Monitoring and Modeling VOCs in Wastewater Facilities

ES&T

ADVANCES IN WATER TREATMENT TECHNOLOGIES

Monitoring and Modeling VOCs in Wastewater Facilities olatile organic compounds (VOCs) have gained prominence in air and water pollution control practice over the past decade as a result of heightened environmental and health concerns and the introduction of new regulations. VOCs are precursors of ground-level ozone, which contributes to smog formation. VOCs in wastewater treatment plant discharges may exert environmental impacts on receiving-water quality. Airborne emissions from treatment facilities may contain VOCs and affect downwind populations. Regulatory initiatives, the most significant of which is the 1990 Clean Air Act Amendments (CAAA) [1-3), have addressed these concerns through control measures designed to minimize VOC emissions. These new r e g u l a t i o n s have spurred the development of VOC fate and transport models in wastewater collection and treatment systems, the inventory of VOC emissions, the development of new VOC sampling and analytical protocols, and research into VOC emission control methods. EPA identified 189 chemicals and chemical categories (referred to as hazardous air pollutants, HAPs) for regulation. VOCs form the majority of these. Prior to the CAAA, the release of the National Emission Standards for Hazardous Air Pollutants (NESHAP) regulation for benzene forced the petroleum refining industry to make multimillion dollar investments to control benzene emissions (4). Title III of the CAAA requires

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that maximum achievable control technology (MACT) standards be developed to control HAPs in both municipal and industrial sectors. To date, EPA has proposed only one MACT standard—the hazardous organic NESHAP (HON)—to regulate the emission of 149 organic HAPs from the synthetic organic chemical manufacturing industry. This paper reviews the advances made in providing the scientific basis for addressing VOC control regulations and focuses on emissions from wastewater collection and treatment systems. Surveys of VOC emissions From 1978 to 1981, EPA, as directed by the Clean Water Act, conducted two major surveys of the removal of 129 priority pollutants in publicly owned treatment works (POTWs). Known as the 25 POTWs (5) and the 40 POTWs (6) studies, these surveys provided data on the concentration and likely occurrence of VOCs in the influents, effluents, and sludges at POTWs. Although considerable variability in influent VOC concentration was observed, only four VOCs were found at concentrations above 1 pg/L at least 50% of the time in the effluents. A survey of priority pollutant data from 105 North American POTWs from 1978 to 1986 showed that only six VOCs were found more than

HENRYK

MELCER

Brown and Caldwell Consultants Seattle, WA 98119

50% of the time in the effluents from secondary POTWs at concentrations of 1-20 pg/L (7). In the early 1980s, a survey of 16 VOC species showed that only a small fraction of nearly 600 Californian POTWs emitted significant quantities of specific VOCs to the air relative to other sources (8). In 1987, the California Air Toxics "Hot Spots" Information Act (known as AB 2588) required emitters to inventory VOC emissions. This program has generated the largest ever statewide VOC emissions database and, although data are still under review, a preliminary conclusion indicates that (contrary to the expectation that POTWs were significant sources of VOCs) the majority of the calculated human health risk in ambient air is contributed by mobile sources [9, 10). In a parallel investigation, the Joint Emission Inventory Program (a working group of 22 of the 23 major POTWs in California's South Coast Air Basin), only two of the facilities tested produced unusually high VOC levels (11). These were the only facilities that accepted petroleum refinery wastewater, although a clear relationship could not be det e r m i n e d between the refinery wastewater and VOC levels. It was also found that POTWs contributed 0.45 megagrams (Mg) per day of VOCs to the south coast atmosphere, which was less than 0.03% of basinwide emissions. Low VOC concentrations have also been reported in the treated effluents from industrial wastewater treatment plants. A 1981 survey of

0013-936X/94/0927-328A$04.50/0 © 1994 American Chemical Society

Recent comparisons of modelled predictions of VOC emissions with actual measurements at specific steps of the process, such as the high-powered surface aerators pictured above, have shown that current models require more

priority p o l l u t a n t s at five organic chemical plants with activated sludge treatment facilities s h o w e d that the high levels of influent VOCs (>1 mg/L) were reduced to low levels in the effluent, although the mechanism for such losses was not determ i n e d {12). Similar investigations conducted at five Canadian refineries between 1980 and 1985 showed that only four VOCs were found in secondary effluents more than 50% of the time at c o n c e n t r a t i o n s greater than 1 pg/L [13, 14). This confirmed the results from similar investigations at U.S. refineries (15, 16). Modeling, mechanisms of VOC emissions At the t i m e t h e a b o v e s u r v e y s w e r e b e i n g c o n d u c t e d , o t h e r researchers w e r e investigating VOC removal m e c h a n i s m s in wastewater treatment plants. Biodegradability of VOCs was demonstrated by EPA

in aerobic static culture flask studies a l t h o u g h significant losses by volatilization were noted (17). Further studies [18) s h o w e d that, at pilot scale, some VOCs were removed at POTWs p r i m a r i l y by stripping. P a r t i t i o n i n g to b i o s o l i d s w a s not found to be a significant removal m e c h a n i s m . Despite the availability of a large VOC monitoring database from three Canadian POTWs, it was not possible to construct stochastic models of VOC removal, confirming the need for a mechanistic approach to fate modeling (19). Early attempts to model VOC removal using biological oxygen dem a n d (BOD) removal kinetics met with limited success (20). It was recognized that the relatively low VOC c o n c e n t r a t i o n in w a s t e w a t e r s did not lend itself to the conventional g r o s s o r g a n i c p a r a m e t e r (BOD, chemical oxygen d e m a n d ) kinetic analysis in the determination of bio-

treatment calibration.

logical degradation coefficients for VOCs (21, 22). T h r e e major fate m e c h a n i s m s were identified: loss to the atmosphere, sorption, a n d biodegradation. M u c h of the research into fate m e c h a n i s m s has focused on the primary m e c h a n i s m : loss to the atmosphere. Loss to the atmosphere In t h e a b s e n c e of c o m p e t i n g m e c h a n i s m s , mass transfer of VOCs to the a t m o s p h e r e can result from n a t u r a l v o l a t i l i z a t i o n across o p e n w a t e r surfaces, v o l a t i l i z a t i o n ind u c e d by mechanical surface aeration, and stripping by diffused aeration. In each case, the rate of mass transfer of a VOC from wastewater to t h e a t m o s p h e r e across a n air— w a s t e w a t e r i n t e r f a c e c a n be described (23) by: dM/dt = -KL (C - Cg/HC)A

(1)

w h e r e KL is the overall mass transfer

Environ. Sci. Technol., Vol. 28, No. 7, 1994 329 A

of mixing (25, 31, 32). Empirical ex pressions to estimate Hr as a func tion of temperature have been de veloped for VOCs (33, 34). Observed values of Hr. vary over a wide range for some VOCs (35). There continues to be debate as to the validity of applying clean-water H(. values to wastewater. For wastewater treatment pro cesses that are open to the atmos phere, accumulation of VOCs in the gas phase is often negligible and mass transfer across the air—waste water interface can be modeled as: (4) dM/dt = -K,C,A This is referred to as an infinite di lution or infinite ventilation condi tion, and can be applied to volatil ization at open surfaces or surfaceaerated processes. Equation 4 is not valid for mass transfer to rising air bubbles, or from poorly ventilated covered treatment or sewage collec tion systems.

Measuring VOC emissions with a floating off gas collector at a publicly owned wastewater treatment works in Waterloo, Ontario. coefficient (m/s), C is the VOC concentration in bulk liquid (pg/L), C„ is the VOC concentration in bulÉ gas (mg/m 3 ), Hc is the dimensionless Henry's law coefficient (m3 liquid/m 3 gas), and A is the interfacial contact area between air and wastewater (m2). KL can be described by two-film theory (24), a flux-matching boundary condition, and the assumption that overall resistance to mass transfer results from resistances through two thin films (gas and liquid) adjacent to the gasliquid interface: HKL = Ilk, + ll(Hcka (2) where kt is the liquid-phase transfer coefficient (m/s) and kg the gasphase transfer coefficient (m/s). The first, second, and third terms in Equation 2 are overall, liquid, and gas-phase resistances, respectively. As kj and kg are directly proportional to molecular diffusivities, and molecular diffusivities of VOCs are generally much greater in gas than liquid, gas-phase resistance has often been assumed to be negligible for Hc > 0.1 (25). For such a condition, liquid-phase resistance to mass transfer is limiting, and KL = k,. Equation 2 also indicates that the liquid-phase resistance will increase the greater the ratio, (kg/k,), and the greater the value of Hc, that is, the more volatile the compound. It has been reported that a value of 330 A

100-150 for (kg/kj) has been used to model VOC stripping in gas—liquid contactors and that gas film resistance is likely more important than previously assumed [26). The ratio of liquid mass transfer coefficients for two VOCs i and j may be described by: (3) Ψ = ku/kn = ÇDJDJP where Ψ is the transfer coefficient ratio for compounds i and j (dimensionless); kn and k,j are liquid mass transfer coefficients for compounds i and /, respectively (m/s); and D, and Dj are liquid diffusion coeffi cients for compounds i and /, re spectively (m 2 /s). The exponent η has been shown to vary from 0.5 for penetration and surface renewal theories (27, 28) to 1.0 for the twofilm theory (24) and depends on tur bulent mixing conditions in the aqueous phase. Oxygen has been the most widely studied volatile compound. It therefore lends itself to serving as reference compound j in Equation 3. Values of kj for oxy gen (kn) can be measured using well-established protocols. Other wise, ka can be estimated using models described in the literature (29, 30). Values of Ψ for VOCs and oxygen have been noted to be ap proximately 0.6 in well-mixed sys tems. They have also been observed to be insensitive to temperature, presence of surfactants, and degree


Volatilization at open surfaces Processes with natural volatiliza tion across open surfaces include clarifiers, conveyance channels, ponds, quiescent portions of acti vated sludge basins, and collection systems. Equation 4 is often used for uncovered processes. For covered processes, Equation 1 is applied with a simultaneous gas-phase mass balance to obtain Cg. For flowing streams with measurable velocities, such as sewers, stream reaeration models can be used to estimate KL. However, clean water oxygen trans fer coefficients must be adjusted by ¥ and a, the transfer coefficient ra tio between wastewater and tap water. Alternatively, a model for oxygen transfer to untreated waste water (29) can be used. A model de veloped to predict emissions of ben zene and toluene from quiescent surfaces with wind waves (36) can be used with Equation 3 to estimate KL for other VOCs with Hc > 0.1. Surface aerated systems For mechanical surface aeration systems, VOC emissions can occur from the agitated surface, airborne wastewater droplets, or mass trans fer to entrained air bubbles. The lat ter has typically been assumed to be negligible, so that Cg - 0 and Equa tion 4 applies. The mass transfer co efficient can then be estimated from the empirical expression derived by Roberts et al. (32). Recent work (37) evaluated the modification of Ψ with a liquidphase resistance term and demon-

strated good agreement between theory and observation for 20 VOCs over a range of hydrodynamic con d i t i o n s . For c o m p o u n d s w i t h Hc < 0.2, gas-phase resistance was significant, confirming results by other researchers [26). It has also been observed that kg/kj was rela tively constant with values of 18 to 26 for full-scale surface aerators, with power inputs of 2.3 to 27.8 W/ m 3 over a temperature range of 20 °C [38). These values were lower than reported above for bench-scale studies and were attributed to a greater contribution of stripping caused by entrained bubbles at the larger scale. This apparent gasphase limitation might be a result of using a model that does not account for stripping caused by entrained bubbles where saturation can be limiting. Subsurface aeration In subsurface aeration, oxygen diffuses from air bubbles to waste water, w h e r e a s VOCs can be stripped from wastewater to rising bubbles, with subsequent release to the atmosphere when the bubble reaches the wastewater surface. Be cause bubbles accumulate VOCs as they rise through wastewater, the gas concentration cannot be as sumed to be zero, and Equation 1 applies. Alternatively, if contami nated air such as that from sewer headspaces or covered tanks is used as the aeration gas, VOC mass trans fer from bubbles to wastewater will occur. Air bubbles, which come to equilibrium with VOCs in the aque ous phase, are said to be saturated with respect to those VOCs, and emissions can be estimated by: E=QgHcC (5) 3 where Qg is the aeration rate (m /s) and Ε the VOC emission rate (mg/s). Such an assumption has been made to estimate VOC emissions from municipal wastewater treatment systems [39). Because rising bub bles may reach only partial satura tion with individual VOCs, Equa tion 5 was modified to include a fractional saturation term [32): E=QgHcC\lexV[-VKLa0VI[HcQg))) (6) where KLa0 is the oxygen transfer coefficient (1/s) and V the reactor volume (m3). As HcQg becomes very small, the b r a c k e t e d term ap proaches unity and Equation 6 ap proaches Equation 5 (i.e., saturated conditions exist). However, as Hc and/or Qg increases, the assumption

where a is the transfer coefficient ratio between wastewater and tap water, and

open to the air and well ventilated, Cg can be assumed to be negligible, and Equation 8 reduces to Cz_voc = Ci-voc^x· Otherwise, Equation 7 should be used to account for re duced mass transfer by VOC accu mulation in the gas phase above falling wastewater [46). Drawing u p o n previous oxygen transfer equations accounting for entrained air bubbles in clean water [47), em pirical expressions have been de veloped for oxygen transfer from wastewater flowing over primary and secondary clarifier weirs [48). Several experiments and model ing efforts have indicated that aver age relative VOC emissions across primary treatment processes can be as high as 10%, with significant contributions from weirs and aer ated grit chambers [18, 48-51). VOC losses can be even greater for sys tems that employ pre-aeration pro cesses and aerated conveyance channels, and can be significantly lower for covered primary treat ment systems and submerged weirs [52, 53). Several mechanisms can contrib ute to gas-liquid mass transfer at drop structures in sewer manholes. It can occur at the free-falling waste water surface, from airborne drops generated by splashing at the sur face of the receiving tailwater, at ag itated tailwater surfaces, and via air bubbles in the tailwater. One of the few investigations of VOC emis sions from sewer drop structures has shown a strong linear correla tion between oxygen transfer and both drop height and wastewater flow rate [54). When this work was extended to address the fate of four VOCs in enclosed sewer drop struc tures, stripping efficiencies and mass transfer coefficients were found to be a strong function of Hc. This suggested the potential impor tance of air entrainment as a mass transfer mechanism [55). Also, kglkt ratios ranged from 0.6 to 15. These values were much lower than dis cussed above and again indicated that gas-phase resistance is proba bly more important than previously considered. In comparing VOC emission data in sewer reaches [56, 57), it is evident that drop structures are likely to be the most dominant source of VOC emissions in munici pal and industrial sewers [55).

rx = (Q.voc ~ CgIHc)l (Q-voc - CgIHr) (8) are where Οτ_νοο and Q,.Voc VOC concentrations immediately up stream and downstream of the drop, respectively. If the drop structure is

Loss by sorption Sorption has not been found to be a major mechanism of removal for VOCs in POTWs [18, 58). Sorption may involve adsorption in which

that air bubbles are fully saturated maybe inappropriate. Diffused bub ble aeration enhances turbulent mixing and, hence, volatilization at the open surface (surface desorption). Although the contribution of surface desorption to overall emis sions from such systems is not well understood, it has been observed that surface desorption accounted for 28 to 59% of total transfer rates of chloroform, carbon tetrachloride, trichloroethylene, and tetrachloroethylene in a laboratory-scale dif fused aeration system [40). For a given oxygen transfer rate, mechanical surface aerators pro duce greater VOC stripping than do subsurface aerators [41). With other conditions constant, VOC stripping increases with increasing aeration rate, and systems that provide the required oxygen transfer at reduced aeration rates will reduce VOC stripping. For example, fine pore diffusers provide greater oxygen transfer per unit air flow rate than coarse bubble diffusers and thus can reduce VOC stripping [41, 42). Pure oxygen systems provide en hanced oxygen transfer and can re duce VOC stripping rates [43). Re cycling of exhaust air from grit tanks and aeration cells can also m i n i m i z e VOC s t r i p p i n g from POTWs [41). Drop structures For volatilization at drop struc tures, such as weirs, it is not possi ble to accurately measure interfacial contact area for droplets formed by falling water in contact with under lying receiving water or entrained air bubbles. Thus, models devel oped to estimate VOC emissions from drop structures are empirical and are based upon equations that predict oxygen reaeration in clean water flowing over weirs. They cor relate the dimensionless oxygen deficit or depletion ratio, ra to drop height, discharge, and pool depth. Researchers have developed em pirical relationships for ra [44, 45). The value of rn is related to an anal ogous term for VOCs according to: rx = r0"*

[7)


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VOCs are bound to solid surfaces such as cell walls, or absorption in which VOCs are dissolved in a second liquid phase. The observation that the sorption properties of primary, secondary, and digested sludges from several treatment plants exhibited no statistically significant differences in sorption properties when the solids concentration was based on volatile solids, suggests that data obtained for one sludge can be used to estimate sorption properties for different sludges and different treatment plants (59). Current models usually do not account for the kinetics of the sorption process. They assume steady state conditions and that VOC sorption takes place instantaneously. It has been found that at the low VOC conc e n t r a t i o n s t y p i c a l l y seen in POTWs, linear i s o t h e r m s adequately describe sorption equilibria; that is, the rate of removal of a VOC is first order with respect to the solids concentration in waste sludge and the liquid-phase VOC concentration [59). Values for the sorption partition coefficient, KP, can be estimated from octanolwater partition coefficients using the empirical relationship derived by Dobbs et al. [59). Activated sludge systems employing powdered activated carbon (PAC) to enhance removal of toxic compounds have been used in industry. Sorption of VOCs is enhanced in the presence of PAC. Improved removal and r e d u c e d stripping of poorly degradable VOCs have been observed when PAC was added to an activated sludge system (60). Loss by biodégradation Biodégradation of organic compounds in wastewater treatment systems is typically modeled by a form of the Monod equation in which the rate of compound disappearance, r (g/m 3 /d), is related to the concentration of the compound, S (g/m 3 ), and the active microbial cell mass, Xa (g/m3). If the substrate concentration is significantly less than the half-saturation coefficient, Ks (g/m3) (which would be the case for VOCs in most wastewaters), the Monod equation reduces to a firstorder expression with respect to substrate concentration: r = -\LmSXj{Y[K„+S)} = -p m SX a /(YiQ = kSXa (9) •where p m is the maximum microbial growth rate (1/day), y is the cell yield coefficient (g cells/g substrate),

and k is the apparent first-order biodegradation rate constant, \Lm/(YKs) (m3/g/d). In a review of biodégradation kinetics models, it was shown that, at the typically low levels at which VOCs are found in m u n i c i p a l wastewaters, inhibition effects are ignored in most models [61). These concentrations may be below those r e q u i r e d to s u p p o r t microbial growth, and degradation may occur by secondary utilization or cometabolism. This is especially critical in biofilm reactors where significant diffusional resistances may occur to mass transfer of nutrients, oxygen, and VOCs (62). Models that describe secondary utilization have been developed (63). Utilization of Equation 9 suffers from the difficulty of interpreting the active cell concentration, Xa. Most models use the total or a fraction of volatile suspended solids to describe the active biomass and all models assume acclimated biomass (61). Others have interpreted X„ as being a large fraction of the total biomass that would degrade VOCs by secondary utilization (39). An additional drawback of Equation 9 is the lack of biodégradation rate data for VOCs. Published biodégradation rate coefficients can vary by up to three orders of magnitude and have been shown to vary according to the degree of biomass acclimation (64). Some researchers have addressed this problem by using a structure—activity relationshipbased approach to estimate biodégradation rate coefficients (65). Generally, in wastewater treatment plant environments, nonchlorinated VOCs, such as benzene, toluene, and xylene, are biodegradable, whereas chlorinated VOCs, such as trichloroethylene, tetrachloroethane, and dichlorobenzene, are regarded as recalcitrant. However, there is increasing evidence that some chlorinated VOCs are amenable to biological degradation (66, 67). Evidence of VOC biodégradation under anaerobic conditions by reductive dehalogenation has been presented (68) but this phenomenon has not been actively exploited in wastewater treatment plants. Low concentrations of chlorinated methanes have been reported in anaerobic municipal sludge digester gas, suggesting minimal sorption to waste sludge streams, destruction during digestion, or both (69). Concentrations of benzene, toluene, xylenes, and ethylbenzene in digester gas were high, suggesting removal in waste sludge streams fol-

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lowed by desorption, formation as metabolic byproducts, or both. VOC fate and transport models The need for VOC fate and transport models was prompted by the complexity and high cost of direct sampling and measurement of VOC emissions from POTWs. Emissions may be estimated from mass balances based on flow and wastewater concentration data. Although adequate for screening purposes, they are conservative because they do not account for sorption, biodégradation, and pass-through. The estimates may be adjusted with an emission factor, which accounts for the fraction of mass loading of loss attributed to volatile emissions. VOC s t r i p p i n g factors are not widely available for many unit processes, and those that are do not account for plant characteristics that will reduce or enhance the emissions. The most comprehensive determination of emission factors for municipal treatment facilities was the Pooled Emission Estimation Program conducted at 21 POTWs for 18 unit processes and 21 VOCs by 24 California agencies in response to the AB 2588 regulation (70). The drawback of these plantwide emission estimation techniques is that they do not quantify the primary sources of VOC emissions (71). The advent of fate and transport models allowed the simulation of competing VOC removal mechanisms in wastewater and should lead to improved estimates of VOC emissions. Such models can also be used to examine the effect of process changes designed to minimize emissions and their impact on effluent VOC concentrations. They have been vised to predict emissions for air permitting purposes (72). Many models have been developed, but not all are computer-based (39, 46, 50, 73-79). They include equations that describe the VOC removal mechanisms described above with differing degrees of refinement relating to mass transfer assumptions and methods of calculating rate coefficients. The attributes of several models have been reviewed (71), as have those of computer-based models (61, 80). Table 1 is based on an extensive review of computer-based fate models (23). None of the currently available models describe a complete liquid treatment train that incorporates headworks, primary and secondary treatment, disinfection, and tertiary treatment; nor do they include the

sludge processing train or sewers. CORAL (81) is the only published computer-based model for VOC stripping in simplified sewer networks. It does not account for sorption and biodégradation. Existing treatment plant models do not simulate the formation of chlorinated products during wastewater disinfection nor the products of hydrogen sulphide oxidation. Most advanced models address activated sludge treatment or aerated stabilization basins. Few data are available to describe VOC fate in other secondary systems. A pilotscale study of fixed-film systems showed that biological aerated filter (BAF) air emissions had the highest mass of VOCs per volume of wastewater treated [82). Emissions were lower for the rotating biological contactor and trickling filter tested. Other work showed comparable masses of VOCs in air emitted from parallel BAF and activated sludge systems (83). Fixed film VOC emission models have not been published to date. The activated sludge models follow the assumption of secondary utilization and require prior solution of the process biomass and oxygen balances. Most fate models assume that the overall removal rate of a target VOC is equivalent to the sum of the removals attributed to each of the mechanisms described above (21, 39). These mechanisms can be competitive, a phenomenon that has been measured in laboratory-scale reactors (64, 84, 85). BASTE (76) has a flexible building block approach to simulating the liquid treatment train as compared to the fixed process flow schematic of the others, which generally includes both primary and secondary clarification. All models are steady state except for TOXCHEM (78), which also simulates the dynamic response of treatment systems. TOXCHEM also includes the fate of VOCs in aerobic and anaerobic sludge digesters. A shortcoming of the models, except for BASTE and TOXCHEM, is the lack of temperature correction for Henry's law coefficients (62). When volatilization is the dominant removal mechanism, this could lead to significant error in the predicted VOC removal even over a small temperature range. Also, all the models neglect interaction between target VOCs except PAVE's treatment of binary mixtures. NOCEPM, PAVE, SIMS, and WATER 7 are based on industrial waste-

TABtE 1

Computer-based fate and transport models Name

Developer/source

Description

BASTE (Bay Area Sewage Toxics Emissions)

Developed by R. L. Corsi of the University of Texas, Austin, and CH2M-HÎII Inc. (Emeryville, CA) for the Bay Area Air Toxics (BAAT) Group. Copyright May 1990. Developed by Research Triangle Institute for the EPA Office of Air Quality Planning and Standards, Research Triangle Park, NO

BASTE is a general fate model that includes split flows, quiescent surfaces, drops, weirs, packed media, aerated processes, biological processes, and covered processes. CHEMDAT 7 is a fate-and-transport model for aerated and non-aerated wastewater treatment processes and impoundments. CHEMDAT 7 includes CHEM 7, a computer program to estimate compound properties. CINCI consists of some conceptual model components selected from the literature, Sorption correlations are unique to CINCI. Model has QSAR capability, albeit limited to biodégradation. Makes use of group contribution and neural network approaches. CORAL is a fate model that simulates two-phase, transient VOC transport and gas-liquid partitioning in enclosed wastewater collection systems, including collection reaches and drop structures. The EPA FATE model is essentially an expanded computerized version of the conceptual model presented by Namkung and Rittmann (39).

CHEMDAT 7 (also known as WATER 7)

CINCI (EPACincinnati Model)

Developed by Richard Dobbs of EPA and Rakesh Govind of the University of Cincinnati with support from the EPA Risk Reduction Engineering Laboratory, Cincinnati, OH.

CORAL (Collection System Organic Release Algorithm)

Developed by R. L. Corsi of the University of Texas, Austin while at University of California-Davis

EPA FATE (Fate and Treatability Estimator)

Developed by ABB Environmental, Inc. (Portland, ME) for the Office of Water Regulations and Standards, Office of Water, EPA, Washington, DC. Developed by D. A. Barton of the National Council of the Paper Industry for Air and Stream Improvement (NCASI), New York, NY.

NOCEPM (NCASI Organic Compound Elimination Pathway Model) PAVE (Programs to Assess Volatile Emissions) SIMS (Surface Impoundment Modelling System-1990) TORONTO A Model of Organic Chemical Fate in a Biological Wastewater Treatment Plant TOXCHEM Toxic Chemical Modelling Program for Water Pollution Control Plants

Developed by the U.S. Chemical Manufacturers' Association, Washington, DC. Developed by Radian for EPA Office of Air Quality Planning & Standards, Research Triangle Park, NO Developed by B. Clark, D. Mackay, L. Tasfi, G.L.H. Henry, and S. Salenieks of the Institute of Environmental Studies, University of Toronto, with support from the Ontario Ministry of the Environment. Developed by Environment Canada's Wastewater Technology Centre, Burlington, Ontario, and Enviromega Ltd., Campbellville, Ontario. Copyright 1990.

water systems, unlike the others, which are focused on municipal wastewater applications. NOCEPM (73) models activated sludge and aerated lagoon systems only. It

NOCEPM consists of various conceptual model components selected by NCASI staff from the literature.

PAVE is a set of computer models for determining volatile emissions from wastewater treatment units and from spills of liquid solutions. CHEMDAT 7 has been incorporated into SIMS. Also, emission models for collection systems from IWW CTC documents are incorporated into SIMS. The TORONTO model is based on the fugacity modelling concept and uses equations similar to those proposed by Blackburn et al. (21) to describe chemical partitioning between phases. TOXCHEM consists of various conceptual model components selected from the literature to address the fate of contaminants through all stages of wastewater treatment. It is the only model with unsteady state capability that allows prediction of the plant's response to a spill condition.

takes account of all VOC removal mechanisms. PAVE simulates only complete-mix activated sludge systems and incorporates complex biological kinetic models (79). It can Environ. Sci. Technol., Vol. 28, No. 7, 1994 333 A

also simulate VOC emissions from chemical spills of either pure components or binary mixtures. SIMS was produced to determine expected air emissions for industrial wastewater treatment facilities on a nation-wide basis as a tool to produce rule-making support data (86). Neither PAVE nor SIMS account for sorption of VOCs. The calibration of overall fate models is costly; this can limit their applicability, although proponents have shown significant success in the predictive qualities of their models. The NOCEPM model can predict the fate of VOCs typically found in pulp and paper wastewater treatment systems (73, 87). The FATE model has been validated at pilot-scale and at POTWs (75). VOC treatability parameters for TOXCHEM have been determined at pilot- and full-scale (88). Model pred i c t i o n s c o m p a r e d well with observations of the fate of selected VOCs at a POTW. The application of BASTE at nine POTWs showed that upstream process modifications tended to increase VOC emissions in downstream processes (53). Conclusion Significant advances have been made over the past decade in understanding the magnitude of VOC emissions from wastewater treatment and collection systems and in m o d e l i n g major VOC removal mechanisms. The development of overall fate and transport models has provided a useful tool to predict VOC emissions from existing instal-

lations and from proposed process modifications intended to minimize emissions. Recent comparisons of fate model predictions with actual measurements show that the models are perhaps not sophisticated enough and/or require more calibration (89). However, the models offer a relatively inexpensive method of making first estimates of emissions and can direct the allocation of scarce funding to the most critical areas. The Achilles heel of any of the overall predictive emission models is the inadequacy of accurate biodégradation coefficient data. The d e v e l o p m e n t of a simple method to measure these coefficients is highly desirable. References (1) (2) (3) (4) (5) (6)

(7)

(8)

(9)

(10)

(11) Henryk Melcer is senior process specialist with Brown &• Caldwell Consultants in Seattle, WA, where he is responsible for the VOC and toxics control program. Previously he conducted research at Environment Canada's Wastewater Technology Centre in Burlington, Ontario, Canada, where he directed fate-of-toxics investigations over the past 12 years. He has taught at McMaster University, Ontario, and Massey University, New Zealand. He received his B.S. degree and Ph.D. in chemical engineering from the University of Birmingham in England.

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(12)

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Monitoring and Modeling VOCs in Wastewater Facilities

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