Phsicochemical Treatment of Hazardous Wastes - Chapter 13 doc

oxi-13.2 Fundamental Theory The reactions in zero-valent iron are heterogeneous due to the strong dence of the reaction rate on the surface area of the iron Burris et al., 1995.The surfa

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so their work was overlooked by the research community (Gillham andO’Hannesin, 1994) In the late 1980s, Reynolds observed that organics dis-appeared from the iron pipes used in his study on the corrosion of PVC andiron pipes by water contaminated with organics Several years later, Gillhamrealized the potential of using the reduction ability of zero-valent iron forpractical purposes and holds several patents for the application of zero-valent iron degradation of organic compounds (Wilson, 1995)

Pump-and-treat and impermeable confinement are frequently used todegrade halogenated and nonhalogenated hydrocarbons in groundwater;however, these remediation techniques have limitations For example,pump-and-treat only transfers contaminants to another media such as airstripping or activated carbon In addition, the discharge of large volumes ofwater and the production of secondary waste may be costly Also, thehydraulic characteristics of the aquifer may be adversely affected Permitsare required for discharge, and groundwater rights have to be purchased forthe disposal of large volumes of treated groundwater, which may result inexcessive operating costs (Cantrell and Kaplan, 1997) The use of a commonalternative, biological degradation, has increased for remediation, but gain-ing an understanding of the biochemical pathways and associated by-prod-ucts involved, as well as developing effective strains of bacteria andmanaging the population of bacteria, can be difficult and the process has notyet been well defined (Gillham and O’Hannesin, 1994)

Zero-valent iron is a promising in situ remediation technology for thedegradation of many common pollutants, as it is comparatively inexpensive,does not restrict land use, and requires no energy for operating Zero-valentiron has been successfully utilized to destroy trichloroethenes, chromate,chlorinated organics, and mixed wastes It is capable of reducing and

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492 Physicochemical Treatment of Hazardous Wastes

dehalogenating a wide variety of halogenated hydrocarbons over wide centration ranges (Hardy and Gillham, 1996) In addition, iron is nontoxicand inexpensive (Gillham and O’Hannesin, 1994) It is easily oxidized byorganic compounds, thereby reducing the contaminant without an addi-tional reactant In addition to being highly reductive to many halogenatedhydrocarbons, zero-valent iron can reduce highly mobile oxycations (UO22+)and oxyanions (CrO42–, MoO42–, TcO4) into insoluble forms (Cantrell andKaplan, 1997)

con-This chapter presents the theory and application of zero-valent iron andincludes the relevant in situ chemical/physical processes To illustrate these

in situ technologies, the basic mechanisms of adsorption reduction and dation processes are discussed for in situ treatment of (1) organic pollutants,(2) heavy metals, and (3) mixtures of organic and inorganic pollutants Thehistory of zero-valent iron, current applications, mechanisms and kinetics ofthe system, system improvements, and advantages and disadvantages forzero-valent iron are also discussed

oxi-13.2 Fundamental Theory

The reactions in zero-valent iron are heterogeneous due to the strong dence of the reaction rate on the surface area of the iron (Burris et al., 1995).The surface reaction proceeds in four steps First, the reactant undergoesmass transport from the groundwater to the iron surface (Matheson andTratnyek, 1994) Second, the contaminant is absorbed onto the surface of theiron, where the chemical reaction occurs Third, the reaction products desorbfrom the surface, which allows the site to become available for anotherreaction (Burris et al., 1996a) Finally, the products of the reaction return tothe groundwater Rate limitation could occur at any step Where it may not

depen-be the sole limitation, mass transport plays an essential role in the kinetics

of dechlorination (Matheson and Tratnyek, 1994)

Essentially, reduction of hazardous wastes by zero-valent iron is due tothe beneficial corrosion of iron This process takes advantage of the chemicalreaction that occurs when iron is oxidized The contaminant is the oxidant(Fairweather, 1996), while zero-valent iron is a strong reductant capable ofdehalogenating several halogenated hydrocarbons (Kaplan et al., 1996).Commercial-grade iron and industrial scrap iron are sufficient to reducechlorinated solvents (Matheson and Tratnyek, 1994) Although iron is actu-ally consumed during the reaction, it remains effective for a long period oftime For example, 1 kg of iron can dechlorinate chloromethane at a concen-tration of 1 mg/L and sufficiently treat 0.5 million liters of water (Gillhamand O’Hannesin, 1994)

The reductive reaction is slow under anaerobic conditions, because ironmay be oxidized by oxygen Chlorinated contaminants possess an oxidizing

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Zero-Valent Iron 493

potential similar to that for oxygen (Tratnyek, 1996) When strong oxidizingcompounds such as chlorinated contaminants are not present, the iron isspontaneously corroded in water (Matheson and Tratnyek, 1994)

Fe0 + 2H2O Fe2+ + H2 + 2OH– (anaerobic corrosion) (13.2)

Fe0 + 2H2O + O2 2Fe2+ + 4OH– (aerobic conditions) (13.3)The redox pair formed from oxidizing the zero-valent iron has a reductionpotential of –0.440 V; therefore, zero-valent iron can reduce hydrogen ions,carbonate, sulfate, nitrate, and oxygen, in addition to alkyl halides (Mathesonand Tratnyek, 1994) Both Equation (13.2) and Equation (13.3) cause the pH

to increase

Zero-valent iron and organic substrate can react with a net result of ironoxidation and reduction of the substrate In such a reaction, iron acts as areducing agent:

The Pourbaix diagram shown in Figure 13.1 illustrates the thermodynamicstability of iron species in aqueous solutions of a few organic substrates Therelative position of each substrate shows that the reaction between iron andthe corresponding organic is thermodynamically favorable Three elemen-tary reactions involved in the reductive dechlorination of organic com-pounds are shown in Figure 13.2

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reaction is a heterogeneous reaction The kinetics of heterogeneous reactionsinvolving Fe0 is determined by the following factors:

• The reaction rate constants from Equation (13.1) to Equation (13.3)

• Physical processes on the surface of the catalyst (reducing agent),including its properties

• Mass transfer limitations and diffusion effects

• Sorption/desorption processes involving the substrate and ability of active reaction sites on the iron surface

avail-• Fluid flow characteristics, including velocity, flow regime

The mass transfer limitations have been shown to be less significant inregard to the kinetics of chlorinated aliphatics based on relatively slow rates

of degradation (Scherer et al., 2000) On the other hand, nitroaromatics andazo dyes have higher reduction rates under which the diffusion and masstransfer effects may become reaction rate-limiting factors (Agrawal and Trat-nyek, 1996) Scherer et al (2000) identified three steps that may imposelimitations on the reduction rates The formation of precursor complex onactive metal sites can be rate limited by the number of reaction sites Burris

et al (1996) suggested that the hydrophobicity of the contaminant maysignificantly affect the sorption rate of substrates Scherer et al (2000) illus-

FIGURE 13.1

Eh–pH diagram (or Pourbaix diagram) showing equilibria with water, iron, and common environmental contaminants including perchloroethene (PCE), nitrobenzene (ArNO 2 ), and chromate (Cr[VI]) Hematite ( a -Fe 2 O 3 ) and magnetite (Fe 3 O 4 ) are assumed to be the controlling phases for iron speciation The stability lines for the reduction of nitrobenzene (ArNO 2 ) to (ArNH 2 ), Cr(VI) to Cr(III), and PCE to TCE are superimposed to show the instability of Fe 0 in the presence of these contaminants (From Scherer, M.M et al., CRC Crit Rev Environ Sci Technol., 30(3), 363–411, 2000 With permission.)

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trated that the transfer of electrons from the surface of the reducing agent

to the substrate could affect the rates by the three major mechanisms shown

in Figure 13.3:

• Electron transfer from bare iron metal exposed by pitting of the oxidelayer, while the pitting mechanism involves localized corrosion andpossible catalytic dissolution pathways

• Electron transfer from conduction bands in the oxide layer

• Electron transfer from adsorbed or lattice Fe(II) surface area, ing reduction of a sorbed or lattice surface site

express-FIGURE 13.2

Scheme showing proposed pathways for reductive dehalogenation in Fe 0 –H 2 O systems: (A) direct electron transfer from iron metal at the metal surface; (B) reduction by Fe 2+ , which results from corrosion of metal; (C) catalyzed hydrogenolysis by the H 2 that is formed by reduction

of H 2 O during anaerobic corrosion Stoichiometries are shown (From Matheson, L.J and nyek, P.G., Environ Sci Technol., 28, 2045–2053, 1994 With permission.)

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The physical properties of the iron metal are, therefore, important factorsassociated with mass transfer limitations of the electron transfer processesand reaction-limiting steps

13.2.3 Adsorption

As a contaminant moves through soil and groundwater, chemical processeswill affect both contaminant concentration and overall hydrogeochemistry(Schoonen, 1998) of the system Different adsorption mechanisms cause pol-lutants to adsorb onto the soil, volatilize, precipitate, and be part of theoxidation–reduction processes Adsorption is loosely described as a process

in which chemicals partition from a solution phase into or onto the surfaces

of solid-phase materials Adsorption at particle surfaces tends to retard taminant movement in soil and groundwater

con-FIGURE 13.3

Conceptual models of electron transfer (ET) mechanisms at Fe 0 –oxide–water interface: (A) ET from bare iron metal exposed by pitting of the oxide layer; (B) ET from conduction bands in the oxide layer; (C) ET from adsorbed or lattice Fe(II) surface sites (From Scherer, M.M et al.,

CRC Crit Rev Environ Sci Technol., 30(3), 363–411, 2000 With permission.)

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Sorption isotherm curves are graphical relationships showing the tioning between solid and liquid form where mass adsorbed per unit mass

parti-of dry solids (S) is plotted against the concentration (C) of the constituent

in solution K is the sorption equilibrium constant; N is a constant describingthe intensity of sorption The linear sorption isotherm can be expressed asfollows:

TABLE 13.1

Sorption Mechanisms in Soils

Hydrophobic expulsion Partitioning Nonpolar organics (e.g., PCBs, PAHs) Electrostatic attraction Outer-sphere

Nonspecific Physisorption Physical Ion exchange

Some anions (e.g., NO 3– ) Alkali and alkaline earth metals (Ba 2+ ,

Ca 2+ )

Complexation reaction Inner sphere

Specific Chemisorption Chemical Ligand exchange

Transition metals (e.g., Cu 2+ , Pb 2+ , CrO 42–)

Source: Adapted from Scherer, M.M et al., CRC Crit Rev Environ Sci Technol., 30(3), 363–411,

2000

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The Freundlich isotherm can be described as:

The Langmuir isotherm is used to describe single-layer adsorption based onthe concept that a solid surface possesses a finite number of sorption sites.When these active sites are filled, the site will no longer sorb solute from thesolution:

where a = absorption constant related to the binding energy (L/mg), and

Gmax = maximum amount that can be adsorbed by the solid (mg/kg).Because adsorption isotherms are equilibrium equations, the rate at whichthe material is adsorbed has to be studied in terms of chemical affinities,

pH, solubility, hydrophobicity, and many other physical and chemical acteristics

char-Because organic nonpolar compounds have stronger attraction to organicmatter than to mineral content, the amount of adsorption of an organiccontaminant is more dependent on the organic content of the soil Theadsorption partition coefficient is generally used to determine this adsorp-tion amount, as it is empirically related to the organic fraction of the soil(f oc), and the normalized partition coefficient K occan be expressed as follows:

pH, the surface chemistry of the Fe, the specific properties of pollutants,and the associated waste matrix The retention mechanisms for pollutantsinclude adsorption of the pollutant by the Fe surfaces and precipitation.The retention of cationic metals by Fe has been correlated with Fe proper-ties such as pH, redox potential, surface area, cation exchange capacity,organic matter content, clay content, iron and manganese oxide content,and carbonate content Anion retention has also been correlated with pH,iron and manganese oxide content, and redox potential In addition to Feproperties, consideration must be given to the type of pollutants, theirconcentration, the competing ions, and the complexing ligands Transport

of metals associated with various wastes may be enhanced due to thefollowing reasons (Puls et al., 1995):

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the wide range of waste characteristics and various ways in which metalscan be adsorbed to the Fe surface, the extent of pollutant retention by a soil

is specific to the type of site, soil, and waste involved

13.2.4 Halogenated Hydrocarbons

Dechlorination is a surface reaction with the zero-valent iron serving as theelectron donor When there is a proton donor, such as water, chlorinatedcompounds will be dehalogenated The reaction kinetics depends upon themass transfer to the surface of the iron, the available surface area, and thecondition of the surface The reaction is pseudo first order, and direct contactwith the surface of the iron is required for degradation to take place (Gillhamand O’Hannesin, 1994) The basic equation for dechlorination by iron metal

is as follows:

The reduction potentials for various alkyl halides range from +0.5 to +1.5V; therefore, when Fe0 serves as an electron donor, the reaction is thermo-dynamically favorable Because three reductants are present in the treatmentsystem (Fe0, H2, and Fe2+), three possible pathways exist Equation (13.9)represents the oxidation of Fe0 by reduction of a halogenated compound Inthe second pathway, the ferrous iron behaves as a reductant, as represented

in Equation (13.10) This reaction is relatively slow because the ability toreduce a pollutant by ferrous iron is dependent on the speciation ferrousions, which is determined by the ligands present in the system The thirdpossible pathway, Equation (13.11), is dehalogenation by hydrogen Thisreaction does not occur easily without a catalyst In addition, if hydrogenlevels become too high, corrosion is inhibited (Matheson and Tratnyek, 1994):

If all three pathways are not possible, then reactions will be limited tional limitations may occur if the reaction is aerobic because Fe3+ could beproduced by further oxidation of Fe2+ and cause precipitation of iron oxides(Helland et al., 1995) The end products such as ferrous chloride and ferricoxide are not capable of reducing chlorinated compounds (Gillham andO’Hannesin, 1994) Burris et al (1995) state that Fe2+ and H2 do not have aneffect on degradation Hardy and Gillham (1996) have suggested the possi-bility that degradation may be due to a catalytic reaction utilizing hydrogenproduced from the reduction of water In developing and improving theperformance of the zero-valent iron technique in the field, detail mechanismsare important and critical for a specific site The dominant process is the

Addi-¨ Ææ

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oxidation of the iron metal (Matheson and Tratnyek, 1994) The reduction of

the halogenated hydrocarbon involves the transfer of an electron and occurs

at the iron surface A carbon-centered radical, R∑, is formed as:

Then, a second electron transfer occurs, and the radical is protonated as in

Equation (13.13)

The result is dehalogenation, and during this step Fe2+ is formed by

dis-solving into solution Hydrocarbon formation also adds perplexity to

mech-anism determination Because C1 to C5 hydrocarbons have been found in

iron and water systems, the hydrocarbons formed may be due to reduction

of aqueous CO2 by zero-valent iron Therefore, iron behaves as both a

reac-tant, by corroding and supplying electrons, and as a catalyst, by promoting

the formation of hydrocarbons Ten hydrocarbons were identified up to C5;

also, iron pretreated with H2 formed hydrocarbons with longer chain lengths

The hydrocarbon concentration and the growth in chain length increase with

time; therefore, if the hydrocarbons are not desorbed, the production of the

hydrocarbons could be rate limiting in the dechlorination of chlorinated

organics Although this does not determine the mechanism by which iron

metal removes halogens from halogenated organics, it does show that the

product distribution, carbon mass balance, and reaction rate may be affected

by catalysts (Hardy and Gillham, 1996)

For modeling purposes, zones are constructed in the column tests

per-formed A subsurface barrier will have three zones: upgradient (area before

the barrier), iron-bearing gradient (the barrier itself), and downgradient (area

after the barrier) In the upgradient zone, the Fe0 is oxidized to Fe2+ and Fe3+

by O2 The reaction increases the pH, and precipitation of the iron oxides is

initiated The precipitation could reduce the permeability of the barrier and

surface area of the iron, thus reducing the reaction rate This also causes the

iron-bearing zone and down gradient zone to be anoxic (Tratnyek, 1996)

This production of precipitates has been observed regularly in the laboratory,

but the degree to which the effect occurs in the field is under much debate

For these reasons, buffers such as pyrite, which lowers the pH, are used

The reactivity of the iron with the contaminants determines the feasibility

and the design of the site Many factors contribute to the reaction rate For

example, the oxide layer that forms on the iron affects the surface of the iron

and inhibits further corrosion, so this layer needs to be minimized

Fortu-nately, iron possesses a “porous and incoherent” nature with oxide film, and

iron usually exhibits satisfactory degradation rates over a long period of

time (Tratnyek, 1996)

Because reaction rates vary widely among contaminants, Fe barrier

design should be dependent on the least reactive contaminant Some

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contaminants are quickly reduced, and others react too slowly for

consid-eration of remediation by iron Although much data exist regarding

deg-radation rates, quantitative predictions concerning the reaction rates

cannot be made because many of the experiments are highly inconsistent

in accounting for the effect of pH, surface area of the iron, mixing rate,

and other experimental variables Comparisons are of better quality if the

rate constants are normalized For example, when rate constants are

nor-malized to the surface area of iron (k SA), the specific rate constant varies

by only one order of magnitude for individual halocarbons Upon further

correlation, the k SA has indicated that dechlorination is more rapid for

saturated carbon than unsaturated carbons, and reaction rates are faster

for perhalogenated compounds By obtaining this quality information, the

amount of iron required to obtain a decrease in contaminant concentration

by a magnitude of three can be calculated (Tratnyek, 1996) Also,

quanti-tative structure–activity relationships (QSARs) could be developed for such

prediction of iron required for different organic pollutants

13.3 Degradation of Hazardous Wastes

13.3.1 Organic Pollutants

Table 13.2 presents organic classes that have been successfully degraded by

zero-valent iron

Trichloroethylene (TCE) and perchloroethylene (PCE) require cleavage of the

carbon–halogen bonds Two methods of cleavage are b-elimination by

dehy-drohalogenation, as shown in Equation (13.14), and nucleophilic substitution

by either water or hydrogenolysis in Equation (13.15) The proposed

path-ways for reduction of chloroethylenes by zero-valent iron are as follows:

The reduction of the triple bond may form an olefin (Equation 13.15) or an

alkane (Equation 13.16) (Burris et al., 1995):

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The overall reaction as shown in Equation (13.17) for the degradation of the

TCE produces ethene and ethane at a ratio of 2:1:

C2HCl3 + 3H+ + 6e–Æ C2H4 + 3Cl– (13.18)

Ethene and ethane account for 80% of the mass of the hydrocarbons

identified as products Trace amounts of methane and acetylene are also

produced (Orth and Gillham, 1996) The reduction of PCE forms cis

-1,2-dichloroethylene (DCE), trans-1,2-DCE, 1,1-DCE, vinyl chloride, ethylene,

dichloroacetylene, acetylene, ethene, ethane, chloroacetylene, methane,

and several alkenes ranging from C3 to C6 The trace amounts of

dichloro-ethylene and vinyl chloride formed during the reduction of PCE and TCE

are further reduced (Burris et al., 1995) Reaction rates vary with substrate,

chemical, and microbiological conditions Selected t1/2 values are provided

1,1,1-Trichloroethane 1,1,2-Trichloroethane 1,1-Dichloroethane Ethenes Tetrachloroethene

Trichloroethene

cis-1,2-Dichloroethene trans-1,2-Dichloroethene

1,1-Dichloroethene Vinyl chloride Propanes 1,2,3-Trichloropropane

1,2-Dichloropropane Aromatics Benzene

Toluene Ethylbenzene Other Hexachlorobutadiene

1,2-Dibromoethane Freon 113

N-nitrosodimethylamine Source: USEPA, EPA/600/R-98/125, U.S

Environmental Protection Agency, Washington, D.C., 1998.

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Helland et al (1995) compared the degradation rates of halogenated aliphaticcompounds by zero-valent iron in both anoxic and oxic reductions Thereactions are as follows:

was pseudo first-order and the reaction constant, k, decreased with each

additional dehalogenation step (Gillham and O’Hannesin, 1994)

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chlo-by the zero-valent iron process

Saturated hydrocarbons such as 1,2-dibromo-3-chloropropane (DBCP)have been used as soil nematicides DBCP has contaminated the ground-water due to its extensive use Although banned in 1977, concentrations

in some parts of the United States exceed the maximum contaminantlevel of 0.2 mg/L The dehalogenation of DBCP has two possible path-ways: (1) a series of three hydrodehalogenation reactions requiring aproton and two electrons for each reaction or (2) the formation of atransitional state of propene followed by hydrogenation of the doublebond, with the product being propane in both mechanisms The onlyintermediate formed is propene, with a reaction rate constant of 0.28 ±0.03 min–1 (Siantar et al., 1996) Table 13.5 shows the effect of pH on thepseudo first-order rate constants and half-lives of DBCP with or withoutmixing Table 13.6 indicates that sulfate and nitrite ions do not signifi-cantly affect the pseudo first-order rate constants and half-lives of DBCP

Degradation of polychlorobiphenyls (PCBs) by zero-valent iron requires peratures of 400°C (Grittini et al., 1995) At 400°C, PCBs are reduced tobiphenyls For further degradation of the biphenyl compound, temperatureshave to exceed 500°C (Chuang et al., 1995) Under normal temperatures,zero-valent iron has little effect on PCBs

Source: Gillham, R.W and O’Hannesin, S.F., Ground Water, 32(6), 958–967, 1994 With permission.

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Nitroaromatic compounds (NACs) are one of the widespread contaminants

in the environments Sources of NACs are numerous; they originate frominsecticides, herbicides, explosives, pharmaceuticals, feedstock, and chemi-cals for dyes (Agrawal and Tratnyek, 1996) Under anaerobic conditions, thedominant action is nitro reduction by zero-valent iron to the amine Otherpathways do exist, such as the formation of azo and azoxy compounds,which is followed by the reduction of azo compounds to form amines Also,

in addition to the possibility of azo and azoxy compounds, ylamine may be an additional intermediate (Agrawal and Tratnyek, 1996).Nitrobenzene reduction forms the amine aniline Known for its corrosioninhibition properties, aniline cannot be further reduced by iron Additionally,

phenylhydrox-it interferes wphenylhydrox-ith the mass transport of the contaminant to the surface of theiron The overall reaction is as follows:

ArNO2 + 3Fe0 + 6H+ Æ ArNH2 + 3Fe2+ + 2H2O (13.25)

TABLE 13.5

Effect of Ions on Pseudo First-Order Rate Constants and Half-Lives of DBCP

Transformation Using 0.1-M HEPES Buffer Solution at pH 7, Mixing at 400 rpm,

and 22°C

Aqueous Phase pH

k (hr–1 ) (No Shaking)

k (hr–1 ) (Mild Shaking)

t1/2 (hr) (No Shaking)

t1/2 (hr) (Mild Shaking)

Concentration k (min–1 ) t1/2 (min)

No sulfate or nitrite added 0.300 2.31

Source: Siantar, D.P et al., Water Res., 30(10), 2315–2322, 1996 With permission.

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Zero-Valent Iron 507The elementary steps are shown in Equation (13.26) to Equation (13.28):

During the reduction of NACs, changes in corrosion and precipitation ofiron products were not significant As long as the diffusivities of the differentsubstitutions on the benzene were similar, the reaction rate did not varysignificantly (Agrawal and Tratnyek, 1996) The rate constants are given inTable 13.7

Because nitrobenzene degradation is faster in batch experimental systemsthan in column studies, a mass-transfer limitation exists; therefore, when

determining the effectiveness of in situ groundwater treatment systems,

hydraulics, mass transfer, and reaction kinetics should be taken into eration (Burris et al., 1996)

consid-Weber (1996) performed additional experimental studies on zobenzene (4-AAB) The compound was chosen based on the hypothesisthat azo compounds capable of reduction by zero-valent iron and an aminogroup would allow for determination of surface-reaction occurrence Thecompound 4-AAB reduces and forms aniline The reduction of 4-AAB wasperformed without cleaning the iron with hydrochloric acid, which is astandard method in most experiments Complete loss of 4-AAB occurred in

4-aminoa-2 hr When the iron was washed with hydrochloric acid, the reaction occurred

so quickly that aniline formation could not be measured An additionalexperiment with bound 4-AAB was conducted to determine if the reaction

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occurred at the iron surface Because no degradation of 4-AAB occurred, thereaction appears to be surface mediated

When NACs are reduced, aromatic amines are formed and are of cant toxicological concern Zero-valent iron cannot accomplish this degra-dation; therefore, remediation has to extend beyond the reduction of theparent NAC (Agrawal and Tratnyek, 1996) Two methods of further reduc-tion are biodegradation and enzyme-catalyzed coupling reactions Biodeg-radation can degrade the NACs further to mineralization, and the reaction

signifi-is more rapid for aromatic amines than for the biodegradation of the originalNAC compound (Burris et al., 1996b); however, biodegradation of NACscan be difficult and can produce toxic metabolites; therefore, further reduc-tion of the NACs can also be accomplished by enzyme-catalyzed couplingreactions, which integrates the amine into organic matter (Monsef et al.,1997) Both methods could be combined with nitro reduction by Fe0 to treatNAC contamination (Agrawal and Tratnyek, 1996)

The presence of nitrates and nitrites in groundwater is of growing concern

as it poses serious health risks The typical sources of nitrate are nitrogenfertilizers, septic tanks, and animal wastes Excessive nitrates present ingroundwater can be reduced by microorganisms to nitrite, which is harmful

to human health and animals, agriculture, and the environment Nitrites andsecondary amines can react to form nitroamines, which cause serious healtheffects, including cancer Such adverse effects with increasing nitrate con-tamination in groundwater draw attention to treatment and removal ofnitrate and nitrite contamination in groundwater Physicochemical removal

by ion exchange resins does not destroy nitrates, frequent resin regeneration

by salt produces much brine, and biological degradation by denitrifyingmicroorganisms is often slow and incomplete The process requires expen-sive maintenance due to production of excessive biomass

The feasibility of chemical degradation of nitrate by reducing agents hasbeen investigated for application in treatment of contaminated groundwater

by Horold et al (1993a,b) The transformation of aqueous nitrate into benignproducts was studied using Fe0 in a pH-buffered anaerobic aqueous mediumfrom groundwater Their results showed that 75 to 85% of the nitrate wasreduced to ammonia at room temperature within an hour in an acidic pH

of 3 The chemical denitrification by organic and inorganic reductants andcatalysts showed a 55% reduction in nitrate within 48 hr with Fe0 powder

in an anaerobic medium at 85°C A rapid catalytic reduction of nitrite withhydrogen gas using Pd metal supported on alumina and of nitrate withhydrogen gas using a Cu–Pd bimetallic catalyst supported on alumina havebeen demonstrated This technique seemed to be efficient at the temperature

and pH ranges of natural groundwater and was recommended for an ex situ

treatment of groundwater

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To gain insight into kinetics, reaction pathways, and reaction end products,laboratory investigations were performed by Rahman and Agrawal (1997).Sodium nitrate and sodium nitrite were selected as model pollutants Reac-tions were carried out at room temperature in the dark with untreated [Fe0]

= 69.4 g/L In some cases, Fe0 was treated with 10% HCl (v/v) for nearly 2min and then washed with deionized water four to six times prior to reaction

The nitrate and nitrite stock solution was nearly 0.16 mM, and mixing was

achieved at 40 rpm

nitrite as an intermediate and ammonia as final product Iron acts as anelectron donor and the reduction is coupled with metal corrosion (Equation13.9) The reduction reaction in the model system was found to proceed intwo sequential steps (Equation 13.30 and Equation 13.31), and the overallreaction was represented as follows:

in its concentration with concurrent decreases in nitrate concentrationthroughout the experiment suggested reduction of nitrate (Equation 13.30)

by zero-valent iron (Rahman et al., 1997) The transformation reaction ofnitrate to nitrite was found to be first-order in substrate concentration, andthe reaction rate constant was obtained These reduction experiments wereperformed with untreated as well as acid-treated Fe0 turnings Experimentswere also performed to investigate if only nitrite was reduced with Fe0 metalunder identical conditions Results showed that nitrite can also be reduced

transformation reaction (Equation 13.31), similar to nitrate reduction, wasfound to be first order in substrate concentration and, unlike nitrate, no by-products of nitrite reduction were evident in ion chromatography This sug-gested that either reduced products (N2 gas) were lost in gaseous forms orthey may be present in solution as NH4+ cation The [NH4+] was estimated

in a batch system after several days as the final reduction product andindicated a mass balance of nearly 80% The hypothesized reaction (Equation13.35) consisted of the following steps:

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Thus, it was observed that the first-order rate constants (k1) for nitratereduction by untreated Fe0 increase due to the pretreatment of iron metalwith HCl; however, observed increases in the rate constant for nitrite reduc-tion have been relatively small under similar acid pretreatment conditions.During the first 12 hr, the rate constant for nitrate reduction showed a gradualdecline, and this decline seems to have been clearly influenced by the pres-ence of chloride The reaction rate constants for nitrate and nitrite reduction

by untreated Fe0 turnings are directly dependent on the concentration of Fe0

used, ranging between 69.4 and 208.2 g/L; thus k1 and k2 increase linearlywith increases in the surface area of the untreated iron Table 13.8 demon-strates that acid-treated Fe0 is more reactive than its untreated counterpart

13.3.2 Reduction of Heavy Metals

Anions or oxyanions of arsenic, selenium, chromium, technetium, and mony are important groundwater contaminants and may occur under nat-ural groundwater conditions Indirect precipitation of inorganic cationsresults from the reduction of an anion-forming species, usually sulfate Sul-fate reduction generates hydrogen sulfide, which combines with metals toform relatively insoluble metal sulfide precipitates Many heavy metals aretreatable using this approach, including Ag, Cd, Co, Cu, Fe, Ni, Pb, and Zn,

anti-as listed in Table 13.9 (Waybrant et al., 1998) Column experiments conductedusing a range of organic substrates demonstrated the potential to remove a

TABLE 13.8

Fe 0 (g/L) Treatment

Rate Constant

Half-Life (hr) R2 Conditions Nitrate reduction

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range of dissolved metals at groundwater velocities similar to those observed

at sites of groundwater contamination A field-scale reactive barrier for thetreatment of acid mine drainage and removal of dissolved Ni was installed

in 1995 at the Nickel Rim mine site near Sudbury, Ontario It was composed

of municipal compost, leaf compost, and wood chips Monitoring of thereactive barrier indicates continued removal of the acid-generating capacity

of the groundwater flowing through the permeable reactive barrier (PRB)and decreases in dissolved Ni concentrations from up to 10 mg/L to <0.1mg/L within the PRB

The mechanism of adsorption implies attachment of the chemical toreactive sites on mineral surfaces These sites usually result from an excesseither positive or negative charge on the surfaces These surface chargescan be constant (fixed) due to ion substitutions in the mineral matrix(isomorphous substitution), variable with pH, or a mixture of both Inaddition, the adsorption can result from either inner-sphere or outer-spherecomplexation Inner-sphere complexation is due to actual covalent andionic chemical bond formation In outer-sphere complexation, adsorptionresults from ion-pair bonding due to electrostatic forces; hydration waterseparates the solvated ion from the surface Many metal oxides and someclay minerals have net surface charges that vary with pH due to theproportion of protonated vs deprotonated surface sites Among these vari-ably charged materials are the iron oxyhydroxides (rusts) that result whenzero-valent iron corrodes These materials are very significant to adsorption

of both inorganic and organic charged solution species (ionic species) Thecharges of both the surface and the solution ion control whether adsorption

TABLE 13.9

Inorganic Contaminants Treated Trace metals Chromium

Nickel Lead Uranium Technetium Iron Manganese Selenium Copper Cobalt Cadmium Zinc Anion

contaminants

Sulfate Nitrate Phosphate Arsenic

Source: Waybrant, K.R et al., Environ.

Sci Technol., 32(13), 1972–1979, 1998.

With permission.

Tiêu đề	Physicochemical Treatment of Hazardous Wastes
Thể loại	Book chapter
Năm xuất bản	2003

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Số trang	42
Dung lượng	1,16 MB