t a m i n a t e d by metal smelters or from biosolids applications was similar for: C d - 20-25%; Cu - 28-43%; and Zn= 34-44%. The proportions of Cd and Cu were approximately 10% less in the smelter affected soils but there was no difference in the proportion of Zn between the two groups of soils. However, the proportion of Ni in the oxide phase was much higher (34%) in the biosolids t r e a t e d soils t h a n those c o n t a m i n a t e d by a smelter (13%). Differences in the chemical forms of Ni a p p a r e n t l y influenced its distribution in the two soil groups, but there was no influence on the distribution of Zn. The identical distributions for Zn m a y have resulted from similar chemical forms in the two sources, which is less likely, or as a result of the strong affinity of Zn for the oxides. 7.5. M e t a l s in t h e o r g a n i c f r a c t i o n The association of metals with soil organic m a t t e r is generally low for Cd and Ni, low to moderate for Pb and Zn and moderate to high for Cu. This relationship would, of course, be different for organic soils or sludges where the majority of the solid m a t e r i a l is organic. In such cases, a high proportion of all metals would occur in the organic phase. No major differences were noted in the distribution of Cd, Ni, Pb and Zn in the organic fraction between a variety of background and c o n t a m i n a t e d soils [56-58,60]. There was little difference in the distribution of Cd, Ni and Zn in the organic fraction of soils affected by two different c o n t a m i n a n t sources but a notable difference in the distribution of Cu between the soils. The following proportions of metal in the organic phase were reported: Cd < 6%; Ni < 9%; Pb and Zn 7-15%. Proportions of Cu were generally 17-33%. There is an effect of added organic m a t t e r on trace metal adsorption and mobility t h a t is not evident from the phase distribution data. Research has indicated t h a t increased levels of dissolved humic materials in leachates from soil or sand columns increased the solubility of Cd, Cu and Zn [64-66]. Simultaneous field and laboratory studies were conducted to investigate the effects on Cd

341 mobility induced by a single liquid sewage sludge application onto a soil [67]. Low-level movement of Cd and soluble organic C from the sludge application site was observed in comparison to the control plot during several weeks following the sludge application. The conclusion was that the mobility of Cd was enhanced by the increased soluble organic m a t t e r from sewage sludge disposal, especially during the period immediately following liquid sludge application. The results of these studies emphasizes the fact that understanding all on-going soil processes is very important in evaluating the exchange, mobility and availability of soil trace metals. 7.6. M e t a l s in t h e r e s i d u a l f r a c t i o n The major concentrations of many metals in native or uncontaminated soils are found in the residual fraction. The proportion in this fraction often decreases significantly when contaminants, usually containing more soluble forms, are applied to soils. High, but variable proportions of trace metals have been reported for the residual fraction for a variety of native soil types: Cd - 23-59%; Ni ~ 80%; Pb ~ 30%; and Zn - 29-70% [56-58]. The proportions of trace metals in the residual phase was usually less in contaminated soils in the same studies" Cd = 16-34%; C u - 11-20%; N i - 50-80%; Pb ~ 10% and Z n - 6-30%. [56-59]. In a study of the phase distribution of metals before and after the application of sewage sludge, the proportion of Cd and Ni in the residual phase changed only slightly but the change in Zn distribution was notable" the proportion of residual Zn decreased from 36 to 6% after sludge application [58]. Similarities exist in the proportion of metals in the residual fractions of native as compared to contaminated soils, but the variation between contaminated soils may be especially notable. A s u m m a r y of metal abundance in the geochemical phases from the cited studies is given in Table 3.

8.

E V A L U A T I O N O F M E T A L B I O A V A I L A B I L I T Y IN S O I L S U T I L I Z I N G SELECTIVE EXTRACTIONS

Plant growth involves interaction of the plant system and soil. Soil is the normal medium for plant growth and the plant's roots absorb nutrients, other elements and water from the soil. The plant's absorption of elements involves processes occurring in both the plant's root and in the soil. Characteristics or mechanisms of either of these two media may notably influence processes of the other. Active ion uptake by the plant roots allows the a t t a i n m e n t of plant metal levels that are in excess of soil concentrations in the mobile metal phases. The depletion of soil concentrations of some elements around the growing roots may induce a redistribution in the soil solution-solid phase partitioning of metals. The mass flow of soil water into the plant may increase the overall concentration of soluble cations, or salt content, at the soil-root interface. This increased concentration will, in turn, affect the distribution of exchangeable cations and

342 could increase the availability of trace metal cations which occupy a low proportion of the exchange sites [68]. There has been considerable discussion over the past several years t h a t metals in the exchangeable phase and a few of the other so-called mobile geochemical phases are the most bioavailable. While this is proving to be correct, m a n y studies have been limited to chemical extractions and have not included plant uptake [47,60,69]. The exceptions have been studies on the use of soil test extractants, many of which measured exchangeable cations [70]. However, this type of extractant was used primarily for determination of the major cations (Ca § Mg § and K § and rarely trace metals. There are several recent investigations which have determined metals in the various soil geochemical phases and compared the values with metal concentrations in the plant. As previously discussed, there are many different selective extraction schemes which make data comparison often difficult. However, the one extraction t h a t is almost universal to all schemes is the one for the exchangeable (adsorbed) cations. As noted in Section 7.1, the proportion of several trace metals in the exchangeable phase may be very small which may explain why plant uptake is often related to phases other t h a n exchangeable. Some researchers have attempted to relate metal uptake to soil solution concentration because they believe that the greatest proportion of metals are absorbed from the soil solution. In an experiment to test this hypothesis, swiss chard (Beta Vulgaris) was grown in soil which had been treated with sewage sludge and various additions of P and N fertilizer [71]. The Cd and Zn levels in the plants were not related to the concentrations in soil saturation extracts (soil solution), P or N application rates. Although metals in the soil phases were not measured, the researchers postulated that the availability and plant uptake of Cd and Zn were reflective of metal desorption from the solid phases. It is reasonable to assume that a high proportion of these metals were desorbed from the solid phase directly into the plant root system. The rhizosphere consists of the soil zone within 1-2 mm of the root surface. There is a much higher soil acidity, bacterial activity and organic content in the rhizosphere than in the bulk soil and there is little question of the significant effect of the rhizosphere on the solid phases in the soil [4]. Low-molecular-weight organic acids (LMWOA) secreted by plant roots were found to modify the mobility of Cd through formation of soluble complexes in the rhizosphere of uncontaminated soils [72]. In a study where organic acids identified in plant root exudates, viz., 0.02 M acetic, citric, fumaric, oxalic or succinic acid were equilibrated with soil samples, the released Cd increased over that from the control soil. The results showed that: (1) Cd was brought into soil solution from the soils as Cd-LMWOA complexes by the LMWOAs secreted by the plant roots; (2) the kinetics of Cd release by LMWOAs was diffusion controlled, and (3) the dynamic release of LMWOAs from the plant roots into the soil rhizosphere would continuously release Cd from the soils, as indicated by the renewal of the LMWOAs. The average diffusion coefficients of Cd release from the soils by

343 LMWOAs and the Cd release by renewal of the LMWOAs followed the same trend as the Cd availability index of the soils. In an earlier study, these workers suggested that NH4Cl-extractable Cd should be used as a Cd bioavailability index based on its highly significant correlation with Cd concentration in the grain of durum wheat [73]. The NHnC1 extraction would measure the exchangeable Cd which is the most likely phase affected by the LMWOAs. For contaminated soils in Sudbury, Ontario, a strong linear relationship was found for Cu and Ni concentrations in birch twigs (dwarf birch, Betula pumila L. var. glandulifera Regel and white birch, Betula paprifera Marsh.) and soil exchangeable metal [61]. There was also a highly significant relationship between Ni concentrations in the twigs and total soil Ni. Levels of Zn in the plants and soil were within normal ranges and there was no significant association between soil and plant concentrations. The Cd concentration in sudax [Sorghum bicolor (L.) Moench] correlated strongly (r 2 = 0.91) with exchangeable Cd where sewage sludges containing high Cd levels were land applied [74]. It was also found that a good correlation existed between Cd plant levels and a Ca(NOa)2/EDTA extractant which was postulated to remove both exchangeable and chelation-bound Cd in the sludge solids. The best correlations between plant content and soil extractions were with soils where the concentration of metal in the applied sludge was relatively high. Associations between Cd in the sudax and soil phases for low-Cd sludges were much poorer and the overall correlation coefficient (r 2) for exchangeable Cd and the sudax concentration was only 0.56. This lack of a strong relationship between plant metal content and soil phases at low soil concentrations of the metal seems to follow a proposed "Threshold Level" theory [75]. The uptake of Cu, Ni and Zn by wheat (Triricum aestivum L.) and soybean (Glycine max L.) grown on soil treated with co-composted sewage sludge under greenhouse conditions was investigated [76]. The selective extraction scheme used by these researchers was somewhat different t h a n the procedure given above. After extracting the exchangeable and organic fractions, EDTA was utilized to remove metals from inorganic precipitates. The results determined that exchangeable Zn correlated with plant uptake but t h a t there was no such correlation with either Cu or Ni. The best uptake models were obtained from a stepwise multiple regression procedure where pH, exchangeable and inorganic precipitate-bound metals were the most important parameters. The authors pointed out t h a t the small increases in tissue concentrations of metals as a function of sludge application rate may have resulted in the poor relationship between metals in plant tissue and in soil phases. However, a relationship could exist even without a notable increase in metal content as a function of t r e a t m e n t rate. Other cited studies [39,61] have noted correlations between soil phase concentrations and plant uptake where there were no t r e a t m e n t effects. Another explanation for the lack of correlation between metals in plant tissue and soil phases is t h a t the overall soil and plant concentrations were below a level where this relationship is clearly expressed.

344 Romaine lettuce (Latuca sativa L.) was grown on soils t h a t had received varying rates and frequencies of biosolids applications over a 20-year period [62]. Applied biosolids where Cd concentrations were high and the Cd was in forms that were easily extracted from soil were readily available for uptake by the lettuce more t h a n 15 years following application. Concentrations of Cd, Cu, Ni and Zn in the lettuce leaves were positively correlated to the total concentrations of respective metals in the soil by either a linear or quadratic regression model. Using plant uptake slopes from the regression analysis equations, the authors suggest that the relative bioavailability of the biosolids-applied metals followed the trend: Cd>>Zn>Ni_>Cu>>Pb. For Cd, the best correlation (r 2 = 0.99) was obtained by including in a regression model the exchangeable, carbonate and Feoxide fractions. The correlation for Cd uptake and the exchangeable fraction alone was r 2 = 0.73. Plant Cu, Ni and Zn concentrations were correlated primarily with the exchangeable soil fraction. Plant Pb levels were generally not correlated to any of the soil geochemical fractions.

0

U S E O F S E L E C T I V E E X T R A C T I O N S TO E S T I M A T E M E T A L L E A C H A B I L I T Y F R O M THE SOIL TO THE G R O U N D WATER.

Few studies have directly investigated the relationship between metal concentrations in geochemical phases and amounts that can potentially leach into the ground water. The exception is the incorporation of exchangeable phase metal concentrations in some transport models. Some of the transport models for u n s a t u r a t e d soils utilize the cation exchange capacity, Freundlich equation exponent, distribution coefficient, Kd, and metal solubility as inputs [77]. The data are often generated from adsorption isotherms or leaching column studies, which are mono-element systems as discussed above. The importance of ion sorption reactions has long been recognized and ion exchange data is included in several models. Many researchers consider that sorption reactions are best described by a cation exchange model. The most rapid reactions are attributed to non-specific ion exchange while slower retention reactions are related to specific sorption of metal ions onto soil solid surfaces [78]. These reactions may also be interpreted in terms of formation of outer- and inner-sphere complexes with soil surfaces [2,3]. Two researchers [79,80] incorporated ion exchange reactions into the tworegion (mobile-immobile) concept. Their approach was generally successful in predicting the overall shape of breakthrough curves for Ca ++ and Mg §247which were obtained from miscible displacement columns that held different sized soil aggregates. Other transport models consider several mechanisms including ion exchange, complexation, dissolution-precipitation and competitive adsorption. Examples of such models include FIESTA [81], CHEMTRAN [82], and TRANQL [83]. Because of their complexity, several of these models have not been fully validated [78]. Results for Cd and Ni breakthrough on a sandy soil utilizing the FIESTA model have been described [84]. The model predictions provided higher

345 retardation of Cd and lower retardation of Ni t h a n empirically observed. Improved predictions were obtained when a kinetic approach was used with data from batch experiments. Model development and validation for movement of trace metal cations in soils has yet a significant period of evolution ahead. A thermodynamic model was developed for Cd, Cu and Zn concentrations/solubilities in soil solutions of sandy textured soils that had received applications of cattle-manure slurry [85]. With the assumption that organic m a t t e r was the dominant exchanger phase for the metals, the model accounted for metal complexation with dissolved organic carbon ligands and with the solid exchanger phase. The authors cautioned that their model is site specific and has yet to be verified however it provides a basis for further research to test assumptions and improve the model. A study of Cd, Pb and Zn in soils that had received surface applications of a metal-rich flue dust revealed that high proportions of Zn were present in the exchangeable phase of one of the soils to a depth of 105 cm [86]. The flue dust was applied annually for six years to raise soil pH and the metal measurements were conducted four years after the last application. Approximately 20% of the total Zn was found in the exchangeable phase in the surface samples, but the exchangeable phase was the dominant fraction at depths greater than 30 cm. Earlier research had reported that Zn in the exchangeable phase in subsurface soils was not the primary fraction in most southeastern soils [43]. The high proportion of exchangeable Zn in the subsurface was caused by the downward migration of the Zn in the soil profile. There were notable amounts of Zn in the organic, amorphous Fe oxide, crystalline Fe oxide and residual phases in the 0-30 cm portion of the profile but only minimal amounts in these phases at greater depths. These results note the stability of metals in the other phases as compared to exchangeable Zn. There also was a slight downward movement of Cd noted in the study, but much less than for Zn. Only about 11% of the Cd was present in the exchangeable form in the surface soil and much smaller proportions at greater depths. 10. S U M M A R Y AND C O N C L U S I O N S The importance of ion exchange reactions to nutrient dynamics in soils was first recognized in the mid-nineteenth century. The ability of soil colloids to reversibly adsorb cations from the soil solution was understood as an ion exchange reaction long before the origin of the negative charge on clay minerals and organic m a t t e r was known. Both non-specific adsorption, rapid ion exchange, and specific adsorption, chemisorption, are important reactions that influence trace metal cation behavior in the soil system. Exchangeable metals are usually the most reactive solid phase metals - most available for plant uptake and most easily leached downward through the soil profile. Other geochemical soil phases are important influencing factors of trace metal behavior in soils. Trace metal cations interact differently with the various soil phases based on

346 both solution and solid phase characteristics. The proportions of metals associated with the individual phases can often provide information pertinent to the contaminant status of the soil. Often, the absolute and relative concentration in each phase, as compared with the total metal concentration, can determine the presence and extent of anthropogenic input as well as the degree of environmental concern over the metals present. With the exception of Cd, the proportion of the trace metal cations in the exchangeable phase is small compared to the total concentration. When larger proportions of exchangeable metals are found, it is a prompt to conduct a detailed investigation of the trace metal geochemistry of the site under study. Exchangeable trace metal cations have been correlated to plant uptake and, in a few instances, related to transport of metals in the soil water. Since all of the geochemical phases are connected through the soil solution in a dynamic state of equilibrium, evaluation of all components is important to fully characterize trace metal cation behavior in soils.

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349 68. S.A. Barber, Soil Nutrient Bioavailability: A Mechanistic Approach, Second Ed., John Wiley & Sons, Inc., New York, 1995. 69.M.P. Levesque and S.P. Mathur, Soil Sci., 142 (1986) 153. 70. B. Zhu and A.K. Alva, Soil Sci., 156 (1993) 251. 71. J.R. Villarroel, A.C. Chang and C. Amrhein, Soil Sci., 155 (1993) 197. 72. G.S.R. Krishnamurti, G. Cieslinski, P.M. Huang and K.C.J. Van Rees, J. Environ. Qual., 26 (1997) 271. 73. G.S.R. Krishnamurti, P.M. Huang, K.C.J. Van Rees, L.M. Kozak and H.P.W. Rostad, Commun. Soil Sci. Plant Anal., 26 (1995) 2857. 74.J. Jing and T.J. Logan, J. Environ. Qual., 21 (1992) 73. 75. J.H. Rule, Use of Small Plants as Phytomonitors with Emphasis on the Common Dandelion, Taraxacum Officinale, in: D.C. Adriano, Z. Chen and S. Yang (eds.), Biogeochemistry of Trace Elements, Environ. Geochem. and Health, Special Issue, 16, 1994, 627. 76. J.T. Sims and J.S. Kline, J. Environ. Qual., 20 (1991) 387. 77. R.N. Yong, A.M.O. Mohamed and B.P. Warkentin, Principles of Contaminant Transport in Soils, Elsevier Science Pub. B.V., Amsterdam, The Netherlands, 1992. 78. H.M. Selim and M.C. Amacher, Reactivity and Transport of Heavy Metals in Soils, Lewis Publishers, Boca Raton, FL, 1997. 79. H.M. Selim, R. Schulin and H. Fliihler, Soil Sci. Soc. Am. J., 51 (1987) 876. 80. R.S. Mansell, S.A. Bloom, H.M. Selim and R.D. Rhue, Soil Sci. Soc. Am. J., 52 (1988) 1533. 81. A.A. Jennings, D.J. Kirkner and T.L. Theis, Water Resour. Res., 18 (1982) 1089. 82. C.W. Miller and L.V. Benson, Water Resour. Res., 19 (1983) 381. 83. G.A. Cederberg, R.L. Street and O.J. Leckie, Water Resour. Res., 21 (1985) 1095. 84. D.J. Kirkner, A.A. Jennings and T.L. Theis, J. Hydrol., 76 (1985) 107. 85. D. Hesterberg, J. Bril and P. del Castilho, J. Environ. Qual., 22 (1993) 681. 86. Z. Li and L.M. Shuman, Soil Sci., 161 (1996) 656.

Adsorption and its Applications in Industryand EnvironmentalProtection Studies in Surface Science and Catalysis,Vol. 120 A. Dabrowski(Editor) 9 1998Elsevier Science B.V. All rights reserved.

351

Application of environmental colloid science in the soil systems J. Szczypa a, I. Kobal b, and W. Janusz a aDepartment of Radiochemistry and Colloid Chemistry, UMCS, 20-031 Lublin, Poland b Josef Stefan Institute, Jamowa 39, 1000 Ljubljana, Slovenia and School of Environmental Sciences, Vipavska 13, 5000 Nova Gorica 1.

INTRODUCTION

Soil is, beside air and water, the most important environment for the living organisms including h u m a n beings. Between these systems, some components exchange, influencing the development and condition of organisms. Up till now most of the h u m a n food comes directly or indirectly from soil cultivation. That is the reason for the special care of soil system quality. The main way is to fully understand complex processes that rule the ecosystem. When air and water systems are simple, the soil is very complicated and variable one, concerning the mineral, grain-size distribution and composition of surrounding solution. Soil is multicomponent, polydispersed system, usually treated as a three-phase system (solid, liquid and gas), although some authors consider organisms the fourth phase. Tens thousands of soils can be distinguished. They can differ by their provenance, parent minerals and age. For example later soils, less weathered, are rich in silica, alumina and iron minerals whereas old weathered ones are without many soluble minerals of alumina and iron and consists of silts and rusty oxides [1]. Looking at composition of soil there can be distinguished primary and secondary minerals. Quartz, orthoclase, plagioclase, muscovite, biotite, pyroxenes and olivine are in the first group and hydrolyzed silicon oxides, aluminum and iron hydroxide, carbonates (calcite and dolomite), hydroxides and silt minerals are in the second one. Half of the most soils is formed by minerals, the rest consists of water solution, air and organic substances (about 5%). In sandy silts the organic substance contents is small and reaches few percent whereas in peat soils or muck the organic substance contents may reach 100% [1]. Among many organic substances in soil very important role play humines, humine acids, humatomelitic acids and fulvic acids. These compounds dissolve in water in a different degree. The solid particles of soil show various size from bigger t h a n l m m treated as gravel and stones, then sand 0.05-lmm, silt 0.002-0.05mm to clays < 0.002mm. The sandy soil, formed by fine particles is concise, plastic, sticky and impermeable. This

352 type of soil is defined as heavy soil. The soil with sand is porous and defined as light soil. In this soil the transportation processes of air and water with nutrient substances are easier. The presence of humus provides proper structure of soil, decides degree and mechanism of soil particle aggregation. Adsorbing on mineral particles, h u m u s provides spongy structure of soil that promotes a transportation process. Soil is a source of heavy metals and their adsorbent also. These factors, which have an influence on the total contents of metals and their amount assimilated by organisms is essential for h u m a n beings life and soil fertility. Some of heavy metals are known as microelements (Cu, Zn, Co, Mn) that are indispensable for growth and their life at specific concentrations and toxic at higher, and other metals that are potentially toxic elements (PTE) (As, Hg, Pb, T1 and U) [2]. Radioactive isotopes, existing in the environment may be divided according to their origin: primary, cosmic and anthropogenic ones. Primary radioisotopes, which came into being during nuclear synthesis of elements have a half-life time comparable or bigger than the age of our planet. To this group belong K-40 and isotopes of U-235, U-238 and Th-232 series. The isotopes of cosmic origin form in atmosphere of the Earth as a result of cosmic radiation for example H-3 and C-14. Man has inconsiderable influence on the formation and spreading of the isotopes from these groups. Unfortunately there is a group of isotopes that were introduced to environment by man. Some of them, by the occasion of nuclear energy production (from mining of radioactive ores to closing down the exploited reactors), other as a result of nuclear weapon tests. The main radioisotopes introduced to environment in this way include the uranium family nucleus' fission products (U, Pu) and products of activation of reactor materials. The list of radioactive isotopes, important for the environment contains elements of the most groups of the periodic table. The list of elements and their origin was presented by Lieser [3]. A physico-chemical behavior of the most radioactive isotopes in soil is the same as their stable equivalents, only for H-3 some differences may be distinguished. Radioisotope concentration may be similar or considerably smaller than stable isotope in environment, depending on its half-life time. The concentration of the hydrolyzable element is connected with a form it exists in the environment. When the concentration is smaller than the solubility product of the respective hydroxide or other insoluble salt and there are no conditions to adsorption on the dispersed particles of other phase, the isotope may form non-ideal solution. When the isotope concentration is so high that exceeds the solubility product then the dispersed phase may form. When the dispersed phase exists and physical conditions enable the adsorption on the solid, then isotope forms pseudocolloid, that means exists in the dispersed phase not forming crystalline structure. From the chemical and physical behavior at the solid/solution interface there is no difference between stable and a radioactive isotope. The only difference is their harmfulness to living organisms. To fully understand the transport phenomena of nutrient substances in such complicated system, the adsorption measurements of the selected element on the

353 soil are not sufficient method. It is necessary to learn surface properties of individual components of soil and on this basis to work out the model of incorporated processes. There are some papers that treat processes running in soil as a physical chemistry of dispersed systems that is ion adsorption and colloid transportation in water [2-7]. In present paper the essential processes, having influence on the ion distribution between solid and solution are presented. They may be related to soil minerals/ aqueous solution system and thus may be useful for the understanding of the transportation and adsorption phenomena in such complex system as the soil is. In the subsequent chapters of this paper the problems of the ion adsorption and surface charge formation are presented.

2.

METAL OXIDE/AQUEOUS SOLUTION SYSTEM

Among metal oxides found in soil, beside silica, there are oxides, hydroxides and hydrated oxides of iron, aluminum and mixes hydroxides. The electrolyte ions may accumulate on the metal oxide surface as a result of nonspecific adsorption, caused by electrostatic interaction, complexing, ion exchange. Another mechanism of accumulation of ions on the soil minerals is heterocoagulation of colloids which have been formed by ions. The adsorption process, connected with the influence of coulombic interaction, is defined as nonspecific adsorption. It is caused by the electric charge on the surface of the metal oxide, which caused distribution of the ions in the surrounding solution layer. In consequence, the ions of the same sign as the sign of surface charge of the particle of the solid will be removed from that region of the solution, whereas the ions of the opposite charge will be accumulated. The layer of the solution that is under the influence of coulombic forces of the solid surface charge is called the diffusion layer. The whole range of the solid with accumulated charge, together with the part of solution with compensating charge is defined as an electrical double layer (edl). The distribution of the charge in the diffusion layer is relatively well described by Gouy-Chapman theory of the edl, where the charge density near the flat plane of the solid is equal:

~d = - ~ / 8 ~ ~

sinh/F~d)2RT

(I)

where: ~d - diffuse layer charge density, ~- relative dielectric constant for water = 78.25, ~o-absolute dielectric constant = 8.85*10~2CemU-~, c - electrolyte concentration, R - g a s constant - 8.314 J*mol~*K-~, T - temperature, ~gd- diffuse layer potential, F - Faraday constant 96500 C'mole 1.

354 Electric charge on the metal oxide surface is formed because of acid-base reactions of the surface hydroxyl groups. In colloid chemistry there are two approaches to an electric charge formation on the surface of the oxide. The first one by the reactions of the ionization surface group, defined as 2-pK model [8]: -SOH~

~

-SOH+H

(2)

§

- S O H +_~ - S O - + H +

(3)

where - S represents some surface of metal oxide irrespective of metal atom n u m b e r that coordinates the oxygen atom of the hydroxyl group. The reaction 2 and 3 constants may be calculated from the charge density data as a function of pH or from ~ potential versus pH dependence. The alternate attitude to the formation of the charge on the metal oxide is proposed by MUltiSIte Complexation model (MUSIC model) where hydroxyl group on the oxide surface are gifted by the charge that depends on the degree of saturation of the oxide valence by coordinating metal atoms and hydrogen, connected by hydrogen bound by donor or acceptor bonding [9]. (--MekO(H)m(HH20)n~ init + H + ~

( - M e k O ( H ) m + l ( H H 2 0 ) n - l ~ fin

(4)

where: Sinit(fin) - charge of the surface group before or after adsorption of hydrogen, which depends on the number and charge of coordinated metal atoms (k and SMe) and number (m) of hydrogen atoms (donor type connection) and (n) number of hydrogen atoms (acceptor type connection), s = ~ SMr + m* s H + n* (1 - s~) + V, SH = 0.8, whereas V = -2. For the number of free orbitals of the oxygen there is restriction for the hydrogen bonds for k = 1 ~ n + m = 2, for k = 2 ~ n + m =2 or n + m = 1 whereas for triple coordinated oxygen atoms (k = 3), n + m = 1. That means that hydrogen atom may be bound to surface oxygen in the acceptor or donor way [9]. Reaction constants of surface groups in MUSIC model are calculated theoretically from crystallographic data. First, from Brown theory, the valence of metal in the lattice of metal oxide is calculated [10], next, the charge of the surface group and finally, on this basis, the constants of the surface groups. The increase of the electrolyte concentration in the metal oxide/electrolyte system causes the increase of the charge density at the interface due to the following reactions: - SOH~An-SOH+Ct

~

+ ~

- SOH + H § + An-

(5)

-SO-Ct ++H +

(6)

According to the site binding theory, anions, which reacts with the hydroxyl group, produce surface complex type compound. The positive charge of this group is

355 located in the surface layer. that occupies position in the concentration of -= SOH ~Ansolution and increase with Similarly, for reaction 6, the

It is compensated by the negative charge of the anion inner Helmholtz plane (IHP). Following reaction 5, the groups should decrease with the increase of pH of the the increase of the concentration of the electrolyte. adsorption of the cation results in the formation of the

surface compounds - S O - C t +- type complex, where negatively charged part is in surface plane of edl, whereas the cation is in the IHP. Because the ions, adsorbing according to the reactions 5 and 6, form the complex type connections not only by electrostatic but also chemical forces, this type of adsorption is called specific adsorption of the ions [11]. The Cs-137 and Cs-134 isotopes adsorb in the same way as the stable caesium. However, because of their low concentrations (for example 1Bq/dm 3 Cs-137 equals to 1"10 -15 mole/dm 3) towards 1:1 salts present in soil water, the processes of specific and nonspecific adsorption will have minor importance. The excess of monovalent cation Na § or K § will lower the adsorption of Cs § because of the competitive adsorption on the same site. The concentration of potassium ions, in aqueous solution of the soil, reaches 2mg/dm 3 (about 50 ~mole/dm~). These ions will adsorb on hydroxyl groups according to reaction 6, whereas the adsorption of Cs § ions from such dilute solution will be limited [12]. For example complexation constant of Na § for TiO2, pKNa=8.2 and Cs § pKc~=7.2. For the concentrations Na+=lmmole/dm ~ and Cs+=lpmole/dm 3 [lkBq of Cs-137/dm 3] the [-TiO-Na ~] to [-TiO-Cs § relation, calculated with neglecting of the radius difference and activity coefficient was 10-s, based on the following equation:

I TiO-Cs TiO-Na-

Kcs * [Cs+]= 10 -7,2 ,10-12 KNa [Na + ]

= 10 -8

(7)

10 .8,2 , 1 0 - 3

That confirms the above opinion of the negligibly low specific adsorption of the Cs-137 or Cs-134 from the soil solutions containing other also alkaline metal cations Na § or K +. Only contamination by Rb-87, whose concentration (1Bq/dm 3 =1 mole/dm~), may cause appreciable specific adsorption. After all, caesium isotopes may adsorb on metal oxides by the exchange reaction with respective ions, presented in the oxide for example as contamination. Another process, responsible for the deposition of the caesium on the solid surface may be heterocoagulaton of the pseudocolloidal form of Cs [13]. This mechanism will be discussed later. The exchangeable adsorption of ions on the metal oxides occurs in the presence of the ion type contaminations. On the surface of the oxide, beside the adsorption of the cation according to reaction 6, the substitution of the contamination for to the Cs takes place. = C t s + C s + ~-

-Cs s+Ct +

(8)

356 As far as the concentration of the cation on the surface of the oxide does not change with pH, the adsorption, according to reaction 8, is independent on the pH of the solution. On the other hand, because in the exchange reaction the H § ions may take part, then a small pH dependence of the Cs sorption can be observed. The investigation of caesium sorption on the titanium, aluminum and silicon oxides, performed by Hakem et al., revealed that the increase of the concentration of the electrolyte lowers the adsorption of the Cs-137 or 1-131 [14]. The pH dependence of the sorption of these radionuclides is typical for the adsorption of the ions on oxides. However, the higher adsorption of the cation at pH>pHpzc and anion at pH>pHpzc was observed. This behavior suggests that ion exchange process has a vivid share in the ion adsorption on the surface of the oxides. The authors of discussed papers characterize applied oxides by mentioning the size of the particles and specific surface, without telling about the existence of the ion contamination. Ionic impurities of metal oxide may have influence on the mechanism of the ion adsorption, especially from very diluted solutions-10-Smole/dm 3. The investigations made by Kosmulski et al. showed that porous glass, containing on its surface borsodium phase, is good adsorbent for Cs-137115-17]. The adsorption of this isotope is promoted by alkaline pH and low ionic strength. Some adsorption of the Cs-137 was observed on the silica gel [15]. Appropriately prepared four component glasses, Vycor-type, showed the good adsorption of Cs-137 [18-19], also not only in the alkaline pH, as it happened for three component glasses. Although the examination of the adsorption on the porous glasses focused on the obtaining the adsorbent for the removing the Cs radioisotopes from water, the achieved results showed that the presence of ions or the ion exchangeable layer on the surface of the oxide increases the adsorption of monovalent ions from the solution. Ion exchange character of the Cs adsorption on the soil sample that mainly consists of the sand was observed by Shenber and Johanson [20]. The adsorption of multivalent ions or monovalent hydrolyzable ions is specific adsorption. Because of the valence of the ion, more than one adsorption site may be occupied. The adsorption of hydrated form may go through dissociation of the hydrogen cation from the hydroxyl group of the adsorbed complexes as well as from the surface hydroxyl group. Because the adsorption of the metal cations on the surface hydroxyl group goes with dissociation of H § then the adsorption of cations in the some range abruptly increases with pH. This effect is called the edge of adsorption. The parameters that characterize the edge of adsorption [22], are explained in Figure 1. The specific adsorption may lead to the formation of inner or outersphere complexes [23,24]. As an innersphere complex is treated surface compound where the cation is directly connected with oxygen from the surface of the metal oxide (Figure 2a). The outersphere complex is formed when the adsorbed cation maintains the hydrated layer of water, Figure 2b. From the pH dependence on the adsorption, one cannot conclude, whether the inner or outersphere complex is formed.

357 ApHlo.9o%

20.0 --

6.4

9 -- ,.L

_--

o~ 15.0 - -

-- 6.0

E "O

_o O

E E

E N

10.0

0 e.0 0




0,2 0,0 0

i

i

i

i

i

0,2

0,4

0,6

0,8

l

C-13

0,0

0

0,2

,

,

,

0,4

0,6

0,8

P/Po

P/Po

Figure 1. Adsorption isotherms of (a) benzene and (b) 2,2-DMB on C-series.

0,8

a

H-30

O,8 ---- H-24

0,6 ~

~

~3

H-13

"~ 0,4 > 0,2

b

H-30

0,6 ~

H

-

2

4

0,4 H-13

H-2

0,2 H-2

0

!

i

i

i

i

0,2

0,4

0,6

0,8

1

P/Po

0

ll'--ll"--r

0

t

0,2

~

I ~ -

0,4

I

I

0,6

0,8

P/Po

Figure 2. Adsorption isotherms of (a) benzene and (b) 2,2-DMB on H-series.

The analysis of the isotherms gives information not only on the adsorption capacities of the carbons but also on the porous structure which is responsible for the behaviour of the adsorbents. Among the several approaches that can be used to analyse the adsorption isotherms, the Dubinin-Radushkevich [32,33] (DR) theory of volume filling of micropores (equation 2), was chosen and the results were compared with those obtained by applying the linear form of the Langmuir isotherm [34]. The DR equation reads:

V=V 0exp-

1 P/ 21

E0~X P 0

(2)

Where Vo is the total volume of micropores, Eo is the characteristic energy of adsorption and ~ is the "similarity coefficient" which depends on the adsorptive. Plots of the DR equation are represented in Figures 3 and 4 for all the hydrocarbons on H-13 and H-30. In general, the linearity of these representations extends up to P/Po = 0.25 except for 2,2-DMB and cyclohexane on H-2 and C-2,

402

-0,1

-0,2 -0,3

-

9

C6H14 C6H6

> -o,4

~

2,2-DMB

-0,5

C6HI2

C6H6

-o,2 ~"

~ ~

a

~

-"

-

~

-0,3

C6H14 ~ " ~

2,2-DMB

"~"---n-----~~..

C6Hi2

-0,6 0

, 1

Ig2 P/Po

-0,4

, 2

0

,

,

1 lg2 P/Po

2

Figure 4. DR plots for H-30.

Figure 3. DR plots for H-13.

which display a very pronounced u p w a r d t u r n before this relative p r e s s u r e was attained. The total volume adsorbed of each adsorbate, Vo, is obtained from e x t r a p o l a t i o n of the l i n e a r r a n g e of these plots. Table 2 shows the Vo values calculated a n d the ratio of the a m o u n t adsorbed at a relative p r e s s u r e of 0.1, V0.1, to t h a t at the s a t u r a t i o n relative pressure, Vs, V01/Vs . It is a p p a r e n t that, with the exception above mentioned, the adsorption process is almost completed at low relative

Table 2 Volumes of adsorption obtained from DR r e p r e s e n t a t i o n s Vo (cm3/g)

Vo.JVs x l 0 0

adsorbates

H-2

H-13

H-24

H-30

H-2

H-13

H-24

H-30

n-hexane

0.30

0.48

0.70

0.66

83

93

93

95

benzene

0.32

0.47

0.63

0.70

78

95

91

90

cyclohexane

0.06

0.43

0.58

0.56

40

93

93

91

2,2-DMB

0.02

0.37

0.57

0.62

20

73

88

92

C-2

C-13

C-24

C-30

C-2

C-13

C-24

C-30

n-hexane

0.40

0.43

0.61

0.56

82

95

95

96

benzene

0.37

0.42

0.52

0.60

85

93

89

91

cyclohexane

0.08

0.37

0.52

0.56

59

94

90

92

2,2-DMB

0.09

0.17

0.62

0.59

43

79

89

96

Reprinted from M. Domingo-Garcia et al. [14].

403 pressures, which indicates t h a t the adsorption is not allowed to develop to many multilayers at higher relative pressures [35] and t h a t the adsorbents have reasonably narrow pore systems. The Vo/V'o ratios for n-hexane/2,2-DMB and benzene/2,2-DMB couples (which have the largest size difference) are 15 and 16 in H-2 sample, and are very close to unity in the rest of the series. For C-series the ratios for C-2 are smaller and also very close to unity in the rest of the series. This means t h a t the discrimination capacity of these carbon samples for these molecules is low, although as explained above, there is some molecular sieve behaviour at low degree of t r e a t m e n t (H-2 and C-2). U p w a r d deviations from linearity of DR plots are generally interpreted as an "additional" adsorption capacity of the adsorbates on supermicropores and small mesopores or multilayer formation on non porous surfaces [35,36]. DR plots for the adsorption of 2,2-DMB on the H-series are represented in Figure 5.

0,0 -

C-30

-0,4 -

C-24

> -0,8

," o--_..__.__~_.~____.__._____.

- 1,2

o~~___....._.~_...,~

-1,6

, 1

0

C-13

-~ C-2 , 2

" ", 3

lg? P/Po Figure 5. DR plots for the adsorption of 2,2-DMB on C-series.

It is a p p a r e n t t h a t with the increased activation time, not only a dramatic rise in the adsorption capacity is produced but also the range of linearity becomes greater and the u p w a r d deviation at high relative pressures is clearly reduced. This behaviour supports the hypothesis t h a t the adsorption of the largest molecules on carbons H-2 and C-2 mainly takes place on the external surface with multilayer formation at high relative pressures. This means that the micropore volume of these two carbons is not accessible to molecules with dimensions of cyclohexane or larger. The small molecular sieve behaviour of these two carbons and the disappearance of this property with the increase of activation time are better demonstrated by plotting the micropore volume, Vo, obtained for each carbon sample with every molecule against the m e a n molecular size of these adsorbates (Figure 6) [8,14]. From this representation it is also possible to conclude t h a t after 24 hours of activation t r e a t m e n t there is no clear improvement in the adsorption capacity on either the H or C carbons. Characteristic adsorption energies for every adsorbent, Eo, derived from the slope of the DR representations of the benzene isotherms (which is generally

404 0,8

0,8 C-24

0,6 []

m

,~0,6

H-24

fi 0,4 o

or-

~

,

,

0,45

0,5

C-13

~0,2

, ~ ~ 0,55

C-30

O

0,2

0,4

2-

E0,4

H-13

0 0,6

d (nm)

0,4

0,45

0,5 d (nm)

0,55

0,6

Figure 6. Volume adsorbed V0, obtained from DR equation, versus the mean molecular size of the adsorbates with six carbon atoms (see Table 16 for molecular dimensions).

taken as s t a n d a r d adsorbate), and the characteristic dimension of the micropores Lo [37] are compiled in Table 3.

Table 3 Characteristic adsorption energy, Eo, and characteristic dimension of the micropores, Lo Adsorbents

Eo (kJ/mol)

Lo (nm)

Adsorbents

Eo (kJ/mol)

Lo (nm)

H-2

16.60

0.78

C-2

21.68

0.60

H-13

21.40

0.62

C-13

22.06

0.58

H-24

21.71

0.59

C-24

18.22

0.71

C-30

22.80

0.57

H-30 22.44 0.58 Reprinted from: M. Domingo-Garcia et al. [ 14].

It is, therefore, apparent that the increase in activation time virtually does not modify the micropore size of the activated carbons because the values of the adsorption energy, Eo, are very similar when the activation time progresses. This could mean that the only effect of activation was to favour the access to smaller micropores. The molecular sieve behaviour shown by the less activated carbons is therefore attributed to constrictions at the entrance of the micropores caused by chemical functionalities linked to carbon atoms at the edges of the entrance [4]. After certain activation time (around 13 hours) the constrictions and, consequently, the discriminative behaviour was lost.

405 Application of the Langmuir model (P/V vs. P) to the type I isotherms gives linear plots in a wide range of pressures. VL/Vs ratios (VL obtained from the slope of the linear plots and Vs obtained from the plateau of the isotherms), are in most cases close to unity (Table 4) except for 2,2-DMB and cyclohexane on H-2 and C-2. Such results suggest that neither multilayer formation or capillary condensation occur on those systems and that the adsorbate-adsorbent interactions are far greater than the adsorbate-adsorbate ones.

Table 4 VL/Vs ratios. VL obtained from Langmuir equation. Vs obtained from DR equation VL/Vs

VL/V~

H-2

H-13

H-24

H-30

C-2

C-13

C-24

C-30

n-hexane

1.00

1.02

0.96

1.00

0.86

1.02

1.03

1.01

benzene

1.00

1.02

0.99

1.00

1.02

1.04

1.03

1.00

cyclohexane

0.56

1.02

0.98

1.00

0.67

1.00

1.03

1.00

2,2-DMB

0.42

0.97

0.98

1.04

0.57

0.90

0.95

1.00

Reprinted from: M. Domingo-Garcia et al. [ 14].

The VL values coincide very closely with the volumes, Vo, calculated by the DR equation (Table 2). It is, therefore, possible to conclude that the micropore size distribution of these carbons is in the range of the molecular size of the organic vapours adsorbed and little, if any, adsorption took place on the external surface of the adsorbents working at relative pressures, P/Po, below 0.8. Surface areas of the carbon materials were calculated using the VL values. These are considerably smaller than those obtained by CO2 measurements at 273 K for the less activated samples, H-2 and C-2, and for these carbons the area increases with a decrease in the mean molecular dimension of the adsorbate. Nevertheless, as the activation increases the areas for all the vapours and that of CO2 become quite similar. This supports the existence of constrictions at the entrance of the pores, earlier commented, hindering the access of the adsorbates. As shown previously agricultural by-products, such as olive stones and almond shells, are good raw materials to obtain carbonaceous adsorbents of organic vapours after the appropriate activation process with quite high adsorption capacities (around 0.6-0.7cm3/g), which is important for their potential application as pollutant removal adsorbents. For these samples, prepared by carbonization and simultaneous activation in C02, the process acts mainly by opening small micropores and eliminating chemical functionalities hindering the access of the adsorbates.

406 2.2.

Carbon

materials

from polyfurfuryl

alcohol

(Glassy

carbons)

The preparation of these carbon materials basically consisted of the polymerization and slow carbonization of furfuryl alcohol [13,21,38,39]. Some t e x t u r a l characteristics like pore volumes and a p p a r e n t surface areas are s u m m a r i z e d in Table 5. I m p o r t a n t differences in the pore size distribution in the region of meso, V2, and macropores, V3, are found for the samples. Moreover, SN2 0,05

0,05

~

b

~ P2

0,00 . 0

.

.

. . 0,00 T 200 300 400 0,8 1,2 1,6 0 100 t (min) t (min) Figure 7. Kinetics adsorption curves on glassy carbons: (a) benzene and (b) methyl iodide. 0,4

i

l

i

!

500

407 Table 6 Apparent adsorption rate, RL, and diffusion parameters, adsorption of methyl iodide and benzene

Dl/2/ro, for the

RL (cm 3 g" min 1/2) P1

P2

P3

Benzene

0.023

0.002

0.043

methyl iodide

0.049

0.015

0.068

D~/ro (min -~)

Benzene

P1

P2

P3

0.178

0.020

0.419

0.097

0.580

methyl iodide 0.572 Reprinted from: M. Domingo-Garcia et al. [12].

These data indicate considerable differences for the adsorption process of these hydrocarbons on each sample, particularly on sample P2 for which these parameters are more than 10 times lower than for P1 and 20 times lower than for P3 when benzene is adsorbed. It is possible, therefore, to assume important differences in the porous system of the three carbon samples due to the different preparation recipes. Actually, from data in Table 5 one can observe that carbon P2 has no macroporosity at all, whereas P1 presents a large one. It is known that macroporosity is very important for the transport and quick penetration of the adsorbate molecules to the micropores. However, the kinetics parameters are lower for carbon P1 t h a n for carbon P3, although the latter shows a less extensive macroporosity. This fact, together with data in Table 5 and the restricted N2 adsorption for carbon P2 due to pore constrictions, suggests that there are also important differences in their microporous system. On the other hand, the adsorption rate and the diffusion parameters of methyl iodide on all the samples are much higher t h a n those of benzene and this is more marked for P1 and P2 than for carbon P3. This means that at the same temperature and relative pressure methyl iodide is kinetically better adsorbed than benzene. The adsorption-desorption isotherms of benzene, n-hexane and 2,2-DMB are depicted in Figure 8. Most of them can be assigned to BDDT type I, i.e. corresponding to predominantly microporous adsorbents; nevertheless, differences in their shape and in the desorption process also indicate a variety in the characteristics of the porous system. Isotherms of benzene are depicted in Figure 8(a). For carbons P1 and P3 the isotherms obtained with 2 hours of adsorption time are coincident with those obtained after 12 hours; therefore, only the isotherms obtained with long adsorption time are depicted in Figure 8. For carbon P2, however, the adsorption-desorption isotherm obtained after 2 hours of adsorption time is very different to that obtained after 12 hours and both have

408

0,3

P3

0,25 P3

a

0,20 0,2

-~-: ~ . _ _. ~ ...... - . .o ~. - -.~ ,.~ - ~ o ~~

Pl

P1

-~ 0,15

E gO,1

"

_,-

A-

> 0,10

P2 (12 h)

P2 I

0,05 zl~*! ** *,

*

0,0 0

0,2

*,

,

0,4

0,6

-! m - -ID-

P*2 (2 h) ,

0,00

0,8

0,0

HI

-I--

!

I

0,2

0,4

I

0,6

0,8

1,0

P/Po

P/Po

0,2 P3

~0,1 E > 0,1

0,0 0,0

i

i

i

i

0,2

0,4

0,6

0,8

i

1,0

P/Po

Figure 8. Adsorption isotherms on the glassy carbons: (a) benzene, (b) n-hexane and (c) 2,2-DMB.

been depicted in this figure. The shape of all these isotherms suggests that sample P1 appears to have predominant adsorption on micropores with little participation on the external surface. However, samples P2 and P3 exhibit an upward deviation in the region of high relative pressure, which is normally interpreted either as capillary condensation in supermicropores and small mesopores or adsorption on the external surface of the adsorbent [35,36]. These two effects follow a different adsorption mechanism to that of adsorption on micropores; capillary condensation often produces an hysteresis loop at high relative pressure when the desorption process is carried out. Desorption isotherms represented in Figure 8(a) show a very small deviation from the adsorption branches on samples P1 and P3, which extends to the lowest relative pressure region. The value of these deviations in volume of adsorbate is less than 0.01 cmS/g. Nevertheless, a very pronounced hysteresis loop in the complete range of relative pressure studied, is found for the two isotherms on sample P2. Of the several hypothesis given [1,42] to explain the appearance of low pressures hysteresis loop, taking into account the high rigidity of glassy carbons, it has to be considered that activated passage of molecules through pre-existing constrictions into wider pores appears to be the most appropriate for these

409 samples. This interpretation is consistent with the restricted adsorption of N2 already mentioned found for carbon P2 (SN2o)

=

--2,00

chloroform .

3,ool

.

6

9

methyl iodide

12

0

Ig2 (P/Po)/IT~

chloroform 5

,

,

10

15

Ig2(e/Po)/~ 2

-0,50 ] c

methyl iodide

~-1,00" ~ ~ n z e n e __

\ I -2,00 . 0

o,',,oro,orrn

9

2,2-DMB . .

.

5

10

15

Ig2 (P/Po)/132

Figure 10. Dubinin-Radushkevich plots of benzene, n-hexane, 2,2-DMB, methyl iodide and chloroform on (a) P 1, (b) P2 and (c) P3.

Plots for carbon P1 (Figure 10 a) seem to indicate that the adsorption of benzene, n-hexane, methyl iodide and chloroform takes place on the same type of micropores. However, the downward deviation from linearity observed for benzene, n-hexane and chloroform at low relative pressures can indicate that in this region equilibrium is not attained for the adsorption of these molecules. This type of deviation is often explained in terms of restricted diffusion into the narrowest pores or because the molecular sieve effect can lead to difficulties in adsorption at low relative pressures[2,43]. It should be noted that for n-hexane this deviation extends up to much higher values of relative pressure than for chlorofom, although the minimum critical size of the latter is somewhat

411 greater.This fact could be due to packing restrictions in the micropores [2] because of the great length of n-hexane. On the contrary, for methyl iodide an upward deviation appears in this region of low relative pressures, suggesting that this molecule has good accessibility into the narrowest pores.

Table 7 Micropore volume, Vo (cm3/g), from DR equation and volume at P/Po =0.9, Vs P1 P2 P3 Vo

V~

Vo

V~

Vo

V~

Benzene

0.173

0.21

0.068

0.14

0.156

0.23

n-hexane

0.169

0.19

0.030

0.06

0.113

0.22

0.092

0.13

2,2-DMB methyl iodide

0.176

0.20

0.093

0.17

0.114

0.23

chloroform

0.177

0.21

0.072

0.13

0.113

0.32

Reprinted from: M. Domingo-Garcia et al. [12].

For carbon P2 the DR curves of the four adsorbates are depicted in Figure 10(b). On this adsorbent the curves of benzene, methyl iodide and chloroform corresponding to isotherms with the same adsorption time are not at all coincident. As previously indicated by the kinetics adsorption measurements, benzene and methyl iodide present very low diffusion parameters on this carbon, although for the latter this is almost five times higher than for the former. Actually, the volume Vs of benzene adsorbed increased from 0.07 cma/g after 2 hours of adsorption time up to 0.14 cma/g after 12 hours, and up to 0.17 cm3/g after 5 days. For methyl iodide the same volume of 0.17 cm3/g is attained after 12 hours. This fact could explain the very different slopes for the corresponding characteristic curves. Therefore, although adsorption on carbon P2 appears to be kinetically restricted for the three adsorptives, this is much smaller for methyl iodide than for benzene and the highest restriction is found for chloroform. As already suggested this restriction seems to be due to constrictions in the entrance to the pores [41]. Another fact to be considered is the high dipole moment of methyl iodide, which can produce specific interactions with the oxygen functional groups, which are on the carbon surface. This contribution accumulates with the dispersion forces leading to a considerable increase of the adsorption energy in the initial region of the adsorption process, i.e. at very low relative pressure. This specific contribution decreases following micropore filling of the adsorbent [6,23]. Chloroform is the less easily adsorbed of the three adsorptives as its critical

412 dimension of 0.43 nm is the greatest. The fact t h a t chloroform, which is also a polar molecule, suffers stronger restriction t h a n benzene to access micropores of P2 indicates t h a t the main p a r a m e t e r s controlling the passage through the pores constrictions is the critical dimension and the shape of the molecule. The molecular sieve effect between benzene and cyclohexane confirms the network of these glassy carbons to be made up by slit-shaped micropores. As pointed out earlier, the micropore system of carbon P3 also appears to have i m p o r t a n t differences with respect to those of carbon P1, P2 and P4. This becomes more obvious when the characteristic curves of the different adsorbates are analysed. The most i m p o r t a n t difference is t h a t the adsorption capacity of cyclohexane and 2,2-DMB on sample P3, although smaller t h a n t h a t of the others adsorbates, turns out to be quite important. Nevertheless, comparing the curves for the different hydrocarbons, it can be observed t h a t the slope of the linear region of the curves obtained with cyclohexane and 2,2-DMB is much steeper t h a n t h a t of the other curves, which usually suggests adsorption on wider pores [33]. Apart from this, the curves of benzene and methyl iodide are coincident as well as those of n-hexane and chloroform. The adsorption of n-hexane seems to suffer some restrictions at low relative pressures while for chloroform only a linear region t h a t extends over a wide range of relative pressures, (P/Po from 10 .3 to 1.5 10 -1) is found. Moreover, at very low relative pressures an upward deviation appears not only for the characteristic curve of methyl iodide but also for t h a t of benzene. On the other hand, observing the Vs values given in Table 7 it is evident t h a t certain molecular sieve behaviour between chloroform and cyclohexane is also shown by carbon P3. The steep slopes of the characteristic curves of cyclohexane and 2,2-DMB suggest t h a t adsorption probably takes place in the supermicropores and small mesopores. It is interesting to note t h a t all plots in Figure 10 show an upward deviation from linearity near the s a t u r a t i o n relative pressures which can be attributed to the filling of supermicropores or small mesopores by a co-operative m e c h a n i s m [44] which involves little, if any, e n h a n c e m e n t of the adsorption energy. Characteristic energies of adsorption, Eo, calculated from the linear region of the characteristic curves of benzene, methyl iodide and 2,2-DMB on carbons P1 and P3 are listed in Table 8. Moreover the corresponding average micropores width, Lo, and the range of validity of r e l a t i v e pressures from which these p a r a m e t e r s were obtained, are also included. The average micropore width, Lo, obtained for carbon P1 with benzene is quite coincident with t h a t calculated with methyl iodide. However, for carbon P3 a slight difference of this p a r a m e t e r is found between these two adsorbates. The very low characteristic adsorption energy, Eo= 6.1 kJ/mol, corresponding to 2,2-DMB on this carbon is typical of adsorption on supermicropores or even small mesopores as pointed out earlier. Therefore one can conclude t h a t these data do not correspond to a specially narrow microporous system, which is really surprising taking into account the high discrimination effect already described.

413 Table 8 Characteristic adsorption energy, Eo, and average micropore width, Lo

P1

benzene

methyl iodide

Eo (kJ/mol)

20.9

21.7

Lo (nm)

1.17

1.10

2,2-DMB

Range of validity for P/Po 4.5 10-3-1.7 10 -1 1.3 102-3.5 10 -1 P3

Eo (kJ/mol)

17.1

19.8

Lo (nm)

1.5

1.26

6.1

Range of validity for P/Po 2.7 10-2-1.6 10 -1 3.0 10-2-3.5 10 -1 4.0 10-2-3.0 10 -1 Reprinted from: M. Domingo-Garcia et al. [ 12].

An u p w a r d deviation from the linearity in the region of low relative pressures (such as the adsorption of methyl iodide on carbons P1 and P3 or benzene on carbon P3) in DR plots could be the consequence of the superposition of two extreme ranges of microporosity [33]. These two ranges can be well approximated by a binomial equation, known as the Dubinin-Isotova (DI) equation:

E/1

V=V01exp-

EOl[3 x RT In

+ Vo2 exp -

E0213

(3)

V01 is the micropores volume obtained from the linear region at very low relative pressures and according to Stoeckli et al. [45], could be related with the micropore filling process which gives a characteristic adsorption energy Eol. V02, and E02 have been associated with the beginning of the secondary micropore filling process, i.e. the volume and energy of adsorption on the walls of relatively large micropores. Therefore the volume V02 could provide information on the monolayer capacity of the largest micropores walls. Table 9 contains all these p a r a m e t e r s for the adsorption of methyl iodide on samples P1 and P3 and of benzene on P3. From these new data one can assume t h a t more t h a n 70% of the total adsorption volume Vs (Table 7) on carbon P1 takes place on micropores with an average dimension, L01, of 0.74 nm. Nevertheless this new value is still not small enough to explain either the drastic molecular sieve behaviour found for cyclohexane or the restricted diffusion of benzene at low values of relative pressures shown earlier. This fact supports the existence of some kind of constrictions at the entrance of the pores. For carbon P3, V01 and E01 data for both benzene and methyl iodide are quite close and indicate t h a t 56% of the total adsorption volume takes place on micropores of about 1 nm or less. This means t h a t 44% of the total micropores volume (i.e., 0.102 cma/g) corresponds to supermicropores or small mesopores which filled up by the secondary mechanism

414 with very low adsorption energy Eo2. This a m o u n t is very similar to the Vo value found for the adsorption of 2,2-DMB on P3, which means t h a t this molecular probe is only adsorbed on the wider pores by a secondary mechanism. Table 9 P a r a m e t e r s obtained from the Dubinin-Isotova equation Adsorbate

Vol (cma/g) Eol(kJ/mol)

Lol(nm) Range of validity for P/Po

P1

methyl iodide

0.146

27.7

0.74

3.0 10-3-3.0 10 .2

P3

benzene

0.128

21.4

1.10

3.0 10-3-3.0 10 .2

methyl iodide

0.120

23.2

1.01

3.0 10-3-3.5 10 .2

V02 (cma/g) Eo2(kJ/mol)

Range of validity for P/Po

P1

methyl iodide

0.040

7.7

3.0 10-2-5.0 10 -1

P3

benzene

0.049

5.0

3.0 10-2-6.0 10 -1

methyl iodide

0.047

6.4

4.0 102-5.0 10 -1

Reprinted from: M. Domingo-Garcia et al. [12].

From the analysis of these four glassy carbons one can conclude t h a t carbons P1, P2 and P4 have a complete discrimination (molecular sieve behaviour) for the adsorption of chloroform, with a critical dimension of around 0.43 nm, and cyclohexane, with a critical dimension of 0.56 nm. Nevertheless, when the isotherms of benzene or methyl iodide on sample P1, P2 and P4 are analysed by the application of DR and DI models the average micropore sizes calculated do not justified theirs molecular sieve properties. This fact, along with the kinetics results, indicates t h a t constrictions or some narrowness at the entrance to the pores are responsible for the molecular sieve character found for these glassy carbons. Carbon P3, however, shows a different behaviour for the adsorption of these molecules, and although it acts as a molecular sieve for cyclohexane and 2,2-DMB at low relative pressure, these two hydrocarbons were adsorbed on supermicropores or small mesopores. As a general conclusion, one can point out t h a t both kind of adsorbents here discussed, activated carbons of lignocelulosic origin and glassy carbons, can be tailored to adsorb selectively organic molecules. For the former carbons this behaviour essentially depends on activation time, while for the latter it depends on the preparation formula. In addition, this discriminative behaviour is much higher on glassy carbons t h a n on those obtained from agricultural by-products. Nonetheless, the adsorption capacities of activated carbons of lignocelulosic origin are always much higher t h a n those of the glassy carbons.

415

2.3. A d s o r p t i o n of CO2 f r o m d i l u t e d e n v i r o n m e n t s A particular case of interest is the adsorption of CO2. This process is currently used as an almost routine measurement in order to determine the surface area of adsorbents. For this purpose the adsorption is carried out from low to high relative CO2 concentration. The data to be reported deal with the adsorption kinetics of CO2 from diluted CO2-N2 mixtures, conditions prevailing in flue gases produced in fossil-fuel-based power plants. Activated carbons prepared by carbonization (H0) and further activation by CO2 (H14, H25 and H35) and by steam (HW20 and HW44) have been used as adsorbents. Additional experimental and theoretical data are given elsewhere [46]. The textural characteristics and the micropore size distributions of these samples are shown in Table 10 and Figure 11.

Table 10 Textural characteristics of the active carbons SN2(m2/g)

Sco2(m2/g)

Vo (cm3/g)

Eo (kJ/mol)

H0

246

813

0.314

22.23

HI4

725

974

0.376

20.92

H25

910

1012

0.391

20.68

H35

1190

1220

0.471

19.01

821

0.317

20.58

HW40 1475 1517 Reprinted from: M. A. Salas-Peregrin et al. [46].

0.586

15.45

HW20

812

3 H25

E2

] |

.-.

b HW20

H35

E

.._... _J

}Q1 0

0,8

"EI

|

|

!

|

|

1

1,2

1,4

1,6

1,8

0

0,8

L (nm)

Figure 11. Micropore size distribution of activated carbons.

,

,

i

1,3

1,8

2,3

L (nm)

416 Diffusion parameters were obtained using equation 1. The values of Ve, Vo and D~/ro at 298 K obtained with a CO2-N2 mixture containing 13.5 % of CO2 by volume for all these activated carbons are compiled in Table 11. These results can be explained on the basis of the micropore characteristics of the adsorbents. Therefore, the carbonized sample, H0, has the smallest value of Dl/2/ro because it possesses the narrowest micropores as can be deduced from its highest adsorption energy, Eo (Table 10) and from the distribution of micropores (Figure 11). The negative value of Vo indicates that there is a retardation in the CO2 adsorption process.

Table 11 Values of Ve, Vo and D~/ro obtained with a CO2-N2 mixture containing 13.5% of CO_9 by volume Sample

Ve

(cma/g)

Vo (cma/g)

D~/ro (s -~)

H0

16.38

-0.09

0.0083

HI4

18.30

5.17

0.0174

H25

18.31

5.96

0.0174

H35

17.11

2.98

0.0174

HW20

17.86

4.67

0.0177

HW44

14.22

3.72

0.0204

Reprinted from: M. A. Salas-Peregrin et al. [46]. When sample H0 is activated in either CO2 (H-series) or steam (HW-series), there is an increase of both Ve and D1/2/ro, and Vo reaches a positive value so that there is no retardation in the CO2 adsorption process. The increase in Ve and D1/2/ro is produced as a consequence of the raised micropore volume and because of an opening of the microporosity produced by the activation process, which can be deduced from the decrease in Eo (Table 10) and from Figure 11. Samples H14 and H25 have coincident Ve and D1/2/ro values because these two samples have almost the same micropore size distribution and similar Eo values. When activation increases, in samples H35 and HW44, there is a decrease in the Ve value, and this is more marked in the case of the most activated sample, HW44. These results, at first, are surprising because H35 and HW44 samples have the highest micropore volume from their respective series. Thus, these results indicate that for a low CO2 concentration in the gas phase (13.5% of CO2 in N2) the higher the degree of activation of the carbon the lower CO2 adsorption capacity at equilibrium. The same trend, as shown in Table 10, is found for other lower C02 concentrations in the C02-N2 mixture. Similar results have been reported [47] for the adsorption of Volatile Organic Compounds (VOC) at trace level onto activated carbon fibres.

417 These results can be explained as being due to a decrease in the number of the smallest micropores available in the most activated samples, which can be deduced from Figure 11 for H35 and HW44, respectively. This occurs because the wider micropores do not benefit from the overlapping adsorption potential of opposite pore walls, and thus, do no experience the enhanced adsorption observed for the narrower micropores. The results found are quite important, because they show that when using activated carbons to remove or store CO2 from low concentration environments, their surface area or pore volume alone is not an adequate design parameter, but micropore size distribution is the controlling factor. Therefore, these results indicate that activated carbon H14 (a sample with a low degree of activation) performs best under these experimental conditions. Moreover, these results together with those reported [47] permit the above conclusions to be applied not only for the adsorption of CO2 but also to the adsorption of VOC at very low concentrations. Consequently, these ideas should be borne in mind when analyzing results in the following sections. 3. A D S O R P T I O N F R O M V E R Y D I L U T E D A T M O S P H E R E S

In the previous sections the adsorption of several molecules at relatively high concentrations measured in gravimetric systems has been considered. In this section the adsorption of VOC under dynamic conditions, at relatively high temperatures and at very low vapour concentration is considered. Among the VOC, several hydrocarbons (linear, cyclic and branched) and some organic compounds with different functionalities as acetone, diethyl ether, tetrahydrofurane, carbon tetrachloryde, chloroform, dichloromethane, methyl iodide and n-alcohols have been chosen. Adsorption has been studied using IGCS. Details of the experimental conditions are give elsewhere [13,22,26,27,29,41,48]. 3.1. A d s o r p t i o n o f V O C o n n o n - p o r o u s c a r b o n m a t e r i a l s

When adsorption is produced on non-porous surfaces the process is controlled by displacement of the electronic density of the molecules produced by action of the electrostatic field of the adsorbent. As a consequence of this the degree of adsorption of linear hydrocarbons increases with the number of carbon atoms of the molecule. Therefore, the specific retention volume Vs (a p a r a m e t e r which measures the degree of adsorption) of, for example, n-nonane is higher than that of n-hexane. Moreover, the degree of adsorption decreases as the temperature increases. These kind of behaviours are shown in Table 12 for the adsorption of linear hydrocarbons at two temperatures on three non-porous carbon materials: a graphitized carbon black (V3G) and two graphites (Pyrolitic and Acheson). For the same reason as that explained above, the relationship at different temperatures between in Vs versus the polarizability of the n-alkanes is linear. Moreover, the differential heats of adsorption are, in absolute values, low and close to the liquefaction heats of the adsorbates [29]. This means that the adsorbate-adsorbent interactions are similar to a d s o r b a t e - a d s o r b a t e ones, i.e.

418 Table 12 Specific retention volumes for the adsorption of n-alkanes on non-porous graphites

Vs (cma/m 2) 343 K

363 K

Adsorbate

V3G

Pyrolitic

Acheson

V3G

Pyrolitic

Acheson

n-C6

4.6

2.6

2.2

2.6

1.5

1.5

n-C7

19.4

7.4

8.4

9.2

3.4

4.3

n-Cs

81.7

28.8

34.0

34.9

12.5

12.9

n-C9

331.6

121.1

109.6

125.5

47.6

44.7

the adsorbate-adsorbent interaction is produced through London dispersion forces. W h a t it is noteworthy is t h a t this kind of behaviour remains in the adsorption on non-porous carbon materials of hydrocarbons which have non-linear shapes. This is the case of benzene, cyclohexane and isooctane. The interaction of these adsorbates on non-porous graphite (Degussa) is similar to t h a t of n-alkanes. Therefore, one can conclude t h a t the adsorption is produced as a consequence of a displacement of the charges of the molecules under the action of the surface of the adsorbent (non-specific interaction). This is shown in Figure 12.

7 363 K

S

if}

>

3

Isooctane Benzene

J

.t,

9 n-alkanes

,~~" Cyclohexane

8

i

i

i

12

16

20

Ct(A3) Figure 12. Variation of In Vs versus the polarizability of the adsorbates for the adsorption on a non-porous graphite.

However, this behaviour is different if the adsorbate has the same shape as t h a t of the n-alkanes but different chemical functionalities. This is the case with the data in Table 13, in which the Vs values of n-alkanes and n-alcohols on graphite (Degussa) are compiled. The degree of adsorption is larger for an

419 Table 13 Specific retention volumes and differential heats of adsorption of n-alcohols and n-alkanes on graphite (Degussa) Vs (cma/m 2)

.AHOA

Adsorbate

333 K

343 K

353 K

363 K

(kJ/mol)

l-propanol

3.6

2.7

2.0

1.5

29.7

l-butanol

7.4

5.6

3.9

2.7

33.1

l-pentanol

13.3

9.1

6.3

4.6

35.1

l-hexanol

60.3

37.7

26.6

16.1

42.4

n-pentane

2.2

1.8

1.3

1.0

28.2

n-hexane

11.4

8.0

5.6

4.0

34.3

n-heptane

42.1

27.7

18.7

12.6

39.6

n-octane

134.3

76.7

50.9

34.5

44.2

n-nonane

487.9

267.7

175.9

104.6

49.6

1465.6

837.2

441.4

55.0

n-decane

n-alcohol t h a n for an n-alkane with the same n u m b e r of carbon atoms. This is because adsorption of n-alcohols depends on their deformation polarizability and on the orientation polarizability (specific interaction) which depends on the dipolar moment, p. Of these two factors, the orientation polarizability is far larger t h a n the deformation polarizability. Consequently, the behaviour of polar molecules u n d e r these conditions is different from the non-polar molecules and depends mainly on the orientation polarization. Moreover, for this reason the differential h e a t of adsorption is also higher, in absolute values, for an n-alcohol t h a n for an n-alkane with the same n u m b e r of carbon atoms. It could be desirable to improve the behaviour of these kind of materials in order to increase their capacity of adsorption. For this purpose the t r e a t m e n t of V3G with oxygen at different degrees of burn-off has been carried out. The results obtained are basically similar to these obtained with the original carbon black [26]: a linear relationship between In Vs and polarizability of the hydrocarbons, such t h a t adsorption increases with the polarizability of the chain length of the n-alkane as well as with the lowering in the adsorption t e m p e r a t u r e . However, the differential h e a t of adsorption for each n-alkane increases in absolute value as the percentage of burn-off increases, as a consequence of the irregularities produced on the surface of the sample [26]. From a practical point of view these data together indicate that the elimination by adsorption on non-porous carbons of hydrocarbons is favoured by the polarizability of the adsorbate such t h a t the adsorption increases with the

420 polarizability of the molecule. Moreover, the adsorption of n-alcohols in the same conditions is more favoured t h a n that of the n-alkanes due to the dipolar moment of the former molecules. 3.2. A d s o r p t i o n o f V O C o n p o r o u s c a r b o n m a t e r i a l s It has already been mentioned that the adsorption on these kind of materials has to be considered taking into account two main factors: the porosity and the chemical functionalities of the adsorbate and adsorbent. The adsorption of linear molecules on activated carbons follows the same trend found for non-porous carbons, i.e. the plot of in Vs versus the n u m b e r of carbon atoms (polarizability) of the adsorbates is a straight line. Therefore, the mechanism of adsorption can be considered as non-specific. Nevertheless, those non-linear hydrocarbons do not follow the same trend as that found for non-porous carbons. This behaviour is shown in Figure 13 in which the adsorption of linear (n-alkanes), branched (2,2-DMB) and cyclic hydrocarbons (benzene and cyclohexane) on two activated carbon series, C and H, is plotted.

H-13 H-2

CeH

9

"J

n-alkanes

0

Y

~ 1

> _c

> --CO

-2

~

06H12 ~,

-1

,

,

,

10

12

14

oc (A 3)

9 06H12

~ t1"

A 2,2-DMB 8

/)P . / n-alkanes

C6H6m

2

-2

8

9 2,2-DMB ,

,

,

,

,

,

9

10

11

12

13

14

o~ (A 3)

Figure 13. In Vs versus the polarizability for the adsorption on activated carbons.

The fact that these non-linear molecules do not follow the behaviour found for n-alkanes suggests that the shape and size of these molecules is an important factor in the adsorption on porous materials. This can be seen in Figure 14 in which the specific retention volumes are plotted versus the mean molecular dimensions of the adsorbates (see Table 16). It is interesting to emphasize that all the adsorbates have different shapes and sizes and the same n u m b e r of carbon atoms (six). The general trend found is that the adsorption decreases as the size of the molecules increases. Moreover, at low degree of t r e a t m e n t of the raw material (C-2 and H-2 carbons) the specific volumes of adsorption are clearly lower t h a n at higher periods of treatments. In all cases the lower volume of adsorption is for 2,2-DMB which is the largest molecule (0.60 nm). In addition the adsorption of this molecule, which is almost negligible at low degrees of treatments (C-2, C-13 and H-2) increases at higher degrees of t r e a t m e n t of the raw materials. The separation ratios, (Vs/Vs'), on H-2 for the n-hexane/2,2-DMB and benzene/2,2-DMB couples are 87.5 and 101.8.

421 20

C-24

20 ~ E

C-13

E

E E ~ {/1

H-13

tar)

>

>

10

10

0 0,40

H-3

0 0,45

0,50 d (nm)

0,55

0,60

0,40

,

,

0,45

0,50 d (rim)

, 0,55

I, 0,60

Figure 14. Specific retention volumes versus the molecular size of hydrocarbons with six carbon atoms (see Table 16 for molecular sizes).

However, for samples obtained with long periods of t r e a t m e n t s , H-30, these ratios are 2.7 and 1.5 respectively. These separation ratios are in all cases higher t h a n those obtained in the adsorption at high relatives pressures above commented. This is because the n u m b e r of molecules to be adsorbed in this case is very low. Consequently, the possibility of finding a n u m b e r of pores with a dimension close to the molecular dimension is high. In contrast to this at high relative pressures the n u m b e r of molecules to be adsorbed is very high and consequently this possibility is clearly lower. On the other hand, the separation ratios found at low concentration support the hypothesis t h a t progressive t r e a t m e n t s of the raw m a t e r i a l open the microporosity of the activated carbons. For this reason it can be concluded that, in general, the activation times higher t h a n 13 hours do not improve the capacity of adsorption of these pollutants from very diluted concentrations. Consequently, on using almond shells or olive stones to prepare activated carbons for removing these organic molecules one can diminish the period of t r e a t m e n t of the raw material. Similar results to those obtained with these series are also reported [15] for activated carbons prepared by activation of chars (HA-series), although the Vs values are slightly lower. The second factor commented above, to be considered in the adsorption of VOC concerns the chemical functionalities of both adsorbate and adsorbent. A case of practical interest is the adsorption of methyl iodide. Iodine-131 is one of the dangerous substances produced in the fission process, but it mainly survives in the atmosphere as methyl iodide. The elimination by adsorption of this substance has been studied using adsorbent materials such as zeolites, silica gel and activated carbons, but the latter seem to show the best results [48-50]. Methyl iodide has an i m p o r t a n t dipolar m o m e n t ~ = 1.62 D. As a result it can be expected to make an i m p o r t a n t contribution of specific interactions in the adsorption of this molecule. One could, therefore, expect an increase in the adsorption capacity in cases in which specific interactions are favoured.

422 To study this effect several activated carbons treated with H202 [41] have been used for the adsorption of methyl iodide. In Table 14 some selected results are recorded. Samples with the symbol O appended at the end of the name have been obtained by treatment the original samples with H202. So, the comparison of data in Table 14 deals with the effect of the oxygen functionalities, on the surface of the active carbons, on the capacity for methyl iodide adsorption. An increase in the adsorption capacity of samples treated with H202 is apparent from these data. Surprisingly, this increase does not mean a more exothermic process as could be expected from a specific interaction. In fact, these data show no energetic differences between the interaction in the original samples and in those obtained after H202 treatment, having a higher differential heat of adsorption, in absolute value, than the liquefaction heat of methyl iodide, -AHL = 27.6 kJ/mol [48].

Table 14 Specific retention volumes, Vs, at 473 K, and differential heats of adsorption, AH~ of methyl iodide on activated carbons Sample

Vs(cma/m 2)

-AHOA(kJ/mol)

Sample

Vs(cma/m 2)

-AH~ (kJ/mol)

C-2

1.2

56.1

C-2-O

1.7

53.6

H-2

0.6

64.8

H-2-O

1.2

52.2

HA2

0.6

49.2

HA2-O

0.9

58.1

HA4

1.0

52.2

HA4-O

0.9

52.5

Reprinted from: F. Carrasco-Marin et al. [48].

Another possible way to increase the capacity of activated carbons for methyl iodide adsorption is by impregnation with KI [41,51]. The data in Table 15 show an increase in the adsorption capacity of the samples treated with KI. In addition they show a decrease in the differential heats of adsorption, although they are also higher, in absolute value, than the methyl iodide liquefaction heat. Hence, the results are similar for the samples treated with H202 and with KI. This could be because the interactions in both cases are produced by a mixed mechanism (specific + non-specific) such that the interaction of the adsorbate with the pores can be more exothermic than with oxygen or with KI. This aspect will be considered later on discussing the adsorption on glassy carbons. These results show that it is possible to increase the adsorption capacity for methyl iodide of activated carbons if different chemical factors capable of specific interactions are introduced. However, these interactions do not appear to be more exothermic than those inside the pores.

423 Table 15 Specific retention volumes at 443 K, Vs, and differential heats of adsorption, AH~ of methyl iodide Vs(cm3/m 2) Sample

-AH~ (kJ/mol)

Original

With KI

Original

With KI

C-13

3.5

3.9

57.3

48.7

C-30

2.8

3.0

53.3

48.3

H-13

3.0

3.2

53.0

51.0

H-24

2.2

3.1

49.8

47.0

Merck

0.3

1.2

57.2

59.9

In relation to carbon materials with a tailored porosity for adsorption of VOC, several adsorption data obtained on carbons prepared from Saran copolymer and from polyfurfuryl alcohol have been reported [10,13,41,52,53]. The behaviour of carbon materials obtained by carbonization of Saran in the adsorption of linear hydrocarbons is the same as that described above, i.e. a linear relationship between In Vs and the number of carbon atoms. Moreover, the Vs values are higher than those obtained with other carbon materials and with a maximum adsorption for the temperature of 1373 K. However, Vs is also affected by the shape and size of the non-linear molecules such that the Vs/Vs' ratio for benzene/cyclohexane couple is close to unity, for treatment temperatures ranging from 973 to 1173 K, while it raises to 210 for the sample obtained at 1573 K. In addition, the adsorption of 2,2-DMB is negligible in all the samples. Therefore, these carbon materials have a large capacity for VOC adsorption and even with high capacity of discrimination of adsorbates depending on the shape and size and on the preparation temperature of the samples. This behaviour of the sample prepared at 1573 K is explained by shrinkage of the carbonaceous structure which permits the adsorption of benzene and inhibits that of molecules larger than cyclohexane. The behaviour of these carbon materials is so versatile that the porous structure can be modified by different treatments. The structure of the porous network of the sample obtained at 1573 K can be progressively opened by mildly gasification at 1 and 6% of burnoff, and the new samples so obtained loose the capacity of discrimination for adsorption of the benzene/cyclohexane couple. The introduction of oxygen chemical groups in these samples by treatment with HNO3 diminishes the capacity of adsorption of hydrocarbons. This is a consequence of fixation of chemical groups at the entrance of the pores [10,54] This decrease is the largest for 2,2-DMB for which the adsorption is almost negligible due to their larger molecular size (0.60 nm).

424 30 180

E

20

.....,

-'~'140

E

o v r

>

E

10

n-pentane

>

100

n-butane 0

973

r"

'

'

'

1173

1373

1573

60

973

T(K)

,

,

,

1173

1373

1573

T (K)

Figure 15. The specyfic retention volumes of linear hydrocarbons versus the carbonization temperature of Saran.

Moreover, the effect produced by the oxygen functionalities can be enhanced if the chemical t r e a t m e n t is carried out with CS2, because in this case the chemical functionalities fixed at the entrance of the pores are sulphur compounds which are larger in size t h a n the oxygen compounds. This effect of the sulphur functionalities is shown in Figure 16. In this figure the separation ratios for the n-hexane/benzene, benzene/cyclohexane and n-hexane/cyclohexane couples are plotted versus the percentage of sulphur content in a sample obtained by carbonization of S a r a n at 1173 K and further t r e a t m e n t s with CS2 . A very high discrimination capacity is shown for 4.3% sulphur content for benzene/cyclohexane and n-hexane/cyclohexane couples, as a consequence of cyclohexane having a larger molecular size t h a n benzene and n-hexane. In addition, the n-hexane/benzene ratios are almost the same in all cases. This suggests t h a t the sulphur functionalities have partially closed the access to the interior of pores [22].

1800 1400

c c

1000 >

benzene/ cyclohexane

600 n-hexane/ benzene

200 -200 0

1

2

3

4

5

s(%)

Figure 16. Effect of the sulphur content on the discrimination capacity of sample S 1173.

425 Concerning the carbon materials obtained by polymerization and further carbonization of furfuryl alcohol the samples used were those whose adsorption behaviour at high relative pressures has been commented above. The following VOC were adsorbed at very low vapour concentration [55]: carbon tetrachloride (CCI4), chloroform (CHCla), dichloromethane (CH2C12), acetone (C3H60), diethyl ether (CnHloO), t e t r a h y d r o f u r a n e (C4H80, THF), n-alkanes (CnH2n+2, from C4 to C7), benzene (C6H6), cyclohexane (C6H12) and 2,2-DMB (C6H14). The change of In Vs for the adsorption of linear hydrocarbons with the n u m b e r of carbon atoms follows, as is usual, a straight line. Again, when the shape and size of the adsorbates are different, the porosity of the samples is a very i m p o r t a n t factor in determining the degree of the adsorption. This is shown in Figure 17 in which Vs values obtained at 473 K for the four carbons are plotted against the critical dimension of the adsorbates with the same n u m b e r of carbon atoms, but with different shapes and sizes. The plots obtained for the Vs values at the other t e m p e r a t u r e s are similar to these.

P1 ~E 4 m

o

P3 0

E o

P4

-A

r~ 2 ;>

P2 0

?

L~

0,4

.- ~

.....

9 -:43---

---4

0,45

0,5 0,55 0,6 d (nm) Figure 17. Specific retention volumes versus the molecular size, for the adsorption of organic molecules with six carbons atoms on glassy carbons (see Table 16 for molecular dimensions).

Sample P1 (Figure 17) can be seen to discriminate between the critical dimension of benzene and cyclohexane. This behaviour is the same as t h a t described above for the adsorption of these adsorbates m e a s u r e d at high relative pressures. One can conclude t h a t sample P1 has such a narrow distribution of micropores t h a t it behaves similarly for adsorption in a wide range of experimental conditions, i.e. from a very low relative pressure (at zero surface coverage) up to P/Po = 1, and from 303 K up to 533 K. The trend of Vs values for P4 is very similar to t h a t for P1. Moreover, P4 has the same discrimination, or molecular sieve effect, for the m i n i m u m critical size between benzene and cyclohexane. These data indicate t h a t P4 also has a very narrow distribution of micropores similar to t h a t of P1. P3 has a molecular sieve effect, but for a m i n i m u m critical size larger t h a n t h a t of P1 and P4. In this case this behaviour appears for the cyclohexane/2,2-DMB couple. Therefore, like for samples P1 and

426 P4, sample P3 has a similar molecular sieve behaviour as that observed for samples P1 and P4, but for larger molecules. The comparison of the adsorption of hydrocarbons to that of organic molecules capable of specific interactions can give useful information to determine the driving forces of the adsorption in each case. Once these are known it would be possible to design the properties of the carbon materials to adsorb these substances. The specific retention volumes for these molecules are compiled in Table 16.

Table 16 Specific retention volumes, Vs (cma/m 2) on glassy carbons at 473 K Vs(cm3/m 2) P4

Mean molecular size (nm)

P1

P2

P3

CC14

0.06

0.052

0.46

CHCla

0.77

0.074

1.39

0.44

0.59

CH2C12

0.66

0.112

1.13

0.51

0.57

CHaCOCH3

1.39

0.113

1.03

1.05

0.33

CH3CH2OCH2CH3

2.12

0.094

1.13

1.26

0.41

THF

0.92

0.I00

2.32

0.52

0.56

C6H14

3.29

0.302

3.84

0.84

0.405

C6H6

4.83

0.131

3.31

3.21

0.52

C6H12

0.07

0.054

2.65

0.05

0.56

2,2-DMB

0.05

0.052

0.20

0.04

0.60

0.64

Reprinted from: M. Domingo-Garcia et al. [55].

For P1 and P4 the polar molecules have Vs values higher t h a n that of cyclohexane although in many cases their minimum critical dimensions are larger or similar to that of cyclohexane. It is, therefore, likely that these molecules are adsorbed in part by chemical surface groups, i.e. by specific interactions. For P2 the values of Vs are much lower than for the other samples. This could be related to the pore constrictions in the microporosity and to the lack of macroporosity and the almost negligible mesoporosity of this sample, above discussed [55]. For P3 the trend of Vs is different to that of P1 and P4. Moreover for P3 a plot of Vs versus the mean critical size of all these molecules shows (Figure 18), with the exception of acetone and diethyl ether, a monotonical decrease of Vs as the molecular dimension increases.

427

o

C6HI4 .. C6H6 ~

t.,,,i

E

~

= THF H12

E

o > v

C3H60 9

"IxHCCI3

CaHIoO 9

CH2C129 ,

0,3

0,4

0,5

2 , 2 - D M B ~ 9 CC14

0,6

0,7

d (nm) Figure 18. Variation of Vs with the mean molecular size of the adsorbates on P3.

It is, thus, difficult to determine whether adsorption of the polar molecules is produced as a consequence of specific interactions or of the wider micropore system of this sample which allows adsorption to occur inside the micropores by a non-specific interaction. It could even result from a combined mechanism in the micropores and on the chemical surface groups. It is also a p p a r e n t t h a t the same molecular sieve behaviour found at high relative pressures for the chloroform/cyclohexane couple appears at very low vapour concentrations. The s t a n d a r d enthalpies of adsorption and the liquefaction heats of these molecules are compiled in Table 17. Comparing the -AHOA values of the hydrocarbons with their -AHL (liquefaction heats), it is noteworthy t h a t for those molecules which, according to the values of V~, can reach the microporosity (n-alkanes and benzene on P1 and P4, and n-alkanes, benzene and cyclohexane on P3) the s t a n d a r d enthalpy of adsorption is much higher in absolute value t h a n the liquefaction heat. The -AHOA values are even more t h a n twofold the -AHL value in m a n y cases. Similar results are always obtained for the adsorption of hydrocarbons from diluted atmospheres on porous carbon materials [9,10,13,15,56-66]. Thus, in all the above described cases the absolute values of -AHOA are much higher t h a n the liquefaction heats. The only cases in which the values of-AH~ and -AHL are very close is when the adsorption occurred in nonporous carbons (V3G and Pyrolitic, Acheson and Degussa graphites) or in the external surface of the porous carbons. If neither n-alkanes nor cyclohexane can be considered to be capable of specific interactions, these high values of-AH~ could be produced as a consequence of a very good fit of the molecules inside the pores [13,52,66], such t h a t the closer the size of the molecule and pore dimension the higher the absolute values of-AH~ This is a consequence of the, so called,

428 surface curvature effect (SCE) and of the proximity of the pore wall increasing the adsorption potential [55,67]. In the case of benzene it is generally accepted t h a t high absolute values of-AH~ are produced either because adsorption occurs in the slit-shaped pores with dimensions similar to the molecular size or possibly specific interactions could be taking place due to the existence of ~ electrons in this molecule [9,67,68-70]. In one of the carbon materials used (P2) the value of -AHOA for benzene is close to -AHL which means t h a t the specific contribution is probably very small.

Table 17 S t a n d a r d enthalpies of adsorption on glassy carbons -AH~176 P1

-AHL

P2

P3

2.9

3.6

36.9

CHC13

38.9

17.4

47.4

43.1

31.4

CH2C12

34.3

25.4

40.5

44.9

31.7

CH3COCH3

49.1

33.5

50.6

53.5

32.0

CH3CH2OCH2CH3

49.9

42.3

51.1

52.6

29.1

THF

44.3

23.4

59.5

49.1

29.8

n-C4H10

29.8

14.0

37.0

27.0

24.3

n-CsH12

45.3

28.5

50.7

52.4

27.6

n-C6H 14

65.2

42.1

65.7

75.9

31.7

n-CTH16

88.1

64.9

82.1

97.8

37.1

C6H6

53.5

36.3

58.8

64.1

34.1

C6H12

6.0

8.9

53.9

16.1

32.8

20.6

4.2

30.4

CC14

2,2-DMB 3.7 3.2 Reprinted from: M. Domingo-Garcia et al. [55].

P4

(kJ/mol) 31.9

With regards the polar molecules and the P1, P3 and P4 samples the -AH% values have a higher absolute value t h a n the liquefaction heats in all cases. This suggests t h a t for P1 and P4, in which Vs is not related to their molecular sizes, the adsorption of these molecules is mainly controlled by specific interactions. In P3, although the interaction could probably also be specific, an a p p a r e n t relationship between Vs and the molecular size has also been shown (Figure 18). Therefore, the high absolute values of-AH~ should be produced as a consequence of the good fit of the molecules in the pores similar to the behaviour observed for

429 hydrocarbons, although some specific contribution, i.e. a combined mechanism of two contributions: specific + non-specific, can not be excluded. W h a t is noteworthy from the comparison of-AH~ values of the hydrocarbons and of the other molecules is t h a t some of the former can be adsorbed more exothermically t h a n polar molecules capable of specific interactions. To u n d e r s t a n d these data one should bear in mind t h a t the classification of these two types of interactions was based [71] on materials which can be considered as basically non-porous. Consequently, the -AHOA values for the non-specific interactions were clearly lower in absolute value t h a n the specific ones and very close to the liquefaction heats because the former were produced on flat surfaces. However, the situation is different in the case of porous materials because adsorption can be produced on the external surface, on the chemical groups or inside the micropores of similar size to the adsorbate. The most plausible of the three possibilities is the latter since this is t h e r m o d y n a m i c a l l y favoured. It is worth noting t h a t the n u m b e r of molecules to be adsorbed is very low (zero surface coverage) hence the probability of finding pores of similar size to the molecule is very high. However, for polar molecules the interaction produced by the dipolar m o m e n t can not be excluded and consequently a combined mechanism of specific + non-specific adsorption should be considered. The data reported in Table 17 clearly indicate t h a t in porous m a t e r i a l s it is unsafe to deduce the type of adsorbate-adsorbent interaction exclusively on the basis of the -AH~ values. In other words, although there can be different driving forces in the adsorption, the non-specific interactions (see for instance -AH~ for C6H14) are not necessarily less energetic t h a n the specific ones, at least when adsorption is produced inside the micropores and at very low coverage. In the case of a combined (specific + non-specific interactions) mechanism, the specific component of the surface free energy can be determined [57,59,60,72,73]. The results are collected in Table 18 for all these molecules except for benzene and carbon tetrachloride because neither carbon tetrachloride nor benzene

Table 18 Specific component of the surface free energy -AGsp(kJ/mol) P1

P2

P3

P4

CHC13

1.4

1.3

2.3

2.5

CH2C12

6.2

5.4

6.3

6.8

14.0

8.0

10.3

13.2

CH3CH2OCH2CH3

8.0

3.5

3.8

8.4

THF

7.2

5.0

7.9

6.8

CH3COCH~

Reprinted from: M. Domingo-Garcia et al. [55].

430 produce specific interactions. This is u n d e r s t a n d a b l e for carbon tetrachloride which has no dipolar m o m e n t and can be considered a spherical molecule. Nevertheless, this behaviour is more unexpected for benzene for which specific interactions are normally expected, because of the unlocalized ~ electrons. Since this behaviour of benzene is repeated for the four samples, it can be concluded t h a t its adsorption is non-specific in all cases and the high absolute values of -AH~ (Table 17) can be explained because adsorption is produced in slit-shaped pores of a similar size as the molecule. This finding again supports the previous suggestion t h a t -AHOA is not a very useful criterion to establish w h e t h e r an interaction is specific or non-specific. Very low values of the specific component of the surface free energy can be observed for CHC13. This suggests t h a t although the adsorption m e c h a n i s m of this molecule is considered a combined mechanism, the porosity is much more i m p o r t a n t t h a n the chemical surface groups, because the specific contribution is never higher t h a n 15 % of the total value of the s t a n d a r d free energy [55]. On the other h a n d the specific contributions for the adsorption of acetone and diethyl ether r e p r e s e n t more t h a n 50% of the total value of the s t a n d a r d free energy for the former and around 40% for the latter. Thus, in the adsorption of these molecules the interactions with the chemical surface groups seem to be as i m p o r t a n t as the porosity of the samples which could explain why these molecules do not follow the general trend shown in Figure 18. The values for the other molecules are lower t h a n those of acetone and in most cases they represent less t h a n 30% of the total value. From a practical point of view, these findings are very i m p o r t a n t because they indicate t h a t the following points should be t a k e n into consideration when one a t t e m p t s to increase the capacity of adsorption of these molecules from diluted atmospheres: i) Slit-shaped pores are very convenient for the adsorption of benzene. They can be produced in carbon materials. ii) Porosity is more i m p o r t a n t for the adsorption of CHC13 t h a n the oxygen functionalities of the adsorbent. iii) The importance of the porosity and of the chemical functionalities is very similar for the adsorption of acetone and diethyl ether. From these data one can conclude t h a t these glassy carbons have very narrow micropores distributions which permit t h e m to behave similarly from zero surface coverage (very low vapour concentration) to high surface coverage. Moreover the criterion frequently used to discriminate between specific and non-specific interactions on the basis of the s t a n d a r d enthalpy of adsorption is not useful when microporous materials are used as adsorbent and when adsorption is carried out at zero surface coverage.

431 ACKNOWLEDGEMENTS

This work has been supported by the DGYCIT under project PB94-0754.

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Adsorption and its Applications in Industryand EnvironmentalProtection Studies in Surface Science and Catalysis,Vol. 120 A. Dabrowski(Editor) 9 1998Elsevier Science B.V. All rights reserved.

435

S e l e c t i v e a n d r e v e r s i b l e a d s o r b e n t s for n i t r i c o x i d e f r o m h o t combustion gases R. Long and R.T. Yang* Department of Chemical Engineering, The University of Michigan Ann Arbor, Michigan 48109-2136, USA

This report provides an updated review and discussion of all selective, reversible sorbents for adsorption of NOx from combustion gases. The sorbents must selectively adsorb NOx over other gas molecules that are also contained in combustion gases: SO2, H20, CO2, 09 and N2.

1. I N T R O D U C T I O N Removal of NOx from exhaust gases is a challenging problem which has been extensively studied worldwide in recent years. The NOx emission is a major cause for the formation of acid rain and for other environmental problems. Selective Catalytic Reduction (SCR) of NOx has been the most effective means for NOx abatement. For large power plants, V2OJTiO2 has been the main commercial catalyst for SCR with NH3 for stationary sources [1]. However, for relatively small scale combustors, such as diesel-fueled and gasoline-fueled engines in vehicles, the use of NH3- based SCR technologies is not practical because of the high cost and NH3 slip. The three way catalyst (Pt-Rh-Pd) is an effective catalyst for SCR (mainly by CO) used in automobiles under rich-burn conditions [2,3], but it suffers from severe loss of activity for NO reduction in the presence of excess oxygen, which is the prevalent condition for diesel or lean-burn gasoline engines. SCR of NOx with hydrocarbons under excess oxygen conditions has been actively studied by many groups most recently [4-8]. A large number of catalysts have been found to be active for these reactions, such as Cu, Fe, Co, Ce, Ga, and H exchanged zeolites, noble metals supported on 7-alumina, metal oxides, pillared clay, and so on. A summary of these catalysts has been published recently by Amiridis et al. [5]. Among them, Cu- and Co-ZSM-5 are the most intensively investigated; however, they are deactivated rapidly by moisture and S02 [6-8]. Noble metal based catalysts appear to be free from deactivation by H20 and SO2, but additional problems, such as narrow window of operation * Address all correspondenceto R.T.Yang

436 temperature, high selectivity for N20 formation and oxidation of SO2 to SO3, inhibit their application in industry [5]. So successful catalyst development is necessary before this technology becomes applicable to diesel or lean-burn gasoline vehicles. A promising alternative approach for the removal of NOx is NOx trapping, or adsorption/absorption of NOx. Adsorption is divided into physical and chemical adsorption. Physical adsorption (surface adsorption and micropore filling) is rapid and reversible, but is less selective for specific gas species. The amount of physical adsorption can greatly exceed a monolayer capacity. It usually occurs as a result of intermolecular forces, such as van der Waals forces and capillary condensation. The normal boiling points of NO and NO2 are 121 K and 294 K, respectively. Therefore, NO2 can be easily condensable on microporous solids by pore filling around room temperature, whereas NO is relatively more difficult. In the presence of oxygen, NO trap by pore filling can be facilitated by the formation of NO2. Chemisorption results from the interaction between the adsorbate molecule and the adsorption site, which is selective to specific gases. The amount of chemisorption is less than the monolayer capacity. Chemisorption can occur at low or high temperatures. In order to remove NOx efficiently from exhaust gases, a very specific sorbent is needed. The sorbent must be able to selectively adsorb NOx from oxygen-rich combustion gases which contain NOx, 02, H20, SO2, CO2 and N2. The desired temperature range for NOx trapping is 300-400~ although temperatures outside this range may be prevalent depending on the specific application. The sorption rates must be high, e.g., suitable for applications at space velocities > 3,0001/h. The sorption must be reversible either by increasing temperature or decreasing pressure, so a desorption stream concentrated in NOx can be obtained [9]. The concentrated stream can be recycled to the combustion zone for NO decomposition into N2. Alternatively, desorption/decomposition can be accomplished by injecting a reducing gas. Still another alternative, applicable to lean-burn engines, is to dope noble metals in the sorbent and to run the engine with pulses of rich-burn conditions, during which time the adsorbed NOx is decomposed into Ne [10,11]. There has been a long search for such a sorbent for NOx as reviewed recently [12,13]. The more promising sorbents have been supported transition metal oxides [14-18], ZSM-5 or MFI zeolites exchanged by Cu 2§ and other cations [19, 20], FeeO~ dispersed on activated carbon fibers (AFC) [21, 22], zeolite [23], Y-Ba-Cu-O [24, 25], mixed metal oxides [10, 26-28] and carbon [31]. The two most promising sorbents, in terms of both NOx capacity and rate of uptake, appear to be Mn-Zr (1:1 molar ratio) mixed oxides [27] and Ce-doped CuO/TiOe [28], reported recently. In this paper, the available literature on adsorption of nitric oxides is reviewed, and the development of NOx removal techniques through sorption on solid materials is discussed.

437 2. NOx A D S O R P T I O N AT NEAR AMBIENT T E M P E R A T U R E

Iwamoto and coworkers [19,20] studied the adsorption of NO on various metal ion-exchanged zeolites with a fixed bed adsorption apparatus. In the adsorption experiment, 1,000 - 2,000 ppm of NO in He was introduced in a stainless steel column containing the adsorbent. After each adsorption run, pure He was introduced into the column to desorb NO from the adsorbent. The amount of reversible adsorption (Qrev) and irreversible adsorption (Qirr) of NO measured at 273K on various cationexchanged MFI zeolites are summarized in Table 1.

Table 1 NO adsorption properties of various cation-exchanged MFI zeolites Amount of NO adsorbed/(cm3g -1) Adsorbent Content of cation/(wt%) Na-MFI(23.3)- 100 Ca-MFI(23.3)-54 Sr-MFI(23.3)- 105 Ba-MFI(23.3)-80 Mg-MFI(23.3)-46 Cu-MFI(23.3)-157 Ag-MFI(23.3)-90 Co-MFI(23.3)-90 Mn-MFI(23.3)- 127 Ni-MFI(23.3)-68 Zn-MFI(23.3)-96 Fe-MFI(23.3)-62 Cr-MFI(23.3)-41 Ce-MFI(23.3)-8 La-MFI(23.3)-7 H-MFI(23.3)-100

2.81 1.32 5.45 6.44 0.69 5.90 10.85 3.06 4.20 2.41 3.79 2.12 0.87 0.43 0.40 0.13

Qrev reversible 0.16(0.006) c 1.81(0.246) 2.71(0.195) 1.50(0.143) 0.69(0.109) 4.28(0.206) 3.38(0.150) 1.52(0.131) 1.19(0.069) 1.03(0.112) 1.01(0.078) 0.52(0.061) 0.38(0.101) 0.34(0.496) 0.25(0.388) 0.12(0.004)

Qirr irreversible 00.(0.000) c 1.56(0.212) 0.20(0.014) 1.44(0.137) 0.22(0.035) 14.90(0.716) 0.54(0.024) 19.69(1.693) 5.81(0.339) 6.64(0.727) 0.50(0.039) 3.08(0.362) 1.16(0.308) 0.34(0.496) 0.24(0.372) 0.32(0.011)

Adsorption time, 45 min; desorption time, 60 min; concentration of NO, 997 ppm; adsorption temperature, 273K; adsorbent weight, 0.5 g; flow rate, 100cm3 min 1. bConcentration of NO, 1,910ppm. cUnit, (NO molecules)'(cation) -1. Reprinted from: Zhang et al. [ 19]. a

438 The Qrev and Qirr changed significantly with the metal ion. For transition metal ion-exchanged zeolites, the values of Qirr were larger t h a n those of Qrev except for Zn-MFI and Ag-MFI. In contrast, Qrev was greater t h a n Qirr on alkaline earth metal ion-exchanged zeolites. The amount of reversible adsorption per cation decreased in the order Ca 2+ > S r 2+ > B a 2+ > M g 2+ The order of Qrev was: Transition Metal Ion - Alkaline Earth Metal Ion > Rare Earth Metal Ion - Alkali Metal Ion - H § Among these zeolites Cu-MFI and Co-MFI showed the largest Qrev and Qirr, respectively. In CuMFI, Qrev and Qirr were found to be proportional to the exchange amount of copper ion, but the Qrev and Qirr per copper ion were constant. IR spectroscopy indicated that most of the reversibly adsorbed NO was NO + adsorbed on Cu 2§ and that the irreversibly adsorbed NO was in the forms of NO +, nitrate (NO~), nitrite (NO~), and NO~. The amounts of reversible and irreversible adsorption of NO were also dependent on the zeolite structure. Table 2 and Table 3 show the results of copper ion-exchanged zeolites and silver ion-exchanged zeolites, respectively.

Table 2 Effect of zeolite structure on NO adsorbability of copper ion-exchanged zeolites

Adsorbent

Content of cation/(wt%)

Amount of adsorption of NO/(cm3.g-1) Reversible

Irreversible

Cu-MFI(23.3)-68

2.63

2.29(0.247) b

7.46(0.805) b

Cu-OFF/ERI(7.7)-81

5.45

2.28(0.146)

5.55(0.270)

Cu-MOR( 10.5)-76

5.26

2.11(0.114)

6.69(0.361 )

Cu-LTL(6.0)-34

3.22

1.23(0.108)

2.38(0.210)

Cu-FER(12.3)-66

3.89

1.42(0.104)

4.82(0.353)

Cu-FAU(2.6)-60

9.27

1.15(0.035)

0.62(0.019)

Cu-FAU(5.6)-83

7.99

0.86(0.031)

1.52(0.055)

Adsorption time 45 min; desorption time, 60 min; concentration of NO, 1,910 ppm; adsorption temperature, 273 K; adsorbent weight, 0.5 g; flow rate, 100 cm 3 min-1. bUnit, (NO molecules) (cation) -1. Reprinted from: Zhang et al. [ 19].

a

439 Table 3 NO adsorption properties of various silver ion-exchanged zeolites amount of adsorption Adsorbent Content of of NO (cma'g -1) ca tio n/(wt%)

Re versib 1e

Irreversible

Ag-MFI(23.3)-104 b

12.38

6.16(0.240 c)

4.11(0.160 c)

Ag-MFI(23.3)- 104

12.38

5.14(0.200)

4.99(0.194)

Ag-MOR(15.0)-112 b

17.09

5.76(0.162)

2.18(0.061)

Ag-MOR( 15.0-112

17.09

5.02(0.141 )

2.39(0.067)

Ag-FER( 12.3)-76

13.66

4.71(0.166)

1.51(0.053)

Ag-OFF/ERI(7.7)

16.71

1.12(0.032)

0.28(0.008)

Ag-LTL(6.0)-37

13.43

0.28(0.010)

0.16(0.006)

Ag-FAU(5.6)- 101

25.76

0.40(0.007)

0.56(0.010)

Ag-FAU(2.6)-98

37.10

4.12(0.053)

3.99(0.052)

Ag-LTA(2.0)- 103

41.22

3.38(0.039)

0.57(0.007)

5.90

4.28(0.206)

14.90(0.716)

Cu-MFI(23.3)-157 b

a Adsorption time, 45 min; desorption time, 60 min; adsorption temperature 273 K. bAdsorption time 60 min; desorption time 120 min. cUnit NO-molecule (Ag ion) -1. Reprinted from: Zhang et al. [20].

The amount of reversible adsorption and irreversible adsorption of NO per copper ion were found to decrease in the following order: MFI > OFF/ERI > MOR > LTL > FER > FAU This result is consistent with the increase in the aluminum content in the zeolites. From Table 3, one can see that Qrev also changed significantly with the zeolite structure among the silver ionexchanged zeolites. Ag-MFI and Ag-MOR showed the highest Qrev. Similar to the Cu-MFI sorbent, Qrev of Ag-MFI and Ag-MOR increased with the ion exchange level, and Qrev/Ag § of Ag-MFI was constant at different exchange levels of silver ion. The real exhaust gases also contain various gases such as NO2, 02, CO2, SO2, CO and H20. Iwamoto et al. [19] further studies their influence on NO adsorption properties in Cu-MFI zeolite. The effects of each gas is shown in Table 4.

440 Table 4 Effect of preadsorbed gases on the adsorption properties of Cu-MFI(23.3) - 147 a Amount of adsorption Preadsorbed gas b

of NO/(cm3-g -1) Reversible

Irreversible

NO2(4,680 ppm)/He

7.14

2.21

02(99.5%)

4.26

14.38

CO2(20%)/He

4.25

12.19

SO2(2,170 ppm)He

3.92

7.86

CO(1,890 ppm)/He

1.39

4.15

H20(3%)/He

0.22

0.45

None

4.35

17.83

Adsorption time, 60 min; desorption time, 120 min; concentration of NO, 1,000 ppm; adsorption temperature, 273 K; adsorbent weight, 0.5 g; flow rate, 100 cm 3 min -~. bThe adsorbent was heated at 773 K for 5 h under helium stream (50 cm3"min-~) before preadsorption treatment. After the preadsorption the sample was purged with helium at room temperature. Reprinted from: Zhang et al. [19]. a

The preadsorption of NO2 resulted in an enhancement of Qrev for NO, which was probably due to the result that the irreversibly adsorbed NO2 provided new sites for NO molecules to produce N203 [19]. When 02, CO2, or SO2 were preadsorbed, almost no change in Qrev for NO was found. CO and H20 poisoned the adsorbability of NO in the Cu-MFI zeolite. The Qirr for NO was also described by the adsorption of CO and H20. Kaneko et al. [21,22,32] prepared cz-FeOOH and Fe203 highly dispersed on ACFs (active carbon fibers), which have very high adsorption capacity for NO near room temperature. For instance, the amount of NO adsorption was about 160 mg/g on cz-FeOOH dispersed ACF at 303 K and 80 kPa NO pressure. This type of NO adsorption seems to have both chemisorption and physical adsorption characteristics. NO was mainly adsorbed by a micropore-filling mechanism, which was inferred by the measurement of pore volume using N2 adsorption on ACF-5 after exposure to NO. The highly dispersed cz-FeOOH particles on the ACF assisted the micropore filling of NO through their chemisorption action. A 13 X molecular sieve was also reported to have a high capacity for NO adsorption in the presence of 02 by micropore filling, the NO capacity reached 75 mg/g at 296 K [23,36]. However, HeO, SO2, COz, all have strong inhibiting effects on this sorbent.

441 Several other research groups also investigated the adsorption of NO on metal oxides. The results are summarized in Table 5.

Table 5 Amounts of NO adsorbed on various adsorbents near room t e m p e r a t u r e Adsorbent

Amount of NO adsorbed a (rag/g)

Ref.

SnO2

5

29

CeO2

5

30

NiO

1

15

Co304

6

15

CuO/7 A1203

36

15

NiO/7 A1203

36

15

Co304/~ A1203

7

15

Fe2OJ7 A1203

45

15

Fe2OJSiO2

5

16

Fe-Y zeolite

18

17

Fe304

22

18

Fe304

24

14

a- Fe203

9

14

Jaosites

3-10

33

10-20

34

4-6

35

13

12

a-FeOOH [3-, 7- FeOOH r

a At 13 kPa and room temperature. Reprinted from: Kaneko and Inouye [ 12].

It is noted t h a t Fe304 and (z-FeOOH showed the highest adsorption capacity for NO at room t e m p e r a t u r e among pure oxides. When oxides such as CuO, NiO, Fe203 were supported on 7-A1203, the adsorption of NO was high. This was due to the high surface area [12].

442

Q

NO A D S O R P T I O N AT HIGHER T E M P E R A T U R E S FROM COMBUSTION GASES

Some sorbents exhibited excellent adsorption capacities for NO at high temperatures [10,11,27,28]. These sorbents are attractive candidates for application to the automobile industry as well as the power industry. The main advantage is that the adsorbed NO can be directly converted to N2 on a catalyst (e.g., the three-way catalyst) under a "rich-burn" condition at the same room temperature. This can be accomplished by a cyclic operation [12]. Another advantage is that the sorbents have the high sorption rate at high temperature, which is suitable for application for the removal of NO from exhaust gases at a high space velocity.

3.1. S u p e r c o n d u c t i n g b a r i u m and y i t t r i u m - c o n t a i n i n g s o r b e n t s Misono et. al. [24] reported that NO and CO could be rapidly adsorbed into superconducting YBa2Cu30~. After pre-evacuation at 300~ the sample adsorbed approximately 2 mol/mol oxide for NO at the same temperature. The adsorbed NO molecules were almost completely desorbed when the temperature was increased to 400~ For these Y-containing oxides, Yamashita and coworkers found that the NO adsorptivity decreased according to the order [37]: YSr2Co3Ox > YBa4CosOx > YSr2Mn.3Ox > YSr2V3Ox. TPD and IR results showed that the adsorbed NO molecules were oxidized to NO~ by lattice oxygen. The adsorbed NO was desorbed as a mixture of NO/O2. Ba-Cu-O mixed oxides have also been reported by Arai et al. [38] to have a high adsorption capacity for NO/N02 at 200~ This adsorption reaction was accelerated by the presence of oxygen. XRD results indicated the formation of Ba(NO3)2/CuO. In the presence of 02, a large amount of NOx was liberated from the sample at temperatures above 500~ However, the NO adsorption capacity for this sorbent was completely vanished by the presence of 8% CO2 because of the formation of surface BaCO3.

3.2. Mixed m e t a l o x i d e s Since the sorbents containing Ba are easily deactivated by CO2, Arai and coworkers developed materials for NOx adsorption which did not contain rare earth and alkaline earth metals. Several mixed-oxide sorbents containing Mn and/or Zr are shown in Table 6 [27]. The uptake of NO in the presence of 02 or absence of 02 was measured at 200~ in a tubular reactor. The presence of 02 promoted the adsorption of NO on these oxides. The Mn-Fe, Mn-Zr, and Mn-Cu systems exhibited a high uptake of NO. The Mn-Zr oxide showed the highest uptake of NO both with and without 02. NO was hardly detected in the effluent gas from the fixed bed adsorption during the initial 60 minutes. After that, the concentration of NO at the outlet

443 gradually increased with time. The Mn-Zr ( 1:1 ) oxide did not show NO removal after 6 hours on-stream. The total a m o u n t of NO removal in 6 hours of operation was 0.133 mol-NO/mol-Zr. In order to compare the removal capacity m e a s u r e d at a fixed gas-phase NO concentration, a gravimetric analysis was also performed for the NO uptake. The a m o u n t of NO adsorption in Mn-Zr oxide (Mn/Zr = 1) after s a t u r a t i o n was 1.43 wt % of the original oxide, which corresponded to 0.047 mol-NO/mol-Zr. The NO adsorption a m o u n t m e a s u r e d by the gravimetric analysis was s o m e w h a t smaller t h a n t h a t from the t u b u l a r reactor by gas phase analysis. The difference from these two m e a s u r e m e n t s was likely due to the diffusion resistance. The gas was forced to pass through the oxide particles in the t u b u l a r reactor, which enhance the diffusion rates.

Table 6 NO removal by mixed oxides containing Mn and/or Zr NO removala(%) Oxide

0% 02

10% 02

MnOx'AlzOa

9.9

14.8

MnOx'Cr202

0

2.2

MnOx'CuO

0

10.2

MnOx-FezO3

11.9

37.4

MnOx'Mo03

5.9

15.2

MnOx-TiOz

0

25.0

MnOx ZrOz

100.0

100.0

ZrO2"AlzO3

0

8.7

ZrO2"CrzO3

0

14.3

ZrO2"CuO

2.7

27.0

ZrOz'FezO3

0.4

8.7

ZrOz'MoO3

0.3

2.4

ZrO2"TiOz

7.6

17.4

Note. Calcination temperature 450~

temperature 200~ W/F =1 g's.cm 3 aNO removal after 30 rain of use. Reprinted from: Eguchi et al. [27].

0.1 vol.% NO, 0 or 10% 02, He balance. Reaction

444 Arai et al. [27] also studied the sorption capacities of NO in the Mn-Zr oxides with different Mn/Zr molar ratios. The results are shown in Table 7.

Table 7 Capacity of Mn-Zr Oxides for NO Removal Sample

Capacity for NO removal mol/mol-Zr

Mn-Zr oxide (Mn/Zr = 5)

0.105

Mn-Zr oxide (Mn/Zr =1)

0.133

Mn-Zr oxide (Mn/Zr =1/5)

0.034

Mn-Zr oxide (Mn/Zr =1/9)

0.029

1 wt% Pt/Mn-Zr oxide (Mn/Zr =1)

0.058

1 wt% Rh/Mn-Zr oxide (Mn/Zr =1)

0.035

1 wt% Ru/Mn-Zr oxide (Mn/Zr =1)

0.042

1 wt% Pd/Mn-Zr oxide (Mn-Zr =1)

0.073

900 ppm NO, 10% 02, He balance. Reaction W/F -1 g s cm-3. Reprinted from: Eguchi et al. [27].

Note. Calcination temperature 450~

temperature 200~

The NO removal was 100% for every Mn-Zr oxide at the start of NO supply and then gradually decreased after 10 to 60 minutes of operation. The adsorption capacity of NO was the largest in the Mn-Zr oxide with Mn/Zr = 1. When decreasing or increasing the Mn/Zr ratio, the amount of NO adsorption decreased. Addition of noble metals also decreased the adsorption capacity of NO, as shown in Table 7. The XRD data indicated that the amorphous phase in the Mn-Zr oxides (Mn/Zr = 1), which had a large surface area, was especially active for NO removal. The adsorbed NO in Mn-Zr oxides was almost completely desorbed under an NO/O2/He atmosphere when the temperature was increased to 400~ indicating that the sorption and desorption was almost reversible. The amount of NO uptake in the Mn-Zr oxides was hardly affected by the presence of CO2 (10%). The NO removal in the first 130 minutes was only slightly affected by H20 and the total amount of uptake was enhanced with H20. The reason of the promoting effect of H20 is not understood [27]. 3.3. CuO - based

sorbents

More recently, the sorbents CuO/Ti02 and Ce-CuO/Ti02, which showed higher adsorption capacities than Mn-Zr oxides, have been prepared in our laboratory [28]. The CuO/Ti02 sample was prepared by using incipient wetness

445 impregnation with aqueous solution of Cu(NO3)2 on TiO2. The Ce dopant was added also by the incipient wetness procedure using aqueous cerium nitrate solution on the CuO/TiO2 sample. The adsorption/desorption experiments of NO were performed in a thermogravimetric analyzer, equipped with a p r o g r a m m e d t e m p e r a t u r e control unit. The results of NO2 uptake at 300~ on CuO/TiO2 and Ce-CuO/TiO2 sorbents are shown in Figure 1 and Figure 2, respectively.

...

o

o

< m

9

f

9 r~

[NO] = 2000 ppm =4%

o

.,..~

9 Adsorption at 300~ 9 Desorptlon at 450 C

9

9

< z •

2:

O

0~,,,,,,,,,,,,,,,,,,,,,,,I,,,,,,,I 0

20

40

60

80

100

L

120

140

160

Time (minute) Figure 1. Adsorption and desorption of NO2 in 5% CuO/TiO2. Desorption was achieved by heating to 450~ in 2 min in the same gas flow. Reprinted from: Li et al. [28]. It can be seen t h a t the m a x i m u m NO adsorption on CuO/TiO2 was 6 mg/g under the conditions NO = 2,000 ppm, 02 = 4%, balance = He. When CeO2 was added to CuO/TiO2, both the NO2 capacity and the uptake rate increased. The sorbent capacity was increased from approximately 6 to 7.7 mg/g, or approximately a 30% increase. The initial sorption rate was increased from 3.4 to 5.0 mg/g, both at t = 10 min, or a 50% increase. The chemisorption rate was clearly the controlling step in the uptake, and the CeO2 dopant substantially increased the chemisorption rate. Since the oxidation of NO is involved in the chemisorption of NO in the presence of 02, the significant increase in both chemisorption rate and sorption a m o u n t as a result of adding Ce to CuO/TiO2 was a t t r i b u t e d to the unique oxygen storage property as well as the redox

446 property of Ce. Results of desorption of NOx at 450~ over the two sorbents, in the same gas flow, are also shown in Figure 1 and Figure 2. Heating from 300~ to 450~ took approximately 2 minutes, during which time a small amount of NOx was desorbed. The results showed that rapid desorption was accomplished at 450~ The working capacity of the sorbent, i.e., the reversible amount, depends on the time of desorption. It is clear from Figure 1 and Figure 2 that well over 95% of the amount adsorbed was desorbed rapidly. 8

6

=z

4


Rb + > K + > Na + > Li + Ba 2+ > Sr 2+ > Ca 2+ > Mg2 + In case of strongly basic anion exchangers with quaternary ammonium groups, the charge of anion exchanger affects the affinity of the anion exchanger as in the case of polystyrenesulphonic cation exchangers. The affinity decreases in the following order:

499 C104- > SCN- > I- > HSO4- > NO3- > Br- > CN- > NO2- > C1- > OH- > FIn case of ion exchangers of different functional groups the problem is more complicated due to the interaction of counterions, with ion exchanger functional groups resulting in formation of ion-pairs and even of covalent bonds. A typical example is very strong affinity of the anion exchangers of the functional groups of primary-, secondary- and tertiary amines for OH- ions contrary to the strongly basic anion exchangers for which the OH- ion is situated almost in the end of the series. The order of affinity of other ions for weakly basic anion exchangers is analogous to strongly basic ion exchangers [10,28-31]. Carboxylic ion exchangers are characterized by strong affinity for the H § ion contrary to strongly acidic cation exchangers [10,28-31]. On carboxylic ion exchangers the affinity series for alkali and alkaline earth metals are as follows: Li § > Na § > K § > Rb § Cs § Mg2 § > Ca2+ > Sr 2§ > Ba2+ that is contrary to those observed for polystyrenesulphonic cation exchangers. Carboxylic ion exchangers are characterized by strong affinity for alkaline earth metals and some transition elements which can be referred to the behaviour of aliphatic or aromatic carboxylic acids forming sparingly soluble salts with these ions. The first resin of the complexing functional groups was synthesized by Shogseid fifty years ago [32]. It was the ion exchanger with dipicrylamine built in the matrix and exhibiting selectivity for potassium. Shogseid used this ion exchanger to remove potassium from the sea water. Based on the solubility series of dipicrylamine with alkali metals, one can expect that the ion exchanger of this type should possess stronger affinity for rubidium and cesium ions. At present there are known several scores of various ion exchangers with complexing groups built in the matrix and exhibiting selectivity for various ions or their groups. Table 1 includes well known and commonly applied both on laboratory and commercial scales selective ion exchangers.

Table 1 Selective ion exchangers Functional group type

Selective for ions

Alamine Amidoxime

U(IV, VI) [34-45], Au(III), Ru(III),

Aminoguanidine 2-amine-4-ketothiazole- 3-acetate

Rh(III), Pd(II, IV), Pt(IV), Ir(III, IV), Cu(II), Fe(III), Cd(II), Hg(II) [46-53] Au(I) [64,65] Ag(I), Hg(II) [66]

Cu(II) [33]

500 Table 1 Selective ion exchangers Functional group type Aminophosphonic

2-amine-4-ketothiazole-3-acetate Anthranilate Arsenazo Arsenic Chitosan Chromotropic Cysteine Diphenylcarbazide Diphenylcarbazone 1,8-dihydroxynaphtalene-0,0-diacetate Dimethylglyoxime Dipicrylamine Dithiocarbamate Dithizone Fluorene Formazane Phosphine Phosphone

Phosphate Glyoxal-bis-2-mercaptoaniline Guanidine 8-hydroxyquinoline o-hydroxyphenylazobenzoate Hydrazide Hydroxamate Imidazole

Selective for ions Sc(III) [30,54], Th(IV) [30,33,55], U(IV,VI) [30,33,56-58], Pb(II),Cu(II), Zn(II), Fe(III) [59-60], Ga(III), In(III) [62-63] Ag(I), Hg(II) [66] Cu(II), Co(II), Ni(II), Fe(II), Zn(II) [67,68] U(VI), Th(IV), Zr(IV), Hf(IV), Ln(III) [33,69] U(VI), Th(IV), Zr(IV), Hf(IV), Ln(III) [30,33] Cu(II), Ni(II), Co(II), Ga(III), Ln(III), Nd(III), Eu(III) [70] Ti(IV) [71], Fe(III), Cu(II), Zn(II)[33] Ag(I), Pt(II), Au(III), Hg(II) [33] Cr(VI) [72-75] Ag(I), Hg(II) [76] Zr(IV), Hf(IV) [77] Ni(II) [78] K(I) [32] Hg(II), Ag(I), Cu(II), Se(IV), Cr(VI), Pb(II) [5,33,79-85] Au(III), Pt(II, IV), Pd(II), Ir(III, IV), Ru(III), Rh(III), Ag(I), Hg(II) [86-88] Sc(III) [89] Pd(II), Ir(IV), Pt(II, IV), Au(III), Ag(I), Hg(II), U(VI)[33,90] Wh(IV), U(VI), Fe(III) [10,33,91] Sc(III) [30,33,54], Th(IV), U(IV, VI) [30,33,44,92,93], Zr(IV), Hf(IV) [94], Fe(III), Zn(II)[30,33,95-97], Ga(III), In(III) [30,33,62,97], Be(II)[30,33] Wh(IV), U(VI) [30,33,98,99] Hg(II) [100] Pd(II), Pt(IV), Au(I) [33,64,65] Cu(II), Fe(III), V(V) [33] Be(II), AI(III), Fe(III) [33] Ag(I), Hg(II) [33] U(VI), Fe(III) [33,85,101-103] Au(I) [104-106]

501 Table 1 Selective ion exchangers Functional group type Iminodiacetate

Crown or cryptands 8-mercaptoquinoline Mercaptide n-methylglucamine Morine Nitriltriacetate r Nitrosorezorcinole Oxime Piridinedicarboxylate Porphyrine Pyrocatechol-0,0-diacetate Rhodamine Salicylate Sarcosinate Tannine Thiazoline Thioglycolate Thiophosphorate Thiourea Thionalide Triazolethiol 2,4,6-triamine- 1,3,5-triazyne

Selective for ions Cu(II), Pb(II), Ni(II) [10,30,33,107,110], Cd(II) [30,33,60,107], Fe(III) [10,30,33,62,107-111], Ln(III) [30,33,107,112-114], Sc(III) [30,33,54], Th(IV) [99], U(VI) [33,98,107], In(III), Ga(III) [30,33,62,115-117], V(IV, V) [118], Se(IV) [119] K(I), Ba(II) [30,33,85] Pd(II), Au(III), Ag(I) [30,33,85] Ag(I), Hg(II), Au(III) [30,33,85] B(III) [33] Hf (IV), Zr(IV), Fe(III) [120] Cu(II), Ln(III) [33] Co(II) [33,107] Fe(III), Cu(II), Co(II), Hg(II)[33] Cu(II), Mo(VI) [33] Cu(II), Ca(II)[33] Ag(I), Cu(II), Co(II), Fe(II)[30,33,84] Zr(IV), Cu(II) [121] Ag(I), Au(III), Pd(II), Pt(IV) [122] U(VI), Fe(III), Cu(II), Hg(II) [10,33,107,123] Cu(II) [33,107] Ge(IV) [33] Hg(II) [33,85] Ag(I), Au(III), Hg(II), Bi(III), Sn(IV) [30,33,85] Hg(II), Pg(II), Cu(II)[33] Ag(I), Au(III), Pt(II, IV), Pd(II), Ru(III), Hg(II) [10,30,33,84,85,107,124], Cd(II), Cu(II), Co(II), Ni(II), Zn(II) [126], As(III, V) [127-128], Sb(III, V) [129,130], Bi(III) [131], Pd(II) [125] Hg(II), Ag(I), Cu(II), Cd(II)[132] Cu(II), Hg(II), Cd(II), Ag(I), P(V), As(V), Cr(VI) [33]

The detailed discussion of physicochemical properties and applications of individual selective ion exchangers presented in Table 1 can be found in many

502 papers [5,8-11,29,30,32-132]. Of various well known chelating ion exchangers only a few types are produced on the commercial scale. These ion exchangers are included in Table 2.

Table 2 Main commercial selective ion exchangers Type

Resin Name

Manufacturer

Amidoxime

Chelite N

Serva, Feinbiochemica GmbH and Co., Germany Rohm and Haas Company, France Not available former USSR

Aminophosphonate

Dithiocarbamate Hydroxyoxime Iminodiacetate

Duolite ES-346 ANKF-2G ANKF-3G ANKF-221 AEF-1 AEF-2 AEF-3 Chelite P Duolite ES-467 Duolite ES-469 Lewatit OC- 1060 Purolite S-940 Purolite S-950 Relite MAC-7 Wofatit ME-55 Nisso ALM-525 Dowex XF-4196 Amberlite IRC-718 ANKB-1 ANKB- 10 ANKB-35 ANKB-50 Chelex 100 Chelite C Diaion CR- 10 Dowex A-1 Dowex XF-4045 Duolite ES-466 Lewatit TP-207 Lewatit TP-208

Serva Feibiochemica GmbH and Co., Germany Rohm and Haas Company, France Bayer, Germany Purolite International Ltd., United Kingdom Resindion, Italy Bayer, Germany Nippon Soda Co. Ltd, Japan Dow Chemical Co., USA Rohm and Haas Company, France Not available former USSR

Bio-Rad, USA; Merck, Germany Serva Feibiochemica GmbH and Co., Germany Mitsubishi Chemical Industries Ltd, Japan Dow Chemical Co., USA Rohm and Haas Company, France Bayer, Germany

503 Table 2 Main commercial selective ion exchangers Type Resin Name Muromac A- 1 Purolite S-930

Isothiouronium

Relite MAC-5 Unicellex UR- 10 Varion CH Wofatit MC-50 Wofatit MC-59 Wofatit Y-66 Duolite ES-345 Ionac SR-3 Ionac SRXL Lewatit TP-214 Monivex Purolite S-920 Relite MAC-3 Srafion NMRR

Picolylamines Poliethylenepolyamine

Polietylenepolyimine Pyridine Thiol

Dowex XF-4195 Dowex XF-43084 Diaion CR-20

Manufacturer Muromachi, Japan Purolite International Ltd., United Kingdom Resindion, Italy Unitika Ltd, Japan Nitrokemia Ipartelepek, Hungary Bayer, Germany Rohm and Haas Company, France Sybron Chemicals Incorporated, USA Bayer, Germany Ayalon Water Conditioning Company Ltd., Israel Purolite International Ltd., U.K. Residion, Italy Ayalon Water Conditioning Ltd., Israel Dow Chemical Co., USA

Mitsubishi Chemical Industries Ltd., Japan Sumichelate MC-10 Sumimoto Chemical Co., Ltd, Japan Mitsubishi Chemical Industries Diaion CR-40 Ltd, Japan Sumimoto Chemical Co., Ltd, Sumichelate CR-2 Japan Serva Feinbiochemica GmbH and Chelite S Co., Germany Rohm and Haas, France Duolite ES-465 Imac TMR Imac GT-73

The selective ion exchangers are produced in the form of granules, membranes, discs, fibres, ion-exchange or ionited paper. As a relatively small number of selective ion exchangers is available commercially, many authors suggest modification of strongly basic anion

504 exchangers by means of aromatic chelating sulpho-derivative reagents like Alizarin S [143-146] Arsenazo I [145], Arsenazo III [145], Beryllon II [145], Bismuthon II [147], Bromopyrogallol Red [143-145], Chromazurol S [145], Chromotrope 2R [148], Chromotropic Acid [144,145], Ferron [144,145], NitrosoR-salt [144,145,149], Pyrocatechol Violet [145], SPADNS [145], 5-Sulphosalicylic acid [145], Tiron [145], Thoron [145], Xylenol Orange [145,150] etc. [151-156]. Complexing organic reagents used for modification transform the anion exchanger into the selective ion exchanger. According to Brajter [143,144] the modified anion exchanger can be characterized by greater selectivity for definite ions or their groups than the corresponding ion exchanger with the analogous functional group built into the matrix. In most cases the separation processes carried out on these ion exchangers are of analytical (separation of micro or miligram amounts, sorption of microgram amounts - trace analysis) or physicochemical character. 2.

S E L E C T I V E REMOVAL OF ION I M P U R I T I E S

It is often sufficient to remove only one type of metal ions from the mixture of impurities contained in sewages. Owing to the ion exchange it is possible to replace undesirable ions with others which are harmless for the natural environment. 2.1. Gold a n d p l a t i n u m m e t a l s Noble metal recovery is a typical example for application of this method. The main aim is total recovery of noble metals but not the reuse of the water which is of less economic importance as the volume of the produced sewages is relatively small. Some used up plating baths contain gold in the form of the anion complex [Au(CN)2-]. Both strongly and weakly basic anion exchangers can be applied to remove this anion from sewages [5,85,105,157]. To remove the complex [Au(CN)2-] in the whole pH range there can be applied strongly basic anion exchangers possessing quaternary ammonium groups:

(R4N+)2SO42- + 2[Au(CN)2]- ~

2R4N+[Au(CN)2]- + 3042-

(1)

The working exchange capacities strongly basic anion exchangers for the complex [Au(CN)2-] are low and equal 55-110 g Au/dm 3 [5,85,105,157]. It is difficult to strip the complex [Au(CN)2-] from the anion exchanger bed by means of alkaline metal hydroxide solutions. Due to high prices of gold, the value of that adsorbed on the anion exchanger exceeds the price of the ion exchanger, ion exchanger combustion at 1000~ in order to recover metal is quite common. Also the complex [Zn(CN)4] 2-, acidified thiourea solution or potassium thiocyanate solution in an organic solvent, can be used for desorption of gold(I) from the anion exchanger bed. The complex [Zn(CN)4] 2- characterized by stronger

505 affinity for the ion exchanger than [Au(CN)2]-is effective for gold recovery from the anion exchanger bed. The regeneration process is as follows: 2R4N+[Au(CN)2] - + [Zn(CN)4]2 ~

(R4N+)2[Zn(CN)4]2- + 2[Au(CN)2]-

(2)

To remove gold from effluents cementation by means of e.g. lead can be used [158]. The cementation process is as follows: 2Au[CS(NHD2]C1 + Pb ~

Pb[CS(NHD2]nC12 + 2Au

(3)

where most probably n equals 4. For desorption of [Au(CN)2]-from the anion exchanger potassium thiocyanate solutions in the water-organic solvent mixtures were used as eluents [159]. The most frequently used mixtures are: water-DMF, water-dimethylacetamide, water-acetone, water-N-methyl-2-pyrrolidine, water-DMSO, water-tetrahydrofurane, water-hexamethylphosphoramide. The 5 M KSCN solution in 50% v/v DMF in water is the most frequently used eluent. The sorption mechanism of the complex [Au(CN)2]- on the weakly basic anion exchanger can be presented in the following way: (R3NH+)2SO4 + 2[Au(CN)2]- ~

R3NH§

- + SO42

(4)

Weakly basic anion exchangers are selective for [Au(CN)2]-, though sorption of cyanide complexes of silver(I), copper(II), iron(II, III), nickel(II) and others can be taken into account [5,85,105,157]. Applicability of some available weakly basic anion exchangers was examined. It proved that the most effective weakly basic anion exchangers are those whose pKa values of functional groups are from 9 to 11. These anion exchangers achieve the maximal sorption of the complex [Au(CN)2]- in the pH range 7-9 which allows for avoiding the problems connected with HCN removal. Regeneration of the weakly basic anion exchanger with deposited gold(I) cyanide complexes is carried out by means of diluted alkaline metal hydroxide solutions. The regeneration process can be presented by means of the following equation: (R3NH+)2SO4 + 2[Au(CN)2]- ~

R3NH+[Au(CN)2] - + 3042.

(5)

The two-column system was applied to remove gold from sewages containing 4-5 mg Au/dm 3 after the plating process [160]. In the first column there was the carboxylic cation exchanger in the H § form (Diaion WK-20) which was probably used to remove other heavy metal ions and to decrease pH. In the second one there was the weakly basic macroporous anion exchanger Lewatit MP-62 on which the gold anion complex adsorbed. The NaCN alkaline solution was used for regeneration of the anion exchanger bed. The regeneration process gave almost 100% gold recovery.

506 In another procedure the post-plating solution containing, besides the gold anion complex, organic acids and their salts was passed through the weakly basic anion exchanger at pH equal about 3 in the three-column system [161]. Gold was adsorbed in the first of them and other substances in the second and third ones. Moreover, selective ion exchangers of various types are used for selective removal of noble metals from sewages. Thus the ion exchanger with the cysteine groups is characterized by high ion-exchange affinity for Au(III) and Pt(II). The ion-exchange capacities for these ions are as follows: Au(III)=l.22 M/kg and Pt(II)=0.39 M/kg [162]. The selective ion exchanger with hydroxamate groups is characterized by great ion-exchange capacity for Au(III) which is 4.0 M Au/kg (pH=l). This sorbent is recommended for selective separation of Au(III) from Cu(II), Ag(I) and Fe(III) and also for removal of Au(III) from sea water and KCN solutions [163]. The ion exchanger with the isonitrosoacetamide groups is selective for Pd(II) [164]. The sorbent of the azoimidazole functional groups is used for sorption of noble metal ions from the solutions of copper(II), nickel(II), cobalt(II), iron(III), aluminum(III) etc. salts [33]. It possesses very great ionexchange capacity for gold which is equal to 660 g Au/kg and is applied for removal of noble metals from the solutions coming from copper and nickel hydrometallurgy as well as from past plating sewages [33]. The sorbent based on polystyreneazorhodamine is stable up to 100~ both in neutral and acidic solutions. Its sorption capacity for the selected noble metal ions in the 1 M HC1 solution is: Pt(II)-19.5 g/kg, Pd(II)-22.2 g/kg and Au(III)-197 g/kg. This ion exchanger is characterized by great selectivity for noble metal ions [33]. The sorbent based on polystyrenazo-8-mercaptoquinoline is stable in the acidic and neutral media. It exhibits great ion-exchange capacity for Pd(II) (which is 300 g Pd(II)/kg in the 1 M HC1 solution) and great selectivity for noble metal ions in the presence of Cu(II), Ni(II), Co(II), Fe(III) etc. [33]. Great selectivity for the complex [Au(CN)2]-is characteristic for the chelating ion exchanger of the guanidine and aminoguanidine functional groups built into various polymer matrices [33,64,65]. It proved that modified hydrophobic copolymers crosslinked with DVB of the expanded structure are the most effective for the sorption of [Au(CN)2]- ions. The ion exchanger of formazane functional groups exhibits great selectivity for gold, silver, platinum metals and mercury ions. Its ion-exchange capacity in 0.01 M HC1 is following: Au(III)0.9 M/kg, Pd(II)-0.75 M/kg, Pt(II, IV)-0.54 M/kg, Ir(IV)-0.43 M/kg,Rh(III)0.1 M/kg and for Ru(III)-0.03 M/kg. This sorbent is recommended for selective removal and separation of noble metal ions as well as for separation of Pd(II) from Ni(II) and Co(II) [33,90,165,166]. Moreover, the selective ion exchangers of dithiocarbamate, thiamide and thioglycole functional groups exhibit great selectivity for gold and platinum metals [30,33,85]. The results of the studies on synthesis and physicochemical properties of the ion exchangers of the dithizone and dehydrodithizone functional groups bonded to the polystyrene matrix by the sulphur atom are very interesting. The ion exchanger of the dithizone functional groups exhibits particularly strong ion

507 exchange affinity for Pd(II), Pt(II) and Au(III). The values of distribution coefficients IQ for Pd(II), Pt(II) and Au(III) in the 0.01-6.0 M HC1 solutions are very high and equal 104-106. This ion-exchanger can be applied both for separation of platinum metals and selective separation of noble metal ions from others [86-88]. Complexes of gold(III) chlorides can be removed from various systems on the sorbents possessing weak ester groups e.g. on Amberlite XAD-7. In this case two types of sorption mechanism can be distinguished [10,167,168]: - solvation: R - CO2 + [nuCl4]- ~ R - CO2[AuC13] + C1-ion-exchange: R - C 0 2 + H20 ~ R - C O O H + + OH-

(6) (7)

R - C O O H + + [AuC14]- ~

(s)

R-COOH+[AuCI4] -

The mixture of hydrochloric acid and acetone can be used for gold(III) desorption. The resins containing thiourea or its derivatives built into matrix exhibit great selectivity for noble metal ions [5,10,30,33,85]. The commercially available ion exchanger selective for noble metals and mercury ions of the isothiourea functional groups (Srafion NMRR) exhibits differentiated ion-exchange capacity for gold and platinum family. Total capacity of this ion exchanger for individual noble metal ions is as follows: Au-340 g/dm ~, Pt-200 g/dm 3, Pd-120 g/dm 3, Ir-120 g/dm ~ and Rh-80 g/dm 3. However, the working capacity of Srafion NMRR for gold ions is small (55-110 g Au/dm3). As the value of the gold retained in the bed exceeds significantly that of the ion exchager used for its recovery, combustion of the ion exchanger at 1000~ is a common procedure, though the 5% NaOH + 5% NaCN solution is also used for its regeneration. The selective ion exchangers of isothiourea groups are used, among others, for separation of platinum metals from other metals. Depending on pH of the medium these goups are of the following forms:

P ~ C H ~ - - S ~ J NH

..

"

P~CH~--S~C

\NH 2 (1)

/NH2 (~ X \NH 2

(9)

(2)

When these groups occur in form (1), they create coordination bonds with metal cations and in form (2) the platinum metal anion complexes are bonded according to the exchange mechanism e.g. [PdC14]2-.

508 NH

2 P ~ CH~-- S--C./ @ 2X

+

_

[ PdC] ]24

\NH 2

(10) / NH P ~ c g T- S - - C \ | NH

H2N \

[PdCJ] 2

H N 2

Srafion NMRR is used for the group removal of noble metals from ores, meteorites, moon, sand, steel, sewages, biological materials etc.[10,33,84,85,124]. 2.2. S i l v e r Silver is another precious metal commonly applied in plating and photographic industries. In plating processes it occurs as the cyanide complex [Ag(CN)2]- with various additions. The ion-exchange methods are recommended for silver recovery from large volumes of sewages of low concentrations of the element being recovered, particularly from the washings. The first technological system for this process was installed in Germany at the beginning of the forties. The silver contained in the washings formed the cyanide complexes which sorbed on the weakly basic anion exchanger in the sulphate form. The NaOH and H2S04 solutions were used successively for the ion exchanger regeneration. The sorption capacity of the ion exchanger was relatively small and was equal to 8 g/dm 3 [169]. Some anion exchangers of various types were applied for the selective removal of silver(I) from cyanide solutions by Riveros [170]. It proved that the anion exchangers like Duolite A-7, Dowex WGR-2, Dowex MWA-1, Dowex XFS-4195 and Amberlite IRA-35 extract silver(I) in large amounts only at pH8 the same ion exchangers extract silver(I) from the solution to a small extent or they do not extract it at all. The comparison of the above mentioned ion exchangers indicates that Duolite A-7 and Dowex MWA-1 possess the highest capacity and effectiveness. Strongly basic anion exchangers like Dowex MSA-1 and Amberlite IRA-910 extract silver(I) in the whole range of pH 4-12. Moreover, it was stated that if the solution under consideration contains only one kind of metal ions, then the given ion exchangers show the highest extraction degree for silver(I). However, if in the feeding cyanide solution the metal ion mixture is present, the percentage of silver(I) removal decreases which depends on the phase contact time. The extent of silver(I) extraction can be increased decreasing the concentation of free cyanides in the solution and also by stopping the extraction before the ion-exchanger reaches the highest saturation [170]. In 1968 Kunin suggested using the strongly basic anion exchanger (type 1) to recover silver(I) being in the form ([Ag($203)2] m from the photographic industry sewages. It was Amberlite IRA-400 in the chloride form regenerated with 1 M

509 hydrochloric acid solution [169]. The studies carried out in the USA in the end of seventies showed that the weakly basic acryldivinylbenzene anion exchanger Amberlite IRA-68 is characterized by better qualities than Amberlite IRA-400 [171,172]. In the first cycles of sorption Amberlite IRA-400 is much better but Amberlite IRA-68 shows greater resistance to impurities. However, its application requires maintaining suitably low values of washings pH (about 5.5). Of many solutions used for silver desorption, 30% (NH4)2S203 solution proved to be the best for regeneration. To recover silver(I) from the post-regeneration solution, the electrolysis was used yielding 98% recovery of silver(I) of 99% purity. The solution remainig after electrolysis was used for another regeneration. The redox ion exchangers were also used for selective removal of silver(I) ions from the photographic sewages. At first the ion exchangers of the polythiomethylenestyrene type were used which enabled achieving exceptionally high capacities (about 22 mmol/g which constitutes about 240% bonded silver in relation to the ion exchanger mass) as for the ion-exchange method. Regeneration was carried out using the solutions Na2SO3 or NaHSO~ [173]. Much better results were obtained using hydroquinoformaldehyde resins of the capacity about 38 mmol/g that is about 400% bonded silver [174]. In this case the ion exchanger was regenerated electrolytically which made it possible to reactivate about 85% initial bed capacity. During the electrolysis silver deposited as pure metal on the cathode after passing into the solution. The sorption effect was 90-98% initial content of silver. Many interesting studies were made using the Russian ion exchanger redox EO-7 which is a polycondensat of sulphonated hydroquinone and phenol with formaldehyde [175]. By its use separation of silver(I) from copper(II), lead(II), zinc(II), bismuth(III), nickel(II), cobalt(II) and aluminium(III) was made [176]. In this case the redox exchange proved to be more selective than the ion-exchange method. Moreover, it enables removal of other components, even trace amounts of silver from the polyionic solutions of high concentrations [177]. Hydroquinone-resorcin-formaldehyde resins were effectively applied in the reactions of various metal ions reduction, among others, of silver(I) ions both in the acidic and neutral media [178]. Sulphonated 2-vinylanthraquinone, styrene and DVB copolymers were used, among others, for selective removal of silver from nitrate solutions [179]. 2.3. Tin a n d c o b a l t Another element used in plating is tin. The sewages containing this element can be purified using a two-column system. One of them is packed with strongly acidic cation exchanger and another one with weakly basic anion exchanger. Tin is retained in the second column and washed with the NaOH solution [5]. Cobalt ions can be recovered from sewages using strongly acidic cation exchangers [5]. Owing to the application of 2 M HNO3 solution for regeneration of the ion exchanger bed in the effluent the cobalt ion concentration reaches about

510 25 g/dm 3. To remove cobalt(II) ions from sewages produced during petrochemical wastes combustion chelating ion exchangers were applied [5]. The chelating ion exchangers of the functional iminodiacetic groups in the sodium form adsorb cobalt(II) ions selectively. The ion-exchange capacity of these ion exchangers for Co 2+ ions is 57 g/kg. In addition, the ion exchange is applied for removal of vanadium from the sewages produced during preparation of the zirconium-vanadium pigment [5]. As well as for recovery of molibdenium ions from sewages using a strongly basic anion exchanger [5].

2.4. Mercury Mercury and its compounds are among the chemical impurities imposing the great t h r e a t for the natural environment due to their ability of translocation in the environment and accumulation in living organisms. Mercury compounds attack the nervous system in organisms. Therefore they are considered to be the most dangerous substances. Its harmful effect consists in inhibiting the protein synthesis in cells. This process is caused by the presence of thiol groups in proteins with great affinity for mercury. Mercury, both metallic and in the form of compounds is widely applied in industry. It is used, among others, in chemical (electrolysis of NaC1 by the mercury method, production of acetic aldehyde, production of some pesticides and plastics), cellulose-paper, pharmaceutical and electronic industries. These branches of industry produce most sewages containing mercury compounds. In the fifties in the area close to M i n a m a t a Bay ( Japanese Sea), serious poisoning with mercury compounds-mainly with methyl mercury took place. Three and half thousand people were poisoned, of whom over 100 died. It was found out that the tragedy was caused by the nearby plastic producing factory which used mercury compounds as catalysts. Post-production sewages with a relatively high content of mercury were dumped into the bay. It has been estimated that only in the period 1951-1970 the factory dumped from 200 to 600 tons of mercury into M i n a m a t a Bay [180]. In the sewages mercury occurs in the forms of metallic, methyl mercury or most frequently of soluble compounds, among others, as undissociated molecules, Hg 2§ and Hg2 e§ ions as well as complex ions. All mercury compounds contained in waters make its self-purification process difficult due to inhibition of biochemical processes. To remove mercury compounds from sewages there are used, among others, reduction, precipitation, extraction and ion-exchange methods [5,10,85,181-196]. Of the above mentioned methods, the ion-exchange is of significant importance because they are technologically simple and enable efficient removal of even trace impurities from solutions. They are particularly useful when it is necessary to treat large volumes of diluted solutions. It is possible to remove mercury(II) from sewages on various types of ion exchangers e.g. strongly acidic cation exchangers, weakly and strongly basic anion exchangers as well as on selective ion exchangers of various types [5,10,85,181-196]. Many West European countries

511 apply the industrial method of mercury ions removal from sewages based on the licence of the Dutch firm Akzo Zout Chemie using the selective ion exchanger Imac TMR of functional thiol groups. This ion exchanger is used for selective sorption of mercury(II) ions from technological solutions, mainly from brines in case of their electrolysis using the mercury method and also from sewages. Imac TMR is a styrenedivinylbenzene copolymer of a macroporous structure. It possesses mainly functional thiol groups and a n u m b e r of sulfone groups [5,182,183,187,191,192]. Owing to the presence o f - S H groups, strong affinity of this cation exchanger for mercury(II) ions can be accounted for by the ability of mercury(II) cations reaction with mercaptans, thiophenols or hydrogen sulfide. Its capacity for mercury (II) is 240 g Hg/dm 3 ion exchanger. Comparing effectiveness of various methods used for mercury removal it was stated t h a t using reduction methods it is possible to decrease the mercury content to 1-3 ppm in sewages, by precipitation of HgS with hydrogen chloride to 1 ppm and applying the ion exchanger Imac TMR to 0.5-5.0 ppb. [5,10,85,182,183,187, 191,192]. Despite of the fact t h a t in the concentrated brines, mercury(II) occurs mainly in the form of the complex ions HgC142-, Imac TMR reacts mainly with Hg 2§ and HgC1 § ions which are in the equilibrium in the solution [5,182,183,191,192]: Hg 2§ + 4Cl-r

HgC1 § + 3C1-r

HgC12 + 2C1- az HgCla- + Cl-r

HgC142- (11)

However, Imac TMR does not react with metallic mercury which can occur in the dispersed form in brines. In this case, metallic mercury should be oxidized with chlorine and then the solution should be deprived of the oxidizer excess before introducing on the column containing Imac TMR as the thiol ion exchanger is readily oxidized according to the reaction: 2 R - S H + oxidizer --> R - S - S - R + 2H § R - S - S - R + oxidizer-~ x R - SOH + yRSO2H + 2 R - SOaH

(12)

(~3)

Where R is the styrenedivinylbenzene copolymer thus losing its precious properties of selective ion exchange. The former reaction can proceed reversibly due to resin reactivation by means of reducing agents but the l a t t e r one gives the ion exchangers of different functional groups which are not selective for mercury (II) ions. Mercury(II) ions sorption on the ion exchanger Imac TMR can be described by m e a n s of the following reactions: 2 R - S H + Hg 2+ -~ R - S - H g - S - R + 2H + R - S H + HgC1 + -~ R - S - H g C 1 + H+

(14)

(15)

512 After the mercury(II) ions sorption, the ion exchanger Imac TMR can be regenerated by means of the concentrated hydrochloric acid solution according to the following reactions: R-S-Hg-S-R + 2HC1 -> 2 R - S H + HgC12 (16) R-S-HgC1 + HC1 -~ R - S H + HgC12 (17) Besides the discussed ion exchanger Imac TMR, the following ion exchangers: Duolite ES-465 (Dia-Prosim, France), Chelite S (Serva, Germany) and Duolite GT-73 (Rohm and Haas, France) possess thiol groups. Of them at present only Duolite GT-73 (Imac GT-73) is produced on a commercial scale [197]. The first industrial system using Imac TMR for mercury ion sorption of 15 m~/h effectiveness was introduced in Delfizji, Holand in 1973. At present the system of 2-60 m3/h effectiveness are in the use in many countries e.g. in Cuba. Akzo Zout Chemie is generally considered to be a leading firm as far as removal of mercury(II) ions from sewages is concerned. Another group of ion exchangers exhibiting great selectivity for mercury(II) are those of isothiourea functional groups [5,10,85] like Ionac SR-3, Imac SRXL (Ionac, USA), Lewatit TP-214 (Bayer, Germany), Monivex (Ayalon, Israel), Purolite S-920 (Purolite, UK), Srafion NMRR (Ayalon, Israel), Sumichelate Q-10 (Sumitomo, Japan), Relite MAC-3 (Residion, USA). Srafion NMRR is most commonly applied for mercury(II) removal from sewages. Its sorption capacity for mercury(II) is 545 g Hg/kg ion-exchanger. 5% thiourea solution containing 0.22% HC1 is used for desorption of mercury (II) from its bed. The ion exchangers of the thiocarbamate functional groups are characterized by strong affinity for heavy metal ions. They allow for the reduction of the ion level from 20-30 ppm to the amount below 1 ppb. Of this type ion exchangers, Nisso ALM-125 ( Society of synthetic Organic Chemistry, Japan) is most commonly applied for selective mercury (II) removal from sewages. Its sorption capacity for mercury (II) is 680 g Hg/kg ion exchanger in the aqueous solution but 340 g Hg/kg ion exchanger in 10% H2SO4 solution. In the continuous process it decreases the mercury (II) content in sewages to 0.1 ppb. The sodium sulphide solution is used for mercury (II) desorption from its bed. In case of the ion exchanger Nisso ALM-126 it was stated that its sorption capacity for mercury(II) ions can be increased by the increase of operating temperature to maximum 50~ However, the bed regeneration is no longer possible. Mercury can be recovered as vapour by the ion exchanger calcination. The ion exchanger with mercury(II) deposited on it can be also mixed with cement in order for its immobilization and storage in such a form. Moreover, it was stated that the macroreticular polystyrene-based resins with the functional aminothiazole, iminothiazole or thiazoline groups exhibit a high selectivity for mercury(II). A thiazoline resin column has been used to concentrate mercury(II) from the sea water adjusted to pH=l with hydrochloric acid. Maximum sorption capacity for mercury(II) was found to be 2.8 M/kg. The

513 sorbed mercury(II) is recovered quantitatively by elution with 5% thiourea containing 0.1 M HC1 [36]. The results of studies on removal of mercury(II) ions from sewages on the anion exchangers modified with complexing organic sulpho-derivative reagents being prototypes of new selective ion exchangers are of significant importance. Immobilization of this type compounds in the ion exchanger phase is possible owing to the affinity of these reagents for anion exchangers. This results from both their aromatic structure and the presence of sulfonate groups [143,144]. For example, strongly basic anion exchangers in the azothiopyrine sulfonate [151], dithizone sulfonate [154] forms and Chromotrope 2R [148], can be effectively applied for removal of mercury(II) ions trace amounts from waters and sewages. Differentiation in mercury(II) ions retention compared with other metal ions on the obtained selective ion exchangers due to their modification is a result of: - affinity of mercury(II) ions for atoms in a ligand, - conditions of proper complexes formation, - stability of the complexes with mercury(II) ions in the ion exchanger phase. These ions exchangers owing to their high selectivity for mercury(II) ions trace amounts can be applied in mercury monitoring in the natural environment. 2.5.

Lead

Lead as a serious pollutant appears from the following sources [3,5]: - waste sludges from petroleum refineries, - waste sludges from the manufacture of alkyl lead compounds, waste solvent-based paint sludges and paint residues, - waste sludges from the manufacture of lead acid batteries, - solvent and water washes from the painting ink production. High toxicity of lead requires that its contents in sewages should be reduced to a minimum (ppb level). For this aim chelating cation exchangers with functional phosphonic or aminophosphonic groups are used [3,5,59,60]. They are selective for lead(II) ions. For example the values of lead(II) distribution coefficients for the aminophosphonic ion exchanger K-AMF depending on the value of pH of the outer solution are as follows: Kd=122 (for pH=2.5) and Kd=1900 (for pH=5) [59,60]. However, the results concerning the kinetics of lead(II) sorption are not very interesting. Weakly basic anion exchangers in the free base or sulphate form can be also applied for selective removal of lead(II) chloride complexes from the solutions of 4-6 pH [3,5]. A cation exchange process combined with precipitation is frequently used for lead removal from sewages [200]. The studies by Dudzinska and Pawlowski [201-204] concerning simultaneous removal from the aqueous solutions of lead(II) and cadmium(II) as well as organic ligands i.e. aminopolycarboxylic acids mainly EDTA in one process on anion exchangers of various types are of significant importance. The anion exchangers Amberlite applied by them differed in basicity of functional groups-

514 strongly basic types 1 and 2 as well as weakly basic ones, matrix porosity: microand macroporous as well as matrix structure: polystyrenedivinylbenzene and polyacrylic copolymers. It was stated that the anion exchangers of functional weakly basic groups exhibit stronger affinity for lead(II) and cadmium(II) complexes with EDTA than strongly basic anion exchangers and that the anion exchangers of a polyacrylic matrix exhibit stronger affinity for the above mentioned complexes t h a n those of the same type with a polystyrene matrix. The authors determined working capacities of selected anion exchangers for lead(II) and cadmium(II) complexes in the presence of various anions. Relatively high values of ion exchange capacity for the above mentioned complexes, their great affinity for the anion exchanger, as well as effective and economical regeneration by means of 1 M NaC1 solution make it possible to use this method in technology of waters and sewages. 2.6. C o p p e r Copper ions due to their toxicity disturb the operation of biological t r e a t m e n t plants and exert a negative effect on self-purification of water and on some organisms living in water reservoirs. Therefore these copper ions should be removed from wastewater by precipitation, electrolysis, ion-exchange and extraction. A very important example of ion exchange application is recovery of copper(II) from sewages formed during leaching which is one of copper production stages. Due to small pH (below 2) of sewages, conventional chelating ion exchangers of functional iminodiacetate and aminophosphonic groups do not practically adsorb copper(II) ions. Special chelating ion exchangers characterized by much greater affinity for copper(II) than for other metal ions were synthesized [5,10,205-208]. These ion exchangers are obtained from N-(hydroxyalkyl)picoliamines. One of them i.e. Dowex XFS-4195 can remove copper(II) even from 1.5 M H2SO4 solution but 5 M sulfuric acid is required for regeneration. Using another chelating ion exchanger Dowex XFS-4196, copper(II) ions are removed from less acidic solutions (pH over 1.5). Dowex XFS-4196 can be easily regenerated by means of sulfuric acid of the concentration 100 g H2SOjdm 3. The post-regeneration liquid contains 33 g Cu2+/dm 3 and 40 g H2SO4/dm 3 and can be applied for copper production using the electrolytic method. Moreover, there was obtained the chelating ion exchanger Dowex XFS-43084 of physicochemical properties similar to those of Dowex XFS-4196 but of stronger affinity for copper ions t h a n for iron ions [5,10,205-208]. Liquid chelating ion exchangers i.e. hydroxyoximes are applied for removal of copper(II) from mine waters, copper ores coming from the mines in P a n g u a n a in New Guinea, Bougainville Copper Ltd [209]. In Germany Copper(II) was also recovered by means of extraction with hydroxyoximes on the commercial scale from the mine waters as well as from the pyrometallurgical processes [210]. Such a recovery was economically justified and financial costs comparable to those of cementation and sorption processes on solid ion exchangers.

515 Copper recovery from the aqueous solutions obtained after adsorption of evaporated off gases originating from chloride calcination of pyrite ashes was proposed [211]. After solution neutralization to p H = l by means of calcium hydroxide and filtering off the sediment, copper(II) was removed using the extraction with hydroxyoximes. Gonczarowa et al. [212] used the ion exchange amphoteric fibres to remove copper(II) ions from acidic sewages. Their structure was based on partially hydrolized polyacrylnitrile and polyethyleneimine. The greatest ion exchange capacity was obtained when the solution pH was increased to 4.5. Ion exchangers of various types were applied for selective removal of copper ions from the ammonium sewages [213-215]. There was studied copper(II) sorption from the ammonium sewages on the following ion exchangers: phenolsulphone, polystyrenesulphone (microporous and macroporous), carboxylic based on acrylic, methacrylic and phenolcarboxylic acids, polyphenol type (Duolite S-30), thiole, chelating ion exchangers of functional iminodiacetate and amidoxime groups as well as on weakly basic anion exchangers of various types. Of the examined ion exchangers, polimerization carboxylic ion exchangers proved to be the most effective for selective removal of copper(II) ([Cu(NH3)4] 2§ from the sewages containing ammonium. It was stated that the maximum concentration of copper(II) 50-70 g Cu(II)/dm 3 in the effluent was obtained during regeneration of these ion exchangers with 2M sulfuric acid solution. Some sewages contain a few grams of ammonium sulphate in a litre and a small number of copper(II) ions. These sewages can be used as fertilizers but copper(II) ions should be removed earlier. The ion exchange method makes it possible to remove copper(II) sulphate solution which can be reused [216]. Carboxylic ion exchangers are most frequntly used for this aim. 2.7. N i c k e l a n d v a n a d i u m Utilization of sewages produced during sulphur removal is another significant example of ion exchange. The contain, among others, nickel and vanadium which can be recovered. Sewages are oxidized, their pH is established from 5 to 10 and then they are passed through the chelating ion exchanger bed on which nickel ions are adsorbed. The solution pH is smaller than 5 after this operation. Then the solution is passed through the chelating ion exchanger bed to adsorb vanadium ions [5,10,216]. Regeneration of beds leads to obtaining concentrated solutions of nickel and vanadium ions which are a source of useful side products. The most serious environmental problem concerning nickel(II) occurs in the metal t r e a t m e n t industry. Nickel is frequently used in the sulphate form. Thus for its removal from sewages there are applied polystyrenesulphone cation exchangers and for their regeneration sulfuric acid. It was stated that under the optimal conditions of cation exchanger bed regeneration with the sulfuric acid solution, a nickel concentration in the effluent can be several scores grams/dm 3. However, ammonium carbonate is used for regeneration of ion exchangers

516 applied in nickel recovery from sewages originating from the nickel refinery [5,10,216]. Polymetallic sewages require chelating ion exchangers e.g. nickel(II) can be selectively removed from ammonium molybdate on the aminophosphonic ion exchanger (Russian ion exchanger ANKF-80). It was stated that its sorption capacity for nickel(If) ions is 19 times as large as that of the conventional cation exchanger. Moreover, chelating ion exchangers are applied in the process of nickel ions selective recovery from the sewages originating from nickel plating and nickel compounds production [5,10,216]. Carboxylic ion exchangers also exhibit great selectivity for nickel(II) ions. Halle et al. [217] applied the macroporous carboxylic cation exchanger Wofatit CA-20 in the sodium form for selective nickel(II) ions removal from washings formed during the nickel plating. For the separation of nickel(II) from metallurgical waste solutions the highest separation efficiency was obtained with an inorganic ion exchanger with sodium titanate, Na4Ti9020 [218,219], which is a layered material, the exchangeable sodium ions being located in the titanium oxide layers. The exchanger takes up nickel(II) at pH values over 5 very efficiently. Column experiments with three types of nickel-containing waste solutions from a metal plating plant showed good performance for sodium titanate in the purification of real waste solution. Metal loadings obtained varied in the range of 0.65 meq/g to 2.7 meq/g and nickel(II) level in the effluent prior to the breakthrough was very low 0.0030.009%. 2.8. C h r o m i u m Removal of chromium(II, VI) from waste waters is absolutely necessary due to toxicity of its compounds [5,10]. Chromium(VI) compounds are believed to be particularly toxic, though all chromium compounds are belived to be carcinogenic. A m a i n source of w a t e r contamination with chromium are sewages from the surface metal t r e a t m e n t plants and from tannery. Sewages coming from baths used for chromium plating, metallic coating passivation and washing already coated surfaces are characterized by low pH caused by the presence of free acids, brown-yellow colour and small transparence. Chromium(VI) concentration in such sewages ranges 5 to 50 mg Cr(VI)/dm 3 (during the periodical change of plating baths it reaches even 200 mg Cr(VI)/dm3). Purification of such a solution consists in passing it through the strongly acidic polystyrenesulphonic cation exchanger and then through strongly basic anion exchanger. Metal ions and other cationic impurities are removed on the cation exchanger and chromates are removed on the anion exchanger: nRSO3H + M + ~-> nRM n+ + nH+ 2ROH + CrO42. ~ R2Cr04 "~- 2 0 H R2CrO4 + Cr042 + H + R2Cr207 + OH-

(18)

(~9) (20)

517 The NaOH solutions are applied for anion exchanger regeneration. To remove chromium(III) from the used up tannery bath (composed of Cr(III)8000 mg/dm 3, Mg(II)-1700 mg/dm ~, Ca(II)-950 mg/dm 3, Na(I)-37000 mg/dm 3, SO42-40000 mg/dm 3, C1-15000 mg/dm 3, HCOO-12000 mg/dm 3 and pH=3.5) there can be applied the polystyrenesulphone cation exchanger Amberlite IR-120. The ion exchange capacity of Amberlite IR-120 does not depend on pH in the range from 1 to 11 and is equal to 35 mg/g. The used up tannery bath contains large number of various ions. Besides chromium(III) ions there are a large quantities of sodium cations. Studies of adsorption selectivity on the cation exchanger Amberlite IR-120 in the H § form showed that the above mentioned cations compete with one another in the process of binding with the ion exchanger functional groups, whereby the ion-exchange capacity of the cation exchanger decreases rapidly with the increase of Na(I) ion concentration in the solution. When the concentration of Na(I) in the solution is about 25 g/dm ~ then the ion exchange capacity of Amberlite IR-120H § decreases to about 5 mg Cr(III)/g, i.e. it decreases by 85% compared with the initial value. The problems of separation selectivity of Na(I) and Cr(III) were overcome using the differences in oxidation reduction properties of both cations. It was shown that the presence of the strong oxidizer which are Cr2072- ions does not cause the loss of ion-exchange capacity by the cation exchanger. The method proposed for removal of chromium(III) from the used up t a n n e r y bath consists of four stages. In the first stage chromium(III) is oxidixed to chromium(VI) by means of sodium persulphate. The obtained solution is passed through the column with Amberlite IR-120 in order to remove sodium, calcium and magnesium ions. Then Cr(VI) is reduced to Cr(III) in the effluent by means of methanol. The last i.e. the fourth stage consists in adsorption of chromium(III) ions on the cation exchanger Amberlite IR-120 which is regenerated with 2 M H2SO4 solution resulting in practically total desorption of chromium(III) [220]. The process of chromium(III) recovery from sewages known as IERECHROM [221] worked out by Petruzzeli et al. is of significant importance. Macroporous carboxylic cation exchangers (Purolite C-106) which, besides chromium ions, retain also trace amounts of other metal ions, among others, of aluminum and iron are used in this process. The regeneration process of the carboxylic ion exchanger with deposited chromium ions proceeds in stages. In the first stage the alkaline H202 solution (pH=12) is used for desorption. As a result the anion forms (chromates and aluminates) being created are quantitatively washed away from the cation exchanger and separated. In the second stage which is purification of the bed in order to wash away iron ions. As a result, iron and aluminum sulphates are formed which are returned into the process as flocculating factors. The remaining chromate solution is then used in the plating industry or in the same t a n n e r y process after reduction of chromium(VI) ions to Cr(III). The ion exchanger of the polyvinylebenzene matrix was synthesized. 1-(4pyridinyl)-2-(1-piperidinyl)ethyl ester of 4-aminobenzoic acid was bonded to this

518 matrix. The ion exchanger is characterized by high selectivity for chromium(VI) ions [222]. The best exchanger for all chemical forms of chromium was found to be a fibrous exchanger FIBAN AK-22 [223]. This selective exchanger contains both carboxylic and imidazole functionalities on polypropylene fibres. This means that the selective exchanger acts both as a cation and an anion-exchanger as well as chelating ion exchanger. At pH=3 Fiban AK-22 removed more than 99.3% of both Cr 3§ and Cr2072 and in the pH range of 5-8 more than 99.6% of CrO42 from 5 ppm Cr solutions. Fiban AK-22 columns took up efficiently chromium from real waste effluents from a metal plating plant. From a lmM solution with pH=6.1 the level of chromium in the effluent was only 0.006% prior to breakthrough. The loading was, however, rather modest 0.4 mmol/g. Liquid anion exchangers of the tertiary amine type can be applied for selective removal of chromium(VI) from the acidic sewages [5,10]. The ion exchangers of this type are produced, among others, by the firms Henkel Co, USA (Alamina 304, Alamina 336), Hoechs, Germany (Hostarex A-324, Hostarex A-327) and others. The mechanisms of Cr2072- ions removal by these ion exchangers is as follows: 2R3N + Cr2072 + 2H § --~ (R3NH)2Cr207

(21)

The liquid ion exchanger and its combination with Cr2072- ions are soluble in kerosene and other organic solvents but in water only in trace amounts. It follows from the above that Cr2072 ions can be extracted from large volumes of the aqueous solution to a small amount of the organic phase by means of such an anion exchanger dissolved in kerosene or in another solvent. Removal of Cr(VI) ions from the organic phase is not difficult. The mechanism of liquid anion exchanger regeneration (removal of Cr2072- ions from the organic phase) is as follows: (RaNH)2Cr207 + 4NaOH -> R3N + Na2CrO4 + 3H20

(22)

Chromium(VI) in the alkaline medium passes back into the aqueous phase. In this way there is obtained a concentrate of the post-regeneration solution Na2CrO4 of the content about 20 g Cr(VI)/dm 3 and the organic solution of liquid anion exchanger is suitable for a direct reuse. Extraction in the liquid-liquid system by means of the liquid cation exchangers like D2EHPA (di(2-ethylhexyl)-phosphoric acid) and Cyanex 272 (di(2,4,4-trimethylpentyl)phosphinic acid) was used for chromium(III) removal from the aqueous solutions of the composition corresponding to the used up tannery bath. The preliminary studies showed that the solutions D2EHPA or Cyanex 272 in kerosene do not extract chromium(III) from the aqueous phase of pH2, extraction is very effective and quick remaining below 2 ppm Zn(II) in the raffinate. Iron is the most serious competitor for zinc(II) in the extraction process and can be washed away from the extractant using NaOH. The Valberg process became depreciated within 4-5 years [228].

521 Lately dialkylthiophosphonic acids are of great interest (Cyanex 301 and Cyanex 302) of the following formulae:

R\fs

R\

/ \ R

s

. . R\

/ \ SH

Cyanex 301 pKa=2.61

R

/P OH

R

\SH

Cyanex 302 pKa=5.63

where R is 2,4,4-trimethylpentyl. Liquid cation exchangers like Cyanex 301 and Cyanex 302 will probably replace D2EHPA used for the extractive removal of zinc(II) ions from the weakly acidic sewages formed during the artificial silk production [229]. Zinc salts are also present in the sewages originating from the ceramic industry. A strongly acidic cation exchanger in the sodium form can be applied for their removal [3,5,10]. Regeneration is carried out by means of the NaC1 solution of the ion exchanger concentration 300 g/dm ~. The ion-exchange capacity of the cation exchanger for zinc(II) ions is 32-40 g Zn(II)/dm~. Zinc(II) removal from the solutions used in cooling towers is also troublsome. Then cation exchangers can be also used [3,5,10]. Scott et. al. Suggest application of phosphonic ion exchangers (Duolite ES-63 or Duolite TSAP-40) for selective adsorption of zinc(II) ions from cooling waters. Strongly basic anion exchangers in the chloride form are used for selective adsorption of zinc(II) chloride complexes [ZnCI4]2-. Pure ZnC12 can be obtained washing the used up anion exchanger with water [3,5,10]. The outline of the process makes it possible to isolate zinc(II) from the postregeneration liquid containing the metal ions mixture. A mixture of chlorides is obtained as a result of regeneration of strongly acidic cation exchanger (used to remove metal ions from mixed plating baths). From the obtained postregeneration liquid [ZnCI4] 2 is selectively removed on the anion exchanger. Anion exchangers are used to recover zinc chloride from digesting solutions used in the steel industry.

3. NEW TRENDS IN ION-EXCHANGE METHOD DEVELOPMENT Though removal of metal ions from the given sewage is not difficult from the economical point of view, the applicable methods are those which make the process economical. The same refers to ion-exchange as a method for the sewage components reuse. In this case effective separation of sewages into concentrated salts and water is often too expensive, particularly when the amounts of sewages are large. Thus, besides new technological systems, the efficiency of unit operations, as well as synthesis of new or improvement of physicochemical properties of already known ion exchangers are important.

522 Keeping this in mind, the scientist from CSIRO [5] worked out magnetic ion exchangers which have many advantages. Their commercial names are Sirotherm (e.g. Sirotherm IR-20 contains weak acid and weak base groups) and Siromag (e.g. Siromag 17-strong acid resin, sulphonic acid type and Siromag 57strong base resin, quaternary ammonium type 1) and they are produced by ICI Operations Pty Ltd., Australia. Application of small ion exchanger grain size (about 200 pm.) caused significant improvement of adsorption rate. As the process rate increases, the same effects can be achieved using a smaller number of ion exchangers in smaller systems. Magnetized ion exchangers can be readily separated from the waters purified from metal ions and translocated which creates perfect conditions for their applications in the flow contact apparatus. In this type of apparatus a continuous transport of the ion exchanger was possible and which, in most existing processes, has a periodical character. Very simple devices from the fluidal counter-current bed to the parallel current tube reactors can be applied to work with magnetic ion exchangers. Studies for removal of various elements particularly of uranium(VI) (about 3 pg/dm 3) from the sea water on the ion exchangers of various types are of special importance [34-45]. In case of the amidoxime ion exchanger [36,45] kinetics of uranium sorption and desorption was studied. Of various type ion exchangers tested for uranium(VI) removal from the sea water, the chelating ion exchanger of functional amidoxime groups Duolite ES-346 (3.6 mg U3Os/g) possesses the greatest working capacity for [U02(C03)3] 4-. The working capacity constitutes about 3% total capacity for uranium(VI) and is several times greater than that of hydrated titanium dioxide used in the English and Japanese experimental systems [36,45]. Contrary to hydrated titanium dioxide Duolite ES-346 is characterized by high mechanical resistance as proved by over six month tests of work in the system fluidal bed using the sea water. According to the preliminary economical analysis, the price of i kg U3Os recovered from sea water by sorption is about 5-10 times higher due to great amounts of uranium in the sea water (about 1000 times greater than the ground resources) its recovery on Duolite E S 346 or on another selective sorbent should be expected [36,45]. Barnes et al. applied the synthesized amidoxime and polydithiocarbamide ion exchangers together with the ISP or ASA method to determine trace amounts of several elements in the sea water, fresh water, biological materials, urea, serum, bones, peritoneum dialysis solutions, D-glucose, high purity graphite as well as NaC1, Li2CO~, H3BO~ [47,80-83]. The chelating amidoxime ion exchanger proved to be very effective for removal of metal ions trace amounts [47-48]. Of new selective ion exchangers Diphonix resin is very important. It is a polyfunctional ion exchanger of the structure:

523

0 HO~ lip/OH ~CH2"'-cH~CH2"cH~CH~::~ 2.~

SO3H

H~CH2~CH~

SO3H

It is characterized by high selectivity for some metal ions including lead(II). The properties of Diphonix resin are a result of introducing the additional functional groups into the matrix of the resin. Examples of modified Diphonix resins are the already mentioned Diphonix-A resin containing the same geminally substituted diphosphonic acid groups bonded to a styrenie-based polymer matrix as the regular Diphonix resin, plus strong base anion exchange groups such as the tetraalkylammonium (Diphonix-A, type 1 resin) or the quaternized pyridinium (Diphonix-A, type 2 resin) groups [230]. In our opinion the future development of these ion exchangers will make it possible to decrease ion-exchange costs as a method for recovery and reuse of the sewage components.

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530 192. R. Bogoczek, E. Kociolek-Balawejder and A. Kogut, Przem. Chem., 68 (1989) 83. 193. F.H. Wang and T. Jung, Huan Ching Pao Hu, 12 (1989) 118; C.A. 113 (1990) 197063d. 194. H. Yoshida and T. Kataoka, AIChE.J., 35 (1989) 318. 195. H. Yoshida and T. Kataoka, Ind. Eng. Chem. Res., 29 (1990) 2152. 196. K.A. Larson and J.M. Wiencek, Ind. Eng. Chem. Res., 31 (1992) 2712. 197. Duolite GT-73 Complexing Resin, Rohm and Haas, Paris, 1995. 198. Anon. Chem. Eng., 82 (1975) 49. 199. A. Sugil, N. Ogawa and H. Hashizume, Talanta, 27 (1980) 623. 200. K. Lee and J. Hong, AIChE J., 41 (1995) 2653. 201. M.R. Dudzinska, Ph.D. Dissertation, M. Curie-Sklodowska University, Lublin, 1992. 202. M.R. Dudzinska and D.A. Clifford, React. Polym., 16 (1991/1992) 71. 203. M.R. Dudzinska and L. Pawlowski, Chemistry for Protection of the Environment 1991, Proc.VIII Int. Conf., L. Pawlowski and W.J. Lacy (eds.), New York, 1992. 204. M.R. Dudzinska and L. Pawlowski, Anion Exchange Removal of Heavy Metal-EDTA Complex, in: Ion Exchange Processes: Advances and Applications, A. Dyer, M.J. Hudson and P.A. Williams (eds.), Royal Society of Chemistry, Cambridge, 1993. 205. K.J. Jones and R.R. Grinstead, Chem. Ind., (1977) 637. 206. R. Grinstead, W.A. Nasutavicus and R.M. Wheaton, New Selective Ion Exchange Resins for Copper and Nickel, in: Extractive Metallurgy of Copper, J.C. Yannopoulos and J.C. Agarwal (eds.), vol.2, New York, 1976. 207. R.R. Grinstead, J. Metals, 31 (1979) 13. 208. J. Melling and D.W. West, A Comparative Study of Some Chelating Ion Exchange Resins for Applications in Hydrometallurgy, in: Ion Exchange Technology, D. Naden and M. Streat (eds.), Ellis Horwood Limited, Chichester, 1984. 209. J. Szymanowski, Hydroxyoximes and Copper Hydrometallurgy, CRC Press, Roca Baton, 1993. 210. G. Barthel, Tech. Mitt. Krup. Werkber, 35 (1977) 73. 211. G.M. Ritcey, B.H. Lucas and K.T. Price, Extraction of Copper and Zinc from Chloride Leach Liquors Resulting from Chlorination Roast-Leach of TineGrained Sulphides, Proc. Int. Solvent Extr. Conf. Liege, 3 (1980) 80. 212. N.A. Goncharova, N.A. Strukova, I.M. Smirnova, E.G. Mabarakshin and L.V. Emets, Z. Prikl. Khim, 55 (1982) 2095. 213. Z. Hubicki and S. Jusiak, Materials Science, 4 (1978) 17. 214. Z. Hubicki, Rudy Metale, 30 (1985) 338. 215. Z. Hubicki and L. Pawlowski, Environment Protection Engineering, 12 (1986) 5. 216. B. Bolto and L. Pawlowski, Przem. Chem., 64 (1985) 569.

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Adsorption and its Applications in Industry and Environmental Protection Studies in Surface Science and Catalysis, Vol. 120 A. Dabrowski (Editor) 9 1998 Elsevier Science B.V. All rights reserved.

533

P h o s p h o r u s r e m o v a l b y slag: e x p e r i m e n t s a n d m a t h e m a t i c a l modeling S. Vigneswaran a and H. Moon b Faculty of Engineering, University of Technology, Sydney, PO Box 123, Broadway, NSW 2007, Australia

a

b Faculty of Applied Chemical Engineering, Chonnam National University, 300 Yongbong, Kwangju 500-757, Korea

1. INTRODUCTION Phosphorus is one of the essential nutrients needed for the growth of plants and animals. When plants and animals excrete wastes or die, microorganisms mineralize the various forms of organic phosphorus in the decaying matter. Both Sewage Treatment Plant(STP) effluent and urban run-off also lead to phosphorus enrichment to the waterways. In country sides, where agriculture and animal husbandry are main industries, wastes from these activities contribute to the accumulation of P in soil and water bodies. These phosphorus compounds dissolved in surface or ground waters are responsible for the eutrophication in the closed water system, especially in lakes and highly enclosed bays where water is stagnant. At present, chemical treatment, biological processes, and land treatment are used to remove phosphorus from water and wastewater. Among them, the land treatment is an attractive solution if the substratum has significant adsorptive capacity of phosphorus [1-3]. Sydney Water, Australia, recommended the land application of sewage effluent in areas where there are serious eutrophication problems (for example in Richmond area in New South Wales, Australia). On the other hand, a constructed wetland system has been suggested as an appropriate technology for P treatment in many parts of the world [4]. In this respect, phosphorus removal by slag media (waste by-products from steel industries) may be a good solution. Several researchers have investigated the utilization of natural soils and amended sand with iron oxides for phosphorus removal [1,5]. These studies were confined mainly to investigate the removal efficiency using those materials. No study, however, have attempted to deal with the P transport in substrata to provide basic information for the design of efficient land treatment facilities.

534 A detailed study on the phosphorus removal using soil and slag media from steel industries [6,7] indicated that slag media have higher sorption capacities for P compared to other substrata, such as natural soils. In particular, the batch and column experiments conducted with natural soil and slag media showed that adsorption was dominant at low pH values less than 6 while chemical precipitation, dominant at higher pH values over 8. The slag particles contain significant amount of soluble metal ions such as magnesium and calcium. Those metal ions are responsible for chemical precipitation and complexation at high pH. The section 2 of this chapter presents the results of column experiments to show the dynamic behavior of P in the column with slag at various operating conditions. Non-equilibrium dynamic models based on surface or pore diffusion inside slag media were used for simulating the adsorption behavior of P. Comparing simulated results with experimental breakthrough curves of P assessed these models. In particular, the effect of kinetic model parameters on breakthrough behavior will be rigorously discussed. At high pH, the phosphorus retention by chemical precipitation is significant. The soluble metal ions react with phosphorus in the solution to form insoluble precipitates such as calcium phosphate. A number of early researches on P chemistry postulated the formation of various insoluble inorganic phosphates of Fe, A1, Mg, and Ca through precipitation reactions. Aluminum ions combine with phosphate ions to form aluminum phosphate. Both ferrous (Fe 2+) and ferric (Fe 3+) ions are also responsible for the precipitation of phosphorus. With Fe a§ the reaction is similar to that of aluminum ion. Fe 2§ shows more complicated reactions, which are not fully understood. In the case of Fe 2§ ferrous phosphate is usually formed at pH values about 8. The precipitation of phosphate from wastewater by calcium is common. The batch precipitation experiments conducted by Lee [7] indicated that a considerable precipitation of P was observed at pH greater t h a n 8. Further, the contribution of precipitation was found to increase with the increase of pH. According to Lee [7], the dominant removal mechanism of P at pH greater than 10 is precipitation. Jenkins et al. [8] observed species such as PO43-, HP042-, P30105-, and P2074- at high pH. Aulenbach and Meisheng [9] showed that calcium present in the wastewater enhanced precipitation of P as a crystal of hydroxy apatite: 3PO34- +5Ca 2+ + O H - --+ Ca5(PO4)3OH At high pH, when phosphorus removal is both by adsorption and precipitation, in order to obtain the exact adsorption isotherms, one has to use the real adsorption amount after subtracting the amount of precipitation from the total uptake. One can also use equilibrium data obtained with washed slag media. In

535

the washed slag media most of most soluble metal ions are removed ( or washed out ) from the adsorbent particles. Because of the presence of the reactive species in the solution, the adsorption model has to incorporate the chemical reaction to calculate the phosphorus removal in slag media. For simplicity reasons, in the present work, the precipitation reaction of phosphate with soluble cations was represented by a simple bimolecular reaction. Since the precipitation reaction occurs both in the bulk solution and in the pore solution, a pore diffusion model incorporates the chemical reaction was developed to interpret the kinetic data of P in batch and column adsorption experiments. The model details are given in Section 3.

2.

PHOSPHORUS

A D S O R P T I O N : C O M P A R I S O N OF K I N E T I C M O D E L S

2.1. K i n e t i c m o d e l s 2.1.1. D y n a m i c m o d e l b a s e d o n s u r f a c e d i f f u s i o n A surface diffusion model(SDM) with external mass transfer resistance was used because of its simplicity and adequacy in describing the adsorption of P from aqueous solutions onto slag media. For model formulation, we assume an isothermal adsorption column, packed with porous spherical particles. The flow pattern is described as an axial dispersed plug-flow model. Another assumption involved in the model is the fast intrinsic adsorption kinetics, resulting in instant equilibrium between the solid phase and liquid phase concentrations at the external surface of particle [10]. If the surface diffusion is dominant, the mass balance inside a spherical porous particle can be described by the following equation:

at

~

~r--2 + -r-~r-r

/

(1)

with the initial and boundary conditions: q(r,t=0)=0

(2)

aq =0 a r r=0

(3)

aq

: kf(C-Cs)

DsPp-~-r r=Rp

(4)

536 where D~ is the surface diffusion coefficient, pp is the particle density, and kf is the external film mass transfer coefficient. The mass balance equation and the relevant initial and boundary conditions for the liquid phase in the column are as follows:

_Da x 02C

0(uC) 0C ( 1 - ~ ; b ) 3 k f ( c C ) 0 + + - s = 0z 0t ~b Rp

a--~ +

C(z,t = 0)= 0

(5)

(6)

- - (Clz o -Clz 0+)

Dax-~-z z:0

aC =0 8z z=L

(8)

The final term surface of particles concentration, Cs, equilibrium theory

of Eq.(5) represents the mass transfer rate to the external which is proportional to the driving force, C - Cs. The surface can be evaluated from the corresponding isotherm or [11].

2.1.2. Pore diffusion model(PDM) If pore diffusion is dominant inside the particle, the intraparticle diffusion model can be represented by the following equations aCp

aq

+ pp

aCp ~pDp Or

=~

lO(

r2Dp

aCp)

=kf(C-Cs) r

(9)

(10)

=Rp

aCP I =0 a r r=0

(II)

Cp(r,t : 0 ) : 0

(12)

where Dp is the pore diffusion coefficient and % is the particle porosity.

537

2.1.3. Simplified m o d e l A simplified model can be formulated from the postulate that the uptake rate by a pellet is linearly proportional to a driving force, defined as the difference between the surface concentration and the average adsorbed-phase concentration. This is called "Linear Driving Force Approximation(LDFA)". According to this approximation, the intraparticle diffusion can be simplified as follows [12]: 0~ = 3k____L(C_Cs)= ks(q s _ ~ ) 0t Rppp

(13)

Here, ks is the solid-phase mass transfer coefficient. For spherical particles, the average adsorbed-phase concentration, ~, is defined as Rp

_ q=

3 Rp

3

]rr2qd r

(14)

0

In the case that the surface diffusion is the controlling mechanism of intraparticle mass transfer, this can be estimated from the following equation. ks

~

15Ds ~ Rp 2

(15)

This relationship has been mathematically derived based on the assumption that the adsorbed-phase concentration profile is parabolic. When adsorbent particles are very small, this simplified version has been successfully applied. The coupled parabolic second-order partial differential equations, Eqs. (1) and (5), can not be solved analytically. Thus, a numerical method must be employed. In this work, a method based on orthogonal collocation [13-15] was used. The method of orthogonal collocation was used to reduce these coupled second-order partial differential equations to a set of first-order ordinary differential equations using several interior collocation points. This technique combines the classical procedure of orthogonal collocation with the high accuracy of the finite element method. The entire column was divided into a finite number of elements, in which the space variables are discretized. The set of ordinary equations was integrated using an integrator named LSODI. LSODI employs the variable-step size, variable-order, and the predictor-corrector techniques that are suitable for stiff equations.

538

2.2. E x p e r i m e n t a l Slag media (dust and cake - waste materials obtained from a steel industry) were used as adsorbents in column experiments. The dust was t a k e n from the secondary bag house deducting system used to control fugitive emissions from vessels in the steel industry while the cake is the gas scrubbing slurry from the blast furnace. In the steel production process, they are produced independently. Slag media were air dried at 105~ for a day and the fraction less t h a n 2 m m size was obtained by sieving. Slag samples were mixed well before being used in the experiments. P solution was prepared by dissolving Na2HPO4 in distilled water. The properties of dust and cake were analyzed by s t a n d a r d procedures [16]. A pycnometer was used to determine the particle and bed densities. Permeability was d e t e r m i n e d by the constant head method. The surface area and pore size were m e a s u r e d by a nitrogen adsorption method (BET, Micromeritics ASAP2400). Total element concentrations were determined in an acidified solution by ICPAES (Inductively Coupled P l a s m a Atomic Emission Spectrometer). Field conditions were simulated in the laboratory-scale column packed with slag media. Column experiments were carried out mainly to find the concentration b r e a k t h r o u g h p a t t e r n under various conditions. A phosphorus solution was supplied to the columns under gravity from overhead containers (each with a capacity of 20 liters). Constant head was m a i n t a i n e d by providing an overflow tube above the adsorbents. Effluent flow was controlled by a flow meter. Concentration of POn-P in the tap w a t e r was less t h a n 0.05 mg/L. Forty liters of the tap w a t e r was infiltrated through the column prior to the introduction of the P solution to ensure a uniform compactness of the adsorbents. Care was t a k e n not to allow any air bubbles in the under drainage section and in the media layer. It is evident t h a t the solid phase concentration (q) and the liquid phase concentration (C) m u s t coexist in equilibrium at the b r e a k t h r o u g h point. The pH of the P solution in the influent was m a i n t a i n e d in the range of 6 to 7. Six columns of 9 cm d i a m e t e r were packed with the dust and cake adsorbents to thickness of 1 and 3 cm, respectively. Depth of the media was m a i n t a i n e d shallow to achieve an early breakthrough. The entire effluent concentration history was recorded. To e s t i m a t e the solid phase concentration, the soil and slag in the column was t a k e n out for air drying at the b r e a k t h r o u g h stage of the adsorbents. It was further dried in an oven at a t e m p e r a t u r e of 105~ for 24 hours. Then the preweighed a m o u n t of the adsorbents was extracted using nitric acid. After extraction, they were diluted with a required volume of distilled w a t e r and the solution was decanted. Then the samples were analyzed for P concentration using a spectrophotometer (Milton Roy Spectronic 20D). 2.3. E x p e r i m e n t a l r e s u l t s and m o d e l v e r i f i c a t i o n 2.3.1. C h a r a c t e r i s t i c s of the a d s o r b e n t s Physical and chemical properties of the adsorbents were m e a s u r e d since P t r a n s p o r t is influenced by the intrinsic characteristics of the media. The physical

539 characteristics of the slag media tested are listed in Table 1. The chemical properties are also shown in Table 2. More detailed properties are given elsewhere [7]. The values of the permeability of slag media are in the range of 1.8 - 3.2"10 .7 m/s. These values are very low so t h a t the slag media are not suitable in rapid infiltration systems. This result would be expected because two slag media used in this study have slit ( TiO2/A1203/SiO2 >A1203 > TiO2//SiO2 > TiO2 > ZnO ~ A1203/SIO2 > SiO2 and the electron acceptor ability (acidity) of the inorganic oxides, as it follows from their KA values, varies as follows: TiO2/A1203/Si02 > Ti02/SiO2 > A1203/Si02 > ZnO > A1203 > TiO2 > Si02 > MgO From the comparison of CA and EA contributions for the oxides it follows that they have more high EA than CA values, i.e. they exhibit "hard" acidic properties. This was explained by a large contribution of the electrostatic field effect of the oxide solid matrix to the adsorption heat of test bases [24]. The most high EA

601 quantity was estimated for alumina/silica with 0.23 wt% A1. This may be due to formation of strong Lewis acid sites and enhancement of the acidity of OH groups which are bound to trigonal aluminium cations on the boundaries between patches of A1203 and SiO2 lattice. Modification of inorganic oxide surface by "soft" organic groups, as in the case of grafted 7-aminopropylsilylated glass beads, gives high decrease of the EA contribution, whereas the CA contribution even increases after the surface treatment. Some q u a n t u m chemical indexes of inorganic oxides and acceptor characteristics (ET) are presented in the last column of Table 7. The q:a.~ and ELUMO quantities were estimated in [24] for silica and alumina/silica surfaces from the relationships between these indexes and CA, and EA contributions. The range of these indexes is close to those computed by quantum chemistry. The experimentally derived ET parameter reflecting the acidic properties of solid surface increases with transition from silica to alumina and decreases in the case of 7-aminopropylsilylated silica in comparison with untreated silica. This order coincides with one determined from change of KA and EA for these solids. In addition, the proton affinities of conjugated bases of the oxide surface OH groups computed by q u a n t u m chemistry (Ea) are presented for some oxides. This quantity characterizes the acidity of the surface OH group in gas phase. The acidity of OH groups of the metal oxide surfaces determined by such manner decreases in the following order 7- A1203 > SiOffamorph) > TiO2(rutile) > MgO It may be seen that this order is close to one determined from change of KA quantity for the metal oxide surfaces. Crude estimation for ability of inorganic oxide surface to interact with organic compounds as acid or base can be found from mean electronegativities of these solids [70]. The mean orbital electronegativities of a solid ()~s) can be calculated using following expressions

~:S =

n i x 7~i Z ni

~i = IPi + EAi

(40)

(41)

where ~i is the Mulliken' orbital electronegativity, first ionization potential and electron affinity of i-th atom in an inorganic molecule, while ni is the number of these atoms in the molecule. These quantities for various inorganic oxides, salts and some minerals are presented in Table 8.

602 Table 8 Mean orbital electronegativities (~) of various solids (taken from [70] Solid

Z, eV

Solid

Z, eV

K20 Na20 CaO SrO BaO Ca3SiO5 MnO CaSiO4

4.06 4.34 4.81 4.87 4.96 5.41 5.44 5.58

MgO 5.68 Ca~Si207 5.71 La203 5.75 A1203 5.82 PbO 5.88 CaSiO3 5.89 Mg2SiO4 6.08 CaCO3 6.19

Solid

)~, eV

Solid

ZnO B203 TiO2 Fe203 MgSiO~ ZrO2 V205 A1PO4

6.14 6.18 6.21 6.22 6.24 6.24 6.44 6.50

Talc MgCO3 Mo3PW1204o SiO2 Kaolinite MoO3 WO3

~, eV 6.51 6.54 > 6.56 6.61 6.61 6.69 6.72

The acidity of the solid surface increases as mean solids electronegativity increases, whereas basic properties change in the opposite direction. Actually, it is known that alkaline, alkaline-earth metals, Mg and Zn oxides exhibit mainly basic properties in various catalytic transformations whereas Zr, Ti, Si, Mo oxides, clays and heteropoly acids typically display acidic properties. The conclusions from Table 8 coincide with observed order of KB change at transition from SiO2 to MgO. Then, a strong interaction should be expected between typical organic bases (ethers, amines, nitriles, etc.) and surface sites of last group of inorganic solids from Table 8. It should be mentioned that the behavior of OH groups at oxide surfaces depends strongly on the composition of oxide and on local chemical environment. It is known that OH groups of the alumina surfaces display wide spectrum of Bronsted acidic and basic properties [134]. The single OH groups of the surfaces exhibit mainly typical basic properties, whereas the bridged OH groups bounded to trigonal aluminium atom behave as typical Bronsted acid sites [135]. On partially dehydroxylated surfaces, they have various acid/base character and interact with organic adsorbates according to their own characteristics (IP, EA quantities, proton affinity and partial charges on the atoms in the surface clusters). Therefore, the acidic and basic properties of oxide surfaces depend on the temperature of surface pretreatment, surface coverage of adsorbates and concentration of physically adsorbed water which effects on the stability of adsorption complexes between organic adsorbates and the surface sites [136-137]. So, the Gutmann' and Drago-Weyland' parameters for different inorganic oxides and carbonaceous materials allow to predict the specific interaction contributions to the adsorption energies of organic compounds on solid surfaces. Also, these parameters may be useful for prediction of the reactivity of surface sites towards various reagents from gas phase. In contrast to descriptors of the molecular structure of organic compounds, which can be computed by quantum chemistry methods or taken from various quantitative relationships derived in

603 the processes, taking place in the solutions or gas phase, similar procedures for q u a n t u m chemical computation or evaluation of experimental descriptors for solid surfaces often run to various serious problems and uncertainties. The G u t m a n n ' and Drago-Weyland' parameters may be useful descriptors reflecting the reactivity of surface sites in various surface acid~ase reactions, including catalytic transformations of adsorbed compounds. For example, OH groups on the silica surface manifested as nucleophiles at interaction with most organosilicon compounds. This reaction causes the substitution of proton of this group by organosilicon residue. On the hypothesis that the reaction mechanism (SEi) is invariant in going from SiO2 to parent and mixed Ti and A1 oxides, one would expect decrease of chemisorption barrier at enhancement of the KB for oxides series. Actually, the chemisorption activation energy of an organosiloxane decreases as the KB of the metal oxide surface increases [83].

6

~

5 ZnO

9 ~O

4

"*=

3 -

~ 3 / S i O

2

.a 2 -

A1203

i

5

10

~'TiO2

I

I

I

15

20

25

*

30

35 -1

K a (kcal mol ) Figure 5. Ozone lifetime on the surface of inorganic oxides as a function of acceptor number for these oxides.

Second example is related to ozone lifetime on the metal oxide surfaces and it presents great interest for heterogeneous atmospheric chemistry. As ozone molecule possess weak basic properties [138], the decrease of ozone lifetime on the solid surfaces determined in [139] should be expected with the increase of the metal oxides acidity characterized by KA quantity. As it follows from Figure 5, the ozone lifetime on the parent A1, Zn, Mg, Ti oxide and alumosilicate surfaces decreases in going from basic ZnO and MgO to more acidic AleO3, TiO2 and A1203/SIO2 surfaces. This may be explained by adsorption of ozone molecule onto

604 stronger Lewis acid surface sites causing the ozone dissociation into free dioxygen and a surface oxygen atom, which remains attached to the metal cation in the oxide lattice. The acidYbase characteristics for other metal oxide surfaces which are absent in Table 7 may be estimated crudely from comparison the sequences for change of KA, KB, E and C parameters for series of metal oxides with order of change of the acid/base parameters determined in other scales or from quantitative relationships between these scales for a given series of solid surfaces. The different methods are proposed for the estimation of the order for relative acidity and basicity of inorganic oxide surfaces. The basicity of oxides increases at enhancement of negative charge on lattice oxygen atom in the following order [140] MgO > MgA1204 > A1203 > Zr02 ~ TiO2 Above basicity order coincides with that determined from increase of the positive charge on metal cations in the oxides lattice A1203 > ZrO2 ~ TiO2 Desorption energies determined from TPD experiments or differential adsorption heats of SO2 possessing acidic properties on the oxide surfaces may be chosen as a measure of their basicity MgO>CoO2 >ZrO2 >MgA1204 >A1203 >TiO2 The Lewis basicity of metal oxide sites decreases as [140] MgO > ZrO2 > TiO2 (rutile) > TiO2 (anatase) > MgA1204 > y- A120~ The temperature-programmed desorption of preadsorbed acidic carbon dioxide is frequently used to measure the number and strength of basic sites. The strength of basic sites determined from temperatures for maximum of the CO2 desorption peak in TPD spectra is in the increasing order of [141] BaO > SrO > CaO > MgO whereas the metal oxide basicity evaluated from differential adsorption heat distributions of C02 varies as [142] 7- A1203 > TiO2 (rutile)

Other different spectroscopic (XPS, UV adsorption and luminescence spectroscopies of the solid surfaces, IR spectroscopy of adsorbed CO2 and pyrrole)

605 and other methods (TPD of hydrogen, titration of solids by typical acid]base indicators, 1so exchange between carbon dioxide and surface oxygen) are proposed for characterization of the basic surface sites. They are reviewed in detail in [143]. Different techniques have been developed to determine the solid acidity and they are briefly described in recent review [144]. Among them the temperatureprogrammed desorption, microcalorimetry and IR spectroscopy of preadsorbed base molecules (NH3, CO, acetonitrile, pyridine, n-butylamine, quinoline, etc.) are widely used in determination the strength of Bronsted and Lewis acid surface sites. In accordance with these data, the acidity of Lewis sites on metal oxide surfaces varies as [140] y- A1203 > TiO2 > ZrO2 > CeO2 The acid/base characteristics of the metal oxide surface sites are changed at addition of minor concentration of extraneous ions. This may be a typical situation in the case of solid atmospheric aerosols having complex surface composition (Table 2). From data of microcalorimetry of base (NH3) and acid (SOD probes on 7-alumina, silica and magnesium surfaces at adding small amount (0.1 wt%) of ions (Ca 2§ Li § Nd 3§ Ni 2§ Zr 4§ SO~-) it was found that the modification of 7-A1203 surface properties changed moderately its amphoteric properties [145]. More substantial changes are observed on MgO which consisted in formation of sites of moderate and weak basic strength. The number of acid/base sites on doped SiO2 is strongly affected by the presence of introduced ions. The acidity of modified oxides increases at enhancement of generalized electronegativity of the metal ions, (Xi), expressed as X i = (l + 2Z)• X,,, where X0 is the electronegativity of neutral atom (Z = 0) given by Pauling and Z is the charge of metal ion [146]. This behavior is much more pronounced on silica series than on alumina series oxides. The cation electronegativity is a parameter determining the ionicity percentage of cation-oxygen bond in oxides. Then the ratio of the oxidation degree to the ionic radius of doping ions is also representative of the Lewis acid strength of cations. It is observed that acidity of modified oxides increases while the charge/radius ratio grows. The Lewis basicity of the modified oxides is associated with the electric charge of the oxygen adjacent to cation [147]. The average heat of SO2 adsorption per basic site increases as the partial oxygen charge associated with the cation grows. This tendency is much more evidenced on the doped silica samples which, because of very weak acidic character of this oxide, reflect much more the doping effect of inorganic ions. Most of organic compounds (alcohols, ethers, nitriles, amines, thio- and phosphor-derivatives, organometallic) are irreversibly adsorbed on most metal oxide surfaces at low temperatures after heat pretreatment of surfaces at moderate and high temperatures. Preliminary partially dehydroxylation of nonreactive cristobalite and pyrogenic silica surfaces at 1073 K also gives the

606 irreversible adsorption of methanol and tert-butanol on the surfaces at 303 K, and their initial differential heats of adsorption increase in comparison with that for the solids pretreated at 423 K [148]. The physical adsorption of gas phase reagent onto solid particle surface and formation of surface H- or donor/acceptor adducts is a first stage of its chemisorption and different surface reactions. The subsequent transformations of the adducts are affected by way of temperature or reagent concentration enhancement as well as by presence of third base or acid component (catalyst). Because different inorganic gases possessing high acid/base characteristics (SO2, HC1, C12, NH3) are adsorbed on the surface of atmospheric solid aerosols, the reactivity of the surface Bronsted and Lewis sites grows and formed H- or donor-acceptor complexes between these sites and organic pollutants are close to the transition state of the chemisorption. Various problems which arise in connection between chemisorption kinetics and adsorption equilibrium of organic compounds on metal oxide surfaces are discussed in [83]. The simple expression for apparent chemisorption activation energy (Ec~hem) with consideration of the H- or donor/acceptor complex formation prior to slow chemical reaction on the solid surface may be derived using the steadyconcentrations assumption for such a complex ~ (ESp + E nsp ) Eapp = E chem-

(42)

where Ec~hemis the activation energy of the chemical reaction between interacting gas and solid surface. Therefore, the apparent reaction barrier reduces with the increase of preliminary adsorption complex stability, and overall reaction rate grows. The overall adsorption energy can be estimated by means of methods discussed above. Moreover, the increase of preliminary adsorption complex stability gives increase of the reagent concentration on the surface and effectiveness of its following monomolecular processes or bimolecular LangmuirHinshelwood surface reactions is grower. Most descriptors of organic compound structures used in their relationships with adsorption energies are also available for quantitative description of chemisorption and surface reactions. However, because of cleavage and formation of new chemical bonds during chemical reactions and large differences between the structure of reagents in the initial and transition states, other descriptors, including the bond dissociation energies, relaxation energies of reagents in an transition state, are proposed for estimation of reaction barriers on metal oxide surfaces [24].

Q

C H E M I S O R P T I O N AND S U R F A C E REACTIONS OF ORGANIC C O M P O U N D S ON METAL OXIDES

The reactions of solid aerosols (NaC1, sulfuric acid-coated solid-propellant rocket motor-exhaust particles, a-A1203, ice, soot particles) with atmospheric

607 gases (NO2, N205, C1ONO2, CO2, SO2, HC1, O2, O3, H20) have been studied for many years [149-159]. However, little attention has been paid to the possibility of organic atmospheric species reacting directly with solid aerosols. These reactions are of interest as a possible means of disposing of stockpiles of organic compounds or products of their photochemical and thermal transformations in Earth atmosphere. Main mechanisms of organic compounds interaction with Bronsted and Lewis acid/base sites of metal oxide surfaces, stages of chemisorption reactions, kinetic equations taking into consideration the chemical and structural heterogeneities of the metal oxide surfaces and some quantitative "structure-reactivity" relationships (QSRRs) for these transformations have been discussed in our recent review [137]. However, most of experimental data considered in [137] are concern to the best investigated reactions of various organosilicon compounds with OH groups of parent and mixed Si, Ti and A1 oxide surfaces. Considered regularities for effects of substituents in these compounds, dissociation energy of bonds and structure of oxide surfaces on the chemisorption activation barrier can be extended to the chemisorption of organic pollutants on solid aerosol surface. But in this case the difference in the electronegativities of Si in the organosilicon compounds and C atoms in most organic pollutant molecules is to be taken into account, as well as the fact that last reactions are more complex and their kinetics is far less studied than that for former processes. Most of quantitative data for products of C, N, O, S, C1, F-containing organic compounds interacting with various metal oxides are derived from FTIR, and X-ray photoelectron spectroscopic, mass-spectrometric and chromatographic studies on the surface and in the gas phase in course of these transformations, whereas main body of kinetic results in this field is derived from temperature-programmed desorption experiments. We will make an attempt to consider here the main mechanisms of possible reactions of organic pollutants with active sites of metal oxide surfaces, experimental kinetic data derived from TPD measurements, kinetic equations, and possible QSRRs in these transformations. 5.1. M a i n m e c h a n i s m s o f c h e m i s o r p t i o n on t h e m e t a l o x i d e s u r f a c e s

The Bronsted acid (various types of OH groups) and base sites (oxygen atom in these groups and M-O bonds of lattice) as well as Lewis acid (metal cations M n§ surface free radical, e. g. = Si ~ and base sites (ones, similar to Bronsted base sites and O 2-, surface radicals - S i O ~ etc. ) are the main reaction sites of the inorganic oxide surfaces. The organic compounds (RX, RH, where R is an hydrocarbon moiety and X is a polar functional group) interact with these sites in accordance with following mechanisms [137, 143-144, 160]. 5.1.1. R e a c t i o n s w i t h B r o n s t e d a c i d / b a s e s i t e s

i. Electrophilic substitution of the surface OH group proton by an organic residue

(SEi)

608 MO-H + R-X --+ MO-R + H-X ii. Nucleophilic substitution of OH group by an organic residue (SNi). W h e n RX is an alcohol, the grafted alkoxy groups m a y be intermediates in its subsequent acid-catalyzed reactions M-OH

+

R-X -~ M-R

+

X-OH

iii. Dissociative electrophilic or nucleophilic addition to a metal-oxygen bond (AdEi or AdNi) M-O-M + R-X -~ M-OR + M-X iv. Proton transfer from acid site to organic compound (AHi). The formed onium cations m a y be i n t e r m e d i a t e s in subsequent acid-catalyzed reactions, e. g. dehydration of alcohols and hydration of olefins, etc. M-OH + R-X -+ Mn+O2- + [RXH] § v. Proton transfer from organic compound to base sites of oxides (EHi) and subsequent base-catalyzed reactions, e. g., double bond migration in olefines CH2=CH-CH2-CH2X + Mn+O 2- ---> [ C H 2 - C H - C H - C H 2 X ] - + Mn+O - H + --+ CH3-CH=CH-CH2X + Mn+O 2vi. Reactions of u n s a t u r a t e d compound with paired Bronsted acid sites M(OH)-OH+ >C=C< --+ >(HO)C-CH< + Mn+O 25.1.2. Reactions

with

Lewis

acid/base

sites

i. Redox reactions of Lewis sites with formation of organic radical-cations, radical-anions or addition of lattice oxigen to an organic compound R + Mn+O2- --> R X +" + M(n+l)+O2-

R + Mn+O2- - , R X " + Mn+O RH + Mn+O 2- --> RHO + M(n-1)+Dii. Dissociative adsorption on the Lewis acid/base pair with formation of Bronsted acid site (Adi) R-H + Mn+O2" -> R-M-OH Formed surface complexes and grafted organic groups undergo to subsequent monomolecular and bimolecular transformations. For example, n-complexes

609 formed between olefins or aromatic hydrocarbons and Lewis acid sites are turned to c-complexes. These complexes may be oxidized by Lewis acid surface sites up to products of their complete oxidation, carbon dioxide and water. Mutual surface diffusion of adsorbed organic free radicals gives their dimerization, polymerization or addition to adsorbed initial organic compounds or to reactive surface sites. It should be mentioned that surface of solid aerosols contains various metal cations, inorganic anions and inorganic salts. Because small size of aerosol particles, strong electric fields arise on their surfaces nearby charged sites. Electronic and spatial structure of organic molecules and products of their interfacial reactions: ions, free radicals, adsorbed on solid surface differ from the structure of these species in gas phase. As a result, the activation barriers of their reactions with other gases or adsorbed components will change and composition of reaction products will differ from those of similar processes in gas phase or in solutions. The charged surface sites induce orientation of reagent molecules. Conformation transitions being possible in molecules, the conformations with higher dipole moment are mainly formed on solid surface. Regioselectivity of the further reactions of adsorbed products is determined by their conformations on solid surface. The influence of electric field on the structure and reactivity of molecules and ions localized nearby solid surface, and the mechanism of interface reactions was exhibited theoretically and confirmed by some experiments on the surface of various heterogeneous oxide catalysts, especially in the zeolites cavity ([161] and references therein). These effects clearly can be manifested by the action of strong electric field with intensity varied from 0.005 to 0.06 a. u. Also, the intensity of non-uniform electric field near solid rough surface of ultrafine particle exceeds the value for flat surface, as F = U/(A r), where U is the outer potential difference and r is the radius of divided crystallites or roughness of the surface. The effects of intensity and vector direction of the electric field on activation barriers of unimolecular interfacial reactions of simple organic compounds were studied by semiempirical quantum chemistry method for orientation of molecules and radical-anions, heterogeneous electron transfer to adsorbed molecules, dissociation of radical-anions, cations and molecules, inversion of anions and molecules, cis-trans isomerization in olefines and chair-boat conformational transition of cyclic molecules in [161]. It was found t h a t electron and spatial structure of neutral particles and ions, height of activation barriers of their reorientation, dissociation, isomerization and conformation transition, the configuration stability to change as affected by the field in the range 0.01 + 0.05 a. u. These effects are to be taken into account at explanation of the surface reaction kinetics and the products composition on the surface of atmospheric solid aerosols. As is evident from Table 2, the surface of most solid atmospheric aerosols contains marked amount of such acidic components, as adsorbed gases NO2, N205, C1ONO2, SO2, and HC1, and their corresponding acids or salts. It is known that sulfate t r e a t m e n t of oxides such as Fe203, A1202, SnO2, TiO2, ZrO2, etc.,

610 gives high increase in the surface acidity and in the catalytic activity, for example, in dehydration of alcohols, olefins hydration, esterification, acylation and isomerization reactions [162]. These solids were claimed to present superacid sites with acid strengths up to Ho < - 16.04. These sulfated metal oxides usually are prepared from their hydroxides by treatment with H2SO4, SO2, SO3 H2S or (NH4)2304 and following calcination [163]. The nature of the high acidity of the sulfated metal oxides is still controversial. It was proposed that the very strong acidity is due to the increase in number and strength of Lewis acid sites [164]. More realistic structure for active site implies the formation of Bronsted acid sites which would allow catalytic reactions to occur at much lower temperatures with the same rate as they occur at higher temperatures without the presence of the Bronsted sites [165]. Also, the addition of chlorine to alumina enhances the activity of solid for skeletal transformations of hydrocarbons [166]. The role of halogen promoter at the surface is to enhance the Lewis or Bronsted acid environments, respectively, through an electroattractive effect of the halogen atom from an adjacent OH group [167]. Quantum chemical studies indicate that the stabilization of LUMO energy occurs with a narrowing of the energy gap, by the substitution of chlorine for oxygen or OH groups in T-alumina cluster models [168]. The Bronsted acid character of the cluster also increases after substitution. Similar enhancement of surface acidity and reactivity in the chemisorption of various organic bases has been observed after substitution of OH groups on silica surfaces by chlorine [169]. Therefore, the oxide surfaces of solid aerosols in the presence of acid additions can exhibit much more high acidity and catalytic activity in the reactions of adsorbed organic compounds than their one-off surface acidity and reactivity suggests. 5.2. K i n e t i c s of c h e m i s o r p t i o n a n d s u r f a c e r e a c t i o n s The rate of adsorption of gas-phase species, Va can be written as [170]

Va : s r ( o ) x F

(43)

where F is the flux of gas reagent and Sr(O) is the sticking coefficient. The superscript r is 1 for non-dissociative and 2 for dissociative adsorption (chemisorption). For direct adsorption (chemisorption) from gas phase, s r ( o ) = sr (0)x (1- O)r

(44)

where Sr(0) is the sticking coefficient at zero coverage O. In accordance with Arrhenius equation

(Ea)

sr(o)= S~ xexp - ~ -

(45)

611 where

S~ and Ea are the pre-exponential factor and activation energy for

adsorption. The normal pre-exponential factors for these reactions according to e s t i m a t e d from transition state theory are varied from 10 -lo to 1 0 1 7 c m 3 s -1 at r = 1 and 2 [171]. The overall non-dissociative adsorption (chemisorption) rate of gas-phase species onto solid aerosol surface (Va(over)) c a n be described by using the simple model for discrete heterogeneous surface [172]

Va(over )

=

F x S Ox

Ea(i) Oti exp - Ea(i) kT x exp - S O x x x exp - kT

i•1 .

(46)

where Oti i s the relative concentration of i-th subsurface, Ea(i) is the chemisorption activation energy on i-th subsurface and x is the reaction time. Because, in the case of dissociative and associative surface reactions, the consideration m u s t be given to correlation effects between E(i) on different sites, the later kinetic equations will be given only for surface reactions t a k i n g place on uniform surface. W h e n chemisorption is a s s u m e d to proceed via the precursor state (H- or donoracceptor surface complexes) discussed above, then rates of non-dissociative and dissociative chemisorption can be written as [173] V a = qxFA x k a ( 1 - O A ) kd* + k a*(1 - | A ) Va

=

g•

• ka(1-|

(47)

(48)

k~ + k a ( 1 - O A ) 2 where ~ is the t r a p p i n g probability from the gas phase into the precursor state, k a is the rate constant for transition from the precursor state into the chemisorbed state, and k d is the rate constant for desorption of the reagent from precursor state. The rate of desorption for a randomly distributed adsorbate can be written as

r@r Vd = kd

(49)

where k~ is the r a t e coefficient for desorption, obeying to Arrhenius equation r

kd = ~d xexp -~-~

(50)

612 where v~ and Ed are the pre-exponential factor and the activation energy for desorption, respectively. The desorption pre-exponential factors range from 1013 to 1019 s 1 for r = 1 and from 1011 to 1019 s -1 for r = 2. Surface bimolecular reactions are classified into L a n g m u i r - H i n s h e l w o o d a n d Eley-Rideal processes. The first includes reactions between two adsorbed species (A and B) or an adsorbed species and a vacant site (V) V 1 - k l A B , or V! = k 1 A V

(51)

where kl is the rate coefficient, 0i is the surface coverage of species i a n d 0v is the fraction of v a c a n t sites. The kl value is obeyed to A r r h e n i u s equation k 1 = v 1 x exp - ~

(52)

The pre-exponential factor for this reaction varies from 10 -n to 104 cm 2 s -1. The second type of the reactions includes the direct interaction of gas-phase species with adsorbed species and the reaction rate can be w r i t t e n as Ve = ke0APB

(53)

where pB is the partial pressure of r e a g e n t B. The ke value can be r e p r e s e n t e d t h r o u g h t reactive sticking coefficient, So

/ Ee)

S~ )1/2 exp k e - (2~mBk T

(54)

where as is the area per reaction site, and mB is the molecular weight of species B. The pre-exponential factor for these reaction varies from 10 .6 to 10 -17 cm 3 s -1.

5.3. Quantitative "Structure-Reactivity" relationships in surface reactions of organic compounds The Evans-Polanyi relation between the activation b a r r i e r (Ea) and the reaction e n t h a l p y (AH) is most often used for description of reaction kinetics on the solid surface Ea = (~ AHr + ~

(55)

where (z is the t r a n s f e r coefficient (0 _< a < 1) and ~ is close to the intrinsic reaction barrier. As is shown [174], this relation may be useful for prediction of Ea values for chemisorption of some organic compounds on hydroxylated silica surface. Also, the bond-order-conservation-Morse-potential (BOC-MP) method [175] t a k i n g into account reaction enthalpy, adsorption h e a t s and bond dissociation energies in

613 reagents and products may be used for prediction of activation barriers for various surface reactions. The Ea value for dissociative adsorption from the gas phase can be written as QAQB E a =0.5 I DAB-(QA +QB)+ QA +QB _QABI

(56)

where QA, QB and QAB are the adsorption heats of species A, B and AB, respectively. For non-dissociative desorption, Ed is given by Ed = QA or Ed = QAB

(57)

whereas for associative desorption, Ed is QAQB +AH) E d = 0.5 QA + QB

(58)

where the enthalpy change for the surface reaction is given by AH = DAB + QAB- QA- QB

(59)

In the case of disproportionation reaction A + BC -~ AB + C, the activation barrier is Ed=0"5/QABQABQC + + Q c AH)

(60)

where the enthalpy change in the reaction is AH = DBC- DAB- QAB + QA- Qc + QBC

(6~)

However, the attempts to estimate the activation energies for the interaction of organosilanes with hydroxylated silica surface using Q values computed by quantum chemistry (Eq. 60) lead to Ea values rather than on 50 +: 100 kJ mol 1 higher than experimental values. Above expressions apparently are suited for surface reactions, which occur through only nonpolar or weakly polar transition state. The static quantum indexes of reactivity may be used for description of the double exchange reactions, which occur through the transition state with high charge separation. If barrier height is controlled by frontier orbital interactions (interaction of a soft acid with a soft base) then the following relationships may be written for SEi reaction Ea = al/(EHOMO(1) - ELUMO(2))+ a2

(62)

614 where a l and ae are coefficients for given series of organic reagents. For the reactions discussed compound 1 is the MOH group of inorganic oxide surface and compound 2 is an organic compound. Actually, the most relationships are observed between the reactivity of organosilicon compounds toward OH groups of hydroxylated silica surface and ELUMO (electron affinity (EA) with an opposite sign) or p a r a m e t e r s which include the EA value for these compounds [137]. For example, the following tendency is observed between activation energy of organosilicon compounds interaction with OH groups of hydroxylated silica surface corrected for preliminary adsorption complex formation, and their electron affinity (EARsix) Ea = - (1.5 + 0.9) 105/(IPsioH - EARsix) + 251 + 84 kJ tool -1

(63)

where IPsioH is the ionization potential of silica surface cluster of OH group (IPsioH = 9.4 eV). As the Ea value is controlled by charge transfer, the choice between SNi and SEi mechanisms can be done by comparison of reagents electronegativity (~ = IP + EA). If inequality ~M-OH> ~RX is satisfied, t h a n the organic compound is a nucleophil, but at ;~M-OH< ~RX, this compound is an electrophil. With allowance for these inequalities and the ionization potential (IP) and EA values of organosilicon compounds, and C, B, Ti, Ge, Sn halides ()~RX = 11 - 14 eV), and clusters of OH groups of silica surface ()~M-OH= 8.8 + 11.7 eV), it follows t h a t these compounds are electrophils in the reactions with OH groups. The ~RX values for ammonia, aliphatic amines, aliphatic alcohol's, phenol and water fall within the range 5.9 + 8.6 eV. According to the first inequality these compounds are nucleophiles in substitution reactions with the SiOH groups. The crude estimation for ability of inorganic oxide to interact with an organic compound as acid or base can be found also from comparison of mean electronegativities of these solids (Table 8) and gas-phase reagents. It was found t h a t the activation energy of organosilanes ((CH3)3SiX) interaction with silica OH groups decreases with lowering the Si-X bond energy (Esi-x) and with enhancement of the inductive coefficient of the substituent X (~I) Ea = 0.29 Esi-x - 274.5 ~i + 62.9 kJ mo1-1

(64)

Then above conclusion about SEi mechanism for the reactions agreed with sign of coefficients in Eq. 64. The main p a r a m e t e r s determining the organic compounds ability to interact with OH groups of inorganic oxide surfaces are bond dissociation energies of reagents and products as well as the energies of frontier orbitals of the reagents. So, simplified Marcus-like equation for the activation energy of the SEi reaction at varying the substituent X in (CH3)3SiX compounds was found [24] Ea = 0.08 (IPHx - EARsix) + 0.025 (DRsi-x - DHX) - 24.7 kJ mo1-1

(65)

615 where DRSi-X and DHX a r e the Si-X and H-X bond dissociation energies in the gaseous reagent and product. First term in this equation takes into account the relative redistribution of electron density between gas-phase species and surface groups for series of species. Several linear relationships for change of Ea in SEi, SNi and AdEi reactions by varying the structure of organic reagents RX and sites of inorganic oxide surfaces (M-OH, M-O-M and M-Y) can be derived using simple thermodynamic cycle and taking into account the difference in the electronegativities of reagents and products [24]. These equations present the combination of kinetic (first term in right side of Eq. 65) and thermodynamic (second term in this Eq.) contributions to the height of the activation barrier for surface reactions. For example, variation of substituents R in RX changes the activation energy for SEi reaction as follows AEa =- a3(EARx + EAROM) + a4(DR-x - DR-OM)

(66)

The change in influence of metal M on the reactivity of M-OH group varies as AEa = a5(IPMoH - EARoM) + a6(DMO-H - DR-OM)

(67)

In the case of interaction RX compound through SNi mechanism with surface MOH group the activation energy changes at varying the substituent X as AEa = aT(IPRx - IPMR) + as(DR-x - DM-R)

(68)

and effect of metal M on the reactivity of M-OH group in this reaction leads to following variations of activation energy AEa = ag(IPRx - EAMoH) + al0(DM-oH - DR-X)

(69)

The reactivity of oxide surfaces grows after replacement of surface OH by more active Y group, where Y = C1, SOn, NH2, NR'R", etc. The change of activation energy for next SNi reaction at varying of this group Y is AEa = al I(EAMY + EAHY) +

a12(DM_Y -

DHY)

(70)

As it seen, the reactivity of surface groups M-Y is determined by electron acceptor ability of substituent Y and difference in strengths of M-Y and H-Y bonds. These factors may be reasons for weak reactivity of Si-F groups on the fluorinated silica surface towards typical organic nucleophiles (alcohols, amines) in comparison with more reactive Si-C1 groups of chlorinated silica. Difference between the activation energies of SEi and AdEi routes at varying substituent X in RX is

616 AEa = a13(IPMx - EARx) + a14(DR-x - DM-X)

(71)

As it follows from [174], the AEa values for AdEi route to the Si-O bond of SiO2 surface at the interaction with (CH~)3SiX (X = N3, NCS and NCO compounds, are 10 + 20 kJ mol 1 lower than the activation energies of SEi reaction of these reagents. From comparison of relations (65) and (71) it is obvious that the maximum AEa value can be observed in the case of lowest reactivity in the SEi reaction. This explains the observed maximum of AEa at the interactions (CHs)sSiNCO and (CHs)sSiN3 as compared with (CH3)3SiNCS. Because most of inorganic oxides-constituents of industrial solid aerosols are active catalysts in various processes of partial or complete catalytic oxidation of organic and inorganic compounds, the activation energy of these transformations is related to bond energy of oxygen in these oxides (Ebo) in accordance with known relationship [176] Ec - a15 Ebo + a16

(72)

As it follows from comparison of oxygen bond energies and activation energies of methane, carbon oxide and hydrogen complete catalytic oxidation, the catalytic activity of metal oxides in these transformations decreases in the following order Co304 > MnO2> NiO > CuO > Cr203> Fe2Os> ZnO > V20~> MgO > A12Os>> SiO2

The expressions presented here may be used as a basis for possible QSRRs between change of the activation barriers for reactions on the solid aerosol surfaces and descriptors of surface sites as well as gas-phase organic species.

5.4. C h e m i s o r p t i o n of o r g a n i c c o m p o u n d s on the i n o r g a n i c o x i d e s u r f a c e s and t h e r m a l d e c o m p o s i t i o n of the surface s p e c i e s Experimental aspects of temperature-programmed desorption with mass spectrometric registration of volatile reaction products as a tool of investigation of disperse oxide surface and reaction mechanisms has been discussed elsewhere [176-180]. Let us consider some experimental results for chemisorption of organic compounds on the inorganic oxide surfaces and for thermal decomposition of surface species. The type of oxide, chemisorbed substance, desorption product, heating rate (~) and temperatures of peaks maxima (Tin) in TPD spectra are presented in Table 9. The position and relative intensity of the peaks depend on the heating rate. In order to compare these spectra, desorption activation energies (Ed) (given in last column in Table 9) have been calculated using Tm and parameters by simple Redhead's peak maximum method [181] at normal preexponental desorption factor Vd = 1013 s 1.

617 Table 9 Products of organic compounds decomposition preadsorbed or bound to inorganic oxide surfaces, temperatures of peak maximum in their TPD spectrum and the apparent desorption energies estimated using the Readhead's method Product of ~(K s -1) decomposition

Substance

Oxide

CH3OH

SiO2 [186] 0.1

n-CnHgOH SiO2 [187] 0.067 C6Hs(CH2)2OH SiO2 [188] 0.1

CH3OH

TiO2 [185] 0.2

C2HsOH

0.2

n-C3H7OH

0.2

i-C3H7OH

TiOe[185]

0.2

CH3OH C=O H2C=O CH3OCH3 CH3C(O)H CH4 CH2CHCH2CH3 C6Hs(CH2)2OH C6HsCH=CH2 C6H6 C6H5 C6H5 C6HsCH3 CH3OH CH3OCH3 H2C=O C=O H2 CH4 C2HsOH C2HsOC2H5 CH3C(O)H H2 C4Hs CO2 n-C~HTOH C3H7OC3H7 C3H7C(O)H C~H6 H2 C=O i-C3H70H CH3COCH3 C3H6 He C=O COe

Tm (K)

Desorption energies (kJ tool -1)

763 723 473, 863* 323, 583* 323, 600* 923 823 573 743 823 823 573 390,645* 635 675 675 675 675 390, 590* 590 600 630 640 710 390, 575* 590 620 625 600 620 390, 490* 540 560 550 560 710

224 212 137, 254** 93, 170"* 93, 175"* 272 245 167 218 242 242 167 110, 185"* 182 194 194 194 194 110, 169"* 169 172 180 183 204 110, 164"* 169 177 179 172 177 110, 139"* 154 160 157 160 204

618 Table 9 Products of organic compounds decomposition preadsorbed or bound to inorganic oxide surfaces, temperatures of peak maximum in their TPD spectrum and the apparent desorption energies estimated using the Readhead's method Product of ~(K s -1) decomposition

Substance

Oxide

HCOOH HCOOH

ZnO [189] 5 ZnO [189] 5

C6HsOH

ZnO [136] 10 FeO[136] 10 SiO2 [191] 0.1

(CH3)eCHNHe

Fe203

0.1

MgO

0.1

CaO

0.1

A1203

0.1

cyclo-C6H11NH2 A1203[191] 0.1

Fe203[191]

0.I

HCOOH CO2 C=O H2 C6HsOH C6HsOH (CH3)2CHNH2 CH3CH=CH2 CH3CN CH4 CH3CN CH4 (CH3)eCHNH2 CH3CH=CH2 NH3 CH3COCH3 COz, H20 (CH3)2CHNH2 CH3CN CH4 (CH3)zCHNH2 HCN (CH3)2CHNH2 CH3COCH3 CH3CN CH4 C6H11NH2 NH3 C6H12 C6HsNH2 1,3-C6H12 C6HIINH2 NH3 c6H12 C6H10=O CH3CH=CH2 C6HsNH2

Desorption energies (kJ mol 1) 180 45 550 142 550 142 500 129 520 132 556 141 403,703* 116,206"* 803 236 843 248 843 248 698 204 698 204 423 122 523 152 523 152 523 152 638, 708* 186, 207** 413 119 548 159 548 159 413 119 883 260 433 126 588 171 548 159 548 159 433 125 598 174 598 174 673 197 653 191 423 122 513 149 513 149 523 152 513 149 533 155

Tm (K)

619 Table 9 Products of organic compounds decomposition preadsorbed or bound to inorganic oxide surfaces, temperatures of peak maximum in their TPD spectrum and the apparent desorption energies estimated using the Readhead's method Product of Substance

Oxide

[~(K s -1)

decomposition

Tm (K)

Desorption energies (kJ mo1-1)

718 210 COz HzO 533,698* 155,204"* C6HsNHz SiOz [192] 0.1 C6H5NH2 413,638" 119,186"* 443,603* 128,176"* A1203 0.1 C6H5NH2 423 122 CaO 0.1 C6H5NH2 423 122 MgO 0.1 C6H5NH2 433,553* 125,161"* Fe203 0.1 C6HsNHz 718 210 CO2 HzO 693 203 403,703* 116,206"* (CH3)3CNH2 SiO2 0.1 (CH3)3CNH2 843 248 (CH3)2CH=CHe NH3 843 248 443 128 A1203 0.1 (CH~)3CNH2 598 174 (CH3)2CH=CH2 NH3 598 174 Fe203 0.1 (CH3)~CNH2 433 125 573 167 (CH3)zCH=CH2 NH3 573 167 638,688* 186,201"* CO2 H20 638,688* 186,201"* CC14 AlzO3[197] 5 CC1 155 39 300 76 CO2 CFC13 5 CC1 130 32 130 32 CFzCI2 5 CC1 360 92 CO2 CF~C1 5 CC1 120 30 390 100 CO2 *Temperature of the second peak maximum in the TPD MS spectrum of this product. **Apparent desorption energy related to the second peak in the TPD MS spectrum of this product.

5.4.1. Oxygen-containing compounds The decomposition of alcohols has been widely used as the probe of the acid~ase properties of metal oxide catalysts [182]. While it is generally assumed that acidic oxides catalyze dehydration and basic oxides catalyze

620 dehydrogenation, recent studies have shown that the selectivity of alcohol decomposition on metal oxides is a function of reagent and surface structures. Oxides which act as solid bases to abstract protons from alcohols can also produce net dehydration products via alkoxide decomposition [183]. It was found that the decomposition kinetics and selectivity of alkoxides on the TiO2 (anatase) surface are dependent upon the alcohol structure [184]. The decomposition temperatures of surface alkoxides are in the order MeO > EtO > n-PrO > i-PrO. The desorption sequence for the decomposition products from ethanol and l-propanol TPD exhibited a common pattern: dialkyl ether (bimolecular interaction) - aldehyde (a-H abstraction) - olefin (~-H abstraction and oxygen deposition). For methanol TPD, aldehyde was followed by methane due to the absence of ~-H. The absence of dialkyl ether formation from 2-propanol TPD again suggested steric constraints on secondary alcohol adsorption and decomposition. The dehydration selectivity increased in the order MeO < EtO < n-PrO < i-PrO. The TPD MS and IR studies on TiO2 (rutile) single-crystal surfaces have suggested that product distributions in the reactions of primary alcohols and carboxylic acids are governed primarily by the coordination environment of individual surface cations. This model would suggest that the reactivity of adsorbed alcohols should be insensitive to bulk structure. In order to test this hypothesis, methanol, ethanol, and 2-propanol were adsorbed at room temperature on anatase and rutile powders [185]. The alcohols were dissociatively adsorbed to form alkoxides and surface hydroxyls. The alkoxide species were removed via two channels, recombination with surface OH groups at 400 K and decomposition at higher temperatures. Dehydration and dehydrogenation pathways are observed for all of the alcohols, with only the primary alcohols yielding bimolecular reaction products. The similarities in product distribution and peak temperatures from aliphatic alcohols on anatase and rutile, particularly with regard to the selectivity for diethyl ether formation from ethanol, indicate that the bulk crystal structure of the oxide does not have a significant influence on the reactions of these molecules. The products of methanol chemisorption on SiO2 surface pretreated at 1020 K were studied by TPD MS method in [186]. Dimethyl ether, acetaldehyde and carbon oxide were observed in the temperature range usual for the effective alcohol chemisorption on the silica surface (~650 K). Their formation may be explained by following reactions -SiOCH~ + CH3OH --+ -SiOH + CH3OCH3 2 -SiOCH3 -+ -SiOH + - S i l l + CH3CHO Thermal decomposition of the grafted methoxy groups in the range from 820 to 1070 K gives formaldehyde, methane, hydrogen and carbon oxide in accordance with supposed reactions

621 -SiOCH3-->-Sill + H2C=O 2 -SiOCH3 --> -SiOH + -Sill + CH4 + CO -SiOCH3 + H -~ -SiO + CH4 H2C=O ~ CO + H2 The decomposition reactions of alkoxide groups -SiO(CH2)3CH3 on dispersed silica surface were examined by TPD MS and quantum chemistry methods in [187]. Main desorption channel corresponds to the elimination of l-butene above 600 K. The TPD data suggest that this reaction is of first order and that it can be viewed as "unimolecular". Chemical transformations of phenylethanol bound to silica gel surface were studied by TPD MS method after the previous partial carbonization of the grafted groups [188]. The main decomposition products of the groups are C6H5(CH2)2OH, C6HsCH=CH2, C8H6, C6HsC6H5 and C6H5CH3. The formation of phenylethylene is due to the unimolecular decomposition of grafted groups with H transfer from CH2 group (nearest to aromatic ring) to O from SiOR group. The formation of benzene and biphenyl was explained by surface migration of intermediate phenyl radicals and H abstraction from surface groups and associative desorption, respectively. The adsorption and reactions of formic acid on the oxygen-terminated ZnO(001)-O surface have been studied by TPD MS and XPS method in [189]. Small amounts of formic acid dissociate at defect surface sites to yield surface formate (HCOO). The surface HCOO decomposes to yield nearly simultaneous CO2 (37 %), CO (63 %) and H2 TPD peaks at 560 K.

5.4.2. Nitrogen-containing compounds The chemical transformations of various amines adsorbed at oxide/vapour interfaces (SIO2, A1203, Fe203, MgO and CaO) were studied by IR spectroscopy and TPD MS methods in [190-192]. It was found that with primary aliphatic amines the main high-temperature reaction on oxides possessing Lewis-acid sites is nitrile formation, but secondary aliphatic amines additionally show CN bond breakages causing desorption of NH3 and propylene (isopropylamine), cyclohexene (cyclohexylamine) as dehydrogenation (cyclohexylamine -~ aniline; isopropylamine ~ adsorbed imine species), and C-C bond breakage (isopropylamine: adsorbed imine species --> desorption of CH4 and acetonitrile). At beam temperature surface complexes formed between aniline and t-butylamine on the oxide surfaces are similar to those in the case of other amines: there is hydrogen bonding and dissociative adsorption on SiOe, and formation of coordination bonds between amine molecules and Lewis sites on the other oxides. With increasing temperature, in the case of aniline only aniline itself desorbs, whereas when t-butylamine is used, in addition to the unchanged amine, isobutene and NH3 are detected as desorption products, indicating the occurrence of CN bond breakage. With amines examined oxidation reactions take place on the surface of Fee03.

622

5.4.3. A r o m a t i c h y d r o c a r b o n s Spectroscopic data about structure of surface complexes formed after adsorption of aromatic hydrocarbons on the metal oxide surfaces are absent. The detailed scheme of toluene transformations on oxide surfaces was proposed in [193] on the basis of reaction products of their partial and complete catalytic oxidation C6H5CH3 + 02 --->C6H5CHO + 02 ~ C6H5COOH --> C6H6 + CO2 C6H6 + H20 --->C6H50H + 02 ~ 0=C6H4-0 + 02 ~ C4H203 + 02 --> CO, CO2, H20 Also, benzophenone is formed in the oxidation transformations with participating two toluene molecules. 5.4.4. H a l o g e n - c o n t a i n i n g c o m p o u n d s The reactions of several atmospherically relevant halomethane compounds (CF3C1, CF2C12, CFC13, and CC14) with heat-treated 7-alumina powders at 1000 K have been investigated in situ FTIR spectroscopy in an attempt to assess the impact of alumina exhaust particles from solid-propellant rocket motors on stratospheric chemistry [194]. The powders were dosed at 100 K with halomethanes and then gradually heated to promote reaction; infrared spectra were recorded as a function of temperature. The spectral features, which appear at temperatures as low as 120 K, are attributed to adsorbed carbonate, bicarbonate, and/or formate species (COn, n = 2, 3), indicating that halomethanes dissociatively chemisorbs on heat-treated 7-alumina at temperatures below that of the lower midaltitude stratosphere. These COn species are stable in the range 250 - 400 K, depending on the degree of chlorination of the halomethane. The decomposition ~eactions apparently proceed through C-X (X = F, C1) bond rupture and AI-X and C-O bond formation and involve the participation of surface active sites such as A13§ ions and AlxOy clusters that are produced during dehydroxylation. The IR adsorption experiments provided no information about absorption features associated with AI-X or O-X bonds, and reaction products that are released into gas phase. Such information is crucial, however, to assessing the impact of surface-mediated decomposition of halomethanes on solid-propellant rocket motor-alumina particles on the stratospheric ozone cycle. The XPS studies suggested that under certain conditions, halogen-containing species are evolved from 7-alumina surfaces exposed to halomethane compounds at stratospheric temperatures [195, 196]. The reactions of CF3C1, CF2C12, CFC13, and CC14 with heat-treated y-alumina powders were studied using TPD MS and XPS methods in [197]. Desorbing species were monitored as a function of substrate temperature using a line-of-sight quadrupole mass spectrometer. Hydrogen chloride and halomethyl fragments, which are indicative of halomethane dissociative chemisorption were observed to desorb below 150 K. Carbon dioxide began to desorb between 240 and 320 K. The CO2 most likely arises from COn (carbonate and/or formate) species which are formed via the low-temperature dissociative

623 chemisorption of the halomethanes. In situ XPS analysis of heat-treated powders t h a t had been dosed at 150 K with halomethanes revealed the presence of both organic and inorganic forms of fluorine. Halogen uptake probabilities, which are estimated to be - 10 .5 from the data, increased as the degree of chlorination of the halomethane increased. These results indicate that halomethanes will probably decompose on solid-propellant rocket motor-alumina particles in the stratosphere, forming adsorbed A1-X (X = C1, F) and COn species and releasing gas phase HC1 and CFxCly fragments. However, the impact of these processes on global stratospheric halomethane and ozone concentrations is likely to be minimal. The localized depletion of halomethanes may occur in the vicinity of the exhaust plume of a booster rocket where particle loading is much larger. The adsorption and thermal decomposition of the polyperfluorinated ether (C2F5)20 on 7-A1203 has been studied by IR spectroscopy in [198]. This ether interacts with the isolated surface OH groups, forming an increasing number of associated OH groups from 150 to 600 K. Surface fluoroacetate and surface fluoroformate species are also formed from the thermal decomposition of the (C2F5)20 layer. At T > 300 K, the surface fluoroacetate converts to surface fluoroformate in accordance with following scheme (C2Fs)20~ds (150-200 K) -~ CF3COOads + FCOOads + HF + A1-F + OH (assoc) (C2F5)2Oads (200-300 K) -~ FCOOads + HF + AI-F + OH (assoc) CF3COOads (300-600 K) ~ FCOOads Oxidation of the ether also occurs on surface preadsorbed with pyridine, indicating t h a t Lewis acid A13+ sites, blocked by pyridine adsorption, are not involved in fluoroacetate and fluoroformate formation.

5.4.5. Organometallics XPS, FTIR spectroscopy and TPD MS methods were used to study the chemisorption and decomposition of trimethylaluminium (TMA) on silica under high vacuum [199]. By annealing series of silicas from 425 to 1573 K prior to TMS exposures at 300 K, the distribution of chemisorption products as a function of the relative concentration of various OH groups types was examined. It was proposed t h a t the monomethylaluminium surface complex and methyl groups bonded to silicon are the majority species on the surface at 300 K. Decomposition of the monomethylaluminium complex begins above 373 K and increases the population of methyl groups bonded to silicon on the surface. The methyl groups react to form methane, ethane and adsorbed hydrocarbon fragments. In addition, methyl groups also react further with the surface to form tetramethylsilane. Small amounts of gas-phase TMA and carbon-contaminated aluminosilicate surface are observed. The mentioned above organic compounds, products of thermal decomposition of the species preadsorbed or bound to inorganic oxide surfaces, temperatures of

624 peak maximum in their TPD spectrum and the apparent desorption energies estimated using the Readhead's method are listed in Table 9. From comparison of Eo data for products of halomethanes chemisorption on y-A12Oz in Table 9 it follows that Eo value for CO2 desorption from A12Oz surface increases as the C-C1 bond dissociation energy in the chemisorbed halomethane grows. This dependence is presented in Fig. 6. Also, the Ed value for desorption of primary aliphatic and aromatic amines from the AlaOz and Fe2Oz surfaces increases with reducing the ionization potential of the amine (Figure 7). This observation agrees with proposed mechanism of the amines adsorption on the oxide surface, including the adsorbate interaction with Lewis acid sites of the oxide surfaces. The amine desorption energy decreases as the acceptor number of oxide surface (KA from Table 7) reduces at transition from A12Oa to SiO2.

Ea (kJ mo1-1) 110

Ed (kJ mo1-1) 130 A1203 O

100 CF2C12 90

80

125

[ Cfl4o

70 300

319

338

357

Ec_cl (kJ mol-l) Figure 6. Plot of desorption energy of carbon dioxide from alumina surface vs C-C1 bond dissosiation energy in halometanes.

120 7

7.5

8

8.5

9 IP (eV)

Figure 7. Desorption energy of aromatic and aliphatic primary amines from oxide surface as a function of amine ionization potential.

The quantum chemical computations of energies for surface reactions taking place on the model clusters of active sites and grafted groups on inorganic oxide surfaces may be useful tool for prediction of activation energies for the examined decomposition reactions. The heat of unimolecular decomposition of n-butyloxy group grafted to the silica surface yielding to the surface OH group and 1-butene (260 kJ moll), computed by AM1 method is close to desorption activation energy

625 for this reaction from Table 9 (245 kJ mol-1). Also, the desorption activation energy of various gas-phase species in thermal decomposition of grafted phenylethoxy groups on the silica surface from Table 9 increases as the enthalpy of the reactions, computed by this method grows [188]. This agrees with expressions 57 and 58, derived for non-dissociative and associative desorption by using BOC-MP method. 5.5. P h o t o c h e m i c a l t r a n s f o r m a t i o n s w i t h p a r t i c i p a t i n g t h e o r g a n i c s on solid a e r o s o l s u r f a c e s

The photochemical reactions of organic pollutants adsorbed on solid aerosol surface with various small atmospheric species, as gases and free radicals in their ground or excited state are of great importance for heterogeneous atmospheric chemistry. The solid surfaces are classified into two categories: (i) nonreactive surfaces, such as silica or alumina which provide an ordered two-dimensional environment for effecting and controlling photochemical processes more efficiently than can be attained in homogeneous phase; (ii) reactive surfaces, such as titania or metal chalcogenides, which directly participate in photochemical reactions by absorbing the incident photon and transferring charge to an adsorbed molecule or by quenching the excited state of this molecule [200]. It is known that aromatic compounds adsorbed on nonreactive solids exhibit a longlived excited state which is beneficial in enhancing the efficiency of such photochemical reactions, as .their photoionization [201], bimolecular electron transfer between adsorbed molecules [202], hydrogen abstraction, etc. The dispersed quartz surface covered by aromatic compounds, e. g. anthracene or organic dyes, offers a scavenger of singlete oxygen 102 (lAg) formed after illumination of the surface by visible light during 2 - 10 minutes at low temperature (273 K or low) [203]. The desorption activation energy of the 102 species from such surfaces varies from 90 to 130 kJ mo1-1. The freezing out of the adsorbent at temperatures from 273 to 323 K gives release of the 102 species into gas phase. Because most of industrial solid aerosols consists of inorganic nuclei, covered by layer of polycyclic aromatic hydrocarbons mixture, the observed effect of the singlete oxygen conservation by dispersed particles may occur in the atmospheric heterogeneous processes, e. g. in depletion of the ozone layer. The presence of reactive subsurfaces is possible on the aerosol surfaces consisting such semiconductor nanoparticles as TiO2, ZnO, ZnS, CdS and other metal chalcogenides. Their role in initiating and controlling such surface transformations of organic compounds as oxidation of olefins, arenes, alkanes, amines, alcohols and the degradation of chlorinated organics, phenols, and haloaromatics has been reviewed recently [204-206]. The reactive oxygen species, such as free radicals HO2, RO2, OH, may be formed on the illuminated semiconductor surfaces in the presence of air via dioxygen reduction by a conduction-band electron in the presence of a suitable adsorbed species possessing electron donor properties. Also, by virtue of producing these radicals via photochemical processes and high concentration of reactive ozone in the

626 stratosphere, they may interact in ground or excited state with organics adsorbed on the solid aerosols via Eley-Rideal or Langmuir-Hinshelwood mechanism. Since ozone transformations have great importance in environmental atmospheric chemistry, let us consider its possible reactions on the solid surfaces. Despite the great importance in environmental chemistry, the ozone adsorption has been studied by IR spectroscopy at low temperature (70 K) on the silica, titania and alumina surfaces only in recent years [138,157,207]. Ozone molecules are shown to form weak hydrogen bonds with OH groups of silica and titania surfaces. Adsorbed ozone reveals comparatively high basicity, close to t h a t of CO and is bound to silica OH group r a t h e r via one of terminal oxygen atoms, t h a n via the central one. The isotope substitution experiments provide evidence for strong deformation of ozone molecules adsorbed on the titania surface, which are bound to titanium ions also via one of terminal oxygen atoms. Although alumina is more acidic t h a n TiO2, IR spectroscopy did not detect 03 species adsorbed on strong Lewis A13+ sites. However, 03 decomposition occurred and pyridine specific poisoning evidenced that such sites are involved in the decomposition. Quantum chemistry calculations confirmed such a result and specified that the remaining O specie resulting from the decomposition, in addition to physisorbed 02, was attached to the aluminum ion constituting the Lewis sites. The oxygen molecules would then gradually be desorbed. Whereas the 102 (lAg) species pruduced via ozone photodecomposition (~ < 1180 nm) possess slow reactivity toward organic molecules, the formed via this process more active oxygen atom O(1D) may abstract hydrogen atom from the organics, surface OH groups or adsorbed water yielding to reactive hydroxyl radicals. The subsequent interaction of these radicals with ozone, organics, carbon oxide or nitrogen oxide gives active species HO2, R, H, and HONO, respectively [208]. These species are responsible for main chemical transformations of organic compounds in the atmosphere. In addition, the ozone decomposition on the solid aerosol surface is possible via its basecatalyzed route (OH-) over basic surface sites: 03 + OH- -> O~- + HO2, with the rate coefficients varying from 70 to 370 1 mo1-1 s 1 [209, 210]. 6.

CONCLUSIONS

The considered here problems in heterogeneous atmospheric organic chemistry are far from such traditional fields as surface science which deals with surface processes on well-characterized solids, as single crystals of metals and their oxides, graphite, etc., industrial adsorption, where main interest is directed to various microporous and mesoporous solids, such as zeolites, activated carbons, silica gels, and industrial heterogeneous catalysis dealing mainly with supported catalysts containing noble metals, zeolites, etc. In contrast to these fields, the study of organic pollutants interaction with solid aerosol surfaces is of no interest for industry. However, the importance of this field for the health and environmental sciences is beyond question. The studies in the heterogeneous

627 atmospheric organic chemistry are initiated only in recent years and methods developed in the surface science and catalysis have given impetus to understanding those complex processes. It is clear that the reactivity of organic pollutants toward the solid aerosol surfaces is controlled by the surface chemistry of the individual constituents of the particles. Significant advances have been made recently in the field of the powders surface characterization due to the development of theory and techniques in inverse adsorption gas chromatography. The dispersive components of the surface free energy and acid/base components of specific interaction' contribution to the adsorption energy of test polar adsorbates on various parent and modified inorganic oxides and carbonaceous materials, determined by this method make possible to estimate the adsorption energies and relative concentrations of nonpolar and polar organic compounds on the complex aerosol surfaces. The quantitative "structure-activity" and "structure-reactivity" relationships between the descriptors of organic compounds, solid surfaces and their activity in adsorption equilibria and reactivity in surface reactions presented here may be used to predict the direction and relative rate for these transformations on solid aerosols. The modeling of industrial aerosols, such as fly ash microparticles through the use of chemically modified pyrogenic metal oxide microparticles of controlled composition and surface structure offers a clearer view of detailed mechanisms for the formation of these aerosols and their interaction with atmospheric organic species. This interaction is the main reason for appearance of secondary organic carbon in the solid aerosols. The study of thermal decomposition of adsorbed organic species or the grafted groups on the model microparticles by temperatureprogrammed desorption mass spectrometry provides the comprehensive information on kinetics and mechanism of the surface reactions and gives insight to the mechanism of formation of fly ash particles.

Acknowledgements Authors wish to thank Professor P. F. Gozhyk (Ukrainian Antarctic Center) for fruitful discussions. One of the authors (V.A.P.) is indebted to Swiss National Science Foundation (Grant 7UKPJ048657) for financial support.

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C o n t r o l of s u p e r c r i t i c a l g a s e s w i t h solid n a n o s p a c e - e n v i r o n m e n t a l aspects K. Kaneko Department of Chemistry, Faculty of Science, Chiba University 1-33 Yayoi,Inage, Chiba 263, J a p a n 1.

SUPERCRITICAL GASES AND THEIR IMPORTANCE

There are many important supercritical gases whose critical temperature is lower than an ambient temperature. 02, N2, NO, CO2, CO, H2 and CH4 are representatives of a supercritical gas near room temperature. 02, N2, CO2, CH4 and NO are intricated with important bioreactions. 02, H2, CH4 and CO2 play an indispensable role in energy technology. The control of CO2, C H 4 and NO is necessary to preserve our earth environment. Thus, these gases are essentially important to h u m a n and earth. Even inert gases such as Ar and Xe play an important role in modern technology. These supercritical gases have been deeply associated with gas separation, gas storage, catalysis, supercritical extraction, supercritical drying, pollution control and life science. The chemistry of supercritical gases should contribute to establish an energy-saving technology of high performance. Accordingly, control of a supercritical gas is strongly requested for establishment of the future technology. Adsorption of supercritical gases by micropores can offer a hopeful control method. Here micropores are the pores whose width is less t h a n 2 nm according to the IUPAC classification, which is shown in Table 1. In this article, we will use the term of nanospaces for micropores whose width is less than about 1 nm. In a nanospace the molecule-pore wall interaction is markedly enhanced, as mentioned later. The adsorption science for a supercritical gas with the nanospace can support the development of fluid chemistry and technology in future. Therefore, we need to understand the nature of adsorption of the supercritical gas by nanospaces. Also we must know molecular properties of these supercritical gases in order to search the best nanospace system for each fluid gas. The above-mentioned molecules have different properties and structures. 02, N2, NO, CO2, CO and H2 are linear molecules and their point groups are C~v and D~h ; only CH4 has Td symmetry. Although their molecular orbitals for valence electrons other t h a n H2 are composed of 2s and 2p atomic orbitals, occupation of

636 Table 1 Classification of pores Micropore

w 50 nm

(Nanopore is not recommended by IUPAC, but it is often used for pore whose width is less than 10 nm).

the molecular orbitals by electrons are different from one molecule to another, providing different intermolecular interaction energy, electrostatic s t r u c t u r e and magnetic strucuture, as shown below. Their molecular sizes and interaction energies are different from each other. L e n n a r d - J o n e s p a r a m e t e r s (~ff and ~ff ) give good m e a s u r e s of the size and interaction energy. H2 of ~ff = 0.292 nm is the smallest molecule, while CO2 of aff = 0.376 nm is the greatest of these molecules. However, aff, being the one-center L e n n a r d - J o n e s p a r a m e t e r , is obtained by the a s s u m p t i o n of the spherical shape and it is not the exact m e a s u r e of the molecular size. Each molecule has a different ~ff value; CO2 has the greatest interaction energy. NO and CO have a small p e r m a n e n t dipole moment. O2, N2, CO2 and H2 have the quadrupole moment, while NO and CO have the quadrupole m o m e n t too. Only CH4 has the octapole moment. Although the p r e d o m i n a n t intermolecular interaction stems from the dispersion interaction, these multipole m o m e n t s contribute to the intermolecular interaction, determining their m u t u a l orientation. 02 and NO are paramagnetic, whereas others are diamagnetic. Hence 02 and NO can interact with magnetically. Thus, these molecules have different properties. Table 2 s u m m a r i z e s physical properties of these molecules. We can use the physical property difference. Here Tb, Tc and Pc are the boiling t e m p e r a t u r e , critical t e m p e r a t u r e and critical pressure, di, qu and oc denote dipole, quadrupole and octapole moments. The units of di and qu are Cm and Cm 2, respectively. Although NO and CO have the quadrupole m o m e n t in addition to the dipole moment, their quadrupole m o m e n t s are ommitted, d i a and p a r a denote d i a m a g n e t i s m and p a r a m a g n e t i s m .

637 Table 2 Physical properties of i m p o r t a n t molecules Molecule H2

Tb K

Tc K

Pc MPa

(~ff nm

s~f/kB K

20.3

33.0

1.29

0.292

38.0

Multipole moment

Magnetism

qu

dia

+2.1.10-40 02

90.2

154.6

5.04

0.338

126.3

qu

para

-1.33.10-4o N2

77.3

126.2

3.39

0.363

104.2

qu

dia

-4.90" 10 -40 NO

121.4

180

6.48

0.347

119

di

para

0.158-10 -3o CO

81.6

132.9

3.50

0.359

110

di

dia

0.112.10-3o CO2

194.7

304.2

7.48

0.376

245.3

qu

dia

-14.9.10-40 CH4

2.

111.6

190.5

4.60

0.372

161.3

oc

dia

W H A T IS A S U P E R C R I T I C A L GAS ?

The state of m a t t e r is described in terms of the pressure P, the molar volume Vm and the t e m p e r a t u r e T. Then three phases of gas, liquid and solid are expressed by the P-Vm projection, as shown in Figure 1. The P-Vm projection indicates the presence of the coexistent region of gas and liquid in equilibrium, which is designated by ( 1 + g ). The broken line parallel to the abscissa of the P-Vm projection at a t e m p e r a t u r e T1 denotes the coexistent region. The coexistent region becomes narrower, as the t e m p e r a t u r e is raised. Finally the coexistent region is reduced to a mere point whose t e m p e r a t u r e is called critical t e m p e r a t u r e Tc. The critical t e m p e r a t u r e Tc is the m a x i m u m t e m p e r a t u r e at which a gas can be liquefied and above Tc liquid cannot coexist. Above Tc there is no vaporization curve and no distinction between liquid and gas. Therefore, we must distinguish the gaseous states above and below Tc. The term vapor is used to describe a gaseous substance when its t e m p e r a t u r e is below Tc and the vapor can be condensed to liquid by pressure alone. However, the gas above Tc, which is called a supercritical gas, cannot be liquefied even at quite high pressures. Vapor has own s a t u r a t e d vapor pressure Po. Then we can use the relative pressure P/P0 for description of adsorption of vapor. On the other hand, the supercritical gas

638

P 1

"'criticalpoint V ~ T V, Figure 1. Phase diagram.

has no s a t u r a t e d vapor pressure. The relative pressure expression cannot be used for description of adsorption of a supercritical gas. Recently not only liquid but also supercritical gas has been used as solvents [2]. The solvent power of a supercritical gas increases with density at a given t e m p e r a t u r e and it increases with t e m p e r a t u r e at a given density. A supercritical gas exhibits physicochemical properties of an intermediate between a liquid and a gas. The relatively high liquid-like density at high pressure affords good solvent power and the mass transfer in the supercritical gas is rapid relative to a liquid. Also the extremely low value of surface tension of the supercritical gas allows better penetration into the sample matrix relative to liquid solvents. The critical t e m p e r a t u r e and pressure of CO2 are 304.21 K and 7.477 MPa, respectively and thereby CO2 has been widely served in supercritical fluid technology. However, the fundamental u n d e r s t a n d i n g of the supercritical state is not necessarily sufficient yet. The supercritical state in the bulk phase has been studied with small angle Xoray scattering, showing the presence of inherent clusters around the critical point for CO2 [3]. The researches on the bulk supercritical state should stimulate the adsorption study on supercritical gases. 3.

NANOSPACE SYSTEMS

The most representative microporous solids are zeolites and activated carbons. Both have different pore structures and adsorptive properties each other. Recently new porous solids have been developed to accelerate the progress in nanopore fluid chemistry. There are two types of pores of intraparticle pores and interparticle pores [4]. The intraparticle pore is in the primary particle itself, while the interparticle pore originates from the interparticle void space, although there is an ambiguous

639 distinction between both types of pores in some systems. Zeolites have welldefined intraparticle pores which arise from the intrinsic crystalline structure [5]. The pore geometry and pore connectability are evaluated by their crystal structure. Aluminophosphates [6] also have cylindrical intraparticle pores inherent to the crystal structures. Even the hydrophobic phosphates [7] having the straight pore of triangular column were synthesized. On the other hand, a new family of mesoporous zeolites [8-11] such as MCM (Mobil Composition of Matter) or modified kanemite FSM (Folded Sheets Mesoporous Material), which have straight cylindrical mesopores, were developed recently. Furthermore, the mesopore size can be controlled by the surfactant molecular size. These mesoporous silica has a regular honeycomb structure, although the pore walls are noncrystalline. The mesoporous silica provided a new problem that the adsorption hysteresis depends sensitively on the mesopore size [12,13]. The classical capillary condensation theory cannot explain the dependence of the adsorption hysteresis, but Inoue et al. gave a new approach to this problem with the extended Saam-Cole theory [14]. The carbon nanotube [15] has also the intraparticle pore stemming from the intrinsic crystalline structure; the tube-wall is composed of graphitic structures. However, the so-called carbon nanotube has a both end-closed cylindrical pore. On the other hand, the catalytic method can produce the one end-open pore in which gas adsorption is available [16]. The CVD technique using the template porous solids produces mesocarbon tubes [17]. Activated carbons are the most popular adsorbent [18-20]. Activated carbons are mainly composed of micrographitic units. The edge carbon atoms of the micrographite are more reactive than carbon atoms on the basal plane, developing pores along the basal plane of the micrographite. Activated carbon fibers (ACFs) [21] have only uniform micropores, while the conventional granulated activated carbons have a wide pore size distribution from micropores to macropores. In particular, pitch-based ACFs have less amount of surface functional groups and the pore width can be well controlled. Superhigh surface-area carbons [22,23] obtained by KOH activation have considerably uniform micropores whose pore width is greater than that of ACFs. The activated carbon film from Kapton film has the oriented structure of slit micropores [24]. The carbon aerogel has uniform mesopores and micropores can be donated to the carbon aerogel [25,26]. Also graphite intercalated compounds can offer micropores [27]. In addition to the above porous systems, pillared clay minerals [28,29] and pore-size controlled glasses [30] can be available for controlling a supercritical gas. Microporous BN is also an attractive solid [31]. An organic metal complex is a new hopeful nanopore system of which width can be variable [32]. As these porous solids have different pore geometry and chemical nature (Table 3), we can choose the best fit nanopore system for control of each fluid.

640 Table 3 Nanoporous s y s t e m s Surface component

Compound Zeolite A l u m i n o p h o s p h a t e ALPO

Si, A1, O

Pore shape nm

Pore width nm

cylinder, cage

0.3 - 1

cylinder

0.8-1.3

A1,P, O

[SAPO: Si, A1, P, O; TAPO: Ti, A1, P, O; FAPO: Fe, A1, P, O, etc] Aluminiummethyl phosphonate A1, CH3PO3 cylinder < 1 triangle prism Mesoporous zeolite Si, O cylinder 2 - 10 [A1, Zn, Ti, Zr, W, Pb etc can be doped] activated carbon fiber carbon aerogel activated carbon aerogel pore-oriented carbon film carbon n a n o t u b e carbon mesotube graphite i n t e r c a l a t i o n comp. microporous BN pillared clay porous glass organic m e t a l complex

4.

C C C C C C C, Ketc B, N, H Si, A1, O Si, O Organic group

slit i n t e r g r a n u l a r void slit + void slit cylinder cylinder slit slit slit cylinder void

0.6 - 1.3 5 - 30 1, 5-30 < 1 > 2 30 - 200 < 0.5 < 1 > 0.5 5 - 104 variable

DEEP POTENTIAL WELL OF NANOSPACE

We will discuss the interaction of a molecule with the graphitic slit pore of the micropore model of activated carbon. The interaction b e t w e e n a molecule and a surface atom as a function q)(r) of the distance r between t h e m can be expressed by the L e n n a r d - J o n e s potential, (I)(r) = 4~sf [(asf/r) 12 - ( a s f / r ) 6 ]

(1)

where ~sf and asf are the well depth and effective d i a m e t e r for the moleculegraphitic carbon atom. These cross p a r a m e t e r s are calculated according to the Lorentz-Berthelot rules, ~sf = (~ss ~ff )l/e ; asf = (ass + aff )/2. Here, (ass , ~s) and (aff, ~ff ) are the L e n n a r d - J o n e s p a r a m e t e r s for a surface atom and a molecule, respectively. The interaction potential (I)(z) for a molecule and a single graphite slab is given by the Steele 10-4-3 potential [33],

641

(I)(z) = 2rtPCg sf (~2sfA{(2/5)((~sf/z)10 - ((~sf/z }4 _(~4sf/bA(0"61A + z)3 ] }

(2)

where z is the vertical distance of the molecule above the surface, A is the separation between graphite layers (=0.335 nm), pc is the n u m b e r density of carbon atom in a graphite layer (=1i4 /nma). As the micropores of activated carbon can be approximated by the slit spaces between the p r e d o m i n a n t basal planes of nanographitic units, the whole interaction potential ~(Z)p of a molecule with the micropore of an inter-graphite surface distance H can be given by eq. (3).

|

: |

(a)

Consequently, we can evaluate the potential profile of the molecule adsorbed in the graphitic micropore. Here H is not the effective pore width w determined by the adsorption experiment. The difference between H and w is a function of ~sf and ~ff [34]. (4)

H - w = 0.85~sf - ~ff

5O0

-500 1

s

s'

~" -1500

-2500

'

'

I

,

,

-0.4

-0.2

0 z/rim

0.2

0.4

Figure 2. The potential profile pore width w.

I

of a N2 molecule in the

graphite slit pore as a function of the

In the case of the Ne-graphitic slit pore system, w is equal to H - 0.24 nm. Figure 2 shows potential profiles of N2 in a slit-shaped graphite pore as a function of w using the one-center approximation. Here, the molecular position in Figure 2 is expressed by a distance z from the central plane between two surfaces. The broken line shows the potential profile for the single graphite surface. The

642 potential becomes deeper with decrease in the w value. The potential profile has double minima for w > 0.6 nm, but a narrower pore has a single potential minimum. Thus, micropores have stronger adsorption fields t h a n flat or mesoporous surfaces. The depth of the potential well for vapor is enough great to give the Type I adsorption isotherm, i.e. an enhanced adsorption at a low relative pressure range, being characteristics of micropore filling [35]. Activated carbon has a b u n d a n t micropores and their adsorption field can be approximated by the above mentioned graphite-slit space model. The depth of the potential well of nanospace is not enough for a supercritical gas to be sufficiently adsorbed, but chemical modification of the pore-walls can deepen the potential well for the supercritical gas, leading to a marked micropore filling. Furthermore, supercritical gas molecules tend to be adsorbed in micropores of the deep potential well and their intermolecular interaction is enhanced to form an organized structure, as if molecules were compressed by a high pressure. For example, when adsorbed molecules of supercritical gas form the dimer in a chemically modified micropore, the interaction profile of the dimer with the micropore becomes completely different from that of the monomer. Hence, the adsorption property for the supercritical gas can be changed dramatically by use of a chemically modified micropore.

0

F U N D A M E N T A L P R O B L E M S IN A D S O R P T I O N OF S U P E R C R I T I C A L GAS

The supercritical gas which has no concept of the s a t u r a t e d vapor pressure cannot be sufficiently physisorbed on the flat surface, macropores and even mesopores with physical adsorption. The adsorption of vapor by micropores, which is called micropore filling, is enhanced at a very low pressure region due to overlapping of the molecule/pore-wall interactions. The deep potential well of the micropore gives rise to adsorption of the supercritical gas to some extent. Still the micropore cannot adsorb a b u n d a n t amount of the supercritical gas. That is, the quite narrow micropore whose width is fit for the size of the adsorbate molecule has a very deep molecular potential well and is effective even for adsorption of the supercritical gas. However, the amount of adsorption by such a narrow micropore is limited to much less t h a n the micropore volume due to the diffusion problem; molecules adsorbed near the entrance in the micropore are bound too strongly to migrate to the inner pore position. Accordingly, the supercritical gas is not an objective gas for a predominant micropore filling [36]. In case of micropore filling of vapor, the potential well of the micropore whose width is even more t h a n trilayer thickness of the molecular size is enough deep for vapor molecules to be sufficiently filled at a low pressure region without any obstacle for the intrapore diffusion. The analytical method for physical adsorption of a supercritical gas by micropores is much less advanced compared with physical adsorption of vapor

643 [37-39], although molecular simulation [40-42] based on the Lennard-Jones is effective for description of adsorption of supercritical gases by micropores. However, we need a simple description method like Dubinin-Radushkevich (DR) equation for adsorption of supercritical gas in micropores, as given by eq.5 [43]. W = Wo exp[-(A/~ E) 2 ]

A : RT ln(P/P 0), E = 13E o

(5)

Here, W0 is the pore volume, E the energy constant, Eo the characteristic adsorption energy and 13 the affinity coefficient. The DR equation includes the saturated vapor pressure Po and thereby it cannot be applied to micropore filling of supercritical gas. As the interaction potential at the mid-point of the slit-pore is the deepest in the pore of w < 0.6 nm and it is not seriously different from the double minima even for w > 0.6 nm, molecular potential for a molecule in the micropore can be approximated by the potential at the mid-point of the slit-pore. The molecular potential indicates the presence of the inherent micropore of the volume of WL for each adsorptive. Here the inherent micropore must have a sufficiently strong molecular field in comparison with the thermal energy at a measuring temperature. The inherent micropore volume WL for vapor molecules is almost equal to the micropore volume Wo obtained from N2 adsorption at 77 K (Gurvitch rule). WL for a supercritical gas which depends on the molecule-pore interaction can be evaluated as the saturated amount of adsorption from the Langmuir plot and WL is less than Wo in general. The supercritical gas is transformed into a quasi-vapor in the micropores of which pore volume is WL. Then, the quasi-saturated vapor pressure Poq can be defined for the quasi-vapor. Both of WL and Poq are determined experimentally, as described using the following supercritical DR equation [37]. [ln(WL/W 0 )]1/2 : (RT/~E 0 )0n Poq _ l n P )

(6)

The plot of [ln(WL/Wo)] 1/e vs. In P leads to both of Poq and ~Eo. This supercritical DR plot is quite useful to obtain the important information on adsorption of supercritical gas. This supercritical DR equation can describe the adsorption isotherm of the supercritical gas using the concepts of Poq and WL, which are related to the intermolecular interaction of supercritical gas in the quasi vapor state and the molecule-pore interaction, respectively. Also adsorption isotherms at different t e m p e r a t u r e s can be reduced to a single isotherm with the aid of these reduced quantities of WL/W0 and P/Poq Hence we can predict the adsorption isotherm at an arbitrary temperature by use of the reduced isotherm.

644

D

CONTROL OF MICROPORE FILLING AND NANOSPACE REACTIVITY OF SUPERCRITICAL GASES

As micropore filling is governed by both geometry of pores and chemical nature of the micropore-wall, we can control it with chemical modification of the micropore-walls. The chemical modification with substances having a weak chemisorptive activity for molecules is ~ffective for enhancement of micropore filling of a supercritical gas; this is called chemisorption-assisted micropore filling [44]. A marked enhancement of micropore filling of supercritical NO and CH4 with the chemical modification is described in this article. Micropores have stronger adsorption fields due to the deep potential well than flat or mesoporous graphitic surfaces. We can estimate the effective pressure from the potential profile; molecules confined in the slit-pore of I nm in width are presumed to be exposed to the high pressure of 100 MPa. Therefore, the graphitic micropore can offer the high pressure field from the macroscopic view. The quasihigh pressure effect was evidenced in the disproportionation reaction of the NO dimer and hydrate mediated micropore filling for NO and CH4, as shown later. Hashimoto et al. [45] reported the presence of the quasi high pressure effect even in an electrochemical reaction using ACF.

6.1. Iron o x i d e - d i s p e r s i o n i n d u c e d m i c r o p o r e filling of s u p e r c r i t i c a l NO The critical temperature of NO is 180 K and NO is a supercritical gas at ambient conditions. Almost all microporous adsorbents cannot sufficiently adsorb supercritical NO, although NO of the representative atmospheric pollutant is desired to be removed with a good adsorbent. Table 4 summarizes NO adsorptivities of activated carbon, zeolite and silica gel at 303 K. Although activated carbon can adsorb more NO than representative zeolites and silica gels, the maximum adsorption does not cope with the saturated adsorption which can be estimated from the pore volume. Zeolite is not a good adsorbent for NO regardless of presence of micropores. Mesoporous silica gel is also not fit for adsorption of supercritical NO. Only activated carbon is considerably effective for NO adsorption due to the assistance by the surface functional groups. Hence the control of surface chemistry of activated carbon is quite essential with the relevance to control of the dimerization of NO. An NO molecule has an unpaired electron and gaseous NO shows paramagnetism. It is well-known that NO molecules are dimerized and show diamagnetism at the condensed phase at low temperature [46]. Consequently, we can understand the molecular state of NO molecules adsorbed in micropores of ACF by the magnetic susceptibility measurement. ACF having less surface functional groups shows diamagnetism, giving a reliable result. The magnetic susceptibility ~ of ACF with adsorbed NO was negative near room temperature irrespective of the adsorption of NO which has a large paramagnetic (positive) susceptibility. The ~ value calculated from the c values of both gaseous NO (of the same amount as adsorbed NO) and ACF was positive and decreased with the measuring temperature. Thus, the calculated temperature dependence of ~ was

645 clearly different from the observed one. This is because the adsorbed NO does not exhibit p a r a m a g n e t i s m ; the negative )~ of ACF with adsorbed NO arises from the dimerization of NO in micropores. The fraction of dimers determined from the a m o u n t of NO adsorption, the observed magnetic susceptibility and the diamagnetic susceptibility of the NO dimer in the literature is 0.98 at 298 K and it is still about 0.9 even at 373 K [47-49]. If we can accelerate the dimerization of NO, NO adsorption can be enhanced, because NO dimer is vapor.

Table 4 Amounts of NO adsorption of activated carbons, zeolites and silica gel at 303 K under the equilibrium NO pressures of 13 and 80 kPa Surface NO mg g-1 area adsorbed 80kPa Substance m2g -1 13kPa 1100

15

28

coconut shell-based activated carbon

860

17

47

molecular sieve carbon (pore width: 5/k)

500

28

60

1100

15

28

860

17

47

g r a n u l a t e d coal-based activated carbon

g r a n u l a t e d coal-based activated carbon coconut shell-based activated carbon

500

28

60

cellulose-based activated carbon fiber(CEL-ACF)

1400

36

65

pitch-based activated carbon fiber

1530

10

55

molecular sieve carbon (pore width: 5/k)

polyacrylonitrile-based activated carbon fiber

830

65

115

iron hydroxide dispersed CEL-ACF

820

120

150

1070

260

320

iron oxide dispersed CEL-ACF molecular sieve 3A

-

0.6

molecular sieve 4A

-

0.6

molecular sieve 5A

-

5

24

Na-mordenite

-

4

22

molecular sieve 13 X

-

2

9

iron hydroxide dispersed molecular sieve 13X

-

5

12

Silica gel

680

2 6

3

5

iron hydroxide dispersed silica gel

600

14

20

Iron oxide*

-50

20

50

*) NO is chemisorbed on iron oxide.

646 NO is weakly chemisorbed on transition metal oxides near room temperature. The electrical conductivity of n-type iron oxides decreases very rapidly upon chemisorption of NO near room temperature [50]. Although NO forms the nitrogen oxide with surface oxygens of iron oxides, almost chemisorbed NO molecules are desorbed by evacuation. The weak chemisorption of iron oxide can assist micropore filling of supercritical NO on microporous carbons; the dispersion of iron oxides near the entrance of micropores of ACF should enhance physical adsorption of supercritical NO due to the concentration increase of NO near the dispersed oxide particles. Figure 3 shows adsorption isotherms of NO on iron oxide-dispersed ACF (Fe-ACF) at 303 K. As iron oxyhydroxide decomposes to iron oxide by heating, iron oxide-dispersed ACF was prepared from iron oxyhydroxide-dispersed ACF by heating at different temperatures. The NO adsorption depends on the preheating temperature of dispersed iron oxyhydroxides. Heating of iron oxyhydroxide dispersed ACF at 873 K for 15 h in vacuo gives the greatest amount of adsorption. The dispersed iron oxides was characterized by EXAFS spectroscopy, showing the presence of ultrafine iron oxides. Fe-ACF can adsorb great amount of NO (maximum: 320 mg/g-adsorbent) in the dimer form at 303 K. Thus, dispersion of ultrafine iron oxide particles enhances markedly NO adsorption. The adsorption isotherm shows a remarkable hysteresis; the adsorbed NO cannot be removed by evacuation with a high

300 (

1 : - . O " ~

200

,~ 100

t

t

i

I

I

I

NO pressure / ~a Figure 3. The adsorption isotherms of NO on iron oxide -dispersed ACF at 303 K as a function of heating temperature of dispersed iron oxyhydroxides in vacuo. (A, A) : 573K, (V,V): 673 K, (O,Q) : 773 K and ( ~ ,'00:823 K. Solid and open symbols denote adsorption and desorption branches, respectively.

647 vacuum pump at 303 K, but it can be desorbed with the ultrahigh vacuum system. This irreversibility arises from dimerization of NO in micropores. As NO dimer is vapor at an ambient temperature, supercritical NO can be adsorbed by micropores through the dimerization with the aid of the electronic interaction and magnetic perturbation due to the high spin Fe a§ ions in the dispersed oxide [51-53]. The importance of the magnetic interaction was evidenced by application of the strong external magnetic field. The application of the magnetic field of 1 T to the adsorption system increased instantaneously NO adsorption on ACF by 1% [54]. Not only the magnetic perturbation but also a weak chemisorptive mechanism should be associated with the enhancement of the NO micropore filling.

160

~" 140 0 Z 120 o

lOO I

I

I

I

0

1

2

3

DopedTi IFe 1 %

Figure 4. The saturated amount of NO adsorption per unit micropore volume at 303 K against the amount of Ti dopant.

Doping of Ti 4§ increases the electrical conductivity of n-type iron oxyhydroxide fine crystals or iron oxide thin film due to formation of a quasi-free electron [55,56]. Hence, the mixed valence formation in the dispersed iron oxides with doping of Ti 4§ enhanced the micropore filling of NO [57]. The adsorption isotherm of NO on Ti-doped iron oxyhydroxide dispersed ACF at 303 K is Langmuirian, which is described by the Langmuir equation. The saturated amount of NO adsorption (WL) from the Langmuir plot as a function of the amount of Ti dopant (Ti/Fe: 0 - 3.6 %) is shown in Figure 4. Ti doping increases markedly WL; Ti doping of 3.6 % enhances WL by about 40 %. Ti doping should be associated with NO dimerization. The NO adsorption isotherms are described by the supercritical DR equation (eq.6). the quasi-saturated vapor pressure P0q can be defined for the

648 quasi-vapor. The determined Poq was in the range of 215 to 480 kPa. Ti doping lowers the P0q value so that the NO micropore filling is enhanced. The supercritical DR plot gives the isosteric heat of adsorption at the fractional filling (~ of e -1 (qst,~=l/e). The qst,~=l/e values (23-28 kJ mol 1) are higher t h a n the dissociation enthalpy of the NO dimer (12-16 kJ mo1-1 ) in the condensed bulk phase at a low temperature. Thus NO dimers are stabilized in the micropores of these ACF samples. If the supercritical DR equation is correctly applicable to the description of the micropore filling for supercritical NO, all adsorption data of different doping samples must be expressed by a single reduced adsorption isotherm having the abscissa of the relative pressure P/P0q. Figure 5 shows such a reduced adsorption isotherm for NO vaporized in the micropore. Almost all the observed points form one curve. Thus, the supercritical DR equation can describe well the micropore filling of supercritical NO. The above facts show a typical chemisorption-assisted micropore filling of supercritical NO.

12o

,N.

80 o.,-~

4o

9 Z I

0

I

0.2

I

0.4

P/P0q(NO) Figure 5. Reduced adsorption isotherm of supercritical NO for iron oxyhydroxide dispersed ACF having different amounts of Ti dopants at 303 K. Ti/Fe %: o none, A 0.2, [] 1 and * 3.

6.2. R e d u c t i o n o f NO to N2 in n o b l e m e t a l - t a i l o r e d n a n o s p a c e Oxides of nitrogen NOx is the inevitable by-products of high temperature combustion and the representative atmospheric pollutants. Increasing automobiles have emitted greater amounts of NOx, giving rise to a serious atmospheric pollution problem in urban areas in particular. Although NO2 can be

649 easily absorbed in soils or dissolved in surface water, a diluted NO is kinetically extremely stable in the absence of suitable catalysts in atmosphere. Consequently an ~fficient removal of NO or reduction of NO to N2 has been desired. There are active studies on catalytic reductions of NO to N2 all over the world [58-60]. The most noticeable catalyst for the decomposition of NO to N2 is Cu-ZSM-5 developed by Iwamoto et al. [60]. This Cu-ZSM-5 decomposes NO into N2 and 02 near 750 K, but coexistent SO2 strongly poisons the catalytic activity. Recent efforts are done on the development of the catalyst for selective catalytic reduction with N-free reductants [59]. NO is quite i m p o r t a n t as not only the atmospheric pollutant but also a biological molecule; recently it is elucidated that NO plays an important role in biomolecular organisms of n a n o m e t e r size [61]. Thus reactivity of NO in nanospace has gathered much attention. The nanospace can work as the high pressure field, as mentioned before. The disproportionation reaction of (NO)2 given by Eq. (7), 3(N0)2 = (N02)2 + 2 N 2 0

(7)

is known as the high pressure gas phase reaction above 20 Mpa [62]. NO molecules dimerized in the micropore of ACF at a subatmospheric pressure of NO give rise to the high pressure disproportionation reaction of the NO dimer in micropores [63]. Furthermore, the produced N20 is reduced to N2 at 423 K with the aid of dispersed transition metal oxides [64]. However, the NO reduction reaction is very slow; it takes more t h a n 10 h. Ru fine particles exhibit a high catalytic activity for NO. Ultrafine Ru particles can be dispersed in micropores of ACF. The Ru fine particle-dispersed ACF is designated Ru-ACF (Ru content: 4 wt. %). The surface areas and micropore volumes determined by N2 adsorption are 1130 m2g-1 and 0.51 mlg-' for Ru-ACF and 1100 m2g-1 and 0.52 mlg -1 for ACF, respectively. The average slit-shaped widths of Ru-ACF and ACF are 0.94 and 0.90 nm, respectively. The reaction extent was determined by the compositional changes of the gas phase with FT-IR and Mass spectrometers. Figure 6 shows the FT-IR spectral change of the gas phase over Ru-ACF at 303 K. The intensity of NO at 1876 cm -1 decreases rapidly; it becomes less t h a n 15% of the initial intensity after 9 min. On the other hand, the bands of NO2 and N20 appear at 1618 cm -1 and 2224 c m -1, respectively. As the absorption intensity of NO is noticeably weak compared with that of the NO2 and N20 bands, the concentration of produced NO2 and N20 after 18h corresponds only less than 0.3% of the residual NO; evolution of CO2 was much less t h a n that of NO2 or N20. IR cannot detect N2, then the whole gas after 10 min and 18 h was analyzed by the mass spectroscopy. The mass analysis elucidated the formation of N2 corresponding to the 77 % yield. Hence, NO is rapidly changed into N2 over RuACF at 303 K. Figure 7 shows the change in the NO concentration ratio vs the initial concentration determined from the corrected IR absorption band intensity of NO at 1876 cm -1. NO is rapidly reduced to N2 at 303 K and 323 K and then the

650

I.) o

a O raej


CCM can a m o u n t to a| = 1.0 g per I g of the adsorbent. Our

686 data presented in [100] show that 1 g of Pyzhevsk montmorillonite in Na-form can extract 0.8 g of OP-7 agent from an aqueous solution. The adsorption capacity of the best grades of activated carbon for non-ionic SAS does not exceed 0.1 g/g [28,100], with their cost being much higher than that of clay minerals.

a, g/g

1 9

0.3

2 0.2

0.1

I

I

I

0.2

0.4

0.6

I

Ce, %

Figure 3. Adsorption isotherms of oxyethylated isooctylphenyol (OP-7) ta Ca-form of." (1) Pyzhevsk montmorillonite, (2) Cherkasy montmorillonite and (3) Cherkasy hydromica.

687 The capability of montmorillonite (bentonite) clays for efficient adsorption of non-ionic surface-active substances (NSAS) from water was utilised in the development of industrial technology for the treatment of stratal water at gas production fields [100]. When the NSAS solution is prepumped into gas wells to reduce the collector properties of the stratum, some portion of the NSAS is carried away to the surface along with stratal water. This water sometimes contains up to I 000 g/m 3 of NSAS, i.e. of OP-7, prevocell EO, prevocell WOF-100, ditalan OTS-45, syntanol DS-10, etc. To remove these from water, commercial bentonite powder can be used. Table 13 generalises the industrial test results of the stratal water treatment of Pynyanu gas deposit (Ukraine, L'viv province) to remove prevocell EO using Cherkasy bentonite powder [100]. It is seen that minimum permissible concentration (MPC) levels of NSAS (0.5 g/m 3) can be achieved by one-stage stratal water treatment with Cherkasy bentonite. In this case the sorbent demand is 17-20 kg per 1 kg of the NSAS removed.

Table 13 NSAS concentration (C) in initial stratal water and water after treating with the aid of bentonite Sorbent consumption, kg/m 3 1.4 4.8 17.0

C, g / m 3 Initial water 48 270 950

Treated water 0.5 0.4 0.5

The order of technological operations is as follows: stratal water which contains NSAS is delivered from the collection tank to the mixer into which bentonite powder was fed. The mixture is stirred until clay is completely dissolved. Then the pulp is moved into the settling tank for natural phase separation. A rather high concentration of salts (8-21 g/dm 3) in stratal water contributes to an efficient precipitation of the solid phase. The clarified water is delivered into the capacity tank of treated water, while the slime (spent sorbent) accumulated on the slime site platform, is disposed for the utilisation as a drilling fluid component.

4.2. Bentonite clays in processes of water purification from cation dyestuffs Cation (basic) dyes: crystal violet, malachite green, fuchsine, rhodamine, safranine, methylene blue etc. exist in water as organic cations. These dyes are widely used in paper industry for paper colouring, due to high affinity of organic cations with respect to negatively charged sites of cellulose macromolecules [1].

688 Publications [103-105] summarise the studies of adsorption properties of various clay minerals with respect to methylene blue, crystal violet and malachite green. It was shown that montmorillonite, the rock-forming mineral of bentonite clays, possess the highest adsorption capacity with respect to these dyes. The adsorption of cationic dyes takes place both at the external surface of montmorillonite, and in its interlayer gaps. After the adsorption of crystal violet, first basal reflection of Na- and Ca- forms of Pyzhevsk montmorillonite doo I = 2.06 nm and 1.96 nm, respectively. Table 14 comprises the data concerning the adsorption of malachite green dye at Oglanly montmorillonite (Turkmenistan), which possesses cation exchange capacity 0. 84 mg-equiv/g. It is seen that for equilibrium concentrations ( 3 + 3 0 ) . 1 0 .4 mol/dm 3, the adsorption of this dye takes place according predominantly to cation exchange mechanism, with high selectivity of the ion exchanger with respect to organic cations. Spectral analysis data [104,105] suggest that the observed differences in the adsorption of dyes and desorption of inorganic cations (a > a') are related to the fact that the dyes are adsorbed not only in the form of individual organic cations, but also as their associates with molecules. Table 14 The amounts of malachite green dye (a) adsorbed by Na form of Oglanly montmorillonite and displaced Na § cations (a') for various equilibrium concentrations (C~) of the dye in the solution C e. 104, mol/dm 3

a, mol/kg

a', g-equiv/kg

3 28 33 39 65

0.52 0.64 0.75 0.88 1.31

0.40 0.53 0.58 0.62 0.62

The contents of dyes in waste waters of paper-making plants does not exceed 10 mg/dm 3 [1]. It was shown in [106] that the introduction of 100 mg/dm 3 of bentonite and 0.6 mg/dm 3 of the flocculant is sufficient for the complete decolouration of such water. The data obtained in the studies [103-106] were used in the development of technology which applies bentonite clays for the purification of waste water discharged by a plant which produces cationic dyes used as analytic indicators [107].

689

4.3. P h y s i c o c h e m i c a l principles of the application of clay m i n e r a l s for the r e m o v a l of high m o l e c u l a r organic c o m p o u n d s from waste waters In this section the advantages of the application of clay minerals for the purification of industrial waste waters from polyvinyl alcohol and protein compounds are discussed. Polyvinyl alcohol (PVA) is a non-ionic water-soluble polymer, which is widely used in various industrial processes; in textile industry, for example, PVA is applied for textile dressing. Macromolecules of this polymer are almost incapable of biologic destruction; therefore, to remove it from industrial waste water, physicochemical methods are to be used, among which the adsorption methods are most promising. It was shown by X-ray studies [94] that the adsorption of PVA from water solutions leads to the increase in first basal reflection of montmorillonite Na- and Ca-form to the values d001 = 3.00 nm and 1.83 nm, respectively. Thus one can expect high values of PVA adsorption on montmorillonite. This conclusion was supported in our adsorption studies [96], see Figure 4. The conditions of adsorption experiment were as follows: polyvinyl alcohol with molecular mass a, g/g

0.5

0.4

0.3

0.2 ( ~ qP

()

0.1 ~-(~ ()

'

I

I

I

0.2

0.4

0.6

Co, %

Figure 4. Adsorption isotherms of PVA on Pyzhevsk montmorillonite: (1) Na-form, (2) Ca-form.

690 M = 63000 and the n u m b e r of acetate groups less t h a n 2.5% was used as an adsorbate, concentration of adsorbent in the solution was 2.5% mass, adsorbent/solution contact time was 24 hours. Dispersion pre-processing of montmorillonite significantly increases its adsorption capacity with respect to PVA. The results presented above show that bentonite clays can be used to remove PVA and accompanying organic substances from industrial waste waters [1]. Sodium bentonites were long been used for the cleaning of vines and beer [108]. In this process the particles of montmorillonite act both as a d a g u l a n t and adsorbent. In fact, for pH values below isoelectric point pH i = 4.5-5, protein macromolecules and their associates are positively charged in water medium, and efficiently interact with negatively charged sites of montmorillonite particles. The neutralisation of charges leads to the aggregation of particles and their sedimentation. Our data [109] concerning the adsorption of bovine serum albumin with molecular mass M ~ 65000 on montmorillonite Na-form (Prodaneshti, Moldova) which possesses cation exchange capacity E = 0 . 9 5 mg-equiv/g show (see Figure 5) that this adsorbent is characterised by high adsorption capacity with

a, g/g o

o

!

o

0.8

0.4

o

3

4

8

C Rb > K > NH 4 > Na > Li) does not agree with its ionic radius (r = 0.143 nm). However, this case is a clear example that the exception proves the rule. In contrast to spherically symmetrical alkali metal ions, NH4 § ion exhibits pronounced tetrahedral structure. Being localised in 8member ring, it is involved in an efficient ion-matrix interaction not with six oxygen atoms, as in the case of Cs § Rb § or K § cations, but with two (for almost flat rings) or three (for puckered rings) oxygen atoms only. This also leads to the decrease in the energy of a m m o n i u m ions interaction with the framework, which results in the decrease in the values of K s' and K a [116]. Calorimetric studies of ion-exchange equilibria on clinoptilolite involving unicharged cations [118] confirm this conclusion. The energies of interaction of the compensating ions with ion exchanger matrix, and therefore the exchange constants, depend both on the

694 commensurability between the exchange cations and silicon-oxygen structural rings, and on the charge of the ion exchanger matrix, i.e. on the level of the AI~+-~ Si 4+ isomorphism. An increase in the charge of the matrix (a decrease in the SiO2 :A1203 ratio) leads to an increase in the contribution of electrostatic repulsion between neighbouring cations WAA into the total energy of their adsorption by ion exchanger. The contribution of WAA is especially significant, when large size cations Cs § Rb § K § are located in exchange positions, this results in the decrease in the Cs (Rb, K)-NaZ exchange constants. This effect is illustrated by Figure 6 reproduced from our study [119].

Ka

\ \

1

b

\ \ Q

80

\ \

\3 \ \ \ \ \ \

40

\ \ \

4

b

\ \ \ \

0.2

I 0.3

I 0.4

I 0.5

(A1)~

Figure 6. Ka values for Cs - Na-zeolite exchange vs equivalent proportion of aluminium in zeolite structure (A1)c: (1)erionite, (2)offretite, (3) low charge chabazite, (4)high charge chabazite.

695 The similarity between the sizes of entering large cations and free section dimensions of oxygen-silicon rings in which they are localised, also produces significant effect on both K a and the charge of the matrix. Thus the dependency of K a o n the ratio SiO 2 : A1203 and (A1)c = NA1/(Nsi + NA1) where NA1 and Nsi are the numbers of aluminium and silicon atoms in an elementary cell, respectively, should be analysed only for those zeolites for which the localisation sites are characterised by similar free section sizes. Chabazite, offretite and erionite are interesting in this regard. Their structure contains only 8-member rings (chabazite, erionite) or 8-member rings in combination with 12-member rings (offretite), which are too large for cations to be localised within. The minimum sizes of 8-member rings free section for all these three zeolites are the same. This fact enables one to postulate that for all these zeolites the structural factor plays a similar role in the determining of K a value. Figure 6, which includes also the data on high-charge synthetic chabazite [120], shows a linear dependency of Cs-NaZ exchange constant on the charge of zeolite matrix. Thus for natural zeolites a clear relation exists between the level of isomorphism in the structure and the exchange constant (free energy of exchange). It is necessary to take into account the composition of natural exchange complex of natural zeolites. The crystallisation of zeolites took place under different geological conditions: some of them were crystallised in the presence of calcium cations, while o t h e r s - in the presence of sodium and potassium cations. It was shown [121,122] that sodium sample of low-silicon calcium clinoptilolite in its natural state reveals a higher selectivity to double-charge cations whose dimensions are similar to those of Ca 2§ ion. On the contrary, sodium sample of high-silicon, sodium-potassium clinoptilolite in its natural state reveals a higher selectivity with respect to unicharged cations possessing large size. The thermodynamic exchange constant for the system Cs-Na-clinoptilolite decreases from K s = 69-70 (Table 15), the value characteristic for Hector (USA) and Dzegvi (Georgia) high silicon sodium-potassium clinoptilolite, to K a = 23-25 [118] for Sokirnitsa (Ukraine) low silicon calcium clinoptilolite. These data show that natural zeolites possess the property of memory with respect to cations which took part in their crystallisation, and to cations which possess charge and size similar to those of ions in initial samples. It is important that high selectivity of clinoptilolite with respect to large size alkali metal cations in both cases takes place. Thus, the sorption of large cations on natural high silicon zeolites, in addition to thermodynamic selectivity factor, is characterised by a more significant crystallochemical (geometric) factor, which results from the localisation of these cations in 8-member structural rings with the dimension of free section being comparable to the sorbed cation size. This conclusion may be considered as one of the formulations for the crystallochemical principle of selectivity of natural high silicon zeolites with respect to large cations.

696 In our earlier publication [115] it was stressed that selective sorption of large cations in 8-member zeolite structural rings, and the formation of stable complexes of alkali and alkali-earth metal cations with macrocyclic ligands are the phenomena of r a t h e r general character. Later [116] we have shown that the relation exists between the crystallochemical principle of selectivity of mineral ion exchangers, and the principle of the closest correspondence of the macrocycle opening to the size of complex-forming metal ion [123].

5.2. Use of z e o l i t e s for the p u r i f i c a t i o n of water from i n o r g a n i c c a t i o n s High selectivity of natural zeolites with respect to large size ions is used for the removal of radioactive 137Cs, 9~ and other ions from nuclear wastewaters, and for the removal of ammonium ions from municipal, industrial and agricultural effluents. A critical analysis of corresponding methods of treating these wastewaters was presented in [1,124]. Here the analysis of most interesting publications concerning the application of natural zeolites for the removal of inorganic ions from waste waters will be presented. Ames [125] was the first to demonstrate the ion-exchange specificity of clinoptilolite for the removal of radioactive caesium and strontium from low-level waste streams of nuclear installations. This process was extensively studied at Hanford and several other nuclear-test stations in the United States in the 1960s. Millions of gallons (1 american g a l l o n - 3.78 dm 3) of low-level 137Cs wastes have been processed through zeolite ion exchangers since that time [124]. The "saturated" zeolite columns were removed from the system, buried as solid waste, and replaced with new drums containing fresh clinoptilolite. Similar process was developed to remove 137Cs from high-level effluents using chabazite-rich ore [126]. Natural zeolites are capable of extracting species such as 9~ 13VCs, 6~ 59'63Ni, 45Ca, ~lCr selectively in the presence of high concentrations of competing ions. Natural zeolites are not only considerably less expensive than organic ionexchange resins; they are much more resistant to nuclear degradation. Due to their silicate nature, these zeolites also react rapidly in cement or glass systems, entraining the radioactive species into the final concrete or vitreous products. Therefore the studies concerning the use of natural zeolites for radioactive waste waters purification are in progress. Here the works of Robinson et al. [127,128] are to be especially mentioned, where it was shown that chabazite zeolites (both natural and synthetic) are the most effective materials for the removing of trace quantities of Cs and Sr from radioactive waste water. The amount of purified water is 3 000 000 dm3/year. The use of natural zeolites for the radioactive waste water t r e a t m e n t have been studied in a number of countries, including Canada, Great Britain, France, Bulgaria, Mexico, Japan, Germany, Russia, Ukraine [124]. The removal of ammonium ions from municipal, agri-industrial and industrial waste waters becomes increasingly important at the present time. Not only is + NH 4 toxic to fish and other forms of aquatic life; it also contributes greatly to the rapid growth of algae and leads to the eutrophication of lakes and rivers.

697 Therefore the concentration of ammonium ions in reclaimed water must not exceed 0.2-0.5 mg/dm 3 [18,129]. House waste water normally contains up to 30 mg/dm 3 of ammonia. A very promising method of the removal of ammonia from waste water is the ion exchange using natural zeolite clinoptilolite as the ion exchanger. Experimental-industrial and industrial testing carried out in the USA [1,124,129] have shown that the use of this mineral makes it possible to decrease the ammonium ion content in waste water by 93-97% and to attain the residual ammonia concentration in water of 0.2-0.5 mg/dm 3. However, in the development of a technology which employs natural zeolites for the t r e a t m e n t of municipal waste water and waste water of industrial plants to remove ammonia nitrogen, in addition to the ion exchange unit it is mandatory to have an efficient design of the unit used to regenerate exchange filters. Presently existing methods for the regeneration of spent natural zeolites may be divided into biological, reagentless and reagent methods. The critical analysis of these methods was presented in our review [54]. Considering the high acid resistance of clinoptilolite, we proposed its regeneration with an acid [130]. This enables the restoration of the sorption capacity of mineral ion exchanger with respect to ammonium ions (by its conversion to the H-form) and production of a solution of ammonium fertiliser, for example ammonium nitrate or sulphate, in a single technological operation. At Cherkasy (Ukraine) Industrial Association "Azot", the efficiency of the application of Sokirnitsa clinoptilolite for the t r e a t m e n t of waste water containing 100-260 mg/dm 3 NH4 § with its subsequent regeneration by 2-2.5% H2SO 4 [54], was industrially tested. It was shown that the most efficient process is the "hungry" (60-70%) regeneration of the spent filter. Here 3 kg-equiv of H2SO 4 are consumed per l k g - e q u i v of clinoptilolite exchange capacity. In particular, 60% regeneration of filter containing 10.6 tons of clinoptilolite had required 66 m 3 of 2.2% H2SO 4 (1462 kg). As a result of the regeneration, 4.91 kgequiv of NH4 § was removed from the entire clinoptilolite media. In 11 cycles of sorption and subsequent regeneration, the dynamic exchange capacity of the clinoptilolite (DEC of approximately 0.5 kg-equiv/kg) was practically unchanged. Animal waste waters possess high concentration of ammonium and potassium ions (up to 350 and 650 mg/dm 3, respectively). A comprehensive technology of such water t r e a t m e n t at swine complexes was developed [130]. This technology incorporates both mechanical and biological pre-treatment, the phosphate precipitation with the aid of calcium hydroxide, and the sorption of potassium and ammonium cations on clinoptilolite filters. Spent clinoptilolite is mixed with the phosphates after their neutralisation, and used as a fertiliser for lowproductivity sod-podzol soils of light granulometric composition [54]. At present, n a t u r a l zeolites clinoptilolite and phillipsite are extensively used in the USA, J a p a n and Italy to remove ammonium ions produced by the vital activity of fish during its transport or cultivation in special ponds [17,131]. We will examine this problem with regard to the t r e a t m e n t of water in a closed

698 system for the incubation of trout spawn [132]. According to the specifications, the concentration in water of NH4 § ions, a product of the metabolism of the spawn, should not exceed 0.4 mg/dm 3. The water should also contain K § Mg 2§ and Ca 2§ with the concentrations not less t h a n 50 mg/dm 3. Since the process of spawn incubation takes place at 6-10~ and this t e m p e r a t u r e is unfavourable for biological water processing, the t r e a t m e n t with n a t u r a l clinoptilolite was chosen as the efficient one. According to the specifications for the concentration of Ca 2§ Mg 2§ and K § in water [28], Sokirnitsa (Ukraine) clinoptilolite whose exchange complex comprises these ions was used for the treatment. From the rate of NH4 § production in a system with volume of 80 dm 3 of water containing 6 000 trout spawns (3 mg of NH4 § per hour), MPC(NH4 § = 0.4 mg/dm 3, the size of the zeolite filter was estimated and optimal conditions of its operation were determined with the consideration of the dynamic exchange capacity of clinoptilolite. 5.3. U s e of n a t u r a l s o r b e n t s as d e c o n t a m i n a t i o n a g e n t s for t h e e l i m i n a t i o n o f t h e c o n s e q u e n c e s o f t h e a c c i d e n t at C h o r n o b y l nuclear power plant High selectivity of clinoptilolite with respect to radioactive isotopes 137Cs, 9~ etc., was used by us to develop in 1986 the technology of the drinking water purification from the radioactive contamination [133]. The technology includes the t r e a t m e n t of water by powdered clinoptilolite at pH 9.5-9.8, followed by the processing with aluminium sulphate. After the precipitation of dispersed impurities in the clarifiers, subsequent tertiary t r e a t m e n t of water was performed by means of filtration through the 2.0 m thick layer of clinoptilolite with graininess 1-3 mm. It was shown by special experiments t h a t this purification method ensured the reduction in water radioactivity from 10 .7 to 3.10 .9 Ci/dm 3. The advantage of this scheme was in the fact t h a t it could be implemented at the existing equipment of the Dnipro water supply station. Since the impermeable "wall in soil" was constructed around the Chornobyl nuclear power plant, the problem had arisen of the purification of drainage waters prior to their discharge into the Prypiat' river. To perform this purification, the process was recommended which employed the filters with clinoptilolite tuff medium. The graininess of clinoptilolite was 2-3 mm, filtrating layer height 2 m, filtration linear rate 2 m/h; 15-20 mm crushed stone served as a bed-layer. Initial water radioactivity being 106-10 .7 Ci/dm 3, the purification degree amounted to 70-80%, and the service life of the filters was one year. The degrees of water purification with respect to various radionuclides, ensured by the filtration, were as follows: 95% for 137Cs, 80% for 9~ 50-60% for radionuclides of heavy elements, 15-20% for l~ The filters were designed in such a way t h a t the worked-out sorbent could be discharged; the resulting waste material which possessed the radioactivity level of 10 .5 Ci/kg was buried. Water t r e a t m e n t by powdered clinoptilolite and brown or porous coal, as well as the clearing with the help of coagulants and flocculants were recommended, in order to achieve higher purification degree.

699 Clay minerals were also extensively used for the decontamination of clothes, machinery and building materials. When formulating the problem, we have proceeded from the fact that the methods developed before 1986 were essentially relied on the use of surfactants, and could not ensure the decontamination required by safety norms. When standard method was applied to process the machinery and building materials, a serious problem of the decontamination of waste waters in large quantities had arisen. Therefore an essentially new sorption-adhesion method employing clay materials was proposed for the decontamination of clothes, machinery and building materials. Clay particles in an aqueous phase carry electrical charges of both signs. Due to this property, clays act as universal adagulants and can efficiently bound radioactive dispersed particles. Clay minerals possess particularly pronounced cation exchange properties, and display increased selectivity with respect to caesium, strontium and barium ions. This selectivity can be significantly enhanced by introducing special complexing additives into the system [134,135]. Thus, the application of clays as decontamination agents makes it possible to remove radionuclides in both dispersed and ion-soluble state from the contaminated surface. Clay minerals are high-disperse substances with developed external surface [1,7]. This property, their high wettability and the trend of particles to form a coagulation structure in aqueous media ensure high adhesive-enveloping properties of aqueous dispersions. At the same time, high wettability of clay particles facilitates the washing of the aggregates which they form with radioactive particles and radionuclide ions by a water jet from contaminated surface. In our work [133], montmorillonite and palygorskite from Dashukovka (Cherkasy province) deposit were used in the form of aqueous pastes (12-15%) and suspensions (2-7%). Clothes were treated with the 2% montmorillonite suspension at 10-12~ during 1 min with permanent stirring in a special washing machine, and then rinsed twice with water at 15-20~ Building materials and tractor machinery were treated with the pressure-feeded jet of 5-7% clay suspension or by clay paste (palygorskite and its mixture with montmorillonite). A paste was applied to the contaminated surface, held for 0.5-1 h and washed away by a water jet. Initial and final contamination levels of clothes, machinery and building materials were measured using the detectors DP-5V and RAM-63 at specially furnished sites or in premises where the radioactive background was below 0.03-0.1 mR/h. It is seen from Table 16, which summarises residual contamination levels and decontamination coefficients K d for the proposed t r e a t m e n t and those obtained using standard SAS solutions, that the use of suspensions and pastes is preferable to traditional decontamination methods. The application of this technology had provided a quite simple solution to the problem of waste water decontamination. Radioactive pollution was concentrated within clay slime, which could be easily precipitated from water; to accelerate the precipitation process, polyacrylamide was added (2.5-3.0 mg/dm3). The slime was

700 buried, while clarified water was subjected to tertiary purification clinoptilolite filters, and then discharged for subsequent natural filtration.

using

Table 16 Decontamination efficiency of standard SAS-containing solutions as compared with clay dispersions and pastes Initial

Processed items Uniform and working clothes Lorries and bulldozers

Concrete slabs, bricks, roofing slate, etc.

contamination level, mR/h 1.2 1.7 2.9 2.0-2.5 8.0-10 50-70 100-200 300-400 6.0-8.0 80-100 200-250

Standard SAS solutions Residual level, mR/h Kd 0.23 0.42 0.44 1.5-2.0 4.0-6.0 25-40 15-20 30-40 2.0-4.0 2.5-3.5 3.5-8.0

4.3 4.0 6.6 1.3-1.2 2.0-1.7 2.0-2.5 6.0-10 10 3.0-2.0 32-30 57-31

Clay suspensions and pastes Residual level, mR/h Kd 0.10 0.07 0.09 0.2-0.3 0.5-0.6 5.0-8.0 3.0-5.0 6.0-8.0 0.7-1.5 0.7-1.5 1.5-2.5

12.0 24.3 25.6 10-3.3 16-10 10-8.0 33-40 50 8.6-5.3 114-66 130-100

To summarise, the comprehensive application of clay minerals in the works concerned with the elimination of the consequences of Chornobyl nuclear accident, involving their properties as sorbents, ion exchange agents, adagulants and filtering materials, was proved to be extremely efficient. 6. A P P L I C A T I O N OF A D S O R P T I O N - A C T I V E M A T E R I A L S B A S E D ON N A T U R A L S O R B E N T S IN T H E P R O C E S S E S OF W A T E R PURIFICATION Extensive studies are performed recently concerned with the development of materials based on natural adsorbents. Sometimes the heat t r e a t m e n t or the substitution of natural exchange complex by other simple inorganic cation is quite sufficient to achieve the required improvement of the physicochemical and technical characteristics of materials based on natural disperse minerals. In the first case the structure of the mineral is changed predominantly (geometric modification); in the second case, the exchange complex is changed (ion-exchange modification). However, the improvement of natural minerals often involves several steps, leading to significant changes both in their structure and surface chemistry; in this way semisynthetic sorbents, a new class of sorption-active materials, are obtained. As defined in [136], semisynthetic sorbents are the

701 composite materials prepared from natural minerals via their chemisorption modification by organic and inorganic compounds, deposition of simple or complex oxides or other treatment, resulting in the formation of sorbents whose surface and porous structure differ from those characteristic for the original mineral, and combining the useful properties of natural and synthetic sorbents. In our recent publication [137], physicochemical principles underlying the production of new materials (sorbents, supports, catalysts, filtering powders, fillers for polymeric media) based on natural minerals are examined. In this section we consider the problems related to the preparation of new materials applicable for water purification, by simple or more complicated chemical t r e a t m e n t of natural minerals. 6.1. M o d i f i c a t i o n by a l u m i n i u m and iron salts In order to impart to natural sorbents the ability to adsorb organic and inorganic anions, these sorbents should be modified. This modification should preferably be performed using the substances and apparatuses which are themselves applied in the water t r e a t m e n t technology. From this point of view, it is desirable to modify the sorbents by aluminium sulphate, because A13§ ions and the products of their hydrolysis are strongly adsorbed by natural dispersed minerals, leading to the change of particle surface charge in water from negative to positive [71]. Perlite and bentonite modified by aluminium salts adsorb anionic SAS and dyestuffs from water rather efficiently [78,138]. In the technology of drinking water conditioning it is very important to remove fluorine ions from the water, because the increased (> 1 mg/dm 3) content of these ions causes various diseases [62]. In Ukraine some water sources (especially underground) contain 2-9 mg/dm 3 of fluorine ions; a number of water sources in the USA, Italy, Spain and England are also polluted with excessive amount of fluorine. All existing methods of removing fluorine-ions from water [62] have some disadvantages. The most promising method is the use of crushed clinoptilolite premodified by aluminium sulphate, acting as a specific absorbent of fluorine ions [56,139]. The testing of this method was performed with artesian water in Lubny town (Ukraine). The filter with clinoptilolite medium was employed, composed of granules 1-3 mm in size; filtering layer was 2 m thick. The modification of clinoptilolite was performed using 0.5% aluminium sulphate solution passed through the layer at a rate of 10 m/h for 1.0-1.5 h. The excessive amount of modifier was then washed out of the granulated medium with initial fluorine-containing water; this washing was terminated when the concentration of aluminium in the filtrate decreased to 0.5 mg/dm 3. The rinse water was collected in a special tank. At the final stage of filtering cycle, when the content of fluorine-ions in the filtrate had increased to 1.0-1.2 mg/dm 3, the rinse water was mixed with the initial water in a proportion (~ 1:14) which corresponded to aluminium concentration in the water supplied to the filters of ~ 1 mg/dm 3. This resulted in a considerable increase of the filter cycle.

702 At the filtration rate 5 m/h, approximately after 32 h of the filter operation, when the concentration of fluorine ions in the treated water increased to 1.01.1 mg/dm 3 (Table 17), the filter was switched over to the regeneration with aluminium sulphate, while the defluorination of water was performed by another filter, connected in parallel. The mean specific capacity of modified clinoptilolite is 500 mg/dm 3.

Table 17 Concentration (C) of fluorine in initial water and filtrate when providing water defluoration using clinoptilolite modified by aluminium sulphate Filter operating time, h 1 9 10 14 18 24 32

C, mg/dm 3 Initial water 2.0 2.2 2.3 2.1 2.2 2.4 2.1

Filtrate 0 0 0.15 0.5 0.8 0.7 1.1

There exist several mechanisms of fluorine ion absorption: anion exchange, formation of alumofluorine complexes and molecular adsorption of fluorine salts. These problems are discussed in more detail in [139]. The inversion of mineral particle charge from negative to positive means that strong sorption of hydroxo cations of aluminium takes place. This leads to the fact that it is not A13§ ions which act as counterions, but r a t h e r the hydroxyl groups of the hydroxo cations: i.e. the minerals acquire the ability for the anion exchange. The use of clinoptilolite modified by aluminium sulphate in the process of defluoration of natural waters is based on this ability. Studies of the interaction of aluminium and iron hydroxides or oxyhydroxides with clay minerals had shown that when a few percent of oxide is added to the clay, surface properties of the clay are controlled by the oxides. The amount of polycation required to produce the charge inversion is small. The results presented in [140] show clearly that A1 polycations are much more efficient in the blocking of negative charge than Fe polycations. It was supposed that this blockage is due not only to the higher charge of A1 polycations (z = 0.49 per gram-atom) as compared with Fe(III) polycations (z - 0.2 per gram-atom), but also to their shape. The spherical shape of Fe polycations limits the blocking of negative sites on the clay surface. On the contrary, planar A1 polycations cover a larger area, thus neutralising and blocking greater number of negative sites. These polycations can also occupy interlayer space of montmorillonite (see below).

703 However, the sorption of A1 polycations does not lead to the changes in the surface area of clay minerals with the rigid s t r u c t u r e - kaolinite, hydromica. On the other hand, the surface area of the kaolinite increases significantly (from 7.7 to 36.0 m2/g) when Fe polycations are added. This can possibly be ascribed to the 'porosity' of clusters formed by Fe polycations. The developed surface of kaolinite and hydromica clays modified by iron (III) salts, and more pronounced complexing properties of Fe 3§ ions as compared with A13§ ions, makes these sorbents promising for water t r e a t m e n t applications. For example, clays with Fe (III) polycations in the exchange complex are good adsorbents for the removal of anionic dyestuffs and non-ionic surfactants from water [2,137]. A remarkable success in the adsorption science was the development of methods for the preparation of semisynthetic microporous sorbents based on layer silicates with an expanded structural cell and basic salts of aluminium, titanium, chromium, etc., the synthesis of so-called pillared clays or pillared interlayer clay sorbents (PILC sorbents). Data were first presented in [141] concerning the sharp increase in the accessible surface and the formation of a thermostable material with open slit-like pores of width ~ ~ 0.8 nm as the result of the implantation of basic aluminium cations into the interlayer spaces of beidellite. In the former Soviet Union, a semisynthetic microporous sorbent based on montmorillonite and basic aluminium salt was obtained for the first time at our Institute [142]. Dozens of papers and patents have been published, dealing with the methods of preparation of PILC sorbents, studies of their properties and their practical application; some of these results are summarised in publications [143-145]. The data extracted from these papers, and also from [146], concerning the specific surface and effective pore radii in PILC sorbents, as compared with aluminosilicate catalysts possessing 13% A1203 produced by Davison Chemical Company (USA) are presented in Tables 18 and 19. It is seen that the sorbents obtained are generally microporous materials.

Table 18 Thickness of interlayer spacing Ad and specific surface area S determined from nitrogen adsorption in PILC sorbents based on montmorillonite Interlayer oligomer

Ad, nm

S, m2/g

[A1,304( ~ H)24] 7+

0.91

400

SisO12(OH)s

0.96

300

[Zr4(O H) ~Clx](~x)+

0.121

290

[TiO(OH)z] x

0.156

300

[Cr2,4(OAc)0,4(O H) 5,s]+

0.71

300

704 Table 19 Effective pore radius r and specific surface area S determined from nitrogen adsorption by All,~ - PILC sorbents and aluminosilica gel S, m2/g r, nm

A l i a - PILC

30 5 - 30 1- 5 K2304 + Cr2Oa + H20 + 3(0)

and Cr203 + H2SO4

Cr2(304)3 + 3H20

The adsorption at lower concentrations of potassium dichromate was found to occur in the first reaction. However, as the concentration of potassium dichromate in the solution increased, the second reaction involving the formation of Cr20~ was observed. Huang and Bowers [20] found that the adsorption and reduction occurred simultaneously and were responsible for removal of Cr(VI) from the bulk solution. These authors also noticed that when the carbon was oxidized with nitric acid and the solution contained chlorinated water with Cr(VI) ions, the removal of Cr(VI) decreased considerably. These observation encouraged us to study the influence of oxidants on Cr(VI) adsorption. In Figure 6 the influence of potassium nitride and potassium perchloride on the Cr(VI) adsorption is shown. The exponential decrease of Cr(VI) adsorption with the concentration increase of the oxidants is observed. In the case of KNO3, for a concentration higher t h a n 0.05 g/dm 3 no adsorption of Cr(VI) was obtained. The slight differences in the interaction of both oxidants on Cr(VI) adsorption into the activated carbon may be caused by the differences in the ion radius of these oxidants. Both oxidants affect redox potential of Cr§ § system. The reduction process of Cr(VI) to Cr(III) is carried out on the carbon surface which possesses negative charge at the acid values of pH. On the other hand, the reduction process is possible when HCrO~ 1 anion is in the acidic bulk solution [18]. Taking into account the analytical application of ion adsorption onto the activated carbon, the influence of matrix on the adsorption capacity is very important. For this reason, the influence of the ion strength on Cr(VI) adsorption was studied. The ion strength was created by proper addition of NaC1 into the Cr(VI) solution. In the first step of the study the influence of NaC1 concentration on Cr(VI) adsorption was determined. In Figure 6 the influence of NaC1 concentration on the Cr(VI) residue in the bulk solution is shown.

788

,--.,

E o

120

(3b

.c. 80-

o~ 300

J

(3r)

=L tO

r~0

40-

.~ >

> O

200

o

._ 00

I

0

0 . ~ -z

0,0

100

0,5

1,0

1,5

2,0

NaCI concentration [mol/dm3 ] Figure 6a. Influence of NaC1 concentration on the Cr(VI) residue in the bulk solution. The system studied: 50 mg of the activated carbon in 50 cm 3 containing 300 ng/cm 3 Cr(VI), (.) denotes carbon MC, (m) carbon MCO, (A) carbon MCN.

0 0,00

0,02

0,04

0,06

Oxidant concentration [g/dm3 ]

Figure 6b. Influence ofKNO3 (o) and KC104 (m) concentration on the Cr(VI) adsorption on carbon MCO. The system studied: 50 mg of the activated carbon in 50 cm 3 containing 300 ng/cm 3 Cr(VI).

The presence of NaC1 does not influence practically Cr(VI) adsorption only for the carbon MCN, in the case of the carbons MC and MCO the Cr(VI) adsorption drastically decreases with the m i n i m u m for 0.025 mol/dm 3 NaC1 concentration, then the adsorption increases asymptotically with the increase of NaC1 concentration. These different behaviors of the studied carbons can be explained by the different adsorption of Cr(III) after the reduction of Cr(VI). For the carbon MCN Cr(III) is adsorbed owing to the complex compounds with humic acids, whereas for the carbons MC and MCO forming the chloride complexes with Cr(III) and the adsorption in this form on the carbon surface seem to take place. Bautista-Toledo et al. [9] studied the presence of NaC1 in the chromium solution on the adsorption capacity of the activated carbon. The studies were carried out on the Merck carbon (sample M) which was modified with HNOa (sample MO). In Figure 7 the adsorption isotherms of Cr(VI) on the activated carbon obtained by Bautista-Toledo [9] are depicted. All adsorption isotherms were obtained out without adding any buffer to control the pH. The final pH of the equilibrium solution was around 6.5 for the sample M and around 3.5 for the sample MO. For the carbon M u p t a k e of Cr(VI) is lower when NaC1 increases. This behavior was explained by the fact t h a t for this carbon the adsorption was nonspecific or through the diffuse layer (competition between Cr024- and CI-) whereas in the case of the carbon MO the increase of Cr(VI) adsorption with increase of NaC1 concentration was observed. This result was a t t r i b u t e d to the fact t h a t an increase in of NaC1 concentration of the solution causes the activity coefficient of the present species to decrease

789

according to the Debye-Hfickel equation which, according to the Nernst equation, causes the reduction potential of the Cr(VI)/Cr(III) system to increase, i.e., the reduction process is more favored which, in turn, causes the a m o u n t of chromium adsorbed as Cr(III) to increase. The oxidation of the carbon by concentrated HNOa, in our case, seems to be more effective and formation of humic acids on the carbon surface strongly influences the mechanism of Cr(VI) adsorption.

9T )

6


0

E w!

l-0

2

o U_

pH Figure 12. Typical pH-adsorption edges for a ligand free system and for separate groups, (After Reed et al. [43]. Reproduced with permission from Marcel Dekker).

They distinguished four groups and typical pH - adsorption edges for the ligand-free system which are described as follows. In the first system, when the general shape of pH-adsorption edge remains unchanged as in the ligand free system except that the adsorption edges move to a higher pH region compared to the ligand free system. A plateau, defined as the leveling off of the pH-adsorption edge at fraction removed 95%) and a small relative standard deviation (< 7%, with relative error < 3%) of analysis. Dithizone has been used as a chelating agent for the combined flow atomic absorption spectrometric system to develop an efficient on-line preconcentrationsolvent elution procedure for determination of trace amounts of cadmium [58]. The authors described a new method whereby the trace amounts of cadmium are preconcentrated as a metal chelate using the chelatingdye dithezone on an activated carbon mini-column included in the flow injection (FI) system. The chelate is eluted with a small volume of a water-immiscible solvent and the analyte does not disperse on transfer to the flame atomic absorption spectrometry (FAAS) instrument, which increases the preconcentration factors. The system was used successfully for determination of cadmium in biological reference material. The enrichment method was developed by Gficer and Yaman [59] for the determination of cadmium and lead in vegetable matter by FAAS after preconcentration with 8-hydoxyquinoline or cupferron on the activated carbon. The enrichment factor of up to 100 was achieved by using both complexling agents. The optimum lowest pH values were found as 4.8 and 4.4 with cupferron and 5.3 and 5.8 with oxine, for cadmium and lead. The described procedure allows to detect 0.06 and 0.48 ng of Cd and Pb respectively, in biological material. The relative standard deviations were found to be 2% for cadmium and 3% for lead. Owing to the complexity of sorption processes, concentration of the ions characterized by different chemical nature is possible. Desorption of metal ions from activated carbons before their further determination by the AAS method (assuming an appropriate choice of a chelating agent) usually permits to attain a high degree of recovery (above 95%). In the analytical applications a special attention was paid to adsorption of mercury on the activated carbon. Koshima and Onishi [23] concentrated the nanogram amounts of mercury and methylmercury contained in the sea water on the activated carbon, then desorbed the concentrated compounds and determined by the flameless AAS after previous deposition on the gold support. Using this method they determined less than 4 ng of mercury contained in 200-300 cm ~ of the sample of artificial sea water using only 100 mg of the activated carbon.

800 During the tests of different carbon adsorbents [23] they concluded that the adsorption process is most effective in the case of the activated carbon. Matsueda [60] determined mercury (II) contained in the river water, after the previous concentration of this metal on the activated carbon. Recently activated carbon fibers were also used for adsorption of mercury (II) from aqueous solutions [61]. A majority of the publications relating to the utilization of activated carbons in the trace analysis by the AAS method describe modification of sorption conditions connected mainly with the changes of chelating agent, eluent and physicochemical parameters of the system e.g. pH. Less attention is paid to physicochemical nature of the surface of activated carbons. Storozhuk and Ivanova [62] paid the attention to significant contribution of surface phenomena occurring on the surface of activated carbons which can be used successfully for separation and concentration of trace amounts of elements. They stated that during the elution of adsorbed silver (I) with different eluents more than 90% of silver adsorbed on the activated carbon was reduced to metallic silver. Thus, the elution of adsorbed species with the eluents of different chemical nature allowed for analytical separation of adsorbed ions depending on their chemical nature. Some interesting surface properties are also found in basic carbons for which mainly physical adsorption of metal ions and surface concentration of these ions are observed [3]. Such properties ensure the increase of selective concentration and this may be used in the process of separation of metal ions on the activated carbon in a wide range of the matrix concentration. The examples of practical application of sorption on the activated carbon and of concentration of trace amounts of the elements from different matrices before their determination by the AAS method are given monographs Bachmann [63] and Mizuike [64]. Flameless atomization [e.g. 46] and flame atomization with injection of an analyte using a micropipette [e.g. 50,54] are especially useful for determination of the metals because they allow to determine a few components contained in the concentrate of a small volume. This allows to obtain very good detection limits (nanogram amounts of element per 1 g of matrix). Apart from the above method consisting in concentration of trace amounts of metals on activated carbons followed by determination using the AAS method one can find the papers suggesting the possibilities of utilization of activated carbons for the concentration of trace amounts of metals and determination of these metals by the atomic emission spectrometry (AES) method in which arc is used as an excitation source. After the preliminary concentration the carbon was washed and dried and then appropriate samples were collected from this material. These samples were usually mixed with appropriate spectral buffer e.g. with NaC1 and then placed in the craters of the carbon auxiliary electrodes. Such a procedure allows for direct excitation of emission spectra of the samples in the direct and alternating current arc. The optimal weight of carbon sample is 25-50 mg. Such an amount of the sample may be transferred completely into the crater of the electrode. The procedure consisting in a direct transfer of carbon samples to the electrodes and excitation of the emission spectra in the arc allows

801 for multielemental determination of the metals deposited on the carbon in the concentration range of 10~-105% with relatively good accuracy (sufficient for practical purposes) and at relatively small consumption of the reagents. This method has been used for determination of impurities in high purity zinc (II) and nickel (II) nitrates [65] and for determination of trace amounts of metals in sea water [66]. The main reasons for toxic ion metals preconcentration on activated carbons are to obtain a better limit of detection and separation of the determined element from the matrix. Toxic metals should be determined at ppb or even ppt levels in different environmental samples. Using modern spectroscopy methods and by preconcentration of metal ions on activated carbon these analytical goals can be reached. Okutani et al. [67] presented a rapid and simple preconcentration method by selective adsorption using carbon as an absorbent and acetylacetone as a complexing agent for beryllium determination by the graphite furnace atomic absorption spectrometry (GFAAS). Beryllium and its compounds are very toxic, especially for the lung, skin, and eyes, and the higher concentration can cause death (the toxic concentration of Be for a person is 0.1mg/m3). Consequently, a reliable method for the detection of beryllium in environmental specimens is required. The authors proposed the method based on adsorption of the beryllium-acetylacetonate complex on the activated carbon from the rainwater and sea-water samples, separation and dispersion in pure water. The resulting suspension was introduced directly into the graphite furnace atomizer. The detection limit was 0.6 ng/dm~, and the relative standard deviation (RSD) at 0.25 ~g/dm 3 was 3.0-4.0% ( n = 4 ) . A similar method was applied for determination of the total selenium content in sediments and natural water [68]. A trace level of Se was collected on the activated carbon (AC) as the Se(IV)-3phenyl-5-mercapto-l,3,4-thiadiazole-2(3H)-thione (Bismuthiol II) complex. The activated carbon was separated from the aqueous phase through a membrane filter (8 pm. pore size). The carbon phase on the filter was then dispersed in water containing the Pd modifier using an ultrasonic device for 30 s. The activated carbon suspension was directly introduced into the graphite tube atomizer and the Se concentration was determined by the GFAAS method. Naganuma and Okutani [69] adopted this method to determine trace amounts of bismuth in the sediments. They preconcentrated bismuth as the bismuthpotassium o-ethyldithiocarbamate (potassium xanthogenate, KetX) complex on the activated carbon and then introduced directly as a carbon suspension into the carbon tube atomizer of AAS spectrometer. The detection limit for the proposed method was 0.005 pg/100 cm 3, and RSD was 2.0% (n=8). The method was applied for determination of bismuth in the river-bottom and submarine sediments. Application of the activated carbon for preconcentration of Cd, Pb, Cr and Co of micro amounts from the reference material, drinking and sea water samples was comprehensively presented by the author [70]. The carbon MN, additionally impregnated with the 1% water so,ution 1 of ammonium pyrrolidinedithiocarbamate (APDC) was used as an adsorbent. Preconcentration of toxic ion

802 metals was carried out by the static method. 50 mg of the carbon was added to a 500 cm 3 sample, pH of the solution was natural and then the solution was mixed mechanically for two hours. The activated carbon was separated from the aqueous phase using an 8 ~m. membrane filter (25 mm in diameter). The carbon was removed from the filter and dispersed in 5 cm 3 of water with the addition of 0.005% Trition X-100 solution, using the ultrasonic device just before injection into the furnace. In the case of Cd and Pb analysis, the palladium nitride was added to the standard and carbon slurry (500 ~g/5 cm 3) due to the fact that carbon impurities (Si and P) influence the atomization mechanism of these elements. The results of toxic metal ions determination in different water samples are shown in Table 5.

Table 5 Determination of toxic ion metals sampling GFAAS technique

Element

Sample

SLRS-3

riverine water

Drinking

Cd

NASS-2 Seawater SLRS-3

NASS-2 Seawater SLRS-3

NASS-2 Seawater SLRS-3

NASS-2 Seawater

12 + 1

sea water

27 + 2

from the Baltic

32 +1.7

riverine water

65 + 3.2

water

83 + 3.3

Reference value (ppt) 13 -I-2a

29 + 4 68+7 b -

slurry

Recovery % 92 93 96 -

39+6

92

180

+7.2

-

-

riverine water

304

+ 24

300 + 40 a

sea water

water

sea water from the Baltic riverine water water

sea water from the Baltic

36 +3.5

using the carbon

from the Baltic

Drinking

Co

Slurry sampling* (ppt)

18 +0.9

Drinking

Cr

samples

water

Drinking

Pb.

in the water

56 + 4.5 173

+ 8.5

192

+ 9.6

25

+ 1.5

II +0.8 5 + 0.6 82 +5.1

175 + I0

I01 99

27 + 3 e

93

-

-

4 + 1 -

125 -

the mean of seven individual determination a - the direct determination by isotope dilution inductively coupled plasma mass spectrometry (ICP MS) b - the acid digestion isotope dilution inductively coupled plasma mass spectrometry (ADIDICP MS) e - the concentration by evaporation, GFAAS *

-

803 The results of determination obtained by the application of the preconcentration procedure and the slurry sampling technique are comparable to the certified values. The standard deviations for these determinations are lower compared to those obtained by different methods used during certification. CONCLUSIONS

Summing up the above considerations we can conclude that: I. Activated carbons are used successfully in the process of concentration and separation of toxic metals before their determination by the atomic spectroscopy methods. Description of adsorption of the electrolytes on the activated carbon is a very difficult task because of complexity of many processes occurring simultaneously such as ion exchange, nonspecific sorption, redox surface reactions, surface precipitation as well as formation of surface complexes of the chelate type. 2. In comparison to the synthetic ion exchangers carbon adsorbents allow for determination of concentrated impurities without the necessity of their isolation from the sorbent. These impurities can be determined directly after their introduction to the electrothermal atomizer e.g. in the form of the suspension (slurry sampling technique) of the activated carbon in diluted acids (carbon powder does not cause any interferences in the case of AAS method). Moreover the inductively coupled plasma atomic emission spectrometry (ICP-AES) [71] or the direct current plasma atomic emission spectrometry (DCP-AES) slurry sampling technique could be used successfully. The samples may be introduced also to the craters of auxiliary electrodes and then determined by the AES method in the arc. 3. Application of activated carbons is recommended especially for the concentration of mercury or methylmercury before their determination. 4. Concentration methods utilizing the carbon adsorbents create the possibilities of the analysis of small volumes of solutions. 5. Concentration and separation of the substances on the activated carbon adsorbents are often of lower costs and simpler than the multielemental extraction method. REFERENCES

I. H.P. Boehm, Advan. Catalysis, 16, (1966) 179. 2. H. Jankowska, A. Swiatkowski and J. Choma, Activated Carbon, WNT, Warsaw, 1985. 3. R. Dobrowolski, M. Jaroniec and M. Kosmulski, Carbon, 24 (1986) 15. 4. R.C. Bansal, J.B. Donnet and F. Stoeckli, Active Carbon, Marcel Dekker Inc., New York, Basel, 1988. 5. A. Capelle and F. De Vooys, Activated carbon...a fascinating material, Norit N.V., The Netherlands, Amersfoot, 1983.

804 6. 7. 8. 9. 10. 11. 12. 13. 14. 15. 16. 17. 18. 19. 20. 21. 22. 23. 24. 25. 26. 27. 28. 29. 30. 31. 32. 33. 34. 35. 36. 37.

H. von Kienle and E. Bader, Aktivkohle und ihre industrielle Anwendung, Ferdinand Enke Verlag, Stuttgart, 1980. A. Gierak, Adsorption Science & Technology, 14 (1996) 47. J. Zawadzki, Carbon, 19 (1991) 19. I. Bautista-Toledo, J. Rivera-Utrilla, M. A. Ferro-Garcia and C. Morebo-Castilla, Carbon, 32 (1994) 93. A. N. Mironov and V.P. Tauskanov, Adsorpcija i Adsorbenty, 2 (1974) 32. D.G. Hager, Chem. Eng. Progr., 72 (1976) 57. I.A. Tarkovskaya, Okislennyj Ugol, Kijev, Naukovaya Dumka, 1981. B.K. Strasko, I.A. Kuzin and A.M. Semusin, Zh. Prikl. Chim., 39 (1966) 2044. V.T. Nikolaev and V.V. Strelko, Gemosorpcija na aktivirovannych uglach, Naukovaya Dumka, Kijev, 1979. A. Macias-Garcia, C. Valenzuela-Calahorro, V. Gomez-Serrano and A. Espinosa-Mansilla, Carbon, 31 (1993) 1249. J.P. Novikov, V.B. Gliva and V.M. Komarevskii, Radiochimija, 19 (1977) 653. W.A. Abbasi and M. Streat, Separation Science and Technology, 29(9) (1994) 1217. H. Yoshida, K. Kamegawa and S. Arita, Nippon Kagaku Kaishi, 3 (1977) 387 B. R. Puri and B. R. Satija, J. Indian Chem. Soc., 45 (1968) 298. R. Leyva Ramos, A. Juarez and R. M. Guerrero Coronado, Wat. Sci. Tech., 30 (1994) 191. C.P. Huang and A.R. Bowers, I0 th Int. Conf. Water Pollution Res., 1977. H. Koshima and H. Onishi, Talanta, 27 (1980) 795. H. Koshima and H. Onishi, Bunseki Kagaku, 31 (1982) 421. M. D. Adams, Hydrometallurgy, 26 (1991) 201. H. Yoshida, K. Kamegawa and S. Arita, Bunseki Kagaku, I0 (1976) 1596. H. Yoshida, K. Kamegawa and S. Arita, The Report of the Government Industrial Research Institute, Kyushu, No. 22, 1979. H. Yoshida, K. Kamegawa and S. Arita, Nippon Kagaku Kaishi, 8 (1977) 1231. Y. Nagasaki, Japan Kokai, 76(122) (1974) 477. R. K. Sinha and P. L. Walker, Carbon, I0 (1972) 754. L. Thiem, D. Badorek and J. T. O'Connor, J. Am. Water Works Ass., 68 (I 976) 447. S. Perfiniemi, M. Ahlgren, Anal. Chem. Acta, 302 (1995) 89. S. Perfiniemi, S. Hannonen, H. Mustalahti and M. Ahlgren, Fresenius Z. Anal. Chem., 349 (1994) 510. A.T. Ellis, D. E. Leyden, W. Wegscheider, B. B. Jablonski and W. B. Bodnar, Anal. Chem. Acta, 142 (1982) 73. S. K. Gupta and K. Y. Chen, J. Water Polution Control Fed., 50 (1978) 493. E. Kobayashi, M. Sugai, I. Satoshi, Nippon Kagaku Kaishi, 23 (1984) 655. L. V. Rajkovic, Sep. Sci. Technol., 27 (1992) 1423. K. Kamegawa, H. Yoshida and S. Arita, Bunseki Kagaku, I0 (1979) 1365.

805 38. Y. Lu, K. S. Subramanian and L. Chuni, J. Environ. Sci. Health, A28(I) (1993) 113. 39. C. P. Huang and F. B. Ostrovic, J. Env. Eng. Div. Proc., ASCE, 104 (EE5) (1978) 863. 40. M. M. Jevtitch and D. Bhattacharyya, Chem. Eng. Commun., 23 (1983) 191. 41. D. Bhattacharyya and C. Y. Cheng, Environ. Progress, 6(2) (1987) II0. 42. O. M. Corapcoiglu and C. P. Huang, Water Research, 21(9) (1987) 1031. 43. B. E. Reed and S. K. Nonavinakere, Sep. Sci. Tech., 27(14) (1992) 1985. 44. C. Chang and Y. Ku, J. Hazardous Materials, 38 (1994) 439. 45. J.A. Korver, Chem. Weekblad, 46 (1950) 301. 46. R. Dobrowolski, Ph.D. Thesis, M. Curie-Sklodowska University, Lublin,1988. 47. A.V. Karyakin, U. Gribovskaya, Metody opticeskoj spektroskopii i luminescencji w analizie prirodnych i stocnych vod, Chimija, Moskva, 1987. 48. E. Jackwerth, Fresenius Z. Anal. Chem., 271 (1974) 120. 49. D.J. Hutchinson and A.A. Schilt, Anal. Chim. Acta, 154 (1983) 159. 50. H. Berndt and J. Messerschmidt, Fresenius Z. Anal. Chem., 308 (1981) 104. 51. H. Berndt, J. Messerschmidt and E. Reiter, Fresenius Z. Anal. Chem., 310 (1982) 230. 52. H. Berndt, E. Jackwerth and M. Kimura, Anal. Chim. Acta, 93 (1977) 45. 53. P. Rama Devi, G. Rama Krishna Naibu, Analyst, 115 (1990) 1469. 54. E. Jackwerth, H. Berndt, Anal. Chim. Acta, 74 (1975) 299. 55. B.M. Vanderberght, R.E. van Grieken, Anal. Chem., 49 311 (1977) 56. T. Aydemir, S. Gticer, Anal. Lett., 29(3) (1996) 351. 57. M. Soylak, I. Narin, M. Dogan, Anal. Lett., 30(I 5) (1997) 28 I0. 58. Y. P. de Pena, M. Gallego and M. Valcarcel, J. Anal. At. Spectrom., 9 (I 994) 69 I. 59. S Gficer and M. Yaman, J. Anal. Spectrom., 7 (1992) 179. 60. T. Matsueda, Bunseki Kagaku, 29 (1980) 110. 61. K. Kaneko, Carbon, 26 (1988) 903. 62. R.K. Storozhuk and C.S. Ivanova, Adsorpcija i Adsorbenty, 7 (1979) 19. 63. K. Bachmann, CRC Critical Reviews in Anal. Chem., 12(1) (1981)1. 64. A. Mizuike, Enrichment Techniques for Inorganic Trace Analysis, Springer Verlag, M~nchen, 1983. 65. K.I. Lazebnik, M.I Obruckii, A.N. Tomasevskaya and I.A. Tarkovskaya, Adsorpcija i Adsorbenty, 3 (1974) 57. 66. Z.P. Suranova, O.J. Grabcuk and A.N. Tomasevskaya, Adsorpcija i Adsorbenty, 2 (1974) 55. 67. T. Okutani, Y. Tsuruta and A. Sakuragawa, Anal. Chem., 65 (1993) 1273. 68. T. Kubota, K. Suzuki and T. Okutani, Talanta, 42 (1995) 955. 69. A. Naganuma and T. Okutani, Anal. Sci., 6 (1990) 77. 70. R. Dobrowolski, The 23rd Annual Conference of the FACSS, Kansas City, USA, 1996. 71. K.D. Ohls, J. Flock and H. Loepp, Winter Conference on Plasma Spectrochemistry, San Diego, USA, 1992.

Adsorption and its Applications in Industryand EnvironmentalProtection Studies in Surface Science and Catalysis,Vol. 120 A. Dabrowski(Editor) 9 1998Elsevier Science B.V. All rights reserved.

807

Air pollution control by adsorption

W.M.T.M. Reimerink, D. v.d. Kleut NORIT Nederland B.V., P.O. Box 105, 3800 AC Amersfoort, The Netherlands 1. I N T R O D U C T I O N The pressure on industry to decrease the emission of pollutants to the air is increasing. The importance to industry is to keep the costs as low as possible. A broad spectrum of techniques is available and is developed to control air pollution. The choice of a technique is determined by the type of pollution and the process conditions. In relation to price/performance, physical adsorption is one of the most important techniques to control air pollution. Both organic and inorganic molecules can be removed from a gas stream by physical adsorption. However, the adsorption affinity increases as the molecules become larger. As a consequence the adsorption capacity of an adsorbent is higher for large molecules t h a n for small molecules. For this reason physical adsorption is extremely suited for adsorption of organic compounds from gas, air, water and liquid streams. 2. A D S O R B E N T S In actual practice only the following adsorbents are applied: Activated carbon Carbon molecular sieves Polymers Silica Alumina - Zeolites Activated carbon can adsorb a broad range of pollutants with varying dimensions by its broad pore distribution of micro- and small meso pores. Activated carbon can adsorb a large a m o u n t of pollutants due to its large pore volume. Due to its hydrophobic character adsorption takes place at high relative humidity. The usability of activated carbon for air pollution control is limited by the risk of ignition at high temperature. Much research has been done to develop polymer adsorbents.In the market polymers as polyad [1,2] are applied for adsorption of high boiling compounds. At -

808 high relative h u m i d i t y these adsorbents have a tendency to swell. The polymers cannot be used at high t e m p e r a t u r e s due to deformation. A great disadvantage of polymer adsorbents is the low adsorption capacity on a volume basis. Alumina and silica [1,2] are meso porous and are not suited for adsorption of small organic molecules by physical adsorption. Unmodified silica and a l u m i n a are hydrophilic, so a high relative humidity disturbs the adsorption of organic pollutants to a large extent. Silica is suitable as adsorbent of water by chemisorption. Zeolites [1,2] and carbon molecular sieves [3-7] have a narrow pore distribution. The pore distribution determines which adsorbate adsorbs well and which adsorbate adsorbs to a less extent. These adsorbents can be applied in purifications of well defined gas s t r e a m s such as are present in gas separation applications. Compared with activated carbon the usability of these adsorbents is limited. Zeolites can be made hydrophobic by increasing the A1 content. These hydrophobic zeolites are suited to purify gas s t r e a m s at high relative h u m i d i t y and high t e m p e r a t u r e . Of the above mentioned adsorbents activated carbon is the most convenient for air pollution control for a broad range of compounds and for a large variation in process conditions. Compared to carbon molecular sieves and hydrophobic zeolites, activated carbon is a relatively cheap adsorbent. 3. A C T I V A T E D C A R B O N Activated carbons are micro porous carbonaceous materials. The activated carbons available in the m a r k e t differ in pore distribution, in form and in chemical composition. To decrease the emission an optimal carbon and system should be chosen dependent of the kind of molecules to be removed and the process conditions. The differences between activated carbons types are a consequence of the choice of activation process, the activation conditions and to some extend the choice of raw material. Activated carbons are produced from raw materials such as peat, wood, lignite, anthracite, fruit pits and shells. The raw materials are converted in activated carbon by steam or chemical activation With steam activation [8] the raw material is carbonised and/or oxidized depending on the carbonisation degree. Activation takes place above 900~ with steam. Process variations as residence time in the kiln, the activation temperature, the type of kiln and other conditions, allow carbonised materials to develops small micro pores which are enlarged up to large micro pores or small meso pores. Activated carbons suitable for gas and air purification are micro porous. When the gas stream contains a low concentration of pollutants, lesser activated carbons with a large a m o u n t of small micro pores exhibiting a high adsorption capacity at low relative pressure are applied. These carbons are produced by steam activation in a rotary kiln after a relatively short residence time. In the case which the gas stream contains a high concentration of pollutants higher activated carbons are applied. These carbons show a higher adsorption capacity at high relative pressures t h a n lower activated carbons. At high concentrations the service time of a filter is relatively short. As a consequence the activated carbon has to be regenerated insitu,

809 otherwise the carbon consumption will be too high. Higher activated carbons have a greater proportion of larger micro pores and small meso pores. These larger pores easily desorb their adsorbate. High activated carbons are also produced by steam activation in a rotary kiln. Only the residence time is longer than for the production of carbons with small micro pores. With chemical activation[9] an activating chemical, normally phosphoric acid, is mixed with a young carbonaceous vegetable material, carbonised at about 500~ followed by recovery of the activation chemical by water washing. The activated carbons produced on this way have less micro pores and more meso pore compared to steam activated carbons and are suited for adsorption of larger molecules such as are present in decolorization steps in the chemical, pharmaceutical and food industries. However special types of chemical activated carbons are suited for a small part of gas phase applications with insitu regeneration. By the choice of the raw material and by modifications in the activation process, more small meso pores and large micro pores can be produced. In Figure 1 the benzene adsorption isotherm of steam activated carbons suitable for gas phase purification with insitu regeneration (SORBONORIT 3) and for gas purification on throw away basis (NORIT R 2030) are given. In this figure a chemical activated carbon for gas phase purification with insitu regeneration has also been involved (NORIT GF 45).The adsorption capacity is given as weight of adsorbate per unit volume of activated carbon since in most applications, adsorbents are compared on performance in an existing filter (fixed volume). IBenzene

o

adsorption

isotherms

I

........i ....iiitil .........tttitii'!t ........tti titil .........i .....iitifl .........i'.i~ ""ti "'_ i i i iiiiii i i i iiiiii i i i iiiiii i i i ii~ill ,.T ~

r

......... i..... iTTIiYYi ......... YYYIYiIIY ........

i !i.~"i:'" 9: ,.,~i'Yiiiil .: ... i iiliii

ii ii iiiiiii ii iiiii

ii ii iili!i i iiiii

......... i ..... !...~...!..~.i .~.~. :: i i iiiiiii

ii ii iiilili i iiiii:

"

i

i

-

~ ii.~!~ii

.~),~D...~I":

i i iilii~

+." 9 i ii~L~"

i

i iiliiii

i i iiiliii

i iiiiiii

i i iiliiii

i i i iiiiii : .~It] iiiiii i i i iiiiii i i ii iiili : : : :::::: .y.,V": :::::: : : : :::::: : : : :::::: ......... i ..... !"" .~"i'~. "i~ .~ .......... ~'" .~"i" .~'i'i'i ~. ........ .~'"" .~'"i'" .~'!" .~ ~. i.~ ......... i ..... ........ i...i..6.66661

0.00001

0.0001

0.001 rel. p r e s s u r e

0.01

i iiiiii

i i iiiiii

i i i iiili : : : ::::: ......... i ..... i...6..i.6.i.i6

0.1

1

(%)

1~ SORBONORIT 3 4,. NORIT R 2 0 3 0

-T. NORIT GF 45

Figure 1. Adsorption isotherm of different carbon types for gas phase applications.

810 With the help of the Kelvin equation the relative pressure can be converted into pores dimensions. Thus Figure 1 shows the indirect relationship between adsorption capacity and pore distributions of different carbon types. The following carbon physical forms are utilised dependent upon the application: - extrudated carbons - granular carbon - powdered carbons - fibres. Extrudates and granular carbons are mostly applied in fixed bed systems. The particle size is chosen dependent on the allowed pressure drop. Especially in gas phase recovery systems, activated carbon is exposed to large pressure differences for long times and is transported on a regular basis to sieve the carbon. In that case the hardness and attrition are important qualities Extrudates especially are extremely hard. Recently in contrast to fixed bed systems, powder injection systems have been applied in gas/air purification. In recovery systems with a large flow and a relatively low concentration loosely woven fibre systems are used. Activated carbons ignite at high temperature in air. The ignition depends of the activation process and the purity of the activated carbon and varies from 200~ for a chemical activated carbon to 500~ for steam activated carbon produced from peat without additives as potassium. In most gas/air applications activated carbons are used at low temperatures (up to 200~ In this case danger of ignition does not exist. Without modification, activated carbon can show chemical interaction with adsorbates or can be catalytically active. This chemical interaction and the catalytic activity can be desired or not. For example, in the recovery of ketones the catalytic activity is undesired. The chemical interaction and the catalytic activity are connected to the presence of functional groups and ash components. Steam activated carbons contain as a consequence of exposing to air after activation, a limited number of varying functional groups, which give the carbon basic qualities. Chemically activated carbons possess by virtue of the production process a much larger number of varying functional groups, which give the carbon acid qualities. The ash content is dictated by the used raw material and can be diminished by washing. For very small molecules, the physical adsorption capacity can be low. Thus for a gas stream with a mixture of very small and larger molecules the low adsorption capacity for the small molecules can dictate the performance of the filter to a large extent. In that case the activated carbon can be modified to increase the removal efficiency of the small molecules by chemisorption or catalytic conversion. For this reason activated carbons are modified.

811 4. GAS P H A S E A D S O R P T I O N ISOTHERMS 4.1. I n t r o d u c t i o n Gas phase adsorption isotherms describe the relationship between the relative pressure (or the concentration) of a component in the gas phase and the maximum loading capacity, that is the loading capacity at equilibrium. A great number of equations have been developed to describe the equilibrium adsorption. For activated carbon as adsorbent none of these equations describe the measured isotherm for all the concentration ranges. For gas phase adsorption the following equations are used in practice [10] - equations based on the theory of Dubinin - the Langmuir equation - the Freundlich equation - the Henry equation - the BET equation. For gas and air purification isotherms based on the equation of Dubinin yield good results and give the most possibilities to predict the adsorption capacity for different compounds and temperatures [11]. 4.2. The a d s o r p t i o n i s o t h e r m of D u b i n i n and R a d u s h k e v i c h The adsorption isotherms of Dubinin and Radushkevich are based on the potential theory of Polanyi and assume filling of the pore volume by means of liquefaction of the gas by physical adsorption. The equation has been modified by a large number of investigators. Investigations carried out by Van Soelen [11] show that these modified equations hardly show an improvement for predicting the adsorption capacity. The equation of Dubinin-Radushkevich, has the following form

in (Av)- ln(W. d ) - B.

0//nl /

9log

(1)

Av

Equilibrium adsorption in terms of weight per volume unit of activated carbon (g/cm 3) W and B Carbon constants d Density of the adsorbate(g/cm 3) T Temperature (K) b Affinity constant of the adsorbate p/po Relative pressure n Exponent, varies from 1 to 3 For n=2 the equation has been suited to micro porous activated carbons. J. Reussien [12] shows that the equation with n=l can be applied for the meso porous part of the pores structure. The carbon constants W and B can be calculated from the intercept and the slope by plotting ln(Av) against [T/b'log(p/p0)] n. To use the

812 above equations, the temperature must be below the critical temperature. For temperatures above the critical temperature an adjusted equation must be applied [12]. Activated carbon is used for the adsorption of a broad range of adsorbates. The advantage of the use of the theory of Dubinin and Radushkevich is the possibility to predicts the adsorption capacity of all kinds of adsorbates on a carbon type using the carbon constants W and B calculated from the adsorption isotherm of a standard adsorbate. The adsorption isotherm of other adsorbates can then be calculated by substituting the liquid density of the adsorbate at the adsorption temperature and the affinity constant. The affinity constant can be calculated from the parachor or from the surface tension as given in standard tables. For gas streams with more than 1 component the adsorption capacities of the combined components are calculated by combining the isotherms of the pure components [13].

5. T H E F I X E D B E D A D S O R P T I O N 5.1.

PROCESS

Introduction

In a fixed bed system a polluted gas stream is passed through a bed of activated carbon. After the start of the adsorption process the activated carbon at the inlet side is loaded. Only after a certain time does the inlet side of the bed reach equilibrium because adsorption in pores does not takes place directly and is subject to transport limitations. Within the bed a mass transfer zone develops (MTZ). After a certain time the MTZ boundary reaches the outlet side of the bed and the emission concentration increases up to the allowed value, when the adsorption process is stopped. For adsorbent - adsorbate systems with a convex adsorption isotherm the length of the MTZ is constant and independent of the bed height [14]. In Figure 2 the course of the MTZ through various bed heights is given. The adsorption capacity of a filter is determined by the equilibrium adsorption and the length of the MTZ. A large equilibrium adsorption capacity and a small MTZ means a long service time of the filter and a low carbon consumption. In the most gas phase applications the MTZ is relatively small, certainly at low relative humidity. Thus the adsorption capacity of a filter is largely determined by the equilibrium adsorption. NORIT has developed an empirical model to calculate service time and carbon consumption on the basis of a standard carbon analysis and process conditions. The carbon analysis used are: the adsorption isotherm of a standard adsorbate the particle size. The process conditions used are: temperature concentration -

flow

-

adsorbate qualities.

813

Mass Transfer zone in a carbon bed 1.2

oo ~

0.8

o~

0.6

start loading

8 N 0

end loading 0.4

0.2

0 0

0.2

0.4

0.6

0.8

1

1.2

bed height Figure 2. The MTZ as a function of the bed height.

5.2. Adsorption kinetics Adsorption of gas molecules does not takes place instantaneously. The t r a n s p o r t is limited by: - axial dispersion - external transport - internal transport I n t e r n a l t r a n s p o r t characteristics are affected by: - pore diffusion - K n u d s e n diffusion a limited adsorption velocity surface diffusion. D e p e n d i n g of the process conditions, the type of a d s o r b a t e a n d the adsorbent qualities one or more steps are dominant. For exact d e t e r m i n a t i o n of the MTZ a set of differential equations h a v e to be solved. To e s t i m a t e the MTZ an empirical equation can be used which is derived from m e a s u r e d d a t a w i t h i n the m a t r i x of process conditions, which exist in practical situations. For activated carbon systems u n d e r relative dry conditions this equation is:

814 MTZ : e S T . (F)"0"054 9(Ci)0"133 9(Dp)1"549 9log Ci-Co Co with MTZ Mass transfer zone (cm) constant CST F Flow (cma/min) inlet concentration (g/cm 3) Ci outlet concentration (g/cm 3) Co particle diameter (cm) Dp

(2)

In solvent recovery applications when the activated carbon is wet after regeneration and cannot be dried during adsorption, the MTZ is 3 times larger.

5.3. The service time and the c a r b o n c o n s u m p t i o n Figure 2 shows that the MTZ curve is symmetrical. Thus half of the MTZ part of the bed can be considered to be in equilibrium with the inlet concentration and half of the MTZ part of the bed can be considered to be completely empty. For a fixed bed with an cross section area S (cm2)and a bed height L(cm) the dynamic loading Aa (g) A d - A v ( L - - - M T Z / $ 2"

(3)

At a flow F and an inlet concentration Ci the service time tb(min) of a bed with volume L'S is equal to tb-

Ad Ci.F

(4)

The carbon consumption CS (cm3/min) is CS- Ci'F'L'S Ad

(5)

5.4. F i x e d bed in insitu r e g e n e r a t i o n s y s t e m s 5.4.1. I n t r o d u c t i o n Insitu regeneration can be applied to gas streams with a high component concentration. In this case, the carbon consumption is too high for throw away basis operation. In insitu-regeneration systems the carbon is loaded in the same way as in fixed bed adsorption system up to break through. After break through the adsorbate is desorbed. Desorption may be followed by a (partly) drying step. During insitu regeneration the adsorbate is desorbed by pressure swing or temperature swing action. With pressure swing, desorption takes place at lower pressure than is present during adsorption. Pressure swing is applied for gas

815 separation. In t e m p e r a t u r e swing, desorption takes place at a higher t e m p e r a t u r e t h a n is present during adsorption.Temperature swing is mostly applied for solvent recovery. S t e a m or inert gas such as nitrogen is used as carrier gas. The benefit of the use of s t e a m is t h a t the installation, including the activated carbon bed,is w a r m e d up very quickly. Inert gas regeneration is applied for components which desorb at relatively high t e m p e r a t u r e and for components which decompose by oxidation with the activated carbon acting as a catalyst. In the last case steam activation is only possible by the use of activated carbon with a low catalytic activity such as the SORBONORIT K4. The motive to recover solvents can be the value of the recovered solvent. Recovery of solvents can also be a method to fulfill emission requirements. In some cases, recovery of a mixture of solvents can be used to effectively concentrate emissions to allow incineration.

5.4.2. The c a l c u l a t i o n of the s t e a m c o n s u m p t i o n With s t e a m regeneration, the solvent is recovered at high t e m p e r a t u r e with steam as the carrier gas. The rest loading on the carbon Ar as a function of the steam consumption is m e a s u r e d to determine the universal steam curve. For calculation of the adsorption capacity of an adsorbent in solvent recovery the effective loading Aef is an i m p o r t a n t factor. The effective loading is the difference between the dynamic loading Ad and the rest loading on the carbon after regeneration. In Figure 3 a steam curve has been given. Figure 3 shows that the recovery of the same a m o u n t of solvent costs much more in steam when starting from a lower dynamic

I

5

Steam curve

i

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

••••••••••••

5

0 0

200

! 400 steam

Figure 3. The steam curve.

I 600 volume

800

1000

816 loading t h a n starting from a higher dynamic loading. In most cases, desorption is stopped before all the adsorbates have been desorbed. By using activated carbons with special pore distribution, desorption can be made more effective.

5.5. F i x e d b e d i n s t a l l a t i o n s In designing an installation the m a x i m u m linear velocity (cm/sec) should be about 100 times the particle diameter, thus preventing fluidisation. The m i n i m u m linear velocity is a few cm/sec preventing axial dispersion. Thus the cross section of an installation is mainly determined by these conditions. The bedheight is mainly determined by the desired service time and the allowed pressure drop. Absorbers with a small bed height are applied for the removal of low concentrations of pollutants (< 1 rag/m3). The bed height is normally about 2 to 5 cm. The contact time is the order of 0.05 to 0.2 sec. Examples of this kind of filter are cylinders and thin rectangle boxes, divided into compartments. In this type of application carbon bounded in sheets such as N O R I T H E N E can be applied. I m p o r t a n t applications for these kind of filters are: - air conditioning - concentration peak smoothing. For higher concentrations (1 mg/m 3 up to 1 g/m 3) larger absorbers with a bed height of 25 up to 50 cm are used. The contact time in this kind of filter is about 0.2 to 2 seconds. Examples of this type of filter are simple steel drums provided with an inlet, an outlet and a base (aeropure filter), rectangular carbon absorbers and vertical as horizontal cylindrical absorbers. Important applications for these kind of filters are: - emission prevention in the chemical and food industries - paint spray installations - sewage air purification. For still higher concentrations (1 up to 50 g/m 3) recovery installations are applied with a bed height of 50 up to 150 cm. The contact time is about 2-4 seconds. Such recovery installations are of m i n i m u m 2 absorbers, one in loading and one in regeneration. Most installations comprise of a large number of absorbers. I m p o r t a n t applications of these kind of filters are: - solvent recovery in printing industry - dry cleaning. 6. I N S T A L L A T I O N S WITH P O W D E R S In a recent development air pollution control systems with injection of powder carbon in the gas stream can be applied. To keep a system in equilibrium, sufficient carbon m u s t be dosed t h a t emission concentration is in equilibrium with the equilibrium adsorption of the carbon. So in gas streams with a pollutant inlet concentration Ci and an emission concentration Co, Z g/cm 3 has to be adsorbed on to the activated carbon.

817 Z = C i - Co At the emission concentration of Co the equilibrium adsorption of the activated carbon Av can be calculated on basis of the theory of Dubinin as shown in 4.2. The carbon consumption CS is CS -

Z.F

(6)

Av The m a x i m u m loading of an activated carbon in such a system may be low compared to the m a x i m u m loading of an activated carbon in a fixed bed system, but by the use of very small particles, the kinetic effect is much faster, an a d v a n t a g e in processes determined by kinetics. Powder injection is applied in gas streams with a high debit and it can be built into a purification train. The system can be relatively cheap compared to fixed bed systems and is flexible concerning carbon dosage. Powder injection systems are used on a large scale at the purification of the flue gas of waste incineration plants [15]. In this way dioxins, dibenzofurans and heavy metals are removed from the flue gas. Typical dosing rates are 50 up to 200 g/m 3. Recently, impregnated powders have also been applied for special applications such as the removal of high concentrations of mercury. Powdered activated carbons have been tested excessively on explosion risks and are considered safe for flue gas conditions.

7. MODIFIED ACTIVATED CARBON The physical adsorption capacity for very small molecules can be low. Thus for a gas stream with a mixture of very small and larger molecules, the low adsorption capacity for the small molecules can dictate the performance of the filter to large extent. In this case the activated carbon m a y be modified to increase the removal efficiency of the small molecules by chemisorption or catalytic conversion. All possible impregnations with metal salts and with organic molecules as well as the modification of the functional groups are mentioned in the literature. To reduce air pollution only a few types of impregnations and modifications are of commercial interest. These i m p r e g n a n t s and the application are given in Table 1.

818 Table 1 Impregnation/modification commercial available activated carbons Component

Impregnant

Application

H2S, methyl mercaptan

- KI - Fe(OH)3 - complexes of transition metals (i.e:Cu,Cr) KOH

sewage air chemical industry

KOH - Na(OH) - KeCO3

sewage air chemical industry air conditioning

- CuO ZnSO4

chemical industry

Hg

-S -KI

purification methane prod. of batteries waste incineration

COS

complexes of transition metals (i.e:Cu,Cr)

chemical industry

HCN, C1CN

complexes of Cu,Cr,Zn and TEDA

gasmasks chemical industry

-

SO2

-

NH3

-

ASH3, PH3

Cu and Cr complexes

radioactive iodide

R

E

F

E

R

E

N

C

E

- TEDA -KI

chemical industry nuclear power plants

S

1. TNO (IMET), Alternatieve Adsorbentia voor het reinigen van koolwaterstoffen bevattende luchtstromen, (1991) (Dutch). 2. Y. Cohen (ed.), Novel adsorbents and their environmental applications, American Institute of Chemical Engineers, (1995). 3. Carbon containing molecular sieves, US Patent No. 3801513 (1971/1974). 4. Carbon containing molecular sieves, US Patent No. 3979330 (1974/1976). 5. Kohlenstoffhaltige Molekularsiebe, German Offenlegungschrift 2305435 (1973/1974). 6. Verfahren zur Gewinning von Stickstoffreichen Gasen aus neber Ne wenigstens 02 enthaltenden Gasen, wie zB. Luft, German Offenlegungsschrift 2441447 (1974/1976).

819 7. Kohlenstoffhaltige Adsorptionsmittel mit einstellbarem unterschiedlichen Porensystem, German Offenlegungsschrift 2624663 (1976/1977). 8. T. Wigmans, Fundamentals and practical implications of activated carbon production by partial gasification of carbonaceous material, NATO ASI series E 105, 559. 9. A. Cameron and J.D. MacDowall, The pore structure of wood based activated carbons, from: Principles and applications of pore structural characterization. Proc. R.I.L.E.M./C.N.R. Symp., Milan, Italy, (1983). 10. R.C. Bansal, J.P. Donnet and F. Stoeckli, Active carbon, New York and Basel

(~988). 1 I. A.C.D. van Soelen, Gas-fysisorptie-isothermen van aktieve kool. RU Utrecht (1991). 12. J.G.J. Reussien, De standaard-benzeenadsorptieisotherm aan actieve kool als een basisgegeven voor de berekening van terugwinnings- en luchtzuiveringsinstallaties, NORIT N.V., (1973). 13. D.M. Ruthven, Principles of adsorption and adsorption processes, New York, (1984). 14. Kel'tsev, Translation chapter 6 and 8 (Dutch), (1984). 15. B.v.d. Akker, D.v.d. Kleut and W.M.T.M. Reimerink, 16th Symposium on Chlorinated dioxins and related compounds, DIOXIN 96, Amsterdam (1996).

Adsorption and its Applications in Industry and Environmental Protection Studies in Surface Science and Catalysis, Vol. 120 A. Dabrowski (Editor) 9 1998 Elsevier Science B.V. All rights reserved.

New developments devices (ELCD)

821

of a c t i v a t e d c a r b o n s for e v a p o r a t i v e loss c o n t r o l

W.M.T.M Reimerink a, J.D. MacDowall, b D. v.d. Kleut a aNORIT Nederland B.V., P.O Box 105, 3800 AC Amersfoort, The Netherlands bNORIT (U.K) Ltd., Clydesmill Place, Cambuslang Industrial Estate. Glasgow G32 8RF, Scotland 1.

INTRODUCTION

Automobiles emit gasoline vapour from their fuel tank and (for old type of automobiles) from their carburettor, when the car is parked and the temperature under the motor hood is still high. Further a car parked at high out-door temperatures can emit a large amount of gasoline vapour by the fuel tank heating up. In addition to emission during parking, gasoline is also emitted during filling the fuel tank. The automobile contributes substantially to man-made hydrocarbon emissions. For example in West Europe the car evaporative emissions amount to about 10%. At refuelling another 3% is emitted. The diagram in Figure 1 shows the contributions to man-made hydrocarbon emissions. In the sixties in California USA legislation was passed to reduce the emission of gasoline from cars. Automobiles had to meet the so-named SHED emission test. Later on in the USA this legislation became more general. In the nineties the legislation in the USA was tightened. Emission of gasoline during refuelling had to decrease. In 1992 legislation in Europe was passed to reduce the emissions from cars by the introduction of an European SHED emission test. Producers of automobiles made arrangements to fulfill to the legal requirements. One of the measurements is the siting of an activated carbon hydrocarbon storage canister under the motor hood. During parking the filter is in contact with the fuel vapour from the tank and in contact with vapour from the carburettor for adsorption of the gasoline. During driving, air is sucked through the filter to desorb the gasoline which is fed to the engine for burning. These filters are described as Evaporative Loss Control Devices (ELCD filters). The extent and stringency of the legislation demanded that new types of canisters and optimized types of activated carbons were developed. In 1990 NORIT started [1-3] the development of an new type of activated carbon for the application in ELCD filters. Important requirements for the nineties and later are a high recovery of gasoline vapour and a low pressure drop. In the future

822

the composition of gasoline will change. ELCD carbons must be able to recover gasoline vapour with more alcoholic substances [4].

I

Contributions to man-made hydrocarbon emission Western E urope (24.8~)

I Car evap. D Car refuelling

(99%)

EI Car exh austs E3 Gasoline distr. I So Ivents

D Ref iner ies c3g.C

[ ] Other

s.8%) .

.

.

.

.

.

.

Figure 1. M a n - made hydrocarbons emissions. 2.

THE S H E D TEST, THE BWC A N D THE GWC TEST

In the SHED [5-8] test an entire vehicle is allowed to emit a certain amount of hydrocarbons during a prescribed test procedure in a test chamber after the car has been driven in. In the test chamber the car is placed on a roller bank. Situations of driving and parking are simulated. The total hydrocarbon emission from all parts of the car are measured, including tail pipe emissions of hydrocarbons. For refuelling special tests are developed. The SHED test is a very expensive and time consuming test. For testing evaporative emissions in the SHED test the activated carbon type has to be approved. For approval the activated carbon has to fulfill requirements on Butane Working Capacity (BWC). A generally useful BWC test is described by ASTM. After approval of a carbon type the BWC test is used as a test for production control. Besides the BWC extended test cycles can be carried out with gasoline to determine the Gasoline Working Capacity (GWC). This is necessary to get approval for a new carbon type. In the ASTM [9] BWC test pure butane is allowed to flow through a tube, filled with activated carbon at a rate of 250 ml per minute for 25 minutes. After

823 completion of the loading step, the activated carbon is purged with air for 40 minutes at 300 ml per minute. Both loading and purging steps are carried out at 25~ The difference between the weights of the carbon tube after loading and after purging divided by the carbon volume in the tube times 100 is the BWC. The BWC is given in g/100 ml. No international standard test method is available for carrying out gasoline tests. Every canister and activated carbon producer has developed his own test which is discussed and adjusted in consultation with the customers (filter and car producers) for correlation with the SHED test. The NORIT GWC test [2,3] is carried out using a carbon bed with diameter 36 mm and bed height of 150 mm. The activated carbon is loaded with a synthetic gasoline vapour, consisting of 50% air, 33% butane and 17% of higher boiling hydrocarbons (RVP(psi)=8.7). Synthetic gasoline vapour is used to obtain optimal reproducibility of the test and to observe small differences in activated carbon performance. The temperature during loading is 30~ and the flow is 0.121 L per minute. The loading is continued up to a breakthrough of 0.5 vol % of the ingoing hydrocarbons. Purging is carried with dry air at the same temperature and at a flow of 2.857 L/min during 16 minutes, which equates to 300 bed volumes. The flow direction during purging is opposite to the direction during the loading step. The GWC is the weight of butane and higher hydrocarbons adsorbed during the subsequent loading step per litre of activated carbon. The test is repeated up to 75 times to measure the decrease in GWC caused by building up a heel of hydrocarbons. In Figure 2 a scheme of the test equipment is shown. Some tests have been carried out with ethanol in the gasoline. In that case the synthetic gasoline vapour composition changed in 15.3% air, 71.1% butane 2.0% ethanol and 11.6% higher hydrocarbons.

Furnace 0, compared to the appropriate values obtained in the basic electrolyte 1M NaC104 comes as a surprise. The effect is most likely the result of fewer water molecules being present in the adsorption layer in the presence of adsorbed PEG molecules which to a more limited extent inhibit pT adsorption compared to water molecules. The linear tests of the F r u m k i n isotherm for PEG - mT systems are presented in Figure 18.

881

PEG 400

PEG 10000

-9

-8

+4

-2 -4 _ I O. 1

I 0.2

0.3

I+2 ' 0.4

'l 0.5

I

0.0

I

0.2

I

I

0.4

I

_ 0+2 I

0.6

Figure 18. Linear test of the Frumkin isotherm for 10.4 M PEG - mT systems, the electrode charges (CYM/10.2 C'm "2) indicated by each line [36].

As indicated by Figure 18, the adsorption properties of mT in the presence of PEG 400 and PEG 10000 are f u n d a m e n t a l l y different. In the presence of 10-4M PEG 400, the value of the A p a r a m e t e r changes along with the increase in the electrode charge f r o m - 4 . 7 t o - 5 . 9 . , while in the presence of 5 9 10-4M PEG 400, the repellent interaction between the adsorbed molecules of mT is more limited and the value of the constant A changes ranging from -2.9 to -3,.4 respectively. In the presence of PEG 10000, the value of the A p a r a m e t e r does not depend on the charge of the electrode and is similar for both c o n c e n t r a t i o n s , - 1 . 9 a n d - 2 . 2 for 10 .4 M PEG 10000 and 5 9 10 .4 M PEG 10000 respectively. The comparison of the value of the impact of the constant A demonstrates t h a t only in the presence of 10-4M PEG 400, the repulsive interaction among the adsorbed mT molecules is stronger compared to the same interaction in the solution of 1M NaC104. In the r e m a i n i n g P E G - mT systems studied, the interaction is weaker. The values of AG ~ determined for mT in the presence of PEG 400 do not depend on the charge of the electrode and indicate a slightly stronger adsorption of mT in the presence of PEG 400 compared to the adsorption in the solution of 1M NaC104. In the presence of PEG 10000, similarly to pT, there is also a linear relationship between the value of the AG ~ and the charge of the electrode. The value of AG O for pT and mT increases along with the increase of the positive charge of the electrode, with a stronger increase found in the presence of 5"10 .4 M PEG 10000 compared to the increase of AG ~ observed in the presence of 10 .4 M PEG 10000. A slightly higher adsorption of mT compared to pT from the 1M NaC104 solution, particularly visible at positive charges of the electrode, finds no

882 c o n f i r m a t i o n in s y s t e m s c o n t a i n i n g P E G in which the value of AG ~ for m T a n d pT are comparable. T h e r e f o r e s t r u c t u r a l differences of toluidine molecules h a v e no i m p a c t u p o n t h e i r a d s o r p t i o n in the p r e s e n c e of P E G polymers. F i g u r e 19 p r e s e n t s a l i n e a r test of the F r u m k i n i s o t h e r m for B U - T U a n d B U - m T systems.

~ (~-7.0

-10.0 \,Q\-. 9 A ",," ~ ~ 9149

= -6.0

-9.0

-5.0

-4.0

. .o

!i

-3.0

0.0

0.2

0.4

0.6

0.8

0.0

|

0.2

0.4

0.6

0.8 |

Figure 19. Linear test of the Frumkin isotherm for systems: a) 0.55 M BU + TU, b) 0.55 M BU + roT, the electrode charges (oM/10 -2 C'm ~ indicated by each line [37].

As i n d i c a t e d by the above Figure, the values of p a r a m e t e r A d e m o n s t r a t e a clear d e p e n d e n c e on the electrode charge only for the B U - TU systems. The effect is i l l u s t r a t e d in F i g u r e 20. 10 8

6 -o---

o

4 2 0 -2

I

I

I

I

0

2

4

6 -2

oM/lO C'm

-2

Figure 20. Variation of the interaction parameter A due to surface change density for TU in the presence of: a) 0.55 M BU, b) 0.88 M BU [37].

883

Figure 20 shows t h a t in general the repulsive interaction among TU molecules decreases with the decrease of the BU concentration and with the increase of the positive charge of the electrode. An exceptional situation is observed for ~ M ---- -t- 0.01C.m-2, at which the value of the p a r a m e t e r A effectively does not depend on the concentration of BU. The m a x i m u m change of the p a r a m e t e r for both concentrations of BU is very similar which can be an indication of similar changes occurring in the TU molecule orientation in both cases. It needs stressing t h a t the values of p a r a m e t e r A for TU in the presence of BU are on the whole either comparable with those obtained in the solution of 1M NaC104, or lower. Exceptionally higher values of the p a r a m e t e r have been obtained in the presence of 0.88 M BU for (~M < 0. The area of surface charges of (YM < 0 in the BU TU system is equivalent to the m a x i m u m adsorption of the substances used and it may be connected with the relatively higher values of p a r a m e t e r A at these charges. Not insignificant is also the possibility of changes in the butanol cluster structure and concentration in the bulk stage which can effect the surface stage as well. The value of the interaction constant for toluidine isomers in the presence of 0.55 M BU o f - 1 . 5 3 for mT a n d - 3 . 3 for pT does not depend on the charge of the electrode. In the presence of 0.44 M BU, the values of p a r a m e t e r A for these isomers indicate weaker repulsive interaction and to a slight degree depend on the charge of the electrode. It needs stressing t h a t in each solution containing BU the values of p a r a m e t e r A are lower in terms of the absolute values t h a n the respective values obtained in the solution of 1M NaC104, with such changes being more a p p a r e n t in the case of mT, as compared to pT. The value of free adsorption energy for 0 = 0, depending on the electrode charge in B U TU and B U toluidine isomer systems was presented in Figure 21. 30 7...., 0

C

26

_

a

b

3~ 22 d

18

_ a

14

I -2

0

2

4

J 6

O"M / 10 .2 C" m "2

Figure 21. Variation of the free energy of adsorption AG 0 due to surface charge density for: a) TU, b) mT, c) pT in the presence of 0.55 M BU and for d) TU in the presence of 0.88 M BU [37].

884 The analysis of the above Figure indicates significantly stronger adsorption of toluidine isomers compared to TU despite much lower concentration of toluidine used in the experiment. The linear nature of the relationship between

AG 0= f((~M) for toluidine isomers is a result of the chemical interaction of the aromatic ring ~ electrons of toluidine with mercury, resulting from partial transfer of the charge [41,42]. Absence of linearity from the relationship presented in Figure 21 for TU in the vicinity of the pzc may be connected with a more physical interaction between TU molecules and the surface of the electrode, as against the adsorption at positive charges. The above effect taking part within the area of the water capacity hump undoubtedly is also related to the relatively loosest structure of the surface water and therefore, the existence of a high dipole polarisability. It should also be noted that in all solutions examined containing BU, the increase in the concentration of BU results in an increase in the value of AG ~ both for TU and toluidine isomers. The effects contrast with the results obtained in systems containing PEG, and in particular PEG 10000. Easier adsorption of the substances under examination in the presence of higher concentrations of BU is most likely connected with increased degree of order in the adsorbed molecules on the surface of the electrode, resulting from selfassociation of BU molecules and also, as indicated earlier, lower number of water molecules displaced by the adsorbed organic molecules.

2.4. Virial a d s o r p t i o n i s o t h e r m s The values of Fs obtained for TU, pT and mT are in the majority of systems at variance with the theoretical values, and therefore to describe the adsorption of these substances, the virial isotherm was applied: J3c = F.exp2BF

(4)

where: B is the second virial coefficient. A linear test of the virial isotherm for PEG - TU systems is presented in Figure 22. Using a linear test of the virial isotherm, the value of the second virial coefficient B was determined based on the slope of the lines, while the value of AG 0 was determined by extrapolation of the lines to the value of F'- 0 in the standard state of 1 mol'dm -3 in the bulk of solution and 1 molecule 9 cm -2 on the surface of the electrode. The value of the second virial coefficient B at pzc in the presence of PEG are slightly lower compared to the value of B = 1.2 nm 2 9 molecule I, obtained in the solution without PEG [34], yet in the presence of PEG I0000, the values are slightly higher. The divergent changes in the value of the B parameter in systems containing PEG 400 or PEG I0000 are connected with the fact that the parameter constitutes the resultant of the inter-molecular repulsive interaction

885 a n d molecule size [43]. In the c h a n g e s of the AG ~ relative to the m o l a r m a s s of PEG, its c o n c e n t r a t i o n as well as the electrode charge, t h e r e is a n a n a l o g y to the c h a n g e s in the r e l e v a n t v a l u e s of AG O o b t a i n e d for the F r u m k i n isotherm. Similar effects were o b s e r v e d in o t h e r s y s t e m s studied.

18

18 -

o~

17

17

16

16 +3 +2 ' ~ ~1

15

14

I 1

0

I 2

,""i-3 3

PEG 10000

15

~

14

I -3 1

0

+2

I0 2 18

-2

F ' v u / 1 0 molec, m

Figure 22. Linear test of the virial isotherm for TU in the presence of 10 -4 M PEG, the electrode charges (CYM/10.2 C'm -2) indicated by each line [22].

Table 1 The values of p a r a m e t e r A b a s e d on the F r u m k i n i s o t h e r m a n d p a r a m e t e r B b a s e d on the virial i s o t h e r m as well as the calculated v a l u e s of p a r a m e t e r B according to P a y n e for the s y s t e m as follows: T U - 10 .4 M P E G B/nm2"molecule -1, ~M/IO-2 C.m-2

A

B (virial isotherm)

B - (2A+l)/2Fs

(JM

P E G 400

P E G 10000 P E G 400

P E G 10000 P E G 400

P E G 10000

+2

4.72

3.60

0.85

1.31

0.84

1.22

+1

5.08

3.80

0.88

1.43

0.90

1.28

0

5.28

4.04

0.91

1.44

0.93

1.34

-1

5.60

4.72

0.98

1.52

0.98

1.52

-2

6.04

4.84

1.00

1.58

1.04

1.55

886 Based on the relationship contributed by Payne [44]: B = (2A +l)/2F s

(5)

the value of B was calculated for selected systems. A satisfactory conformity of the values of B so calculated and the values of B determined based on the virial isotherm was found. Table 1 presents an example of the values of A and B p a r a m e t e r s calculated for the P E G - TU systems.

2.5. E l e c t r o s t a t i c

parameters

of the inner layer

An insight into the potential drop changes in the inner layer (I)M2 at a constant charge caused by adsorption offers information on the structure of the double layer. The changes are the resultant of the contribution of the free charges and oriented dipoles. Experimental separation of the effects is in principle impossible [45]. According to the electrostatic model of Parsons [34], the potential ~M-2 is a sum of two components depending on the surface density of the charge: (I)M -2 _ 4~x----L1cyM + 4~pp F'p

(6)

where: pp is the dipole m o m e n t of the isolated molecule of the accelerating agent: for TU, ~ = 16.31"10 .30 C'm, for pT, ~ = 4.43"10 .30 C'm and for mT ~ = 4.76"10 ~0 C'm. The values of the dipole moments in the inner layer undergo usually certain changes caused by the field of the electrode and additional interactions with the adjacent dipoles. The analysis of equation (6) ignores these effects. Other variables used in equation (6) or arising out of it include: Xl - the inner layer thickness; ~ i - electrical permittivity of the inner part of the double layer; K i - the integral capacity of the inner layer. The value of potential (I)M-2 was calculated based on the relationship: (I)M-2 = E - E z - (I)2-S where E stands for the potential equivalent to the given value of F'p and CyM and Ez stands for the potential of the zero charge for the solution which does not contain the accelerating agent, while the drop of the potential in the diffusion layer (I)2.s was calculated based on the theory of G o u y - C h a p m a n [46]. The dependence of the value of (I)M-2 on F'TU at a constant surface charge density and in the presence of 10 .4 M PEG is presented in Figure 23. The relationships presented in the above Figure are linear in nature, similarly to the results obtained in other systems containing TU [47, 48]. The linearity of the dependence of ~M-2 on F'TU is additionally confirmed by the congruency of the adsorption isotherms described relative to the charge. An analysis of these relationships was conducted using the method applied previously [48]. The p a r a m e t e r s obtained which describe the inner layer are presented in Table 2.

887 PEG 400 -0.4

PEG 10000 -3

-2

-1

-0.4 - 3 - 2 -1 0 +1 +2

o

~e

1 +2

-0.3

-0.3

+3

-0.2

-0.2

-0.1

-0.1

0.0

0.0

0.1

0.1 0

1

2

3

0-

1

2 -18

10

-2

F'wu/molec" m

Figure 23. Potential drop across the inner layer q)M-2 as a function of the quantity of TU adsorbed at constant electrode charges (CYM/10.2 C'm -2) in the presence of 10.4 M PEG [22].

Table 2 I n n e r l a y e r p r o p e r t i e s for TU a d s o r b e d at the m e r c u r y / w a t e r - N a C 1 0 4 -

PEG

m i x t u r e interface 102 ~M/C'm -2, xl/nm, 10-~SF'Tu/molecule-m-2 102K iF ' /F'm-2122] P EG 400 (~M CPEG=COnSt

Ei

10 -4 M -3 9.3 - 2 10.2 - 1 11.0 0 12.4 +1 13.3 CPEG----5" 10 -4 M -3 7.1 -2 7.5 -1 8.4 0 9.4 +1 10.0

F'TU = 0 Ki

P EG 10000 F'TU - 1

xl

Ki

Xl

26.3 0.31 30.3 0.30 32.3 0.30 3 - 8 . 5 0.29 47.6 0.28

17.9 20.2 22.2 26.3 31.3

0.46 0.45 0.44 0.42 0.38

35.7 38.5 45.5 50.0 58.8

18.9 21.3 23.8 27.8 32.3

0.33 0.31 0.31 0.30 0.27

Ei

F ' T U -- 0

F ' T U -- 1

Ki

x1

Ki

Xl

5.2 6.2 6.8 7.5 8.0

20.8 24.4 27.0 30.3 38.5

0.22 0.22 0.22 0.22 0.18

9.8 13.3 15.6 18.5 20.8

0.47 0.44 0.39 0.36 0.34

4.3 5.1 6.1 6.7 7.8

23.3 28.6 33.3 35.7 42.0

0.16 0.16 0.16 0.16 0.16

8.9 11.4 14.7 16.4 19.2

0.43 0.40 0.38 0.36 0.36

C P E G ----

0.18 0.17 0.18 0.17 0.15

The v a l u e s of Ei for TU o b t a i n e d in the p r e s e n c e of P E G are in g e n e r a l lower c o m p a r e d to the v a l u e of Ei = 11.4 o b t a i n e d for TU in w a t e r solution at aM = 0 [34]. The fact t h a t the electrical p e r m i t t i v i t y of the i n n e r l a y e r i n c r e a s e s along

888 with the increase of the positive charge of the electrode, may prove the existence of free PEG molecules in the inner layer, subject to partial rotation. The effects are in conformity with changes of the constant value of the interaction of A determined on the basis of the F r u m k i n isotherm. The drop in the integral i with the increase of TU concentration is connected with the capacity of K F, increase of the thickness of the inner layer Xl. Small values of Xl for F~,U = 0 indicate fiat orientation of PEG molecules adsorbed on the surface of the mercury electrode, as previously suggested [31]. The values of K iF, determined in the presence of PEG 10000 are lower t h a n the respective values determined in the presence of PEG 400. The effect is due to lower values of ~i in the presence of PEG 10000, as compared to PEG 400. The results obtained indicate a more condensed structure of the inner layer in the presence of PEG 10000. Figures 24 and 25 present changes in the value of (I)M-2 relative to F'p for toluidine isomers in the presence of PEG.

>

-0.4

-0.4

m

PEG 10000 -0.3 _

-4

-0.2

-0.2

-2

-0. l

+2

0

"-a. -2

-0.1 _

-4

-0.3

~

j_oj_----o-o-~-__~__>___q0 0.0

0.0

0.1 0.4

I

1

I

I

I

0.8

1.2

1.6

2.0

2.4

0.1

I 0.5

1 1.0

1

[

1.5

2.0 18

2.5 -2

F ' ~ j / 1 0 molec, m

Figure 24. Potential drop across the inner layer (I)M-2 as a function of the quantity of pT adsorbed at constant electrode charges (C~M/10.2 C'm -2) in the presence of 10 -4 M PEG [23].

The relationships presented in the above Figures are linear in n a t u r e effectively only in the presence of PEG 10000. Unfortunately, p a r a m e t e r s of the inner layer calculated in these systems vary significantly from rational values. The general absence of linearity in the relationship between (I)M-2 and F'p in the presence of PEG 400 in the case of mT may be due to the changed orientation of mT molecules on the surface of mercury, which is also indicated by the values of

889

>

e

-0.6

-0.20

P E G 400

-2

-0.16 -0.4 -0.12 -0.2

..- .::-

_

0

~~----__~ -0.08 " F'TU and all this in much lower concentrations of toluidine (conditioned by its solubility in the solutions studied), compared to the concentration of TU. 9 The increase in the positive charge of the electrode in all systems studied results in the increase of the value of free adsorption energy AG 0 which may be connected according to Payne [57] with squeezing w a t e r molecules out of the surface of the electrode at aM > 0.

9

9

The linear dependence of AG o on aM accompanying mainly the ion adsorption arrived at in the majority of systems indicates t h a t the effects connected with the polarity of molecules (stable dipole moment) constitute the dominating impact of the field of the electrode on the adsorbate molecule. Larger area occupied by a single molecule calculated on the basis of the Fs value compared to the area determined based on the molecular size indicate flat positioning of molecules on the surface of the electrode and their m u t u a l repulsion, confirmed by relatively larger values of the virial coefficients. It can also result from residual presence of other molecules. Negligible differences in the adsorption p a r a m e t e r s determined for toluidine isomers both in the 1M NaC104 solution as well as in the presence of inhibitors indicate t h a t the location of amino groups in the aromatic ring has a limited impact upon the adsorption process. The stronger adsorption of

898 toluidine isomers compared to TU confirms an exceptionally strong impact of the aromatic ring upon the adsorption process [58]. 9 The values of the A parameter determined on the basis of the Frumkin isotherm indicate repulsive interaction between the adsorbed molecules of the accelerating agent. In all systems studied containing TU, the increase of the positive charge of the electrode was accompanied by the reduction in the value of the A parameter resulting from the penetration of C104 anions between the positively charged TU amino groups at ~M > 0. In solutions containing toluidine isomers, the value of the A parameter in general does not depend on the charge of the electrode which suggests stability of the orientation of the adsorbed toluidine molecules. 9 The linear dependence of the potential drop in the inner layer ~M-2 to F'p found in the majority of systems confirms the congruence of the determined adsorption isotherms in relation to the charge. The values of Xl arrived at confirmed the fiat positioning of BU and PEG molecules on the surface of the electrode postulated earlier. Relatively large values of electric permittivity of Ei for F'p = 0 (~i > 10) obtained in some systems indicate a considerable freedom of inhibitor molecule rotation in the field of the double electric layer. The results of the Zn(II) ion electroreduction measurement as an ion piloting the adsorption equilibrium in mixed adsorption layers at a potential distant from the pzc indicate significantly higher efficiency of toluidine isomers in eliminating the inhibiting impact of PEG as compared to TU. In systems containing BU, the relationship between the concentration of the accelerating agent and the inhibitor with a view to the compensation effect is weak, i.e. it does not depend on the concentration of BU, while in the presence of PEG, such relationship does occur. The results arrived are indicated a higher BU adsorption lability compared to PEG at Zn(II) ion reduction potentials which additionally enables the adsorption of the accelerating agent. It is worth noting that at such potentials, the surface concentrations of the low-molecule substances studied are undoubtedly small and occur in the area of the initial, approximately linear, course of the isotherm which explains the stability of the relationship between the surface concentration and the bulk concentration. The above relationship does not apply to high-molecular compounds. The examination of the Zn(II) ion reduction mechanism as an adsorption probe using other inhibitors will undoubtedly allow expanding the information available on the properties and structure of mixed adsorption layers at potentials distant from the pzc. The results of research presented include information which can be used to understand better the adsorption equilibrium generated in complex corrosion systems comprising inhibitors with a specific structure and properties. Effectiveness of toluidine isomers as potential inhibitors of corrosion should be greater than that of TU owing to stronger adsorption of toluidine isomers. This adsorption is conditioned by the m e t a l - aromatic ring ~ electrons which are the effects of a partial charge transfer between the adsorbate and metal causing on

899 the whole, fiat orientation of the adsorbed toluidine molecules. An additional favourable effect of using toluidine as the inhibitor of corrosion is the fact that this substance causes a local increase of pH value which does not occur while using TU. Comparison of h i g h - and low molecular inhibitors activity allows for making the statement that the adsorption layers formed by high molecular inhibitors, as a rule, are not tight which enables their penetration by various factors which can affect the protective function change. Formation of mixed adsorption layers created by the adsorbate molecules with chemical interactions between them (e.g. hydrogen bonds), undoubtedly causes the increase of the adsorption layer tightness. In metal corrosion taking place in natural conditions, besides basic processes occurring on the metal surface and leading to oxidation of metal atoms, and their possible transition into the aqueous phase an essential role is also played by the redox processes caused by active components of the solution and leading to formation of the mixed potential. After its establishment there is formed on the metal a surface layer of a complex structure which can be stimulated by adsorption processes which can be conditioned by both electrostatic and donor-acceptor interactions. These considerations lead to a general conclusion that reliable results of the investigations of organic inhibitor effects on metal corrosion can be expected in the model systems very close to natural conditions and the fundamental investigations enable better understanding of the nature of these complex dependences.

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D.Jonhua, S.Guangling, L.Haicao, C.Chunan, Zhongguo Fushi Yu Fanghu Xuebao, 16 (1996) 53. M.Metikos-Hukovic, R.Babic, Z.Grubac, S.Brinic, J.Appl.Electrochem., 26 (1996) 443. R.M.Souto, M.M.Laz, S.Gonzalez, J.Chem.Soc.Faraday Trans., 92 (1996) 2725. N.Rajendron, K.Ravichandran, S.Rajeswari, Anti-Corros.Methods Mater., 42 (1959) 9. Q.J.M.Slaiman, H.M.AI-Saaty, Ann.Univ.Ferrara, Sez.5 Suppl. 1995, I0 (8 th European Symposium on Corrosion Inhibitors, 1 (1995) 333. R.H.Rahim Abdel, B.Hassan Hanaa, M.W.Khalil, Materialwiss.Werkstofftech., 28 (1997) 198. M.S.Abdel-Aal, M.S.Morad, Z.A.Ahmed, Ann.Univ.Ferrara, Sez. 5 Suppl. 1995, 10 (8 th European Symposium on Corrosion Inhibitors 1 (1995) 343. A.A.El-Shafei, M.N.H.Moussa, A.A.EI-Far, J.Appl.Electrochem., 27 (1997) 1075. A.M.Ai-Mayouf, Corros.Prev.Control, 43 (1996) 68. A.A.Aksut, A.N.Onal, Bull.Electrochem., II (1995) 513.

900 11. R.T.Vashi, V.A.Champanerii, Trans.SEAST, 32 (1975) 5. 12. S.Muralidharan, B.Babu Ramesh, S.Iyer Venkatakrishna, S.Rengamani, J.Appl.Electrochem., 26 (1996) 291. 13. M.Elachouri, M.S.Haji, M.Salem, S.Kertit, J.Aride, R.Coudert, E.Essassi, Corrosion (Houston), 52 (1996) 103. 14. E.Khamis, H.A.Al-Lohedan, A.Al-Mayouf, Z.A.Issa, Materialwiss.Werkstofftech., 28 (1997) 46. 15. Wang Jianming, Cao Chunan, Lin Haichao, Wuli Huaxue Xuebao, 12 (1996) 1036. 16. E.Kalman, J.Lukovits, G.Palinkas, ACH-Models Chem., 132 (1995) 527. 17. T.Suzuki, H.Nishihara, K.Aramaoki, Corros.Sci., 38 (1996) 1223. 18. M.Th.Makhlouf, M.H.Wahdan, Polish J.Chem., 69 (1995) 1072. 19. K.Sykut, J.Saba, B.Marczewska, G.Dalmata, J.Electroanal.Chem., 178 (1984) 295. 20. G.Dalmata, J.Nieszporek, Polish J.Chem., 68 (1994) 2009. 21. A.Baars, J.W.J.Knapen, M.Sluyters-Rehbach, J.H.Sluyters, J.Electroanal.Chem., 368 (1994) 293. 22. J.Saba, Electrochim.Acta, 41 (1996) 297. 23. J.Saba, Monats.Chem., 128 (1997) I. 24. M.R6sch, Kolloid Z., 147 (1956) 80. 25. M.R6sch, Kolloid Z., 150 (1957) 153. 26. R.Parsons, Can.J.Chem., 59 (1981) 1898. 27. B.B.Damaskin, S.L.Diatkina, V.K.Venkatesan, Elektrochimija, 5 (1969) 524. 28. K.M.Joshi, S.J.Mahajan, M.R.Bapat, J.Electroanal.Chem., 54 (1974) 371. 29. M.R.Moncelli, M.L.Foresti, R.Guidelli, J.Electroanal.Chem., 295 (1990) 225. 30. A.F.Nesterenko, W.S.Burykina, A.J.Aprasjuchin, M.A.Loszkariew, Elektrochimija, 23 (1987) 425. 31. H.Jehring, Elektrosorptionsanalyse mit der Wechselstrompolarographie, Akademie Verlag, Berlin 1974. 32. J.Saba, Collect. Czech. Chem. Commun. , 60 (1995) 1457. 33. J.Saba, Electrochim.Acta, 39 (1994) 71 I. 34. R.Parsons, Proc.Roy.Soc., A 261 (1961) 79. 35. R.Parsons, Trans. Faraday Soc., 51 (1955) 158 I. 36. J.Saba, Gazz.Chim.ltal., 127 (1997) 53. 37. J.Saba, Collect. Czech. Chem. Commun. , 61 (1996)999. 38. R.Parsons, P.G.Symons, Trans.Faraday Soc., 64 (1968) 1077. 39. J.D.Garnish, R.Parsons, Trans. Faraday Soc., 63 (1967) 1754. 40. G.Pezzatini, M.R.Moncelli, R.Guidelli, J.Electroanal.Chem., 301 (1991) 227. 41. D.Rolle, J.W.Schultze, J.Electroanal.Chem., 229 (1987) 141. 42. S.Trasatti, Electrochim.Acta, 37 (1992) 2137. 43. S.Trasatti, J.Electroanal.Chem., 53 (1974) 335. 44. R.Payne, J.Chem.Phys., 42 (1965) 3371. 45. S.Trasatti, J.Electroanal.Chem., 64 (1975) 128. 46. D.C.Grahame, Chem.Rev., 41 (1947)441.

901 47. S.Minc, J.Jastrz@ska, M.Jurkiewicz-Herbich, J.Electroanal.Chem., 152 (1983) 223. 48. M.Jurkiewicz-Herbich, J.JastrzQbska, Pol.J.Chem., 58 (1984) 1125. 49. O.Ikeda, H.Jimbo, H.Tamura, J.Electroanal.Chem., 137 (1982) 127. 50. M.Jurkiewicz-Herbich, Pol.J.Chem., 66 (1992) 1695. 51. W.R.Fawcett, M.D.Mackey, J.Chem:Soc., Faraday Trans. I, 69 (1973) 634. 52. J.H.Sluyters, J.J.C.Oomen, Rec.Trav.Chim., 79 (1960) 1101. 53. J.H.Sluyters, Rec.Trav.Chim., 79 (1960) 1092. 54. R.S.Nicholson, J.Shain, Anal.Chem., 36 (1964) 706. 55. D.Zytner, Elektrochimija, 11 (1975) 1368. 56. R.Andreu, M.Sluyters-Rehbach, J.H.Sluyters, J.Electroanal.Chem., 171 (1984) 139. 57. R.Payne, J.Electroanal.Chem., 41 (1973) 145. 58. E.Dutkiewicz, J.Stuczyflska, Electrochim.Acta, 34 (1989) 1513.

Adsorption and its Applications in Industry and Environmental Protection Studies in Surface Science and Catalysis, Vol. 120 A. Dabrowski (Editor) 9 1998 Elsevier Science B.V. All rights reserved.

903

T h e b i o l o g i c a l s i g n i f i c a n c e of H a l o b a c t e r i a o n n u c l e a t i o n a n d s o d i u m chloride crystal growth A. LSpez-Cort@s and J. L. Ochoa The Center for Biological Research, P.O. Box 128, La Paz 23000, BCS, Mexico

ABSTRACT

The participation of halobacteria in halite formation has been previously considered by a number of authors. In fact, the release of bacterial nucleation factors has been suspected but never properly documented. Some studies show that certain chemicals are capable of modifying sodium chloride crystal habit, and this possibility is of the outmost importance for the correct management of salterns. For example, from over 500 compounds tested, only 12 have been found able to modify the crystal arrangement of halite. The halobacteria can also get entrapped in fluid inclussions within the crystal and retain viability for several years. Thus, they may play some role in contamination of salt preserved foods. Here we review the subject of halobacteria participation in sodium chloride crystal formation and present some results with regard the influence of cell surface layer (S-layer) components on crystal habit. As appears, the halobacteria may influence crystal growth rate and crystal habit, allowing the optimization of saltern management. 1. I N T R O D U C T I O N Cell morphology and cell surface components have been implicated in a wide variety of biological and adhesion phenomena, including mineral and organic crystal formation. Crystallization is used in industry to recover and purify many inorganic and organic materials, but very little work has been done to investigate the feasibility of bulk crystallization for the recovery and purification of proteins and mineral crystals [1]. Crystallization of large molecules, such as proteins, is less known than crystallization of mineral salts, and their mutual effects are poorly understood. For example, it is known that for each protein there appears to be a set of mineral substrates that promote nucleation of protein crystals at lower critical levels of supersaturation than required for spontaneous growth [2,3]. Judged et al. [1] studied the crystallization of ovalbumin in ammonium sulfate and observed that crystal growth could occur without nucleation at a relative supersaturation value of 20. The crystal size distribution was measured and the

904 crystal growth rate was found to be of second-order depending on ovalbumin supersaturation. There was only a slight effect of the ammonium sulfate concentration at pH above 4.9, while the effect of pH itself allowed a 10th-fold increase in the crystal growth rate constant in the range of 4.6-5.4. Under these conditions, the presence of other protein components did not effect the crystal growth rate of ovalbumin. In general, protein supersaturation can eliminate impurity effects [4] and temperature seems to favor crystallization. As a rule, slow crystallization of proteins is due primarily to impedance of the elementary act of entering the growth site, indicating a decisive role of entropy, not of energy barriers, in the crystallization of biological macromolecules. Pressure may actually inhibit crystal yields since, at equilibrium, two conformations of the protein (in the case of lysozyme) have been postulated: one capable of being incorporated into the nuclei of the crystal, and the other which is not [5]. The addition of polyethylene glycol (PEG) into the crystallization medium favors crystal growth by disordering protein aggregates either, in the medium, or at the surface of the crystal, therefore allowing the formation of much larger crystals [6]. In humans, the organic matrix of renal calculi influence the crystal growth that occurs in such pathological mineral deposits and the use of antibodies against these molecules has allowed to visualize their distribution in a variety of normal and pathological mineralized tissues [7]. In the case of cholesterol, nucleation time referred as the time of the first appearance of cholesterol crystals from isotropic crystal-free biles on light microscopy, is being used to assess the potency of nucleating agents such as immunoglobulins [8]. As demonstrated by this test, biliary IgM (M i n m u n o g l o b u l i n s ) , are more potent t h a n biliary IgG (G i m m u n o g l o b u l i n s ) as a potential nucleating agent. A 42 kilo D a l t o n s (KDa) biliary glycoprotein has been also shown to be related with cholesterol crystallization promotion on the pathogenesis of gallstone disease [9]. The protein is an extensively glycosylated (37%) monomer with an i s o e l e c t r i c p o i n t , pI, of 4.1, probably due to its sialic content. Enzymatic N-deglycosylation removes the carbohydrate moiety and inactivates the promoting activity. Enzymatic proteolysis results in both a complete structure degradation and functional inactivation. Biomineralization is the term used to refer to biological metal precipitation, however it should be borne in mind that not necessarily true crystal minerals are formed through this process. Thus, the concept may be missleading. The metal precipitates can exist either intra- or extracellularly, attached or unattached to the cell surface. The adsorption of metals to living or dead cells, which does not involve metabolic energy or transport processess, has been termed "biosportion" [10]. Biosorption has been the metal-microbe interaction most widely studied and covers both the terrestrial and the marine environment. It should be emphasized, however, that biosorption by marine bacteria may have special implications since they occur in a high ionic medium with high concentrations of some metals. Aside of its use for detoxification of heavy metals in aqueous systems [11], very little

905 information is available about the influence of bacteria in metal crystal formation. One important exception may be the cycling of manganese in the marine environment. In general, metal-precipitating bacteria are abundant in hydrothermal vents and the mechanism of metal precipitation may differ greatly among the different types of microorganisms and metals. Metal-microbe interaction can be classified into two classess [12]: a passive process which does not require the direct participation of living organisms and can occur whether the microbes are alive or dead; and an active process, in which some metabolic or enzyme activities are involved. In the first type interaction results as a consequence of the negative charge of microbial cell surfaces and their exopolymers, or through reactions with extracellular complexing agents that can be attached to the cell surface, or released into the medium. This passive sorption of metals is mediated by various functional groups which promote adsorption, ion-exchange, chelation, and/or covalent binding through carboxyl, hydroxyl, sulfhydryl, amino, imino, imidazole, sulfate and/or sulfonate groups, present in the cell surface, in the cell wall, or within the cytoplasm, as polyssachhcarides or glycoproteins. A classical application of this mechanism of sorption is wastewater treatment. Microorganism that can produce extracellular complexing agents such as "siderophores" do so to trap the metals by a passive mechanism that later are internalized by an active process. The active removal of metals occurs via extracellular precipitation, redox reactions, intracellular accumulation, or volatilization [12]. The biomineralization of gypsum, for example, seems to be a two-step process initiated by the binding of calcium to the cell surface following of the binding of sulfate to the calcium. In photosinthesizing cells, the sulfate is eventually replaced by carbonate to form calcite at the cell surface as the pH increases. Thus, the cells provide essential nucleation sites and the chemical conditions necessary for mineral formation [13].

2. THE P H Y S I O L O G I C A L ROLE OF CRYSTAL F O R M A T I O N In polar and sub-polar fishes, for example, some glycopeptides protect their body fluids from freezing. Such peptides prevent the growing of ice by a noncolligative process thus inducing the development of unusual and strikingly similar crystal habits suggesting that the peptides show some affinity for similar crystal faces of ice [14,15]. Ideally, the protein has an exact octapeptide repetition and is assumed to have an helical conformation to control crystal formation [3]. Certain bacteria promote the formation of ice in super-cooled water by means of ice nucleators; the opposite effect, inhibition of ice formation, is common for a group of glycoproteins found in different fish and insect species. These substances termed anti-freeze glycoproteins promote the supercooling of water with no appreciable effect on the equilibrium freezing point, or melting temperature, by binding to a growing ice crystal and slowing crystal growth [16]. In a similar fashion, glycosaminoglycans (GAGs) and some sulphated polysaccharides are

906 involved in preventing urinary stone formation by inhibiting crystal growth and agglomeration, and possibly also nucleation. They can prevent crystal adherence, correct an abnormal oxalate flux, and avoid renal tubular cell damage [17]. Other synthetic polyanions, including peptide analogs of naturally occurring proteins, inhibit the nucleation and growth of calcium salt crystals [18] under physiological conditions.

3. CRYSTAL PROMOTION Bacteria interact with metals, not only because they are needed as nutrients, but also as important agents in their geochemical cyling. Such metal cycles are driven by diverse chemical and biological processes and may be biotechnologically important. For example, in mining industry the bacteria may participate in the oxidation and solubilization of metal sulfide ores and to recover valuable metals from low grade ores. Metal precipitation by bacteria refers to the transformartion of a soluble metal to an insoluble form. Usually, in the first stage of the process the formation of amorphus, highly hydrated, precipitates are obtained, but with aging they can be transformed into crystals [19]. In the crystallization of biomolecules two critical steps, the nucleation of the initial seed and the enlargement of this seed, determine the quality of the final crystal. However, the degree of supersaturation required to nucleate crystals is often higher than the optimal concentration necessary for enlargement. Thus, even under conditions suitable for crystal growth, kinetic factors may prevent the onset of nucleation and crystal growth. Also, spontaneous nucleation may occur at such frequencies that the resulting microcrystalline precipitates are indistinguishable from their amorphous counterparts. It many situations, it is advisable to decouple crystal growth from nucleation in order to grow large, regular crystals. One must not only control the number of seeds, but also reduce the supersaturation level and, therefore, decrease the incorporation of defects detrimental to crystal quality. Seeding techniques provide a preformed, regular crystal surface onto which further molecules may be aggregated in an orderly form, generally at a lower degree of supersaturation than is required for nucleation. Such techniques are ideally suited to bypass the nucleation step, and hence accomplish the decoupling between nucleation and crystal growth. Three aspects for seeding should be considered: 1) preequilibration of the solution to be seeded and determination of the proper supersaturation level for seeding; 2) the environment and necessary precaution for seeding; and 3) the streak seeding technique and how it can be used in conjunction with microseeding and macroseeding [20]. In the case of some proteins, crystallization may be affected by the aminoterminal segment that sticks out to interact with a symmetry related molecule through an intermolecular salt-bridge. For example, removal of the r e s i d u e lysine in position 38 (Lys 38), in the case of an endonuclease from Clostridium thermocellum, or the substitution of its bridge-forming residues by site-directed

907 mutagenesis, promotes crystal packing arrangements different from the wild type enzyme [21]. Flexible amino-, and/or carboxy-terminal extensions, influence crystal nucleation but not crystal growth [21]. Following this idea, an attempt to facilitate crystallization has been the use of engineering cysteins to promote formation of a "back-to-back" dimmer that occurs in different crystal forms of wild-type and m u t a n t lysozymes [22]. The designed double m u t a n t in which the surface residues of a s p a r a g i n e Ash68, and a l a n i n e Ala93, were replaced by cysteines, formed dimers in solution and crystallized isomorphously to wilde-type but a much faster rate. Overall, the m u t a n t structure remained very similar to wild-type despite the formation of two intermolecular disulfide bridges. The results suggest that the formation of the lysozyme dimmer is a critical intermediate in the formation of more than one crystal form and that covalent cross-linking of the intermediate accelerates nucleation and facilitates crystal growth. Another role seems to be played by polysaccharides in the case of the calcifying algae Pleurochrysis carterae, which produces structures known as coccoliths in homogeneous cell cultures. The polysaccharides PS-1 and PS-2 have been localized in the crystal coats of mature coccoliths, and in electron dense Golgi particles. These polyanions are synthesized in medial Golgi cisternal and co-aggregate with calcium ions into discrete 25 nm particles. The polysaccharides remain with the mineral phase after the coccoliths are extruded from the cells [23]. The effect of many different compounds on calcium stone formation has been evaluated [24]. A few appear to inhibit the nucleation rate, growth and suspension density (crystal mass produced) in proportion to its concentration. On the other hand, glucose, glycerol and certain amino acids which are recognized as osmoregulators, and are produced by halophylic microorganisms, have been also evaluated as halite crystal habit modifiers [25]. In the same way, silica gel as an inert particle, and the ferrocyanide ability to form dendrite crystals of halite, have been already reported [26,27]. Altogether these data point to a dramatic effect of different compounds on mineral and protein crystallization and have induced us to study the influence of halobacteria, and of its components, in the nucleation, shape and growth of NaC1 crystals under natural conditions. 4.

I N F L U E N C E OF HALOBACTERIA ON HALITE FORMATION

It seems that since halobacteria have not being connected to any dreadful disease, and because they are easy to grow, albeit very slowly for practical purposes, interest on their biotechnological use is very scant. In fact, their ecological role is still poorly understood [28]. The term halobacteria refers to the halophylic Archaea, not to halotolerant bacteria, and this distinction is of outmost importance. The order Halobacteriales contain six genera and the number of newly found species is increasing. Phylogenetic data indicate that they are among the most modern Archaea, as their strong preference for aerobic life

908 would suggest. Halobacteria require a minimum of 1.5 M (9%) sodium chloride for growth, and in most cases the optimum lies between 3.5-4.5 M (21-27%) NaC1. Such salt concentration exceeds by far the total saltiness of sea water (which is about 0.6 M or 3.5% of dissolved salts). As a particular feature, halobacteria exhibit active growth and motility in saturated salt solutions, only reduced when entrapped into the salt crystals they bump into. In the absence of salt, all except the coccal forms of halobacteria disrupt promptly and dramatically. In any location where the basic requirement of salt is met, halobacteria will be found. They may become the dominant microflora in what appears to be a classical ecological succession [28]. The halobacteria can be grouped into three major types reflecting their natural source: in the first category are habitats in which the salt mixture derives from evaporated seawater. The commercial salterns consist of shallow evaporating pools containing brines of steadily increasing salinity. They are an ideal ecosystem to observe and study microbial succession. As the brine concentrates under the hot sun and wind, its density and, therefore, thermal storage capacity increase, while its ability to hold dissolved oxygen declines. Certain ions reach saturation earlier than others and precipitate out. For example, calcium and sulfate crystallize out as gympsum. In the less-saline stages ( F +

O~

(14)

Then, due to the interaction with the quencher, the fluorophore comes back to the ground state, without the emission of fluorescence. In both cases, the relationship between the fluorescence intensity I and the concentration of quencher [Q] is:

932 I

1

I0

1+ K[Q]

(15)

where Io is the fluorescence in the absence of the quencher; K is a constant equal to the dissociation constant of equation (13) in the case of static quenching, and is the Stern-Volmer constant in the case of dynamic quenching (the Stern-Volmer constant is equal to the product kq.to, between the diffusion-controlled rate constant kq and the fluorescent lifetime to of the excited state F* in the absence of the quencher). A decrease in the intensity of the fluorescence may also be due to an energy transfer from fluorophore F* in the excited state to another molecule, acceptor A, whose absorption spectrum, modulated by the chemical species under investigation, overlaps the emission spectrum of the fluorophore. Therefore, fluorescence and absorption can be combined to detect a chemical parameter. In this case, the fluorescence intensity in presence of acceptor, I is given by: I

--=I-~] Io

(16)

where Io is the intensity of the fluorescence in the absence of the acceptor, and ~l is a term depending on the distance between fluorophore F and acceptor A. If other chromophores are present in the solution being tested, a decrease can be observed in the fluorescence, caused by the absorption of the excitation light (primary inner filter effect) or of the emission light (secondary inner filter effect) by these chromophores. It is obvious t h a t in this case, the previous equations are no longer valid, but t h a t corrective terms are necessary. In the case of dynamic quenching, it is more convenient to consider the timedependent decay [7]. In the presence of an interaction with the excited state, the lifetime of the fluorophore is decreased: the higher the concentration of the quencher, the greater the decrease in the lifetime. This is not the case for static quenching, in which the lifetime of the fluorophore is not affected by a change in the concentration of the quencher. Typical fluorescence decay times are in the range between 2 and 20 nsec, while phosphorescence decay times are in the 1 psec + 10 sec range. According to Stern and Volmer, the relationship between the decay time and the concentration is: T to

:

1

(17)

1 + Ksv[Q ]

where t and to are the decay times of the excited state of the fluorophore in the presence and in the absence of the quencher, respectively. Lifetime can be measured either in the time domain or in the frequency domain. In the former case, the fluorophore is excited with a narrow pulse and the

933 fluorescence decay is monitored. In the latter case, a modulated excitation is used: the fluorescence emission is still modulated at the same frequency, but is diminished in amplitude and phase shifted. The extent of the amplitude decrease and of the phase shift depends on the lifetime of the fluorophore. When ground-state reactions are involved, lifetime measurements can be performed if reagents and products Of the reactions are fluorescent species characterized by different decay times, as occurs in some acid-base reactions [8,9]. The advantage of this approach lies mainly in the fact that there is no more dependence on loading or photobleaching of the chemical transducer fixed at the end of the optical fibres, which is one of the greatest drawbacks of intensity-modulated chemical sensors. Moreover, no problems arise from possible fluctuations or drift in the source intensity or photodetector sensitivity which, on the contrary, heavily affect intensity of modulated sensors. 3.3.

Evanescent

wave field

Interaction of the chemical species with the evanescent wave field can be exploited for sensing purposes [10]. The electromagnetic field which propagates along the fibre inside the core extends also in the cladding region. The solution to Maxwell's equations shows that, in the presence of total internal reflection, a standing wave (called evanescent wave) exists in the cladding, propagates in the direction of the fibre axis and decays exponentially in the direction perpendicular to the core/cladding interface. The penetration depth of the evanescent wave is a key parameter for sensing purposes; it is the distance, from the cladding, at which the amplitude of the electromagnetic is decreased by a factor equal to 1/e and is expressed by the following formula (valid for a step-index fibre): dp=

~-0 1 2~n1 [sin2~_sin2~c~/2

(18)

Typical values of penetration depth are in the order of the utilized wavelength. For example, if n1=1.5 and n2=1.33 (aqueous medium) the minimum value of the penetration depth ( 3 = 90 ~ is about U5 and increases upwards by about 1 wavelength for angles 1~ greater than the critical angle. The penetration depth goes to infinity in correspondence of the critical angle. However this fact can be disregarded since, for angles close to the critical angle, the fibre is characterized by losses due to the scattering coming from the surface roughness. In practice, the evanescent wave field is limited to within few microns or less from the core surface. The efficiency of the approach depends on the fraction r of the optical power carried by the fibre which propagates in the cladding. The optical power carried by the core represents a background since it is not modulated by the analyte: higher this background is, lower are the performances of the evanescent wave sensors. This fraction is clearly high for monomode fibres (r>0.5) and, in multimode fibres, for the modes close to the cut-off condition (the so-called higher modes). In

934 multimode fibres the average power in the cladding, coming from the contribution of all modes, has to be considered. It has been shown [11] that for weakly guiding multimode fibres (nl -~ n2) the value of r is given by the following relationship: r = Pcla____dd= 4x/-2 Ptot 3V

(19)

For multimode fibres typical values of r are in the range of 0.01. Any changes in the microenvironment close to the fibre core which is ascribed to chemical species and modifies the evanescent field distribution can be used in the development of optical sensors. Modification of the refractive index of the cladding due to the penetration of the chemical species in the region close to the core is the sensing mechanism most followed. This approach is not characterized by selectivity, which can be reached by combining the evanescent wave analysis with the absorption or fluorescence coming either directly from the analyte or from the proper chemical transducer, which is located in the proximity of the fibre core. 4.

THE OPTODE

The term optode (or the more used word optrode: the first is the correct term, however, since its etymology comes from the Greek " 6 ~ ! 65o~") denotes the probe of the optical fibre chemical sensor: that is, the chromophore with its mechanical envelope, if existing, and the portion of the fibre in contact with this chemical transducer. The simplest case is when the investigated species has optical properties: in this case, development of the optode is reduced to the manufacturing of an optimized optical cell to be connected to the fibre. As explained above, if a chemical transducer is utilized, use is made of a special reagent, whose optical properties vary in accordance with the variation in the concentration of the p a r a m e t e r under examination. The reaction may be direct as in pH sensors in which the hydrogen ions react with an acid-base indicator or a fluorophore, causing a variation in the absorption or fluorescence, respectively. In other cases, the reaction which takes place in the optode gives an opticallydetectable compound as final product. An example is the detection of carbon dioxide, which is based on the detection of the pH of a carbonated solution, the acidity of which changes according to the quantity of CO2 dissolved therein. Enzyme-based sensors are included in this latter category. Detection is based on a selective conversion, catalysed by a special enzyme, of the p a r a m e t e r under investigation. Among the reagents or the products of the enzymatic reaction, there is a species which is optically detectable. In glucose sensors, the consumption of oxygen is detected when glucose reacts in the presence of glucose oxydase; in a penicillin sensor, instead, the production of hydrogen ions in the presence of the enzyme penicillase is exploited.

935 Two types of optode can be distinguished: - extrinsic optodes: the chemical transducer is immobilized on an external support, such as glass or polymeric matrix; in this case, a mechanical envelope is necessary for attaching the support to the tip-end of the fibre; - intrinsic optodes: the chemical transducer is directly immobilized on the fibre. This can be done at the fibre tip or along the fibre on the core: in this case, a compact and highly-miniaturized structure is attained, since the probe is practically the fibre itself. In an optode, a fast and reliable interaction with the investigated analyte m u s t be guaranteed. Moreover, if a chemical transducer is used, an appropriate immobilization process should be developed. Adsorption can play an i m p o r t a n t role in both cases, either by assuring a good exchange of the analyte between the probe and the external environment, or by offering a simple immobilization process for the sensitive reagent. 5.

A D S O R P T I O N AS AN E X C H A N G E M E C H A N I S M F O R T H E DETECTED ANALYTES

Adsorption of the chemical compounds under investigation on appropriate layers deposited along the fibre core can offer some advantages on traditional bulk procedures.

cladding

core

'

............................ " ';

I

modified cladding Figure 4. Sketch of the modified fibre for sensing purposes by means of evanescent wave technique.

Figure 4 provides a sketch of the fibre modified for sensing purposes. The deposited layer works as new cladding for the fibre, provided t h a t it has a refractive index smaller t h a n the refractive index of the fibre core. Since the interaction between the light carried in the fibres and the external environment occurs by means of the evanescent field, the sensing area is limited to a few microns from the core/cladding interface. If the new cladding has a thickness greater t h a n the penetration depth, the investigated volume is inside the cladding; therefore, the effect of interfering elements which may be found in the investigated sample can be

936 eliminated through the choice of an appropriate layer. A clear a p p a r e n t example is in the case of the analysis of species in aqueous samples, such as dissolved gases or nonpolar and organic compounds. If the m e a s u r e m e n t is performed with bare fibres, the evanescent field extends in the aqueous phase, and the presence of dissolved species cannot be detected, because of the strong absorption of the water. The choice of hydrophobic and organophilic layer can prevent the diffusion, close to the fibre core, of w a t e r molecules. These cannot interfere with the m e a s u r e m e n t , which makes it possible to detect the dissolved species, which is adsorbed inside the layer (Figure 5). 0 0.,,,.

0

~

0

oo o o,O,o , o Oo&, o, o.O, o d_.,.o o o o . o o

o

water

9~ a n a l y t e

modified cladding uJ

fibre c o r e

Figure 5. Adsorption of the analyte in the modified cladding: on the left side the exponential decay of the electric field outside the fibre core along with the penetration depth is shown.

For sensing purposes, the ideal layer should have the following characteristics: - a refractive index smaller t h a n the refractive index of the fibre core; - selectivity to the investigated compounds; - efficient adsorption and desorption processes, which would lead to a good reversibility of the sensor; - fast diffusion of the investigated compounds, which would m e a n a fast response time of the sensor. Modulation of the light carried by the fibre can be induced in two ways: 9 a change in the refractive index of the cladding after the adsorption, which gives rise to a change in the penetration depth of the evanescent field. The selectivity of the sensor is determined only by the selectivity of the cladding material, since the sensor responds to all the adsorbed species. Its a d v a n t a g e lies in the simplicity of the optoelectronic system, since no requirements are given regarding the choice of the wavelength; 9 light absorption in correspondence with the absorption bands of the analysed species. Higher selectivity can be attained, but a multiwavelength system should be used either in the NIR region, if the overtones of chemical compounds are utilized, or in the infrared region, if the vibrational bands are exploited.

937 Clearly, t e m p e r a t u r e changes can affect this type of m e a s u r e m e n t [12]. A variation in temperature has two effects: i) a change in the cladding density, due to the volume change, and ii) a change in the refractive indices of the core and of the cladding. The first effect gives rise to a change in concentration of the polymer cladding, while the second gives rise to a change in the penetration depth of the evanescent field. In any case, a change in the detected signal is observed, and a t e m p e r a t u r e compensation is necessary. 5.1.

Sensors based on refractive index changes

Since, in this case, the sensor is practically a refractometer, a requisite characteristic for the deposited layer is its high capability to change its refractive index with the adsorption of the investigated chemical compound. Refractive index changes of the order of 10 .5 can be detected with optical fibres [13]. As already pointed out, the selectivity of the sensor in this case depends only on the selectivity of the cladding; therefore its choice is crucial. The first system based on this principle was proposed for the detection of hydrocarbons in water [14]. An unclad fibre (core diameter 140 ~m) was silylated for a length of 80 cm. Different chemical reagents (RnSiX4-n) were used to realize the organophilic cladding: in particular, two silylating agents having the same R group (R = -ClsH37) but different leaving group (X=OCH2CH3, X=C1), namely, octadecyltriethoxysilane and octadecyltrichlorosilane. Other reagents used have the same leaving group (X=C1) but different R groups (R--Ph or R=n-C10H21), diphenyldichlorosilane and n-decyltrichlorosilane. An He-Ne laser (632.8 nm) was used as source, and the detector consisted of appropriate electronic circuitry with a simple pin photodiode. Laboratory tests on different contaminants indicated that the type of coating and the method of applying it affected the capacity for absorbing hydrocarbons. The detection limit was variable (400 rag/1 for p-xylene, 3 mg/1 for crude oil). Reversibility of the sensor was obtained by washing the coated fibre with acetone and methanol. Heteropolysiloxane polymers were also used for the detection of chemical vapours [15]: by incorporating different functions (amino, vinyl, glycidoxypropyl) inside the polymers, different sensitivities and selectivities were obtained for different chemical vapours. In the experimental tests, light from a laser diode 0~=670 nm) was coupled into two fibres (core diameter 600 pm). Angular excitation of the fibre was performed in order to excite the higher modes of the fibre and to increase the fraction of the optical power carried by the cladding. One of the fibres was stripped for a length of about 2.5 cm and coated with the vapour-sensitive coating; the other one was used as a reference, as it was insensitive to the vapour in contact with this fibre. Both fibres were inserted into a flow-cell. The signals coming out from the two fibres were detected by two photodiodes; their ratio resulted less sensitive to temperature changes. A detection limit of 100 ppm was obtained for toluene. The same system was tested for methane using a polyoxyethylene lauryl ether polymer as cladding material. A detection limit of 2% (in vol) in the air was obtained [13].

938 Selectivity still remains a problem when this approach is followed, but in principle the analysis of compound mixtures could be performed with several sensors, each coated with a different organophilic compound and properly calibrated. An appropriate processing system should be capable of discriminating the different compounds. On the other hand, selectivity is not a fundamental requirement for some applications. An example of this is the detection of hydrocarbon leakages from storage tanks, or remediation efforts. Klainer et al. developed an optical fibre sensor for the analysis of petroleum hydrocarbons, both in water and in vapour [16]. The new version of the system, called Petrosense CMS 5000, is at present distributed by Whessoe Varec [17]. Figure 6 shows a photo of the optical probe: the length of the probe is 25.5 cm, and the sensing area is limited to 2.5 cm of a glass fibre covered with a proprietary coating which attracts C6 and higher petroleum hydrocarbons. Table 2 describes the technical specifications as declared by the distributor.

Figure 6. Photo of the optical fibre probe for direct monitoring of hydrocarbons.

Table 2 Technical specifications of the hydrocarbon sensor Vapor Operating range

Water

0-20,000 ppm as TPH

Lower detection limit

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1013

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1014

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1015

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1016

ION EXCHANGE,

CHROMATOGRAPHY,

AND RELATED

SEPARATIONS

1967 Copeland, J.P.; Henderson, C.L., and Marchello, J.M., Influence of resin selectivity on film diffusion-controlled ion exchange, AIChE J., 13(3), 449-452 (1967). Gilwood, M.E., Saving capital and chemicals with countercurrent ion exchange, Chem. Eng. (N.Y.), 18 December, 83-88 (1967). Klein, G.; Tondeur, D., and Vermeulen, T., Multicomponent ion exchange in fixed beds, Ind. Eng. Chem. Fund., 6(3), 339-364 (1967). Lai, C.L., and Roth, J.A., Dynamic simulation of gas chromatographic column, Chem. Eng. Sci., 22(10), 12991304 (1967). Solt, G.S., Continuous countercurrent ion exchange, Brit. Chem. Eng., 12(10), 1582-1586 (1967). Tallmadge, J.A., Ion exchange treatment of mixed electroplating wastes, Ind. Eng. Chem. Process Des. Dev., 6(4), 419-423 (1967). 1968 George, D.R.; Riley, J.M., and Ross, J.R., Potassium recovery by chemical precipitation and ion exchange, Chem. Eng. Prog., 64(5), 96-99 (1968). Hall, G.R.; Streat, M., and Creed, G.R.B., Ion exchange in nuclear chemical processes, Trans. IChemE, 46, T53T59 (1968). Lifshutz, N., and Dranoff, J.S., Inversion of concentrated sucrose solutions in fixed beds of ion exchange resin, Ind. Eng. Chem. Process Des. Dev., 7(2), 266-269 (1968). Michalson, A.W., High quality water via ion exchange, Chem. Eng. Prog., 64(10), 67-73 (1968). Ryan, J.M.; Timmins, R.S., and O'Donnell, J.F., Production-scale gas chromatography, Chem. Eng. Prog., 64(8), 53-59(1968). Schneider, P., and Smith, J.M., Adsorption rate constants from chromatography, AIChE J., 14(5), 762-771 (1968). Tuichiev, I.S.; Rizaev, N.U.; Merenkov, K.V., and Yusipov, M.M., Hydrodynamic properties of ion exchange resins during fluidization, Int. Chem. Eng., 8(2), 221-223 (1968). Turner, J.C.R., and Snowdon, C.B., Liquid-side mass-transfer coefficients in ion exchange: Nernst-Planck model, Chem. Eng. Sci., 23(3), 221-230; 23(9), 1099-1104 (1968). Turner, J.C.R.; Snowdon, C.B.; Jones, D.C., and Ward, J.W.C., Estimation of ion-exchange equilibrium diagrams involving weakly dissociated electrolytes, Trans. IChemE, 46, T232-T235 (1968). 1969 Cooke, J.P., Understanding a gas chromatograph, Chem. Eng. (N.Y.), 10 March, 134-144 (1969). Copeland, J.P., and Marchello, J.M., Film-diffusion controlled ion-exchange with a selective resin, Chem. Eng. Sci., 24(9), 1471-1474 (1969). Eberly, P.E., Diffusion studies in zeolites and related solids by gas chromatography, Ind. Eng. Chem. Fund., 8(1), 25-30(1969). Kadlec, V., and Matejka, Z., Mixed-bed deionisation by weak electrolyte ion-exchange resins regenerated in-situ by carbon dioxide, J. Appl. Chem., 19, 352-355 (1969). Kuong, J.F., Maximising ion-exchanger throughput, Chem. Eng. (N.Y.), 15 December, 160 (1969). Pollio, F.X.; Kunin, R., and Petralia, J.W., Treat sour water by ion exchange, Hydrocarbon Process., 48(5), 124126 (1969). Timmins, R.S.; Mir, L., and Ryan, J.M., Large-scale chromatography: New separation tool, Chem. Eng. (N.Y.), 19 May, 170-178 (1969). 1970 Campbell, D.O., and Buxton, S.R., Rapid ion exchange separations, Ind. Eng. Chem. Process Des. Dev., 9(1), 89-99 (1970). McGovern, T.J., and Dranoff, J.S., Sucrose inversion by partially deactivated ion-exchange resin beds, AIChE J., 16(4), 536-538 (1970). Streat, M., and Brignal, W.J., Representation of ternary ion exchange equilibria, Trans. IChemE, 48, T151-T155 (1970). Turner, J.C.R., and Snowdon, C.B., Liquid-side mass transfer coefficients in ion exchange, Chem. Eng. Sci., 25(11), 1673-1678 (1970).

1017

Weber, O.W.; Miller, I.F., and Gregor, H.P., Absorption of carbon dioxide by weak-base ion exchange resins, AIChE J., 16(4), 609-614 (1970). 1971 Colwell, C.J., and Dranoff, J.S., Nonlinear equilibrium anJ axial mixing effects in intraparticle diffusioncontrolled sorption by ion-exchange resin beds, Ind. Eng. Chem. Fund., 10(1), 65-70 (1971). Danes, F., Batch process application to ion-exchange unit operation, Chem. Eng. Sci., 26(8), 1277-1288 (1971). Gardiner, W.C., and Munoz, F., Mercury removal from waste effluent via ion exchange, Chem. Eng. (N.Y.), 23 August, 57-59 (1971 ). Gondo, S.; Itai, M., and Kusunoki, K., Computational and experimental studies on a moving ion-exchange bed, Ind. Eng. Chem. Fund., 10(1), 140-146 (1971). Heines, V., A history of chromatography, Chemtech, May, 280-285 (1971). Hsu, H.W., Optimum adsorbent volume in liquid adsorption chromatography, Sep. Sci., 6(5), 645-652 (1971). Kunin, R., and Downing, D.G., Ion-exchange system boasts more pulling power, Chem. Eng. (N.Y.), 28 June, 67-69 (1971). Maldacker, T.A., and Rogers, L.B., Effect of loading on separation efficiency using steric exclusion chromatography, Sep. Sci., 6(6), 747-758 (1971). Mir, L., Comparison of static bed and moving bed chromatography, Sep. Sci., 6(4), 515-536 (1971). Moreland, A.K., and Rogers, L.B., Effects of slow mass transfer using microporous adsorbents in gas-solid chromatography, Sep. Sci., 6(1 ), 1-24 (1971 ). Smuts, T.W.; Jordaan, J.T., and Pretorius, V., Phenomenological plate height equation for packed chromatographic columns, Sep. Sci., 6(5), 653-684 (1971). Various, Gel permeation chromatography (topic issue), Sep. Sci., 6(1), 47-164; 6(2), 207-330 (1971). 1972 Cloete, C.E., and de Clerk, K., Distillation vs. chromatography: Comparison based on purity index, Sep. Sci., 7(4), 449-456 (1972). Conrard, P.; Caude, M., and Rosset, R., Separation of close species on ion exchangers, Sep. Sci., 7(5), 465-490 (1972). Dodds, J.A., and Tondeur, D., Design of cyclic fixed-bed ion-exchange operations, Chem. Eng. Sci., 27(6), 1267-1282; 27(12), 2291-2298 (1972). Golden, L.S., and Irving, J., Osmotic and mechanical strength of ion-exchange resins, Chem. Ind. (London), 4 November, 837-844 (1972). Holliday, D.C., Continuous ion exchange: Design and development, Chem. Ind. (London), 16 September, 717723 (1972). Parker, K.J., Ion exchange in the sugar industry, Chem. Ind. (London), 21 October, 782-790 (1972). Qureshi, M.; Qureshi, S.Z.; Gupta, J.P., and Rathore, H.S., Progress in ion-exchange studies on insoluble salts of polybasic metals, Sep. Sci., 7(6), 615-630 (1972). 1973 Buys, T.S., and de Clerk, K., Effect of temperature on production rate in chromatography, Sep. Sci., 8(5), 551566(1973). Conrard, P.; Caude, M., and Rosset, R., Separation of close species on ion exchangers, Sep. Sci., 8(1), 1-10; 8(2), 269-278 (1973). Johnson, J.F.; Macphail, M.G.; Cooper, A.R., and Bruzzone, A.R., Effect of column length on chromatographic fractionation of polymers, Sep. Sci., 8(5), 577-584 (1973). Lal, B.B., and Douglas, W.J.M., Techniques for measuring sorption of water by ion-exchange resin spheres, lnd. Eng. Chem. Fund., 12(3), 381-384 (1973). Letan, R., Continuous ion-exchanger, Chem. Eng. Sci., 28(3), 981-985 (1973). Martin, J.R., and Johnson, J.F., Cost-efficiency comparisons of some polymer chromatographic fractionation techniques, Sep. Sci., 8(5), 619-622 (1973). Meares, P., Characteristics and uses of ion exchange membranes, Chem. Ind. (London), 1 December, 103-107 (1973). Metzger, V.G.; Barford, R.A., and Rothbart, H.L., Chromatography and countercurrent distribution, Sep. Sci., 8(2), 143-160 (1973). Millar, J.R., Fundamentals of ion exchange, Chem. Ind. (London), 5 May, 409-413 (1973).

1018

Ouano, A.C., and Barker, J.A., Computer simulation of linear gel permeation chromatography, Sep. Sci., 8(6), 673-700(1973). Stevens, B., Chromatographic refining unit, Process Eng. (London), March, 82-84 (1973). Weiss, G.H., and Dishon, M., Resolution in nonuniform chromatographic systems, Sep. Sci., 8(3), 337-344 (1973). Williams, R.C., Ion exchange resins in power stations, Chem. Ind. (London), 19 May, 465-470 (1973). 1974 Braud, C., and Selegny, E., Interrelation of swelling and selectivity of ion-exchange resins, Sep. Sci., 9(1), 13-26 (1974). Bull, P.S.; Evans, J.V., and Nicholson, F.D., Condensate polishing performance of powdered ion-exchange resins, J. Appl. Chem. Biotechnol., 24, 475-486 (1974). Caude, M.; Conrard, P., and Rosset, R., Displacement development on ion exchangers, Sep. Sci., 9(4), 269-286 (1974). Dodds, J.A., and Tondeur, D., Design of cyclic fixed-bed ion-exchange operations, Chem. Eng. Sci., 29(2), 611620(1974). Kirchner, J.G., Thin-layer chromatography, Chemtech, February, 79-82 (1974). Lal, B.B., and Douglas, W.J.M., Equilibrium water sorption and volumetric behavior of ion-exchange resin spheres, Ind. Eng. Chem. Fund., 13(3), 223-227 (1974). Nikelly, J.G., Porous-layer open-tubular gas chromatography columns, Sep. Purif. Methods, 3(2), 423-441 (1974). Scott, C.D., High-pressure ion-exchange chromatography applied to separation of complex biochemical mixtures, Sep. Purif. Methods, 3(2), 263-298 (1974). Singhal, R.P., Separation and analysis of nucleic acids and their constituents by ion-exclusion and ion-exchange column chromatography, Sep. Purif. Methods, 3(2), 339-398 (1974). Various, Liquid-liquid extraction and ion exchange in analytical chemistry (topic issue), Chem. Ind. (London), 17 August, 639-647 (1974). Various, Chromatographic separations (topic issue), Sep. Purif. Methods, 3(1), 1-86, 133-244 (1974). Whitlock, L.R., and Siggia, S., Fusion reaction gas chromatography, Sep. Purif. Methods, 3(2), 299-338 (1974). 1975 Bolto, B.A., Sirotherm ion-exchange desalination, Chemtech, May, 303-307 (1975). Braud, C., and Selegny, E., Interrelation of swelling and selectivity of ion-exchange resins, Sep. Sci., 10(1), 47110; 10(2), 175-244; 10(3), 331-358 (1975). Farkas, E.J., and Himsley, A., Fundamental aspects of behavior of ion exchange equipment, Can. J. Chem. Eng., 53,575-585 (1975). Kataoka, T.; Nishiki, T., and Ueyama, K., Mass transfer with liquid anion exchange, Chem. Eng. J., 10(3), 189196(1975). Pawlowski, L., and Zytomirski, S., Influence of ion exchange capacity and total concentration of solution of ions of different valency on their chromatographic separation, Sep. Sci., 10(1 ), 33-38 (1975). Prengle, H.W., et al., Recycle wastewater by ion exchange, Hydrocarbon Process., 54(4), 173-184 (1975). Rendell, M., Future of large-scale chromatography, Process Eng. (London), April, 66-70 (1975). Vermeer, D.J.; Lynn, S., and Vermeulen, T., Cation-exchange column behavior in desalination process with regenerant recovery, Ind. Eng. Chem. Process Des. Dev., 14(3), 290-297 (1975). 1976 Cooper, A.R., and Lynn, T.R., Coiled high-efficiency liquid chromatography columns, Sep. Sci., 11(1), 39-44 (1976). de Rosset, A.J.; Neuzil, R.W., and Korous, D.J., Liquid column chromatography as a predictive tool for continuous countercurrent adsorptive separations, Ind. Eng. Chem. Process Des. Dev., 15(2), 261-266 (1976). Ito, Y., and Bowman, R.L., Foam countercurrent chromatography, Sep. Sci., 11(3), 201-206 (1976). Kadlec, V., and Hubner, P., Ion exchange deionisation with recirculation of regenerant by heat, Chem. Ind. (London), 4 September, 744-746 (1976). Kataoka, T.; Nishiki, T., and Ueyama, K., Simultaneous mass transfer of acid and ions in a liquid anion exchanger, Chem. Eng. J., 12(3), 233-238 (1976). Mostafa, H.A., and Said, A.S., Theoretical-plate concept for fixed-bed adsorption and ion-exchange, Trans. IChemE, 54, T132-T134 (1976).

1019

Roland, L.D., Ion exchange: Operational advantages of continuous plants, Processing (Sutton, Engl.), January, 11-12 (1976). Slater, M.J., and Lucas, B.H., Flow patterns and mass transfer rates in fluidized-bed ion-exchange equipment, Can. J. Chem. Eng., 54, 264-270 (1976). Smirnov, N.N., Mathematical models of ion-exchange process, Int. Chem. Eng., 16(2), 234-240 (1976). Sussman, M.V., Continuous chromatography (review paper), Chemtech, April, 260-264 (1976). Talmon, Y., and Rubin, E., Chromatographic separation by foam, Sep. Sci., 11(6), 509-533 (1976). Weatherley, L.R., and Turner, J.C.R., Ion-exchange kinetics: Comparison between a macroporous and a gel resin, Trans. IChemE, 54, T89-T94 (1976). 1977 Holl, W., and Sontheimer, H., Ion exchange kinetics of the protonation of weak-acid ion-exchange resins, Chem. Eng. Sci., 32(7), 755-762 (1977). Pauls, R.E., and Rogers, L.B., Comparisons of methods for calculating retention and separation of chromatographic peaks, Sep. Sci., 12(4), 395-415 (1977). Pauls, R.E., et al., Experimental variables in recycle gas chromatography, Sep. Sci., 12(3), 289-306, (1977). Pusch, W., Ion-exchange membranes, Int. Chem. Eng., 17(1), 62-75 (1977). Shah, D.B., and Ruthven, D.M., Measurement of zeolitic diffusivities and equilibrium isotherms by chromatography, AIChE J., 23(6), 804-809 (1977). Umbreit, G.R., Chromatographic anomalies, Chemtech, February, 101-106 (1977). Various, Novel ion exchangers (topic issue), Chem. Ind. (London), 6 August, 634-652 (1977). 1978 Barker, P.E.; Ellison, F.J., and Hatt, B.W., Countercurrent chromatographic unit for continuous fractionation of dextran, Ind. Eng. Chem. Process Des. Dev., 17(3), 302-309 (1978). Chihara, K.; Suzuki, M., and Kawazoe, K., Adsorption rate on molecular sieving carbon by chromatography, AIChE J., 24(2), 237-246 (1978). Danesi, P.R., and Chiarizia, R., Mass transfer rate with liquid ion exchangers, J. Appl. Chem. Biotechnol., 28, 581-598(1978). De, A.K., and Sen, A.K., Synthetic inorganic ion-exchangers, Sep. Sci. Technol., 13(6), 517-540 (1978). Hubner, P., and Kadlec, V., Kinetic behavior of weak-base anion exchangers, AIChE J., 24(1), 149-154 (1978). Marra, R.A., and Cooney, D.O., Multicomponent sorption operations: Bed shrinking and swelling in an ionexclusion case, Chem. Eng. Sci., 33(12), 1597-1602 (1978). Smith, R.P., and Woodburn, E.T., Prediction of multicomponent ion exchange equilibria for ternary systems from binary systems data, AIChE J., 24(4), 577-587 (1978). Wiley, J.R., Decontamination of alkaline radioactive waste by ion exchange, Ind. Eng. Chem. Process Des. Dev., 17(1), 67-71 (1978). 1979 Abe, M., and Kasai, K., Synthetic inorganic ion-exchange materials, Sep. Sci. Technol., 14(10), 895-908 (1979). Agarwal, J.C., and Klumpar, I.V., Role of liquid ion exchange in processing of complex solutions, J. Chem. Technol. Biotechnol., 29, 730-740 (1979). Erickson, K.L., and Rase, H.F., Fixed-bed ion exchange with differing ionic mobilities and nonlinear equilibria, Ind. Eng. Chem. Fund., 18(4), 312-317 (1979). Goto, S.; Sato, N., and Teshima, H., Periodic operation for desalting water with thermally regenerable ionexchange resin, Sep. Sci. Technol., 14(3), 209-218 (1979). Gupta, A.R., Isotope effects in ion-exchange equilibria in aqueous and mixed solvent systems, Sep. Sci. Technol., 14(9), 843-858 (1979). Hadzismajlovic, D.E., et al., Mass transfer in liquid spout-fluid beds of ion exchange resin, Chem. Eng. J., 17(3), 227-236 (1979). Knaebel, K.S.; Cobb, D.D.; Shih, T.T., and Pigford, R.L., Ion-exchange rates in bifunctional resins, Ind. Eng. Chem. Fund., 18(2), 175-180 (1979). Various, Ion exchange in the water industry (topic issue), Chem. Ind. (London), 3 March, 142-165 (1979).

1980 Brown, J.M., and Wilson, D.J., Macroreticular resin columns, Sep. Sci. Technol., 15(8), 1533-1555 (1980). Burfield, D.R., and Smithers, R.H., Desiccant efficiency in solvent drying: Applications of cationic exchange resins, J. Chem. Technol. Biotechnol., 30, 491-496 (1980).

1020

Calmon, C., Explosion hazards of using nitric acid in ion-exchange equipment, Chem. Eng. (N.Y.), 17 November, 271-274 (1980). Curtis, M.A., et al., Liquid chromatographic fractionations of mixtures of polystyrene oligomers, Sep. Sci. Technol., 15(7), 1413-1428 (1980). Kennedy, D.C., Predict sorption of metals on ion-exchange resins, Chem. Eng. (N.Y.), 16 June, 106-118 (1980). MacLean, G.T., Effect of synthetic flocculant on ion-exchange resin, Sep. Sci. Technol., 15(8), 1555-1563 (1980). Omatete, O.O.; Clazie, R.N., and Vermeulen, T., Column dynamics of ternary ion exchange, Chem. Eng. J., 19(3), 229-250 (1980). Ruthven, D.M., and Kumar, R., An experimental study of single-component and binary adsorption equilibria by a chromatographic method, Ind. Eng. Chem. Fund., 19(1), 27-32 (1980). Soldatov, V.S., and Bichkova, V.A., Ternary ion-exchange equilibria, Sep. Sci. Technol., 15(2), 89-1 l0 (1980). Takahashi, T., and Gill, W.N., Hydrodynamic chromatography, Chem. Eng. Commun., 5(5), 367-380 (1980). Various, Advances in ion-exchange water treatment (topic issue), Chem. Ind. (London), 20 September, 712-743 (1980). Various, Chromatographic processes (symposium papers), Sep. Sci. Technol., 15(3), 587-696; 15(4), 697-798 (1980). 1981 Annino, R.., Chromatographs can run on air, Chemtech, August, 482-487 (1981). Barker, P.E., and Chuah, C.H., A sequential chromatographic process for the separation of glucose/fructose mixtures, Chem. Eng. (Rugby, Engl.), August, 389-393 (1981). Dyer, A.; Enamy, H., and Townsend, R.P., Plotting and interpretation of ion-exchange isotherms in zeolite systems, Sep. Sci. Technol., 16(2), 173-184 (1981). Gomez-Vaillard, R.; Kershenbaum, L.S., and Streat, M., Performance of continuous, cyclic ion-exchange reactors, Chem. Eng. Sci., 36(2), 307-326 (1981). Goto, S.; Goto, M., and Teshima, H., Simplified evaluations of mass-transfer resistances from batch-wise adsorption and ion-exchange data, Ind. Eng. Chem. Fund., 20(4), 368-375 (1981). Huang, J.C.; Forsythe, R., and Madey, R., Gas-solid chromatography of methane-helium mixtures: Transmission of step increase in concentration of methane through activated carbon adsorber bed at 25~ Sep. Sci. Technol., 16(5), 475-486 (1981). Matsuda, H.; Yamamoto, T.; Goto, S., and Teshima, H., Periodic operation for desalination with thermally regenerable ion-exchange resins (dynamic studies), Sep. Sci. Technol., 16(1), 31-42 (1981 ). Moharir, A.S.; Saraf, D.N., and Kunzru, D., Effect of crystal size distribution on chromatographic peaks in molecular sieve columns, Chem. Eng. Commun., 11(6), 377-386 (1981). Phillips, J.B.; Wright, N.A., and Burke, M.F., Probabilistic approach to digital simulation of chromatographic processes, Sep. Sci. Technol., 16(8), 861-884 (1981). Rahman, K., and Streat, M., Mass transfer in liquid fluidized beds of ion exchange particles, Chem. Eng. Sci., 36(2), 293-306 (1981). Raman, M.S., Polymer resins for water treatment, Chemtech, April, 252-255 (1981). Rice, R.G., and Foo, S.C., Continuous desalination using cyclic mass-transfer ion exchange with bifunctional resins, Ind. Eng. Chem. Fund., 20(2), 150-155 (1981). Said, A.S., Theory of nonlinear chromatography, Sep. Sci. Technol., 16(2), 113-134 (1981). Various, Pharmaceutical applications of ion exchange and solvent extraction (topic issue), Chem. Ind. (London), 3 October, 677-690 ( 1981). Various, Advances in chromatography (topic issue), Chem. Ind. (London), 17 October, 710-732 (1981). 1982 Clifford, D., Multicomponent ion-exchange calculations for selected ion separations, Ind. Eng. Chem. Fund., 21(2), 141-153 (1982). Graham, E.E., and Dranoff, J.S., Application of Stefan-Maxwell equations to diffusion in ion exchangers, Ind. Eng. Chem. Fund., 21(4), 360-369 (1982). Graham, E.E., and Fook, C.F., Rate of protein absorption and desorption on cellulosic ion exchangers, AIChE J., 28(2), 245-250 (1982). Huang, J.C., et al., Gas-solid chromatography of methane-helium mixtures: Moment analysis of breakthrough curves, Sep. Sci. Technol., 17(12), 1417-1424 (1982).

1021

Husain, S.W., et al., Synthesis and ion-exchange properties of lanthanum tungstate, Sep. Sci. Technol., 17(7), 935-944 (1982). Koff, F.W.; Sifniades, S., and Tunick, A.A., Ion-exchange process for recovery of hydroxylamine from Raschig synthesis mixtures, Ind. Eng. Chem. Process Des. Dev., 21 (2), 204-216 (1982). Kojima, T., et al., Fundamental study on recovery of copper with a cation-exchange membrane, Can. J. Chem. Eng., 60, 642-658 (1982). Novosad, J., and Myers, A.L., Thermodynamics of ion exchange as an adsorption process, Can. J. Chem. Eng., 60, 500-503 (1982). Pelosi, P., and McCarthy, J., Preventing fouling of ion-exchange resins, Chem. Eng. (N.Y.), 9 August, 75-78; 6 September, 125-128 (1982). Rao, M.G., and Gupta, A.K., Ion exchange processes accompanied by ionic reactions, Chem. Eng. J., 24(2), 181190(1982). Rousar, I., and Ditl, P., Kinetic characteristics of batch adsorber or ion exchange device operated under nonisothermal conditions, Chem. Eng. Commun., 18(5), 341-354 (1982). Schenk, H.J., et al., Development of sorbers for recovery of uranium from seawater, Sep. Sci. Technol., 17(11), 1293-1308 (1982). Slater, M.J., The relative sizes of fixed bed and continuous countercurrent flow ion exchange equipment, Trans. IChemE, 60, T54-T58 (1982). Various, Ion exchange in the petrochemical industry (topic issue), Chem. Ind. (London), 21 August, 561-573 (1982). 1983 Abe, M., and Hayashi, K., Synthetic inorganic ion-exchange materials, Solvent Extr. Ion Exch., 1(1), 97-112 (1983). Altshuller, D., Design equations and transient behaviour of the countercurrent moving-bed chromatographic reactor, Chem. Eng. Commun., 19(4), 363-376 (1983). Barba, D.; del Re, G., and Foscolo, P.U., Numerical simulation of multicomponent ion-exchange operations, Chem. Eng. J., 26(1), 33-40 (1983). Barker, P.E.; England, K., and Vlachogiannis, G., Mathematical model for the fractionation of dextran on a semicontinuous countercurrent simulated moving bed chromatograph, Chem. Eng. Res. Des., 61, 241-247 (1983). Begovich, J.M.; Byers, C.H., and Sisson, W.G., A high-capacity pressurized continuous chromatograph, Sep. Sci. Technol., 18(12), 1167-1192 (1983). Bobman, M.H.; Golden, T.C., and Jenkins, R.G., Ion exchange in selected low-rank coals: Equilibrium and kinetics, Solvent Extr. Ion Exch., 1(4), 791-826 (1983). Choppin, G.R., and Ohene-Aniapam, F., Equilibrium sorption of Am(IIl), Ce(III), and Eu(III), on Biorex-70 ionexchange resin, Solvent Extr. Ion Exch., 1(3), 585-596 (1983). Fujine, S.; Saito, K.; Shiba, K., and Itoi, T., Liquid mixing in a large-sized column for ion exchange, Solvent Extr. Ion Exch., 1(1), 113-126 (1983). Goto, M.; Hayashi, N., and Goto, S., Separation of electrolyte and nonelectrolyte by ion retardation resin, Sep. Sci. Technol., 18(5), 475-484 (1983). Shih, C.K., et al., Large-scale liquid chromatography separation system, Chem. Eng. Prog., 79(10), 53-57 (1983). Turner, J.C.R., and Murphy, T.K., A CSTR method for determining ion-exchange equilibria, Chem. Eng. Sci., 38(1), 147-154 (1983). Various, Uses of ion exchange in the food industry (topic issue), Chem. Ind. (London), 7 November, 804-824 (1983). 1984 Bailly, M., and Tondeur, D., Reversibility and performances in productive chromatography, Chem. Eng. Process., 18(6), 293-302 (1984). Bonmati R., et al., Industrial gas chromatography process applied to essential oils, Sep. Sci. Technol., 19(2), 113-156(1984). Costa, E.; Lucas, A., and Gonzalez, M.E., Ion exchange: Determination of interdiffusion coefficients, Ind. Eng. Chem. Fund., 23(4), 400-405 (1984). Jenkins, I.L., Ion exchange in the atomic energy industry with particular reference to actinide and fission product separation (review paper), Solvent Extr. Ion Exch., 2(1), 1-28 (1984). Jepson, B.E., and Shockey, G.C., Calcium hydroxide isotope effect in calcium isotope enrichment by ion exchange, Sep. Sci. Technol., 19(2), 173-182 (1984).

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Larin, A.V., Criterion for the quantitative assessment of ideal conditions in chromatography, Adsorpt. Sci. Technol., 6(4), 212-218 (1989). Lee, W.C.; Huang, S.H., and Tsao, G.T., Design equations of chromatography by perturbation method, Chem. Eng. J., 40(3), 165-174 (1989). Lin, B., and Guiochon, G., Numerical simulation of chromatographic band profiles at large concentrations: Length of space increment and height equivalent to a theoretical plate, Sep. Sci. Technol., 24(1), 31-40 (1989). Lin, B.; Ma, Z., and Guiochon, G., Influence of axial dispersion on the bald profile in nonlinear chromatography using the Lax-Wendroff method, Sep. Sci. Technol., 24(11), 809-830 (1989). Liu, X.; Liu, J.C., and Cheng, J.K., New inorganic ion exchangers containing phosphorus, Sep. Sci. Technol., 24(1), 63-78 (1989). McCoy, B.J., Adsorption chromatography of a heterogeneous mixture, Chem. Eng. Sci., 44(4), 993-996 (1989). Mustafa, S., et al., Temperature effect on ion exchange sorption of phosphate, Solvent Extr. Ion Exch., 7(4), 705720 (1989). Oh, M.; Smith, J.M., and McCoy, B.J., Diffusion and adsorption in arrested-flow chromatography, AIChE J., 35(7), 1224-1226 (1989). Ostman, C.E., and Colmsjo, A.L., Separation of polycyclic aromatic compounds from complex oil samples using bonded phase backflush HPLC and GC-MS techniques, Fuel, 68(10), 1248-1250 (1989). Perrut, M., and Jusforgues, P., A new fractionation process: Preparative chromatography with a supercritical eluent, Int. Chem. Eng., 29(4), 646-653 (1989). Sagara, F., et al., Preparation and adsorption properties of manganese oxide-cellulose hybrid-type ion-exchanger for lithium ion: Application to enrichment of lithium ion from seawater, Sep. Sci. Technol., 24(14), 12271244(1989). Saunders, M.S.; Vierow, J.B., and Carta, G., Uptake of phenylalanine and tyrosine by a strong-acid cation exchanger, AIChE J., 35(1), 53-68 (1989). Schaeffer, S.T.; Zalkow, L.H., and Teja, A.S., Supercritical fluid isolation of monocrotaline from Crotalaria spectabilis using ion-exchange resins, Ind. Eng. Chem. Res., 28(7), 1017-1020 (1989). Sheth, A.C.; Prasad, J., and Butler, W.A., Desulfurization of alkali metal sulfates using anion-exchange resins, AIChE J., 35(3), 519-523 (1989). Strelow, F.W.E., Distribution coefficients and ion exchange behavior of some chloride complex forming elements with bio-rad AG5OW-X8 cation exchange resin in mixed nitric-hydrochloric acid solutions, Solvent Extr. Ion Exch., 7(4), 735-747 (1989). Venkateswarlu, K.S., et al., Use of strong-base organic anion exchangers for removal of suspended alumina particles in light water-heavy water systems, Sep. Sci. Technol., 24(5), 467-474 (1989). Yoshida, H., and Kataoka, T., Recovery of mercuric chloride using chloride-form ion exchanger, AIChE J., 35(2), 318-320 (1989). Yoshida, H., and Ruthven, D.M., Adsorption of gaseous ethylamine on H-form strong-acid ion exchangers, AIChE J., 35(11), 1869-1875 (1989). Yu, Q., and Wang, N.H.L., Computer simulations of dynamics of multicomponent ion exchange and adsorption in fixed beds: Gradient-directed moving finite element method, Comput. Chem. Eng., 13(8), 915-926 (1989). 1990 Allen, R.M., and Addison, P.A., Ion exchange equilibria for ternary systems from binary exchange data, Chem. Eng. J., 44(3), 113-118 (1990). Anon., Solid electrolyte separation offers pure gases, Chem. Eng. (Rugby, Engl.), March, 30 (1990). Bain, P.E., A model predicting equilibrium for plutonium sorption by anion exchange resin, Solvent Extr. Ion Exch., 8(2), 341-352 (1990). Bi, Y., and Wen-Bin, H., Two-barrel bile-acids-sensitive microelectrodes based on liquid ion exchanger, Biotechnol. Prog., 6(1), 62-66 (1990). Bolden, W.B., and Groves, F.R., Amine recovery by ligand exchange: Pore diffusion model, Ind. Eng. Chem. Res., 29(1), 116-121 (1990). Breeman, D.J., Backflush column removes GC sample water, Hydrocarbon Process., 69(3), 90-91 (1990). Byers, C.H., et al., Sugar separations on a pilot scale by continuous annular chromatography, Biotechnol. Prog., 6(1), 13-20 (1990). Carta, G., and Bauer, J.S., Analytic solution for chromatography with nonuniform sorbent particles, AIChE J., 36(1), 147-150 (1990).

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Mardan, A., and Liljenzin, J.O., Sulfonation and performance of surface-sulfonated (5 and 10 mm diameter) highly crosslinked porous and (15 micron diameter) 10% crosslinked gel-type resins, Solvent Extr. Ion Exch., 8(1), 137-150 (1990). Mardan, A.; Alstad, J., and Liljenzin, J.O., Elution behavior comparison of dyno-resins for rare-earths/2-hydroxy isobutyric acid system, Solvent Extr. Ion Exch., 8(1), 151-172 (1990). Mustafa, S.; Hussain, S.Y., and Ahmad, R., Phosphate/hydroxide exchange studies on Amberlite IRA-400, Solvent Extr. Ion Exch., 8(2), 325-340 (1990). Ray, A., et al., The simulated countercurrent moving bed chromatographic reactor, Chem. Eng. Sci., 45(8), 24312438(1990). Rice, R.G., and Heft, B.K., Radial flow chromatography in compressed pancake-shaped beds, Chem. Eng. Commun., 98, 231-240 (1990). Rovere, C.E., et al., Chemical class separation of shale oils by low-pressure liquid chromatography on thermallymodified adsorbants, Fuel, 69(9), 1099-1104 (1990). Sun, Y.D.; Grevillot, G., and Tondeur, D., Modelling and optimization of the cyclic regime of an ion-exchange process for sugar juice softening, Biochem. Eng. J., 43(2), B53-B66 (1990). Takahashi, Y., and Goto, S., Adsorption isotherms of amino acids and kinetic analysis of ion-exchange chromatographs by the moment method, Sep. Sci. Technol., 25(11), 1131-1140 (1990). Trobajo, C., et al., Lamellar inorganic ion exchangers: Li+, Na +, H+ ion exchange in gamma-titanium phosphate, Solvent Extr. Ion Exch., 8(4), 729-740 (1990). Usuda, S., et al., Desorption behavior of plutonium from anion-exchange resin with HNO3-HI mixed acid solution, Sep. Sci. Technol., 25(11), 1225-1238 (1990). Vasheghani-Farahani, E., et al., Swelling of ionic gels in electrolyte solutions, Ind. Eng. Chem. Res., 29(4), 554560 (1990). Yoshida, H., and Kataoka, T., Recovery of mercury from a mercury(II)-form chelate resin by electrolytic desorption, Ind. Eng. Chem. Res., 29(10), 2152-2154 (1990). Zuyi, T., and Jinlong, N., Shell-progressive model with changing bulk concentration and exchanger volume in ion exchange, Solvent Extr. Ion Exch., 8(1), 99-116 (1990). 1991 Ackley, M.W., and Yang, R.T., Diffusion in ion-exchanged clinoptilolites, AIChE J., 37(11), 1645-1656 (1991). Adams, R.J.W., and Hudson, M.J., Reversible extraction of ionic species using electrochemically assisted ion exchange: Cobalt(II) using alpha-zirconium hydrogen phosphate, Solvent Extr. Ion Exch., 9(3), 497-514 (1991). Akintoye, A.; Ganetsos, G., and Barker, P.E., The inversion of sucrose on a semicontinuous countercurrent chromatographic bioreactor-separator, Food Bioprod. Process., 69(C 1), 35-44 (1991 ). Alen, R.; Sjostrom, E., and Suominen, S., Application of ion-exclusion chromatography to alkaline pulping liquors: Separation of hydroxy carboxylic acids from inorganic solids, J. Chem. Technol. Biotechnol., 51(2), 225-234 (1991). Baksh, M.S.A., and Yang, R.T., Chromatographic separations by pillared clay, Sep. Sci. Technol., 26(10), 13771394(1991). Barker, P.E., and Bridges, S., Continuous annular chromatography for the separation of beet molasses, J. Chem. Technol. Biotechnol., 51(3), 347-360 (1991). Barker, P.E., and Joshi, K., The recovery of fructose from inverted sugar beet molasses using continuous chromatography, J. Chem. Technol. Biotechnol., 52(1), 93-108 (1991 ). Bloomingburg, G.F., et al., Continuous separation of proteins by annular chromatography, Ind. Eng. Chem. Res., 30(5), 1061-1067 (1991). Bridger, N.J.; Jones, C.P., and Neville, M.D., Electrochemical ion exchange, J. Chem. Technol. Biotechnol., 50(4), 469-482 (1991). Bruening, M.L., et al., A novel, highly selective anion-exchange column prepared by binding Pd 2+ to an immobilized ligand, Sep. Sci. Yechnol., 26(6), 761-772 (1991). Campbell, D., and Foundos, A., Chromatographs meet environmental needs, Hydrocarbon Process., 70(2), 63-65 (1991). Cerny, J.; Sebor, G., and Mitera, J., Comparison of the selectivity of extrographic and chromatographic fractionations, Fuel, 70(7), 857-860 (1991). Das, N.R., and Lahiri, S., Liquid ion exchangers and their uses in the separation of zirconium, niobium, molybdenum, hafnium, tantalum and tungsten, Solvent Extr. Ion Exch., 9(2), 337-350 (1991).

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Ganetsos, G., and Barker, P.E., Large-scale chromatography in industrial processing, J. Chem. Technol. Biotechnol., 50(1), 101-108 (1991). Goetz, V., and Graves, D.J., Axial dispersion in a magnetically stabilized fluidized-bed liquid chromatography column, Powder Technol., 64(1), 81-92 (1991). Gu, T.; Tsai, G.J., and Tsao, G.T., Multicomponent adsorption and chromatography with uneven saturation capacities, AIChE J., 37(9), 1333-1340 (1991). Gu, T.; Tsai, G.J., and Tsao, G.T., A theoretical study of multicomponent radial flow chromatography, Chem. Eng. Sci., 46(5), 1279-1288 (1991). Harkins, D.A., and Schweitzer, G.K., Preparation of site-selective ion-exchange resins, Sep. Sci. Technol., 26(3), 345-354 (1991). Harries, R.R., Ion exchange kinetics in ultra-pure water systems, J. Chem. Technol. Biotechnol., 51(4), 437-448 (1991). Helfferich, F.G., The h- and w-transformations in multicomponent fixed-bed ion exchange and adsorption: Equivalent mathematics, different scope, Chem. Eng. Sci., 46(12), 3320-3323 (1991). Hsu, T.B., and Pigford, R.L., Mass transfer in a thermally regenerable ion-exchange resin by continuous cycling, Ind. Eng. Chem. Res., 30(5), 1067-1075 (1991). Huang, H., et al., The sorption behavior of boric acid on weak-base anion exchange resin, Solvent Extr. Ion Exch., 9(2), 319-336 (1991 ). Hufton, J.R., and Danner, R.P., Gas-solid diffusion and equilibrium parameters by tracer pulse chromatography, Chem. Eng. Sci., 46(8), 2079-2092 (1991). Jacobson, S.; Golshan-Shirazi, S., and Guiochon, G., Isotherm selection for band profile simulations in preparative chromatography, AIChE J., 37(6), 836-844 (1991). Jaroniec, M.; Madey, R., and Lu, X., Application of gas-solid adsorption chromatography for characterizing adsorbent heterogeneity, Sep. Sci. Technol., 26(2), 269-278 (1991). Jeng, C.Y., and Langer, S.H., Rate process analysis in the liquid chromatographic reactor: An application of the first statistical moment, Ind. Eng. Chem. Res., 30(7), 1489-1499 (1991). Kawakita, T., and Matsuishi, T., Elution kinetics of lysine from a strong cation-exchange resin with ammonia water, Sep. Sci. Technol., 26(7), 991-1004 (1991). Kawakita, T., et al., Breakthrough curve of lysine on a column of a strong cation-exchange resin of the ammonium form, Sep. Sci. Technol., 26(5), 619-636 (1991). Kawakita, T.; Matsuishi, T., and Koga, Y., Optimization of lysine adsorption process using strong cationexchange resin, Sep. Sci. Technol., 26(6), 869-884 (1991). Landau, I.; Belfer, A.J., and Locke, D.C., Measurement of limiting activity coefficients using non-steady-state gas chromatography, Ind. Eng. Chem. Res., 30(8), 1900-1906 (1991). Lobarzewski, J.; Kowalska-Pylka, H., and Cybulski, W., A simple affinity chromatography method for the separation of gastric proteases from mucous substances, J. Chem. Technol. Biotechnol., 52(3), 359-368 (1991). Lukac, M., and Perina, Z., A dynamic model of physical processes in chromatographic glucose-fructose separation, Chem. Eng. Sci., 46(4), 959-966 (1991). Ma, Z., and Guiochon, G., Application of orthogonal collocation on finite elements in the simulation of nonlinear chromatography, Comput. Chem. Eng., 15(6), 415-426 (1991). Mannhardt, K., and Novosad, J.J., Chromatographic movement of surfactant mixtures in porous media, Chem. Eng. Sci., 46(1), 75-84 (1991). Maranon, E., and Sastre, H., Ion exchange equilibria of heavy metals onto chemically modified apple residues, Solvent Extr. Ion Exch., 9(3), 515-532 (1991). Masoliver, J., and Weiss, G.H., Transport equations in chromatography with a finite speed of signal propagation, Sep. Sci. Technol., 26(2), 279-290 (1991). Moon, J.K., and Lee, W.K., Adsorption characteristics of cresols with eluent composition in adsorption chromatography, Sep. Sci. Technol., 26(5), 675-688 (1991). Oi, T., et al., Fractionation of lithium isotopes in cation-exchange chromatography, Sep. Sci. Technol., 26(10), 1353-1376(1991). Orford, C.D.; Adlard, M.W., and Perry, D., Isolation of gamma-(L-alpha-aminoadipyl)-L-cysteinyl-D-valine from culture broths by covalent chromatography, J. Chem. Technol. Biotechnol., 50(4), 523-534 (1991). Rice, R.G., and Heft, B.K., Separations via radial flow chromatography in compacted particle beds, AIChE J., 37(4), 629-632 (1991).

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Row, K.H., and Raw, J.I., Parameter estimation of cyclobutane pyrimidine dimers and monomers of uracil and thymine in reversed-phase high-performance liquid chromatography, Sep. Sci. Technol., 26(1), 15-24 (1991). Surakitbanharn, Y.; Muralidharan, S., and Freiser, H., Separation of palladium(II) from platinum(II), iridium(Ill), and rhodium(IIl) using centrifugal partition chromatography, Solvent Extr. Ion Exch., 9(1), 45-60 (1991 ). Suwondo, E., et al., Simulation via orthogonal collocation on finite element of a chromatographic column with nonlinear isotherm, Chem. Eng. Commun., 102, 161-188 (1991 ). Takahashi, Y., and Goto, S., Continuous separations of amino acids by using an annular chromatograph with rotating inlet and outlet, Sep. Sci. Technol., 26(1), 1-14 (1991). Takase, H., and Yoshimura, Y., Mass transfer from a slurry adsorbent to an ion-exchange resin, Int. Chem. Eng., 31(2), 351-358 (1991). Ysuji, M., and Komarneni, S., An evaluation method of chromatographic parameters from the ion-exchange isotherm of A13+-substituted tobermorite cation exchanger, Sep. Sci. Technol., 26(5), 647-660 (1991). Whitley, R.D., et al., Effects of protein aggregation in isocratic nonlinear chromatography, AIChE J., 37(4), 555568(1991). Wong, J.W.; Albright, R.L., and Wang, N.H.L., Immobilized metal ion affinity chromatography (IMAC): Chemistry and bioseparation applications, Sep. Purif. Methods, 20(1), 49-106 (1991). Worthy, W., New perfusion-chromatography separation method, Chem. Eng. News, 18 November, 25-26 (1991). Yang, B.L., and Goto, S., Complete separation of albumin and hemoglobin by metal chelate affinity chromatography, Sep. Sci. Technol., 26(5), 637-646 (1991 ). Yasuda, S., and Kawazu, K., Separation of germanium from ethylene glycol distillates by N-methylglucamine resin, Sep. Sci. Technol., 26(9), 1273-1278 (1991). Yu, Q., and Do, D.D., Reversed displacement chromatography of adsorptions with unfavourable equilibrium isotherms, Biochem. Eng. J., 46(3), B93-B98 (1991). Zecchini, E.J., and Foutch, G.L., Mixed-bed ion-exchange modeling with amine-form cation resins, Ind. Eng. Chem. Res., 30(8), 1886-1892 (1991). 1992 Alexandratos, S.D., and Kaiser, P.T., Reaction kinetics of polystyrene-based phosphinic acid ion exchange/redox resins with metal ions, Solvent Extr. Ion Exch., 10(3), 539-550 (1992). Anon., Advances in ion exchange, Chem. Eng. (N.Y.), September, 63-71 (1992). Anon., Ion exchange for esterification, Chem. Eng. (Rugby, Engl.), 10 December, 14-15 (1992). Barker, P.E., et al., Bioreaction-separation on continuous chromatographic systems, Biochem. Eng. J., 50(2), BZ3-BZ8 (1992). Bauza, R., et al., Separation of mono-, di-, and tri-stearin from an industrial mixture of glycerides by normal- and reverse-phase HPLC, Sep. Sci. Technol., 27(5), 645-662 (1992). Bhandari, V.M.; Juvekar, V.A., and Patwardhan, S.R., Sorption studies on ion exchange resins: Sorption of strong acids on weak base resins, Ind. Eng. Chem. Res., 31 (4), 1060-1073 (1992). Bhandari, V.M.; Juvekar, V.A., and Patwardhan, S.R., Sorption studies on ion exchange resins: Sorption of weak acids on weak base resins, Ind. Eng. Chem. Res., 31(4), 1073-1080 (1992). Bhandari, V.M.; Juvekar, V.A., and Patwardhan, S.R., Modified shrinking core model for reversible sorption on ion-exchange resins, Sep. Sci. Yechnol., 27(8), 1043-1064 (1992). Bhattacharyya, D.K., and Dutta, N.C., Role of hydrous titanium oxide on the uptake of several tracer cations, and separation of carrier-free 125mTefrom I25Sb and 13Zlfrom 132Te,Sep. Sci. Technol., 27(3), 399-408 (1992). Binous, H., and McCoy, B.J., Chromatographic reactions of three components: Application to separations, Chem. Eng. Sci., 47(17), 4333-4344 (1992). Bricio, O.; Coca, J., and Sastre, H., A comparative study of kinetic models for ion-exchange using macroporous resins and concentrated solutions, Solvent Extr. Ion Exch., 10(2), 381-400 (1992). Bridges, S., and Barker, P.E., Modelling continuous chromatographic separations, Chem. Eng. Sci., 47(5), 12991306 (1992). Brooks, C.A., and Cramer, S.M., Steric mass-action ion exchange: Displacement profiles and induced salt gradients, AIChE J., 38(12), 1969-1978 (1992). Calvarin, L.; Roche, B., and Renon, H., Anion exchange and aggregation of dicyanocobalamin with quaternary ammonium salts in apolar environment, Ind. Eng. Chem. Res., 31(7), 1705-1709 (1992). Carta, G., et al., Chromatography of reversibly reacting mixtures: Mutarotation effects in sugar separations, Chem. Eng. Sci., 47(7), 1645-1658 (1992).

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Cavender, M.R.; Chiang, H.L., and Myers, K., Optimize ion exchange resins replacement, Chem. Eng. Prog., 88(9), 56-59 (1992). Chase, H.A., and Draeger, N.M., Expanded-bed adsorption of proteins using ion-exchangers, Sep. Sci. Technol., 27(14), 2021-2040 (1992). Chiarizia, R.; Horwitz, E.P., and Dietz, M.L., Acid dependency of the extraction of selected metal ions by a strontium-selective extraction chromatographic resin: Calculated vs. experimental curves, Solvent Extr. Ion Exch., 10(2), 337-362 (1992). de Bokx, P.K.; Baarslag, P.C., and Urbach, H.P., Calculation and experimental verification of solute retention in liquid chromatography using binary eluents, Sep. Sci. Technol., 27(7), 875-900 (1992). De Lucas, A.; Zarca, J., and Canizares, P., Ion-exchange equilibrium of Ca2+, Mg 2+, K +, Na +, and H+ ions on Amberlite IR-120: Experimental determination and theoretical prediction of the ternary and quaternary equilibrium data, Sep. Sci. Technol., 27(6), 823-842 (1992). Durao, M.I.G.; Costa, C.A.V., and Rodrigues, A.E., Saturation and regeneration of ion exchangers with volume changes, Ind. Eng. Chem. Res., 31 (11), 2564-2572 (1992). Eccles, H., and Greenwood, H., Chelate ion-exchangers: The past and future applications, a user's view, Solvent Extr. Ion Exch., 10(4), 713-728 (1992). Economopoulos, N., et al., A plant kinetic study of alcoholic fermentation using reversed-flow gas chromatography, Sep. Sci. Technol., 27(15), 2055-2070 (1992). EI-Naggar, I.M., and Aly, H.F., Kinetics of Cs +, Sc3+, and Eu 3+ exchange on crystalline atimonic acid, Solvent Extr. Ion Exch., 10(1), 145-158 (1992). Gu, T., et al., Modeling of gradient elution in multicomponent non-linear chromatography, Chem. Eng. Sci., 47(1), 253-262 (1992). Heininger, M.W., and Meloan, C.E., A resin with selectivity for the removal and recovery of chromate from contaminated water, Solvent Extr. Ion Exch., 10(1), 159-172 (1992). Horwitz, E.P.; Chiarizia, R., and Dietz, M.L., A novel strontium-selective extraction chromatographic resin, Solvent Extr. Ion Exch., 10(2), 313-336 (1992). Hossain, M.M., and Do, D.D., The effects of denaturation in the displacement chromatographic behaviour of proteins, Biochem. Eng. J., 49(3), B29-B39 (1992). Huang, S.Y., and Jin, J.D., Operation strategy for displacement chromatography: Selection of optimum mobile phase for separation of weak adsorptive nucleotides, Chem. Eng. Sci., 47(1), 21-30 (1992). Kaur, P., et al., Studies on the sorption behaviour of some amino acids on silica gel pretreated with alkalis in relation to chromatography, Adsorpt. Sci. Technol., 8(3), 157-173 (1992). Kim, S.U., et al., Peak compression in stepwise pH elution with flow reversal in ion exchange chromatography, Ind. Eng. Chem. Res., 31(7), 1717-1730 (1992). Larson, K.A., and Wiencek, J.M., Liquid ion exchange for mercury removal from water over a wide pH range, Ind. Eng. Chem. Res., 31(12), 2714-2722 (1992). Lee, K.N., and Lee, W.K., A theoretical model for the separation of glucose and fructose mixtures by using a semicontinuous chromatographic refiner, Sep. Sci. Technol., 27(3), 295-312 (1992). Leung, B.K.O., and Hudson, M.J., A novel weak-base anion-exchange resin which is highly selective for the precious metals over base metals, Solvent Extr. Ion Exch., 10( 1), 173-190 (1992). Levy, D., et al., Immobilization of quaternary ammonium anion exchangers in sol-gel glasses, Sep. Sci. Technol., 27(5), 589-598 (1992). Lewandowski, R., and Lameloise, M.L., Study of exclusion equilibrium between a sucrose-NaCl solution and an ion exchange resin, Chem. Eng. Process., 31(4), 207-212 (1992). Mijangos, F., and Diaz, M., Metal-proton equilibrium relations in a chelating iminodiacetic resin, Ind. Eng. Chem. Res., 31(11), 2524-2532 (1992). Miyabe, K., and Suzuki, M., Chromatography of liquid-phase adsorption on octadecylsilyl-silica gel, AIChE J., 38(6), 901-910 (1992). Mohammad, A.W.; Stevenson, D.G., and Wankat, P.C., Pressure drop correlations and scale-up of size exclusion chromatography with compressible packings, Ind. Eng. Chem. Res., 31(2), 549-561 (1992). Oi, T., et al., Fractionation of strontium isotopes in cation-exchange chromatography, Sep. Sci. Technol., 27(5), 631-644 (1992). Olson, K.C., and Gehant, R.L., Applications of ultrafast HPLC to process development of recombinant DNAderived proteins, Biotechnol. Prog., 8(6), 562-566 (1992).

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Zhong, G., and Guiochon, G., Analytical solution for the linear ideal model of simulated moving bed chromatography, Chem. Eng. Sci., 51 (18), 4307-4320 (1996). Zurer, P.S., Chromatography and mass spectrometry, Chem. Eng. News, 18 March, 38-46 (1996). 1997 Achuthan, P.V.; Janardanan, C., and Ramanujam, A., Water sorption isotherms and ionic hydration of uranium and thorium forms ofDowex 50W resins, Solvent Extr. Ion Exch., 15(4), 631-646 (1997). Anklam, M.R.; Prudhomme, R.K., and Finlayson, B.A., Ion exchange chromatography laboratory: Experimentation and numerical modeling, Chem. Eng. Educ., 31 (1), 26-31 (1997). Bartos, B.; Bilewicz, A., and Delmas, R., Synthesis and ion exchange properties of various forms of manganese dioxide for cations of the I and II groups, Solvent Extr. Ion Exch., 15(3), 533-546 (1997). Bhandari, V.M.; Juvekar, V.A., and Patwardhan, S.R., Ion-exchange studies in the removal of polybasic acids: Anomalous sorption behavior of phosphoric acid on weak base resins, Sep. Sci. Technol., 32(15), 2481-2496 (1997). Bortun, A.I.; Bortun, L.N., and Clearfield, A., A novel layered zirconium phosphate Zr203(HPO4): Synthesis and characterization of properties, Solvent Extr. Ion Exch., 15(2), 305-328 (1997). Bortun, A.I.; Bortun, L.N., and Clearfield, A., Synthesis and characterization of ion exchange properties of spherically granulated titanium phosphate, Solvent Extr. Ion Exch., 15(3), 515-532 (1997). Bortun, A.I.; Bortun, L.N., and Clearfield, A., Evaluation of synthetic inorganic ion exchangers for cesium and strontium removal from contaminated groundwater and wastewater, Solvent Extr. Ion Exch., 15(5), 909-929 (1997). Bortun, A.I.; Bortun, L.N., and Khainakov, S.A., Modified titanium phosphates as cesium selective ion exchangers, Solvent Extr. Ion Exch., 15(5), 895-907 (1997). Bricio, O.; Coca, J., and Sastre, H., Effect of the heterogeneity of macroporous styrene-DVB resins on ionexchange equilibria, Solvent Extr. Ion Exch., 15(4), 647-664 (1997). Chiarizia, R.; Horwitz, E.P., and Alexandratos, S.D., Diphonix(R) resin: A review of its properties and applications, Sep. Sci. Technol., 32(1), 1-35 (1997). Clearfield, A., et al., Synthesis and characterization of a novel layered sodium titanium silicate Na2TiSi2OT.2H_~O, Solvent Extr. Ion Exch., 15(2), 285-304 (1997). Cortina, J.L., and Miralles, N., Kinetic studies on heavy metal ions removal by impregnated resins containing di(2,4,4-trimethylpentyl) phosphinic acid, Solvent Extr. Ion Exch., 15(6), 1067-1083 (1997). Cumming, I.W.; Tai, H., and Beier, M., A model to predict the performance of an electrochemical ion exchange cell, Chem. Eng. Res. Des., 75(1 ) 9-13 (1997). Dave, S.M.; Patil, S.S., and Suresh, A.K., Ion exchange for product recovery in lactic acid fermentation, Sep. Sci. Yechnol., 32(7), 1273-1294 (1997). Defilippi, I.; Yates, S., and Sedath, R., Scale-up and testing of a novel ion exchanger for strontium, Sep. Sci. Technol., 32(1 ), 93-113 (1997). Delucas, A.; Canizares, P., and Rodriguez, J.F., Ion-exchange kinetics for the removal of potassium from crude polyols on strong acid resins, Sep. Sci. Technol., 32(11), 1805-1820 (1997). Dickson, M.L.; Norton, T.T., and Fernandez, E.J., Chemical imaging of multicomponent viscous fingering in chromatography, AIChE J., 43(2), 409-418 (1997). Ernest, M.V.; Bibler, J.P.; Whitley, R.D., and Wang, N.H.L., Development of a carousel ion-exchange process for removal of cesium-137 from alkaline nuclear waste, Ind. Eng. Chem. Res., 36(7), 2775-2788 (1997). Ernest, M.V., et al., Effects of mass action equilibria on fixed-bed multicomponent ion-exchange dynamics, Ind. Eng. Chem. Res., 36(1), 212-226 (1997). Farkas, T.; Sepaniak, M.J., and Guiochon, G., Radial distribution of the flow velocity, efficiency and concentration in a wide HPLC column, AIChE J., 43(8), 1964-1974 (1997). Farnan, D.; Frey, D.D., and Horvath, C., Intraparticle mass transfer in high-speed chromatography of proteins, Biotechnol. Prog., 13(4), 429-439 (1997). Frey, D.D., Mechanism for glutamic acid adsorption on a weak-base ion exchanger, Chem. Eng. Sci., 52(7), 1227-1231 (1997). Greenstein, E.M., Filters for resins, Chem. Eng. (N.Y.), June, 155 (1997). Gu, D.; Nguyen, L., and Philip, C.V., Cs+-ion exchange kinetics in complex electrolyte solutions using hydrous crystalline silicotitanates, Ind. Eng. Chem. Res., 36(12), 5377-5383 (1997). Habbaba, M.M., and Ulgen, K.O., Analysis of protein adsorption to ion exchangers in a finite bath, J. Chem. Technol. Biotechnol., 69(4), 405-414 (1997).

1043

Hasnat, A., and Juvekar, V.A., Dynamics of ion-exchange involving multivalent cations, Chem. Eng. Sci., 52(14), 2439-2442 (1997). Hasnat, A., and Juvekar, V.A., Ion exchange in weak acid resin: Diffusion in shrinking core, AIChE J., 43(10), 2605-2608 (1997). Jansen, M.L., et al., Effect of dissociation equilibria on ion-exchange processes of weak electrolytes, AIChE J., 43(1 ), 73-82 (1997). Juang, R.S., and Chen, M.L., Comparative equilibrium studies on the sorption of metal ions with macroporous resins containing a liquid ion-exchanger, Sep. Sci. Technol., 32(5), 1017-1035 (1997). Juang, R.S., and Ju, C.Y., Equilibrium sorption of copper(II)-ethylenediaminetetraacetic acid chelates onto crosslinked, polyaminated chitosan beads, Ind. Eng. Chem. Res., 36(12), 5403-5409 (1997). Jyo, A.; Yamabe, K., and Egawa, H., Metal ion selectivity of a macroreticular styrene-divinylbenzene copolymerbased methylenephosphonic acid resin, Sep. Sci. Technol., 32(6), 1099-1106 (1997). Kawamura, Y.; Yoshida, H., and Asai, S., Effects of chitosan concentration and precipitation bath concentration on the material properties of porous crosslinked chitosan beads, Sep. Sci. Technol., 32(12), 1959-1974 (1997). Kitakawa, A.; Yamanishi, Y., and Yonemoto, T., Complete separation of amino acids using continuous rotating annular ion exchange chromatography with partial recycle of effluent, Ind. Eng. Chem. Res., 36(9), 38093814(1997). Kuhr, J.H., et al., Ion exchange properties of a Western Kentucky low-rank coal, Energy Fuels, 11(2), 323-326 (1997). Kulikov, N.S., Molecular modelling in chromatostructural analysis: A new approach to the GC/MS study of isomers, Adsorpt. Sci. Yechnol., 15(2), 115-124 (1997). Lee, D.D.; Walker, J.F., and Taylor, P.A., Cesium-removal flow studies using ion-exchange, Environ. Prog., 16(4), 251-262 (1997). Lee, J.G.; Lee, W.C., and Wang, F.S., Simulation of pH elution in high-performance affinity chromatography using non-porous adsorbents, Chem. Eng. J., 65(3), 175-186 (1997). Lilga, M.A,; Orth, R.J., and Sukamto, J.P.H., Metal ion separations using electrically switched ion exchange, Sep. Purif. Yechnol., 11(3), 147-158 (1997). Lucas, A.D., et al., Potassium removal from water-methanol-polyol mixtures by ion exchange on Amberlite 252, Chem. Eng. J., 66(2), 137-148 (1997). Luo, R.G., and Hsu, J.T., Rate parameters and gradient correlations for gradient-elution chromatography, AIChE J., 43(2), 464-474 (1997). Luo, R.G., and Hsu, J.T., Optimization of gradient profiles in ion-exchange chromatography for protein purification, Ind. Eng. Chem. Res., 36(2), 444-450 (1997). Ma, Z., and Wang, N.H.L., Standing wave analysis of SMB chromatography: Linear systems, AIChE J., 43(10), 2488-2508 (1997). Mardan, A., Enrichment of boron-10 by inverse-frontal chromatography using quaternized 4-vinylpyridinedivinylbenzene and anion-exchange resin, Sep. Sci. Technol., 32(13), 2115-2125 (1997). Marquez, N.; Subero, N., and Anton, R.E., Effect of alkylate isomerism upon surfactant retention in an HPLC column and partitioning between water and oil, Sep. Sci. Technol., 32(6), 1087-1098 (1997). Matijasevic, L.; Vasic-Racki, D., and Pavlovic, N., Separation of glucose/fructose mixtures: Analysis of elution of profiles, Chem. Eng. J., 65(3), 209-212 (1997). McNulty, J.Y., The many faces of ion-exchange resins, Chem. Eng. (N.Y.), June, 94-100 (1997). Milan, Z., et al., Ammonia removal from anaerobically treated piggery manure by ion exchange in columns packed with homoionic zeolite, Chem. Eng. J., 66(1), 65-72 (1997). Miyabe, K., and Takeuchi, S., Surface diffusion phenomena in reversed-phase liquid chromatography with methanol/water and acetonitrile/water mixtures, Ind. Eng. Chem. Res., 36(10), 4335-4341 (1997). Miyoshi, H., Diffusion coefficients of ions through ion-exchange membranes for Donnan dialysis using ions of the same valence, Chem. Eng. Sci., 52(7), 1087-1096 (1997). Muralidharan, P.K., and Ching, C.B., Determination of multicomponent adsorption equilibria by liquid chromatography, Ind. Eng. Chem. Res., 36(2), 407-413 (1997). Nakayama, M., and Egawa, H., Recovery of gallium(III) from strongly alkaline media using a Kelex-100-loaded ion-exchange resin, Ind. Eng. Chem. Res., 36(10), 4365-4368 (1997). Nikolaev, N.P.; Muraviev, D.N., and Muhammed, M., Dual-temperature ion-exchange separation of copper and zinc by different techniques, Sep. Sci. Technol., 32(1), 849-866 (1997).

1044

Noriega, J.A.; Tejeda, A., and Magana, I. Modeling column regeneration effects on dye-ligand affinity chromatography, Biotechnol. Prog., 13(3), 296-300 (1997). Oi, T.; Shimazaki, H., and Ishii, R., Boron isotope fractionation in liquid chromatography with boron-specific resins as column packing material, Sep. Sci. Technol., 32(11), 1821-1834 (1997). Pais, L.S.; Loureiro, J.M., and Rodrigues, A.E., Separation of 1,1'-bi-2-naphthol enantiomers by continuous chromatography in simulated moving bed, Chem. Eng. Sci., 52(2), 245-258 (1997). Porter, C.E.; Riley, F.D., and Vandergrift, R.D., Fermium purification using Teva resin extraction chromatography, Sep. Sci. Technol., 32(1), 83-92 (1997). Prazeres, D.M.F., A theoretical analogy between multistage ultrafiltration and size-exclusion chromatography, Chem. Eng. Sci., 52(6), 953-960 (1997). Ramirez-Vick, J.E., and Garcia, A.A., Recent developments in the use of group-specific ligands for affinity bioseparations, Sep. Purif. Methods, 25(2), 85-130 (1997). Rastogi, R.K.; Mahajan, M.A., and Chaudhuri, N.K., Separation of thorium from uranium product at the tail end of thorium fuel reprocessing using macroporous cation-exchange resin, Sep. Sci. Technol., 32(10), 1711-1723 (1997). Rendueles, M.; Fernandez, A., and Diaz, M., Coupling of ion exchange with industrial processes: Application in fertilizer production and modeling of the key elution step, Solvent Extr. Ion Exch., 15(1), 143-168 (1997). Rendueles, M.; Fernandez, A., and Diaz, M., Sorption of counter and co-ions at high concentration in ion exchangers, Solvent Extr. Ion Exch., 15(4), 665-688 (1997). Rincon, J., et al., Selection of a cation exchange resin to produce lactic acid solutions from whey fermentation broths, Solvent Extr. Ion Exch., 15(2), 329-346 (1997). Robichaud, M.J.; Sathyagal, A.N., and Can', P.W., An improved oil emulsion synthesis method for large, porous zirconia particles for packed- or fluidized-bed protein chromatography, Sep. Sci. Technol., 32(15), 25472559 (1997). Roddick, F.A., and Britz, M.L., Production of hexanoic acid by free and immobilised cells of Megasphaera elsdenii: Influence of in-situ product removal using ion exchange resin, J. Chem. Technol. Biotechnol., 69(3), 383-391 (1997). Rogers, R.D.; Griffin, S.T., and Horwitz, E.P., Aqueous biphasic extraction chromatography (ABEC): Uptake of pertechnetate from simulated Hanford tank wastes, Solvent Extr. Ion Exch., 15(4), 547-562 (1997). Shalliker, R.A., et al., Examination of various pore size zirconias for potential chromatographic applications, Powder Yechnol., 91 ( 1), 17-24 (1997). Shelley, S., Ion exchange curb water and chemical use, Chem. Eng. (N.Y.), December, 117-118 (1997). Simon, G., et al., Preparative-scale separation of amino acids by using thermal ion exchange parametric pumping, Chem. Eng. Sci., 52(4), 467-480 (1997). Sujatha, V.; Sarma, C.B., and Raju, G.J.V.J., Studies on ionic mass transfer with coaxially placed helical tapes on a rod in forced convection flow, Chem. Eng. Process., 36(1), 67-74 (1997). Tanaka, Y., and Ysuji, M., Thermodynamic study of alkali metal ions/proton exchanges on an alpha-type manganese dioxide, Solvent Extr. Ion Exch., 15(4), 709-729 (1997). Tsaur, Y., and Shallcross, D.C., Comparison of simulated performance of fixed ion exchange beds in linear and radial flow, Solvent Extr. Ion Exch., 15(4), 689-708 (1997). Tsaur, Y., and Shallcross, D.C., Modeling of ion exchange performance in a fixed radial flow annular bed, Ind. Eng. Chem. Res., 36(6), 2359-2367 (1997). van Buel, M.J.; van der Wielen, L.A.M., and Luyben, C.C.A.M., Effluent concentration profiles in centrifugal partition chromatography, AIChE J., 43(3), 693-702 (1997). Various, New adsorbents and ion exchange materials (topic issue), Adsorption, 3(1), 5-105 (1997). Varotsis, N., and Pasadakis, N., Rapid quantitative determination of aromatic groups in lubricant oils using gel permeation chromatography, Ind. Eng. Chem. Res., 36(12), 5516-5519 (1997). Vlasenko, E.V.; Gavrilova, T.B., and Daidakova, I.V., Intermolecular interactions in gas chromatography on carbon black coated with monolayers of hydrocarbons with different electronic structures, Adsorpt. Sci. Yechnol., 15(2), 79-90 (1997). Warshawsky, A., et al., Solvent-impregnated resins via acid-base interaction of poly(4-vinylpyridine) resin and di(2-ethylhexyl) dithio-phosphoric acid, Solvent Extr. Ion Exch., 15(2), 259-284 (1997). Williams, C.J., and Edyvean, R.G.J., Ion exchange in nickel biosorption by seaweed materials, Biotechnol. Prog., 13(4), 424-428 (1997).

1045

Wolfgang, J.; Prior, A., and Byers, C.H., Continuous separation of carbohydrates by ion-exchange chromatography, Sep. Sci. Technol., 32(1 ), 71-82 (1997). Wu, D.J., et al., Recovery and purification of paclitaxel using low-pressure liquid chromatography, AIChE J., 43(1), 232-242 (1997). Yun, T.; Zhong, G.M., and Guiochon, G., Experimental study of the influence of the flow rates in SMB chromatography, AIChE J., 43(11), 2970-2983 (1997). Zagorodni, A.A.; Muiraviev, D.N., and Muhammed, M., The separation of Zn and Cu using chelating ion exchangers and temperature variations, Sep. Sci. Technol., 32(1), 413-429 (1997). Zheng, L.G., Some practical properties of sec-alkyl (C11-13) hydrogen styrylphosphonate, Solvent Extr. Ion Exch., 15(6), 1043-1049 (1997). Zheng, Z.; Anthony, R.G., and Miller, J.E., Modeling multicomponent ion exchange utilizing hydrous crystalline silicotitanates by a multiple interactive ion exchange site model, Ind. Eng. Chem. Res., 36(6), 2427-2434 (1997). Zhong, G., and Guiochon, G., Simulated moving bed chromatography: Effects of axial dispersion and mass transfer under linear conditions, Chem. Eng. Sci., 52(18), 3117-3132 (1997). Zhong, G., and Guiochon, G., Simulated moving bed chromatography: Comparison between the behaviors under linear and nonlinear conditions, Chem. Eng. Sci., 52(23), 4403-4418 (1997). Zhong, G.M.; Smith, M.S., and Guiochon, G., Effect of the flow rates in linear, ideal, simulated moving-bed chromatography, AIChE J., 43(11), 2960-2969 (1997). Zurer, P.S., Chromatography and mass spectrometry, Chem. Eng. News, 31 March, 42-47 (1997).

LIST OF J O U R N A L S S U R V E Y E D Abbreviation

Adsorption Journal Adsorption Science and Technology Advances in Chemical Engineering American Institute of Chemical Engineers Journal Ammonia Plant Safety Biotechnology Progress Canadian Journal of Chemical Engineering Catalysis Reviews in Science and Engineering Chemical Engineering (McGraw-Hill, New York) The Chemical Engineer (IChemE, UK) Chemical Engineering in Australia Chemical Engineering Communications Chemical Engineering Education Chemical Engineering Journal (including Biochemical Engineering Journal) Chemical and Engineering News Chemical Engineering and Processing Chemical Engineering Progress Chemical Engineering Research and Design Chemical Engineering Science Chemistry and Industry Chemtech Computers in Chemical Engineering Developments in Chemical Engineering and Mineral Proceesing Energy and Fuels Energy World Environmental Progress Food and Bioproducts Processing Fuel

Adsorption Adsorpt. Sci. Technol. Adv. Chem. Eng. AIChE J. Ammonia Plant Safety Biotechnol. Prog. Can. J. Chem. Eng. Catal. Rev. Sci. Eng. Chem. Eng. (N.Y.) Chem. Eng. (Rugby, Engl.) Chem. Eng. Aust. Chem. Eng. Commun. Chem. Eng. Educ. Chem. Eng. J. (Biochem. Eng. J.) Chem. Eng. News Chem. Eng. Process. Chem. Eng. Prog. Chem. Eng. Res. Des. Chem. Eng. Sci. Chem. Ind. (London) Chemtech Comput. Chem. Eng. Dev. Chem. Eng. Mineral Process. Energy Fuels Energy World Environ. Prog. Food Bioprod. Process. Fuel

1046

Gas Separation and Purification Hydrocarbon Processing Industrial and Engineering Chemistry Process Design and Development Industrial and Engineering Chemistry Research International Journal of Heat and Mass Transfer Journal of Applied Chemistry Journal of Chemical Technology and Biotechnology Journal of Food Engineering Journal of Institute of Energy Journal of Institute of Fuel Journal of Loss Prevention in the Process Industries Powder Handling and Processing Powder Technology Plant/Operations Progress Process Engineering Processing Process Safety and Environmental Protection Process Safety Progress Separation and Purification Methods Separation Science Separation Science and Technology Separation Technology Solvent Extraction and Ion Exchange Transactions of IChemE

Gas Sep. Purif. Hydrocarbon Process. Ind. Eng. Chem. Process Des. Dev. Ind. Eng. Chem. Res. Int. J. Heat Mass Transfer J. Appl. Chem. J. Chem. Technol. Biotechnol. J. Food Eng. J. Inst. Energy J. Inst. Fuel J. Loss Prev. Process Ind. Powder Handling Process. Powder Technol. Plant/Operations Prog. Process Eng. (London) Processing (Sutton, Engl.) Process Safety Environ. Prot. Process Safety Prog. Sep. Purif. Methods Sep. Sci. Sep. Sci. Technol. Sep. Technol. Solvent Extr. Ion Exch. Trans. IChemE

1047

Subject Index A Activated carbon for enrichment and separation of toxic metals, 771 modification for the analytical purposes, 772-783 Activated Carbon Fibres (ACF), 740 Adhesion soot particles to polymeric surface, 194-198 work, 178 Adsorbent activated carbons, 639,640, 808-810,817 activated carbons for ELCD, 821 et seq. BWC test, 822,824 GWC test, 822,824 SHED test, 822,824 production, 825-830 adsorption - active based on natural sorbents, 700-715 adsorption of NOx, 435 boehmite, 394 carbon materials for vapour pollutants, 397 et seq. activated carbons from lignocelulosic origin, 398,406-414 glassy carbons, 398, 407-415 carbon nanotubes, 639,640 dessicant materials, 839 elutrilithe, 382,383 for tobacco smoke, 852 GAC, 763 gibbsite, 394 natural, their structure and properties, 659-670

new composite, 381 et seq. selective and reversible for adsorption of NOx CuO - based, 444-449 mixed metal oxides, 442-444 sulfated metal oxides, 449-481 superconducting, 442 slag media, 533 zeolites, 245 et seq., 639,640 Adsorption and space craft applications carbon dioxide removal (Skylab), 460-463 future directions, 466,467 trace contaminant control, 458-460 water recovery, 463-466 CFCs in faujasite zeolite, 275-280 in polarographic/voltammetric environmental analysis, 111 et seq. from very diluted atmospheres by carbons differential heats, 419,422,423 under dynamic conditions, 417-430 in life support systems, 455 et seq. inside zeolites - molecular modelling, 246,259-265,273-275 in space environmental control, 455 et seq. microbial, 843 of cations on carbonates, 367-369 of cations on clay minerals, 362-367 of corrosion inhibitors, 867-89determining surface excess, 872-877

1048 differential capacity curves of EDL, 867-872 of exchangeable of ions on the metal oxides, 355-362 of metals to living or dead cells, 904 of microorganisms to surfaces, 842 of phosphorus by slag media, 533 et seq. of the toxic ions by activated carbons, 783- 797 of trace metal cations in soils, see also Sorption, 319 et seq. of vapour polutants by carbons, 397 et seq. nitric oxides, 435 et seq. from combustion gases, 435 et seq. on ion - exchanged zeolites, 437-441 on electrode, 112,113 on ideal activated carbon for ELCD test, 824 on layer deposited along optical fibre, 935-937 on powdered activated carbon, 726,727 on solid aerosol surfaces, 571 et seq. of organic compounds on inorganic oxides, 580-606 of organic pollutants on heterogeneous surfaces, 577-580 phenomena in electroanalysis, 111,112 sampling and sample preparation, 8-14 surfactants application in carbody experimental techniques, 178-180 on diesel engine soot, 177,178, 190-194 washing, 177 et seq. supercritical gases by nanospaces, 635 et seq.

Adsorption excess, 179 of surfactants at interface, 203 Adsorption heat pumps, 949 general idea, 949 operating principles, 950,951 prototypes and tests, 971-975 refrigerants and adsorbents, 951-953 thermodynamic cycles, 953-963 thermodynamic performances, 963-970 Adsorption isotherm DR, 643 for phenol compounds, 386-389 Frumkin, 867,877 Gibbs, 179, 873 on unwashed diesel engine soot, 188,189 virial, 867,884 Adsorption potential Lennard-Jones, 640,643 Steel, 640 Adsorption, see also Sorption phenomena in environmental systems, 285 Adsorptive Stripping Potentiometry (AdSP),112 Adsorptive Stripping Tensammetry (AdST),122 Adsorptive Stripping Voltammetry (AdSV),ll2 Air Purification (AP), 213 Air Revitalization System (ARS), 457 Airborne microbes, 834-836 Amino-Methyl Phosphonic Acid (AMPA), 724 Analysis of surfactants in the aquatic environment, 135 et seq. solved and unresolved questions, 135-168 in the aquatic environment initial steps, 137,138 ionic surfactants, 140-150 non-ionic surfactants, 150-162

1049 separation of surfactants and their metabolites, 138-140 Anodic Stripping Voltammetry (ASV),II3 Apollo, 463 Aquatic environment, 135 concentrations of surfactants, 149 Artificial kidney, 77 Artificial liver, 77 Atomic Absorption Spectroscopy (AAS), 798 Atomic Emission Spectroscopy (AES), 8OO

B Boyd-Adamson-Myers (BAM) theory, 746 Bed Capacity Factor (BCF), 216 Benzoic Acid(BA),82 Benzoylperoxide(BPO),82 Bibliography, on adsorption and adsorptive type separations for environmental protection, 979 et seq. Biological fluids, 77-85 Biomineralization, 904,905 Biosorption and detoxification of heavy metals, 904 and metal-bacteria interactions, 904-906 marine environmental, 904 Bismuth Active Substances (BIAS) method, 152 Blood Urea Nitrogen(BUN),83 Breakthrough curves, 545 for different beds, 848,853 for GAC beds, 729-733 for slag media, 545-547 Breakthrough moment, 733 Butane Working Capacity (BWC) test, 822

C Capillary Electrochromatography (CEC), 31 Capillary Electrophoresis (CE), 31, 83 Capillary Gel Electrophoresis(CGE),31 Capillary Isoelectric Foeusing(CIEF),31 Capillary Isotachophoresis(CITP),31 Capillary Zone Electrophoresis (CZE),31, 84 Carbody surfaces contact angle measurements, 201-204 wetting by adsorption surfactants, 198-203 Carbody washing adsorption of surfactants as condition, 208-210 Cation Exchange Capacity (CEC), 323 Cell surface layer (S-layer), 903 CFCs, 246 Chemical transducer, 926,940 Chemisorption and surface reactions on metaloxides, 606-626 on solid aerosol surfaces, 571 et seq. Chlorofluorocarbons (CFCs), 214, 949 Chornobyl nuclear power plant, 698 natural sorbents as decontamination agents, 698-700 Chromatography definitions and applications, 14-31 Clapeyron diagram, 954 Cobalt Thiocynate Active Substances (CTAS) method, 154 Coefficient Of Performance (COP), 963 Colloid science in the soil systems, 351 et seq. Combustion gas, 435,436 Complexation behaviour of organophosphorus polymer-supported reagents, 475-492 selective of metal ions, 473 et seq.

1050 Computational studies on the design of zeolite catalysts and novel adsorbents, 245 for pollution control, 246 et seq. Computer simulation methods, 246 Molecular Dynamics (MD) method, 246 Monte Carlo (MC) method, 246 Condensation Nucleus Counter (CNC), 846 Contaminant pesticides in the surface water, 723,724 Control of air pollution by adsorption, 807 adsorbents, 807 adsorption isotherms, 811,812 fixed bed probes, 812-816 of environment in space station, 455-458 Corrosion mechanism, 863 Corrosion inhibitors, 863-867 Critical Micelle Concentration (CMC), 181 Crystal growth the influence of bacteria in, 903 Crystal promotion, 906,907 Crystallization environmental applications, 903-905 physiological role, 905

D Detergency carbody surface, 204-207 Diesel- Exhaust Particles (DEP), 178 Diesel engine soot characterization, 185-189 Diffusion inside zeolites - molecular modelling, 246,259-265 Diffusion coefficient, 399

Dimethyl Methylphosphonate (DMMP), 213 Double Electrical Layer, see also Electrical Double Layer basis for electrochemical analysis, 111 Dual Mechanism Bifunctional Polymers (DMBPs), 481-484

E Electric Double Layer (EDL), see also Double Electrical Layer, 287, 288, 867 charge density, 886 diffuse charge density, 353 Gouy- Chapman (GC) model, 353 Gouy-Chapman-Stern-Grahame (GCSG) model, 288 Inner Helmholtz Plane (IHP), 355 MUSIC model, 354 soil/soil solution interface, 322 Stern layer model, 322 zeta potential, 354 Electroanalytical methods polarography/voltammetry, 111-113 Adsorptive Stripping Potentiometry (AdSP),112 Adsorptive Stripping Tensammetry (AdST), 122 Adsorptive Stripping Voltammetry (AdSV),112,113 Anodic Stripping Voltammetry (ASV),III-113 Electrochemical Detector (ECD),78 Electrode charge, 873 potential, 873 Elutrilithe, 381 chemical composition, 383 Energy Dispersive X-Ray Fluorescence (EDXRF), 794 Engine diesel, 435 gasoline, 435

1051 catalysts, 435,436 lean- burn, 436 reach- burn, 442 Enrichment (E), 216 Enthalpy of adsorption on glassy carbons, 428 Environmental analysis, 770,771 evaluation of the environmental profiles, 772-774 Environmental analysis, 78 adsorption phenomena, 3 et seq. adsorption in polarographic/ voltammetric, 111 et seq. Adsorption Stripping Voltammetry (AdSV) methods,112-125 chromatographic methods, 14-33 miscellaneous electroanalytical methods, 125-131 electrocapillary, 129-131 other methods, 131 tensammetry, 127-129 micellaneous methods, 31-33 sampling methods 4,8-14 steps, 3,14 Environmental Control and Life Support System (ECLSS), 456 design of, 456-458 Environmental media, 3 Environmental pollution diesel engines and DEP, 178 general problem, 571-574 Environmental Protection Agency (EPA), 463 Environmental Tobacco Smoke (ETS), 838 Equation Dubinin - Isotova (DI), 413 Dubinin - Radushkevich (DR), 401 Freundlich, 725 Kelvin, 195 Koryta, 113 Young, 203

Equivalent Background Concentration (EBC) model, 725 Evaporative Loss Control Devices (ELCD) filtres, 821 Extraction biovailability of metals, 341-344 metal leachability from the soil to phosphorous - based complexants, 474-492 the ground water, 344,345 selective sequential from soils, 331-335 solvent with soluble complexants, 473

F Fibre Optical Chemical Sensor (FOCS), 925 adsorption-based optical transduction in, 925 et seq. applications, 940-946 construction, 925,926 working principles, 930-934 Flame Atomic Absorption Spectroscopy

(FANS) FTIR, 780,783

G G immunoglobulins (IgG), 904 Gas Chromatography (GC),21-26 Gas-Liquid Chromatography(GLC),78 Gasoline Working Capacity (GWC), 822 equipment, 823 Gemini, 463 Global environmental problems, 214 emission control of greenhouse gases, 214 emission control of ozone depletion gases, 214 recovery of CFCs and VOC, 214 Glycosaminoglycans (GAGs), 905 Granular Activated Carbon (GAC), 724

1052 Graphite Furnace Atomic Absorption Spectrometer (GFAAS), 786 Greenhouse effect, 949 Groundwater treatment, 766

H Halobacteria, 903 Hamaker constant, 197, 375 Hanging Mercury Drop Electrode (HMD E), 155 Heterogeneous atmospheric organic chemistry, 572,627 Heterogeneous catalysts environment- friendly, 246 Friedel-Crafts, 246 hazardous, 246 molecular shape selectivity, 259 zeolites, 245 et seq. High Performance Liquid Chromatography(HPLC),15-21 Homogeneous Surface Diffusion Model (HSDM), 728 HPLC - Mass Spectrometry(MS),82 HPLC,78

Ideal Adsorbed Solution Theory (IAST), 725 In Situ Resource Utilization (ISRU), 466 Indirect Tensammetric Method (ITM),156 Indirect Tensammetric Technique (ITT),156 Indoor air pollutants, 833-836 Inductively Coupled Plasma Atomic Emission Spectrometer (ICPAES), 538 Inductively Coupled Plasma-Atomic Emission Spectroscopy(ICP-AES), 123 Inductively Coupled Plasma-Mass Spectroscopy(I CP-MS), 124 Infrared Spectroscopy (IRS), 442

International Space Station (ISS), 456-458 Inverse Gas Solid Chromatography (IGSC), 399 Ion exchange, 449-451 and sewages, 504-521 in soil materials, 320-328 new trends, 521-523 removal of metallic ions from water kinetics model in natural water system, 745 selecticity, 498-504 water recovery, 504-521 with crosslinked polymer beads, 473-492 Ion exchangers active carbon, 772 selective (tabulated data), 489-503 Ionic surfactants anionic surfactants, 140-146 cationic surfactants, 146-150

K KINEQL program, 285,294 Kinetics gas adsorption on activated carbons, 813 ionic solute adsorption, 745 main steps, 745 mathematical model, 748-755 of chemisorption and surface reactions, 610-612 of ion sorption processes, 299-306 factors influencing, 299,300 models, 300 KINEQL, 300-306 of phosphorus adsorption, 535-548 of physisorption on carbon materials, 399,400,407 of zinc ion reduction, 890

L Lambert - Beer law, 180, 930

1053 Langmuir-Blodgett films, 940, 941 Life Cycle Assessment (LCA), 763 backgrounds, 764 Linear Alkylobenzene Sulphonate (LAS),I39 Linear Driving Force (LDF), 730 Liquid-Liquid Extraction(LLE), 37, 84 Liquid-Solid Extraction(LSE),37 Local environmental problems, 214 removal SOx and NOx from flue gas, 214 solvent vapor fractionation, 214 SVR, 214 Low Molecular Weight Organic Acid (LMWOA), 342

Molecular modelling, see also Computation Studies adsorption and diffusion in zeolites, 245 in relevance to environmental multitechnique methods, 246, 247-266 protection, 245 et seq. Molecular shape selectivity, 259 Monitoring, 8-14 general problem, 3 monitored substances, 2-7 sampling methods, 8-14, 37-39 M-Toluidine (MT), 867 MUltiSIte Complexation (MUSIC) model, 354

M M immunoglobulins (IgM), 904 Man-made hydrocarbon emissions, 822 Mars atmosphere, 466 Marshall Space Flight Center (MSFC), 462 Mass Transfer Zone (MTZ), 812 Mercury, 463 Methyl Methacrylate(MMA),82 4,4' Methylenedianiline(MDA), 78 Methylene Blue Active Substance (MBAS) method, 140 Methylenediisocyanite(MDI),78 Micellar Electrokinetic Capillary Chromatography(MECC), 31 Micellar Electrokinetic Chromatography (MEKC),83 Micellar Electrokinetic Gas Chromatography (MEGC),84 Micropore filling mechanism, 644 MINEQL program, 291 Mixed adsorption layers formed by corrosion inhibitors, 863 et seq. Modelling of metal ion sorption phenomena, 285 et seq.

N N,N-Dimethyl P-Toluidine(DMPT),82 Nanoparticles or nanomaterials, 638-640 Nanopore Molecular Engineering, 654 Nanospace system, see also Nanoparticles, 635,638-640 adsorption in, 640-642 National Aeronautics and Space Administration (NASA), 457 Natural Organic Matter (NOM), 724 n-butanol (BU), 866 Non-ionic surfactants, 150-169 Nucleation and sodium chloride crystal growth, 903 et seq.

O Opthode adsorption based, 934,935,946 Optical fibre description, 926-930 Ozone depletion, 949 Ozone/UV, 763

1054

P

R

Parson's function, 873 Percolation HCH/chlorobenzenes contaminated groundwater, 763 water from landfill, 763 Pollutants anionic organic and inorganic, 381 applications of the adsorption phenomena for their analyses, 3 et seq. cationic organic and inorganic, 381 environmental, 5-7 anthropogenic, 5 natural, 5 secondary, 5 neutral organic and inorganic, 381 phenol compounds, 382,383 vapour, 400 et seq. Polycyclic Aromatic Nitrogen Heterocyclic (PANHs) compounds, 13 Polyethylene Glycol (PEG), 139, 904 Polymethyl Methacylate(PMMA),82 Polyurethane surface, 178 Polyurethane (PU),78 Pore Diffusion Model (PDM), 537,538 Potential zero charge (pzc), 867,872 Potentially Toxic Elements (PTE), 352 Pressure Swing Adsorption (PSA), 213 application for the environment, 213 et seq. Air Purification (AP), 213, 216, 219, 230-232 defence applications, 213 fundamental of environmental PSA processes, 214,215-218 mathematical modelling, 218-241, PSA-SVR, 219,220-230,232-241 Production of shape selective of alkylaromatics, 259-265 Pseudosurfactant, 144,145 P-Toluidine (PT), 867

Rapid Small Scale Column Test (RSSCT), 728 Raw material almond shells, 398 lignocellulase, 398 olive stones, 398 organic copolymers, 398 Recovery of CFCs and VOC, 214 of gasoline vapour, 821 of gold and platinum metals, 504-508 of humidity condensate, 455 of metal ions from aqueous solutions, 473 et seq. of organics from chemical processes and storage-tanks, 214 of proteins, 903 of silver, 508-509 of water, 463-466 Relative Standard Deviation (RSD), 801 Removal nitrogen oxides, 269 by selective catalytic reduction, 269-275 of contaminants in defense applications, 214 of disperse impurities from water, 670-682 of hazardous compounds in biological fluids, 77 et seq. of inorganic cations from water, 691-700 of microorganisms and particulates from indoor air, 833 adsorbent based removal systems, 839,840 control strategies, 838,839 fundamentals and mechanisms, 841-844 laboratory test, 844-861

1055 of CFCs, 246 of liquid soils from a surface, 177 of metallic ions from water and pollutants from wastewater, 381 et seq. of nitric oxides, 436 et seq. of oil and petroleum pollutants from water, 710-715 of organic molecules and ions form water, 682-691 of pesticides from the surface water, 733-737 of phosphorus by slag, 523 et seq. characteristics of the adsorbents, 538-540 from water and wastewater, 533 kinetic models, 535-548 Pure Diffusion Model (PDM), 536, 548,549 using soil and slag media, 534 of SOx and NOx from flue gases, 214 of trace gas-phase contaminations, 455,460-463 sewages, 504-521 chromium, 516 copper, 514 lead, 513 mercury, 510-513 nickel and vanadium, 515 tin and cobalt, 509 zinc, 519-521 Respirable Suspended Particulates (RSP), 833

Selective Catalytic Reduction (SCR), 435 nitric oxide with hydrocarbons, 435,436 Separation of surfactants and their metabolites, 138-140 gas stripping technique, 139 ion - exchange, 139

liquid/liquid extraction, 138 solid/phase extraction, 139 Sequential Quadratic Programming (SOP), 294 Sewage Treatment Plant (STP), 533 SHED emission test, 821 Skylab, 460,461 Slag media, 533 Soil adsorbing and complexing properties, 369-373 environmental general problems, 351-353 geochemical phases, 329-331 phase distribution of metals, 336-341 solids, 319-323 solution, 321,322 trace metal cation sorption in, 319 et seq. types, 320 Solid aerosol composition, 573-577 industrial, 571-573 Solid Phase Extraction (SPE), 11, 37 application to environmental analyses, 37,65-73 impact of various factors, 51-64 overview, 37-39 basic steps of SPE, 45-47 chemical characteristic of the sorbent, 41,42 physical characteristic of the sorbent, 39,40 sorbent selection, 43,44 SPE format, 44,45 Solvent Vapor Recovery (SVR), 213 Sorption of trace metal cations in soils, 319 et seq. phenomena in environmental systems, 285 et seq. Sorption equilibria of metal ions, 286-299

1056 factors influencing, 286-288 models, 288 SCFM, 288 SPE procedure for blood MDA,82 SPE,77 Super Critical Fluid Chromatography (SCFC),30,31 Super Fluid Critical Extraction (SFCE),77 Supercritical gas, 637 control with solid nanospaces, 635 et seq. practical importance, 635,654 physical properties, 637 Surface Active Substance (SAS), 111, 682 Surface charge, 771 Surface Complex Formation Model (SCFM), 285,288Surface Diffusion Model (SDM), 535 Surface pressure, 873 Surface tension critical of a carbody surface, 198,199 surfactant solutions, 179,182-185 Surfactant applications, 136,137 main types and mixtures, 136 Surfactant solution aquatic, 135 characterization, 180 Surfactants, 180,181 Synthesis of organophosphorus polymersupported reagents, 475-492 Synthesis of acyclic enones by solid acid catalysts, 246-253

T Temperature Programmed Desorption (TPD), 442 Temperature Swing Adsorption (TSA), 220 Tetrahydrofurane (THF), 399

Thin Layer Chromatography(TLC), 26-30 Thiourea (TU), 867 Toxic compounds to human health selective retention, removal and elution for analysis of,85-106 Trace analysis by atomic spectroscopy, 797-803 slurry sampling technique, 802 Transport of metal ions in natural systems, 285, 306-312 HYDROGEOCHEM model, 285, 308-312

U Ultraviolet Detector (UVD),78

V Vapor Compression Distillation (VCD), 463 Volatile Organic Compounds (VOC), 213, 416

W Wastewater treatment, 381 et seq., 728-742 advanced adsorption techniques, 763 advanced oxidation by means of ozone/UV, 763,7655,769 environmental analysis, 770,771 environmental profiles, 772-774 GAC on reactivation basis, 763,765,767 by adsorptive slag media, 533 by filtration, 737 by ion exchange, 473 in space craft, 463-466 ion exchange, 497, 745 natural sorbents, 659-700 adsorption- active based on, 700-715

1057 new composite adsorbents, 382-394 Water purification, 659 et seq. treatment, 745 Water production by activated carbon filtration, 723 et seq. GAC filtration, 737-742 general problems, 725-727 HSDM, 728,729 membrane/GAC filtration, 738-740 ozone-GAC - filtration, 741 Water Recovery System (WRS), 457

Zeolites model of cages and windows in zeolite Y, 253,254 pore dimensions and architecture, 249,250