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Waste Management 25 (2005) 281–289

Laboratory studies of the remediation of polycyclicaromatic hydrocarbon contaminated soil by in-vessel composting

Blanca Antizar-Ladislao *, Joseph Lopez-Real, Angus J. Beck

Department of Agricultural Sciences, Imperial College London, High Street, Wye campus, Wye, Ashford, Kent TN25 5AH, UK

Accepted 11 January 2005

Abstract

The biodegradation of 16 polycyclic aromatic hydrocarbons (PAHs), listed as priority pollutants by the USEPA, present in a

coal-tar-contaminated soil from a former manufactured gas plant site was investigated using laboratory-scale in-vessel composting

reactors to determine the suitability of this approach as a bioremediation technology. Preliminary investigations were conducted

over 16 weeks to determine the optimum soil composting temperature (38, 55 and 70 �C). Three tests were performed; firstly, soil

was composted with green-waste, with a moisture content of 60%. Secondly, microbial activity was HgCl2-inhibited in the soil green-

waste mixture with a moisture content of 60%, to evaluate abiotic losses, while in the third experiment only soil was incubated at the

three different temperatures. PAHs and microbial populations were monitored. PAHs were lost from all treatments with 38 �C being

the optimum temperature for both PAH removal and microbial activity. Calculated activation energy values (Ea) for total PAHs

suggested that the main loss mechanism in the soil-green waste reactors was biological, whereas in the soil reactors it was chemical.

Total PAH losses in the soil-green waste composting mixtures were by pseudo-first order kinetics at 38 �C (k = 0.013 day�1,

R2 = 0.95), 55 �C (k = 0.010 day�1, R2 = 0.76) and at 70 �C (k = 0.009 day�1, R2 = 0.73).

� 2005 Elsevier Ltd. All rights reserved.

1. Introduction

There are three main reasons for the growth of the

composting industry in the UK: legislation for biode-

gradable municipal solid waste, environmental benefitsand economic benefits. Green-waste comprised the

majority (92% in 1998) of municipal wastes produced

in the United Kingdom. The three main regulatory driv-

ers for composting are the EU landfill directive (EC,

1999), the UK Waste Strategy 2000 (DETR, 2000) and

the EU Animal Byproducts Regulations (EC, 2003).

These have increased interest in composting of garden,

tree, and food-processing organic wastes. Compostingof yard wastes, municipal wastewater sludges, and mu-

0956-053X/$ - see front matter � 2005 Elsevier Ltd. All rights reserved.

doi:10.1016/j.wasman.2005.01.009

* Corresponding author. Tel.: +44 20 759 42779; fax. +44 20 759

42640.

E-mail address: b.antizar@imperial.ac.uk (B. Antizar-Ladislao).

nicipal solid wastes are long established; however, com-

posting of soils contaminated with hazardous materials

is still an emerging ex situ biotreatment technology.

Composting conditions differ from other ex situ soil

treatment systems in that bulking agents are added tothe compost mixture to increase porosity and serve as

sources of easily assimilated carbon for biomass growth

(Haug, 1993). Aerobic metabolism generates heat,

resulting in significant temperature increases that bring

about changes in the microbial population and physiol-

ogy in the compost mixture. The conventional aerobic

compost process passes through four major microbio-

logical phases identified by temperature: mesophilic(30–45 �C), thermophilic (45–75 �C), cooling, and matu-

ration. The greatest microbial diversity has been ob-

served in the mesophilic stage. The thermophilic stage

is characterised by spore-forming bacteria and thermo-

philic fungi. Microbial recolonisation during the cooling

phase is characterised by the appearance of mesophilic

Table 1

Physicochemical properties of the green-waste used

282 B. Antizar-Ladislao et al. / Waste Management 25 (2005) 281–289

fungi whose spores withstand the high temperatures of

the thermophilic stage. In the final compost stage (mat-

uration), most digestible organic matter has been con-

sumed by the microbial population, and the

composted material is considered stable (Epstein, 1997;

Sela et al., 1998).Composting has been demonstrated to be effective in

biodegrading PAHs (McFarland and Qiu, 1995; Potter

et al., 1999; Canet et al., 2001), chlorophenols (Laine

and Jørgensen, 1997), polychlorinated biphenyls (PCBs)

(Block, 1998), explosives (Gray, 1999) and petroleum

hydrocarbons, especially diesel fuel (Namkoong et al.,

2002) at both the laboratory and field scales. It is widely

accepted that temperature is an important environmen-tal variable in composting efficiency (Joshua et al., 1998;

Namkoong et al., 2002). Temperature affects not only

the physiological reaction rates and population dynam-

ics of microbes, but also most of the physicochemical

characteristics of the environment.

Temperature increase within composting materials is

a function of initial temperature, metabolic heat evolu-

tion and heat conservation. Temperatures of compo-sting material below 20 �C have been demonstrated to

significantly slow or stop the composting process (Paul

and Clark, 1996). Temperatures in excess of 60 �C have

also been shown to reduce the activity of the microbial

community, and microbial activity declines when the

thermophilic optimum of microorganisms is exceeded.

If the temperatures reach 82 �C, the microbial commu-

nity is severely inhibited (Paul and Clark, 1996). Mac-Gregor et al. (1981) found that optimum composting

temperatures, based on maximising the decomposition

of raw sewage sludge mixed with woodchips were in

the range of 52–60 �C. However, some researchers have

found that such high temperatures are not required to

produce a high quality product (Miller et al., 1990).

Other studies have indicated that lower temperatures

might allow more microbial activity (Liang et al., 2003).The objectives of this study were to: (i) determine the

potential for losses of the 16 USEPA-listed PAHs from

a coal-tar-contaminated soil during composting, (ii) elu-

cidate the impact of temperature on the (bio)degrada-

tion of these 16 PAHs, (iii) study the rates of

(bio)degradation of 16 PAHs at different temperatures,

and (iv) monitor the changing microbial populations

in relation to temperature.

Green waste Moisture

content (%)

Incinerable

matter (%)

Foodstuff, which contains: carrot (16.7%),

cucumber (16.7%), lettuce (16.7%),

onion (16.7%), potato (16.7%),

tomato (16.7%).

90.6 ± 0.2 99.4 ± 0.0

Sawdust 10.4 ± 0.10 99.7 ± 0.0

Leaves 46.5 ± 4.9 97.3 ± 0.0

Grass 64.3 ± 19.7 97.0 ± 0.1

Wheat straw 9.9 ± 0.5 94.5 ± 0.4

Foodstuff, leaves and wheat straw were blended, grass was cut.

2. Materials and methods

Nine experimental conditions were tested in tripli-

cates using 189 laboratory-scale composting reactors.

The standard composting reactors comprised a soil to

green-waste ratio of 0.6:1 on a dry weight basis. TheHgCl2-inhibited composting reactors comprised a soil

to green-waste ratio of 0.6:1 on a dry weight basis with

2% HgCl2 used as a microbiological inhibitor. The con-

trol reactors consisted of 100% soil. Batches of 63 reac-

tors were placed in three different incubators at a

constant temperature equal to 38, 55 and 70 �C,respectively.

2.1. Contaminated soil

The coal-tar-contaminated soil was obtained from a

manufactured gas plant site commissioned in 1838 at

Clitheroe, Lancashire, United Kingdom. An extensive

description of the site and the procedures for soil sam-

pling and preparation is provided by Birnstingl (1997).

The soil samples were selected and composited from sev-eral areas on site. Stones and oily materials were re-

moved, the soil was then air-dried and homogenised

by passing through a 5-mm sieve followed by a 2-mm

sieve and stored in the laboratory at room temperature.

Before experimentation the soil was diluted by homoge-

nizing with silver sand (sharp fine sand of silvery appear-

ance) (1:1) to provide a more homogeneous distribution

of the coal-tar residue. Soil organic content was4.79 ± 0.16% (wt/dry wt); soil pHw was 7.3 ± 0.1. The

soil was conditioned with green-waste at a ratio of

0.6:1 on a dry weight basis. The green-waste was pre-

pared by mixing foodstuff (mixture of carrots, cucum-

ber, lettuce, onions, potatoes and tomatoes in equal

amounts) (3% dw), sawdust (38% dw), leaves (18%

dw), grass (27% dw) and wheat straw (14% dw)

(Table 1).

2.2. Reactors design

One hundred and eighty nine 200 ml glass compo-

sting reactors were made to provide closely monitored

and controlled conditions (Fig. 1). These fully enclosed

bench-scale reactors each held about 65 g total compost

mixture. The reactor units stood vertically with air, sat-urated with water vapour, flowing continuously up

through the compost mixture. Constant air-flow to the

composting reactors was provided by 100% oil-free dia-

phragm pumps (Model PXW-600-DIOV, VP1, Fisher

Scientific) and vented outdoors. In order to maintain

Fig. 1. Design of laboratory-scale composting reactors.

B. Antizar-Ladislao et al. / Waste Management 25 (2005) 281–289 283

similar air-flow in the 189 reactors, they were separated

in batches of 63 reactors per incubator (and tempera-

ture), 21 standard-composting reactors, 21 HgCl2-inhib-

ited composting reactors, and 21 soil reactors. Air waspumped to an air/water reservoir kept at the same tem-

perature as the reactors (i.e., 38, 55 or 70 �C) where it

was saturated with water. The air/water reservoir had

42 exits, which were connected to each reactor (standard

and HgCl2 reactors). Soil reactors were not aerated, but

open to the aerobic atmosphere.

Compost moisture content was measured weekly to

ensure it was maintained at 60%. The air inlet was bub-bled through a water reservoir to avoid excessive water

evaporation during aeration. The cylindrical reactor de-

sign permitted a better distribution of the air flow inside

the reactors, preventing the creation of anaerobic pock-

ets in the compost mixture. Streams of inlet and exhaust

gas were occasionally monitored for carbon dioxide pro-

duction as evidence of aerobic biodegradation.

2.3. Sample analysis

Destructive sampling (in triplicate) for each treatment

occurred at time 0 and after 7, 21, 35, 54, 66, 102, 111 d for

PAH analyses, and after 21, 54 and 102 d for biomass

analyses. Ash content was determined using a loss-on-

ignition procedure. Triplicate 5 g samples were dried for

24 h at 110 �C (moisture content) and then transferredto a muffle furnace at 550 �C for 12 h to burn the organic

matter. Moisture content was expressed on a wet basis,

defined as the mass of the water in a sample divided by

the total wet mass of the sample (Agnew and Leonard,

2003). Ash content was calculated from the ratio of pre-

and post-ignition sample weights.

2.4. PAH Analysis

PAH extraction from compost mixtures and soil

was by Accelerated Solvent Extraction (ASEe) 200,

with 22 mL stainless steel extraction cells meeting the

requirements for the extraction of PAHs from solid

waste as described in the USEPA Method 3545.

Briefly, glass fibre disks were placed at the outlet

end of the extraction cells and a 7-g sample of com-

post was mixed with 3 g of sodium sulphate and 7 gof Hydromatrixe and introduced into each extraction

cell. Surrogate standards (1-fluoronaphthalene, 2-fluo-

robiphenyl, purity >97%, Greyhound Chromatogra-

phy & Allied Chemicals (UK)) were added to the

cells prior to extraction to monitor PAH losses.

Extraction cells were placed into the auto-sampler tray

with copper turnings to remove sulphur. ASEe 200

conditions for PAH extraction were: 14 MPa(2000 psi), 100 �C, oven heat-up time = 5 min, static

time = 5 min, solvent dichloromethane/acetone (1:1),

(v/v), flush volume = 60% of extraction cell volume,

nitrogen purge = 1 MPa (150 psi) for 60 s.

The extracts were purified on chromatographic col-

umns packed with 1 g of activated-florisil (SiO2,

84.0%; MgO, 15.5%; Na2SO4, 0.5%; 60/100 mesh;

130 �C; 12 h) and 2 g of Na2SO4. In order to removehydrophobic impurities, the columns were washed with

10 ml dichloromethane, then 5 ml of extracts (or more

according to the removal rates) were eluted, and left to

dry for 1 min. The PAHs were then eluted with 10 ml

dichloromethane. Internal standards (naphthalene-d8,

acenaphthene-d10 in a mixture with chrysene-d12, 1,4-

dichlorobenzene-d4, perylene-d12, phenanthrene-d10,

purity >97%, Greyhound Chromatography & AlliedChemicals (UK)) were added to the clean extracts prior

to analysis.

A Hewlett–Packard 6890 series gas chromatograph

with a 7673 series auto-sampler and a 5973 series mass

selective detector was used for the analysis. Data acqui-

sition and processing was achieved using a Hewlett–

Packard MS Chemstation (G1034C Version C.02.00).

The GC inlet was operated in pulsed (0.90 min,30.0 psi) splitless mode at 270 �C with helium as carrier

gas. The injection volume was 1 ll and the inlet purged

at 50 ml min�1 1 min after injection; inlet pressure was

varied by electronic pneumatics control (EPC) to main-

tain a constant column flow of 1 ml min�1. Separation

was achieved using an HP-5MS column (19091S-433

30 m · 0.25 mm · 0.25 lm). The temperature program

comprised 70 �C for 2 min, 10 �C min�1 to 300 �C,which was maintained for 10 min to allow late eluting

peaks to exit the column. The MS transfer line was

280 �C providing conductive heating of the MS source

to about 230 �C. The instrument was tuned using perflu-

orotributylamine. The MS was operated in selective ion

monitoring (SIM) mode. The GC–MS system was cali-

brated prior to the analysis of samples using seven cali-

bration standards. The calibration was frequentlychecked during the analysis of samples by the repeated

analysis of quality control standards. The 16 USEPA

Table 2

Quantification and confirmation ions of 16 USEPA PAHs, internal

standards and surrogates

Compound Quantification ion Confirmation ions

Naphthalene 128 127, 129, 102

Naphthalene-d10 136 137, 134, 108

1-Fluoronaphthalene 146 120, 125

2-Fluorobiphenyl 172 171, 170

Acenaphthylene 152 151, 153, 76

Acenaphthene 154 153, 152

Acenaphthene-d10 164 162, 160, 163

Fluorene 166 139, 165

Phenanthrene 178 165, 163, 82, 176

Anthracene 178 179, 176, 89

Fluoranthene 202 200, 101, 203

Pyrene 202 200, 201, 101, 203

Benzo[a]anthracene 228 226, 229

Chrysene 228 226, 230, 113

Chrysene-d12 240 236, 241

Benzo[b]fluoranthene 252 250, 253, 126

Benzo[k]fluoranthene 252 253, 250, 126

Benzo[a]pyrene 252 207, 253, 250, 126

Indeno[1,2,3-c,d]pyrene 276 277, 279, 138

Dibenzo[a,h]anthracene 278 279, 139, 276

Benzo[g,h,i]perylene 276 138, 137, 277

Table 3

PAH concentrations (mg PAH kg�1 dry soil) in reactors at the

beginning and end of treatment (% removal in parenthesis)

Compound Initial Temperature

38 �C 55 �C 70 �C

111 d 107 d 105 d

Standard composting reactors

2 + 3 rings 32.5 2.7 (91.8%a) 8.4 (72.8%a) 5.9 (81.9%a)

4 rings 46.4 10.4 (77.6%b) 13.2

(71.7%b)

18.4

(60.3%b)

5 + 6 rings 21.4 6.1 (71.4%c) 6.2 (70.9%c) 11.7

(45.1%c)

Total PAHs 100.3 19.2 28.2 36.1

Percent removal 80.9% 71.9% 64.1%

HgCl2-composting reactors

2 + 3 rings 32.5 5.7 (82.4%a) 5.5 (82.9%a) 1.5 (95.5%a)

4 rings 46.4 24.5 (47.3%b) 11.8

(74.7%b)

5.9 (87.3%b)

5 + 6 rings 21.4 6.5 (69.4%c) 5.4 (74.6%c) 2.7 (87.6%c)

Total PAHs 100.3 36.4 22.7 10.0

Percent removal 63.4% 77.3% 90.0%

Soil reactors

2 + 3 rings 32.5 20.0 (38.6%a) 10.7

(67.1%a)

3.7 (88.7%a)

b

284 B. Antizar-Ladislao et al. / Waste Management 25 (2005) 281–289

PAHs, internal standards and surrogates for SIM GC–

MS mode are summarised in Table 2.

2.5. Biomass

Analysis of bacteria, fungi and actinomycetes were by

the dilution and spread-plate method following the

‘‘Standard Methods for the Examination of Water and

Wastewater’’ (APHA-AWWA-WPCF, 1998) with

minor modifications. Briefly, 10 g of the soil green-wastemixture sample were mixed with 90 ml of Ringers� solu-tion and shaken for 10 min. Consecutive 1:10 dilutions

were prepared, starting with 1 ml of sample to produce

eight dilutions of each sample. Then 0.1 ml of each dilu-

tion were spread onto five plates of nutrient agar (with

cycloheximide) for bacteria, five plates of starch casein

(with cycloheximide) for actinomycetes and five plates

of potato dextrose agar (with rose bengal) for fungi.Cycloheximide was used to inhibit the growth of fungi

from the soil, and rose bengal was used to suppress

the growth of bacteria. Samples from the soil green-

waste mixtures treated at 38, 55 and 70 �C were

incubated at 38, 55 and 70 �C, respectively, for 72 h.

Following incubation, plates were counted.

4 rings 46.4 41.2 (11.3% ) 30.6

(34.0%b)

19.1

(58.8%b)

5 + 6 rings 21.4 18.3 (14.3%c) 16.9

(21.0%c)

11.0

(48.6%c)

Total PAHs 100.3 79.5 58.2 33.8

Percent removal 20.8% 42.0% 66.3%

a 2 + 3 rings percent removal.b 4 rings percent removal.c 5 + 6 rings percent removal.

3. Results and discussion

The 16 USEPA-PAHs (total PAHs) under investiga-

tion were grouped as two- and three-ring PAHs (naph-

thalene, acenaphthylene, acenaphthene, fluorene,

anthracene, phenanthrene), four-ring PAHs (fluoranth-

ene, pyrene, benzo[a]anthracene, chrysene) and five-

and six-ring PAHs (benzo[b]fluoranthene, benzo[k]fluo-

ranthene, benzo[a]pyrene, dibenzo[a,h]anthracene,

indeno[1,2,3-c,d]pyrene, benzo[g,h,i]perylene) and thus

defined as small, medium and large PAHs, respectively,

for ease of discussion. The initial total PAH concentra-tion in the investigated soil after dilution with silver

sand (100 mg PAH kg�1 air dried soil) was lower than

those concentrations (about 450 mg PAH kg�1 soil/sed-

iment) reported in a manufactured gas plant site by

Erickson et al. (1993), however, they are above the

Dutch List action level of 40 mg PAH kg�1 air dried soil

and thus they should be treated.

3.1. Removal of PAH

The concentrations of the 16 USEPA-listed priority

pollutant PAHs investigated in the standard reactors be-

fore treatment and after 111, 107 and 105 d at 38, 55 and

70 �C, respectively, (as mg PAH kg�1 dry soil) are pre-

sented (Table 3, Fig. 2(a)). Losses of total PAH were ob-

served during all temperature treatments, although PAH

Time, days

55 0C83 0C 07 0C

(a)

(b)

(c)

moc-dradnatS srotcaer gnitsop

55 0C 07 0C83 0C

lCgH 2 moc- otcaer gnitsop rs

0

20

40

60

80

100

120

0 20 40 60 80 100 120

Con

cent

rati

on, m

g/K

g

Time, days

0

20

40

60

80

100

120

0 20 40 60 80 100 120

Con

cent

rati

on, m

g/K

g

Time, days

0

20

40

60

80

100

120

0 20 40 60 80 100 120

Con

cent

rati

on, m

g/K

g

Time, days

0

20

40

60

80

100

120

0 20 40 60 80 100 120

Con

cent

rati

on, m

g/K

g

Time, days

0

20

40

60

80

100

120

0 20 40 60 80 100 120

Con

cent

rati

on, m

g/K

g

Time, days

0

20

40

60

80

100

120

0 20 40 60 80 100 120

Con

cent

rati

on, m

g/K

g

Time, days

0

20

40

60

80

100

120

0 20 40 60 80 100 120

Con

cent

rati

on, m

g/K

g

Time, days

0

20

40

60

80

100

120

0 20 40 60 80 100 120

Con

cent

rati

on, m

g/K

g

Time, days

0

20

40

60

80

100

120

0 20 40 60 80 100 120

Con

cent

rati

on, m

g/K

g 83 0C 55 0C 07 0C

er lioS ca srot

Fig. 2. Evaluation of temporal concentrations of small (e), medium (h), large (n) and total PAHs at 38, 55 and 70 �C in (a) standard-composting

reactors, (b) HgCl2-composting reactors, and (c) soil reactors. Plots show average values for triplicate reactors.

B. Antizar-Ladislao et al. / Waste Management 25 (2005) 281–289 285

losses decreased with increasing hydrophobicity of the

PAHs. Large PAHs have higher octanol–water partition

coefficients and lower water solubilities than medium

and small PAHs (Antizar-Ladislao et al., 2004), thus

bioavailability (Carriere and Mesania, 1995; Potter

et al., 1999; Lee et al., 2001) and toxicity (Sverdrup

et al., 2002) may have limited their (bio)degradation,

resulting in their persistence. The majority of smallPAHs were removed by the end of the composting treat-

ment resulting in a concentration removal of 91.8% at

38 �C, 72.8% at 55 �C and 81.9% at 70 �C. Medium

and large PAHs were also removed to a great extent at

38 �C, as compared to their removal at 55 and 70 �C(Table 3). Increasing the temperature from 38 to 70 �Cresulted in a significant decrease in total PAHs removal

(P < 0.01), from 80.9% to 64.1%, respectively.

Comparing the final removal of PAHs in the three

different types of composting reactors at 38 �C (Table

3), the highest removal percentage of total PAHs was

observed in the standard composting reactors (80.9%).

The concentration of total PAH in the HgCl2-inhibited

composting reactors remained constant during the first

21 d of treatment at 38 �C (Fig. 2(b)), and then fell, cul-

minating in 63.4% removal of total PAH after 111 ds ofcontinuous composting treatment. In the soil reactors, a

20.8% removal of total PAH occurred over 111 d,

mainly due to the removal of small PAHs. At 55 �C(Fig. 2), the temporal concentration of total PAH

started to decline in the standard and HgCl2-inhibited

composting reactors after 21 d of composting treatment

resulting in similar final removals of total PAH (74.6%

average) in both reactor types after 107 d of continuous

286 B. Antizar-Ladislao et al. / Waste Management 25 (2005) 281–289

composting treatment. In the soil reactors, a 42% final

removal of total PAH occurred, mainly due to the re-

moval of small and medium PAHs (Table 3). At 70 �C(Fig. 2), the temporal concentration of total PAH varied

in the standard and HgCl2-inhibited composting reac-

tors during the length of the experiment resulting in afinal higher removal of total PAH in the HgCl2-compo-

sting reactors (90.0%) than in the standard composting

reactors (64.1%). In the soil reactors a removal of

66.3% occurred (Table 3).

Increases in PAH concentration (mg PAH kg�1 dry

soil) during composting were occasionally observed in

the reactors. The experimental variation of moisture

content or flow rate during the composting treatmentwould potentially affect the biodegradation extent and

rate of PAHs in the composting mixtures, although they

would not explain an increase in PAH concentration.

Thus, an occasional increase in PAH concentration

might be a consequence of a selective biodegradation

of organic matter within the soil to green waste mixture,

where components of the green waste would have de-

graded faster than components in the soil, changingthe ratio of soil to green waste in the mixture and there-

fore in the calculation of the concentration of PAHs in

the mixture.

Removal of PAHs observed in the HgCl2-inhibited

composting reactors may indicate that the biocidal ef-

fects of 2% HgCl2 were reduced over time, thus some

foreign microorganisms may have been able to colonise

the medium again. Difficulties found with the use of achemical inhibitor in this and previous studies (Canet

et al., 2001), suggest that a better option might be the

use of non-amended soil as an abiotic control. Thus, re-

moval of PAHs from the original aged-soil without

green waste, water or air supply amendment at different

temperatures would better represent the abiotic losses in

this type of experiments. In the soil reactors 20.8%,

42.0% and 66.3% removal of total PAH was achievedat 38, 55 and 70 �C, respectively, which clearly showed

a direct temperature influence on the removal of total

PAHs.

In order to predict the relative contributions of chem-

ical and biological processes to the removal of PAHs,

activation energy values (Ea) were calculated from data

obtained in the reactors using the Arrhenius equation,

lnðrÞ ¼ lnðAÞ � ðEa=RT Þ;where r is the removal of PAHs (%),A is an empirical con-

stant,T is temperature (K),R is the universal gas constant

(8.3145 J K�1 mol�1) and Ea is expressed in kJ mol�1.

The percent removal (%) calculated at each temperature

in standard composting reactors and soil reactors (Table3) was used to determine Ea. On the basis of regression of

the percent removal with temperature, an Ea was calcu-

lated for the removal of total PAHs in all reactors. Previ-

ous studies have suggested that Ea values less than

30 kJ mol�1are likely to represent biologicalmechanisms,

whereas values greater than 60 kJ mol�1have been re-

ported for chemical reactions in soil (Taylor-Lovell

et al., 2002). To explain this, it is assumed that catalysed

reactions such as enzyme-mediated biological processes

have a lower activation energy requirement, causing themtobe less responsive to temperature compared to chemical

reactions. The activation energy in this study indicates

that biological mechanisms govern the removal of PAHs

from composting mixtures in the standard-composting

reactors (Ea = �6.43 kJ mol�1,R2 = 0.99) and that chem-

ical reactions lead the mechanisms of removal in the soil

reactors (Ea = 32.25 kJ mol�1, R2 = 0.99).

Additionally, at the highest temperature investigated,most of the microorganisms would be rendered inactive

(Antizar-Ladislao et al., 2004), and thus, the removal of

PAHs would occur mainly due to volatilisation. This

would indicate that the leading mechanism of removal

at 38 �C was biological, whereas at 70 �C it was volatil-

isation (Table 3), and most likely a combination of these

two mechanisms at 55 �C. Other authors have also re-

ported removal of PAHs from contaminated wastesdue to a combination of abiotic and biotic mechanisms

(Civilini, 1994; McFarland and Qiu, 1995). Neverthe-

less, abiotic losses are more important for the small,

more volatile PAHs than for larger PAHs. McFarland

and Qiu (McFarland and Qiu, 1995) reported no loss

of benzo(a)pyrene (large PAH) through volatilisation

or mineralization during composting of soil with corn

cobs at 39 �C, which is consistent with our findings at38 �C. Thus, temperature plays an important role in

the removal of PAHs during composting. In this study

it appears that a temperature of 38 �C enhances the bio-

logical removal of PAHs, which might occur due to a

promotion of the native microbial population and activ-

ity. In addition, higher temperatures may facilitate

desorption (Lee et al., 1998) and volatilisation (Lazzari

et al., 1999) of PAHs. Desorption of PAHs at highertemperatures from the soil-composting matrix may have

increased their availability to the present thermophiles

but also may enhance inhibition of biological activity

as reported elsewhere (Carriere and Mesania, 1995).

3.2. Kinetics of removal

Most of the PAH losses occurred within the first 21 dof treatment, slowing thereafter with little change being

observed by the end of the composting treatments. The

pseudo-first-order kinetic approximation was applied

using the linear integrated form of

lnðC=C0Þ ¼ �k � t;whereC is the concentration at time t,C0 is the concentra-tion at t = t0, k is the first-order constant of removal (ob-

tained by linear regression) and t is time. First-order

kinetic analyses were performed for the standard-compo-

B. Antizar-Ladislao et al. / Waste Management 25 (2005) 281–289 287

sting mixtures (Fig. 3). A good fit was observed at 38 �C(k = 0.013 day�1, R2 = 0.95), 55 �C (k = 0.010 day�1,

R2 = 0.76) and at 70 �C (k = 0.009 day�1, R2 = 0.73) for

the removal of total PAHs. Removal rates of small, med-

ium and large PAHs were also investigated (Table 4),

which indicated that a higher removal rate at 38 �C wasmainly due to the approximately two times faster removal

rate of small PAHs at 38 �C (k = 0.028 day�1, R2 = 0.83)

than at 55 �C (k = 0.011 day�1, R2 = 0.75) or 70 �C(k = 0.012 day�1, R2 = 0.57).

Other investigators have reported better fitting of

pseudo-first order kinetics using two separate regression

analysis in the two apparent phases (Admon et al.,

2001). This two-phase model approach was additionallytested in the standard-composting mixtures (Fig. 3). Dif-

ferences in the removal rate during the first three weeks

k = 0.013 day-1

, R2 = 0.95

-2

-1.5

-1

-0.5

0

0 20 40 60 80 100 120

nl(C

C/0)

k = 0.010 day-1

, R2 = 0.76

-2.0

-1.5

-1.0

-0.5

0.0

0 20 40 60 80 100 120

time, days

time, days(a)

(b)

(c)

nl(C

C/0)

Time, days

k = -0.009 day-1

, R2 = 0.73

-2.0

-1.5

-1.0

-0.5

0.0

0 20 40 60 80 100 120

nl(C

C/0)

Fig. 3. Kinetics of the removal of total PAHs in the standard-composting re

constant (+), and k1 and k2 represent the first (�) and second (�) phase rate

of treatment at 38 �C (k1 = 0.030 day�1, R2 = 0.94),

55 �C (k1 = 0.023 day�1, R2 = 0.79) and 70 �C(k1 = 0.022 day�1, R2 = 0.76) and after the first three

weeks of treatment at 38 �C (k2 = 0.013 day�1,

R2 = 0.98), 55 �C (k2 = 0.004 day�1, R2 = 0.35) and

70 �C (k1 = 0.004 day�1, R2 = 0.34) where found usingthe two-phase model. The model of Admon et al.

(2001) did not improve the fitting of pseudo-first order

kinetics to our experimental results when considering

only the first phase, while the fitting of the second phase

at 55 and 70 �C was very poor. However, the use of their

suggested two-phase model indicated that approxi-

mately 2.4 times higher removal rates might be found

during the first three weeks of treatment as comparedto the use of the one-phase model. The reduction in bio-

degradation over time in the kinetic study can be

k1 = 0.030 day-1

, R2 = 0.94

k2 = 0.013 day-1

, R2 = 0.98

-2.0

-1.5

-1.0

-0.5

0.0

0 20 40 60 80 100 120

nl(C

C/0)

k1 = 0.023 day-1

, R2 = 0.79

k2 = 0.004 day-1

, R2 = 0.35

-2.0

-1.5

-1.0

-0.5

0.0

0 20 40 60 80 100 120

time, days

Time, days

time, days

nl(C

C/0)

k1 = -0.022 day-1

, R2 = 0.76

k2 = -0.004 day-1

, R2 = 0.34

-2.0

-1.5

-1.0

-0.5

0.0

0 20 40 60 80 100 120

nl(C

C/0)

actors at (a) 38 �C, (b) 55 �C and (c) 70 �C. k represents one-phase rate

constants, respectively.

Table 4

Degradation rate constants in the standard-composting reactors at 38,

55 and 70 �C

Compound Temperature

38 �C 55 �C 70 �C

2 + 3 rings 0.028 (0.83) 0.011 (0.75) 0.012 (0.57)

4 rings 0.010 (0.78) 0.009 (0.52) 0.004 (0.15)

5 + 6 rings 0.011 (0.86) 0.012 (0.80) 0.008 (0.68)

Total PAHs 0.013 (0.95) 0.010 (0.76) 0.008 (0.53)

k represents one-phase rate constant, and R2 is the correlation coeffi-

cient obtained for the regression analyses (R2 in parenthesis).

Table 5

Colony forming units in the standard-composting reactors at 38, 55

and 70 �C

Microorganisms 21 d 54 d 111 d

38 �CBacteria 2.9 · 108 3.2 · 108 n.d.

Actinomycete 3.9 · 108 n.d. n.d.

Fungi 5.1 · 107 1.4 · 107 3.1 · 107

21 d 51 d 107 d

55� CBacteria 6.7 · 106 3.9 · 104 n.d.

Actinomycete 1.9 · 106 n.d. n.d.

Fungi 9.1 · 102 n.d. n.d.

21 d 54 d 105 d

70 � CBacteria n.d. n.d. n.d.

Actinomycete n.d. n.d. n.d.

Fungi n.d. n.d. n.d.

n.d., not detected.

Data show average values for triplicate reactors.

288 B. Antizar-Ladislao et al. / Waste Management 25 (2005) 281–289

explained by reduced bioavailability of PAHs due to

immobilisation in micropores or changes in binding

forms (McFarland et al., 1992).

Although data has been analysed using the one-phase

and two-phase models, the variable nature of compost

complicated the fitting of the second phase of the two-

phase model. Thus, the use of the one-phase model is

more appropriate in the present study, and recom-mended to be used to fit short-term experimental data.

First-order kinetics proves convenient since the rate of

degradation is proportional to the amount of substrate

available, allowing a half-life time, to describe the degra-

dation pattern over the entire duration of decay of a gi-

ven substance. For this reason, regulatory agencies often

favor this approach even when more complex mechanis-

tic models fit the data more closely (Wolt et al., 2001).

3.3. Biomass

During composting, the amount of biomass was high-

er in the reactors incubated at 38 �C than at 55 �C, andat 70 �C no biomass was detected using the dilution and

spread plate method (Table 5). Additionally, the greatest

amount of biomass appeared within the first three weeksof composting treatment at 38 �C. Higher biomass pop-

ulation at 38 �C supports our assertion that PAH bio-

degradation was greater at 38 �C than at 55 or 70 �C.No biomass was apparently present at 70 �C using the

dilution and spread plate method, indicating that the re-

moval of PAHs at this temperature was mainly due to

abiotic mechanisms. However, only a small fraction

(possibly <0.1%) of the soil microbial community isamenable to investigation using traditional culturing

techniques using a variety of culture media designed to

maximize the recovery of diverse microbial populations

(van der Merwe et al., 2002). To overcome these

problems, other methods such as phospholipid fatty

acids (PLFA) analysis may prove more appropriate to

study a greater proportion of the soil microbial commu-

nity, and they are currently being applied in the investi-gation of the rapidly changing microbial community in

active composting mixtures (Baath and Anderson,

2003;Ranneklev and Baath, 2003).

4. Conclusions

This study used laboratory-scale in-vessel compo-

sting reactors to investigate the (bio)degradation of 16USEPA-listed PAHs from coal-tar-contaminated soil.

Our findings indicated that in-vessel composting can re-

duce PAH concentration in a contaminated soil, and

thus it might have useful potential as a bioremediation

technology. Optimal removal occurred at 38 �C where

the highest microbial activity was also observed. The

main mechanism of removal of PAHs in the standard

composting reactors at 38 �C was biological, althoughabiotic mechanisms also played a role. Additionally,

the use of the one-phase model is recommended to de-

scribe the degradation pattern of PAHs in short-term

studies. The highest removal rate of total PAHs during

in-vessel composting was observed at 38 �C (k = 0.013

day�1, R2 = 0.95). Future challenges for research on

in-vessel composting of PAH contaminated soils in-

volves understanding how other parameters such asmoisture content or soil to green-waste ratio may also

influence the optimal environmental conditions for

maximum removal. These questions will be addressed

in future experiments.

Acknowledgements

We are grateful to Cleanaway Ltd and London Re-

made for providing support for this study through the

Entrust scheme. We also thank Miss Jennifer Gosling

for the biomass analysis, and Dr. Jeremy Birnstingl for

providing the coal-tar-contaminated soil.

B. Antizar-Ladislao et al. / Waste Management 25 (2005) 281–289 289

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