RESTORATION AND MANAGEMENT OF LAKES AND RESERVOIRS - CHAPTER 8 pot

8 Phosphorus Inactivation and Sediment Oxidation 8.1 INTRODUCTION Nuisance algal blooms can be reduced or eliminated if phosphorus (P) concentrations are lowered to growth-limiting levels by diversion of external loading, by dilution, or a combination of these methods. In cases where loading reduction is significant, where the lake flushing rate is relatively fast, and where recycling from sediments is unimportant, in-lake P can be reduced and trophic state significantly and rapidly improved. The case of Lake Washington (Edmondson, 1970; 1994) is probably the most recognized example of this response (Chapter 4). For many lakes, however, internal P release prolonged the lake’s enriched state and supported continued algal blooms, even though diversion removed a significant fraction of external loading (Cullen and Forsberg, 1988; Sas et al., 1989; Jeppesen et al., 1991; Welch and Cooke, 1995; Scheffer, 1998). Lakes that experience significant internal loading of P to their water columns are the rule rather than the exception. Lakes with extensive littoral and wetland areas (Wetzel, 1990), close proximity between the epilimnion and anoxic sediments (Fee, 1979), or shallow lakes with enriched sediments from a history of high external loading (Jeppesen et al., 1991), will have extensive P recycling. In those lakes, additional in-lake steps may be necessary, following nutrient diversion, to prevent a prolonged eutrophic state. For example, Shagawa Lake, Minnesota (MN), did not respond as rapidly as expected to the reduction of a large fraction of external loading, and it is predicted to require decades to reach equilibrium (Larsen et al., 1981; Chapra and Canale, 1991; Chapter 4). In some lakes, even without reduction of external loading the major input of P, and cause for summer algal blooms, is from sediments (Welch and Jacoby, 2001). In-lake treatments have been effective in such lakes. Phosphorus inactivation is an in-lake technique, designed to lower the lake’s P content by removal of P from the water column (P precipitation) and by retarding release of mobile P from lake sediments (P inactivation). Usually an aluminum salt, either aluminum sulfate (alum), sodium aluminate, or both, is added to the water column to form aluminum phosphate and a colloidal aluminum hydroxide floc to which certain P fractions are bound. The aluminum hydroxide floc settles to the sediment and continues to sorb and retain P within the lattice of the molecule, even under reducing conditions. Alum has been used for coagulation in water treatment for over 200 years and is probably the most commonly used drinking water treatment in the world (Ødegaard et al., 1990). Polyaluminum chloride is another coagulant used in water treatment that has a more favorable floc-forming pH range than alum (Ødegaard et al., 1990), and has been used in lakes (Carlson, personal communi- cation). Iron and calcium salts have also been used to precipitate or sorb P. Summer lake trophic state is improved when the control of internal P release significantly lowers P concentration in the photic zone, which is the whole water column of polymictic lakes, and the epilimnion and sometimes the metalimnion of eutrophic, dimictic lakes and reservoirs. This technique has been mistakenly classified as an algicide or herbicide by some agencies. Phosphorus inactivation provides long-term control of algal biomass by significantly reducing the supply of an essential nutrient rather than through poisoning of algal cells. Algicides work by direct toxic action, and are effective only during the brief period when the toxic active ingredient Copyright © 2005 by Taylor & Francis (frequently copper) is present in the water column (Chapter 10). Phosphorus inactivation with aluminum salts is effective for years while algicides are effective for days. A different method to control internal P loading from anaerobic lake sediments, called sediment oxidation, was developed by Ripl (1976). With this procedure, Ca(NO 3 ) 2 is injected into lake sediments to stimulate denitrification, where nitrate acts as an electron acceptor. This process oxidizes the organic matter. At the same time, ferric chloride is added, if natural levels are low, to remove H 2 S and to form Fe(OH) 3 to which P is sorbed. 8.2 CHEMICAL BACKGROUND Aluminum, iron, and calcium salts have been used for centuries for drinking water clarification, and their use today, particularly aluminum, is essential in the treatment of wastewater and drinking water. Lund (1955) appears to be the first to suggest that the addition of aluminum sulfate (alum, Al 2 (SO 4 ) 3 . 14 H 2 O) to streams and lakes could be a successful means to control algal blooms. The first published account of such a treatment is Jernelöv (1971), who applied dry alum to the ice of Lake Långsjön, Sweden, in 1968. Iron and calcium are major controllers of the P cycle in lakes, and like aluminum, have been used extensively in wastewater and potable supply treatments, but less frequently than alum in lakes. The first report for iron in lakes to control P was in Dordrecht Reservoir, The Netherlands (Peelen, 1969) and for calcium in a Canadian hard water lake (Murphy et al., 1988). 8.2.1 ALUMINUM The chemistry of aluminum is complex and incompletely understood (Dentel and Gossett, 1988; Bertsch, 1989). The reactions in water have been reviewed by Burrows (1977), Driscoll and Letterman (1988), and Driscoll and Schecher (1990), among others. The following is drawn from these reports, and from the first detailed lake and laboratory studies of aluminum salts for P inactivation (Browman et al., 1977; Eisenreich et al., 1977). When aluminum sulfate or other aluminum salts are added to water they dissociate, forming aluminum ions. These are immediately hydrated: Al +3 + 6 H 2 O (8.1) A progressive series of hydrolysis (the liberation of hydrogen ions) reactions occurs leading to the formation of aluminum hydroxide, Al(OH) 3 , a colloidal, amorphous floc with high coagulation and P adsorption properties: Al(H 2 O) 5 OH 2+ + H 3 O (8.2) (8.3) etc. Omitting coordinating water molecules from the equations, the following occurs: (8.4)  Al (H O) 26 3+ Al(H O) H O 26 3+ 2 +  Al(H O) OH + H O Al(H O) OH + H O 25 2+ 2242 + 3  Al H O Al OH H 3 2 2+++ ++ () Copyright © 2005 by Taylor & Francis (8.5) (8.6) where (s) = a solid precipitate. Al(OH) 3 is a visible precipitate or floc that settles through the lake’s water column to the sediments. A surface application produces a milky solution, which quickly forms large, visible particles. The floc grows in size and weight as settling occurs and particles within the water column are incorporated. Within hours, water transparency increases dramatically. The pH of the solution determines which aluminum hydrolysis products dominate and what their solubilities will be (Figure 8.1). At the pH of most lake waters (pH 6 to 8), insoluble polymeric Al(OH) 3 dominates and P sorption and inactivation proceeds. At pH 4 to 6, various soluble intermediate forms occur, and at pH less than 4, hydrated and soluble Al 3+ dominate. When alum is added to poorly buffered waters, their acid neutralizing capacity (ANC) decreases, pH falls, and soluble aluminum species dominate if ANC is exhausted. At higher pH levels (>8.0), the amphoteric nature (having both acidic and basic properties) of aluminum hydroxide results in the formation of the aluminate ion: Al(OH) 3 + H 2 O ↔ Al(OH) 4 ¯ + H + (8.7) At increasing pH levels above 8, as would occur during intense photosynthesis for example, solubility again increases, which could lead to a release of P sorbed to an aluminum salt. Aluminum salts in water have a time-dependent component to their chemistry (Burrows, 1977). The concentration of monomeric forms (Al 3+ , Al(OH) 2+ , and stabilizes within 24 h. But crystallization takes over a year to complete as larger and larger units of polymeric Al(OH) 3 are formed. In lakes, this continued reaction occurs in the sediments, though its consequences to the control of P release are poorly understood. Toxicity studies carried out with a freshly prepared solution of buffered aluminum present a different array of potentially toxic aluminum species than an aged solution with a lower concentration of monomeric species and intermediate polymers (Burrows, 1977). This also may be the reason why continuous exposure to the early hydrolysis products of alum, as would occur in a continuous addition to flowing waters, may be deleterious to biota versus the single treatment to lake sediments (Barbiero et al., 1988). Exposure time to the floc is shorter in a lake treatment due to its relatively quick transport to the bottom. Properties of Al(OH) 3 of greatest interest to lake managers are its apparent low or zero toxicity to lake biota (see later section), its ability to adsorb large amounts of particulate and soluble P, and FIGURE 8.1 Fractional distribution of aluminum species as a function of pH (concentration 5.0 × 10 –4 M). (Courtesy C. Lind, General Chemical Inc., Parsippany, NJ. With permission.) 0.0 0.2 0.4 0.6 0.8 1.0 4567891011 Fraction Aluminum pH Al(OH) 3 (c) Al(OH) 4 - Al 3+ Al 8 (OH) 20 4+ AlOH 2+ Al OH H O Al OH H() () 2 22 +++ ++ Al OH H O Al OH s H() ()() 22 3 ++ ++ Al OH(), 2 + Al OH()) 4 − Copyright © 2005 by Taylor & Francis the binding of P to the floc. In contrast to iron, low or zero dissolved oxygen (DO) concentrations in lake sediments do not solubilize the floc and allow P release, although P may be released from the floc if high pH occurs. Particulate organic P (cells, detritus) is removed to some extent from the water column by coagulation and entrapment in the Al(OH) 3 floc. The settling of the floc through the water column clarifies the water in this manner. That is the reason alum is used extensively in water treatment plants. However, Al(OH) 3 may be less effective in removing dissolved organic matter (Browman et al., 1977). Treatment timing may vary depending on local conditions. Because Al(OH) 3 is so sorptive of inorganic P, a seemingly ideal time for treatment is just after ice-out, or in early spring in warmer climates, before the spring bloom of algal cells and corresponding uptake of P occurs. However, temperature alters the rate and extent of reactions of aluminum salts in water (Driscoll and Letterman, 1988). At low temperatures, coagulation and deposition are significantly reduced and high quantities of species such as Al(OH) 2 +, that are toxic to some organisms, might occur. This suggests that aluminum solubility is temperature dependent as well as pH dependent. There are other reasons why early spring may not be an ideal time, in spite of high inorganic P. These include (1) sediment P release, not water column P content, is the primary target of P inactivation, (2) early spring months may be windy, making application difficult, (3) wind mixing may distribute the floc to one area of the lake, or scour it from the sediments before the floc consolidates into those sediments, and (4) silicon content, a major complexer of soluble and possibly toxic aluminum species, may be low following a spring diatom bloom. Thus, summer, before blue-green algal blooms appear, or early fall months, may be the most appropriate periods for application. On the other hand, an early spring treatment may avoid the problem of macrophytes, in spite of other risks. Because hydrogen ions are liberated when an aluminum salt is added to water, H + increases in proportion to the decline in alkalinity. In lakes with low or moderate alkalinity (< 30 to 50 mg CaCO 3 /L), treatment produces a significant decline in pH (increase in H + ) at a low or moderate alum dose, leading to increasing concentrations of toxic, soluble aluminum forms, including Al(OH) 2 + and Al 3+ . This limits the amount of alum that can be added safely. This problem has been addressed by adding a buffer to the lake or to the alum slurry as it is applied. The work of Dominie (1980) for Lake Annabessacook, Maine, Smeltzer (1990) for Lake Morey, Vermont, and Jacoby et al. (1994) for Green Lake, Washington are examples. Buffering compounds were tested, including sodium hydroxide, calcium hydroxide, and sodium carbonate. The buffer chosen was sodium aluminate (Na 2 Al 2 O 4 ⋅ N H 2 O), a high alkalinity compound with the added benefit of having a high aluminum content (Smeltzer, 1990). Much of this compound’s alkalinity comes from the NaOH used in its production (Lind, personal communication). Sodium aluminate and alum should be added to the lake separately to avoid damage to pipes from overheating if mixed together. Sodium carbonate was also successfully used to buffer the treatment of soft water (35 mg/L alkalinity) Long Lake, Washington (Welch, 1996). A mixture of alum and lime has also served the purpose of buffering in soft waters (Babin et al., 1992). In summary, the primary objective of an in-lake alum treatment is to cover the sediment with Al(OH) 3 . Mobile P, which otherwise would diffuse into the water column, is sorbed, thereby reducing internal loading. The formation of Al(OH) 3 also removes particulate organic and inorganic matter with P from the water column, a secondary objective. The formation of large amounts of Al(OH) 3 and negligible amounts of other hydrolysis products depends upon maintaining water column pH between pH 6 and 8. Because lakes differ in alkalinity and sediment mobile-P content, the dose to a lake is lake specific. In some cases, a buffer must be added. Dose determination is discussed in a later section. 8.2.2 IRON AND CALCIUM Phosphorus forms precipitates and complexes with iron and calcium, and these elements can be used to lower P concentration with less concern for pH shifts and/or the appearance of toxic forms. Copyright © 2005 by Taylor & Francis The chemistry of these metals with regard to P is probably better understood than that of aluminum (Stumm and Lee, 1960; Stumm and Morgan, 1970). Inorganic iron exists in solution in lake water and lake sediments in either the oxidized ferric (Fe 3+ ) or reduced ferrous (Fe 2+ ) forms, depending on solution pH and oxidation-reduction potential. Changes in the redox state of iron in lake sediments has an important effect on the P cycle (Mortimer, 1941, 1971). In oxygenated, alkaline conditions, a common state of the entire water column during spring and fall mixing, the redox potential is high and iron is oxidized to the ferric form. (8.8) Fe(OH) 3 sorbs P from the water column, and forms part of an oxidized “microzone” over the sediment surface, providing high sediment P retention. FePO 4 also forms, but the primary means of P removal and retention in sediments is sorption to Fe(OH) 3 , and it is greatest at pH 5 to 7 (Andersen, 1975; Lijklema, 1977). The generally accepted cycle of Fe and P in lakes is as follows. During periods of thermal stratification in eutrophic lakes, hypolimnetic waters are dark and isolated from mixing for periods of days or weeks (polymictic lakes) to periods of months (dimictic lakes). Without net photosyn- thesis or aeration by mixing, pH and particularly dissolved oxygen (DO) concentrations decline in the water overlying the sediments. As DO in the overlying water drops below 1.0 mg/L, the oxidized microzone is eliminated, and iron is used by the microbial community as an alternate electron acceptor to oxygen. In the reduced state, ferrous iron (Fe 2+ ) is soluble, and previously iron-bound P (part of mobile P) is free to be released to the water column. This change occurs rapidly so that even brief periods of thermal stability in a shallow eutrophic lake with high sediment oxygen demand can lead to substantial P release. Thermal stability in shallow lakes is largely controlled by wind (see Chapter 4). Also, sediment P release may occur on a diurnal basis in littoral areas of some eutrophic lakes so that P is sorbed to ferric complexes during the day and released during the night (Carlton and Wetzel, 1988). In spite of persistent anoxia in dimictic stratified lakes, the magnitude of sediment P release rates are as great in shallow unstratified lakes and are related more to trophic state than to depth (Nürnberg, 1996). This generally accepted Fe–P redox cycle may not always hold. Sulfate reduction and the formation of insoluble FeS can remove iron from the cycle, effectively decreasing the Fe:P ratio, producing a greater fraction of P that remains soluble and can be released to overlying water (Smolders and Roelofs, 1993; Søndergaard et al., 2002). However, work by Caraco et al. (1989) seems to contradict the effect of S. In a sample of 23 lakes, only those with an intermediate (100 to 300 μm) sulfate concentration conformed to the iron redox model. Low sulfate (60 μm) systems had low P release under both oxic and anoxic conditions, and high (> 3000 μm) sulfate systems had high P release under both conditions. There are many lakes with low sediment P release under anoxic conditions (Caraco et al., 1991a, b). Phosphorus may be released when OH – is exchanged for PO 4 3– on the iron-hydroxy complex during periods of high pH, even under aerobic conditions (Andersen, 1975; Jacoby et al., 1982; Boers, 1991a; Jensen and Andersen, 1992). This process enhances P release during resuspension events, especially if the suspended particle concentration is relatively low (Koski-Vähälä and Hartikainen, 2001; Van Hullenbusch et al., 2003). That is, the equilibrium shifts from desorption of P from particles to sorption by particles as particle concentration increases. Iron’s reaction to redox and pH conditions means that its addition to lakes as a P inactivant may have to be accompanied by a technique (aeration or artificial circulation) to prevent breakdown of the oxidized microzone or a photosynthetically caused increase in pH. Even aeration may not reduce P release if the sediments have a low Fe:S ratio (Caraco et al., 1991a, b). There is some evidence, however, that iron enrichment of lake sediments may inhibit P release even under anoxic conditions (Quaak et al., 1993; Boers et al., 1994). Fe O OH H O Fe OH s 2 223 14 2 12+++ → − //()() Copyright © 2005 by Taylor & Francis Calcium compounds also affect P concentration. Calcium carbonate (calcite) and calcium hydroxide can be added to a lake from allochthonous sources, or produced in hard water lakes during periods of CO 2 uptake during photosynthesis, as follows: (8.9) As plants assimilate CO 2 , pH increases and CaCO 3 precipitates. Calcite sorbs P, especially when pH exceeds 9.0 (Koschel et al., 1983), and results in significant P removal from the water column (Gardner and Eadie, 1980). At high levels of pH, Ca 2+ , and P, hydroxyapatite forms, as follows: (8.10) Hydroxyapatite, unlike Fe(OH) 3 and Al(OH) 3 , has its lowest solubility at pH >9.5, and P sorbs strongly to it at high pH (Andersen, 1974, 1975). The solubility of calcite and hydroxyapatite increases sharply as CO 2 concentration increases and pH falls, as would be expected in a hypolim- nion or dark littoral zone with intense respiration. This will lead to P release. Thus, as with iron, effective P removal and inactivation is possible with calcium, but conditions conducive to continued P sorption can be lost unless an additional management step is taken to maintain an alkaline pH in deep water. Phosphorus inactivation, by definition, is an attempt to permanently and extensively bind P in lake sediments and thereby lower or essentially eliminate sediments as a P source to the water column. Phosphorus is strongly sorbed to Al(OH) 3 and this complex is apparently inert to redox changes, thus providing the possibility of a high degree of treatment permanence. However, the addition of aluminum salts to lakewater produces H + ions, and pH falls at a rate dictated by lake water alkalinity and dose of the salt. This can lead to high concentrations of soluble and potentially toxic aluminum species. Thus, unless the lake is well buffered, or buffers are added, the use of aluminum salts may not be appropriate. Sorption of P to iron and calcium complexes can also lead to significant P removal and to P retention in the sediments, but without toxicity problems. However, solubility of these compounds, and hence P sorption, is highly sensitive to pH and redox changes. Anoxia can occur very rapidly in productive lakes, even in shallow water sediments. Aeration or complete mixing would be needed on a continual basis if iron or calcium are employed for P inactivation. 8.3 DOSE DETERMINATION AND APPLICATION TECHNIQUES 8.3.1 A LUMINUM There have been two approaches to the use of metal salts to control P concentration in lakes. Phosphorus precipitation emphasizes P removal from the water column, while P inactivation emphasizes longer-term control of sediment P release with P removal a secondary objective. During early Al applications, no basis for dose using either procedure existed (see Cooke and Kennedy, 1981 for a review and tabular summary of the first 28 treatments). Phosphorus removal or precipitation is achieved by adding enough Al to the lake surface to remove the P in the water column at the time of application. Dose is determined by adding increments of Al 2 (SO 4 ) 3 to lake water samples until the desired removal of P is achieved. This dose is then used to calculate the amount needed to remove P from the entire lake. Small amounts of Al are usually needed to bring about P removal. However, the goal of nearly all alum treatments in recent years is long-term control of internal loading and that will not be achieved with low doses. Ca HCO CaCO s H O CO() () 32 3 2 2  ++ 10 6 2 1 34 2 210462 CaCO HPO H O Ca PO OH s++ + −  ()()()00 3 HCO − Copyright © 2005 by Taylor & Francis Low dose and continued high external loading are reasons why some of the first treatments were short-lived (e.g., Långsjön). Also, some P fractions, particularly the dissolved organic fraction, may be incompletely removed, leaving a substrate for algal assimilation and growth. Thus, P precipitation to control algae is usually not recommended. There are three procedures for determining dose to inactivate sediment P. The first is the alkalinity procedure developed by Kennedy (1978) and it has had widespread use. Kennedy believed that P inactivation should provide the greatest control possible. Treatment duration was assumed to be related to Al(OH) 3 concentration in lake sediments, because the amount of mobile P bound as Al–P is proportional to Al added. The goal then was to apply as much Al to the sediments as possible, consistent with environmental safety. As described in an earlier section, the form of Al in water is dictated by pH (Figure 8.1). Between pH 6 and 8, most is in the solid Al(OH) 3 form. As pH falls below pH 6.0, other forms, including Al(OH) 2 + and Al 3+ become increasingly important. These forms are toxic in varying degrees, particularly dissolved Al 3+ (Burrows, 1977). Everhart and Freeman (1973) found that rainbow trout (Salmo gairdneri) could tolerate chronic exposure to 52 μg/L dissolved Al with no obvious changes in behavior or physiological activity. The observation led to adopting 50 μg Al/L as a safe upper limit for post treatment dissolved Al concentration (Kennedy, 1978). Maximum dose was thus defined as the maximum amount of Al that, when added to lake water, would ensure that dissolved Al concentration is less than 50 μg/L (Kennedy and Cooke, 1982). As shown in Figure 8.1, 50 μg Al 3+ /L should not occur as long as pH remains between pH 6.0 and 8.0. Maintaining pH ≥ 6.0 should prevent toxicity from dissolved monomeric Al as judged from experiments with alum applications to lake inflows (Pilgrim and Brezonik, 2004). There is an added safety factor in that a lake treatment, unlike continuous exposure bioassay, produces a single maximum dose exposure to organisms, followed by a rapid decline in concen- tration, because the alum floc settles through the water column rather quickly (∼ 1 hour). Also, the maximum formation of Al(OH) 3 occurs in the 6–8 pH range, leading to maximum deposition and removal of most Al from the water column, and to maximum formation of the P-retaining floc at the sediment-water interface. Dissolved Al concentrations have remained at 100–200 μg/L following treatments of some Washington lakes, without adverse effects, so the Al was probably in an organic, non-toxic form (Welch, 1996). Addition of natural organic matter was shown to reduce toxicity of Al by a factor of two (Roy and Campbell, 1997). The alkalinity procedure (Kennedy and Cooke, 1980, 1982) provided a toxicological basis for adding enough alum to provide long-term control of sediment P release in lakes with adequate alkalinity (> 35 mg/L CaCO 3 ). There were several problems inherent in developing this procedure. First, few toxicity studies (see later paragraphs) had been conducted on the effects of Al on lake community processes and structure in non-acidified lakes. However, trout are highly sensitive to metals and that may provide a safety factor for the community as a whole. Second, low pH itself can be detrimental. Adverse effects in acidified lakes have been observed to begin at pH ≤ 6.0 without Al added (Schindler, 1986). However, those were long-term chronic effects, while alum treatments produce only short-term reductions in pH. Third, some lakes have low alkalinity and only small amounts of alum could be added before pH 6.0 is reached. As noted earlier, this last problem is overcome by the use of buffers. The following step-by-step procedure, from Kennedy (1978) and Kennedy and Cooke (1982), describes dose determination based on lake water alkalinity. 1. Obtain water samples over the range of lake water alkalinities. Normally this means a series of samples from surface to bottom. Determine alkalinity to a pH 4.5 endpoint. 2. The dose for each stratum is approximated from Figure 8.2, which uses pH 6.0 as the endpoint of alum addition to the lake, rather than 50 μg Al/L. At the determined lake dose, dissolved Al will remain below this limit as long as the pH is between 6.0 and 8.0. Kennedy and Cooke (1982) selected this pH range to provide a safety margin with Copyright © 2005 by Taylor & Francis respect to dissolved Al and excessive hydrogen ion concentration. This allows the addition of sufficient Al to lakes to give long-term control of P release, when lake alkalinities are above about 35 mg CaCO 3 /L. A more accurate dose determination is made by titrating water samples with a stock solution of aluminum sulfate of known Al concentration (a solution that contains 1.25 mg Al/ml is made by dissolving 15.4211 g technical grade Al 2 (SO 4 ) 3 ⋅ 18H 2 O in distilled water and diluting to 1.0 liter). Adding 1.0 ml of this stock to a 500-ml water sample is a dose of 2.5 mg Al/L. As alum is added, the samples are mixed with a stirrer and pH changes are monitored. Optimum dose for each sample is the amount that produces a stable pH of 6.0. Linear regression is used to determine the relationship between dose and alkalinity. The resulting equation is then used to obtain dose for any alkalinity for this particular lake or reservoir, over the alkalinity range tested. To be cautious, a slightly higher pH (e.g., 6.2–6.3) may be chosen to determine dose (see Jacoby et al., 1994). The maximum dose for each depth interval from which the alkalinities are obtained is calculated by converting the dose in mg Al/L to lbs (dry) alum/m 3 , using a formula weight of 666.19 [Al 2 (SO) 4 ) 3 ·18 H 2 O] and a conversion factor of 0.02723 to change mg Al/L to lbs (dry) alum/m 3 , because English units are used with commercial alum. A conversion factor of 0.02428 is used for Al 2 (SO 4 ) 3 ⋅ 14H 2 O. 3. If liquid rather than granular alum is to be used, as is the usual case, further calculations are necessary to express the dose in gallons of alum/m 3 . Alum ranges from 8.0 to 8.5% Al 2 O 3 , which is equivalent to 5.16 to 5.57 pounds dry alum per gallon at 60°F (Lind, personal communication). It is shipped by tank truck at about 100°F and will thus have lower density. The percent Al 2 O 3 at 60°F will be stated by the shipper. Convert this to density, expressed as degrees Baumé, using Figure 8.3. Then obtain the shipment tem- perature and adjust the 60° Baumé number by subtracting the correction factor (using Figure 8.4) from the 60° Baumé number. Pounds per gallon is then obtained from Figure 8.5, using the adjusted Baumé number. FIGURE 8.2 Estimated aluminum sulfate dose (mg Al/L) required to obtain pH 6 in treated water of varying initial alkalinity and pH. (From Kennedy, R.H. and G.D. Cooke. 1982. Water Res. Bull. 18: 389–395. With permission.) Total alkalinity (mgCaCO 3 /ᐍ) 250 200 150 100 50 0 6.5 7.0 7.5 8.0 Intitial pH Aluminum dose (mgAl/ᐍ) to obtain pH 6.0 30 25 20 15 10 5 Copyright © 2005 by Taylor & Francis Maximum dose for each depth interval sampled for alkalinity was calculated earlier as pounds dry alum/m 3 . This is converted to gallons/m 3 by dividing pounds (dry)/m 3 by the value in pounds per gallon obtained from Figure 8.5. Total dose to the lake is then the sum of the individual depth interval doses. 4. Accuracy in treating the lake is obtained by dividing the lake into areas marked with buoys or, by using a barge equipped with a satellite guidance system. The volume and alkalinity in each area is measured and the gallons per treatment area determined. This FIGURE 8.3 Relationship of Baumé (60°F) and percent Al 2 O 3 . (From Cooke et al. 1978.) FIGURE 8.4 Temperature correction factors for 32 to 36° Baumé liquors. (From Cooke et al. 1978.) Be′ (60°F) 40 35 30 25 20 4.0 5.0 6.0 7.0 8.0 9.0 % Al 2 O 3 (Allied Chemical Corp.) Correction factor (F) 2.8 2.4 2.0 1.2 1.0 0.8 0.4 0 1601401201008060 (Allied Chemical Corp.) Copyright © 2005 by Taylor & Francis approach prevents under-dosing deep areas and overdosing shallow ones, as would occur if an even application were made to the whole lake or reservoir area. Distribution of the alum with respect to lake volume can be accomplished automatically with a barge equipped with an electronic sounding device. In soft-water lakes, only small amounts of aluminum sulfate can be added before the pH falls below 6.0. Gahler and Powers (personal communication) were perhaps the first to suggest that sodium aluminate, which supplies alkalinity and increases the pH of an aqueous solution, could be used with Al 2 (SO 4 ) 3 to maintain a pH between 6.0 and 8.0. Dominie (1980) was apparently the first to successfully use this buffered dose approach on a large scale, when Annabessacook Lake, Maine (alkalinity 20 mg CaCO 3 /L), was treated with this mixture in an empirically determined ratio of 0.63:1 sodium aluminate to alum. Sodium carbonate was also used successfully to treat Long Lake, Washington (alkalinity 35 mg/L) in 1991 to maintain a pH above 6.2 (Welch, 1996). By adding a buffer, it is possible to add an amount of alum to lake sediments that is limited only by available funds. While this procedure is based on alkalinity and does not consider the quantity of internal loading, high-alkalinity lakes dosed by the alkalinity method have experienced substantial treatment longevity. The second procedure for determining dose is based on estimated rates of net internal P loading from the sediments as determined from a mass balance equation. In the first use of this approach, the net internal loading per year in Eau Galle Reservoir, Wisconsin was determined and multiplied by 5, with the goal of controlling P release for 5 years as a desirable target, assuming that the P complexed by Al would ultimately attain a stoichiometric ratio of 1.0 (Kennedy et al., 1987). The Al dose was therefore determined as the quantity of Al equivalent to five times the average summer internal P load. This quantity was then doubled to account for any underestimate of internal loading (release rate can vary year-to-year; Figure 3.5) giving a final dose of 14 g/m 2 Al. Had the alkalinity procedure been used the dose would have been 45 g/m 2 , greatly increasing the ratio of Al–P added:Al–P formed. A ratio of 5–10 appears to be appropriate based on observations from core analyses in alum-treated lakes (Rydin et al., 2000). Dose, using the internal loading rate, was expressed as a mass-areal unit (Al/m 2 ), which is actually more appropriate than is concentration resulting from the alkalinity procedure. The dose expressed as an areal unit is what the sediments should actually receive regardless of water column depth. Eberhardt (1990) has used a similar dose calculation, modified to account for application efficiency of the equipment. FIGURE 8.5 Curve to determine pounds of alum per gallon, based on adjusted Baumé. (From Cooke et al. 1978.) lbs./gal. 6.0 5.0 4.0 3.0 2.0 20 25 30 35 40 Adjusted °Be (Allied chem. corp.) Copyright © 2005 by Taylor & Francis [...]... treatment and then again in 1 988 , 1 989 , and 1990 (Garrison and Ihm, 1991) and briefly again in 1991 (Welch and Cooke, 1999) A destratification system was installed in Mirror Lake and operated in spring and fall to ensure full circulation Its operation could confound the data interpretation Figures 8. 12 and 8. 13 illustrate the volume-weighted mean P concentrations in the two lakes following diversion and. .. 1977 19 78 1979 Ju Se De Ma Ju Se De Ma Ju Se De 1 988 1 989 1990 FIGURE 8. 12 Volume-weighted mean P concentrations in Mirror Lake before, immediately following, and a decade after completion of restoration work (From Garrison, P.J and D.M Ihm 1991 First Annual Report of Long-Term Evaluation of Wisconsin’s Clean Lake Projects Part B Lake Assessment Wisconsin Dept Nat Res., Madison.) 8. 4.2.1 Mirror and Shadow... 5/ 78 AS 6.6 Hypolimnion 0.05 13.1 7 .8 222 Dimictic 10 Shadow, Waupaca, WI 5/ 78 AS 5.7 Hypolimnion 0.17 12.4 5.3 188 Dimictic Copyright © 2005 by Taylor & Francis 6 /86 Ref Spittal and Burton, 1991 Cooke et al., 19 78 Cooke et al., 19 78 Garrison and Knauer, 1 984 Garrison and Ihm, 1991 Garrison and Ihm, 1991 11 Snake, Woodruff, WI 5/72 12 Horeshoe, Manitowoc, WI 5/70 AS:SA Ratio unknown AS 13 14 5 /86 9 /80 ... epilimnion is the treatment of a 4.6 ha section of 89 ha sandpit Lake Leba, Nebraska (Holz and Hoagland, 1999) The isolated section has mean and maximum depths of 4.2 and 9 m, respectively, and it stratifies strongly (the Osgood Index, or OI = 19 .8; Chapter 3) The section was dosed with 10 mg/L Al in 1994 Hypolimnetic SRP and epilimnetic TP remained below pretreatment levels by 97% and 74%, respectively,... WA 18 19 Pattison, Tumwater, WA Wapato, Parkland WA 15 12 (80 % of lake v) Surface 0.05 5.5 2.0 50 Dimictic Garrison and Knauer, 1 984 2.6 Surface 0.09 16.7 4.0 2 18 2 78 Dimictic AS AS 4.5 5.5 Hypolimnion Surface 0.60 1.40 9.0 3.7 3.2 2.0 144 10–40 Dimictic Polymictic Garrison and Knauer, 1 984 Barko et al., 1990 Welch, et al., 1 982 9 /83 AS 7.7 Surface 1.30 6.4 3.6 45 Polymictic Entranco, 1 987 b 9 /85 AS... (5–6) 24 (4–5) 42 Chl a 81 49 29 0 93 45 47 30 (17– 18) (17– 18) (17– 18) (8 13) (5 8) (4 8) (5–6) (4–5) 42 Note: Years of observation in parentheses Lakes showing availability of hypolimnetic TP to the epilimnion are indicated with +, and those that did not with – Source: Modified from Welch, E.B and Cooke, G.D 1999 Lake and Reservoir Manage 15 With permission 57%) in epilimnetic TP and chlorophyll (chl)... Years Lake TP Chl a TP Chl a Erie (> 8) a Campbell (> 8) Long (T) North (> 8) South (5) Pattison North (7) South (< 1) Long (K) (> 11) Mean of 6b 77 43 91 44 75 46 83 28 60 32 89 68 56 50 39 0 43 –7 48 51 40 6 65 66 29 — 30 48 — — 49 40 a b Longevity in years in parentheses Unsuccessful treatment of Pattison-South excluded Source: From Welch, E.B and Cooke, G.D 1999 Lake and Reservoir Manage 15: 5–27 With... Jacoby, 2001) The lake was treated with alum in 1 980 TP declined from a three-summer mean of 65 μg/L to a four-summer mean of 32 μg/L, but then returned abruptly to pretreatment levels during 1 985 –1 986 , and declined again in 1 986 –1 987 , and averaged 41 μg/L from 1 987 to 1991 (Figure 8. 17) By summer 1990, TP had slowly returned to near pretreatment levels and a second Al application was carried out in October... 84 Year 85 86 87 Harvest (69%) Drawdown 30 Harvest (43%) 40 88 89 90 91 92 FIGURE 8. 17 Whole-lake mean total phosphorus concentration during summer in Long Lake in relation to various lake treatments (From Welch et al., 1994 Verh Int Verein Limnol 25.With permission.) Copyright © 2005 by Taylor & Francis 10 Alum Drawdown 0 Alum TPL - TPi, μg L−1 (June– September averages) 20 −10 −20 76 – 78 80 81 – 84 ... 35 (1) 55 43 (1–2) –7 (1–2) 48 (1–4) 68 (1) b –24 (1–2) –26 (1) 51 Latest % Reduction in Rate Release Initial% Latest % Longevity (yr) 82 (5–6) 64 (5–6) >8 >8 56 (7 8) 50 (4–5) 84 (1–2) 79 (7 8) Macrophytes >8 5 29 (5–7) – 30 (7–11) 81 (1–2) 73 (5–7) Macrophytes 62 (1–4) 40 (7–10) 7 . hard water lake (Murphy et al., 1 988 ). 8. 2.1 ALUMINUM The chemistry of aluminum is complex and incompletely understood (Dentel and Gossett, 1 988 ; Bertsch, 1 989 ). The reactions in water have been. the rate and extent of reactions of aluminum salts in water (Driscoll and Letterman, 1 988 ). At low temperatures, coagulation and deposition are significantly reduced and high quantities of species. With permission.) Lake Delavan ( 0-1 cm) 0 400 80 0 1200 1600 1 10 100 1000 Al+1(mg g -1 DW) P (μg g -1 DW) Al-P Extractable biogenic-P Fe-P Ca-P Loosely sorbed-P Copyright © 2005 by Taylor &