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The influence of sediment redox chemistry on chemically active forms of arsenic, cadmium, chromium, and zinc in estuarine sediment
Environment
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THE INFLUENCE OF SEDIMENT REDOX CHEMISTRY ON CHEMICALLY ACTIVE FORMS OF ARSENIC, CADMIUM, CHROMIUM, AND ZINC IN ESTUARINE SEDIMENT Tingzong Guo, R.D. DeLaune*, and W.H. Patrick, Jr. Wetland Biogeochemistry
Institute, Louisiana State University, Baton Rouge, LA 70803, USA
EI 9608-I 76 M (Received 14 August 1996; accepted IO February 1997)
Kinetics and chemical fractionation procedures were used to quantify the effect of the sediment redox (Eh) condition on the behavior of As, Cd, Cr, and Zn in the bottom sediment collected from a Louisiana coastal site receiving produced water discharge. Sediment samples were incubated in microcosms in which Eh-pH conditions were controlled. Sediment was sequentially extracted to determine metals in various chemical fractions (water soluble, exchangeable, bound to carbonates, bound to iron and manganese oxides, bound to insoluble organic and sulfides) and the chemical inactive fraction (mineral residue). Under oxidizing conditions, As, Zn, and Cr behavior was governed by redox chemistry of Fe(II1) and Mn(IV) oxides. Cd transformations were controlled by both Fe(II1) and Mn(IV) oxides and carbonates. Under a reducing condition, the behaviors of Zn and Cr were controlled primarily by insoluble large molecular humic material and sulfides; the behavior of Cd was controlled by carbonates. When sediment redox potential increased, the affinity between Fe(II1) and Mn(IV) oxides and As, Cd, Cr, and Zn increased. When sediment redox potential decreased, the affinity between carbonates and Cd and Zn increased; the affinity between insoluble sulfides, large molecular humic matter and As, Cd, Cr, and Zn increased; the soluble Cd and Zn decreased; the soluble As and Cr remained constant. Results suggest reducing sediment conditions would reduce Cd and Zn toxicity. Under reducing or anaerobic conditions, the solibilization rate constants (mg kg-’d-‘) for As, Cr, Cd, and Zn bound to Fe(II1) and Mn(IV) oxides were -0.88, -0.32, -0.01, and -6.5, respectively; the rate constants (mg kg’ de’) for dissolved Cd and Zn were -0.09 and -1.78, respectively. 81997Elsevier Science Ltd
INTRODUCTION Koons et al. 1977; Lyssj 1981). As the produced water enters wetland environments, heavy elements can enter the sediment column. Analysis of total element concentration in sediment can quantify the degree of trace element enrichment. Total element content does not provide information on transformation and mobilization of trace elements. Speciation studies can: 1) provide an insight into element distribution patterns; 2) identify metal bioavailability and toxicity in ecosystems; and, 3) explain
Petroleum production and recovery activity in the Louisiana coastal zone can discharge a considerable amount of produced waters (Boesch and Rabalais 1989). The produced waters can contain elevated concentrations of heavy metals compared to the receiving water. Element content in produced water can include barium, cadmium, chromium, iron, mercury, manganese, strontium, and thallium (Tillery et al. 1981; *Corresponding author.
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transformation and mobility of metal species. Trace element species distribution has been studied in various ways (DeLaune and Smith 1985; Gambrel1 et al. 199 1a and b; Giblin et al. 1986; Giesy et al. 1977; Keller and Vedy 1994). These techniques have included multistep extractions in which a chemical solution removes various element forms. Such fractionation schemes can provide information on the general behavior of heavy elements in sediment and provide an estimate of their potential mobility (Keller and Vedy 1994). Metals in sediment are generally considered to be present in the following forms: water soluble, exchangeable, carbonate bound, ferric and manganic oxide bound, organic matter and sulfide bound, silicate bound, and residual. The oxidation/reduction state (redox potential) of sediment is an important parameter affecting heavy metal transformation. The redox conditions (Eh) of wetland soil and sediments vary widely from approximately +500 mV (surface soils) to approximately -300 mV (strongly reducing soils) (Patrick and DeLaune 1977). Sediment redox levels can greatly affect toxic metal uptake by plants (Gambrel1 and Patrick 1989; Giblin et al. 1986), leaching losses of toxic metals by runoff or ground water (Folsom et al. 1988; Palermo et al. 1989), but there is little information on redox chemistry of toxic metals in different geochemistry forms in sediment. Heavy element species kinetics in wetland soil kinetics are influenced by many factors. These factors include temperature, organic matter, surface activity of Fe and Mn compounds, microorganism activity and other sediment characteristics. Since it is difficult to study these factors individually, the factors were combined in this study into one parameter-the rate constant of the assumed zero-order reaction, the socalled pseudo-zero-order reaction model. The pseudozero-order reaction model was used in this study. When the rate constant is positive, the metal content increases in the specific fraction. When the rate constant is negative, the metals are removed from the specific fraction. When considering trace elements bound to the organic matter and sulfide phase, it was assumed that the formation of insoluble organic matter (complexation of insoluble, large molecular weight humic fraction) and sulfides was independent (Gambrel1 and Patrick 1978; Gambrel1 et al. 1980). At sediment Eh = -130 mV or Eh > -130 mV, the changes in heavy elements in this fraction were due to insoluble, large molecular weight humic substances.
T. Guo et al.
Surface of clays, organic matter, and iron oxides in sediment will absorb or desorb heavy elements when the ionic composition or Eh-pH changes (Keller and Vedy 1994; Khalid et al. 1981; Gambrel1 et al. 1980). Significant heavy element content is also associated with sediment carbonates (Ramos et al. 1994; Gambrel1 1994). This fraction would be susceptible to pH change. Iron and manganese oxides existing as nodules and concretions, cemented between particles or on particle coatings in sediment are excellent scavengers for heavy elements and are affected by sediment Eh and pH change (Feijtel et al. 1988; Levy et al. 1992). Heavy elements are also bound to various insoluble organic forms such as living organisms, detritus, and humic material (Gambrel1 et al. 1980; Ramos et al. 1994). Sediment redox conditions can affect the degradation and solubility of such organic material and then influence the release of heavy elements. Heavy elements can also exist as sulfides under anaerobic conditions (Gambrel1 et al. 1980; 199 1b) which are susceptible to Eh and pH changes. Heavy elements found in primary and secondary minerals are relatively stable in a natural sediment environment (Gambrel1 1994). This research examines the kinetics and transformations of As, Cd, Cr, and Zn in estuarine sediment at a site in coastal Louisiana receiving produced water discharge. The effects of the redox potential (Eh) on the kinetics of transformation of different forms of the heavy elements in sediments are detailed. MATERIAL AND METHODS Sediment
Sediment samples were collected from a produced water discharge located near a brackish marsh (Fig. 1). The sediment pH was 7.0. The sediment contained 20 g kg-’ Fe, 0.4 g kg-’ Mn and 1 g kg-’ Ba. The metal content of the sediment was analyzed and four times the original metal content of Cd, Zn, Cr, and As was added to the sediment. The total metal content (mg g-’ dry sediment) of the sediment after spiking was Cd=2 1, Zn=SOO, Cr=2 10, and As=3 70. Incubation
of the soil
Wet sediment equivalent to 200 g of the dry sediment was placed in each of 5 microcosms. Sufficient water (5 g L-’ salinity) was added to the microcosm to obtain a sediment to water ratio of 1: 10. The microcosm originally described by Patrick et al. (1973) was fitted with two bright Pt electrodes, a calomel reference elec-
301
Sediment redox chemistry in estuarine sediment
Fig. 1. Location of the Lirette (LRT) site from which sediment samples were collected.
trode connected to the suspension by a saturated potassium chloride-agar salt bridge, a serum cap, a thermometer, an inlet tube for nitrogen or oxygen, and an outlet which was submerged in water to prevent atmospheric oxygen from diffusing into the microcosm. The soil suspensions were maintained at 26°C. Sediment pH was adjusted and maintained at 7.0 with either hydrochloric acid or sodium hydroxide. The bright Pt electrodes immersed in the suspension were connected to a potentiometer for redox potential measurements. Sediment was incubated for 60 d. Element contents in the sediment suspension were selectively extracted as described by Shannon and White (1991):
Sequentialextractionprocedures Watersolublephase. Samples from the sediment suspensions were taken at selected intervals. The suspension samples were centrifuged and the supernatant filtered through a 0.45 pm membrane filter. This supernatant was assumed to be water soluble. The remaining sediment was extracted sequentially as described below. The sediment sample was kept under nitrogen or oxygen free atmosphere during extraction. Following removal of the water soluble phase, the sediment was sequentially extracted into fine fractions (Fl to F5) described below.
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2000
1500
+
Fl
G
........0 ........
F2
? y
----o----
F3
---B---
Soluble
e
1000
5
5OO
0 0
10
20
430
0
-80
30 -130
40 -150
Fig. 2. The effect of Eh on the distribution
1) Fl-Exchangeable phase: The solid phase from the water soluble fraction was extracted at room temperature for 30 min with 8 mL 0.5 M Mg(NO,)Jg dry weight sediment, adjusted to pH 7.0 with nitric acid. The samples were agitated continuously. 2) F%-Bound to carbonate phase: The sediment residue from Fl was leached at room temperature for 5 h with 8 mL, 1M NaOAc, adjusted to pH 5.0 with acetic acid for 1 g dry weight sediment. These samples were also agitated continuously. 3) F3-Bound to iron and manganese oxides phase: The sediment residue from F2 was extracted at 96°C for 6 h with 20 mL 0.08 M NH,OH HCl in 250 mL/L acetic acid for 1 g dry weight sediment. These samples were occasionally agitated. 4) F4-Bound to organic matter and sulfides phase: For 1 g of dry weight sediment, the sediment residue from F3 was extracted at 85°C for 2 h with 3 mL 0.02 M HNO, and 5 mL 300 mL/L H202 (adjusted to pH = 2.0 with IWO,) was added, and extraction continued at 85°C for another 3 h. The sample was then cooled, 5 mL 3.2 M NI&OAc in 20 mL/L HNO, was added, and the sample was diluted to 20 mL with deionized water. The samples were agitated continuously for 30 min. NHOAc was added to prevent adsorption of extracted metals onto the oxidized sediment. 5) FS-Mineral matrix phase: The sediment residue from F4 was extracted with 25 mL concentrated HNO, for 1 g of dry weight sedi-
50
60 -170
T(day) Eh(mV)
or content of Fe in chemical fractions.
ment at 105 “C, the sediment was digested until 5 mL solution was left, and the sample was diluted to 25 mL with deionized water. The above sequential extractions were conducted in 250 mL centrifuge tubes which prevented any loss of sediment between the successive extractions. Separation was conducted by centrifuging at 5000 rpm for 30 min. Supernatants were filtered using 0.45 urn millipore filters and then analyzed for metals. The residues were rinsed with 8 mL deionized water for 1 g dry weight sediment and centrifuged at 5000 rpm for 30 min. These second supernatants were discarded. Metal concentrations in water soluble and the chemical extracts were determined by ICP. Quality assurance was conducted by spiking extracts with certified element standards. RESULTS AND DISCUSSION
Fe and Mn behavior The effects of Eh (redox potential) on the contents of Mn and Fe in the various chemical fractions are shown in Figs. 2 and 3, and 4 and 5, respectively. As sediment Eh decreased from +430 mV to 0 mV, Mn associated with carbonates decreased (K= -6.7 mg kg’ d-l). Also, Mn(IV) oxides were apparently reduced to soluble Mn(I1) under more reducing sediment conditions (the pseudo-zero-order constant of Mn(IV) oxides reduction being - 1.3 mg kg’ a’). As the result of the above two reactions dissolved, Mn(I1) concentration increased
Sediment redox chemistry in estuarine sediment
309
e _.I.
Soluble
0 .I....
Fl
-II_ 0 _I__
F2
__- & _--
F3 F4
0
lo
20
430
0
-80
30
40 -130
50
-150
60
T(day)
Eh(mV)
-170
Fig. 3. The effect of Eh on the percentage of Fe in the various chemical fractions.
0
10
20
430
0
-80
30 -130
40 -150
Fig. 4. The effect of Eh on the distribution
significantly (K=l2.7 mg kg’ d’). In the Eh range between 0 mV and -130 mV, dissolved Mn(I1) decreased as the result of Mn becoming associated with carbonates. At sediment Eh below -130 mV, sulfides are produced. At such low Ehs, dissolved Mn(I1) concentration was lowered through Mn becoming associated with carbonates and sulfides and insoluble organic matter (complexes of the Fe and Mn with insoluble large
50
60 -170
__(I_
Fl
I._.
F2
0 ...
I___ 0 _I__
F3
-__ A _I_
F4
___ w ___
Soluble
T(W) Eh(mV)
of Mn in the various chemical fractions.
molecular humic material). The rate constant for the removal of the dissolved Mn(I1) at this Eh range was 0.94 mg kg-’d-‘. The rate constants for the formation of insoluble carbonates, sulfides, and Mn with humic substances were 0.49, 1.8, and 0.07 mg kg’ d-‘, respectively. Under oxidizing sediment conditions, Mn(IV) oxides and carbonates were the predominate form of Mn pre-
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Fl F2 F3 F4 w-v
a
m-w
Soluble
000 0
10
20
430
0
-80
30 -130
40 -150
Fig. 5. The effect of Eh on the percentage
sent. Under a reducing condition, Mn(I1) associated with exchangeable fractions and sulfides were the primary Mn forms present. Figure 2 shows the effect of sediment Eh (redox potential) on the content of Fe in the various chemical fractions. As Eh decreased, Fe(II1) oxides were microbially reduced to soluble Fe(II), therefore increasing soluble Fe concentration (K=42.5 mg kg-’ d-‘) in soluion. When Eh further decreased to - 130 mV, a reduction in iron oxides was observed which was attributed to FeS precipitation. Based on the fractionation data at all redox levels studied, dissolved Fe(I1) was also removed through Fe becoming associated with insoluble organic matter. The above two factors resulted in decreased dissolved Fe at Eh < - 130 mV (K= -25.8 mg kg’ d-l). The rate constants of the removal reactions for dissolved Fe(I1) are 40.7 mg kg’ d’ and 3 1.4 mg kg’ d-l for the formation of sulfides and complexation of insoluble, large molecular humic, respectively. Under a reducing condition (Eh > - 130 mV), the reduction of Fe(II1) oxides likely occurred by direct microbial reduction (K= -42.4 mg kg’ d-l) involving organic carbon turnover. This part of Fe(II1) reduction was equivalent to approximately 207 1 mg kg’ (20% of total reducible Fe(II1)). At sediment (Eh < -130 mV), Fe(II1) reducing microorganisms can inhibit sulfate reduction by out competing sulfate reducer for electron donors (Lovley and Phillips 1987). A portion of Fe(II1) reduction was assumed to have occurred by sulfides
50
60 -170
Way) Et-W4
of Mn in the various chemical fractions.
reducing Fe(II1) oxides. Only a small amount of the Fe(II1) reduction at these redox levels can occur by direct bacteria reduction involving organic carbon turnover (Jacobson 1994). Accordingly, in this study, 8571 mg kg-’ (80% of total reducible Fe(II1)) was apparently reduced by the sulfide oxidation (K= - 17 1.5 mg kg’ d-l). The rate constant of indirect Fe(II1) oxide reduction by sulfides is significantly greater than that by direct bacteria reduction. During anaerobic incubation, approximately half of the reduced Fe(II1) was converted to carbonate bound Fe(I1). Under an oxidizing sediment condition, Fe(II1) oxide predominated with dissolved Fe concentrations controlled by Fe(II1) oxides. Under reducing sediment conditions, sulfide and insoluble large molecular humic bound Fe(I1) was the dominant fraction controlling iron behavior.
As behavior
Figure 6 shows the effect of Eh on the level of As in the water soluble chemical fraction. When Eh decreased to 0 mV, As(V) was reduced to As(II1) (K= -0.21 mg kg’ d-l). At sediment Eh=O to -100 mV, dissolved arsenic concentration was essentially zero. This may be due to the fact that the fresh As(II1) which was formed from As(V) reduction became insoluble. Following manganic oxide and ferric oxide reduction, the As bound with these oxides decreased (K= -0.88 mg kg’ a’). Similar results were reported by
Sediment redox chemistry in estuarine sediment
311
Fl F2 F3 F4 Soluble
0
10
20
430
0
-80
30 -130
40 -150
Fig. 6. The effect of Eh on the distribution
50
T(day)
60
Eh(mV)
-170
of As in the various chemical fractions.
. . 0 ...“.
F2
I.....
.1.1
0
.
-_-
A
I-
F4
m-m
a
-m_
Soluble
.
.
.
F3
0
Way) 430
0
-80
-130
-150
Fig. 7. The effect of Eh on the percentage
McGeehan and Naylor (1994). Parallel increases in As bound to insoluble large molecular humic compounds increased with the reduction of manganic and iron oxide. The rate constant of pseudo-zero-order reaction for the formation of As with insoluble large molecular humic substances was 0.97 mg kg’ dS’. There was no evidence to show that As associated with sulfides was formed.
-170
EWW
of As in the various chemical fractions.
At Eh levels from 430 mV to - 130 mV, As bound to carbonates decreased (K= -0.51 mg kg’ d-l). Further decreases in Eh (< -130 mV) caused the As bound to carbonates to increase (K= 0.59 mg kg-’ d“). Figure 7 shows the effect of Eh on the percentage of As in the various chemical fractions. The dominant active As fraction is As bound to Fe(II1) and Mn(IV) oxides.
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0.65
0.6 Fl F2 F3 F4 Soluble
0 430
lo 0
20 -80
30 -130
40
50
-150
Fig. 8. The effect of Eh on the distribution
T(day)
60 -170
Eh(mV)
of Cr in the various chemical fractions.
,
600 ,
4
- 2.5
...... ..0 .I.....
F2
111_0 .--.
F3
____fi ____
F4
___ a ___
Soluble
I 0
10
20
430
0
-80
30 -130
40 -150
Fig. 9. The effect of Eh on the percentage
Cr behavior
Cr fractions were also affected by sediment redox conditions (Figs. 8 and 9). As Eh decreased to 0 mV, Fe(II1) and Mn(IV) oxides in the sediment were reduced to more soluble Fe(I1) and Mn(I1) and the Cr adsorbed on Fe(II1) and Mn(IV) oxides was apparently released increasing dissolved Cr concentration. At Eh < 100 mV, soluble Cr(V1) was apparently reduced to
50
60 -170
2 T(day) Eh(mV)
of Cr in the various chemical fractions.
insoluble G-(111) (mostly as Cr(OH&) (Masscheleyn et al. 1992). Cr(II1) also apparently reacted with organic matter to form dissolved Cr(IV)-organic complexes (Masscheleyn et al. 1992). The above three reactions controlled dissolved Cr concentration in the sediment. The combination of these three reactions resulted in little difference in dissolved Cr content being measured under the various redox conditions studied.
Sediment redox chemistry in estuarine sediment
313
5
12.5
?? ? ? ?? ? ?
m, \\
10 G +zJ fi z s 8 U 3
4
7.5
3
2
5
S +!J
+
Fl
: Y z
. .... 0 .......
F2
z
----
0
F3
me--
1
2.5
h
----
F4
-me-
_-- ? ? ---
Soluble
0
0 0
10
20
430
0
-80
30
40 -130
50 -150
Fig. 10. The effect of Eh on the distribution
0
10
430
0
20 -80
30 -130
40 -150
Fig, 11. The effect of Eh on the percentage
As sediment Eh decreased, Cr associated with Fe(II1) and Mn(IV) oxides decreased (K= -0.32 mg kg’ d“), while Cr associated with insoluble large molecular humic increased (K= 0.81 mg kg’ d-l). There was no evidence to show that any Cr associated with sulfides was formed. Figure 9 shows the effect of Eh on the percentage of Cr in the various chemical fractions. Under oxidizing conditions, Cr associated primarily with Fe(II1) and
60 -170
T(day) Eh(mV)
of Cd in the various chemical fractions,
50
60 -170
.... ....0 . .......
F2
----
0 ----
F3
---- A ----
F4
_-- a ---
Soluble
T(day) Eh(mV)
of Cd in the various chemical fractions.
Mn(IV) oxides, with Cr activity being controlled by chemical adsorption of Cr on Fe(II1) and Mn(IV) oxides. Under reducing soil conditions, Cr was bound to insoluble large molecular humic substances. Cd behavior Figure 10 shows the effect of Eh on the levels of water soluble Cd and Cd in the various chemical fractions. As sediment Eh decreased, dissolved Cd decreased
314
T. Guo et al.
_c)_
Fl
...0 ..
F2
__S. 0 __I.
F3
.
____& ____ F4 ___ a ___
0
10
430
0
20 -80
30 -130
40
50
-150
Fig. 12. The effect of Eh on the distribution
60
Soluble
Day)
-170
Eh(mV)
of Zn in the various chemical fractions,
500 Fl
400
. . 0
- 100
..I.
____
300
..I....
o-
sm_
F2 F3
-__h ____F4 200
-
50
-
0
s-v
?? e-v
Soluble
100
0 0 430
10 0
20 -80
30 -130
40 -150
Fig. 13. The effect of Eh on the percentage
from 4.6 mg kg’ to 0.3 mg kg’ (K= -0.09 mg kg’ d-‘, Fig. 14), and Cd associated with Fe(II1) and Mn(IV) oxides also decreased (K= -0.01 mg kg’ d-l). In contrast, Cd associated with carbonates (K= 0.01 mg kg-’ d-l), and Cd associated with insoluble large molecular humic substances and sulfides increased as Eh decreased. The rate constants of the pseudo-zero-order reaction on the formation of Cd associated with insoluble sulfides and large molecular humic were 0.16 mg kg’ d-i
50
60 -170
T(day) Eh(mV)
of Zn in the various chemical fractions.
and 0.01 mg kg’ d -I, respectively. These results are similar to Kerner and Wallman’s (1992) finding for Cd associated with dissolved and sulfide forms. Figure 11 shows the effect of Eh on the percentage of Cd in the water soluble and chemical fractions. Under an oxidizing sediment condition, Cd was associated with Fe(II1) and Mn(IV) oxides, carbonates, and soluble phase Cd. Soluble Cd accounted for 230 g kg-* of the total concentration under oxidized conditions.
Sediment redox chemistry in estuarine sediment
315
Zn Cd
0
10
430
0
Fig. 14.
20 -80
30
40
-130
50 -150
60 -170
T(day)
Eh(mV)
The effect of Eh on dissolved concentrationof Cd and Zn.
Under reducing sediment conditions, Cd bound to the carbonates fraction accounted for most of the Cd, and water soluble Cd accounted for only 15 g kg-’ of total Cd concentration under reducing conditions. As Eh decreased, the decreased Cd associated with Fe(II1) and Mn(IV) oxides and soluble phases was transformed into Cd associated with insoluble carbonates and sulfides. Zn behavior
Figure 12 shows the effect of sediment Eh on the distribution of Zn in the water soluble and chemical fractions. As Eh decreased, dissolved Zn decreased from 100 mg kg’ to 0.8 mg kg-’ (K= - 1.78 mg kg’ d-‘, Fig. 14), Zn associated with Fe(II1) and Mn(IV) oxides also decreased (K= -6.5 mg kg’ d-l). In contrast, Zn associated with carbonates increased (K= 3.3 mg kg’ d-‘) and Zn associated with insoluble large molecular humic and sulfides increased as sediment Eh decreased. The rate constants of the pseudo-zero-order reaction for the formation of Zn associated with insoluble sulfides and large molecular humic were 5.4 mg kg’ d-’ and 2.9 mg kg’ d-‘, respectively. These results are similar to those reported by Kerner and Wallman (1992) who determined Zn existed primarily in dissolved and sulfide forms.
Figure 13 shows the effect of sediment Eh on the percentage distribution of Zn in the various fractions. Under oxidizing conditions, Zn was associated with Fe(II1) and Mn(IV) oxides and soluble phases. Soluble Zn accounted for 140 g kg’ of the total content of all fractions. Under reducing conditions, Zn was found to be associated with insoluble sulfide, large molecular humic compounds and carbonates. Soluble Zn accounted for 1 g kg-’ of the total content found in the fractions. As sediment Eh decreased, the Zn associated with Fe(II1) and Mn(IV) oxides and soluble phases was transformed into Zn fractions associated with insoluble carbonates, sulfide, and large molecular humic compounds. The data collected show that Zn becomes less mobile under reducing sediment conditions. These studies demonstrate that sediment redox potential is an important parameter affecting heavy element solubility and mobility in coastal environments. Reduction or aeration status of sediment should be considered in predicting the impact of these elements to the aquatic environment.
Acknowledgment-The research was supported by Louisiana State University Coastal Marine Institute funded by the Minerals Management Service (Contract # 14-35-0001-30660, T.U. 19907).
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REFERENCES Boesch, D.F.; Rabalais, N.N., ed. Produced waters in sensitive coastal habitats: An analysis of impacts, central coastal Gulf of Mexico. OCS ReportiMMS89-00331. New Orleans, LA: U.S. Dept. of Interior, Minerals Management Service, Gulf of Mexico OCS Regional Office; 1989: 157. DeLaune, R.D.; Smith, C.J. Release of nutrients and metals following oxidation of freshwater and saline sediment. J. Environ. Qua]. 14: 164-168; 1985. Feijtel, T.C.; DeLaune, R.D.; Patrick, W.H., Jr. Biogeochemical control on metal distribution and accumulation in Louisiana sediments. J. Environ. Qual. 17: 88-94; 1988. Folsom, B.L., Jr.; Skogerboe, J.G.; Palermo, M.R.; Simmers, J.W.; Pranger, S.A.; Shafer, R.A. Synthesis of the results of the field verification program upland disposal alternative. Tech. Rep. D88-7. Vicksburg, MS: U.S. Army Corps of Eng. Waterways Exp. Stn.; 1988. Gambrel], R.P.; Patrick, W.H., Jr. Chemical and microbiological properties of anaerobic soils and sediments. In: Hook, D.D.; Crawford, R.M.M., eds. Plant life in anaerobic environments. Ann Arbor, MI: Ann Arbor Science; 1978: 375-423. Gambrel], R.P.; Khalid, R.A.; Patrick, W.H., Jr. Chemical availability of mercury, lead and zinc in Mobile Bay sediment suspension as affected by pH and oxidation-reduction conditions. Environ. Sci. Technol. 14: 431-436; 1980. Gambrell, R.P.; Patrick, W.H., Jr. Cu, Zn, and Cd availability in a sludge-amended soil under controlled pH and redox potential conditions. In: Bar-Yosef, B. et al., eds. Inorganic contaminants in the vadose zone. Berlin: Springer-Verlag; 1989: 89-106. Gambrel], R.P.; DeLaune, R.D.; Patrick, W.H., Jr. Redox processes in soils following oxygen depletion. In: Jackson, M.B. et al., eds. Plant life under oxygen deprivation. The Hague, The Netherlands: SPB Academic Publishing; 1991 a: 101 - 117. Gambrel], R.P.; Wiesepape, J.B., Patrick, W.H., Jr.; Duff, M.C. The effects of pH, redox, and salinity on metal release from a contaminated sediment. Water Air Soil Pollut. 57-58: 359-367; 1991b. Gambrell, R.D. Trace and toxic metals in wetlands - A review. J. Environ. Qual. 23: 883-891; 1994. Giblin, A.E.; Luther, G.W. III; Valiela, I. Trace metal solubility in salt marsh sediments contaminated with sewage sludge. Estuar. Coast. Shelf. Sci. 23: 477-498; 1986. Giesy, J.P.; Leversee, G.L.; William, D.R. Effects of naturally occurring aquatic organic fractions on cadmium toxicity to simocephalus serrulatus (daphnidae) and Gambusia afjnis (Polciludae). Water Res. 11: 1013-1020; 1977. Jacobson, M.E. Chemical and biological mobilization of Fe(III) in marsh sediments. Biogeochemistry 25: 41-60; 1994.
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Keller, C.; Vedy, J.C. Distribution of copper and cadmium fractions in two forest soils. J. Environ. Qual. 23: 987-999; 1994. Kemer, M.; Wallman, K. Remobilization events involving Cd and Zn from intertidal flat sediments in the Elbe Estuary during the tidal cycle. Estuar. Coast. Shelf Sci. 35: 371-393; 1992. Khalid, R.A.; Gambrel], R.P.; Patrick, W.H., Jr. Chemical availability of cadmium in Mississippi River sediment. J. Environ. Qua]. 10: 523-529; 1981. Koons, C.B.; McAuliffe, C.D.; Weiss, F.T. Effects of produced waters on the marine environment. Oceans 2: 45-79; 1977. Levy, D.B.; Barbarick, K.A.; Siemer, E.G.; Sommers, L.E. Distribution and partitioning of trace metals in contaminated soils near Leadville, Colorado. J. Environ. Qua]. 21: 185-195; 1992. Lovley, D.R.; Phillips, E.J.P. Competitive mechanisms for inhibition of sulfate reduction and methane production in the zone of ferric iron reduction in sediment. Appl. Environ. Microbial. 53: 2636-2641; 1987. Lyssj, I. Oil and gas extraction operations. Report to EPA Municipal Environmental Research Laboratory, Cincinnati, OH; 1981. Masscheleyn, P.H.; Pardue, J.H.; DeLaune, R.D.; Patrick, W.H., Jr. Chromium redox chemistry in a Lower Mississippi Valley bottomland hardwood wetland. Environ. Sci. Technol. 26(6): 1217-1226; 1992. McGeehan, S.L.; Naylor, D.V. Sorption and redox transformation of arsenite and arsenate in two flooded soils. Soil Sci. Sot. Am. J. 58: 337-342; 1994. Palermo, M.R.; Shafer, R.A.; Brannon, J.M.; Myers, T.E.; Truitt, C.L.; Zappi, M.E. Evaluation of dredged material disposal alternatives for US Navy homeport at Everett, Washington. Tech. Rep. EL-89-l. Vicksburg, MS: U.S. Army Corps of Eng. Waterways Exp. Stn.; 1989. Patrick, W.H., Jr.; Williams, B.G.; Moraghan, J.T. A simple system for controlling redox potential and pH in soil suspension. Soil Sci. Am. J. 32: 331-332; 1973. Patrick, W.H., Jr.; DeLaune, R.D. Chemical and biological redox systems affecting nutrient availability in coastal wetlands. Geosci. Man 18: 131-137; 1977. Ramos, L.; Hemandez, L.M.; Gonzalez, M.J. Sequential fraction of copper, lead, cadmium and zinc in soils from or near Donana National Park. J. Environ. Qua]. 23: 50-57; 1994. Shannon, R.D.; White, J.R. The selectivity of a sequential extraction procedure for the determination of iron oxyhydroxides and iron sulfides in lake sediments. Biogeochemistry 14: 193-208; 1991. Tillery, J.B.; Thomas, R.E.; Windom, H.L. Trace metal studies in sediment and fauna. In: Ecological investigation of petroleum production platforms in the Central Gulf of Mexico. Vol. 1, Part 4. Prepared by Southwest Research Institute, San Antonio, TX, for the Bureau of Land Management, New Orleans OCS; 198 1.
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THE INFLUENCE OF SEDIMENT REDOX CHEMISTRY ON CHEMICALLY ACTIVE FORMS OF ARSENIC, CADMIUM, CHROMIUM, AND ZINC IN ESTUARINE SEDIMENT Tingzong Guo, R.D. DeLaune*, and W.H. Patrick, Jr. Wetland Biogeochemistry
Institute, Louisiana State University, Baton Rouge, LA 70803, USA
EI 9608-I 76 M (Received 14 August 1996; accepted IO February 1997)
Kinetics and chemical fractionation procedures were used to quantify the effect of the sediment redox (Eh) condition on the behavior of As, Cd, Cr, and Zn in the bottom sediment collected from a Louisiana coastal site receiving produced water discharge. Sediment samples were incubated in microcosms in which Eh-pH conditions were controlled. Sediment was sequentially extracted to determine metals in various chemical fractions (water soluble, exchangeable, bound to carbonates, bound to iron and manganese oxides, bound to insoluble organic and sulfides) and the chemical inactive fraction (mineral residue). Under oxidizing conditions, As, Zn, and Cr behavior was governed by redox chemistry of Fe(II1) and Mn(IV) oxides. Cd transformations were controlled by both Fe(II1) and Mn(IV) oxides and carbonates. Under a reducing condition, the behaviors of Zn and Cr were controlled primarily by insoluble large molecular humic material and sulfides; the behavior of Cd was controlled by carbonates. When sediment redox potential increased, the affinity between Fe(II1) and Mn(IV) oxides and As, Cd, Cr, and Zn increased. When sediment redox potential decreased, the affinity between carbonates and Cd and Zn increased; the affinity between insoluble sulfides, large molecular humic matter and As, Cd, Cr, and Zn increased; the soluble Cd and Zn decreased; the soluble As and Cr remained constant. Results suggest reducing sediment conditions would reduce Cd and Zn toxicity. Under reducing or anaerobic conditions, the solibilization rate constants (mg kg-’d-‘) for As, Cr, Cd, and Zn bound to Fe(II1) and Mn(IV) oxides were -0.88, -0.32, -0.01, and -6.5, respectively; the rate constants (mg kg’ de’) for dissolved Cd and Zn were -0.09 and -1.78, respectively. 81997Elsevier Science Ltd
INTRODUCTION Koons et al. 1977; Lyssj 1981). As the produced water enters wetland environments, heavy elements can enter the sediment column. Analysis of total element concentration in sediment can quantify the degree of trace element enrichment. Total element content does not provide information on transformation and mobilization of trace elements. Speciation studies can: 1) provide an insight into element distribution patterns; 2) identify metal bioavailability and toxicity in ecosystems; and, 3) explain
Petroleum production and recovery activity in the Louisiana coastal zone can discharge a considerable amount of produced waters (Boesch and Rabalais 1989). The produced waters can contain elevated concentrations of heavy metals compared to the receiving water. Element content in produced water can include barium, cadmium, chromium, iron, mercury, manganese, strontium, and thallium (Tillery et al. 1981; *Corresponding author.
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transformation and mobility of metal species. Trace element species distribution has been studied in various ways (DeLaune and Smith 1985; Gambrel1 et al. 199 1a and b; Giblin et al. 1986; Giesy et al. 1977; Keller and Vedy 1994). These techniques have included multistep extractions in which a chemical solution removes various element forms. Such fractionation schemes can provide information on the general behavior of heavy elements in sediment and provide an estimate of their potential mobility (Keller and Vedy 1994). Metals in sediment are generally considered to be present in the following forms: water soluble, exchangeable, carbonate bound, ferric and manganic oxide bound, organic matter and sulfide bound, silicate bound, and residual. The oxidation/reduction state (redox potential) of sediment is an important parameter affecting heavy metal transformation. The redox conditions (Eh) of wetland soil and sediments vary widely from approximately +500 mV (surface soils) to approximately -300 mV (strongly reducing soils) (Patrick and DeLaune 1977). Sediment redox levels can greatly affect toxic metal uptake by plants (Gambrel1 and Patrick 1989; Giblin et al. 1986), leaching losses of toxic metals by runoff or ground water (Folsom et al. 1988; Palermo et al. 1989), but there is little information on redox chemistry of toxic metals in different geochemistry forms in sediment. Heavy element species kinetics in wetland soil kinetics are influenced by many factors. These factors include temperature, organic matter, surface activity of Fe and Mn compounds, microorganism activity and other sediment characteristics. Since it is difficult to study these factors individually, the factors were combined in this study into one parameter-the rate constant of the assumed zero-order reaction, the socalled pseudo-zero-order reaction model. The pseudozero-order reaction model was used in this study. When the rate constant is positive, the metal content increases in the specific fraction. When the rate constant is negative, the metals are removed from the specific fraction. When considering trace elements bound to the organic matter and sulfide phase, it was assumed that the formation of insoluble organic matter (complexation of insoluble, large molecular weight humic fraction) and sulfides was independent (Gambrel1 and Patrick 1978; Gambrel1 et al. 1980). At sediment Eh = -130 mV or Eh > -130 mV, the changes in heavy elements in this fraction were due to insoluble, large molecular weight humic substances.
T. Guo et al.
Surface of clays, organic matter, and iron oxides in sediment will absorb or desorb heavy elements when the ionic composition or Eh-pH changes (Keller and Vedy 1994; Khalid et al. 1981; Gambrel1 et al. 1980). Significant heavy element content is also associated with sediment carbonates (Ramos et al. 1994; Gambrel1 1994). This fraction would be susceptible to pH change. Iron and manganese oxides existing as nodules and concretions, cemented between particles or on particle coatings in sediment are excellent scavengers for heavy elements and are affected by sediment Eh and pH change (Feijtel et al. 1988; Levy et al. 1992). Heavy elements are also bound to various insoluble organic forms such as living organisms, detritus, and humic material (Gambrel1 et al. 1980; Ramos et al. 1994). Sediment redox conditions can affect the degradation and solubility of such organic material and then influence the release of heavy elements. Heavy elements can also exist as sulfides under anaerobic conditions (Gambrel1 et al. 1980; 199 1b) which are susceptible to Eh and pH changes. Heavy elements found in primary and secondary minerals are relatively stable in a natural sediment environment (Gambrel1 1994). This research examines the kinetics and transformations of As, Cd, Cr, and Zn in estuarine sediment at a site in coastal Louisiana receiving produced water discharge. The effects of the redox potential (Eh) on the kinetics of transformation of different forms of the heavy elements in sediments are detailed. MATERIAL AND METHODS Sediment
Sediment samples were collected from a produced water discharge located near a brackish marsh (Fig. 1). The sediment pH was 7.0. The sediment contained 20 g kg-’ Fe, 0.4 g kg-’ Mn and 1 g kg-’ Ba. The metal content of the sediment was analyzed and four times the original metal content of Cd, Zn, Cr, and As was added to the sediment. The total metal content (mg g-’ dry sediment) of the sediment after spiking was Cd=2 1, Zn=SOO, Cr=2 10, and As=3 70. Incubation
of the soil
Wet sediment equivalent to 200 g of the dry sediment was placed in each of 5 microcosms. Sufficient water (5 g L-’ salinity) was added to the microcosm to obtain a sediment to water ratio of 1: 10. The microcosm originally described by Patrick et al. (1973) was fitted with two bright Pt electrodes, a calomel reference elec-
301
Sediment redox chemistry in estuarine sediment
Fig. 1. Location of the Lirette (LRT) site from which sediment samples were collected.
trode connected to the suspension by a saturated potassium chloride-agar salt bridge, a serum cap, a thermometer, an inlet tube for nitrogen or oxygen, and an outlet which was submerged in water to prevent atmospheric oxygen from diffusing into the microcosm. The soil suspensions were maintained at 26°C. Sediment pH was adjusted and maintained at 7.0 with either hydrochloric acid or sodium hydroxide. The bright Pt electrodes immersed in the suspension were connected to a potentiometer for redox potential measurements. Sediment was incubated for 60 d. Element contents in the sediment suspension were selectively extracted as described by Shannon and White (1991):
Sequentialextractionprocedures Watersolublephase. Samples from the sediment suspensions were taken at selected intervals. The suspension samples were centrifuged and the supernatant filtered through a 0.45 pm membrane filter. This supernatant was assumed to be water soluble. The remaining sediment was extracted sequentially as described below. The sediment sample was kept under nitrogen or oxygen free atmosphere during extraction. Following removal of the water soluble phase, the sediment was sequentially extracted into fine fractions (Fl to F5) described below.
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2000
1500
+
Fl
G
........0 ........
F2
? y
----o----
F3
---B---
Soluble
e
1000
5
5OO
0 0
10
20
430
0
-80
30 -130
40 -150
Fig. 2. The effect of Eh on the distribution
1) Fl-Exchangeable phase: The solid phase from the water soluble fraction was extracted at room temperature for 30 min with 8 mL 0.5 M Mg(NO,)Jg dry weight sediment, adjusted to pH 7.0 with nitric acid. The samples were agitated continuously. 2) F%-Bound to carbonate phase: The sediment residue from Fl was leached at room temperature for 5 h with 8 mL, 1M NaOAc, adjusted to pH 5.0 with acetic acid for 1 g dry weight sediment. These samples were also agitated continuously. 3) F3-Bound to iron and manganese oxides phase: The sediment residue from F2 was extracted at 96°C for 6 h with 20 mL 0.08 M NH,OH HCl in 250 mL/L acetic acid for 1 g dry weight sediment. These samples were occasionally agitated. 4) F4-Bound to organic matter and sulfides phase: For 1 g of dry weight sediment, the sediment residue from F3 was extracted at 85°C for 2 h with 3 mL 0.02 M HNO, and 5 mL 300 mL/L H202 (adjusted to pH = 2.0 with IWO,) was added, and extraction continued at 85°C for another 3 h. The sample was then cooled, 5 mL 3.2 M NI&OAc in 20 mL/L HNO, was added, and the sample was diluted to 20 mL with deionized water. The samples were agitated continuously for 30 min. NHOAc was added to prevent adsorption of extracted metals onto the oxidized sediment. 5) FS-Mineral matrix phase: The sediment residue from F4 was extracted with 25 mL concentrated HNO, for 1 g of dry weight sedi-
50
60 -170
T(day) Eh(mV)
or content of Fe in chemical fractions.
ment at 105 “C, the sediment was digested until 5 mL solution was left, and the sample was diluted to 25 mL with deionized water. The above sequential extractions were conducted in 250 mL centrifuge tubes which prevented any loss of sediment between the successive extractions. Separation was conducted by centrifuging at 5000 rpm for 30 min. Supernatants were filtered using 0.45 urn millipore filters and then analyzed for metals. The residues were rinsed with 8 mL deionized water for 1 g dry weight sediment and centrifuged at 5000 rpm for 30 min. These second supernatants were discarded. Metal concentrations in water soluble and the chemical extracts were determined by ICP. Quality assurance was conducted by spiking extracts with certified element standards. RESULTS AND DISCUSSION
Fe and Mn behavior The effects of Eh (redox potential) on the contents of Mn and Fe in the various chemical fractions are shown in Figs. 2 and 3, and 4 and 5, respectively. As sediment Eh decreased from +430 mV to 0 mV, Mn associated with carbonates decreased (K= -6.7 mg kg’ d-l). Also, Mn(IV) oxides were apparently reduced to soluble Mn(I1) under more reducing sediment conditions (the pseudo-zero-order constant of Mn(IV) oxides reduction being - 1.3 mg kg’ a’). As the result of the above two reactions dissolved, Mn(I1) concentration increased
Sediment redox chemistry in estuarine sediment
309
e _.I.
Soluble
0 .I....
Fl
-II_ 0 _I__
F2
__- & _--
F3 F4
0
lo
20
430
0
-80
30
40 -130
50
-150
60
T(day)
Eh(mV)
-170
Fig. 3. The effect of Eh on the percentage of Fe in the various chemical fractions.
0
10
20
430
0
-80
30 -130
40 -150
Fig. 4. The effect of Eh on the distribution
significantly (K=l2.7 mg kg’ d’). In the Eh range between 0 mV and -130 mV, dissolved Mn(I1) decreased as the result of Mn becoming associated with carbonates. At sediment Eh below -130 mV, sulfides are produced. At such low Ehs, dissolved Mn(I1) concentration was lowered through Mn becoming associated with carbonates and sulfides and insoluble organic matter (complexes of the Fe and Mn with insoluble large
50
60 -170
__(I_
Fl
I._.
F2
0 ...
I___ 0 _I__
F3
-__ A _I_
F4
___ w ___
Soluble
T(W) Eh(mV)
of Mn in the various chemical fractions.
molecular humic material). The rate constant for the removal of the dissolved Mn(I1) at this Eh range was 0.94 mg kg-’d-‘. The rate constants for the formation of insoluble carbonates, sulfides, and Mn with humic substances were 0.49, 1.8, and 0.07 mg kg’ d-‘, respectively. Under oxidizing sediment conditions, Mn(IV) oxides and carbonates were the predominate form of Mn pre-
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Fl F2 F3 F4 w-v
a
m-w
Soluble
000 0
10
20
430
0
-80
30 -130
40 -150
Fig. 5. The effect of Eh on the percentage
sent. Under a reducing condition, Mn(I1) associated with exchangeable fractions and sulfides were the primary Mn forms present. Figure 2 shows the effect of sediment Eh (redox potential) on the content of Fe in the various chemical fractions. As Eh decreased, Fe(II1) oxides were microbially reduced to soluble Fe(II), therefore increasing soluble Fe concentration (K=42.5 mg kg-’ d-‘) in soluion. When Eh further decreased to - 130 mV, a reduction in iron oxides was observed which was attributed to FeS precipitation. Based on the fractionation data at all redox levels studied, dissolved Fe(I1) was also removed through Fe becoming associated with insoluble organic matter. The above two factors resulted in decreased dissolved Fe at Eh < - 130 mV (K= -25.8 mg kg’ d-l). The rate constants of the removal reactions for dissolved Fe(I1) are 40.7 mg kg’ d’ and 3 1.4 mg kg’ d-l for the formation of sulfides and complexation of insoluble, large molecular humic, respectively. Under a reducing condition (Eh > - 130 mV), the reduction of Fe(II1) oxides likely occurred by direct microbial reduction (K= -42.4 mg kg’ d-l) involving organic carbon turnover. This part of Fe(II1) reduction was equivalent to approximately 207 1 mg kg’ (20% of total reducible Fe(II1)). At sediment (Eh < -130 mV), Fe(II1) reducing microorganisms can inhibit sulfate reduction by out competing sulfate reducer for electron donors (Lovley and Phillips 1987). A portion of Fe(II1) reduction was assumed to have occurred by sulfides
50
60 -170
Way) Et-W4
of Mn in the various chemical fractions.
reducing Fe(II1) oxides. Only a small amount of the Fe(II1) reduction at these redox levels can occur by direct bacteria reduction involving organic carbon turnover (Jacobson 1994). Accordingly, in this study, 8571 mg kg-’ (80% of total reducible Fe(II1)) was apparently reduced by the sulfide oxidation (K= - 17 1.5 mg kg’ d-l). The rate constant of indirect Fe(II1) oxide reduction by sulfides is significantly greater than that by direct bacteria reduction. During anaerobic incubation, approximately half of the reduced Fe(II1) was converted to carbonate bound Fe(I1). Under an oxidizing sediment condition, Fe(II1) oxide predominated with dissolved Fe concentrations controlled by Fe(II1) oxides. Under reducing sediment conditions, sulfide and insoluble large molecular humic bound Fe(I1) was the dominant fraction controlling iron behavior.
As behavior
Figure 6 shows the effect of Eh on the level of As in the water soluble chemical fraction. When Eh decreased to 0 mV, As(V) was reduced to As(II1) (K= -0.21 mg kg’ d-l). At sediment Eh=O to -100 mV, dissolved arsenic concentration was essentially zero. This may be due to the fact that the fresh As(II1) which was formed from As(V) reduction became insoluble. Following manganic oxide and ferric oxide reduction, the As bound with these oxides decreased (K= -0.88 mg kg’ a’). Similar results were reported by
Sediment redox chemistry in estuarine sediment
311
Fl F2 F3 F4 Soluble
0
10
20
430
0
-80
30 -130
40 -150
Fig. 6. The effect of Eh on the distribution
50
T(day)
60
Eh(mV)
-170
of As in the various chemical fractions.
. . 0 ...“.
F2
I.....
.1.1
0
.
-_-
A
I-
F4
m-m
a
-m_
Soluble
.
.
.
F3
0
Way) 430
0
-80
-130
-150
Fig. 7. The effect of Eh on the percentage
McGeehan and Naylor (1994). Parallel increases in As bound to insoluble large molecular humic compounds increased with the reduction of manganic and iron oxide. The rate constant of pseudo-zero-order reaction for the formation of As with insoluble large molecular humic substances was 0.97 mg kg’ dS’. There was no evidence to show that As associated with sulfides was formed.
-170
EWW
of As in the various chemical fractions.
At Eh levels from 430 mV to - 130 mV, As bound to carbonates decreased (K= -0.51 mg kg’ d-l). Further decreases in Eh (< -130 mV) caused the As bound to carbonates to increase (K= 0.59 mg kg-’ d“). Figure 7 shows the effect of Eh on the percentage of As in the various chemical fractions. The dominant active As fraction is As bound to Fe(II1) and Mn(IV) oxides.
312
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0.65
0.6 Fl F2 F3 F4 Soluble
0 430
lo 0
20 -80
30 -130
40
50
-150
Fig. 8. The effect of Eh on the distribution
T(day)
60 -170
Eh(mV)
of Cr in the various chemical fractions.
,
600 ,
4
- 2.5
...... ..0 .I.....
F2
111_0 .--.
F3
____fi ____
F4
___ a ___
Soluble
I 0
10
20
430
0
-80
30 -130
40 -150
Fig. 9. The effect of Eh on the percentage
Cr behavior
Cr fractions were also affected by sediment redox conditions (Figs. 8 and 9). As Eh decreased to 0 mV, Fe(II1) and Mn(IV) oxides in the sediment were reduced to more soluble Fe(I1) and Mn(I1) and the Cr adsorbed on Fe(II1) and Mn(IV) oxides was apparently released increasing dissolved Cr concentration. At Eh < 100 mV, soluble Cr(V1) was apparently reduced to
50
60 -170
2 T(day) Eh(mV)
of Cr in the various chemical fractions.
insoluble G-(111) (mostly as Cr(OH&) (Masscheleyn et al. 1992). Cr(II1) also apparently reacted with organic matter to form dissolved Cr(IV)-organic complexes (Masscheleyn et al. 1992). The above three reactions controlled dissolved Cr concentration in the sediment. The combination of these three reactions resulted in little difference in dissolved Cr content being measured under the various redox conditions studied.
Sediment redox chemistry in estuarine sediment
313
5
12.5
?? ? ? ?? ? ?
m, \\
10 G +zJ fi z s 8 U 3
4
7.5
3
2
5
S +!J
+
Fl
: Y z
. .... 0 .......
F2
z
----
0
F3
me--
1
2.5
h
----
F4
-me-
_-- ? ? ---
Soluble
0
0 0
10
20
430
0
-80
30
40 -130
50 -150
Fig. 10. The effect of Eh on the distribution
0
10
430
0
20 -80
30 -130
40 -150
Fig, 11. The effect of Eh on the percentage
As sediment Eh decreased, Cr associated with Fe(II1) and Mn(IV) oxides decreased (K= -0.32 mg kg’ d“), while Cr associated with insoluble large molecular humic increased (K= 0.81 mg kg’ d-l). There was no evidence to show that any Cr associated with sulfides was formed. Figure 9 shows the effect of Eh on the percentage of Cr in the various chemical fractions. Under oxidizing conditions, Cr associated primarily with Fe(II1) and
60 -170
T(day) Eh(mV)
of Cd in the various chemical fractions,
50
60 -170
.... ....0 . .......
F2
----
0 ----
F3
---- A ----
F4
_-- a ---
Soluble
T(day) Eh(mV)
of Cd in the various chemical fractions.
Mn(IV) oxides, with Cr activity being controlled by chemical adsorption of Cr on Fe(II1) and Mn(IV) oxides. Under reducing soil conditions, Cr was bound to insoluble large molecular humic substances. Cd behavior Figure 10 shows the effect of Eh on the levels of water soluble Cd and Cd in the various chemical fractions. As sediment Eh decreased, dissolved Cd decreased
314
T. Guo et al.
_c)_
Fl
...0 ..
F2
__S. 0 __I.
F3
.
____& ____ F4 ___ a ___
0
10
430
0
20 -80
30 -130
40
50
-150
Fig. 12. The effect of Eh on the distribution
60
Soluble
Day)
-170
Eh(mV)
of Zn in the various chemical fractions,
500 Fl
400
. . 0
- 100
..I.
____
300
..I....
o-
sm_
F2 F3
-__h ____F4 200
-
50
-
0
s-v
?? e-v
Soluble
100
0 0 430
10 0
20 -80
30 -130
40 -150
Fig. 13. The effect of Eh on the percentage
from 4.6 mg kg’ to 0.3 mg kg’ (K= -0.09 mg kg’ d-‘, Fig. 14), and Cd associated with Fe(II1) and Mn(IV) oxides also decreased (K= -0.01 mg kg’ d-l). In contrast, Cd associated with carbonates (K= 0.01 mg kg-’ d-l), and Cd associated with insoluble large molecular humic substances and sulfides increased as Eh decreased. The rate constants of the pseudo-zero-order reaction on the formation of Cd associated with insoluble sulfides and large molecular humic were 0.16 mg kg’ d-i
50
60 -170
T(day) Eh(mV)
of Zn in the various chemical fractions.
and 0.01 mg kg’ d -I, respectively. These results are similar to Kerner and Wallman’s (1992) finding for Cd associated with dissolved and sulfide forms. Figure 11 shows the effect of Eh on the percentage of Cd in the water soluble and chemical fractions. Under an oxidizing sediment condition, Cd was associated with Fe(II1) and Mn(IV) oxides, carbonates, and soluble phase Cd. Soluble Cd accounted for 230 g kg-* of the total concentration under oxidized conditions.
Sediment redox chemistry in estuarine sediment
315
Zn Cd
0
10
430
0
Fig. 14.
20 -80
30
40
-130
50 -150
60 -170
T(day)
Eh(mV)
The effect of Eh on dissolved concentrationof Cd and Zn.
Under reducing sediment conditions, Cd bound to the carbonates fraction accounted for most of the Cd, and water soluble Cd accounted for only 15 g kg-’ of total Cd concentration under reducing conditions. As Eh decreased, the decreased Cd associated with Fe(II1) and Mn(IV) oxides and soluble phases was transformed into Cd associated with insoluble carbonates and sulfides. Zn behavior
Figure 12 shows the effect of sediment Eh on the distribution of Zn in the water soluble and chemical fractions. As Eh decreased, dissolved Zn decreased from 100 mg kg’ to 0.8 mg kg-’ (K= - 1.78 mg kg’ d-‘, Fig. 14), Zn associated with Fe(II1) and Mn(IV) oxides also decreased (K= -6.5 mg kg’ d-l). In contrast, Zn associated with carbonates increased (K= 3.3 mg kg’ d-‘) and Zn associated with insoluble large molecular humic and sulfides increased as sediment Eh decreased. The rate constants of the pseudo-zero-order reaction for the formation of Zn associated with insoluble sulfides and large molecular humic were 5.4 mg kg’ d-’ and 2.9 mg kg’ d-‘, respectively. These results are similar to those reported by Kerner and Wallman (1992) who determined Zn existed primarily in dissolved and sulfide forms.
Figure 13 shows the effect of sediment Eh on the percentage distribution of Zn in the various fractions. Under oxidizing conditions, Zn was associated with Fe(II1) and Mn(IV) oxides and soluble phases. Soluble Zn accounted for 140 g kg’ of the total content of all fractions. Under reducing conditions, Zn was found to be associated with insoluble sulfide, large molecular humic compounds and carbonates. Soluble Zn accounted for 1 g kg-’ of the total content found in the fractions. As sediment Eh decreased, the Zn associated with Fe(II1) and Mn(IV) oxides and soluble phases was transformed into Zn fractions associated with insoluble carbonates, sulfide, and large molecular humic compounds. The data collected show that Zn becomes less mobile under reducing sediment conditions. These studies demonstrate that sediment redox potential is an important parameter affecting heavy element solubility and mobility in coastal environments. Reduction or aeration status of sediment should be considered in predicting the impact of these elements to the aquatic environment.
Acknowledgment-The research was supported by Louisiana State University Coastal Marine Institute funded by the Minerals Management Service (Contract # 14-35-0001-30660, T.U. 19907).
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