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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 Moisturecontent (%)
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
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Con
cent
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on, m
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cent
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cent
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cent
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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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