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Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements (1985)

Chapter: 2. State of Knowledge of Metal-pH-Ecosystem Interactions

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Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
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Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 5
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 6
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 7
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 8
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 9
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 10
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 11
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 12
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 13
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 14
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 15
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 16
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 17
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 18
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 19
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 20
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 21
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 22
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 23
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 24
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 25
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 26
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 27
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 28
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 29
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 30
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 31
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 32
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 33
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 34
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 35
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 36
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 37
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 38
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 39
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 40
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 41
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 42
Suggested Citation:"2. State of Knowledge of Metal-pH-Ecosystem Interactions." National Research Council. 1985. Acid Deposition: Effects on Geochemical Cycling and Biological Availability of Trace Elements. Washington, DC: The National Academies Press. doi: 10.17226/808.
×
Page 43

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2 STATE OF KNOWLEDGE ME:TAL-pH-ECOSYSTEM INTERACTIONS DI SCUSSTON OF FRAME:WORE QUESTIONS The following discussion is organized into four parts (atmosphere, terrestr ial ecosystems , aquatic ecosystems, and sediments) in accordance with the framework depicted in Figure 1.1. The questions listed in each part refer to the questions shown in the figure and addressed in the Appendix (Tables A.1 through A.18~. Atmospher e Question 1: Is Deposition Controlled by Human Processes? Whether human activity or natural emission processes control the current concentrations of metals in precipitation can be assessed in four ways: 1. by comparing the metal emission rates of human sources and natural sources, 2. by comparing the ratios of metal concentrations in the atmosphere to the ratios of metal concentrations in the natural sources, 3. by determining historical trends of metal concentrations in atmospheric deposition, 4. by comparing concentrations of metals in wet deposition in urban, rural, and remote areas. Using these four criteria, Galloway et al. (1982) divided the metals of concern in this report into three groups: · metals whose rates of atmospheric deposition in eastern North America are controlled by anthropogenic processes. · metals whose rates of atmospheric deposition in eastern North America are still controlled by natural processes. · metals for which there is insufficient evidence to determine the effect of anthropogenic activities on the rate of atmospheric deposition. 4

5 A review of the literature suggests that deposition rates for AS, Cd, Cu. Pb, Mn, Hg, Ni, Se, Ag, V, and Zn are controlled by anthropo- genic activities (Galloway et al., 1982; Lantzy and MacKenzie, 1979~. There are insufficient data from which to draw conclusions on atmos- pheric deposition rates for Be, Co, Mo, Sn, Te, and T1. Of the 18 elements considered in this report, the sole one for which the atmospheric deposition rate is known to be controlled by natural sources is A1, due to its prevalence in geological materials (Galloway et al., 1982; Jeffries and Snyder, 1981~. The magnitude of the effect of anthropogenic activities is illustrated by a comparison of the concentrations of metals in wet deposition in urban, rural, and remote sites (Table 2.1~. Question 2: Do We Know the Speciation of the Metals in Atmopsheric Deposition? This question may be divided into three parts: (a) Do we know the physical speciation, i.e., dissolved versus particulate? (b) Do we know the chemical speciation? (c) Which is most important, wet or dry deposition? Physical Speciation. There have been few studies on the distribution of metals between the dissolved and particulate phases of wet deposition. For Ag, AS, Be, Co, Hg, Mo, Se, Sn, Te, T1, and V, we were unable to find any references. The few papers on the subject discussed A1, Cd, Cu. Me, Ni, Pb, and/or Zn. Lindberg et al. (1977) used 0.5-pm filters to separate dissolved from particulate material in summer rain from Walker Branch, Tennessee. They found that 90%, 75%, and 92% of the Ma, Pb, and Zn, respectively, were in the dissolved phase. Tanaka et al. {1981) used a 0.4-pm filter to differentiate between dissolved and particulate Ni, Zn, Cu. Pb, and A1 in winter rain from Tallahassee, Florida. Their results showed that A1 occurred exclusively in the particulate fraction, while Ni, Cu. and Zn were present principally as dissolved elements. Pb was more evenly distributed between the two fractions. Gatz et al. {1984) separated dissolved Zn , Cd, Cu. and Pb from particulate forms using a 0.4-pm filter on 49 weekly wet deposition samples collected in a Chicago suburb. They found that the medians of the percent dissolved distributions were 95%, 95%, 90%, and 83% for Zn, Cd, Cu. and Pb, respectively. - - Chemical Speciation. To be able to link metal deposition with ecological changes, it is necessary to know not only the rate of deposition but also the chemical Speciation of the metal. Its geochemical mobility and biological availability may both depend on the form of metal deposited. Dissolved metals in wet deposition can exist in a variety of ionic forms depending on the characteristics of the individual metals, pH of the precipitation, and kinetics of transformation from one chemical

6 TABLE 2.l Median Concentrations of Metals in Total Wet Deposition (pg/~) Metal Urban Rural Remote As 5.8 0.286 0.019 Cd 0.7 0.5 0.008 Co 1.8 0.75 -- Cu 41 5.4 0.060 Pb 44 12 0.69 Mn 23 5.7 0.194 Hg 0.745 0.09 0.079 Mo 0.20 —- - - Ni 12 2.4 -- Ag 3.2 0.54 0.007 V 42 9 0.163 Zn 34 36 0 .22 NOTE: If only one concentration was available, it was used as the median value. If the median fell between two numbers, the average of the two was used. If a range was reported, the midpoint of the range was used. Insufficient data exist for Be, Se, Sn, Te, and T1. No bulk data were used. SOURCE: Galloway et al. (1982~. species to another. There are two methods to estimate the chemical speciation of metals in precipitation: direct measurement and thermodynamic modeling. Both approaches have disadvantages. Existing analytical methods are often limited, either by their detection limits or because they involve perturbation of the sample that could alter the speciation of the metal. Thermodynamic modeling gives the speciatio~r assuming that all metals are at equilibrium. This ignores the fact that, at times, the kinetics of thermodynamically favorable reactions are too slow to result in equilibrium.

7 TABLE 2 .2 Calculated Speciation of Dissolved Trace Metals in a Typical Rainwater (pa 4.14, pC1 4.88) Total Percent Concentration Aquo Ion Other Dissolved Species Metal ~lo-8M, MZ~ (H2O) n (>1%) Ag 0.50 98 AgCl+ (2%) Cd 0.45 100 Co 1.27 100 Cu 8.50 100 Hg 0.45 <1 - HgC12 (60%), HgCl+ (2%), HgOHCl (26%), Hg(OH)2 (118) Mn 10.4 100 Ni 4.09 100 Pb 5 . 7 9 10 0 Zn 55.1 100 As 0.38 Mo 0.21 V 17.7 H2As0~2 (100 % ) MMoO4 (3496), MOO42 (66%~ HVO4 (100%) We have been unable to find any references on measurements of the chemical speciation of metals in wet deposition and have found only one reference on the use of thermodynamic models to predict the speciation. Sposito et al. (1980) used the computer program GEOCHEM to calculate the equilibrium speciation of acid precipitation from New England for a few metals (Pb, Mn, Al, Cd, Cu. Zn, and Ni). In order to check their analyses and to include additional metals (Ag, Co, Hg, As, Mo, and V), we used the MINEQL thermodynamic model (Westall et al., 1976) to determine the chemical speciation of the metals in precipitation typical of central eastern North America. The data used are from the annual volume weighted concentrations of major inorganic ions in precipitation at the Hubbard Brook Exper imental Forest for the period 1963 to 1974 (Likens et al., 1977), and the median concentrations of trace metals measured in rural locations (from Table 2.1~. The results of our analyses (Table 2.2) agree very well with those of Sposito et al. (1980~. Specifically, based on the MINEQL and GEOCHEM models, Al, Ag, Cd, Co, Cu. Mn, Ni, Pb, and Zn are present as the aqua ion (M +tH2O)n). Mercury exists primarily as HgC12 (60~) and HgOHC1 (26%~. Arsenic exists as H2ASO4; Mo exists as MoO4 (66%) and HMoO4~34~; and V exists as HOOD. Changes in chloride concentration would influence the speciation of Ag and Hg.

8 TABLE 2.3 Mean Values of Data Reported from All Seasons for Dry Fraction of Total Deposition Metal Mar ine Rural urban As -- -- 0.2 Cd 0.4 0.4 0.6 Cu 0.5 -- __ Pb 0.6 0.3 0.2 Mn 0.5 O.S 0.5 Ni 0.6 0.5 O.5 V 0.4 -- __ an 0.7 0.4 0.5 SOURCES: Duce, 1979 (marine); Feely and Larsen, 1979 (rural and urban); Lindberg and Harriss, 1981 (rural). Importance of Wet Versus Dry Deposition. Metals may be deposited from the atmosphere via wet or dry deposition. Dry deposition can occur in three ways: by gravitational settling (generally particles greater than 2 to 10 Am, depending on meteorological conditions), by aerosol diffusions and/or impaction (<10 ~m), and by gaseous diffusion. The rate of gravitational settling is typically estimated by what falls into an open container. To measure the rates of aerosol impaction and gaseous adsorption, estimates of atmospheric concentrations and deposition velocities are required. For the metals in this study, AS, Hg, Se, and perhaps Cd are the only ones with large vapor pressures. Therefore most metals will be in aerosol form and will be deposited by gravitational settling or aerosol impaction. Measurements of toxic metals in deposition suggest that the dry fraction is substantial. Few cases in the literature have indicated that the dry fraction is less than 0.1 of the total deposit. For the most part, the mean dry fraction lies between 0.3 and 0.6. The further the site is away from a source area, the less important dry deposition will be compared to wet deposition. Table 2.3 summarizes data for several toxic metals from deposition in marine, rural, and urban areas. No systematic trend between the different environments is evident. In summary, based on these initial studies, we are convinced that dry deposition of metals is important relative to wet deposition, but the degree of importance has yet to be determined primarily because of inadequate techniques to determine the rate of dry deposition. In addition, the deposition of fog, dew, frost, and cloud water has the potential to be a significant mechanism for transferring metals to forests. Lovett (1984) and others have shown that there is a signifi- cant amount of cloud water deposition to forests, but little work has

9 been done on the metal composition of this cloud water. This is clearly an area for research. Terrestrial Ecosystems Question 3: Does the Metal Accumulate in the System? Question 4: Does the Metal Move in Solution to the Lake? Question 5: IS There an Interaction of the Metal with Acidification over the pH Range 7 to 4? In the general area of trace metal mobility in the terrestrial ecosystem (Questions 3 and 4) and possible interactions with acidification (Question 5), three types of investigation can be identified: 1. laboratory experiments with soils, or with individual soil components (e.g., clays, iron and manganese oxides, humic acids), in which interactions between a particular metal and the solid phasets) have been studied under different pH conditions (usually in stirred suspensions); 2. laboratory or field experiments in which soil columns have been subjected to simulated or real precipitation events, and the chemical composition of the percolate has been followed over time; 3. regional surveys of soils, vegetation, and/or surface waters in areas of relative geologic homogeneity where there exists a gradient in precipitation chemistry (increasing acidity). Each of these approaches is discussed below. Soil and Soil Components in Aqueous Suspension. In the first type of study, if the percent metal adsorbed is plotted as a function of pH, a relationship similar to that shown in Figure 2.1 is usually obtained; for those metals existing as cationic species, an abrupt increase in the amount of adsorbed metal occurs over a narrow pH range. In such a plot, pH50, the pH at which 50% of the original cation concentration is adsorbed, is a useful parameter for comparing the adsorption of several different cations on a given solid, or the adsorption of a particular metal on several different solid phases (Rinniburgh and Jackson, 1981~. (See Table 2.4 for an example with two different solid phases, Fe(OH)3 gel and A1(OH)3 gel.) These laboratory studies are useful in illustrating the potential effects of a relatively small change in pH on trace metal mobility and demonstrating the selectivity of certain solid phases for different trace metal cations. However, in a given metal-solid system, the position and shape of the pa adsorption sedges (i.e., the value of peso) will depend on the following: · the composition of the solid phase; · the concentration of the solid phase;

10 100 c: UJ m a: o In a 50 Ct: _ l _ _— / / 1 pH FIGURE 2.1 Typical pH adsorption curve for divalent cations on hydrous metal oxides (pH50 values may range from about 3 to 8~. . the initial concentration of the metal, [M]; · the nature and concentration of any competing metals present; · the nature and concentration of any ligands present, [L] (Benjamin and Leckie, 1981a,b; Davis and Leckie, 1978~. Relatively few studies have been performed with natural soils (or even with two or more competing solid phases), and virtually none have been carried out at the high (solid/solution) ratios that prevail in a soil system. Accordingly, the quantitative transposition of the results of this type of laboratory study to f ield conditions is unwarranted. In a qualitative sense, however, it could be anticipated that the larger the fraction of a particular metal present on exchange sites in a soil, the more sensitive the metal will be to minor pH changes. Soil Columns. In the second type of study, involving the leaching of soil columns, it in not surpr ising that trace metals have been found to migrate at different rates and to exhibit different sensitivities to pa changes. Three representative studies are discussed here. Fuller et al. (1976) compared 11 soils representing 7 different soil types and reported that Al, Cu. Cd, Fe, Mn, Ni, and Zn were more highly solubilized by dilute acid (H2SO4, pH 3.0) than by deionized water; this increased mobility was particularly noticeable for Al' Fe, and En, whereas Co and Pb remained immobile even when subjected to the acid leach. The relative order of mobility was thus as follows:

11 TABLE 2.4 Adsorption of Trace Metals on Colloidal Metal Hydroxides Fe (OH) ~ gela Al(OH)~. gel metalb: pE50 c metal: pH50 Pb < Cu < Zn < Ni < Cd - Co 3.0 4.3 5.3 5.7 5.9 Cu < Pb ~ Zn < Ni < Cd - Co 4.8 5.2 5.6 6.3 6.6 aGel concentration: 9.3 x 10-2M A1 or Fe. brace metal concentration: 1.25 x 10-4M. CpH50: pH at which 50% of original metal concentration is adsorbed. SOURCE: Kinniburgh and Jackson (19811. A1, Fe, Zn > Cd, Cu. Mh, Ni ~ Co, Pb. In a similar exper iment, Tyler (1978) leached two organic spruce forest soils with artificial rainwater (pH 4.2, 3.2, and 2.8) and established the following ranking (decreasing percent leached): Mn _ Zn > Cd ~ Ni > Cu > V > Cr > Pb. The residence times for all the above elements except V and Cr decreased with increasing acidity of the simulated precipitation. In a subsequent two-and-one-half-year field study, Tyler (1981) quantified the amounts of metal (A1, Fe, Mn, Cd, Cr. Cu. Ni, Pb, V, and Zn) leached from the A-horizon of a podzolic spruce forest soil in southern Sweden and identified two distinct seasonal patterns. One metal group (Cr. Pb, Ni, V, A1, and Fe) exhibited maximum concentrations In late summer and autumn and minimum values in winter and early spring; considerable losses of these elements occurred under conditions favoring the leaching of organic matter (high soil temperature and moisture content). Tyler notes that the Pb reprecipitates in the B-horizon and thus is not lost to the ground water. The second group (Cd, Mh, and Zn) attained maximum concentrations during the winter months. This latter group, comprising those metals having an appreciable fraction present in exchangeable form, was susceptible to minor pH changes in the soil percolate and the concentration variations were positively correlated with [H+~. Tyler speculated that a lowering of soil pH would reduce the residence times of these elements in particular. Regional Surveys. With reference to the possible accumulation of airborne metals in the terrestrial system, Steinnes (1983) sampled 500 natural soils in a nationwide survey of metals in Norway and found a

12 distinct geographical distribution of As, Cd, and Pb, which closely resembled the pattern of measured atmospher ic deposition. Be proposed that the higher levels in southern Norway were due to long-range atmospheric transport. Data from peat profiles in 13 Norwegian ombrotrophic bogs (Hvatum et al., 1983) supported Steinnes's contention for the distribution of As, Cd, and Pb and showed similar patterns, also suggesting long-range transport and accumulation, for Co, Cr. Cu , Fe, Mn, and Ni. Local natural sources of airborne Se added to the levels of Se or iginating from long-range atmospher to transpor t of this element. The moss Hylocomium spl endens was also collected from the sites used for soil analysis, and its metal content showed very large geographical differences for As, Pb, and Sb, with much higher values in southern Norway. A similar but less pronounced pattern of elevation was shown for Ag, Cd, Se, V, and Zn (Rambaek and Steinnes, 19801. In general, the net amount of metal added to soil from the atmosphere as a result of long-range transport is not considered a major source of metal contamination for terrestrial systems. Although there is good evidence of increased inputs, any environmental effects are likely to be a result of changes in mobility, because absolute increases in metal levels are relatively small in comparison with background levels. In regional surveys designed to evaluate the mobility of metals in the terrestrial system, changes in soil, water, or sediment chemistry noted along a precipitation pH gradient have been attributed to the interaction of the acid precipitation with the terrestrial system. Several examples are presented below. In the course of a regional soil survey in Scotland, McLaren and Crawford (1973a) determined the relative distr ibution of Cu among the forms: Cu-exchangeable ~ Cu-adsorbed · Cu-adsorbed - _ (inorganic) ~ (organic) As the natural pH of the soil decreased, this equilibrium was observed to shift to the left; i.e., a greater proportion of Cu was found in the easily exchangeable form. In their general review of the ecological effects of acid precipitation, Almer et al. (1978) present data for waters from different lakes in Sweden showing an increase in the concentrations of Al, Mn, Cd, and Hg along a spatial gradient toward lower pH (see Figure 2.2 for the Al data). More recently, Dickson (1980) has added Pb to this group of metals, noting that these trends apply to differently acidified lakes in Sweden subjected to similar atmospheric loadings. This Similarity of atmospheric loadings," a key point for Cd, Hg, and Pb, could not be verified as the relevant data were neither presented nor cited. Henriksen and Wright (1978) reported similar results for Pb and Zn in a series of small lakes in southern Norway, but they noted that they could not distinguish between Acidification + mobilization" or Acidification + concomitant increased atmospheric loading. as pass ible mechanisms.

13 700 600 - o 500 - ~r CC 400 z o i) 300 i 200 J 100 o . . · ·:: ·. i. :-. · — . · . - - . _ . . . I I · · Hi · ; ~ · 1 4 5 6 7 8 pH FIGURE 2.2 Relationship between pH and total (unfiltered) Al concentrations in Swedish lakes (redrawn from Dickson, 1980~. For well-drained organic soils in North America, Hanson et al (1982) studied a number of metals (Al, Ca, Cd, Mg, Mn, Na, Pb, and Zn) and reported that they are differentially leached at an accelerated rate as precipitation pH decreases. The results of the study showed the following: Mn > Ca > Mg ~ Zn > Cd; Na > Al. In this system, nearly 100% of the incoming Pb in precipitation is retained in the organic soil layer (see also Smith and Siccama, 1981~. The authors suggest that Zn is also accumulated in soils where the pH values for precipitation and soil solution are generally Greater than 5.0-5.5~; at lower pH values Zn and other Chemically similar elements are leached at an accelerated rate. Recent studies of trace metal profiles in forest floors in remote regions showed similar vertical profiles for Pb, Cu. Zn, Ni, and percent organic matter in three different forest types, with highest concentrations of each of these metals 2 to 4 cm below the surface. Metals concentrated in the He horizon (Friedland et al., 1984a). Studies of the metal cycling by forest vegetation in the Solling

14 project in Germany Showed annual inputs of Cr. Mn, and Ni from the atmosphere to be low (<30%) compared with the amount stored in the annual increment of biomass. For Fe the percentage was higher (40 to 60%) , while for Cu. inputs were 100% from the atmosphere (Heinrichs and Mayer, 1980~. Time trends of metal accumulation in soils are rarely available except for high loadings, e.g., smelter studies. However, a recent study by Friedland et al. (1984b) on Camels Hump Mountain indicated an increase of Pb, Cu. and Zn in the forest floor over the period of 14 years since 1966; this increase was consistent with annual atmospheric deposition rates reported in the literature for these metals. Johnson et al. (1982} estimated a 5- to 10-fold increase in Pb in forest floors of the northeastern United States over the last century. Summary. Based on an analysis of the available experimental results (types 1, 2, and 3), the 18 trace metals of current concern are tentatively classified according to their potential mobility in a terrestrial environment subject to acid precipitation: high mobility: A1, Cd, Me, Zn moderate mobility: Cu. Ni low mobility: Co, Pb, V Insufficient data are available for Ag, As, Be, Hg, Ma, Se, Sn, Te, and T1. Question 6a: Is the Metal Bioavailable? Question fib: Does the Metal Bioaccumulate? There is no strong evidence for accumulation by terrestrial plants of metals derived from long-range atmospheric transport, although metals do accumulate in litter and organic layers of forests in areas remote from point sources (Johnson et al., 1982~. However, if soils acidify, there is some potential for plant uptake to increase with the mobilization and increased availability of the following metals in soils: A1, Mn, Fe, Zn, Cu. and Ni (Hutchinson and Collins, 1978~. These metals may occur naturally in soils, or result from atmospheric deposition. In naturally acid soils from California, A1 and Mn in soil solution ranged from O. . 1 to 108 mg/L and 2.6 to 200 mg/L r respectively (Straughan et al., 19811. It has long been known that one of the beneficial effects of liming is decreased solubility of Al and Mn (Vlamis, 1953~. Foy et al. (1978) found that the limitation to the growth of certain calcicolous plants on acid soils is directly related to A1 toxicity. In the absence of measurements of metal uptake and accumulation by plants from soils exposed to acid precipitation, the information cited in the Tables A.1 through A.18 has been taken from recent articles and reviews that deal with relatively heavy loadings; for the most part these have addressed either sewage sludge application or local point

15 sources of metals such as smelters and highways. From the sewage sludge literature one may obtain information on soil as a source of metals (e.g., Adriano and Page, 1981), and from smelter studies one may obtain information on both airborne material and contaminated soils as sources of metals (e.g., Whitby, 1974~. For the airborne sources, it would appear that lichens (and to some extent bryophytes) are the most useful monitors. A large body of information on this subject has been published for Canada by Rao and LeBlanc (1967) and Nieboer et al. (1977) and for Scandinavia by Pilegsard (1978~. Bioaccumulation in a lichen may differ mechanisti- cally from that in higher plants. For the most part, these lower plant biological monitors indicate a gradient of concentration from source to background. They also give historical patterns in some instances. One recent study in the Canadian arctic and in Ontario has shown for radionuclides that lichens are ideal long-distance monitors as well (Hutchison-Benson, 1982~. For metals that are common in the earth's crust or are at very low levels in deposition, however, enrichment of nonradioactive isotopes above background will be much more difficult to detect in lichens or indeed in any other biological mater ial. For h igher plants to accumulate metals from soils , the chemical form of the element as well as physiological properties of the plant species need to be prescribed. Even leaving aside the so-called accumulator species (e.g., A~tragalus spp. for Se), the literature indicates that most metals accumulate in roots and some in shoots of higher plant tissues (Foy et al. , 1978~ . In the absence of studies designed to relate metal accumulation in plants to long-range transport of metals in the atmosphere, one can Speculate that only under extreme conditions will metal accumulation be a problem for terrestrial plants. Assimilation studies have been performed on foliar uptake of Hg vapor in wheat by Browne and Fang (1978~. The model provided by these authors could be applied to the much lower levels of Hg in the atmosphere, which, until very recently, was not amenable to direct measurement because of the very low concentrations. Airborne particles may also be incorporated into plant tissue; e.g., particulate Pb has been shown to accumulate in plant tissue close to point sources (Roberts et al., 1974~. While there is potential for forage plants or edible vegetables to accumulate metals in acid soils, normal soil amendment practices such as liming and addition of organic matter will counteract the tendency for metals to be taken up from soils by vegetation (e.g., see gingham et al., 1979~. If the soils acidify, it is highly probable that plant growth will fail so accumulation in food or forage crops becomes an academic problem. For perennial plants, especially forest trees, there is little information on metal accumulation, although it could be occurring from direct deposition onto leaves, or from soils. Tree-ring analysis for metals may be informative (Johnson and Siccama, 1983~.

16 Question 7: Dces the Metal Have Biological Effects on the Terrestrial System? The elements that are known to be required as micronutrients may be leached from acid soils. Eventually, deficiencies might be expected to occur (for example, in the case of an, Cu. and Mo), but no such effect" have been demonstrated to date. Excessive metal leaching usually occurs in soils heavily impacted by smelter emissions, but often such soils have altered physical and chemical structures. On these soils, plant growth fails for a number of reasons; direct micronutrient deficiency is not likely to be one of them. Direct uptake of metals through leaves is another possible source of effects on plants. For example, foliar uptake and translocation of Cd and Pb was demonstrated by Hemphill and Rule (1975), but no effects were observed until 50 ppm of Pb and 10 ppm of Cd were applied. Since such levels far exceed those typically found in precipitation, this route may only be significant in areas affected by elevated metal concentrations from point sources. Whether metal toxicity results from the mobilization of metals in acidified soils appears to be a much more urgent and compelling question. To date, the most notable discussions of this aspect on a global or large regional scale relate to A1. A1 toxicity to plants in acid soils has long been known (Foy, 19741. Hutchinson (1983) point" out that the older literature of agronomy has been largely ignored in the context of acid precipitation. It has been proposed that the poor- condition of many trees in acid-stressed regions is a result of A1 toxicity to roots (Tomlinson, 1983; Ulrich et al., 1980), and investigations are under way to determine whether this hypothesis is tenable (G. Abrahamsen, Norwegian Forest Research, personal communication, 1983; Hutchinson, 19837. The extensive dieback and defoliation of trees in Europe and North America is sufficiently serious that such an hypothesis requires complete evaluation. It must be emphasized that there are alternative hypotheses concerning the cause of the dieback, and that these do not all implicate soil acidification (Dunnett, 19831. For other metals, the references to effects given in the tables of the Appendix are for stimulation or intoxication in experimental or highly polluted situations. For agricultural soils, liming would counteract any increased availability of metals such as Cd that may be added via precipitation (e.g., see gingham et al., 1979) so potential toxicity would be alleviated. TO date there has been no suggestion that metals other than A1 are toxic to plants in soils receiving acid precipitation. However, microorganisms may be affected, as suggested by increased organic matter paralleling increases in metal" in the forest soils in eastern North America {Friedland et al., 1984b; Siccama et al., 19807. These studies provide circumstantial evidence for toxicity to decomposers in soils at concentrations of metals that result from long-range transport. Mycorrhizal fungi should also be considered as potentially metal sensitive, or, alternatively, protective of their hosts; the literature is inconclusive on this point. Killham and Firestone (1983) found enhanced metal uptake by endomycorrh~zal plants

17 for the grass species Ehrharta calycina compared with nonmycorrhizal plants, where metals were applied with simulated acid rain. When metal deposition was combined with acid rain, shoot yields of the grass were more adversely affected when mycorrhizae were present. The applied level of metals, however, was quite high, ranging from 0.l to 4.2 mg/L. In contrast, Bradley et al. (1981) reported protection against Conner toxicity in Calluna vuloaris infected with the ericoid mYcorrhiza1 fungus Pezizella ericae when compared with noninfected _ . , ~ ~ ~ controls. These findings need to be pursued, and more studies done at levels of metals that are realistic in the context of metals added from deposition away from point sources. In the general context of metal toxicity, the role of calcium {Ca) deserves mention. Ca (as well as magnesium (Mg) and potassium (R)) has a protective effect on plants exposed to potentially toxic metals (Hutchinson and Collins, 1978; Jowett, 19641. From work in solution culture, and from work with simple aquatic plants, it would appear Ca protection is at least in part the result of an effect at the cell membrane level (Mierle, 1982~. Therefore excessive losses of Ca or of Mg from soils, which may be a reality within decades (Abrahamsen, 1980), may indirectly increase the risk of metal toxicity to plants. Ca and Mg may also be leached directly from foliage (E. Cowling, North Carolina State University, personal communication, 1984), and this effect would also increase the potential for metal uptake and toxicity. Aquatic Ecosystems Question 8: Do We Know the Speciation of the Metal in Solution (Measured)? Examples of possible physicochemical forms of trace metals in natural waters are shown in Table 2.5; note that the operational distinction between dissolved, colloidal, and particulate forms is convenient {particularly to the analytical chemist) but arbitrary. AS these various forms will often exhibit different chemical reactivities, the measurement of the total concentration of a trace metal provides little indication of the metal's potential interactions with the abiotic or biotic components present in its environment; the corollary, of course, is that knowledge of the speciation of a trace metal is necessary for understanding its geochemical behavior (mobility/ transport) and biological availability. Possible approaches to the determination of the speciation of trace metals in natural waters can be grouped into two major categories (Florence, 1982; Florence and Batley, 19801: 1. thermodynamic calculations, involving the use of measured total concentrations and published stability-constant values, to compute the equilibrium contribution of the various species (e.g., Jenne, 19793; 2. experimental methods, involving the determination of the physical properties (e.g., size) and/or chemical reactivity (e.g., electrochemical lability, ion-exchange behavior) of the trace metal of interest.

18 0 X to : A In 0 £ so En o A o V · - s 0 S V) o . m En o 0 P' o O · - ~ V En us go _ V ~ H ~ O ~ ~ ~ :: O —£~1 — P ~ ~ ~ ~ a: 0 0 H ^ ~ _1 _ V H ~ ' 41) 1 16 0 H H ^ + Jo ~ ~ ~ V _ _ ~ o ~ ~ ~ ~ ~ ~ ~ -- ~C V O O t: ~ V ~ - ~ ~ ' ' U 0= ~ V ^ O O ~ ^~ ' C~ ~ _1 ~ ~ ~ V O ~ H ^ :t O O. ~ _ ~ + + H H O + 3 ~ ~ N N N C) ~ C) ~ C) ~ ~ ~ ~ ~; 0 ~n O ~ 0 O ~ U] O _I ~ 41) ~ V~ O a, x 0 0 V o. - X X ~' - O O 0 0 ~ ~ V V _~ ~ It_~ C O X ~ ~ .C ~ ~ ~ ~ ~ O £ ~ ~ ~ ~ ~ 0 ~ V X O ~ O ~X ~ ' V V ~ ~ ~ ~ ~ ~ ~ V 0 ~ ~ ~ ~ O ~ ~ ~ = o X O ~ ~ O C: ~ 0 o —~ ~ V ~ ~ ~ ~ ~ ~ Q ~c · · _1 0 ~ O O O ~ O ~ 1 o o Q. V V ~ ~ 0 ~ O V ~ ~ · - ~ ~ ~ ~ ~ ~ ~ ~ Q O O ~ V ~ ~ ~ 0 ~ 0 ~ 0 V ·- ~ ' V ~ ~ ~ Q Q Q Q C: ~ ·rl X ~ O · · ~ c~ O ~ ~ ~ ~ O 81 0 ~ ~ ~ ~ O O O O . . ~ 0 ~ x ~ 0 ~n 0 0 ~ v m 0 ~ ~ ~ " - O ~ ~ ~ ~ ~ V ~ ~ X ~ ~ ~ ~ - O ~ V ~ ~ ~ ~ ~ O O ~ O O 0 ~ - 3 ~ ~ ~ · - V ~ ~ V O 00 0 0 0 ~ ~ U] u) ~ ~ t O ~ ~ O V Pd ~ V ~C 0 0 c: ~ O O 0 ~ ~ m V ~ V ~ ~ ~ ~ ~C O ~ ~ _ ~ c ·rl 0 ~ 0 ~ 0 ~ 0 c 0 . V 0 X 0 ~ O O ~ O ~ - ~ ~ O 0 ~ c ~ ~ ~ V ~ ~ 0 ~ ~ _ l V U] S V S U

19 In the former case, where the assumption of equilibrium may itself be misleading, the major current limitation is the absence of reliable thermodynamic data for many of the species suspected to be present in natural waters (notably those involving natural organic matter, present as dissolved ligands, as colloidal organic material, or as an organic coating on inorganic particles); Nordstrom et al. (1979) have recently examined the sensitivity of these equilibrium models to the choice of stability constants serving as input data. The second or experimental approach is often limited by the lack of experimental techniques sufficiently sensitive and selective to detect individual trace metal forms at the concentrations normally found in natural waters. Current techniques allow only the classification of trace metal forms into various operationally defined categories, according to their physical properties or chemical reactivity (for recent examples of such speciation schemes, see Figura and McDuffie (1980~; Florence (1977~; Hart and Davies (1981a,b); and Laxen and Harrison (1981a,b)~. It should be recognized that such techniques may cause shifts in the prevailing equilibrium and thus alter the trace metal speciation in the sample under analysis. In other words, the analytical results tend to be method-dependent (Florence, 1982) and difficult to compare. Factors influencing the speciation of a particular trace metal in the water column will include the following: 1. the nature of the ligands (L1, L2, L3, ..., Ln) and solid phases (S1, S2, S3, ..., Sn) present (i.e., the stability of the various ML and MS combinations); 2. the concentrations of ligands and suspended solids; 3. the concentrations of the hardness cations (Ca and Mg) and of certain other metals (e.g., Fe and Mn); 4. the redox potential; 5. the pH; 6. the degree to which equilibrium is attained (i.e., reaction kinetics). Several excellent reviews of trace metal speciation have been published recently (Florence, 1982; Florence and Batley, 1980~. Examination of these exhaustive compilations suggests that the 18 trace metals considered in the present report can be grouped into three categories: · well studied: · moderately studied: · little studied/unstudied: Cd, Cu. Me, Pb, Zn; Al, As, Co, Hg, Ni, Se, Sn; Ag, Be, Mo, Te, T1, V. However, even in the case of those elements referred to as "well studied," there have been very few studies of different waters with the same technique, and even fewer investigations of the same waters with different techniques. Consequently, whereas we may answer with some confidence that we know how to determine the speciation of Cd, Cu. Mh, Pb, and Zn, it unfortunately does not follow that we ~know" their speciation {as asked in Question 8~. Indeed, it has recently been suggested (Laxen and

20 . to v - a: Eve 0 erl o En o o · - Q o the A: 0 tRe Eve S X H A 0 4 0 em o em ID · ~ ~ . to En _ £ Eve X ;~ 0 0 the eQ Ed Cal ED At Ed em e. C) En n, tOe fix, ~ 0` ~ co co _t _1 _ Me ~ We Cal F F ~ ~ ~ ~ ~e 0 0 ~ 0 ~ U'e ~ 0 ~ 0 e. e" ~ ~ 0 0 0 0 0 4 0 4 dJ _I ~ e~ ~ e - 1 ~J e~ X ~1 X ~ ~ ~ ~ 1 ~ ~ :n ~ :n :n ~ :~: ~ :n ~ n: a 1 1 ~ ~ ~ o I I t~ C~e U.t ~) O ~ O O ~ ~ ~ 1 1 1 +1 _t ce_| ~ ~ ~ t1 ) CD U~ cn c~ 0 ~ a, ~r 0\ O ~ ~ 1 1 +1 0 1 t— ~ ~ ~ ~ 1 1— C~ +~ +~e + +n5 +~ ~ C) ~ ~ Z Z Z ~ ~ ~ ~e ~ m m m m m m 4 0 ~J 0 t.) e~ e~ e~ e~ 0 e~ 3 0 —~ ~ ~ I ~— e4 _I ~ t~ ~ ~ AS ~ e~ H _t eY 4 0 eY 4 e. e4 ~ 4 0 · · ~ (V ~ ~ 0 ~ O J~ —~ ~; ~ P~ ~ 0 te~ (U ~ U) ~~ Q~e 0 · ~ 4 eO ~ 4 eO dV eO L' e~ ·3 ~ ~l C ) 45 ~ ~ ~ e~~ P. ~ £ ~ ~ ~

21 an A ,0 ~ o o .- CD ' ~5 ~ ~ 0 ~ ~ ~ 0 o u' ~ a, ~ ~ ~ ~ ~ ~ ~ 0 ~ 0 0 = - = - ~ ~ ~ ~ ~ :3 ~ 1 ma ma x C) :: ~ ~ ~ ~ ~ ~ ~ 0 a: O O ~ 0 s 1 1 1 0 o o C: ~ a, en ~ ~ ~ 1 1 1 +1 O ~ ~ ~ Go O Us ~ ~ 1 1 1 +1 ~ ~ O 0 Up HI t— 0:) ~ ~ a) O . ~ ~ Us 1 ' 1 +1 o t— 1 _I to o —I to 1 0 Cal +~ + +~ ~ +~ +a Z :~ Z C) C) ~ ' ~ ~ ~ ~ m m m c' ~ cat · ~ · ~ 0 up to · ~ ~ 3 ' - ~ ~ ~ O ~ O 0 ~ ~ ' ~ ~ ~ ~ ~ ~ " 'A ~ ~ ~ ~ ~ ~ ~ ~ ~ ~ S O ~ ~ O 0 ,,y ~ ~ 0 ~ al 0 ~ ~ ~ ~ ~ ~ ~ O ~ ~ ~ ~— ~ ~ ~ ~ ~ ~ ~ ~ ~ O ~ ~ E~ ~ E~ g . u, U) 0 0 Sv - - 8 o _ :~: Z o _ + m ~ _ Z o · - ' V~ ~ JJ + U] Q C) - aQ · 0 ~ ~ ~ V ~ 0 O ~ P' · - ~ 0 Q' O ~Q O ~ ~ ~ · - ~ Q. S 1 "Q X ~ ~ 11 {Q O O ~ H V % a; Q ~ +1

22 Harrison, 1981a) that such generalizations may never be possible, the speciation of a given trace metal being so closely linked to other water quality parameters that it will inevitably vary greatly from one sample to another. An example of this var lability is provided by Table 2.6, in which the proportion of ion-exchangeable, "filterable. metal (i.e., the <0.4-pm fraction reacting with Chelex-100 resin) is presented for a number of natural waters. The order of relative exchangeability of Cd, Cu. Pb, and Zn varies considerably from one sample to another. Question 9: Does the Metal Increase in Concentration in the Aquatic System? Question 10: IS There an Interaction of the Metal with Acidification of the Water Column over the pa Range 7 to 4? aqua Ion: Regional surveys of water quality in areas affected by acid precipitation together with results from the experimental acidification of lakes (Schindler and Turner, 1982; Schindler et al., 1980a,b) and streams (Hall et al., 1980) indicate that the concentrations of the following trace metals normally increase in response to acidification: Al, Mn, Zn, Cd, Hg, and Pb (see Table 2.77. As mentioned earlier, the evidence for the latter three elements is somewhat ambiguous as it is difficult to distinguish between increases in concentration that result from acidification of the watershed and those that are caused by an accompanying increase in trace metal at~nospher ic loadings (Henr iksen and Wr ight, 1978 ~ . With regard to the possible interaction with acidif ication over the pH range 7 to 4, we refer here not to the changes in trace metal concentration mentioned above, but rather to possible changes in speciation occasioned by the decrease in pH, and to the accompanying changes in metal reactivity. Even at constant total metal concentration, changes in metal speciation can be anticipated as a result of a decrease in pH, e.g., · shift in the hydrolysis equilibrium to the left, favoring the MZ+(H2O)n _ M(H2O)n-l(OH}(Z-l)+ + H+ · shift in complexat~ion e~quilibr ia (competition between MZ+ and H+ for the same ligands): MZ+ + L -1, MLZ+ HI + L ' HL+ - shift in specific adsorption equilibria (competition between HZ~ and H+ for the same surface sites): MZ~ + - S H+ + S M-S H-S+

23 TABLE 2.7 Trace Metals Exhibiting Increases in Concentration in Response to Acidification Metal Type of Evidence Reference Al Regional survey lakes Response headwater streams Regional survey soil solution Experimental acidification lakes Experimental acidification stream Regional survey lakes Experimental acidification lakes Regional survey lakes Experimental acidification lakes Regional survey lakes Pb Regional survey lakes Hg Regional survey lakes Almer et al., 1978 Dickson, 1980 Johnson et al., 1981 Cronan and Schofield, 1979 Schindler et al., 1980a,b Hall et al., 1980 Almer et al., 1978 Dickson, 1980 Schindler et al., 1980a,b Henriksen and Wright, 1978 Schindler et al., 1980a,b Almer et al., 1978 Dickson, 1980 Dickson, 1980 Henriksen and Wright, 1978 Almer et al., 1978 In these equations, MZ~ = generalized metal cation of charge +z; L = ligand (uncharged for simplicity) ; S = a solid surface (uncharged for simplicity) with sites for cation adsorption (e.g., clay, metal hydroxide, and cell membrane). The degree to which a particular metal is affected bv a decrease in OH will denend on the acid-base properties _ _ , ~ ~ ~ _ ,= _ ~ of L and S. and on the stability of the various ML and MS species. Chemical equilibrium models can be used to identify those metals most likely to be affected by pH changes. For the present purposes, the composition of an "average surface water. was derived from data available for 10 lakes located in the Laurentide Park area on the Canadian Precambrian Shield north of Quebec City (Campbell et al., 1983~. Trace metal concentrations were assigned either average observed values (Al and Mn) or arbitrary but representative values (Table 2e8 ~ column 2~; provided that ligands are present in excess (i.e., [L] >> [M]), the calculated speciation of dissolved M will be independent of And. The MINEQL-1 chemical equilibrium model (Westall et al., 1976) was used to calculate the theoretical speciation of those trace metals for which the necessary thermodynamic data were available (only Te and T1 had to be eliminated). Simulations were performed for aerobic conditions (redox potential, pE, of 12) at fixed pH values of

24 U] £ ~ _ Al .e ~ Ed a' U) ,e~ . ~ Ed em o o Z - ~5 · - S V \*e ,e_| U em hey o 0 ~ S O em C) em em em V o Sat O V Ed eme V ~ ~e_t V CD ewe · ~d E~ `:: EH e,e~ ~e e o tl~e - o - + :r: o ~ m ~1 0 e~ ~ - 0 + ~ tJ U) V dP _~ + ~) :z + S ~e ~ _t ~) ~1 ~ O ~ e~ p~ ~ I o a on— ~ m cr co ~ 0 0 0 {S. ~ ~ 0 0 0 · - v o 5: - + N ~: o H e~ ~5 V Q~ C Q. ~: Q. e'~ ¢,4 o e" £ ~ le~ ,e - t (V I ~ V O O O X 1— <_) o ' ^O 'O I: + ~ ~ + ~0 ~ ~ O ~e_t _~ _~ /? ~ ~ ~ ~ C~ C) C) ~ ~ ~ ·= Q a~ O O O ~t ~e O O O V O Q ~ O ~e_| ,e_ V Q ,e_| V O 0\ · ~ O en ~ ~: m ~ t - ) ~ O e~ O ,e_t V ~e_ V O O O O ~ O ~ O O · · · · ~ O ~ O ~ O - H H - o C) e :t

25 O O ~ In +-o~ :t O o o o o o o o o v o o o o o o o o v Q o o ~ o o ~ o .,' V In V V O C~ - - ' 1 - ' 1~0 ~ ~ 1~0 O o, 1 - 1~r 0 0 ~ ~ #¢ 0 0 0 cn 1 - ~ ~ ~ ~ ~ ~ g o U~ o o o o o o o o ~ o o ~ ~ ~ · e ~ · · · · · ~ O O U~ ~ O O O ~ ~ ~ ¢: ~: £ Z P. u, c~ O~ ~ U) ~ o - - OQ . 3 0 ~ tn O o 0 0 ~n - S~ s ~ . - o 0 0 · u, ~n ~ ~ ~ · - 3 0 0 0 0 U] o . - 0 v o t) o 0 tQ 0 s n ~ ~n S~ ~ · - ~ n ~ Ll E~ Q

26 7, 6, 5, and 4 for the simplest inorganic system (no organic ligands present; no adsorbing solid sur faces) . Based on the results of these simulations (Table 2.8), we can define four classes of behavior in the pH r ange 7 to 4: 1. trace metals existing predominantly (>99%) as aqua ions throughout the studied pa range--Ag, Cd, Co, Hi, and Zn; 2. trace metals existing as oxyanions, for which the pH change merely affects the degree of protonation of the anion--As, Ho, Se, and V; 3. trace metals existing predominantly (>99%) as a solid phase throughout the studied pH range--Se; 4 . trace metals exhibiting signif leant changes in speciation in the pH r ange 7 to 4--A1, Be, Cu. Hg, Mn, and Pb . In this latter class we find both metals that show signif leant changes in solubility as the pa decreases (Al and Ma) and metals for which the equilibrium distribution among dissolved species shifts {A1, Be, and Hg; to a lesser degree Cu and Pb). A second series of simulations was performed for a system containing model organic ligands (argin~ne, lysine, ornithine, and Saline; citric, maleic, phthalic, and salicylic acid) at a total ligand concentration of 1.8 x 10-5 M (Sposito, 19811. For A1, Be, and Cu the sensitivity of metal speciation to pH changes is enhanced in the presence of these ligands. A field study to evaluate pH-induced changes in metal speczation was carried out in limnocorrals in the Experimental Lakes Area (Jackson et al., 1980; Schindler et al., 1980a). Lowering the pH within the enclosures from 6.7 to 5 .7 to 5 .1 affected both the physical speciation ~ r at lo of d issolved metal: par t iculate metal ~ and the r eactiv ity ~ loss rate from the water column) of some but not all of the added r adioisotopes . For specif ic metals, the results are as follows: · proportion of metal dissolved; increased--Be, Co, Fe, Mn, Zn decreased--Hg, V unchanged--As, Cs, Cr. Se, Th · loss rate from water column; increased--Be, Cs, Se, V deer eased--Me, Zn unchanged--As, Co, Cr. Fe, Hg, Th Question lla: IS the Metal Bioavailable in the Aquatic System? Question lib: Does the Metal Bioaccumulate? Question llc: I s the Metal Biomagnif led? Question lid: What Is the Inherent Toxicity of the Metal? Although knowledge of the chemical speciation of a metal gives some insight into its potential bioavailabil~ty, the best means to measure

27 availability is still the organism itself. For the metals under consideration, those that are taken up by biota are therefore assumed to be bioavailable. Essential trace elements from the list include Co, Cu. Mn, Mo, Ni, Se, V, and Zn. Until deficiencies of these have been demonstrated, one cannot assume that they are not bioavailable. For the remaining elements, uptake by aquatic biota of Al, Hg, Pb, Cd, and Ni has also been demonstrated, so these too are assumed to be bioavailable. Bioaccumulation is defined as uptake and storage in biota resulting in concentration factors (i.e., concentration in biota:concentration in water of >1~. More precisely, for multicellular organisms, the terms bioconcentration (concentration factor of >1) and bioaccumulation (increase of concentration factor with age} should be distinguished. Given the available data, it is feasible to use only one term, bioaccumulation, since the two can rarely be distinguished. Most metals are bioaccumulated by aquatic plants, and to a lesser extent by organisms in higher trophic levels. For acid-stressed waters, the following metals have been shown to bioaccumulate in algae: Al, Cu. Hg , Mn, Ni , Pb, and Zn (Stokes , 1983a). Accumulation in fish of Cd, Hg, and Pb has been demonstrated (K. Suns, Ontario Ministry of the Environment , personal communication ~ 1982 ; Suns et al., 1980; Wiener, 1983 ~ . Mn accumulates in the bones of f ish at low pa (Harvey and Frazer, 1982) . Of greatest concern is Hg, which may exceed levels in f ish considered safe for human consumption (0 .5 ug/g) (Rose, 1979) and which in field studies is negatively correlated with pH (Hakanson, 1980 ; Scheider et al., 1979 ; Suns et al., 1980 ; C . Wren, University of Guelph, Department of Zoology, personal communication, 1982~. Preliminary studies indicate a correlation between Hg in algae and Hg in fish {Stokes et al., 1983b), and between Hg in zooplankton and Hg in fish ha. Hultberg, Swedish Water and Air Pollution Research Institute, Goteborg, personal communication, 19841. From preliminary data on yearling perch, Cd and Pb also show pH-related accumulation in fish tissue (K. Suns, Ontario Ministry of the Environment, personal communication, 1982; Wiener, 1983~. In order to predict whether other metals will bioaccumulate in acid lakes, an understanding of the mechanism is required. For Hg, methylation with an increase of the bioevailable monomethyl species found at low pH has been proposed (Fagerstrom and Jernelov, 1972) , but laboratory studies have not confirmed that low pH increases methylation of Hg (Rudd, 19821. Other physical and chemical parameters need to be evaluated for their role in Hg bioaccumulation. A recent review of the Hg budgets of Swedish lakes in forested areas indicates that Hg in water is influenced by deposition from the atmosphere onto the lake surface and by influx from the entire catchment area (Lindqvist et al.' 19841. An increase from either of these sources would result in the observed increase of He concentration in pike. The authors suggest the main source is Hg from humus in the catchment area that accumulated during high levels of emission from burning of fossil fuels during the 1950s and 1960s. Cd and Pb can be methylated in vitro, but the methylated forms are too unstable to be candidates for bioaccumula- tion. There is a pressing need to investigate these phenomena. Concern for metals in fish muscle and human health is quite serious,

28 since fish from waters with almost undetectable Hg concentrations may contain Hg at levels considered unsafe for human consumption (R. Suns, Ontario Ministry of the Environment, personal communication, 1982~. Food chain accumulation (biomagnification) may occur in aquatic systems for Hg, but the existing evidence for other metals is not convincing. Even for Hg, there is conflicting evidence concerning its tendency to undergo biomagnification. In view of the scanty evidence, even for Hg, this is an area where more research is needed. Food chain transfer of Al, but not necessarily its biomagnifica- tion, appears to occur in passerine birds via the benthic biota (Eriksson et al., 1980; Nyholm, 1981~. The effect of increased levels of Al may be eggshell thinning and decreased clutch size (Nyholm, 1981~. Studies in Ontario on passerine birds show a relationship between nesting parameters, increased body burden of metals (including Al), and acidification in aquatic systems (V. Glooschenko, Ontario Ministry of Natural Resources, personal communication, 19821. Recent work on herbivores feeding on aquatic vegetation suggests that Cd levels in moose livers increases with acidification (H. Hultberg, through N. D. Yan, Ontario Ministry of the Environment, personal communication, 1982~. Wren et al. (1983) also demonstrated metal transfer from the aquatic system into terrestrial mammals feeding on fish or aquatic macrophytes. With reference to the classification of the 18 metals according to their Inherent toxicity,. we have used the empirical equations derived by Raiser {1980), which only relate metal toxicity to various ion-specific physicochemical parameters and ignore other possible modifying factors. These equations take the following form: AN pT = aO + al log _p 2 0 where aO ~ a1, a2 = pT = negative logarithm of the metal concentration causing a particular toxic response; AN = atomic number; TOP = difference between the fonts ionization potential with oxidation number n, and the ionization potential with the next lower oxidation number, n - 1; AEo = absolute value of the difference in electrochemical potential between the ion and the first stable reduced state; constants obtained by regression analysis, values of which depend on the ion group, on the biota considered, and on the particular toxic effect determined. Raiser suggests that the equations can be interpreted as showing the toxicity of an ion as a function of its atomic ionization potentials (AIP} and its observed electronic behavior (6Eo) in the solvent/ ligand system in question. Inclusion of the atomic number allows the successful correlation of elements having similar AIP and AEo

29 TABLE 2.9 Calculated Inherent Toxicities. of Some Metals PTc Inherent Toxicity Metalsa <5~0 low 5.0-7.0 moderate >7 . O high Be(II), Mh(II), A1(III) As(V), SntII), Zn(II), Cu(II) Ni{II), Co(II), Pb(IT) Se(VI), Ag(I), V<V), Hg{II), Cd(II), T1(III) aNumbers in parentheses indicate valence state of metals. values but significantly different atomic sizes. Using the calculated values of pT (i.e., pTc), we arbitrarily assigned the metal ions of concern to classes of low, moderate, or high Inherent toxicity. (Table 2.9~. Question 12: Does the Metal Have Biological Effects in the Aquatic System? The toxicity to fish of A1 is now reasonably well established (NRCC, 1981a); to date A1 is the only example of a metal that increases in concentration and whose toxicity is implicated in a biological response to acidification. Baker and Schofield (1982) also showed an interaction between low pH and A1 in laboratory bioassays with fish. They found that at a pH range of 4.2 to 4.8, A1 was beneficial to egg survival. Thus it is still necessary to be cautious when making statements on the toxicity of A1 in acid waters; the mechanisms are not yet completely understood. It is possible that other pH-related changes in community structure in acid-stressed water, e.g., in phytoplankton or zooplankton, result from toxic effects of A1 or other metals, but to date no causal relationship has been demonstrated. Furthermore, the levels of zinc in acid waters are approaching those that have resulted in toxicity to plankton in field experiments (Marshall et al., 1981~. The complex effects that have been shown in laboratory experiments when the toxicity of mixtures of metals are determined illustrates the potential for synergistic or antagonistic effects in surface waters susceptible to increased concentrations of more than one trace metal. Summarizing Questions 11 and 12, there is sufficient circumstantial evidence regarding the accumulation of metals in aquatic blots and the toxicities of these metals (at least that of A1) to justify a greater research emphasis on the relationship between metals and acidification of water.

30 Sediments Question 13: Do We Know the Partitioning of the Metal in the Sediments? Possible forms of trace metals in sediments include the dissolved and colloidal species present in the interstitial water and also those particulate forms that have already been indicated as part of the overall speciation scheme presented in Table 2.5. 1. adsorbed; 2. carbonate-bound; 3. occluded in Fe and/or MA oxides; 4. organic bound/sulfide bound; 5. matrix bound (detrital). Intuitively, it might be expected that chemical reactivity, geochemical mobility, and biological availability will decrease in the order 1 > 2, 3, 4 > 5; the relative rank ing of forms 2, 3, and 4 is, however, more difficult to predict. In principle, the speciation of sediment-bound trace metals could be determined both by thermodynamic calculations (provided equilibrium conditions prevail) and by exper imental techniques . The modeling of sediment systems is far less advanced than.is that of water samples, primarily because the thermodynamic data needed for handling concen- trated sediment-water systems are not yet available. Thus the only realistic means of studying trace metal partitioning in sediments at the present time is to fractionate the sediment physically and/or chemically. Conceptually, the solid material can be partitioned into specific fractions; sequential extractions with appropriate reagents can then be devised to leach successive fractions ~selectively" from the sediment sample. Several such experimental procedures, varying in experimental complexity, have been proposed (Engler et al., 1977; Forstner and Patchineelam, 1980; Gibbs, 1977; Tessier et al., 1979), and or it~cally reviewed (Jenne and Luoma , 1977; Luoma and Davis, 1983~. They have been applied to a wide variety of stream and lake sediments. For stner and Wittmann (1981) have reviewed the increasingly prolific literature in this area and have proposed a Standard extraction methods (see also Salomons and Forstner, 19807. Examination of this literature suggests that the 18 trace metals of current concern can be grouped into three categories: · well studied: Al, Cd, Co, Cu. Hg, Mn, Ni, Pb, an; · moderately studied: Ag, AS, Be, Mo, V; · little studied/unstudied: Se, Sn, Te, T1. The partitioning of trace metals present in a natural sediment will be affected by the physicochemical conditions prevailing within the sediment, and hence by the biological activity occurring at the sediment-water interface (Rrantzberg and Stokes, 1981~. For a particular trace metal' the factors influencing its partitioning will include the following:

31 · the nature and concentration of the ligands present in the interstitial water (L1, L2, L3, ..., Ln); the nature and concentration of the other cations present in . the interstitial water; · the redox potential; · the pa; · the nature and concentration of the solid phases present. The experimentally observed distribution of a trace metal obtained with any extraction scheme will not necessarily reflect the relative scavenging action of discrete sediment phases; hence the partitioning should be considered as operationally defined by the method of extraction (Rendell et al., 1980; Tessier et al., 1979~. Nevertheless, comparison of the results obtained for a variety of sediments with different extraction procedures reveals a certain coherence and suggests that trace metal partitioning is less variable in sediments than in the water column. For some of the ~well-studied" metals identified above, certain distribution patterns tend to recur in the forms that are not matrix bound (i.e., fractions 1 through 4 of a variety of sediments). Metal Cd Co Cu Hg Ni Pb Zn Dominant fractions 2 (specific adsorption; carbonate bound) 3 (Fe, ME oxides) 4 (organic bound) 4 (organic bound) 3 (Fe, Mn oxides) 3 (Fe, Mr1 oxides ~ 2 (specific adsorption; carbonate bound) 3 (Fe, Mn oxides) Metals found in large proportions in the readily exchangeable fraction 1 would be expected to be sensitive to pH changes in the interstitial water. Question 14: IS the Accumulated Metal Readily Recycled Between the Sediments and the Lake? Question 15: Is There an Interaction of the Metal with Acidification Over the pa Range of 7 to 4? With regard to trace metal release from aquatic sediments (Question 14) and possible interactions with acidification (Question 15), three complementary lines of investigation have been followed. 1. laboratory experiments with sediments, or with sediment components, in which interactions between a particular metal and the solid phased) have been studied under different pH conditions (either in stirred suspensions or quiescent microcosms);

32 2. regional sediment surveys in area. of relative geologic homogeneity, where there exists a gradient in surface water chemistry (increasing acidity); 3 . f ield exper foments involving the experimental acidification of lakes or streams and their associated pediments. Each of these approaches is discussed below. Laboratory Simulations. Typical sediment components acting as trace metal sinks (e.g., clays, iron and manganese oxides, and humic acids) are, of course, also found in soils, and laboratory studies of interactions between trace metals and these components are pertinent both to soil systems and to aquatic sediments. As noted earlier, if the percent metal adsorbed is plotted against pa, a ~igmoid relation- ship is normally obtained (Figure 2.11. However, these determinations are normally carried out in the laboratory during short-term experiments, with the assumption that equilibrium conditions are rapidly established. The existence of such equilibrium conditions in the natural environment is debatable. The exact position of the pE adsorption ~edge. will vary as a function of the following: · the nature and total concentration of the metal being considered, CHIT; · the nature and concentration of the solid phase, [S]; · the nature and concentration of any competing metals present; · the nature and concentration of any ligands present, [L]. Thus, in a present at adsorbed. sediments although 2.101. From laboratory results it might be anticipated that under quiescent conditions a long-term decrease in pa in the water column, and hence in the sediment interstitial water, would affect the mobility of certain trace metals more than others. In this connection, Brannon et al. (1980) have recently published the results of a study of the long-term (8 months) release of trace metal" (AS, Cd, Cu. Fe, Eg, Mn, Pb, and an) from 32 sediment samples held in laboratory microcosms under quiescent conditions. The pa of the overlying waters was not controlled but rather allowed to vary naturally (range pE 6.5 to 8.7~. Statistical analyses of the results suggested that only for three elements was the exchange with the overlying water related to changes in [H+~; the exchange of Fe and Pb (both cationic) was inversely related to the hydrogen ion concentration, whereas that of As (anionic) varied directly with changes in [H+~. These results are contrary to explanations based on chemical principles and other exper immoral data (see Table A.2 and A.7 for AS and Pb, respectively); the authors concluded that the mechanisms by which pa affected long-term net mass release from the sediments were not readily apparent. defined medium ([S], [Ll, pa fixed), different trace metals similar initial concentrations will often be differentially Such behavior is observed with suspensions of natural studied in the laboratory under controlled conditions, relatively few such studies have been reported (see Table

33 ~2 0 U] ~4 As o A Sit EN o o o A s o to o U] 0 to . ED 0 c 0 0 fir: to $ U] - o _' a: 0 A: . ~ V] · - 5 to O o ~ a' ~ ~ e ~ ~ ~ ~ o ~ ~ ~ ~ ~ co ~ e ~ ~ ~ ~ IS, C ' ~ ~ e.^ ~ ~ en ~ ~ ~ - ~ · 0 0 . - ~ ' 3 ~5 ~ O ~ - - - ~ O ~ ~ ~ ~ O ~ O S c) t) 0 c; ~ U] ~ U) ~r ~ ~ c~ · · · ~ ~r ~ ~ ~r 1 1 1 1 ~r ~ ~ c~ ~ ~ O · · · e · · · O . · · 1 O O O O O ~ ~ ~ O ~) In 11~) o o x: :1E: x: :lE~ U~ kD X 1 1 1 1 1 O O O O O I ~ t~ 1 X X X X X o 1 cn · ~ ~ X O Q ~ ~ Q O C) C) ~ ~ C) C~ ~ C) . ~ · U) · ' U) ~ ~ ~ ~ _I _t ~ c) ' ·— ' ~ s ~ · S ~ e~ ~ . - ~ O P~ C .' - ~1 ~q ~ 0 ~ · Z Z 0 C ~Q ~ ~ O ~ O ~ C ~ U) ~1 ~ $¢ O 0 o IO U] o erl 0 o o · - - 11~ e~ · - o S o dP o U, - S ut o eO _~.

34 TABLE 2.11 Metals Influence of pa on the Speciation of Sedbment-8Ound Trace Metal ApH Effect Reference Cd 8 ~ 5 Relatively mobile; organic forms Rhalid decreased as pH lowered; et al., 1981 accompanying increase in dissolved and exchangeable forms; no significant changes in Cd - associated with reducible fraction. Pb 8 ~ 5 Immobile; little or no dissolved Pb Gambrell detected at any pa; exchangeable et al., 1980 forms increased under moderately acid conditions (pa 5.0~; no Significant changes in Pb associated with reducible fraction. Zn 8 ~ 5 Relatively mobile; dissolved and Ga~hrell exchangeable Zn increased et al., 1980 markedly as pH decreased; concomitant decrease in Zn associated with reducible fraction. Hg 8 ~ 5 Immobile, little or no dissolved fig Gambrell detected at any pa; exchangeable et al., 1980 Hg increased slightly under moderately acid, reduced (pa 5.0, -150 mV) and weakly alkaline, oxidized (pE 8.0, +500 mV) conditions. A different laboratory approach has been to determine the effect of a change in pH on trace metal partitioning in the sediment, an deter- mined experimentally by sequential selective extractions. Several such studies on river sediments have been carried out by Gambrell et al. (1980) and Khalid et al. (1981) at the Laboratory for Wetland Soils and Sediments, Louisiana State University {Table 2.113. In their closed systems, a decrease in pa from 8 to 5 led to an increase in the levels of dissolved and exchangeable Cd and Zn, i.e., to an increase in the geochemical mobility of these metals. In a natural (i.e., open) system, Such an increase in mobility would result in a net loss of the metal from the sediment. An interesting field corroboration of this suggestion is provided by the recent work of Reuther et al. {1981), who determined the partitioning of several trace metals in sediment cores from lakes Hov~atn (pH 4.4) and Langtjern (pa 4.95) in Norway. They

35 TABLE 2.12 Trace Metals Released from Sediments in Response to Acidification Metal Type of Evidence Reference Al Experimental acidification Schindler et al., 1980a,b of lakes Experimental acidification Hall et al., 1980 of streams Paleolimnological study of Schofield, 1980 lake sediments Mn, Fe Experimental acidification Schindler et al., 1980a,b of lakes Zn Experimental acidification Schindler et al., 1980a,b of lakes Regional survey of lake Hanson et al., 1982 sediments Paleolimnological study Hanson et al., 1982 of lake sediments Norton et al., 1981 Reuther et al., 1981 Cd, Co, Ni Paleolimnological study Reuther et al., 1981 of lake sediments found that Cd and Zn had been remobilized from the recent sediment strata in the lake of lower pH, mainly from the organic and the easily reducible fractions, respectively. Co and Ni also showed evidence of remobilization, but to a lesser degree, whereas Pb and Cu were little affected. Regional Sediment Surveys/Experimental Acidification Experiments. Both regional sediment surveys in areas affected by acid precipitation and the results from the experimental acidification of lakes and streams suggest that the acidification of an overlying water column will increase the geochemical mobility in the sediments of the following trace metals: Al, Cd, Co, Fe, Mn, Ni, and Zn. AS indicated in Table 2.12, the evidence for Cd, Co, and Ni is somewhat less persuasive than for the remaining elements because it is based on a single observation (Reuther et al., 1981~. This latter report suggests that remobilization of Cd, Co, Ni, and an from the sediment does not occur until the pH is below 4.95. The concept of a critical or threshold pH value has also been suggested for Zn remobilization (Norton et al., 1981~. Paleo- 1 Prolog ical data from sediment cores collected from lakes with relatively undisturbed watersheds revealed concentration versus depth profiles for Pb that increased monotonically toward the sediment-water

36 interface. In contrast, the concentration of Zn first increased and then decreased in the most recent strata near the top of the core. This general relationship holds for some 20 lakes in the northeastern United States and in Norway where the pH is less than 5.5 (Norton et al., 1981~. The authors suggest that at pa values of >5.5 lake sediments either act as a sink for incoming particulate En or chemically scavenge dissolved Zn from the water column. In acidified lakes (pH of <5.5), not only do the sediments not scavenge Zn from the water column, but they also apparently yield dissolved Zn to the overlying waters. It has been suggested id. C. Kennedy, USGS, Menlo Park, California, personal communication, 1984) that the Remobilization threshold. may mark the boundary between desorption of exchangeable ions as the major process contributing to metal exchange, and dissolution of carrier phases containing occluding trace metals (e.g., Fe or Mn oxides). The paleolimnological data available for lake sediments in Norway, the northeastern United States, Ontario (Dillon and Evans, 1982), and Quebec (Ouellet and Jones, 1983) all suggest that Pb is strongly bound in lake sediments. The critical pH value for release of Pb lies some- where below the lowest observed water column pa values. Experimental confirmation of this immobility was recently provided by Davis et al. (1982), who subjected the upper strata from two sediment cores to acidification in the laboratory; significant lead release {>5% of total concentration) only occurred at pH < 3.0 in the sediment from Woods Lake and pH < 2.0 in that from Sagamore Lake. Question 16a: Is the Metal Bioavailable? Question lab: Does the Metal Bioaccumulate? Question 17: Does the Metal Have Biological Effects on the Benthic System? It is commonly accepted that most of the metals transported into an aquatic system are scavenged or precipitated and accumulated in sediments. This fact is the basis for using downcore profiles of metals for estimating historical patterns of deposition. Yet for the biota, which can only assimilate soluble material, only a fraction of the sediment metal is potentially available, even to sediment dwelling (benthic) organisms. For example, Tessier et al. (1984) showed that the Cu. Pb, and Zn levels in various tissues of the Ellintio comulanata a suspension-feeding freshwater mollusk, were best related not to total metal concentrations in the adjacent sediment, but rather to one or more of the relatively easily extracted fractions. Body burdens of Cu. Pb, and Zn were also influenced by the protection or competitive effect of other sediment constituents, notably amorphous iron oxyhydroxides and, to a lesser extent, organic matter. Benthic organisms can assimilate metals from dissolved forms in pore water or from particles ingested into the gut. Studies with tubificids in contaminated sediments indicated uptake mainly through

37 the integument, but also showed deposition of ingested metals in the foregut (Back, 1983; Prosi, 1983~. The metal, released from Sediment by chemical or biochemical processes, is bioavailable for direct uptake, and metals in ingested sediment, if weakly bound, may be assimilated in the gut. Not surprisingly, it is technically difficult to distinguish between metals obtained by the biota directly from the water column or pore water, and those obtained via sedimented materials. If the sediment should be a major source for metals in the water column, this distinction for the biota may be academic. Studies on the geochemical characteristics of metals in sediments and on the relationship between pa and metal release from sediments suggest that there is a reasonable basis for expecting some metals in sediments to show increased bioavailability as pH declines. The literature is, however, quite deficient in definitive experiments on biological uptake made in conjunction with geochemical studies; those that exist tend to relate to heavily polluted sediments . and Anew ial lv ~ ~ = ~ , ~——_ _ = _ ~ ~ ~ ~ to marine systems (e.g., Luoma and Bryan, 1979). In responding to the questions on bioavailability and bioaccumulation, we found that the most frequently encountered studies used correlation analysis on field collections to determine relationships between metals in sediment (total or identified fractions) and metals in biota; and only one of these studies dealt specifically with acidified systems {Schindler et al., 1980a}. Significant correlations would provide circumstantial evidence for the sediment as a source of metal. Copper in benthic blots exhibits a relationship with Cu in sediments, but this has only been shown for relatively contaminated water bodies such as sites affected by mining activities (e.g., Tessier et al., 1982), or systems spiked with Cu (e.g., DikS and Allen, 1983~. Stokes et al. (1983a) found significant accumulation of Cu in benthic algae related to sediment contamination, but there was no relationship for softwater lakes undergoing acidification unless they were close to _ point sources ot CU. zn also chows sediment-related accumulation in biota (Forstner and Wittmann, 1981; Tessier et al., 1982), and a release of Zn from sediments of relatively low contamination occurred in the presence of an active infauna (Rrantzberg and Stokes, 1982~. Mercury mobilization from sediments has not been reported from experimental studies on acid-stressed systems. However, Beijer and Jernelov (1979) suggested that methyl mercury was formed in sediments. At pH below 5.0, the monomethyl form of mercury is expected to predominate (Fagerstrom and Jernelov, 1972~. Transplant experiments with freshwater clams (Karbe en al., 1975) showed Hg uptake to be related to a number of parameters including organic content and oxygen in sediment, but not to total Hg in sediments. Hakanson (1980) produced an empirical model that related Hg in fish to Hg in sediment, pH of the overlying water, and the nutrient status of the sediment (referred to as the bioproduction index). Stokes et al. (1983b) found no significant correlation between Hg in benthic algae and total Hg in sediments in a series of acidic and neutral lakes lacking any point source of Hg. Hg levels in algae were, however, as high as or higher than those in sediments and up to 10,000 times higher than in water. Algal Hg showed a significant correlation with Hg in yearling perch from the same lakes.

38 These studies for Hg illustrate the complexity of the task of resolving the interacting factors determining biological uptake of sedimented metals. It is clear that bioaccumulation of a metal by a sediment-dwelling organism does not necessarily relate in any direct or simple way to the geochemically mobile forms of metals in sediments. The dilemma faced by the researcher is quite obvious: biota can only assimilate metals from solution, which they can either obtain by ingesting solid material and absorbing weakly bound metals from the solid, or by absorbing metal directly from the water or from the water that is ingested with the sediment. Since there is a flux of metals between sediment and water, it may be in fact Impossible to determine whether sediment-dwelling organisms are accumulating metal from water or sediment. Furthermore, it becomes impossible to determine whether observed effects of acidification on benthic biota result from metals in sediments, or directly from H+. The question on effects therefore has to be given the response ~unknown. for all the metals of concern. Benthic faunal communities show certain broad responses to acidification, but data for blots are available for the most part only from studies of acidification of the water column. and not from studies of the conditions in the ~:~ imprint. · . ~ ~ * ~ e itself {Krantzberg, 1982~. Faunal biomass and species composition of benthos may show quite wide variations among lakes of apparently comparable pa and alkalinity in the water column. The obvious question to address concerns conditions in the sediment. Sediment pH and metal availability may have been measured in geochemical studies, but these have not been linked to faunal or floral response=. Baker and Schofield (1982) were able to look at the effects of pH and A1 on fish and found, for example, that the presence of A1 at low pa was beneficial to certain life stages. They stated that Simple generalizations concerning the effect of increasing A1 concentrations with acidification are not possible. ~ For benthic biota, where our knowledge of effects is much more rudimentary, it has to be concluded that while the potential exists for metal toxicity in sediments of acid-stressed systems, there are as yet no available studies that separate the respective influence of low pa, low nutrients, and increased metal availability. The implications for increased availability of metals to the benthic biota are not trivial. Increased body burden may lead to transfer of metals to the terrestrial food chain for birds that feed on benthos or emerging insects (Erikason et al. , 1980), while toxicity to benthos may result in depletion of food supplies for these animals at subsequent trophic levels. Bioaccumulation of metals in benthic algae may be an important factor in the cycling of metals in acid-stressed systems. Finally, the long-term fate of sedimented metals in post- acidification (neutralization or recovery) scenarios is likely to be not only chemically but also biologically determined. Local effects of bioturbation (Rrantzberg and Stokes, 1981) may provide for release over extended periods of time, reinforcing the concept of the sediment as a reservoir rather than as a sink for metals. . . . . . . —

39 SCARY This section contains a summary of the information presented in the preceding sections. Table 2.13 summarizes the materials presented in the Appendix (Tables A.1 through A.18~. The reader can either look down the table to determine quickly how much is known about individual metals, or across the table to determine how much Is known about specific questions indicated in Figure 1.1. To synthesize the results of our analysis and to extract an answer to the general question--Which metals should be of primary concern in the context of acid deposition?--we condensed the original framework of seventeen questions (Figure 1.1) to five factors: 1. Are metal concentrations in the atmosphere controlled by anthropogenic activities? 2. Do metal concentrations increase in response to acidification, in either the soil or the aquatic environment? 3. Does a pH change in the critical range 7 to 4 cause significant changes in metal speciation? 4. What is the inherent toxicity of the metal, and has toxicity to biota been observed in relation to acidification? 5. Does the metal have a tendency to bioconcentrate? The results of this summary analysis are presented in Table 2.14.

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