l A dissolved criterion treats all the toxic and nontoxic dissolved metal in the site water as if its average toxicity were the same as the average toxicity of all the toxic and nontoxic dissolved metal in the toxicity tests in laboratory dilution water on which the criterion is based.

l A dissolved m is a measurement of the actual ratio of the average toxicities of the dissolved metal and replaces the assumption that the ratio is 1. In both cases, use of a criterion without a WER involves measurement of toxicity in laboratory dilution water but only prediction of toxicity in site water, whereas use of a criterion with a wER involves measurement of toxicity in both laboratory dilution water and site water.

When WERs are used to derive site specific criteria, the total recoverable and dissolved approaches are inherently consistent. They are consistent because the toxic effects caused by the metal in the toxicity tests do not depend on what chemical measurements are performed; the same number of organisms are killed in the acute lethality tests regardless of what, if any, measurements of the concentration of the metal are made. The only difference is the chemical measurement to which the toxicity is referenced. Dissolved WERs can be derived from the same pairs of toxicity tests from which total recoverable WERs are derived, if the metal in the tests is measured using both the total recoverable and dissolved measurements. Both approaches start at the same place ( i. e., the amount of toxicity observed in laboratory dilution water) and end at the same place ( i. e., the amount of toxicity observed in site water). The combination of a total recoverable criterion and WER accomplish the same thing as the combination of a dissolved criterion and WER. By extension, whenever a criterion and a WER based on the same measurement of the metal are used together, they will end up at the same place. Because use of a total recoverable criterion with a total recoverable m ends up at exactly the same place as use of a dissolved criterion with a dissolved m, whenever one WER is determined, both should be determined to allow ( a) a. check on the analytical chemistry, ( b) use of the inherent internal consistency to check that the data are used correctly, and ( c) the option of using either approach in the derivation of permit limits. . An examination of how the two approaches ( the total recoverable approach and the dissolved approach) address the four relevant forms of metal ( toxic and nontoxic particulate metal and toxic and nontoxic dissolved metal) in laboratory dilution water and in site water further explains why the two approaches are inherently consistent. Here, only the way in which the two approaches address each of the four forms of metal in site water will be considered:

114 a. Toxic dissolved metal: This form contributes to the toxicity of the site water and is measured by both chemical measurements. If this is the only form of metal present, the two WERs will be the same. b. Nontoxic dissolved metal: This form does not contribute to the toxicity of the site water, but it is measured by both chemical measurements. If this is the only form of metal present, the two WERs will be the same. ( Nontoxic dissolved metal can be the only form present, however, only if all of the nontoxic dissolved metal present is refractory. If any labile nontoxic dissolved metal is present, equilibrium will require that some toxic dissolved metal also be present.) c. Toxic particulate metal: This form contributes to the toxicological measurement in both approaches; it is measured by the total recoverable measurement, but not by the dissolved measurement. Even though it is not measured by the dissolved measurement, its presence is accounted for in the dissolved approach because it increases the toxicity of the site water and thereby decreases the dissolved WER. It is accounted for because it makes the dissolved metal appear to be more toxic than it is. Most toxic particulate metal is probably not toxic when it is particulate; it becomes toxic when it is dissolved at the gill surface or in the digestive system; in the surface water, however, it is measured as particulate metal. d. Nontoxic particulate metal: This form does not contribute to the toxicity of the site water; it is measured by the total recoverable measurement, but not by the dissolved measurement. Because it is measured by the total recoverable measurement, but not by the dissolved measurement, it causes the total recoverable WER to be higher than the dissolved WER. In addition to dealing with the four forms of metal similarly, the WERs used in the two approaches comparably take synergism, antagonism, and additivity into account. Synergism and additivity in the site water increase its toxicity and therefore decrease the WER; in contrast, antagonism in the site water decreases toxicity and increases the WER.

Each of the four forms of metal is appropriately taken into account because use of the WERs makes the two approaches internally consistent. In addition, although experimental variation will cause the measured WERs to deviate from the actual WERs, the measured WERs will be internally consistent with the data from which they were generated. If the percent dissolved is the same at the test endpoint in the two waters, the two WERs will be the same. If the percent of the total recoverable metal that is dissolved in laboratory dilution water is less than 100 percent, changing from the total recoverable measurement to the dissolved measurement will lower the criterion but it will 115 comparably lower the denominator in the WER, thus increasing the WER. If the percent of the total recoverable metal that is dissolved in the site water is less than 100 percent, changing from the total recoverable measurement to the dissolved . measurement will lower the concentration in the site water that is to be compared with the criterion, but it also lowers the numerator in the WER, thus lowering the WER. Thus when WEI& are used to adjust criteria, the total recoverable approach and the dissolved approach result in the same interpretations of concentrations in the site water ( see Figure D3) and in the same maximum acceptable concentrations in effluents ( see Figure D4).

Thus, if WEFU are based on toxicity tests whose endpoints equal the CMC or CCC and if both approaches are used correctly, the two measurements will produce the same results because each WER is based on measurements on the site water and then the WER is used to calculate the site specific criterion that applies to the site water when the same chemical measurement is used to express the site specific criterion. The eguivalency of the two approaches applies if they are based on the same sample of site water. When they are applied to multiple samples, the approaches can differ depending on how the results from replicate samples are used: a. If an appropriate averaging process is used, the two will be equivalent. b. If the lowest value is used, the two approaches will probably be equivalent only if the lowest dissolved WER and the lowest total recoverable WER were obtained using the same sample of site water.

There are several advantages to using a dissolved criterion even when a dissolved WER is not used. In some situations use of a dissolved criterion to interpret results of measurements of the concentration of dissolved metal in site water might demonstrate that there is no need to determine either a total recoverable WER or a dissolved WER. This would occur when so much of the total recoverable metal was nontoxic particulate metal that even though the total recovdrable criterion was exceeded, the corresponding dissolved criterion was not exceeded. The particulate metal might come from an effluent, a resuspension. event, or runoff that washed particulates into the body of water. In such a situation the total recoverable WER would also show that the site specific criterion was not exceeded, but there would be no need to determine a WER if the criterion were expressed on the basis of the dissolved measurement. If the variation over time in the concentration of particulate metal is much greater than the variation in the concentration of dissolved metal, both the total recoverable concentration and the total recoverable WER are likely to vary so much over time that a dissolved criterion would be much more useful than a total recoverable criterion.

116 Use of a dissolved criterion without a dissolved WER has three disadvantages, however: 1. Nontoxic dissolved metal in the site water is treated as if it is toxic. 2. Any toxicity due to particulate metal in the site water is ignored. 3. Synergism, antagonism, and additivity in the site water are not taken into account. Use of a dissolved criterion with a dissolved WER overcomes all three problems. For example, if ( a) the total recoverable concentration greatly exceeds the total recoverable criterion, ( b) the dissolved concentration is below the dissolved criterion, and ( c) there is concern about the possibility of toxicity of particulate metal, the determination of a dissolved WER would demonstrate whether toxicity due to particulate metal is measurable.

Similarly, use of a total recoverable criterion without a total recoverable WER has three comparable disadvantages: 1. Nontoxic dissolved metal in site water is treated as if it is toxic. 2. Nontoxic particulate metal in site water is treated as if it is toxic. 3. Synergism, antagonism, and additivity in site water are not taken into account. Use of a total recoverable criterion with a total recoverable WER overcomes all three problems. For example, determination of a total recoverable WER would prevent nontoxic particulate metal ( as well as nontoxic dissolved metal) in the site water from being treated as if it is toxic.

pelationshim between WERs and the Forms of Metal%

Probably the best way to understand what WERs can and cannot do is to understand the relationships between WERs and the forms of metals. A WER is calculated by dividing the concentration of a metal that corresponds to a toxicity endpoint in a site water by the concentration of the same metal that corresponds to the same toxicity endpoint in a laboratory dilution water. using the classification scheme given in Figure Dl: Therefore,

mR= R,, + N, + T, + dV' + AT, Rt + NL + Tt + AN, + AT,

The subscripts IS. and . t' denote site water and laboratory dilution water, respectively, and:

R = the concentration of Befractory metal in a water. definition, ( By all refractory metal is nontoxic metal.)

117 N = the concentration of Nontoxic labile metal in a water.

T = the concentration of xoxic labile metal in a water.

AN = the concentration of metal added during a WEX determination that is Nontoxic labile metal'after it is added.

AT = the concentration of metal added during a WER determination that is sxic labile metal after it is added.

For a total recoverable WER; each of these five concentrations includes both particulate and dissolved metal, if both are present; for a dissolved WER only dissolved metal is included.

Because the two side by side tests use the same endpoint and are conducted under identical conditions with comparable test organisms, T,+ AT,  Tt + AT, when the toxic species of the metal are equally toxic in the two waters. If a difference in water quality causes one or more of the toxic species of the metal to be mOre toxic in one water than the other, or causes a shift in the ratios of various toxic species, we can define II= Tn + AT, T&+ AT, '

Thus H is a multiplier that accounts for a proportional increase or decrease in the toxicity of the toxic forms in site water as compared to their toxicities in laboratory dilution water. Therefore, the general WE3 equation is:

WERt Rs+ N,+ aN,+ H( TL+~ TL) R,+ N,+ aN,+ ( T,+ AT,) l Several things are obvious from this equation: 1. A WER should not be thought of as a simple ratio such as H. H is the ratio of the toxicities of the toxic species of the metal, whereas the WER is the ratio of the sum of the toxic and the nontoxic species of the metal. Only under a very specific set of conditions will wg~  H. If these conditions are satisfied and if, in addition, H= 1, then wg~  I. Although it might seem that all of these conditions will rarely be satisfied, it is not all that rare to find that an experimentally determined WER is close to 1. 2. When the concentration of metal in laboratory dilution water is negligible, RL  NL = Tt  0 and wm RH + NB + dV" + HAT,) A& + AT, .

118 Even though laboratory dilution water is low in Tot and TSS, when metals are added to laboratory dilution water in toxicity tests, ions such as hydroxide, carbonate1 and chloride react with some metals to form some particulate species and. some dissolved species, both of which might be toxic or nontoxic. The metal species that are nontoxic contribute to A&, whereas those that are toxic contribute to AT,. Hydroxide, carbonate, chloride, Tot, and TSS can increase AN'. Anything that causes AN', to differ from AN' will cause the WER to differ from 1. 3. Refractory metal and nontoxic labile metal in the site water above that in the laboratory dilution water will increase the WER. Therefore, if the WER is determined in downstream water, rather than in upstream water, the WER will be increased by refractory metal and nontoxic labile metal in the effluent. Thus there are three major reasons why WERs might be larger or smaller than 1: a. The toxic species of the metal might be more toxic in one water than in the other, i. e., H l 1. b. AN might be higher in one water than in the other. C. R and/ or N might be higher in one water than in the other.

The last reason might have great practical importance in some situations. When a WER is determined in downstream water, if most of the metal in the effluent is nontoxic, the WER and the endpoint in site water will correlate with the concentration of metal in the site water. In addition, they will depend on the concentration of metal in the effluent and the concentration of effluent in the site water. This correlation will be best for refractory metal because its toxicity cannot be affected by water quality characteristics; even if the effluent and upstream water are quite different so that the water quality characteristics of the site water depend on the percent effluent, the toxicity of the refractory metal will remain constant at zero and the portion of the WER that is due to refractory metal will be additive.

The Denendence of WERs on the Sensitivitv of Toxicity Tests It would be desirable if the magnitude of, the WER for a site water were independent of the toxicity test used in the determination of the WER, so that any convenient toxicity test could be used. It can be seen from the general WER equation that the WER will be independent of the toxicity test only if:

which would require that R, N,  A&  RL  NL  tiL  0. ( It would be easy to assume that Tt  0, but it can be misleading in some situations to make more simplifications than are necessary.)

119 This is the simplistic concept of a WER that would be advantageous if it were true, but which is not likely to be true very often. Any situation in which one or more of the terms is greater than zero can cause the WER to depend on the sensitivity of the toxicity test, although the difference in the WERs might be small.

Two situations that might be common can illustrate how the WER can depend on the sensitivity of the toxicity test. For these illustrations,. there is no advantage to assuming that H= 1, so a will be retained for generality. 1.

2. The simplest situation is when R,> 0, i. e., when a substantial concentration of refractory metal occurs in the site water. If, for simplification, it is assumed that N, aN,= RL= NL tiL Or then:

Nmi R, + H( T, + AT,) I R6 ( T, + AT,) ( TL + AT,) + H.

The quantity T,+ AT, obviously changes as the sensitivity of the toxicity test changes. When R, 0, then MER H and the WER is independent of the sensitivity of the toxicity test. when R#> 0, then the WER will decrease as the sensitivity of the test decreases because TL + AT, will increase.

More complicated situations occur when ( N, + a~,) > 0. If, for simplification, it is assumed that R6  RL  N,  AN,  0, then:

=  ( N6 + d7,) + H( T, + A?`,)  ( 4 + fl#) ( T + AT + H. L L) ( T, + AT,)

a. If ( N, + AN,) > 0 because the site water contains a substantial concentration of a complexing agent that has an affinity for the metal and if corqplexation converts toxic metal into nontoxic metal, the complexation reaction will control the toxicity of the solution ( Allen 1993). A complexation cume can be graphed in several ways, but the S shaped curve presented in Figure DS is most convenient here. The vertical axis is '% uncomplexed', which is assumed to correlate with '% toxic'. then the '% nontoxic'. The ratio of toxic metal is:

% nontoxic, % ccnl@ ad % toxic % Llncoilplm! Bd For the complexed nontoxic metal: The '% complexed' is nontoxic metal to  v.

vt concentration ofnontoxdcmetal concentration of toxicmetal l 120 In the site water, the concentration of complexed nontoxic metal is ( N, + ti6) and the concentration of toxic metal is ( T,+ AT,), so that:.

( Iy, + u#) ( N4 + a#) v6  ( T6 + AT,)  H( T, + AT,) l and wBRN V&( TL + AT,) + H(=, + AT,)

( TL + AT,)  vfi+ H= H( V,+ 1) .

If the WRR is determined using a sensitive toxicity test so that the % uncomplexed ( i. e., the % toxic) is 10 %, then v, ( 90 %)/( I0 %)  9, whereas if a less sensitive test is used so that the % uncomplexed is 50 %, then v, ( 50 %)/( 50 %)  1. Therefore, if a portion of the WER is due to a complexing agent in the site water, the magnitude of the WER can decrease as the sensitivity of the toxicity test decreases because the % uncomplexed will decrease. In these situations, the largest WER will be obtained with the most sensitive toxicity test; progressively smaller WERs will be obtained with less sensitive toxicity tests. The magnitude of a WER will depend not only on the sensitivity of the toxicity test but also on the concentration of the complexing agent and on its binding constant ( complexation constant, stability constant). In addition, the binding constants of most complexing agents depend on pH.

If the laboratory dilution water contains a low concentration of a complexing agent,

VL  ru, + flL TL + AT,

and mR v& T, + AT,) + H( T, + A?',) = v,$ + H I H( v, + 1) .

VL( TL + AT,) + tTL + AT,) v, + 1 v, + 1 The binding constant of the complexing agent in the laboratory dilution water is probably different from that of the complexing agent in the site water. Although changing from a more sensitive test to a less sensitive test will decrease both V, and v,, the amount of effect is not likely to be proportional.

If the change from a more sensitive test to a less sensitive test were to decrease v, proportionately more than v,, the change could result in a larger WER, rather 121 than a smaller WER, as resulted in the case above when it was assumed that the laboratory dilution water did not contain any complexing agent. This is probably most likely to occur if H? 1 and if v# < V,, which would mean that lygR< l. Although this is likely to be a rare situation, it does demonstrate again the'importance of determining WERS using toxicity tests that have endpoints in laboratory dilution water that are close to the CMC or CCC to which the WER is to be applied.

b. If ( N,+ AN,) > 0 because the site water contains a substantial concentration of an ion that will precipitate the metal of concern and if precipitation converts toxic metal into nontoxic metal, the precipitation reaction will control the toxicity of the solution. The ' precipitation curve. given in Figure D6 is analogous to the ' coxnplexation curve. given in Figure D5; in the precipitation curve, the vertical axis is '% dissolved., which is assumed to correlate with ' 8 toxic'. If the endpoint for a toxicity test is below the solubility limit of the precipitate, ( N,+ AN,)  Or whereas if the endpoint for a toxicity test is above the solubility limit, ( N, + ti#) > 0. If WERs are determined with a series of toxicity tests that have increasing endpoints that are above the solubility limit, the WER will reach a maximum value and then decrease. The magnitude of the WER will depend not only on the sensitivity of the toxicity test but also on the concentration of the precipitating agent, the solubility limit, and the solubility of the precipitate.

Thus, depending on the composition of the site water, a WER obtained with an insensitive test might be larger, smaller, or similar to a WER obtained with a sensitive test. Because of the range of possibilities that exist, the best toxicity test to use in the experimental determination of a WEX is one whose endpoint in laboratory dilution water is close to the CMC or CCC that is to be adjusted. This is the rationale that was used in the selection of the toxicity tests that are suggested in Appendix I.

The available data indicate that a less sensitive toxicity test usually gives a smaller WER than a more sensitive test ( Hansen 1993a). Thus, use of toxicity tests whose endpoints are higher than the CMC or CCC probably will not result in underprotection; in contrast, use of tests whose endpoints are substantially below the CMC or CCC might result in underprotection.

The factors that cause ~~ and ( N, + AN,) to be greater than zero are all external to the test organisms; they are chemical effects that affect the metal in the water. The magnitude of the WER is therefore expected to depend on the toxicity test used only in regard to the sensitivity of the test. If the endpoints for two 122 different tests occur at the same concentration of the metal, the magnitude of the WERs obtained with the two tests should be the same; they should not depend on ( a) the duration of the test, ( b) whether the endpoint is based on a lethal or sublethal effect, or ( c) whether the species is a vertebrate or an invertebrate.

Another interesting consequence of the chemistry of complexation is that the % uncosnplexed will increase if the solution is diluted ( Allen and Hansen 1993). The concentration of total metal will decrease with dilution but the % uncomplexed will increase. The increase will not offset the decrease and so the concentration of uncomplexed metal will decrease. Thus the portion of a WER that is due to complexation will not be strictly additive ( see Appendix G), but the amount of nonadditivity might be difficult to detect in toxicity studies of additivity. A similar effect of dilution will occur for precipitation.

The illustrations presented above were simplified to make it easier to understand the kinds of effects that can occur. The illustrations are qualitatively valid and demonstrate the direction of the effects, but real world situations will probably be so much mOre complicated that the various effects cannot be dealt with separately.

. Pther ProDertles of WQ&

1. Because of the variety of factors that can affect WERs, no rationale exists at present for extrapolating WERs from one metal to another, from one effluent to another, or from one surface water to another. Thus WERs should be individually determined for each metal at each site.

2. The most important infonaation that the determination of a WER provides is whether simulated and/ or actual downstream water adversely affects test organisms that are sensitive to the metal. A WER cannot indicate how much metal needs to be removed from or how much metal can be added to an effluent. a. If the site water already contains sufficient metal that it is toxic to the test organisms, a WER cannot be determined with a sensitive test and so an insensitive test will have to be used. Even if a WER could be determined with a sensitive test, the WER cannot indicate how much metal has to be removed. For example, if a WER indicated that there was 20 percent too much metal in an effluent, a 30 percent reduction by the discharger would not reduce toxicity if only nontoxic metal was removed. Thenext WER determination would show that the effluent still contained too much metal. Removing metal is useful only if the metal removed is toxic metal. Reducing the total recoverable concentration does not necessarily reduce toxicity.

123 3. b. If the simulated or actual downstream water is not toxic, a WER can be determined and used to calculate how much additional metal the effluent could contain and still be acceptable. Because an unlimited amount of refracmry metal can be added to the effluent without affecting the organisms, what the WER actually determines is how much additional toxic metal can. be added to the effluent.

The effluent component of nearly all WEFU is likely to be due mostly to either ( a) a reduction in toxicity of the metal by TSS or ' pot, or ( b) the presence of refractory metal. For both of these, if the percentage of effluent in the downstream water decreases, the magnitude of the WER will usually decrease. If the water quality characteristics of the effluent and the upstream water are quite different, it is possible that the interaction will not be additive; this can affect the portion of the WER that is due to reduced toxicity caused by sorption and/ or binding, but it cannot affect the portion of the WER that is due to refractory metal.

4. Test organisms are fed during some toxicity tests, but not during others; it is not clear whether a WER determined in a fed test will differ from a WER determined in an unfed test. Whether there is a difference is likely to depend on the metal, the type and amount of food, and whether a total recoverable or dissolved WER is determined. This can be evaluated by determining two WERs using a test in which the organisms usually are not fed  one WER with no food added to the tests and one with food added to the tests. Any effect of food is probably due to an increase in TOC and/ or TSS. If food increases the concentration of nontoxic metal in both the laboratory dilution water and the site water, the food will probably decrease the WER. Because complexes of metals are usually soluble, cqlexation is likely to lower both total recoverable and dissolved WEF&; sorption to solids will probably reduce only total recoverable WERs. The food might also affect the acute chronic ratio. Any feeding during a test should be limited to the minimum necessary.

The acceptable WERs found by Brungs et al. ( 1992) were total recoverable WERs that were determined in relatively clean fresh water. These WERs ranged from about 1 to 15 for both copper and cadmium, whereas they ranged from about 0.7 to 3 for zinc. The few WEF& that were available for chromium, lead, and nickel ranged from about 1 to 6. Both the total recoverable and dissolved WERs for copper in New York harbor range from about 0.4 to 4 with most of the WERs being between 1 and 2 ( Hansen 1993b).

124 Figure D2: An & amp10 of the Empirical S% trapolation Proces8 Assume the following hypothetical effluent and upstream water:

Effluent: T,: 100 ug/ L D6: 10 ug/ L U*: 24 cfs Upstream water: T,: 40 ug/ L Do: 38 ug/ L Pa: 48 cfs Downstream water: Tn : 60 ug/ L 4.8: 36 ug/ L QD: 72 cfs where:

T = concentration of D = concentration of u = flow. ( 10 8 dissolved)

( 95 % dissolved)

( 60 % dissolved)

total recoverable metal. dissolved metal.

The subscripts E, U, and D signify effluent, upstream water, and downstream water, respectively.

By conservation of flow: & Q6+ PO.

By conservation of total recoverable metal: TIpo Tp,+ Trpo.

If P = the percent of the total recoverable metal in the effluent. that becomes dissolved in the downstream water,

pI loo(~ D DI&)

W6 '

For the data given above, the percent of the total recoverable metal in the effluent that becomes dissolved in the downstream water is:

p I 100 [ ( 36 ug/ t) ( 72 CfB)  ( 38 W/ t) ( 48 Cfd 1 . 32 Q , ( 100 u& L) ( 24 Cf8)

which is greater than the 10 % dissolved in the effluent and less than the 60 % dissolved in the downstream water.

125 Figure 03: The Intormal Coa8imtuxcy of the Two Approachem The internal consistency of the total recoverable and dissolved approaches can be illustrated by considering the use of WERS to interpret the total recoverable and. dissolved concentrations of a metal in a site water. For this hypothetical example, it will be assumed that the national CCCs for the metal are: 200 ug/ L as total recoverable metal. 160 ug/ L as dissolved metal. It will ' also be assumed that the concentrations of the metal in the site water are: 300 ug/ L as total recoverable metal. 120 ug/ L as dissolved metal. The total recoverable concentration in the site water exceeds the national CCC, but the dissolved concentration does not.

The following results might be obtained if WERs are determined:

atorv . . Dilution Water Total recoverable LCSO = 400 ug/ L. % of the total recoverable metal that is dissolved = 80. ( This is based on the ratio of the national CCCs, which were determined in laboratory dilution water.) Dissolved LCSO = 320 ug/ L.

Total recoverable LCSO = 620 ug/ L. % of the total recoverable metal that is dissolved = 40. ( This is based on the data given above for site water). Dissolved LCSO = 248 ug/ L.

Total recoverable WEB = ( 620 ug/ L)/( 400 ug/ L) = 1.55 Dissolved WER = ( 248 ug/ L)/( 320 ug/ L) = 0.775 Tmalrocowrabl~ WgR nissolvedclrgR 91.55. 0.775 lab water % dissolvd * 80 N 2 site hater % dfssolwd 40 Total recoverable ssCCC = ( 200 ug/ L)( l. Sf) = 310 ug/ L. Dissolved ssCCC = ( 160 ug/ L)( O. 775) = 124 ug/ L.

Both concentrations in site water are below the respective sscccs.

126 In contrast, the following results might have been obtained when the WERs were determined:

In Laboratorv Dilution Wate Total recoverable LCSO = r400 ug/ L. % of the total recoverable metal that is dissolved = 80. Dissolved LCSO = 320 ug/ L.

In Site ate TotalWre& erable LC50 = 580 ug/ L. % of the total recoverable metal that is dissolved = 40. Dissolved LC50 = 232 ug/ L.

S Total recoverable WER = ( 580 ug/ L)/( QoO ug/ L) = Dissolved WER = ( 232 ug/ L)/( 320 ug/ L) = 0.725 Checkina the Calculations mta1 reBcoverable WER 1.45 t 9 lab water 0 dissolwd Dissolved WHZ 0.725 site water 0 ciissolv8d 1.45 980.2 40 Site snecific CCCs ( ssCCCs)

Total recoverable ssCCC = ( 200 ug/ L) ( 1.45) = 290 ug/ L. Dissolved ssCCC = ( 160 ug/ L) ( 0.725) = 116 ug/ L.

In this case, both respective ssCCCs. concentrations in site water are above the In each case, both approaches resulted in the same concerning whether the concentration in site water site specific criterion. conclusion exceeds the The two key assumptions are: 1. The ratio of total recoverable metal to laboratory dilution water when the WERs . the ratio of the national CCCs. 2. The ratio of total recoverable metal to site water when the WERs are determined the concentrations reported in the site Differences in the ratios that are outside experimental variation will cause problems dissolved metal in are determined equals dissolved metal in equals the ratio of water. the range of for the derivation of site specific criteria and, therefore, with the internal consistency of the two approaches.

127 Figure 01: The AOplicrtion of tha SW0 Approrchm Hypothetical upstream water and effluent will be used to demonstrate the equivalence of the total recoverable and dissolved approaches. The upstream water and the effluent will be assumed to have specific properties in order to allow calculation of the properties of the downstream water, which will be assumed to be a 1: l mixture of the upstream water and effluent. It will also be assumed that the ratios of the forms of the metal in the upstream water and in the effluent do not change when the total recoverable concentration changes.

YEfFEergiF = 3 cfs) Refractory par& ulate: 400 ug/ L 200 ug/ L Toxic dissolved: 200 ug/ L ( 50 % dissolved)

Bffluea ( Flow = 3 cfs) Total recoverable: 440 ug/ L Refractory particulate: 396 ug/ L Labile nontoxic particulate: 44 ug/ L Toxic dissolved: 0 ug/ L ( 0 % dissolved) ( The labile nontoxic particulate, which is 10 % of the total recoverable in the effluent, becomes toxic dissolved in the downstream water.)

( Flow = 6 cfs) Downstream wateq Total recoverable: 420 ug/ L Refractory particulate: 298 ug/ L Toxic dissolved: 122 ug/ L ( 29 % dissolved)

The values for the downstream water are calculated from the values for the upstream water and the effluent: Total recoverable: [ 3( 400) + 3( 440)]/ 6 = 420 ug/ L Dissolved: [ 3( 200) + 3( 44+ 0)]/ 6 = 122 ug/ L Refractory particulate: [ 3( 200) + 3( 396))/ 6 = 298 ug/ L Assumed National CCC ( CCC) Total recoverable tn300 ug/ L Dissolved = 240 ug/ L 128 Uostream site snecific CCC ( ussCCC)

Assume: Dissolved CCCWER = 1.2 Dissolved ussCCC = ( 1.2)( 240 ug/ L) = 288 ug/ L By calculation: TR ussCCC = ( 288 ug/ L)/( O. S) = 576 ug/ L Total recoverable CCCWER = ( 576 ug/ L)/( 300 ug/ L) = 1.92 30: ccc ECCWER ussccc co . Total recoverable: ug/ L 1.92 576 ug/ L Dissolved: 400n: g, L 240 ug/ L 1.2 288 ug/ L 200 ug/ L % dissolved 80 %  50 % 50 % Neither concentration exceeds its respective ussCCC.

lmalreowrablemR 91.92. lab water 0 dissolwd Dissolved HER Sit8 diseolwd N 1.2 water % 80 50 N1 l 6 Downstream site soecific CCC ( dssCCC)

Assume: Dissolved cccWER = 1.8 Dissolved dssCCC = ( 1.8) ( 240 ug/ L) = 432 ug/ L By calculation: TR dssCCC = (( 432 ug/ L I( 200 ug/ L)/ 2l)/ O. ll+(( 400 ug/ L)/ 2) = 3520 ug/ L This calculation determines the amount of dissolved metal contributed by the effluent, accounts for the fact that ten percent of the total recoverable metal in the effluent becomes dissolved, and adds the total recoverable metal contributed by the upstream flow. Total recoverable cccWER = ( 3520 ug/ L)/( 300 ug/ L) = 11.73 ccc Total recoverable: 30: ug/ L cccl+ mR dssCCC co c. 11.73 Dissolved: 3520 ug/ L 420nug, L 240 ug/ L 1.80 432 ug/ L % dissolved 122 ug/ L 80 %  12.27 % 29 % Neither concentration exceeds its respective dssCCC.

Total recowrable HER 80 Dissolved MER N 11.73 N lab water 0 disoolwd 1.80 site   water 0 dissolwd 12.27  6.52 Calculatina the Maximum Accentable Concentration in the Effluent Because neither the total recoverable concentration nor the dissolved concentration in the downstream water exceeds its respective site specific CCC, the concentration of metal in the effluent could be increased. Under the assumption that the ratios of the two forms of the metal in the effluent do not change when the total recoverable concentration changes, the maximum acceptable concentration of total recoverable metal in the effluent can be calculated as follows:

129 Starting with the total recoverable dssCCC of 3520 ug/ L ( 6 cfs) ( 3520 w/ L)  ( 3 cfs)( 400 up/ =) 3 cfs  6640 ug/ L Starting with the dissolved dssCCC of 432 ug/ L ( 6 cfs) ( 432 u& L)  ( 3 cf8) ( 400 ug/=) ( o 5) N 6640 ug,~ ( 3 cfs) ( 0.10)

Total recoverable:

( 3 cfs) ( 6640 u#/ L) + ( 3 cfs) ( 400 ug/ =' 6 cfs  3520 u& L.

Dissolved:

( 3 cf8) ( 6640 ug/ L) ( 0.10) + ( 3 Cf8) ( 400 Ug/ L)( 0* 50) 6 cfs  432 ug/ L .

The value of 0.10 is used because this is the percent of the total recoverable metal in the effluent that becomes dissolved in the downstream water.

The values of 3520 ug/ L and 432 ug/ L equal the downstream site specific CCCs derived above.

Another Wav to Calculate the Maximum Accentable Concentration The maximum acceptable concentration of total recoverable metal in the effluent can also be calculated from the dissolved dssCCC of 432 ug/ L using a partition coefficient to convert from the dissolved dssCCC of 432 ug/ L to the total recoverable dssCCC of 3520 ug/ L:

16 cfi] [ 4; 21g; L  ( 3 cfs) ( 400 I& L)] . 3 cfs  6640 tag/ L.

Note that the value used for the partition coefficient in this calculation is 0.1227 ( the one that applies to the downstream water when the total recoverable concentration of metal in the effluent is 6640 ug/ L), not 0.29 ( the one that applies when the concentration of metal in the effluent is only 420 ug/ L). The three ways of calculating the maximum acceptable concentration give the same result if each is used correctly.

130 Figure D5: A Qenoralimd Coqplexation Cufpe The curve is for a constant concentration of the complexing ligand and an increasing concentration of the metal.

100 h W ti . . .

. . .

.

.

.

.

.

.

.

.

. . . . I I I LOG OF CONCENTRATION OF METAL 131 ? iguro D6t A Gumralir~ Precipitation CURO The cume is for a constant concentration of the precipitating ligand and an increasing concentration of the metal.

loo* l l .

l 0 I I I I LOG OF CONCENTRATION OF METAL 132 Allen, H. E. 1993. Importance of Metal Speciation to Toxicity. Proceedings of the Water Environment Federation Workshop on Aguatic Life Criteria for Metals. Anaheim, CA. pp. 55 62.

Allen, H. E., and D. J. Hansen. 1993. The Importance of Trace Metal Speciation to Water Quality Criteria. Paper presented at Society for Environmental Toxicology and Chemistry. Houston, TX. November 15.

Borgmanxi, U. 1983. Metal Speciation and Toxicity of Free Metal Ions to Aquatic Biota. IN: Aquatic Toxicology. ( J. O. Nriagu, ed.) Wiley, New York, NY.

Brungs, W. A., T. S. Holderman, and M. T. Southerland. 1992. Synopsis of Water Effect Ratios for Heavy Metals as Derived for Site Specific Water Quality Criteria. U. S. EPA Contract 68 CO 0070.

Chapman, G. A., and J. K. McCrady. 1977. Copper Toxicity: A Question of Form. In: Recent Advances in Fish Toxicology Tubb, ed.) EPA 600/ 3 77 085 or PB 273 500. National T&~~ c~ l . Information Service, Springfield, VA. pp. 132 151.

Erickson, R. 1993a. Memorandum to C. Stephan. July 14.

Erickson, R. 1993b. Memorandum to C. Stephan. November 12. . French, P., and D. T. E. Hunt. 1986. The Effects of Inorganic Complexing upon the Toxicity of Copper to Aquatic Organisms ( Principally Fish). IN: Trace Metal Speciation and Toxicity to Aguatic Organisms  A Review. ( D. T. E. Hunt, ea. 1 Report TR 247. Water Research Centre, United Kingdom.

Hansen, D. J. 1993a. Memorandum to C. E. Stephan. April 29.

Hansen, D. J. 1993b. Memorandum to C. E. Stephan. October 6.

Nelson, H., D. Benoit, R. Erickson, V. Mattson, and J. Lindberg. 1986. The Effects of Variable Hardness, pH, Alkalinity, Suspended Clay, and Humics on the Chemical Speciation and Aquatic Toxicity of Copper. PB86 171444. National Technical Information Service, Springfield, VA.

Wilkinson, K. J., P. M. Bertsch, C. H. Jagoe, and P. G. C. Campbell. 1993. Surface Complexation of Aluminum on Isolated Fish Gill Cells. Environ. Sci. Technol. 27: 1132 1138.

133 Appmdix St U. S. SPA Aquatic Lifo Criteria Dccunntm for Hotrlm Aluminum Antimony Arsenic Beryllium Cadmium chromium Copper Lead Mercury Nickel Selenium Silver Thallium Zinc EPA 440/ S 86 008 EPA 440/ S 80 020 EPA 440/ S 84 033 EPA 440/ S 80 024 EPA 440/ S 84 032 EPA 440/ S 84 029 EPA 440/ S 84 031 EPA 440/ S 84 027 EPA 440/ S 84 026 EPA 440/ S 86 004 EPA 440/ S 87 006 EPA 440/ S 80 071 EPA 440/ S 80 074 EPA 440/ S 87 003 PB88 245998 PB81 117319 PB85 227445 PB81 117350 PB85 227031 PB85 227478 PB85 227023 PB85 227437 PB85 227452 PB87 105359 PB88 142237 PB81 117822 PB81 117848 PB87 153581 All are available from: National Technical Information Service ( NTIS) 5285 Port Royal Road Springfield, VA 22161 TEL: 703 487 4650 134  ix F: Coruidorrtionm Concoming Multiple Metal, Ibultiplo Dischugo, axad Special Flowing Water SitUatiOM Multinle Metal Situation9 Both Method 1 and Method 2 work well in multiple metal situations, although the amount of testing required increases as the number of metals increases. The major problem is the same for both methods: even when addition of two or more metals individually is acceptable, simultaneous addition of the two or more metals, each at its respective maximum acceptable concentration, might be unacceptable for at least two reasons: 1. Additivity or synergism might occur between metals. 2. More than one of the metals might be detoxified by the same complexing agent in the site water. When WERs are determined individually, each metal can utilize all of the complexing capacity; when the metals are added together, however, they cannot simultaneously utilize all of the complexing capacity. Thus a discharger might feel that it is cost effective to try to justify the lowest site specific criterion that is acceptable to the discharger rather than trying to justify the highest site specific criterion that the appropriate regulatory authority might approve.

There are two options for dealing with the possibility of additivity and synergism between metals: a.

b. WERs could be developed using a mixture of the metals but it might be necessary to use several primary toxicity tests depending on the specific metals that are of interest. Also, it might not be clear what ratio of the metals should be used in the mixture. If a WER is determined for each metal individually, one or more additional toxicity tests must be conducted at the end to show that the combination of all metals at their proposed new site specific criteria is acceptable. Acceptability must be demonstrated. with each toxicity test that was used as a primary toxicity test in the determination of the WEXs for the individual metals. Thus if a different primary test was used for each metal, the number of acceptability tests needed would equal the number of metals. It is possible that a toxicity test used as the primary test for one metal might be more sensitive than the ' CMC ( or CCC) for another metal and thus might not be usable in the combination test unless antagonism occurs. When a primary test cannot be used, an acceptable alternative test must be used. The second option is nreferred because it is more definitive; it provides data for each metal individually and for the mixture. The first option leaves the possibility that one of the metals is antagonistic towards another so that the toxicity of the mixture would increase if the metal causing the antagonism were not present.

135 Because the National Toxics Rule WTR) incorporated WERs into the aquatic life criteria for some metals, it might be envisioned that more than one criterion could apply to a metal at a site if different investigators obtained different WERs for the same metal at the site. . . . . . ; EI1 mzqdictrons sublect to the NTR. as well, l . ens, EPA intends that Dere s d be . no  re one w& szaon for a Dolluta t at a Dolnt in a bide Thus whenever a site specificncriterion is to be of water. derived using a WER at a site at which more than one discharger has permit limits for the same metal, it is important that all dischargers work together with the appropriate regulatory authority to develop a workplan that is designed to derive a site specific criterion that adequately protects the entire site.

Method 2 is ideally suited for taking into account more than one discharger.

Method 1 is straightforward if the dischargers are sufficiently far downstream of each other that the stream can be divided into a separate site for each discharger. Method 1 can also be fairly straightforward if the WERs are additive, but it will be complex if the WERs are not additive. Deciding whether to use a simulated downstream water or an actual downstream water can be difficult in a flowing water multiple discharge situation. Use of actual downstream water can be complicated by the existence of multiple mixing zones and plumes and by the possibility of varying discharge schedules; these same problems exist, however, if effluents from two or more discharges are used to prepare simulated downstream water. Dealing with a multiple discharge situation is much easier if the WERs are additive, and use of simulated downstream water is the best way to determine whether the WERs are additive. Taking into account all effluents will take into account synergism, antagonism, and additivity. If one of the discharges stops or is modified substantially, however, it will usually be necessary to determine a new WER, except possibly if the metal being discharged is refractory. Situations concerning intermittent and batch discharges need to be handled on a case by case basis.

Method 1 is intended to apply not only to ordinary rivers and streams but also to streams that some people might consider ' special., such as streams whose design flows are zero and streams that some state and/ or federal agencies might refer to as ' effluent dependent', ' habitat creating', ' effluent dominated', etc. ( Due to differences between agencies, some streams whose design flows are zero are not considered ' effluent dependent',

136 etc., and some ' effluent dependent' streams have design flows that are greater than zero.) The application of Method 1 to these kinds of streams has the following implications: 1. If the design flow.. is zero, at least some WRRs ought ti be determined in 100% effluent. 2. If thunderstorms, etc., occasionally dilute the effluent substantially, at least one WER should be determined in diluted effluent to assess whether dilution by rainwater might result in underprotection by decreasing the WRR faster than it decreases the concentration of the metal. This might occur, for example, if rainfall reduces hardness, alkalinity, and pH substantially. This might not be a concern if the WER demonstrates a substantial margin of safety. 3. If the site specific criterion is substantially higher than the national criterion, there should be increased concern about the fate of the metal that has reduced or no toxicity. Even if the WRR demonstrates a substantial margin of safety ( e. g., if the site specific criterion is three times the national criterion, but the experimentally determined WE% is 111, it might be desirable to study the fate of the metal. 4. If the stream merges with another body of water and a site specific criterion is desired for the merged waters, another WER needs to be determined for the mixture of the waters. 5. Whether WET testing is reguired is not a WER issue, although WET testing might be a condition for determining and/ or using aWER. 6. A concern about what species should be present and/ or protected in a stream is a beneficial use issue, not a WER issue, although resolution of this issue might affect what species should be used if a WER is determined. ( If the Recalculation Procedure is used, determining what species should be present and/ or protected is obviously important.) 7. Human health and wildlife criteria and other issues might restrict an effluent more than an aquatic life criterion. Although there are no scientific reasons why ' effluent dependent', etc., streams and streams whose design flows are zero should be subject to different guidance than other streams, a regulatory decision ( for example, see 40 CFR 131) might require or allow some or all such streams to be subject to different guidance. For example, it might be decided on the basis of a use attainability analysis that one or more constructed streams do not have to comply with usual aquatic life criteria because it is decided that the water quality in such streams does not need to protect sensitive aquatic species. Such a decision might eliminate any further concern for site specific aquatic life criteria and/ or for WET testing for such streams. The water quality might be unacceptable for other reasons, however.

In addition to its use with rivers and streams, Method 1 is also appropriate for determining cmcWERs that are applicable to near field effects of discharges into large bodies of fresh or salt water, such as an ocean or a large lake, reservoir, or estuary:

137 a.

b. The near field effects of a pipe that extends far into a large body of fresh or salt water that has a current, such as an ocean, can probably best be treated the same as a single discharge into a flowing stream. For example, if a mixing zone is defined, the concentration of effluent at the edge of the mixing zone might be used to'define how to prepare a simulated site water. A dye dispersion study ( Kilpatrick 1992) might be useful, but a dilution model ( U. S. EPA 1993) is likely to be a more cost effective way of obtaining information concerning the amount of dilution at the edge of the mixing zone. The near field effects of a single discharge that is near a shore of a large body of fresh or salt water can also probably best be treated the same as a single discharge into a flowing stream, especially if there is a definite plume and a defined mixing zone. The potential point of impact of near field effects will often be an embayment, bayou, or estuary that is a nursery for fish and invertebrates and/ or contains conmercially important shellfish beds. Because of their importance, these areas should receive special consideration in the determination and use of a WER, taking into account sources of water and discharges, mixing patterns, and currents ( and tides in coastal areas). The current and flushing patterns in estuaries can result in increased pollutant concentrations in confined embayments and at the terminal up gradient portion of the estuary due to poor tidal flushing and exchange. Dye dispersion studies ( Kilpatrick 1992) can be used to determine the spatial concentration of the effluent in the receiving water, but dilution models ( U. S. EPA 1993) might not be sufficiently accurate to be useful. Dye studies of discharges in near shore tidal areas are especially complex. Dye injection into the discharge should occur over at least one, and preferably two or three, complete tidal cycles; subsequent dispersion patterns should be monitored in the ambient water on consecutive tidal cycles using an intensive sampling regime over time, location, and depth. Information concerning dispersion and the comunity at risk can be used to define the appropriate mixing zone( s), which might be used to define how to prepare simulated site water.

Kilpatrick, F. A. 19' 92. Simulation of Soluble Waste Transport and Buildup in Surface Waters Using Tracers. Open File Report 92 457. U. S. Geological Survey, Books and Open File Reports, Box 25425, Federal Center, Denver, CO 80225.

U. S. EPA. 1993. Dilution Models for Effluent Discharges. Second Edition. EPA/ 600/ R 93/ 139. National Technical Information Service, Springfield, VA.

138  ix 0: Additivity and the Two Compoxmnta of a WER Determined Using Downstrwm Water The Concent of Additivitv of WERs In theory, whenever samples of effluent and upstream water are taken, determination of a WRR in 100 % effluent would quantify the effluent WRR ( eWER) and determination of a WRR in 100 % upstream water would quantify the upstream WRR ( uWRR); determination of WRRs in known mixtures of the two samples would demonstrate whether the eWER and the uWRR are additive. For example, if eWRR = 40, uWRR = 5, and the two WERs are additive, a mixture of 20 % effluent and 80 % upstream water would give a WER of 12, except possibly for experimental variation, because:

2O( emR) + 8O( umR) t 20( 40) + 80( S) t 800 + 400 100 100 100 dE! L~ 2 100 l Strict additivity of an eWRR and an uWER will probably be rare because one or both WERs will probably consist of a portion that is additive and a portion that is not. The portions of the ewER and uWER thut are due to refractory metal will be strictly additive, because a change in water quality will not make the metal more or less toxic. In contrast, metal that is nontoxic because it is complexed by a complexing agent such as RDTA will not be strictly additive because the % uncomplexed will decrease as the solution is diluted; the amount of change in the % uncomplexed will usually be small and will depend on the concentration and the binding constant of the complexing agent ( see Appendix D). Whether the nonrefractory portions of the UWER and eWRR are additive will probably also depend on the differences between the water quality characteristics of the effluent and the upstream water, because these will determine the water quality characteristics of the downstream water. If, for example, 85 % of the eWRR and 30 % of the uWER are due to refractory metal, the WRR obtained in the mixture of 20 % effluent and 80 % upstream water could range from 8 to 12. The WRR of 8 would be obtained if the only portions of the eWER and uWRR that are additive are those due to refractory metal, because:

ZO( O. 85) ( eMEN + 8OtO. 30) ( WER)  20( 0.85) ( 40) + SO( O. 30) ( 5) = 8 100 100 The WRR could be as high as` 12 depending on the percentages of the other portions of the WERs that are also additive. Even if the eWER and uWER are not strictly additive, the concept of additivity of WRRs can be useful insofar as the eWER and uWER are partially additive, i. e., insofar as a portion of at least one of the WERs is additive. In the example given above, the WER determined using downstream water that consisted of 20 % effluent 139 and 80 % upstream water would be 12 if the eWER and uwER were strictly additive; the downstream WER would be less than 12 if the eWER and uWER were partially additive.

The major advantage of additivity of WERs can be dmnstrated using the effluent and upstream water that were used above. TO sir& ify this illustration, the acute chronic ratio will be assumed to be large, and the eWEF! of 40 and the uWER of 5 will be assumed to be cccWERs that will be assumed to be due to refractory metal and will therefore be strictly additive. In addition, the coap? lete mix downstream water at design flow conditions will be assumed to be 20 8 effluent and 80 % upstream water, so that the downstream WER will be 12 as calculated above for strict additivity.

Because the ewER and the USER are cccWERs and are strictly additive, this metal will cause neither acute nor chronic toxicity in downstream water if ( a) the concentration of metal in the effluent is less than 40 times the CCC and ( b) the concentration of metal in the upstream water is less than 5 times the CCC. As the effluent is diluted by mixing with upstream water, both the ewER and the concentration of metal will be diluted simultaneously; proportional dilution of the metal and the eWER will prevent the metal from causing acute or chronic toxicity at any dilution. When the upstream flow equals the design flow, the WER in the plume will decrease from 40 at the end of the pipe to 12 at complete mix as the effluent is diluted by upstream water; because this WER is due to refractory metal, neither fate processes nor changes in water quality characteristics will affect the WER. When stream flow is higher or lower than design flow, the complete mix WER will be lower or higher, respectively, than 12, but toxicity will not occur because the concentration of metal will also be lower or higher.

If the eWER and the uWER are strictly additive and if the national CCC is 1 mg/ L, the following conclusions are valid when the concentration of the metal in 100 % effluent is less than 40 xng/ L and the concentration of the metal in 100 % upstream water is less than 5 mg/ L: 1. This metal will not cause acute or chronic toxicity in the upstream water, in 100 % effluent, in the plume, or in downstream water. 2. There is no need for an acute or a chronic mixing zone where a lesser degree of protection is provided. 3. If no mixing zone exists, there is no discontinuity at the edge of a mixing zone where the allowed concentration of metal decreases instantaneously. These results also apply to partial additivity as long as the concentration of metal does not exceed that allowed by the amount 140 of additivity that exists. It would be more difficult to take into account the portions of the eWER and uWER that are not additive.

The concept of additivity becomes unimportant when the ratios, concentrations of the metals, or WRRs are very different. For example, if eWER = 40, uWER = 5, and they are additive, a mixture of 1 % effluent and 99 % upstream water would have a WRR of 5.35. Given the reproducibility of toxicity tests and WERs, it would be extremely difficult to distinguish a WER of 5 from a WER of 5.35. In cases of extreme dilution, rather than experimentally determining a WER, it is probably acceptable to use the limiting WER of 5 or to calculate a WRR if additivity has been demonstrated.

Traditionally it has been believed that it is environmentally conservative to use a WER determined in upstream water ( i. e., the uWRR) to derive a site specific criterion that applies downstream ( i. e., that applies to areas that contain effluent). This belief is probably based on the assumption that a larger WER would be obtained in downstream water that contains effluent, but the belief could also be based on the assumption that the uWER is additive. It is possible that in some cases neither assumption is true, which means that using a uWRR to derive a downstream site specific criterion might result in underprotection. It seems likely, however, that WRRs determined using downstream water will usually be at least as large as the uWER.

Several kinds of concerns about the use of WERs are actually concerns about additivity: 1. Do WERs need to be determined at higher flows in addition to being determined at design flow? 2, Do WERs need to be determined when two bodies of water mix? 3. Do WERs need to be determined for each additional effluent in a multiple discharge situation. In each case, the best use of resources might be to test for additivity of WERs.

Mixinu Zones In the example presented above, there would be no need for a regulatory mixing zone with a reduced level of protection if: 1. The eWRR is always 40 and the concentration of the metal in 100 % effluent is always less than 40 mg/ L. 2. The uWER is always 5 and the concentration of the metal in 100 % upstream water is always less than 5 mg/ L. 3. The WERs are strictly additive. If, however, the concentration exceeded 40 mg/ L in 100 % effluent, but there is some assimilative capacity in the upstream water, a regulatory mixing zone would be needed if the discharge were to be allowed to utilize some or all of the assimilative 141 capacity. The concept of additivity of WERs can be used to calculate the maximum allowed concentration of the metal in the effluent if the eWER and the uWER are strictly additive.

If the concentration of metal in the upstream water never exceeds 0.8 mg/ L, the discharger might want'to determine how much above 40 mg/ L the concentration could be in 100 % effluent. If, for example, the downstream water at the edge of the chronic mixing zone under design flow conditions consists of 70 % effluent and 30 b upstream water, the WER that would apply at the edge of the mixing zone would be:

70( e) + 3o( uwim)  70( 40) + 30( s) I 2800 + 150 I 100 100 100 2g . 5 .

Therefore, the maximum concentration allowed at this point would be 29.5 lng/ L. If the concentration of the metal in the upstream water was 0.8 m/ L, the maximum concentration allowed in 100 % effluent would be 41.8 mg/ L because:

70( 41.8 a& L) l 3OfO. 8 n& L) I 2926 m/ q/ L + 24 I& 100 100 L = 29.5alg/ L*

Because the eWER is 40, if the concentration of the metal in 100 0 effluent is 41.8 mg/ L, there would be chronic toxicity inside the chronic mixing zone. If the concentration in 100 % effluent is greater than 41.8 mg/ L, there would be chronic toxicity past the edge of the chronic mixing zone. Thus even if the eWER and the uWER are taken into account and they are assumed to be completely additive, a mixing zone is necessary if the assimilative capacity of the upstream water is used to allow discharge of more metal.

If the complete mix downstream water consists of 20 % effluent and 80 8 upstream water at design flow, the complete mix WER would be 12 as calculated above. The complete mix approach to determining and using downstream WERs would allow a maximum concentration of 12 xng/ L at the edge of the chronic mixing zone, whereas the alternative approach resulted in a maximum allowed concentration of 29.5 mg/ L. The complete mix approach would allow a maximum concentration of 16.8 mg/ L in the effluent because:

70( 16.8 n& L) + 3OtO. 8 a& L) I 1176 m& L + 24 m& L  100 100 12 n& L.

In this example, the complete mix approach limits the concentration of the metal in the effluent to 16.8 mg/ L, even though it is known that as long as the concentration in 100 % effluent is leas than 40 mg/ L, chronic toxicity will not occur inside or outside the mixing zone. If the WER of 12 is used to derive a site specific CCC of 12 mg/ L that is applied to a site 142 that starts at the edge of the chronic mixing zone and extends all the way across the stream, there would be overprotection at the edge of the chronic mixing zone ( because the maximum allowed concentration is 12 mg/ L, but a concentration of 29.5 mg/ L will not cause chronic toxicity), whereas there would be underprotection on the other side of the stream ( because the maximum allowed concentration is 12 mg/ L, but concentrations above 5 mg/ L can cause chronic toxicity.)

The Rxnerimental Determination of Additivitv Experimental variation makes it difficult to quantify additivity without determining a large number of WERs, but the advantages of demonstrating additivity might be sufficient to make it worth the effort. It should be possible to decide whether the eWER and uWER are strictly additive based on determination of the eWER in 100 % effluent, determination of the uWER in 100 % upstream water, and determination of WERs in 1: 3, l: l, and 3: l mixtures of the effluent and upstream water, i. e., determination of WERs in 100, 75, 50, 25, and 0 % effluent. Validating models of partial additivity and/ or interactions will probably require determination of more WERs and more sophisticated data analysis ( see, for example, Broderius 1991).

In some cases chemical measurements or manipulations might help demonstrate that at least some portion of the eWER and/ or the uWRR is additive: 1. If the difference between the dissolved WER and the total recoverable WER is explained by the difference between the dissolved and total recoverable concentrations, the difference is probably due to particulate refractory metal. 2. If the WERs in different samples of the effluent correlate with the concentration of metal in the effluent, all, or nearly all, of the metal in the effluent is probably nontoxic. 3. A WEX that remains constant as the pH is lowered to 6.5 and raised to 9.0 is probably additive. The concentration of refractory metal is likely to be low in upstream water except during events that increase TSS and/ or WC; the concentration of refractory metal is more likely to be substantial in effluents. Chemical measurements might help identify the percentages of the eWRR and the uWER that are due to refractory metal, but again experimental variation will limit the usefulness of chemical measurements when concentrations are low.

The distinction between the two components of a WER determined using downstream water has the following implications: 1. The magnitude of a WER determined using downstream water will usually depend on the percent effluent in the sample.

143 2.

3.

4.

5.

6. Insofar as the eWRR and uWRR are additive,. the magnitude of a downstream WER can be calculated from the eWER, the UWER, and the ratio of effluent and upstream water in the downstream water. The derivation and implementation of site specific criteria should ensure that each component is applied only where it occurs. a. Underprotection will occur if, for example, any portion of the eWER is applied to an area of a stream where the effluent does not occur. b. Overprotection will occur if, for example, an unnecessarily small portion of the eWER is applied to an area of a stream where the effluent occurs. Even though the concentration of metal might be higher than a criterion in both a regulatory mixing zone and a plume, a reduced level of protection. is allowed in a mixing zone, whereas a reduced level of protection is not allowed in the portion of a plume that is not inside a mixing zone. Regulatory mixing zones are necessary if, and only if, a discharger wants to make use of the assimilative capacity of the upstream water. It might be cost effective to quantify the eWER and uWER, determine the extent of additivity, study variability over time, and then decide how to regulate the metal in the effluent.

Peference Broderius, S. J. 1991. Modeling the Joint Toxicity of Xenobiotics to Aquatic Organisms: Basic Concepts and Approaches. In: Aquatic Toxicology and Risk Assessment: Fourteenth Volume. ( M. A. Mayes and M. G. Barron, eds.) ASTM STP 1124. American Society for Testing and Materials, Philadelphia, PA. pp. 107 127.

144 Appendix EI: Spmcirl Conddorationm Concerning the Deteminatiozh of WHIl with Saltwater Spociem 1. The test organisms should be compatible with the salinity of the site water, and the salinity of the laboratory dilution water should match that of the site water. Low salinity stenohaline organisms should not be tested in high salinity water, whereas high salinity stenohaline organisms should not be tested in low salinity water; it is not known, however, whether an incompatibility will affect the WER. If the community to be protected principally consists of euryhaline species, the primary and secondary toxicity tests should use the euryhaline species suggested in Appendix I ( or taxonomically related species) whenever possible, although the range of tolerance of the organisms should be checked. a. When Method 1 is used to determine cmcWRRs at saltwater sites, the selection of test organisms is complicated by the fact that most effluents are freshwater and they are discharged into salt waters having a wide range of salinities. Some state water quality standards require a permittee to meet an LCSO or other toxicity limit at the end of the pipe using a freshwater species. However, the intent of the site specific and national water quality criteria program is to protect the communities that are at risk. Therefore, freshwater species should not be used when WRRs are determined for saltwater sites unless such freshwater species ( or closely related species) are in the community at risk. The addition of a small amount of brine and the use of salt tolerant freshwater species is inappropriate for the same reason. The addition of a large amount of brine and the use of saltwater species that require high salinity should also be avoided when salinity is likely to affect the toxicity of the metal. Salinities that are acceptable for testing euryhaline species can be produced by dilution of effluent with sea water and/ or addition of a commercial sea salt or a brine that is prepared by evaporating site water; small increases in salinity are acceptable because the effluent will be diluted with salt water wherever the cozununities at risk are exposed in the real world. Only as a last resort should freshwater species that tolerate low levels of salinity and are sensitive to metals, such as Danhnia mauna and Hvalella azteca, be used. b. When Method 2 is used to determine cccWERs at saltwater sites: 1) If the site water is low salinity but all the sensitive test organisms are high salinity btenohaline organisms, a commercial sea salt or a brine that is prepared by evaporating site water may be added in order to increase the salinity to the minimum level that is acceptable to the test organisms; it should be determined whether the 145 salt or brine reduces the toxicity of the metal and thereby increases the WER. 2) If the site water is high salinity; selecting test organisms should not be difficult because many of the sensitive test organisms are compatible with high salinity water.

2. It is especially important to consider the availability of test organisms when saltwater species are to be used, because many of the cwnly used saltwater species are not cultured and are only available seasonally.

3. Many standard published methodologies for tests with saltwater species reconzaen d filtration of dilution water, effluent, and/ or teat solutions through a 37 m sieve or screen to remove predators. Site water should be filtered only if predators are observed in the sample of the water because filtration might affect toxicity. Although recommended in some test methodologies, ultraviolet treatment is often not needed and generally should be avoided.

4. If a natural salt water is to be used as the laboratory dilution water, the samples should probably be collected at slack high tide (& 2 hours). Unless there is stratification, samples should probably be taken at mid depth; however, if a water quality characteristic, such as salinity or TSS, is important, the vertical and horizontal definition of the point of sampling might be important. A conductivity meter, aalinometer, and/ or transmissometer might be useful for determining where and at what depth to collect the laboratory dilution water; any measurement of turbidity will probably correlate with TSS.

5. The salinity of the laboratory dilution water should be within f 10 percent or 2 mg/ L ( whichever is higher) of that of the site water.

146 appendix It Su~ 3u; ftod Toxicity Tostn for Dotozminhg WBIU for Selecting primary and' secondary toxicity tests for determining WERs for metals should take into account the following: 1.

2.

3,

4. 5.

6.

7.

8.

9. WERs determined with more sensitive tests are likely to be larger than WERs determined with less sensitive tests ( see Appendix D). Criteria are derived to protect sensitive species and so WRRs should be derived to be appropriate for sensitive species. The appropriate regulatory authority will probably accept WRRs derived with less sensitive tests because such WERs are likely to provide at least as much protection as WRRs determined with more sensitive tests. The species used in the primary and secondary tests must be in different orders and should include a vertebrate and an invertebrate. The test organism ( i. e., species and life stage) should be readily available throughout the testing period. The chances of the test being successful should be high. The relative sensitivities of test organisms vary substantially from metal to metal. The sensitivity of a species to a metal usually depends on both the life stage and kind of test used. Water guality characteristics might affect chronic toxicity differently than they affect acute toxicity ( Spehar and Carlson 1984; Chapman, unpublished; Voyer and McGovern 1991). The endpoint of the primary test in laboratory dilution water should be as close as possible ( but must not be below) the CMC or CCC to which the WER is to be applied; the endpoint of the secondary test should be as close as possible ( and should not be below) the CMC or CCC. Designation of tests as acute and chronic has no bearing on whether they may be used to determine a cmcwER or a cccWRR. The suggested toxicity tests should be considered, but the actual selection should depend on the specific circumstances that apply to a particular WER determination.

Regardless of whether test solutions are renewed when tests are conducted for other purposes, if the concentrations of dissolved metal and dissolved oxygen remain acceptable when determining WRRs, tests whose duration is not longer than 48 hours may be static tests, whereas tests whose duration is longer than 48 hours must be renewal tests. If the concentration of dissolved metal and/ or the concentration of dissolved oxygen does not remain acceptable, the test solutions must be renewed every 24 hours. If one test in a pair of side by side tests is a renewal test, both of the tests must be renewed on the same schedule.

Appendix H should be read if WRRs are to be determined with saltwater species.

147 Suggested Tests 1 for Determining cmcWERs and cccWERs2 ( Concentrations are to be measured in all tests.)

JlY&!& d CRICWERS' CCCWERS' Netal Aluminum Arsenic( II1)

Cadmium Chrom( II1)

chmm( vx)

Capx= r Lead Mercury Nickel Selenium Silver Zinc X CDC Fw DA X FMC Em F MC X CDC GM FM AR X X Y Y FMC BMC Y X FMC BMC FMC BMC Fw SW DA BM GM CR CDC MYC SLs or FM CR Fw SW DA MY CDC MYC Fw GM SL or DA FMC Fw SW DA MY GM NE CDC MYC Tw SW DA EM FM or GM AR CDC BMC Fw SW GM MYC CDC MYC DA BM Fw SW DA MY GM EM Y Y Fw SW DA MY FX EM CDC MYC Fw SW Y CR Y MYC Y MYC Fw SW DA BM FMC CR CDC MYC Fw SW DA BM FM MY CDC MYC The description of a test specifies not only the test species and the duration of the test but also the life stage of the species and the adverse effect( s) on which the endpoint is to be based.

Some tests that are sensitive and are used in criteria documents are not suggested here because the chances of the test organisms being available and the test being successful might be low. Such tests may be used if desired.

148 FW= Fresh Water; SW = Salt Water.

Two letter codes are used for acute tests, whereas codes for chronic tests contain three letters and end in ' C'. One letter codes are used for cosunents.

In acute tests on cadmium with salmonids, substantial numbers of fish usually die after 72 hours. Also, the fish are sensitive to disturbance, and it is sometimes difficult to determine whether a fish is dead or immobilized.

AR. A 48 hr EC50 based on mortality and abnormal development from a static test with embryos and larvae of sea urchins of a species in the genus Arbacia ( ASTM 1993a) or of the species gtronwlocentrotus ournuratus ( Chapman 1992).

EM. A 48 hr EC50 based on mortality and abnormal larval development from a static test with embryos and larvae of a species in one of four genera ( Crassostrea, Mulinia, Mvtilus, flercenaria) of bivalve molluscs ( ASTM 1993b).

CR. A 48 hr EC50 ( or LC50 if there is no immobilization) from a static test with Acartia or larvae of a saltwater crustacean; if molting does not occur within the first 48 hours, renew at 48 hours and continue the test to 96 hours ( ASTM 1993a).

DA. A 48 hr EC50 ( or LC50 if there is no immobilization) from a static test with a species in one of three genera ( U. S. EPA?; 93 ( CeriodaD ' g DaDhnia SimoceDhalus) in the family Daphnidae a; ASTM i993a).

FM. A 48 hr LC50 from a static test at 25OC with fathead minnow . gro elas) larvae that are 1 to 24 hours old ( ASTM EPAm1993a) The embryos must be hatched in the laboratory dilution water, except that organisms to be used in the site water may be hatched in the site water. The larvae mumt not be fed before or during the test and at least 90 percent muat survive in laboratory dilution water for at least six days after hatch. Note: The following 48 hr LC5Os were obtained at a hardness of 50 mg/ L with fathead minnow larvae that were 1 to 24 hours old. The metal was measured using the total recoverable procedure ( Peltier 1993) :

CS LC50 ( us/ L) 13.87 Copper 6.33 Zinc 100.95 149 FX. A 96 hr LC50 from a renewal test ( renew at 48 hours) at 25OC with fathead minnow ( PimeDhaleS nro elas) larvae that are 1 to 24 hours old ( ASTM 1993a; U. S. EFA 1993a). The embryos atit be hatched.. in the laboratory dilution water, except that organisms to be used in the site water may be hatched in the site water. The larvae m; ut rrdt be fed before or during the test and at least 90 percent ru8t survive in laboratory dilution water for at least six days after hatch. Note: A 96 hr LC50 of 188.14 w/ L was obtained at a hardness of 50 mg/ L in a test on nickel with fathead minnow larvae that were 1 to 24 hours old. The metal was measured using the total recoverable procedure ( Peltier 1993). A 96 hr LC50 is used for nickel because substantial mortality occurred after 48 hours in the test on nickel, but not in the tests on cadmium, copper, and zinc.

GM. A 96 hr EC50 ( or LC50 if there is no izzaobilization) from a renewal test ( renew at 48 hours) with a species in the genus Gamnaw ( ASTM 1993a).

MY. A 96 hr EC50 ( or LC50 if there is no inznobilization) from a renewal test ( renew at 48 hours) with a species in one of two . . genera ( JWsidoDsis, Holmeg& mvsis [ nee Acanthomvsisl) in the family Mysidae ( U. S. EPA 1993a; ASTM 1993a). Feeding is required during all acute and chronic tests with mysids; for determining WEF&, mysids should be fed four hours before the renewal at 48 hours and minimally on the non renewal days.

NE. A 96 hr LCSO from a renewal test ( renew at 48 hours) using juvenile or adult polychaetes in the genus Nereidae ( ASTM 1993a).

SL. A 96 hr EC50 ( or LC50 if there is no immobilization) from a renewal test ( renew at 48 hours) with a species in one of two salmo) in the family Salmonidae ( ASTM CHRONIC m BMC. A 7 day IC25 from a survival and development renewal test ( renew every 48 hours) with a species of bivalve mollusc, such as a species in the genus Mulinig. One such test has been described by Burgess et al. 1992. [ Note: When determining WEBS, sediment must not be in the test. chamber.] [ Note: This test has not been widely used. 1 CDC. A 7& y IC25 based on reduction in survival and/ or reproduction in a renewal test with a species in the genus in the family Daphnidae ( U. S. EPA 1993b). The 150 test solutions mum+ be renewed every 48 hours. ( A 21 day life cycle test with DaDhI'& manna is also acceptable.,)

FMC. A 7 day IC25 from a survival and growth renewal test ( renew every 48 hours) with larvae ( s 48 hr old) of the fathead minnow ( PimeDhales promela%) ( U; S. EPA 1993b). When determining WERs, the fish mumt be fed four hours before each renewal and minimally during the non renewal days.

MYC. A 7 day IC25 based on reduction in survival, growth, and/ or reproduction in a renewal test with a species in one of two genera ( Mvs'd DS'S family Mysiiag ( i. k. Holmesimvsis [ nee Acanthomvsis)) in the EPA 1993c). Mysids mwt be fed during all acute and chronic tests; when determining WERs, they must be fed four hours before each renewal. The test solutions must be renewed every 24 hours.

NEC. A 20 day IC25 from a survival and growth renewal test ( renew every 48 hours) with a species in the genus Neanthes ( Johns et al. 1991). [ Note: When determining WERs, sediment mwt not be in the test chamber.] [ Note: This test has not been widely used.]

COMMENTS X. Another sensitive test cannot be identified at this time, and so other tests used in the criteria document should be considered.

Y. Because neither the CCCs for mercury nor the freshwater criterion for selenium is based on laboratory data concerning toxicity to aquatic life, they cannot be adjusted using a WER.

REFERENCES Asm. 1993a. Guide for Conducting Acute Toxicity Tests with Fishes, Macroinvertebrates, and Amphibians. Standard E729. American Society for Testing and Materials, Philadelphia, PA.

ASTM. 1993b. Guide for Conducting Static Acute Toxicity Tests Starting with Embryos of Four Species of Saltwater Bivalve Molluscs. Standard E724. American Society for Testing and Materials, Philadelphia, PA.

Burgess, R., G. Morrison, and S. Rego. 1992. Standard Operating Procedure for 7 day Static Sublethal Toxicity Tests for Mulinia lateralis. U. S. EPA, Environmental Research Laboratory, Narragansett, RI.

151 Chapman, G. A. 1992. Sea Urchin ( Stronwlocent otus Fertilization Test Method. U. S. EPA, Newport, f; R.  1 Johns, D. M., R. A. Pastorok, and T. C. Ginn. 1991. A Sublethal Sediment Toxicity Test using Juvenile EJeantheg sp. ( Polychaeta: Nereidae) . In: Aquatic Toxicology and Risk Assessment: Fourteenth Volume. ASTM STP 1124. ( M. A. Mayes and M. G. Barron, eds.) American Society for Testing and Materials, Philadelphia, PA. pp. 280 293.

Peltier, W. H. 1993. Memorandum. to C. E. Stephan. October 19.

Spehar, R. L., and A. R. Carlson. 1984. Derivation of Site Specific Water Quality Criteria for Cadmium and the St. Louis River Basin, Duluth, Minnesota. Environ. Toxicol. Chem. 3: 651 665.

U. S. EPA. 1993a. Methods for Measuring the Acute Toxicity of Effluents and Receiving Waters to Freshwater and Marine Organisms. Fourth Edition. EPA/ 600/ 4 90/ 027F. National Technical Information Service, Springfield, VA.

U. S. EPA. 1993b. Short term Methods for Estimating the Chronic Toxicity of Effluents and Receiving Waters to Freshwater Organisms. Third Edition. EPA/ 600/ 4 91/ 002. National Technical Information Service, Springfield, VA.

U. S. EPA. 1993c. Short term Methods for Estimating the Chronic Toxicity of Effluents and Receiving Waters to Marine and Estuarine Organisms. Second Edition. EPA/ 600/ 4 91/ 003. National Technical Information Service, Springfield, VA.

Voyer, R. A., and D. G. McGovern. 1991. Influence of Constant and Fluctuating Salinity on Responses of WsidoDsis bahia Exposed to Cadmium in a Life Cycle Test. Aquatic Toxicol. 19: 215 230.

152 nix 3: Rmc mod Saltm of Hotala The following salts are recommended for use when determining a WEH for the metal listed. If available, a salt that meets American Chemical Society ( ACS) specifications for reagent grade should be used.

Aluminum * Aluminum chloride ( j hydrate: AlCl,= 6H2O Aluminum sulfate 18 hydrate: A12( S0,) 1* 18H, 0 Aluminum potassium sulfate la hydrate: AlK( S0,),= 12HI0 ite: NaAs02 enate ' I hydrate, dibasic: Na2HAs0,= 7H20  Cadmium chloride 2.5~ hydrate: CdC12= 2.5H20 Cadmium sulfate hydrate: 3CdS0,= 8H20 . Sk0 um( I * Chrtkic % oride 6 hydrate ( Chromium chloride): CrCl,= 6HiO * Chromic nitrate g hydrate ( Chromium nitrate): Cr( NO,),* 9& 0 Chromium potassium sulfate 12 hydrate: CrK( S0,12* 12H20 Potassium chromate: K2 04 Potassium dichromate: K, Cr, O, * Sodium chromate I hydrate:  Na2Cr0, 4H, O Sodium dichromate 2 hydrate: Na2Cr20,= 2H20 izQRRss * Cupric Cupric Cupric chloride a hydrate ( Copper chloride): CuC12* 2Ha0 nitrate 2.5 hydrate ( Copper nitrate): CU( NO,)~~~.~ H~ O sulfate 5 hydrate ( Copper sulfate): CuS0,* 5H20 * Lead chloride: PbC12 Lead nitrate: Pb( NOB) 2 Pe CUT Mrrcuric chloride: HgC12 Mercuric nitrate monohydrate: Hg( N03) 2* H20 Mercuric sulfate: HgSO,

153 Nickel l Nickelous chloride 6 hydrate ( Nickel chloride) : NiC12= 6H20 l Nickelous nitrate 6 krydrate ( Nickel nitrate): Ni( NO,),* 6H, O Nickelous sulfate 6 hydrate ( Nickel sulfate): NiS0,= 6H20.

* Sodium selenite 5 hydrate: Na, SeO, 5H, O * Sodium selenate lo hydrate: Na2Se0,= 10H20 Silver Silver nitrate: AgNO, ( Even if acidified, standards and samples containing silver rwt be in amber containers.)

Zinc Zinc chloride: ZnClr * Zinc nitrate 6 hydrate: Zn( N0,12* 6H, 0 Zinc sulfate 7 hydrate: ZnS0,= 7H20 * Note: ACS reagent grade specifications might not be available for this salt.

No salt should be used until information concerning the safety and handling of that salt has been read.

154


[Federal Register Volume 76, Number 210 (Monday, October 31, 2011)]
[Notices]
[Page 67167]
[FR Doc No: 2011-28111]

-----------------------------------------------------------------------

DEPARTMENT OF ENERGY

Federal Energy Regulatory Commission

[Project Nos. 14172-000, 14178-000 and 14190-000]

Riverbank Hydro No. 6 LLC, Lock Hydro Friends Fund XLVII, FFP
Project 52 LLC; Notice of Competing Preliminary Permit Applications
Accepted for Filing and Soliciting Comments, Motions To Intervene, and
Competing Applications

    On May 2, 2011, Riverbank Hydro No. 6 LLC (Riverbank), and on May
3, 2011, Lock Hydro Friends Fund XLVII (Lock Hydro), and FFP Project 52
LLC (FFP 52) filed preliminary permit applications, pursuant to section
4(f) of the Federal Power Act, proposing to study the feasibility of a
hydropower project at the U.S. Army Corps of Engineers' (Corps) Toad
Suck Ferry Lock and Dam, located on the Kentucky River, in Perry and
Faulkner Counties, Arkansas. The sole purpose of a preliminary permit,
if issued, is to grant the permit holder priority to file a license
application during the permit term. A preliminary permit does not
authorize the permit holder to perform any land-disturbing activities
or otherwise enter upon lands or waters owned by others without the
owners' express permission.
    Riverbank's Project No. 14172-000 would consist of: (1) A forebay;
(2) an intake structure; (3) a powerhouse containing two generating
units with a total capacity of 37.8 megawatts (MW); (4) a tailrace
structure; and (5) a 12.6-mile-long, 69 kilo-volt (kV) transmission
line. The project would have an estimated average annual generation of
78.8 gigawatt-hours (GWh), and operate run-of-river utilizing surplus
water from the Toad Suck Ferry Lock and Dam, as directed by the Corps.
    Applicant Contact: Mr. Kuo-Bao Tong, Riverbank Power Corporation,
Royal Bank Plaza, South Tower, P.O. Box 166, 200 Bay Street, Suite
3230, Toronto, ON, Canada M5J2J4.<PHONE_NUMBER> x 154.
    Lock Hydro's Project No. 14178-000 would consist of: (1) Three lock
frame modules, each frame module will be 109-feet long, 40-feet-high
and contain ten generating units with a total combined capacity of 22.5
MW; (2) a new switchyard containing a transformer; and (3) a proposed
2.0-mile-long, 115 kV transmission line to an existing distribution
line. The proposed project would have an average annual generation of
147.926 GWh, and operate run-of-river utilizing surplus water from the
Toad Suck Ferry Lock and Dam, as directed by the Corps.
    Applicant Contact: Mr. Wayne F. Krouse, Hydro Green Energy, 5090
Richmond Avenue 390, Houston, TX 77056.<PHONE_NUMBER> x 709.
    FFP 52's Project No. 14190-000 would consist of: (1) An 900-foot-
long, 300-foot-wide approach channel; (2) a powerhouse, located on the
west side of the dam, containing four generating units with a total
capacity of 40.0 MW; (3) a 1,560-foot-long, 320-foot-wide tailrace; (4)
a 7.2/69 KV substation; and (5) a 1.25-mile-long, 69 kV transmission
line. The proposed project would have an average annual generation of
160.0 GWh, and operate run-of-river utilizing surplus water from the
Toad Suck Ferry Lock and Dam, as directed by the Corps.
    Applicant Contact: Ms. Ramya Swaminathan, Free Flow Power Corp.,
239 Causeway Street, Suite 300, Boston, MA 02114.<PHONE_NUMBER>.
    FERC Contact: Michael Spencer<EMAIL_ADDRESS>(202) 502-
6093.
    Deadline for filing comments, motions to intervene, competing
applications (without notices of intent), or notices of intent to file
competing applications: 60 days from the issuance of this notice.
Competing applications and notices of intent must meet the requirements
of 18 CFR 4.36. Comments, motions to intervene, notices of intent, and
competing applications may be filed electronically via the Internet.
See 18 CFR 385.2001(a)(1)(iii) and the instructions on the Commission's
Web site http://www.ferc.gov/docs-filing/efiling.asp. Commenters can
submit brief comments up to 6,000 characters, without prior
registration, using the eComment system at http://www.ferc.gov/docs-filing/ecomment.asp. You must include your name and contact information
at the end of your comments. For assistance, please contact FERC Online
Support. Although the Commission strongly encourages electronic filing,
documents may also be paper-filed. To paper-file, mail an original and
seven copies to: Kimberly D. Bose, Secretary, Federal Energy Regulatory
Commission, 888 First Street, NE., Washington, DC 20426.
    More information about this project, including a copy of the
application, can be viewed or printed on the ``eLibrary'' link of the
Commission's Web site at http://www.ferc.gov/docs-filing/elibrary.asp.
Enter the docket number (P-14172-000, P-14178-000, or P-14190-000) in
the docket number field to access the document. For assistance, contact
FERC Online Support.

    Dated: October 25, 2011.
Kimberly D. Bose,
Secretary.
[FR Doc. 2011-28111 Filed 10-28-11; 8:45 am]
BILLING CODE 6717-01-P