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7 Streams and Lakes fames R. Kramer, Anders W. Andren, Richard A. Smith Arthur H. Johnson, Richard B. Alexander, and Gary OehZert INTRODUCTION Various attempts have been made to assess changes in the acidity of surface waters that may have occurred over the past 50 years. An excellent review of the methods used in the analyses as well as the basic conclusions from a large number of investigations may be found in the EPA's Critical Assessment Review Papers (Environmental Protection Agency 1984). For example, Henriksen (1979, 1980, 1982), Almer et al. (1978), Dickson (1980), Christophersen and Wright (1981), Thompson (1982), Galloway et al. (1983a), and Wright (1983) used basic geochemical concepts to derive general models that are designed to predict both the sensitivity and the degree of response of a water body to acid inputs. More complex mechanistic and time-dependent models that are intimately synchronized to hydrological, biological, and geomorpho- logical factors have also been used (for example, Chen et al. 1982, Schnoor et al. 1982, Booty and Kramer 1984). These models have provided much useful information, but currently none can provide detailed, quantitative assess- ments of past or future trends in lake and stream acidification from acid deposition. (For an evaluation of the strengths and limitations of these models see Church (1984).) Another approach has been to compare the acidic status of lakes, rivers, and ponds in North America using his- torical and current data on pH, alkalinity, and sulfate. Studies in the United States include, for example, those of Schofield (1976) for New York lakes, Davis et al. (1978) for lakes in Maine, Johnson (1979) for streams in New Jersey, Pfeiffer and Festa (1980) for lakes in New York, Hendrey et al. (1980) and Burns et al. (1981) for 231

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232 headwater streams in North Carolina and New Hampshire, Arnold et al. (1980) for Pennsylvania streams, Crisman et al. (1980) for lakes in northern Florida, Norton et al. (1981) for lakes in New Hampshire, Maine, and Vermont, Lewis (1982) for lakes in Colorado, Haines et al. (1983) for lakes in New England, and Eilers et al. (1985) for lakes in northern Wisconsin. Similar studies have also been published on lakes in Canada (Beams and Harvey 1972, Beamish et al. 1975, Dillon et al. 1978, Watt et al. 1979, Thompson et al. 1980) and Scandinavia (Wright and Gjessing 1976, Rebadrosf 1980). These studies analyzed water quality data using a variety of methods. Some sets include continuous data records and others only data at discrete points in time. Many of the analyses account for such complicating factors as changes in analytical methods and land use practices. The conclusions from a variety of studies are that at least some lakes exhibit a decrease in pH and alkalinity when historical and recent data are compared. Similarly, a review by Turk (1983) compares historical and geographical trends in the acidity of precipitation with trends in North American surface waters. Based on the synthesis of a variety of data collected during discrete time periods (mainly the 1930s and late 1970s) and examined by Schofield (1976), Pfeiffer and Festa (1980), Davis et al. (1978), Norton et al. (1981), Burns et al. (1981), and Haines et al. (1983), he concludes that decreases in pH and alkalinity have occurred in lakes in New York and New England. In addition, water quality data are available from the U.S. Geological Survey's Hydrologic Bench-Mark Network. These data, collected since the middle 1960s, provide a continuous record of high-quality data for pa, alkalinity, and sulfate as well as the major_cations and anions for small watersheds in the United States. Smith and Alexander (1983) have studied these records and conclude that "In the northeastern quarter of the country, SO2 emissions have decreased over the past 15 years and the trends in the cited chemical characteristics (alkalinity and sulfate) of Bench-Mark streams are consistent with a hypothesis of decreased acid deposition in that region. Throughout much of the remainder of the country, SO2 emissions have increased and trends in stream sulfate, alkalinity and alkalinity/total cation ratios are con- sistent with a hypothesis of increased acid deposition. There are, however, several acknowledged problems in the interpretation of "historical n lake water chemistry

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233 data. Perhaps the chief drawback is the lack of docu- mentation of sampling and analytical procedures. A major problem encountered in our study has been the lack of sufficient documentation, before 1950, of the procedures for calculating pH and alkalinity. For example, whereas the current method for measuring pH and alkalinity employs standard electronic instruments, before about 1950 the method of choice was calorimetry using indicator dyes, which, when added in small quantities to solutions of unknown pH, cause color changes at specific pHs. Methyl orange, a commonly used indicator for measuring the alkalinities of lake waters in the 1930s, undergoes subtle color changes over a range of pH values, and the exact endpoint used by analysts is often not clearly specified in the historical records. In this chapter, we review some of these records and assess trends based on the most likely endpoints employed in the historical studies. Furthermore, for reasons that are described in detail in Appendix D, titration of lake water solutions to a methyl orange endpoint requires a correction for "over- titration." Assumptions about the magnitude of the correction are especially important when analyzing low- alkalinity waters. In most previous studies, researchers have assumed that the correction factor is a constant independent of the original alkalinity. Values ranging from zero to greater than 100 microequivalents per liter (peq/L) have been employed in such studies. A uniform method for analyzing these types of samples is highly desirable, and one goal of this chapter is to present a method for determining the historical corrections by calculations rather than by assumption. The method accounts for the fact that the relation between the correction factor and alkalinity is sensitive to the alkalinity of the test solution. In addition, the method provides a means for checking the internal consistency of the data set. Additional factors that must be considered in inter- preting current and historical data include (1) clima- tology, since the hydrology of lakes and rivers may vary considerably from year to year in response to fluctuations in precipitation and climatic factors affecting evapora- tion, (2) regional patterns of emissions of acid pre- cursors, since natural waters downwind of sources may be expected to exhibit greater changes in water quality than those situated upwind, (3) changes in land use (although Drably et al. (1980) found that land use changes in Norway did not seem to be related to regional acidifica-

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234 tion patterns, local changes in alkalinity and pH regimes were evident), (4) the statistical validity of using data separated by 20 to 40 years to examine trends, and (5) representativeness of data sets for various regions. Inferences about trends obtained from data on a limited number of lakes within a region should reflect the distribution of values of alkalinity and pH (and other physical, chemical, and biological parameters) for the whole region. In the current study, we have tried to overcome some of the difficulties mentioned above by examining temporal and spatial variations in several water quality parameters--pH, alkalinity, sulfate, and selected major anions and cations--that simultaneously give some information about the acidic status of lakes. The analysis is restricted to a few sets of data of high quality selected to provide both a determination of regional trends in lake and stream acidification and an illustration of a consistent methodologic approach. In this chapter, we first examine the hypothesis that sulfate levels in lakes and wet precipitation in the northeastern United States and eastern Canada are correlated. Many concepts of lake acidification focus on the titration of a surface water body by sulfuric acid deposited from the atmosphere (e.g., Henriksen 1980). The result of this simple titration is to substitute sulfate ion for bicarbonate ion, thus lowering the alkalinity and the pH of the affected surface water. Wright (1983, 1984) has shown a linear correlation between averages of sulfate concentrations for selected lakes and average sulfate concentration in precipitation in eastern North America, and Thompson and Hutton (1985) show a similar relationship in eastern Canada. To link sulfate ion concentration in atmospheric deposition to the sulfate ion concentration of surface waters, a direct relationship must exist between atmo- spheric and lake sulfate fluxes for a steady state assumption. A fundamental condition for this relation- ship to be valid is that there be no overriding internal sources or sinks of sulfate ions in the watersheds. If such sources and sinks do occur, then the method of averaging sulfate concentrations of large sets of lakes may mask the large variability in the data, particularly if averaging cancels factors equal in magnitude but opposite in sign. We examine data from a set of 626 lakes and assign to each a value of precipitation sulfate flux. We draw general conclusions about spatial

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235 associations between precipitation and lake sulfate fluxes from a statistical analysis of both the aggregated data and the data subdivided by regions. Next we address sulfate-driven changes in stream-water chemistry, using data from the U.S. Geological Survey Bench-Mark Network. The network provides a continuous record of monthly data on major ions in stream water beginning in the middle 1960s. Data are from pre- dominately undeveloped watersheds in 37 states. Con- tinuous sampling and unchanged analytical methods for the period make the data particularly well suited for examining atmospheric influences on water quality as a function of spatial and temporal SO2 emission patterns. The study of sulfate-driven changes in surface water chemistry is accomplished with the aid of a general paradigm that links acidic deposition to changes in the water quality in watersheds with acidic soil. Finally, we compare pH and alkalinity data of lakes taken at discrete time intervals, i.e., from the 1920s and 1930s to the 1970s and l980s. Three lake surveys were chosen for analysis; two in the northeastern United States (New York and New Hampshire) and one in the upper Midwest (Wisconsin). These lakes were selected because they provide, in our judgment, some of the most complete historical and recent data available, allowing us to test them for internal consistency according to the protocol of Kramer and Tessier (1982). We present this analysis to illustrate a chemically rigorous interpretation of historical pH and alkalinity measurements and to estimate the magnitude of changes in the acidic status of these lakes that may have occurred over the past 50 to 60 years. In our approach we limited our data to those lakes for which historical data on pH, alkalinity, and free CO2 acidity were available. The availability of simultaneous data for the three parameters allows us to perform a check for internal consistency of the data as an indicator of their reliability. Without this constraint a larger historical data base is available over wider geographic areas. Although our method allows us to screen the data for internal consistency, a larger (unscreened) data base may provide more expansive regional information and be amenable to more powerful statistical tests, although the data may be less reliable and hence the results of the tests less meaningful. Differences that may exist between the results obtained in this chapter and elsewhere must be viewed in light of the assumptions made and how the assumptions affect the selection and analysis of the data.

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236 SULFATE FLUXES IN PRECIPITATION AND LAKES In this section, we test the hypothesis that the fluxes of sulfate into lakes from atmospheric deposition generally determine the fluxes of sulfate outflow from lakes over a broad geographical area. Such a relationship must be demonstrated to exist before we can assume with any confidence that water chemistry and related data on fisheries records and lake sediment stratigraphy (de- scribed in the following chapters) are accurate indicators of atmospheric acid deposition. Studies of the mass balance of sulfate in a few water- sheds have been reported for which parameters such as atmospheric deposition, streamflow, and geology are sufficiently well determined to permit more detailed analyses (e.g., the Hubbard Brook watershed, Likens et al. 1985; and three lakes in the Adirondacks, Galloway et al. 1983b). The studies conclude that sulfate concentra- tions in these waters are controlled by atmospheric deposition of sulfur. In a different type of analysis, Wright (1983) found an approximately linear relationship between the mean concentration of sulfate in 15 groups of lakes in North American and mean excess (i.e., nonmarine) sulfate in wet precipitation. In our analysis, we compare sulfate input fluxes in wet deposition with the estimated sulfate output fluxes from 626 lakes in New York, New England, Quebec, Labrador, and Newfoundland. Fluxes of sulfate from wet deposition were obtained by multiplying the concentration of sulfate in rain by the intensity of rainfall (meters/year). Lake output fluxes of sulfate were estimated as the product of sulfate concentration in lake water and the intensity of net precipitation (i.e., meters/year measured as rainfall minus evaporation). (See Appendix B for specific methods and references.) It was possible to perform a detailed statistical analysis to test the relation between sulfate inputs and outputs for linearity (or some other monotonic relationship) and regional variance on subdividing the area into smaller regions. Such an analysis suited the particular needs of this study, which addresses the question of broad regional patterns in eastern North America. However, we note that this analysis is subject to uncertainty because for each of these lakes we generally lack important information on dry deposition, transpiration, observed streamflow from the watershed, soil chemistry, and processes within the watershed that are sources of sulfur.

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237 Results Figure 7.1(a) is a plot of sulfate inputs to 626 faxes from wet-only deposition as a function of annual sulfate outputs from the lakes based on data for 1980-1982. It is apparent that there is a greater variability in lake sulfate outflow fluxes than in precipitation sulfate fluxes. The scatter is greatest for lakes in Massachusetts (Figure 7.1(b)) owing to some extreme outliers and is of similar, smaller magnitude for other regions. In some cases, the output flux of sulfate from [ekes is greater than the flux from wet precipitation, suggesting that significant sources of sulfur other than from wet deposition exist in these watersheds. In other cases, the reverse is apparently true, perhaps suggesting that atmospheric sulfur may be retained in the watershed, or that sulfate is chemically reduced in the lake and deposited in the sediment. The apparent lake sulfate output flux normalized to the wet-only precipitation sulfate input flux is analyzed by geographic region in Table 7.1. Deficiencies in lake sulfate outputs compared with inputs from wet-only deposition Generally occur in remote regions such as Newfoundland and Labrador, whereas excesses occur in more populated areas such as New York, Connecticut, and Massachusetts. The origins of the deficiencies and excesses are not known on a lake-specific basis, but dry deposition, watershed processes involving sulfur cycling, and other sulfur sources of unknown origin are expected to be important contributors. Analysis In an effort to stabilize the variability in the sulfate output fluxes of lakes, the data were analyzed on the logarithmic (base e) scale (Figure 7.2). As a whole the data do not follow the y = x line (i.e., the case in which sulfate outputs equal sulfate inputs). Linear regression yields an estimate of a line log y = a log x + mi' in which the intercept, ml', is -0.306 (standard error, 0.041) and the slope, a, is 1.593 (standard error, 0.046). The slope of this line is significantly different from one, so the overall relationship on the natural scale, y = mixa in which mi is +0.736 (antilogarithm (base e) of ml') and a is 1.593, is significantly different from linear.

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238 1 5 14 13 12 11 ., 10 Q 9 ° 8 J 7 IL u~ 6 ~ 5— u7 4- y 3 2 1 o o o o ° 0 °~o o~ ~; O 0 ~ ~o 0.5 (a) o o o o o8° - ° :~10 0 fi~°C~ O 0 0 O 0 0 ° o o e} c ~ 0 3~ ~ _ '~ ~p ~ 0 0 ~uO - : o 1.5 2.5 WET PRECIPITATION SULFATE FLUX (input) 9 m~2 yr~1 36 - 34 - 32 - 30- E 28- ~ 26— Q 241 O 22— x 20- 18— 16- 1 4— ~ 12- y 10- 8— 6- 4 4.5 o 0 0 0 oB o o o ° ° ° C~ O o 2- 1 1 1 1 1 r 2.1 2.3 2.5 2.7 (b) o c~ o o 0 0 2.9 3.1 WET PRECIPITATION SULFATE FLUX (input) 9 m~2 yr 1 FIGURE 7.1 (a) Comparison of sulfate input fluxes to lakes from wet-only deposition with sulfate output fluxes from lakes in the northeastern United States and eastern Canada. (b) A similar plot for lakes in Massachusetts only.

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239 o U: ._ o a~ ~ Ct ~Q c~.. .C) Ct o q:: ~ O au U) ~ _ ~ ^ V) .= 3 ~ ' ·_ Ce ~ G =,, .~ O Ct ~ p~,, o z o _ _ .= · C~ .c · ~ Ce C~ (D 8< ~ o E~ Z ~ (,, ~ .= m ~ ~ zo O C) ~L) _ ~ . _ ·_ e~ := C) X ~ C) C~ ~ mX 3 o A Ch X o Vl oo X V o o Vl ~ X V U~ - U~ - Vl ~o X V o - o - Vl ~ X V o V~ o Vl X V o o Vl ~ X V O. O. Vl X Vl ol ol v - x ~0 oD 0.4 'e '~ c~ ~ ko ~n ~ ~ ~ cr 0 oo ~ - - ~ ~ - - - ~ o o ~ - - - ~ o o oX ~ ~ ~ - x - - t - ) ~ ~ ~ ~ — ~D ~ ~ ~ ~t {~) — c~ - ) - - - ~ u~ ~ ~ ~) - ~ x ~ ~ ~ ~ ~ oo - - ~ ~ - - ~ - oC o o ~ - o ~ ~ ~ x ~ o ~ ~ x - ~ 0 ~ t~ } ) ~ ~ ~ ~D ~ O ~ ~ - - os ~ ~ ox ~ - ~ ~ ~ ~ ~ ~ t - ~ ') ') - O' m~ = - - O ~ r 0 ~ _ _ _ cM r~~ ~ 0 Cx ~o 0 ~ U~ ~ _ _ ~ ~ —) ~ _ ~ _ ~ ~ — O O ~ ~ ~ — m_ _ _ ~ _ ~ ~ cr~ r ~c v~ 0 0 — ~ ~ o~ _ X ~ ~ ~ _ - ] _ t ~ _ ~ oo _ 0 o0 r~ cx~ _ ~ _ ~ —1 oo 0 x O == — '] _ .= '_ ~0 O ~ C) mm ~ ~ ~ z ~ z ~ z ~ ~ 3 ~ 9 ~ ~ ~ =g, ~ e 2 E . s ~ ' 2 s s ~ u E e ~ c 0 9 u ° u =°gEc o~!c ^c 0 0 ~ s ~ sc 0= ~ 9 9-sE 9 ~ ,, 3 3 3 D U ~ — O O E ,., v s O ~ ~ u 9 s u ~ 2 0 s c E E -3 c ,~ _= _ 5 3 c ,~ 2 -e ~ ; ~ ~ ~ ~-2g-~!~-g~Ei~ce

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240 41 1 @ 3 r x 2 CL 3 1 a: J ~ O UJ A a: J o 1 o . . -2 · — . |.'r~ . · _ . _3 ~ 1 1 1 -0.5 0.0 0.5 1.0 1.5 LOG OF WET PRECIPITATION SULFATE INPUT FLUX FIGURE 7.2 Comparison of the logarithm (base e) of sulfate fluxes to lakes from wet-only deposition with the logarithm (base e) of sulfate fluxes from lakes in the northeastern United States and eastern Canada. Solid line is the y = x line (i.e., condition in which sulfate flux from the lakes equals the sulfate flux to the lakes from wet-only deposition); dashed line is the line estimated from linear regression with slope of 1.593. As noted earlier there is a tendency for lakes in some regions to lie above the line of equal fluxes (excess lake sulfate flux in Table 7.1) and other regions to lie below it (excess precipitation sulfate flux in Table 7.1). This tendency may result from regional differences in geology, dry deposition, or other factors causing the regional ratios of input to output fluxes to differ from the overall relationship. To account for regional differences we performed an analysis of covariance in which we allowed each region to have its own intercept, but fitted the same slope for all regions. To analyze the data on a regional basis, we have chosen the following six regions: Connecticut, Massachusetts, and Rhode Island (CT-MA-RI); Maine, Vermont, and New Hampshire (ME-VT-NH); New York (NY); Newfoundland (NF); Labrador (LB); Quebec (QU). The linear regressions shown for each region in Figure 7.3 yield the following estimates:

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241 Region Intercept (mi') Standard Error CT~ RI 0.691 ME—VT—NH 0 . 2 57 NY 0.280 NF -0.004 LB -0 .515 QU 0.098 Slope (a) 1.089 0.0791 0.0729 0.1082 0.0626 0.0407 0.0781 0.0712 ~ The improvement in fit in going from a single regres- sion line to the six regional lines is highly significant, having a p value less than 10-6. Furthermore, the common slope (a) is not significantly different from one. Thus, there is no evidence that the relationship on the natural scale (y = mica) is nonlinear once regional differences have been taken into account. The multipliers mi' are regionally specific. Thus lake sulfate output flux y is estimated to be a multiple of precipitation sulfate input flux x. (Note, that on the natural scale, our model will not in general estimate the arithmetic mean of the lake sulfate flux.) The multiple depends on such factors as the ratio of total deposition to wet-only deposition, the rate at which sulfur is demobilized in the watershed or sediments, the magnitude of sulfur sources in the watershed, and the magnitude of any bias in our estimate of wet-only sulfate flux. The values of mi may be calculated from the antilogarithms of the intercepts (mi') shown in Figure 7.3. The values are as follows: 2.00 (CT-MA-RI), 1.32 (NY), 1.29 (ME-VT-NH), 1.10 (QU), 1.00 (NF), and 0.60 (LB). These multiples are greatest in regions close to strong sulfur source areas and smallest in remote areas, suggesting that dry deposition plays a major role. Since the 0.60 value for Labrador indicates that output fluxes are smaller than input fluxes, watershed processes that retain sulfate may be important. To obtain more detailed information of sulfur mass balance from these data, each location and perhaps each watershed needs to be considered separately. In some cases internal sources of sulfate may be important contributing factors to the sulfur mass balance (Wagner et al. 1982). For example, in Table 7.2, the geological setting is described for the locations of lakes in Massachusetts and Maine having large excesses of lake sulfate fluxes. Since strike zones, amphibolites,

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289 literature of the pH of this endpoint range between about 4.0 to higher than 4.3. The color of the endpoint in this range has been variously described as "faint pink," "pink," and "salmon pink." The Wisconsin survey provides the best documentation of the pH of the MO endpoint, citing a range of pH values from 4.06 to 4.18 for the "fainter" pink color and recommending that a value of 4.18 be used in calculating titrations. Our analysis shows that changes in alkalinity and pH in New York lakes are sensitive to the assumption made about the pH of the MO endpoint. Assuming an endpoint pH of 4.2 (or greater), New York lakes, on average, appear to have increased substantially in acidic status over the past 50 years, as reflected in reductions in both alkalinity and pH. Alternatively, if a value of the pH endpoint close to 4.0 is assumed and if the historical data are compared with the 1984 data set, then there appears to have been little change, on average, in the acidic status of New York lakes over the past 50 years. However, if the historical data are compared with the 1980 New York data set, there may have been an overall decline in alkalinity. New Hampshire lakes do not appear to be so sensitive to the choice of tne endpoint pH of MO. If an endpoint pH close to 4.2 is assumed, these lakes, on average, show little or no decrease in alkalinity and show a slight increase in pH. With an endpoint near 4.0 there is again little or no change, on average, in alkalinity, but an increase in pH of about 0.4 pH unit. Wisconsin lakes, on average, appear to show a significant increase in alkalinity (38 peq/L) and pH (+0.5 unit) if an MO endpoint of 4.2 is assumed. The increases are even larger if an endpoint close to 4.0 is assigned. Currently, the question regarding the correct endpoint pH in historical alkalinity titrations is unresolved. In the judgment of the authors of this chapter the endpoint lies in the range from 4.19 to 4.04, but evidence is not sufficient to specify a most likely value. Tests done with indicator dyes which have been preserved for the past 50 years may be valuable in resolving this question. We have not attempted to indicate the mechanisms for change in acidic status of these lakes. Any discussion of mechanisms must consider each lake separately and include all the hydrological and biogeochemical factor For most lakes, data are not available currently to permit this kind of analysis.

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290 One possible avenue of future research is to conduct a detailed study of the outliers in the three data sets. If the maximum change in lake alkalinity from deposition of atmospheric sulfur is about 100 peq/L, then study of the larger changes subjected to the most careful scrutiny for quality may offer a rational explanation for other acidification mechanisms. REFERENCES Almer, B., W. Dickson, C. Eckstrom, and E. Hornstrom. 1978. Sulfur pollution and the aquatic ecosystems. Pp. 273-311 of Sulfur in the Environment, Part Il. Ecological Impacts. J. O. Nriagu, ed. New York: J. Wiley and Sons. Arnold, D. E., R. W. Light, and V. J. Dymond. 1980. Probable effects of acid precipitation on Pennsylvania waters. U.S. Environmental Protection Agency. EPA-600/3-80-012. 21 pp. Baker, J., and T. Harvey. 1985. Critique of acid lakes and fish population status in the Adirondack region of New York State. Draft final report for NAPAP Project E3-25, U.S. Environmental Protection Agency. Beamish, R. J., and H. H. Harvey. 1972. Acidification of the La Cloche Mountain Lakes, Ontario and resulting fish mortalities. J. Fish. Res. Board Can. 29:1131-1143. Beamish, R. J., W. L. Lockhart, J. C. van Loon, and H. H. Harvey. 1975. Long-term acidification of a lake and resulting effects of fishes. Ambio 4:98-102. Blakar, I. A., and I. Digernes. 1984. Evaluation of acidification based on former calorimetric determination of pH: The effect of indicators on pH in poorly buffered water. Verh. Int. Ver. Limnol. 22:679-685. Booty, W. G., and J. R. Kramer. 1984. Sensitivity and analysis of a watershed acidification model, Phil. Trans. R. Soc. London Ser. B 305:441-449. Burns, D. A., J. N. Galloway, and G. R. Hendry. 1981. Acidification of surface waters in two areas of the eastern United States. Water Air Soil Pollut. 16: 277-285. Chen, C. W., J. D. Dean, S. A. Gherini, and R. A. Goldstein. 1982. Acid rain model: hydrologic module. J. Env. Eng. Div. Am. Soc. Civ. Eng. 108:455-472.

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291 Christophersen, N., and R. F. Wright. 1981. Sulfate budget and a model for sulfate concentrations in stream water at Birkenes, a small forested catchment in southernmost Norway. Wat. Resour. Res. 17:377-389. Church, M. F. 1984. Predictive modeling of acidic deposition on surface waters. Pp. 4-113--4-128 of The Acidic Deposition Phenomenon and Its Effects. Critical Assessment Review Papers. Vol. II. Effects Sciences. A. P. Altshuller and R. A. Linthurst, eds. EPA-600/8-83-016 BE. U.S. Environmental Protection Agency. Clark, W. M. 1928. The Determination of Hydrogen Ions. Baltimore, Maryland: The Williams and Wilkins Company. Cobb, E. D., and J. E. Biesecker. 1971. The national hydrologic Bench-Mark network. Circular 460-D, U.S. Geological Survey. 38 pp. Colquhoun, J., W. Kreter, and M. Pfeiffer. 1984. Acidity states of lakes and streams in New York State. New York Department of Environmental Conservation, Raybrook, New York. 49 pp + viii app. Crisman, T. L., R. L. Schulze, P. L. Brezonik, and S. A. Bloom. 1980. Acid precipitation: the biotic response in Florida lakes. Pp. 296-297 of Ecological Impact of Acid Precipitation. D. Drably, and A. Tollan, eds. Proceedings of an international conference, Sandefjord, Norway. Sur Nedb~rs Virkning Pa Skog Og Fisk (SNSF) Project, Oslo. Critical Assessment Review Papers (CARP). 1984. The acidic deposition phenomenon and its effects. A. P. Altschuller and R. A. Linthurst, eds. Report EPA-600/8-83-016AF. U.S. Environmental Protection Agency. Cronan, C. S., and C. L. Schofield. 1979. Aluminum leaching response to acid precipitation: effects on high-elevation watersheds in the Northeast. Science 204(20):304-306. Davis, R. B., M. O. Smith, J. H. Bailey, and S. A. Norton. 1978. Acidification of Maine (U.S.A.) lakes by acidic precipitation. Verh. Int. Ver. Limnol. 10:532-537. Dickson, W. 1980. Properties of acidified waters. Pp. 75-83 of Proceedings of an International Conference. Ecological Impact of Acid Precipitation, D. Drabl OCR for page 231
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