A study of the degradation of phenoxyacid herbicides at different sites in a limestone aquifer

Pergamon

Chemosphere,Vol.36, No. 6. pp. 1211-1232, 1998 © 1998ElsevierScienceLtd All fightsreserved.Printedin GreatBritain 0045-6535/98 $19.00+0.00

PH: S0045-6535(97)10043-1 A STUDY OF THE DEGRADATION OF PHENOXYACID HERBICIDES AT DIFFERENT SITES IN A LIMESTONE AQUIFER.

IAN HARRISON*, RACHEL U. LEADER, JENNY J.W. HIGGO & GEOFFREY M. WILLIAMS Fluid Processes & Waste Management Group, British Geological Survey, Kingsley Dunham Centre, Keyworth, Nottingham, U.K. (Receivedin Germany29 April1997;accepted10 September1997) ABSTRACT The biodegradation of three phenoxyalkanoic acid herbicides, viz. MCPA, dichlorprop and mecoprop, under aerobic and anaerobic conditions was investigated using microcosm techniques. The field studies were conducted in a limestone aquifer that had suffered contamination from leaking landfill sites in which phenoxyaikanoic acid herbicides (mostly mecoprop) had undergone disposal. The results from/n situ field and laboratory microcosms indicated that under microbially active aerobic conditions the biotransformafion of all three herbicides was rapid and that lag phases were short. Under fully aerobic conditions the concentration of each of the three herbicides was reduced from 2,000 lag/l to below the detection limit (approx. 10 lag/l) of the HPLC system, used for their analysis, within 14 days. However, under microbially active anaerobic conditions no degradation of the herbicides could be discerned over the 100-200 day duration of the experiments. This finding has significant implications for the disposal of phenoxyalkanoic acid herbicides particularly in situations where any resulting leachate may find its way into underlying water resources. ©1998 Elsevier Science Ltd INTRODUCTION During the 1940s, chlorinated phenoxyacetic acids, e.g. 2,4-D, 2,4,5-T and MCPA (Fig. 1), were introduced as selective herbicides against broadleaf weeds. Typically, they were used for weed control on cereal crops, grasslands and lawns. They remain among the most commonly applied herbicides and are still commercially available as lawn weed killers. In the 1950s and 60s chlorinated phenoxypropionic acids, e.g. 2,4-DP (dichlorprop) and MCPP (mecoprop) (Fig. 1), were found to be more effective against some weed varieties than the chloro-phenoxyacetic acids and were used in conjunction with them. Chloro-phenoxyalkanoic acid herbicides are applied either in the form of salts or esters, and are rapidly metabolized by plants. The salts have high water solubilities and are used in the form of their aqueous concentrates. Because they are anionic species that do not sorb appreciably to soil or its constituents, they tend to leach into the subsurface. The conclusion of a study of the leachates from six municipal landfills in the USA was that chlorinated phenoxypropionic acids, particularly mecoprop, are more persistent than the chlorinated phenoxyacetic acids. They were found in all the samples at concentrations up to 90 lag/l [Gintautus et al., 1992]. * Corresponding author 1211

1212 CH2 - COOH

CH2 - COOH

CH2- COOH

I

I

I

CH3- CH -COOH

I

CH3- CH - COOH

I

O

O

O

O

O

CI

CI

CI

C1

C1

2,4-D 2,4-dichlorophenoxyaceticacid

2,4,5-T MCPA 2,4-DP MCPP 2,4,5-trichloro- 2-methyl-4-chloro- 'Dichlorprop' 'Mecoprop' phenoxyaceticacid phenoxyaceticacid 2-(2,4-dichloro- 2-(4-chloro-o-tolylphenoxy)propionicacid oxy)propionicacid

Figure I.

Phenoxyalkanoic acid herbicides.

While studies of their degradation in the topsoil suggest that collectively they degrade fairly readily [Smith, 1989], the fate of the material leached into the subsurface is less well understood and comparatively limited information is available on transformation in groundwaters [e.g. Gibson & Suflita, 1990; Agertved et al., 1992; Heron & Christensen, 1992; Hughes & Gardner, 1992; Klint et al., 1993]. As a potential source of contamination to aquifers and groundwater resources, there is a pressing need to investigate the fate of phenoxyalkanoic acid herbicides in subsurface environments, particularly as there is evidence of mutagenicity in certain of the compounds [Seiler, 1978]. In this study the degradation of three widely-used phenoxyalkanoic acid herbicides, i.e. MCPA, 2,4-DP (dichlorprop) and MCPP (mecoprop), in a limestone aquifer was monitored in both the field and the laboratory using a variety of microcosm techniques.

At the same time the behaviour of a range of priority organic

pollutants (aromatic hydrocarbons, substituted aromatics and chlorinated hydrocarbons) was followed. The priority pollutant degradation behaviour is described elsewhere [Higgo et al., 1994; Harrison et al., 1995; Higgo et al., 1996]. This paper focuses on the herbicides. The aquifer, in which the experiments were conducted, occurs in the Lincolnshire Limestone and is extensively exploited by Anglian Water for the supply of drinking water in the north Peterborough area. In 1987 groundwater abstracted from one of the company's boreholes, located in the confined aquifer, began to show increasing concentrations of the herbicide mecoprop [ENDS Report, 1992]. By 1990 the concentration had reached 3.9 I.tg/1 and two years later was 8 I.tg/1, well above the 0.1 I.tg/l limit for individual pesticide compounds set in the 1980 EEC directive on drinking water (80/778/EEC). A granular activated carbon (GAC) treatment plant was installed and this reduced concentrations in the abstracted water to 0.01 I.tg/1. The source of the mecoprop was eventually traced to two landfills, situated in an outcrop area of the limestone, that had taken in pesticide washings until the mid-1980s [ENDS Report, 1994]. Mecoprop concentrations of up to 39,000 I,tg/1 were detected in and around the landfills. A geological fault, considered to act as a hydraulic barrier between the outcrop area and the confined section of the aquifer, did not in fact perform this function. Groundwater flow across the fault was confirmed and herbicide concentrations of up to 1,000 I.tg/1 were found in abstractions to the north of the fault [ENDS Report, 1992].

1213 The effect of redox conditions upon the degradation of the three chosen herbicides and a range of priority pollutants was investigated in two boreholes sunk into the limestone (Bhl & Bh2).

The first (Bhl) was

adjacent to the landfill site. The groundwater was reducing with no measurable dissolved oxygen and an appreciable concentration of Fe(U). The second borehole (Bh2) was on the other side of the geological fault. Here conditions were aerobic and the groundwater contained 2.0 mg/l dissolved oxygen and little or no Fe(II). The compositions of both groundwaters are given in Table 1. Substitution of the groundwater data into the EQ3NR computer code [Wolery, 1992] indicated that for Bhl the system was just saturated with respect to siderite (saturation index = 0.016) whereas for Bh2 it was undersaturated (saturation index = -1.621). However, the presence of organic matter influences the behaviour of iron via complex formation and chelation [Hem, 1960] and it is probable that both groundwaters are undersaturated with respect to siderite.

Table 1.

Groundwater characteristics at the boreholes in the limestone aquifer.

Reducing Bh 1 pH 7.0

Aerobic Bh2

7.0

7.0

< 0.1

2.0

C1- (mg/1)

640

98

SO42- (mg/l)

670

356

HCO3- (mg/l)

520

378

HPO42-(mg/l)

< 0.2

< 0.2

NO3- (mg/l)

< 1.0

7.9

NO2- (mg/l)

< 0.3

0.2

NH4÷(mg/l)

48

2.2

Fe2+ (rag/l)

21

< 0.6

Mn2÷ (mg/l)

4.4

0.04

Zn2* (rag/l)

0.079

0.005

Ca2+ (rag/l)

285

247

Mg2÷ (mg/l)

43

13

Na÷ (rag/l)

297

49

K÷ (mg/1)

12

3.2

Dissolved 02 (mg/1)

TOC (mg/l) Depth to water (m below ground level)

26

2.8

6.5 - 8.5

6.5 - 13.5

Degradation experiments were carried out in the field using specially designed "borehole microcosms" (BHMs) and in the laboratory, on a smaller scale, using laboratory column microcosms (LCMs).

The latter

were designed to simulate the BHMs as closely as possible and had the advantage of economy and convenience but the disadvantage that, using them, it was more difficult to maintain in situ conditions.

1214 MATER/ALS AND METHODS

Herbicides and Priority pollutants MCPA, 2,4-DP and MCPP were purchased from Promochem, Welwyn Garden City, Herts., U.K. Synthesized by the Institute of Organic Industrial Chemistry, Warsaw they were certified to be of > 99.5% purity. The following priority pollutants were added to the groundwaters:benzene, toluene, o-xylene, naphthalene, phenol, o-cresol, nitrobenzene, 2,4-dichlorophenol, 2,6-dichlorophenol, o-nitrophenol, p-nitrophenol, 1,4-dichlorobenzene, 1,2-dichlorobenzene, tetrachloromethane, 1,1,1-trichloroethane, trichloroethylene, tetrachloroethylene. They were all analytical or chromatography grade.

FIFE Siphon tl

undwaterReservoir ml SyringeBarrel blanked off)

'Quick-couplin[ SamplingValw Loading and plenishmentLine SamplingLine

StainlessSteel End-piece

ranulatedAquifer Material

tainless Steel Tube 65 cm x 5 cm i.d.) Stainless Steel "I (lmm i.d.) tainless Steel Mesh Bed Support

StainlessSteel End-piece

Figure 2.

Drilled Galleries

Borehole Microcosm (BHM)

Field Microcosms The BHMs (Figure 2) were adaptations of the in situ microcosm (ISM) design of Gillham and co-workers [1990]. They were essentially columns of geologic media encased in stainless steel cylinders (6.5 cm x 65 cm) which could be loaded with spiked groundwater and installed in boreholes where degradation under

1215 in situ conditions could be followed. The main difference between an ISM and a BHM is that the former is

open at the base and is driven into unconsolidated fine- to medium-grained sediments, whereas the latter is an enclosed cylinder that may be filled with any type of solid granular matrix. This allows the BHM to be used in boreholes in coarse-grained or consolidated recks. Experiments were conducted in which BHMs were filled with either indigenous limestone (crushed and sieved to give particles nominally 1 mm to 1.5 mm in diameter) or with acid-washed DansandTM, a coarse sandy material, predominantly quartz particles with nominal diameters 0.9 mm to 2.0 mm (Sand No. 3, Dansand Silkeborg A/S, Horsens, Denmark). The acid washing removed easily dissolved acid-soluble substances but left some of the iron compounds and the organic carbon content intact. The primary purpose of the aquifer solids was to provide a substrate for microbial colonisation and previous work [Higgo et al, 1996] had shown that it was not essential to use material from the aquifer being studied. Limestone tended to disintegrate and block up the tubing and was only used in one pair of BHMs in the aerobic zone. Degradation experiments were generally carried out using pairs of microcosms, one microbially active and one sterile control. They were slurry packed with the selected matrix and then flushed with groundwater which was pumped from the recently purged borehole into a Tedlar® bag and then into the microcosm via the sampling line. After flushing the sterile (control) and non-sterile BHMs were harnessed together, lowered into the borehole and left immersed in the groundwater. Thereafter, they were brought to the surface every two weeks or so, for the next three months, and flushed with fresh groundwater. This allowed the material within both microcosms to equilibrate with the groundwater and also permitted their colonisation by the indigenous microbial community. Loading of the BHMs with the compounds of interest was achieved by a modification of the flushing procedure.

Thus, once the Tedlar® bag had been filled with fresh groundwater, a concentrated aqueous

cocktail of the compounds together with tritiated water (to act as a conservative tracer), was injected into the bag through a self-sealing, PTFE-faced septum. For the sterile BHM, sodium azide was additionally incorporated into the loading mixture (to give a final concentration of 0.2% w/v in the BHM). After agitation of the bag contents to provide mixing, the spiked groundwater was pumped, by peristaltic pump, at 50 ml/min, through the loading/replenishment line. Effluent samples were taken at regular intervals during loading and breakthrough curves were plotted. The initial concentration, Co, of each compound in the microcosm was determined from samples taken directly from the Tedlar® bag (herbicides = 2,000 ktg/1 in the aerobic zone or 7000 ktg/1 in the anaerobic zone, priority pollutants =150 p.g/1). Finally, after the 5 1of spiked groundwater had been passed through the microcosm, the BHM was reinstalled in the borehole and degradation followed by sampling at preset intervals. At predetermined times, the microcosms were raised for sampling. A glass syringe with a suitable quick-coupling on its hub was attached to the sampling valve. A sufficient amount (5 ml) of water, to flush out the dead volume of the sampling line etc., was withdrawn from the BHM and discarded.

1216 A further sample (20 ml) was withdrawn and dispensed into suitable containers for laboratory analysis l. During sampling the water withdrawn from the bottom of the microcosm was replaced at the top from the groundwater reservoir. Typically, a total of 350 to 400 ml of water could be taken from the BHMs before a significant decrease in C/C0 of tritium indicated that dilution with unspiked groundwater was beginning. When this occurred the experiments were terminated.

Mininert® V aafter lv~ for sampling Loading Stopcock

Mininert® Valve for sampling Post-column during Loading

~im~

Measuring Cylinder to collect effluent during Loading ~

Refrigerator/ 4°C Tedlar@ Bag with ~''~aplKeaGroundwater

1

~'-

~:::~

.~

m

m

"'~: " / x. Mininert® Valve for nn

l

Jacketed Column (Pharmacia XK 50/30) with Stainless Steel Mesh Bed Supports packed with Granulated Aquifer Material.

~

~

Cooling water 10°C

Thermostatted Water Bath 10°C

i oie ainles Steel Tubing

Figure 3. Laboratory Column Microcosm (LCM) Laboratory Column Microcosms

Designed to simulate the BHMs on a smaller scale, the LCMs (Figure 3) consisted of 30 cm x 5 cm i.d. jacketed glass chromatography columns (Pharmacia U.K. Ltd.) and were packed with the same media used in the field microcosms.

! (a) For GC and HPLC analysis a Quickfit® glass test tube (ca. 12 ml), containing 10 I.tl of 10M NaOH was filled leaving no headspace. (b) For tritium analysis a 2.5 ml of filtered (0.45 p_m)sample was placed in a plastic vial. (c) For Fe(II) determinations 4.5 ml of filtered (0.45 [am) sample was added to 0.5 ml of 1% w/v aqueous 2,2"-dipyridyl in a graduated acrylic tube. (d) Microbial analysis was performed on 1 ml unfiltered samples placed in a Sterilin tube (aerobic zone) or a 'no headspace' sample in an autoclaved 3 ml Quickfit® glass-stoppered test tube (reducing zone). The samples were maintained at ca. 4°C in cool boxes while in transit to the laboratory. All analyses were performed as soon after collection as possible, usually within 24 hours. When not in use samples were stored in the dark at 4°C in a refrigerator.

1217 The bed supports supplied were replaced with stainless steel meshes and Mininert valves were installed at the entrance and exit of each column. Equilibration of columns with aerobic groundwater from Bh2 was undertaken in the laboratory using groundwater collected from the borehole in amber glass Winchesters (maintained at 4°C in the dark until required). Equilibration of the column with the anaerobic groundwater from Bhl was performed in situ by attaching a small caravan submersible pump (Whale Supersub 921) to the column inlet. The column was lowered into Bhl and twice a week one litre (ca. 4 pore volumes) of groundwater was pumped through. After two months, the column was retrieved from the borehole, plugged at the ends, and placed in a cool box through which a continuous stream of nitrogen was passed. After retrieving the LCM the borehole was purged and then fresh groundwater was pumped into a 5 I nitrogen-filled stainless-steel kettle. Inserted in the filling line was a septum through which an aqueous spike concentrate was injected during filling.

BHM and kettle were returned to the laboratory where they were stored overnight in a

nitrogen-flushed incubator at 10°C (i.e. in situ groundwater temperature). The next day the LCM was loaded with spiked groundwater directly from the kettle, which was pressurised with nitrogen to establish a flow

(ca. 2 ml/min) through the column. During loading, samples for analyses were taken from the column inlet and outlet, in a similar manner to the field microcosms, to permit the construction of breakthrough curves. After loading, a small 1 1 Tedlar® bag containing groundwater was connected by means of a length of nylon tubing to the inlet of the LCM to act as the groundwater reservoir. At the outlet was a Mininert® valve (Fig. 3) that could be used for taking samples at set intervals. Before and after sampling the column was stored in the nitrogen-flushed incubator at 10°C. Because of the technical difficulties encountered with running an anaerobic LCM only a non-sterile experiment was performed. The behaviour of a sterile LCM was inferred from the behaviour of the sterile BHM in Bhl. The LCMs, used to simulate the BHMs in Bh2, were loaded in a similar fashion except that the spiked groundwater was contained in a Tedlar® bag, as air exclusion was not necessary.

Analytical Procedures The gas chromatographic method, used to determine the seventeen priority pollutants, present in addition to the herbicides in most of the microcosms, has been described in detail [Harrison et al., 1994]. During the GC method two extractions are performed, one into pentane and the other, after derivatisation, into a mixture of npentane and diethyl ether. This has the benefit of removing low polarity compounds and phenols that would otherwise interfere with the reversed-phase HPLC analysis of the herbicides. The aqueous phase remaining after these extractions (or a sample that contained only the herbicides) was adjusted to pH 3.0 + 0.5 with 50% v/v aqueous o-phosphoric acid. A 500 I.tl sample was then injected (Rheodyne 7125) onto a reversed-phase column (Nucleosil 5 I.tm C18, 250 m m x 4.6 mm i.d.). The herbicides were eluted with a mobile phase composed of 380 ml 1:1 acetonitrile/methanol and 620 ml of 5mM aqueous potassium dihydrogen orthophosphate containing 20 I.tl of glacial acetic acid. The mobile phase flow rate was 1 ml/min and the herbicides were detected at 226 nm. The HPLC method was adapted from a report on phenoxyacid herbicide analysis [Schuster & Gratzfeld-Hiisgren, 1990]. Tritium was determined on a liquid scintillation counter (LKB 1219 Rackbeta). Bacterial numbers of heterotrophs were estimated using a conventional spread-plate technique based on PTYG

1218 medium [Balkwill & Ghiorse, 1985]. A small sample of spiked groundwater (100 Ixl) was inoculated onto sterile plates of media which were then incubated at 25°C in aerobic, and where appropriate, anaerobic atmospheres. The plates were examined for bacterial growths over a five to fourteen day period and the number of colony-forming units (CFU) noted. Geochemical analysis (Table 1) for cations and anions was performed on suitably preserved samples. Cations were determined by inductively-coupled plasma optical emission spectroscopy (ICP-OES) and anions by ion chromatography (Dionex Ltd.). Iron (II) was determined by measuring the absorption of the red Fe(II) dipyridyl complex at 520 nm on a double-beam, UV/visible spectrophotometer (Pye Unicam SP8-100). The instrument provided a linear calibration for standards in the range 0 - 5 mg/1 Fe(II). Dissolved oxygen and pH were measured in the field, the former both electrochemically (Orion Ltd.) and titrimetrically by Alsterberg's modification of Winkler's method [Rodier, 1975]. Limits of detection were about 0.02 ppm and 0.5 ppm for Fe(II) and oxygen, respectively. In general either oxygen or Fe(II) were detectable but not both. One pair (sterile and non-sterile) of Dansand-filled BHMs was installed in each of the two boreholes, Bhl (reducing zone) and Bh2 (aerobic zone). In addition a pair of limestone-filled BHMs was installed in Bh2. The LCMs were all filled with Dansand. Two pairs used aerated groundwater from Bh2 and one, designed to simulate the reducing BHM, was equilibrated in situ in Bhl before loading with spiked groundwater from the borehole. Since the MCPP concentration in Bhl was already 7,000 lxg/i it was not used to spike water from this borehole. In Bh2 MCPP levels were about 40 ~tg/l and sufficient MCPP was added to water from this borehole to give a concentration of about 2,000 ~tg/l in the loaded microcosms. RESULTS

Breakthrough Curves Breakthrough curves were plotted during loading of all the microcosms (Fig. 4). In all cases the breakthrough of the herbicides (except MCPP in Bhl since it was already present in the groundwater) corresponded closely to that of the tritium, indicating little if any sorption. However, C/C0 values rose much more slowly in the limestone-filled microcosms than in the Dansand-filled microcosms. This reflected the greater intragranular porosity of the limestone particles compared with the sand particles.

Degradation Curves (i). Borehole Microcosms in the Reducing Zone (Bhl ) Throughout loading, the Fe(II) concentration remained steady, at about 20 mg/l, in both microcosms. However, it fell rapidly to 2 mg/l during the first 35 days of the degradation period. Thereafter, it rose slowly and was constant at 15 mg/I between days 110 and 140, before falling again to ca. 2 mg/l. The fate curves for the herbicides, plotted together with Fe(U) concentrations, are shown in Figures 5a & 5b. In the sterile BHM, C/Co for the herbicides remained constant at about 1.0 within the bounds of experimental and analytical errors (ca. + 20%), and the Fe(II) concentration fell slowly from 18 mg/l to 8 mg/1 over a 190 day period.

1219 C/Co Tritium & Herbicides

Fe(II) mg/1 3O

1.4

1.2

-25

1.0 -20 0.8 -15 [].

0.6

---__

_ --

0.4

E

I

E-

tli-

_

,[]

-10

5

0.2

0.0

J_

0 10 '00

2 0 0' 0

3 0 0' 0

4 0'00

5000

Volumeof SpikedGroundwaterLoaded(ml) .... • .... o-------

-

Figure 4

Tritium

•

MCPA

2,4-DP

a

MCPP

-

-

m -

-

-

F e ( I I )

Typical breakthrough curve (non-sterile anaerobic LCM containing Dansand equilibrated in situ at Bhl).

The decrease in concentration was probably attributable to abiotic reduction of some of the organic compounds in which Fe(II), acting as a reducing agent, was oxidised. The added nitro- compounds, viz. nitrobenzene, o- and p-nitrophenols, disappeared in both the sterile and non-sterile BHMs, as did tetrachloromethane. Reduction of the nitro- group to the amino- group is a six electron process [Weber, 1994] which means that 1 mg of nitrobenzene requires about 3 mg Fe(II) for complete reduction. Thus, nitro-reduction of added nitrocompounds would account for about a 1.5 mg/l drop in the concentration of Fe(II). The reductive dehalogenation of tetrachloromethane would require very much less, ca. 0.1 mg/l Fe(II). The non-sterile BHM was found to be microbiologically active throughout the experiment and samples typically gave counts of the order of 103 to 104 CFU/ml. However, the herbicide concentration remained constant for the first 140 days indicating no degradation. In fact, only one compound, toluene, showed clear evidence of biodegradation. In the non-sterile microcosm C/C0 decreased from one to zero during the first 110 days (see Figure 5a) but in the sterile microcosm it remained constant (C/C0 = 1).

1220 C / C o Tritium Herbicides & Toluene

1.4

Fe(II) mg/l

III ~

25

(a)

1.2 -20 1.0 r

'

~

r

.

~

l

~

.

.

I

-

'-m

-

i

-

...........

~

.m . . . . .

~

~

__,.,

15

0.8 0.6

10

0.40.20.0

I

I

50

100

-

-r

-~ ,4._

I 150

0 200

Time from Loading (days) Fe(n) mg/l

C/Co Tritium & Herbicides 1.4

25

1

.

2

-

~

(b) -20

1.015

0.8-

i---_ ..... 0.6-

10 .... • ....

Tritium

•

MCPA

2,4-DP

A

MCPP

Fe(ll)

~-

Toluene

--!

0.40.2-

- - - []- - -

0.0

0 5'0

' 100

' 150

200

Time from Loading (days)

Figure 5

Degradation curves for (a) non-sterile anaerobic Dansand BHM and (b) sterile anaerobic Dansand BHM, using groundwater from Bhl.

1221 After 140 days only two samples were taken because by this time strong suction had to be applied in order to remove samples from the BHM2 and it was feared that, as a result, air was entering the microcosms. The fact that the concentrations of herbicides and Fe(II) were low in these samples supports this theory, i.e. aerobic degradation of the herbicides and aerial oxidation of Fe(ll) was occurring. There was evidence also of biodegradation of compounds other than the herbicides, thus, C/C0 for benzene, naphthalene, phenol, and o-cresol fell significantly during this period.

(ii). BHMs in the aerobic zone (Bh2) In both experiments in this borehole the herbicide concentration remained constant (C/C0 --- 1) in the sterile microcosms (Figures 6b & 7b), but the non-sterile BHMs behaved quite differently. In the non-sterile sand-filled BHM, the herbicide concentrations remained constant for the first seven days and then fell rapidly from C/C0 = 1 to a C/C0 between 0.1 and 0.2.

Thereafter, they rose slowly and steadily

until, after 150 days, C/C0 values were between 0.8 and 1.0. Some of the other organic pollutants showed a similar, if somewhat less pronounced, pattern to the herbicides. Of these 2,4-dichlorophenol was the most markedly similar and, after the initial fall in concentration, rose from C/C0 = 0.2 to 0.8, whereas the others, i.e. benzene, toluene and o-xylene displayed much more modest rises. In this BHM the Fe(II) concentration remained below detection limits for the first 14 days but then rose to 5 mg/1, in parallel with the rise in concentration of the herbicides and the other organic compounds (Figure 6a). In the sterile BHM there was no sign of degradation and the Fe(II) concentration remained zero throughout the experiment (Figure 6b). In the non-sterile limestone-filled BHM (Figure 7a) there were no significant changes in the herbicide concentrations during the 200 day degradation period. There was a small increase in Fe(II) concentration (from 0 to 2 mg/l ) during the first 14 days. This slowly decreased to 1 mg/l by the end of the experiment.

( iii). Anaerobic laboratory microcosm The fate curves (Figure 8) show no degradation of the herbicides and, unlike the corresponding BHM in which toluene degraded, no biodegradation of any of the other added organics. There was evidence of abiotic transformation of tetrachloromethane, nitrobenzene, o-nitrophenoi and p-nitrophenoL

The Fe(II)

concentration fell rapidly during breakthrough from 20 mg/l levelling out at about 8 mg/l by the time 5 1 of spiked groundwater had been loaded (Figure 4). During the degradation period it rose steadily from 8 mg/l to 17 mg/1 (Figure 8). Microbial counts were of the order of 10 CFU/ml after loading, rising to 102 to l03 CFU/ml after 35 days, and maintaining that level thereafter.

2 This was caused by the gradualdepositionof silty material,present in the borehole,which settled in the groundwaterreservoiras an impermeablelayer. The reservoirsin the Bhl BHMs were directlyconnectr.xlto the loading replenishmentlines by the LuerTM fittings oftbe syringe barrels. In subsequentBHMs this problem was overcome by couplingthe reservoirto the BHM by the PTFE siphon tube, as shown (Fig. 2) and blankingoff the LuerTM fittings.

1222 C/Co Tritium & Herbicides

Fe(U) mgn

1.2~

_ •

1.0

~

a

)

3O

z n .

"" " l " - ' i - - I " "

""'8

.....

n--

_

.....

-25 -20

0.6

15

0.4

10

0.2

\W" ,,/-

--'~-

0.0

_ _-=---

i

0

-i-

- ~ -

=-"

= -. . . . .

i

50

5

=

0

i

100 Time fromLoading(days)

150

2OO

C/Co T r i t i u m & Herbicides

Fe(II)

mg/!

1.2

30

1.0 -

~

(b)

-25

0.8-

-20

0.60.4-

.... • ....

Tritium

#

MCPA

----o

2,4-DP

&

MCPP

- - - n-

- -

-15 -10

Fe(lI)

0.20.0

5 n . . . . _. . .

0

Figure 6

n

. . . . . -. . . . i

50

i - - - ~

. . . . . . .~ . . . I . . . . . . . • . . . .

100 Time fromLoading(days)

t

. I. . . . . . . .

150

0

•

2OO

Degradation curves for (a) non-sterile aerobic Dansand BHM and (b) sterile aerobic Dansand BHM, using groundwater from Bh2.

1223

C/Co Tritium & Herbicides

Fe(II) mg/l

1,2 ,

(a)

. 30

1 . 0 ~

(8)

~25

0.8-

~

-

-

'

l

'

'

-

'

~

20

0.6-

-15

0.4-

-I0

0.2-

5 n--n--

0.0

0

l-~-t---t---n---

. . . . . . t. . . . . . . . . . . . . t. . . . . . . . . . . . t. . . .

I

I

I

50

100

150

0 200

Time from Loading (days) C/Co Tritium & Herbicides

Fe(H) mgh

1.2

30 (b)

1.o

-25

0.8

-20

0.6

-15 .... g ....

0.4-

o---0.2-

Tritium

•

MCPA

2,4-DP

&

MCPP

-10

5

---U--- Fe(II)

0.0

........... 0

-'--i- . . . . . . . . . . . . 50

"----~ .... 100

-". . . . . . . . . .

-'I2"l~~ - - - - I 150

0 200

Time from Loading (days)

Figure 7

Degradation

c u r v e s f o r (a) n o n - s t e r i l e a e r o b i c l i m e s t o n e B H M

(b) sterile aerobic limestone BHM, using groundwater

and

from Bh2.

1224

C/Co Tritium & Herbicides

~(/I)

mg/l

1.4

30

=25

1.0

. ". . . . . .

"2:'2---20

............................

0.8.

m

[]

-15

0.6-

Tritium

o

MCPA

2,4-DP

,0,

MCPP

0.4. o-------

-10

5

0.2. - - - m- - -

Fe(lI)

0.0 2i0

40

6i0

I 80

0 100

TimefromLoading(days)

Figure 8

Degradation curve for non-sterile anaerobic Dansand LCM equilibrated in situ at Bhl.

(iv). Aerobic laboratory microcosms In these LCMs only the three herbicides and tritium were loaded. Aerated groundwater was used so that the effect of a fully aerobic redox environment upon the degradation of the herbicides could be examined. Accordingly, the dissolved oxygen (dO2) content of samples from the columns, was used to monitor the prevailing redox conditions within the LCMs. In the sterile LCM (Figure 9b) the dOz content remained in the range 8 - 10 mg/1 during the whole experimental period, but in the non-sterile LCM (Figure 9a) it dropped rapidly after loading to reach 1 mg/1 after 14 days. Thereafter, it appeared to diminish slowly and was ca. 0.5 mg/l after 98 days. The herbicide concentrations in the non-sterile LCM stayed constant for the first four days after loading but by the 6th day had all fallen to C/C0 = 0.8 and by the 14th day to C/C0 = 0. In the sterile LCM they remained constant at C/C0 = 1 throughout the duration of the experiment.

1225 C ~ o Tritium & Herbiei~s

Oxygen ppm

1.2

10

~1[-

(a) "m- ..... • ..... • ............ -m- ................................ • ........

1.0. am

....................

I

0.80.60.40.2m . . . . . . . . . . . . .

ua 20

0.0 0

~

a

-

u

u

40

60

a

II

u

80

100

Time from Loading (days) Oxygen ppm

C/Co Tritium & Herbicides

10

1.2 ~

.

(b)

............. ~ _ . = _ _ = . . . . . . . . . . . . . . . . ~;l

,.0 0.8 0.6 0.4

.... • ....

Tritium

•

MCPA

o

2,4-DP

A

MCPP

0.2 - - - II- - 0.0

Oxygen ppm 0

!

20

4o

60

8'0

1oo

Time from Loading (days)

Figure 9

Degradation

c u r v e s f o r (a) n o n - s t e r i l e a e r o b i c D a n s a n d

(b) sterile aerobic Dansand

LCM,

using groundwater

LCM from

and Bh2.

1226 DISCUSSION The overall general pattern of biodegradation of the three herbicides under different redox conditions is fairly easily discerned from the fate curves, i.e. that under microbially active anaerobic conditions they tend to be persistent, whereas under microbially active aerobic conditions they biodegrade quite rapidly. Thus, in Bhl no herbicide degradation was observed during the 140 day anaerobic period. After 140 days it appears that air entered the microcosm whereupon the herbicides degraded, as did other species known to degrade aerobically, viz. benzene, naphthalene, phenol and o-cresol [Nielsen et al., 1996]. The drop in the Fe(II) concentration after 140 days supports the idea of the introduction of air and it is likely that the fall in concentration corresponded to the formation of insoluble Fe(m) oxyhydroxide flocs. During the period up to 140 days clear evidence of biodegradation had only occurred for one compound, i.e. toluene. Degradation of toluene anaerobically under a variety of redox conditions is fairy well documented [Salanitro, 1993] and of the simple aromatic hydrocarbons it is probably the most susceptible to anaerobic biodegradation. Its maximum rate of disappearance corresponds with the minimum in the Fe(I/) plot (at ca. 35 days). The behaviour of Fe(II) in the microcosms provides useful insights into the processes occurring and it is possible to postulate a series of events. Although there are high concentrations of Fe(II) in the groundwater from Bhl this does not necessarily imply that dissimilatory iron reduction is the principal terminal electron-accepting process for microbial species. Being soluble Fe(II) may well be in transit from an iron-reducing zone elsewhere in the aquifer [Vroblesky & Chapelle, 1994]. In fact during the experiment the outside of the microcosm became coated with black material (found by chemical analysis to contain ferrous sulphide) and this together with the groundwater's high sulphate concentration seemed to indicate sulphate reduction. This process occurring in the microcosm would account for the early loss of Fe(lI) as precipitated ferrous sulphide (from the reaction with sulphide ions generated by sulphate reduction). The Fe(II) concentration minimum then results when the sulphate-reducing bacteria are exhibiting their maximum activity. It is probable, therefore, that toluene was being degraded by sulphate-reducing microorganisms. There are some reported instances of toluene degradation by sulphate-reducing bacteria [Belier et al., 1992]. The fall in Fe(II) concentration was considerably greater than would account for complete mineralisation of the toluene and the abiotically transformed compounds (nitros and tetrachloromethane). The TOC content of the groundwater at Bhl was 26 mg/1 and it may be that this represents the primary carbon source for microbiological activity. As sulphate reduction subsided, indicated by the slowing of the rate of toluene degradation, then the rate of production of Fe(II) started to increase. A rise in the concentration of Fe(II) usually accompanies microbially-mediated dissimilatory Fe(m) reduction [Chapelle, 1993]. Viable communities of sulphate- and iron-reducers are known to coexist in aquifers, though iron-reducers tend to outcompete sulphate-reducers if abundant Fe(III) oxyhydroxides are present in the sediments [Chapelle & Lovley, 1992]. In the BHM there probably was initial suppression of iron reduction because of a lack of bioavailable Fe(m) from the surfaces of the acid-washed Dansand. Once the iron-reducers had adapted to use the more crystalline forms of Fe(III) oxyhydroxides, their population rose with time, the sulphate-reducers were outcompeted and thus the Fe(II) concentration increased. During the period in which iron reduction was assumed to be the dominant microbial

1227 process (35 days to 140 days) no significant degradation of any other added organic species was observed. It was hoped that comparable behaviour would be found with the anaerobic LCM. However, it appeared that attempts to minimise the oxidation of the spiked groundwater were not wholly successful. Thus, the rapid fall in Fe(II) concentration during column loading indicated that some oxidation of the spiked groundwater had occurred (Figure 4). As a result, a relatively high concentration of amorphous Fe(III) oxyhydroxide in suspension would have been pumped into the column. Unlike the BHM, therefore, there was from the outset a readily bioavailable source of Fe(l]/) and consequently, with sulphate reduction inhibited, iron reduction was the predominant microbial process throughout the experiment. As with the BHM, during its iron-reducing phase, no degradation of any added species was noted, though the rise in Fe(II) concentration and microbial numbers, after loading, implied microbial Fe(III) reduction was taking place. It was assumed that the high concentration of indigenous organic matter served as the primary carbon source for such bacterial activity. The initial loss of viable microbial numbers may point to the loss of some obligate anaerobes, e.g. organisms of the GS-15 type [Lovley, 1991].

But the recovery tends to show the presence of aerotolerant, possibly

facultatively anaerobic, iron-reducing species [Lovley, 1991]. In contrast to the lack of biodegradation in the anaerobic microcosms, some of the aerobic microcosms revealed transformation of many compounds including the herbicides. The behaviour of the LCMs, in which aerated groundwater (dO2 ca. 9 mg/1) was used for equilibration and loading, demonstrated just how rapid the biodegradation of herbicides could be under fully aerobic conditions. Thus, after a lag period of about 4 days rapid microbially-mediated degradation of the herbicides began. herbicides were below the detection limit (10 lxg/1).

Ten days later the concentrations of the

This is lower than original, unspiked groundwater

concentration indicating that, either degradation was occurring continuously in this region or that the addition of other organics had lowered the threshold value e.g. by cometabolism. The existence of a threshold value is supported by the presence of 8 ~tg/1Mecoprop at the pumping station.

Either, the aquifer is totally confined

and conditions are anaerobic along the 10 Km pathway or degradation rates slow dramatically as the concentration decreases. The lag phase was shorter than lag phases reported by other workers investigating herbicide degradation. In some laboratory batch experiments adaption periods of between 20 and 110 days were required before MCPP was degraded [Heron & Christensen, 1992], and in others a 7 day lag was followed by complete degradation in 22 to 27 days [Klint et al., 1993]. A field experiment conducted in an aerobic aquifer (Borden) revealed a 42 - 56 days lag phase [Agertved et al., 1992]. However, in all these cases groundwater and sediments had been previously unpolluted by herbicides. The microbial community in the aerobic borehole at Helpston was adapted to MCPP and this probably explains the short lag and degradation periods. Klint et al. [1993] found that once the initial input of MCPP had degraded in their laboratory microcosms, then further inputs were degraded very rapidly without a lag phase. The behaviour of the non-sterile aerobic sand BHM was somewhat puzzling. Initially, it behaved in a similar manner to the non-sterile aerobic LCM.

Thus, profound and rapid biotransformation of the herbicides

occurred. This seemed incompatible with the relatively low (ca. 2 mg/l) dissolved oxygen content of the groundwater. By 21 days the herbicides were at C/C0 = 0.1 but thereafter their concentrations rose steadily.

1228 Similar behaviour was not observed in the non-sterile limestone BHM. Possible explanations were that :1).

Initial biodegradation used up the dissolved oxygen and herbicide concentrations fell. The BHM thus

became anaerobic and the Fe(III) oxyhydroxide associated with the sand began to undergo microbial iron reduction and go into solution as Fe(II). Any herbicides sorbed to the hydrous iron oxides (or sorbed to the biofilm clinging to such) may then have been gradually released back into solution. 2).

During packing a pocket of air may have been trapped at the bottom of the BHM leading to

non-uniform redox conditions along the length of the column. Thus, fully aerobic degradation occurred in the bottom section of the microcosm and the C/C0 of the herbicides rose as samples were obtained from the increasingly oxygen depleted zones further into the column. The former was dismissed after batch sorption experiments on both clean sand and sand that had been equilibrated in-situ, so that a biofilm might develop (i.e. by several weeks immersion in the groundwater at Bh2), failed to reveal any propensity for herbicide sorption by either type of material. When the equilibrated sand samples were examined for biofilm, by optical, scanning electron and epifluorescent microscope techniques, none was apparent [Harrison et al., 1995]. Given the initial similarity to the fully aerobic LCM, the latter was considered to offer the more satisfactory explanation. This view was also supported by the fact that to a lesser extent other compounds prone to aerobic degradation mimicked the behaviour of the herbicides. Thus, initial degradation followed by apparent recovery was exhibited, (in descending order of degree of recovery), by 2,4-dichlorophenol, p-nitrophenol, benzene, toluene, o-xylene, trichloroethylene and o-cresol. Naphthalene and phenol underwent rapid initial degradation to C/C0 = 0, but thereafter showed no recovery. This may indicate that their aerobic biodegradation is particularly easy and proceeds to completion very quickly. In the non-sterile limestone BHM, naphthalene and phenol, behaved in much the same way, though C/Co for naphthalene only fell to about 0.2, but remained steady thereafter. The behaviour of the aerobic non-sterile limestone BHM was less complicated. After loading there was a short period of aerobic biodegradation during which C/C0 for some of the added organics fell slightly. Thus, after 20-30 days benzene, toluene, o-xylene, o-cresol and phenol fell in concentration owing to aerobic degradation. As they fell the Fe(II) concentration went up to about 2 mg/l, then steadied and remained virtually constant for the rest of the experiment. However, microbial counts remained throughout at 103-104 CFU/ml. The lack of a rise in Fe(II) concentration may be due to its use as a reductant for the processes of microbially-assisted reduction that were found to occur for tetrachloromethane, nitrobenzene, and o-nitrophenol.

Such

bioreduction by Fe(II) has been noted elsewhere [Klausen et al., 1995]. These reductive degradations were the only form of biological transformation that was noted after the cessation of the short aerobic phase. Again no noticeable herbicide degradation was found to occur throughout the duration of the experiment.

CONCLUSIONS The findings of our experiments appear to underline the importance of redox conditions on the degradationof phenoxyacid herbicides that find their way into aquifers. The variety of redox and microbiological conditions encountered in the investigations all pointed to the same conclusions. Under aerobic conditions high

1229 concentrations of the three herbicides studied were found to be readily biodegraded. However, there are indications that there are threshold values below which degradation ceases or stops. The shortness of the lag phases and rapidity of transformation were possibly attributable to acclimatisation of the indigenous microbial populations to MCPP. However, under microbially-active anaerobic conditions very little or no degradation, biotic or abiotic, of the three phenoxyacid herbicides occurred. The aerobic experiments tend to support the widely-documented observation that substantial phenoxyacid herbicide degradation may occur in topsoils or other aerobic regions. However, the solubility and lack of sorption of these compounds to geologic media, can permit them to percolate undegraded from aerobic zones into underlying anaerobic aquifers.

They may also enter such aquifers by the leaching of phenoxyacid

herbicides disposed of in older (unlined) landfills. Under these anaerobic conditions there is the potential for their transport over considerable distances. ACKNOWLEDGEMENTS This study was part of a major research programme focussing on the effects of waste disposal on groundwater carried out in collaboration with the Technical University of Denmark (TUD). Funding was provided by the Commission of the European Communities and the UK Department of the Environment. The results will be used in the formulation of Government Policy, but do not necessarily represent that policy. The authors would like to thank their colleagues at TUD for ongoing discussions and advice, also staff at the BGS for analytical support without which this work would not have been possible. This paper is published with the permission of the Director of the British Geological Survey (NERC).

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