ࡱ>  Z\ WXYZoG Abjbjَ ȔY:]T**** *Lc" 0^1^1^1^168:a H+ + Cl- (2.11) CH3COOH -----> CH3COO- + H+ (2.12) Z = 2, for diprotic acids (e.g., sulfuric acid) H2SO4 --------> 2H+ + SO42 (2.13) Z = 3, for triprotic acids (e.g., phosphoric acid) H3PO4 --------> 3H+ + PO43 (2.14) For bases, Z is the number of hydroxide ions liberated, because one hydroxide neutralizes one proton to form water. Z = 1, for mono-hydroxides (e.g., sodium hydroxide) NaOH -----------> Na+ + OH- (2.15) Z = 2, for di-hydroxides (e.g., calcium hydroxide) Ca(OH)2 ----------> Ca+ + 2OH- (2.16) For species undergoing oxidation or reduction (redox) reactions, Z is the number of electrons transferred per molecule. Z = 1, for single electron transfers, such as the oxidation of ferrous iron to ferric iron. Fe+2 -------------> Fe+3 + e- (2.17) Z = 6, for six electron transfers, such as the reduction of potassium dichromate to trivalent chromium. K2Cr2O7 + 7H+ + 6e- ----------> 2K+ + 2Cr+3 + 7OH- (2.18) For other reactions such as precipitation or complexation, Z will depend on the particular stoichiometry. However, in most cases the value of Z will be equal to the value of the oxidation state of the atom (or group) that will be reacting times the number of these atoms (or groups) that are bound in the reacting molecule. For example there are two atoms of Al(+III) in one molecule of alum (Al2(SO4)3 18H2O). Therefore, the Z value for alum is equal to six. Balancing Equations The first step in working with redox reactions is identifying the nature of the reacting species. Which one is being oxidized and which one is being reduced. This is done by writing expected products and analyzing the oxidation states of all species. The following is a step-by-step method for balancing any general redox equation. 1. Identify substances undergoing reduction and those undergoing oxidation 2. Prepare the oxidation half reaction by the following procedure identify the substance containing the element being oxidized and bring it down so that it is on the left hand side of the oxidation reaction identify the substance containing the oxidized form of this element and pull it into the right hand side of the oxidation half reaction. determine its stoichiometric coefficient based on the number of atoms of this type on the left hand side determine the change in oxidation state of the oxidized element, multiply it by the number of atoms of this element (often one) and add this number of electrons to the right hand side of the equation balance the number of oxygen atoms by adding water molecules to one side or the other balance the number of hydrogen atoms by adding hydrogen ions to one side or the other verify that the charges balance on both sides of the equation 3. Prepare the reduction half reaction by the following procedure identify the substance containing the element being reduced and bring it down so that it is on the left hand side of the reduction reaction identify the substance containing the reduced form of this element and pull it into the right hand side of the reduction half reaction. determine its stoichiometric coefficient based on the number of atoms of this type on the left hand side determine the change in oxidation state of the reduced element, multiply it by the number of atoms of this element (often one) and add this number of electrons to the left hand side of the equation balance the number of oxygen atoms by adding water molecules to one side or the other balance the number of hydrogen atoms by adding hydrogen ions to one side or the other verify that the charges balance on both sides of the equation 4. Find the lowest common denominator for the number of electrons in each equation. Then multiply each equation (i.e., multiply the stoichiometric coefficients) by the factor required in each case to give the same number of electrons (electrons consumed in the case of the reduction reaction, and electrons produced in the case of the oxidation reaction) 5. Now add the two half reactions, and the overall REDOX reaction will be balanced As an example let's consider the oxidation of manganese by ozone. The substance being oxidized is manganese, and the one doing the oxidizing (i.e., being itself reduced) is ozone.  EMBED Equation.2  (xx5) Next we need to evaluate the products formed. We know from experience that reduced soluble manganese (i.e., Mn+2) can be oxidized in water to the relatively insoluble manganese dioxide. We also know that ozone ultimately forms hydroxide and oxygen after it becomes reduced.  EMBED Equation.2  (xx5) The next step is to determine the oxidation state of all atoms involved.  EMBED Equation.2  (xx5) From this analysis, it is clear that manganese is oxidized from +II to +IV, which involves a loss of 2 electrons per atom. On the other hand, the ozone undergoes a gain of 2 electrons per molecule, as one of the three oxygen atoms goes from an oxidation state of 0 to -II. Therefore the two half reactions can be written as single electron transfers. These half reactions are balanced by adding water molecules and H+ ions to balance oxygen and hydrogen, respectively.  EMBED Equation.2  (xx5) By convention, when hydroxide appears in a half-reaction, we add additional H+ ions until all of the hydroxide is converted to water. This is done to the reduction half-reaction.  EMBED Equation.2  (xx5) From this point it is a simple matter of combining the equations and canceling out terms or portions of term that appear on both sides. At the same time we can combine the standard electrode potentials, to ge the overall cell potential. Note that the sign on the potential for the oxidation half reaction must always be reversed, because by convention, all half-cell reactions are given as reductions.  EMBED Equation.2   EMBED Equation.2  (xx5)  EMBED Equation.2   EMBED Equation.2  now rearranging equation xx7, we get:  EMBED Equation.2  (xx5) so for this reaction  EMBED Equation.2  (xx5)  EMBED Equation.2  (xx5) and because the activity of solvents (i.e., water) and solid phases are by convention equal to one.  EMBED Equation.2  (xx5) Furthermore, if we pick a pH of 7.0 and maintain an oxygen dissolved oxygen concentration of 10 mg/L and an ozone concentration in the contactor of 0.5 mg/L, we calculate and equilibrium manganese concentration of So if we want to determine is ozone can possibly oxidize manganese dioxide forming permanganate, we combine the above ozone equation with the reverse of the permanganate equation. O3(aq) + H+ + e- ( O2(aq) + H2O 2/3 H2O 1/3 MnO2 ( 1/3 MnO4- + 4/3 H+ + e- O3(aq) + 1/3 MnO2 ( 1/3 MnO4- + 1/3 H+ + O2(aq) + 1/3 H2O  EMBED Equation.2   EMBED Equation.2  And now we can formulate the equilibrium quotient directly from the balanced equation. Note that neither manganese dioxide (MnO2) nor water (H2O) appear in this quotient. This is because both are presumed present at unit activity. Manganese dioxide is a solid and as long as it remains in the system, it is considered to be in a pure, undiluted state. The same may be said for water. As long as the solutes remain dilute, the concentration of water is at its maximum and remains constant.  EMBED Equation.2  So under typical conditions where the pH is near neutrality (i.e., [H+] = 10-7), dissolved oxygen is near saturation (i.e., [O2(aq)] = 3x10-4) and the ozone residual is 0.25 mg/L (i.e., [O3(aq) = 5x10-6); the expected equilibrium permanganate concentration should be:  EMBED Equation.2   EMBED Equation.2  and solving for permanganate  EMBED Equation.2  Obviously one cannot have 327 moles/liter of permanganate. Nevertheless, the system will be forced in this direction so that all of the manganese dioxide would be converted to permanganate. Once the manganese dioxide is gone, the reaction must stop. The previous thermodynamic analysis only addresses the equilibrium condition. Kinetics also plays an important role in these processes. Many redox reactions that are expected to occur actually proceed at such a slow rate as to be considered non-reactive.  EMBED JandelGraphicObject.1  Figure  STYLEREF 1 \n 2- SEQ Figure \* ARABIC \r 1 1 Stoichiometry of Manganese Oxidation at pH 8.0 in the Absence of Organic Matter (after  QUOTE "{Reckhow, Knocke, et al. 1991 #230}" ADDIN  ADDIN  {Reckhow, Knocke, et al. 1991 #230})  EMBED JandelGraphicObject.1  Figure  STYLEREF 1 \n 2- SEQ Figure \* ARABIC 2 Stoichiometry of Manganese Oxidation in the Presence of 5 mg/L DOC (after  QUOTE "{Reckhow, Knocke, et al. 1991 #230}" ADDIN  ADDIN  {Reckhow, Knocke, et al. 1991 #230}) Kinetics and Mechanism Types of Reactions (include. oxidation vs substitution reactions) To this point we have considered whether or not a certain redox reaction can occur, and perhaps how fast it can occur. However, it is sometimes even more important to know how the reaction occurs, and by what mechanism or pathway it goes from products to reactants. For example the problem of disinfection byproducts is one of chemical pathways. There is no inherent problem with oxidizing natural organic matter with chlorine. However, when that reaction occurs through addition and substitution reactions, rather than simple electron transfers, we get chlorinated organic byproducts such as the THMs. Redox reaction can occur through a wide range of pathways. They may be generally categorized as those involving electron transfer and those involving transfer of atoms and groups of atoms. Aqueous chlorine presents an interesting array of reaction types. Show diagrams of: electron transfer, substitution addition combined reaction (oxidative decarboxylation?) Reaction Pathways Oxygen, in its most stable form (triplet oxygen) has two odd electrons with similar spin (i.e., it is a diradical) and it is paramagnetic. Oxygen reactions are often catalyzed by free radical species, because oxidation reactions with oxygen are themselves often free-radical reactions. Prominent among these catalysts are the transition metals, and metal oxides. Hydrogen peroxide is available commercially in solutions up to 30%. Stabilizers are added to prevent rapid decomposition, which can occur from trace concentrations of aluminum and many transition metals. Despite their similarities, the water treatment oxidants will react with different groups. This is evident in the observations that preozonation of a water has little or not effect on the permanganate demand ( QUOTE "{Wilczak, Knocke, et al. 1993 #1660}" ADDIN  ADDIN  {Wilczak, Knocke, et al. 1993 #1660}) and only a small impact on the chlorine demand ( QUOTE "{Reckhow, Legube, et al. 1986 #310}" ADDIN  ADDIN  {Reckhow, Legube, et al. 1986 #310}). Reaction Kinetics Consider a reaction in the gas phase between molecules of hydrogen and iodine. H2 + I2 -----> 2 HI (4.1) The rate at which this reaction occurs will depend on a number of factors such as the concentration of hydrogen and iodine in the reacting gas. In order for a molecule of hydrogen and a molecule of iodine to combine to form hydrogen iodide, the two molecules must come into contact with each other (contact meaning approach within a certain distance so that bonding forces can play a role). Imagine a chamber containing 100 molecules of hydrogen and only 1 molecule of iodine. Eventually the molecule of iodine will, by random motion, contact a hydrogen molecule in such a way as to form hydrogen iodide. This may take some time, however, due to the small number of hydrogen molecules available for contact. If the number of hydrogens is doubled, one would expect the formation of hydrogen iodide to occur twice as fast. On the other hand, if the number of iodine molecules was increased to 10, the rate at which any one of the iodine molecules forms hydrogen iodide would be unchanged, but the overall rate of hydrogen iodide formation would increase by a factor of 10. Thus, a general rate law for hydrogen iodide formation can be written, based on the simple concept of probability of molecular contact. That is, the rate of formation of the reaction product is proportional to the number of molecules of one reactant times the number of molecules of the second reactant. This is the kinetic law of mass action. This principle is the same whether we're dealing with numbers of molecules or molar concentrations of molecules. Equation 4.2 is the common form for this reaction rate law.  EMBED MSDraw  (4.2) where the reactants and products are expressed in molar concentrations and kf is called the forward reaction rate constant. The units for kf are liters/mole per unit time. The reaction tate is going to be a function of such things as the rate of movement of the molecules, and the probability of HI formation given that a collision between hydrogen and iodine has already occurred. Most of the simple reactions discussed here occur in a single phase. These are termed Homogeneous Reactions and they usually involve dissolved species. Many environmentally important reactions involve species in different phases, and these are termed Heterogeneous Reactions. An example of a heterogeneous reaction is the dissolution of calcitic minerals in acidic waters. Such reactions involve both a transfer step and a reaction step. If the transfer step is slow, they are said to be transport-limited. If the chemical reaction step is slow, they are reaction-limited. Consider a homogeneous reaction of the following type: aA + bB ------> cC + dD (4.3) where the capital letters represent chemical species participating in the reaction and the small letters are the stoichiometric coefficients (i.e., the numbers of each molecule or ion required for the reaction). In most cases the rate law will be of the form:  EMBED MSDraw  (4.4) Note that, depending on your frame of reference, equation 4.4 could also be written as:  EMBED MSDraw  (4.5) where:  EMBED MSDraw  (4.6) The overall order of reaction 4.3 is (a+b). The order with respect to species A is a, and the order with respect to species B is b. Homogeneous reactions may be either elementary or non-elementary. Elementary reactions are those reactions that occur exactly as they are written, without any intermediate steps. These reactions almost always involve just one or two reactants. The number of molecules or ions involved in elementary reactions is called the molecularity of the reaction. Thus, for all elementary reactions, the overall order equals the molecularity. Non-elementary reactions involve a series of two or more elementary reactions. Many complex environmental reactions are non-elementary. In general, reactions with an overall reaction order greater than two, or reactions with some non-integer reaction order are non-elementary. Rates of reaction for the various oxidants are often positively correlated. In other words, a compound favored for oxidation by one oxidant is generally favored by others as well. Those that are relatively resistent to oxidation by one, will likewise be unreative to others. A good case study is the data set available for oxidation of phenolic compounds shown in  REF _Ref358533591 \* MERGEFORMAT Figure 2-3 ( QUOTE "{Tratnyek 1995 #7850}" ADDIN  ADDIN  {Tratnyek 1995 #7850}). These highlight the similarities in the effects of substitutent groups on reactivity from one oxidant to another. Figure  STYLEREF 1 \n 2- SEQ Figure \* ARABIC 3 Correlation Matrix of Rate Constants for the Oxidation of Phenolic Compounds by Four Different Chemical Oxidants (after  QUOTE "{Tratnyek 1995 #7850}" ADDIN  ADDIN  {Tratnyek 1995 #7850}). Chemists have used such relationships to develop quantitative structure-activity relationships (QSARs) A very common groups of such relationships are the Hammett equations. These incorporate the effect of substituents on reaction rates into a single factor (() which is substituent-specific, but independent of the oxidant. These are useful, because the rate of a series of reactions with similar mechanisms are often directly related to the free energy of that reaction. For this reason, some have used free energy changes directly (Marcus theory). The Hammett equation has been successfully used for substituted phenols  REF _Ref358551375 \* MERGEFORMAT (Eq 2-2). It is a log expression of the rate constant for any given phenolic compound as equal to the rate for the parent, unsubstituted phenol plus some product of a substituent constant (() and a reaction constant ((). These latter two constants are totally independent of each other, and are therefore perfectly general for all reactions in the case of (, and all substitutents in the case of (.  EMBED Equation.2  (Eq  STYLEREF 1 \n 2- SEQ Eq \* ARABIC 2) While very useful within its intended scope, the Hammett equations have some serious limitations. First, they fail to account for steric effects from ortho substituents. Although, there are some corrections that can be applied for use with such compounds, these systems are clearly more difficult to work with. Second, reactions with compounds other than substituted phenols are not directly included. These are also more difficult to model. Table  STYLEREF 1 \n 2- SEQ Table \* ARABIC \r 1 1 Selected Linear Free Energy Substituent Constants Substituent( -( +metapara, orthodimethylamino-1.7amino-1.3O--0.82-0.47OH0.13-0.38-0.92methoxy0.11-0.28-0.78(-alanyl-0.23t-butyl-0.19methyl-0.06-0.16-0.31ethyl-0.15-0.30propyl-0.28butyl-0.26phenyl-0.18H000F-0.07Br0.220.15Cl0.370.220.11I0.14carboxylato0.37sulfonato0.58carboxyl0.78acetyl0.84CN0.88formyl1.04nitro1.26  EMBED JandelGraphicObject.1  Figure  STYLEREF 1 \n 2- SEQ Figure \* ARABIC 4 Hammett Plot for Reaction of Substituted Phenols and Phenolates with Chlorine Dioxide and Ozone (Data from  QUOTE "{Tratnyek & Hoigne 1994 #4160}" ADDIN  ADDIN  {Tratnyek & Hoigne 1994 #4160}, and  QUOTE "{Hoign & Bader 1983 #7870}" ADDIN  ADDIN  {Hoign & Bader 1983 #7870})  EMBED JandelGraphicObject.1  Figure  STYLEREF 1 \n 2- SEQ Figure \* ARABIC 5 Linear Free Energy Relationship for Reactions of Substituted Benzenes with Ozone (Data from  QUOTE "{Hoign & Bader 1983 #7860}" ADDIN  ADDIN  {Hoign & Bader 1983 #7860}) While this is generally true, there are many exceptions . Analysis Of Kinetic Data Determination of Reaction Rate Constants Reaction rate constants can be evaluated from experimental data by any one of four techniques: the integral method; the excess method; the differential method; and the initial rate method. Integral Method of Analysis This method allows one to use most or all of the experimental data in determining rate constants. However, one must first be certain of the reaction stoichiometry before attempting this kinetic analysis. In the following paragraphs, a series of linearizations are presented depending on the reaction order. If the exact reaction order is uncertain, one may try several different linearizations. The best fit should occur with the linearization that is appropriate for the data (i.e., indicating the correct reaction order). For zero order reactions, the rate is simply a constant.  EMBED MSDraw  (4.7) Integrating equation 4.7 gives:  EMBED MSDraw  (4.8) Thus, the reaction rate constant is obtained from the slope of a plot of the molar concentration of "C" (i.e., [Ct]) versus reaction time. The intercept is, of course, the initial concentration ([C]o). For first order reactions, the rate is proportional to one of the reactants to the 1st power.  EMBED MSDraw  (4.9) Integrating equation 4.9 gives:  EMBED MSDraw  (4.10) Thus, the experimental data are plotted with natural log of the concentration on the y-axis and time on the x-axis. The reaction rate constant is simply the negative of the slope of this line, and the intercept is the natural log of the initial concentration. For second order reactions, the rate is proportional to one of the reactants to the 2nd power, or to two reactants, each to the 1st power. For the former case the following rate equation holds.  EMBED MSDraw  (4.11) Integrating equation 4.11 gives:  EMBED Equation.2  (4.12) Thus, the experimental data are plotted as the reciprocal concentration (y-axis) versus time (x-axis). The reaction rate constant is simply the slope of this line, and the intercept is the reciprocal of the initial concentration. The Excess Method Some second and higher order reactions are more easily examined when one reactant is essentially held constant. This can be done by using a large excess of one of the reactants such that fractional change in concentration over the course of reaction is negligible. For example, if reaction 4.3 is allowed to proceed with "B" originally present at a concentration 50 times greater than "A", then the rate expression reduces to:  EMBED MSDraw  (4.13) where "k" can be treated as a constant with respect to time:  EMBED MSDraw  (4.14) The rate law expressed by equation 4.13 is said to be pseudo-ath order. This means that although the reaction is fundamentally of order (a+b), it appears to be of order "a" in the experiments. Kinetic analysis of 4.13 is much easier than 4.4. This also relieves the need to measure "B" throughout the experiment (i.e., it is unchanging). The Differential Method Some complicated kinetic systems cannot be analyzed by the integral method, with or without the use of an excess. In these cases either the differential, or the initial rate methods should be employed. The differential method has the advantage of allowing one to use all of the experimental data. The simplest case would be for a reaction which is some non-integer order with respect to a reactant. If the reaction is dependent on another reactant, the excess method can be employed to suppress this effect. Experimental data for the reactant are plotted as a function of reaction time. A smooth curve is drawn through these data, and tangents are drawn to the curve at various points. The slope of each one of these tangents gives the instantaneous reaction rate. These rates are then plotted versus concentration of the species being followed on a log-log scale. The slope of the line formed gives the reaction order with respect to this constituent. The Initial Rate Method. Use of the initial rate method requires that many separate experiments be run. However, for reactions of changing order and dependency, this is the best method of analysis. As with the differential method, a smooth curve is drawn through the concentration vs time data. Here, only a tangent to the curve at the origin is constructed. The slope gives the initial rate of this reaction, the rate at time=0, when the solution composition is well known. Experiments with different starting concentrations are run and a single initial rate is determined for each. This is a very versatile method, and it is not subject to competitive or catalytic pathways initiated by reaction products. Determination of Activation Energy & Temperature Dependence As mentioned previously, the reaction rate constant, k, is a function of temperature. The Arrhenius equation (4.15) is the classic model.  EMBED MSDraw  (4.15) where ko is the called the frequency factor, or the pre-exponential factor, Ea is the activation energy, R is the universal gas constant (199 cal/K-mole), and T is the temperature in K. The natural log of the reaction rates are plotted as a function of the reciprocal absolute temperature. The slope is then -Ea/R and the intercept is lnko. In environmental engineering, equation 4.16 is often used to describe the relationship between temperature and reaction rate constants.  EMBED MSDraw  (4.16) where k20 is the value of the rate constant at 20C, T is the temperature in C, and R is an empirically derived constant, usually between 1.0 and 1.2. Although, equation 4.16 is not based in chemical theory, as 4.15 is; it is more convenient to use. Over short spans of temperature equation 4.16 gives results that are similar to 4.15. Catalysis (include. pH effects) Process Evaluation and Monitoring (10 pages) Bench-scale Testing General principles for testing of oxidants (when can we use bench-scale testing, and how should it be done) Bench-scale testing is almost always appropriate when working with purely homogeneous reactions. These are systems where all of the reactants are in the same phase, usually, the aqueous phase. The scale of these reactions is so small that all approaches to process evaluation will be of extreme macro-scale and should give equivalent results. Thus the more convenient, less expensive bench-scale approach should be used. However, all reactions of concern to water treatment involve at the very least, some mixing processes. Reactants, even if they're all in the aqueous phase, must be mixed at the time of addition. These mixing processes can have large length scales, and may not be well simulated in bench-scale. If there are heterogeneous reactions, the need for larger-scale experimentation may exist. Systems with reactions occurring across two phases include: oxidation and precipitation of iron and manganese Control of Iron and Manganese Removal of manganese can be complicated by secondary reactions of the oxidized, particulate forms. Bench-scale studies have shown that ozone will readily oxidize reduced manganese ( QUOTE "{Reckhow, Knocke, et al. 1991 #230}" ADDIN  ADDIN  {Reckhow, Knocke, et al. 1991 #230}). However, pilot-scale research has shown that high levels of soluble manganese will persist through ozonation, coagulation and filtration ( QUOTE "{Reckhow, Knocke, et al. 1991 #230}" ADDIN  ADDIN  {Reckhow, Knocke, et al. 1991 #230}). This is probably due to resolubilization of colloidal manganese by reaction with reduced substances in the water and on the filter media. Complex multi-phase systems such as those encountered in manganese removal must be studied in pilot scale. Nevertheless, bench-scale studies are useful for determining oxidant stoichiometry, and initial oxidation rates. Taste & odor control Oxidation of taste and odor compounds is generally a homogeneous process. It has been very successfully studied in bench-scale. Perhaps the only significant advantages of using pilot-scale is the opportunity of catching transient T&O events, and testing the process under those conditions. Another advantage might be in determining the effectiveness of biological filtration at removing T&O compounds. Oxidation may improve the effectiveness of biological filtration and thereby help remove T&O compounds through a secondary effect. Elimination of Color Like T&O control, the elimination of color may be studied quite effectively at bench-scale. Also like T&O control, color removal may occur as a secondary effect of oxidation during subsequent biological filtration. If this is the case, it is better to use pilot scale processes. Enhancing subsequent processes Bench-scale processes can be used, and they are only constrained by the scale requirements of the subsequent process being investigated. For example, a wide range of studies have been conducted on impacts of ozonation on subsequent biological filtration. These have included fundamental studies aimed at enhancing the theoretical background (e.g.,  QUOTE "{Hozalski, Goel, et al. 1995 #7810}" ADDIN  ADDIN  {Hozalski, Goel, et al. 1995 #7810}), to studies examining a particular application at a particular site. While valuable information can be obtained at bench-scale, there remain some important scale-up problems that require pilot-scale studies for more conclusive design information (e.g., see:  QUOTE "{Hozalski, Goel, et al. 1995 #7810}" ADDIN  ADDIN  {Hozalski, Goel, et al. 1995 #7810}). Filtration processes are sometimes difficult to scale-up. For general filter performance, it is important to minimize wall effects. A filter-diameter-to-media (effective size) ratio of 50 or greater is recommended for filtration studies ( QUOTE "{Lang, Giron, et al. 1993 #1630}" ADDIN  ADDIN  {Lang, Giron, et al. 1993 #1630}). Oxidation of Synthetic Organic Chemicals The case for oxidation of SOC's is very much the same as for the control of color. Bench-scale study is quite acceptable, except where removal by biological filtration is expected. Control of Nuisance Aquatic Growths Pilot & Full-scale testing General principles Whenever a pilot-plant study is undertaken, it is critical that it be carefully planned. Issues such as sample location, type of chemical analysis, level of accuracy needed and use of controls must be addressed. Pilot-scale studies are inherently more expensive than studies conducted at a bench-scale, so the cost savings of a good experimental plan can be quite substantial. In fact, it is often advisable to conduct related bench-scale tests prior to piloting or in parallel with piloting, so that the piloting methods and objectives may be better focused. Details of pilot study design are beyond the scope of this work, but there are many references that can be consulted (e.g.,  QUOTE "{Logsdon, LaBonde, et al. 1996 #7480}" ADDIN  ADDIN  {Logsdon, LaBonde, et al. 1996 #7480}) Applications (refer back to Bench-scale testing wherever possible) Analysis and Monitoring sampling design 2. UV Absorbance Most natural organic matter will absorb sufficient ultraviolet (UV) light to be easily detected by a standard UV-Vis spectrophotometer. By convention, we have chosen 254 nm as the wavelength to measure UV absorbance. This parameter is quite important because: (1) it is inexpensive, rapidly measured, and requires a minimum of training; and (2) it has been found to correlate with certain water quality characteristics, such as DOC and THMFP. UV absorbance has been successfully used as a means of estimating DOC and THM precursor levels in raw waters. However, its most important contribution is to process monitoring. For a single raw water source, coagulation effectiveness can be effectively monitored by UV absorbance. One can generally develop good linear correlations between UV abs and DOC for raw and treated waters from the same plant (e.g., Edzwald et al., 1985). The interpretation changes, however, whne a disinfection or oxidation step is encountered. When monitored across oxidation/disinfection, UV absorbance provides information on the degree of oxidation of the natural organic matter in the water. The specific absorbance, which corresponds to the absorbance per mg/L of DOC, is a useful tool for rapidly assessing the "humic/non-humic nature" of a water. Some specific absorbances for extracted humic and non-humic fractions are shown in Figure 2. Note that humic acid and fulvic acid show the highest SUVA (6.3 and 4.4, respectively). In another study, averages of 10 aquatic humic substances showed SUVA values of 5.8 for the humic acids and 3.6 for the fulvics (Reckhow et al, 1990). Other fractions, especially the hydrophilic acids, show lower SUVA values. For this reason, waters with a high SUVA generally have higher humic contents, and more amenable to DOC removal by coagulation. Table 2 summarizes some attempts to correlate UV absorbance (254 nm) to DOC for raw waters. Note that the reciprocal of the slopes in Table 2 correspond to the specific absorbances in Figure 2, and that the average slope (~25) gives a specific absorbance of 0.004/cm or 4/m which is similar to those reported for fulvic acids. 1. Formation Potentials vs Simulated Distribution System Tests One can distinguish 3 categories of DBP precursor tests: (1) the high-dose formation potential (HDFP) tests; (2) the low-dose formation potential (LDFP) tests; and (3) the SDS tests. The HDFP is characterized by two attributes (a) it is not used to simulate any particular water system or water treatment scenario; and (b) it uses a sufficiently high chlorine dose so that the residual remains high and nearly constant with contact time, and independent of chlorine demand (e.g., many use a dose of 20 mg/L for samples with a chlorine demand of 8 mg/L or less, so that the residual is always between 12 and 20 mg/L). The HDFP is trying to be an unbias precursor test. This means it exhibits the same percent precursor recovery regardless of precursor concentration. Changing TOC concentrations in waters of high bromide does not present a complication with the HDFP, because the TOC "sees" the same oxidizing environment, regardless of what the actual TOC concentration is. This is not the case for the other tests. The LDFP is also characterized by two attributes (a) it is not used to simulate any particular water system or water treatment scenario; and (b) it uses a low chlorine dose so that the residual at the end of the contact time is close to what is commonly found in the taps of most US public water systems. The exact chlorine dose is adapted in some way to the sample's chlorine demand. Note that the Uniform Formation Test (Summers, 1993) is a type of LDFP. Tests that fall into this group have an inherent bias toward higher precursor recoveries for more highly colored waters. On the other hand, if carefully run, these tests can provide more accurate information for assessing the expected DBP concentrations at consumer's taps. The third type of test, the SDS, is characterized by a single attribute: it is designed to simulate the formation in a particular system on a particular day. This test uses a site-specific chlorine dose, pH, temperature and contact time. The values chosen are either based on an existing system on a particular day; or they are based on a very specific scenario intended to simulate a postulated system. Each of these three types of tests has its own appropriate use. The HDFP is most useful for the assessment of process performance. It can tell you what level of precursor removal is being achieved. The LDFP is most appropriate for comparing finished waters from parallel and alternative treatment trains. It is generally used when the precise disinfection scenario or distribution system characteristics are uncertain. It is not intended for assessing precursor removal across processes, just comparative precursor levels in the finished waters. The SDS is what should be used when the most accurate information about compliance and real-world concentrations are needed. The HDFP is the easiest test to run, and it is the most precise. This is because it is nearly independent of chlorine dose, so errors in dosing or excessive demands will introduce very little error. The LDFP and SDS tests are more labor-intensive, and they are prone to larger uncertainties. In summary, the three types of precursor tests should be used as follows: SYMBOL 183 \f "Symbol" \s 10 \h HDFP- for studies of isolated process performance; for understanding the behavior of complex treatment systems; for extrapolating findings at one utility to systems elsewhere in the country SYMBOL 183 \f "Symbol" \s 10 \h LDFP- for comparisons of parallel treatment trains at a single pilot plant. SYMBOL 183 \f "Symbol" \s 10 \h SDS- for estimating whether a new or midified treatment system will achieve compliance. 2. Instantaneous DBPs, Terminal DBPs, and Available Precursors The concentration of trihalomethanes or any other DBP in a sample of water taken from a process stream in a treatment plant or from a tap in a distribution system is referred to as simply the THM concentration or the instantaneous-THM concentration. This is the concentration that existed at the time of sampling. In many cases it is important to know how much effective DBP precursor organics are left in the sample. For this one typically holds the sample without quenching the residual oxidant, and measures the DBP concentration at some later date. The results of this measurement are called the terminal concentration. The exact conditions used (e.g., holding time, pH, additonal chlorine dose, temperature) will depend on the specifics of the study (refer to section 2.a.1). The difference between the terminal concentration and the instantaneous concentration is then the available precursor content. This is the amount of DBP precursors, present at the time of sampling, that will react to form DBPs under the test conditions used.  EMBED JandelGraphicObject.1  Chloramination Scenarios: pre-chlorine pre-ammoniation simultaneous addition pre-formation D. BIODEGRADABILITY TESTS Biodegradable or assimilable organic carbon, depending on the analytical method employed, is defined as either the fraction of dissolved organic carbon (DOC) that can be used by bacteria for growth and cell maintenance or the degree to which microbial growth is stimulated by this DOC. Because of the complexity of natural organic matter, it is impossible to identify the entirety of the DOC. The characterization of biological availability of dissolved organic matter requires the either the use of bioassay techniques or possibly the use of chemical surrogates. In heavily polluted waters (e.g., municipal wastewater), biochemical oxygen demand (BOD) or chemical oxygen demand (COD) tests are traditionally used for assessing the effectiveness of biological treatment. However, these methods are not sufficiently sensitive for use with treated drinking waters (Rittmann & Huck, 1989). In the past decade, several methods have been proposed to determine the easily assimilable organic carbon (AOC) or the biodegradable dissolved organic carbon (BDOC) in drinking waters. Several of these methods have been review recently by Huck et al. (1990). 1. AOC tests The first and still the most widely-used method for measuring assimilable organic carbon was developed by van der Kooij et al. (1982). It requires one to follow the growth of a pure culture by a plate count on gelose. The bacteria are inoculated in the water sample, after heating to 60`C. The bacteria commonly used are Pseudomonas fluorescens P17 and Spirillum NOX. The maximum plate count generally achieved after 4 to 8 days of incubation (15`C) is related to the maximum growth of a strain on a specific substrate. It is converted into units of acetate equivalents or oxalate equivalents by referring to a calibration test on a pure acetate or oxalate standard (van der Kooij et al., 1982, Van der Kooij, 1987). van der Kooij's pure culture approach suffers from several disadvantages. Pure culture of micro-organisms has a more limited capacity for biodegration than a heterogenous bacteria population. Maximum growth counts may be influenced by the physiological state of the Pseudomonas fluorescens culture. The use of pure cultures requires special techniques and highly skilled analysts. The AOC method as it is now commonly used, may also be subject to certain types of bias and large random errors (e.g., LeChevallier et al., 1993). For example, Prvost and co-workers (1992) have noted that difficulties in determining the maximum cell count can lead to significant errors, especially at high AOC concentrations. Possible growth inhibition or delayed growth caused by ozone byproducts and aluminum coagulants has also been reported (Huck et al., 1990). Perhaps these or other factors are responsible for some of the large anomalous increases in AOC-P17 across filtration as noted by Miltner et al. (1992). Nevertheless, the pure culture approach, as exemplified by van der Kooij's, should be subject to less interlaboratory variation, because the innoculum is always the same. Some laboratories, such as the Water Research Center in England, have modified the method of van der Kooij et al. (1982) by using an inoculum of an autochthonous bacterial population. They have also followed the growth rate through measurements of adenosine triphosphate (ATP) concentration, rather than using plate counts (Jago and Stanfield, 1984). Rice and co-workers (Reasoner & Rice, 1989; Rice et al., 1990) have developed another pure culture procedure that attempts to measure the ability of coliforms to grow. In this method, a pure culture of coliform (Enterobacter cloacae) is seeded and the ratio of log growth after 5 days (20SYMBOL 176 \f "Symbol"C) and time zero is determined. A higher ratio indicates a higher coliform growth response (CGR) and suggests that regrowth of coliforms may be possible in the distribution system. As with many of these alternative bioassays, the correlations with AOC and other tests are poor. 2. BDOC Tests The methods of van der Kooij et al. (1982), Jago and Stanfield (1984), and Rice (Reasoner & Rice, 1989; Rice et al., 1990) are based on bacterial growth. Other methods have been developed whereby the loss in organic substrate is measured. When the organic substrated is characterized by DOC measurements, the methods are termed biodegradable dissolved organic carbon (BDOC) assays. Servais and co-workers (Hascoet et al., 1986; Servais et al., 1987; Servais et al., 1989) developed a bioassay that incorporates some aspects of the van der Kooij and Werner methods. The water sample containing dissolved organic matter is sterilized by membrane filtration, inoculated with a second sample of the same water and incubated for 3 weeks at 20`C. After this time the reduction of the dissolved organic carbon (DOC) concentration is measured. In order to circumvent the long incubation times proscribed by Servais, Joret and co-workers (Joret & Levi, 1986; Joret et al., 1988) proposed that water samples could be tested in the presence of an inoculum of bacteria biomass attached to sand, a more active inoculum, that could reduce the response time to 3 or 4 days. 3. Chemical Surrogates A very different approach to assessing biodegradability is to use chemical surrogates for AOC or BDOC. As previously discussed, the most widely-used measure of biodegradability in the US is probably the Assimilable Organic Carbon (AOC) assay. Its chief disadvantages are its cost, labor requirements, and long analysis time, and lack of robustness. For these reasons it is not likely to be used as a routine water quality measurement, like dissolved organic carbon (DOC) or trihalomethane formation potential (LeChevallier et al., 1990). Several researchers have examined alternative chemical assays in order to avoid some of the drawbacks of AOC analysis. Reckhow et al. (1992, 1993) suggested that aldehydes or keto-acids could serve in this capacity. The use of aldehydes has also been supported by Zhou et al. (1992) and Krasner et al. (1993). Both groups of compounds can be measured by gas chromatography following sample extraction and chemical derivatization. Analyses are reproducible, and fast; usually taking only a few hours from start to finish. The Keto-Acids, and probably the aldehydes are be components of the AOC-NOX, but not the AOC-P17. This is an important characteristic for a proposed surrogate, because these two strains behave very differently. It has been observed that when the two strains are added together, NOX growth is most closely aligned with the concentration of biodegradable organic matter produced by ozonation (e.g., LeChevallier et al., 1992; Shukairy et al., 1992). In contrast, P17 either cannot utilize the ozone byproducts (e.g., oxalate) or it grows more slowly on them, so that its growth more closely reflects the level of naturally-occurring biodegradable organic compounds. Since the precise relationship between AOC levels in finished water and problems related to bacterial regrowth in the distribution system is not known, the desirable level of AOC in drinking waters remains speculative. Many engineers have chosen as a treatment goal the reduction of post-ozonation AOC levels to their pre-treatment levels. This often means that the AOC-NOX produced by ozonation, must be removed by subsequent biological filtration. Therefore, a chemical surrogate specific for AOC-NOX could be especially useful. Studies conducted by Reckhow et al. (1993) have shown that the keto-acids represent only the most biodegradable of the AOC compounds. Furthermore, there is a fraction of the AOC (even as fraction of the AOC-NOX) that is not readily degraded through water treatment. Similar observations have been made for some of the low molecular weight aldehydes (Miltner et al., 1992; Krasner et al., 1993). analysis of oxidant residuals General Principles Titrimetric methods Redox titrations are based on oxidation-reduction reactions between the analyte and the titrant or some intermediate redox carrier. Common oxidants used in redox titrations include dichromate (Cr2O7-2), iodate (IO3-), iodine (I2), and permanganate (MnO4-). Common reducing agents are arsenite (AsO3-3), ferrocyanide (Fe(CN)6-4), ferrous (Fe+2), sulfite (SO3-2) and thiosulfate (S2O3-2). Although many of the oxidizing agents are relatively stable, the reducing agents are often susceptible to oxidation by atmospheric oxygen, and therefore their titer must be checked regularly against a standard. Many redox titrations utilize iodine. When a reducing analyte is added to I2 to form I-, the process is called Iodimetry. Instead, when an ozidizing analyte converts I- to I2, it is Iodometry. In the presence of iodide (I-), iodine (I2) will form a triiodide complex (I3-) which greatly enhances its solubility in water. For the determination of triiodide, sodium thiosulfate is almost always used as the titrant. Although triiodide can be self-indicating, starch is generally used as an end point indicator. It forms an intense blue color with triiodide which can increase the sensitivity of endpoint detection by a factor of ten. It is important not to add starch until just before the endpoint. This allows one to see a more more gradual color change in the early part of the titration, and it avoids problems with excessively strong triiodide binding by the starch and over shoot. Thiosulfate reacts with triiodide rapidly under neutral or acidic conditions to give iodide and titrathionate. Commercial thiosulfate is not of sufficient purity to be a primary standard. Instead, it must be titrated against a standard triiodide solution prepared by reaction of iodide with some primary standard oxidant (e.g., KIO3). Thiosulfate is also readily oxidized by atmospheric oxygen at neutral to acidic conditions. It must be stored in a slightly alkaline buffered solution; often carbonate is used for this purpose.  EMBED Equation.2  Eq  STYLEREF 1 \n 3- SEQ Eq \* ARABIC \r 1 1 Chlorine There have been many methods proposed for the measurement of free and combined chlorine. Some of these include the iodometric methods, the DPD methods, the FACTS method, chemiluminescence method, LCV method, ultraviolet absorbance, membrane electrode methods and amperometric methods ( QUOTE "{Gordon, Cooper, et al. 1988 #3100}" ADDIN  ADDIN  {Gordon, Cooper, et al. 1988 #3100};  QUOTE "{Gordon, Cooper, et al. 1992 #970}" ADDIN  ADDIN  {Gordon, Cooper, et al. 1992 #970}). DPD Titrimetric Method Either chlorine or triiodide react with N,N-Diethyl-p-phenylene diamine (DPD) to form a relativly stable free radical species with an intense red color. This is then back titrated to the original colorless form with ferrous iron. The detection limit is about 18 ug/L. Oxidized manganese species will interfere. The DPD methods are the most widely used of the chlorine residual procedures (Gordon et al., 1988). Free residual chlorine is measured first via direct reaction with DPD. Monochloramine will react slowly with DPD through a direct pathway at a rate of about 5% per minute depending on concentration. For this reason mercuric chloride is added (HgCl2). It apparently inhibits the reaction between monochloramine and DPD. Although the reaction mechanisms isn't known, it is presumed to act by formation of an unreactive complex with monochloramine. Combined residuals are measured after addition of iodide. Monochloramine will react quickly with trace quantities of iodide to form triiodide which then reacts with the DPD. For dichloramine, the reaction is much slower, and relatively large amounts of iodide are needed for the reaction to go to completion. Both hydrogen peroxide and persulfate will also oxidize iodide, and can therefore interfere with the combined residual chloring determination. The DPD reagent is also subject to base catalyzed oxidation by atmospheric oxygen. For this reason it is kept in an acidified state, and replaced every month. Since the reaction with chlorine and triiodide is best when carrier out at neutral pH, a neutral buffer is used and the DPD must be stored separately from the buffer and added at the last minute.  Figure  STYLEREF 1 \n 3- SEQ Figure \* ARABIC \r 1 1 DPD Titrimetric Determination of Chlorine Procedure 1. Place 5 ml of both the buffer reagent and the DPD indicator solution in the titration flask and mix. 2. Add 100 ml sample and mix. 3. Free Residual Chlorine (FRC): Titrate rapidly with standard ferrous ammonium sulfate (FAS) titrant until the red color disappears (Reading A). 4. Monochloramine (MCA): Add one very small crystal of Potassium Iodide (KI) to solution from step 3 and mix. Continue titration until the red color again disappears (Reading B). 5. Dichloramine (DCA): Add several crystals of KI (about 1 g) to the solution titrated in step 4 and mix to dissolve. Allow to stand for 2 minutes and then continue titration until the red color is again discharged (Reading C). For very high dichloramine concentrations, allow an additional 2 minutes standing time if color driftback indicates incomplete reaction. When dichloramine concentrations are not expected to be high, use half the specified amount of potassium iodide. 6. CALCULATIONS The various forms of chlorine residual are best calculated according to the following scheme. Note that for a 100 ml sample, 4.00 ml standard FAS titrant = 1.00 mg/l residual chlorine. SpeciesFormulaHOCl + OClA/4 NH2Cl(B-A)/4NHCl2(C-B)/4 Ozone There have been many methods proposed for aqueous and gas-phase measurement of ozone. Some of these include the iodometric methods, arsenic direct oxidation, FACTS method, indigo trisulfonate, o-tolidine, carmine indigo, ultraviolet absorbance, and amperometric methods ( QUOTE "{Gordon, Cooper, et al. 1988 #3100}" ADDIN  ADDIN  {Gordon, Cooper, et al. 1988 #3100};  QUOTE "{Gordon, Cooper, et al. 1992 #970}" ADDIN  ADDIN  {Gordon, Cooper, et al. 1992 #970}). Iodometric method: Gas Phase Gas-phase concentrations of ozone are most easily measured iodometrically. A portion of the gas stream is directed to a gas bubbler filled with 2% KI solution for an exact period of time. Ozone reacts stoichiometrically to form an equivalent amount of iodine. O3 + 2KI + H2O ---------> I2 + O2 + 2OH- + 2K+ (Eq  STYLEREF 1 \n 3- SEQ Eq \* ARABIC 2) The iodine formed is then titrated with sodium thiosulfate using starch as an indicator to accentuate the endpoint (APHA et al., 1985). Ozone is powerful oxidant that is widely used in Europe for the treatment of drinking water. Its use in the U.S. is less common, but it is a rapidly growing technology. Principal applications of ozone include disinfection, oxidation of Fe and Mn, oxidation of industrial pollutants, bleaching of color, and improvement of coagulation/filtration. a. Procedure 1. Bubble ozone gas through a gas washing bottle containing a convenient volume of BKI solution (usually 250-500 mL). Record the bubbling time and gas flow rate or settings. A sample of the original (time zero) BKI solution should be saved for determination of UV blank. For low-level measurements, it is recommended that the BKI solution be very slightly preozonated. This removes small amounts or reducing materials that are invariably present in commercial potassium iodide. 2. At the end of the ozone trapping period, disconnect the gas washing bottle, and pour a small sample (e.g., 10 mL) into a 50 mL beaker. 3. Measure absorbance of the ozonated BKI at 352 nm after 1 minute (Abss). If the absorbance is greater than 1.2, the sample must be diluted. The final absorbance value must then be corrected for this dilution. 4. Obtain a blank measurement by determining the absorbance at 352 nm of an aliquot of the same BKI solution that was present in the gas washing bottle at time zero (Absb). 5. Calculate concentration with equation #8. Obtain the calibration factor, SYMBOL 98 \f "Symbol"1, by the standardization procedure in "b". Ozone Concentration (mg/L as O3) = (Abss-Absb)/SYMBOL 98 \f "Symbol"1 (17.38) b. Standardization 1. Prepare a set of standard triiodide solutions. This is done by first cleaning a series of 100 mL volumetric flasks, and half filling them with the Acidic KI solution. Add aliquots ("V" mL) of the Standard Potassium Iodate Solution to each and fill to the mark with additional acidic KI solution. Each mL of iodate will be roughly equivalent to 1.5 mg/L ozone. 2. Measure the absorbance of each of the standard triiodide solutions at 352 nm. Plot absorbance vs equivalent ozone concentration. Abs = SYMBOL 98 \f "Symbol"1 (Equiv. Conc.) (17.39) where the equivalent concentration is given by: Equiv. Conc. (mg/L as O3) = MIO3 (100) (3Mole IO3) (48,Mole O3O3) (17.40) = 1,440 MIO3 V Note that MIO3 is the exact molar concentration of the Standard Potassium Iodate Solution (see equation 11). Try to work only in the linear range, or the range from 0-1.2 absorbance units. c. Reagents 1. BKI Reagent (0.1 M Boric Acid, 1% Potassium Iodide): Add 6.2 g H3BO3 and 10.0 g KI to 1 liter of distilled water. Stir to dissolve. 2. Standard Potassium Iodate Solution: Dry the primary standard at 120oC for 2 hours. Then, weigh out about 0.021 g of the dried material. Record the exact weight to 4 significant figures. Dissolve in distilled water in a 100 mL volumetric flask and fill to the mark. Calculate exact molar concentration: MIO3 = wt in grams / 21.402 (17.41) 3. Acidic KI Solution (1% potassium iodide in 0.1N acid): Add 5.6 mL of concentrated H2SO4 to a 1 liter volumetric flask. Slowly fill the flask about half-way with distilled water and stir. Then add 10 g KI, stir, and fill to the mark with distilled water. Strictly speaking, all of these iodometric methods are non-selective. That is, they measure a wide range of oxidants, not just ozone. For this reason, they should not be used to measure aqueous ozone concentrations. Because ozone is by far the major oxidant species produced by corona ozone generators, and because ozone is far more easily stripped from water, the measurement of ozone in the gas phase is not subject to significant problems with interferences. Thus, iodometric methods may be used in this case without reservation. Direct UV Absorption: Gas Phase Commercial gas phase ozone monitors are based on the direct measurement of ultraviolet absorbance. With bench-scale studies, it is often convenient to use a laboratory UV-Vis spectrophotometer equiped with a flow-through quartz cell (0.1-0.2 cm pathlength) as a substitute for a dedicated ozone gas monitor. The ozone concentration may be calculated based on Beer's Law and the Ideal Gas Law. EMBED Equation  (Eq  STYLEREF 1 \n 3- SEQ Eq \* ARABIC 3) where SYMBOL 97 \f "Symbol"o is the absorptivity in atm-1cm-1 of ozone at the wavelength of measurement, Ta is the absolute temperature of the gas being measured, P is the pressure in mm Hg of the gas, and L is the cell pathlength in cm. At 253.7 nm, the absorptivity of ozone in the gas phase (at 760 torr, 293oK or 20oC) is 134 atm-1cm-1 (Inn & Tanaka, 1959; Hearn, 1961; DeMore & Patapoff, 1976). There is quite a bit of fine structure to ozone's broad UV absorbance band in the gas phase. For this reason, narrow band widths are preferable. When expressed as mass per volume, the terms for pressure and temperature drop out and one gets equation 17.34a. EMBED Equation  (17.34a) which for a wavelength of 253.7 nm and a pathlength of 0.2 cm reduces to equation 17.34b. EMBED Equation  (17.34b) Use of equations 17.32 and 17.33 is more convenient in the laboratory. If necessary, one can convert back to percent-based concentrations by equation 17.34c. EMBED Equation  (17.34c) Commercial analyzers generally use a mercury lamp, having a strong band at 254 nm ( QUOTE "{Eltze 1996 #5780}" ADDIN  ADDIN  {Eltze 1996 #5780}). This allows the use of wavelength filter, rather than a more expensive monochromator. Both single-beam and double-beam designs are used. Direct UV Method: Aqueous Phase Aqueous ozone concentrations in pure (e.g., distilled) water may be conveniently determined by direct spectrophotometric measurement at 260 nm. CO3 (mg/L as O3) = 14.59*(Abs @260 nm) (17.42) Equation 12 is based on a molar absorptivity of 3290 M-1cm-1 (Hart et al., 1983). Unfortunately, most solutes will interfere at this wavelength, so with actual environmental samples another method must be used. Since the iodometric method is too nonspecific for aqueous determinations, the indigo method of Bader and Hoigne (1981) is recommended. Indigo Method This colorimetric procedure uses solutions of indigo trisulfonate (Bader & Hoigne, 1981). Ozone will stoichiometrically bleach this intense blue dye, and the loss in absorbance at 600 nm may be translated directly into an ozone concentration. The reaction product is relatively unreactive to further ozonation. The reaction is best carried out at low pH to minimize ozone decomposition, and preserve the 1:1 stoichiometry. Bader and Hoigne (1981) report a sensitivity factor or apparent absorptivity for indigo trisulfonate of 20,000 M-1cm-1. This is based on an aqueous ozone molar absorptivity of 2900 M-1cm-1. If one adopts the higher value reported by Hart et al. (i.e., 3290 M-1cm-1), the sensitivity factor for indigo trisulfonate becomes 22,700 M-1cm-1. This method is quite selective, however, it is subject to a few notable interferences. The presence of residual chlorine will cause a positive bias. Addition of 500 mg/L malonic acid to the Indigo Reagent solves this problem by out-competing the indigo for the chlorine. Oxidized manganese species will also result in bleaching of the indigo. Here it is recommended that duplicate samples be analyzed, one according to the standard procedure, and one following addition of glycine. The glycine selectively reduces residual ozone without affecting oxidized manganese species. The true ozone concentration may then be estimated from the difference of these two measurements. a. Procedure 1. Prepare an indigo blank by adding 1 mL of the Standard Indigo Stock to a 25 mL volumetric flask and filling to the mark with Phosphate Buffer. Stopper and mix. Measure the absorbance of this solution at 600 nm (Absi). When the Indigo Stock is new, it should be about 0.650. With time this value will drop. When it falls below 80% of the original value, prepare a new Indigo Stock and repeat procedure. If low ozone concentrations are anticipated (i.e., < 0.3 mg/L) prepare a Secondary Indigo Stock by adding 20 mL of the Standard Indigo Stock to a 100 mL volumetric flask and diluting to the mark with Super-Q water. 2. Soak a series of volumetric pipets in a dilute ozone solution for several minutes. These pipets are to be used to transfer the solution to be measured to the indigo-containing flasks. They must therefore be rendered ozone-demand-free. The capacities of the pipets will depend on the range of anticipated ozone concentrations (see table below). 3. Assemble a series of clean, glass-stoppered 25 mL or 50 mL volumetric flasks (see table below) and fill each with 1.00 mL of Standard Indigo Stock (25 mL flasks) or Secondary Indigo Stock (50 mL flasks) using a volumetric pipet. Wash this down from the inner surfaces of the neck with about 10 mL of Phosphate Buffer. 4. Quickly pipet the recommended sample volume to an indigo-containing volumetric flask (see table below). Be sure that the tip of the pipet is below the meniscus. Fill the flask to the mark with Phosphate Buffer, cap and invert several times to mix. (Vs)(Vt)(L)Anticipated OzoneRecommendedRecommendedRecommended Concentration Sample Volume Total Volume Pathlength0 - 0.2 mg/L25 mL50 mL *10 cm 0.1 - 0.3 mg/L15 mL50 mL *10 cm0.2 - 0.5 mg/L10 mL50 mL *10 cm0.3 - 2.0 mg/L15 mL25 mL1 cm1.5 - 3.0 mg/L10 mL25 mL1 cm1.8 - 3.5 mg/L8 or 9 mL25 mL1 cm2.3 - 4.5 mg/L6 or 7 mL25 mL1 cm3 - 6 mg/L5 mL25 mL1 cm4 - 7 mg/L4 mL25 mL1 cm5 - 10 mg/L3 mL25 mL1 cm7 - 15 mg/L2 mL25 mL1 cm15 - 30 mg/L1 mL25 mL1 cm *Secondary Indigo Stock must be used with 10 cm pathlength cells 5. Measure absorbance (Absf) of each sample at 600 nm using cells of the indicated pathlength (L). Concentration is calculated from a slope or calibration factor determined by calibration against the direct UV method (see b. Calibration). Ozone Conc. (mg/L as O3) = Vt(Absi Absf)/b1VsL (17.43) b. Calibration 1. Prepare a series of aqueous ozone standards in slightly acidified (0.1 mM HNO3) Super-Q water. This is generally done by first bubbling ozone gas through about 500-1000 mL of the acidified water until near saturation (~1 hour). Under optimal conditions for high ozone output (i.e., high voltage, low gas flow, low temperature) this should result in aqueous concentrations of about 10 mg/L. Then, aliquots of this water are removed and diluted with varying amounts of un-ozonated, acidified Super-Q water. The degree of dilution will depend on the range of ozone concentrations anticipated in the samples of interest. 2. One-by-one measure each of these diluted solutions by the indigo method above and by the direct UV method (equation 17.42). Use the same sample volume, Vs, for all standards. 3. Plot absorbance (from indigo method) versus aqueous ozone concentration (from direct UV method). The slope of this line multiplied by Vt/Vs gives b1L (see equation 17.44). Based on the presumed sensitivity factor for indigo trisulfonate of 22,700 M-1cm-1, the calibration factor, b1, should be about 0.47 abs/cm per mg-O3/L. Absf = Absi - (b1LVs/Vt) (Ozone Conc.) (17.44) c. Reagents 1. Standard Indigo Stock (1 mM in 20 mM phosphoric acid): Dissolve 1.36 mL conc. H3PO4 in 1 liter of super-Q water and mix. To this add 0.6 g indigo trisulfonate, mix and store in a brown glass bottle. 2. Phosphate Buffer (pH 2): Dissolve 28 g NaH2PO4.H2O and 20.6 mL (35 g) conc. H3PO4 in Super-Q water and dilute to 1 liter. Chloramines Chlorine Dioxide from; The Chemists Companion: A Handbook of Practical Data, Techniques and References. A.J. Gordon & R.A. Ford, J. Wiley & Sons Publ., New York, 1972. The old mercury vapor lamps, once common in spectrophotometers, had an especially high intensity at 254 nm. 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R|_918820083F+(4+(4Ole TCompObjUfObjInfoWEquation Native Xh_918820174F(4`(4Ole ^LlJ qJlJ 1.29x10 6 ={O 2(aq) } 0.5 {MnO 2(s) } 0.5 {H + }{O 3(aq) } 0.5 {Mn +2 } 0.5 {H 2 O} 0.5dJ FMicrosoft Equation 2.0 DS EqCompObj_fObjInfoaEquation Native b_920201944;F@(4@(4uation Equation.29qlJ qJlJ 1.29x10 6 ={O 2(aq) } 0.5 {H + }{O 3(aq) } 0.5 {Mn +2 } 0.5Ole gCompObjhfObjInfojEquation Native k FMicrosoft Equation 2.0 DS Equation Equation.29qӬlJ$qJlJ E cello =E oxo +E redo =(-1.68V)+(+2.037V)=+0.36V_920202054F (4@O(4Ole oCompObjpfObjInfor FMicrosoft Equation 2.0 DS Equation Equation.29qӸlJ$qJlJ logK=10.059E cello =10.059(+0.36V)=6.1K=10 6.1Equation Native s_920204131F (4 (4Ole wCompObjxf FMicrosoft Equation 2.0 DS Equation Equation.29qӸlJ$qJlJ K={A red } a {B ox } b {A ox } a {B red } b K=MnO 4- [] 0.33 H + [] ObjInfozEquation Native {_920204148F((4((4Ole 0.33 O 2 [] 0.5 O 3 [] 0.5 =10 6.1 FMicrosoft Equation 2.0 DS Equation Equation.29qӸlJ$qJlJ K={A rCompObjfObjInfoEquation Native 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