Cycling of Manganese and Iron in Lake Mendota, Wisconsin Robert E. Stauffert
Water Chemistry Laboratory, University of Wisconsin, Madison, Wisconsin 53706 ~~
On the basis of field evidence from 1966-1967 and 1977-1979, Delfino’s laboratory studies, and thermodynamic calculations, Mn2+accumulation in the hypolimnion of Lake Mendota is regulated neither by rhodochrosite equilibrium nor by pH-dependent desorption from hydrous oxides. Instead, the recurring patterns of Mn2+accumulation are best explained by microstratigraphy and redox shifts in the upper sediments, coupled with seasonally dependent dispersion into the overlying water column. During each summer Mn2+accumulation in and below the thermocline is equivalent to 2.5-years’ supply from external sources. Because of rapid removal by oxidation accompanying turnover, Mn concentrations are low in the epilimnion and the long-term sediment retention coefficient (RMn)is 0.94. On the basis of mass budget analysis, sedimentology, thermodynamic calculations, and experimental confirmation, the low concentrations of soluble Fe in the hypolimnion in late summer ( 15 m) at each of six stations, representing both Middleton Bay and the Central Basin (cf. Figure 1). This expanded sampling coverage was designed to test the hypothesis that Mn concentrations a t the Deep HoIe station were representative of the hypolimnion as a whole. In March 1978 a special sampling apparatus equipped with an optico sediment sensor (11,12) was deployed from the ice platform. This apparatus guarantees the sampling levels (to -1 cm) in the 2-m interval immediately above the profundal sediment interface (11,12). By using the peristaltic pump (all dates), it was possible to obtain samples from a thin (several centimeters) stratum and quite close to the sediment-water interface (5-10 cm even under ice-free conditions) without disturbing the interface and introducing visible turbidity into the collected sample. Nevertheless, filtered sample "splits" (0.4-pm Nucleopore) were collected near the sediment interface, without the sample ever contacting the atmosphere, using a pressure filter apparatus (Nucleopore) connected via a T valve and a valve made of Teflon, both going to the outflow tube of the peristaltic pump. Samples for sulfide determinations (250 mL) were collected by peristaltic pump without filtering in September 1978. These were "fixed" immediately in the field by adding zinc acetate. Samples for pH determinations were collected by being pumped into thoroughly flushed 200-mL Winkler bottles. These were tightly plugged, taped, and stored temporarily in a cool dark box until returned to the laboratory. All samples for nutrient and cation determinations were acidified after collection to pH 1.0 using analytical grade HC1. Analytical Methods. Ca, Mg, Na, K, Mn, and Fe were all determined on acidified samples by flame atomic absorption spectrometry using a Perkin-Elmer Model 603 instrument, following standard PE procedures. Alkalinity was estimated (for thermodynamic calculations) as the sum of milliequivalents of cations (including NH,+; known from independent studies) after correcting for the minor charge contributions of C1-, Sod2-, and NO3- (13). These computed alkalinities agree closely with independent earlier 450
Environ. Sci. Technol., Vol. 20, No. 5, 1986
ORIFICES TO TO WHITNEV THERMISTOR
v
L O 1 Crn ' EeC
.
PUMP FLOW
-
THRU WEIGHT
Figure 2. Geometry of combined temperature-pump sampling probe. Pump is rated at 11 mL s-'. Flow-through weight is stainless steel. Note the low mean intake velocity ( 0 )across the base of a cone of approach.
measurements for this lake (4,14). Monomeric silicic acid was determined spectrophotometrically by using molybdenum blue (15). Total phosphorus was also determined spectrophotometrically by using molybdenum blue (16), after first digesting the samples with persulfate in an autoclave (17). The fixed sulfide in the spiked hypolimnetic samples was determined iodometrically (9), after first collecting the ZnS precipitate on a glass fiber filter. pH measurements were made with an Orion meter, after first allowing the tightly stoppered anoxic samples to warm to room temperature in the dark. After a standard buffer check, the combination glass electrode was briefly preconditioned in Lake Mendota surface water before final transfer into the narrow neck of the Winkler bottle. The time then required for electrode equilibration in the bottle was too short to permit significant oxidation of dissolved H2S or degassing of C02. A temperature correction for the bicarbonate system was applied to estimate the pH in situ, at the lower known temperature of the bathylimnion (18; see also ref 19). Computational Methods. Saturation indexes (SI (mineral) = log (IAP/KT), where IAP denotes the ion activity product and KT is the temperature-adjusted thermodynamic constant) were computed for various minerals by use of the Fortran version, by Plummer et al. (20), of WATEQ (21). The thermodynamic constants adopted from Plummer et al. were (log K at 25 "C with AHr0 in parentheses) calcite, -8.41 (-3.19); siderite, -10.55 (-5.33); rhodochrosite, -10.54 (-2.08); vivianite, -36.00 (0.0). The K value for the dissolution of amorphous FeS was adopted from Berner (22). Davison (23) made an environmental study of FeS solubility and found evidence supporting Berner's relatively high solubility for dimictic lakes. The K value for MnS is based on AH,", and although of low reliability, it is close to values used elsewhere (5, 24, 25). The thermodynamic constants for MnC03
apply to well-ordered rhodochrosite (26). Johnson (27) has confirmed these minimum solubility values. Vertical fluxes of manganese within the hypolimnion and metalimnion were computed by the flux-gradient algorithm, using estimates of the vertical eddy conductivity (K,) and concentration gradients measured a t the Deep Hole station, or a t satellite stations in the central basin (some dates in 1966-67; cf. 4, 5). The K, estimates were computed following ref 28, as modified in ref 29 to include the differential effects of heat penetration into sediments as a function of time and overlying water depth (30, 31). The mass of Mn (or Fe) in the water column below a designated plane (1 - 0.5 m) was calculated as
12
N
In
N
mass =
Te(z)~(z)
6
z=1
where 6 ( z ) is the measured concentration a t depth z (integral meters) in the central basin and V(z)is the volume of a l-m-thick slice centered at depth z. The “retention coefficients” for Mn and Fe in Lake Mendota (defined as the fraction of the total external “loading” that is permanently retained in the lake basin; cf. 32) were computed based on estimates of annual inputs vs. losses a t the lake’s outlet. Because the stream inputs of Mn and Fe were not monitored directly, loads were computed indirectly based on the chemical composition of the postcultural sediments (33),coupled with reliable measurements of external P loading (34;R. Lathrop, WI, D.N.R., unpublished). This application of “ratio estimation” (cf. 35) depends on the close geochemical association among Fe, Mn, and P in the postcultural sludge sediments (called gytta) in Mendota (33). Annual losses of Mn and Fe a t the outlet were computed based on seasonal discharge (cf. 34 and references cited therein) times surface concentrations as measured at the Deep Hole (data from 1977-1979 and 1966-1967 were pooled to improve precision).
Results pH and Redox Shifts in the Bathylimnion-Winter. Because of ita exposed continental location and long wind fetches, Mendota freezes (late December) when T (water column) 5 1“C. Water column DO is then a t saturation, following the long autumnal circulation period (cf. 36). Because of the low water temperature (below 4 “C or T(pmax)) and because the ratio of sediment contact area to layer volume (dA/dVin Hesslein’s (37) notation, in m-l) increases monotonically with depth within the hypolimnion, the lower water column (bathylimnion, z 1 18 m) is stabilized by the release of heat and solutes from the profunda1 sediments during ice cover. Because of Mendota’s high alkalinity (3.5 mequiv/L), the pH of the water column is >8.5 during autumnal circulation. The pH drops in the bathylimnion during winter stratification because of increases in Pcozunder the ice. Measured pHs (5) were 7.80 a t 18 m and 7.50 a t 22 m on March 29, 1967. Based on net oxygen consumption at those two levels during 3 months of ice cover (6.2 and 13.0 mg of DO L-l, respectively) and the assumption that ACOz = - A 0 2 (aerobic oxidation of carbohydrate), the calculated in situ pHs (eq 1;3, 18) are 7.82 and 7.47 a t 18 and 22 m, pH(T) = pK1’
+ log [HCOS-] - log [HzC03*]
(1)
respectively. The agreement between the measured and calculated (carbonate system) values is thus excellent. The downward shift in bathylimnetic pE is comparatively slight in winter, because DO remains above the
0
Figure 3. Total solute profiles (mg L-‘) under ice at Deep Hole station on March 9. 1978.
analytical detection limit (Figure 3; see also 4 , 5 , I I ) , and pE is notably insensitive to DO provided DO >0.05 mg/L (3). However, Figure 3 shows that reducing conditions are likely in the uppermost sediment “crust” in late winter. pH and Redox Shifts in the BathylimnionSummer. The downward shifts in bathylimnetic pH and pE are more pronounced in summer than winter because summer “stagnation” lasts longer (-5 months) and because increased primary productivity in the epilimnion markedly increases the BOD “loading” on the hypolimnion. The observed trends are the usual ones for highly productive calcareous lakes of moderate depth (3, 36). The measured pH at 22 m dropped steadily from 8.5 at the end of spring overturn in late May 1967 to 7.50 in mid-July, at the time DO disappeared from the bathylimnion (5). Based on ADO and eq 1,the calculated in situ pH was 7.45 on July 17, 1967, i.e., in close agreement with the measured value. The rate of pH decline slows markedly following the exhaustion of DO (5) because the next four redox couples are alkalinity-producing reactions (eq 3-6 in Table I). Several lines of evidence indicate that bathylimnetic pH ranges from 7.10 to 7.40 in this lake, prior to downward migration of the main thermocline in September. This range includes the majority of late summer measurements in Mendota (this work; 4 , 5 ; a few of Delfino’s earlier measurements could have been biased low by sulfide oxidation), pH measurements in similar calcareous lakes (25,38;G. Holdren, unpublished; R. Stauffer, unpublished), predictions of eq 1 based on late summer sulfide and CHI profiles for Mendota (13),and the stoichiometries listed in Table I. The bathylimnetic pH in early September is 0.10-0.40 unit lower than in late winter. The Mendota hypolimnion passes through several redox stages following the disappearance of DO in mid-summer. By early-mid August nitrate has disappeared completely from the hypolimnion (39) and H2S has commenced its rapid accumulation (6,13). Although sulfate fails to disappear completely from the bottom waters (13), methane also accumulates in the hypolimnion, undoubtedly because fermentation (eq 7 in Table I) is taking place in the sedEnviron. Sci. Technol., Vol. 20, No. 5, 1986 451
29JUL
Table I. Staged Redox and Carbonate Dissolution Reactions in the Hypolimnion of Lake Mendota
170CT
7SEP
'
I
1
acceptance capacity"
reaction 2. CHzO + Oz = COz + HzO 3. 5CHz0 + 4NOS- = COZ + 4HC0; + 2Nz + 3HzO 4. CHzO + 2Mn02 + 3coz + HzO = 4HCOy + 2Mn2+ 5. CHzO + 4Fe(OH)3 + 7C0, = 8HC03- + 4Fez+ + 3HzO 6. 2CHz0 + SO:= 2 H C Q + HzS 7. 2CHz0 = COz + CHI . 8. MeC03 + COP+ H20 = Mez++ 2HCO;
7.0 1.1
0.05b 0.7b 16.0 C
nEquiv e m-2 below the 15.5-m plane (hypolimnion) at onset of stratification. *Calculation based on estimated net annual flux of oxide downward through 15.5-m plane. Not armlieable.
I30
I70
210
1
1
250
290
JULIAN DAY
iments below the zone of sulfate reduction. Profile Development--Winter. By late winter the Mn concentration increases exponentially as the sedimentwater interface is approached in the Deep Hole region (Figure 3). Virtually all of the Mn in the 2-m benthic boundary is soluble (5, I ] ) , coexisting with low concentrations of DO (50.5 mg L-l), at low temperature (95 mg S m-2 per day) exceeds summer Fe external inputs (>6X) and exceeds by a factor of 200 the potential upward diffusion rate for Fez+ in the profunda1 sediments (the latter calculated using Dpe = 2.4 X lo4 cm2 s-l and measured gradients for the top 2 cm of the sediment (54)).Because of this immobilization of Fe as FeS in the sludge sediments, Mendota has been called a “Sulfuretum” (8). After incorrectly assuming that Delfino’s total Fe data represented soluble Fe in the hypolimnion, Hoffman and Eisenreich (7) concluded that organic ligands were necessary to stabilize these “reduced Fe concentrations in the presence of oxygen (early summer) or sulfide (late summer). This hypothesis is tenable if humic and related organic substances are abundant in the lake (55). However, Mendota’s anoxic bottom water is colorless in late summer, and this lake, like most calcareous lakes of moderately long hydraulic residence time set in rich agricultural drainage basins, is generally free of true (nonalgal) color (cf. 3,56). Also, Davison and Woof (38) found no evidence that organic ligands were significantly stabilizing Fe(I1) in the bottom waters of calcareous Rostherne Mere, U.K. Because Rostherne Mere has 3X higher levels of organic-C in the surface sediments as compared to Mendota (cf. 38, 53) and is, on other evidence, more subject to influxes of humic substances, organic ligands are likely to be even less significant in stabilizing Fe(I1) in Mendota. Thus, neither Environ. Sci. Technol., Vol. 20, No. 5, 1986 455
the need nor the basis exists for assuming that the Fe(I1) cycle is influenced by organic complexation in this lake. On the basis of statistical analysis of 35 sediment samples, Delfino et al. (33) concluded that Mn and P were much more closely related in recent Mendota sediments than Fe and P. Because of the long-known association between Fe and P in certain environments undergoing redox shifts (cf. 2,36,50), this Mn-P “association”has been labeled a statistical “artifact” (57). In fact, Figures 3, 5, and 6 (and all of Delfino’s work) directly support the conclusion of Delfino et al., in that the seasonal mobilities of Mn and P are closely comparable in the hypolimnion, whereas Fe is at all times of the year a comparatively immobile element. Thus, the chemical residence times for Mn and P in the lake are 3 f 0.5 years vs. only 1-4 weeks for Fe. Because Mn and P are recycled more frequently from the sediment, one would expect stronger statistical relationships between sediment Mn vs. P composition (a major conclusion reached by Delfino et al.). Based on the very high Mn and P concentrations in the benthic boundary layer in late winter (Figure 3; Table 11), calculations by Nriagu and Dell (58),and laboratory findings of Tessenow (2), it is even conceivable that reddingite (the Mn(I1) analogue of vivianite) forms at times in the surface microlayer of Mendota’s profunda1 sediments. Below the zone of sulfide influence, within the diagenetic zone where vivianite forms (25), the mobility of P declines and its chemistry is once again more directly linked to Fe (25). In other Fe-rich, S-poor noncalcareous lakes, iron dominates sulfur in the sediments and in the water column (23, 47,50,51). Here, the classic linkage between Fe and P (cf. 2, 50) is likely to be observed.
Acknowledgments I thank especially J. Colman for assistance in the field, B. Buhr for assistance in the laboratory, L. N. Plummer for providing a deck and program listing for WATEQF, and N. Dominquez for assistance in adapting WATEQF to the UNIVAC 1110 a t MACC. Through his work and correspondence, W. Davison offered important insights. This paper is dedicated to J. J. Delfino and G. C. Bortleson. Registry No. CaC03, 471-34-1; MnC03, 598-62-9; MnS, 18820-29-6; FeS, 1317-37-9; Fe, 7439-89-6; Mn, 7439-96-5; Si, 7440-21-3; P, 7723-14-0; vivianite, 14567-67-0.
Literature Cited Jenne, E. A. Adv. Chem. Ser. 1968, no. 73, 338-387. Tessenow, U. Arch. Hydrobiol. Suppl. 1974, 47, 1-79. Stumm, W.; Morgan, J. J. “Aquatic Chemistry”, 2nd ed.; Wiley-Interscience: New York, 1981. Delfino, J. J. Ph.D. Thesis, University of Wisconsin, Madison, WI, 1968. Delfino, J. J.; Lee, G. F. Enuiron. Sci. Technol. 1968, 2, 1094-1100. Delfino, J. J.; Lee, G. F. Water Res. 1971, 5, 1207-1217. Hoffman, M. R.; Eisenreich, S. J. Enuiron. Sci. Technol. 1981,15, 339-344. Nriagu, J. 0. Limnol. Oceanogr. 1968, 13,430-439. APHA (American Public Health Association) “Standard Methods for the Examination of Water & Wastewater”, 13th ed.; Washington, DC, 1971. Stewart, K. M. Int. Rev. Gesamten Hydrobiol. 1976, 61, 563-579. Colman, J. Ph.D. Thesis, University of Wisconsin, Madison, WI, 1979. Colman, J.; Armstrong, D. E. Limnol. Oceanogr., in press. Ingvorsen, K.; Brock, T. D. Limnol. Oceanogr. 1982,27, 559-564. Hawley, J. E. M.S. Thesis, University of Wisconsin, Madison, WI, 1967. Environ. Sci. Technol., Vol. 20, No. 5, 1986
(15) Fanning, K. A.; Pilson, M. E. Q. Anal. Chem. 1973, 45, 136-140. (16) Stauffer, R. E. Anal. Chem. 1983,55, 1205-1210. (17) Menzel, D. W.; Corwin, N. Limnol. Oceanogr. 1965, 10, 280-282. (18) Harned, H. S.; Davis, R. J. Am. Chem. SOC.1943, 65, 2030-2037. (19) Davison, W.; Heaney, S. I. Limnol. Oceanogr. 1978, 23, 1194-1200. (20) Plummer, L. N.; Jones, B. F.; Truesdell, A. H. Water Resour.-Invest. U.S. Geol. Suru. 1976, 76, 13. (21) Truesdell, A. H.; Jones, B. F. J. Res. U S . Geol. Suru. 1974, 2, 233-48. (22) Berner, R. A. Am. J. Sci. 1967, 265, 773-785. (23) Davison, W. Geochim. Cosmochim. Acta 1980,44,803-808. (24) Berner, R. A. “Principles of Chemical Sedimentology”; McGraw-Hill: New York, 1971. (25) Emerson, S. Geochim. Cosmochim. Acta 1976,40,925-934. (26) Robie, R. A.; Hemingway, B. S.; Fisher, J. R. U.S. Geol. Surv. Bull. 1978, 1452. (27) Johnson, K. S. Geochim. Cosmochim. Acta 1982, 46, 1805-1809. (28) Jassby, A.; Powell, T. Limnol. Oceanogr. 1975,20,530-543. (29) Stauffer, R. E. “The Estimation of Heat Budgets, Thermocline Migration, and Vertical Eddy Conductivities in Stratified Lakes”; monograph, Water Chemistry Laboratory, University of Wisconsin: Madison, WI, 1983. (30) Birge, E. A.; Juday, C.; March, H. W. Trans. Wis. Acad. Sci. Arts Lett. 1928, 23, 187-232. (31) Likens, G. E.; Johnson, N. M. Limnol. Oceanogr. 1969,14, 115-135. (32) Cross, P. M.; Rigler, F. H. Can. J . Fish. Aquat. Sci. 1983, 40, 1589-1597. (33) Delfino, J. J.; Bortleson, G. C.; Lee, G. F. Enuiron. Sci. Technol. 1969, 3, 1189-1192. (34) Lathrop, R. “Water Quality Conditions, Dane County, Appendix B”; Dane County Regional Planning Commission and Wisconsin Department of Natural Resources, Madison, WI, 1978; 174 pp. (35) Cochran, W. G. “Sampling Techniques”;Wiley: New York, 1963. (36) Hutchinson, G. E. “A Treatise on Limnology, Vol. 1”;Wiley: New York, 1957. (37) Hesslein, R. H. Can. J. Fish. Aquat. Sci. 1980,37, 552-558. (38) Davison, W.; Woof, C. Water Res. 1984, 18, 727-734. (39) Brezohik, P.; Lee, G. F. Enuiron. Sci. Technol. 1968, 2, 120-125. (40) Stauffer, R. E. Limnol. Oceanogr. 1980, 25, 513-528. (41) Hem. J. D. “Redox Coprecipitation Mechanisms of Manganese Oxides”;U.S. Geological Survey: Menlo Park, CA, 1979. (42) Bricker, 0. P. Am. Mineral. 1965, 50, 1296-1354. (43) Morgan, J. J. “Principles and Applications of Water Chemistry”; Faust, S. D., Hunter, J. V., Eds.; Wiley: New York, 1967; pp 561-622. (44) Stauffer, R. E. Ph.D. Thesis, University of Wisconsin, Madison, WI, 1974. (45) Fallon, R. D.; Brock, T. D. Limnol. Oceanogr. 1980, 25, 72-88. (46) Li, Y. H.; Gregory, S. Geochim. Cosmochim. Acta 1974,38, 703-714. (47) Davison, W. Nature (London)1981,290, 241-243. (48) Spencer, D. W.; Brewer, P. G. J. Geophys. Res. 1971, 76, 5877-5892. (49) Sundby, B.; Silverberg, N.; Chesselet, R. Geochim. Cosmochim. Acta 1981, 45, 293-307. (50) Mortimer, C. H. Ecology 1941-1942,29 and 30, 280-329, 147-200. (51) Kjensmo, J. Arch. Hydrobiol. Suppl. 1967, 32, 137-312. (52) Sasseville, D. R.; Norton, S. A. Limnol. Oceanogr. 1975,20, 699-714. (53) Bortleson, G. C.; Lee, G. F. Enuiron. Sci. Technol. 1972, 6, 799-808. (54) Holdren, G. C. Ph.D. Thesis, University of Wisconsin, Madison, WI, 1977.
Environ. Sci. Technol. 1886, 20,457-463
Theis, T. L.; Singer, P. C. Environ. Sei. Technol. 1974,8, 569-573. James, H. R.; Birge, E. A. Trans. Wis. Acad. Sei. Arts Lett. 1938, 31, 1-154. Jones, B. F.; Bowser, C. J. In "Lakes: Chemistry, Geology, Physics"; Lerman, A., Ed.; Springer-Verlag: New York, 1978; Chapter 7.
(58) Nriagu, J. 0.;Dell, C. I. Am. Mineral. 1974, 59, 934-946.
Received for review October 28, 1983. Revised manuscript received May 10,1985. Accepted October 21,1985. Partial research support was provided by U S E P A Research Grant R805281 and by the private K A M Foundation.
Distribution of Chlorophenolics in a Marine Environment Tlan-Mln Xle,+ Katarlna Abrahamsson, Ellsabet Fogelqvlst, and Bjorn Josefsson" Department of Analytical and Marine Chemistry, Chalmers University of Technology and University of Goteborg, S-41296 Goteborg, Sweden
w The distribution of some chlorophenols, chloroguaiacols, and chlorocatechols, which were discharged from a sulfate pulp mill, were studied in the Gulf of Bothnia. These compounds were determined both in the water column and in the bulk sediments to understand their degradation pathways and accumulation patterns in the sea. Chloroform was used as a tracer to monitor the distribution and dilution of the effluent plume. The results showed that the transport of chlorophenolics was dominated by dilution and adsorption processes. Their behavior in the receiving water was correlated to the lipophilicity of the chlorophenolics, described by their partition coefficients in an octanol-water system (corrected for pH). A strong influence of bioactivity on the fate of chloroguaiacols and chlorocatechols was observed in sediments containing large amounts of organic carbon.
Introduction The release of industrially derived halogenated organic compounds into the aquatic environment is of great concern, mainly because of their toxicity, resistance to degradation, and tendency to bioaccumulate. The wood pulp industries, which use different chlorine bleaching processes, are one of the main sources of organic chlorinated compounds ( I and references cited therein). Chlorophenolic compounds have been found to be major constituents produced from lignin residues (2-6). Some of them are toxic and can be accumulated in living organisms (7-9). The transport and concentrations of some chlorophenols have been investigated in rivers, lakes, and estuaries (10-12). Recently, some chlorocatechols were determined in seawater and sediment close to pulp mills (13). Still, there is little known about the marine biogeochemistry of these types of compounds. Pulp mill effluent plumes are traditionally investigated by using artificially introduced tracers, such as radioactive bromide or dyes, e.g., rhodamine (14). Recently, Fogelqvist et al. utilized chloroform as a tracer substance in marine receiving waters (15). Chloroform is produced during chlorine bleaching in pulp mills in such a high concentration that it is possible, with this method, to monitor dilutions up to 10 000 times. Chloroform determinations are simple and fast, thereby, a large recipient area can be investigated in a short time period to provide adequate statistical information. The adsorption of organic pollutants onto particles is mainly dependent on their lipophilicity, described by their log P values-the logarithm of the partition coefficient in Permanent address: Academy of Science of the Ministry of Light Industry, Beijing, China. 0013-936X186/0920-0457$01.50/0
the octanol-water system (16-19). The use of log P values to investigate the transportation of organic pollutants in a river has been reported (20). The log P concept is valuable in explaining the behavior of the chlorophenolic constituents discharged into the sea. It should be noticed that the log P values of the chlorophenolic compounds investigated are in the range 3-5, which are considerably lower than the corresponding values of other organic pollutants such as DDT, PCBs, and hydrocarbons (21). The biogeochemistry of chlorophenolics is therefore expected to be different. Investigations of the distribution of chlorophenolics in a relatively large area in a marine environment are rare, mainly due to the difficulty of analyzing large numbers of samples at low concentrations under the often complex hydrologic conditions (compared to rivers). The lipophilicity parameters of chlorophenolics (22,23), as well as the simple and sensitive GC methods for determining trace amounts of chlorophenolics in water and sediments, presented recently (24,25), offer the opportunity to conduct such investigations. The objective of this work was to establish the distribution pattern of different chlorophenolic compounds over a large marine recipient area, both in water and sediment, and compare this pattern with the effluent distribution pattern determined by using chloroform as a tracer.
Experimental Section Nature of the Sampling Area. The sampling area, located in the Gulf of Bothnia off the East Coast of Sweden, is characterized by several islands and skerries. This complicates the determination of a distribution pattern of the effluent. However, the following circumstances dominate the transport mechanism. The effluent (1m3/s) from the plant is mixed with seawater in a shallow receiving area. Into this pond seawater flows from the south and is pumped to the north at a rate of 20 m3/s. The effluent water, thus diluted 20 times, enters a bay surrounded by small islands. The seabed in the near recipient is dominated by transportation bottoms. However, accumulation of sediment takes place in patches all over the area. The sampling was performed on four occasions: (1) September 1982 (seawater and sediment); (2) March 1983 (effluent); (3) May 1983 (sediment); and (4) November 1983 (effluent, seawater, and sediment). Details of sampling locations, dates, and types are given in Figure 1. To avoid contamination artifacts or adsorptive losses, unfiltered water samples were used. Materials. The compounds measured were as follows: 2,4-dichlorophenol (DCP), 2,4,6-trichlorophenol (TrCP), 2,3,4,6-tetrachlorophenol(TeCP), pentachlorophenol
0 1986 American Chemical Society
Environ. Sci. Technol., Vol. 20, No. 5, 1986
457
