Fs and non-o-PCBs in digested U.K. sewage sludges

Chemosphere,Vol. 30, No. 1, pp. 51-67, 1995

Pergamon 0045 -6535(94)00376-9

Copyright © 1995 Elsevier Science Ltd Printed in Great Britain. All rights reserved 0045-6535/95 $9.50+0.00

PCDD/Fs AND NON-o-PCBs IN DIGESTED U.K. SEWAGE SLUDGES

A. Sewart 1, S. J. Harrad 2, M. S. McLachlan3, S. P. McGrath 4 and K. C. Jones I

1 Institute of Environmental and Biological Sciences, Lancaster University, Lancaster LA1 4YQ, UK. 2 Scientific Analysis Laboratories, Business & Technology Centre, Green Lane, Eccles, Manchester, M30 0RJ, UK. 3 Ecological Chemistry and Geochemistry, University of Bayreuth, 95440 Bayreuth, Germany. 4 Soils and Crop Science Division, Institute of Arable Crops Research, Rothamsted Experimental Station, Harpenden, Herts, AL5 2JQ, UK. (Received Jn Germany 15 August 1994; accepted 27 September 1994)

ABSTRACT Twelve digested sewage sludges from rural and urban waste water treatment works in the north-west of England were analysed for PCDD/Fs and non-o-PCBs. The PCDD/F analysis of eight s a m p l e s was repeated using high-resolution mass spectrometry, which enabled detection of the lower chlorinated congeners and calculation of TE values. ~TEQ values for these eight samples ranged from 19-206 ng/kg with the higher values detected in the samples from urban/industrial areas. Examination of the congener/homologue profiles for the more contaminated samples suggests a major input from the use of pentachlorophenoL Archived sewage sludge samples collected and stored from one sewage treatment works in the south of England between 1942 and 1960 were analyzed to gain some insight into temporal trends and possible variations in source inputs. These provide some evidence of changing sources of PCDD/Fs over time and a decline in YTEQs since the 1950s.

INTRODUCTION Nearly 50% of the 1.22 x 106 t (dry weight) of sewage sludge generated annually in the UK is currently applied to agricultural soil and the banning of its disposal at sea after 1998 under new EC leglisation m a y result in an even higher proportion of sludge being added to agricultural soils in the future. Sewage sludges contain trace amounts of PCDD/Fs and PCBs

51

52 (Nestrick and Lamparski, 1983; Weerasinghe et al., 1985; Broman et al., 1990; McLachlan and Reissinger, 1990; Alcock and Jones, 1993), and it is therefore important that the environmental implications of this disposal method are fully understood. Concerns centre around the possible transfer of these compounds from sludge into the human food chains, principally via grazing livestock (McLachlan et al., 1994; Wild et al., 1994). Attempts have been made to identify and quantify sources of PCDD/Fs in sewage sludges (Horstmann and McLachlan, 1994}. Gihr and co-workers (1991) estimated PCDD/F sources to sewage sludge and made two important observations.

Firstly, known sources of P C D D / F s (e.g.

pentachlorophenol [PCP]) cannot explain the total burden in sewage sludge and secondly, that while PCP-associated PCDD/Fs can account for the presence of several key congeners, the broad and relatively high background contamination observed cannot be wholly attributed to PCP use. They therefore concluded that the most likely source of PCDD/Fs in sludges was atmospheric deposition onto roads etc., followed by runoff. In contrast Horstmann and McLachlan (1994) identified household wastewater as a more important source of PCDD/Fs than runoff, implicating laundry wastewater as a major source in household wastewater. Naf et al. (1990) conducted a PCDD/F mass balance for sewage treatment plants but were unable to account for all the load observed. Oberg et al. (1990) and Oberg and Rappe (1992) have reported the de novo formation of 13C-PCDDs from 13C-PCP in municipal sewage sludge and concluded that biological activity was the cause of this transformation; this may partially explain the discrepancy between known sources and the observed burden in sludge. Gihr et al. (1991) used data from Svenson et al. (1989) to estimate that the annual biogenic formation of XPCDD/F in Germany from chlorophenols and PCP in the presence of hypochlorite was 2.4 kg. It may therefore be appropriate to consider sewage sludge as a primary as well as a secondary source of PCDD/Fs.

This study involved analysis for all seventeen of the 2,3,7,8-substituted PCDD/Fs, the nonortho substituted PCB congeners (numbers 77, 126 and 169), together with total tetra through hepta PCDD/F homologue concentrations in twelve sewage sludges collected from various sewage treatment works (STW) in the north west of England. Sludges from these sites have already been analysed for o-substituted PCBs (Alcock and Jones, 1993), chlorophenols (Wild et al., 1993), chlorobenzenes (Wang and Jones, 1994) and a range of low molecular weight

solvent type compounds (Wilson et al., 1994). Comparison of congener profiles with known sources (i.e. PCP and combustion profiles) enables the importance of various sources to the

53 catchments to be inferred. Archived sewage sludge samples collected and stored from one STW near London between 1942 and 1960 were also analysed retrospectively for all 17 2,3,7,8substituted P C D D / F s and their homologue groups, for comparison with contemporary samples. It is well known that sewage sludge samples present particularly difficult problems for PCDD/F extraction, clean-up and analysis. For this reason we were particularly interested in investigating the feasibility of using a low resolution MSD instrument for their analysis at Lancaster University, so data obtained for PCDD/Fs on this instrument have been compared with that obtained by a well established procedure used at the University of Bayreuth, where quantification is routinely carried out by high resolution MS (Horstmann and McLachlan, 1994). EXPERIMENTAL PROTOCOLS The sampling and analytical protocols employed for the determination of PCDDs, PCDFs and the non-ortho IUPAC congener numbers 77, 126 and 169 in sewage sludge samples are described in this section.

Sampling Contemporary Samples: sewage sludges were collected from 12 STWs in the north west of England in 1992. Details are given in Table 1. These STWs represented a range of catchment sizes from rural/domestic to urban and industrial areas, with the works receiving varying proportions of industrial effluents. Approximately 125,000 tonnes of dry sludge solids are produced each year in the north west region of England from about 630 STWs serving a population of about 9 million people. Currently 3 major disposal routes are employed in the region, disposal to sea (51%), application to agricultural land (23%) and landfill (16%). Sludge samples were air dried and ground before extraction.

Archived Samples: These samples were collected, air dried and kept in glass storage containers for a long term agricultural experiment in southern England by the Rothamsted Experimental Station. The experiment, at the Woburn Market Garden plots, has been used recently to quantify the inputs of heavy metals to agricultural soils in sewage sludge, and to investigate their fate and effects in the soil and crop plants (McGrath, 1987).

54

Table I : Details of sewage treatment works and sludge characteristics Sludge

No

Treatment

STW

Site Type

1 2 3 4 5 6 7 8 9 10 11 12

P F F AP F P AF A A A F P

Urb Urb Rur/Dom Ind/Urb Dom/Rur Ind/Urb Ind/Urb Ind/Urb Urb Ind/Urb Urb Urb

Population Current Equiv. (x 1000)

Disposal Option

Solids Content (%)

Industrial Effluent (%)

89 94 18 ? 22 465 282 491 64 1344 375 143

Land Land Land Sea Land Sea Land Sea Land Sea Land Land

2.1 4.63 3.17 1.73 3.09 2.98 2.89 3.53 4.18 2.61 8.61 8.38

3 9 12 3 1 13 34 14 5 11 24 46

P= Primary sedimentation treatment only A= Activated sludge F= Filter units

Ind=industrial Urb=urban

Dom=domestic

Extraction and Preliminary Purification Ten g dry weight (DW) samples were treated with a mixture of 13C12-PCDDs and 13C12-PCB congeners 77, 126 and 169 and extracted in a soxhlet apparatus for 8 hrs with hexane/acetone (2/3 v/v). Following extraction, crude extracts were shaken with a 50 ml aliquot of 1 M NaOH (aqueous), followed by a 50 ml aqueous wash prior to concentration of the organic layer and purification via a 3-stage column chromatographic procedure.

Column A contained (from the bottom), layers of Na2SO4, silica gel, NaHCO3/Na2SO4 and H 2 S O 4 - i m p r e g n a t e d silica. The column was eluted with hexane and the entire eluate concentrated before application to column B, which was packed with activated (350°C, 4 hrs) alumina. Elution of column B with (a) CH2C12/hexane (1:50 v / v ) and (b) CH2Cl2/hexane (1"1 v / v ) , y i e l d e d two fractions, one containing o-chlorinated PCBs, the other the non-ochlorinated PCBs and PCDD/Fs. This latter fraction was concentrated and separation of nono-PCBs and PCDD/Fs achieved by application to column C, which was packed with Florisil® (1.2% w / w deactivated). Non-o-PCBs were eluted with CH2C12/hexane (1:20 v / v ) and PCDD/Fs with CH2C12. Both fractions were concentrated to incipient dryness and treated

55 with 10 ~l of a dodecane solution of PCB congener 155, used as a recovery determination standard.

Archived sludge samples analysed by HRMS were prepared differently. Toluene soxhlet extraction of 2-3 g (DW) for 48 hrs was followed by purification via a 2-stage column chromatographic procedure. Column A contained, from the bottom, layers of: NaOH/silica, silica gel and H2SO4 impregnated silica. The column was eluted with hexane and concentrated before application to column B, which was packed with activated alumina. Elution with benzene, CH2C12/hexane (1:50 v/v) and CH2Cl2/hexane (1:1 v / v ) yielded the PCDD/Fs in the third fraction. 13C12 labelled 2,3,7,8-TCDD was used as the recovery determination standard on this occasion. Samples were finally concentrated to incipent dryness, prior to addition of 30 ]al toluene before analysis.

Gas Chromatographic-Mass Spectrometric Protocols GC-MSD analyses were conducted on all contemporary samples, using a HP 5970B mass selective detector operated in selected ion mode (unit mass resolution; EI mode, 70 eV) and interfaced to a HP 5890A gas chromatograph fitted with a HP Ultra 2 capillary column (50 m x 0.2 mm id, 0.11 ~m film thickness).

GC temperature conditions for the chromatographic

separation of PCB congeners 77, 126 and 169 were as follows: 190°C for 2 min, then 10°C/min to 210°C hold for 1 min, 5°C min to 280°C and hold for 1 min. Both injector and detector temperatures were 300°C and 2 ~l of sample was injected in splitless mode. For PCDD/Fs, the oven temperature program employed was: 190°C for 2 min, then 10°C/min to 210°C, hold for 1 min, 2°C/min to 230°C, hold for 20 min, 15°C/min to 300°C and hold for 8.33 min. Both injector and detector temperatures were 300°C and 2 Ill of sample was injected in splitless mode.

The archived sludge samples and eight contemporary north-west sludge samples were analysed by HRGC/HRMS on a HP 5890 gas chromatograph fitted with a Restek Rtx 2330 capillary column (60 m x 0.25 mm id, 0.20 ~m film thickness) coupled with a VG Autospec Ultima mass spectrometer at a resolution of 10,000, at the University of Bayreuth. Chromatographic separation of PCDD/Fs was achieved as follows: 130°C for 1 min, then

56 30°C/min to 200°C, 2°C/min to 248°C, 20°C/min to 270°C, held for 11 minutes.

Quantification criteria

Before quantification of any compound, the following criteria had to be met: (a) simultaneous detection of a peak for both ions monitored within the expected retention window for each congener; (b) ion intensity ratios of sample peaks within 20 % of the mean values for calibration standards; (c) signal to noise ratios exceeding 3:1; (d) satisfactory method blanks - one blank was run with every 5 samples and where analyte concentrations in the blank exceeded 5% of that detected in a sample, the blank level was subtracted and (e) satisfactory internal standard recoveries (i.e. within the range 30 - 130%).

Method recoveries Recoveries for each sample were checked by reference to the ratios of 13C labelled internal standards relative to the recovery determination standard. The range (mean) of recoveries for the non-ortho PCBs were 54-98 (68%), 62-98 (79%) and 40-99 (73%) for congeners 77, 126 and 169 respectively. The range (mean) of recoveries for tetra through octa CDD/Fs in the contemporary samples were 31-55 (39%), 40-58 (45%), 61-116 (75%), 68-113 (71%) and 63-118 (85%), respectively.

Comparison between GC-MSD and GC-HRMS data The 8 sludge samples analysed by GC-HRMS were measured twice, once on a DB5 MS column for homologue sums and once on the Restex RTX2330 column for the 2,3,7,8 substituted congeners. A comparison of GC-MSD and HRGC data for 2,3,7,8 substituted PCDD/F levels is presented in Table 2. GC-MSD provided detectable levels only for the hexa through octa CDD/Fs. Hence comparisons can only usefully be made for these constituents. Clement et al. (1992) carried out a comparison of GC/HRMS and GC-LRMS performance between 18 laboratories in an inter-laboratory study of the determination of PCDD/Fs in ambient air. They concluded that analytical systems achieved comparable performance under the following limitations: (1) when PCDD/F levels are relatively high and (2) when comparing the higher chlorinated congeners. As Table 2 shows, the findings of this study with sludge support the conclusions of Clement et al. (1992) in that GC/HRMS and GC-MSD give comparable results for the hepta and octa CDD/Fs only.

57

Table 2 : Comparison of the 2,3,7,8-substituted PCDD/F levels (ng/kg) determined by both GC-MSD a and GC-HRMS b 4

2

Sample Location 7 b a b

Con~eners

a

b

a

123678 HxCDD 123789 HxCDD

23 11

5.3 2.8

65 14

22 8.7

14 ND

1234678 H p C D F

870

790

300

190

1234678 H p C D D

1040

860

4500

OCDF

1980

1540

350

OCDD

13500 11000 30000 23200

9

8 a

b

4.6 2

140 80

81 39

170

156

1120

670

4050

270

340

5730

5540

280

250

240

1100

1470

460

650

59000

63000

10

11

12

Congeners

a

b

a

b

a

b

a

b

123678 HxCDD 123789 HxCDD

40 15

16 5.8

110 55

46 19

25 15

25 11

13 ND

4.7 2.1

1234678 H p C D F

340

220

690

350

510

300

80

56

1234678 H p C D D

480

520

2000

2010

700

730

660

580

OCDF

170

225

420

470

390

530

40

35

OCDD

7100

6500

5500

7500

2300

1820

21000 15400

RESULTS A N D GENERAL D I S C U S S I O N

Contemporary Sludge Samples: The concentrations of the three non-o-substituted PCBs are r e p o r t e d in Table 3. These toxic coplanar PCBs constitute a small percentage of commercial PCB m i x t u r e s (Schulz et al., 1989), a n d hence their environmental occurrence is likely to be w i d e s p r e a d . Site 10 has the highest concentration of non-o-substituted PCBs. Alcock and Jones (1993) r e p o r t e d that s l u d g e from this STW also c o n t a i n e d the h i g h e s t total PCB concentration of the sites studied. Various studies have c o n c l u d e d that the toxic n a t u r e of technical PCB m i x t u r e s is associated w i t h the presence of the n o n - o - s u b s t i t u t e d PCBs. A h l b o r g et al (1994) h a v e p r o p o s e d TEFs for all ortho s u b s t i t u t e d PCBs, t h o u g h the

58

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59 toxicological significance of these congeners is not yet fully recognised. In man, PCB congener 126 is the most toxicologically significant, whilst PCB 77 is more readily metabolised (Duarte-Davidson et al., 1993). The abundance of these 3 non-o-substituted PCBs in sewage sludge relative to XPCB does not vary significantly from ratios in commercial mixtures (Schulz et al., 1989), indicating that these congeners are not subject to selective enrichment in sewage sludge. The concentrations of the tetra through octa C D D / F congeners and homologue groups in the 12 sewage sludges analysed by GC-MSD and the 8 by GC/HRMS are given in Tables 3 and 4, respectively. McLachlan (1992) has discussed the limitations of applying the toxicity equivalent (TEQ) concept to environmental samples but it is commonly done and indeed is used for regulating sewage sludge applications to agricultural land in Germany (Schulz, 1993). TEQs are therefore presented for these samples in Table 4. A general increase in concentration with increasing degree of chlorination can be observed, with OCDD the predominant congener. This is consistent with previously reported data for sewage sludges (see Tables 5 and 6). Concentrations in the rural sludges are also in line with those published by other workers. However, the sludges from industrial areas (samples 4, 8 and 10) have relatively high PCDD/F concentrations. There is a substantial database on P C D D / F concentrations in German sewage sludges. Hagenmaier (1988) reported a wide range of TEQ values in sludges from the Waldshut and Lorach provinces of Germany where P C D D / F levels are unusually high. Samples 4, 8 and 10 are close to the average concentrations reported in that study (see Table 6). Interestingly, the high TEQs in samples 4, 8 and 10 are not due to a general increase in each congener group, but to elevated concentrations of hexa-octa C D D / F s (see Table 4). Concentrations of TCDD/Fs and PeCDD/Fs were similar in all samples. This implies that there is a c o m m o n range of sources, giving a general background level which is supplemented by specific source(s) of the hexa-octa congeners.

60

Table 4 : 2,3,7,8-substituted PCDD/Fs determined in 8 UK sewage sludges (ng/kg) by GC/HRMS Congener

S a m p l e Location 8 9_

2

4

_7

2378-TCDF 2378-TCDD 12378-PeCDF 23478-PeCDF 12378-PeCDD 123478-HxCDF 123678-HxCDF 123789-HxCDF 234678-HxCDF 123478-HxCDD 123678-HxCDD 123789-HxCDD 1234678-HpCDF 1234789-HpCDF 1234678-HpCDD OCDF OCDD *TEQ

36 1.9 9.3 22 1.6 4.5 5.3 ND 23 1.2 5.3 2.8 790 5.4 860 1540 11000 51

30 3.8 7.9 14 2.9 15 19 ND 12 2.2 22 8.7 190 8.8 4050 280 23200 90

35 1.4 12 22 4 13 24 ND 9 2 4.6 2 156 3.4 340 240 650 30

66 2.9 38 56 8.3 59 67 4.7 81 13 81 39 670 40 5540 1470 63000 206

PCB77 PCB126 PCB169 **TEQ

1700 280 ND 29

1700 280 ND 29

1100 220 15 23

1100 ND 15 0.7

10

11

12

22 1 12 28 3.6 31 37 ND 34 4.3 16 5.8 220 8 520 225 6500 47

38 2.1 28 35 4.7 40 36 ND 33 6.6 46 19 350 17 2010 470 15400 85

61 1 13 19 2.9 16 17 ND 14 3.4 25 11 300 8.3 730 530 7500 46

24 2.9 5.8 4.7 1 6.3 5.9 ND 5.4 1 4.7 2.1 56 1.8 580 35 1820 19

1500 90 10 9.9

2700 270 55 29

2400 140 15 15

800 ND ND 0.4

* based on the N A T O / C C M S (1988) ** based on the W H O / I P C S scheme (Ahlborg et al, 1994)

C o m p a r i s o n of congener profiles of hexa and hepta C D D / F for p e n t a c h l o r o p h e n o l (PCP) (Rappe et al., 1989) a n d these sewage sludges clearly suggests PCP as the source of these constituents in s a m p l e s 2, 4, a n d 8. PCP signature congeners are 1,2,4,6,8,9-HxCDF and 1,2,3,4,6,8,9-HpCDF (Harrad et al., 1991). The low ratio of 1,2,3,4,6,7,8-HpCDF to total H p C D F s in these s a m p l e s is a strong indication of PCP contamination.

Arehived Sludge Samples: The total P C D D / F h o m o l o g u e concentrations a n d the 2,3,7,8substituted congeners in the 7 archived sludge samples analysed by G C / H R M S are presented in Table 7. The Y~TEQ v a l u e s in the later samples are higher than those m e a s u r e d in the c o n t e m p o r a r y sludges r e p o r t e d earlier. This m a y reflect a general decline in P C D D / F inputs

6! to the environment, due to tighter controls on organochlorine use and disposal.

There is very limited information regarding PCDD/F levels in archived sewage sludges in the literature. Lamparski et al. (1984) reported PCDD/F levels in two 1933 archived sludge samples and compared them to contemporary levels from the same STWs. Levels were similar to contemporary concentrations in their study, with the only significant elevation over time being for 2,3,7,8 TCDD.

Table 5 : PCDD/F concentrations reported in sewage sludge in some selected studies (ng/kg) Congeners

2378-TCDF Y.TCDFs 2378-TCDD Y.TCDDs 12378-PeCDF 23478-PeCDF YPeCDFs 12378-PeCDD ~.PeCDD 123478-HxCDF 123678-HxCDF 123789-HxCDF 234678-HxCDF ~HxCDFs 123478-HxCDD 123678-HxCDD 123789-HxCDD ,~j-IxCDDs 1234678-HpCDF 1234789-HpCDF YHpCDFs 1234678-HpCDD ~/-IpCDDs OCDF OCDD

Ref a

Ref b

Ref c

Ref d

4.2 23 0.72 31 1 1.6 26 0.56 40 4 1.5 ND(0.5) ND(0.3) 47 1.3 9.3 3 77 56 ND(18) 160 900 1500 150 9100

9.7 67 1.5 77 5.2 7.1 87 6.2 322 12 10 1.7 11 200 4.7 49 13 480 85 1.1 89 2170 4230 5 32900

4.9 37 1 27 4.7 7.3 58 5.2 97 9.9 13 0.6 17 160 5.1 25 12 260 46 0.6 51 680 1300 ND(1) 7500

12 140 1.1 64 8.2 15 350 4.9 84 16 10 3.3 15 180 4.9 31 20 340 110 10 190 910 2000 400 4400

ND =non detect (detection limits) a Rappe et al., 1989 c Naf et al., 1990

b Broman et al., 1990 d McLachlan and Reissinger, 1990

62

Table 6 : Summary of reported PCDD/F levels in sewage sludge from selected studies (ng/kg) Reference Congeners

a

b

_c

d

e

2, 3, 7, 8-TCDD mean range

0.6

1.7 1.3-2.2

-

Total PCDD ( n g / g ) mean range

9.8

29 16-42

38 8-281

70 23-195

5.9 2.5-12.5

0.45

0.52 0.45-0.67

2.9 0.3-19

0.3 0.6-10

0.56 0.28-1

23 -

79 41-133

142 37-776

341 64-1560

28 14-66

b e

B r o m a n et al, 1990 H u t z i n g e r et al, 1992

1 0.2-3.8

Total PCDF (rig/g) mean

range TEQ mean range a

c&d

Rappe et al, 1989 H a g e n m a i e r , 1988

Three distinct p a t t e r n s were noted over time in the W o b u r n samples: Pattern A : Tetra a n d p e n t a C D D / F a n d H x C D F concentrations r e m a i n e d quite constant b e t w e e n 1942-1960. Source correlation indicates atmospheric inputs i.e. c o m b u s t i o n sources, being the major source a n d these have remained fairly constant over the s a m p l i n g period. Pattern B : Hexa- to octa- CDD congeners increased by up to a factor of 200 b e t w e e n 1942 and 1956, followed b y a sharp (60%) decrease in levels over the next 4 years. Pattern C : 1,2,3,4,7,8-HxCDD a n d HpCDF-OCDF congeners increased by u p to a factor of 20 between 1942 to 1958, subsequently decreasing by 40%.

The increases in patterns B a n d C can probably be attributed to PCP contamination; PCP and its d e r i v a t i v e s are w i d e l y u s e d to p r e v e n t f u n g a l / m i c r o b i a l infection in the w o o d a n d textiles i n d u s t r y . A w i d e r a n g e of textiles s o l d in G e r m a n y h a v e been a n a l y s e d b y H o r s t m a n n a n d McLachlan (1994), who noted 2 distinct h o m o l o g u e profiles. One h a d a p a t t e r n w i t h little f u r a n c o n t a m i n a t i o n , the other h a d a h i g h f u r a n content. Such a difference could explain the change in the pattern C contribution. Different sources of PCP

63 m a n u f a c t u r e a n d also differences b e t w e e n the two synthetic p r o d u c t i o n r o u t e s i.e. direct c h l o r i n a t i o n of p h e n o l a n d a l k a l i n e h y d r o l y s i s of h e x a c h l o r o b e n z e n e , c o u l d r e s u l t in v a r i a t i o n s in i n d i v i d u a l P C D D / F

c o n g e n e r c o n t a m i n a t i o n levels. A shift f r o m the

p r e d o m i n a n t i n p u t of one PCP source to another in the W e s t L o n d o n s e w a g e c a t c h m e n t area, for example, m a y have been responsible for the change from 'pattern B to 'pattern C'. The s a m p l e s from early 1940s had a 1,2,3,4,6,7,8-HpCDF to Y H p C D F ratio is a r o u n d 0.5. By 1949 it h a d d e c r e a s e d to 0.36, thereafter to a r o u n d 0.27. A low 1,2,3,4,6,7,8-HpCDF : Y.HpCDF ratio is a strong indication of PCP contamination.

Table 7 : PCDD/F Concentrations in A r c h i v e d Sewage S l u d g e from the Isleworth S.T.W., West London (ng/kg).

Congener

1942

1944

2378-TCDF 30 Y.TCDFs 130 2378-TCDD 1.3 XTCDDs 32 12378-PeCDF 16 23478-PeCDF 11 ~PeCDFs 95 12378-PeCDD 1.7 ~,PeCDD 36 123478-HxCDF 11 123678-HxCDF 6.6 123789-HxCDF ND 234678-HxCDF 7.4 Y.HxCDFs 98 123478-HxCDD 1.1 123678-HxCDD 5.5 123789-HxCDD 2.8 ZI-IxCDDs 73 1234678-HpCDF 63 1234789-HpCDF 4 YHpCDFs 130 1234678-HpCDD 83 XHpCDDs 160 OCDF 89 OCDD 1000 Y.TEQ 18

35 500 2.5 89 19 19 440 6.2 130 28 13 ND 18 480 3.7 23 9.9 300 270 11 540 190 370 200 1500 36

Year of collection 1949 1953 90 450 2.9 71 47 23 340 4.3 92 31 15 ND 14 520 5.3 32 11 340 350 18 980 1000 2000 540 9800 61

21 400 3 170

14 16 230 6.8 120 36 15 ND 14 660 12 95 25 930 410 29 1500 4000 8300 940 45000 127

1956

1958

1960

17 260 4 110 13 15 240 5.6 130 100 20 ND 24 1100 15 240 72 2200 660 76 2400 16000 31000 1200 170000 402

21 270 3.6 110 16 17 270 7.6 170 100 19 ND 20 1300 17 190 53 1700 820 110 3000 7500 15000 1600 85000 229

13 100 2.5 51 11 10 140 5.7 100 46 15 ND 12 590 13 110 29 970 440 57 1700 5200 10000 1100 73000 166

64

Discussion of possible fate on soil Sewage sludge application to agricultural land is currently under scrutiny to assess it's potential to increase human exposure to various organics, including PCDD/Fs and non-oPCBs. There are limited field data on the effects of sludge application on levels of PCDD/Fs and non-o-substituted PCBs in the foodchain. Wild et al. (1994) used literature and pathways analyses to show that human exposure to PCDD/Fs could potentially increase under some realistic sludge application scenarios.

During the last few years, the German Federal Government has taken far reaching legislative action to reduce PCDD/F release into the environment.

No restrictions are

applied to soils containing up to 5 ng ~TEQ/kg in Germany. Certain restrictions apply when the soil TEQ value falls in the range of 5-40 ng/kg, including recommendations to monitor potential increased transfer into the foodchain. Above 40 n g / k g there are restrictions on the uses of the soil for agriculture. In July 1992, a limit value of 100 ng TEQ/kg dry matter was established for P C D D / F in sludges used in German agriculture. This limit was set in conjunction with a sludge application rate limit of 5 tonnes of sludge (DW) per hectare over a 3 year period. Some UK sludges that go to land may clearly be expected to exceed this 100 n g / k g TEQ limit, on the basis of the data presented in Table 4. Typical sludge application rates in the north west of England are 8.5 tonnes (DW) per hectare annually.

Various scenarios can be considered. For illustrative purposes, if a plough depth of 15 cm and a soil bulk density of 1 g / c m 3 is assumed, the addition of 8.5 t s l u d g e / h a / y e a r to land would result in a 'dilution' of sludge-borne PCDD/Fs by a factor of 180. Losses can reasonably be assumed to be negligible within a year (Sewart et al., unpublished data). A typical application of sludge from site 12 (19 ng Y~TEQ/kg) which is routinely applied to agricultural land (see Table 1) would increase the soil concentration by - 0.1 ng T E Q / k g / y e a r . Alternatively, the same rate of sludge from site 9 (47 n g / k g TEQ) would increase the soil TEQ concentration by ~ 0.27 ng/kg/year. Typical UK rural and non-industrial urban soil TEQ values have been estimated at 3.4 and 13 ng YTEQ/kg, respectively, and therefore often exceeds the 5 ng Y.TEQ/kg limit anyway. Assuming zero loss from soils, the limit could be exceeded even in rural areas receiving routine sludge applications for just a few years. Particular attention has been focussed on the application of sewage sludge to agricultural

65 land in the context of P C D D / F additions, given the propensity for the 2,3,7,8-substituted c o m p o u n d s to bioaccumulate into livestock. McLachlan et al. (1994) studied the influence of sludge application on concentrations of P C D D / F s in cow's milk on four dairy farms and measured an increase in milk P C D D / F concentrations from one farm where sludge had been applied. Theoretical calculations point to the potential increase in livestock tissue and milk P C D D / F concentrations under various realistic scenarios through the application of sewage sludge to land (Wild et al., 1994). Further detailed studies are required to assess the potential transfer of P C D D / F s (and other organochlorines) to grazing animals under various agricultural m a n a g e m e n t settings, to enable a fuller evaluation of these issues.

Acknowledgments We are grateful to the UK Ministry of Agriculture Fisheries and Food for funding related studies on trace organic contamination in agricultural soils receiving sewage sludge. The comments expressed here are those of the authors and should not be taken as indicative of or reflecting UK G o v e r n m e n t policy.

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