AtmospkL Emirmmenf Vol. 22. No. 2, pp. 41 I-416. 1988. oou4d981,88 S3.00~0.M)

Printed in Great Britain. ,Q 1988 Pergamon Journals Ltd.

ENVIRONMENTAL LEVELS OF CADMIUM AND LEAD IN THE VICINITY OF A MAJOR REFUSE INCINERATOR

M. HUTTON Monitoring and Assessment Research Centre (MARC), King’s College London,The Octagon Building, 4.5944

Fulham Road, London SW10 OQX, U.K.

A. WADGE*?

Department of Human Environmental Science, King’s College London, Hortensia Road, London SW10 OQX, U.K.

and

P. J. MILLIGAN

Computer Centre, King’s College London, Pulton Place, London SW6 5PR, U.K.

(First received 6 May 1987 and received for publication 14 August 1987)

Abstract-This study examines whether the atmospheric release of Cd and Pb from a refuse incinerator in London has caused contamination of the local environment. Sampling networks were established for street dusts, surface soils, vegetation and total deposit gauges in areas up to 5 km downwind and upwind from the incinerator. Measurements of Cd and Pb indicate there is neither a marked nor an extensive contamination by these metals in the downwind area. However, dust Cd values decreased with distance from the incinerator in this area and Cd deposition rates were higher than in the upwind area. Nevertheless, most Cd values obtained in the downwind area were similar to those previously reported for other parts of London while Pb values were often lower. It is considered that the 100 m stack of the incinerator minimizes the deposition of particu~te emissions in the vicinity of this source. Appreciable Cd ~ont~ination was found within the grounds of the incinerator, the values being abom 4-SO-foId higher than in either the upwind or downwind areas. The extent of Pb contamination in this smali area was more limited, with values being about twice those found in the two study areas. The source of this contamination is ascribed to fugitive releases arising from the storage and transport of the ashes produced by the incinerator.

Key word index: Deposition, cadmium, lead, incinerators, fugitive emissions, dust, contamination of soil, vegetation.

INTRODUCTION

It has been known for many years that certain industrial activities release trace metals to the atmos- phere in sufficient quantities to cause marked accumu- lation in soils and vegetation in the vicinity of the source. Field surveys have shown that this local contamination is most commonly associated with the mining and smelting of non-ferrous metal ores, to- gether with industrial processes which consume these metals (Goodman and Roberts, 1971; Burkitt et al., 1972; Lee, 1972; De Koning, 1974; Waliin, 1976, Johnson et al., 1978).

More recently, attention has been paid to the construction of emission inventories for the release of trace metals and other contaminants to the environ- ment. This approach has revealed that activities other than the production and consumption of non-ferrous

* Present address: Toxicology and Environmental Protection Division, Department of Health and Social Security, Hannibal House, London SE1 6TE, U.K.

t The contents of this article represent the author’s views and in no way commit the Department of Health and Social Security.

metals can also be appreciable sources of environ- mental trace metal release. For example, globa in- ventories indicate that coal combustion is an import- ant atmospheric source of Hg and As (Chadwick et al., 1987; Chilvers and Peterson, 1987), while iron and steel production is a significant source of airborne Pb (Nriagu, 1979). Another high-temperature process, refuse incineration, is a major source of atmospheric Cd release at both the global level (Nriagu, 1979) and also in Europe (Hutton, 1983). Furthermore, a detailed inventory of metal release concluded that refuse incineration is the single largest source of airborne Cd in the U.K., being responsible for about a third of the country’s total emissions from human activities (Hutton and Symon, 1986). Lead emissions from this source are also considerable and the third largest in the U.K.; they are, however, small compared with motor vehicle emissions.

Despite the importance of refuse incineration as a source of Cd release, there have apparently been no investigations into whether the atmospheric emissions from modern refuse incinerators exert any local en- vironmental impact. One in-stack investigation of refuse incinerators in the U.S.A. suggested that Cd

411

412 M. HUTTON er al

emissions from such facilities would be a major source of the metal in the urban atmosphere (Greenberget al., 1978). This~onclusion was, however, based on asimple comparison between the extent of Cd enrichment on stack particulates emitted and the composition of the urban aerosol. Nevertheless, it is interesting to note that the closure of all refuse incinerators in domestic apartments in Rio de Janeiro, Brazil, coincided with a

decline in the ambient air Cd levels in the city (Trindade et ai., 1981). It is unlikely that these inciner- ators employed any form of particulate control.

This paper describes the results of a survey of environmental Cd and Pb levels in the vicinity of a major refuse incinerator in the U.K. A network of total deposit gauges was established around the plant, and sampling surveys were also carried out for street dusts, surface soils and vegetation in the area. Earlier investigations into the concentrations and chemical characteristics of trace elements in the fly ash and suspended particulate matter generated by the plant have been published elsewhere (Wadge ef al., 1986; Wadge and Hutton, 1987).

MATERIALS AND METHODS

The refuse incinerator at Edmonton. North London, was commissioned in 1971 and is one of the lar est in the U.K.. with an annual consumption of about 4 x 10 ! t waste. The five combustion units are equipped with rolling furnace grates and operate at about lOO+C. The flue gases and suspended particulate matter from the combustion units are cooled to about 3OO”C, then passed through electrostatic precipitators before being discharged through a 100 m stack. Furnace ash and precipitator fly ash are stored temporarily on site.

The incinerator is located on the outskirts of the London conurbation in the broad valley of the River Lea (Fig. 1). Major roads are found 2 km to the west (AlO)and 0.5 km to the south (A406) of the plant. The land surrounding the incinerator is open and relatively Rat with no other tail buildings nearby. The area to the west of the plant is urbanized; to the east there are reservoirs. some grazing land and housing.

The London Weather Centre is 11 km southwest of the incinerator and was the nearest site where wind data were available. Annual average frequencies of wind direction for the period 1970-1980 indicated that the incinerator plume is most commonly dispersed into the area between due north (0 f and due east (90’) of the facility. Examination of the same

Average annual wind frequency for the Lovdon Weather Centre I1 970-841

0 SOlIS Et plants

T Total deposit gauges

T .

Fig. 1. Map of the Edmonton area, north London, showing location of the incinerator, the sampling design and annual wind frequency.

Environmental levels of cadmium and lead in the vicinity of a major refuse incinerator 413

data on a monthly basis also revealed that this quadrant would receive the incinerator plume most frequently during the 3 months (September-November) of sampling. This quadrant, the “downwind area”, was selected for sampling as it was the area where Cd and Pb emissions from the plant were likely to be deposited in greatest amounts.

It was desirable to obtain control data from the same part of London and the area upwind of the incinerator was considered most suitable. Data from the Meteorological Office revealed that winds which originate from the sector 120-140” were only recorded about 5% of the time on an annual basis and about 3% of the time during the study period. Thus, the control area was selected in the sector 300-320” from the incinerator.

Sampling, sample preparation and chemical analysis

The sampling design for roadside dusts in the downwind area consisted of seven transects from the stack at 15” intervals from 0 to 90”. Along each transect, I5 sampling points were located at 0.15-5.0 km from the stack. For the control area, sample points were located at the same intervals along two transects, at 300 and 320” from the stack (Fig. 1). It was not possible to obtain dusts where the sampling points coincided with the reservoirs near to the incinerator. Dust samples were also not collected if the sampling point was located on a major road, as proposed by Schwar (1983). In these cases, samples were taken from the nearest adjacent area but at the same distance from the incinerator.

The sampling scheme for plants and soils consisted of three transects in the downwind area (30,60 and 90”) and one in the upwind area (310”). Ten sample points were located at 0.2-5.0 km from the stack (Fig. 1); collections of dusts, plants and soil samples were made in August 1985. If any of the sampling points were unsuitable, then the nearest convenient site at the same distance from the stack was selected.

Five soil cores (0-20cm) were combined to produce a composite sample at each site. These were dried at 100°C for 24 h, ground, sieved and the c 2 mm fraction retained for metal analysis. Above-ground grass samples were collected and precautions were taken to avoid condonation with soil and dust. After washing each plant sample several times in double-distilled water (DDW) to remove surface adhering material. the samples were dried at 100°C for 24 h and finely ground prior to metal analysis.

Samples (1 g) of dust, soil and plant material were extracted in concentrated HNO.+ Cadmium and Pb concentrations were deter~n~ by flame or gameless AAS using a Varian 1475 spectrophotometer and GTA-95 graphite furnace with background correction and the double beam mode. For flameless work, absorbance calibration was achieved by the method of multiple standard additions to samples. Further analytical details, including the furnace programmes used, are given elsewhere (Wadge, 1985).

Metal deposition rates were measured using conventional total deposit gauges (Duggan and Burton, 1983). To prevent contamination, deposit bottles were soaked in 10% HNOJ for 24 h prior to use. Five sampling points were selected at 0.2-5.0 km from the stack along two transects (30and 60”) in the downwind area and one in the upwind area (310”). It was not always possible toconform to this strategy and the actual locations of the sampling sites are shown in Fig. 1. The gauges were positioned 1.5 m above ground and left for 27-28 days over three sampling periods during September to November 1985. Aftercoll~tion, thecontents of each total deposit bottle were transferred to a conical flask and the empty bottle was rinsed twice with I y0 HNOa and the washings poured into the conical flask. The contents of the Bask were heated to near-dryness. After adding 5 ml concentrated HNOJ, the flask contents were again heated to near-dryness. The sample digest was transferred quantitatively with DDW and allowed to stand prior to analysis.

The funnel of the deposit gauge was wiped with two filter

papers moistened with DDW to collect any adhering par- ticulate matter. These were digested with a 5: 1 mixture of concentrated HNO, and HCIO,. The nylon gauze covering the funnel was also digested in the same acid mixture. The Cd and Pb concentrations of each of the three sample digests weredetermined by AAS. The total deposition rates were then calculated by adding these data together and taking into account the funnel area and number of days the gauges had been in the field.

Statistical analysis

Two-way analysis ofvariance was used to examine whether the metal data obtained in this study showed any relationship with distance from the stack or if differences existed in the metal values between the downwind and upwind areas. In the analysis of the dust, soil and plant data, results obtained within the incinerator boundary (i.e. Q 0.2 km from the stack) were omitted. Generally, these values were elevated and considered to result from fugitive releases rather than from stack emissions. The anomalous elevated values obtained elsewhere in the study areas were also omitted using the method described by Velleman and Hoaglin (1981): for each of two areas, the Hinges (essentially Quartiles) were com- puted and the H-spread (Upper Hinge-Lower Hinge) was calculated. Those values greater than (Upper Hinge) f 3 x (H-spread) were taken as outliers and omitted from the calculations. To preserve the symmetry of the ANOVA design, the “missing-value” technique was used to replace the excluded outlier observations with calculated estimates (Cochran and Cox, 1957). This is an acceptable approach provided the number of outliers are small, as in this case.

RESULTS

Table 1 summarizes the metal data obtained in the downwind and upwind areas of the study. The values found in the immediate vicinity of the incinerator are also shown, together with concentrations considered to be representative of urban locations in the U.K. For both Cd and Pb in dusts, there was no evidence of an overall difference between the upwind and downwind areas. Dust Cd values in the downwind but not the upwind area showed a significant (P c 0.001) dif- ference with distance (Fig. 2). No distance effect was found for Pb in dusts in either the downwind or upwind areas. Cadmium data for soils and plants showed no signi~cant differences between and within the two areas. Plant and soil Pb levels showed significant (P < 0.01) differences between the two areas. However, there was no consistency as plant Pb values were higher in the downwind area while soil levels were higher in the upwind area. In both areas there were significant (P < 0.01) differences in soil Pb with distance from the incinerator. Analysis of the deposition data revealed that the only significant effect was that Cd values differed (P < 0.05) between the two areas, with levels in the downwind area being greater.

DISCUSSION

The results obtained in this study indicate that there is neither a marked nor extensive Cd and Pb contami- nation downwind of the Edmonton Incinerator. There is some evidence for a moderate decrease in dust Cd

Tab

le

1. L

ead

and

cadm

ium

val

ues

(pg

g ’ d

ry w

eigh

t ex

cept

as

indi

cate

d) i

n th

e st

udy

area

saro

und

the

Edm

onto

n in

cine

rato

r co

mpa

red

with

dat

a pr

evio

usly

re

port

ed

in o

ther

pa

rts

of L

ondo

n

Are

a ar

ound

in

cine

rato

r D

ata

from

V

icin

ity

Dow

nwin

d ar

ea

Upw

ind

area

ot

her

Lon

don

site

s M

ean

Ran

ge

(N)

Mea

n R

ange

W

) M

ean

Ran

ge

(IV

) M

ean

Ran

ge

(N)

~--

Stre

et d

usts

C

d Pb

Soils

C

d Pb

Gra

ss

Cd

Pb

Dep

ositi

on

Cd$

ra

te&

Pb

f

19.6

1.

5-56

(9

) 4.

5 28

34

6365

170

(9)

814

35.8

5.

2-12

6 (4

) 2.

5 53

6 36

6612

(4

) 13

3

26

0.35

-54

(4)

0.63

17

7.

9-29

(4

) 14

18

40

9

0.50

-13.

5 (9

6)

loo-

3330

(9

5)

0.86

4.5

(27)

51

-269

(2

6)

0.26

1.2

(27)

4.

626

(27)

2.

869

(27)

11

7-77

6 (2

7)

-Not

de

term

ined

. *

No

suita

ble

data

ava

ilabl

e.

tTot

al

depo

sitio

n in

gha

-’

a-‘.

1

Dat

a as

sign

ed

to d

ownw

ind

or u

pwin

d ar

ea.

“Hut

ton

(197

9). R

ange

of

six

mea

ns f

rom

six

mid

to

inne

r ur

ban

site

s bC

ulba

rd

et a

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983)

. Sev

en o

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and

inn

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rban

site

s.

c Dug

gan

and

Bur

ton

(198

3). F

ive

oute

r an

d in

ner

urba

n si

tes.

4.0

921 2.

8 30

3 0.46

6.

8

6.5

345

1.tk

lO.O

(2

7)

3.1-

7.5

(6W

41

6248

0 (2

7)

- 96

7-21

21

(6O

)a

1.84

.0

(7)

1.6

< I

-40

(579

)b

96-7

29

(9)

1086

<

l-1

3680

(5

79)b

0.18

-1.0

(8

) *

2.8-

7.8

(9)

* 1.

4-16

(1

2)

8.8

4 61

2 (4

7F

153-

641

(12)

96

0 60

0-18

00

(47)

c

Environmental levels of cadmium and lead in the vicinity of a major refuse incinerator 415

, 2 3 4 5

Distance from incinwator (km)

Fig. 2. Cadmium in street dusts as a function of distance from the incinerator in the area downwind of the plant. Each point is the mean of seven samples; the bar represents + 1 SD.

levels with distance from the incinerator in the down- wind area and Cd deposition rates were higher. For Pb, the data are less convincing and even though plant Pb values were higher in the downwind area, soil Pb levels were lower.

Data obtained from within the grounds (G 0.2 km from the stack) of the incinerator suggest that ap- preciable Cd and Pb contamination exists in this limited area. This effect was most noticeable for Cd, with concentration increases over the two study areas being about 4-fold for dusts, 1Zfold for soils and 50- fold for vegetation. In the latter case, two markedly elevated plant concentrations (47 and 54 fig g-’ Cd) were found. The increases in Pb concentrations in the vicinity of the incinerator were more modest, at around twice those found in the do~w~d and upwind areas.

It is considered that concretion in the grounds of the incinerator does not result from stack emissions but rather from fugitive releases arising from the storage, handling and transport of ashes produced by the incinerator. Certainly the fly ash generated by refuse incineration contains elevated concentrations of metals, and samples from the Edmonton incinerator have previously been shown to contain about 500 ggg- ’ Cd and 10,000 pgg- ’ Pb (Wadge er al., 1986). Surprisingly little attention has been paid to the importance of fugitive emissions from other point sources of metal discharge. In the case of a Pb/Zn smelter in the U.K., it has been suggested that fugitive emissions of coarse particles are the main source of airborne Cd and Pb within the works boundary (Coy, 1984).

Previously published Cd and Pb data from other parts of London may be compared with the findings of the present study. Street dust cd and Pb levels from the downwind and upwind areas are broadly similar to

those reported from other London locations (Harrison, 1979; Hutton, 1979). Culbard et al. (1983), however, obtained average dust Cd levels that were about 3 times as great as those found in the Edmonton study areas. Dusts collected from the grounds of the incinerator contained elevated Cd and Pb levels com- pared with other London locations.

Little information is available on the metal content of soils in London. The data obtained by Culbard et al.

(1983) and summarized in Table 1 indicate that the soils from the open ground of Edmonton contained similar Cd but considerably lower Pb concentrations than were found in garden soils from other London areas. This disparity suggests that garden soils may have received lead inputs additional to the atmospheric fallout of motor vehicle emissions. Davies (1978) found an age-related increase in the Pb content of U.K. garden soils in both rural and urban areas and suggested this cumulation was a “habi~tion effect”.

There are apparently no pub~shed data on metal levels in grass from London or other urban areas; most work of this nature has been conducted on roadside areas and around point sources of emission. One U.K. study found that grass growing close to a motorway contained higher Pb concentrations than were found in the Edmonton samples (Crumpet al., 1980). Pasture grass contained elevated Cd and Pb levels at sites downwind of a U.K. Pb/Zn smelter, with values of 50 and 225 pig g- ’ , respectively, at about 0.3 km from the facility (Burkitt et at., 1972). The Cd value is similar to some of the higher concentrations in grass samples from the grounds of the incinerator but the smelter Pb level is much greater. Even at a distance of about 5 km from the smelter, Cd and Pb values remained elevated, with levels of 9.5 and 160 pg g- I. The corresponding values 5 km downwind of the incinerator were 0.63 pg g- 1 Cd and 9.9 pg g- ’ Pb.

Examination of the total deposition data for London reported by Duggan and Burton (1983) indicates that rates of Pb deposition in Edmonton are lower than those found in the centre of the city. In the case of Cd, deposition rates in the downwind area are moderately higher than in other parts of London, while values from the upwind area are similar.

This study has shown that the atmospheric release of Cd and Pb from the Edmonton Incinerator has not caused a marked contamination downwind of the plant. The urban location of the incinerator caused difficulties when examining distance effects from the facility. This is because of the characte~stically high b~kground levets of metals present in urban areas, particularly Pb. Nevertheless, the estimated annual Cd and Pb emissions from the Edmonton Incinerator at 0.5 and 10 t (Wadge and Hutton, 1987) are not insubstantial. In comparison, the U.K. Pb/Zn smelter referred to earlier is annually emitting an estimated 3.5 t Cd and 170 t Pb (Hutton and Symon, 1986) and the contamination at this site is extensive. It is considered that the 100 m stack of the incinerator is responsible for the wide dispersion and dilution of the

416 M. HUTTON et al

plume and this has minimized the deposition of Greenberg R. R., Zoller W. H. and Gordon G. E. (1978) particulate emissions in the vicinity of the source. Composition and size distributions of particles released in

refuse incineration. Enl;iron. Sci. Technol. 12. 566-573.

Acknowledgements-This study was funded by the Scientific Branch of the Greater London Council. We are grateful to the staff of the Edmonton Incinerator for their co-operation in this project. Mr A. Cullen of Human Environmental Science provided technical assistance during the study. Dr M. S. Johnson and Dr M. H. Martin commented on earher drafts of the paper.

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