Conclusions The significance of the organic compounds detected in this indoor environment warrants comment. Actual concentrations of the various substances are relatively small, in part because total levels of suspended particulates are low (between 10 and 14 pg/m3) (6).E x h of the identified organics contributes on the order of O.Olo’0 of this figure. The important features of this study are the spectrum of organics identified and the information this spectrum reveals. Phthalate esters have been observed in numbers and relative abundances that appear greater than those of outdoor samples (7,9);the detected alcohols and phosphates have not been reported in ambient aerosols ( 1 0 ) . A number of these compounds are used as additives rather than entities, and yet they have become significant constituents of an indoor aerosol. The salient point is that stable substances introduced into a confined environment are likely to accumulate in that environment. Materials used in such situations should be chosen accordingly. Acknou ledgmerlt
I thank T. E. Graedel, J. D. Sinclair, P. C. Milner, and C. A. Russell for helpful discussions during this work. Literature Cited (1) Halpern M , J Air Pollut Control Assoc , 28,689 (1978).
( 2 ) Cautreels, W.. Van Cauwenberghe, K.. Water, Air Soii Poiiut., 6,103 (1976). ( 3 ) LVeschler, C. J., Enciron. Sci. Technol., 12,923 (1978). (4) Bauer, E. J . , Reagor, B. T., Russell, C. A,, A S H R A E J . , 13, 53 (1973). (5) Weschler, C. J., “Abstracts of Papers”, 176th National Meeting of the American Chemical Society, Miami Beach, Fla., Sept 1978, American Chemical Society, Washington, D.C., 1978, ENVT225. (6) Walker, M. V.,Weschler, C. J., Bell Laboratories, unpublished data, 1978. (7) Cautreels, R.,Van Cauwenberghe, K., Atmos. Eni’iron., 10,447 (1976). (8)Hoffman. D.. Wvnder. E. L.. “Air Pollution”. Vol. 2. Academic Press, New York, 1968. p p 192-222. (9) Karasek, F. W., Dennv. D. LV., Chan, K. h-., Clement. R. E.. Ana/ Chem , 5 0 , 8 2 (1978) (10) Graedel, T. E., “Chemical Compounds in the Atmosphere”. Academic Press, S e n York, 1979. (11) Giam, C. S., Chan, H. S.. Neff, G. S., Atlas, E. L., Science, 199, 419 (1978). (12) Lundgren, D. A,, Paulus, H. ,J., J . Air Pollut. Control Assoc., 25, 1227 (1975). (13) Cadle. R. D.. “Particle Size”, Reinhold, New York, 1965, pp 33-42, (14) Friedlander. S. K.. “Smoke. Dust and Haze”. LVilev. New York. 1977, p p 1-3. (15) Van i’aeck, L.. Van Cauwenbernhe. K.. Atmos Enuiron , 12. 2229 (1978).
Receiced for rerieu June 18, 1979 Accepted December 21, 1979.
Chemical Availability of Mercury, Lead, and Zinc in Mobile Bay Sediment Suspensions as Affected by pH and Oxidation-Reduction Conditions Robert P. Gambrell”, Rashid A. Khalid, and William H. Patrick, Jr. Laboratory for Wetland Soils and Sediments, Center for Wetland Resources, Louisiana State University, Baton Rouge, La. 70803
Mobile Bay sediment material was incubated under conditions of controlled pH (5.0,6.5, and 8.0) and redox potential (-150, +50, +250, and +500 mV) to determine the effects of these physicochemical parameters on the chemical form of added labeled mercury and lead as well as indigenous lead and zinc. After equilibrating for 2 weeks under controlled pHredox potential conditions, suspension aliquots were subjected to a chemical fractionation procedure to determine levels of metals in total :soluble, noncationic soluble, exchangeable, reducible, and chelate extractable forms. The soluble and/or exchangeable fractions, representing readily available forms, were found to represent only a small proportion of each metal. A large proportion of each metal was found in potentially available forms. An altered physicochemical environment influenced levels of the metals in most of the general chemical forms studied. The responses of metal levels in the selected chemical forms to controlled p H and redox potential conditions tended to differ for the three metals studied, indicating that the immobilizing processes do not affect the metals to the same degree. It was concluded that the physicochemical condition of contaminated dredged materials a t a disposal site may influence transformations of these metals among readily available and potentially available chemical forms.
The possible release of sediment-bound toxic metals to bioavailable forms as a consequence of dredging and dredged material disposal has received considerable attention in recent years (2-3). Potentially toxic metals exist in several chemical 0013-936X/80/0914-0431$01.00/0
@
1980 American Chemical Society
forms in surface waters, soils, and sediments. These forms range greatly in their chemical and biological availability. Metals dissolved in soil solution, surface, and interstitial waters and metals adsorbed to the solid phase by cation exchange processes are readily available to aquatic and benthic organisms as well as to plants. Trace metals bound as impurities within the crystalline lattice structure of primary and secondary minerals by isomorphic substitution are essentially unavailable to organisms and become available only over geologic time as a consequence of mineral weathering. Between these availability extremes are a number of chemical forms which are potentially available and are subject to several mobilization/immobilization processes affecting their availability. In many soils and sediments, and especially those which are contaminated, most of the trace and toxic metals present are in potentially available forms. Chemical and biological processes affecting the mobility of metals may be initiated by altering the physicochemical environment (salinity, pH, oxidation-reduction conditions) of a soil or sediment-water system ( 4 ) .Gambrell et al. ( 5 )have discussed how the pH and oxidation-reduction conditions of dredged sediments may change depending on the chemical nature of the dredged material and disposal methods. Important regulatory processes influencing the chemistry and availability of trace and toxic metals include: (a) precipitation as insoluble sulfides under highly reduced conditions (6-8), (b) formation of discrete metal oxides and hydroxides of low solubility ( 8 ) ,(c) adsorption to colloidal hydrous oxides of iron and manganese, primarily in aerobic, neutral, or alkaline environments (9-15), and (d) complex formation with Volume 14, Number 4, April 1980
431
Calomel half cell
Millivolt
1. pH electrode 2. Platinum electrodes
3.Salt bridges to
In1
-
calomel half cells
'
/ 4.Gas outlet
5. Serum cap 6. Air inlet 7. N2 inlet 8. Tfiermometer
9. Stirring bar
I
Figure 1. pH and redox potential control apparatus for Mobile Bay sediment suspensions soluble and insoluble organic matter (13, 16-19) under all conditions of pH and oxidation levels. The objective of this report is to demonstrate the influence of pH and redox potential on the chemical forms of readily available and potentially available mercury, lead, and zinc in Mobile Bay sediment material using a selective chemical fractionation procedure.
Methods
Collection and Characterization of Sediment Material. Sediment and surface water material were collected from Mobile Bay, Mobile, Ala. The bulk sediment material was mixed and stored under a nitrogen atmosphere a t 4 "C until needed for incubation studies. Clay mineralogy, organic and carbonate carbon content, particle size distribution, and cation exchange capacity were determined to characterize the sediment material. Glass and platinum electrodes were inserted directly into the sediment material after obtaining a mixed aliquot to determine indigenous pH and redox potential. Incubation of Sediment Material Containing Labeled and Indigenous Metals under Conditions of Controlled pH and Redox Potential. A predetermined amount of reduced, wet sediment material, equivalent to 200 g of oven dry solids, was placed in each of four incubation flasks. Sufficient stored surface water from the sampling site was added to each incubation flask to produce a sediment-water mixture with a so1ids:water ratio of a typical dredged slurry (20). A 2-L, three-necked, flat-bottomed flask was used to contain sediment suspensions incubated under conditions of controlled pH and redox potential (Figure 1).A motor driven magnetic stirrer was used to maintain a suspension of the sediment-surface water mixture. Each flask was fitted with two bright platinum electrodes, a combination electrode for measuring pH, calomel reference electrodes connected to the suspension by saturated potassium chloride-agar salt bridges, one serum cap, a thermometer, separate inlet tubes for air and oxygen-free nitrogen, and an outlet tube. The outlet tube was submerged in an acid, potassium permanganate solution to prevent gaseous oxygen diffusion into the flask and to trap labeled, elemental mercury, should it be volatilized from the system. The suspensions were maintained a t 30 "C, fl "C. Suspension pH was adjusted and maintained a t the selected value by injecting 1 N hydrochloric acid or 1 N sodium hydroxide through the serum cap located in the center rubber stopper cap. Bright platinum electrodes immersed in the sediment suspension were connected to a potentiometer for redox potential measurements. A meter relay (General Electric, Type 196) was connected to the recorder output of the 432
Environmental Science & Technology
potentiometer. This relay activated an aquarium pump when the redox potential of the suspension dropped below the desired level. The air flow from the aquarium pump was regulated to 0.5-2.0 mL/min to permit slow oxidation of the suspension during the aeration cycle. When the suspension was again oxidized to the proper level, the meter relay automatically switched off the aerator. Because sediments exhibit a natural tendency to become more reduced in the absence of oxygen, redox potential could be maintained over a considerable range of oxidation-reduction levels by regulating the addition of air. An additional gas inlet to the flask supplied oxygen-free nitrogen gas. Nitrogen gas was effective in purging excess oxygen at the end of the aeration cycle or from possible small leaks and in preventing a buildup of gaseous decomposition products such as carbon dioxide and hydrogen sulfide. Each experimental run included four redox potentials ranging from strongly reduced to well-oxidized (-150, +50, +250, and +500 mV) a t one pH value. The pH levels studied were 5.0, 6.5, and 8.0. After obtaining desired pH and redox potential conditions, the suspensions were incubated an additional 4-5 days and carrier-free radioisotopes of 203Hgand 210Pb(0.3 pCi/g of oven dry solids) were added. The sediment suspensions were incubated for 2 weeks with added radiotracers before being sampled for chemical fractionation. Chemical Fractionation. The sediment suspensions were sequentially fractionated according to the following scheme: (a) total water soluble, (b) soluble, but bound as uncharged or negatively charged complexes, (c) exchangeable, (d) reducible, (e) DTPA extractable, and (f) total elements. The following paragraphs briefly describe the fractionation procedures. Suspension aliquots were centrifuged and the supernatant filtered through a 0.45-pm membrane filter into a receiving flask to obtain the total water-soluble fraction. The receiving flask contained sufficient distilled nitric acid to lower the pH of the filtrate to 2 for sample preservation. The dissolved, noncationic form was collected as described for the total water-soluble fraction, except that the filtered supernatant was passed through a cation chelating resin bed (Chelex-100) connected to the outlet of the filtration apparatus. Exchangeable metals were extracted by shaking the residual sediment solids from the total water-soluble fraction for 2 h with a 1 N sodium acetate solution ( 1 O : l extractant to solids ratio), which had been adjusted to the pH of the sediment suspension. The extracted solution was centrifuged, filtered, and stored for analysis as described for the total water-soluble fraction. The residual solids from the material extracted for exchangeable metals were extracted with 0.15 M oxalic acid and 0.25 M ammonium oxalate (20:l extractant to solids ratio) to remove reducible metals as previously described for the exchangeable fraction. This procedure is thought to solubilize poorly crystalline metal oxides and hydroxides and particularly poorly crystalline hydrous oxides of iron with which trace and toxic metals may be coprecipitated (21,22). An extracting solution of 0.05 M DTPA (diethylenetriaminepentaacetic acid) and 0.2 M sodium acetate was prepared and the pH adjusted to 7.0. The residual solids from the reducible fraction were extracted with this reagent (1O:l extractant to solids ratio) as described for the exchangeable fraction, except that refrigeration at 4 "C was the only sample preservation required. This extractant is more concentrated than the synthetic organic chelate extractant normally used for soil testing procedures (23). Synthetic organic chelates have been used to solubilize some soil organic matter and/or metals bound with naturally occurring humic materials (24-26). As the final step in a sequential chemical fraction-
Table l. Physical and the Mobile Bay Sediment Material
PropertiesOf
7.2 -160 mV
PH redox potential predominant clay mineralsa silt coarse clay (2-0.2 prn) fine clay (50 p m ) silt (2-50 prn) clay ( < 2 prn) cation exchange capacityb
illite, potassium feldspars, quartz kaolinite, montmorillonite, illite montmorillonite, kaolinite
1.8YO 0.1% 23 yo 41 ?o' 36 Yo 33 mequivl100 g 600 p g / g
total sulfidesb iron total potentially reactive total Hg total Pbb total Zn
37 400 p g l g 19 000 p g l g
a Listed in decreasing order of abundance based on relative peak areas of X-ray diffractograms (30). Oven dry solids basis.
ation procedure, this extractant is thought to primarily solubilize a portion of the metals bound to large molecular weight humic materials ( 2 7 ) . Aliquots of the incubated sediment suspensions were sampled, extracted, centrifuged, and filtered under a nitrogen atmosphere. All chemical extractants were deoxygenated prior to shaking with the sediment suspensions. Also, the sodium acetate and ammonium oxalate reagents were treated t o remove trace metal contamination. The 203Hg and 210Pbactivities in the various chemical forms were measured with a n Ortec y counting system equipped with a NaI crystal. Details of materials and methods are given elsewhere ( 2 8 ) . Results and Discussion
Selected physical and chemical properties of the Mobile Bay sediment material are given in Table I. The material had a near neutral p H and was moderately to strongly reducing. Predominant clay minerals were determined from X-ray diffractograms (29) and are listed in decreasing order of abundance (30). Potentially reactive iron represents that part of the total iron content recovered by the selective chemical fractionation procedure employed. Much of this iron was in the reducible form ( 2 8 ) .
Mercury. The influence of controlled pH and redox potential on the recovery of added 203Hg in total water-soluble and the soluble, complexed forms is indicated in Table 11. Certain pH-redox potential combinations enhanced soluble mercury levels, though we can only speculate on the processes involved in this study. Moderately acid, reduced (pH 5.0, -150 mV) and weakly alkaline, oxidized (pH 8.0, 500 mV) conditions were found to enhance soluble levels of mercury. The similarity of dissolved mercury in both total soluble and the soluble complexed forms indicated that most of the dissolved mercury was bound as inorganic, nonionic, or negatively charged complexes (31), or complexed with soluble organics. Lindberg and Harriss (32) have suggested that the formation of polysulfide complexes of relatively high solubility may account for greater levels of dissolved mercury under the pH and sulfide conditions found in their study of' mercury in Everglades sediments. Although soluble sulfides were not measured in this study, the increasing total sulfide content measured (28) with decreasing incubation pH under strongly reducing conditions may contribute a favorable environment for soluble mercury polysulfide complexes. Wollast, Billen, and Mackenzie (33)reported that although the extremely insoluble mercuric sulfide (cinnabar) will form in reducing environments. dissolved levels of mercury may be increased in more strongly reducing conditions by conversion of the mercuric ion t o the free metal form. Although this Mobile Bay sediment-water system may differ from the Belgium River water described by Wollast e t al., it is interesting to note that a t p H 5.0, -150 mV, where we found the highest level of soluble noncationic mercury, their Eh-pH diagram showed the comparatively soluble free metallic form to be stable compared to less soluble sulfide forms a t -150 mV and pH 6.5 and 8.0. The reasons for the increase in recovery of labeled soluble mercury under oxidized, weakly alkaline (500 mV, p H 8) conditions are also unknown. However, these conditions may favor: (a) the formation and solubility of smaller molecular weight organics which may complex with mercury, or (b) the formation of the relatively soluble mercuric hydroxide (31, 33). The response of exchangeable *03Hglevels to changes in pH and redox potential was very similar to that found for the water-soluble fractions. At -150 mV at p H 5.0 and 6.5, oxalate extractable (reducible) mercury was greater than at 50 mV (Table 11).Though the magnitude of this difference was small, this greater recovery a t -150 mV than at 50 mV was thought to be related to the generally larger recovery of poorly crystalline iron under strongly reduced conditions in this sediment material ( 2 8 ) . Except for the anomaly noted above at -150 mV, levels of
Table 11. Effect of pH and Redox Potential on the Chemical Form of *03Hg Added to Mobile Bay Sediment Suspensions % fraction
total dissolved dissolved, noncat ionic exchangeable reducible DTPA extractable a
+50
0.43a fO.12C 0.29 f0.16 0.22 f0.05 2.02 f0.21 1.44 f0.21
Mean of duplicate?samples.
Of *03Hg
PH 5.0
-
+250
n.d.
n.d.
0.01
0.01 fO.OO 0.05 fO.O1 2.36 f0.24 5.87 f0.62
fO.OO
0.03 fO.OO 0.62 f0.24 1.74 f0.08
Not detectable.
-
+500
0.01 fO.OO 0.01 fO.OO 0.08 fO.O1 3.90 f0.38 6.38 f0.59
recovered at pH and redox potential, rnV PH 6.5 +50 +250 +500 -
0.12 0.02 fO.O1 f O . 0 1 0.15d O.Old
0.07 f0.04 0.02d
0.03 f0.02
0.09 f0.02 1.84 f0.46 1.89 f0.87
0.05 f0.02 3.52 f0.16 9.39 f3.22
0.07 f0.03 6.14 f0.58 11.42 f 1.85
0.03d
0.03 40.00 0.04 fO.OO
Standard deviation.
0.02 fO.O1 1.12 f0.22 4.84 f2.18
0.04 f0.03 0.42 f0.14 3.75 fO.ll
+50
PH 8.0
0.04 f0.02 0.03 fO.OO 0.07 fO.OO
1.28 f0.14 7.68 f0.46
+250
+500
0.05 fO.O1 0.06 kO.01 0.08 f0.02 1.82 f0.62 9.14 f0.46
0.19 fO.O1 0.17 f0.02 0.19 f0.04 3.28 fO.01 9.31 f0.77
Lost one sample.
Volume 14, Number 4, April 1980
433
203Hg,Mobile Bay, PH 8.0 I
gl
Mobile Bay.
pH 6.5
50t
Extractable
8
210Pb,
7H
u 654-
f
;
U
01
I
321
Exchangeable
0.15-
0.05-,
?
I
I
I
1
1
Potential, mv Figure 2. The effectsof redox potential on the distribution of 2 0 3 ~ among selectedchemical forms i n Mobile Bay sediment suspensions incubated at pH 8.0 Redox
reducible mercury increased with increasing oxidation intensity. DTPA extractable 203Hgincreased as oxidation intensity increased, such that approximately 10% of that added was recovered under well-oxidized conditions. While results of this study are in basic agreement with a previous published report ( 3 4 )on the relative distribution of sediment-bound mercury between ferromanganous oxide (reducible) and organic forms, our research demonstrates that oxidation-reduction conditions affect the amount and/or bonding strength of mercury in these forms. Figure 2 summarizes the effects of redox potential on the sequentially extracted forms of 203Hga t pH 8.0. At all pH levels, the labeled mercury recovered in potentially available (reducible and DTPA extractable) forms was greatest under well-oxidized conditions. The increasing recovery of DTPA extractable mercury with increasing oxidation conditions indicates that large molecular weight humic materials may be qualitatively and quantitatively altered by sediment oxidation such that its mercury binding capacity or bonding strength is reduced as the sediment-water system becomes oxidized, as suggested by Feick, Johanson, and Yeaple (35). Most of the labeled mercury was not recovered by the sequential fractionation procedure and was thought to still be bound to large molecular weight organics not solubilized by the DTPA extraction, though sulfide precipitation could be contributing to this low recovery under strongly reducing conditions. Less than 1-2% of the added labeled mercury was recovered in acid permanganate traps attached to the outlet gas stream. Lead. The sediment suspensions were amended with carrier-free zloPb to determine the effects of controlled pH and redox potential on the distribution of lead in dissolved and exchangeable forms where levels too low for detection by flame atomic absorption were anticipated. For the remaining fractions of the sequential chemical extraction procedure, recoveries of both labeled and indigenous lead are reported (Table 111). Little, if any, dissolved labeled lead was detected at any pH-redox potential combination. However, exchangeable labeled lead was measured under moderately acid pH conditions. I t is apparent that pH influenced the recovery of labeled lead in the exchangeable form more than oxidation intensity. Reducible lead, both labeled and indigenous, increased with sequential increases in redox potential and was less affected 434
Environmental Science & Technology
Redox Potential, mv The effectsof redox potential on the distribution of *loPb among selected chemical forms in Mobile Bay sediment suspensions incubated at pH 6.5
Figure 3. ~
by pH. Approximately half of the labeled lead was recovered in the reducible fraction under oxidized conditions. Leland, Shukla, and Shimp (13) also reported that much of the potentially available lead in Lake Michigan sediments was bound in the reducible form. If reducible levels of indigenous lead are compared with the total lead content of this sediment material (36 pg/g), a lower proportion of the indigenous lead was recovered in this chemical form compared to the added labeled lead. This is reasonable since essentially all of the added labeled lead was in potentially mobile chemical forms, whereas a considerable proportion of the indigenous lead was likely bound in unavailable forms within the crystalline lattice structure of primary and secondary minerals. Much of the remaining labeled and potentially reactive indigenous lead was recovered in the DTPA extractable form (Table 111).Increasing the redox potential from -150 to +50 mV resulted in a considerable increase in DTPA extractable lead a t pH 5.0 and 6.5, while subsequent increases in oxidation resulted in little change to moderate decreases in chelate extractable lead. The initial increase in chelate extractable lead with the first oxidation increment probably reflects a reduction in the lead binding activity of natural organics as the sediment materials become oxidized. The lack of a substantial, sequential increase in DTPA extractable lead as redox potential increased above 50 mV was due to most of the potentially available lead being bound and previously extracted in the reducible form under oxidized conditions. Figure 3 summarizes the effects of redox potential on the recovery of labeled lead incubated at pH 6.5 in the Mobile Bay material. Unlike the results for mercury, most of the labeled lead was recovered under oxidized conditions by the sequential chemical fractionation procedure in potentially available forms as was approximately half of the total indigenous lead. Zinc. Suspension pH had a large effect on total dissolved and exchangeable levels of zinc (Table IV). Total dissolved and exchangeable zinc were greatest at pH 5.0 and decreased to undetectable levels at pH 8.0 (except for the 500-mV treatment where exchangeable zinc found in one of the two subsamples was attributed to contamination). All of the dissolved zinc was retained by a cation chelating resin, indicating the soluble zinc was likely present in free cationic form. Under strongly reduced conditions (-150 mV), levels of exchangeable zinc were substantially lower than under more oxidized conditions a t both pH 5.0 and 6.5. This was also
~~
~~
Table 111. Effect of pH and Redox Potential on the Chemical Form of Indigenous and Added *lOPbin Mobile Bay Sediment Suspensions fraction
total dissolved dissolved, noncationic exchangeable reducible DTPA extractable
reducible DTPA extractable a
Not detectable.
-
redox potentlai, mV, at pH 5.0 +50 +250 +500
n.d.a
0.026 0.01 f0.00C fO.00
0.02 fO.O1
0.01
n.d.
0.01 fO.OO
n.d.
5.33 f0.33 27.64 f1.63 27.17 f3.17
4.81 f0.13 44.62 f0.20 41.68 f2.43
fO.O1 4.01 f0.06 51.86 f0.24 32.20 f0.60
n.d.a
6.0b f1.2 6.0 fO.0
9.9 f2.4 4.8 fO.0
4.1 f0.3
Mean of duplicate samples.
3.14 f0.04 57.19 f1.54 30.60 f0.48
redox potential, mV, at pH 6.5 +50 +250 +500
-
% of 210PbRecovered 0.02 n.d. 0.02 f0.02 f0.02 n.d.d
0.11 f0.04 23.18 f7.16 27.53 110.97
n.d.d
n.d.d
0.28 fO.O1 46.75 f14.50 37.37 f10.66
0.12 40.01 49.21 12.20 34.26 13.52
n.d.
fO.l
4.0 f0.4
3.9 f1.5
f4.5 5.3 f0.9
f6.7 4.6 41.0
redox potential, mV, at pH 8.0 +50 f250 f500
0.01
fO.O1 n.d.d
0.11 f0.03 49.86 f2.06 35.28 f2.26
p g of Pb/g of Oven-Dry Solidse n.d. 3.2 4.7 10.3
11.6
-
n.d.
0.04 f0.02 32.13 f1.39 34.78 f0.27 5.5
f1.2 5.9 f0.9
fO.l
6.4 fl.O
0.14 f0.04 0.07 fO.O1 0.04 f0.02 37.41 f0.81 36.95 f2.41
0.04 10.05 0.06 f0.02 0.01 10.01 44.60 50.07 39.38 f6.21
0.06 f0.03 f0.02 0.02 fO.01 50.90 f1.77 34.96 f1.26
6.4 fl.1 6.8 f1.5
5.6 fO.0 6.2 f1.5
10.2 f1.2 6.1 10.5
0.05
Standard deviation. dLost one sample. e Indigenous Pb determined by flame atomic absorption.
Table IV. Effect of pH and Redox Potential on the Chemical Form of Indigenous Zn in Mobile Bay Sediment Suspensions a fraction
-
total dissolved
0.13b f0.18d
dissolved, noncationic exchangeable
n.d. e
reducible DTPA extractable a
1.2 f0.7 104.8 f9.3 10.3 f2.5
+50
PH 5.0 +250
8.81 f0.88 0.13 f0.18 18.4 f0.2 76.6 f1.7 7.7 62.2
Determined by flame atomic absorption.
pg of Znlg of oven-dry solids at pH and redox potential, mV PH 6.5 +500 +50 +250 +500 -
12.33 40.96
11.72 f0.18
1.26c f1.41
n.d.
n.d.
n.d.'
25.2 f0.7 70.8 f3.3 5.9 fl.0
24.8 f0.2 74.4 f5.2 6.0 fO.l
0.7 fO.l
25.7 f8.6 20.1 f6.1
Mean of duplicate samples.
Conclusions This study demonstrates that changes in p H and oxidation-reduction potential may result in changes in the chemical
+500
1.01 40.64
0.57 410.11
0.98 10.12
n.d.
n.d.
n.d.
n.d.
n.d.'
n.d.'
n.d.'
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
n.d.
118.8 52.5 11.8 f2.0
114.0 411.9 9.6 f2.0
107.1 45.1 14.4 f6.8
1.2c 41.6 112.5 f4.3 12.1 41.1
3.0 f0.3 62.2 f12.0 11.5 f5.9
4.6 40.2 71.0 f7.1 8.8
f4.5
Possible contamination.
found a t pH 5.0 for dissolved zinc. Suspected contamination in one subsample masked treatment effects, if any, a t pH 6.5, -150 mV. In the Mobile Bay sediment suspension, considerable sulfide was measured a t -150 mV a t all pH levels, and no sulfide was detected under more oxidized conditions. Thus, the formation of sparingly soluble zinc sulfide was thought to be contributing to the low levels of readily available zinc in the strongly reducing sediment suspensions. Complex formation with insoluble, large molecular weight humic materials may have also contributed to a decrease in soluble zinc under strongly reduced conditions. A similar decrease in soluble and exchangeable zinc levels at -150 mV was found for a Mississippi River sediment suspension which contained no measurable sulfide (:?8). More of the zinc was recovered in the reducible form than other chemical forms studied (26-119 pg/g). Levels of zinc recovered in the DTPA fraction were much less than levels recovered in the reducible form. This again suggested that adsorption or coprecipitation with colloidal hydrous oxides was a major process immobilizing potentially available zinc in the sediment-water system. The importance of this process in regulating zinc availability has been reported by a number of investigators (36-38).The effect of redox potential on the chemical form of' zinc is summarized in Figure 4.
PH 8.0 f50 +250
4
7.9 41.8 95.3 f0.8 11.1 f2.5
Standard deviation. e Not detectable.
\
'Lost one sample.
Z n , Mobile Bay, pH 5 . 0 (Total Zn, 2 0 5 p g / g )
301
Exchangeable
Total Dissolved
10-
DTPA Extractable A
1
I
I
I
L
Figure 4. The effects of redox potential on the distribution of indigenous zinc among selected chemical forms in Mobile Bay sediment suspensions incubated at pH 5.0
availability of mercury, lead, and zinc in sediment-water systems. Such changes could occur during dredging and disposal of contaminated sediments. Where dredged material is released into quiescent open waters for disposal, it is likely Volume 14, Number 4, April 1980
435
that the pH-redox p o t e n t i a l condition of the bulk solid material is not significantly affected. During the relatively short t r a n s i t t i m e of hydraulically d r e d g e d solids dispersed in open waters, and particularly w h e r e mechanical dredging is used, t h e r e m a y b e little o p p o r t u n i t y for most of the r e d u c e d , b u l k solid m a t e r i a l t o oxidize before settling to the b o t t o m at a disposal site. HoCever, some changes i n dredged material p H and r e d o x p o t e n t i a l m a y occur t o interstitial w a t e r and susp e n d e d colloidal solids m i x e d w i t h t h e oxygenated surface water for e x t e n d e d periods, s u c h as at a disp,osal s i t e w i t h a high-energy water column. Similar changes m a y occur t o bulk d r e d g e d solids where u p l a n d disposal m e t h o d s are used. T h i s could affect toxic m a t e r i a l availability to a q u a t i c and benthic organisms at an o p e n water disposal s i t e as well as increase leaching and p l a n t availability of m e t a l s at upland disposal sites. M o s t s t u d i e s t o date have s h o w n little o r n o s h o r t - t e r m release of toxic m e t a l s to a c t u a l o r laboratory s i m u l a t e d disposal s i t e w a t e r columns. T o x i c m e t a l release resulting f r o m long-term a l t e r a t i o n s of the d r e d g e d materials physicochemical condition m a y b e of greater environmental concern. Thus, selecting a disposal a l t e r n a t i v e based o n long-term changes in the physicochemical e n v i r o n m e n t of a c o n t a m i n a t e d s e d i m e n t could be a valuable m a n a g e m e n t tool i n minimizing a d v e r s e e n v i r o n m e n t a l effects f r o m toxic m e t a l s in dredged sediments. F o r t u n a t e l y i n t h e Mobile B a y s e d i m e n t material, m o s t of the chemical t r a n s f o r m a t i o n s i n d u c e d b y a c h a n g e in t h e physicochemical e n v i r o n m e n t occurred between potentially available f o r m s and not between readily available forms. Comparatively s m a l l changes were f o u n d in the readily available soluble and exchangeable levels of these metals. However, i t s h o u l d be kept i n m i n d that relatively s m a l l changes in mobilization and availability m a y b e magnified in food webs. It s h o u l d also b e noted that t h i s is a chemical availability s t u d y suggesting possible effects on biological availability. U p t a k e s t u d i e s using a similar e x p e r i m e n t a l app r o a c h should be c o n d u c t e d to assess the biological signific a n c e of the observed physicochemical effects on transform a t i o n of m e t a l s a m o n g various chemical forms.
Literature Cited (1) Boyd, M. B., Saucier, R. T., Keeley, J. W., Montgomery, R. L.,
Brown, R. D., Mathis, D. B., Guice, C. J., “Disposal of Dredge Spoil. Problem Identification and Assessment and Research Program Development”, Technical Report H-72-8, U.S. Army Engineers Waterways Experiment Station, Vicksburg, Miss., 1972. (2) Lee, G. F., Enuiron. Sci. Technol., 10,334-8 (1976). (3) Smith, D. D., Enuiron. Sci. Technol., 10,328-33 (1976). (4) Gambrell, R. P., Khalid, R. A., Patrick, W. H., Jr., in “Dredging and Its Environmental Effects”, Krenkel, P. A,, Harrison, J., Burdick, J. C., 111, Eds., American Society of Civil Engineers, New York, 1976, pp 418-34. (5) Gambrell, R..P., Khalid, R. A., Collard, V. R., Reddy, C. N., Patrick, W. H., Jr., in “Dredging: Environmental Effects and Technology”, Proceedings of WODCON VII, World Dredging Conference, San Pedro, Calif., 1976, p p 579-604. (6) Krauskopf, K. B., Geochim. Cosmochim. Acta, 9,1-328 (1956). ( 7 ) Engler, R. M., Patrick, W. H., Jr., Soil Sci., 119, 217-21 (1975). (8) Morel, F., McDuff, R. E., Morgan, J. J., in “Trace Metals and Metal-Organic Interactions in Natural Waters”, Singer, P. C., Ed., Ann Arbor Science Publishers, Ann Arbor, Mich., 1974, pp 157200. (9) Taylor, R. M., McKenzie, R. M., Aust. J . Soil Res., 4, 29-39 (1966). (10) Jenne, E. A., Adu. Chem. Ser., No. 73,337-88 (1968).
436
Environmental Science & Technology
(11) Lee, G. F., paper presented at the Symposium on Heavy Metals in the Environment, Vanderbilt University, Nashville, Tenn., December 1973. (12) Windom, H. L., “Investigations of Changes in Heavy Metals Concentrations Resulting from Maintenance Dredging of Mobile Bay Ship Channel, Mobile Bay, Alabama”, Report for Contract No. DACW01-73-C-0136, U.S. Army Corps of Engineers, Mobile District, 1973. ( 1 3 ) Leland, H. V., Shukla, S. S., Shimu. N. F.. in ref 8. DU 89-129. (14) Loganathan, P., Burau, R. G., F u e k e n a u , D. W., So;/ Sci SOC. Am J., 41,57-62 (1977). (15) Huang. C. P.. Elliott. H. A,. Ashmead. R. M.. J. Water Pollut. Contr. F&., 49,’745-56’(1977j. (16) Schnitzer, M., Skinner, S. I. M., Soil Sci., 102,361-5 (1966). (17) Schnitzer, M., Skinner, S. I. M., Soil Sci., 103,247-52 (1967). (18) Stevenson, F. J., Ardakani, M. S., in “Micronutrients in Agriculture”, Dinauer, R. C., Ed., Soil Science Society of America, Madison, Wis., 1972, pp 79-114. (19) Shimp, N. F., Schleicher, J. A., Ruch, R. R., Heck, D. B., Leland, H. V., Enuiron. Geol. Notes, Ill. State Geol. Surv., No. 41 (1971). (20) Keeley, J. W., Engler, R. M., “Discussion of Regulatory Criteria for Ocean Disposal of Dredged Materials: Elutriate Test Rationale and Implementation Guidelines”, Miscellaneous Paper D-74-14, US.Army Engineers Waterways Experiment Station, Vicksburg, Miss., 1974. (21) Hem, J. D., Water Resources Res., 8,662-79 (1972). (22) McKeague, J. A,, Day, J. H., Can. J. Soil Sci., 46, 13-22 (1966). (23) Lindsay, W. L., Norvell, W. A., “Development of a DTPA Micronutrient Soil Test”, from Agronomy Abstracts of 61st Annual Meeting of American Society of Agronomy, Madison, Wis., 1969. (24) Evans, L. T., J. Soil Sci., 10, 110-8 (1959). (25) Rashid, M. A,, Chern. Geol., 13,115-23 (1974). (26) Cox, F. R., Kamprath, E. J., in ref 18, pp 289-317. (27) Verloo, M. G., Faculty of Agronomy, University of Gent, Belgium, private communication, 1974. (28) Gambrell, R. P., Khalid, R. A., Verloo, M. G., Patrick, W. H., Jr., “Transformations of Heavy Metals and Plant Nutrients in Dredged Sediments as Affected by Oxidation-Reduction Potential and pH. Volume 11: Materials and MethoddResults and Discussion”, Contract Reuort D-77-4. U S . Armv Engineers Waterwavs Exueriment Statidn, Vicksburg, Miss., 1677. (29) Whittie. L. D.. in “Methods of Soil Analvsis. Part I”. Black. C. A., Ed., American Society of Agronomy, Inc:, Madison, Wis., 1965, pp 671-98. (30) Caldwell, A. G., Department of Agronomy, Louisiana State University, private communication, 1975. (31) Hem, J. D., in “Mercury in the Environment”, Geological Survey Professional Raper 713, U S . Government Printing Office, Washington, D.C., 1970. (32) Lindberg, S. E., Harriss, R. C., Enuiron. Sei. Technol., 8,459-62 (1974). (33) Wollast, R., Billen, G., Mackenzie, F. F., in “Ecological Toxicology Research”, McIntyre, A. R., Mills, c. F., Eds., Plenum Publishing Co., New York, 1975, pp 145-66. (34) Crecelius. E. A.. Bothner. M. H..,Caruenter. . , R.., Enuiron. Sci. ‘ Technol., 9,’325-33 (1975). (35) Feick. G.. Johanson. E. E., Yeade. D. S..“Control of Mercurv Contamination in Freshwater Sediments”, EPA-R2-72-077, p r i pared for Office of Research and Monitoring, U S . Environmental Protection Agency, Washington, D.C., 1972. (36) Kinniburgh, D. G., Jackson, M. L., Syers, J. K., Soil Sci. SOC. Am. J., 40,796-9 (1976). (37) Shuman, L. M., Soil Sci. SOC.Am. J., 41,703-6 (1978). (38) Kalbasi, M., Racz, G. J., Loewen-Rudgers, L. A., Soil Sci., 125, 146-50 (1978).
Received for review July 17,1978. Accepted December 31,1979. This study was supported by the Dredged Material Research Program of the Enuironmental Effects Laboratory, U.S. Army Corps of Engineers Waterways Experiment Station, Vicksburg, Miss. (Contract NO.DACW-39-74-CO076).
