‘lraceorganics in groundwater Processes affecting their movement and f a t e in the subsurface environment, and research needs in this field are discussed
Perry L. McCarty Martin Reinhard Department of Civil Engineering Stanford University Stanford, CaliJ: 94305 Bruce E. Rittmann Department of Civil Engineering University of Illinois Urbana, Ill. 61801
More than 40% of the U S . population use groundwater for drinking, often without any treatment other than disinfection. Moreover, about 25% of fresh water for all purposes comes from the ground. Generally, groundwater has been considered a pristine source, but recent evidence indicates that some groundwater is contaminated with synthetic organic materials, especially in urban and industrial areas. The extent of contamination is presently unknown, but this has the potential of becoming one of the country’s most serious problems, largely because groundwater does not have the natural cleansing mechanisms common in surface water. Once contaminated, it may remain so for years, decades, or even longer. Monitoring to determine the location, extent, and source of contamination of groundwater is also far more difficult and expensive than with surface waters. Questions of importance are: How effectively are organic pollutants removed during travel of water into and through subsurface systems? What are the mechanisms of removal or trans40
EnvironmentalScience & Technology
formation? What are the end products of possible transformations? What is the transport speed of the pollutants with respect to that of the associated water? These questions on the behavior of organic pollutants are being addressed through a research project connected with the 0.088-m3/s, 2-mgd (million gallons per day) Palo Alto water-reclamation facility, which is operated by the Santa Clara Valley (Calif.) Water District and supplies water for a groundwater injection system (1). Basic laboratory studies are also being conducted so that from the field data, general conclusions may be drawn and applied elsewhere. Fate of organic materials Figure 1 illustrates how various phenomena affect the movement of organic compounds in a subsurface system. In this system, water radiates horizontally from an injection -well and, after a certain time of travel, tHzO, reaches an observation well. An organic compound contained in the injection water may spread out because of dispersion, may move slower than the injected water because of sorption, and may decrease in mass because of chemical and biological degradation. A conservative Contaminant would behave just as the water in which it is contained, but the contaminant’s concentration patterns would be spread out because of dispersion. If a solute, i, in the injection water is partitioned between the water and the material containing the aquifer, then the movement of i through the system is retarded and it arrives at the observation well later than the con-
servative contaminant. Dispersion would broaden the contaminant response, and its concentration would gradually increase at the observation well until it approached the value COas the sorptive capacity was saturated. Biodegradation would act to destroy the organic contaminant, so its concentration at the observation well would never reach CO.If only dispersion and biodegradation were operative, the concentration at the observation well would initially be governed by dispersion alone, until sufficient time had passed to establish an acclimated bacterial population with sufficient mass to change measurably the organic concentration. The contaminant concentration would then decrease and level out at some minimum value. If sorption were also involved, then the process would be similar, but contaminant arrival at the observation well would be delayed still further. Sorption A method for interpreting field results on the movement of contaminants in an aquifer was given previously (2). Figure 2 illustrates the characteristics of breakthrough at an observation well following injection of a conservative tracer, as well as a compound which is subjected to .sorptive and dispersive forces only. Vlw (volume) represents the volume of water injected into the aquifer since the step-change in the contaminant’s concentration from 0 to CO.For a constant flow rate of injection water, Ql (volume/time), the mean residence time of the injected water can be estimated from the response of the conservative tracer at the
0013-936X/81/0915-0040$01,00/0
@ 1981 American Chemical Society
observation well:
(1)
Also, the mean residence time for organic compound i is:
wherefiw andfi represent the ratios of output to input concentrations of conservative tracer and compound i, respectively, at time t after injection began. V, and Vi are represented by the areas above the respective breakthrough curves shown in Figure 2. The relative residence time ( t J i in the aquifer for the contaminant, compared to the residence time of water, is defined as:
mum quantity of contaminant that can to exist as a mixture of N independent be accumulated in the subsurface phases, (Kp)i represents the averageof system when contaminated water N individual sorption constants, K,. flows through it. This can be important Provided that the Kj values and the in estimating the time required to fldsh composition of the aquifer materials the system of the contaminant after it are known, ( K J i can be calculated by adding all individual contributions: is removed from the applied water. The specific aquifer retention capacity is a measure of the concentra- (Kp)i =fi(Ki)i +.fdK2)i +A(Kj)i + . ” + f N ( K N ) j (7) tion of contaminant which would be sorbed to the aquifer material when wheref, is the fraction of phase j and there is equilibrium with the input (Kj)i is the partition coefficient for concentration of contaminant in the solute i onto phase j. AssumingJhe injection water (CO)~. Karickhoff, Brown, and Scott (3) system is in equilibrium, fi* and ( C O ) ~ evaluated the sorption of several hyare related to the partition coefficient, drophobic organic contaminants on (Kpk natural sediments. They (and others (Kph= ri*/(Co)i ( 6 ) referevced in their article) found that partition coefficients for sediment/ The units used to express q* and (co)i water systems tend to increase proportionally with the organic content of should be the same (pg/kg or ppb). Because of the heterogeneity of aq- the matrix. Sorptive effects of the inuifer material, which may be thought organic matrix were negligible at an
FIGURE 1
Another parameter of interest is the specific aquifer retention capacity, q* (mass of compound i retained per unit mass of dry aquifer material at a constant input concentration, CO),which is defined by:
ri* =
(Fdi Jw0 (fiw -fi)dViw
Effects of dispersion, sorption, and biodecomposition on the time change in concentration of an organic compound at the aquifer observation well’ Expected responses lo a SI
1 concentratlo I
1
r’
C’C,
(1 -fiw)dviw
&g caq
=(
Co)i%35
(4) VP where e, is the “effective” porosity of theaquirer and equals the ratio of the volume of injected water in an aquifer element at steady state to the total volume of the aquifer element; pa9 is the average bulk density of aquifer material under field conditions. Equations 3 and 4 can he combined to give a relationship between relative residence time and specific aquifer retention capacity: Paq
ri* = (Co)i % [(tr)i - 11
Sorption and dispersion
.
0
0
1
2
3
Biodegradation. sorption, and dispersion
4
5
6
FIGURE 2
Comparative response at an obsewation well, following i n p c t i i of a consetvative tracer and a sorbable compound, i’
(5)
Pas
Values for (tJi and ri* can be obtained from plots of field data that give fractional breakthrough values for the injected water (determined by tracer experiments) and for thecompound i, as illustrated in Figure 2. A value of (tr)i tells how long it would take a contaminant to pass from one point to another in a subsurface system, given the travel time of water between the two points. The relative retention capacity allows estimation of the maxiVolume 15. Number 1. January 1981 41
organic content level of 1% and above. Hence, only one term had to be considered and the partition coefficient with respect to the organic phase, (K,& could be calculated: ( K d i = (Kp)i/foe (8) where foeis the fraction of organic carbon in the solid matrix. (K,)i is a function of particle size and is about the same for clay and fine, medium, and coarse silt. Values for sand fractions are generally 20-25% of those for the finer materials. However, since little of the organic material was associated with the sand, this difference was not considered significant. These workers (3) also evaluated the relationship between the organic partition coefficient and various properties of organic solutes, and found a highly significant correlation with the octanol/water partition coefficient, (Kow)i,for a variety of organic compounds:
(K& = 0.63(KOw)i ( r 2 = 0.96) (9) They reported that the organic carbon partition coefficient changed significantly if the concentration of the organic compound in water approached the compound's solubility limit. Therefore, Equation 9 appeared to be applicable only when the contaminant concentration is less than one-half of its solubility in water. Within this range, the adsorption was linear. Mixtures of hydrophobic compounds sorbed independently through the linear portions of their respective isotherms. Assuming that the retention is solely due to the organic phase and assuming the relationship (Equation 9) found by Karickhoff et al. holds for the organic carbon fraction of aquifer materials, a relationship can be developed between the relative retention time of a compound and its octanol/water partition coefficient by combining Equations 5, 6, 8, and 9: (fr)i
=1
+ 0.63f~(K.w)i(~.q/e.q) (10)
Similarly, a value for the specific aquifer retention capacity can be obtained by combining Equations 6, 8, and 9: Q*
= 0.63(Cj)fs(KOw)j
(11)
Table 1 presents a comparison between relative retention times for various organic contaminants, as calculated from Equation IO, and measured values for the field injection experiments at the Palo Alto water reclamation facility, as described earliei 42
Environmental Science 8 Technology
retention times for various organic compounds
Chlorobenzene 1,4-DichIorobenzene
(2). The (t,)i value for chloroform was not given previously, but was estimated similarly. The predicted values are based upon an assumed average value for& of 0.01 for the Palo Alto system. The values for paq and caq were estimated to be 2 kg/kg and 0.22 kg/kg, respectively (2). The ri* value calculated from Equation 1 1 for chlorobenzene at the average concentration (4.1 pg/kg) in the injection water was 18 pg/kg, which compares favorably with the previously reported field value of 16 pg chlorobenzene removed per kg of aquifer material. The above comparison suggests that the organic content in the aquifer of the Palo Alto well system is the only stationary phase of significance retaining and accumulating hydrophobic trace organics. However, sorption studies using pure minerals, such as sand, limestone ( 4 ) , and montmorillonite clay ( 5 ) . have demonstrated that hydrophobic solutes also are sorbed by inorganic surfaces. Consequently, there must be a critical level of organic matter in inorganic matrices at which sorption to organic and inorganic materials is the same, and below which the organic phase is not dominant. This critical level can be calculated by rearranging the expression for (Kp)j for a hypothetical two-phase system: (12) where (Ki,)j andfj, are the partition coefficients for solute i with the inorganic phase and the inorganic fraction, respectively. Assuming (K& >> (Ki,)i andhi, N 1, we obtain the critical organic fraction for solute i, v,,,*)i: (Kp)i = f i o ( K i o ) i
+fw(Kx)i
To determine (foc*)iin a system consisting of organic matter and silica, the sorptive characteristics of silica for various hydrophobic solutes were determined in a column packed with chromatography-grade silica (Merck Si-60) (6). The silica surface was found to exhibit selective properties, in that it retained compounds with equal
or similar (Kowh values to markedly different degrees. Plotting the log of the sorption coefficient with silica, (KJi, for the nine compounds against their log (Kow)ivalues, and fitting the data points to a linear equation (the linear free-energy relationship), revealed a positive correlation which was significant only at the 0.25 level ( r 2 = 0.47). Thus, the correlation was poorer than that found with organic matter partitioning. The compounds tested included benzene, chlorinated benzene isomers, naphthalene, and tri- and tetrachloroethylene; their (KO& values ranged from 102-104.The expression derived from the free-energy relationship to approximate sorption by silica surfaces was:
where (KSs)iis the constant relating the surface to the solution concentration; S is the silica-specific surface area (mZ/g). By combining Equations 9 and 14, general expressions for (fs*)i were obtained, in which S and (KO& were the only variables: S l (15) = 200 (~,,)~0.84 This equation reveals that the soil organic matter becomes more important as (Kow)iof the solute increases; but its importance decreases as S increases. For benzene, with (Kow)i= 102 and S = 1 3 m*/g, the critical organic fraction would be on the order of 0.1%; for trichlorobenzene isomers, with Kow = lo4, (fm;)i 0.001%. Thus, if aquifer materials have a very low organic content, it is possible that sand and clay may dominate the retardation effect. Such cases can be expected to exhibit selective organic solute retention. Clearly, some of the assumptions made in arriving at these conclusions are tenuous. Numerous other factors need to be investigated, such as the effects of multiple dissolved organic solutes on one another, effects of
metal-organic complexes, interactions of ionized organic molecules, and the specific role of the various aquifer minerals, including clay and silica. In spite of these limitations, the reasonable correlation of relative retention time and specific aquifer retention time with the octanol/water partition Coefficient is most attractive, especially when the aquifer organic content is above about 0.1%. (Kow)i values have been tabulated for numerous compounds (7-9). or can be estimated based on substituent constants (7). Thus, information to make general predictions for many compounds is readily available or can be obtained with relative ease (3). From the above discussion, it appears that compounds with octanol/water partition coefficients less than about IO3 can move through the subsurface environment quite readily, and that migration of organic compounds from points of application to points of water withdrawal appears likely. Biodegradation Transformations of organic compounds in the subsurface environment can occur through the action of microorganisms which are attached to particles or contained within the void spaces between. This biodegradation can be modeled by biofilm kinetics. Figure 3 illustrates the characteristics of an idealized biofilm, which is a homogeneous matrix of bacteria and their extracellular polymers that bind them toeether and to the inert surface (10-12f
The ideal biofilm has a uniform cell and a lodensitv. ~.Xt. (mass/volume). . cally uniform thibkness, i;(length). Substrate concentration within the biofilm changes only in the z direction, which is normal to the surface of the biofilm. Substrate gradients in the x and y directions are small and are taken as zero. Bacteria require many nutrients for growth; but for biofilm modeling, all required nutrients generally are considered to be in excess concentration, except one which is called the rate-limiting substrate, or simply the substrate. The model illustrated in Figure 3 successfully describes the kinetics of bacterially mediated reactions. The model employs three basic processes: mass transport from the bulk liquid, biodecomposition within the biofilm, and biofilm growth and decay. Substrate transport from the bulk liquid through a diffusion layer of thickness L, and to the biofilm surface, occurs in response to a concentration gradient, [Si (Ss)i]/L. The substrate then diffuses into and through the biofilm
-
in response to the substrate gradient within the biofilm. The diffusional flux, (Jz)i (mass/area.time), through the diffusion layer and the biofilm is given by Fick's law of diffusion:
The rate of substrate utilization within the biofilm, (a(Sf)i/at), is described by Monad bacterial kinetics. It is proportional to the concentration of microorganisms present, and is a function of substrate concentrations:
where ki (time-') is the maximum utilization rate for substrate i per unit mass of bacteria; (K,)i (mass/volume) is the half-velocity coefficient for substrate i; and ( X & is the biofilm concentration of bacteria which utilize substrate i. In response to substrate utilization, bacteria grow and the biofilm increases in thickness. Bacteria also require energy from substrate utilization for maintenance. As the biofilm thickness increases, so do the overall maintenance requirements. At some thickness, the energy re-
quirements for biofilm maintenance equal the energy available from sub-, strate utilization; the net growth rate of the biofilm then equals zero. The biofilm is then in steady state, since its thickness is constant with time. The growth of the biofilm can be described by the following equation: a[4(xr)il = at
where Yi is the yield coefficient (mass/mass) for the bacteria utilizing substrate i; bi, the bacterial decay coefficient (time-'), is a function of the maintenance energy requirements for the bacteria. In order for substrate to arrive at the biofilm surface, it must pass from the bulk liquid phase through the diffusional layer. This flux, Ji, is given as:
where Di is the diffusion coefficient for substrate i in the liquid. Once at the surface, substrate maVolume 15, Number l , January 1981 43
terial diffuses into the biofilm and is consumed. by microorganisms. Equations 16 and 17 have been combined and integrated by Williamson and Chung (13) to give the flux of substrate i into the biofilm (JJi at steady state as a function of the concentration at the biofilm surface, (Ss)i:
where (Dr)i is the coefficient of diffusion for substrate i within the biofilm. From a m a s balance, it follows that the flux of substrate i through the diffusional layer must equal its flux into the biofilm: Ji = (Jdi (21) For biofilm outer-surface substrate concentrations well above (KJi, the biofilm thickness at steady state can become substantial, and the concentration of substrate i at the inert media surface (Sw)iwould approach zero. In this case, Equations 19,20, and 21 can be readily solved, either explicitly (12) or iteratively (10. 1 1 ) . However, at the low-substrate concentrationswhich are typical of subsurface waters, this may not be the case. Then, application of Equation 18 may be needed to determine biofilm thickness and the resulting concentration of substrate at the inert media surface. Procedures for these calculations are also presented elsewhere ( I 2). Concept of Sma. An important question for biodecomposition of trace organic compounds is whether or not a minimum substrate concentration exists, below which bacteria cannot utilize the substrate. Based upon Equation 18, a substrate concentration may be small enough so that the rate of energy extraction from the substrate’s consumption is not sufficient to maintain the bacterial cells; if this happens, the bacteria will decay. The concentration at which this will occur can be found by solving Equation 18 for the case where the net rate of growth for an infinitely thin biofilm equals zero (12). The substrate con centration to which the biofilm is ex. posed would then be the same as at the biofilm surface. Also, since substrate utilization essentially would be zero, there would be no net movement of substrate across the diffusion layer. So Si = (SJi = (Smin)i.For this case:
44
Environmental Science (L Technolcgy
from which
Equation 22b suggests that an important limitation exists in the biodecomposition of trace organic compounds by biofilms. This equation has an implied assumption of no interactions between substrates or microorganisms. In other words, it was developed on the basis that organism i obtains energy for growth and maintenance only from substrate i. This is not always the case. Also. Equation 22b applies to steady-state conditions; its applicability and the effects of deviations from the inherent assumptions were evaluated in laboratory studies. Laboratory studies. Experiments were conducted to evaluate the applicability of the biofilm model for describing the kinetics of organic compound removal in the subsurface environment and for evaluating the concept of a minimum substrate concentration. The details of the studies and results have been presented elsewhere (11. 12. 14. 15) and are only summarized here. A small laboratory-scale column (I2 cm long and 2.5 cm in diameter) was filled with 3mm-diameter glass beads to simulate a subsurface medium for which organic sorption would be minimal. Acetate and the required trace inorganic nutrients were contained in a feed solution that was continually passed through the column, which was seeded with a very small concentration of an aerobic, acetate-enrichment culture. The feed solution contained dissolved oxygen in sufficient concentration so that the rate of organic substrate utilization would not be limited. Appropriate coefficients for this system are summarized in Table 2. The biofilm model, Equations 16-19, was used to predict the decrease in substrate concentration
through the column, under steadystate conditions, for several inlet concentrations and flow rates. The numerical solution considered substrate advection and dispersion in the column. A steady-state biofilm took about three weeks to develop. Equation 19 and independently determined kinetic parameters were used to estimate (Smin)ifor acetate. A comparison of results from influent concentrations of 2.2 mg/L and 7.2 mg/L of acetate are illustrated in Figure 4. The steady-state concentrations predicted by the biofilm model and experimental data are shown. The superficial velocity of flow (flow rate per cross-sectional column area) was 3.2 m/d. For both cases in Figure 4, the Concentration decreased to a plateau at values near (Smin)i.The shape of the curves remained essentially constant over several weeks of operation, thus confirming the steady-state nature of the observations. Figure 5 illustrates a further comparison between predicted and observed substrate concentrations at higher flow rates with an influent concentration of 3.6 mg/L. With the lower flow rate, the concentration approached the minimum value within the column; but at the higher flow rate, the column length was insufficient for this. The higher flow rate is of the same order of magnitude as that near an injection well, whereas the lower rates may be more typical of those for movement into the ground from a spill, a hazardous waste disposal site, or at points farther from an injection well. Even with high velocities of flow, the substrate concentration decreased quite rapidly in the steady-state system and approached the minimum value well within 1 m. Thus, most removal of readily degradable organic compounds should occur very near the water’s point of entrance into the subsurface environment, provided the environ-
ment there is not inhibitory to bacterial action. This has important implications both in conducting field studies for evaluating biotransformations and for monitoring hazardous waste disposal sites. Of course, if other nutrients (such as oxygen) are limiting, the picture may change. That aspect needs more consideration. These preliminary studies illustrate that for a single substrate there is a minimum concentration below which a steady-state biofilm cannot be supported. This concentration is predictable from coefficients of bacterial growth and decay. Since these values are functions of bacterial species, the electron donors and acceptors for the reaction, temperature, pH, and other environmental factors, the value for (S,,,in)iwill also be a function of these parameters. Calculated values for other systems are summarized in Table 3 for comparison. Two important conclusions are that the minimum values for aerobic systems appear lower than for anaerobic systems, which is in accord with normal expectations, and that the biological processes are quite fast even at low substrateconcentrations; thus for biologically active systems, the reactions tend to be complete within a meter of entrance of water into the subsurface environment. Biodecomposition below The steady-state biofilm model predicts that a single substrate will not be reduced in concentration below some minimum level that is required for maintenance of the bacteria. This concentration may be on the order of a fraction of a mg/L to several mg/L depending upon substrate donor and acceptor. But is it possible, through biodecomposition, to reduce the concentration of a trace organic compound for which the initial concentration may be in the pg/L range? Experimental studies have illustrated that it is possible, either with a nonsteady-state biofilm or when the compound is being used as a secondary substrate. In the nonsteady-state case, if a biofilm is first developed with a high substrate concentration so that it becomes relatively deep, and then the concentration is suddenly dropped, the concentration can fall below the minimum value. This is illustrated in Figure 6. Here, a steady-state biofilm was first developed at an influent concentration of 1.2 mg/L. The concentration was then suddenly changed to 3.6 mg/L, with the result that the concentration fell to a level of about onehalf (Sm& Compared with the steady-state biofilm, the deep non-
1 FIGURE 4
Experimental evaluation of steady-state biofilm model, and concept of (.Smin),for a single subswat@
Length along column. crn ''Circles represent erperlmenlal pomt5. lines represent model pcediclions ior Column influent acetale ConCentralmnS 01 2 2 mgiL and 7 2 mgit. reSpeClWely
FIGURE 5
Experimental evaluation of steady-statebiofilma 1
4
0
2
4
6
8
10
12
Volume 15. Number 1,Januarl 1981 45
steady-state biofilm was capable of greater concentration reduction even though it had a lower initial substrate concentration. The effect can be even more dramatic if a secondary substrate is utilized (Figure 7). The biofilm was maintained by using the other substrates, but it also had the capability of using acetate. This phenomenon is termed secondary utilization, since the trace organic material is not the primary source of energy for maintenance of the biofilm. The concept has important implications for evaluating the potential for biodegradation of trace organic materials in the subsurface environment. Indeed, secondary utilization may be the main mechanism by which such materials are biodegraded. Competiriue substrate utilization. Water passing through a subsurface system may contain a variety of substrates, or electron donors, some of which are utilized by different bacterial species. If the concentration of a common nutrient, such as oxygen, is limited, then competition for this nutrient may exist. An example is water that contains a biodegradable organic material and ammonia. Here, the competition would be between heterotrophic bacteria, which use the orOB
Environmental Science & Technology
ganic material for energy, and autotrophic nitrifying bacteria, which oxidize ammonia to nitrite or nitrate for energy. A common view is that ammonia cannot be utilized in the presence of organic material; however, a more accurate description is that the autotrophic growth rate is lower than the heterotrophic growth rate. For this reason, the autotrophs lose out in the competition. The effect of growth rate is illustrated in Figure 8. Initially, it is assumed that there are an equal number of dark and light bacteria on the media surface, and that the dark bacteria utilizing substrate A grow faster than the light bacteria using substrate B. During time interval a, a new layer of bacteria is produced which contains more dark bacteria. The greater percentage of dark bacteria in the second layer gives them an even greater advantage in increasing their population. The dark bacteria keep increasing until, finally, the outer layer of bacteria contains essentially all dark bacteria and no light bacteria. Since growth rate is a function of substrate concentration, as well as of the various kinetic coefficients for bacterial growth and decay, it can be hypothesized that the particular group of microorganisms that will dominate
is determined to some extent by the relative concentrations of substrates. Calculations were made for rates of heterotrophic and autotrophic growth in a very thin biofilm for varying substrate concentrations and using the typical coefficients listed in Table 4. The initial population density of each group of bacteria was assumed to be equal. Figure 9 illustrates the results of these calculations. The two lines are the result of different assumptions on autotrophicorganism decay rate. Above the lines, the heterotrophs grow faster and dominate. Below tbe lines, the autotrophs grow faster and dominate. The curves are quite sensitive to organism decay rate, and the upper curve is more indicative of commonly reported values. Based on these curves, whenever the biodegradable organic concentration is equal to or greater than about 3 mg/L, the heterotrophs will dominate, but when the concentration is less than about 1-2 mg/L, the autotrophs will dominate in the biofilm. When there is competition between organic and ammonia oxidation for an excess of dissolved oxygen and when the influent organic concentration exceeds 3 mg COD/L, then the following sequence is likely. At the point where water enters the system, the outer layers of the biofilm will be predominately heterotrophic, and organic oxidation will be rapid. With oxygen in excess, it and the ammonia will diffuse down to the lower biofilm layers, where nitrifying bacteria will have less competition, and some oxidation of ammonia can take place. However, as the water passes through the system and the concentration of organic matter decreases. a paint will be reached where the autotrophs can dominate and OCCUDV the outer lavers of the biofilm. On the other hand, if the dissolved oxygen supply is limited, then it may be depleted within the biofilm before it can penetrate to the depth at which the nitrifying population could compete. Ammonia could still penetrate to that depth, but would not be oxidized. One can envision several other variations with competitive substrates. If the relative substrate concentrations were such that ammonia oxidation dominated at the entrance to the system, then the nitrite and nitrate produced could serve as electron donors for organic oxidation at depths within the biofilm, deeper within the aquifer system after the ammonia concentration was reduced sufficiently by oxidation, or after the dissolved oxygen
.,
FIGURE 7
How sewndwy metabdim reducas the concantration of a substrate below (S,,.,,,ha
FIGURE 8
Biotilm growth with wtnmtltlve
FIGURE 9
RelathreCOncentratlonSofCOO and ammonia nitrogen‘
OGmwonA 0 Grow on B
cg’
Lines of equal growth rate
cr.
t=o
A B t=a
A B t
= 2a
A B
0
1
2
3
Surface NH-N,
4
mg/L
I Volume 15. Number 1, January 1981
47
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48
EnvironmentalScience & Technology
supply had become depleted. Such a system would result in some nitrogen loss by denitrification. The biofilm model is capable of making such predictions, and is therefore useful, in a general sense, for explaining many observed biological phenomena within subsurface systems.
What governs the fate The movement and fate of organic materials that enter the subsurface environment are governed largely by sorption and biodegradation. Sorption affects the rate of travel of organic material, relative to that of water, through subsurface systems. Sorption also allows for the accumulation of organic compounds on subsurface solid media. Hydrophobic organic materials are either adsorbed or absorbed by particulate organic material contained within or on the solid medium; therefore, the extent of sorption appears to be a function of the fraction of organic carbon within the subsurface matrix. The partitioning of hydrophobic organics between the water phase and the subsurface organic particulates can be estimated from the octanol/water partition coefficient for each organic compound. This is an important finding, since such values are readily available for many organic compounds or can be determined with relative ease in the laboratory. Good correlations between predictions and field measurements of relative retention times were found for several organic compounds. The significance of the organic matter for hydrophobic solute retention depends upon the sorptive characteristics of the solutes. While interactions of hydrophobic solutes with organic matter have been shown to be reasonably predictable over a wide range, interactions with inorganic surfaces do not generally follow a simple pattern. The specific nature of these interactions needs further exploration. The role of the inorganic matrix is likely to dominate not only the solute transport of polar and ionized species, but also of nonionized, hydrophobic solutes if the organic content is below the critical level cf,,*),. The concept of treating an aquifer as a mixed phase system is still very hypothetical considering the complexity of aquifer materials. Much needs to be learned about the interactions of the many different aquifer components. With the generation of more and better data on solute-mineral surface interactions, a good basis appears to be emerging for understanding and predicting the movement of organic solutes through
the subsurface environment. The ultimate fate of organic compounds in the subsurface system depends strongly upon their biodegradability. Biofilm models are useful for explaining the extent of biodegradation that can occur and the effects of competition between different electron donors and acceptors. Evidence available indicates there is a minimum concentration to which a single organic material can be decomposed under steady-state conditions. This concentration is a function of the growth and decay coefficients of the bacteria, which in turn are functions of environmental variables such as temperature, pH, bacterial species, and available electron acceptors. Trace organic materials may be biodegradable, but are often below the minimum concentration described above. Biodegradation of such materials can occur in general only if they are used as secondary substrates; that is, if there is an abundant primary organic substrate available and bacteria capable of decomposing both the primary and secondary substrates. An alternative is decomposition within a nonsteady-state system that has a sufficiently large population of bacteria previously grown on a primary substrate. Biodecomposition is also possible if several organic substrates are present in a sufficiently large total concentration. These concepts are just emerging, and relatively little is known of the kinetics of degradation of mixed substrates at low concentrations. Possible interactions among primary and secondary substrates and bacteria are not known, nor are the effects of sorption on rates of transformation. The breadth of organic compounds to which these concepts are applicable is also unknown, as is the ultimate concentration level to which they may be reduced by biological activity. Methods for clearly assessing the potential of a given aquifer for microbial activity are also presently lacking. Surprising, perhaps, is the relative speed of biological transformations that can occur in aquifer systems; biological degradation as described above tends to occur in minutes, rather than in days or years. Yet retention times of water in aquifers may be exceedingly long, and almost no studies on long-term transformations of the more biologically refractory materials in aquifer systems are available. Within such time frames, chemical and biological processes may transform trace organic compounds into other materials that are more or less hazardous; the processes involved have
not been clearly elucidated. More information on the relative roles of chemical and biological factors and on the effects of the aquifer materials and environments on reaction rates is certainly needed if the long-term fate of trace organic materials in aquifer systems i s to be well understood.
Acknowledgment This research was supported by the Robert S. Kerr Environmental Research Laboratory of the U.S. EPA under Research Grant No. EPA-R-804431. Additional financial support was provided by the Water Resources Control Board and the California Department of Water Resources. The authors are grateful to the Santa Clara Valley Water District for assistance and for providing use of their facility for the field measurements. This article was read and commented upon by George L. Baughman, chief, Environmental Processes Branch, Environmental Research Laboratory. US.EPA. Athens. Ca. 30613: and by John T. Wilson, microbiologist. Ground Water Research Branch, Robert S. Kerr Environmental Research Laboratory. U.S. EPA. Ada, Okla. 74820.
w 4
4’ \. 43 Perry L. McCarty is rhe chairman o/the Department of Civil Engineering at
Sran/ord University. H e is active in the field.? o/waterpollution control and sanitary engineering. and has wrirren more than 100 research papers and reports in enuironmental engineering. McCarty has coauthored “Chemistry of Environmental Engineering” ( M c C r a w - H i l l ) . and has been the recipient o/seueral enuironmenral engineering awards.
M a r t i n Reinhard ( 1 . ) is a .seiiior rp.\wrch nssociare ar Sran/ord Unireniry. t i h c w he has been incolved in rcscarch on rhe analysis andjare o/rrace orpanics in water lrearment and in the encironmenl. H e hold.$ a Ph.D. in chemistry /rom E T H . Zurich, Switzerland. Bruce E. Rittmann ( r . ) is assisIan1 professor o/encironmenfal engineering at the Unicersity :0 1llinoi.s at Urbana-Champaign. H e receiced his Ph.D. degree/rom Sran/ord Universiry. and his B.S. and M.S. dexrees/rom Washinglon Uniuer:siry (SI. Lor4i.s. Mo.). Riltmann’s principal research interests are in rhe kinetics ofbioJilms and in rhe hiodepadation o/tracelerel orxanicr.
References (I) Roberts.
P. V.: P. L. McCarty: W. M. Roman. J . Am. So?. Ciu. Eng. 1978, 104(EES), 889. (2) Roberts. P. V.: J. Schrcincr: M. Reinhard: P.L. MFCarty. J . Wovr Pullur. Conrrol Fed. 1980.52, 161. (3) Karickhoff, S . W.: D.S. Brown: T. A. Scott. W a r ~ Re.% r 1979. 13, 241. (4) Ciccioli. P.: W. T.Cooper; P. M. Hammer: 1. M. Hayes. Wolw Rcsourres R m 1980, 161(1).217. ( 5 ) Rogers. R. D.; J. C. McFarlane:A. J. Cross. E m . Sri. Techno/. 1980. 14(4). 457. ( 6 ) Reinhard. M.: P. V . Roberts. “Factors Govcming Adsorption of Trace Organic Contaminants During Groundwater Rccharge by Injection of Reclaimed Watcr.” Report for the California Department of Water Resources. Sacramento. Calif. (7) Lea. A.:C. Hansch: D. Elkins. Chon. Reo. 1976, 71. 525. (8) Veith. G. D.; N. M. Auntin: R. T.Morris. Ware, Res. 1979, I J , 43 (1979). ( 9 ) Kenage. E. W.; C. A. I . Caring. ”Relationship Between Water Solubility. SoilSorption, Octanol-Water Partitioning, and Concentration of Chemicals i n Biota.” Dow Chemical Co.. Midland. Mich. (March 2.
Easv as 1.2.3. MeasureCI,, NH,, and Fthe fussless ORION way. These free applications sheets spell out, step-by.step. how to measure ammonia. fluoride, and total residual chlorine in water and wastewater. They show you just how fussless analysis can be. So why cope with math, color. turbidity, or titration when you can enjoy fast, accurate measurements with an ORION electrode-based system? Send for literature or, better yet. call our Technical Service Department toll.free at 800-225-1480. They’ll arrange for a demonstration using your own samples. What better way to find out why the ORION way is worth switching to?
1979)
( I O ) Williamson. K.: P. L. McCarty. J . Wort, Pollur. Conrrol Fed. 1976.48. 9. ( I i ) Rittmann. B. E.: P. L. McCarty. J . Am Soe. C/o. Eng. 1978, I O I ( E E 5 ) . 889. (12) Rittmann. B. E.: P.L. McCarty.“A M d c l of Steady-Slate-Biafilm Kinetics.” To he puhlirhcd in Biorrchnology ond Bioengineering.
(13) Williamson. K. J.: T. H. Chung. ”Dual Limitation of Suhstratc Ulilimlion Kinetics Within Bacterial Films.” Presented at the 49th National Meeting of the American Institute of Chemical Engineers, Houston. Tex (March ~. .. .... 19751 . . .,. (14) Rittmann. B. E.: P. L. McCarty.“Evaluation of Steady-State-Biofilm Kinetics.” To be published in Biotmhnology and Biomgi-
ORION RESEARCH
840 Memorial Drive. Cambridge. MA 02139 Telephone in Massachusetts. 617.864.5400 ~~~
Circle 35 f o r demonstration: Circle 36 f o r information
“cering.
P. L. McCarty. “Substratc Utilization by Nan-Steady-State Bio. films.” Submitted far whlication.
( I S ) Rittmann. E. E.:
Volume 15. Number 1. January 1981 5 1
