- Email: [email protected]
stabilized cement waste forms
Waste Management, Vol. 17, No. 1, pp. 2~32, 1997 © 1997 ElsevierScienceLtd All rights reserved. Printed in Great Britain 0956-053X/97 $17.00+ 0.00
Pergamon PII: S0956-053X(97)00031-7
ORIGINAL CONTRIBUTION
DETERMINATION OF ACUTE Hg EMISSIONS FROM SOLIDIFIED/STABILIZED CEMENT WASTE FORMS
William P. Hamilton and Alan R. Bowers* Vanderbilt University, Civil and Environmental Engineering, Nashville, TN 37235, U.S.A.
ABSTRACT. The chemical form of mercury in wastes to be solidified/stabilized may lead to volatile losses from the finished solidified/stabilized monolith. Elemental mercury vapor (Hg vapor) was detected in the headspace of batch reactors that contained solidified/stabilized ordinary Portland cement doped with mercuric oxide (HgO) or liquid elemental mercury [Hg°(l)]. Vapor concentrations increased as a function of time and temperature; the headspace over the HgO samples was saturated in about one hour, while the samples containing Hg°(1) reached approx. 20% of saturation in about two hours. Increased temperatures due to cement hydrolysis lead to increased Hg vapor evolution. Mercury solidified/stabilized as mercuric sulfide (HgS, black) emitted no Hg vapor. Data for the HgO and Hg°(1) experiments was fit to a reversible firstorder rate expression. Samples containing HgO displayed the greatest volatility as a result of the rapid dissolution of HgO and the subsequent formation of a strong driving force across the air-water interface. The evolution of Hg vapor from samples solidified/stabilized as Hg°(1) is limited by mass transfer resistances that kinetically limit the dissolution of Hg°(l) into the aqueous phase. The inert character of HgS (extremely low solubility and resistance to oxidative dissolution) prevents the evolution of detectable Hg in wastes solidified/stabilized as HgS. The findings of these studies may be important when considering treatment and disposal scenarios for Hg-containing wastes. © 1997 Elsevier Science Ltd INTRODUCTION
Although most metals are amenable to solidification/stabilization, mercury (Hg) may offer some unique challenges. Unlike most metals under standard environmental conditions, Hg enjoys a complex chemistry in four phases (solid, liquid, aqueous, gaseous), and the range of pH and E . typical of cured cement may favor the generation of a Hg species that is unsuitable for solidification/stabilization. Of particular concern is the unique potential of Hg to volatilize even in slightly oxidized environments; a volatile Hg compound would likely result in the failure of the monolith to contain the waste. In spite of this potential, investigations regarding the solidification/stabilization of Hg wastes rarely consider the vapor phase. Roy eta/. 5,6 used a variety of microscopic and X-ray diffractive techniques to study the microstructure and microchemistry of a Hg-containing sludge that had been solidified/stabilized in ordinary Portland cement (OPC). They were unable to detect any Hg in their solidified/stabilized samples, and did not analyze for Hg vapor. Poon et al. 7,8 studied the solidification/stabilization of Hg in a variety of solidifying agents and concluded that Hg was only physically encapsulated within the cement matrix, and therefore was capable of rapid mobilization once in contact with water. No
Solidification/stabilization technology addresses the mandates of several federal laws (SWDA, HSWA, RCRA, CERCLA) that ban the disposal of free liquids and require the containment of hazardous residues prior to disposal in regulated landfills. 1,2 Although the use of solidification/stabilization results in a more massive and voluminous waste and subsequent increased disposal costs, it is regarded as a Best Demonstrated Available Technology (BDAT) because of its low materials and operating costs, ease of use, and capacity to a contain w~tste and prevent its migration in the natural environment. 1,3,4 Because of these benefits, 20% of remedial decisions at Superfund sites in 1990 involved the use of solidification/stabilization. 1
RECEIVED 8 JANUARY 1997; ACCEPTED 29 APRIL 1997. *Author to whom correspondence may be addressed. Fax: 615-3223365; E-mail: [email protected] or bowersa@vuse. vanderbilt.edu Acknowledgements--The authors wish to thank Dr R.R. Turner for the use of his laboratory and Jerome model 431-X Hg vapor analyzer, and Mr M.O. Barnett for thoughtful discussions of Hg chemistry and Hg analysis. William Hamilton is supported by the Civil and Environmental Engineering Department of Vanderbilt University.
25
26
W.P. HAMILTON AND A. R. BOWERS
consideration was given to the vapor phase. Cheeseman et al. 9 studied the effect of speciation on heavy metal leaching from solidified/stabilized monoliths using three leaching protocols. They found that Hg concentrations in aqueous leachates were lowest when Hg was solidified in the sulfide (HgS) form, but their Toxicity Characteristic Leachate Procedure (TCLP) analysis did not include a headspace analysis. Yang l° also investigated Hg leaching from solidified/ stabilized monoliths using a modified TCLP analysis without including a headspace analysis. Bonen and Sarkar 1~ investigated the effects of CO2 attack on solidified/stabilized monoliths by submerging cured cement bars into carbonated (CO2-bubbled) deionized water. While they observed adequate reproducibility with the more conservative metals (Ni, Cd, Pb), Hg analyses were not reproducible, possibly due to volatile Hg losses. Cocke and co-workers 12,13 used X-ray photon spectroscopy and energy dispersive spectroscopy to conclude that the final chemical form of Hg initially solidified as aqueous Hg(NO3)2° was HgO, but no mention was made of Hg vapor analysis. Since Hg vapor is a significant health risk, and since little attention has been given to its potential presence in solidification/stabilization, it is the goal of this work to assess the extent of Hg volatilization during solidification/stabilization; specifically, this work will focus on the initial phase (acute emissions) of solidification/stabilization of Hg-containing wastes. EXPERIMENTAL
Materials
Commercially available Type I OPC that conforms to ASTM C-150 standards was used (Holnam, Inc., Theodore, AL). Type I OPC conforming to these standards has the following general composition: 14 Component
wt%
CaO AI203 Fe203 SiO2 MgO SO3 Na20 + K20 TiO2 P205
60-67 3-8 0.~6.0 17-25 0.1-5.0 1-3 0.5-1.3 0.1-0.4 0.1-0.2
Commercially pure, triple distilled commercial grade elemental Hg [Hg°(l)] was supplied by Goldsmith Brothers Smelting and Refining Company (Chicago, IL). Mercuric oxide powder (HgO, red) was reagent grade and supplied by Malinckrodt
Chemical, Inc. (St Louis, MO). Granular mercuric sulfide (HgS, metacinnibar) was supplied by Aldrich Chemical Co. (Milwaukee, WI). Headspace Analysis
Headspace vapor concentrations were measured using a Jerome model 431-X portable Hg vapor analyzer (Arizona Instruments, Phoenix, AZ). This instrument exploits the propensity of Hg vapor to form amalgams with elemental gold (Au). When Hg vapor is passed over a thin Au film carrying an electrical current, it amalgamates and the concomitant increase in electrical resistance experienced by the Au film is proportional to the mass of the Hg vapor in the sample. Multiple samples can be taken before the Au film is saturated. The device provides two methods of sampling: an undiluted method, in which the entire volume of air drawn into the device is analyzed; and a diluted method, in which smaller volumes of Hg-laden air are injected through a sampling port with a vapor syringe, diluted with Hg-free influent air, and analyzed. The latter method allows analysis of more concentrated samples without rapidly saturating the instrument. Once saturated, the Au film can be regenerated by thermally desorbing the amalgamated Hg. Although there are more sensitive analytical options for the detection of Hg vapor such as cold vapor atomic absorption (CVAA) or cold vapor atomic fluorescence (CVAF) spectrometry, these methods are more expensive. The Jerome model 431-X provides ample precision and accuracy at a fraction of the cost of other analytical methods. Because of its flexibility and low cost, the use of the Jerome model 431-X for the analysis of Hg vapor is not unprecedented. Kriger and Turner 15 employed the model 431-X as a field screening tool to determine dissolved and soil-bound Hg concentrations at a contaminated site in Oak Ridge, Tennessee. Their technique was found to be quick, portable, and inexpensive relative to the more expensive analytical options (CVAA or CVAF). Callaghan 16 used the Jerome to monitor for Hg°(l) in liquefied natural gas operations. Detection of Hg in the process equipment prevented corrosion and the formation of explosive HgN compounds. Procedure
Experiments were performed in 1-1 HDPE bottles (I-CHEM Certified; Nalge Company, Rochester, NY). Twenty-five ml of deionized/distilled water was combined with 50 g of OPC (w/c = 0.5), and blended with a spatula. Once well blended (,,~30s), the slurry was spiked with Hg [either as Hg°(l), or as a divalent mineral (HgO or HgS)], resulting in a total Hg concentration of 0.2% (by weight dry solids) which is in the range of previous Hg solidification/stabilization work. 7,8 The mixture was blended again, the head
27
DETERMINATION OF ACUTE Hg EMISSIONS
Cherrasorl~ Adsorpti( Ion Exch~ Isomorphic St Diodochy
1
Pore Space
afion,
Cement Matrix
FIGURE 1. Possiblephysicochemicalprocessesinvolvedin the solidification/stabilizationof Hg-containingwaste/2,Is
space was sampled, and the bottles were sealed with Parafilm lab film. Spiked samples were allowed to hydrolyze (about 4 min), and then the headspace was sampled again. Typically, vapor samples were withdrawn from the headspace of the reaction vessels by penetrating the Parafilm seal with a 2ml vapor syringe; however, samples with low Hg vapor concentrations (undetectable using this method) were sampled directly (no dilution). In both cases, the vessels were resealed with another layer of Parafilm after sampling. All experiments were performed in triplicate to assess reproducibility. A process blank indicated no detectable Hg originated from the OPC or water. Ambient air samples were drawn prior to the analysis of each sample, and Hg vapor was never detected in any of these samples. The calibration of the instrument was periodically verified by analyzing a known volume of Hg-saturated vapor injected in the syringe sampling port. The instrument remained calibrated throughout the entirety of the experiments. The temperature of the curing cement was measured with a thermometer inserted in the center of the curing cement blank. The temperature of the cement increased 2°C (from 296 to 298K) in the first 30min of the experiment, and remained constant thereafter.
equilibrium conditions can offer some insight into chemical speciation within the cured cement. Central to the behavior of Hg is its propensity to disproportionate 17 according to the reaction: Hg 2+ *+ Hg ° + Hg 2+
(1)
15
-.. -.°
W a t e r Oxidized
1
4
A
n ~
q
0.5
0
~
-0.5
RESULTS AND DISCUSSION Water Reduced
The cement environment is complicated due, in part, to the variety of chemical phases present, the change in system temperatures due to cement hydrolysis, and the extreme pH and ionic strength conditions. In addition, a number of physicochemical processes can affect the fate of solidified/stabilized Hg (Fig. 1), making system modeling challenging. However,
-]
b
I
I
I
I
I
2
4
6
8
10
12
I4
pn
FIGURE 2. Pourbaixdiagramfor the systemHg-S-H20 at 25°C and 1atm.
28
W. P. HAMILTON AND A. R. BOWERS
The equilibrium speciation is determined by the oxidation-reduction (redox) conditions of the system, which are illustrated with the use of a Pourbaix diagram (Fig. 2). Superimposing on the Pourbaix diagram for Hg the domain of E n and pH characteristic of cured cement 18 reveals that both Hg(II) and Hg(0) are thermodynamically possible species. This has serious implications for the capacity of solidified/stabilized monoliths to retain Hg waste. If equilibrium conditions favor the formation of the volatile species, then the capacity of the waste form
118
to retain the wastes is only a function of the kinetics of the redox reaction and the overall diffusion constant of the monolith. A variety of approaches are applied to solidification/stabilization, i.e. some waste is precipitated as an oxide, silicate, or sulfide species, 9Al some is solidified as a dissolved component, 7,8,13 and some is solidified as a 'sludge', or a mixture of aqueous, solid and semisolid phases, s,6:° Due to the variety of possible chemical species, cement samples were spiked with one of three different chemical forms of Hg
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20
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16 14 12 10
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* . . . . . . . . . .
O
.......................................
HgO~t~. 298 K ............H..g.°[email protected]... . . . . . .
" .
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.
Hg ° from SIS HgO
6 k O
~
~ I ~ . . . . ........y.,
HgO blank
0---0
20
a
I
I
i
i
i
i
40
60
80
100
120
140
time (min)
20 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
14 "" 12 l0
....
. °blank
2o-If' " ' ~,, 0
t
I
I
I
[
I
bI
20
40
60
80
100
120
140
time (rain)
20 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
m
,-, 15 10 Hg* from S/S HgS ~ a n d HgS blank (both non detect)
O t-
/
O
~ e 0
C
', I0
e
', 20
e
',
^ I
I^
30
40
50
time (rain)
v
w
v
60
FIGURE 3. Evolution of Hg vapor from solidified/stabilized OPC monoliths. Initial Hg species (a) HgO(s); (b) Hg°(l); (c) HgS(s). Solid lines represent fit to the first-order reversible model. Data points represent the average of triplicate experimental runs. Blank data is included to illustrate the effect of cement on Hg° evolution.
29
DETERMINATION OF ACUTE Hg EMISSIONS chosen to represent a continuum of the likely equilibrium species possible under different redox conditions. HgO was used to represent an oxidized species, Hg°(l) represented the transitional redox condition, and HgS was used to represent the reduced phase. Note that although the valence of Hg in HgS is + 2, from a thermodynamic standpoint, HgS represents the most reduced form of Hg in the system Hg-O-S at 25°C (see Fig. 2). The evolution of Hg vapor as a function of time for three different initial Hg species is depicted in Fig. 3. No detectable Hg vapor evolved from samples solidified as HgS [Fig. 3(c)]; however, samples that initially contained Hg°(1) or HgO emitted significant amounts of Hg vapor. Samples containing HgO solidified in OPC evolved the most vapor [Fig. 3(a)]; the headspaces of the batch samples were saturated with Hg vapor in about one hour. By contrast, blanks prepared with the same initial mass of HgO (0.1 g) and the same water:solid ratio (H20:HgO = 0.5) but without cement exhibited little Hg ° evolution over the same time period. The vapor concentration of Hg ° in HgO samples solidified with cement approached 96% of saturation after 25 min while the vapor concentration of Hg ° in blanks without cement achieved only about 3.5% of saturation over the same time period. The samples that contained Hg°(1) also evolved significant Hg vapor [Fig. 3(b)], but the headspace never reached saturation, even after considerable time. By comparison, a bead of Hg°(1) (same amount of Hg as was added to the solidified/ stabilized samples, but without water or cement) saturated the headspace of the 1-1 vessel in less than one hour [Fig. 3(b)]. The effect of temperature on saturation vapor pressure is described by the Clausius-Clapeyron equation: 19
din(p) AnvapdT dT
R
T2
(2)
Where p is the vapor pressure, T is the temperature, R is the gas constant, and AHwp is the enthalpy of vaporization. Assuming the quotient AHv~p/R is a constant (valid for environmental conditions) the integrated form of the Clausius-Clapeyron equation is the Antoine equation: 2° B In(p) = A - -~ (3) where p (atm) and T(K) are as defined before, and A and B are element-specific constants. For Hg, A= 12.044 and B=7447.4 (adapted from Ref. 20). Cement hydrolysis is an exothermic process, and temperatures of curing cement can increase up to 40°C. 14 This increased temperature will result in a higher Hg vapor pressure. Figure 3(a) illustrates
this effect. While the Hg vapor concentration appears to exceed the saturation vapor pressure of Hg ° at the initial ambient temperature (Hg°sat@296 = 16.62 mgm-3), the increase in temperature of 2°C due to cement hydrolysis explains this deviation (Hg°sat@298 = 19.54mgm-3). Due to the complex system environment and the variety of physicochemical processes affecting Hg in solidified/stabilized samples, a number of competitive and parallel reactions may occur (Fig. 4). These reactions are defined as: 1. Dissolution/precipitation of oxide and sulfide minerals; 2. Aqueous hydrolysis [Hg(OH)2%-~Hg(OH)3-]; 3. Oxidation/reduction [Hg(II) ~ Hg(O)]; 4. Volatilization [Hg°(aq) ~ Hg°(g)]; or 5. Condensation [Hg°(aq) ~ Hg°(1)]. Aqueous hydrolysis reactions (rxn 2) for divalent metals are generally fast,2~ and the aqueous reduction reaction (rxn 3) is likely fast due to (a) the reducing conditions of the system and because (b) dissolved species are more likely to be reduced than solid species. Dissolution rates of oxide minerals (rxn 1) at a constant temperature are a function of (a) the concentration of H ÷ (pH) and other sorbing ligands which, when adsorbed, can polarize metal-oxygen bonds near the mineral surface and expedite dissolution, and (b) the concentration of surface functional groups, the discrete sites where these dissolution reactions take place, a2-24 It is likely that because of the high pH of concrete, surface functional groups will be completely deprotonated, resulting in strong polarization of metal-oxygen bonds and rapid dissolution. The rate of isothermal mass transfer (rxn 4 and 5) is a function of the rate of diffusion of the species through a mass transfer zone;25 the diffusion rate depends on, among other things, the number of mass transfer zones, since the individual effects of these are added, similar to resistances in series. Assuming the rate of Hg volatilization is due to mass transfer limitations, then the time dependent vapor phase Hg concentration can be described by the following first-order expression: 26 dHg = k(Hg~at _ Hgl) dt
(4)
where Hg~at = saturation vapor phase concentration of Hg, mgm -3 Hg I = vapor phase concentration of Hg, mg m -3, at time t t = time, min k = first order rate constant, min-% (sam ples of constant volume and cross sec tional area)
30
W. P. H A M I L T O N A N D A. R. B O W E R S
Hg°(g)
Hg~s) + H200) ~
Hg(OH)2 ° ~
Hg°(aq)
1lc Hg(OI-I)5"
Hg~)
FIGURE 4. Possible parallel and competitive reactions involving aqueous Hg. HgX(s) represents either HgO(s) or HgS(s). The integrated form of this differential equation relates the concentration of Hg ° in the vapor phase as a function of time: Hg~ = Hgsat°(1 - e -kt)
(5)
The data for the solidification/stabilization of HgO and Hg°(l) was fit to this equation using a nonlinear least-squares program. 27 The values of Hg~at and k derived from this fit for each experiment are summarized in Table 1, and values of Hg~at predicted with this model are compared to Hg~at values calculated with the Antoine equation [eqn (3)] to evaluate the assumption that mass transfer is the mechanism controlling Hg volatility. The values of Hg°sat for the bead of Hg°(l) as well as the solidified/stabilized HgO sample indicates that mass transfer limitations across the water-air interface are controlling the volatilization of Hg vapor. The values of Hg~at predicted with the first-order model are similar to the values predicted by the Antoine equation, particularly considering the other possible mechanisms that may be operating on the system. While the data for the solidification/stabilization of Hg°(l) resulted in a good fit, the value of Hg~at predicted with the least-squares fit of eqn (5) deviated from the value of Hg~at determined with the Antoine equation because insufficient data was collected (Hg vapor concentration did not approach saturation during the experiment); however, an order-of-magnitude approximation of the rate constant can be made by assuming that as t approaches infinity, the vapor pressure of Hg approaches saturation. Including the value of Hg~at (16.62mgm -3) at t > 1900rain as a point to the fit of eqn (5) results in a new approxi-
mation of k (k -- 0.0282 min-l), without sacrificing goodness-of-fit [Fig. 3(b)]. The volatilization of Hg vapor is controlled by the driving force, i.e. concentration gradient, between phases. 25'26 The evolution of Hg vapor from solidified/stabilized HgO is faster than the evolution rate from solidified/stabilized Hg°(l) due to the dissolution rates of the two species. Both the dissolution of HgO and subsequent aqueous reduction of Hg(II)(aq) to Hg°(aq) are assumed fast, resulting in a nearly instantaneous saturated aqueous Hg°(aq) concentration, i.e. the concentration of Hg°(aq) approaches its solubility (60.78/Lg/l), 2s and maximum driving force. The rate constants indicate that additional limitations on the volatility of solidified/stabilized Hg°(l) may be controlling the overall rate of Hg vapor evolution. As a non-aqueous phase liquid (NAPL), an interface develops between Hg°(1) and water. It is likely that during hydrolysis, this interface becomes increasingly complex due to the formation of a crystalline Ca(OH)2 sheath around any Hg°(l) droplets in the matrix, consistent with the gel model of cement hydration, ls,29 This sheath forms an additional mass transfer interface and increases the overall transfer resistance, resulting in limited Hg°(1) dissolution and Hg vapor evolution. HgS is one of the most insoluble of the metal sulfides 17 (Ksp = 10-52), and, as shown in Fig. 2, is the thermodynamically stable species under reducing conditions. However, even under oxidizing conditions, HgS is very resistant to weathering, 3° i.e., the oxidation reaction: HgS(s) + 202(aq) 4- 2H20 Hg(OH)2(aq) 4- SO 2- 4- 2H +
(6)
is very slow, with a pseudo first-order half-life of ~180 years. 31 The low solubility and inert redox character of HgS will prohibit a significant concentration gradient from developing, resulting in a particularly stable and innocuous Hg phase, at least over the time frame addressed by these experiments (< 3h). Figure 5 illustrates the proposed conceptual model. The dissolution of the oxide is fast resulting in a relatively high concentration of Hg(II)(aq), which
TABLE 1 Summary of Rate Data for the Evolution of Hg ° from SoHditled/StabiUzed cement Hgsat
HgO(s) solidified/stabilized Hg°(l) untreated Hg°(1) solidified/stabilized Corrected*
[eqn (3)] (mgm-3)
Hg~at
[eqn (5)] (mgm-3)
19.54 16.62 19.54 19.54
"Kinetic data based on a model fit [eqn (5)] including H g t
16.75 16.56 5.756 19.54 ° =
19.54mg m-3 [eqn (3)].
k [eqn (5)] (min-l)
R2
0.153 0.0686 0.0131 0.00230
0.967 0.999 0.982 0.974
31
DETERMINATION OF ACUTE Hg EMISSIONS a
HgO(s) + H20(1) + xOH" ~
Hg(OH)X-2+x~
Hg°(aq) ~
Hg vapor
rate limiting
Hg°(l
Hg°ca(OH)2
Hg°'a )~'J ~ q/.'-~
Hg vapor
rate l i m i t i n /
HgS(s) ~
Hg 2+ + S2"
AND HgS(s) + O2(aq) + 2H20(1) ~
Hg(OH)2(aq) +SO42" +21-1+
FIGURE 5. Mechanisms controlling the transport of Hg vapor in solidified/stabilized monoliths. Hg°ca(oH)2represents Hg diffusing through the Hg°(I)-Ca(OH)2(s)interface. Pairs of curved arrows signifymass transfer processes. due to the reducing conditions in the matrix, is quickly reduced to Hg°(aq). The subsequent evolution of Hg vapor is limited only by the rate of mass transfer across the air-water interface. In the elemental form, Hg°(l) is subject to additional mass transfer limitations at the Hg°(1)-water interface that are compounded by the precipitation of Ca(OH)2 along this interface. These mass transfer limitations impede the rate of dissolution of Hg°(1) resulting in a lower Hg°(aq) concentration with respect to time, a smaller concentration gradient between the water and the air, and thus a slower evolution rate of Hg vapor from the monolith. In the case of HgS, the Hg(II)(aq) concentration is controlled both by the low solubility of HgS and the exceptionally slow oxidation of HgS. The extremely low dissolved Hg(II)(aq) concentration results in an insignificant driving force, and therefore little Hg vapor is evolved. While no studies to date have focused on losses of Hg vapor from solidified/stabilized cement wastes, Lindberg and Turner 32 observed volatile Hg ° losses from Hg-containing sludges associated with a chloralkali facility. Measurements as well as predictions from a Gaussian diffusion model indicated that, depending on temperature, steady state Hg vapor concentrations in ambient air could be significant (approaching 1/zgm-3). In addition, attempts to solidify/stabilize wastes containing volatile organic compounds (VOCs) have experienced volatile losses. Weitzman e t al. 3 observed losses of up to 90% during mixing and curing of VOC wastes. Studies conducted by Spence e t al. 33 experienced volatile losses of up to 50%, in spite of the fact that their experiments were
designed to minimize VOC losses. While this study did not focus on the fraction of total Hg lost from the solid, it can be said qualitatively that the losses of Hg (as a fraction of initial mass) during the initial stages of solidification/stabilization were considerably less than the losses of VOCs from similar solids; this may be due to the lower saturation vapor pressure of Hg vapor relative to VOCs, and the fact that experiments were conducted in closed, batch reactors. Experiments in open systems would likely provide an environment that is conducive to greater Hg losses. The conclusions of this study are consistent with the findings of these VOC studies, namely that volatile compounds are lost from solidified/stabilized waste forms. The findings of this study may be important to those making disposal decisions for Hg-contaminated waste. Hg vapor is a neurotoxin, 34 and chronic exposure to Hg vapor has caused effects ranging from micromercurialism (weakness, fatigue, weight loss, gastrointestinal dysfunction) to mercurial erethism (severe behavioral and personality changes, memory loss, insomnia, depression, delirium) and mercurial tremors. If Hg losses are significant in a toxicological sense, worker exposure should be minimized and steps should be taken to reduce the potential for Hg volatilization. While these experiments only examine losses during the initial phase of solidification/stabilization, the results may point to an additional, chronic problem. Further investigation into losses of Hg from solidified/stabilized waste over longer time periods, as well as investigations into the total fraction of Hg lost,
32
s h o u l d be c o n d u c t e d to d e t e r m i n e a p p r o p r i a t e disp o s a l s c e n a r i o s for H g - c o n t a m i n a t e d wastes,
CONCLUSIONS
Mercury has always provided unique challenges due to its c o m p l e x c h e m i s t r y , a n d this i n c l u d e s the treatm e n t o f H g - c o n t a i n i n g wastes. B e c a u s e o f its a b i l i t y to volatilize, the s p e c i a t i o n o f H g is a n i m p o r t a n t f a c t o r w h e n c o n s i d e r i n g d i s p o s a l o p t i o n s . T h e solidif i c a t i o n / s t a b i l i z a t i o n o f H g as e i t h e r H g O o r Hg°(1) m a y lead to the e v o l u t i o n o f H g v a p o r f r o m solidified/stabilized m o n o l i t h s u n d e r t y p i c a l c o n d i t i o n s o f E / - / a n d p H . C o n v e r s e l y , the s o l i d i f i c a t i o n o f H g as H g S s h o w e d n o p r o p e n s i t y to volatilize, likely d u e to its low s o l u b i l i t y a n d i n s e n s i t i v i t y to r e d o x c o n d i t i o n s o v e r the t i m e scale c o n s i d e r e d . T h e s e results m a y be important when considering treatment and disposal o p t i o n s for H g - c o n t a i n i n g wastes.
REFERENCES
1. Barth, E. F. An overview of the history, present status, and future direction of solidification/stabilization technologies for hazardous waste treatment. Journal of Hazardous Materials 24: 103-109 (1990). 2. Wiles, C. C. A review of solidification/stabilizationtechnology. Journal of Hazardous Materials 14:5-21 (1987). 3. Weitzman, L., Hamel, L. E. and Barth, E. F. Evaluation of solidification/stabilization as a best demonstrated available technology for contaminated soils. EPA-600/S2-89/013, Risk Reduction Engineering Laboratory. U.S. Environmental Protection Agency, Cincinnati, OH, (1990). 4. Jones, L. W. Interference mechanisms in waste stabilization/ solidification processes. Journal of Hazardous Materials 24: 83-88 (1990). 5. Roy, A., Eaton, E., Cartledge, F. K. and Tittlebaum, M. E. Solidification/stabilization of a heavy metal sludge by a Portland cement/fly ash binding mixture. Hazardous Waste and Hazardous Materials 8:3341 (1991). 6. Roy, A., Eaton, E., Cartledge, F. K. and Tittlebaum, M. E. Solidification/stabilization of hazardous waste: evidence of physical encapsulation. Environmental Science and Technology 26:1349-1353 (1992). 7. Poon, C. S., Peters, C. J., Perry, R., Barnes, P. and Barker, A. P. Mechanisms of metal stabilization by cement based fixation processes. Science and the Total Environment 41:55-71 (1985). 8. Poon, C. S., Clark, I. A., Peters, C. J. and Perry, R. Mechanisms of metal fixation and leaching by cement-based fixation processes. Waste Management Research 3:127-142 (1985). 9. Cheeseman, C. R., Butcher, E. J., Sollars, C. J. and Perry, R. Heavy metal leaching from hydroxide, sulphide and silicate stabilised/solidified wastes. Waste Management 13:545-552 (1993). 10. Yang, G. C. C. Durability study of a solidified mercury-containing sludge. Journal of Hazardous Materials 34:217-223 (1993). 11. Bonen, D. and Sarkar, S. L. The effects of simulated environmental attack on immobilization of heavy metals doped in cement-based materials. Journal of Hazardous Materials 40: 321-335 (1995). 12. Cocke, D. L. The binding chemistry and leaching mechanisms of hazardous substances in cementitious solidification/stabilization systems. Journal of Hazardous Materials 24:231-253 (1990).
W . P . H A M I L T O N A N D A. R. B O W E R S
13. McWhinney, H. G., Cocke, D. L., Balke, K. and Ortego, J. D. An investigation of mercury solidification and stabilization in Portland cement using X-ray photoelectron spectroscopy and energy dispersive spectroscopy. Cement & Concrete Research 20:79-91 (1990). 14. Lea, F. M. The Chemistry of Cement and Concrete, 3rd Vol. Chemical, New York, (1970). 15. Kriger, A. A. and Turner, R. R. Field analysis of mercury in water, sediment and soil using static headspace analysis. Water, Air, and Soil Pollution 80:1295-1304 (1995). 16. Callaghan, K. W. Mercury analysis in natural gas using a gold film sensor based analyzer. Arizona Instrument Corporation, Phoenix, AZ, (1993). 17. Hepler, L. G. and Olofsson, G. Mercury: Thermodynamic properties, chemical equilibria, and standard potentials. Chemical Reviews 75:585-602 (1975). 18. Yousuf, M., Mollah, A., Vempati, R. K., Lin, T. -C. and Cocke, D. L. The interracial chemistry of solidification/stabilization of metals in cement and pozzolanic material systems. Waste Management 15:137-148 (1995). 19. Atkins, P. W. Physical Chemistry, 4th edn. Freeman, New York, (1991). 20. Lide, D. R. (ed.) CRC Handbook of Chemistry and Physics, 77 edn. CRC Press, Boca Raton, (1996). 21. Margerum, D. W., Cayley, G. R., Whetherburn, D. C. and Pagenkopf, G. K., Kinetics and mechanisms of complex formation and ligand exchange. In Coordination Chemistry, A. E. Martell (ed.), American Chemical Society, Washington, DC, pp. 1-220 (1978). 22. Furrer, G. and Stumm, W. The coordination chemistry of weathering: I. Dissolution kinetics of &AI203 and BeO. Geochimica et Cosmochimica Acta 50:1847-1860 (1986). 23. Wieland, E., Wehrli, B. and Stumm, W. The coordination chemistry of weathering: III. A generalization on the dissolution rates of minerals. Geochimica et Cosmochimica Acta 52: 1969-1981 (1988). 24. Stumm, W. and Sulzberger, B. The cycling of iron in natural environments: considerations based on laboratory studies of heterogeneous redox processes. Geochimica et Cosmochimica Acta 56:3233-3257 (1992). 25. Mackay, D. Multimedia Environmental Models: The Fugacity Approach. Lewis, New York (1991). 26. Thibodeaux, L. J. Chemodynamics: Environmental Movement of Chemicals in Air, Water, and Soil. John Wiley & Sons, New York (1979). 27. Hyams, D. CurveExpert: A Curve Fitting System for Windows, version 1.3. Microsoft Corporation, Clemson SC, (1996). 28. Clever, H. L., Johnson, S. A. and Derrick, M. E. The solubility of mercury and some sparingly soluble mercury salts in water and aqueous electrolyte solutions. Journal of Physical Chemical Reference Data 14:631-680 (1985). 29. Hansen, J. The delicate architecture of cement. Science 82 3: 49-55 (1982). 30. Barnett, M. O., Harris, L. A., Turner, R. R., Stevenson, R. J., Henson, T. J., Melton, R. C. and Hoffman, D. P. Formation of authigenic mercuric sulfide in soil. Submitted to Environmental Science and Technology. 31. Barnett, M. O., Turner, R. R. and Singer, P. C. Oxidation of authigenic mercuric sulfide. Abstracts of the Geological Society of America 27:35 (1995). 32. Lindberg, S. E. and Turner, R. R. Mercury emissions from chlorine-production solid waste deposits. Nature 268:133-136 (1977). 33. Spence, R. D., Gilliam, T. M., Morgan, I. L. and Osborne, S. C. Stabilization/solidification of wastes containing volatile organic compounds in commercial cementitious waste forms. In Stabilization and Solidification of Hazardous, Radioactive, and Mixed Wastes, T. M. Gilliam and C. C. Wiles (eds).. American Society of Testing and Materials, Philadelphia, pp. 61-72 (1992). 34. Berlin, M. Mercury. In Handbook on the Toxicology of Metals, L. Friberg, G. F. Nordberg and V. B. Vouk (eds), pp. 387-445. Elsevier, Amsterdam (1986).
Pergamon PII: S0956-053X(97)00031-7
ORIGINAL CONTRIBUTION
DETERMINATION OF ACUTE Hg EMISSIONS FROM SOLIDIFIED/STABILIZED CEMENT WASTE FORMS
William P. Hamilton and Alan R. Bowers* Vanderbilt University, Civil and Environmental Engineering, Nashville, TN 37235, U.S.A.
ABSTRACT. The chemical form of mercury in wastes to be solidified/stabilized may lead to volatile losses from the finished solidified/stabilized monolith. Elemental mercury vapor (Hg vapor) was detected in the headspace of batch reactors that contained solidified/stabilized ordinary Portland cement doped with mercuric oxide (HgO) or liquid elemental mercury [Hg°(l)]. Vapor concentrations increased as a function of time and temperature; the headspace over the HgO samples was saturated in about one hour, while the samples containing Hg°(1) reached approx. 20% of saturation in about two hours. Increased temperatures due to cement hydrolysis lead to increased Hg vapor evolution. Mercury solidified/stabilized as mercuric sulfide (HgS, black) emitted no Hg vapor. Data for the HgO and Hg°(1) experiments was fit to a reversible firstorder rate expression. Samples containing HgO displayed the greatest volatility as a result of the rapid dissolution of HgO and the subsequent formation of a strong driving force across the air-water interface. The evolution of Hg vapor from samples solidified/stabilized as Hg°(1) is limited by mass transfer resistances that kinetically limit the dissolution of Hg°(l) into the aqueous phase. The inert character of HgS (extremely low solubility and resistance to oxidative dissolution) prevents the evolution of detectable Hg in wastes solidified/stabilized as HgS. The findings of these studies may be important when considering treatment and disposal scenarios for Hg-containing wastes. © 1997 Elsevier Science Ltd INTRODUCTION
Although most metals are amenable to solidification/stabilization, mercury (Hg) may offer some unique challenges. Unlike most metals under standard environmental conditions, Hg enjoys a complex chemistry in four phases (solid, liquid, aqueous, gaseous), and the range of pH and E . typical of cured cement may favor the generation of a Hg species that is unsuitable for solidification/stabilization. Of particular concern is the unique potential of Hg to volatilize even in slightly oxidized environments; a volatile Hg compound would likely result in the failure of the monolith to contain the waste. In spite of this potential, investigations regarding the solidification/stabilization of Hg wastes rarely consider the vapor phase. Roy eta/. 5,6 used a variety of microscopic and X-ray diffractive techniques to study the microstructure and microchemistry of a Hg-containing sludge that had been solidified/stabilized in ordinary Portland cement (OPC). They were unable to detect any Hg in their solidified/stabilized samples, and did not analyze for Hg vapor. Poon et al. 7,8 studied the solidification/stabilization of Hg in a variety of solidifying agents and concluded that Hg was only physically encapsulated within the cement matrix, and therefore was capable of rapid mobilization once in contact with water. No
Solidification/stabilization technology addresses the mandates of several federal laws (SWDA, HSWA, RCRA, CERCLA) that ban the disposal of free liquids and require the containment of hazardous residues prior to disposal in regulated landfills. 1,2 Although the use of solidification/stabilization results in a more massive and voluminous waste and subsequent increased disposal costs, it is regarded as a Best Demonstrated Available Technology (BDAT) because of its low materials and operating costs, ease of use, and capacity to a contain w~tste and prevent its migration in the natural environment. 1,3,4 Because of these benefits, 20% of remedial decisions at Superfund sites in 1990 involved the use of solidification/stabilization. 1
RECEIVED 8 JANUARY 1997; ACCEPTED 29 APRIL 1997. *Author to whom correspondence may be addressed. Fax: 615-3223365; E-mail: [email protected] or bowersa@vuse. vanderbilt.edu Acknowledgements--The authors wish to thank Dr R.R. Turner for the use of his laboratory and Jerome model 431-X Hg vapor analyzer, and Mr M.O. Barnett for thoughtful discussions of Hg chemistry and Hg analysis. William Hamilton is supported by the Civil and Environmental Engineering Department of Vanderbilt University.
25
26
W.P. HAMILTON AND A. R. BOWERS
consideration was given to the vapor phase. Cheeseman et al. 9 studied the effect of speciation on heavy metal leaching from solidified/stabilized monoliths using three leaching protocols. They found that Hg concentrations in aqueous leachates were lowest when Hg was solidified in the sulfide (HgS) form, but their Toxicity Characteristic Leachate Procedure (TCLP) analysis did not include a headspace analysis. Yang l° also investigated Hg leaching from solidified/ stabilized monoliths using a modified TCLP analysis without including a headspace analysis. Bonen and Sarkar 1~ investigated the effects of CO2 attack on solidified/stabilized monoliths by submerging cured cement bars into carbonated (CO2-bubbled) deionized water. While they observed adequate reproducibility with the more conservative metals (Ni, Cd, Pb), Hg analyses were not reproducible, possibly due to volatile Hg losses. Cocke and co-workers 12,13 used X-ray photon spectroscopy and energy dispersive spectroscopy to conclude that the final chemical form of Hg initially solidified as aqueous Hg(NO3)2° was HgO, but no mention was made of Hg vapor analysis. Since Hg vapor is a significant health risk, and since little attention has been given to its potential presence in solidification/stabilization, it is the goal of this work to assess the extent of Hg volatilization during solidification/stabilization; specifically, this work will focus on the initial phase (acute emissions) of solidification/stabilization of Hg-containing wastes. EXPERIMENTAL
Materials
Commercially available Type I OPC that conforms to ASTM C-150 standards was used (Holnam, Inc., Theodore, AL). Type I OPC conforming to these standards has the following general composition: 14 Component
wt%
CaO AI203 Fe203 SiO2 MgO SO3 Na20 + K20 TiO2 P205
60-67 3-8 0.~6.0 17-25 0.1-5.0 1-3 0.5-1.3 0.1-0.4 0.1-0.2
Commercially pure, triple distilled commercial grade elemental Hg [Hg°(l)] was supplied by Goldsmith Brothers Smelting and Refining Company (Chicago, IL). Mercuric oxide powder (HgO, red) was reagent grade and supplied by Malinckrodt
Chemical, Inc. (St Louis, MO). Granular mercuric sulfide (HgS, metacinnibar) was supplied by Aldrich Chemical Co. (Milwaukee, WI). Headspace Analysis
Headspace vapor concentrations were measured using a Jerome model 431-X portable Hg vapor analyzer (Arizona Instruments, Phoenix, AZ). This instrument exploits the propensity of Hg vapor to form amalgams with elemental gold (Au). When Hg vapor is passed over a thin Au film carrying an electrical current, it amalgamates and the concomitant increase in electrical resistance experienced by the Au film is proportional to the mass of the Hg vapor in the sample. Multiple samples can be taken before the Au film is saturated. The device provides two methods of sampling: an undiluted method, in which the entire volume of air drawn into the device is analyzed; and a diluted method, in which smaller volumes of Hg-laden air are injected through a sampling port with a vapor syringe, diluted with Hg-free influent air, and analyzed. The latter method allows analysis of more concentrated samples without rapidly saturating the instrument. Once saturated, the Au film can be regenerated by thermally desorbing the amalgamated Hg. Although there are more sensitive analytical options for the detection of Hg vapor such as cold vapor atomic absorption (CVAA) or cold vapor atomic fluorescence (CVAF) spectrometry, these methods are more expensive. The Jerome model 431-X provides ample precision and accuracy at a fraction of the cost of other analytical methods. Because of its flexibility and low cost, the use of the Jerome model 431-X for the analysis of Hg vapor is not unprecedented. Kriger and Turner 15 employed the model 431-X as a field screening tool to determine dissolved and soil-bound Hg concentrations at a contaminated site in Oak Ridge, Tennessee. Their technique was found to be quick, portable, and inexpensive relative to the more expensive analytical options (CVAA or CVAF). Callaghan 16 used the Jerome to monitor for Hg°(l) in liquefied natural gas operations. Detection of Hg in the process equipment prevented corrosion and the formation of explosive HgN compounds. Procedure
Experiments were performed in 1-1 HDPE bottles (I-CHEM Certified; Nalge Company, Rochester, NY). Twenty-five ml of deionized/distilled water was combined with 50 g of OPC (w/c = 0.5), and blended with a spatula. Once well blended (,,~30s), the slurry was spiked with Hg [either as Hg°(l), or as a divalent mineral (HgO or HgS)], resulting in a total Hg concentration of 0.2% (by weight dry solids) which is in the range of previous Hg solidification/stabilization work. 7,8 The mixture was blended again, the head
27
DETERMINATION OF ACUTE Hg EMISSIONS
Cherrasorl~ Adsorpti( Ion Exch~ Isomorphic St Diodochy
1
Pore Space
afion,
Cement Matrix
FIGURE 1. Possiblephysicochemicalprocessesinvolvedin the solidification/stabilizationof Hg-containingwaste/2,Is
space was sampled, and the bottles were sealed with Parafilm lab film. Spiked samples were allowed to hydrolyze (about 4 min), and then the headspace was sampled again. Typically, vapor samples were withdrawn from the headspace of the reaction vessels by penetrating the Parafilm seal with a 2ml vapor syringe; however, samples with low Hg vapor concentrations (undetectable using this method) were sampled directly (no dilution). In both cases, the vessels were resealed with another layer of Parafilm after sampling. All experiments were performed in triplicate to assess reproducibility. A process blank indicated no detectable Hg originated from the OPC or water. Ambient air samples were drawn prior to the analysis of each sample, and Hg vapor was never detected in any of these samples. The calibration of the instrument was periodically verified by analyzing a known volume of Hg-saturated vapor injected in the syringe sampling port. The instrument remained calibrated throughout the entirety of the experiments. The temperature of the curing cement was measured with a thermometer inserted in the center of the curing cement blank. The temperature of the cement increased 2°C (from 296 to 298K) in the first 30min of the experiment, and remained constant thereafter.
equilibrium conditions can offer some insight into chemical speciation within the cured cement. Central to the behavior of Hg is its propensity to disproportionate 17 according to the reaction: Hg 2+ *+ Hg ° + Hg 2+
(1)
15
-.. -.°
W a t e r Oxidized
1
4
A
n ~
q
0.5
0
~
-0.5
RESULTS AND DISCUSSION Water Reduced
The cement environment is complicated due, in part, to the variety of chemical phases present, the change in system temperatures due to cement hydrolysis, and the extreme pH and ionic strength conditions. In addition, a number of physicochemical processes can affect the fate of solidified/stabilized Hg (Fig. 1), making system modeling challenging. However,
-]
b
I
I
I
I
I
2
4
6
8
10
12
I4
pn
FIGURE 2. Pourbaixdiagramfor the systemHg-S-H20 at 25°C and 1atm.
28
W. P. HAMILTON AND A. R. BOWERS
The equilibrium speciation is determined by the oxidation-reduction (redox) conditions of the system, which are illustrated with the use of a Pourbaix diagram (Fig. 2). Superimposing on the Pourbaix diagram for Hg the domain of E n and pH characteristic of cured cement 18 reveals that both Hg(II) and Hg(0) are thermodynamically possible species. This has serious implications for the capacity of solidified/stabilized monoliths to retain Hg waste. If equilibrium conditions favor the formation of the volatile species, then the capacity of the waste form
118
to retain the wastes is only a function of the kinetics of the redox reaction and the overall diffusion constant of the monolith. A variety of approaches are applied to solidification/stabilization, i.e. some waste is precipitated as an oxide, silicate, or sulfide species, 9Al some is solidified as a dissolved component, 7,8,13 and some is solidified as a 'sludge', or a mixture of aqueous, solid and semisolid phases, s,6:° Due to the variety of possible chemical species, cement samples were spiked with one of three different chemical forms of Hg
24
20
•
.
16 14 12 10
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* . . . . . . . . . .
O
.......................................
HgO~t~. 298 K ............H..g.°[email protected]... . . . . . .
" .
.
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.
.
Hg ° from SIS HgO
6 k O
~
~ I ~ . . . . ........y.,
HgO blank
0---0
20
a
I
I
i
i
i
i
40
60
80
100
120
140
time (min)
20 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
14 "" 12 l0
....
. °blank
2o-If' " ' ~,, 0
t
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I
I
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bI
20
40
60
80
100
120
140
time (rain)
20 . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
m
,-, 15 10 Hg* from S/S HgS ~ a n d HgS blank (both non detect)
O t-
/
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~ e 0
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', I0
e
', 20
e
',
^ I
I^
30
40
50
time (rain)
v
w
v
60
FIGURE 3. Evolution of Hg vapor from solidified/stabilized OPC monoliths. Initial Hg species (a) HgO(s); (b) Hg°(l); (c) HgS(s). Solid lines represent fit to the first-order reversible model. Data points represent the average of triplicate experimental runs. Blank data is included to illustrate the effect of cement on Hg° evolution.
29
DETERMINATION OF ACUTE Hg EMISSIONS chosen to represent a continuum of the likely equilibrium species possible under different redox conditions. HgO was used to represent an oxidized species, Hg°(l) represented the transitional redox condition, and HgS was used to represent the reduced phase. Note that although the valence of Hg in HgS is + 2, from a thermodynamic standpoint, HgS represents the most reduced form of Hg in the system Hg-O-S at 25°C (see Fig. 2). The evolution of Hg vapor as a function of time for three different initial Hg species is depicted in Fig. 3. No detectable Hg vapor evolved from samples solidified as HgS [Fig. 3(c)]; however, samples that initially contained Hg°(1) or HgO emitted significant amounts of Hg vapor. Samples containing HgO solidified in OPC evolved the most vapor [Fig. 3(a)]; the headspaces of the batch samples were saturated with Hg vapor in about one hour. By contrast, blanks prepared with the same initial mass of HgO (0.1 g) and the same water:solid ratio (H20:HgO = 0.5) but without cement exhibited little Hg ° evolution over the same time period. The vapor concentration of Hg ° in HgO samples solidified with cement approached 96% of saturation after 25 min while the vapor concentration of Hg ° in blanks without cement achieved only about 3.5% of saturation over the same time period. The samples that contained Hg°(1) also evolved significant Hg vapor [Fig. 3(b)], but the headspace never reached saturation, even after considerable time. By comparison, a bead of Hg°(1) (same amount of Hg as was added to the solidified/ stabilized samples, but without water or cement) saturated the headspace of the 1-1 vessel in less than one hour [Fig. 3(b)]. The effect of temperature on saturation vapor pressure is described by the Clausius-Clapeyron equation: 19
din(p) AnvapdT dT
R
T2
(2)
Where p is the vapor pressure, T is the temperature, R is the gas constant, and AHwp is the enthalpy of vaporization. Assuming the quotient AHv~p/R is a constant (valid for environmental conditions) the integrated form of the Clausius-Clapeyron equation is the Antoine equation: 2° B In(p) = A - -~ (3) where p (atm) and T(K) are as defined before, and A and B are element-specific constants. For Hg, A= 12.044 and B=7447.4 (adapted from Ref. 20). Cement hydrolysis is an exothermic process, and temperatures of curing cement can increase up to 40°C. 14 This increased temperature will result in a higher Hg vapor pressure. Figure 3(a) illustrates
this effect. While the Hg vapor concentration appears to exceed the saturation vapor pressure of Hg ° at the initial ambient temperature (Hg°sat@296 = 16.62 mgm-3), the increase in temperature of 2°C due to cement hydrolysis explains this deviation (Hg°sat@298 = 19.54mgm-3). Due to the complex system environment and the variety of physicochemical processes affecting Hg in solidified/stabilized samples, a number of competitive and parallel reactions may occur (Fig. 4). These reactions are defined as: 1. Dissolution/precipitation of oxide and sulfide minerals; 2. Aqueous hydrolysis [Hg(OH)2%-~Hg(OH)3-]; 3. Oxidation/reduction [Hg(II) ~ Hg(O)]; 4. Volatilization [Hg°(aq) ~ Hg°(g)]; or 5. Condensation [Hg°(aq) ~ Hg°(1)]. Aqueous hydrolysis reactions (rxn 2) for divalent metals are generally fast,2~ and the aqueous reduction reaction (rxn 3) is likely fast due to (a) the reducing conditions of the system and because (b) dissolved species are more likely to be reduced than solid species. Dissolution rates of oxide minerals (rxn 1) at a constant temperature are a function of (a) the concentration of H ÷ (pH) and other sorbing ligands which, when adsorbed, can polarize metal-oxygen bonds near the mineral surface and expedite dissolution, and (b) the concentration of surface functional groups, the discrete sites where these dissolution reactions take place, a2-24 It is likely that because of the high pH of concrete, surface functional groups will be completely deprotonated, resulting in strong polarization of metal-oxygen bonds and rapid dissolution. The rate of isothermal mass transfer (rxn 4 and 5) is a function of the rate of diffusion of the species through a mass transfer zone;25 the diffusion rate depends on, among other things, the number of mass transfer zones, since the individual effects of these are added, similar to resistances in series. Assuming the rate of Hg volatilization is due to mass transfer limitations, then the time dependent vapor phase Hg concentration can be described by the following first-order expression: 26 dHg = k(Hg~at _ Hgl) dt
(4)
where Hg~at = saturation vapor phase concentration of Hg, mgm -3 Hg I = vapor phase concentration of Hg, mg m -3, at time t t = time, min k = first order rate constant, min-% (sam ples of constant volume and cross sec tional area)
30
W. P. H A M I L T O N A N D A. R. B O W E R S
Hg°(g)
Hg~s) + H200) ~
Hg(OH)2 ° ~
Hg°(aq)
1lc Hg(OI-I)5"
Hg~)
FIGURE 4. Possible parallel and competitive reactions involving aqueous Hg. HgX(s) represents either HgO(s) or HgS(s). The integrated form of this differential equation relates the concentration of Hg ° in the vapor phase as a function of time: Hg~ = Hgsat°(1 - e -kt)
(5)
The data for the solidification/stabilization of HgO and Hg°(l) was fit to this equation using a nonlinear least-squares program. 27 The values of Hg~at and k derived from this fit for each experiment are summarized in Table 1, and values of Hg~at predicted with this model are compared to Hg~at values calculated with the Antoine equation [eqn (3)] to evaluate the assumption that mass transfer is the mechanism controlling Hg volatility. The values of Hg°sat for the bead of Hg°(l) as well as the solidified/stabilized HgO sample indicates that mass transfer limitations across the water-air interface are controlling the volatilization of Hg vapor. The values of Hg~at predicted with the first-order model are similar to the values predicted by the Antoine equation, particularly considering the other possible mechanisms that may be operating on the system. While the data for the solidification/stabilization of Hg°(l) resulted in a good fit, the value of Hg~at predicted with the least-squares fit of eqn (5) deviated from the value of Hg~at determined with the Antoine equation because insufficient data was collected (Hg vapor concentration did not approach saturation during the experiment); however, an order-of-magnitude approximation of the rate constant can be made by assuming that as t approaches infinity, the vapor pressure of Hg approaches saturation. Including the value of Hg~at (16.62mgm -3) at t > 1900rain as a point to the fit of eqn (5) results in a new approxi-
mation of k (k -- 0.0282 min-l), without sacrificing goodness-of-fit [Fig. 3(b)]. The volatilization of Hg vapor is controlled by the driving force, i.e. concentration gradient, between phases. 25'26 The evolution of Hg vapor from solidified/stabilized HgO is faster than the evolution rate from solidified/stabilized Hg°(l) due to the dissolution rates of the two species. Both the dissolution of HgO and subsequent aqueous reduction of Hg(II)(aq) to Hg°(aq) are assumed fast, resulting in a nearly instantaneous saturated aqueous Hg°(aq) concentration, i.e. the concentration of Hg°(aq) approaches its solubility (60.78/Lg/l), 2s and maximum driving force. The rate constants indicate that additional limitations on the volatility of solidified/stabilized Hg°(l) may be controlling the overall rate of Hg vapor evolution. As a non-aqueous phase liquid (NAPL), an interface develops between Hg°(1) and water. It is likely that during hydrolysis, this interface becomes increasingly complex due to the formation of a crystalline Ca(OH)2 sheath around any Hg°(l) droplets in the matrix, consistent with the gel model of cement hydration, ls,29 This sheath forms an additional mass transfer interface and increases the overall transfer resistance, resulting in limited Hg°(1) dissolution and Hg vapor evolution. HgS is one of the most insoluble of the metal sulfides 17 (Ksp = 10-52), and, as shown in Fig. 2, is the thermodynamically stable species under reducing conditions. However, even under oxidizing conditions, HgS is very resistant to weathering, 3° i.e., the oxidation reaction: HgS(s) + 202(aq) 4- 2H20 Hg(OH)2(aq) 4- SO 2- 4- 2H +
(6)
is very slow, with a pseudo first-order half-life of ~180 years. 31 The low solubility and inert redox character of HgS will prohibit a significant concentration gradient from developing, resulting in a particularly stable and innocuous Hg phase, at least over the time frame addressed by these experiments (< 3h). Figure 5 illustrates the proposed conceptual model. The dissolution of the oxide is fast resulting in a relatively high concentration of Hg(II)(aq), which
TABLE 1 Summary of Rate Data for the Evolution of Hg ° from SoHditled/StabiUzed cement Hgsat
HgO(s) solidified/stabilized Hg°(l) untreated Hg°(1) solidified/stabilized Corrected*
[eqn (3)] (mgm-3)
Hg~at
[eqn (5)] (mgm-3)
19.54 16.62 19.54 19.54
"Kinetic data based on a model fit [eqn (5)] including H g t
16.75 16.56 5.756 19.54 ° =
19.54mg m-3 [eqn (3)].
k [eqn (5)] (min-l)
R2
0.153 0.0686 0.0131 0.00230
0.967 0.999 0.982 0.974
31
DETERMINATION OF ACUTE Hg EMISSIONS a
HgO(s) + H20(1) + xOH" ~
Hg(OH)X-2+x~
Hg°(aq) ~
Hg vapor
rate limiting
Hg°(l
Hg°ca(OH)2
Hg°'a )~'J ~ q/.'-~
Hg vapor
rate l i m i t i n /
HgS(s) ~
Hg 2+ + S2"
AND HgS(s) + O2(aq) + 2H20(1) ~
Hg(OH)2(aq) +SO42" +21-1+
FIGURE 5. Mechanisms controlling the transport of Hg vapor in solidified/stabilized monoliths. Hg°ca(oH)2represents Hg diffusing through the Hg°(I)-Ca(OH)2(s)interface. Pairs of curved arrows signifymass transfer processes. due to the reducing conditions in the matrix, is quickly reduced to Hg°(aq). The subsequent evolution of Hg vapor is limited only by the rate of mass transfer across the air-water interface. In the elemental form, Hg°(l) is subject to additional mass transfer limitations at the Hg°(1)-water interface that are compounded by the precipitation of Ca(OH)2 along this interface. These mass transfer limitations impede the rate of dissolution of Hg°(1) resulting in a lower Hg°(aq) concentration with respect to time, a smaller concentration gradient between the water and the air, and thus a slower evolution rate of Hg vapor from the monolith. In the case of HgS, the Hg(II)(aq) concentration is controlled both by the low solubility of HgS and the exceptionally slow oxidation of HgS. The extremely low dissolved Hg(II)(aq) concentration results in an insignificant driving force, and therefore little Hg vapor is evolved. While no studies to date have focused on losses of Hg vapor from solidified/stabilized cement wastes, Lindberg and Turner 32 observed volatile Hg ° losses from Hg-containing sludges associated with a chloralkali facility. Measurements as well as predictions from a Gaussian diffusion model indicated that, depending on temperature, steady state Hg vapor concentrations in ambient air could be significant (approaching 1/zgm-3). In addition, attempts to solidify/stabilize wastes containing volatile organic compounds (VOCs) have experienced volatile losses. Weitzman e t al. 3 observed losses of up to 90% during mixing and curing of VOC wastes. Studies conducted by Spence e t al. 33 experienced volatile losses of up to 50%, in spite of the fact that their experiments were
designed to minimize VOC losses. While this study did not focus on the fraction of total Hg lost from the solid, it can be said qualitatively that the losses of Hg (as a fraction of initial mass) during the initial stages of solidification/stabilization were considerably less than the losses of VOCs from similar solids; this may be due to the lower saturation vapor pressure of Hg vapor relative to VOCs, and the fact that experiments were conducted in closed, batch reactors. Experiments in open systems would likely provide an environment that is conducive to greater Hg losses. The conclusions of this study are consistent with the findings of these VOC studies, namely that volatile compounds are lost from solidified/stabilized waste forms. The findings of this study may be important to those making disposal decisions for Hg-contaminated waste. Hg vapor is a neurotoxin, 34 and chronic exposure to Hg vapor has caused effects ranging from micromercurialism (weakness, fatigue, weight loss, gastrointestinal dysfunction) to mercurial erethism (severe behavioral and personality changes, memory loss, insomnia, depression, delirium) and mercurial tremors. If Hg losses are significant in a toxicological sense, worker exposure should be minimized and steps should be taken to reduce the potential for Hg volatilization. While these experiments only examine losses during the initial phase of solidification/stabilization, the results may point to an additional, chronic problem. Further investigation into losses of Hg from solidified/stabilized waste over longer time periods, as well as investigations into the total fraction of Hg lost,
32
s h o u l d be c o n d u c t e d to d e t e r m i n e a p p r o p r i a t e disp o s a l s c e n a r i o s for H g - c o n t a m i n a t e d wastes,
CONCLUSIONS
Mercury has always provided unique challenges due to its c o m p l e x c h e m i s t r y , a n d this i n c l u d e s the treatm e n t o f H g - c o n t a i n i n g wastes. B e c a u s e o f its a b i l i t y to volatilize, the s p e c i a t i o n o f H g is a n i m p o r t a n t f a c t o r w h e n c o n s i d e r i n g d i s p o s a l o p t i o n s . T h e solidif i c a t i o n / s t a b i l i z a t i o n o f H g as e i t h e r H g O o r Hg°(1) m a y lead to the e v o l u t i o n o f H g v a p o r f r o m solidified/stabilized m o n o l i t h s u n d e r t y p i c a l c o n d i t i o n s o f E / - / a n d p H . C o n v e r s e l y , the s o l i d i f i c a t i o n o f H g as H g S s h o w e d n o p r o p e n s i t y to volatilize, likely d u e to its low s o l u b i l i t y a n d i n s e n s i t i v i t y to r e d o x c o n d i t i o n s o v e r the t i m e scale c o n s i d e r e d . T h e s e results m a y be important when considering treatment and disposal o p t i o n s for H g - c o n t a i n i n g wastes.
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